Handbook of Pesticide Toxicology, Two-Volume Set: Principles and Agents: Pt. 1 & 2

Handbook of Pesticide Toxicology Second Edition VOLUME 1 PRINCIPLES EDITED BY Robert I. Krieger U~iW!" iIJ o/CQlijom...

Author: Robert Krieger

127 downloads 1749 Views 201MB Size Report

This content was uploaded by our users and we assume good faith they have the permission to share this book. If you own the copyright to this book and it is wrongfully on our website, we offer a simple DMCA procedure to remove your content from our site. Start by pressing the button below!
Report copyright / DMCA form

DOWNLOAD PDF

Handbook of Pesticide Toxicology Second Edition VOLUME 1 PRINCIPLES

EDITED BY

Robert I. Krieger U~iW!" iIJ

o/CQlijomia, Ri",,.iII.

ACADEMIC PRESS San Diego San Frnndoco N<w York Il 200' .. ACADF.\!IC "'RiS All "-"" "",,,,,,,-

,", ,..., .(,", , """*"'" - .......- ,. """""'"...... ...,. to"'" .... ..,.- ,~

.

... ""

- " " . " , ,.."" ~ ,,........ "",,., pt ........... - - .... .....

..,.--"" pot» ......

00's......."of""' ..... _ ......

" " - ' " fo< .,......,.,.., ...... """, of L>opo.--. _ _ 1 H..-, """". Jl

00~"""',.9l10 1 ....,9S ,

,ISA

)am•••,... Rood. t.o.>.Ioo NW, nw.~ •

*"""""""""".-

.. ,~.I"'.•.

~ o/~ .. C...""C,""

S _ !look S,"""", "-, .... 2>.'00-00'811"

III

TiU; ..

U~""FJ)

'"

$"fA"I1iS Of' ~ "f.• 1CA

...

""

•

I

,

"

,

•

,

!

I

C ;.pvrlghted material

Contents of the Handbook \ '01 .... 1

Dose. TIme. and Olhiolosit I), I'I>1rdW!. l'lf~i"'""' , Chi""", T~, _ M=-" Ywi>ida, and M"""""""", UeloiJa

30 Mod<m A!?',..,,,,,, '" "ooIy,;, of Pe5tidd, R"KI ... in Foods.oo ,tie Em';"""""", LW O. R"'PdotJ n.o..... CIlill

671

31 R;,J; ..........."""" aOO Ri'"

Monaf: I"HLIl T';',;o< n,,_idc.< -lam""

~~::,i•. """ Tho:nnoqu.. u..~nJ~ .

IOIJ

1"9 Lod. aNI Martu. f Wilt,

71 Diqu>
of DNA H iooo)~ , hIOis:

"" c:.oo..:ll'ulpI. O·Ma''''

_

~f".,..,

no-- H. _

0/1,.,.•.

.un.r.... J. 1IJbi/4!Md _>TJ _

'"' Md"JI_

lohibiom of Ar.-io: Acid B....,.........

(.Iilt ~'. F.J~. ~ ..-I

8' D i o I k y _ IEBoc;.) s.- H. ... ~ Gail -,,,,,,. ~ Il.oO.

&6 """'rhi... I( F. """1 .od .t.

~ ~

J7 M

Figure 1.7 Bladder neoplasms in dead, moribund, and sacrificed mice fed 2-AAF continuously: (e) 33 months; (x) 24 months; (0) 18 months. The graph is based on data by Farmer et al. (1980).

Figure 1.8 Liver neoplasms in dead, moribund, and sacrificed mice fed 2-AAF continuously: (e) 33 months; (x) 24 months; (0) 18 months. The graph IS based on data presented by Farmer et al. (1980).

1.3 Quantitation of Dosage-Response Relationships

question of whether the values observed above the noise levels determined by the controls were consistent with a cumulative lognormal distribution. The report of the workshop held in September 1981 (Hart et aI., 1983) was apparently concerned mainly with integrating time with dosage response. It was concluded (Hughes et aI., 1983) that "the ED 01 Study demonstrates the observed risk is more adequately expressed in a time and dose continuum rather than simply as dose." The conclusion undoubtedly is correct and was supported by very sophisticated calculations. However, the question of whether the observed increase in incidence above control values was cumulative lognormal in distribution was neglected. The conclusion was that "even with a study as large as the ED 01 study, statistical uncertainty makes it impossible to establish the true shape of the dose-response curve at low tumor rates. Neither can such studies prove or disprove the existence of thresholds." The ED 01 study, also called the megamouse experiment, is indeed very important for toxicology. It provides several lessons beyond those identified by Hayes (1991). The ED 01 study was designed with a lack of kinetic considerations but the initiators were lucky that the dynamic half-life of 2-AAF-induced damage is such that feeding it in the diet provided a steady state of injury (Rozman et aI., 1996). It was also fortuitous that there was not enough toxicity as the experiment progressed and therefore the 24- and 33-month sacrifice schedules were added. Otherwise a reconstruction of the liver dose response from the time response would not be possible. There was much greater emphasis on statistical considerations than on toxicological theory in the design of this very large experiment. However, it is remarkable how little attention was paid to Druckrey's c x t studies on cancer (Druckrey et aI., 1963; Druckrey et aI., 1967). The interpretation of the study was also driven by statistics and not by the science of toxicology. Revisiting the megamouse study revealed that the occurrence of both bladder cancer and liver cancer was highly consistent with the thesis of Druckrey that c x t = k (Rozman et aI., 1996). For the controversy about linearity or nonlinearity of the dose response there is a straightforward biological explanation. 2-AAF is a more potent bladder carcinogen and a less potent liver carcinogen. Therefore, the ED 01 study generated a fairly complete dose response with a shallow and a steep part in terms of bladder cancer with an identifiable threshold. Both the dose response and the time response for liver cancer occur to the right of the corresponding bladder cancer response, mainly toward the end of the animal's life span. Therefore, in terms of liver cancer only the shallow part of the dose response is documented by data, because the steep part of it was prevented from developing by the natural life span of the mice. Thus, the low dose linearity is an illusion, which is inconsistent with the theory of toxicology that under conditions of toxicokinetic or toxicodynamic steady state at constant time all dose responses must have also a steep part of the slope. However, if the time response of a particular effect is truncated by the life span of the species then it is impossible to establish a complete dose response for that particular end point of toxicity. Nevertheless, it

17

is possible to construct a hypothetical dose-response curve beyond the natural life span of a species using the c x t = k x W conversion. Such a hypothetical dose response for 2-AAF becomes also very steep for liver cancer beyond the natural life span of mice. It is important to note that this biological interpretation of the ED 01 study is entirely consistent with statistical approaches as exemplified by the Hartley-Sielken model using both dose and time as variables of toxicity (Hartley and Sielken, 1977a, b), but is completely at odds with currently used linear extrapolation models using only dose as a variable of toxicity. Dosage Response to Radiation At least one study involving a very large number of animals exposed to varying doses of gamma radiation is available (Ullrich and Storer, 1979a, b, c). Including controls, there were 17,587 mice, of which 15,558 were female. All graphs were on plain paper, with incidence of tumors plotted against dosage in rads. No attempt was made to explore the logprobit relationship, an omission that, because Gaddum's famous paper was published in 1933, appears unjustified. Because Ullrich and Storer presented raw data, it is possible to explore how their results fit the cumulative lognormal concept. The curves for reticular cell neoplasms in female mice are shown in Fig. 1.9. Here again, the results for thymic lymphoma and for myeloid leukemia are entirely consistent with a linear logprobit relationship that intersects the control level. However, the graph also indicates that, at least within the range of 25-150 rads, increasing doses of radiation caused a decrease in the incidence of reticular cell sarcoma with the result that the total number of reticular cell neoplasms did not begin to increase until the dosage exceeded 100 rads. The authors discussed this phenomenon but failed to consider whether the decrease of one kind of tumor and the increase in other kinds of tumors were independent phenomena or whether the reticular cells that otherwise would become sarcomas were somehow 99.99

..-..,.......,""TT.....--,.-.-TTTrm-~.....-r"TTTm

99.9 99

95

~ 80

-zf

50 J - - - -

20

5

1----.............._... _.._.-4............_

....................,

0.1 O~,I~~~~w,O--~~wIOLO~~~~wO

Dosage (rails)

Figure 1.9 Incidence of thymic lymphoma (0), myeloid leukemia (600

Increased storage; no clinical effect

Hayes et al. (1956)

Rat (M, F)

0.24

161

Histopathological changes of the liver

Laug et al. (1950)

Human

2

56

No effect

Stein et al. (1965)

Rat (M, F)

4.87

750

No effect level

Lehman (1965)

Human

0.05

24

15% reduction of plasma ChE only

Moeller and Rider (1962b)

Rat (F)

0.05

112

No significant depressions of ChE

Barnes and Denz (1954)

Rat (F)

0.14

112

30% inhibition of ChE

Barnes and Denz (1954)

Rat (F)

0.24

66

60% reduction of plasma; 40% reduction of

T. B. Gaines, unpublished

RBCChE

Dimefox

Reference

Dog (M,F)

0.025

Dog (M,F) Human

results (1962) Frawley and Fuyat (1957)

168

No significant depression of ChE

0.047

168

Significant depression of ChE

Frawley and Fuyat (1957)

0.002

70

No effect on ChE

Edson (1964)

Human

0.0034

70

25% reduction of whole blood ChE

Edson (1964)

Rat (F)

0.024

28

About 50% inhibition of RBC ChE;

Edson (1964)

Rat (F)

0.095

28

75% reduction of RBC ChE; 25% reduction of

no effect on plasma ChE Edson (1964)

plasmaChE Rat (F)

0.475

28

Almost complete inhibition of RBC ChE;

Edson (1964)

75% reduction of plasma ChE Dioxathion

Human

0.05-0.075

59

No inhibition of RBC and plasma ChE

Frawley et al. (1963)

Human

0.15

28

Slight inhibition of plasma ChE; no effect on


RBCChE Rat (M, F)

0.22

91

No significant effect on ChE


Rat (M, F)

0.78

91

Significant reduction of RBC and plasma ChE


Dog (M, F)

0.25

12

Marked effect on plasma ChE; no effect on


RBCChE Dog (M, F)

0.8

12

Marked effect on plasma ChE; no effect on


RBCChE Malathion

Human

0.34

56

Maximal reduction of 25% plasma and RBC ChE

Moeller and Rider (1962a)

Rat (F)

3.2

90

29% reduction in RBC and no reduction of


plasma ChE on 30th day;

results (1968)

recovery by 90th day Methyl parathion Paration

Rat (M)

4.5

730

10-30% inhibition of plasma and RBC ChE

Hazleton and Holland (1953)

Human

0.1

24

15% reduction of plasma ChE only

Moeller and Rider (1962b)

Dog (M, F)

0.94

84

Significant depression of plasma and RBC ChE

Williams et al. (1959)

Human

0.1

42

33% reduction of whole blood ChE;

Edson (1964)

16% inhibition of RBC ChE; 37% inhibition of plasma ChE Rat (F)

0.07

90

No effect


Rat (F)

0.26

84

80% reduction of RBC ChE; slight

Edson (1964)

Rat (F)

0.35

90

37% reduction of plasma and 44% reduction of

results (1968) inhibition of plasma ChE RBCChE Dog

0.047

168

60% inhibition of plasma ChE


results (1968) Frawley and Fuyat (1957) Edson (1964)

Pig

4.0

49

80% inhibition of RBC ChE; no inhibition of plasma ChE

(continues)

60

CHAPTER 1 Dose, Time, and Other Factors Influencing Toxicity

Table 1.9 (continued) Species

Dosage

Duration

Compound

(sex)

(mg/kg/day)

(days)

Schradan

Human

0.014

44

25% reduction of blood ChE

Edson (1964)

Human

0.06

60

77% inhibition of RBCD ChE; 50% inhibition of

Edson (1964)

Rat (M, F)

0.045

112

Substantial reduction of ChE; no effect on

Results b

Reference

plasma ChE Edson (1964)

plasma ChE Rat (M)

0.22

14-85

Complete inhibition of RBC ChE

Edson (1964)

pig (F)

0.1

102

55% inhibition of RBC ChE; slight reduction of

Edson (1964)

plasma ChE Arsenic trioxide

Warfarin

Human

0.44

Frequent mild poisoning

Sollmann (1957)

Sheep

10

Tolerated without symptons

Reeves (1925)

Horse

4.7

Tolerated without symptons

Reeves (1925)

Human

0.14

Indefinite

Maintenance therapeutic dose

Friedman (1959)

Human

0.29-1.45

IS

Hemorrhage in 12 people (4-70 yr)

Lange and Terveer (1954)

Human

1.7

6

Hemorrhage in 22 yr man followed

followed by recovery Holmes and Love (1952)

by recovery

2,4-D

Human

0.83-2.06

IS

Fatal to boy (19 yr) and girl (3 yr)

Lange and Terveer (1954)

Rat (M, F)

0.08

40

Killed 5 of 10

Hayes and Gaines (1959)

Rat (M, F)

0.39

IS

Killed 10 of 10 rats

Hayes and Gaines (1959)

Human

14-37

18

Tolerated

Seabury (1963)

Human

66

I

Coma, hyporeflexia, incontinence

Seabury (1963)

Rat (F)

IS

112

Tolerated

Hill and Carlisle (1947)

Dog

9

84

Tolerated

Drill and Hiratzka (1953)

aFrom Hayes (l967a), by permission of the Royal Society, London. All doses are oral unless otherwise noted.

bRBC, red blood cells; ChE, cholinesterase.

to another and must be taken into account in selecting the duration of chronic studies. Sprague-Dawley rats are characterized by a fairly high incidence of mammary tumors, and the life span of the male rat is commonly limited by the background incidence of renal disease. The F344 rat, though widely used in carcinogenesis studies, has an extremely high occurrence of interstitial cell tumors of the testis, precluding the use of the testis as an organ of evaluation of testicular tumors of that cell type. The B6C3FI mouse, particularly the male, has a high and variable incidence of liver tumors (Haseman et aI., 1984). There is no single strain or species of laboratory animal that is obviously most predictive of chronic toxicity or carcinogenicity for humans. That statement can also be made for most other end points of toxicity. Thus, acceptance of one or multiple species of animal for testing must recognize the limitations of each species and strain, and interpretation of results must be made accordingly. Individual Differences Individual differences are apparent in every toxicological test, including those carried out in people. A paper by Gaines (1969), in which he reports the acute oral toxicity of pesticides, shows that for 69 compounds the LD 50 value for male rats was l.20-7.14-fold greater than the corresponding LD 1 value. The average factor of difference was

2.42. The corresponding factors of difference for the dermal toxicity of 42 pesticides were l.37-14.93 with an average of 3.00. In other words, judged in this way, individual variation, although very real, is usually relatively small. In studies of storage and excretion, the greatest individual average excretion of malathion-derived material differed from the smallest individual average excretion at the same dosage by factors of only 2.2-8.7 for different groups of people (Hayes et aI., 1960). Thus, the degree of difference was relatively constant in tests carried out at different dosage levels or at different times. A similar observation was made regarding the storage of DDT and the excretion of DDA in humans. In separate tests, the maximal storage of DDT was l.3-5.9 times the minimal storage at the same dosage level. For a single dosage group, the maximal rate of excretion of DDA by one man in anyone day was 18.0 times greater than his own minimal rate, and the difference between the lowest minimum and highest maximum within the group was a factor of 2l.5 (Hayes et aI., 1971). In all of these tests, the relative constancy of one individual compared with another was noted. Although individual differences may be described in statistical terms, physiological understanding of these differences is lacking almost entirely. If a popUlation is sufficiently heterozygous, the differences between individuals may depend

1.5 Factors Influencing Toxicity of Any Kind on their genetic diversity. However, individual differences persist to some degree in a homozygous population (see Section 1.3.1.4 for discussion on Gaussian distribution). This is illustrated by the failure of the LD 50 values of four pesticides in a particular population of mice to change in the course of 12 or more generations even though each succeeding generation was bred from mice that had survived an LD 50 dose (Guthrie et aI., 1971). Sex Chiefly, because of its convenience, the rat is used more than any other species for studies in toxicology. The rat also has the apparent distinction of showing more variation between the sexes in its response to chemicals than any other species. This fact may have led to more concern than is justified regarding possible differences in the susceptibility of men and women to chemicals. In any event, calculations from data provided by Gaines (1960, 1969) for the oral toxicity of 69 pesticides showed that the difference in the oral LD 50 for male and female rats ranged from 0.21 (indicating greater susceptibility of the female) to 4.62 (indicating greater susceptibility of the male), and averaged 0.94. The corresponding factors for the dermal toxicity of 37 pesticides were from 0.11 to 2.93, with an average of 0.81 (Hayes, 1967a). The differences in the susceptibility of male and female rats are associated to a large degree with differences in their liver microsomal enzymes. In contrast to the situation in rats and to a lesser degree in other rodents, significant differences between the sexes of other species in their susceptibility to poisons usually have not been reported. Such differences were looked for but not found in studies of the storage ofDDT in monkeys (Durham et aI., 1963; Ortega et aI., 1956). Such differences between men and women are small or lacking entirely. Pregnancy Susceptibility to a particular compound may be either greater, less, or identical in pregnant females than in nonpregnant ones of the same strain and age. For example, pregnancy exaggerates the danger of anticoagulants but reduces the danger of paraquat. For some differences such as susceptibility to anticoagulants the reason for the difference is clear. In most instances the reason is obscure. In a systematic study of 19 drugs, given by different routes, pregnant mice were more susceptible than nonpregnant ones by factors ranging from 0.74 to 14.55 and averaging 1.90, or 1.27 if the single high value is excluded (Beliles, 1972). Lactating rats consume approximately 3-fold (Hayes, 1976) or 2.5-3-fold (Yang et aI., 1984a) more feed than the same rats before or after lactation; thus lactating rats are subject to a marked increase in the dosage of all compounds in their diet. The extent to which this is true of other species apparently is not documented. Other Endocrines There is considerable evidence that the pituitary adrenal axis may be influenced by photoperiodicity and in turn may influence susceptibility to toxicants (see Section 1.5.17).

61

q \

70 60

>:

i so

S40

j

S ~

30

20 10 0

2

4

6 Months

8

10

12

Figure 1.19 Calculated DDT intake (mg/kg body weight) in rats receiving various levels ofDDT in the die (male, - ; female, ---). From Fitzhugh and Nelson (1947), by permission of the WiIIiams & Wilkins Co., Baltimore.

Lipsett (1983) has reviewed the relationship between hormones and cancer. Sex and other hormones usually have either pulsatile release or some cyclicity (diurnal or circadian cycles) in their secretion. In either event, an additional timescale needs to be incorporated into studying these phenomena in the context of toxicity in addition to the three independent time scales discussed earlier. Gender differences will become manifest in toxicology only when the aforementioned timescale becomes rate-determining (or -limiting). Pregnancy introduces still another timescale with known consequences for altered hormonal timescales for a limited period in an individual's life. Age Children and young animals are often more susceptible than adults to poisons in food. The most common reason is that children and other young animals eat more than adults in proportion to their weight. Thus, when given the same contaminated food, young animals receive a higher dosage of toxicant. The relationship for rats is shown in Fig. 1.19. Although the figure is based on DDT, it applies equally to any compound that does not cause a reduction of food intake. However, other factors may be involved. It is now well known that a number of drugs are poorly metabolized by infants, particularly those born prematurely (Fouts and Adamson, 1959). Although it is seldom possible to quantitate the difference, it is clear that a dosage of some drugs easily tolerated by human adults may lead to severe illness or even death in very young children. Calves and sometimes lambs are markedly more susceptible than adult cattle or sheep to sprays or dips of chlordane, dieldrin, and lindane (Radeleff, 1970). Systemic study of drugs (Hoppe et aI., 1965; Yeary and Benish, 1965) and pesticides (Brodeur and DuBois, 1963; Gaines and Linder, 1986; Lu et aI., 1965) indicates that newborn an-

62

CHAPTER 1 Dose, Time, and Other Factors Influencing Toxicity

imals are generally more susceptible than adults of the same species regardless of route of administration. However, the difference may be small and there are some exceptions in which the newborn are actually less susceptible than adults (Brodeur and DuBois, 1963; Gaines and Linder, 1986; Hoppe et al., 1965). As reviewed by Durham (1969), the difference, no matter what its direction, can often be explained in terms of the recognized activity of microsomal enzymes in activating or deactivating the chemical in question. Other factors that may be of importance are renal function and membrane permeability. Fortunately, compounds of similar pharmacological action tend to show similar differences in their toxicity to young adult animals (Yeary and Benish, 1965). In studying drugs, Yeary and Benish found that newborn rats were 0.6-10.0 times more susceptible than adults. Hoppe et al. (1965) found a similar range of 0.7-6.2. Goldenthal (1971) studied a much larger number of drugs and a few other compounds (290 in all) and reported a much wider range of factors of differences: 2,500 >1500

Oleoresin!

Shimkin and Anderson, 1936

102

Oleoresin!

Shimkin and Anderson, 1936 Malone and Brown, 1968

77.8%

Malone and Brown, 1968 Lawrence and Casida, 1982

Guinea pig

oral

Guinea pig

ip

Rabbit

dermal

> 19, 800

77.8% purea

Rabbit

dermal

>2,000

57.6% pureb

Schoenig, 1995

Dog

iv

LD = 6-8 mg/kgC

Negherbon, 1959

Chicken

pvg

1,565

Bobwhite quail

oral

>2,000

Mallard duck

5-day

Honeybee

contact

Frog

sc

Rainbow trout

48-h

Steelhead trout

96-h

0.000022

77.8% purea

Malone and Brown, 1968

57.6% pure

Gabriel and Mark, 1995

LCso > 5,620 ppm h


57.6% pure


5.8

Cole and Casida, 1983

Catfish

48-h

Catfish

96-h

Bluegill

48-h

Bluegill

96-h

= 54 ppbi LCso = 25 ppb j LCso = 82 ppbi LCso = 114 ppb j LCSO = 74 ppbi LCSO = 4 I ppb j

Bluegill

96-h

LCso

LCso

Bridges and Cope, 1965 Mauck et aI., 1976 Bridges and Cope, 1965 Maucketal.,1976

Bridges and Cope, 1965 Maucketal.,1976

soft water, pH 6.5

= 87 ppb l

Mauck et aI., 1976

soft water, pH 9.5 Bluegill

96-h

LCSO

= 46 ppbj

Mauck et aI., 1976

very hard water, pH 8.2 (continues)

3.2 Insect Control Agents

115

Table 3.1 (continued) LDso

Notes,

(mg/kg)

other data

Reference

Animal

Assay

Coho salmon

96-h

LCso = 39 ppb i

Pillmore, 1973

Coho salmon

96-h

LCso

Mauck et aI., 1976

American lobster

48-h

Stonefly

48-h

Daphnia pulex

48-h

= 39 ppb f LCso = 0.73 ppbk LCso = 6.4 ppbi LCso = 25 ppb i

Burridge and Haya, 1997 Bridges and Cope, 1965 Pillmore, 1973

aRefined nitromethane concentrate of pyrethrum extract. bThe ratio of pyrethrins I to pyrethrins H of the sample was 1.85. C Approximate lethal dose of pyrethrins. dlntracerebral injection. eMixture of 40% pyrethrins I and 46% pyrethrins H. fRecalculated from original mortality data. Oleoresin contained 14% pyrethrins I and 11. gPerivisceral injection. hDietary toxicity; lethargy and reduced body weight but no mortality were observed. i Static tests with extract containing 24.6% pyrethrins. f Static tests at 12°C with extract containing 20% pyrethrins. kPor the most sensitive, stage IV larvae. Twenty-five percent pyrethrum extract with piperonyl butoxide synergist was used.

3000 ppm. The highest concentration without treatment-related reproductive toxicity was 100 ppm. No oncogenic effects were seen in rodent studies at 100 ppm dietary concentrations of pyrethrins (California EPA, 1996; Schoenig, 1995). Pyrethrum extract was not mutagenic in the Ames test, no increases in chromosomal aberrations were seen in Chinese hamster ovary cells, and no increase in unscheduled DNA synthesis was observed (California EPA, 1996; Schoenig, 1995; see also Ashwood-Smith et aI., 1972; Moriya et aI., 1983).

Endocrine Effects The binding of pyrethrins and their synthetic analogs to various steroid hormone receptors was recently examined. In a human skin fibroblast androgen receptor assay, a 20% pyrethrum extract inhibited methyltrienolone binding with a Ki of 1.5 x 10-5 M and also inhibited testosterone binding to sex hormone binding globulin at 10-4 M (Eil and Nisula, 1990). According to Soto et al. (1995), pyrethrum extract lacks estrogenic activity. Among a series of pyrethroids tested in T47D human breast cancer cell line, 30 J..lM d-trans-allethrin antagonized the progestagenic effect of 5 nM progesterone (Garey and Wolff, 1998). Treatment The toxic symptoms caused by intravenous administration of permethrin at 40 mg/kg (ED95) in mice could be partially alleviated by intraperitoneal pretreatment with diazepam (10 mg/kg), aminooxyacetic acid (50 mg/kg), or cycloheximide (1 mg/kg) (Staatz et aI., 1982). Diazepam at a 3 mg/kg intraperitoneal dose was more specific and potent than phenobarbital in protecting mice from LD95 doses of Type 11 pyrethroids and increased the intracerebroventricular LD50 values of deltamethrin or permethrin six- to ninefold (Gammon et aI., 1982).

Mephenesin prevented choreoathetosis at moderate doses but gave full protection against both types of pyrethroids only at high doses that also caused profound muscle relaxation (Bradbury et al., 1983). Methocarbamol, alone or in combination with atropine that only blocked salivation, reduced poisoning symptoms and mortality in pyrethroid-treated rats (Hiromori et aI., 1986). Toxicity to Humans

Poisoning Incidents Acute pyrethrum poisoning cases are extremely rare. In fact, pyrethrum preparations were once recommended as internal anthelmintic agents (McLellan, 1964). Accidental ingestions by children, including a fatal case of a 2-year-old girl, were reported in the late 19th century (see Ray, 1991). Irritation and Sensitization Allergy from inhalation of or direct dermal contact with pyrethrum flowers during harvesting and processing or with unrefined pyrethrum extracts is not uncommon (Barthel, 1973; Garcia-Bravo et aI., 1995; McCord et aI., 1921). The symptoms of pyrethrum dermatitis are initially mild erythema covering hands and face, which, on further contact, can develop into edema and blistering (see RickeU et aI., 1972). Understandably, the allergenic properties of the flower, its constituents, and the synthetic pyrethroids were thoroughly investigated (Taplin and Meinking, 1990). Testing immediate and delayed hypersensitivity reactions in guinea pigs previously sensitized to pyrethrum, Rickett and Tyszkiewicz (1973) identified the major allergens in the 0.9% saline extract of pyrethrum flowers. The allergenic agents were tentatively assigned as 60-200 kDa glycoproteins. The refined pyrethrum extract and pyrethrin 11 were not allergenic. Crude oleoresin

116

CHAPTER 3 Pest Control Agents from Natural Products

contained traces of another allergen, pyrethrosin, a sesquiterpene lactone possessing an a-methylene moiety (Mitchell et aI., 1972; Rickett and Tyszkiewicz, 1973). Sensitization to pyrethrum frequently occurs with individuals sensitive to the pollen of Ambrosia spp. (ragweed) (Feinberg, 1934; see also Carlson and Villaveces, 1977). Pyrethrosin-related lactones are the major allergens of noninsecticidal ornamental chrysanthemum species (Hausen and Schulz, 1973).

Treatment Because pyrethrins are readily metabolized and excreted, treatment of pyrethrin poisoning is mainly symptomatic and supportive. Pneumonitis, resulting from aspiration of kerosene or other hydrocarbons used in the insecticide formulation, may complicate poisoning incidents. Because of the potential irritancy of pyrethrins, proper decontamination following exposure is important. Dermatitis and allergic reactions caused by occupational exposure on pyrethrum-growing farms and in production facilities were prevented by minimizing exposure to the irritant (Gnadinger, 1945; Moore, 1975). Pyrethrin spray-related hypersensitivity pneumonitis could be treated with prednisone and ampicillin (Carlson and Villaveces, 1977). Tucker et al. (1983) recommended vitamin E (a-tocopherol) oil for the treatment of the cutaneous sensation or paresthesia among individuals exposed to synthetic pyrethroids. The mode of action of vitamin E appears to be the selective block of pyrethroid-modified Na+ channels as was shown in situ in rat cells (Song and Narahashi, 1995). 3.2.1.2 Nicotine Introduction Nicotine, a structurally simple alkaloid (Fig. 3.2), is a most notorious botanical insecticide. It is the main bioactive component of the tobacco plants Nicotiana tabacum, N. glauca, and N. rustica (Solanaceae). Nicotine is also found in the leaves of the Australian shrub Duboisia hopwoodi (Solanaceae), which is used by the Aborigines as a stimulant and hunting aid, and is also present in a number of other plants of the families of Lycopodiaceae, Crassulaceae, Leguminosae, Chenopodiaceae, and Compositae (Leete, 1983). The alkaloid is biosynthesized in the roots and thereafter translocated to the aerial parts, reaching, on a dry-weight basis, 1-8% content in the leaves of N. tabacum and 2-18% in N. rustica.

o,y N (S)-nlcotlne

uV

nomicotlne

coIInlno

anabasine

anatabine

N

myosmlno

Figure 3.2

Structures of tobacco alkaloids.

The pure alkaloid was first isolated by Posselt and Reimann in 1828 and the structure was determined by Pinner in 1893. Identity, Physicochemical Properties, and Uses IUPAC name: (S)-3-(1-methylpyrrolidin-2-yl)pyridine Chemical Abstract name: (S)-3-(I-methyl-2-pyrrolidiny1)pyridine CAS Registry Number: (S)-nicotine: [54-11-5] CAS Registry Numbers for other tobacco alkaloids (Fig. 3.2): (R)-nicotine: [25162-00-9]; racemic nicotine: [22083-74-5]; nicotine sulfate: [65-30-5]; (S)-anatabine: [581-49-7]; (S)-cotinine [486-56-6]; myosmine [532-12-7]; (S)-nornicotine [494-97-3] Empirical formula: ClOH14N2; molecular weight: 162.2

Physicochemical Properties of Nicotine Pure nicotine is a colorless liquid that boils at 246-247°C; its freezing point is below - 79°e. Its density is 1.009 g/cm3 at 20°e. The free base is fairly volatile with a vapor pressure of 4.25 x 10- 2 mmHg. The concentration of nicotine in the vapor phase, as determined by the "air bubbling method," is about 28 ppm at 25°C (see Jackson, 1941). Nicotine is hygroscopic and freely miscible with water, ethanol, ethyl ether, and most organic solvents. Its pKaJ = 3.09 and pKa2 = 8.18. The log P value of the nonionized alkaloid is 0.93 (Chamberlain et aI., 1996). Pure natural nicotine is levorotatory: [a]~o = -163.9°. Hereinafter, "nicotine" will generally refer to the natural product, that is, the (S)-isomer. Stability Upon exposure to air, pure nicotine turns brown with a characteristic odor reminiscent of tobacco. Its water-soluble salts are more stable. At 30°C and in the presence of oxygen, the alkaloid is degraded into several oxidized products, including cotinine, myosmine, nicotinic acid, methylamine, and ammonia. In the field, however, the persistence of nicotine is sufficiently long for insecticidal purposes. For example, nicotine sprayed on mustard greens had a half-life of 4.5 days at 10-50 ppm treatments, indicating that the alkaloid penetrated into the wax layer of the leaf and thus was protected from aerial dissipation (Gunther et aI., 1959). Formulations and Agricultural Uses The tobacco plant was introduced to Europe in 1559 from the Americas where it had long been cultivated by the American Indians primarily for smoking; from 1690 on, tobacco dust and extracts were used to repel and kill insects. Tobacco smoke was also used for fumigation. Tobacco was reintroduced as an insecticide in the United States in 1814. Nicotine is mainly obtained as a by-product of cigarette manufacturing. The commercial insecticide may contain traces of accompanying tobacco alkaloids. Nicotine has a systemic action and is used on fruits, vegetables, and ornamentals in the field and in greenhouses against a wide range of insects, including aphids, thrips, and whitefiies. It also has anthelmintic

3.2 Insect Control Agents properties. Nicotine and related insecticides are reviewed by Schmeltz (1971) and Ujvary (1999). Nicotine free base is commercialized as a concentrated e.g., 40%, aqueous solution or fumigant formulation of 0.05-4.0% alkaloid content. Nicotine sulfate is sold as a dispersible powder and as a 40% aqueous solution (Black Leaf 40). Nicotine is formulated also on an inert material support such as bentonite. Nicotine can also be used to capture and restrain dangerous, unmanageable, or wild animals. This method relies on a projectile-type syringe loaded with nicotine solution and fired from a special C02-powered rifle. Cattle weighing up to 450 kg could thus be safely immobilized with up to 3 mg/kg doses within minutes (Hayes et aI., 1959). The high amount of nicotine that should be handled, however, represents a risk to users. Biological Properties Because nicotine is also the major psychoactive component of tobacco used for smoking or chewing by hundreds of millions of people worldwide, its biochemistry, pharmacology, and toxicology have been thoroughly investigated. Early works on the various biological activities of nicotine are summarized by Larson et al. (1961). For a modem pharmacological treatment, the reader is referred to Taylor (1996) as well as to recent reviews by Benowitz (1996) and Brioni et al. (1997).

Mode of Action Nicotine exerts its pharmacological and neurotoxic effect in animals and humans by binding to a subset of cholinergic receptors, the nicotinic acetylcholine receptors (nAChRs). Several nAChR subtypes are known. These receptors are 1igand-gated ion channels formed from pentameric arrangements of four different homologous peptide subunits (Changeux et aI., 1992). Receptors composed of different subunit combinations have distinct physiological and pharmacological properties and are differentially distributed in autonomic ganglia, skeletal muscles, and the central nervous system. The activation of nAChRs causes a rapid increase in cellular permeability to Na+ and Ca2+, leading to depolarization and excitation. Prolonged application of nicotine or other agonists results in desensitization of the cholinergic receptor site and a lasting blockade. Activation of nAChRs facilitates the release of neurotransmitters, including acetylcholine, norepinephrine, dopamine, serotonin, ,B-endorphin, among others. The effects of nicotine on nAChRs are not antagonized by atropine but can be selectively blocked by other agents (e.g., tubocurarine or a-bungarotoxin). Nicotine can also act at noncholinergic sites. Mammalian neural nAChR subtypes are found in both the central and the peripheral nervous systems, whereas, in insects, nAChRs were detected only in the central nervous system (EIdefrawi and Eldefrawi, 1997). The binding of nicotine is stereose1ective. (S)-Nicotine was lO to 60-fold more active than the nonnatura1 (R)-enantiomer in vitro in some (Barlow and Hamilton, 1965; Romano and Goldstein, 1980; Zhang and Nordberg, 1993) but not in other (Ikushima et aI., 1982) bioassays. The acute toxicities of the isomeric alkaloids also differ; for example, the intravenous LDso value of the (S)- and (R)-stereoisomer in mice

117

was 0.38 mg/kg and 2.75 mg/kg, respectively (Aceto et aI., 1979). Recent developments on the pharmacology of the nAChR, with special regard to insect versus mammalian systems, were reviewed in a monograph by Yamamoto and Casida (1999).

Absorption, Metabolism, and Excretion The mammalian metabolism of nicotine is well understood (Gorrod, 1993) as is the pharmacokinetics (Rotenberg, 1982). Nicotine base is readily absorbed through the skin, the mucous membranes, and, when inhaled, the lungs. Absorption is less from acidic solutions, rendering the commercial sulfate salt safer on dermal contact (Faulkner, 1933). Nicotine is not readily absorbed from the stomach unless intragastric pH is raised but intestinal absorption is far more efficient. Up to 90% of the absorbed nicotine is rapidly metabolized in the liver, the kidneys, and the lungs. The metabolites and any unaltered nicotine are eliminated by the urine. The half-life of nicotine in the body is in the range of 2 to 3 h. The primary metabolite of nicotine is cotinine (Fig. 3.2) formed by cytochrome P-450-mediated microsomal oxidation both in insects and in mammals. Cotinine is also a trace tobacco constituent. It is a poor insecticide (Yamamoto, 1965) and inactive at the mammalian nAChR (Benowitz, 1995). Nevertheless, cotinine influences neurotransmitter release in the brain and affects the cardiovascular system and a number of enzymes. Cotinine levels are at least lO-fold higher than those of nicotine and its half-life is 4-8 times longer than that of the parent alkaloid; thus, its contribution to the overall pharmacology and toxicity of nicotine cannot be discounted (see Benowitz, 1996; Vainio et al., 1998). Cotinine is also the major metabolite in crops treated with nicotine insecticide (Gunther et ai., 1959). Nornicotine, formed by oxidative demethy1ation, is another metabolite of nicotine. It is also present in Solanaceae usually as a mixture of the (S)- or (R)-stereoisomers, either of which may predominate. In studies with rat brain nAChRs, (S)nicotine binding was effectively displaced by both isomers of nornicotine with ICso values of approximately 0.9 nM (Zhang and Nordberg, 1993). An additional urinary metabolite in many species, including humans is the N' -oxide. Although the N' -oxide at large doses shared the pharmacological profile with nicotine in dogs (see Larson et aI., 1961), this could be due to the parent nicotine regenerated from the N' -oxide by gastrointestinal reduction as observed both in vitro and in vivo (Beckett et aI., 1970; Crooks, 1993). The subcutaneous LDso of N'-oxide is 940 mg/kg in mice. A novel metabolic intermediate, the reactive alkylating agent nicotine-~ l'(S') iminium ion (Nguyen et aI., 1979), was recently implicated in the pharmacological and toxicological effects of nicotine in the brain (Gorrod, 1993; Jacob et aI., 1997). The rapid metabolism of nicotine was illustrated in experiments with cats given 40 ).Lg/kg intravenous dose of [14C]nicotine (Turner, 1969). Tissues, including the brain, liver, kidneys,

118


lungs, skeletal muscles, and stomach, showed maximum nicotine content in about 5 min after injection. Cotinine, appearing in the blood and liver within minutes of injection, was continuously transformed to other metabolites. Fifty-five percent of the radioactivity was excreted in the urine within 24 h, but only 1% of the radioactivity was unchanged nicotine. In 3 days, 70% of the injected radioactivity was excreted via urine, whereas feces contained less than 1% of radioactivity. Additional metabolites identified in the urine and liver were nomicotine, demethylcotinine, pyridylacetic acid, and nicotine N' -oxide. Pharmacological Actions and Poisoning Syndromes The mammalian pharmacology of nicotine in vivo is discussed in recent reviews (Brioni et aI., 1997; Taylor, 1996) so it will only be summarized here briefly. Nicotine has a multitude of pharmacological and physiological effects, and the responses depend greatly on the dose and rate of absorption. The alkaloid targets both the central and the peripheral nervous systems. In general, the initial stimulation is followed by depression. Both enantiomers of nicotine and nornicotine show stimulant and depressant activities. For example, injection of nicotine into rats produces a biphasic effect on locomotor activity in a dose-dependent manner. First, ataxia is seen, which, at doses larger than 0.8 mg/kg, takes the form of prostration. About 10-20 min later, a stimulant effect can be observed. The stimulant effect was shown to involve nAChR-mediated dopamine release. The motor activation is followed by tremor, and seizures are also common. Typical stimulatory responses of the cardiovascular system to nicotine are increases in heart rate, myocardial con tractility, and blood pressure. However, the drug can also slow the heart rate by paralysis of sympathetic or stimulation of parasympathetic cardiac ganglia. Nicotine also initiates the release of catecholamines in a number of isolated organs, causing additional cardial responses (Haass and Kiibler, 1996). The effect on the gastrointestinal tract is overall parasympathic stimulation, which results in increased bowel activity. Individuals not exposed previously to nicotine often experience nausea, vomiting, and diarrhea. The alkaloid has marked effects on the exocrine glands: The initial stimulation causing increased salivation is followed by inhibition of secretion. Neurochemical and behavioral assays suggest that nicotine metabolites are also pharmacologically active (see Crooks and Dwoskin, 1997). In summary, the overall response to nicotine manifests as the summation of stimulatory and inhibitory effects. There is increasing evidence that nicotine elicits these effects by binding to different receptor subtypes in the brain and in the periphery (Brioni et aI., 1997).

Toxicity to Laboratory Animals Acute Toxicity Over the past century, a plethora of data have accumulated on the toxicity of tobacco alkaloids. Table 3.2 lists representative acute toxicity data for various test species. Nicotine has a rapid contact and vapor action. With lethal exposures, renal failure, hypotension, paralysis, and coma may

precede death, which is usually caused by respiratory failure due to both central paralysis and peripheral blockade of the muscles of respiration. Nomicotine is generally less toxic in most species but could be more toxic to some, depending on the mode of administration. In guinea pigs, for example, the subcutaneous LDso values for nicotine and nomicotine are 32 and 28 mg/kg, respectively, but, for rats, the subcutaneous LDso values are 23.5 mg/kg for both alkaloids (Negherbon, 1959; see also Larson et aI., 1945). In mice, nomicotine had intraperitoneal and intravenous LDso values of 21.7 and 3.4 mg/kg; in rabbits, the respective values were greater than 13.7 and 3.0 mg/kg (Larson et aI., 1945). Myosmine is a trace component in Nicotiana species. In rats, the oral and intraperitoneal LDso values are 1875 and 190 mg/kg, respectively (Ambrose and DeEds, 1946). Other Pharmacological and Biochemical Effects It was noted (Ruppert, 1942) that a single dose of nicotine resulted in tolerance to a subsequent dose of nicotine (tachyphylaxis). In mice, for example, an intravenous sublethal dose of 0.8 mg/kg nicotine hemitartrate, when given 5 min before toxicity determination, raised the intravenous LDso up to 20.8 mg/kg, which is over 10 times higher than the corresponding value without pretreatment (Barrass et al., 1969; see also Benowitz et aI., 1987). Nicotine N' -oxide was almost as effective when given 40 min before nicotine. Nicotine is both a substrate and an inhibitor of cytochrome P-450 enzymes involved in glucocorticoid and sex steroid biosynthesis. For example, nicotine, cotinine, and anabasine inhibit cytochrome P-450-mediated adrenal aldosterone synthesis in rats (Skowronski and Feldman, 1994) and estrogen synthesis by human aromatase in vitro (Barbieri et aI., 1986; see also Kadohama et aI., 1993). Carcinogenesis and Mutagenicity Recent investigations established that the carcinogenicity of tobacco and nicotine is associated with N' -nitrosamines of secondary amines, such as N' -nitrosonomicotine or certain nitrosaminoketones, either already present in tobacco or formed from nicotine during storage, processing, or biotransformation in the body (reviewed by Hecht, 1998; Hoffmann et al., 1985). Cotinine was carcinogenic in rats (Truhaut et aI., 1964). Nicotine sulfate was nonmutagenic in Salmonella typhimurium reversion-assay systems (Moriya et al., 1983). Treatment Artificial respiration alone before cessation of circulation or artificial respiration together with intracardial injection of epinephrine rescued nicotine-poisoned dogs (Franke and Thomas, 1936). Nicotine-induced convulsions can be blocked by certain anticholinergic drugs such as diethazin and diphenhydramine but not with cholinergic antagonists. The prostration response produced by injecting nicotine in rat brain was prevented by subcutaneous injection of mecamylamine (Shoaib and Stolerman, 1994). Tetraethylammonium chloride proved to be a useful antidote against acute doses of nicotine in mice


119

Table 3.2 Acute Toxicity of Nicotine LDso (mg/kg)

Notes, other data

Reference

Animal

Route

Rat

oral

188a

Rat

oral

50--60

Rat

oral

52.5

Lazutka et aI., 1969

Rat, female

oral

83 b

Gaines, 1969

Rat

oral

90b

Rat, female

dermal

Rat, male

sc

47 c

Hoick et aI., 1937

Rat, female

sc

37 c

Hoick et aI., 1937

Rat

sc

33.5

Rat

im

Rat, male

ip

14.6

Rat

iv

2.8

Rat

iv

Mouse

oral

Mouse

oral

Ambrose and DeEds, 1946 Negherbon, 1959

Bond et aI., 1973

285 b

Gaines, 1969

Negherbon, 1959 LDlOO

= 15 mg/kgd

Feurt et aI., 1958 Blum and Zacks, 1958 Larson et aI., 1949a

LDlOO

= I mg/kgd

Negherbon, 1959 Heubner and Papierkowski, 1938

24

Lazutka et aI., 1969

3.34

Heubner and Papierkowski, 1938

Mouse

sc

Mouse

im

16

Mouse

iv

7.1

Larson et aI., 1949b

Mouse, female

iv

2.0e

Barrass et aI., 1969

LDlOO

= 8.0 mg/kgd

Feurt et aI., 1958

Negherbon, 1959

26

Guinea pig

sc

Guinea pig

im

LDlOO

Guinea pig

iv

LDlOO

Rabbit

im

Rabbit

dermal

50

Negherbon, 1959

Rabbit

ip

14

Larson et aI., 1945

Rabbit

iv

Cat

im

Cat

iv

Cat

iv

Dog

im

Dog

iv

Pig

= 15 mglkgd = 4.5 mg/kg LDIOO = 30 mg/kgd

Feurt et aI., 1958 Negherbon, 1959 Feurt et aI., 1958

Larson et aI., 1949b

9.4

LDlOO

= 9.0 mg/kgd

Feurt et aI., 1958 Larson et aI., 1949b

2.0

= 6.1 mglkgd LDlOO = 15 mglkgd

Feurt et aI., 1958

im

LDlOO > 14 mglkgd

Feurt et aI., 1958

Goats

im

Feurt et aI., 1958

Deer

im

Cattle

im

= 13 mg/kgd LDlOO = 9.0 mg/kgd LDlOO = 9.0 mg/kgd LDlOO = 8.8 mglkgd LDlOO = 6.0 mg/kgd LDlOO = 9.0 mg/kgd LDlOO = 40 mg/kgd LC50 = 4 ppm LC50 = 0.24 ppm

Horse

im

Monkey

im

Pigeon

im

Frog

sc

LDlOO

LDIOO

Daphnia pulex a Applied

in aqueous solution adjusted to pH bNicotine su1fate. c Calculated from mortality data. d Approximate minimal lethal dose. eNicotine hydrogen tartrate.

Larsonetal.,1949b

5.0

Rainbow trout

= 7 with concentrated hydrochloric acid.

Negherbon, 1959

Feurt et aI., 1958 Feurt et aI., 1958 Feurt et al., 1958 Feurt et aI., 1958 Feurt et aI., 1958 Negherbon, 1959 Tomlin, 1997 Tomlin, 1997

120


(Andrews and Miskus, 1968). The behavioral, respiratory, and electrocortical effects of nicotine infused into the brain of fowl could be prevented or abolished by systemic or local administration of pempidine (Marley and Seller, 1974). Poisoning Incidents Application of a 5% aqueous nicotine solution on cattle to get rid of ectoparasites produced classical signs of nicotine toxicosis: accelerated respiration, salivation, sweating, diarrhea, and tremor. Of the 18 animals receiving nicotine, one died 2 h and another one 14 h after treatment. While 14 of the animals were symptomless on the second day, two showing serious symptoms were given digitalis, veratrine, and arecoline to recover (Kamaras, 1936). Plumlee et al. (1993) described a case of toxicosis of cattle that was caused by grazing on tree tobacco, N. glauca, a shrub found in the southwestern United States. Toxicity to Humans The often-cited acutely fatal dose of nicotine for an adult is about 60 mg (one drop) of nicotine, yet individuals have ingested larger quantities and recovered (see Franke and Thomas, 1936). In the first half of the 20th century when this readily available insecticide was extensively used, both accidental and intentional nicotine poisoning cases were rather common. Severe accidents occurred due to dermal absorption. Symptoms of acute nicotine poisoning occur rapidly and include nausea, salivation, abdominal pain, vomiting, diarrhea, cold sweat, headache, dizziness, disturbed hearing and vision, mental confusion, and overall weakness. The blood pressure falls and faintness and prostration occur. Breathing becomes difficult; the pulse is weak, rapid, and irregular. Collapse may be followed by convulsions. Death may result within a few minutes from respiratory failure (Taylor, 1996). In dermal poisoning, however, the onset of symptoms could be delayed by several hours (Benowitz et aI., 1987). Effect on Reproduction The effect of nicotine, especially smoking in pregnant women, has adverse consequences for the mother, her fetus, and the newborn (reviewed by Lambers and Clark, 1996). Nicotine increases spontaneous abortions or premature delivery rates and decreases birth weight. Nicotine was shown to concentrate in fetal blood, amniotic fluid, and breast milk, causing various effects discussed earlier in both the fetus and the neonate. Poisoning Incidents Green-tobacco sickness is an occupational illness occurring regularly in tobacco fields (McBride et aI., 1998; Quandt et aI., 2000). Symptoms usually occur several hours after harvesting of wet green tobacco begins and last no longer than 24 h if contact with tobacco leaves is avoided. The initial headache and nausea usually lead to severe vomiting, pallor, and prostration. Smoking appears to protect against the illness. Accidental poisonings by solutions and vapors of nicotine insecticide were rather common in the early part of the 20th century (Faulkner, 1933; Wehrlin, 1938).

Garcia-Estrada and Fischman (1977) described an unusual case of nicotine poisoning by homemade tobacco enema for which an aqueous concoction of 5-10 cigarettes was used. The patient's hypotension and bradycardia were reversed about 10 h after the enema by intravenous atropine sulfate. Weiss (1996) described a case of acute nicotine toxicosis for a patient hospitalized for lithium therapy for bipolar disorder. The patient wanted to give up smoking and was given a transdermal nicotine patch (21 mg/day). He still smoked intermittently and 4 days later complained of nausea, diaphoresis, and hand tremor. The symptoms were first misdiagnosed as lithium toxicity but soon corrected as a typical case of nicotine toxicity and the transdermal nicotine treatment discontinued. The universal availability of nicotine has made it a common means of suicidal poisoning as illustrated by several cases in the literature. In a fatal suicide case, described by Lavoie and Harris (1991), a 17-year-old smoker ingested an estimated 5 g of a concentrated nicotine insecticide solution then vomited and collapsed, pulseless. Following immediate CPR, the initial asystole was converted to ventricular fibrillation with intravenous epinephrine. Subsequent orotracheal intubation and defibrillation produced a synapse tachycardia and the patient was placed on lidocaine for premature ventricular contractions. In the emergency room, mechanical ventilation and dopamine infusion were applied. Gastric lavage with normal saline followed by charcoal was implemented. The severe convulsions that occurred were controlled with diazepam, dilantin, mannitol, and dexamethason. The urine nicotine screen was positive and the serum level was 13,600 ng/ml, enough to be lethal [the average peak level for smokers is 49 ng/ml (Russell et aI., 1976)]. Computer tomography showed cerebral edema and electroencephalogram (EEG) revealed no cortical function. On the second day of hospitalization, intractable hypotension set in and the patient died 64 h after ingestion. Besides insecticide formulations, the traditional sources of nicotine were snuff or cigarette butts (Saxena and Scheman, 1985). The recently developed and now easily accessible transdermal nicotine patches, which typically contain 7-114 mg of the alkaloid, are modern sources of the poison for suicide attempts among adults (Engel and Parmentier, 1993; Kemp et aI., 1997) or for accidental intoxication among children (Woolf et aI., 1997). Pathology In contrast to acute poisoning cases where there are no specific pathological changes attributable to nicotine, chronic exposure to nicotine is implicated in the pathogenesis of coronary and peripheral vascular disease, chronic lung disease, cancer, and various endocrine disturbances (Benowitz, 1986,1996). Treatment Early removal of nicotine, aggressive respiratory support, and treatment of shock are important countermeasures (Franke and Thomas, 1936; Taylor, 1996). Vomiting should be induced; gastric lavage could also be performed. Activated charcoal is a valuable adjunct in neutralizing ingested nicotine


in the stomach, but tannic acid solution, although useful in precipitating some other alkaloids (e.g., strychnine), is of little use for this alkaloid (Hayes, 1975). Potassium permanganate diluted 1O,000-fold in the lavage fluid can also be used. Alkaline solutions, which facilitate absorption, should be avoided. Patients exhibiting convulsions may require sedation (e.g., by intravenous diazepam) and vasopressor drugs if hypotension fails to respond to the usual therapy. Patients surviving for more than a few hours after ingesting nicotine are likely to recover because nicotine is detoxified fairly rapidly.

121

followed by depression of respiration and complete muscular paralysis. Anabasine, however, was found to be less excitatory and more depressing than nicotine. The animals completely recovered from sublethal doses of anabasine in about an hour. The teratogenicity of N. glauca plants and its primary alkaloid, anabasine, in pigs, sheep, goats, calves, or chicken was extensively studied (Bush and Crowe, 1989). The principal defects in piglets were deformities of fore and rear limbs (arthrogryposis) and cleft palate that could be induced at 2.6 mg/kg doses of anabasine twice daily during the 43rd to 53rd days of gestation (Keeler et aI., 1984).

3.2.1.3 Anabasine Introduction Anabasine (Fig. 3.2), a close structural relative of nicotine, was isolated from the toxic Asian plant Anabasis aphyUa (Chenopodiaceae) by Orechoff and Menschikoff in 1931. The same year, Smith isolated this compound as an insecticidal trace constituent of synthetic dipyridyl oil and named it "neonicotine." The anabasine content of A. aphylla could be as high as 2.6% and the alkaloid is also present in Nicotiana species (e.g., the wild tree tobacco, N. glauca). Anabasine sulfate was a widely used botanical insecticide in the former Soviet Union until the 1970s. Identity, Physicochemical Properties, and Uses IUPAC name: (S)-3-(piperidin-2-yl)pyridine Chemical Abstract name: (S)-3-(2-piperidinyl)pyridine CAS Registry Number: (S)-anabasine: [494-52-0]; racemic anabasine: [13078-04-1] Empirical formula: ClOH14N2; molecular weight: 162.2 Anabasine is a clear liquid with a boiling point of 281°C. Its freezing point is 9°C. The density of anabasine is 1.048 g/cm3 at 20°C. The alkaloid is miscible with water and most organic solvents. Anabasine from A. aphylla is levorotatory [a]5° = -82.20°, but the alkaloid obtained for practical purposes from the American N. glauca is mostly racemic.

Stability Anabasine is somewhat more stable than nicotine, but in air it also turns brown. Biological Properties

Mode of Action Similar to nicotine, anabasine is also a nAChR agonist alkaloid. Toxicity to Animals

Acute Toxicity Anabasine was more toxic than nicotine when given intravenously to rabbits: the minimal fatal doses for anabasine and nicotine were 3 and 9 mg/kg, respectively (Haag, 1933). For guinea pigs, the difference between anabasine and nicotine was less with respective fatal doses of 22 and 26 mg/kg on subcutaneous administration. The symptoms of anabasine treatments were the same as those for nicotine: initially increased respiration, then hyperexcitability, muscular twitching,

Poisoning Incidents Habib et al. (1974) reported that an anabasine-containing plant, Haloxylon persicum (Chenopodiaceae), growing wild in Saudi Arabia, caused death among grazing animals. A case of N. glauca toxicosis in a herd of cattle ingesting leaves of the tree tobacco was reported by Plumlee et al. (1993). Symptoms of poisoning were ataxia, depression, anorexia, and mild colic. Necropsy of one of the dead animals revealed edematous and congested lungs with peracute aspirative pneumonia. Histology indicated a mild, multifocal, suppurative rumenitis. Anabasine was found in the liver (2 ppm), urine (3 ppm), and rumen contents (20 ppm). Human Poisoning Incidents A poisoning case due to homemade enema with an aqueous solution of anabasine sulfate was reported by Danilin and Shabaeva (1968). The initial symptoms 30 min after administration of the enema were general weakness, nausea, cardiac disturbances with an arterial blood pressure of 80/50, pulse rate of 42 per min, a body temperature of 35.2°C, and acrocyanosis. Gastric lavage was performed. After an additional 35 min, respiration became severely depressed, and the patient sweated and lost consciousness. Resuscitation and intubation were employed and subcutaneous atropine and nikethamide were given and the patient's condition slowly improved over the next 6 h. Anabasine was subsequently identified in the urine but not in the gastric lavage fluid because the alkaloid had been removed by defecation that immediately followed the enema. Complete recovery took several days. A fatal anabasine poisoning due to ingestion of N. glauca leaves was described by Castorena et al. (1987). A young adult male was found in the field where the plant was also identified. Autopsy revealed only moderate congestion and edema of both lungs. Anabasine concentration of the gastric content, urine, and blood was 113.4, 73.8, and 1.15 mg/l, respectively. The heart, kidney, brain, and lung contained 10.4-15.8 mg/l of the alkaloid. An extract of the leaves collected in the field contained 2.0 mg/g of anabasine.

Treatment ing.

Treatment is the same as that for nicotine poison-

3.2.1.4 Rotenone Introduction Roots of the East Asian Derris (tuba) plants, particularly of Derris elliptica and D. chinensis (Leguminosae),

122


have long been used to stupefy fish for easy collection and also as insecticides and hunting aids. In South America, Lonchocarpus utilis, L. urucu and L. nicou (cube, timbo, and barbasco), as well as preparations from Tephrosia, Mundela, and Millettia species, have been used for the same purpose. The major and most studied bioactive principle of the tropical plants is rotenone, first isolated from L. nicou by Geoffroy in lS95. The rotenone content in D. elliptica and L. uti/is roots ranges from 3 to 11 %. The structure of rotenone (Fig. 3.3) was established by several research groups simultaneously in 1932 (reviewed by Crombie, 1963; LaForge et al., 1933). Additional notable rotenoids from leguminous plants are deguelin, ellipton, malaccol, sumatrol, and a-toxicarol (Crombie, 1963; Fukami and Nakajima, 1971) as well as the two recently identified oxahomologs (Fang and Casida, 1997).

Identity, Physicochemical Properties, and Uses IUPAC name: (2R,6aS,12aS)-1,2,6,6a,12,12a-hexahydro2-isopropenyl-S, 9-dimethoxychromeno [3 ,4-b ]furo[2,3-h ] chromen-6-one Chemical Abstract name: [2R-(2a,6aa,12aa)]-1,2,12,12atetrahydro-S, 9-dimethoxy -2-( 1-methyletheny1)-[1 ]benzopyrano[3,4-b ]furo[2,3-h] [1 ]benzopyran-6( 6aH)-one CAS Registry Number: [S3-79-4] Empirical formula: C23H2206; molecular weight: 394.4

The decomposition of rotenone in solution, on plants and glass surfaces was studied in detail by Cheng et al. (1972), who found that irradiation of rotenone in oxygenated methanol with ultraviolet (UV) light yielded a mixture of over 20 photodegradation products with lower mammalian toxicity than the intact parent compound. The most toxic and prevailing degradate isolated was the l2a,B-hydroxylated derivative, 6a,B,12a,Brotenolone, having an intraperitoneal LD50 of 4.1 mg/kg to male mice.

Formulations and Uses Cube, derris, and tuba preparations are commercialized alone or in combination with other botanical insecticides and synergists. Products containing pure rotenone as the sole active ingredient are used as agricultural insecticides and acaricides in orchards and vegetable cultivations as well as to control fire ants and household insects. Rotenone is also used to treat scabies and head lice on humans as well as various ectoparasites on livestock and pet animals. It is formulated as dust, wettable powder, and emulsifiable or soluble concentrates typically containing 0.4-8% of the active ingredient. In fishery management, rotenone is used to kill unwanted fish in ponds either alone or with a synergist at concentrations of 1-4 ppm. A recent application was to eradicate the predatory northern pike in Lake Davis, California.

Biological Properties Physicochemical Properties The melting point of rotenote in its orthorhombic form is 165-166°C; in its dimorphic form, the melting point is lS5-1S6°C. At 20°C, rotenone is practically insoluble in water (0.2 mg/l) (see, however, Loeb and EngstromHeg, 1970) but readily soluble in chloroform (472 g/l) and acetone (66 g/l) and only slightly soluble in ethyl ether (4 g/l) and ethyl alcohol (2 g/l). Rotenone is highly lipophilic with a log P of 4.26 (Gingerich and Rach, 1985). It is levorotatory: [a]5° = -228° (c = 2.22 in benzene). Stability Derris and cube preparations as well as rotenone are stable on storage. When exposed to light and air in the field, however, rotenone decomposes rapidly, losing toxicity within days. Rotenone is readily isomerized by bases, even by alkaline glass surfaces. After use in fish management programs, residues of the easily oxidizable rotenone can be neutralized by dilute solutions of potassium permanganate (Lawrence, 1956) or chlorine.

I o

"0

rotenone

Figure 3.3

Structure of rotenone.

Mode of Insecticidal Action Rotenone is a classical inhibitor of the respiratory chain of mitochondria and acts by inhibiting electron transport at reduced nicotinamide adenine dinucleotide (NADH):ubiquinone oxidoreductase (Complex I). Complex I of the well-characterized bovine heart mitochondrial membranes consists of more than 40 protein subunits with a total molecular weight of over 900 kDa (Pilkington et al., 1993). Complex I contains one flavin coenzyme and at least four ironsulfur centers and it is linked to additional polypeptide electron carrier complexes that involve iron-sulfur proteins and flavin or cytochrome coenzymes. The inhibition of Complex I ultimately results in loss of oxidative phosphorylation, so that adenosine 5' -triphosphate (ATP) levels fall rapidly and cell death ensues. Complex I is also the target of other structurally different natural and synthetic pesticides (Hollingworth and Ahammadsahib, 1995; Hollingworth and Gadelhak, 1998; Ltimmen, 1998). Papaverine, a noninsecticidal spasmolytic, was also a respiratory inhibitor in vitro with the same type of activity as rotenone (Fukami, 1976; Santi et al., 1963, 1964). Rotenone has feeding deterrent activity against various insects (Nawrot et al., 1989). Metabolism and Excretion The selectivity of rotenone appears to originate mainly from differences in the detoxification rates in various organisms and not from differences in the target enzyme system. Rotenone is effectively transformed by microsomal mixed-function oxidases in the liver, and the metabolic


pathways as well as the chemical and biological nature of the products are well defined for mammals, insects, and certain fish (Fukami et aI., 1969; Yamamoto et aI., 1971). The principal metabolites formed both in vivo and in vitro from rotenone for these species are basically the same and are, in decreasing order of mammalian toxicity, as follows: 8'-hydroxyrotenone, 6a,B, 12a,B-rotenolone, 6a,B,12aa-rotenolone, and 6' ,7'-dihydro6',7'-dihydroxyrotenone (Fukami et aI., 1967). The hydroxylated metabolites also have reduced inhibitory activity to insect and rat liver mitochondria (Horgan et aI., 1968). The formation of water-soluble conjugates, which were more abundant in mammalian than in insect tissues, was also noted in these studies. A phenolic metabolite resulting from 3-0-demethylation was also identified (Unai et aI., 1973). Synergists that block oxidative metabolism enhance the toxicity of rotenone to both insects and vertebrates. Early experiments with rabbits and dogs fed with rotenone indicated retention and/or metabolism of the toxicant (Ambrose and Haag, 1937). No intact material was obtained from the urine, but the feces contained rotenone for at least 8 days after administration. A somewhat different picture emerged for mice (Fukami et aI., 1967): 48 h after treatment with 14C-Iabeled rotenone, 20% of the 14C was in the urine, 0.3% was expired, 5% remained in the body, and the rest was in the feces. When the fate and distribution of 14C-Iabeled rotenone in different organs were followed in mice (Fukami et aI., 1967), 21.6% of the radioactivity was found in the small intestine, 19.5% in the urine, and 4.4% in the liver. In the urine, more than 82% of the product was water soluble, 17% was 6',7'dihydro-6',7'-dihydroxyrotenone, and only 1% was unchanged rotenone. In the liver and small intestine, 6',7'-dihydro-6',7'dihydroxyrotenone was the predominant metabolite. However, water-soluble products also formed in almost equal amount, and 10-16% rotenone and a few percentage of other monohydroxylated metabolites were also present. In a similar experiment with carp, about 20% of the administered rotenone could be recovered from the tissues analyzed, and the bulk of the metabolites (>45%) consisted of unidentified water-soluble products. The metabolism of rotenone in fish was studied in detail. For example, the half-life of rotenone in the head, viscera, and carcass of bluegills was about 22, 11, and 28 days, respectively, and the major metabolites identified were rotenolone and 6',7'dihydro-6',7'-dihydroxyrotenone (Gingerich and Rach, 1985; see also Gingerich, 1986; Rach and Gingerich, 1986). Notable differences exist among fish species both in vivo and in vitro: Carp, rainbow trout, and bluegill produced the various hydroxylated metabolites in different relative ratios (Erickson et aI., 1992; Gingerich and Rach, 1985). The fate of intravenous 14C-labeled rotenone (120 IJ-g/kg) in rainbow trout was studied by Gingerich (1986). After 20 min of injection, over 98% of the rotenone was cleared from the plasma and started to accumulate in the heart ventricle, lateralline red muscle, and posterior kidney, tissues that are highly dependent on aerobic metabolism, and in the liver, pyloric caeca, and small intestine. After 18 h, over 40% of the radioactivity

123

of the liver, kidney, and muscle tissues was associated with the mitochondrial fractions. The estimated half-life of rotenone in the whole body was 68.5 h, and the major metabolite was 6',7'dihydro-6',7'-dihydroxyrotenone. Environmental temperature affects the rate at which rotenone is degraded: The toxicity of a 2-ppm aqueous solution is lost in 3 days at 20°C but in 11-16 days at 11°C as determined by the survival time of the roach, Rutilus rutilus (Meadows, 1973). Biochemical Effects and Pharmacology There are numerous reports on the inhibitory effects of rotenone on contractile responses in isolated guinea pig muscle preparations, including negative inotropic and chronotropic effects on atria, blocking of barium chloride- and bradykinin-induced ileum contraction, inhibition of epinephrine-induced contraction of seminal vesicles, as well as antagonism of the chemically induced slow-reacting substance of anaphylaxis in isolated ileum (see, e.g., Ashack et aI., 1980; Haley, 1978, and the references therein). Because the dysfunction of Complex I has been implicated in the pathogenesis of Parkinson's disease, binding of rotenone and its interference with l-methyl-4-phenylpyridinium (MPP+) at this site were examined in several laboratories. Heikkila et al. (1985) reported that administration of rotenone to rat brain resulted in nigrostriatial cell death similar to that of MPP+ -caused parkinsonism. Later, Ramsay et al. (1991) showed that rotenone and I-methyl-4-phenylpyridinium share a common binding site at Complex I supporting the validity of the mitochondrial inhibition hypothesis for chemically induced parkinsonism. Similarly, chronic infusion of 2-3 mg/kg daily doses of rotenone to rats caused specific nigrostriatal dopaminergic cell degeneration and parkinsonian behavior (Betarbet et al., 2000). Since this dose is sufficient to inhibit Complex I, but too low to significantly impair oxidative phosphorylation in the brain, oxidative damage of dopaminergic neurons might be, at least partly, responsible for the toxic effects of the compound. A somewhat different conclusion was reached by Ferrante et al. (1997), who examined histologically the brains of rats treated intravenously with rotenone. They found selective bilateral lesions within the striatum and globus pallidus but not in the substantia nigra, suggesting that factors other than Complex I, for example, selective dopaminelMPP+ transporters, were also involved in the chemically induced neurotoxicity. It may be noted that rotenone was also shown to selectively and efficiently block the uptake of the neurotransmitter dopamine in nigrostriatial dopaminergic neurons (Marey-Semper et aI., 1993). Rotenone thus appears to be a unique neurotoxin. The fatty acid changes seen in the liver after chronic feeding (Haag, 1931) could be explained by the blocking ofNADH oxidation and thus the inhibition of fatty acid oxidation, leading to the accumulation of a long-chain acyl-coenzyme A (CoA; e.g., palmitoyl-CoA), which was shown for rabbit heart mitochondria (Hull and Whereat, 1967). In cultured rat hepatocytes, the mitochondrial membrane permeability transition caused by rotenone via this mechanism could be prevented by L-camitine presumably converting palmitoyl-CoA to palmitoyl

124


carnitine, which does not induce the transition (Pastorino et al., 1993). Furthermore, rotenone-induced anoxic death could be prevented in vitro by either L-carnitine or the immunosuppressive drug, cyclosporin A.

Cytotoxicity and Mutagenicity The cytotoxic and antitumor activity of rotenone is well established, but its carcinogenic potential has been a matter of controversy. On the one hand, Gosalvez (1983) found that chronic administration of rotenone, given either a total oral dose of 13.5 mg/kg during 60 days or total intraperitoneal doses of 7.1-9.1 mg/kg during 42 days, induced mammary gland tumors in female albino rats. On the other hand, several studies showed lack of carcinogenicity for pure rotenone (California DFA, 1997); in particular, a detailed investigation by Greenman et al. (1993) could not reproduce previous experimental results on mammary gland carcinogencity. Microscopic investigations with cultured mammalian cells showed that rotenone delayed cell progression in all phases of cell development and reversibly inhibited mitotic spindle microtubule assembly (Barham and Brinkley, 1976). Rotenone was found to be toxic to a number of human cancer cell lines, including solid tumor types, with EDso values ranging from 0.05 to 0.15 J.Lg/ml (Blask6 et aI., 1989). Fang and Casida (1998) recently showed that rotenone and 28 other rotenoids blocked phorbol-ester-induced ornithine decarboxylase activity in human epithelial breast cancer cells in vitro, which correlated well with the inhibition of bovine heart NADH:ubiquinone oxidoreductase activity by these compounds. These results could explain the anti proliferative effects noted earlier for cube in animals (Hansen et aI., 1965) and for rotenoids in vitro (Konoshima et aI., 1993). Rotenone treatment did not increase chromosome aberrations in Chinese hamster ovary cells (California EPA, 1996). In another test with human lymphocyte cultures, rotenone increased the frequency of binucleated micronucleated cells and caused a delay in cell cycle but did not influence chromosome aberrations and sister-chromatid exchanges (Guadano et aI., 1998). Rotenone was nonmutagenic in bacterial reversion tests (California EPA, 1996; Moriya et al., 1983). Toxicity to Animals

Acute Toxicity Results of various toxicity studies with derris and cube preparations were published (Ambrose and Haag, 1936, 1937, 1938; Cutkomp, 1943; Mathews and Lightbody, 1936; Negherbon, 1959) as were data for individual rotenoids (Fukami and Nakajima, 1971; Negherbon, 1959). Data for the acute toxicity of purified rotenone are shown in Table 3.3. Rotenone was reported to be emetic to dogs (California EPA, 1996; Lightbody and Mathews, 1936) and lethal to pigs at 3.7 mg/kg oral dose (see Oliver and Roe, 1957). Large doses of derris elicited convulsion in rabbits (Ambrose and Haag, 1936). When inhaled, derris was more toxic than pure rotenone, indi-

cating high toxicity for the other constituents of the plant (Ambrose and Haag, 1936). Derris and rotenone do not seem to be toxic to birds. For 12 different nestling birds, including chicken, pigeon, lark, sparrow, and pheasant, the oral LDso values ranged from 0.1 to 0.3 g/kg (Cutkomp, 1943). The exposure time required for death at 50 ppb rotenone concentration was 3 h for yellow perch and bluegill but 11.25 h for the more resistant common carp (Rach and Gingerich, 1986). To kill fish in lakes and reservoirs, rotenone is applied at low rates; thus, consumption of poisoned fish by either wildlife or humans is not dangerous.

Irritation Derris powder was shown to be a mild local dermal irritant. However, it produced intense, but reversible irritation to the rabbit eye as well as severe pulmonary irritation to animals when inhaled (Ambrose and Haag, 1936; Haag, 1931). Rotenone itself is less irritating. In humans, it produces a sensation of numbness when applied over the mucous membrane of the mouth (Haag, 1931). Poisoning Symptoms Depending on the route of administration, the poisoning symptoms appear between 2 and 20 min and include initial increased respiratory and cardiac rates, incoordination, clonic and tonic spasms, and muscular depression. Temperature changes are not characteristic. Later the respiration slows down until it ceases but the heart continues to beat (Haag, 1931; Santi and T6th, 1965). Death generally occurs in a few minutes after intravenous administration and within two days after intraperitoneal and oral treatments. Typical symptoms of rotenone poisoning in the common carp were observed by Fajt and Grizzle (1998). Within minutes, carp increased respiration rates and activity, such as rising to the surface and gulping air, and displayed other behavioral changes. Blood oxygen concentration, first only in arteries and after a short lag time also in veins, increased up to lO-fold, whereas while C02 concentration was unchanged. Plasma pH decreased significantly, indicating a switch from aerobic to anaerobic metabolism. Death ensued after 35-40 min at 0.1 ppm rotenone concentration. Chronic Toxicity In a 37-day toxicity study with rats, 5 or 10 mg/kg daily oral doses of rotenone decreased food intake and reduced the weight and the survival rate of the animals as compared to the control animals (Lightbody and Mathews, 1936). In a 70-week study with rats, an emulsifiable concentration (Pro-Noxfish®), containing 2.5% rotenone and 2.5% sulfoxide synergist, at 100 ppm in the drinking water showed no ill effect (Brooks and Price, 1961). In a 6-week trial, rabbits receiving food containing 150 mglkg showed no symptoms of poisoning (Haag, 1931). For rotenone at 7.5, 37.5, or 75 ppm concentration in food fed to rats and at 0.4, 2, or 10 mg/kg given orally to beagles for 26 weeks, lower food intake and lower body weight gains at the highest doses only were noted (California EPA, 1996).


125

Table 3.3 Acute Toxicity of Rotenone LDSO (mglkg)

Animal

Assay

Rat

oral

25 a

Rat

oral

150"

Rat

oral

Rat, male

oral

60 b

102

Notes, other data

Reference

In olive oil

Lightbody and Mathews, 1936

As fine crystals


LD70 = 600 mg/kga

Ambrose and Haag, 1937

In acetone

Santi and T6th, 1965

In corn oil

D.S. EPA, 1997

In corn oil

D.S. EPA, 1997

Rat, male

oral

Rat, female

oral

Rat, male

ip

1.6b


Rat

iv

6a


Rat, male

iv

0.2b


Mouse

oral

Mouse, male

ip

Guinea pig

oral

12a

In olive oil

Guinea pig

oral

50a

In starch paste


Guinea pig

oral

MLDd = 75 mg/kg

Haag, 1931

Guinea pig

oral

Guinea pig

sc

LD70 = 60 mg/kga MLD = 16 mg/kg d

Haag,1931

39.5

350

Tomlin, 1997

2.8 c

Cheng et aI., 1972 Lightbody and Mathews, 1936


Guinea pig

ip

MLD = 2.0 mg/kgd

Haag,1931

Rabbit

oral

MLD = 1500 mg/kgd

Haag,1931

Rabbit

oral

LD70 = 3000 mglkga


Rabbit

sc

MLD = 20 mg/kg d

Haag,1931

Rabbit

iv

MLD = 0.35 mg/kgd

Haag,1931

Cat

iv

MLD = 0.65 mg/kgd

Haag,1931

Dog

iv

MLD = 0.65 mg/kgd

Haag,1931

Chicken

oral

Pheasant

oral

Pigeon

oral

Pigeon

iv

Eastern robin

oral

996c 850C

Cutkomp,1943

~100c

Cutkomp, 1943 MLD = 1.0 mg/kgd

195"

Frog Bluegill

Cutkomp,1943

96-h acute

Haag, 1931 Cutkomp, 1943

LCSO =2ppm

Haag,1931

LCso = 23 ppb

Bridges and Cope, 1965

sunfish Rainbow trout

48-h acute

LCso = 28 ppb

Bridges and Cope, 1965

Rainbow trout

48-h acute

LCso = 1.2-5.8 ppm, Various formulations

Tooby et al.. 1975

Daphnia pulex

I-h acute

LC 100 = 250 ppb

Negherbon, 1959

White perch

24-h acute

Wujtewicz et aI., 1997

Crayfish

24-h acute

LClOo = 150 ppbe LCO = 3.0 ppme,f

Bluegill

24-h acute

LCso = 26 ppb

Gingerich and Rach, 1985

Bluegill

96-h acute

LCso = 11 ppbg

Gingerich and Rach, 1985

Wujtewicz et aI., 1997

aRecorded after 15 days. bRecorded after 7 days. CRecorded after 24 h. dMinimallethal dose. eDetermined at 20-23°C water temperature. fMaximum nonlethal concentration. gDetermined at BOC water temperature.

Pathology Dogs receiving rotenone at a chronic daily dietary dose of 10 mg/kg for a month displayed toxemia chiefly manifest in hepatic fatty metamorphosis and some one-third of the bulk of liver occupied by fat (Haag, 1931).

Effect on Reproduction Pregnant guinea pigs fed 150 ppm dietary rotenone had litters that were either born dead or died within 5 days after birth (Haag, 1931). The 75 ppm dietary concentration was tolerated better, but the weight gain of

126

CHAPTER 3

Pest Control Agents from Natural Products

the surviving young was slower than for the control animals. Khera et al. (1982) found that administration of a daily oral dose of 2.5 mg/kg cube extract containing 87% rotenone to female rats on days 6-15 of gestation was without effect, but the 5 mg/kg dose resulted in increased frequency of skeletal aberrations of the fetuses such as extra rib and delayed ossification of sternebrum. The 10 mg/kg dose reduced maternal body weight, 60% of the dams were killed, and it also produced a high incidence of re sorption in the surviving dams. A series of reproductive studies with 95-98% pure rotenone were summarized (California EPA, 1996). In a chronic feeding study with male and female rats, the effects of dietary rotenone on generations Fo-F2, including breeding, gestation, lactation, and weaning, were observed. At 75 ppm, the highest concentration tested, decreased mean live litter sizes in the Fo and F 1 generations decreased mean birth weights of the F 1 and F2 pups at 37.5 and 75 ppm, and decreased body weight gain of the FJ and F2 pups at 37.5 and 75 ppm were observed. In a related 3-month one-generation study with Syrian golden hamsters, dietary rotenone at 1000 ppm concentration resulted in smaller testicles and infertility; at 500 ppm and above, poor litter survival, maternal death, and cannibalism were observed. In a teratology study with mice, involving rotenone at a 24 mg/kg/day oral dose, given on days 6-17 of gestation, no maternal toxicity was observed, but decreased live litter size and increased fetal resorption were noted. Treatment Extending the observations of Ernster et al. (1963); see also Wijburg et al. (1990) that inhibition by rotenone of rat liver mitochondrial respiration in vitro could be overcome by the addition of catalytic menadione (vitamin K3), Santi and T6th (1965) demonstrated that an intravenous dose of2.5 mg/kg menadione could rescue a rabbit from a 0.4 mg/kg lethal intravenous dose of rotenone (see also T6th et aI., 1966). Human Poisoning Incidents Acute rotenone poisoning is rare. De Wilde et al. (1986) described a fatal case of a 3.5year-old girl who accidentally ingested approximately 0.6 g rotenone in about 10 ml of mixed ethereal oil insecticide formulation. Half an hour after ingestion, the girl vomited and felt drowsy. The initial symptoms rapidly developed to slow and irregular respiration, coma, and apnea. In spite of artificial respiration and gastric lavage, the girl died of respiratory arrest about 8.5 h after ingestion. Significant postmortem findings included anoxic hemorrhages in the lung, heart, and thymus; anoxic damage to the cerebrum; bloody stomach content; and renal tubular necrosis, although the latter was suggested to be caused by the ethereal oils in the formulation. Analysis for rotenone in various tissues showed 1260 ppm in the stomach content, and 2-4 ppm in the blood, liver, and kidney but none in the brain, muscle, and thymus. The estimated oral dose was 40 mg/kg. It is noteworthy that the label on the insecticide, which was recommended for plants and external use on animals, stated "Natural Product-Non Toxic." Rotenone was once also a common means of suicide in some parts of New Guinea. Victims taking small dosages usually

recovered after gastric lavage and stimulants. For fatal cases, vomiting before death was usual and autopsy revealed acute congestive heart failure (Holland, 1938). 3.2.1.5 Sabadilla Alkaloids Introduction Liliaceous plants are known both for their poisonous properties and for their medicinal value. The crushed seeds of Schoenocaulon officinale (earlier Veratrum sabadilla), indigenous to the northern region of South America and Central America, is known as sabadilla powder and was used by the American Indians in pre-Columbian times as an insecticide (e.g., louse powder) (Crosby, 1971). The seeds contain 2-4% of a biologically active alkaloid mixture, called veratrine. Veratrine was first isolated by Pelletier and Caventou in 1819, and since then it has been the subject of extensive chemical and pharmacological studies. Vigorous alkaline hydrolysis of veratrine affords the steroid alkamine cevine, the structure of which was elucidated in 1954 (Barton et aI., 1954). Veratrine is a mixture of cevadine and veratridine (Fig. 3.4), present in an approximately 2: 1 ratio, which are esters of the same, practically inactive veracevine (i.e., the 3,B-epimer of cevine), obtainable from veratrine by hydrolysis under controlled conditions. Sabadilla alkaloids belong to the ceveratrum group of Veratrum alkaloids characterized by a C-nor-D-homosteroid skeleton containing 6-8 hydroxyl groups of which at least one is esterified (Kupchan and Flacke, 1967). Alkaloids of the jerveratrum group contain only 1-3 oxygens and occur as free alkamines or as glycosides in plants. Identity and Physicochemical Properties

Cevadine Chemical Abstract name: [3,B(Z),4a,16,B]-4,9-epoxy-cevane3,4,12,14,16,17,20-hepto13-(2-methyl-2-butenoate) CAS Registry Number: [62-59-9] Empirical formula: C32H49N09; molecular weight: 591.7 The melting point of cevadine is 208-21O°C (decomposition); the optical rotation is [a]~} = +10.7° (c = 6.0 in ethanol).

veracevine R = H cevadine R = (Z)-CH,CH=C(CH,)CO veratridine R = 3,4-(CH,O),PhCO cevaclne R = CH,CO 3-O-vaniiloylveracevine R = 3-CH,C>-4-0H·PhCO

Figure 3.4 Structures of sabadilla alkaloids.

3.2 Insect Control Agents Veratridine Chemical Abstract name: [3,B,4a,16,B]-4,9-epoxy-cevane-3,4, 12,14,16,17 ,20-hepto13-(3,4-dimethoxybenzoate) CAS Registry Number: [71-62-5] Empirical formula: C36HSINOll; molecular weight: 673.8 The melting point of veratridine is 167-184°C (decomposition); the optical rotation is [a]51 = +7.2° (c = 3.9 in ethanol). Veratridine is a weak base with a pKa of 9.54 The solubility of veratridine in a 150 mM aqueous NaCl solution is 12.5 g/l at pH = 8.07. The sabadilla alkaloids are freely soluble in dilute acids but decompose in solutions with a pH greater than 10. They are readily soluble in alcohols, ether, and chloroform but not in hexane. The CAS Registry Number of sabadilla or the veratrine mixture is [8051-02-3]; CAS Registry Numbers of minor sabadilla alkaloids: veracevine: [5876-23-3]; cevacine: [28111-33-3]; sabadine: [124-80-1]; 3-0-vanilloylveracevine: [187237-90-7]. History, Formulations, and Uses Powdered rhizomes of the white hellebore (Veratrum album), growing in Europe and Asia, and the "green (or false) hellebore" or Indian poke (V. viride), indigenous to the eastern part of North America, were used to cure herpes, toothache, rheumatism, and catarrh, and drugs from these plants were also important hypotensive agents although of low therapeutic index (Kupchan and Flacke, 1967). Preparations from hellebore roots were once commercial insecticides against the hemipteran and homopteran pests of fruits and vegetables (Fisher, 1940; Shepard, 1951). Due to its low persistence and compatibility with beneficial insects (Bellows and Morse, 1993), sabadilla formulations reappeared in the late 1970s and are now used against thrips in citrus and avocado (Hare and Morse, 1997). Sabadilla is formulated as dust, wettable powder, or watersoluble concentrate, which might contain sugar as an insect feeding stimulant, with 0.2-25% alkaloid content. Typical use rates are 20-100 g total alkaloidlha.

Stability When exposed to air and sunlight, sabadilla formulations rapidly lose activity. Field trials in a citrus plantation recently showed that the alkaloid level on leaves declined to 60% of the initial deposit within 20 h after spraying. Veratridine persisted somewhat longer than cevadine; nevertheless, degradation and rainfall resulted in no residual toxicity to citrus thrips 7 days after treatment (Hare and Morse, 1997). Biological Properties

Mode ofAction Cevadine, veratridine, and related ceveratrum alkaloids activate the voltage-sensitive Na+ channels of nerve, heart, and skeletal muscle cell membranes similar to pyrethrins (see Section 3.2.1.1). The receptor for these alkaloids has not been isolated, but various experiments indicate they evidently bind at a site distinct from that of pyrethrins. Both veratridine

127

and cevadine cause persistent activation and alter the ion selectivity of Na+ channels and share a binding site (site 2) with the botanical steroid alkaloid aconitine, the frog steroid alkaloid batrachotoxin, and the botanical diterpenoid grayanotoxins (reviewed by Bloomquist, 1996; Catterall, 1980; Honerjager, 1982). Structurally, veratridine and cevadine differ only in their acyl group (RCO in Fig. 3.4) and this difference is enough to cause quantitative and qualitative variations in insecticidal activity (AlIen et aI., 1945; Ikawa et aI., 1945; Ujvary et aI., 1991), mammalian toxicity (Swiss and Bauer, 1951; Ujvary et aI., 1991) as well as in their pharmacological (Honerjager et aI., 1992; Mendez and Montes, 1943;) and electrophysiological (Leibowitz et aI., 1987; Nanasi et aI., 1990; Ohta et aI., 1973; Shanes, 1952) properties. Physiological and Pharmacological Activities Concoctions from Veratrum, Schoenocaulon, and Zygadenus genera have been used for centuries in the treatment of fever and circulatory disorders, and their pharmacology has been thoroughly investigated (Honerjager, 1982; Krayer and Meilman, 1977; Kupchan and Flacke, 1967). Purified veratridine, which recently replaced veratrine mixture, is a commonly used and well characterized pharmacological tool in various electrophysiological studies. The ceveratrum alkaloids have a characteristic hypotensive effect not directly involving the CNS: They slow the heart and lower arterial blood pressure by reflexly stimulating medullary vasomotor centers without decreasing cardiac output (Bezoldlarisch effect). The electrophysiological aspects of the cardiac activity of veratridine and related steroid alkaloids were discussed in detail (Honerjager, 1982; Honerjager et aI., 1982). The alkaloids also affect respiration; low doses induce tachypnea and higher doses cause respiratory depression and apnea. Ceveratrum alkaloids increase salivation and have a reflex emetic effect in mammals able to vomit. The low therapeutic index of the ceveratrum alkaloids limits their use as hypotensive agents. The acute hypotensive action of Veratrum alkaloids can be blocked by barbiturates and the reflex action is counteracted by atropine (Frey and Weigmann, 1943). It was also established that local anesthetics, including cocaine (Matthews and Collins, 1983; Zimanyi et aI., 1989; see also Ragsdale et aI., 1994) and procaine (Bir6 and Gabor, 1969; Nishizawa et aI., 1988; Ohta et aI., 1973) antagonize veratrine action in vitro. Because ischemic conditions produce simultaneous Na+ and Ca2+ influx through cardiac Na+ channels, veratridine-induced intoxication was recently suggested as an experimental model of ischemia in animals (see Wermelskirchen et aI., 1991). Lakics et al. (1995) showed that cell death evoked by veratridine in rat cerebrocortical cell cultures could be inhibited by submicromolar concentrations of the cerebroprotective agent vinpocetine more effectively than by the prototype Na+ channel blocker anticonvulsant phenytoin. The Ca2+ entry blocker flunarizine also provided protection against neuronal damage induced by veratridine (Pauwels et aI., 1989).

128

CHAPTER 3


A broad range of pharmacological and biochemical effects of veratridine, including the release of various neurotransmitters such as ACh, norepinephrine, dopamine, and GABA, were shown to be triggered by the alkaloid-elicited increase in Na+ permeability of neural cells in vitro (see Cunningham and Neal, 1981). Veratrum alkaloids, in general, are easily absorbed through the skin and mucous membranes and upon ingestion. The metabolism and excretion of veratridine and cevadine have not been studied. On oral administration to humans of a structurally related steroid alkaloid (3-0-acetylgermine), slow urinary excretions of both the parent ester and its hydrolytic product, germine, were observed (Biich, 1971). In insects, the toxicity of sabadilla alkaloids can be synergized by the cytochrome P-450 inhibitor PB, indicating oxidative detoxification (Blum and Keams, 1956; Ujvary et aI., 1991).

Metabolism and Excretion

Toxicity to Laboratory Animals Acute toxicity data for the natural sabadilla alkaloids are listed in Table 3.4. Intravenously administered veratridine kills mice usually within 4 min; on intraperitoneal injection, the alkaloid causes death within 1.5 h after injection due to respiratory failure. Salivation and cyanosis are also notable and a short period of convulsion can precede death. The poisoning

Acute Toxicity

Table 3.4 Acute Toxicity of Sabadilla Alkaloids

LD50 Animal

Assay

(mg/kg)

Reference

Veratrine Mouse

ip

7.5 a

Swiss and Bauer, 1951

Mouse

ip

8.5 h


Rat

ip

4.8

Mendez and Montes, 1943

ip

>100

Ujvary et aI., 1991

Veracevine Mouse Cevadine Mouse

ip

3.5


Mouse

ip

5.8

Ujvary et aI., 1991

Rabbit

sc

0.5-1.3

Krayer et aI., 1944

Frogs

sc

1.5-30

Krayer et aI., 1944

Veratridine Mendez and Montes, 1943

Rat

ip

3.5

Mouse

ip

0.42

Krayer et aI., 1944

Mouse

ip

1.35


Mouse

ip

9.0

Ujvary et aI., 1991

symptoms of other Veratrum alkaloids also unfold almost immediately, but paralysis followed by severe convulsions could occur (Krayer et aI., 1944; Mendez and Montes, 1943). Teratogenicity The ceveratrum-type sabadilla alkaloids do not appear to be teratogenic. Other, especially the jerveratrum alkaloids, are, however, teratogenic. The bizarre birth defects (cyclopia) in livestock grazing plants such as V. californicum (Keeler, 1986, 1988) are caused by inhibition of cholesterol biosynthesis and transport in embryos by certain Veratrum alkaloids of the jerveratrum group (Cooper et aI., 1998).

Toxicity to Humans Sabadilla and related Veratrum alkaloids have a stemutatory action, a property that was nevertheless a cause of several bizarre intoxications among children. A recent report summarized seven such poisoning cases and warned of the potential danger of V. album-based sneezing powders marketed in some European countries in the 1980s (Carlier et aI., 1983; Fogh et aI., 1983). Eye contact could result in severe irritation, lacrymation, and inflammation of the conjunctiva.

Irritation

Despite the long and widespread use of ceveratrum alkaloids for the treatment of tachycardia and various circulatory disorders, especially hypertension, no fatal poisoning has been documented (Krayer and Meilman, 1977; Rokin and Kustovskii, 1997). Documented sabadilla poisoning cases are also rare (see Ray, 1991). Because sabadilla alkaloids and the steroid alkaloids of other Veratrum plants have a similar mode of action, poisoning cases reported for the latter could be instructive. Veratrum plants are sometimes mistaken for other edible plants causing intoxication. laffe et al. (1990) described six poisoning cases due to accidental ingestion of V. viride. The typical signs of intoxication are nausea, vomiting, diaphoresis, bradycardia, and hypotension developing from 0.5 to 4.5 h after ingestion. A similar V. album poisoning case was described by Festa et al. (1996). Quatrehomme et al. (1993) surveyed 32 Veratrum poisoning cases, one of which was an accidental ingestion of veratrine antilouse preparation.

Poisoning Incidents

Treatment The treatments employed in Veratrum alkaloid poisoning cases are supportive (Jaffe et aI., 1990). Charcoal and a cathartic can be administered after nausea and vomiting have subsided. Seizures, if they occur at all, may be treated with the usual anticonvulsants. Bradycardia is responsive to intravenous atropine, and hypotension could be treated with dopamine or metaraminol. The patients usually recover within 48 h.

3-0- Vanilloylveracevine Mouse

ip

>SC

Merck & Co. bFrom S. B. Penick & Co. cNo mortality at this exploratory dose (Ujvary and Casida, unpublished observations). a From

3.2.1.6 Ryania

Introduction The botanical insecticide ryania is the ground stem wood of Ryania speciosa (Flacourtiaceae), a tropical tree growing in Central America and the Amazon Basin (Crosby, 1971). Related Ryania species in that region were used as


OH

ryanodoJ R = H ryanodine R =

Qy H

Figure 3.5

0

Structures of ryanodine and ryanodol.

the source of arrow poisons. Ryanodine was isolated from the roots and the stemwood of R. speciosa by Rogers et al. (1948), and the structure of the alkaloid was established by Wiesner (1972) as the pyrrolecarboxylate ester of the diterpene ryanodol (Fig. 3.5). Recently, another major insecticidal component, 9,21-didehydroryanodine, was also isolated from the wood (Waterhouse et aI., 1984). The total active alkaloid content of ryania insecticide is about 0.22%. Identity, Physicochemical Properties, and Uses Chemical Abstract name of ryanodine: ryanodol 3-( 1H -pyrrole-2-carboxylate) Chemical Abstract name of ryanodol: [3S-(3a,4,'I,4aS*,6a, 6aa,7a,8,'I,8aa,8b,'I,9,'I,9aa)]-hexahydro-3,6a,9-trimethyl7-( I-methylethyl)-6,9-methanobenzo[ 1,2]-pentaleno[ 1,6bc ]furan-4,6, 7 ,8,8a,8b,9a,(6aH,9 H)-heptol CAS Registry Numbers: ryania: [8047-13-0]; ryanodine: [15662-33-6]; 9,21-didehydroryanodine: [94513-55-0]; ryanodol [6688-49-9]; 9,21-didehydroryanodol: [106821-54-9] Empirical formula of ryanodine: C2SH3S09; molecular weight: 493.6 Ryanodine melts with decomposition and the reported melting points are 219-220°C (Rogers et aI., 1948) and 235237°C (Waterhouse et aI., 1984). Ryanodine is soluble in water, ethanol, acetone, ethyl ether and chloroform, but practically insoluble in hexane. Ryanodine is dextrorotatory: [a]¥; = +26° (c = 1.02 in methanol). Stability Ryania is relatively stable on exposure to light and air and has a longer residual activity in the field than rotenone and the pyrethrins. Formulations and Uses Commercial ryania is formulated as a dust or water-dispersible powder. Ryania is a slow-acting contact and stomach insecticide and is often mixed with other botanical preparations, but its use is now limited. The usual application rates of ryania are 10-72 kg/ha (20-145 g alkaloid equivalent/ha) against lepidopteran pests (Crosby, 1971).

Biological Properties Mode of Action Reviews on ryanodine pharmacology in mammals (Sutko et aI., 1997) and in insects (Iefferies and

129

Casida, 1994; Schmitt et aI., 1996) as well as on receptor structure (Coronado et aI., 1994) are available. Ryanodine acts by binding to a family of intracellular Ca2+ release channel proteins, the so-called ryanodine receptors (RyRs), associated mainly with the sarcoplasmic reticulum of skeletal and cardiac muscles but also detected in the brain and liver. The alkaloid at submicromolar concentrations locks the Ca2+ channels in a fractional conductance state, whereas at higher concentrations the channel is transformed to a nonconducting, closed state. The RyRs are large polypeptides with a molecular weight of 550-565 kDa, forming the ion channel in a homotetrameric arrangement (Takeshima et aI., 1989). Different RyR isoforrns were isolated from various organisms and, in some cases, also from different cell types of the same organism. Biochemical Effects and Pharmacology In the cat, intraarterial injection of 50 J.lg of the alkaloid induced tetanic contractures in skeletal (tibialis anticus) muscle, which could be abolished by intraarterial injection of 1 mg atropine (Procita, 1956). At nanomolar concentrations, ryanodine caused an irreversible loss of contractile tension in cat cardiac muscle that could be temporarily reversed by sympathomimetics or excess calcium (Hill yard and Procita, 1959). The potency of ryanodine (ICso = 10 nM) and analogs in blocking the rabbit skeletal muscle RyR proved to be a good indicator of their toxicity to mice, establishing the toxicological relevance of these assays (Iefferies et al., 1992; Pessah et aI., 1985; Waterhouse et aI., 1987; see also Sutko et aI., 1997). In studies in vitro, millimolar concentrations of Mg2+ inhibited, whereas caffeine stimulated eHJryanodine binding to skeletal receptors. Cardiac receptors were only slightly affected by these agents. The binding of eH]ryanodine to both skeletal and cardiac sarcoplasmic reticulum was stimulated by increasing the pH, temperature, or concentration of NaCl or KCl (see Coronado et aI., 1994). Metabolism and Excretion Except for one report showing no detectable insecticide in the urine of rats receiving a 600 mg/kg oral dose of ryania powder (Kuna and Heal, 1948), the metabolism of ryanodine and other ryania alkaloids in mammals has not been studied.

Toxicity to Laboratory Animals Acute Toxicity The acute toxicity of ryania powder and pure ryanodine to various animals is shown in Table 3.5. The pharmacological actions of ryanodine to various mammalian species were studied by Procita (1958). In the dog, for example, injection of lethal doses (50-300 J.lg/kg) of ryanodine produced enophthalmos, followed by general spastic muscle rigidity, salivation, vomiting, and defecation. Spastic rigidity then progressed and death was due to asphyxiation, although in some cases cardiac arrest could also be observed. A somewhat different picture emerged when animals were anesthetized with pentobarbital: A single lethal dose of ryanodine, without affecting other muscles, produced circulatory failure before respira-

130


Table 3.5 Acute Toxicity of Ryania and Ryanodine

Animal

Assay

LD50 (mg/kg)

Other data

Reference

Ryania Rat

oral

Mouse

oral

Guinea pig

oral

Rabbit

oral

Dog

oral

Monkey

oral

Chicken

oral

1200 650 2500 650 150 >400 >3000

Kuna and Heal, 1948 Kuna and Heal, 1948 Kuna and Heal, 1948 Kuna and Heal, 1948 Kuna and Heal, 1948 Kuna and Heal, 1948 Kuna and Heal, 1948

Ryanodine Mouse

ip

0.10

Waterhouse et al., 1987

Rat

ip

0.32a

Procita, 1958

Mouse

ip

Guinea pig

ip

0.26 0.21a

Procita, 1958

Rabbit

iv

0.026a

Procita, 1958

Cat

ip

0.071 a

Dog

oral

Dog

iv

Procita, 1958

Procita, 1958 LD

0.075

LD

mg/kg b

= 0.4 = 0.1 mg/kgb

Procita, 1958 Procita, 1958

aCalculated from mortality data. bLethal dose.

tory difficulties arose and artificial respiration was of no value. The hypotension produced by sublethal doses of ryanodine was unaffected by atropine and cardiac glycosides but epinephrine reversed normal circulation. Ryanodol, the hydrolysis product of ryanodine, has low toxicity to mice (intraperitoneal LD50 > 20 mg/kg) and little activity at the mammalian sarcoplasmic reticulum RyR (ICso = 35 !-1M), yet it is a potent insecticide (Waterhouse et aI., 1987), demonstrating that selective toxicity could be achieved among ryania alkaloids. Whether the selectivity is due to speciesdependent differences in distribution, detoxification, or properties of the Ca2+ channel is not clear (Jefferies and Casida, 1994). It was also demonstrated that both ryanodine and the non-ester 9,21-didehydroryanodol, which is insecticidal but has low mammalian toxicity to mice (intraperitoneal LDso > 20 mg/kg), affected the excitability and ion selectivity of K+ channels in insects but not in mammals (Usherwood and Vais, 1995; Vais et al., 1996). Ryanodine influences acetylcholine receptor synthesis by modulating the cytoplasmic Ca2+ level in a concentrationdependent manner. At (sub)micromolar concentrations, the alkaloid decreased, whereas, while at millimolar concentrations, it increased the synthesis of acetylcholine receptor in cultured chicken muscle cells (Pezzementi and Schmidt, 1981). Chronic Toxicity and Pathology Kuna and Heal (1948) studied the chronic toxic effects of the powdered stemwood of R. speciosa on various animal species. Rats, guinea pigs, and chickens remained symptomless for 5 months when fed a diet containing 1% ryania powder. Rats fed 5% ryania showed de-

creased weight gain, and some deaths occured within 25 days after the start of the treatment. Rats exposed to ryania dust or a spray of 1% aqueous ryania suspension 8 h daily for 22 days did not show any signs of toxicity. Dogs, however, showed signs of toxicity (vomiting, irritated eyes and respiratory passages) when exposed to ryania powder for 2 h. Autopsy of rats receiving 2-5% ryania powder in the diet revealed hemorrhages in the pancreas and intestinal tract, pulmonary complications, and pleural exudation. Treatment Based on the similarities between some of the effects of ryanodine and malignant hyperthermia (MH), a potentially fatal genetic disorder of skeletal muscle characterized by uncontrolled Ca2+ release involving ryanodine receptors, ryanodine toxicity is considered to be a model of MH. Thus, the muscle relaxant dantrolene used for the clinical treatment of MH could protect mice and rats from lethal intraperitoneal doses (135 !-1g/kg) of ryanodine (Fairhurst et aI., 1980; see also Fruen et aI., 1997). 3.2.1. 7 Azadirachtin Introduction The neem tree, Azadirachta indica (Meliaceae), also known as "nim" or "margosa," is indigenous to the arid parts of India and Burma and is now grown in Africa and other tropical and subtropical regions of the world. Indian folklore and medicinal literature, including "Ayurveda," consider it a miracle tree. Traditional neem preparations from all parts of the tree, including the bark, roots, flowers, and seeds, have been used for centuries for medical, agricultural, hygienic, and cos-


~2,JL

,i

Its log P value is 1.09. Azadirachtin is levorotatory: [a]D -53° (c = 0.5 in chloroform).

1

o

,)lo"'" /

o

o azadirachtln

nimbin

saianoin

Figure 3.6

131

Structures of azadirachtin and other limonoids from the neem tree.

metic purposes. Reviews on the medical and botanical aspects of neem use were collected in a book by Randhawa and Parmar (1993). Neem seeds can be pressed to give 20-40% neem oil, a dark-yellow liquid of disagreeable, garlic-like odor. The oil has a bitter taste due to the presence of some 2% limonoid constituents. Azadirachtin (or azadirachtin A) (Fig. 3.6), a potent insect feeding inhibitor (antifeedant), was isolated as the major bitter limonoid from the seeds of A. indica (Butterworth and Morgan, 1968). Dried seed kernels may contain up to 0.7% azadirachtin. The full structure of the highly oxygenated tetranortriterpenoid was elucidated fully in 1987 (BiIton et al., 1987; Kraus et al., 1987; Turner et al., 1987). The chemistry of azadirachtin and related limonoids was thoroughly reviewed (Champagne et al., 1992; Devakumar and Sukh Dev, 1993; Ley et al., 1993).

Identity, Physicochemical Properties, and Uses Chemical Abstract name: [2aR,-[2aa,3,B,4,B(laR*,2S* ,3aS*, 6aS*,7 S* ,7aS*),4a,B,5a,7aS* ,8,B(E),IO,B, lOaa, IOb,B]]-IO(acetyloxy)octahydro-3,5-dihydroxy-4-methyl-8-[(2methyl-l-oxo-2-butenyl)oxy]-4-(3a,6a, 7, 7a-tetrahydro-6ahydroxy-7a-methyl-2,7-methanofuro[2,3-b ]oxireno [e ]oxepin-la(2H)-yl)-IH,7 H -naphtho[1,8-bc:4,4a-c'] difuran-5, lOa(8H)-dicarboxylic acid dimethyl ester CAS Registry Number: [11141-17-6] CAS Registry Numbers of some related compounds: 22,23-dihydroazadirachtin: [108189-58-8]; 2' ,3' ,22,23-tetrahydroazadirachtin: [108168-76-9]; nimbin: [5945-86-8]; nimbolide: [25990-37-8]; salannin: [992-20-1] Empirical formula: C3SH44016; molecular weight: 720.7 Physicochemical Properties The meIting point is 155-158°C; at 20°C, azadirachtin is slightly soluble in water, readily soluble in ethanol, ethyl ether, acetone and chloroform, but insoluble in hexane.

=

Stability Crystalline azadirachtin is a relatively stable substance if stored in the dark. Its laboratory half-life in mildly acidic solutions (pH 4 and 5) is about 50 days at room temperature, but rapid decomposition occurs at higher temperatures and in alkaline and strongly acidic media (Jarvis et al., 1998; Sundaram et al., 1995; Szeto and Wan, 1996). Azadirachtin is light sensitive (Ermel et al., 1987), but the (photo )degradation products retain some insecticidal activity (Bamby et al., 1989). Neem formulations retain their azadirachtin content for at least a year when stored at 25°C. Studies on the behavior of various azadirachtin formulations in the environment were recently reviewed (Sundaram, 1996). In the field, the residual life of azadirachtin in neem seed kernel extracts is 8-10 days; commercial formulations, however, usually contain stabilizers that retard both hydrolytic and photodegradation. The half-life of azadirachtin on foliage could be as short as 17 h, whereas in the soil, due to the absence oflight, the half-life could be as high as 25 days. Two more stable hydrogenated derivatives of azadirachtin, namely 22,23-dihydro- and 2',3' ,22,23-tetrahydroazadirachtin (Bamby et al., 1989; Immaraju et al., 1994; V.S. EPA, 1997), are also used in insect control. Formulations and Uses The most important neem products traditionally used in agriculture are aqueous neem seed kernel and leaf extracts, alcoholic seed kernel and leaf extracts, enriched and formulated seed kernel extracts, and neem seed oil and neem seed cake that remains from the kernels after pressing the oil there from. Neem seeds and, consequently, their extracts, from different geographic regions were shown to vary considerably in their limonoid composition (Ermel et al., 1987; Isman et al., 1990; Stark and Waiter, 1995). The commercial and standardized formulations are mostly refined extracts made from neem oil and sold as emulsifiable liquid concentrates or powder with 0.3-20% specified azadirachtin content. Azadirachtin-based products are used against a broad range of insect pests in orchards, vegetables, mushrooms, herbs, tea, coffee, cotton, turf, and ornamentals as well as for disease vector control. Typical use rates are low (10-40 g/ha), but due to the instability of the active ingredient(s) frequent applications may be required (Immaraju, 1998).

Biological Properties Mode of Insecticidal Action Neem extracts and azadirachtin are nonneurotoxic pest control agents exhibiting unparalleled insect selectivity. In spite of considerable research efforts, the mode of action of azadirachtin has not been clarified at the cellular or biochemical level. Both neem oil extracts and pure azadirachtin demonstrated a wide range of physiological and behavioral activities in insects, and the subject has been reviewed extensively (Ascher, 1993; Mordue and Blackwell,

132


1993; Schmutterer, 1987, 1990). Neem preparations have deterrent or antifeedant activities against many insect species, and azadirachtin inhibits feeding at 0.01-1 ppm concentrations. More important, azadirachtin markedly affects insect metamorphosis and reproduction, including fecundity, but these effects manifest slowly. Depending on the dose, azadirachtin causes growth inhibition, malformation, and mortality in insect larvae. The steroid-like compound disturbs insect development, apparently by interfering with the release or action of ecdysteroids and/or other hormonal regulators of insect molt (see, e.g., Marco et aI., 1990). The insect toxicity of azadirachtin, however, cannot be entirely explained by its effect on the endocrine system alone. Because commercial neem formulations contain not only azadirachtin but also other minor, but potentially bioactive limonoid components, the insecticidal effect of the preparation is more complex than that observed for pure azadirachtin. Neem preparations have other agriculturally important biological activities, including nematicidal (Alam, 1993) and antifungal (Parveen and Alam, 1993) effects.

fertility in cycles subsequent to treatment (Mukherjee and Talwar, 1996). Single intrauterine administration of 100 J.LI neem oil caused lasting infertility by apparent induction of leukocytic infiltration in the uterine epithelium during the preimplantation period (between days 3 and 5 postcoitum). Fertility was regained 5 months after treatment without apparent teratogenic effects (Upadhyay et aI., 1990). Daily intramuscular injection of 250 and 500 mg/kg doses of neem oil for 8 days to male rats caused significant decreases in sperm counts, epididymial weight, and glycogen levels; reduced acid phosphatase and influenced lactate dehydrogenase activities; and increased alkaline phosphatase activity. Marked structural changes in the testes and impaired spermatogenesis were also observed. It was suggested that neem oil impaired the androgen supply to the testes and epididymial tissues (Manoranjitham et aI., 1993; Sampathraj et aI., 1993). Investigations of the antifertility (spermicidal) property of neem oil components in rats and humans culminated in the commercialization of human contraceptive formulations in India (see Jacobson, 1989; Riar and Alam, 1993; Talwar et aI., 1997).

Metabolism and Excretion Information on the fate of azadirachtin in animals is scarce. Injection of a tritiated azadiToxicity in Animals rachtin derivative, [22,23- 3H]dihydroazadirachtin into locusts indicated fast clearing of radiolabeled material from the blood. Acute and Chronic Toxicity As with the pharmacological Ninety percent of the applied radioactivity was excreted during studies, toxicological evaluations rarely used pure ingredients 3 the first 7 h with the feces, whereas the remaining [22,23- H]dibut tested neem oils, extracts from various parts of the tree, or hydroazadirachtin accumulated in the Malpighian tubules where formulated insecticide products instead (Jacobson, 1989; Kait could be detected even 24 days after treatment. After the first 24 h, feces contained at least three polar, unidentified metabo- nungo, 1993). The acute toxicity values of several neem prepalites but no [22,23- 3H]dihydroazadirachtin (Rembold et aI., rations and some for pure azadirachtin for laboratory animals and some nontarget species are listed in Table 3.6. For addi1984, 1988). tional data on nontarget aquatic and terrestrial organisms, see Effects of Neem Preparations on Mammals The pharma- Darvas and Polgar (1998) and Kreutzweiser (1997). Neem seed-based animal feed supplements were found to be cological and toxicological properties of neem extracts, especially neem oil that is a commercial product in India, have safe to chicken and cattle but not to sheep. Furthermore, aflabeen studied extensively. The broad spectrum of activity of toxin contamination of improperly treated seeds is a potential neem products includes anti-inflammatory, antipyretic, anal- hazard (Jacobson, 1989; see also Hansen et aI., 1994). Neem gesic, cardiovascular, hypoglycemic, immunostimulant, derma- leaf was reported to be toxic to sheep (Ali and Sa1ih, 1982), tological, antimicrobial, antimalarial, and antifertility effects goats, and guinea pigs (Ali, 1987) but not to rabbits (Thompson (reviewed by Dhawan and Patnaik, 1993; Jacobson, 1989; and Anderson, 1978). Based on a three-generation toxicology study with rats, debitterized neem oil, which is obviously deKanungo, 1993). pleted of limonoid ingredients, was recommended as suitable Effects on Reproduction The antifertility activity (and thus for human consumption (Chinnasamy et aI., 1993). Methanolic extracts of neem leaf and bark had oral LDso the potential contraceptive application) of neem preparations has received special interest. In one of the early studies (Sadre values of about 13 mg/kg in mice, and the poisoning signs were et al., 1984), daily oral doses of neem leaf extract to male discomfort, gastrointestinal spasms, loss of appetite, hypothermice and rats did not affect normal development but caused re- mia, and, ultimately, convulsion leading to death within 24 h versible infertility. Spermatogenesis was not affected and the (Okpanyi and Ezeukwu, 1981). The subacute toxicity of Vepacide, an enriched neem oilobserved infertility was attributed to the decreased motility of spermatozoa. The extract, however, was toxic to guinea pigs based preparation containing 12% azadirachtin plus additional and rabbits. By contrast, neem seed kernel extracts lacked ac- terpenoids, upon oral administration of 80, 160, and 320 mg/kg daily insecticide doses for 90 days was studied in male rats tivity in a similar test with rats (Krause and Adami, 1984). Oral administration of neem seed extract to female rats (Mahboob et aI., 1998). On the 90th day, the high and medium from days 8 to 10 of pregnancy caused complete resorption doses caused significant decreases in (1) cytochrome P-450 of embryos by day 15 of pregnancy and the animals regained concentration in the liver, lungs, and kidneys but not the brain;


133

Table 3.6 Acute Toxicity of Neem Preparations

Test material

Species, route

Neemrich I

Rat, oral

Neemoil

Rat, oral

Margosan-Ob

Rat. oral

NeemAzal- T/Sc

Rat. oral

Neemrich I

Rat, dermal

Vepacided

Rat, oral

Leaf and bark extract

Mouse, oral

Neemrich I

Mouse, oral

NeemAzal-T/Sc

Mouse, oral

LDSO (g/kg)

Other data

Sharma et aI., 1984

8.7 14a

Gandhi et aI., 1988

>5 a

Larson, 1989 Trifolio-M, 1995

>10

Sharma et al., 1984

11.2

Mahboob et aI., 1998

1.57

Okpanyi and Ezeukwu, 1981

::::13

Sharma et aI., 1984

6.8

Trifolio-M, 1995

>10 24a

Gandhi et aI., 1988

Neemoil

Rabbit, oral

Margosan-Ob

Rabbit, dermal

LCso > 2 ml/kg

Margosan-Ob

Rabbit, I-h inhalation test

LCSO > 43.9 mg/1

Neemrich I

Chicken, oral

39.9 6.3

Neemrich I

Pigeon, oral

Margosan-Ob

Bobwhite quail, 5-day

Margosan-Ob Margosan-Ob

Reference

Larson, 1989 Larson, 1989 Sharma et aI., 1984 Sharma et al., 1984

LCso > 7000 ppm

Larson, 1989

Mallard duck, 5-day

LCso > 7000 ppm

Larson, 1989

Rainbow trout, 96-h

LCso = 8.8 ppm

Larson, 1989

Margosan-Ob

Rainbow trout, 96-h

LCso = 29ppm

Wan et aI., 1996

Azatine

Rainbow trout, 96-h

LCso =4ppm

Wan et aI., 1996

Azadirachtin f

Rainbow trout, 96-h

LCso =4ppm

Wan et aI., 1996

Azadirachting

Rainbow trout, 96-h

LCso =61 ppm

Wan et aI., 1996

Margosan-Ob Azatine

Coho salmon, 96-h

LCso = 38ppm

Wan et aI., 1996

Coho salmon, 96-h

LCso =5 ppm

Wan et aI., 1996

Azadirachting

Coho salmon, 96-h

LCso = 81 ppm

Wan et aI., 1996

Margosan-Ob

Bluegill, 96-h

LCso = 37ppm

Larson, 1989

Neem stem bark extract

Aphyosemon giardneri, 96-h

LCso = 15.1 ppm

Osuala and Okwuosa, 1993

Margosan-O b

Daphnia, 48-h

LCso = 13 ppm

Larson, 1989

Margosan-Ob

Daphnia, 48-h

ECso = 125 ppm

Scott and Kaushik, 1998

Azatin h

Mayfly, I-hi

Neemrich I

Honey bee, topical

LCso = 1.12 ppm 0.0735

Kreutzweiser, 1997 Sharma et aI., 1984

aMilliliter per kilogram. bFormulation containing 14-20% neem oil giving a total of 3 gll azadirachtin. cFormulation containing 1% azadirachtin. dContains 12% azadirachtin and 88% other neem constituents. eFormulation containing 3% azadirachtin and approximately 27% other neem constituents. f Contains 49% azadirachtin and 51 % other neem constituents, including salannin, nimbin, etc. gFormulation containing 4.6% azadirachtin and 15% other neem constituents. hContains 3% azadirachtin. iOne-hour flow-through test, followed by a 21-day period for mortality observation.

(2) cytochrome bs in the brain; and (3) cytochrome P-450 reductase level in the liver and brain. The highest dose also caused 10% mortality; the medium dose elicited toxic signs, including behavioral abnormalities, lacrymation, reduced feeding, and loss in body weight. No toxicity was seen for the low dose. The toxic symptoms disappeared by the 28th day after cessation of treatment. Cytotoxic Effects In insect and mammalian cell culture tests, nimbolide was found to be the most active ingredient among the cytotoxic limonoids of neem seed extracts. Azadirachtin as

well as neem oil containing less than 1% of totallimonoids was essentially devoid of cytotoxicity (Cohen et aI., 1996a, b; see also Cui et aI., 1998; Kigodi et aI., 1989). Toxic Effects in Humans Poisoning Incidents Neem seed oil produced occasional diarrhea, nausea, and general discomfort when given orally as an anthelmintic (see Jacobson, 1989). Sinniah and Baskaran (1981) summarized 13, including 2 fatal, poisoning cases due to neem seed (margosa) oil, a tra-

134


ditional remedy in India and Malaysia. Five to ten milliliters of the oil given orally to children against minor ailments caused vomiting, drowsiness, tachypnea with acidotic respiration, and polymorphonuclear leukocytosis, and encephalopathy developed within hours of ingestion. Seizures, associated with coma, also developed in some cases. Autopsy demonstrated pronounced fatty acid infiltration of the liver and proximal renal tubules, with mitochondrial damage and cerebral edema, changes consistent with Reye's syndrome. Follow-up model studies with mice suggested that the syndrome could be associated with the long-chain fatty acid or lipid components of the oil (see Skellon et aI., 1962) provoking mitoses of hepatocytes within 30 min after ingestion, hypertrophy of endoplasmic reticulum, and loss of liver glycogen, which was consistent with fat accumulation in the liver cells (Sinniah et aI., 1989). In model studies with rat liver, Trost and Lemasters (1996) proposed that the pathogenesis of Reye's syndrome, caused by chemicals such as salicylic acid, valproic acid, and neem oil, is associated with the induction of mitochondrial permeability translation. The induction effect could be blocked by cyclosporin A.

Treatment The recommended treatment for margosa oil poisoning is control of seizures by diazepam, respiratory support, correction of acidosis, reduction of cerebral edema by dexamethasol and/or mannitol, and hydration of the patient (Sinniah and Baskaran, 1981). 3.2.2 MICROBIAL INSECTICIDES

3.2.2.1 Bacillus thuringiensis Endotoxins Introduction Bacillus thuringiensis (Bt) is an aerobic, sporeforming, gram-positive, rod-shaped bacterium distributed widely in the natural environment from the Arctic to the Tropics (Martin and Travers, 1989). The entomopathogenic and insecticidal action of the bacterium was first noted by Ishiwata in Japan in 1901. In 1915, another strain was found in Thuringia, Germany, by Berliner, who named it B. thuringiensis. A Btbased microbial insecticide was commercialized first in France in 1938. Bt was registered in the United States in 1961. Early works on this microbial insecticide were reviewed by Burges (1981). Identity, Properties, and Nomenclature During the past century, thousands of Bt isolates have been obtained from sources as diverse as living and dead insects, soil, grass, grain dust, and water. These can be divided into 58 serologically different Bt subspecies (or varieties), each producing one or more crystalline inclusions during sporulation. The crystals (or 8-endotoxins) are toxic against economically important insect pests and vectors of animal and human diseases. The following Bt subspecies have practical importance: kurstaki and aizawa (against lepidopteran larvae), israelensis (against mosquitoes and blackflies), and tenebrionis and japonensis (against beetles).

The insect-specific Bt toxins were traditionally classified according to their host, size, and crystal shape (HOfte and Whiteley, 1989). The four major groups (and their specific targets) are as follows: CryI (Lepidoptera), CryII (both Lepidoptera and Diptera), CryIII (Coleoptera), and CryIV (Diptera). The recently proposed nomenclature, based solely on amino acid identity, replaces Roman numerals with Arabic numerals (e.g., Cryl) to accommodate the growing number of new proteins, giving altogether 31 systematically arranged primary ranks (Crickmore et aI., 1998,2000). Each Bt subspecies may synthesize more than one class of 8-endotoxin. In addition to the Cry toxins, some Bt contain another crystal toxin (Cyt), which has a specific action on Diptera in vivo and also a broad spectrum cytolytic activity in vitro, being hemolytic for most eukaryotic cells, including horse, sheep, rat, mouse, rabbit, and human erythrocytes (Knowles et aI., 1992; Thomas and Ellar, 1983). Bt is taxonomic ally closely related to Bacillus cereus and also to Bacillus anthracis (see, e.g., Henderson et aI., 1994), which, however, lack the 8-endotoxin and are thus noninsecticidal and can be harbored in insects. However, the 8-endotoxin crystal-coding Bt plasmid could be transferred to B. cereus, yielding transcipients that produce crystals of the same antigenicity as the donor strain (Gonzalez et aI., 1982).

Structure The molecular weights of Cry proteins vary from 27 to 140 kDa, with regions of homology or high similarity interspersed with variable regions (HOfte and Whiteley, 1989). The x-ray crystal structure of the coleopteran-specific Cry3Aa (old name: CryIIIA) toxin (Li et al., 1991), the lepidopteranspecific Cry1Aa (old name: CryI) toxin (Grochulski et aI., 1995), and the dipteran-specific Cyt2Aa (old name: CytB) toxin (Li et aI., 1996) were recently solved. Along with 8-endotoxin, many variants of Bt, the subspecies aizawai in particular, produce a low-molecular-weight, thermostable, and water-soluble insecticidal compound, known as ,B-exotoxin [23526-02-5] (see Perani et aI., 1998; Sebesta et al., 1981). This exotoxin, also called thuringiensin, is a structural analog of ATP (Farkas et aI., 1969) (Fig. 3.7), inhibiting both prokaryotic and eukaryotic DNA-dependent ribonucleic acid (RNA) polymerases (Bond et al., 1969; McClintock et al., 1995; Sebesta and Horska, 1970). Although Bt insecticides containing ,B-exotoxin or its persistent salts are of commercial interest (see Abrosimova et al., 1985; Gingrich, 1990; Haufler

H$COO: H H

H

cD

HO

O-~P(OH)'

OH 0

\

11

0

~

.>

0

eOOH

NH,

N

""I0'(jN

""N

N

, HO

~OH

•.

~

~

~-exotoxln

Figure 3.7

Structure of Bacillus thuringiensis ,B-exotoxin.


and Kunz, 1985; Hsu et aI., 1997), most endotoxin-based Bt products are required to be free from ,B-exotoxin. Formulations and Uses Bt insecticides are produced by fermentation and formulated as water-soluble granules or liquid concentrates, emulsifiable suspensions, wettable powders, and slow release rings. A suspension containing Bt encapsulated in dead Pseudomonas fluorescens is also available (Copping, 1998). The formulations are standardized to contain a dose of toxin that is expressed in terms of international units active against the target pest per milligram of product. Repeated applications at rates ranging from 0.1 to 4 kg of formulated product are used. As already mentioned, the various subspecies of Bt have a narrow activity spectrum, being active against a few selected species of Lepidoptera, Diptera, or Coleoptera. Bt preparations are used to control insect pests in cotton, vegetables, orchards, maize, forests, turf, and ornamentals, as well as against mosquito and blackfly larvae in water bodies and sewage filters. Genes that code for the biosynthesis of the insecticidal proteins (cry genes) were cloned and engineered into maize, cotton, soybean and potato varieties that have been commercialized or undergoing registration in the United States and other countries. The complete toxicological evaluation of the transgenic plants producing the proteinaceous insecticide at ppm level, however, is a challenging task requiring the development of novel test methodologies and new safety criteria. Stability The proteinaceous Bt a-endotoxin is not soluble in water and organic solvents but can be dissolved in dilute alkalis such as aqueous NaOH solutions. Both the Bt bacterium and the toxin can be inactivated by the usual physical (heat) and chemical (formaldehyde, hypochlorite, or strong acid solution) sterilization methods. Under field conditions, Bt spores and crystals have, in general, low persistence, depending on the type of formulation and abiotic factors. Field half-lives ranging from 0.5 to 4 days were reported for various Bt preparations (Beegle et aI., 1981; Ignoffo et aI., 1974, 1977; Pinnock et aI., 1974). Sunlight alone or in combination with high temperature and rain is responsible for the rapid inactivation of Bt spores in the field (Ishiguro and Miyazono, 1982; Leong et aI., 1980; Raun et aI., 1966; van Frankenhuyzen and Nystrom, 1989). Bt-containing products lose activity in solution at pH > 8. Spectroscopic studies with purified protein crystals pointed out the role of exogenous photosensitizers, such as singlet oxygen, in the breakdown of tryptophan side chains in sunlight-mediated inactivation (Pozsgay et aI., 1987). A recent study, however, found Bt spores persisting in forest soil for up to 2 years after a 5-year intensive use of the insecticide against the gypsy moth (Smith and Barry, 1998).

Biological Properties Mode ofAction The insecticidal mode of action of Bt is complex (reviewed by Knowles, 1994; Schnepf et aI., 1998). Bt-

135

based insecticide formulations kill insects not because of their infectivity but because of their crystalline toxins that disrupt the midgut epithelium of susceptible insects. There are four steps involved in the mode of action of the toxin.

1. Ingestion. Bt has no contact activity and must be ingested to be toxic. 2. Solubilization and proteolytic activation. Within the digestive tract, the parasporal crystals dissolve in the alkaline environment and, because most Bts are actually inactive protoxins, are activated by proteolysis within the insect midgut. To exhibit toxicity against cell cultures in vitro, the protoxin requires prior enzymatic activation. 3. Binding to target site(s). The active toxins, with molecular weights ranging from 60 to 70 kDa, then bind to high-affinity sites, apparently glycoprotein receptors (Knight et aI., 1994; Vadlamudi et aI., 1995), on the microvilli of the midgut epithelial cells. 4. Formation of toxic lesions. The prevailing hypothesis for the cell membrane effect of the toxin is the formation of pores. After specific binding on the membrane of epithelial cells, the activated toxins insert rapidly and irreversibly into the plasma membrane and form a pore or lesion. This increases cell membrane cation permeability and initiates an influx of cations, especially K+. Osmotic balance is thus disturbed, which causes the cell to swell and burst by a process called colloid-osmotic lysis, eventually leading to larval death within 1 or 2 days (Knowles and Ellar, 1987). Another mechanism, not requiring pore formation for the K+ influx, involves inhibition of the (Na,K)-ATPase pump from the cytoplasmic side of the plasma membrane (English and Cantley, 1986). Distribution and Excretion The toxicity, fate, and infectivity of Bt subsp. israelensis preparations in mice, rats, and rabbits were studied in detail by Siegel et al. (1987). Viable bacteria could be recovered at the injection site and from the spleen of mice 14 days after subcutaneous administration of 109 Bt organisms. Aerosol exposure of rats to a spray containing 2.05 x 106 bacterial organisms/ml for 30 min showed that viable Bt from the lungs cleared completely within 7 days without any lesions, and no bacteria could be detected in the spleen. Twenty-six of 42 athymic mice died within 5-10 h after receiving 3.4 x 107 bacteria intraperitoneally; colony-forming units (CFUs) in the spleen of the surviving animals declined with time but persisted as long as 7 weeks. No mortality was seen in euthymic mice that received a comparative dose. In rat intracerebral injection experiments with different Bt preparations, only the highest doses of 107 Bt organisms per animal produced mortality (79-83%), and clearance from the spleen and brain of rats receiving 1.15 x 105 organisms was essentially complete within 3 weeks. Because recovery of CFUs decreased rather than increased over time in all experiments, the Bt subsp. israelensis preparations tested were clearly not infective. The involvement of phagocytic cells such as macrophages,

136

CHAPTER 3


the lymphatic system, and the blood stream in clearance of the bacteria was also proposed. Subsequent tests with mice and rabbits confirmed and extended these findings (Siegel and Shadduck, 1990). On intraperitoneal injection into mice, Bt subsp. israelensis CFUs were also recovered from heart blood, and their disappearance from heart blood coincided with their clearance from the spleen. Immunodeficient mice cleared Bt preparations at a slower rate. However, mice failed to remove one preparation of Bt subsp. israelensis from their enlarged spleen and a constant number of bacteria (1.6-2.0 x 106 CFUs) was recovered even after 10 weeks. On ocular installation, the bacteria persisted in both flushed and unflushed rabbit eyes for only 1 week. McClintock et al. (1995; see also U.S. EPA, 1998) reviewed unpublished studies on the clearance of viable Bt spores from rodents. Microbial clearance of Bt through the digestive tract of rats was complete in some instances in 2 days. Clearance of inhaled Bt subspecies aizawai and kurstaki from mouse brain, blood, liver, kidney, lung, lymph nodes, and spleen was complete in 2-3 weeks. Intravenous doses of Bt cleared at a slower rate from these tissues in mice and a similar pattern was observed for rats; viable CFUs could be recovered even after 50 days in some of the tissues of these rodents.

Toxicity to Animals and Laboratory Studies Acute Toxicity of Endotoxins As mentioned before, the crystalline Bt endotoxins require activation by alkalis and/or digestion, conditions absent in the mammalian stomach but present in the insect midgut, providing a basis for selective toxicity. For example, in mice, intravenous administration of crystalline a-endotoxin of Bt subsp. israelensis at 1 mg per animal produced no toxic symptoms, but the solubilized toxin had an LDso of approximately 0.49 mg per animal (Thomas and Ellar, 1983). Additional acute toxicity data are given in Table 3.7 and by Lamanna and Jones (1963). The mammalian toxicology studies submitted to the EPA on Bt-based insecticides were recently summarized by McClintock et al. (1995), relating the now internationally used "colonyforming unit" (CFU) dose data to exposure doses used in earlier studies. A CPU is defined as a single, viable propagule that produces a single colony (a popUlation of the cells visible to the naked eye) on an appropriate semisolid growth medium. The effects of various Bt preparations on nontarget aquatic and terrestrial organisms were recently compiled (Darvas and Polgar, 1998; see also U.S. EPA, 1998). Bt preparations containing 0.1 % of the water-soluble f3-exotoxin had an intraperitoneal LDso value of 364-387 mg/kg (5.9-6.2 x 109 cells/kg) in rats but were devoid of acute toxicity at 10,000 mg/kg (1.6 x lOll cellsikg) oral, at 6000 mg/kg (9.6 x 1010 cells/kg) dermal, and at 300 mg/m 3 (4.8 x 1010 cells/kg) inhalation doses (Khalkova et al., 1993). Chronic Toxicity In rat chronic studies with Bt subsp. kurstaki, daily doses of 8.4 g/kg oral administration for 90 days and feeding for 2 years did not show treatment-related effects.

A 13-week study with Bt subsp. kurstaki (Dipel), daily oral administration of 1.3 x 109 spores/kg, showed no toxicity or infectivity in rats. With Bt subsp. israelensis, the no-observed-effect level was daily 4 g/kg in a 3-month study with rats (McClintock et aI., 1995; see also California DFA, 1998). Sheep fed with 500 mg/kg daily doses of Dipel or Thuricide insecticide for 5 months showed no treatment-related effects (Hadley et aI., 1987). The bacterium could be cultured from blood and tissue samples taken at the end of the trial. The only pathological finding was mild lymphocytic hyperplasia in Pleyer's patches of the cecum of some animals. Irritation and Sensitization No allergenic response to Thuricide Bt preparation was evident in mice by inhalational exposure to 9 x 1010 viable spores for 10 min, and in guinea pigs by subcutaneous injection of 10 doses of approximately 9 x 105 spores during 3 weeks or topical application of approximately 4.5 x 107 spores to intact or abraded skin (Fisher and Rosner, 1959). Bt subsp. israelensis dry-powder preparations caused slight ocular irritation, whereas pastes caused severe conjunctival congestion and corneal injury in rabbit eyes (Siegel and Shadduck, 1990; Siegel et aI., 1987). Mutagenicity At 0.5-1000 Il-g/plate concentrations, the endotoxin-based Thuricide was not mutagenic in the 48-h Salmonella thyphimurium mutagenicity assay with or without a metabolic activator, although the validity of the Ames test for this type of product was questioned (California DFA, 1998). The same assay showed no mutagenicity for f3-exotoxin, either (Carlberg et aI., 1995). Poisoning Symptoms, Biochemistry, and Pathology The solubilized Bt subsp. israelensis endotoxin crystals caused poisoning symptoms and death in mice receiving 16.6 or 33.3 mg/kg doses on intravenous but not on oral administration (Thomas and Ellar, 1983). Within 1 h of injection, the animals developed paralysis in their hindquarters and became relatively immobile within 3 h. Breathing and heart rate increased. Death occurred after 12 h or 36-48 h after injection at the higher and lower doses, respectively. In mice, solubilized endotoxin and immunoaffinity-purified toxin fractions, a 28-kDa protein fraction in particular, of Bt subsp. israelensis given at 10, 7.5 and 2.5 mg/kg respective intraperitoneal doses caused hypothermia and bradycardia (Mayes et aI., 1989). Cytolysis of mouse red and white blood cells was not detectable after a 2.5 mg/kg intraperitoneal dose of solubilized endotoxin. Pathological and histological examinations of rats and mice treated intraperitoneally with 5 mg/kg solubilized endotoxin revealed focal-to-segmental reddened and edematous areas within the small intestine with major lesions in the jejunum. The toxic symptoms in the Japanese quail receiving more than 10 mg/kg intraabdominal doses of soluble Bt subsp. israelensis endotoxin were loss of alertness, loss of activity in the legs, and substantial volumes of fluid on the cloacal excreta


137

Table 3.7 Acute Toxicity of Bacillus thuringiensis Preparations Bt subspecies, a

Test species,

toxin type

route

Bti

Rat, oral

>2670 mg/kg

McClintock et al., 1995

Bti

Rat, oral

1.2 x 1011 spores/kg

McClintock et aI., 1995

Bti

Rat, dermal

>2000 mg/kg


Bti o-endotoxinb

Rat,ip

1.95 mg/kg

Roe et aI., 1991

Bti o-endotoxinb Bti o-endotoxinb

Rat, iv

Roe et aI., 1991

Rat, sc

>21 mg/kgC >9 mg/kgC

Bti

Rat, inhalation

Bti 8-endotoxinb Bti o-endotoxinb

Mouse, male, oral

>30ppm

Mouse, male, ip

1.31 ppm

Cheung et aI., 1985

Bti 8-endotoxind

Mouse, male, ip

0.77 ppm

Mayes et aI., 1989

Bti 8-endotoxinb Bti o-endotoxin b

Mouse, male, ip

2.33 mg/kg

Mayes et aI., 1989

Mouse, iv

~16mg/kge

Thomas and Ellar, 1983

Other data, comments

LD50

Reference

Roe et aI., 1991 LC50 = 8 x 107 spores/rat

McClintock et aI., 1995 Cheung et aI., 1985

Bti o-endotoxin!

Mouse, iv

>33 mg/kg C


Bti 8-endotoxinb

Suckling mouse, sc


Bti

Rabbit, oral

2.7-6.6 mglkg8 > 2 x 109 spores/rabbit


Bti

Rabbit, dermal

>6280 mglkg


Bti 8-endotoxinb Bti o-endotoxinb

Japanese quail, ip

22.8 mg/kg

Kallapur et aI., 1992

Japanese quail, intranasal

>50 mglkgC


Bti o-endotoxinb Bti 8-endotoxinb

Quail, iv

>100mg/kg C >lOOmg/kg C

Roe et aI., 1991

Bti

Brook trout, 48-h

LC50

= 2321 ppm

Wipfli et aI., 1994

Bti

Brown trout, 48-h

LC50 = 1691 ppm

Wipfli et aI., 1994

Btk

Rat, oral

>4.7 x 1011 spores/kg

McClintock et ai., 1995

Btk

Rat, dermal

Btk o-endotoxinb

Mouse, iv

>3.4 x 1011 spores/kg >33 mg/kgC


Btk 8-endotoxin! Btk 8-endotoxinb

Mouse, iv

>33 mglkgC


Mouse, male, oral

>30ppm

Cheung et aI., 1985 Cheung et aI., 1985

Quail, sc


Btk 8-endotoxinb

Mouse, male, ip

>30ppm

Btk

Rat

> 108 CFU/animal

Btk

Daphnia, 21-day

Bta

Rat

Bta

Daphnia, 21-day

Btt

Rat

Btt

Daphnia, 48-h


McClintock et aI., 1995 LC50

= 5-50 ppm

> 108 CFU/animalc

U.S. EPA, 1998 McClintock et aI., 1995

EC50

= 0.8-2.7 ppm

>2 x 108 CFU/animal

U.S. EPA, 1998 McClintock et aI., 1995

EC50 > 100 ppm

U.S. EPA, 1998

aBta, Bt subsp. aizawai; Bti, Bt subsp. israelensis; Btk, Bt subsp. kurstaki; Btt, Bt subsp. tenebrionis. bSolubilized endotoxin. CNo mortality. d A 28-kDa polypeptide fraction of solubilized endotoxin. eCalculated from original mortality data. ! Crystalline endotoxin. 8Lethal doses.

within 2 h of injection (Kallapur et aI., 1992). Bradycardia and hypothermia were observed for the 30 mg/kg treatment. This dose reduced serum lipid and alkaline phosphate levels and increased serum glucose, creatine phosphokinase, and lactate dehydrogenase. Serum calcium, alanine transaminase, blood urea nitrogen, bilirubin, and protein levels were the same as for the control.

Toxicity of ,B-Exotoxin ,B-Exotoxin showed delayed toxicity to mammals. The pure substance obtained from Bt subsp. gelechiae gave an LD50 of 18 mg/kg in mice, as estimated on the third day after intraperitoneal application (Sebesta et aI., 1969). The exotoxin is much less toxic when given orally to mice and upon dephosphorylation toxicity is lost completely (see Sebesta et aI., 1981).

138

CHAPTER 3


,B-Exotoxin, obtained from the culture supernatant of Bt subsp. morrisoni, gave subcutaneous LDso values of 184.5 and 135.6 mg/kg for male and female mice, respectively (Haufler and Kunz, 1985). The oral LDso of purified ,B-exotoxin was about 170 mg/kg in rats and caused dermal toxicity to rabbits at 0.4 mg/kg (see McClintock et al., 1995). The toxin also inhibited mitosis and impaired microtubules of the spindle and phragmoplast, effects characteristic of colchicine and vinblastine (Sharma and Sahu, 1977). Toxicity to Humans Allergenic reactions, infection and toxicity problems with Bt can arise during manufacture (fermentation), handling and field use. Physical and laboratory examination of human volunteers who inhaled 3 x 108 viable spores of Bt Berliner as powder for 5 days and ingested 3 x 109 viable spores of the bacterium daily for 5 days revealed no genitourinary, gastrointestinal, cardiorespiratory, or nervous system anomalies (Fisher and Rosner, 1959). Oral and dermal administrations of 105 _10 9 cells/g of Bt var. galleriae preparations produced nausea, belching, vomiting, tenesmus, colic-like pain, diarrhea, and fever, symptoms similar to those caused by B. cereus food poisoning (Pivovarov et a!., 1977) but these effects could have been related to the ,B-exotoxin present in the preparation (see Ray, 1991). Volunteers receiving 1 x 1010 Bt spores daily for three days showed no treatment-related symptoms, and viable Bt spores could be recovered from half of the patients for 30 days after treatment. Irritation and Sensitization Two reported incidents of possible allergic reactions to Bt-based products were unrelated to the bacterium but likely caused either by a previously undiagnosed disease (Kawasaki syndrome) or by an existing food allergy elicited by a formulating ingredient (see McClintock et a!., 1995). A recent study examining the allergenic potential of farm workers exposed to various levels of Bt insecticide revealed no adverse respiratory symptoms but skin allergy responses lasting up to four months and the occurrence of Bt-specific IgG and IgE antibodies were present especially in high-exposure groups (Bernstein et a!., 1999). Poisoning Incidents The first reported occurrence of an infection caused by Bt in humans was due to the accidental splashing of Dipel insecticide suspension in the eye of a farmer (Samples and Buettner, 1983a, b). In spite of immediate rinsing with water and application of antibiotic ointment, the eye was still irritated 3 days later when local corticosteroid treatment was begun. Ten days after the accident, a corneal ulcer was noted. The bacillus could be cultured from the eye and proved to be susceptible to gentamicin, which cured the patient. Another case of infection occurred in a laboratory when a student working with spores and endotoxin of Bt subsp. israelensis and Acinetobacter calcoaceticus var. anitratus accidentally stuck his finger on a needle (Warren et a!., 1984). Within 2 h, the finger became painful and soon became discolored and

the hand was swollen. In spite of immediate antibiotic therapy consisting of intravenous gentamicin first then benzylpenicillin (2.5 g every 4 h), lymphangitis developed. After 24 h, the flexor tendon sheath required decompression, over the finger joint close to the inoculation site. The patient recovered after 5 days. It was also discovered that the crystalline israelensis o-endotoxin protoxin could be activated in vitro at room temperature or 30°C within 2-4 h by proteases present either in Bt or in culture filtrates from unrelated bacteria, including the A. calcoaceticus var. anitratus (see also Damgaard et al., 1997). Green et al. (1990) described the results of an epidemiological study conducted in connection with an extensive Bt spraying program for gypsy moth control in Oregon. Of the 55 cultures from human specimens positive for Bt, only four cases could be related to the insecticide treatment. One of these was a sprayer who accidentally splashed the Bt spray mixture on his face and eyes and developed dermatitis, pruritis, burning, swelling, and erythema, with conjuntival injection. His eyelid and skin were treated with steroid cream. Bt was cultured from 18 different body sites or fluids from the other 54 cases, suggesting that the bacterium was appearing as a contaminant or commensal, rather than a pathogen, because there was no consistent pattern of disease associated with its presence. The authors also pointed out that immunocompromised persons could be at risk when exposed to Bt-based insecticides. Damgaard et a!. (1996) suggested, but did not prove, that the presence of insect pathogenic Bt strains, including subsp. kurstaki, in food and raw agricultural commodities, could be due to Bt insecticide residues containing live spores. Treatment Treatment is symptomatic and in case of infection appropriate antibiotics should be used. Bt is known to be resistant to several antibiotics. The resistance is associated with the plasmid that codes for the toxin crystal and the resistance to penicillin and cephalosporin antibiotics is lost with the loss of the plasmid (see Green et a!., 1990).

3.2.2.2 Spinosad Introduction Spinosad is a new insecticide contammg a structurally unique glycosylated macrolactone with selective activity against a wide variety of insect pest species. It is isolated from tht;! fermentation broth of the aerobic, gram-positive soil bacterium Saccharopolyspora spinosa (Actinomycetes). Identity, Physicochemical Properties, and Uses Spinosad consists of about 80% spinosyn A and 20% spinosyn D (Fig. 3.8) with traces of other structurally related macrolides as minor components. SpinosynA

Chemical Abstract name: spinosyn A: 2-((6-deoxy-2,3,4-triO-methyl-a-L-mannopyranosyl)oxy)-13-((5-(dimethylamino )tetrahydro-6-methyl- 2H -pyran-2-yl)oxy )-9-ethyl2,3,3a,5a,5b,6,9, 10, 11, 12, 13, 14,16a,6b-tetradecahydro-14methyl-1 H -as-indaceno[3,2-d]oxacyclododecin-7, 15-dione


139

to control lepidopteran and thysanopteran pests on cotton, vegetables, and omamentals at rates of 12-150 g/ha. Biological Properties

=

spinosyn A R H spinosyn 0 R = CH,

Figure 3.8

Structures of the components of spinosad insecticide.

CAS Registry Number: [131929-60-7] Empirical formula: C41H6SN016; molecular weight: 732.0

SpinosynD

Chemical Abstract name: 2-«6-deoxy-2,3,4-tri-O-methyl-a-Lmannopyranosy l)oxy )-13-( (5-( dimethylamino )tetrahydro-6methyl- 2H -pyran-2-yl)oxy )-9-ethy 1-2,3,3a,5a,5b,6,9, 10, 11,12,13,14, 16a, 16b-tetradecahydro-4, 14-dimethyl-l H -asindaceno[3,2-d]oxacyclododecin-7, 15-dione CAS Registry Number: [131929-63-0] Empirical formula: C42H67N016; molecular weight: 746.0 Physicochemical Properties The physicochemical properties of the active ingredients in spinosad were recently published (Spinosad Technical Guide; DeAmicis et aI., 1997). Pure spinosyn A melts at 118-124°C. Its solubility in water is 290, 235, and 16 mg/kg at pH 5.0, 7.0, and 9.0, respectively; its log P values are 2.8 and 4.0 at pH 5.0 and 7.0, respectively. The pKa of spinosyn A is 8.1. Optical rotation: [a]D = -135.3° (c = 1.0 in ethanol). Pure spinosyn D melts at 169-174°C. Its solubility in water is 29, 0.332, and 0.053 mg/kg at pH 5.0, 7.0, and 9.0, respectively; its log P values are 3.2 and 4.5 at pH 5.0 and 7.0, respectively. The pKa of spinosyn D is 7.87. Optical rotation: [a]D = -156.7° (c = 1.0 in ethanol). Stability Both spinosyns decompose when mixed with strong acids (pH < 2). The aqueous solution of the technical-grade material is stable in the dark and has a pH value of 7.74. Due to rapid photolysis, however, spinosad solutions in water at pH = 7.0 and 25°C have a half-life of about 1 day when exposed to light. The half-life of spinosad in soil is about 2 weeks (Saunders and Bret, 1997). History, Formulations, and Uses The spinosyns are a group of macrolide insecticides that were isolated from the soil bacterium S. spinosa (Kirst et aI., 1991, 1992; reviewed by DeAmicis et aI., 1997; Thompson et aI., 2000). The commercial product, spinosad, is a mixture of two active components, spinosyn A and spinosyn D (Fig. 3.8). The commercial material is sold as a water-based suspension concentrate or as waterdispersible granules. The microbial insecticide is currently used

Mode of Action Spinosad kills insects quickly with a speed comparable to most neurotoxic insecticides. The actual mode of action of these macrocyclic lactones has not been established yet. The compounds failed to show any significant effect for more than 65 known insecticide, drug, and toxin target sites (Salgado, 1997). In insects, the poisoning symptoms of either spinosad or spinosyn A treatment are muscle contractions due to excitation of the central nervous system, leading initially to postural changes, typically elevation of the body and straightening of the legs (Salgado, 1997, 1998; Salgado et aI., 1998). After many hours of excitation, the movements and fine tremors finally cease and the ensuing paralysis is apparently due to neuromuscular fatigue. Spinosyn A has no direct depressant effect on the neuromuscular system and at high concentrations neuromuscular transmission is actually enhanced. In electrophysiological studies with cockroach neurons, 20 nM spinosyn A was shown to activate nAChRs and this action could be blocked by the selective nicotinic receptor antagonist a-bungarotoxin. In addition, spinosyn A also prolonged the action of ACh. The compound also affected GABA receptors in isolated insect neurons (Salgado, 1997). Spinosyn A increased neural activity in other insect species examined, including the tobacco budworm and the housefly (Salgado et aI., 1998). In contrast to conventional macrolides such as erythromycin, spinosyns are devoid of antibacterial activity. Metabolism and Excretion Metabolism studies with 14C_ labeled spinosyn A in rats showed that 80-86% of the administered radioactivity was eliminated in the feces and 7-10% of the dose was eliminated in the urine within a day. Among the metabolites, glutathione conjugates and N - and O-demethylated macrolides were identified (Spinosad Technical Guide). In the cockroach Periplaneta americana, over 60% of radiolabeled spinosyn A was metabolized by 64 h (Salgado et aI., 1998).

Toxicity to Laboratory Animals Acute Toxicity Acute toxicity data for spinosyn A are summarized in Table 3.8 (Bret et aI., 1997). Because of the low human and environmental risks it presents, spinosad meets the EPA's criteria as a reduced-risk material. In acute and subchronic tests, spinosad did not demonstrate any neurotoxic or reproductive effects in rats, mice, and dogs (Bret et aI., 1997; Copping, 1998). In subchronic feeding studies in male rats, daily doses above 68.5 mg/kg (lowest observed effect level, LOEL), decreased body weight gain, anemia, and vacuolation in kidney, liver, heart, spleen, adrenals, and thyroid were observed (Anonymous, 1998). In a chronic feeding study

140


Table 3.8 Acute Toxicity of Spinosada

LDSO Species

Route

(mg/kg)

Rat, male

oral

3738

Rat, female

oral

>5000

Mouse

oral

>5000

Rabbit

dermal

>5000

Rat

inhalation

Other test results

3.2.3.1 Insect Sex Pheromones LCso >5.18 mg/! air

Rabbit

Slight conjunctival

eye

irritation, clearing in48h No sensitization

Guinea pig

dermal

Rabbit

dermal

Bobwhite quail

oral

Bobwhite quail

5-day dietary

Mallard duck

oral

Mallard duck

5-day dietary

LC so 2: 5000 ppm

Rainbow trout

96-h

LC so

Bluegill

96-h

LC so

No irritation >2000

LC so 2: 5000 ppm >2000

Daphina magna

48-h

= 30ppm = 5.9 ppm LCsO = 5.0 ppm LCso = 7.9 ppm LCso = 92.7 ppm

Green alga

7-day

EC so > 105 ppm

5-day

EC so

96-h

LCso > 9.8 ppm

Carp

96-h

Sheepshead minnow

96h

Selenastrum capricornutum

Blue green alga

= 8.1 ppm

Anabaena flosaquae

Grass shrimp

which give advantage to the emitter (e.g., defensive secretions), and kairomones, which give advantage to the receiver (e.g., secretions that can be detected by predators or parasites). This section will deal only with two types of semiochemicals: (1) insect sex pheromones, which are volatile compounds indispensable in mate finding, and (2) kairomones, which are important cues in locating hosts on which the insect feeds.

aFrom Bret et aI., 1997.

in dogs, the LOEL was 8.22 mg/kg/day based on increased liver enzymes, triglycerides, vacuolated parathyroid cells, and arteritis (Anonymous, 1998). A mutagenic battery consisting of the Chinese hamster ovary, mouse lymphoma cell, mouse bone marrow micronucleus, rat hepatocyte unscheduled DNA synthesis, and Ames tests did not show mutagenic activity (Spinosad Technical Guide). According to the EPA classification for pollinators, spinosad is "highly toxic," based on an acute 48-h topical LDso of 0.0025 jl.g/bee (Bret et aI., 1997). However, spinosad is still significantly less toxic than many synthetic insecticides, especially when considering its photolability.

3.2.3 SEMIOCHEMICALS Semiochemicals, in general, are natural compounds involved in animal communication. These natural products can be divided into two main groups, pheromones and allelochemicals. Pheromones mediate communication between individuals of the same species, whereas allelochemicals act between different species. The latter can be further divided into allomones,

Introduction and General Properties Insect sex pheromones are emitted by one sex, usually females, of a particular species in miniscule amounts, and are perceived by members of the opposite sex, eliciting a complex of behavioral responses including searching and mating. Sex pheromones are highly species specific and, as a rule, a unique blend of the natural pheromone components is needed for attraction. The first sex pheromone identified was (E,Z)-1O,12-hexadecadienol or bombykol, the sex pheromone of the silkworm moth, Bombyx mori (Butenandt et aI., 1959). The history of pheromone research was reviewed by Schneider (1992). Structurally, the sex pheromones encompass a diverse group of predominantly lipid-like, highly volatile compounds, but several polycyclic oxygenated structures have also been identified. Structures of selected sex and aggregation pheromones are depicted in Fig. 3.9. Several hundreds of compounds have been isolated and identified from volatile secretions of insect pests and disease vectors. Pheromones reported until 1988 were compiled by Mayer and McLaughlin (1991). An up-to-date list of known pheromones and attractants, "The Pherolist," is available on the Internet (Am et aI., 1999). Uses and Formulations Pheromones are applied in four major ways (Howse et aI., 1998; Jutsum and Gordon, 1989) (1) population monitoring with traps baited with the pheromone, (2) mass trapping using a large number of high-capacity trapping devices, (3) pheromone plus insecticide combination (lureand-kill), and (4) mating disruption by permeating the area with specially formulated pheromones. In practice, only synthetic compounds manufactured on a large scale (> 1000 kg) are used. The amount of the pheromone blend in traps generally ranges from 0.1 to 1000 mg, whereas effective and lasting mating disruption typically requires 50-200 g/ha of the pheromone. Stability Because of their particular mode of action, the pheromones are volatile substances and require special formulations providing even emission of the pheromone blend for weeks. Pheromones containing C=C double bonds, especially conjugated ones, and aldehyde groups are vulnerable to oxygen and sunlight (see, e.g., Dunkelblum et at., 1984; Shani and Klug, 1980; Shaver and Ivie, 1982), and the attractivity of the pheromone is lost unless formulated with antioxidants and UV screeners. Environmental Fate Using gas chromatography-mass spectrometry (GC-MS) detection, Spittler et at. (1992) could not


muscalure gossyplure

~CHO (Z)-11-Hexadecenal n = 2 (Z)-13-0cladecenal n = 3

(+)-23,070 >10,000 >5,000 >34,600 :=:: 1I, 730 >3,250 >5,600 :=::13,430 >15,000 >5,000 >5,000 >3,460 >15,000 >5,000 >5,000 >5,000 >5,000 >5,000 >5,000 >15,000 4,556

>2,025 >20,000 >2,000 >2,025

>26.6

~3,400

>2,025 >3,000 >2,000

>5.0 >5.0 >6.7

>5

>2,025 >3.33

>5,000

>16.9

>3,000 >2,025 a

>2.9

Reference Beroza et aI., 1975 Hodosh et aI., 1985 U.S. EPA, 1994 Beroza et aI., 1975 Beroza et aI., 1975 Copping, 1998 Inscoe and Ridgway, Beroza et aI., 1975 Copping, 1998 Copping, 1998 Inscoe and Ridgway, Beroza et aI., 1975 Hodosh et aI., 1985 Inscoe and Ridgway, Inscoe and Ridgway, Hodosh et aI., 1985 Inscoe and Ridgway, Inscoe and Ridgway, Inscoe and Ridgway, Hodosh et aI., 1985 Beroza et aI., 1975

1992

1992

1992 1992 1992 1992 1992

death out of four animals treated with 2025 mg/kg.

98.7% pure pheromone in the diet showed adverse effects for 3-week-old embryos, normal hatchlings, and l4-day-old survivors; the technical-grade material, however, produced abnormalities at 2 ppm. The pheromone was nontoxic to freshwater fish, but highly toxic to Daphnia magna (LCso = 1.08 ppm). The metabolism and environmental fate of (Z)-9-tricosene is less studied. In housefly cuticle, it undergoes cytochrome P-450--mediated oxidation forming ketone and epoxide metabolites (Ahmad et aI., 1987).

Disparlure (±)-Disparlure (racemic) Chemical Abstract name: cis-2-decyl-3-(5-methylhexyl) oxirane CAS Registry Number: [29804-22-6] (+ )-Disparlure Chemical Abstract name: (2S-cis)-2-decyl-3(5-methylhexyl)oxirane Other name: (7 R,8S)-7,8-epoxy-2-methyloctadecane CAS Registry Number: [54910-51-9] Empirical formula for both: C19H380; molecular weight: 282.5 Females of the gypsy moth, Lymantria dispar, emit a powerful sex pheromone, attracting males of the same species from several hundred meters. The compound was identified, without

establishing the absolute configuration, by Bierl et at. (1970). It was subsequently found that the synthetic (+ )-isomer was the attractive component in the synthetic racemic mixture (Plimmer

et at., 1977). Disparlure traps are used in forestry and in orchards for population monitoring and mass trapping. Special formulations were developed for mating disruption (Cameron, 1979). Disparlure is essentially nontoxic and nonirritating (Table 3.9) (Beroza et aI., 1975). In male antennae, the pheromone undergoes enzymatic hydrolysis, converting the epoxide into a behaviorally inactive diol (Prestwich et aI., 1989). Cameron (1983) reported that years after being regularly exposed to synthetic disparlure, he became attractive to males of the gypsy moth, suggesting an unusual persistency of the lipophilic material in human skin.

Gossyplure Chemical Abstract name: (Z,E)- and (Z,Z)-7,11-hexadecadien-l-01 acetate CAS Registry Numbers: unspecified stereochemistry: [50933-33-0]; (Z,E)-isomer: [51607-94-4]; (Z,Z)-isomer: [52207-99-5] Empirical formula: C18H3202; molecular weight: 280.4 The sex pheromone of the pink bollworm, Pectinophora gossypiella, a serious pest of cotton, was identified by Hum-


mel et al. (1973) as an approximately 1:1 mixture of (Z,E)and (Z,Z)-7,II-hexadecadien-1-yl acetates. It is now widely used in traps for monitoring as well as in special formulations for mating disruption alone (Jackson, 1989) or in combination with a neurotoxic insecticide (Haynes et al., 1986). Gossyplure is insoluble in water and readily soluble in most organic solvents. Henson (1977) found that the half-life of gossyplure in soil was 1 day, while in water 7 days. The loss was attributed to volatilization and hydrolysis to the corresponding alcohol mixture.

(+)-a.-copaene

Roelofs et al. (1971) identified this dienol as a sex attractant of the codling moth, Cydia pomonella. The compound is extensively used in traps for population monitoring and also for mating disruption (Jutsum and Gordon, 1989; Kirsch, 1997).

o

~o

HO

o cucur1litadn A: 2~-OH. R = CH.OH cucurbilacin B: 2~-OH. R CH, cucurbilacinE: A"'. R=CH,

=

Codlemone Chemical Abstract name: (E,E)-8,1O-dodecadien-l-ol CAS Registry Number: [33956-49-9] Empirical formula C12H220; molecular weight: 182.3

143

Figure 3.10

Structures of natural and synthetic kairomones.

test. In static fish tests, bluegill sunfish became flaccid, with shallow respiration, and lay on the bottom of the tank. The 96-h LCso was 12.1 ppm. With trout, trimedlure evoked dark discoloration of the integument, rapid and shallow respiration, excessive swimming with gyrating, and later lying on the bottom of the tank. The 96-h LCso for this fish was 9.6 ppm.

3.2.3.2 Kairomones If the sex pheromone of a practically important insect species is not available or not suitable for practical purposes, another attractant semiochemical could offer an alternative solution. Two such compounds, a synthetic attractant and a plant-derived, nonvolatile phagostimulant, as representative kairomones will be mentioned. Trimedlure

Introduction During the search for attractants of the Mediterranean fruit fly, Ceratitis capitata, a cheap synthetic analog of one of its postulated natural kairomones, (+ )-a-copaene, emerged as the most powerful substance (Mc Govern and Beroza, 1966). The commercially available trimed1ure, in fact, is a mixture of 16 isomers altogether of which the (1 S ,2S,4 R)4-chloro isomer (Fig. 3.10) is the most attractive (see Warthen et al., 1995). Identity Chemical Abstract name of trimedlure mixture: 4(or 5)-chloro-2-methylcyclohexanecarboxylic acid 1, I-dimethylethyl ester CAS Registry Number: isomeric mixture: [12002-53-8] Empirical formula: C12H21 C102; molecular weight: 232.7

Toxicity to Animals In one of the first acute toxicity studies with attractants (Beroza et al., 1975), trimedlure was slightly toxic to mammals (Table 3.9). Trimedlure caused local skin reactions characterized at the end of the 24-h contact period by erythema and edema. It was not an eye irritant in the Draize

Formulations and Uses Traps containing up to 1-2 g of the stereoisomeric mixture of this attractant are used to detect the insect and for mass trapping. Pharmacological Studies The 4-chloro-trans isomer of trimedlure effectively replaced the GABA-gated chloride channel probe eSS]t-butylbicyclophosphorothionate ([ 3s S]TBPS) from receptors of housefly brain membrane preparations. Rat brain eSS]TNPS receptors, however, were not sensitive to trimedlure (Cohen and Casida, 1985). Cucurbitacins

Introduction Cucurbitacins are highly oxygenated, tetracyclic triterpenes present in the fruits and roots of cucurbits such as watermelon, squash, and zucchini at 0.1-0.3% concentrations. These compounds are notable for their extreme bitterness. The detection level for humans is about 1 ppb in solution. They also have a broad range of pharmacological properties, including purgative, hepatoprotective, antifungal, cytotoxic, and antineoplastic activities (reviewed by Mir6, 1995). Importantly, they are locomotor arrestants and phagostimulants to Diabrotica species (Chrysomelidae), which are major coleopteran insect pests of corn in the United States, Mexico, and parts of Europe (Metcalf, 1994). Cucurbitacincontaining, sprayable insecticide baits, developed recently by Metcalf et al. (1987), greatly reduce insecticide application rates. Although it is obvious that the poisonous properties of this combination are due to the insecticide content, the plant material is also toxic and thus poses a risk during manufacture and handling.

144


Identity Of the dozens of cucurbitacins identified (Lavie and Glotter, 1971) representative examples, for which toxicological data exist, are discussed only. Cucurbitacin A Chemical Abstract name: (2,B,9,B,lOa,16a,23E)-25(acetyloxy)-2, 16,20-trihydroxy-9-(hydroxymethyl)-19norlanosta-5,23-diene-3, 11 ,22-trione CAS Registry Number: [6040-19-3] Cucurbitacin B

3.3 DISEASE CONTROL AGENTS 3.3.1 FUNGICIDES 3.3.1.1 Blasticidin-S Introduction Blasticidin-S, a member of the aminohexose pyrimidine nucleoside group of antibiotics, is produced by the soil bacterium Streptomyces griseochromogenes (Actinomycetes). Its structure was elucidated by Otake et al. (1966) (Fig. 3.11). Blasticidin-S was found to be fungicidal in 1958, and it is now used for the preventive and curative control of Pyricularia oryzae, the causative agent of rice blast (reviewed by Yamaguchi, 1995, 1996). Identity, Physicochemical Properties, and Uses

Chemical Abstract name: (2,B,9,B,lOa,16a,23E)-25(acetyloxy)-2, 16,20-trihydroxy-9-methyl-19-norlanosta5,23-diene-3, 11 ,22-trione CAS Registry Number: [6199-67-3] Cucurbitacin E Chemical Abstract name: (9,B,lOa,16a,23E)-25-(acetyloxy)2, 16,20-trihydroxy-9-methyl-19-norlanosta-l ,5,23-triene3,11,22-trione CAS Registry Number: [18444-66-1] This compound (also called a-elaterin) appears to be one of the most abundant in squash. Formulations and Uses The semiochemical-insecticide bait combination is formulated as either dry-flowable microspheres or polymer-based tank mixes. The source of the phagostimulant cucurbitacins used is usually the wild-growing buffalo gourd, Cucurbitafoetidissima. The Slam™ formulation contains 13% carbaryl and less than 1% plant material. Acute Toxicity to Laboratory Animals The LDso values upon intraperitoneal administration of cucurbitacin A were 1.2 mg/kg in male mice and 2.0 mg/kg in female rats; the LDso value of cucurbitacin B was 1.0 mg/kg in mice (David and ValIance, 1955). Lethal doses caused respiratory distress and pathology showed acute pulmonary edema. Cucurbitacin E had an intraperitoneal LDso of 2.0 mg/kg in mice (see Rymal et al., 1984). Stoewsand et al. (1985) reported diarrhea, anemia, and mortality in mice receiving a diet containing 1% cucurbita fruit of cultivars rich in cucurbitacin. Human Poisoning Incidents Ferguson et al. (1983), Kirschman and Suber (1989), and Rymal et al. (1984) summarized cucurbitacin-related food poisonings that occurred in the United States and Australia in the early 1980s.

IUPAC name: 1-(4-amino-l,2-dihydro-2-oxopyrimidin-l-yl)4-[(S)-3-amino-5-( I-methylguanidino)valeramido]-1,2,3,4tetradeoxy-,B-D-erythro-hex-2-enopyranuronic acid Chemical Abstract name: (S)-4-[[3-amino-5-[(aminoiminomethyl)methylamino]-I-oxopentyl]amino]-1-[4-amino-2oxo-l (2H)-pyrimidinyl]-1 ,2,3,4-tetradeoxy-,B-D-erythrohex-2-enopyranuronic acid CAS Registry Number: [2079-00-7] Empirical formula: C17H26NgOS; molecular weight: 422.4 Physicochemical Properties Pure blasticidin-S forms colorless crystals melting at 253-255°C with decomposition; the melting point of technical-grade material is 235-236°C with decomposition. Blasticidin-S is dextrorotatory: [a]g = + 108.4° (c = 1.0 in water). Blasticidin-S is a weak base with pKaj = 2.41 (carboxyl), pKa2 = 4.6, pKa3 = 8.0, and pKilj > 12.5 (three bases). It is readily soluble in water and acetic acid (>30 g/l in each NH,

")~~~-§~O yNH

o blasticklln-S

kasugamycin

HN }-NH H,N

HO

HO

0

o

~ HO HN)HO

~:~~

OH O-R HO

o

mildiomycln

Figure 3.11

OH

validamycin A R = ~-D-Glc validoxylamlne A R = H

Structures of microbial antifungal agents.

3.3 Disease Control Agents

at 20°C) but practically insoluble in common organic solvents. The pH of the aqueous solution ofthe free base is 9.3. The commercially available N -(4-benzylamino)benzenesulfonic acid (BABS) salt of the antibiotic has a pH of 6.0.

Stability Blasticidin-S is stable in solution at a pH of 5-7, unstable at a pH less than 4, and decomposes under alkaline conditions with the loss of ammonia (Otake et al., 1966). As dry film, blasticidin-S BABS salt is photostable. History, Formulations, and Uses Blasticidin-S, the first antibiotic developed for agricultural use, was isolated from the culture broth of S. griseochromogenes in 1955 and its unique fungicidal properties were discovered in 1958 (Takeuchi et al., 1958; reviewed by Misato, 1969; Yamaguchi, 1995). This nucleoside derivative possesses a wide range of biological activities, including antimicrobial (Takeuchi et al., 1958), antiviral (Hirai and Shimomura, 1965; Kummert and Semal, 1971), and antitumor (Tanaka et al., 1961) properties. Its only use is in rice to control P. oryzae, a pathogen of rice. Spraying a concentration of 10-40 ppm (10-40 g/ha) blasticidin-S in the field gives excellent control of this disease. The stable and non-phytotoxic blasticidin-S BABS salt is sold as dispersible powder, emulsifiable concentrate, or wettable powder formulations containing 1.4-6% active ingredient. To alleviate eye irritation (see following discussion), an improved formulation containing 5% calcium acetate additive was introduced (Yamaguchi, 1995). Biological Properties

Mode of Action The mode of action of blasticidin-S is not fully understood, but inhibition of protein biosynthesis in both prokaryotes and eukaryotes by interference with peptidyl transfer at the specific binding sites of ribosomes is mainly re-

145

sponsible for biological activity (Kinoshita et al., 1970; Pestka et al., 1972; see also Yamaguchi, 1995). In a cell-free system of P. oryzae, the target pathogen, incorporation of amino acids into protein was inhibited, whereas other metabolic pathways, including glycolysis, electron transport, oxidative phosphorylation, and nucleic acid synthesis, were not affected. Recent studies with the water mold, Achlya bisexualis, suggested inhibition of DNA synthesis as an additional effect (Sullia and Griffin, 1977).

Degradation, Metabolism, and Excretion On the plant surface, blasticidin-S is decomposed by sunlight and eventually gives rise to cytosine as the main degradation product (Yamaguchi et al., 1972). Common microbes in the field also contribute to the inactivation and disappearance of the antibiotic, and an aminohydrolase (blasticidin-S deaminase), selectively catalyzing the deamination of the cytosine nucleus, was isolated and characterized from resistant Aspergillus (Seto et al., 1966; Yamaguchi et al., 1975) and Bacillus (Endo et aI., 1987) strains. The residue level ofblasticidin-S in rice was below 0.05 ppm a week after application, whereas the soil half-life of blastic id inS was about 2 days under flooded conditions (Ebata, 1983). Upon oral application to the mouse, blasticidin-S and metabolites were excreted in the urine and feces within 24 h. Toxicity to Laboratory Animals

Acute and Chronic Toxicity In general, blasticidin-S is rather toxic to mammals but has low toxicity to fish. Acute toxicity data for several species are listed in Table 3.10. (Some reports fail to specify whether the material tested was the free base of blasticidin-S or its BABS salt, which could explain the variations in the reported LD50 values.)

Table 3.10 Acute Toxicity of Blasticidin-S a

Species, sex

Assay

LDSO (mg/kg)

Rat

oral

39.5

Rat

oral

16.3

Yamashita et aI., 1987

Rat, male

oral

56.8 b

Tomlin, 1997

Rat, female

oral

55.9b

Tomlin, 1997

Rat

dermal

>500

Tomlin, 1997

Mouse

oral

10.1

Yamashita et aI., 1987

Mouse

oral

33.0b

Yang and Deng, 1996

Mouse, male

oral

51.9b

Tomlin, 1997

Mouse, female

oral

60.1b

Tomlin, 1997

Mouse

iv

Carp

48-h

LCso > 40ppm

Copping, 1998

Daphnia pulex

3-h

LCSO >40ppm

Copping, 1998

a Free

base unless otherwise noted. bFor blasticidin-S N -(4-benzylamino )benzenesulfonate.

Other data

Reference Misato, 1969

2.82

Takeuchi et aI., 1958

146

CHAPTER 3


In rats given blasticidin-S orally at 3 mg/kg or higher, alkaline phosphatase activity in serum and small intestine was temporarily reduced (see Ray, 1991). Pathology The main pathological findings of blasticidin-S poisoning in animals occur on the mucous membrane and skin. Upon topical application, conjunctivitis, keratitis, nasal bleeding, and skin lesions, including hyperemia, edema, and ulceration, are observed. Peritoneal adhesion involving intraabdominal organs and occasional gastrointestinal perforation are thought to be due to the lesions on the mucous membrane. Diarrhea is frequent and considered to be caused by irritation of the mucous membrane (Yamashita et aI., 1987). Intratracheal injection of blasticidin-S into rabbits produced pneumonitis, characterized by focal destruction of tissues. Within 5-6 days, these proliferations formed glandular structures extending from the bronchiole. Within two more days, blood capillaries began to surround the glandular cells and then the glandular cells began to differentiate from the alveoli. By 12-14 days after treatment, both types of cells could be distinguished (Ebe, 1969). In a 2-year study with rats given 1 ppm blasticidin-S in the diet, no adverse effects were observed (Tomlin, 1997). The fungicide was nonmutagenic in a battery of S. typhimurium and E. coli bacterial reversion-assay systems (Moriya et aI., 1983).

above 120 beats/min. Death occurred about 1 day after ingestion (Yamashita et al., 1987). Toyoshima et al. (1994) reported an unusual case of acute interstitial pneumonia caused by inhalation of blasticidin-S powder. Yang and Deng (1996) gave a detailed analysis of 24 suicidal, 3 occupational, and 1 accidental poisoning cases that occurred in Taiwan between 1985 and 1993. None of the five fatal cases was work related. The characteristic symptoms of blasticidin-S poisoning again were gastrointestinal disorders; redness of the conjunctiva; hypotension, occasionally preceded by hypertension, with tachy- or bradycardia; and aspiration pneumonia. Occasionally, neurological manifestations of poisoning could be seen. As little as 70 mg of the active ingredient was capable of producing symptoms. Death resulted from cardiovascular collapse or aspiration pneumonitis with possible bronchospasm. Treatment Because blasticidin-S is water soluble, contaminated skin should be washed. The oral and nasal cavities should be cleaned as well. Prompt symptomatic treatment is important and should include intravenous fluid administration and management of body water and electrolyte balance. Hydration and adequate urinary output are essential to facilitate renal elimination of the toxic ant. Prevention of aspiration pneumonitis and support of ventilatory function are also vital. For workers with chronic exposure, steroids and antibiotics could be helpful.

Toxicity to Humans

3.3.1.2 Kasugamycin

Irritation Blasticidin-S causes irritation and inflammation upon contact with eyes and mucous membranes. A survey of ophthalmic disturbances conducted in the 1960s showed that in some years one-third of the applicators suffered some damage, including skin eruptions all over the body. Newer, calcium acetate-containing formulations are safer in this respect (reviewed by Ray, 1991; Yamashita et aI., 1987).

Introduction Kasugamycin (Fig. 3.11) is an aminoglycoside antibiotic produced by Streptomyces kasugaensis (Umezawa et aI., 1965). Its structure was established by Suhara et al. (1966). Kasugamycin has systemic activity and is widely used to control the rice blast disease caused by P. oryzae (reviewed by Yamaguchi, 1995, 1996).

Poisoning Incidents Yamashita et al. (1987) described four incidents of acute suicidal poisoning from ingested blasticidinS. In the three fatal cases, nausea, vomiting, and severe diarrhea appeared almost immediately after ingesting 100-250 ml of undiluted formulation containing 2-5 g blasticidin-S BABS salt. Pain in the oral cavity and pharynx was noted in all patients. In one case, the vomitus was bloody and esophageal pain was claimed. All patients were completely conscious and restless during the 6-10 h after ingestion. Hypotension associated with tachycardia became more pronounced as time passed. Marked cold, pale, and perspired extremities were usually noted. Thermal symptoms indicated insufficient peripheral circulation. No cardiac anomalies were seen. The average hematocrit and hemoglobin concentrations for the three fatal cases were 52.9 and 17.4 g/lOO ml, respectively. Remarkable hemoconcentration was consistently noticed. Laboratory findings revealed moderate hepatic dysfunction. In the terminal phase of the fatal cases, blood pressure dropped and pulse rates increased

IUPAC name: 1,3,4/2,5,6-1-deoxy-2,3,4,5,6-pentahydroxycyclohexyl 2-amino-2,3 ,4,6-tetradeoxy-4-(a - iminoglycino )-a - 0arabino-hexopyranoside Chemical Abstract name: 3-0-[2-amino-4-[( carboxyiminomethyl)amino]-2,3,4,6tetradeoxy-a-o-arabino-hexopyranosyl]-o-chiro-inositol CAS Registry Number: kasugamycin [6980-18-3]; kasugamycin hydrochloride hydrate: [19408-46-9] Empirical formula: kasugamycin: C14H2SN309; molecular weight: 379.4; kasugamycin hydrochloride hydrate: C14H28CIN301O; molecular weight: 433.8

Identity, Physicochemical Properties, and Uses

Physicochemical Properties The HCI hydrate forms sweet, colorless crystals melting at 202-204°C (with decomposition). At room temperature, the solubility of kasugamycin HCI hydrate is 125 g/l in water and 2.76 mg/l in methanol; it is insoluble in acetone, ethyl acetate, and chloroform. Kasugamycin


is a weak base with pKaj < 2 (carboxyl), pKa2 = 7.1, and pKa3 = 10.6 (two bases). Kasugamycin is dextrorotatory: [a]6' = +120° (c = 1.6 in water).

147

1.0 g kasugamycin into humans, about 63% of the fungicide was excreted unchanged with urine in 8 h. Toxicity to Laboratory Animals

Stability Kasugamycin HCI hydrate is more stable than the free base and does not deteriorate upon storage at 50°C for 10 days. It tolerates weak acids, decomposes slowly at a pH of 7, and decomposes faster within weeks in alkaline solutions at ambient temperature. Formulations and Uses Kasugamycin is usually sold as hydrochloride hydrate. Typical formulations for spraying, dusting, or seed treatment are wettable powders and soluble liquid concentrates alone (0.3-3% active ingredient) or in combination with other pesticides (Copping, 1998; Tomlin, 1997). In addition to controlling rice blast and a few other fungal diseases, kasugamycin is active against Pseudomonas, Erwinia, Xanthomonas, and Corynebacterium bacterial species (Ogawa, 1992). Biological Properties

Mode of Action In cell-free systems, kasugamycin inhibited protein synthesis in P. oryzae and Pseudomonas fiuorescens markedly but much less so in rat liver preparation (Tanaka et aI., 1965). The antibiotic was shown to interfere with aminoacyl-tRNA binding to the 30-S ribosomal subunit of E. coli (Okuyama et aI., 1975). Interestingly, kasugamycin is inhibitory to P. oryzae in acidic (pH = 5) but not in neutral media (Hamada et aI., 1965). Metabolism and Excretion Oral administration of 100 mg/kg kasugamycin to mice indicated rapid absorption and 43-68% excretion with urine in 6 h (Takeuchi et aI., 1965). When a rabbit was subcutaneously injected with the same dose of kasugamycin, the fungicide disappeared from the blood within 8 h, and 96% of the injected material was excreted into urine in 8 h after injection; kasugamycin concentration in the urine was highest (43 mg/ml) after 45 min. On intramuscular injection of

Acute and Chronic Toxicity Acute toxicity data for kasugamycin HCI hydrate are shown in Table 3.11. Doses of 2000 mg/kg kasugamycin, administered intravenously, subcutaneously, or intraperitoneally to mice, caused neither observable effects nor death (Takeuchi et aI., 1965; see also Matsuzaki et aI., 1968). Furthermore, an 800 mg/kg dose intravenously injected into a monkey failed to show any toxic effects. The blood analysis of monkeys receiving repeated intramuscular doses up to 800 mg/kg of the fungicide was normal. Similarly, no effect on the blood chemistry was found upon administration of a total of 10 g kasugamycin to a dog (10 kg) during a 45-day trial. Kasugamycin, belonging to the aminoglycoside antibiotics known for their potential nephrotoxic, ototoxic, and neuromuscular paralytic activity, did not induce polymerization of rabbit muscle actin in vitro (Someya and Tanaka, 1979). Kasugamycin was nonmutagenic in a battery of S. typhimurium and E. coli bacterial reversion-assay systems (Moriya et aI., 1983). Toxicity to Humans Based on the remarkable low toxicity of kasugamycin, it was tested and proven to be effective against Pseudomonas aeruginosa urinary infections in humans (Takeuchi et aI., 1965). 3.3.1.3 Mildiomycin Introduction The antibiotic mildiomycin was isolated from the culture broth of the actinomycete Streptoverticillium rimofaciens (Iwasa et aI., 1978). The structure of mildiomycin was identified as a 5-(hydroxymethyl)cytosine-containing nucleoside derivative (Harada and Kishi, 1978; Harada et aI., 1978) (Fig. 3.11). It is a potent and selective fungicide used against various pathogens causing powdery mildews in fruits and vegetables (Iwasa, 1983; Kusaka et aI., 1979).

Table 3.11 Acute Toxicity of Kasugamycin Hydrochloride Hydrate

Species, sex

Assay

LDSO (mg/kg)

Rat

oral

>5000

Rat

dermal

>5000

Rat

4-h inhalation

Rabbit

dermal

Monkey

iv

Other data

Reference Ogawa,1992 Ogawa,1992

LCso > 2.4 mg/l air >2000

Ogawa,1992 Copping, 1998

>800

Takeuchi et aI., 1965

Japanese quail

oral

Carp

48-h

LCso > 40ppm

Copping, 1998

Goldfish

48-h

LCso > 40ppm

Copping, 1998

Daphnia pulex

6-h

LCso > 40ppm

Copping, 1998

LDSO >40 J.!glbee

Copping, 1998

Honeybee

>4000

Copping, 1998

148


Identity, Physicochemical Properties, and Uses Chemical Abstract name: (S)-4-amino-l-[4-[(2-amino-3hydroxy-l-oxopropyl)amino ]-9-[(aminoiminomethyl) amino]-6-C-carboxy-2,3,4,7,9-pentadeoxy-a-L-talo-non-2enopyranosy 1] _5-(hydroxymethy1)-2( 1H)- pyrimidinone Code numbers: Antibiotic B-98891, TF-138 CAS Registry Number: [67527-71-3] Empirical formula: C 19H30NS09; molecular weight: 514.5

Physicochemical Properties The melting point of mildiomycin hydrate is greater than 300°C (decomposition). Mildiomycin is hygroscopic, readily soluble in water and acids, but sparingly soluble in dioxane, dimethyl sulfoxide, and pyridine. Mildiomycin is a weak base with pKa] = 2.8 (carboxyl), pKa2 = 4.2, pKa3 = 7.2, and pK 12 (two bases). The antibiotic is dextrorotatory: [a]5° = +100° (c = 0.5 in water); [a]5° = + 78S (c = 0.5 in 0.1 N HCI). Stability In aqueous solutions, mildiomycin is stable at pH = 7, but slowly decomposes in alkaline (pH> 9) and strongly acidic (pH < 2) media. Formulations and Uses The fungicide is formulated as a wettable powder or aqueous solution containing 8% active ingredient. It was introduced against powdery mildews in cucumber, apple, grape, barley, green pepper, strawberry, mulberry, tobacco, and rose. It is applied to the foliage as a spray at concentrations of 40-80 ppm (Kusaka et aI., 1979). Biological Properties

Mode ofAction Mildiomycin selectively inhibits protein synthesis by blocking the peptidyltransfer in human HeLa cells. RNA or DNA synthesis is not affected. When HeLa cells are permeabilized by animal viruses, the inhibitory effect on protein synthesis increases, indicating that the basis for the selective antibacterial action of mildiomycin could be due at least in part to poor penetration into the cell (Feduchi et aI., 1985). Mildiomycin suppressed the growth of E. coli by inhibiting bacterial polypeptide synthesis without interfering with respiration, oxidative phosphorylation, nucleic acid synthesis, or lipid and steroid biosynthesis. Polypeptide synthesis in a mammalian cell-free system from rabbit reticulocytes was less sensitive than the bacterial system from E. coli (Om et aI., 1984). Toxicity to Laboratory Animals

Acute and Chronic Toxicity Although mildiomycin is a structural relative of blasticidin (Fig. 3.11), the compounds differ greatly in their toxicities to plants and mammals. Mildiomycin has a very low acute toxicity to test animals (Table 3.12). In 30-day feeding trials, no treatment-related adverse effects were observed in mice or rats at 200 mg/kg daily doses. In a 3-month subacute study, the highest no-ob served-effect level was 50 mg/kg/day in rats. The antibiotic was nonmutagenic in

Table 3.12

Acute Toxicity of Mildiomycina

Species, sex

Assay

Rat, male/female Rat, male/female Rat, male/female Rat, male/female Rat Mouse, male/female Mouse, male/female Mouse, male/female Mouse, male Mouse Carp Japanese kiIIifish

oral sc iv ip dermal oral sc iv ip dermal 72-h 7-day 6-h

Daphnia pulex aFrom

Kusaka et aI.,

LDSO (mg/kg)

Other

4300/4120 463/684 885/700 679/842 >5000 5060/5250 II90/I 150 645/599 1020/1050 >5000 LCso > 40 mg/l LCSO > 40 mg/l LCSO > 20 mg/l

1979.

the Ames test with or without rat liver homogenate (Kusaka et aI., 1979). 3.3.1.4 Validamycin A Introduction Validamycin A (Fig. 3.11) is the major antifungal component of the validamycin complex isolated from the culture broth of Streptomyces hygroscopicus subsp. limoneus (lwasa et aI., 1971a). The first proposed structure of this aminosugar antibiotic (Horii and Kameda, 1972) was recently revised by Suami et al. (1980). Identity, Physicochemical Properties, and Uses IUPAC name: (IS)-(1,3,4/2,6)-2,3-dihydroxy-6-hydroxymethyl-4-[(IS,4R,5S,6S)-4,5,6-trihydroxy-3-hydroxymethylcyclohex-2-enylamino ]cyclohexyl f3 -D-glucopyranoside Chemical Abstract name: [IS-(la,4a,5f3,6a)]-1,5,6-trideoxy3-0- f3- D-glucopyranosyl-5-(hydroxymethyl)-I-[[ 4,5,6-trihydroxy-3-(hydroxymethyl)-2-cyclohexen-l-yl]amino]-Dchiro-inositol CAS Registry Number: [37248-47-8] Empirical formula: C20H3SN013; molecular weight: 497.5

Physicochemical Properties Validamycin A is a colorless hydrophilic powder without a sharp melting point: It softens at approximately 100°C and decomposes at approximately 135°C. The antibiotic is readily soluble in water, methanol, and dimethyl sulfoxide; sparingly soluble in ethanol and acetone; and insoluble in ethyl acetate and ethyl ether. Validamycin A is dextrorotatory: [a]54 = + 110° ± 15° (c = 1 in water). Its pKa is 6.0. Validamycin A monohydrochloride is a colorless crystalline powder with a melting point of 95°C (decomposition). The salt


is soluble in water, methanol, and dimethyl sulfoxide, slightly soluble in acetone and ethanol; and insoluble in ethyl acetate and ethyl ether. The optical rotation of the salt is [a]52 = +49° ± 10° (c = 1 in water). Stability Validamycin A is stable in mild alkaline and acidic solutions. It is stable in sunlight, but in soil microbial degradation is rapid with a half-life ofless than 2 h (Asano et aI., 1984; Matsuura, 1983). Formulations and Uses Validamycin A is formulated as 3% liquid concentrate and 0.3% dust for seed dressing. It is widely used for the treatment of sheath blight of rice, black scurf on potatoes, bottom rot on lettuce, and against other diseases caused by Rhizoctonia solani and other basidiomycetes fungi. Biological Properties Mode of Action and Biochemical Effects The mode of action of validamycin A was reviewed by Yamaguchi (1995). The antibiotic alters the morphology of R. solani by inhibiting the biosynthesis of myo-inositol, thereby reducing the pathogenicity of the fungus. Validamycin A was found to inhibit trehalases from R. solani (Asano et aI., 1987), as well as from various other organisms, including rat, rabbit, pig, yeast, and insect with an ICso ranging from 10-8 to 10-6 M (Kameda et aI., 1987). Trehalase (a,a-trehalose glucohydroxylase, EC 3.2.1.28) is widespread among many organisms and is important in regulating D-glucose transport into the intestines, reserving the supply of energy, germinating of spores, etc. The antibiotic was a poor inhibitor (lCso > 10-3 M) of other sugar-hydrolyzing enzymes such as porcine intestinal maltase, isomaltase, and sucrase (Kameda et aI., 1986). In R. solani, validamycin A was shown to be efficiently transported into the mycelia and hydrolyzed therein by a ,B-glycosidase to validoxylamine A [38665-10-0], a more potent inhibitor of trehalase of the fungus both in vitro and in vivo (Asano et aI., 1987). Validoxylamine A was also a strong inhibitor of mammalian intestinal, yeast, and insect trehalases (Kameda et aI., 1987; Kyosseva et aI., 1995). Certain sugars, such as L-sorbose (Trinci, 1985), fructose, glucose, sucrose, lactose, and mannose (Robson et aI., 1991), antagonized the growth inhibitory effect of validamycin A in Rhizoctonia species.

The intraperitoneal and intravenous LDso values in mice are greater than 13,000 and 10,000 mg/kg, respectively. No irritative effects on the skin at 10 mg/cm2 and on the cornea at 10 mg/eye of rabbits were observed. Acute exposure of rats to 12.46 mg/l air of validamycin aerosol caused no death or untoward reaction during exposure or 14 days afterwards. Oral administration of 12.5 g/kg doses to chicken and quail showed no treatment-related effects. The antibiotic at 10,000 ppm in the diet did not show any effect in rats and mice in 23-day subchronic studies. In a 2-year feeding study with rats, no treatment-related effects were seen for 1000 ppm (40.4 mg/kg) daily doses. Validamycin A at 40 and 10,000 ppm did not affect carp and killifish, respectively, in a 72-h toxicity assay (Anonymous (no year); Matsuura, 1983; Tomlin, 1997; see also Iwasa et aI., 1971b). Onishi and Miyaji (1973) summarized the results of 3-month feeding experiments with rats and mice given validamycin in food at 0.1-10% concentrations. For both species, the highest dose evoked an increased tendency of diarrhea from the third day, lasting for about 2 months. Analyses of the blood and urine of rats showed only minor treatment-related changes, generally in males. Pathological examinations revealed hypermucosecretion of the cecum, pneumonia, focal granulomatous formations in cardial tissues, and hepatic congestions for both rodent species, but these abnormalities were slight and sporadic. Reproduction, Teratology, and Mutagenicity Studies When rats were fed 500 and 10,000 ppm validamycin A, no death or abnormalities were seen in the Fo and FI generations or in teratology studies of the F 1 (b) and F2 (b) progeny. Validamycin was nonmutagenic in a battery of S. typhimurium and E. coli bacterial reversion-assay systems (Moriya et aI., 1983). 3.3.2 BACTERICIDES 3.3.2.1 Streptomycin Introduction Streptomycin (Fig. 3.12) was discovered as a fermentation product of the soil actinomycete Streptomyces griseus in 1943 (see Waksman, 1953). The aminoglycoside structure of the antibiotic was shortly established (Kuehl et aI.,

HN~NH,

Metabolism Rats, upon oral administration, rice plant, bacteria, and soil metabolize validamycin A into validoxylamine A and D-glucose. Rats, upon intravenous administration, however, excrete the intact antibiotic in the urine (Kameda et aI., 1975; Matsuura, 1983).

NH,

I

HN== 100 J.l.g/bee) (Tomlin, 1997). The inconsistency between poisoning incidents and the reported safety of the pure active ingredient prompted Koyama et al. (1997) to examine GLA and the anionic surfactant, SPAS, present in commercial formulations, for their cardiovascular effects in rats in vitro and in vivo. Whereas GLA had no effect on isolated atria and aortas, both the herbicide formulation and SPAS produced negative chronotropic responses in isolated atria and exerted significant vasodilative activity in phenylephrine-pretreated aortic ring segments. Intravenous administration of either the herbicide formulation or SPAS at 0.330 mg/kg reduced blood pressure in a dose-dependent manner. Additional symptoms noted were a slight increase in heart rate for the low doses and a marked decrease at the 30 mg/kg dose. In contrast, GLA failed to produce any of these effects, strongly suggesting that the hypotensive effect of the commercial formulation is caused by the surfactant SPAS. Toxicity Symptoms and Pharmacology Hack et al. (1994) reported the results of a detailed mammalian pharmacological and neurotoxicological study. Intracerebroventricular administration of 10 J.l.g GLA to male rats elicited slight spasms of

154

CHAPTER 3


Table 3.13 Acute Toxicity of Glufosinate-ammonium and Its Formulated Products LDSO Species, sex

Route

Rat, male/female

oral

1660/1510

Ebert et aI., 1990

Rat, male/female

oral

2170/191Oa

Ebert et aI., 1990

Rat, male/female

sc

73/61

Ebert et aI., 1990

Rat, male/female

ip

96/83

Ebert et aI., 1990

Rat, male/female

dermal

>4000/4000

Ebert et aI., 1990

Rat, male/female

dermal

14001l380"

Ebert et aI., 1990

Mouse, male/female

oral

436/464

Ebert et al., 1990

Mouse, male/female

oral

1420/1570"

Ebert et aI., 1990

Mouse, male/female

sc

881104

Ebert et aI., 1990

(mg/kg)

Other data

Reference

Mouse, male/female

ip

Rabbit

oral

Trout

96-h

LCso = 710 ppm

Trout

96-h

LCso :=:: 15 ppma

Hoerlein, 1994

Bluegill

96-h

LCso :=:: 320 ppm

Hoerlein, 1994

103/82

Ebert et aI., 1990

1550"

Ebert et aI., 1990 Hoerlein, 1994

Bluegill

96-h

LCso = 56-75 ppma

Hoerlein, 1994

Daphnia

48-h

Daphnia

48-h

LCso :=:: 560 ppm LCso :=:: 15 ppma

Hoerlein, 1994

Green algae Zooplankton community aA

bA

6-day

Hoerlein, 1994

LCso:=:: 37 ppm

Hoerlein, 1994

ECso = 0.24 ppmb

Faber et aI., 1998

200 gII glufosinate-ammonium liquid formulation containing alkylether sulfate wetting agent, propylene glycol ether, defoamer, dye, and water. 137 g/l glufosinate-ammonium liquid formulation.

the forelimbs and opisthotonos (arched-back body), which was responsive to diazepam (10 mg/kg). A 20 ).lg dose produced general convulsions after a latency period of approximately 3 h. The symptoms could be alleviated by an intraperitoneal injection of diazepam but recurred within 24 h. An analysis of catecholamine levels in various parts of the brain revealed a significant increase in the striatal dihydroxypheny lacetic acid level and a decrease in the norepinephrine level of the frontal cortex for animals receiving the 20 ).lg dose. No effects were seen for the 10 ).lg dose treatment. Brain glutamine synthetase activity, however, showed slight to moderate dose-dependent inhibition not only after intracerebroventricular but after intravenous (100 mg/kg) application as well. Oral application of a single dose of 1600 mg/kg of glufosinate to female rats caused poisoning symptoms, starting with diarrhea 6 h after treatment. Convulsions, restlessness, and piloerection were also observed. Intoxication reached a maximum on days 2 and 3 after treatment, with tonoclonic convulsions, squatting position, lagophthalmos, drowsiness, reduced respiration, and blood-encrusted eyelids and snouts. Some of the animals succumbed. At lower doses, the symptoms were less severe. The sign of intoxication receded 3.5 days after treatment. For the orally treated animals, a decrease in glutamate synthetase activity occurred in the liver and the kidneys. In the kidneys, enzyme inhibition was detectable 4 h after dosing, was maximum on day 1, and disappeared completely within 7.5 days; in the liver, it was more lasting. The brain enzyme

was less sensitive. The glutamate level slightly increased in the liver and decreased in the brain. GLA had no effect either on Ca2+ ion channels from rat frontal cortex or on GABA, benzodiazepine, norepinephrine, dopamine, and serotonin rat or bovine receptors in vitro. A circadian-stage-dependent toxicity study in mice reported that mortality was highest when the animals received the GLA formulation (Basta®) (oral 1500 and 3000 mg/kg doses) at 9 PM, the beginning of the light-on phase. Mortality was lowest at 9 AM (Yoshiyama et aI., 1995). Mutagenicity, Oncogenicity, and Reproductive Toxicity GLA was nonmutagenic in bacterial, yeast, and in vitro and in vivo mammalian genotoxicity assays. There was no evidence for oncogenic potential in mice in a 2-year study with a maximum of 160 ppm (males) or 320 ppm (females) dietary concentration of the herbicide. In a two-generation reproductive toxicity study with rats, the no-effect-level for fertility and reproductive performance was 120 ppm dietary GLA, equivalent to an average daily dose of 12 mg/kg for the dams during pregnancy and lactation. In embryotoxicity studies, maternal toxicity was noted for rats at the highest doses of 50 and 250 mg/kg, whereas, for rabbits, at the highest dose of 20 mg/kg. The herbicide was not teratogenic (Ebert et aI., 1990). GLA caused growth retardation, various morphological abnormalities, and lethality in developing mouse embryos in vitro at 10-5 M concentrations and in cultured embryonal cells of the midbrain and the limb bud at 10-6 M concentrations (Watan-

3.5 Rodenticides

1~~

abe, 1997; Watanabe and Iwase, 1996). GLA specifically affected the neuroepithelium of the brain vesicle and neural tube, leading to apoptosis by an unknown mechanism; nevertheless, the observed excitotoxic cell death was similar to that caused by glutamate observed in other studies.

The convulsion occurring on day 2 was responsive to thiopental sodium and diazepam. Magnetic resonance imaging demonstrated slight ischemia-induced changes in the white matter of the lateral regions of the brain. During recovery, retrograde and anterograde amnesia appeared.

3.4.2.5 Toxicity to Humans

Treatment Treatment is generally symptomatic as was illustrated previously. Activated charcoal is expected to adsorb the surfactant but not GLA in the herbicide formulation. Diuresis is believed to be essential, but the clinical usefulness of hemopurification remains to be demonstrated. It is essential to monitor vital signs closely during hospitalization for several days after poisoning regardless of the amount ingested, to provide respiratory support, and to expect delayed symptoms, even when the primary treatment is complete (Koyama, 1999). Based on a study with bovine blood contaminated with 268280 ppm herbicide formulation, Tanaka et al. (1995) preferred hemodialysis to direct hemoperfusion because the former removed over 99% of GLA, whereas the latter extracted merely 3% of GLA within 2 h; the removal of the surfactant, however, was not examined. The efficacy of hemodialysis and hemoperfusion in removing GLA was also demonstrated in a human poisoning case with initial serum concentration of 1.79 g/l GLA (Shinohara et aI., 1997).

Poisoning Incidents and Treatments Suicidal ingestions of the liquid herbicide formulations containing GLA caused respiratory failures and delayed nervous system disorders with a mortality rate of 19% in Japan (Koyama, 1999). The pathophysiology of human GLA poisoning has not been clarified. In general, the initial poisoning symptoms are basically due to the gastroirritant effect of surfactant in the herbicide, whereas the neurological and circulatory failures occurring after a latency period of about 1 day could be related to the direct or indirect effect of GLA. Koyama et al. (1994) describe a poisoning case of a woman ingesting 500 ml of Basta® containing 18.5% GLA plus 35% anionic surfactant. Immediately after ingestion, she developed nausea and vomiting, which subsided within a few hours. Treatment consisted of gastric lavage, administration of charcoal and a cathartic, and forced diuresis with furosemide. She, however, gradually lost consciousness and was in deep coma with general cyanosis 9 h after ingestion. Intubation followed by artificial ventilation was initiated. Hemoperfusion for 4 h did not improve clinical signs. Consciousness and spontaneous respiration were slowly regained and extubation was possible on day 8. The patient had generalized edema from days 1-5 and elevated body temperature up to 40°C from days 1-8. Endoscopy showed erosion of the gastric mucous membranes. No convulsions developed in this case. Toxicity was attributed to the anionic surfactant in the herbicide formulation. Based on this and 10 other poisoning cases, the human acute oral toxic dose that causes delayed consciousness disturbance was estimated to be 1.6-1.8 ml/kg of Basta® (296-333 mg/kg of GLA) (Koyama et aI., 1995). Tanaka et al. (1998) reported two suicidal ingestion cases, which were complicated by general convulsions that developed 8.5 and 33 h after ingestion. Watanabe and Sano (1998) described in detail the suicidal poisoning case of a man who ingested 180 ml herbicide formulation (corresponding to 33.3 g GLA). Symptoms, developing shortly after ingestion, included vomiting, diarrhea, and impaired consciousness. Upon hospitalization, metabolic acidosis, a body temperature of 35.4°C, a pulse rate of 110 beats/min, and no detectable diastolic blood pressure were determined. Emergency treatment consisted of intubation and gastric lavage, followed by the administration of charcoal, diuretics, and a purgative. White blood cell, glucose, and urea nitrogen levels were elevated during the first 5 days, whereas lactate dehydrogenase activity peaked on day 4 of hospitalization. The cholinesterase level was reduced during the first 5 days. The urine glufosinate level was higher than 40 J.1g/ml on day 2 but undetectable on day 3 after direct hemoperfusion.

3.5 RODENTICIDES Rodenticides are used to kill mammals that compete for our food, are vectors of fatal diseases such as rabies and the plague, and damage buildings, dams, or underground cables. Rodent control agents occupy a unique place among pest control agents due to their high vertebrate toxicity and also because these poisons have been among the most frequently misused pesticides (Barnett and Fletcher, 1998; Parsons et aI., 1996). The appearance of resistance to the synthetic anticoagulant rodenticides requires alternative vertebrate control products. In the case of an outbreak of disease, single-dose toxicants of natural origin, such as strychnine and red squill, provide economic and rapid reduction in the rodent population. Some of the rodenticides surveyed in this chapter appear to have become obsolete; nevertheless, their unique toxicology and the fact that, from time to time, they still resurface in some countries justifies a brief discussion. General reviews on the history of rodent control (Chitty, 1954) and on the chemistry of current important natural and synthetic rodenticides (Elliott, 1995) are available.

3.5.1 STRYCHNINE 3.5.1.1 Introduction The extremely poisonous alkaloid strychnine (Fig. 3.14) was isolated in pure form by Pelletier and Caventou in 1818 from St. Ignatius beans, Strychnos ignatii (Loganiaceae), a woody vine native to the Philippines. It is now obtained from the ripe

156


birds. Wild bear population can also be effectively controlled by strychnine nitrate baits (Inukai, 1969). Due to its bitterness, however, the alkaloid is not suitable as a rat poison (bait shyness). In the United States, strychnine-containing products are generally classified as restricted use for below-ground applications such as for control of gophers (U.S. EPA, 1996). Typical formulations in baits are pellets, grain, or eggs containing 0.25-1 % of the alkaloid. Strychnine has also been listed in pharmacopeias of many countries as a tonic and stimulant in veterinary and human medicine.

o strychnine

°

3.5.1.3 Biological Properties

0y °

scilliroside R = II-D-Glc

Figure 3.14

Structures of rodenticides.

and dried seeds of a related plant, S. nux vomica, growing in India, Sri Lanka, and Southeast Asia. The seeds contain 1.0-1.5% strychnine and about the same amount of its 2,3-dimethoxy derivative, brucine.

3.5.1.2 Identity, Physicochemical Properties, and Uses Chemical Abstract name: strychnidin-lO-one CAS Registry Number: [57-24-9] Empirical formula: C21H22N202; molecular weight: 334.4

Physicochemical Properties Strychnine is an odorless base (pKa = 8.26) forming, in pure form, colorless or white crystals that melt at 275-285°C with decomposition. The solubility of the free alkaloid and its sulfate in water is 143 mg/l at ambient temperature and 30 g/l at 15°C, respectively (Tomlin, 1997). The solubility of the free base in ethanol and chloroform is 6.7 and 200 g/l, respectively. The log P of the free base is 4.0 at pH = 7. Strychnine is levorotatory: [a]& = -139 (c = 0.4 in chloroform). Strychnine is very bitter with a taste threshold of 1.4 ppm in solution (Budavari, 1996). A dilute solution of strychnine in 80% sulfuric acid gives a reddish-violet to bluish-purple color on the addition of a trace amount of potassium dichromate solution (Quo reaction). Under abiotic conditions, strychnine is a relatively stable compound. It is photostable and does not hydrolyze at pH 5-9. The alkaloid is immobile in the soil where degradation is entirely microbial (Rogers et aI., 1998b; U.S. EPA, 1996). 0

History of Use and Formulations The use of the seeds of S. nux vomica as a rat poison was introduced in Germany in the late 17th century. Since then, strychnine, in one form or another, has been used worldwide to kill vertebrate pests, including moles, skunks, gophers, mice, rabbits, coyotes, and various predators, as well as sparrows, pigeons, and other unwanted

Mode of Action Strychnine is a strong convulsant. The alkaloid excites the CNS by specifically antagonizing the inhibitory neurotransmitter amino acid, glycine, at postsynaptic receptors (reviewed by Breitinger and Becker, 1998). Inhibitory glycine receptors are abundant in the spinal cord and brain stem where they are mainly involved in regulation of motor functions. Strychnine-binding glycine receptors were also found in the cortex, as well as in the auditory and visual systems. When inhibition is blocked, ongoing neuronal excitability is increased and sensory stimuli produce exaggerated reflex effects. Glycine receptors in higher brain centers such as the substantia nigra, neostriatum, and hippocampus are commonly insensitive to strychnine, explaining why strychnine symptoms are largely spinal in origin (Bristow et aI., 1986). Strychnine can also depress nicotinic-cholinergic responses through interaction with nicotinic receptors and at high concentrations in vitro binds to other receptors as well. In rodent spinal cord, there are two major isoforms of the inhibitory glycine receptor. The receptor variant of newborn rodents is a homopentamer made up of polypeptides with a molecular weight of 49 kDa (Becker et al., 1988). The adult glycine receptor isoform, however, is a complex glycoprotein consisting of three polypeptides with molecular weights of 48, 58, and 93 kDa (Pfeiffer et aI., 1982). The neonatal receptor is predominantly expressed around birth and has low strychninebinding affinity; within 2 weeks after birth, it is replaced by the adult receptor form, which is strychnine sensitive (Becker et aI., 1988; Briining et aI., 1990). Absorption, Metabolism, and Excretion Strychnine is rapidly absorbed from the gastrointestinal tract and nasal mucosa but not through the skin. Distribution of the drug in tissues is also rapid as is its metabolism to several nontoxic polar products by hepatic enzymes (Adamson and Fouts, 1959; Oguri et aI., 1989). Only 5-20% of the intact alkaloid is excreted in urine. The metabolism of the alkaloid was inhibited by the cytochrome P-450 blocker SKF-525 in rodents (Adamson and Fouts, 1959; Kato et aI., 1962) but was induced by phenobarbital (Kato et aI., 1962). The observed different oral toxicities of strychnine to guinea pigs and to rats were attributed to different metabolic rates in these rodents (Kato et al., 1963; see Table 3.14). The observation that female rats were more

3.5 Rodenticides

157

Table 3.14 Acute Toxicity of Strychnine to Laboratory Animals

Species

Route

LDSO (mglkg)

Note

Reference

Rat, male

iv

0.57

Rat, female

iv

0.57

Kato et aI., 1962

Rat, male

ip

2.82

Kato et al., 1962 Kato et aI., 1962

Kalo et aI., 1962

Rat, female

ip

1.62

Rat, male

sc

4.01

Kato et aI., 1962

Rat, female

sc

1.81

Kato et aI., 1962

Rat, male

ip

3.1

Rat

ip

1.5

at 26°C

Keplinger et aI., 1959

Rat

ip

0.25

at 8°C and 36°C

Keplinger et aI., 1959

Blum and Zacks, 1958

Rat

iv

0.96

Setnikar and Magistretti, 1967

Rat

sc

2.73a

Kamel and Afifi, 1969

Rat, male

oral

6.4

U.S. EPA, 1996

Rat, female

oral

2.2

U.S. EPA, 1996

Mouse

oral

8.0

Schafer and Bowles, 1985

Mouse, male

sc

1.45

Kretzschmar et aI., 1970

Mouse, male

ip

Lakatos et aI., 1964

Mouse, female

ip

1.9b 1.6b

Guinea pig, female

ip

10.9b

Guinea pig, female

iv

0.39 b

4.8 b

Kato et aI., 1963

>2000c

U.S. EPA, 1996

Guinea pig, female

sc

Rabbit

dermal

Mongrel dog

sc

0.46a

Lakatos et aI., 1964 Kato et aI., 1963 Kato et aI., 1963

Kamel and Afifi, 1969

a Strychnine

hydrochloride. sulfate. cNo signs of toxicity observed. b Strychnine

susceptible to strychnine than males (Poe et aI., 1936) was explained by more efficient hepatic metabolism in the latter sex (Kato et aI., 1962). The oxidative nature of the metabolism was demonstrated in rats where strychnine 21,22-epoxide and strychnine N -oxide were identified as the respective major and minor urinary metabolites (Oguri et aI., 1989). In nonfatal human poisoning cases, strychnine disappearance followed first-order kinetics with a half-live of 10 h (Edmunds et aI., 1986) or 16 h (Palatnick et aI., 1997). Degradation of strychnine by various soil microbes was recently shown in laboratory studies done in connection with a major mouse plague control program in South Australia during 1993 (Rogers et aI., 1998a, b). 3.5.1.4 Toxicity to Animals Acute Toxicity The toxicity of strychnine and its salts was thoroughly studied, and representative acute toxicity data are listed in Tables 3.14 and 3.15 for experimental and wildlife animals, respectively. For additional data, see Ray (1991). The acute toxicity of strychnine to rats was shown to be influenced by environmental temperature (Keplinger et aI., 1959) and altitude (Moore and Ward, 1935).

The poisoning syndrome in animals is essentially the same as that observed in humans and is described in Section 3.5.1.5. Although strychnine is generally less toxic to avian species than to mammals, the way it is commonly applied poses a danger to nontarget birds (Warnock and Schwarzbach, 1995; Wobeser and Blakley, 1987). Pathology Autopsy findings are nonspecific and reflect only the presence of violent convulsions and anoxia. Hemorrhages were sporadically observed in the brain of poisoned rats (Pensa and Ceccarelli, 1968) and in the myocardium and intestines of poisoned aquatic birds (Sterner et aI., 1998; Wobeser and Blakley, 1987). Treatment of Poisoning in Animals Because strychnineinduced death is mainly due to respiratory failure, artificial respiration will protect animals from an otherwise fatal dose of strychnine. Pentobarbital, mephenesin, and diazepam are traditionally useful drugs in treating strychnine poisoning (summarized by Ray, 1991). Kavapyrone constituents of the roots of the Polynesian kava plant, Piper methysticum, were recently shown to antagonize

1:'8


Table 3.15 Toxicity of Strychnine to Wildlife Species

Species

LCso (ppm)

Other data

Reference

LDSO < 5.oa

Schafer and Bowles, 1985

Acute toxicity Starling (oral) (Sturnus vulgaris) European ferret (dietary, 5-day) (Mustella putorius)

198

U.S. EPA, 1996 LDJOO = 31 mg/egg/skunk

Striped skunk (dietary)

U.S. EPA, 1996

(Mephitis mephitis) Red fox (dietary, 5-day) (Vulpes fulva)

70

U.S. EPA, 1996 Lethal dose

Bear, (acute oral)

= ~0.5 mg/kgb Inukai, 1969

(Ursus arctos yesoensis) Subacute dietary toxicity Northern bobwhite quail (Colinus virginianus) Northern bobwhite quail (28-day)

3536

U.S. EPA, 1996

4974

Sterner et ai., 1998

Mallard duck (Anas platyrhynchos) Mallard duck

212

U.S. EPA, 1996

680

Sterner et ai., 1998

Black-billed magpie (Pica pica)

99

U.S. EPA, 1996

234

U.S. EPA, 1996

American kestrel (Falco sparverius) aStrychnine sulfate. bStrychnine nitrate.

the convulsant and lethal action of strychnine in mice (Kretzschmar et aI., 1970). Upon intraperitoneal pretreatment at 300 mg/kg, methysticin (LDso = 530 mg/kg) raised the subcutaneous LDso of strychnine from 1.45 to 7.3 mg/kg, making it more active than mephenesin and less active but safer than phenobarbital. The kavapyrone anticonvulsants were shown to exhibit neuroprotective activity against experimentally induced ischemia in rats and mice (BackhauB and Krieglstein, 1992). Strychnine-induced tonic extensor seizures in mice were selectively and effectively blocked by a novel anticonvulsant aryltriazole derivative (MDL 27,531), administered either orally or intraperitoneally (Kehne et aI., 1992). Interestingly, the compound did not affect eH]strychnine binding in mouse brain stem and upper spinal cord membranes. Picolinic acid and its methyl ester showed anticonvulsant activity against strychnine-induced seizures in mice at 200 mg/kg intraperitoneally. The compounds also had a muscle-relaxant effect on rat decerebrate rigidity (100 mg/kg iv) and depressed spinal reflexes in cats at cumulative doses of 25-200 mg/kg iv (Tonohiro et al., 1997). Mylacemide (2-n-pentylaminoacetamide), an orally active anticonvulsant glycine prodrug that is able to cross the bloodbrain barrier, selectively inhibited strychnine-induced allodynia in rats at doses of 100-600 mg/kg iv (Khandwala and Loomis, 1998).

3.5.1.5 Toxicity to Humans Acute Toxicity The typical lethal dose of strychnine is between 50 and 100 mg for adults and 15 and 30 mg for children; however, much higher doses were reportedly tolerated (reviewed by Perper, 1985; Ray, 1991). In adults, symptoms can develop from doses as low as 30 mg. Poisoning Syndrome and Laboratory Findings The clinical syndrome of strychnine poisoning is very characteristic (Smith, 1990; Swissman and Jacoby, 1964). The initial symptoms are stiffness and twitching of face and neck muscles and movements may be abrupt. Reflex excitability is heightened and sudden tactile, visual, or acoustic stimuli induce violent motor responses. Within 30 min after ingestion, full tetanic convulsion and opisthotonos develop. The jaws are fixed (risus sardonicus) and froth gathers at the mouth. Seizures usually occur at 10- to 15-min intervals, last 30 s to 2 min, and are accompanied by loud moans of pain. During seizures, the patient remains conscious, which can arouse panic. Hyperthermia may occur and body temperature as high as 43°C was reported (Boyd et aI., 1983). The contractions of the diaphragm and thoracic and abdominal muscles halt respiration and cause cyanosis and marked anoxia. Death is due to brain damage secondary to apnea from

3.5 Rodenticides uncontrolled seizures (Dittrich et al., 1984) or cardiac arrest (O'Callaghan et al., 1982). For patients recovering from strychnine poisoning, lactic acidosis (arterial pH = 6.55) and rhabdomyo1ysis with an increased creatine phosphokinase level are typical (Boyd et al., 1983). In a serious, but nonfatal case, the classical clinical poisoning syndrome was complicated by acute chemical pancreatitis (Hemandez et al., 1998). Metabolism and Excretion Within a few minutes of ingestion, strychnine appears unchanged in the urine, but the major route for removal and detoxification is oxidative hepatic metabolism (Perper, 1985; see also Boyd et aI., 1983). The urine and gastric aspirate are the most useful specimens for confirming the diagnosis. A strychnine serum level as high as 2.17 I-Lg/ml at 6 h after poisoning has been reported (Hemandez et al., 1998). Poisoning Incidents Because of the limited use of strychnine today, the number of fatal accidents caused by this nonselective poison has decreased and most of the reported cases are suicidal ingestions. There were 1000 reported cases and at least 4 deaths due to strychnine rodenticides in the United States between 1985 and 1990 (Klein-Schwartz and Smith, 1997). Of the 73 strychnine rodenticide poisoning cases recorded in France between 1973 and 1994, 12 were fatal (Fran

CL>

Of)

0 "Cl

'C

£

e CL>

0:

~ U

0 0

~

Of)

§

e CL>

"Cl 0

~

'§ il

CL>

C:i

~::s

ro

'C

"5

~

:0 {i

~

0.-

x

x

x

x

x

x

x

x

x

x

x

x

0

C

~

Cl

::s

Z

0:::

Anticoagulants Brodifacouma

x

x

x

Bromadiolonea

x

x

x

Chlorophacinone

x

x

x

Difethialone"

x

x

x

Diphacinone

x

x

x

Pindone

x

x

x

Warfarin

x

x

x

Bromethalin

x

x

Cholecalciferol

x

x

x x

x

x

x

Non-anticoagulants

Strychnine Zinc phosphide a Second-generation

x x

x

x

x

x

x

x

X

anticoagulants,

or vegetables. Muskrat and nutria populations that must be reduced may be controlled with zinc phosphide baits prepared with cut-up chunks of apples, sweet potatoes, or carrots. These are prepared by coating the chunks with a vegetable oil and then blending them with the appropriate amount of zinc phosphide concentrate. Such perishable fruit or vegetable baits must be prepared fresh and applied shortly thereafter, prior to deterioration. Such baits are relatively expensive to prepare and therefore are only used for the control of those pest species that have little fondness for cereals. Liquid baits may be an appropriate alternative to food baits for some situations. Because most animals are attracted to water and utilize it, liquid baits are not very selective for the target species. For this reason, their use is limited to the control of commensal rats and house mice in select situations such as warehouses and manufacturing facilities where cereal-based baits may not be well accepted for various reasons. Water or liquid baits are prepared with water-soluble sodium salts of anticoagulants such as diphacinone, pindone, or warfarin. 7.9.1.2 Methods of Bait Use

In practice, toxic rodent baits are used or applied in various ways, depending on the pest and the situation or crop. In urban environments, rodent baits are placed as close as possible to the location where rodent sign or evidence of their activities is present. For safety concerns, baits are placed in locations inaccessible to children and pets or placed in tamper-resistant bait stations. Tamper-resistant bait stations are enclosed boxes usually made of metal or sturdy plastic and are of a sufficient

size to permit rats or mice to enter through one of two small holes located on opposite sides of the station. The toxic bait placed within these stations is not only protected from access by larger nontarget species, but the rodents are provided a secluded location to feed on the bait. The amount of bait placed in each station may vary from about 6 oz for house mice to several pounds for rats. The criteria for a tamper-resistant bait station are dictated by the EPA and require that its design be such that a young child cannot reach into the bait station through the holes to touch the reservoir of bait. To be tamper resistant, the stations must, by some means, be anchored to the floor or ground so that bait cannot be shaken out, and the lid must be secured so that a child cannot gain access. Professional structural pest control operators routinely use tamper-resistant bait stations when controlling commensal rats and mice. Tamper-resistant bait stations, because of the added cost, are rarely used by nonprofessionals conducting their own rat or mouse control, and this, unfortunately, is a contributing factor to accidental child and pet exposure to rodent baits. Toxic water or liquid baits are often presented to the rodents in shallow containers placed within a bait station. Liquid rodent baits are generally used within industrial-type buildings such as warehouses and other enclosed facilities, providing all other animals are excluded. Water baits are sometimes offered in reservoir-type chick fountains, which hold a larger supply of liquid bait. Water baits are very rarely used and then mostly by structural pest control operators. They are not advantageous in most situations; they are time consuming to prepare, and evaporation and deterioration are relatively rapid. Ingredients to prepare water baits are not marketed to the public.

7.9 Lethal Vertebrate Pesticides

In rural settings, anticoagulant-filled bait stations are commonly used around farm buildings for the control of house mice, deer mice, and rats, as they prevent pet and livestock access to the poison bait. In agricultural crops, bait stations, although not necessarily of the tamper-resistant type, are frequently used when anticoagulant-type baits are applied for ground squirrels, Norway and roof rats, and muskrat. When baiting ground squirrels, specially designed bait stations have been devised to exclude the endangered San Joaquin kit fox as well as endangered kangaroo rats. Other designs exclude threatened deer mice when baiting for roof rats. Baits for some agricultural pests are placed within the burrows, as is done for pocket gophers, or in the burrow entrances, as is sometimes done for Norway rats and ground squirrels. As previously mentioned, perishable baits of cut-up apples, carrots, or sweet potatoes find limited use. When used for the control of muskrats, they are placed in floating bait stations; when used for nutria, they are offered on floating rafts anchored away from the bank. Perishable-type baits prepared with zinc phosphide fall under the "Restricted Use" category and require some expertise to prepare and safely use. Native muskrats are not a serious pest problem in most regions, and the introduced nutria is principally limited in distribution to the southeastern states, especially Texas and Louisiana. A common method of bait application used in agriculture is called "spot baiting" in which a small amount of bait is scattered on bare ground near the burrow entrance, as is done with zinc phosphide baits for prairie dog or ground squirrel control, or placed in the burrow openings or trails of voles. Baiting of voles in apple orchards, vineyards, sugar beet fields, and on non-crop land is commonly conducted by broadcasting, using some type of tractor-mounted power seeder calibrated to deliver the precise amount of bait per acre. Baits for ground squirrels are also sometimes applied by broadcasting. In some instances, the bait for vole control may be broadcast by airplane, when the acreages are large or when the ground may be too wet and soft to accommodate a vehicle-mounted broadcaster. Aircraft are also used to broadcast baits for the control of various rats that damage sugarcane, including native cotton rats, rice rats, and Florida water rats, as well as the introduced Norway and roof rats. Rat control in sugarcane grown in Hawaii includes the Polynesian rat, which is not found on the mainland. Maturing sugarcane is too tall and dense to accommodate any other type of bait application. 7.9.2 FUMIGANTS Fumigants are either toxic gases or substances that produce toxic gases that are lethal when inhaled. In vertebrate pest control, fumigants are principally used to control rodents in one of two ways, as a building or transportation vehicle fumigant or as a burrow fumigant. Fumigants have many advantages over other control methods because they do not require any particular behavior or action on the part of the target animal. Fumigation of buildings, rail cars, etc. is often conducted for

259

insect control and, depending on which fumigant is used, the process can also provide rat and house mouse control. Fumigation of buildings specifically for rodent control with methyl bromide and chloropicrin is sometimes conducted but it is generally prohibitively expensive. Building fumigation can only be conducted by licensed pest control operators under a strict set of regulations. Burrow fumigants are used outdoors against a wide variety of burrowing rodents, including Norway rats, chipmunks, ground squirrels, prairie dogs, woodchucks (marmots), and pocket gophers. Fumigants are also used, to a limited extent, as burrow or den fumigants to control certain carnivorous species such as coyotes, foxes, and skunks. Some are registered for use against moles; however, moles are not easily controlled with burrow fumigants. There are two fumigants that are commonly used in vertebrate pest control, aluminum phosphide and ignitable gas cartridges. Aluminum phosphide, a "Restricted Use Pesticide" to be used only by certified applicators, comes in tablet or pellet form. When the prescribed number of tablets or pellets are placed well within the burrow or den, they react with the soil or atmospheric moisture to produce lethal phosphine gas. The burrows or dens are sealed off with soil immediately following treatment to retain as much of the toxic gas as is possible and for as long as possible. The other and more commonly used fumigant is the ignitable gas cartridge, which is sold over the counter to the public. There are several manufacturers of these cartridges, but all generally contain two basic ingredients, sodium nitrate and charcoal, combined with smaller amounts of active or inert ingredients. The formulated ingredients are compressed into a cardboard tube with a fuse inserted in one end. When ignited, they produce a toxic suffocating smoke that is lethal to animals in a confined space. To use, the cartridges are placed in the burrow or den entrance and the fuse is lit. Once lit, the cartridge is pushed deep into the burrow or den with a shovel handle and the opening is sealed off with a soil plug and tamped tightly to retain the smoke. When used in accordance with the directions, gas cartridges present little hazard to the user. 7.9.3 TRACKING POWDERS Toxic tracking powders for commensal rodent control are applied in a thin layer on a solid surface where rats or house mice travel. When a rodent runs over a patch of toxic powder, the fine particles adhere to its feet and fur. Because rodents characteristically groom themselves by licking their paws and fur, sufficient toxic ant is ingested to be lethal. Ingestion is the means of exposure to these pesticides, as skin adsorption and inhalation are negligible. Tracking powders, sometimes referred to as grooming toxicants, are frequently formulated in fine clay at concentrations of active ingredient substantially greater than concentrations used in baits. This is because "tracking and grooming" is not a highly efficient method of delivering a toxicant to the target species. Zinc phosphide and the anticoagulants

260

CHAPTER 7

Vertebrate Pest Control Chemicals and Their Uses

(chlorophacinone and diphacinone) are the toxicants commonly used in formulating tracking powders. Tracking powders are applied with some type of duster over a rectangular area of a few inches wide, 12-18 in. in length, and about 1/16 in. depth. The treated spots are commonly referred to as "tracking patches" and are usually placed along walls where sign and other evidence suggests the rodents are traveling. Tracking powders can also be placed in specially made tubes with open ends, which allow the rodent to run through them, or they can be placed in shallow metal trays to facilitate easy cleanup. Such powders are also blown into wall voids where rodents are known to thrive. When used outdoors, tracking powders are not as effective; however, they are sometimes placed in the burrow entrance of Norway rats. They are usually ineffective if they become wet. These toxic powders are more effective on house mice than on Norway or roof rats because, in proportion to their size, more toxic powder adheres to house mice than to rats. Mice also spend more time grooming than do rats. They are not recommended for use in food-processing facilities or other critical areas where toxicant contamination of living spaces or food commodities may be possible. Tracking powders are "Restricted Use Pesticides" and not sold to the public. They are used mostly by licensed structural pest control operators. 7.9.4 GLUE BOARDS

Glue boards are designed to capture rats and house mice and consist of a 3/16- to 5/ 16-in. layer of sticky nonhardening adhesive with extraordinary holding properties. This adhesive is applied to a rectangular-shaped piece of cardboard or a similar tough material or placed in shallow plastic trays. The size of the glue board varies but generally ranges in the area of 2 1/2 x 6 in. for mouse-sized boards and 3 1/2 x lOin. for rat glue boards. The chemical formulas used to prepare these essentially nontoxic glues (i.e., adhesives) vary and remain trade secrets of the manufacturers. Glue boards are considered a type of trap and therefore do not come under the EPA's registration process. For use, glue boards are placed along walls or in areas where rodents are known to travel. When the rats or mice run over the glue boards, they are entrapped. In the struggle to escape, the rodents get their muzzles in the glue and suffocate. The glue boards and their contents are disposed of following use; they are not intended to be reused. Glue boards are used almost entirely indoors or under some kind of cover because they become less effective if they get damp or wet. Temperature extremes and an accumulation of dust on the surface reduces effectiveness. Special glue board covers are marketed that permit rodent entry and exclude pets. Covers also keep the entrapped rodents out of sight until disposal. Glue boards offer a nontoxic means of rodent control and can be very effective, especially for trapping house mice. They are extensively used by the public as well as professional structural pest control operators. Their use has increased greatly over the past 20 years and continues to grow.

7.9.5 LIVESTOCK PROTECTION COLLARS

The livestock protection (LP) collar, more commonly referred to as a toxic collar, is a relatively new device used to selectively kill livestock-depredating coyotes. Each collar, constructed to surround the neck of a goat or lamb, has within it two sealed pouches that together contain a small quantity (300 mg) of the "Restricted Use Pesticide," Compound 1080 (sodium ftuoroacetate), formulated in a liquid carrier. Collars are temporarily placed around the necks of a target group of 25-50 sacrificial young goats or lambs. Coyotes in making their kill usually attack the throat of their prey; in doing so, they puncture the collar and, in the process, ingest a sufficient amount of the toxicant to be lethal. The coyote may not exhibit symptoms and die for several hours and therefore may have fed on its kill and left the scene before it succumbs. Toxic collars are highly specific, targeting only those coyotes actually preying on livestock. Unfortunately, toxic collars have many limitations as to where they can be most effectively used, in addition to the fact that they are expensive and that a kid or lamb must be sacrificed for every coyote taken. Only a few states have been authorized by the EPA to use toxic collars, and then only in very rural areas. Those individuals utilizing them must go through an extensive approved training program to be specially certified to use the devices. The use restrictions and recordkeeping requirements have made this selective tool so burdensome to employ that it is used only as a tool of last resort, where the depredating coyote cannot be controlled by the use of traps, snares, or other standard methods. 7.9.6 TOXICANT EJECTOR DEVICE

The M-44 is a spring-activated device used to propel an orally active toxic ant into the mouth of a coyote when the device is triggered by the biting and pulling behavior of the targeted animal. The relatively small stakelike device consists of a capsule holder, a spring-activated ejector mechanism, a capsule containing 0.9 g of powdered sodium cyanide mixture, and a 5- to 7-in.-long specially designed hollow stake into which the ejector mechanism is inserted. Sodium cyanide is a "Restricted Use Pesticide." The M-44 is positioned and set just off the trails showing evidence of use by coyotes when entering the livestock area. In addition to coyotes, the device and sodium cyanide are also registered for taking red and gray fox. To set the device, the hollow stake is first driven into the ground. The trigger ejection mechanism is cocked, inserted, and secured inside the stake. The capsule holder, which has been wrapped with an absorbent material and loaded with a cyanide capsule, is screwed onto the positioned below-ground ejector unit. When set, only a few inches of the device projects above the ground surface. A small amount of fetid meat bait is applied to the absorbent wrapping surrounding the capsule holder. In addition to the baited device, a dab of coyote lure may be placed on a nearby bush to draw coyotes from a greater distance. Coyotes attracted by

7.9 Lethal Vertebrate Pesticides

the bait will try to bite the baited capsule holder. In the process, they will pull on the exposed capsule holder and trigger the device. The spring-activated plunger forcefully propels the sodium cyanide from the capsule through the open end of the holder and into the coyote's mouth. Death results within a few seconds. The M-44 is very selective for canids because of the baits and species-specific lures used and because the device is designed so that it can only be triggered by an upward pulL The device can be used with relative safety in pastures where livestock are present. Where M-44s are employed, the property is posted with warning signs to alert individuals to their presence. A special training program is required before the M-44s can be used. In some states, only federal employees involved in predator control are permitted to use the devices. In certain other states, the M-44s can be used by trained and certified livestock producers. The EPA has authorized the use of M -44s only in certain states that have a demonstrated need and have developed an appropriate training program. 7.9.7 FLOCK DISPERSAL AGENT

Avitrol® (4-aminopyridine) is registered as a flock-frightening repellent and is used in a bait form to frighten pest bird species such as pigeons, house sparrows, and certain blackbirds and cowbirds from structures and the vicinity of structures. In agricultural situations, the Avitrol bait may be used for a somewhat broader group of birds. The material is formulated on grain baits. This treated grain is then blended with untreated grain to give the appropriate dilution. The dilution ratio may vary depending on the pest species. Such diluted baits are placed in trays or on rooftops accessible to the target pest species. A period of prebaiting with a placebo precedes exposure of the treated bait. The ingestion of an active amount of Avitrol by a small proportion of birds causes the affected birds to emit distress calls and display erratic behaviors that frighten away the remaining birds of the flock. The use of diluted baits limits the number of birds affected. The material is sufficiently toxic that some of the birds that are affected will succumb. Dead birds are immediately picked up following treatment. When the flocks are adequately frightened, the birds may not return to that area for months. Avitrol is a "Restricted Use Pesticide" and can be used by or under supervision of government agencies, by licensed and certified structural pest control operators, and by certified applicators. It is not available to the public. Its use is relatively limited because a few birds may die outside of the treated property and this often results in an adverse public reaction. The use of the material is prohibited in some cities. 7.9.8 TOXIC BIRD PERCHES

Toxic bird perches have been used for some time to control nonnative pest birds, such as pigeons, house sparrows, and star-

261

lings, which often nest or roost on or in buildings. Toxic perches are horizontally mounted short tubes with a wicking system along the top edge. The tubes, which are closed at each end, can be filled with a liquid formulation of contact toxicant. The perches are fastened to level roof ridges or ledges or they are attached to vertical surfaces by a bracket so that the perch projects horizontally outward from the side of the building. Because perches are only about 24-27 in. long, they must be strategically placed so that the birds will alight on them repeatedly. It is essential that those installing toxic perches are very familiar with the pest bird species, especially its flight patterns and perching habits. Fenthion, which is scheduled to be phased out by the EPA, is the only toxicant currently registered for use in these perches. The liquid contact toxicant is taken up by the wick from the tubular reservoir to the top side of the perch. When the bird alights on the perch, its feet come in contact with the toxic ant-saturated wick and absorption takes place through the feet. The bird may have to alight on a perch more than once to accumulate a lethal dose. For specific pest bird problems, toxic perches are reasonably effective. Proper placement, as stipulated on the label, is essential so that only the target species is taken. Toxic perches are restricted to professional pest control operators and to persons trained in bird controL They are not used in situations where nontarget birds may be present and at risk or where secondary hazards may occur as a result of a predator or scavenger consuming sufficient contaminated birds. Because of these limitations, perches receive relatively little use in pest bird controL 7.9.9 POISON BIRD BAIT

Starlicide® (3-chloro-p-toluidine hydrochloride) is registered and marketed for starling and blackbird control in and around livestock- and poultry-raising facilities. It is formulated as a grain-based pellet and used by or under the direction of personnel trained in bird damage controL The active ingredient is frequently referred to as DRC-1339 or Compound DRC1339. The U.S. Department of Agriculture-Animal and Plant Health Inspection Service (USDA-APHIS) presently maintains the registration of the DRC-1339 concentrate, and its use is restricted to APHIS personneL The USDA-APHIS registration has been expanded to include several other pest species, namely, pigeons, gulls, ravens, crows, and magpies. In livestock feedlots and poultry operations, the bait is placed in feeding stations. Applications are made before the starlings and blackbirds arrive for their first morning feeding. Starlicide is ineffective for house sparrows and several other pest birds, as there is a wide variation in sensitivity to the toxicant among bird species. Hawks and mammals are relatively resistant to the materiaL APHIS personnel have used this material to control crows, ravens, and magpies that prey on the eggs or young of federally designated threatened or endangered species, as well as those preying on newborn livestock. Baits for these purposes may be prepared with eggs or meat.

262

CHAPTER 7

Vertebrate Pest Control Chemicals and Their Uses

The use of Starlicide and baits prepared with DRC-1339 concentrates is fairly limited; most of it is used in rural situations.

REFERENCES Hygnstrom, S. E., Timm, R. M., and Larson, G. E., eds. (1994). "Prevention and Control of Wildlife Damage." Nebraska Cooperative Extension Service, University of Nebraska, Lincoln, USDA-APHIS-Animal Damage Control, and Great Plains Agriculture Council.

Marsh, R. E. (1986). Vertebrate pest management. In "Advances in Urban Pest Management" (G. W. Bennett, and J. M. Owens, eds.), pp. 253-285. Van Nostrand-Reinhold, New York. Mason, J. R. (1998). Mammal repellents: Options and considerations for development. In "Proceedings of the 18th Vertebrate Pest Control Conference" (R. O. Baker and A. C. Crabb, eds.), pp. 325-329. Univ. of California, Davis. National Academy of Sciences (1970). "Vertebrate Pests: Problems and Control." Natl. Acad. Sci., Washington, DC.

C HA PTER

8 Pesticide Use in Veterinary Medicine Frcderick W. Oehme and Shajan Mannala Kan sas Slale Unive,,; ly

Do",.,{k animal, at< raised.nd IIoosed in • ".ri«y of cn";ron_ men ... AltOOugh ok>m<stic "nim.l•• ,. fie" cl'»e ly .ss' of our l ar~ •• nd ,mall JomesIk animal> ..,'<ml ti ...., each yo.,. These expot of p.. ficide, is offen .sth"ir. to IL l.! CATTLE keep Jog, .od c.~, from carrying "n, i~hlly .nd anno),ing lie.., Md tick •. but pe"idde, .1"" ore used to confrol rodents .nd The hc"'iesl expo'o", of ."i mal, to pticitb """"f'5 in beef infernal parasi ... _ The ' .. r)' n"ure of livestock operation, "'"", callIe ,hat are ",i.. d fur "",id wdght g.in.rod meat productivdi,pore "",nario of obeep aod goat. ne.rly J»nIlkb

I~_'

6 0rketO'O'ing to envimnmen,,1 pc"i,te"". ;" . 00 c,:nomic' ;n ,he pas! h..e led to lhe use of org,nochlorin< compouods in ,he", an_ im.I., A, lhc I"=n' ,ime, """" of ,he ori:."ochlorinc In · pounds are lMnned .,Id U", few remoin irlll 0"'" . re being ph. sed OU1, 00 e'P'r< rou· tinely ""wurmcd aod 'reated for p .... 'il;o ;nfo>,a,ion •. 1lJe prodIlClS in u>e i",,'ode vano.., pen""'Mn. coum.p/Ios. 00rameetin. amitraz. and tetnochlorvinphal of dogs, ... nlajot" of in,ecticid (KI"; . 1997). to OIher domestic pesticide' 11>< u", of corJIm,erci.ti dOll'oo ca' foo,b h •• ""arty elimDichlo"",. an Ued "' part of rouli". helmin,hie c"",ml in hors MO in cu""n' use. in · 8. 1.5 OOC S cluding ",'ico.agul.n' rudtnl;oiile>. zinc p!>OSphide •• trychni"" (oot cun-entlr ,,'ail.bl. 10 ,he B.... ral pobli.). 1080 (>Odium In'em.1 para.i,. medic .. ion .od u"'mally .pplied fl •• and ,ick mono"uoroact commonly"sed in dog ..... nd """"pta""" .mong pe' >OOiogs in oorthc:m Greec. frum 1'!90 ID 1995, 926 animal

cl."

'0

cl.""

'0

,ha,

'0

.'most

Iu,'.

,o.

,o. """d.

,i"".. , W~re analy,,,d hy chmrna'ogra ph;'; ,«bniques, I\:"idde, cau",d n% of ,be poi"",ingc"-"",. ",be'euall ",be,,",;.; ,ubsl:.neCS c.used 22%. The .0i"",1< affected were mainly c,". dogs, shtcp. bin!s, . nd bet:; (Alllonioo " ai" 1997), De'pile lhe lack 0( mono .pecifoc SI.,i"":,, il secm. crnain I",,' poiSo hO'" I ,hott.r life_span. ",hich m.kt, lbem k» li.bk 10 Iher cau>cs of dc .. h . ""h .. degcllC .... i"c di ....... . nd cancer, Anim." will bo injured"-,. re,ul, of inlOn,ion.1 applk ..ioo of pc"icide • • , .. ~rooli,.,,! moinly by ~Ju · C",<midali"" in the li,..,r. The se' of colonic obstipolion in pooi. probably .. a '"'luel 10 lrealmont ... ith.n amit ... , for· mul'tion. ha,< been "'portoo (Mu ..... "'.od ,·",,·der·V.lden. In8). Sick""" .Iso occurred in three of fou, ........... i,hin 24 hours of ... ing 'p"'y1 . ign, of I,.nquili,ati<m. depre"ion. a.. ,i •• muscular ineoorllination •• od impaction colic l.sting up 10", day" S ubcu,"neoo, edcma of the foe. """ur=! in one ........ Mild dchydr"i"" and ",-"ompani.d ,he .yn_ dl'tKl"l«rl vain on 1"'0 oollo"D MA U C IOUS PO lSO:>l lNG

Paraqu.u i, a re'lrided use herbi,-ide tha! i. "-"t",,,,,,ly ""i< to companion "nim.l •• nd I;"~ .. ock "'I,,,n illj!<sied. Etc,'on heif.... wrn: mode"''''l)' poi1Or>ed by p. raqu.t'p"'yod un grass along a dilch be.ide which !he heif.... ",".!Ud on Ul1UrC (Pin"" and JOoI'dan. 1'onoo Id """'..- ,,d ,he f«>d I"'tdl>l;.

A pad (I01e "'-'i"l typical dooIi ....uc:noe. inltibitioot P"i_inc'fuor 211 heir.,. co in """"'...., anima!o reiul, from itnorance or miirnamg""",n,-i n shott. ·". iOOlI' piiOftinll' am "'ill due to h.piu.llrd use f cllemicll , or rai lure "':gni •• the signific~""" f mud ifi.",ion, in their .miX"i,ilicido '" to' it:oIocic.1 Conc.n'B';"'" in .... appropriat" I)ioJollitll or ticid.. ~" i' in ",u" toxic df"", •. Clinic.1 .iln. ofl." appear wi,hin hour:< ofl., "'JlOOUl'C ODd J"umtiuliy demon...... ,..., 1>ioIot:. iul .««t~"" .. of tbo", poisom, Sy"""","", can ~ fmm Iht ...... roIoJ:iocaJ si"" ...,n .. illl lIIO>I of Iho i~ ID Iho bIof,.1,nl ond ... ~ .bftonnality ...... wi, h .......... u .... rodtnticide W"h ~ •• <eplioni.. "'" anendina >'CIerinari.. find, ",,~I .....NI! ol..-ly «l in ""m.ns for the 'pocific in~icitlc. funlitid'n)I and tile.,.."..·.

"""""itr,.

(o.",.ilo,.,

,i.,.

fi,·.

pal"""

..".,....

."""'at

.«ecI•.

..". SYS'.mic orll'""""",ph." empoU nd. p""."' ",me .p"d.1 problem. if '''''y lire DOt applied .. _wrop. If ac>PIied lOO cmy. "'" . . ,.. an: .... if applied ID late. ,be Iolenlly. 'hey rilise concern 01>001 the m.ny , .. ,iable f"'-"Un!h" impact '0' icily in dome>l", . "imals. o."",k Jiffe",nce" ",·en bnwe(:n iooiv;ciu.l,of tt..,.me bre.:d, and ...,urolugic.1 sen';li" ily from JIIOviou. chern"" ) exposure, contrihute to O>'entWlI pesticide llC"roto.icity (B"""" ., ~L J985).

.11"

'0

ha.

8.3.4 BJO LOGICAL ~10RAG[, EXCIU:1'10." . ANI) RESIlIUI'.s The "lIim,,,, challenge in u"ng P"'llicides in do."..,ic ani· m.l, i. to "void !he occurrellCe of ct..mic.l ""idu., in .nimal Jlloo",," intended for human con,umption. To !hi, end. .00 ncrrlion are studied in dome"ic ani"",l. early in pesliddc er degr.>da. ,j,.• rro<e~[>OornpI< '" """'" u;ft"..... .. ~._y 11'1o!tydrne b.... kdown, Gratinll.,i"",I. h,,'e." almost neu!ral ",men pH bccau"" tbe rum ingo,led by co"I • . >h~p• .,,- ~o." immediately"" ri," in tbi. I"'~. f.nnenting.OO reducing aCfe.
""are

'""cc

0""'"

II.4.S AGG IU:SSIO N ANI)

ANTJCII 0 l.ISIu", This iovol, .. remo ....1 frum ,bo offending ,ubslane< 01 1'Iu"... of 'he makes lhese """". ban" less eflki'''' in ,uminants, la.troi"",,,inal .y'tem m.1;c ""'.. I... lmenl procwu .. , Ies, C'.

REFERENCES

r1""""

AJo ' =~ A. A, M",<W)' " .... "'" .. the - ,IttJ,<J. F. ~ ..... C. ~_ .""-~ 1>04 .... ,\w. "- ... J_ .... l . F.h.,;".. 10._, 8 _ , K.. 1kV_,o..,,,,,. B. <J.,." N.. !;!w"",,,- w, C" Si;. I .......

Mpot ......

A.,,,,,. 1T7.

A."""

Muohom./, W" ...r Coop, M , ~ . !, ,m. pp, . _. """.,ad. 0....

""""".y

o/S'-"·. A........ p. I. " Vrt· m ...... 0 ..," T",,,,,,,,,,,- \\11, IOJ, W' lO9.j '9, F\>oqriod, o--u. ...

.. _

""- Sri ..

d _

J.'" lll-JJJ

"'~""'-A""

.~ .""

C,

1' I'

.. _ l_ _ _ ....... _ I _ U, llWIl.

f'

11._ ,

"

.....

_"'* _

......

CHAPTER

9 Pesticide Use Practices in Integrated Pest Management Frank G. Zalom University of California, Davis

Integrated pest management (IPM) has become broadly accepted as an approach to effectively manage insect, disease, nematode, weed, and vertebrate pests in many parts of the world. However, the complexity of IPM is not always appreciated, and its definition can be distorted to reflect the agendas of organizations and individuals that have come to embrace it. IPM is the management of pests in a systems framework, rather than a tactic or group of tactics for a specific pest or pest group. Many IPM tactics, although they may reduce chemical use, remain chemically intensive. IPM has been described as a continuum, with IPM systems ranging from those that are chemically intensive to those that embrace measures that prevent or avoid pest problems and utilize more biologically based tactics. A minimum level of IPM requires the use of scouting and decisions based upon established action thresholds. Moving IPM along the continuum toward the use of more biologically based methods of managing pests remains a challenge that requires extensive interaction between agricultural scientists, consultants, growers, and regulators to ensure relevant development and effective implementation of these increasingly more complex IPM systems. Crop consultants can play an important role in this process, but their role can become even more significant if the use of a broader array of pest controls is linked to their certification. A wide array of pests including insects, mites, weeds, nematodes, disease-causing organisms, and vertebrates lower the quality and the yield of agricultural products, affect the health of humans and other animals, invade structures and landscapes, and adversely affect natural ecosystems. Managing pests has always been a challenge. Before the introduction of synthetic organic pesticides in the 1940s, which allowed reduction of pest abundance and pest damage to levels that were not previously possible, farmers and others responsible for pest control typically employed multiple tactics, such as sanitation, crop rotation, crop diversity, bait trapping, and mechanical pest or host removal, which were applied preventatively based on knowledge of pest biology. Weeds were removed by hand hoeing and tillage; chemical herbicides were seldomly applied. They Handbook of Pesticide Toxicology

Volume 1. Principles

also used inorganic materials, such as copper, lead, antimony, and arsenic, or botanical compounds, such as nicotine and pyrethrum, which were available at the time. These materials were toxic and expensive to produce in quantity; therefore, availability was limited. Equipment for their application was relatively unsophisticated or lacking. Overall, pesticide use was low relative to contemporary levels. The chemical control paradigm was developed effectively by industry, government, and university researchers, and became widely implemented. Along with modem plant breeding, fertilization, and irrigation methods, the introduction of synthetic pesticides reduced on-farm labor requirements, facilitating the transition of agricultural production in developed countries to a highly mechanized system with relatively more concentrated production that is characterized by increased yields and reduced variability in production. Arguably, this transition has been beneficial in that fewer people must work on farms to produce the food and fiber products required to sustain an ever-growing population. The cost of food and fiber remains low as a proportion of income, and food supplies are relatively stable in developed countries. Unfortunately, in spite of an extensive regulatory system for registration, the increased use of pesticides has been accompanied by unintended social and environmental consequences. These consequences include documented cases of pest resistance and pesticide-induced pest outbreaks, environmental contamination, worker exposure, and public concern for residues on food. The only way to totally eliminate the risk of using pesticides is to prohibit their use, but at what cost? Pesticides are legally classified as economic poisons and are defined as substances used to control, prevent, destroy, or mitigate any pest. Pesticides include inorganic products like sulfur and naturally occurring botanical products like pyrethrum, both of which are acceptable for use by organic growers. Pesticides include vegetable and petroleum oils, fertilizers, and certain fatty acid soaps when they are used for pest control. Naturally occurring microbes, such as Bacillus thuringiensis and Trichoderma harzinium, are considered pesticides when they are pro-

275

Copyright © 200 I by Academic Press. All rights of reproduction in any form reserved.

276

CHAPTER 9

Practices in Integrated Pest Management

duced commercially and marketed as pest control agents. Many pesticides are very specific in their actions, acting as growth regulators, repellents, pheromones, desiccants, and defoliants. However, the general public seems most concerned with the use of certain synthetic pesticides, particularly those with broader activity and those with which they are unfamiliar. Agricultural uses in particular are not well understood by the public, so questions abound concerning the safety of these products and the need for their use.

9.1 INTEGRATED PEST MANAGEMENT In the 1950s, identification of pest resistance, pesticide-induced pest outbreaks, and the resurgence of pests that had been under controlled some researchers to call for "integrated control," that is, the combination of compatible biological and chemical control tactics. One of the first citations for integrated control in the literature involved walnut production, and it described the need for integration of pesticides so as to preserve parasites of the walnut aphid [Chromaphis juglandicola (Kaltenbach)] (Michelbacher and Bacon, 1952). Other researchers (e.g., Smith and AlIen, 1954; van den Bosch and Stem, 1962) supported the concept of integrated control to reconcile the use of insecticides and biological controls for insect pests. The concept was expanded to include economic thresholds by Stem et al. (1959), who called their approach integrated pest management or IPM. Economic thresholds are the pest densities at which the value of resulting damage exceeds the cost of applying a control. Their description of IPM added the requirements of pest monitoring and risk assessment before justifying the application of therapeutic measures such as pesticides. Environmental contamination by organochlorine insecticides was recognized in the 1960s, following the publication of the book Silent Spring by Carson (1962). Pesticide use became a political issue, and IPM was promoted as an acceptable approach for managing agricultural pests among some scientists and growers who were interested in applying "supervised control" rather than using strictly preventative pesticide treatments, which had become prevalent by that time. However, concerns about the slow rate of IPM adoption by farmers were raised by IPM researchers (e.g., van den Bosch, 1964). Funding for IPM research greatly increased during the 1970s and early 1980s, with increasing efforts to implement IPM practices through extension services, governmental agencies, and community-based programs. As the philosophy of integrated pest management matured, there grew an ever greater appreciation for integrating the management of weeds, pathogens, and nematodes as well as insects in a cropping systems context, recognizing that fundamental differences exist in the biology of these pests and, therefore, in the preventative and therapeutic measures that can be applied for their control. IPM strategies and tactics have gradually been adopted as alternatives to the conventional chemical control paradigm, and the breadth of institutions and organizations promoting IPM as the most effective way to reduce the risks of using pesticides has dramatically increased.

The United Nations Food and Agriculture Organization's (UNFAO) Panel of Experts on Integrated Pest Control (UNFAO, 1967) defined IPM as " ... a pest management system that, in the context of the associated environment and the population dynamics of the pest species, utilizes all suitable techniques and methods in as compatible a manner as possible and maintains the pest populations at levels below those causing economic injury." In the United States, several recent administrations have endorsed IPM. The United States Department of Agriculture (USDA) Council on Environmental Quality (1972) in its publication Integrated Pest Management wrote that IPM is " ... an approach that employs a combination of techniques to control the wide variety of potential pests that may threaten crops. It involves maximum reliance on natural pest population controls, along with a combination of techniques that may contribute to suppression--cultural methods, pest-specific diseases, resistant crop varieties, sterile insects, attractants, augmentation of parasites or predators, or chemical pesticides as needed." In urging IPM adoption in an environmental message, President Carter (1979) said that " ... IPM uses a systems approach to reduce pest damage to tolerable levels through a variety of techniques, including natural predators and parasites, genetically resistant hosts, environmental modifications and, when necessary and appropriate, chemical pesticides. IPM strategies generally rely first upon biological defenses against pests before chemically altering the environment." Attention to IPM in the United States increased again following the Clinton Administration's 1993 pledge to have 75% of cropland acreage under IPM by the year 2000. During the late 1980s and 1990s, several European countries set goals for reducing pesticide use by 50-75%, often suggesting IPM or integrated crop management as the preferred means to achieving the goals. In some countries, these goals have been met with documented use reductions. However, pesticide use reductions documented to date in these countries may be more the result of using products that are applied at reduced rates or by changes in application methodology than of using nonchemical approaches (Matteson, 1995).

9.2 WHAT IS IPM? Forty years after the term "integrated pest management" first appeared in the literature, a single definition has yet to be universally adopted. This is not unexpected because IPM can be as much a philosophy as a science. Because IPM has diverse proponents, the term has been adapted to support a variety of objectives and agendas. This adaptation has tended to permit narrow definitions of IPM to be proposed, in which it is mentioned primarily in terms of tactics, such as chemical controls or biological controls, which have particularly strong advocates. What was largely promoted as an ecologically based view of pest management by a relatively small group of academics and certain agricultural interests in the 1950s and 1960s has become a term for reaching consensus among government

9.3 The IPM Continuum

institutions, many mainstream agricultural and environmental organizations, agrichemical industry leaders, and sustainable agriculture advocates. Allowing diverse groups to reach common ground is indeed a strength of IPM. However, in many ways it has also become one of its major shortcomings. Depending on its interpretation, IPM can be used to justify current pest control practices, even those that are chemical intensive, without emphasizing reduced-risk alternatives or, more importantly, management of the pest species within an ecosystem framework. Cate and Hinkle (1993) stressed the ecological basis of IPM rather than the tactical emphasis of many IPM definitions in their report "Integrated Pest Management: The Path ofa Paradigm," by correctly stating that IPM is about the manner by which communities are managed. Perhaps the phrase itself has resulted in misinterpretation. Kogan (1988) identified "integrated" as the most ambiguous component of the term "integrated pest management." To many people, integrated refers to the use of multiple control tactics integrated into a single pest control strategy (e.g., Metcalf and Luckmann, 1982). This strategy most typically targets only one species of pest or a single class of pest and, in this sense, focuses upon control measures for the target species, prevention of natural enemy disruption and secondary pest outbreaks, and delaying development of pesticide resistance. A broader interpretation refers to management of the complex of pests that attack a crop, considering the combined effects of weeds, plant diseases, insects, and nematodes (e.g., Newsom, 1980). At its highest level, IPM incorporates interactions among pests, the crop, and the environment within the context of a social, political, and economic matrix. Prokopy (1994) likened the increasing levels of IPM complexity, from integration of control methods for a specific pest to its incorporation into a socioeconomic matrix, to the steps of a ladder, where progressing up the steps represents increased level of integration in a systems context. The word "management" as opposed to control also presents an important IPM concept. Flint and van den Bosch (1981) stated that the word "management" implies acceptance of pests as inherent components of an agricultural system. Indeed, some would say that acceptance of pests in an agricultural system is essential to permit their natural enemies to survive in an ecosystem. The IPM approach is to apply controls to suppress pest populations when necessary to reduce damage to an acceptable level, rather than to eradicate the pest.

9.3 THE IPM CONTINUUM Recently, IPM systems have been characterized as falling along a continuum (Sorenson, 1993) ranging from those that are more chemically intensive, where pesticides are applied based on scouting and the use of thresholds, to these that at biologically intensive, where reduced-risk pesticides may be applied, but in which biological controls and biologically based preventative approaches are predominant.

277

The USDA formalized the continuum concept in quantifying IPM adoption by creating categories for no IPM use and three additional levels, which represent progressively greater use of biological or cultural practices instead of conventional pesticides (Vandeman et aI., 1994); other groups have accepted and modified this approach (e.g., Benbrook et aI., 1996; Hoppin et at., 1996; Kogan, 1998). When presented as a continuum, the minimum criteria that constitute the use of IPM are field scouting for both pests and natural enemies, and using action thresholds where they exist to make pesticide use decisions. When an action is warranted, those people who employ a minimum level of IPM would apply selective or the "least disruptive" pesticides available. Although IPM emphasizes a systems approach to management, it is impossible to discuss the practice of managing pests without mentioning tactical intervention. 9.3.1 PESTICIDES

Pesticides often represent the first line of defense in situations of pest outbreak or where a specific pest must be eradicated for quarantine or public health purposes. As mentioned previously, pesticides may be used in an IPM system when applied based on scouting and strict consideration of available action thresholds. However, where choices of pesticides exist, those which are least toxic and present the lowest potential for disruption should be selected for use. Wiggles worth (1950) pointed out that it is sometimes " ... through the activities of the entomologists themselves that entomological problems arise." He also stated that "the public loves the hospital, the doctor, and the bottle of physic; while the advances in preventative medicine which have transformed our lives are scarcely noticed. So too it creates a greater impression on the mind to destroy an infestation of insects that can be seen, than by some simple change in practice prevent any infestation from developing." Pest resistance to specific pesticides, and pest outbreaks that result from applications of broad spectrum pesticides can occur and are well documented. Pest resistance to a chemical can develop rapidly, particularly when the life cycle of a pest species is relatively short and the chemical is repeatedly applied. In a pest population, there are always some individuals that will be genetically resistant to a pesticide. Even when a high percentage of the population is killed, those few individuals that possess the resistant traits will survive and reproduce, passing their genes to the succeeding generation. Thus, a pest population develops that can be controlled only by higher chemical dosages. Eventually, the pesticide being applied can become ineffective against the pest. Thus, most pesticides have a finite effective life. Pesticide resistance has been documented in hundreds of species of insects and mites, plant pathogens, weeds, rodents, and nematodes (see Georghiou, 1986). The best way to manage pest resistance is to apply pesticides less frequently. IPM tactics such as scouting or utilizing nonchemical approaches that reduce the need to apply pesti-

278

CHAPTER 9


cides may be helpful. When pesticides must be used, alternating classes of pesticides applied to reduce selection pressure on the pest population can delay the development of resistance. Approaches for monitoring susceptibility of pest populations to specific pesticides have been developed for several key arthropod and disease species. Such technology is relatively common for research applications and when applied by the pesticide manufacturer, but commercial implementation of resistance monitoring by consultants is rare. Many plant-feeding insects do not significantly damage agricultural crops because they are kept under natural control by predators and parasites. However, these natural control agents can be inadvertently disrupted by chemical applications that target the bonafide pest species. This situation can result in the emergence of secondary pests, which have been released from natural control. For example, it is widely believed that spider mites have emerged as serious agricultural and forest pests primarily because their predators have been reduced in abundance by chemical sprays for primary pests.

Insect monitoring, which incorporates the use of bait or pheromone traps, is an approach that has become widely used (see Flint and Klonsky, 1989) for monitoring various pest species, providing information on the mobile adult stage. When used in conjunction with phenomenological models, monitoring can be used to predict pest development and, ultimately, to accurately time pesticide applications. Recent advances in technologies for monitoring temperature and leaf wetness have led to commercial implementation of risk assessment models for several key diseases, including late blight of potato (Phytophthora infestans) (e.g., Krause and Massie, 1975; Stevenson, 1983), and grape powdery mildew (Uncinula necator) (Gubler, 1991; SaIl, 1980). Model predictions are usually made by first predicting when conditions are met that are favorable to disease development, and then assessing the severity using a disease risk index. Commercial validation of risk assessment models has shown potential for reducing the number of applications, depending on year, geography, and disease pressure (e.g., Weber et a!., 1996).

9.3.2 FIELD SCOUTING

9.3.3 REDUCED-RISK PESTICIDES

9.3.2.1 Monitoring

When chemical tactics are deemed necessary in an IPM system, the choice of a selective pesticide that kills only the target species is desirable because it is the least disruptive to the crop ecosystem. Examples include some acaricides that are applied for spider mite control that do not affect beneficial predatory mites and microbial controls such as Bacillus thuringiensis, which target particular types of insects. Microbial antagonists have become available for certain pathogens, and selective herbicides are also available. Postemergence herbicides provide the opportunity for growers and consultants to use herbicides in an IPM system by first scouting for weeds to determine their composition and relative abundance before deciding on the control tactic to be employed.

Field scouting or monitoring includes proper identification of pests through surveys or scouting programs, and may incorporate trapping, weather monitoring, and soil testing where appropriate. It may be supported through the use of phenology or risk assessment models, or other types of decision support. In practice, monitoring can be done by either the grower or consultants who check the fields for growers, but there is a labor cost associated with monitoring that is not associated with the preventative use of pesticides. The challenge for researchers is to develop commercial monitoring plans that are economically implementable as opposed to sampling regimes developed for research purposes. Lack of practical monitoring procedures and use of those procedures results in poor timing of applications and an excessive use of pesticides. In many instances (e.g., National Research Council, 1989), pesticide use for controlling a given pest has been reduced 40% without affecting quality or yield simply by using quantitative monitoring procedures in combination with realistic control action thresholds. 9.3.2.2 Decision Support One focus of IPM research for many years has been the development of models that present a framework for integrating information from the various biological disciplines, meteorology, and the field monitoring of pest populations. These models have served to bring disciplines together in analyses of production systems and have yielded tools that can be implemented to support the monitoring or scouting process. IPM research has pioneered many applications for computer technology in agriculture and helped to bring about the early use of electronic instruments for field data gathering (Zalom and Strand, 1990).

9.3.3.1 Behavioral Chemicals Pheromones are highly specific chemicals released by insects to affect the behavior of members of their own species, usually as attractants for mating, but also as signals for aggregation, alarm, or feeding. Synthetically produced pheromones are frequently used in IPM programs as described earlier to monitor adult insect flights. The direct use of pheromones as control agents has also met with some success, usually when the chemical is released over the field from dispensers with the intent of confusing males and preventing mating by inhibiting their ability to locate females. This technique has been applied for control of such key pests as the oriental fruit moth [Grapholitha molesta (Busck)] in Australia and California (e.g., Rice and Kirsch, 1990), the tomato pinworm (Keiferia lycopersicella) in Mexico and the United States (e.g., Jimenez et a!., 1988), the codling moth (Cydia pomonella) (Brunner, 1994), and the pink bollworm (Pectinophora gossypiella) (Flint et a!., 1993). Recently, areawide programs have been established by the USDA

9.3 The rPM Continuum to implement mating disruption for certain key pests on extensive crop acreage in the United States. 9.3.3.2 Conventional Products and Risk Reduced-risk pesticides, as opposed to those that are selective, are usually considered to be safer than traditional pesticides in terms of toxicity to humans and the environment. As regulatory pressures increase the potential for eliminating older classes of pesticides, those pesticides that appear to have a reduced-risk profile are being favored by regulatory agencies as replacement products. Like selective pesticides, those pesticides that have less potential for harming humans or the environment are favored in IPM systems over those that are known to possess such characteristics. One type of risk occasionally ignored when promoting risk reduction in environmental and health terms is the financial risk associated with using less effective controls. Risk is probably the most important financial obstacle to IPM adoption. Growers value pesticides for reducing production risk as well as contributing to profit. For more biologically intensive IPM systems to be voluntarily adopted, it is very important that IPM can be shown to decrease this risk (see Antle and Park, 1986; Gruys, 1982; Way, 1977). In reality, IPM strategies such as monitoring can be tools for managing risk. The more growers learn about pests and their damage potential under an IPM scenario, the less is the uncertainty in their minds about the state of their crop and the more likely it becomes that they will not choose to make a preventative pesticide application. 9.3.4 CULTURAL AND PHYSICAL SUPPRESSION Cultural controls have been used historically to manage many pests, but were often abandoned in favor of pesticides that were less labor intensive. Such controls include a broad range of production practices that render the crop environment less favorable for the pest. Tillage and water management are effective cultural controls in the management of weeds. Furthermore, increased mortality in many insects that overwinter in the soil may result from particular tillage practices. Narrow row plant spacing or optimal in-row spacing can also suppress weeds under certain cropping systems. The destruction of crop residues is important in the management of many pests, such as navel orangeworm in almond, late blight of potato, stem rot of rice, and pink boIl worm and boIl weevil in cotton, for which there are compulsory plow-down dates in several regions. Physical suppression tactics may include cultivation or mowing for weed control, and temperature management or controlled atmospheres for postharvest pests. 9.3.5 PREVENTION Pests are managed in an IPM system in part by preventing their occurrence. Prevention includes those practices that keep

279

pests from invading a crop or field and then becoming established. It includes such tactics as using pest-free seeds or transplants, excluding pests by screens or barriers, preventing weeds from reproducing by disking or mowing, and choosing plant cultivars with genetic resistance to insects, nematodes, or diseases, as well as benefits that result from transgenic gene insertion, irrigation scheduling to avoid situations conducive to disease development, cleaning tillage and harvesting equipment when moving between fields, using sanitation procedures to remove an incipient infestation, and eliminating alternate hosts or sites of pest organisms. Even applying fertilizer with the seed of annual crops or through drip irrigation systems can provide a measure of weed control, especially in contrast to broadcast application of fertilizers, which stimulates weed growth. 9.3.6 AVOIDANCE Avoidance is practiced when pest populations exist in a field or site, but the impact of the pest on the crop can be avoided through some cultural method. Examples of avoidance tactics include crop rotation to break the life cycles of pest species, using trap crops, choosing plant cultivars with maturity dates that may allow harvest before pest populations develop or that have a sufficiently short season to permit planting after conditions are conducive to infestation, fertilization programs to promote rapid crop development, and simply not planting certain fields or areas within fields where damaging pest populations are most likely to develop. 9.3.7 PESTICIDES AND BIOLOGICAL CONTROLS Biological control is the augmentation, conservation, and importation of natural enemies, including predators, parasites, and pathogens, to reduce a pest popUlation. This may involve either introduction of a natural enemy or augmentation of one that already exists in the crop ecosystem. Biological control is generally considered to be the cornerstone of any IPM program. Mass culture and release of predatory lacewings, various species of parasitic wasps, and insect pathogens such as Bacillus thuringiensis have been effective in certain insect pest management programs. The release of certain species of plantfeeding insects and pathogenic fungi has been successful in controlling some range land weed pests as well. Biological control has the advantage of generally being safe to nontarget organisms, although there is concern that biological control agents be specific so as not to disrupt native systems. Classical biological control, the release of an imported natural enemy to control a pest species, when successfully established remains more stable in the environment than other pest control tactics. Natural enemies used in augmentative releases generally do not persist and must be rereleased periodically. Conserving natural enemies by avoiding disruptive sprays has become an essential practice in IPM cropping systems.

280

CHAPTER 9 Practices in Integrated Pest Management

Prior to the 1940s, spider mites were considered to be sporadic pests in most perennial crops. Following the introduction and widespread use of broad spectrum pesticides, spider mites became annual pests in many crops. One of the best examples of conserving natural enemies through the careful use of pesticides involves apple production in the Pacific Northwest, where the spider mite Tetranychus mcdanieli (McGregor) can be a primary arthropod pest. Beginning in the mid 1960s, an effective mite management approach based upon the conservation of the western orchard predator mite, Galendromus occidentalis (Nesbitt), using selective insecticides for control of orchard pests was developed and implemented by Hoyt (1969). In this system, organophosphate cover sprays could be used for the codling moth (c. pomonella), the key pest of apples if an alternate prey, the apple rust mite [Aculus schlechtendali (Nalepa)], was encouraged to support populations of G. occidentalis. Implementation of this program reduced the average mite control cost for Washington state growers from $24 per hectare in 1967 to $8-12 per hectare in 1985 (Croft, 1990). As organophosphates and other older classes of insecticides are being replaced by newer materials, there exists a danger of disruption of IPM systems that have successfully integrated the use of chemical and biological control tactics. The increased use of pyrethroids in many cropping systems presents such a danger because they have been shown to be highly disruptive in orchards by killing predator mites (e.g., AliNiazee, 1984; Croft and Hoyt, 1978). Residues of pyrethroids have been shown to be persistent and remain biologically active against predatory mites long past their initial application (Zalom et al., 1998), and there is some indication that the residual effects of pyrethroids persist on orchard trees into the subsequent growing season (Bentley et aI., 1987). The release of natural enemy strains selected for resistance to disruptive insecticides allows the selected natural enemies to persist even when the disruptive materials are applied for control of key pests. A laboratory selected strain of the predator mite G. occidentalis that is resistant to carbaryl and organophosphates has been successfully used to manage spider mites in California almond orchards (Hoy et aI., 1984). Extensive research was conducted on economical mass-rearing (Hoy et aI., 1982), sampling (Wilson et aI., 1984; Zalom et aI., 1984), and applications of selective acaricides at lower than label rates to help adjust the ratio of spider mites to predator mites in favor of the predator mites, thus enabling the integration of this approach with other almond orchard practices.

9.4 PROGRESS ALONG THE CONTINUUM A USDA Economics Research Service study (Vandeman et aI., 1994) estimated that as many as 52% of U.S. vegetable growers were using some level of IPM to manage insect pests; fewer used any IPM practices for plant diseases (~42%) or weeds (~34%). The same study found that 50% of fruit and

nut acreage was under some level of IPM. By definition in this study, a minimum level of IPM use was making pest management decisions based upon sampling and thresholds. Higher level IPM required the use of multiple nonchemical control practices. A subsequent analysis by Benbrook et al. (1996), who used similar definitions for IPM use, produced similar estimates, but placed the use of more biologically intensive IPM approaches at only about 6% ofU.S. crop acreage. The National Agricultural Statistics Service (NASS) completed a national survey of producers of major crops and crop groupings to determine the percentage of crop acreage using various approaches to pest prevention, avoidance, monitoring, and suppression. Although a variety of practices were used to some extent, certain practices were more common. In the NASS survey (USDA NASS, 1998), alternating pesticides to delay resistance (a chemically intensive IPM tactic) reportedly was used on barley (41%), corn (43%), cotton (50%), soybeans (40%), wheat (30%), fruits and nuts (68%), and vegetables (68%). Scouting for pests was used on cotton (75%), corn (47%), soybeans (45%), wheat (30%), alfalfa (24%), fruits and nuts (80%), and vegetables (81%). Tillage or plowing (a cultural or physical IPM tactic) to control pests was used on barley (30%), corn (40%), cotton (63%), soybeans (41%), wheat (37%), alfalfa (21%), and fruits and nuts (74%). Crop rotation (a more strategic IPM approach) was a leading pest management practice to control pests of barley (59%), corn (69%), soybeans (69%), wheat (53%), alfalfa (32%), and vegetables (74%). All values in parentheses represent percentage of total acreage. Although many or even a majority of growers have adopted a minimum level of IPM, why has there not been more progress toward less chemically intensive IPM systems? A significant body of literature (e.g., Wearing, 1988; Zalom, 1993) discusses the technical, financial, educational, institutional, and social constraints to IPM use. The National IPM Forum held in Washington, DC, in 1992 (see Sorenson, 1994) asked participants representing a variety of public and private interests to identify and rate the constraints. Among the top issues identified were the lack of a national commitment to IPM, lack of funding for IPM research and extension activities, perceived problems with the regulatory process that affects registration of new technologies, and the shortage of well-trained, independent IPM consultants. Since the Forum, the national commitment to IPM increased with the Clinton Administration's 1993 pledge to have 75% of cropland acreage under IPM by the year 2000. However, increased funding for IPM research and extension activities has only marginally materialized through narrowly targeted partnership programs established by federal and state agencies. The U.S. Environmental Protection Agency has made significant changes in the way it approaches registrations of new technologies by establishing a fast track for certain compounds, notably insect pheromones. Little new attention has been paid to the issue of training, certification and promoting the use of IPM consultants.

9.6 Conclusion

9.5 ADVISORY SERVICES The development of a pest management consulting industry can be viewed as one of the most positive results of IPM implementation efforts, and consultants have become a major force in the delivery of pest management information to growers (Blair, 1986; Frisbie and McWhorter, 1986). As pest management becomes more complex and business requirements of the farming enterprise compete for their time, growers increasingly rely on the advice of third parties in the pest management decision process. Crop consultants play a significant role in implementing IPM systems, particularly those that are more biologically intensive and, therefore, require a greater degree of information about the status of pests and their natural enemies in the context of the cropping system. References to crop consultants providing pest management information to their grower clientele can be found in the scientific literature over the past 50 years (see Michelbacher, 1945). In the early 1970s, a program for licensing pest control advisers was initiated by the state of California, and over 4000 individuals are now so licensed. The law requires anyone who recommends pesticides or any other pest control method or device for agricultural use to be licensed. This law has had a far-reaching effect on increasing the number of growers who use a minimum level of IPM; for most crops, a higher proportion of acres are scouted in California than elsewhere in the United States. In spite of this, the potential effect of this program perhaps has not been realized because the licensing program does not distinguish private consultants from the majority of consultant who work for farm supply dealers or other chemical retailers (Wearing, 1988). Although most individuals (none in California) no longer receive commissions or bonuses based on their sales of farm chemicals, an incentive to consider alternative practices, including taking no action, is often lacking. In fact, there is a certain margin of safety in deciding to use a product if there is any question about a pest's damage potential. In human medicine, a regulatory system has been adopted whereby drugs, which (like pesticides) are chemicals intended to kill or neutralize the impact of target organisms or deleterious biological processes, are registered to insure their safe use. Some drugs are recognized as safer to apply than others, and these are available "over the counter" to consumers for selftreatment. Other drugs, the application of which have a greater risk of unintended consequences, are available only by prescriptive use as recommended by physicians who have met specific requirements of the profession. At present, most growers practice a form of self-treatment with pesticides, controlling the choice of chemicals and treatment schedules. As long as pesticides are used according to label restrictions, there are few additional restrictions on their availability or use. With few exceptions, there is no requirement that treatments be based on an accurate diagnosis of a problem or whether, in fact, a problem exists. There is no requirement that alternative treatments be considered or that knowledge of alternative treatments exist. This undoubtedly contributes to the public's negative attitude toward agricultural chemicals

281

and continuing demands for new regulations. Increasingly more stringent regulations will likely lead to the loss of certain higher risk pesticides or uses of pesticides for which alternatives are less effective or more costly to growers. Prescriptive use of drugs has not eliminated cases of injury due to their application, the development of resistance to drugs, or other ancillary problems. However, there is public confidence in the regulatory system for medicine and drugs that does not exist in the regulatory system for pesticides. Would the prescriptive use of pesticides by licensed practitioners help to improve public confidence in the use of pesticides? Coble et al. (1998) addressed this issue by proposing a model similar to that used in the medical profession whereby relatively low-risk chemicals may be self-prescribed, but highrisk chemicals may be prescribed only by a trained and licensed professional. This proposal is one mechanism by which certain valuable pesticide uses could be maintained, while addressing the public's concern for safe use of those products. Already, pesticides are not treated equally in the registration process. For example, pesticides that present the greatest risk to human health or the environment have various restrictions placed on their use. Pesticides that are believed to be "safe" may be put on a fast track for registration.

9.6 CONCLUSION There are many challenges to the development and implementation of IPM systems, but an excellent framework exists in the scientific literature and in experiences with successful field implementation. The concept of integration began with the realization that the use of synthetic pesticides, which helped to make pest control more predictable and less labor intensive, brought about certain unintended consequences such as pest resistance, secondary pest outbreaks, and the resurgence of pests that previously had been under good control. Integrated control suggested that by utilizing pesticides in such a manner as to preserve naturally occurring biological control, more effective and, in the longer term, more economical pest control could be achieved. Integrated pest management incorporated the concept that pesticides should be used only when needed based upon careful assessment of the risk posed by specific pest densities and the potential for control of those pests by naturally occurring beneficial organisms or other factors in the environment. IPM became more interdisciplinary, incorporating an ecosystem approach. As concern about the impact of pesticides on the environment and on human health became elevated in society, IPM gained favor as an acceptable strategy for managing pests. With wide acceptance of the paradigm came a particular emphasis on IPM tactics within the range of practices that can be utilized in an IPM system. In fact, many practices can be and are utilized in an IPM system to prevent, avoid, and suppress the range of pests that threaten crops, human and animal health, or other elements of

282

CHAPTER 9


the landscape. Some observers refer to the range of IPM practices as falling along a continuum from those that are more chemically intensive to those that are more biologically intensive. As older chemistries are replaced with newer classes of pesticides and materials that are biologically derived, users of pesticides can expect to enjoy more possibilities for using products that are selective and are presumed to present less risk to human health and the environment, while providing the opportunity for development and implementation of economically sound IPM systems.

REFERENCES AliNiazee, M. T. (1984). Effect of two synthetic pyrethroids on the predatory mite, Typhlodromus arboreus, in the apple orchards of western Oregon. In "Acarology VI" (D. A Griffiths and C. E. Bowman, eds.), pp. 655-658. Wiley Interscience, New York. Antle, J. M., and Park, S. K (1986). The economic ofIPM in processing tomatoes. Calif. Agric. 40(3/4), 31-32. Benbrook, C. M., Groth, E., Halloran, J. M., Hansen, M. K., and Marqnartdt, S. (1996). "Pest Management at the Crossroads." Consumers Union, Yonkers, NY. Bentley, W J., Zalom, F. G., Barnett, W W, and Sanderson, J. P. (1987). Population densities of Tetranychus spp. (Acari: Tetranychidae) after treatment with insecticides for Amyelois transitella (Lepidoptera: Pyralidae). J. Econ. Entomol. 80, 193-200. Blair, B. D., and Parochetti, J. V. (1982). Extension implementation of pest management systems. Weed Sci. 30, 48-53. Brunner, J. F. (1994). Integrated pest management in tree fruit crops. Food Rev. Internat. 10, 135-157. Carson, R (1962). "Silent Spring." Houghton Mifflin, Boston, MA. Cate, J., and Hinkle, M. (1993). "Integrated Pest Management: The Path of a Paradigm." National Audubon Society, Washington, DC. Cob le, H. D., Bonanno, A R, McGaughey, B., Purvis, G. A., and Zalom, F. G. (1998). "Feasibility of Prescription Pesticide Use in the United States." Issue Paper 9, Council Agric. Sci. Tech. Council on Environmental Quality (1972). "Integrated Pest Management." Council on Environmental Quality, Washington, DC. Croft, B. A. (1990). "Arthropod Biological Control Agents and Pesticides." Wiley, New York. Croft, B. A., and Hoyt, S. C. (I978). Considerations for the use of pyrethroid insecticides for deciduous fruit pest control in the U .S.A Environ. Entomol. 7,627-630. Flint, H. M., Yamamoto, A K, Parks, N. J., and Nyomura, K. (1993). Aerial concentrations of gossyplure, the sex pheromone of the pink bollworm (Lepidoptera: Gelechiidae) within and above cotton fields treated with longlasting dispensers. Environ. Entomol. 22, 43-48. Flint, M. L., and KIonsky, K. (1989). IPM information delivery to pest control advisors. Calif. Agric. 43(1), 18-20. Flint, M. L., and van den Bosch, R. (1981). "Introduction to Integrated Pest Management." Plenum, New York. Fl1sbie, R. E., and McWhorter, G. M. (1986). Implementing a statewide pest management program for Texas, USA. In "Advisory Work in Crop Pest and Disease Management" (J. Palti, and R Ausher, eds.), pp. 234-262. Springer-Verlag, Berlin. Georghiou, G. P. (1986). The magnitude of the resistance problem. In "Pesticide Resistance: Strategies and Tactics for Management." National Academy Press, Washington, DC. Gruys, P. (1982). Hits and misses. The ecological approach to pest control in orchards. Entomol. Exp. Appl. 31, 70-87. Gubler, W. D. (1991). Powdery mildew: Epidemiology and Control. In "Proceedings Nelson J. Shaulis Viticultural Symposium," pp. 44-47. New York State Agric. Exp. Station, Geneva, NY.

Hoppin, P., Liroff, R A, and Miller, M. M. (1996). "Reducing Reliance on Pesticides in Great Lakes Basin Agriculture." International Policy Program, World Wildlife Fund, Washington, DC. Hoy, M. A., Barnett, W W, Hendricks, L. c., Castro, D., Cahn, D., and Bentley, W. J. (1984). Managing spider mites in almonds with pesticide-resistant predators. Calif. Agric. 38(7/8), 18-20. Hoy, M. A., Barnett, W W, Reil, WO., Castro, D., Cahn, D., Hendricks, L. c., Coviello, R, and Bentley, W. J. (1982). Large scale releases of pesticideresistant spider mite predators. Calif. Agric. 36(1/2), 8-10. Hoyt, S. C. (1969). Integrated chemical control of insects and biological control of mites on apples in Washington. J. Econ. Entomol. 62, 74-86. Jimenez, M. J., Toscano, N. C., Flaherty, D. L., Ilic, P., Zalom, F. G., and Kido, K (1988). Controlling tomato pinworm by mating disruption. Calif. Agric. 42(11/12), 10-12. Kogan, M. (1988). Integrated pest management theory and practice. Ann. Rev. Entomol. 49, 559-570. Kogan, M. (1998). Integrated pest management: Historical perspectives and contemporary developments. Ann. Rev. Entomol. 43, 243-270. Krause, R A., and Massie, L. B. (1975). Predictive systems: Modem approach to disease control. Ann. Rev. Phytopath. 13,31-47. Matteson, P. C. (1995). The "50% pesticide cuts" in Europe: A glimpse of our future? Amer. Entomol. 41(4), 210-220. Metcalf, R. L., and Luckmann, W. H. (1982). "Introduction to Insect Pest Management." Wiley, New York. Michelbacher, A. E. (1945). The importance of ecology in insect control. J. Econ. Entomol. 38,129-130. Michelbacher, A. E., and Bacon, O. G. (1952). Walnut insect control in northern California. J. Econ. Entomol. 45, 1020-1027. National Research Council (1989). "Alternative Agriculture." National Academy Press, Washington, DC. Newsom, L. D. (1980). The next rung up the integrated pest management ladder. Bull Entomol. Soc. Amer. 26, 369-374. Prokopy, R. J. (1994). Integration in orchard pest and habitat management: A review. Agric. Ecosyst. Environ. 50, 1-10. Rice, R E., and Kirsch, P. A (1990). Mating disruption of oriental fruit moth in the United States. In "Behavior-Modifying Chemicals for Insect Management" (R L. Ridgway, R M. Silverstein, and M. N. Inscoe, eds.), pp. 193-211. Dekker, New York. SaIl, M. A. (1980). Epidemiology of grape powdery mildew: A model. Phytopath. 70, 338-342. Smith, R F., and AlIen, W W (1954). Insect control and the balance of nature. Sc;' Amer. 190(6), 38-92. Sorenson, A. A. (1993). "Regional Producer Workshops: Constraints to the Adoption of Integrated Pest Management." National Foundation for Integrated Pest Management Education, Austin, TX. Sorenson, A. A (1994). "Proceedings of the National Integrated Pest Management Forum." Center for Agriculture in the Environment, American Farmland Trust, De Kalb, IL. Stem, v., Smith, R. F., van den Bosch, R. F., and Hagen, K. S. (1959). The integrated control concept. Hilgardia 29, 81-97. Stevenson, W. R (1983). An integrated program for managing potato late blight. Plant Dis. 67, 1047-1048. United Nations Food and Agriculture Organization (UNFAO) (1967). "Report of FAO Panel of Experts on Integrated Pest Control." United Nations Food and Agriculture Organization, New York. USDA National Agricultural Statistics Service (USDA NASS) (1998). "1997 Pest Management Practices." Special Circular 1(98), U.S. Dept. Agric., Washington, DC. van den Bosch, R (1964). Practical application of the integrated control concept in California. In "Proc. Intern. Congo Entomol." Vol. 12, pp. 595-597. van den Bosch, R, and Stem, V. M. (1962). The integration of chemical and biological control of arthropod pests. Ann. Rev. Entomol. 7, 367-386. Vandeman, A, Fernandez-Cornejo, J., Jans, S., and Lin, B. H. (1994). "Adoption of Integrated Pest Management in U.S. Agriculture." Agricultural Information Bulletin 707, United States Department of Agriculture, Economic Research Service, Washington, DC.

References

Way, M. J. (1977). Integrated control-Practical realities. Outlook Agric. 9, 127-135. Wearing, C. H. (1988). Evaluating the IPM implementation process. Ann. Rev. Entomol. 33, 17-38. Weber, E., Gubler, W. D., and Derr, A. (1996). Powdery mildew controlled with fewer fungicide applications. Winegrowing Jan/Feb, 13-16. Wigglesworth, V. B. (1950). The science and practice of entomology. Adv. Sci. 7, 154-161. Wilson, L. T., Hoy, M. A., Zalom, F. G., and Smilanick, J. M. (1984). The within-tree distribution and clumping pattern of mites in almond orchards: Comments on predator-prey interactions. Hilgardia 52(7), 1-13. Zalom, F. G. (1993). Reorganizing to facilitate the development and use of integrated pest management. Agric. Ecosystems Environ. 46, 245-256.

283

Zalom, F. G., Walsh, D., Stimmann, M. W, PickeI, c., Krueger, W, Buchner, K, and Brazzle, J. (1998). Impact of pyrethroids on beneficial mite predators. In "Proceedings of the California Plant and Soil Conference," pp. 62-67. Agronomy Society of America, California Chapter, Sacramento, CA. Zalom, F. G., Hoy, M. A., Wilson, L. T., and Bamett, W W (1984). Presenceabsence sequential sampling for web-spinning mites in almonds. Hilgardia 52(7), 14-24. Zalom, F. G., and Strand, J. F. (1990). Expectations for computer decision aids in rPM. AI Appl. Nat. Res. Manag. 4(1), 53-58.

CHAPTER

10 Toxicity Testing Donald J. Ecobichon Queen's University

10.1 INTRODUCTION Although governmental concerns about the safety of chemical pesticides were evident in the early 1900s in many agricultural nations, these concerns did not prompt legislation and regulations until the late 1940s when organochlorine insecticides were introduced into agricultural use. Important changes in pesticide legislation, particularly concerning regulations about testing, were promulgated in the late 1950s and early 1960s when the much more toxic organophosphorus and carbamate ester insecticides appeared and concomitant increases in the number of poisonings were reported among agricultural workers, bystanders, and produce consumers. An examination of the development of pesticide testing requirements in different nations reveals a parallel with the evolution of tests required for drugs, industrial chemicals, and home and health care products. Most industrialized countries have developed legislation/regulations for pesticide testing, registration, and use. Now companies that wish to register a pesticide product are required to submit a toxicity data base that is comparable to that required by the U.S. Environmental Protection Agency (EPA) under the Federal Insecticide, Fungicide and Rodenticide Act (FIFRA) as is shown in Table 10.1. In addition to these specific test procedures, a considerable amount of environmental testing is required-a topic beyond the scope of this chapter and this book. Although distinct national requirements have evolved over year of test development, recent emphasis has been on international harmonization of testing requirements for pesticides, thereby permitting a chemical registered in one country to be registered in others, given that the toxicity data base meets common standards. Such a change would simplify the regulatory procedures for industries, which currently must meet variable, individual, and national regulatory requirements, a:1d thereby reduce the efforts, cost, and time for multinational registrations. If the flow of toxicological information is coordinated internationally, it would be extremely advantageous to emerging nations, some of whom have shortages of trained scientific, technical, and legal professionals to develop their own legislation. The availability of this information would allow developing countries to adopt a common regulatory framework for pesticides. Such harmonization is being encouraged by the European Economic Handbook of Pesticide Toxicology

Volume I. Principles

Community (EEC), the Organization for Economic and Cooperative Development (OECD), the World Health Organization (WHO) via the International Programme on Chemical Safety (IPCS), and the United Nations Food and Agricultural Organization (FAO).

10.2 TESTING STRATEGIES As defined by Hayes (1975), toxicology is "the qualitative and especially the quantitative study of the injurious effects of chemical and physical agents, as observed in alterations in structure and response in living systems; it includes the application of the findings of these studies to the evaluation of safety and to the prevention of injury to humans and all useful forms of life." The aim of any of the studies listed in Table 10.1 is to develop a quantitative dose-response (effect) relationship between level of exposure and measured biological effects, hopefully, thereby, permitting extrapolation of the study data from animal surrogate models to humans. Within reason, extrapolations and predictions are possible for all classes of insecticides, as well as some rodenticides and fungicides, because target sites and mechanisms of action in most, if not all, mammalian systems and species are comparable. However, for many other classes of pesticides, the toxicity encountered has less to do with the active agent and more involvement with low levels of contaminants and by-products that are capable of causing unique, adverse, biological effects. The extent of the testing required for any new pesticide must address any and all possible toxic events, both predictable and idiosyncratic, as well as the concerns of product safety raised by workers, bystander populations, and the general public-a tall order indeed! As is found with any chemical, all pesticides possess some degree of inherent toxicity-a property of the agent that is as distinctive as the physical and chemical properties of the molecule. This is a basic tenet of toxicology. Thus, anyone member of a class of chemicals should behave qualitatively in the same manner as any other homolog, although the concentration required to elicit a given biological effect may vary considerably. A second basic tenet is exposure, that is, the

285

Copyright © 2001 by Academic Press All rights of reproduction in any form reserved.

286

CHAPTER 10 Toxicity Testing

Table 10.1 Federal Toxicity Testing Requirements for a Pesticide

Table 10.2 Toxicity Data Base Development for Chemicals-Progression of Studies

Acute orallethality

First wave

Acute dermallethality Acute inhalation lethality

Acute (lethality) studies, via routes of exposure anticipated for the human

Primary dermal irritation

Irritation studies (ocular, dermal)

Dermal sensitization

Dermal sensitization studies

Primary ocular irritation

Mutagenicity studies with in vitro microbial and

Acute-delayed neurotoxicity

mammalian cell lines

21-Day dermal exposure

Teratogenicity studies (mouse, rat, rabbit) with agent

90-Day dermal exposure

administered to normal, pregnant animals during

90-Day feeding study 90-Day inhalation exposure 90-Day neurotoxicity assessment Chronic feeding study Oncogenicity study Teratogenicity study Reproduction study Gene mutation Chromosomal aberration studies

organogenesis Second wave Subchronic studies, 21-90-day feeding studies in rodent and nonrodent species; may use other routes of exposure Chronic/oncogenicity studies of 6-month (rodent) duration for the former and 24-month (or life-span) for the latter in rodents Reproductive studies, generally in rodents: effects are studied in both males and females

Other genotoxic effects General metabolism

Source: Ecobichon (l997a-d).

Domestic animal safety Source: V.S. Code of Federal Regulations, Title 40, Part 158 (40CFR Part 158).

amount of agent necessary to elicit an adverse biological effect. Exposure consists of two components: the concentration of the agent and the duration (time) of exposure, which reflects the possibility of both acute, high-level, short duration as well as low-level, long duration exposures. The third basic tenet of toxicology arises from the other two, that is, the previously mentioned dose-response relationship, where the degree of adverse response is proportional to the level of exposure. This evaluation entails a knowledge of the levels at which adverse biological effects begin to appear. A threshold toxicant is presumed to pose no risk below some experimentally determined concentration. In contrast, a nonthreshold toxicant is presumed to pose some risk at all dosages above zero. The exorbitant costs of conducting extensive toxicological assessments of pesticides makes it imperative that any potential toxicity be discovered as early as possible in the development program. Toxicity studies are carried out in two waves (Table 10.2). The first wave consists of studies that can be completed in a relatively short time period (approximately two months) and at minimal cost, giving good indications of what hazards might be encountered following acute exposure and providing clear options whether to abandon the agent or proceed. The second-wave package consists of the more costly, time-consuming, longer-term studies required to examine chronic effects, including fertility, reproductive outcomes, and carcinogenicity, on various target organs. The results of these studies will be in hand only after some 3-4 years.

10.2.1 FIRST-WAVE STUDIES

Given the distinct possibility that, at some time in the production and use of the pesticide, some individual(s) (factory workers, handlers, mixer-loaders, sprayers, bystanders, general public) will be exposed accidentally or by intent to toxic levels of the pesticide, studies that involve exposure to high concentrations of the agent are essential to determine its potency relative to other, known agents tested in the same species. The potency will be assessed in three ways: (1) acute lethality of the agent along with other selected endpoints of acute toxicity; (2) irritancy to both skin and eyes; (3) dermal sensitization. This information is important for the occupational health and safety of workers and will be incorporated into the material safety data sheet (MSDS), which, by law, must be prepared and made available for every chemical manufactured and used. In addition, the potential of a chemical to cause mutations in the nucleic acid of microbial and mammalian cell systems will be examined at this time, because an extension of this biological activity is the occurrence of adverse reproductive outcomes (including birth defects) and cancer. Teratogenicity studies are generally conducted at this stage, providing possible confirmation of positive mutagenicity results. 10.2.1.1 Acute Lethality

The potency of a chemical is frequently expressed in terms of the median lethal dose (LD50). This index is defined as a "statistical estimate of the acute lethality of an agent administered under defined and controlled conditions to a certain sex, age and strain of species of animal." The route of administration (oral, dermal, inhalation) is selectively based on that route by

10.2 Testing Strategies which humans are expected to be exposed. Standardized protocols are plentiful for designing and conducting this test (Chan and Hayes, 1994; Ecobichon, 1997a). Unless it is indicated, the LDso is assumed to represent the median lethal dose for deaths that occur in the first 24 hours after treatment. In the case of inhalation and dermal studies, the LDso is actually an LCso, that is, the concentration in the breathing air or applied to the skin rather than the known amount of agent taken up by the body via the lungs or through the skin. The value of the LDso rests in its use as an index of relative toxicity of the unknown chemical compared to the toxicity of known chemicals administered by the same route to the same species, strain, age, and sex of test animals. Within a certain class of chemicals that have a similar mechanism(s) of action, direct comparisons between various members can be made concerning relative potency. Such agent-to-agent comparisons should not be done for chemicals of diverse structure and mechanisms of action, that is, insecticides, herbicides, fungicides, avicides, molluscicides, etc. Although such comparisons are commonly made, it is far more complex than comparing apples, oranges, and bananas, and is a serious limitation of the LDso value obtained experimentally. However, it is important to emphasize that acute toxicity studies are not limited to lethality alone. Classic oral LDso determinations are conducted in an appropriate surrogate animal model (mice, rats, etc.) to ascertain the total adverse biological effects during a finite period of time following the administration of the test agent, by an appropriate route, as single, frequently large, doses to various groups (n = 10 per treatment) of preselected animals. Lethality, expressed as the LDso (mglkg body weight), is one endpoint determined at 24 hours posttreatment. The appearance and behavior of the animals is closely observed during the 24-hour period to detect possible mechanisms of action of the test agent. All dead animals and those moribund animals that are euthanized to prevent suffering will be subjected to gross anatomical examination and tissues are preserved for microscopic study. The animals that survive the toxic insult will be monitored for 14 more days. These surviving animals provide a vast amount of information on behavior, growth and development, recovery, persistence of signs and symptoms, secondary target organ toxicity and delayed toxicity as well as biochemical parameters from blood and urine. On day 14 posttreatment, the surviving animals are euthanized and undergo gross examination. Terminal samples of biological fluids are collected for analysis and tissues are secured for preservation and microscopic examination. Dermal exposure to pesticides is a constant hazard and concern. National and international regulatory agencies require that estimates of lethality via the dermal route be determined for active ingredients as well as for concentrated solutions and end-use formulations (Ecobichon, 1997a; Patrick and Maibach, 1994). In such studies, the test agent is applied to the closely clipped dorsal skin, one dose per group, to several groups of male and female rodents (mice, rats) and nonrodents (guinea pigs, rabbits). The test agent, as a powder, paste, or liquid (neat

287

or dissolved in a vehicle), is spread uniformly over the test site, which should be equivalent to approximately 10% of the total body surface (3.0 x 4.0 cm for rats; 5.0 x 6.0 cm for rabbits; Ecobichon, 1997a). A number of doses are used, and the range is spaced to elicit toxic effects and lethality. The test agent is left in place for 24 hours with the site occluded by a gauze pad taped securely to the skin and the entire region covered by a rubber sheet or a form-fitting elastic cloth sleeve. After 24 hours of exposure, the unabsorbed material is washed off with warm water and soap, avoiding vigorous scrubbing. The acute lethality is expressed as the LCso (median lethal concentration, mglkg). If chemical-related toxicity and/or mortality are not observed on reaching a dose of 2000 mglkg body weight, further study at higher doses is unnecessary. The likelihood of exposure to vapors, aerosols, and/or particulates of pesticide formulations is high among agricultural workers both during and after spraying; thus determination of the acute lethality by the inhalation route is necessary. Groups of experimental animals-rodents and nonrodents, males and females of each-will be exposed to the test substance for a defined period, usually 4 hours, at a graduated range of three concentrations, one concentration per group. The animals may be accommodated within an inhalation chamber (whole body exposure) or in a head-only or nose-only apparatus. The desired concentration of aerosolized active ingredient or complete formulation is generated via a nebulizer system. At the end of the test period, the animals are removed from the test apparatus and are returned to their cages for further monitoring as well as determining the LCso, the concentration in the breathing air of the animal that resulted in acute lethality in 24 hours. The surviving animals are monitored biochemically and physiologically for 14 more days, as previously described. The multitude of problems inherent in acute lethality studies by inhalation have been discussed (Ecobichon, 1997a). Most national and international regulatory agencies no longer demand determination of a precise LDso, but do require an estimation of the acutely lethal dose that can be obtained using a limited number (n = 6 or 8) of animals rather than the 50-60 animals required in the preceding test. What is desired is a rough estimate of the toxicity of the agent so that suitable ranges of doses can be defined, thereby avoiding excess early mortality in future subchronic and chronic studies. The up-anddown or staircase method uses only one animal per dose; a second animal will receive a higher or lower dose, depending on whether the first treated animal dies (Ecobichon, 1997a). Although it is used mostly for oral LDso determination, the same approach would be valid for dermal and inhalation routes of exposure. The fixed dose approach avoids the death of the animals as an endpoint, relying on clear observations of toxicity at one or another of a set of preselected or fixed doses (5.0, 50, 500, and 2000 mglkg), where the results permit classification or ranking of the chemical according to the system used by the EEC as shown in Table 10.3. Oral and dermal LDso values have been utilized by the WHO to prepare a classification of pesticides that distinguishes between the more and the less hazardous forms of each pesticide.

288


Table 10.3 Ranking System Used by the European Economic Community for Acute Oral Toxicity of Chemicals Category

LDSO (mg/kg)

Very toxic

25 25-200 200--2000 2000

Toxic Harmful Unclassified Source: European Commission Directive 83/467IEEC (1983).

Such a distinction pennits formulations to be classified according to percentage of active ingredient and physical state (Table 10.4; Copplestone, 1988). It must be emphasized that the quoted LD50 values are not the median values, but are the lower confidence limit values for the most sensitive sex, thereby ensuring that a large safety factor has been built into the classification. The use of a single value avoids confusion or uncertainties in the case of products that have confidence limits that lie across a class boundary. The WHO is of the opinion that this classification scheme works well in practice, reflecting the idea that acute toxicity is the most effective parameter by which to judge the hazards of pesticides to human health. 10.2.1.2 Dermal Irritation Any chemical that may come into contact with the skin must undergo primary irritation studies using suitable animal models. Pesticides are no exception, because occupational dennatitis is a major problem among pesticide users. Several standard protocols have been described for such studies; most are only slight modifications of the original Draize test (Draize et aI., 1944; Ecobichon, 1997a; Patrick and Maibach, 1994). All tests require closeclipping of the animal's dorsal hair to provide a suitable test site, preparing several areas of intact as well as Table 10.4 World Health Organization Recommended Classification of Pesticides by Hazard LDsO for the rat (mg/kg body weight) Dermal

Oral Class la

Extremely

Solids

Liquids

Solids

Liquids

:sS

:s20

:s10

:s40

5-50 50--500

20--200 200--2000

10--100 100--1000

40--400 400--4000

>500 >2000

>2000 >3000

> 1000

>4000

hazardous lb

Highly hazardous

II

Moderately hazardous

III

Slightly hazardous

III

Unlikely to present hazard in normal use

Source: Copplestone (1988).

abraded epidennis, and applying suitable amounts of the test agent (not exceeding 1.0-2.0 ml/kg neat if liquid or dissolved in a vehicle; not exceeding 2000 mg/kg if a powder or paste) uniformly over the test sites (3 x 4 cm for rats; 5 x 6 cm for rabbits). The site is usually occluded to prevent the animals from removing the test substance during the 24-hour exposure period. After 24 hours, the residual, unabsorbed material is washed off using warm soapy water and the dermal reaction (erythema or redness, edema or puffiness) is assessed immediately and at 48 and 72 hours, using a standardized, subjective scoring system that has numerical values from 0 (no effect) to 4.0 (severe effects; McCreesh and Steinberg, 1983). Thus, for any pesticide, it can be detennined whether the components of a formulation (co-solvents, emulsifiers, adjuvants, etc.) play any role in the dennal penetration of the active ingredient or exert any effect on toxicity. Two scenarios could occur: a formulation component could either enhance or reduce the potency of the test agent. Valid concerns about the infliction of pain and suffering on animals in primary irritation studies has led to substitution of in vitro studies that use established mammalian cell lines and/or epidennal surrogates and examine a number of biological endpoints that measure cell penneability, cytotoxicity, and inflammation (Ecobichon, 1997a). Many of the assay systems listed in Table 10.5 have been validated, and good agreement has been obtained between in vitro results and previously determined in vivo effects. In the future, greater use will be made of test batteries of in vitro assays to screen new chemicals to eliminate highly to moderately toxic agents, leaving those of apparent low toxicity to be evaluated in animals. 10.2.1.3 Dermal Sensitization Repeated dermal exposure to a number of chemicals results, over time, in a hypersentitive, allergy-like reaction similar to that seen for certain foods, pollens, plants, venoms, etc. The initial exposure elicits little or no reaction, but repeated exposure may cause dennal swelling and redness at the site of exposure. Each time exposure occurs, the reaction occurs more quickly and persists for a longer period of time. This mechanism appears to be similar to the classical allergic reaction, where the agent that acts as a hapten binds with proteins to form an antigenic complex that is recognized as foreign to the body and triggers cell-mediated production of IgE antibodies. A number of test procedures have been described, usually using the guinea pig as the surrogate model (Chan and Hayes, 1994; Ecobichon, 1997a; Seabaugh, 1994). All tests are similar in that an attempt is made to develop an immunological "awareness" by daily application of the test agent to the shaved skin of the animal over a period of 24 days, followed by a rest period of 10-14 days. A challenge dose, usually smaller than that used to raise the sensitivity, is administered to a fresh, untreated dermal site, and the severity of the redness and swelling is scored at 24, 48 and 72 hours posttreatment, the same as mentioned for dermal irritation. Depending upon the severity of the immunological reaction, the test agent can be classified as a weak-to-strong sensitizer.

10.2 Testing Strategies Table 10.5 Alternative Test Systems to Assess Dermal Irritation Using Cell Cultures

Table 10.6 Test Methods for Skin Sensitization Acceptable to the EEC

Assessment

Test

Test endpoint

Draize Test

Cytotoxicity

Trypan blue

Dye inclusion in damaged

Neutral red

Dye sequestered in lyso-

cells somal membranes of viable cells Cell viability

Kenacid blue

Total protein content

Coomassie blue

Buehler Test Guinea Pig Maximization Test Mauer Optimization Test Freund's Complete Adjuvant Test (FCAT) Open Epicutaneous Test Split Adjuvant Test Information sources: OECD Test Guideline 406 (1981); Patrick and Maibach (1994); Seabaugh (1994); Ecobichon (1997a).

Lowry protein Rhodamine or

289

Total protein content

Nile blue Hexosaminidase

Cell activity metabolism

Leakage of cytosolic

Lactic acid

enzymes through damaged

dehydrogenase 14C-Ieucine

cell membranes Transport and accumulation

3H-uridine

intracellularly and in-

3H-thymidine

corporation into proteins or nucleic acids

Mitochondrial

Reduction of yellow tetra-

function

zolium salt (MTT) to blue-black, insoluble formazan

Cell metabolism

Microtox™

Luminescent bacterium that emits light during metabolism (reduction of light)

Irritation

Skintex™

Test agent induced alterations of macromolecular matrix of keratin-collagendye membrane, with release

shaved, tape-stripped site for 3 consecutive days; (2) a challenge dose on day 10, applied to one ear and vehicle applied to the other; (3) measurement of the ear thickness by micrometer at 24 and 48 hours posttreatment. The detection of weak haptens can be enhanced by using abraded skin (Botham et aI., 1991). A more promising test, the murine local lymph node assay (LLNA), has shown remarkable responses to sensitizing agents (Kimber and Basketter, 1992). The protocol, using 8-week-old mice, involves daily, topical treatment with agent in solution (25 J.!l) on the dorsum of both ears for 3 consecutive days and, 5 days later, an injection (IV) of 20 J.!Ci of 3H-methylthymidine or 3H-thymidine in saline. After an additional 5 days, the mice are euthanized and the auricular lymph node cells are recovered. After incubating overnight in suspension, the lymph node cells are subjected to ,B-scintillation counting to determine a treated/control ratio. Validation of this test has demonstrated the correct identification of human sensitizers (Basketter et aI., 1994).

of dye Inflammatory response

Cell culture

Measurement of eicosanoids

(chemical

and cytokines from keratino-

mediators)

cytes incubated with test agent

Currently, seven tests, all using the guinea pig, are accepted by the OECD and EEC (Table 10.6). Guinea pigs are used as the test animal because of their recognized susceptibility to a wide variety of chemical sensitizers. All of these tests involve repeated, intermittent, topical or intradermal application of agent, with or without Freund's adjuvant, examination of the site of application for erythema and edema after a 2-3-week "rest" period, and administration of a challenge dose of the test agent. Details of these individual tests were given by Patrick and Maibach (1994) or the references listed in the table may be consulted. The mouse ear-swelling test (MEST) was developed as an accurate, sensitive, alternative test to evaluate sensitization (Gad et aI., 1986). The test includes (1) an induction phase with topical application of the agent in liquid form to an abdominal,

10.2.1.4 Ocular Irritation The inadvertent, accidental splashing of a diluted or concentrated pesticide formulation during handling (owing to the frequent absence or lack of use of appropriate protective equipment) makes it mandatory that the potential of an agent to cause ocular irritation be tested. The Draize eye test has been the standard method used to assess product safety (Draize et aI., 1944). Few modifications have been made to the basic methodology with the exception of the low-volume test, but additional, quantifiable parameters have been introduced, including erythema, thickness of the eyelids and nictitating membranes, edema, discharge, corneal opacity, capillary damage, and pannus (vascularization) of the cornea (Buehler, 1964; Conquest et aI., 1977). The standard ocular irritation test is no longer in favor because of perceived trauma and pain to the eyes of test animals. In reality, if a thoroughly conducted dermal evaluation reveals irritational or corrosive properties for the test agent, there is no need to conduct the eye test. However, given the fact that the corneal membrane and the epidermis are quite different in structure, false negative results with the skin could mask serious ocular problems. A number of in vitro tests that use mammalian cell lines, isolated, intact eyes from chickens, or isolated bovine

290


corneas have been designed and introduced to reduce and replace the use of animals and to refine the ocular irritation test (Ecobichon, 1997a). These tests assess (1) cellular cytotoxicity, which affects the uptake or exclusion of dyes by cell cultures, (2) opacity, that is, light transmission through the isolated bovine or rabbit corneas, and (3) inflammation, which measures the release of inflammatory mediators (histamine, serotonin, leukotrienes, prostaglandins) from the bovine corneal epithelium (Ecobichon, 1997a). Although validation of these tests points to their utility, it is likely that these in vitro assay systems will be used to screen moderate to severe irritants, reserving the animal test for those agents being contemplated for use in formulations. All final products will continue to be tested by the Draize test to assure the consumer that damage (reversible or irreversible) to the eye cannot occur. In the standard ocular irritation test, 0.1 ml of liquid or 100 mg of solid is instilled into the pouched, lower conjunctival sac of one eye of each of six rabbits and the eyelids are held together for 1.0 second and then released. The treated eye is not washed; rather, the animal's own tear secretions are allowed to flush the eye. The untreated eye serves as a control. Generally, the eyes are irrigated with 0.9% physiological saline after 24 hours. Both eyes are examined at 1, 24,48, and 72 hours following treatment. Such studies are usually terminated at 72 hours, but if residual injury is present, examination of the eyes may be prolonged at the discretion of the investigator. Damage to the cornea, conjunctiva, and iris is scored subjectively according to the numerical system described by Draize et al. (1944). To improve the reproducibility of the scoring, slit-lamp examination and staining of the eye with fluorescein dye are frequently employed. A low-volume test (LVET), using 0.01 ml of liquid or 0.01 g of solid has been introduced, wherein the agent is applied directly to the cornea and the eyelids are not held shut following treatment. Using only three rabbits, good correlations were obtained between the LVET results and the human response. The LVET predicted human ocular exposure more accurately than the standard test (Bruner et aI., 1992; Freeberg et aI., 1986). In contrast to the preceding study design, more recent investigations have demonstrated that high-level accuracy can be obtained with fewer rabbits. Two- or three-rabbit tests have been shown to be accurate in classifying the irritational potential of known agents to levels of 88-91 and 93-94%, respectively (De Sousa et aI., 1984; Talsma et aI., 1988).

10.2.1.5 Mutagenicity Many chemical mutagens interact with cellular deoxyribonucleic acid (DNA) by alkylating reactions. The results of interaction are manifested as cellular lethality (severe cases) or as chromosomal breaks, dysfunctions, deletions, fragments, exchanges, point mutations, etc., that are replicated faithfully through descendent cells. It is accepted dogma that the consequences of subtle alterations in chromosomal material may be expressed as infertility, embryolethality, embryotoxicity, spontaneous abortion, congenital anomalies, altered resistance to infections, reduced life-span, and even carcinogenesis. To attempt meaningful "mutagenicity" studies in surrogate animal

models and to measure the above endpoints collectively is possible, but a lot of time may be consumed, the costs will be excessive, and the results will often be clouded by confounding factors. These problems led to the search for relatively simple, short-term, inexpensive, sensitive, and reproducible biological tests that were capable of predicting the mutagenic potential of chemicals. The search has been successful to a degree: a plethora of rapid, sensitive assays currently are available to explore a variety of mutational end points and even to give insight into the mechanism(s) by which these events are produced. The utility of these tests will be to screen the vast array of untested chemicals for mutagenic properties. A variety of microbial systems using prokaryotic, auxotrophic bacteria (Salmonella typhimurium, Escherichia coli), yeasts (Saccharomyces cervisiae), and fungi (Neurospora crassa) have been developed based on extensive knowledge of the genetics of various strains of the organisms. The experiments are conducted rapidly and reproducibly, and the organisms are responsive to the agents being tested (Ecobichon, 1997b). The prototype assay was the Ames test (Ames et aI., 1973, 1975), which used genetically defective strains of S. typhimurium as the indicator organism. These strains depend on provision of an essential nutrient for growth (histidine for S. typhimurium; other organisms require sugars, amino acids, purines, pyrimidines, etc.). When a standard mixture of organisms, top agar, and a small amount of essential nutrient was incubated with a range of concentrations of the suspect mutagen and then grown for 24 hours on agar plates that were deficient in the essential nutrient, a quantitative, dose-related increase in the number of revertant colonies was visible that indicates mutagenic reversion (reverting to prototypic or wild-type organisms capable of synthesizing the essential nutrient). At high concentrations, reverse mutation would be obscured because of excessive lethal mutations that result in cell death. With this simple assay, a quantitative relationship can be developed between the concentration of suspect mutagen and the extent of reverse mutation of the organism, at minimal cost and in a relatively short time (Ames et aI., 1975). Many mutagenic substances exist as promutagens, which are inactive per se and require biotransformation into reactive intermediates before any mutagenic activity is exhibited. Microorganisms are incapable of carrying out such reactions rapidly. Hence modifications in the microbial assay to include an active metabolizing enzyme system are needed. The usual modification involves adding an aliquot of a 9000-g (S-9 fraction) supernatant obtained from homogenized rodent (mouse, rat) liver from a donor animal that was pretreated with a suitable enzyme-inducing agent (polychlorinated biphenyls, phenobarbital, ,B-naphthoflavone, etc.) to provide maximally induced, hepatic, microsomal Phase I enzymes. Before the test mixture is spread on the minimal agar, the S-9 aliquot plus suitable cofactors are incubated with the microorganism and test substance to ensure conversion of the promutagen to a reactive, possibly mutagenic, intermediate that is capable of causing a mutagenic effect in the nucleic acids of the microorganism. Such an "activity system" can be used with any of the prokaryotic organisms

10.2 Testing Strategies

and also with eukaryotic, mammalian cells as described subsequently. The apparent supersensitivity of the Ames test (almost everything was positive) and the prudence of not relying on anyone particular bioassay, resulted in the development of other in vitro assays that use a variety of immortalized, standardized, eukaryotic, mammalian cell lines [Chinese hamster ovarian cells (CHO), hamster pulmonary cells (V79), mouse lymphoma cells, fibroblasts, leukocytes] to examine various endpoints such as clastogenicity (chromosomal aberrations, breaks, fragments, chromatid exchanges, micronucleus formation) and gene mutations (forward or reverse) that result in unscheduled DNA synthesis, the loss or appearance of specific enzymatic functions, etc. Mutagenic assay systems have also included plants (onion root tip, allium, tradescantia) and insects (Drosophila melanogaster), but such tests take somewhat longer to complete and yield results that are more difficult to interpret beyond the observation that a mutation has occurred, and growth and development are affected. The overall assessment of the mutagenic potential of a chemical is based on the spectrum of positive and negative results from five or six assays selected from a test battery such as that shown in Table 10.7. All the tests in Table 10.7 are representative, reproducible, and well-established assays that examine different facets of genetic toxicology and allow exploration of the mechanism(s) of action of the test substance. Such a group of assays can be completed within 30 days at moderate cost.

Table 10.7 Test Battery for Mutagenic Evaluations of Chemicals Type Bacterial Mammalian (in vitro)

Test system/assay

Function

Ames salmonella-liver S9

Reverse mutation

Escherichia coli

Reverse mutation

Mouse lymphoma L51778 (TK + / - )

Forward mutation

Chinese hamster ovary HGPRT

Forward mutation

Pulmonary V79 cells HGPRT

Forward mutation

Mammalian cell lines

Chromosome aberrations Sister chromatid exchange

Mammalian

Drosophila melanogaster

Sex-linked recessive

Rodent bone marrow stem cells

Chromosomal

(in vivo)

lethal assay aberrations Micronucleus formation Rodent (mouse, rat) dominant

Lethal mutations

lethal assay Source: Data from National Research Council, Toxicity Testing, Strategies to Determine Needs and Priorities, National Academy Press, Washington, DC (1984).

291

10.2.1.6 Teratogenicity Teratology may be defined as the study of the generation, causes, and manifestation of structural (anatomical) and functional (metabolic or physiological dysfunction, psychological, behavioral) alterations in development. The rapid growth and development of embryonic and fetal tissues significantly increase the susceptibility of cellular DNA to toxicant-induced changes in gene expression or in the retardation and/or arrest of cellular development. It is important to appreciate that biological effects can be elicited in embryonic tissues by chemicals at concentrations far below those that cause target organ toxicity in adults. Such chemicals, called teratogens, are a major concern to the public, to industry, and to regulatory agencies because of the low levels required to initiate cellular damage (in some cases, at levels on the order of those detected in various environments). The fundamental aim of screening studies used to evaluate the teratogenic potential of chemicals is to predict the absence of a teratological hazard for humans. Toxicity may be elicited in the pre- and postimplantation embryo, causing death (embryolethality) or mild to severe dysmorphogenesis in one or more organ systems, resulting in structural malformations and physiological or biochemical dysfunction, as well as psychological, behavioral, and cognitive deficits in the offspring at birth or in a defined postnatal period (Schardein, 1985; Ecobichon, 1997c). Timing is critical in the design of teratogenic studies because exposure must occur during the period of organogenesis, beginning after the implantation of the fertilized ovum. This event varies between animal species, as is shown in Table 10.8 (Ecobichon, 1997c). In the human, organogenesis occurs in the first 8 weeks of pregnancy. Different organ systems develop at different but sometimes overlapping rates and time periods. Frequently, the critical "toxic window" of time, when an agent can elicit an effect, is only a few days duration. Agents acquired before and after the "window" show no effect on one particular target organ but, perhaps, cause some adverse effect on another developing organ system. The concept of the toxic window becomes crucial when chemicals are tested in surrogate animal models because of the shortened gestation time and windows as narrow as 24 hours in which the appropriate concentration of agent must be in the right place to elicit an effect (Ecobichon, 1997c). Upon completion of organogenesis, further chemical exposure will not cause teratogenicity except in the still developing central nervous system, which, in most species continues to develop throughout gestation and well into the postnatal period. Frequently, behavioral and cognitive deficits are detected. The reader is referred to the chapter by MacPhail (Volume 1, Chapter 12) on neurobehavioral toxicity for further details. Assessment of teratogenicity is routinely conducted in two species, a rodent (mouse, rat or hamster) and a nonrodent (rabbit), and involves the daily administration of a range of three appropriate dosages to different groups of timed-pregnant animals throughout the period of organogenesis (e.g., days 6-15 for rats and days 6-18 for rabbits). The pregnant animals are euthanized 24 hours prior to the calculated day of parturition

292


Table 10.8 Species Differences in Gestational Endpoints (in days) Species

Implantation

Organogenesis

Gestation

10--56

270

Human

7-8

Rat

5.5-6

6-15

21-22

Mouse

4.5-6

6-15

19-21

Hamster

4.5-5

4--14

16

Rabbit

7

6-18

32

Guinea pig

6

6-20

67

Monkey (Rhesus)

9

9-40

165

7-35

114

Dog

13-14

Pig

10--12

63

Source: Ecobichon (1997 c).

and undergo a complete autopsy. The numbers of dead (late fetal deaths) and live pups are determined, the uterine muscle is examined for reabsorption sites indicative of embryonic deaths, and the ovaries are examined for the number of corpora lutea to determine the number of ova released prior to fertilization. The latter provides an index of fertility in the female. The position of each pup in the uterine horn is recorded, each is weighed, the sex is determined, and each fetus is examined for external abnormalities prior to dissection to detect internal malformations. Whole fetuses may be fixed in special fluids for histological examination or for the detection of skeletal anomalies. The design and conduction of teratogenicity studies has been described in detail elsewhere (Ecobichon, 1997c; Manson and Kang, 1994). Concerns about subtle changes in behavior and cognitive functions have led to a modification of teratological study protocols wherein some of the treated females are allowed to give birth and to rear their young for 6 weeks or longer after parturition, at which time the pups can be examined by testing strategies designed to assess behavioral and cognitive development. In such protocols, treatment may begin on the last day of implantation and continue throughout the gestational period and even into the postpartum lactational period up to weaning. With this modification in dosage regimen, the developing young receive the agent both transplacentally and via the milk throughout the longer period of neurological development-an important consideration with pesticides and their known covert actions On neuronal membranes, receptors, and neurotransmitters. 10.2.1.7 Acute Neurotoxicity Unfortunate experiences with a few of the early organophosphorus ester insecticides, principally the agent leptophos (0-4bromo-2,5-dichlorophenyl O-methyl phenylphosphonothioate, M PHOSVEC ), led to the mandatory requirement by the U.S. EPA that all pesticides undergo an acute delayed neurotoxicity assessment using the domestic chicken (Callus domesticus) as the preferred, experimental animal (Abou-Donia, 1981; U.S. EPA, 1978). The study is "acute" in the sense that, frequently, a single dose of the test substance is administered to the animal,

but "delayed" in aspect because the animals must be studied for some 7-21 days before peripheral and central neurological effects can be detected and quantified both physiologically and morphologically (Durham and Ecobichon, 1984, 1986; Slott and Ecobichon, 1984). Chemicals used as positive control standards and known from extensive study to produce the characteristic effects of leptophos include such organophosphorus esters as mipafox (N, N-di-isopropylphosphorodiamidic fluoride), DFP (O,O-di-isopropylphosphorofluoridate), and TOTP (tri-o-tolyl phosphate). The subject of organophosphate-induced delayed polyneuropathy (OPIDP) has been extensively reviewed and discussed (Abou-Donia and Lapadula, 1990; Cranmer and Hixon, 1984; Ecobichon, 1996) and is the subject of a lengthy chapter in this text by Ehrich and Jortner (Volume 2, Chapter 52). The discussion in this section will be restricted to the format of the test protocol. Adult hens of 2.0-2.5 kg body weight are housed in individual stainless steel, mesh-bottomed cages and acclimatized for 3--4 days in an environment that has a 12-hour light/dark photoperiod cycle. The birds are assessed visually for neurological competence (locomotion, posture, equilibrium-coordination, and walking strength). Groups of birds are fasted for 18 hours prior to receiving single doses of agent, usually by oral gavage using a fine rubber catheter attached to a syringe. Appropriate doses of test agent and a positive control chemical are prepared in a suitable vegetable oil solution. In situations where signs of acute poisoning are observed following treatment, atropine sulfate (30 mg/kg body weight) can be administered subcutaneously as required to alleviate the signs/symptoms. A physical examination of the control and treated birds is conducted every 24 hours by removing the hens from their cages, placing them On the floor, and observing any changes in behavior and walking ability. The responses of the hens are scored independently by two individuals, who have nO knowledge of the treatment regimen, using the scoring systems of Cavanagh (1954) and Sprague et at. (1980). At selected posttreatment time intervals over a period of 18-21 days, subgroups of hens are euthanized by anesthesia with an intravenous injection of 5.0% chloralose in 5.0% sodium borate followed by exsanguination by cardiac puncture. At dissection, sections of brain, spinal cord, and sciatic nerve are either fixed in 4.0% neutral-buffered formaldehyde prior to histological preparation and staining or frozen at -20°C in tightly sealed glass bottles for enzymatic analysis of acetylcholinesterase (AChE) and neuropathic target esterase (NTE). The morphological, physiological, and biochemical changes caused by an OPIDP-inducing organophosphorus ester are beyond the scope of this chapter and the reader is referred to the relevant chapters by Ehrich and Jortner (Volume 2, Chapter 52), Johnson (Volume 2, Chapter 50), and Wilson (Volume 2, Chapter 51). The classic, positive, morphological effects are a typical, dying back, Wallerian peripheral and central axonal degeneration with a concomitant myelinopathy (AbouDonia, 1981). The birds will show ataxia, a high stepping gait, toe dragging, an inability to balance even with the aid of the wings, and, in the most severe state, an inability to move, al-


ways remaining in a squatting position. Biochemically, an initial inhibition of nervous tissue AChE will be observed, with a slow recovery over the next 18 days. In contrast, inhibition of NTE (if the agent has the capability to do so) will be observed up to 24 hours posttreatment, but quickly recovers to normal levels of activity within 2 or 3 days. The main drawback of this test of neurotoxic potential is that it behaves properly only with a few organophosphorus ester insecticides, none of which, on the basis of structure-activity relationships, would ever be marketed. Theoretically, all phosphate esters and those phosphorothioate esters that can be metabolically desulfurated to form phosphate analogs have the potential to cause the foregoing effects. The fact that they do not do so is perhaps due to (1) high toxicity and lethality, (2) rapid detoxification, or (3) slow desulfuration to the phosphate, all of which suggest that insufficient levels of the toxic agent were acquired in vivo to elicit the effects. However, this does not mean that other classes of insecticides (organochlorines, carbamates, and pyrethroid esters) and, indeed, other classes of pesticides do not cause neurotoxicity. They do, but are quite different in their actions, a topic that was discussed in depth by Ecobichon and Joy (1994). When some of these agents are in formulations, some thought must be given to effects caused by co-solvents, emulsifiers, or adjuvants, rather than by the pesticide. Completion of the test battery listed in the first-wave studies (Table 10.2) will provide adequate information concerning the acute toxicity of a test chemical in occupationally exposed individuals, in those accidentally exposed, and intentional, suicidal attempts, all situations that reflect high-level, short-duration exposures. Some indication of the relative potency of the unknown chemical in comparison to other agents in the same chemical class can be obtained as can possible mechanisms of action. A measure of the mutagenic potential can be obtained from a number of assay systems, and identified mutagenicity possibly can be confirmed by teratogenic effects in two sensitive animal models. These studies, completed at relatively low cost, are the basis upon which decisions will be made either to shelve the potential pesticide, halt further development, or proceed to the more expensive and time-consuming studies that comprise the second-wave list. 10.2.2 SECOND-WAVE STUDIES Considerably more humans are exposed to pesticides at relatively low levels over a much longer period, even a lifetime. The origins of such pesticide exposure are air- and water-borne remnants of agricultural or forestry operations or residues in foods above, at, or below tolerances established by regulatory agencies. The simulation of such exposures requires the development of other testing strategies, that is, short-term (subacute, subchronic) and long-term (chronic) studies (Ecobichon, 1997d). The endpoints of toxicity in such studies may differ considerably from those of acute intoxication and they include: 1. Reproductive studies in which target organ toxicity in the various phases of male and female reproduction are

293

explored. Many of these phases are as susceptible to chemical-induced damage and/or dysfunction as would be expected in the developing fetus. 2. Cancer, a threat to the well-being of all humanity and a "disease" that has been linked to pesticide residues as part of the "environmental contamination/exposure" scenario. A degree of flexibility must be maintained in these studies because since it is impossible to predict if and when toxicity will appear. Although anticipated target organ toxicity can be predicted from physiological properties of the test agent and experience with similar chemicals, serendipitous observations are the rule rather than the exception. Studies should be kept as open-ended as possible to permit the exploration of results, which necessitates the inclusion of sufficient numbers of animals at the "front end" to (1) demonstrate a dose-related toxicity without too high a mortality rate, (2) identify suspicious toxicological events, and (3) allow enough animals in each treatment group to study the possible permanence or reversibility of the toxicity. As shown in Table 10.9, there are specific goals to be achieved in addition to detecting toxicity. The experimental design of such studies has been discussed in detail by Arnold et al. (1990), and by Ecobichon (1997c), among others, and will only be summarized here. Subacute toxicity studies, usually of 2-4 weeks duration, are conducted as range-finding studies to choose dosage levels to be used in longer-term, subchronic (up to 90 days) and chronic (6 months to 2 years) studies. However, a recent examination of newly introduced OECD guidelines indicates that the subacute/subchronic terminology is being phased out; particular tests designate the duration of the study and leave no doubts in investigators' minds as to how long the study should last (e.g., OECD Guideline 407-Repeated Dose 28-Day Oral Toxicity Study in Rodents. There has also been a reduction in the duration of chronic studies from 2 years to 6 months for rodents, although chronic studies in dogs are still conducted for 2 years.

Table 10.9 Objectives to Be Achieved in Subchronic and Chronic Toxicity Studies 1.

Examine the biological nature of the toxic effects elicited from low dosages, monitoring a range of biological parameters

2.

Ascertain the variation in species response(s) to repeated exposure to the agent, looking for commonality of responses and/or distinct species differences

3.

Assess possible cumulative effects of the repeated exposure as body burdens of the agent are acquired with time

4.

Determine the nature of macroscopic and microscopic organ or tissue damage as it develops

5.

Identify the approximate dosage at which the altered physiological, biochemical, and morphological changes might occur

6.

Predict the long-range adverse health effects in the species arising from intermittent, repeated, or chronic exposure to the agent

294

CHAPTER 10

Toxicity Testing

10.2.2.1 Subchronic Studies

Such studies are designed to explore possible mechanisms of action in animals over a longer time period and at a lower range of doses than those reported to be lethal. Both sexes of the same age, strain, and species are used. At least three dosage levels (low, moderate, and high) as well as control (untreated or vehicle-treated) are included. The chosen dosage levels, generally derived from acute toxicity results, are usually fractional dosages, such as 1.0,5.0, and 10.0% of the LDso, and are administered via the route by which humans would be expected to acquire the agent, for example, in drinking water or in the food. To ensure that the entire dose is received, oral gavage with a feeding needle may be chosen. Dermal application may be required to satisfy occupational exposure concerns. Although a wide variety of physiological and biochemical parameters can be used to monitor the general well-being of the animals and to detect any adverse biological effect(s), it is imperative to determine when the first signs of toxicity appear at each dose level. A plan should be made to euthanize subgroups of each treatment group at 30, 60, and 90 days during treatment and carry out a full morphological and biochemical study to detect the appearance and the progression of a lesion rather than waiting for the treated animals to become ill. Frequently, a time- and a dose-effect relationship can be established. All dead and severely moribund animals, as well as those selected for euthanasia at fixed time points in the study, should undergo careful necropsy and examination. Tissues should be preserved for embedding, sectioning, staining, and microscopic examination. A large number of body organs are removed from each animal at necropsy for examination or are held for special study (Ecobichon, 1997d). Frequently, the initial tissue study involves only control and high dose animals' tissues to establish that there is a difference. Having done this, the low and moderate dose treated animals' tissues will be examined to establish a sUbjective dose-effect relationship. Available techniques allow microscopic quantification of tissue cell size and volume subcellular organelle volume, etc. Frequently seen cytologicai reactions to a chemical insult appear to adhere to a dose-effect relationship.

conducted with longer-lived species would require an adjustment (increase) in the length of the study if the same three-dose range was to be used. Such a study was a costly endeavour, considering that a 2-year study is only 20% of the life-span of a dog and 13% of the life-span of a monkey. The alternative approach, taken by regulatory agencies, is to adjust (increase) the dosage administered to longer-lived species so that the animals acquire a "life-span dosage" in 2 years. The pitfalls of this approach are obvious: Exposure levels are so high that the animals are unable to efficiently biotransform and excrete the test agent. Currently, most national and international regulatory agencies will accept a 6-month chronic study in rodents, but require a 24-month study in nonrodents and primates. A controversial point is the perspective that no new chronic toxicity has been detected in 24-month studies that was not seen in "appropriately designed" 6- or 12-month studies. All chronic studies, regardless of duration, suffer from the same design defect-an insufficient number of animals (usually 50 per sex per treatment group) to detect the effects associated with an agent that has a low incidence of toxicity in humans. An incidence rate of 1.0 in humans would require some 299 animals to be certain that the effect was not overlooked (Zbinden, 1973). An incidence rate of 0.1, would require some 2995 animals in the study. At best, the observations or lack thereof are "guesstimates" of toxicity/safety. As with the subchronic studies, groups of animals (both sexes, same strain and species) should receive one of a range of three dosages (low, moderate, and high) via a suitable route of exposure. These dosage levels should be chosen on the basis of the results of the subchronic studies (Ecobichon, 1997d). Once again, the investigator-selected subgroups should be euthanized at suitable intervals during the treatment period to detect the appearance and progression of lesions. The general well-being of the animals should be monitored by suitable parameters of

Table 10.10 Duration of Studies in Experimental Animals and Time Equivalents in the Human Duration of study (in months)

10.2.2.2 Chronic Studies

Species

A major challenge in designing long-term toxicologic experiments is to calibrate exposure levels to permit a reasonable, normal laboratory life (health, appearance, growth, and development) for the animals, while guaranteeing obvious evidence of chronic toxicity over and above that typically seen in aged animals (Huff et aI., 1991). As with the subchronic studies the objectives of chronic studies are to characterize the mech~ anism(s) by which an agent induces some toxic effect(s) when administered over a considerable portion of the life-span of the test animal. The greatest controversy centers on just how long that study should be. Originally, in regulatory parlance, a chronic mammalian study had a duration of 2 years, which represents the approximate life-span of a laboratory rodent (Table 10.10). Studies

Percent of life-span

6

3

24

12

Rat

4.1

12.0

2S.0

49.0

99.0

Rabbit

I.S

4.S

9.0

18.0

36.0

Dog

0.82

2.5

4.9

9.8

20.0

Pig

0.82

2.5

4.9

9.8

20.0

Monkey

0.55

1.6

3.3

6.6

13.0

Human equivalents (in months) Rat

34

101

202

404

808

Rabbit

12

36

72

145

289

Dog

6.S

20

40

81

162

Pig

6.5

20

40

81

162

Monkey

4.5

13

27

61

107

Source: Ecobichon (1997d).


blood chemistry and urinalysis. The required minimum number of animals should be expanded to allow for possibly high, toxic doses and (unexpected) observations of toxicity. In addition, sufficient numbers of test subjects permit a recovery period study following termination of treatment. Extensive morphological examination of a variety of tissues should be conducted (Ecobichon,1997d). 10.2.2.3 Carcinogenicity Studies The design of a carcinogenicity study, which usually is conducted in rodent species, must permit detection of one (or more) of the properties in the operational definition of a carcinogen (Table 10.11; Ecobichon, 1997c). Generally, carcinogenic chemicals cause an effect at a low incidence rate, creating difficulties with the numbers of animals that must be used costeffectively in the study. Prolonged exposure (in excess of 12 months) of the animals is required to demonstrate the carcinogenic potential of most test agents. The normal 6- or 12-month chronic toxicity study would not be of sufficient duration to demonstrate this properly. The exorbitant costs of running parallel chronic toxicity and carcinogenicity studies has resulted in their combination, where high exposure levels are used to identify chronic toxicity and additional groups of animals receive much lower daily doses of the test agent for a much longer time to detect carcinogenicity (Ecobichon, 1997b). Typically, carcinogenicity studies are of 24-month duration and the highest dose of agent selected will be one that neither produces clinical signs of toxicity nor affects the long-term health or the normal longevity of the animals. Appropriate designs for such studies have been discussed (Ecobichon, 1997b; Robens et aI., 1994). There has been considerable controversy over the selection of appropriate dosage levels for carcinogenicity studies. The dosage, defined in the previous paragraph as the highest level, is called the maximum tolerated dose (MTD) and is usually chosen as a consequence of the results obtained in subchronic studies. It elicits no adverse biological effects other than cancer. Normally, two additional dose levels, lower than the MTD, are used in an attempt to establish a dose-related incidence of tumors, decreased latency period, etc. Butterworth et al. (1991) provided a rational discussion of this controversy. Along with the routine biochemical monitoring of general health, particular attention is focused at necropsy on detection of "lumps and bumps" and, at later microscopic examination, on the detection and identification of cell hyperplasia, preneoTable 10.11 Operational Definition of a Carcinogen

An agent that has the ability to induce tumors as evidenced by: I.

An increased incidence of tumor types found in controls

2.

Occurrence of tumors earlier than in controls

3.

Development of tumor types not seen in controls

4.

An increased multiplicity of tumors in individual animals

Source: Ecobichon (1997b).

295

plastic nodules, and tumors (benign and malignant) to satisfy the criteria listed in Table 10.11 for a carcinogen. 10.2.2.4 Short-Term Carcinogenicity Studies The discovery and promotion of several strains of mice and rats that have predispositions toward developing high incidences of certain tumor types has been of considerable assistance in short-term studies for carcinogenicity. Chemical exposure for 90-120 days is frequently sufficient to elicit an increased incidence and/or earlier appearance of tumors. Although this type of testing does not replace the full-scale studies, these animal models can be used to screen chemicals at considerably lower cost, which saves funds for application to extended studies of those chemicals that do cause positive results. 10.2.2.5 Reproductive Studies Reproductive toxicology is the study of the occurrence, causes, manifestations, and sequelae of the adverse effects of exogenous agents on reproduction (John son, 1986). Reproductive hazards encompass adverse health effects to the prospective mother and father (loss of libido, infertility, sterility) as well as to the developing offspring (abortion, fetal and/or perinatal death, teratogenesis). Many of these events are considered to be associated with cellular mutations. Although mutagenicity and teratogenicity were assessed by the test protocols in the first-wave studies, toxicological evaluation of any chemical must encompass the entire breadth of the life cycle. To date, there are no tests to examine the direct effect(s) of the agent on gametogenesis in either adult male or female mammals. The gametes (spermatozoa, ova) as well as the fertilized ovum, and the pre- and postimplantation embryo are exquisitely susceptible to physical and chemical insult, primarily because of the frequency of cell division and replication of cellular DNA in repeated mitotic and meiotic division. These activities are sensitive to exceptionally low concentrations of agents. In the past, most emphasis was placed on studies that examined the effects of agents on the female in her role as the vehicle for the susceptible, developing embryo. Essentially, these studies are the Segment 11 or teratogenic studies already discussed. More recently, attention has focused on pretreating either male or female animals before mating to examine any effects on gametes, because any damage to the gamete may play a vital role in the viability of the developing embryo. Experimental paradigms have been described to examine the mechanisms of toxicity in either or both of the sexes (Ecobichon, 1997c; Manson and Kang, 1994; Zenick et aI., 1994). In a classical Segment I (fertility, reproduction) study, groups of either male or female test animals are treated prior to mating with a range of dosages (usually three levels) for one gametogenic cycle (60-days for the male or 15 days for the female). If the treated female animals are mated with normal, untreated males, the pregnant females continue to receive the same daily treatment for either the duration of the pregnancy or beyond parturition if there is concern about effects in the lacta-

296

CHAPTER 10

Toxicity Testing

tional time period. If the postnatal development period is being studied, the pregnant animal will be allowed to give birth and to rear the young past the weaning stage, thereby permitting study of postnatal viability and behavioral/cognitive development. The pretreated male animals are mated with healthy, untreated, virgin females throughout the next gametogenic cycle. Essentially, a group of new females (n = 4 or 5) are placed with each male every 5 days. Those females that are confirmed to be pregnant are removed and placed in individual cages until the day before parturition, when they are euthanized so that the effects on the developing embryo can be studied. The focus of the study is on whether the test agent can produce an effect on spermatogenesis, which is expressed as alterations in a number of indices (i.e., fertility, mating, fecundity, gestational time, live births, dead fetuses) that are calculated at the termination of the pregnancy (Ecobichon, 1997c). In this bioassay, the reproductive biology of the female is assumed to be normal; only that of the male is presumed to be affected. A 60-day treatment with an active mutagenic agent might produce disastrous effects on spermatogenesis at several different stages, although the results are unclear except in the cases of sterility or reduced fecundity (Ecobichon, 1997c). A variant of the previously discussed test is the dominant lethal test in which selected doses of test agent are administered either as a pulse dose or single dose for a short term (5 to 7 days) to adult, breeding-age males. The dosage(s) selected is estimated to cause severe chromosomal damage or lethal germ-cell mutations, resulting in fetallethality. Whereas gamete production is a continuous synthesis from stem cells in the testes, a pulse dose may elicit limited damage at one or only a few stages. A breeding program is established where, for the next cycle after treatment, females are bred with the treated male every 4-5 days. These females are then followed to the end of their pregnancies and the fetuses are examined on the day before parturition. Some 50 females will be bred with each male over the 60-day period following the pulse dose of test agent. The main advantage of this assay is the ability to test the chemical sensitivity of the germ cells in vivo at different (premeiotic, meiotic, postmeiotic) stages of spermatogenesis. The disadvantage is that if the chemical is only a weak mutagen, it will not produce a significant amount of chromosomal damage, thereby "escaping" the screening assay (Ecobichon, 1997c). The classical reproductive, two-generation study has, in the past, been one in which groups of young, healthy, female animals (Fo generation) are treated with the test agent (three different exposure levels) from the time of conception (mated with normal males), throughout the gestation period, the rearing of the postnatal progeny (F lA generation), and at least through one repeated breeding cycle (FIB generation) (Ecobichon, 1997c). The FIA progeny are euthanized for gross and microscopic study and, following a 2-week rest period, the female is bred again to produce an FIB generation, the female offspring of which are the source of animals for the next generation. The selected FIB females receive the same level of exposure to the test agent as their dams from weaning throughout their breeding cycle to produce F2 generation pups. Normally, a two-generation

study is conducted under regulatory requirements, but the study can be extended to any number of generations, keeping in mind that the progeny receive the test agent transplacentally and via the milk until weaned, at which time they must begin to receive the agent via their food or drinking water. It must be emphasized that in both teratological and reproductive studies, the toxicity observed may be related to maternal and/or nutritional toxicity, especially if high concentrations of a rather nontoxic chemical are being administered via the diet or the test agent has an unpleasant odor or taste. In such scenarios, the female may reject the food, thus depriving the developing fetuses of nutrients essential for normal growth and development or affecting fertility, fecundity, and/or the ability of the female to maintain the pregnancy. All too frequently, adverse reproductive outcomes are falsely attributed to the test chemical, when, in effect, the defect is nutrition related.

REFERENCES Abou-Donia, M. B. (1981). Organophosphorus ester-induced delayed neurotoxicity. Am. Rev. Pharmacal. Taxical. 21, 511-548. Abou-Donia, M. B., and Lapadula, D. M. (1990). Mechanism of organophosphorus ester-induced delayed neurotoxicity: Type I and Type n. Ann. Rev. Pharmacal. Taxical. 30, 405-440. Ames, B. N., Durston, W. E., Yamasaki, E., and Lee, E. D. (1973). Carcinogens are mutagens: A simple test system combining liver homogenates for activation and bacteria for detection. Proc. Natl. Acad. Sci. U.S.A. 70, 22812285. Ames, B. N., McCann, J., and Yamasaki, E. (1975). Methods for detecting carcinogens and mutagens with the salmonella/mammalian microsome mutagenicity test. Mutat. Res. 31, 347-364. Arnold, D. L., Grice, H. C., and Krewski, D. R., eds. (1990). "Handbook of In Vivo Toxicity Testing." Academic Press, San Diego. Basketter, D. A., Scholes, E. W., and Kimber, 1. (1994). The performance of the local lymph node assay with chemicals identified as contact allergens in the human maximization test. Faad Chem. Taxical. 32,543-547. Botham, P. A., Basketter, D. A., Maurer, T., Mueller, D., Potokar, M., and Bontinck, W. J. (1991). Skin sensitization-A critical review of predictive test methods in animals and man. Faad Chem. Taxicol. 29, 275-286. Bruner, L. H., Parker, R. D., and Bruce, R. D. (1992). Reducing the number of rabbits in the low-volume eye test. Fundam. Appl. Taxica!. 19, 330-335. Buehler, E.V. (1964). A new method for detecting potential sensitizers using the guinea pig. Taxical. Appl. Pharmacal. 6, 341. Butterworth, B. E., Goldsworthy, T. L., Popp, J. A., and McClellan, R. O. (1991). The rodent cancer test: An assay under seige. ClIT Activities 11, 1-8. Cavanagh, J. B. (1954). The toxic effects of tri-a-cresylphosphate on the nervous system: An experimental study in the hen. J. Neural. Neurosurg. Psychiatry 17, 163-172. Chan, P. K., and Hayes, A. W. (1994). Acute toxicity and eye irritancy. In "Principles and Methods of Toxicology" (A. W. Hayes, ed.), 3rd ed., pp. 579648. Raven Press, New York. Conquest, P., Durand, G., Laillier, J., and Blazonnet, B. (1977). Evaluation of ocular irritation in the rabbit: Objective vs subjective assessment. Taxicol. Appl. Pharmacal. 39, 129-139. Copplestone, J. F. (1988). The development of the WHO Recommended Classification of Pesticides by hazard. Bull. WHO 66, 545-551. Cranmer, J., and Hixon, J., eds. (1984). "Delayed Neurotoxicity." Intox Press, Little Rock, AR. De Sousa, D. J., Rouse, A. A., and Smolon, W. J. (1984). Statistical consequences of reducing the numbers of rabbits utilized in eye irritation testing: Data on 67 petrochemicals. Taxical. Appl. Pharmacal. 76, 234-242.

References

Draize, J. N., Woodward, G., and Calvery, H. O. (1944). Methods for the study of irritation and toxicity of substances applied topically to the skin and the mucous membranes. J. Pharmacol. Exp. Ther. 82, 377-389. Durham, H. D., and Ecobichon, D. J. (1984). The function of motor nerves innervating slow tonic skeletal muscle in hens with delayed neuropathy induced by TOTP. Can. 1. Physiol. Pharmacol. 62, 1268-1273. Durham, H. D., and Ecobichon, D. J. (1986). An assessment of the neurotoxic potential offenitrothion in the hen. Toxicology 41, 319-332. Ecobichon, D. J. (1996). Toxic effects of pesticides. In "Casarett and Doull's Toxicology. The Basic Science of Poisons," 5th ed., (c. D. Klaassen, ed.), Chap. 22, pp. 643-689. McGraw-Hill, New York. Ecobichon, D. J. (1997a). Acute toxicity studies. In "The Basis of Toxicity Testing," 2nd ed., Chap. 3, pp. 43-86. CRC Press, Boca Raton, FL. Ecobichon, D. J. (1997b). Mutagenesis-{:arcinogenesis. In "The Basis of Toxicity Testing," 2nd ed., Chap. 6, pp. 157-190. CRC Press, Boca Raton, FL. Ecobichon, D. J. (1997c). Reproductive toxicology. In "The Basis of Toxicity Testing," 2nd ed., Chap. 5, pp. 117-156. CRC Press, Boca Raton, FL. Ecobichon, D. J. (1997d). Subchronic and chronic studies. In "The Basis of Toxicity Testing," 2nd ed., Chap. 4, pp. 87-116. CRC Press, Boca Raton, FL. Ecobichon, D. J., and Joy, R. M. (1994). "Pesticides and Neurological Diseases," 2nd ed., CRC Press, Boca Raton, FL. European Commission Directive 83/4671EEC (1983). Freeberg, F. E., Nixon, G. A., Reer, P. J., Weaver, J. E., Bruce, R. D., Griffith, J. F., and Sanders, Ill, L. W (1986). Human and rabbit eye responses to chemical insult. Fundam. Appl. Toxicol. 7, 626-634. Gad, S. c., Dunn, B. J., Dobbs, D. W, Reilly, c., and Walsh, R. D. (1986). Development and validation of an alternative dennal sensitization test: the mouse ear swelling test (MEST). Toxical. Appl. Pharmacol. 84, 93-114. Hayes, Jr., W J. (1975). "Toxicology of Pesticides," p. I. Williams and Wilkins, Baltimore. Huff, J., Haseman, J., and Rail, D. (1991). Scientific concepts, values and significance of chemical carcinogenesis studies. Ann. Rev. Pharmacal. Toxicol. 31,621-652. Johnson, E. M. (1986). Perspectives on reproductive and developmental toxicology. Toxicol. Indus. Health 2, 453-482. Kimber, I., and Basketter, D. A. (1992). The murine local lymph node assay: A commentary on collaborative studies and new directions. Food Chem. Toxicol.30, 165-169.

297

Manson, J. M., and Kang, Y. J. (1994). Test methods for assessing female reproductive and developmental toxicology. In "Principles and Methods of Toxicology," 3rd ed. (A. W Hayes, ed.), pp. 989-1038. Raven Press, New York. McCreesh, A. H., and Steinberg, M. (1983). Skin irritation testing in animals. In "Dennatotoxicology," 2nd ed., (F. N. Marzulli and H. I. Maibach, eds.), pp. 147-166. Hemisphere Publishing, New York. National Research Council, Toxicity Testing, Strategies to Detennine Needs and Priorities, National Academy Press, Washington, DC (1984). Patrick, E., and Maibach, H. I. (1994). Dennatotoxicity. In "Principles and Methods of Toxicology," 3rd ed., (A. W Hayes, ed.), pp. 767-803. Raven Press, New York. Robens, J. F., Calabrese, E. J., Piegorsch, W. W., Schuler, R. L., and Hayes, A. W (1994). Principles of testing for carcinogenicity. In "Principles and Methods of Toxicology," 3rd ed., (A. W Hayes, ed.), pp. 697-728. Raven Press, New York. Schardein, J. L. (1985). "Chemically Induced Birth Defects." Dekker, New York. Seabaugh, V. M. (1994). EPA's requirements for dennal irritation and sensitization testing. Food Chem. Toxicol. 32, 93-95. Slott, v., and Ecobichon, D. J. (1984). An acute and subacute neurotoxicity assessment of trichlorfon. Can. J. Physiol. Pharmacol. 62, 513-518. Sprague, G. L., Sandvik, L. L., Bickford, A. A., and Castles, T. R. (1980). Evaluation of a sensitive grading system for assessing acute and subchronic delayed neurotoxicity in hens. Life Sci. 27, 2523-2528. Talsma, D. M., Leach, C. L., Hatoum, N. S., Gibbons, R. D., Roger, J. c., and Garvin, P. J. (1988). Reducing the number ofrabbits in the Draize eye irritancy test: A statistical analysis of 155 studies conducted over 6 years. Fundam. Appl. Toxicol. 10, 146-153. U.S. Code of Federal Regulations, Title 40, Part 158 (40CFR Part 158). U.S. Environmental Protection Agency (EPA) (1978). Acute delayed neurotoxicity study and subchronic neurotoxicity studies. Fed. Register 43, 37,36237,363,37,374-37,375. Zbinden, G. (1973). "Progress in Toxicology. Special Topics," Vol. I, pp. 1719. Springer-Verlag, New York. Zenick, H., Clegg, E. D., Perreault, S. D., Klinefelter, G. R., and Gray, L. E. (1994). Assessment of male reproductive toxicity: A risk assessment approach. In "Principles and Methods of Toxicology," 3rd ed., (A. W. Hayes, ed.), pp. 937-988. Raven Press, New York.

CHAPTER

11 Regulatory Evaluation of the Skin Effects of Pesticides MichaelO'Malley University of California, Davis

11.1 INTRODUCTION 11.1.1 BASIC PATTERNS OF SKIN REACTION Clinical effects of pesticides on the skin include both systemic and topical reactions. Systemic effects, such as urticaria, chloracne, and porphyria cutanea tarda, may occur following ingestion, inhalation, or topical exposure. Direct topical effects include acute irritation and corrosion, subacute (gradual-onset) irritation, and delayed-onset allergies. Any of the injuries described above may damage the pigment-producing basal layer of the skin, resulting in either an increase or a decrease in epidermal melanin production. Typically, both injuries and residual effects occur in a pattern that coincides with the site of contact. Depending upon the time interval between exposure and the onset of lesions, recognizing the source of the skin injury may be simple or complex. Protocol for clinical patch testing

In the regulatory arena, evaluation of the capacity of individual compounds to cause irritant or allergic reactions depends upon animal testing as well as analysis of human use experience. Protocols for evaluating allergy and irritation in experimental animals are discussed.

11.1.2 TESTING REQUIREMENTS AND TEST PROTOCOLS The requirements for skin testing of pesticides for the purposes of federal and state registration vary with the government jurisdiction. In the United States, primary dermal irritation studies and sensitization studies are required for each manufacturinguse product and each end-use product. The tests performed in this manner are considered part of the regulatory database and are not available in the public literature. The irritation testing requirements are similar to the standard Draize tests (Inset 2). Dermal sensitization tests may be done according to one of several standard protocols (Inset 3).

Application of previously identified nonirritating concentration of test substance for 48 hours, followed by removal of patch and initial reading. Follow up reading at 96 hours. Simplified scoring system for grading patch tests: O--no visible reaction 1+ -erythema 2+ -erythema and blistering 3+ -necrotic reaction

Dermal irritation tests

Distinguishing between allergic and irritant effects is a primary goal of both clinical and regulatory evaluation of the skin effect of pesticides. Clinically, irritant reactions tend to develop within a short time of exposure, whereas skin allergies are typically delayed in onset. Exceptions to this simple rule occur-some irritant reactions are cumulative and some allergic reactions occur within minutes of contact with the offending allergen (urticaria). A summary of the clinical protocol for provocation tests (or patch tests) is shown (Inset 1). Handbook of Pesticide Toxicology


Irritation test using albino rabbits (Draize test) A single dose of the technical material with detailed characterization of contaminants, or an end-use product, is applied to the skin of one or several experimental animals (depending upon the possibility of a corrosive reaction) for 4 hours. The irritation is scored at intervals until the irritation has resolved or is considered permanent. Based upon scores at 72 hours and persistence of irritation for more than 14 days, materials are categorized as corrosive (category 1-72 hr dermal irritation score >7), severe irritants (category II-72 hr dermal irritation score 5-7), moderate irritants (category III-72 hr dermal irritation score 2-5), or minimally irritant (category IV-72 hr dermal irritation score 0-2) -may be divided into compounds that produce no irritation (nonirritants) and those that produce transient, mild irritation.

A newer technique, not yet approved by the EPA, but probably more reproducible than any of the currently approved methods for testing dermal sensitization, is the regional lymph

299

Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.

300

CHAPTER 11

Regulatory Evaluation of the Skin Effects of Pesticides

Allergy testing protocols Sensitization test protocols Following initial exposure to a test substance, tbe animals are challenged, to establish whether the animal has developed hypersensitivity. This is evaluated by comparing scores during the induction period with those during the challenge period and witb those of control animals who received the challenge witbout initial exposure. For ambiguous results on challenge, a rechallenge phase is used. Epicutaneous test methods: Buehler test-A closed patch is applied for six hours, weekly, during a 3-week induction period; tbe test concentration for induction is chosen to be approximately lO-fold higher than the expected human exposure concentration. Challenge, with a nonirritating concentration, takes place weeks 5,6,7. Open epicutaneous test-After establishing the concentration that produces minimal irritation and no irritation threshold, induction is begun at the latter concentration. Applications are repeated daily for tbree weeks or five times weekly for four weeks, always on the same skin site. The challenge is conducted on day 21 using the minimal irritant and some lower concentrations-skin reactions are read after 24, 48, and/or 72 hours. Rechallenge, if necessary, is done on day 35. Intradermal methods: Guinea-pig maximization test (GPMT)-Induction is begun on day 0 with 0.1 ml test material intradermal (by injection) togetber with 0.1 cc Freund's Complete Adjuvant (FCA). Control animals receive only tbe injection of 0.1 ml FCA. On day 7 induction is boosted by occluding the test material against the skin for 48 hours. On day 21 tbe challenge is performed on a shaved 4-cm2 area on the left flank, using a nonirritating concentration of the test material. Controls are treated with occluded vehicle only. If challenge reactions are ambiguous, animals are rechallenged on day 28. Freund's complete adjuvant (FCA) test-Induction day 1,5,9: 0.1 ml test material in FCA in shoulder of animals of the control group treated witb FCA only. Day 21 + 36 challenge and rechallenge: A: a minimum irritating concentration; B: max nonirritating concentration both tested in test animals and controls. Days 22-24; 36-38 skin sites read 24, 28, 72 h after challenge and rechallenge. The test is simple to perform and involves low material and operational expenses.

node assay (Ashby et aI., 1995; Ikarashi et aI., 1994, 1996). The test involves use of a 3-day induction period, followed by monitoring of the uptake of H3 -labeled thymidine in regional lymph nodes (excised and placed in cell culture after sensitization) as a marker of sensitization. In addition to H3 -labeled thymidine, sensitization can be monitored using cell number or levels of interleukin-2 produced in cell culture (Hatao et aI., 1995). The results compare favorably to the more cumbersome, 3- to 7-week-long in vivo assays (Ikarashi et aI., 1994), but pesticides have not been systematically studied with this technique. For interested readers, additional details of predictive skin testing procedures have recently been described by Bashir and Maibach (Bashir and Maibach, 2000).

11.1.3 REGULATORY DECISIONS The principal regulatory decision dependent on the results of preregistration animal testing is the content of precautionary statements on the pesticide product label. Materials showing corrosive effects or reversible skin irritation are labeled as such.

Labels for products not found to be corrosive or irritating generally carry statements advising the user to minimize the degree of skin contact. Materials judged as sensitizers in animal tests or reported as sensitizers in the public domain scientific literature are required to indicate the possibility of skin sensitization on the product label.

11.1.4 USE OF ILLNESS-SURVEILLANCE DATA WITH EXPERIMENTAL DERMAL IRRITATION AND SENSITIZATION TESTS Apart from tests required for registration, additional regulatory information is obtained from postregistration surveillance of pesticide illness reports and from public literature reports on adverse skin effects of pesticides. A summary of information on irritation, sensitization, and postregistration illness-surveillance information derived from California Department of Pesticide Regulation (CDPR) data for 178 active ingredients is shown in Table 11.1. The data are discussed below by use and structural category. The sensitization data are based on a review of study summary memoranda for 427 formulations or active ingredients reviewed between January, 1989 and July, 1997. The irritation data derive from a review of the publicly available product label database.' Probable or definite skin illness or injury cases involving the same products are also shown in the table. Because the pesticide illness database (1982-1995) contained more than 10,000 reports of skin injury or illness from both agricultural and nonagricultural use, the review was limited to 991 cases involving probable or definite skin effects to pesticide applicators from a single active ingredient. None of the 991 cases had patch testing performed. For purposes of this review, these cases are referred to in the text below as the "pesticide handler database" or "handler database." Further details regarding the cases in the database are included in a recent review (O'Malley, 2000).

11.2 REVIEW OF USE CATEGORIES 11.2.1 ANTIMICROBIAL AGENTS Many of the antimicrobial agents registered as pesticides are corrosive or markedly irritant. This biological property may be correlated with underlying chemical reactivity linked in many cases to an ability to provoke sensitization.

11.2.1.1 Isothiazolins (Kathon® Compounds) In animal irritation studies, several of the isothiazolin compounds (octhilinone, 2682-20-4, 82633-79-2, 26172-55-4, 226530-20-1) are acutely irritant or corrosive in their concentrated forms. Although no dermal sensitization studies were reviewed for these compounds, all are labeled as sensitizers because of numerous reports in the public literature documenting their capacity to sensitize (Bruze and Gruvberger, 1988; I http://www.cdpr.ca.gov.

under database resources.

11.2 Review of Use Categories

301

Table 11.1

Compound

CAS #

Predictive tests in animals

Data from pesticide handler data base 1982-1995

Draize irritation test

# of cases

Sensitivity

Case examples

Antimicrobials

Isothiazolin (Kathon®) compounds 1,2-benzisothiazolin3-one

2634-33-5

Corrosive

Sensitizer per public domain literature

2-methyl-4-isothiazolin3-one

2682-20-4

Labeled as a corrosive


5-chloro-2-methyl-4isothiazilone-3-one and 2-methyl-4,5-trimethylene4-isothiazolin-3-one

82633-79-2 26172-55-4

Corrosive


2-n-octyl-4-isothiazolin3-one

26530-20-1



13

89-1312: A worker sanitizing cooling towers with a Kathon® compound accidentally spilled some of the solution on himself and suffered a chemical bum. 90-552: An employee adding concentrate (of Kathon®) to washing solution spilled some of the material onto trouser leg of work pants, causing a chemical bum.

Quarternary ammonium compounds alkyl dimethyl benzyl ammonium chloride (multiple compounds alkyl groups = cl4,cl6 etc.)

112-18-9

Corrosive at 50% concentration

Nonsensitizer in modified Maguire method (variant of the FCA test)

dioctyl dimethyl ammonium chloride

5538-94-3

Corrosive at 50% concentration

Sensitizer by Buehler test

trimethyl ammonium chloride (didecyl dimethyl ammonium chloride)

7173-51-5

Corrosive

Nonsensitizer in Buehler test; sensitizer per public domain literature

107

88-1893: A worker splashed a sanitizer containing quaternary ammonium compounds onto his face. Four hours after exposure, he developed 6-10 macular lesions at the site where the material splashed on him.

Chlorine compounds and chlorine stabilizer---Cases reported represent irritant reactions in end-users of sanitizers/disinfectants cyanuric acid

108-80-5

No data

No data

16

hexahydro-l,3,5-triethyls-triazine

108-74-7

Corrosive

Weak sensitizer by Buehler test

0

sodium hypochlorite

7681-52-9

Labeled as an irritant


132

91-2327: A custodial employee splashed material on her right arm while cleaning a toilet and subsequently developed itchy, red, and swollen area at the site of contact.

87-1468: A pet store employee developed a severe, painful rash on her hands from using a 12.5% sodium hypochlorite product to clean kennels without wearing gloves. The product proved to be a swimming pool sanitizer used in violation of the product label.

Phenolic compounds: 128 phenolic compounds registered for use as disinfectants; typical examples include ortho-phenylphenol, p-tert-butyl phenol o-phenylphenol

90-43-7

11 % formulation is a category II irritant

No data

p-tert-butylphenol

98-54-4

No data


Phenol

108-95-2

Undiluted phenol is an irritant

No data

tributyltin oxide

56-35-9

Labeled as a minimal irritant

No data except on mixtures

tributyltin methacrylate and tributyltin fluoride

91745-52-7

Corrosive

Equivocal reaction in Buehler test

29

88-909: A hospital janitor got disinfectant on the hand through a hole in disposable glove, caused burning of the skin, diagnosed as irritant contact dermatitis.

Organotins 90-728: An employee added a mildewcide to a can of paint. When she pounded the lid back on the paint can, some of the material splashed on her, resulting in a rash on the face and neck. 0

(continues)

302

CHAPTER 11


Table 11.1 (continued)

Compound

CAS #




Sensitivity

# of cases

Case examples

Miscellaneous

2,2-dibromo-3-nitrilpropionamide

10222-01-2

Category II irritant

Inadequate data

0

2-(hydroxymethylamino)ethanol

34375-28-5


N onsensitizer

0

Bronopol 2-bromo-2-nitro-1 ,3propanediol

52-51-7


Sensitizer per public literature

0

1,2-dibromo-2,4-dicyano butane

35691-65-7

Labeled as an irritant


0

iodine

7553-56-2

Technical material corrosive

Nonsensitizer

acephate

30560-19-1

Technical material causes transient irritation

Nonsensitizer

3

89-2500: An applicator developed a rash, described as urticaria and contact dermatitis, on both arms soon after acephate. The symptoms disappeared soon after he showered, but reappeared when he next applied acephate. 83-2409: An applicator spraying acephate on trees was exposed to liquid insecticide soaking through his clothes from a leaking fitting which allowed material to soak through his clothes; developed contact dermatitis. 86-1084: A structural pest control operator was trying to attach a crack and crevice injector to a spray can containing acephate. He sprayed his face and hands, and contaminated the respirator he was wearing and subsequently developed erythemaous papules and vesicles on forearms, hands, and ears.

chlorpyrifos

2921-88-2


41 % agricultural formulation labeled as a sensitizer; animal studies negative

11

diazinon

333-41-5

Minimal irritant

Sensitizer

5

85-343: An applicator treating a large carpet area for fleas came using a backpack, noticed that some of the spray material (chlorpyrifos) had leaked, soaking his lower back and upper legs. He had burning at the site of contact, but did not develop overt dermatitis. 87-2537: An applicator developed a rash on his arms and chest after a hose ruptured. He changed his shirt, but did not shower.

dichlorvos

62-73-7

Labeled as a ninimal irritant

No data

dimethoate

60-51-5

Technical material causes transient, minimal irritation

Nonsensitizer

2

83-1880: Even though protective clothing was worn, applicator developed a rash after application of dimethoate on grapes. He also had nonspecific symptoms of systemic poisoning (nausea and headache).

malathion

121-75-5

Technical material causes transient, minimal irritation

Nonsensitizer in animal studies: 20% product labeled as sensitizer

5

94-401: An employee of a small central valley city was pumping up a spray tank containing malathion when a hose coupling broke, spraying the material on his face and neck. Despite washing immediately, he developed a mild erythema in the exposed areas.

methamidophos

10265-92-6

Labeled as minimal irritant

Nonsensitizer in Buehler test

methidathion

950-37-8

25% formulation nonirritant

Nonsensitizer

87-2288: While disinfecting with an iodine product, an employee developed pruritic rash on arm.

Insecticides Organophosphates

83-1870: An apartment manager was spraying dichlorvos for roach control when the hose broke on his hand sprayer, splashing material on his right arm and left ear. On exam he had severe dermatitis of the right elbow and forearm and left ear, complicated by a possible secondary infection.

84-129: A mixerlloader splashed mixture of methamidophos and buffer on himself while transferring material and developed blisters in the exposed area. 0

(continues)


303





Sensitivity

# of cases

2

Compound

CAS #

naled

300-76-5


Sensitizer

oxydemeton-methyl

302-12-2

53% formulation causes transient irritation

Labeled as sensitizer

83-304: A worker was cleaning the tip of a spray rig when



86-318: A San Diego pet shop employee developed a rash on

3689-24-5



22781-23-3



phosmet

sulfotep

732-11-6

Case examples 88-942: A worker hand poured naled (Dibrom®) for application on strawberries, without wearing rubber boots, gloves, respirator, or eye protection. After spilling the material on his leather boots, he wore them the rest of the day. He developed severe blister on foot, which did not improve with home treatment and eventually required medical attention. 88-2330: A worker was using naled for fly control, when a hose broke, spraying him in the face. He developed a rash on the ears despite wearing coveralls, gloves, respirator, and goggles. At the time of treatment, he was noted to have a chemical contact dermatitis with a secondary infection.

oxydemeton-methyl (Metasystox®) splashed into face underneath the shield he was wearing. He developed erythema at the site of contact. her hands after she began using a phosmet flea dip.

o

Carbamate bendiocarb

82-1278: While treating for cockroaches with bendiocarb, a

hotel employee got his fingers into the material. He later stuck his fingers in his mouth, causing a condition described as a mild allergic reaction to the lips and tongue.

carbaryl

63-25-2

Technical material nonirritant


2

fenoxycarb

72490-01-8

Minimal irritant effects ascertained from 21-day dermal study


o

methiocarb

2032-65-7


o

methomyl

16752-77-5


Nonsensitizer in Buehler test Nonsensitizer in Buehler test

propoxur

114-26-1



4

121-21-1

57% technical material causes transient irritation

Positive Buehler tests for some formulations

10

2

82-2634: A turkey farm employee developed dermatitis after a hose broke during an application of carbaryl. 82-2703: An applicator applying carbaryl dust developed dermatitis after getting the material on his hands and arms.

84-1512: While spraying methomyl on corn, an applicator cleaned clogged nozzles on his equipment with his bare hands and developed a bad rash. 88-297: A structural pest control worker was spraying propoxur, and wiped his hands on shirt. He then developed rash on chest where he wiped his hands.

Pyrethrins

pyrethrins

90-2621: A fairgrounds employee suffered chemical bum to

right leg while applying a pyrethrin insecticide to livestock bams. "The fogger" machine he was using had a loose cap on the reservoir tank causing insecticide concentrate to come in contact with his leg. 86-385: A kitchen employee set off a fogger and remained in the treated area for 10 minutes in violation of and thus was cited (nov) for conflict with labeling "leave area immediately after setting off fogger." He developed skin irritation on his face. (continues)

304

CHAPTER 11



Compound

CAS #




# of Sensitivity

cases

Case examples

Pyrethrin synergists piperonyl butoxide

51-03-6

92% formulation category II irritant

Inadequate-studies on mixtures with pyrethrins

o

Listed in registry data only as a component of mixtures.

n-octylbicycloheptenedicarboximide

113-48-4

20% concentration, mixed with pyrethrins causes transient irritation

Studies with mixtures only

o

Listed in registry data only as a component of mixtures.

allethrin

584-79-2 42534-61-2 (cis/trans allethrin)


Nonsensitizer

bifenthrin

82657-04-3


Nonsensitizer

2

94-1052: An applicator bumped his right arm against a spray nozzle and got some of the bifenthrin spray solution on the arm and then developed numbness and tingling in the right arm. 95-1210: A mixerlloader, employed by a professional agricultural pest control company to treat cotton, splashed bifenthrin on his arms, face and eyes while transferring product from a closed system holding tank into a I-gallon container. The container overfilled and the pressure created forced the product out. He develped redness, and a burning sensation on the face, chest, and shoulder.

permethrin

52645-53-1

Technical material category III irritant

sensitizer

3

92-1381: Worker was mixing material and small amounts kept getting under gloves and shirt. A dermatitis problem developed as a result. Pain, swelling, and blisters on hands and forearms.

phenothrin

26002-80-2


mixture with tetramethrin sensitizer in Buehler study

o

resmethrin

10453-86-8

85% technical material is category III irritant

3% formulation weak sensitizer in Buehler study

o

tetramethrin

2117279

21 % mixed with 21 % resmethrin causes transient irritation

mixture with phenothrin is a weak sensitizer in Buehler test

o

cyfluthrin

68359-27-5



o

cyhalothrin

91465-08-6 68085-85-8


Inadequate data

o

cypermethrin

52315-07-5


Nonsensitizer in animal tests; 40% wettable poweder labeled as a sensitizer

Type I pyrethroids 94-138: A worker spilled a mixture of pip butoxide and allethrin on the outside of a backpack sprayer and his arms while mixing a tank load and wiped off the sprayer with a paper towel. After applying the material, he noticed itching, redness, and swelling of the lower back and arms.

Type II pyrethroids

88-2388: A mixerlloader handling cypermethrin developed burning in the groin area, shortly after going to the bathroom without thoroughly washing his hands.

(continues)


305





# of cases

Sensitivity

Compound

CAS #

esfenvalerate

66230-04-4

35% formulaton causes transient, mild irritation

Nonsensitizer by Buehler method

0

fenpropathrin

395151-41-8

Technical material nonirritant based upon observations in acute dermal toxicity study

Weak sensitizer in human repeated insult patch test

0

fenvalerate

51630-58-1

Category II irritant-persistent erythema at day 7

3.5% formulation nonsensitizer in Buehler test; 25% formulation labeled as a sensitizer

dicofol

115-32-2


Sensitizer

3

84-954: An employee was mixing dicofol for an aerial application on corn, wearing gloves and face shield, when some material splashed up on his neck. He saw doctor 3 days later when the burning and itching on the front of neck did not improve after initial treatment with first aid ointment. The condition was recorded as a second degree burn. 84-1454: A mixerlloader/applicator splashed dicofol on the arms that soaked through his protective clothing. Erythema of the forearm was noted when he sought treatment 3 days later.

lindane

58-89-9

No data

Borerlleaf miner formulation is a sensitizer in the Buehler assay.

2

86-309: A hose split during an application under a house and the material sprayed onto the applicator's hands. He made repairs without gloves, washed off, but noticed a rash later in the day.

methoxychlor

72-43-5

25% formulation labeled as minimal irritant

No data

0

borax

1303-96-4

99% formulation labeled as nonirritant

Inadequate data

disodium octaborate tetrahydrate

3791278

Labeled as a nonirritant

Nonsensitizer

Case examples

86-1191: Concentrated pydrin® spilled from a measuring cup onto a mixerlloader's arm, but he did not wash or change clothes. A contact burn developed at the site, as recorded by the treating physician.

Organochlorines

Borates 84-197: A restaurant employee applying boric acid as a crack and crevice treatment, suffered a reported mild skin reaction.

Bacillus insecticides & other biologicals Azadiracthin-from Neem extract

1141-17-6

Nonirritant

Nonsensitizer

0

avermectin (abamectin)-mixture of avermectin Bla and BIb

71751-41-2

Nonirritant


2

Bacillus thuringiensis

No CAS #

Nonirritant


0

Capsicum oleoresin

404-86-4

Nonirritant

Labeled as a sensitizer

0

92-520: An employee developed skin problem on arm after spraying roses with Abamectin. He was wearing a rubber rainsuit, but felt wetness on his arm, and did not wash the affected area immediately. Examination showed ulcerative lesions with mild surrounding erythema on right proximal forearm. 92-2243: A worker developed a rash on his right hand while applying abamectin to roses. He developed a similar rash the previous year after spraying the same pesticide.

(continues)

306

CHAPTER 11



Compound

CAS #




Sensitivity

# of cases

Case examples

Miscellaneous chemical structure amitraz

33089-61-1

50% wettable powder labeled as corrosive

50% EC sensitizer in maximization study

o

benzyl benzoate

120-51-4



o

butoxy polypropylene glycol

9003-13-8

Category III irritant

Inadequate

o

Diethyl toluamide (DEET)

134-62-3



3

hydramethylnon (aminohydrazine) bait

67485-29-4

Minimal irritant

Nonsensitizer

o

Imidacloprid

105827-78-9

Minimal irritant

Nonsensitizer

k salts of fatty acids (1596)

61790-44-1

49% formulation Irritant/corrosive

Inadequate

o o

oxythioquinox

2439-01-2


40% fomulation sensitizer in Buehler test

propargite

2312-35-8

Corrosive

Nonsensitizer

82-1871: Worker had an allergic reaction (hives) after treating himself with an insect repellent according to the label directions. 93-1422: Worker sprayed an insect repellent on her exposed skin before collecting a lab sample from treated sewage water. She suffered an apparent allergic reaction to the repellent a short time later.

Details of case not specified

55

82-1667: A mixerlloader splashed a few drops of concentrated propargite on his neck while opening a can for closed system loading. Over the following day, he developed a rash, which persisted for a week until he got it treated. 85-1642: During a mixlload operation, the nurse tank overflowed, dousing worker's arm with material and a chemical burn subsequently developed at the site of contact.

sulfluramid (bait)

o

4151-50-2

Nonirritant


Captafol

2939-80-2

Labeled as irritant


captan

133-06-2



6

folpet

133-07-3


Sensitizer in the maximization test

o

17804-35-2



6

Fungicides

Phthalimido compounds 83-1783: An applicator developed a rash after using difolatan on tomatoes. He did not seek treatment until after the rash became infected. 84-1668: A worker was loading captan dust into a helicopter for application on grapes when he developed contact dermatitis. 87-260: After spraying captan using adequate protective equipment, an orchard applicator developed a rash on face, neck, and arms, thought to be an an allergic reaction. 87-694: A worker removed lids from 5 gal containers of captan, wearing gloves and goggles. He developed a rash, but did not get any of the material on his skin via splash or spill.

Carbamates benomyl

85-448: An applicator splashed benomyl onto his face and neck while spraying pruning cuts in vineyard. He suffered burning, erythema, irritation, and swelling of eyes and face. The condition was described as a first degree chemical burn.

(continues)


307

Table 11.1 (continued) Predictive tests in animals



Sensitivity

# of cases

Compound

CAS #

thiophanate methyl

7912735


50% formulation labeled as sensitizer; others not

mancozeb

8018-01-7


Sensitizer in Buehler study in combination with thiophanate

88-1764: During hot weather, mixerlloader developed a rash after he started handling Dithane®. Examination showed a rash on the arms and abdomen characterized as contact dermatitis.

maneb

12427-38-2

No data

Sensitizer

84-811: An employee developed an allergic rash after spraying dithane on grapes

thiram

137-26-8

Technical material category IV


84-1488: A worker treating seeds with thiram dust developed lesions around the respirator line on the day of the application.

zineb

12122-67-7

Inadequate


ziram

137-30-4

96% industrial formulation labeled as an irritant


2

copper

7440-50-8

85% formulation category III irritant


2

copper ammonium carbonate

33113-08-5

24% formulation minimal irritant


0

copper hydroxide

20427-59-2


Labeled as sensitizer; Buehler study on CuOH negative

3

94-561: A worker applied copper hydroxide to walnuts using a high-volume sprayer. His face began to itch and bum a few hours after he finished the application. The affected area was on the unprotected portion of his face. 90-1368: A worker applying fungicide to nut orchard, wearing all protective gear, got wet from rain blowing in around his hood and down his gloves. He then began itching in areas that had gotten wet.

copper naphthenate

1338-02-9

80% formulation category II irritant

68% CuNaphthenate Nonsensitizer in Buehler test

2

87-1724: A wood worker was painting a copper naphthenate wood preservative to the cut end of wood and developed a chemical bum to his arms. 90-1127: A student employee wearing rubber boots, gloves, goggles, respirator, and ransuit treated wooden benches with preservative. He accidentiy rubbed his neck and face while wearing the rubber gloves and developed contact dermatitis.

copper oxide

1317-39-1

97% technical material nonirritant

Animal tests do not show clear evidence of sensitization

0

copper oxychloride

1332-40-7



0

copper sulfate

7758-98-7 7768-98-7

99% technical material minimal irritant


4

Case examples

0

Thiocarbamates

83-298: A worker applied ziram with no hand or face protection and developed contact dermatitis. 84-518: An applicator spraying almonds with ziram developed a rash after a hose broke on his spray equipment.

Copper compounds

90-2391: After adding copper sulfate to water, a worker developed a rash on his forearms and itching all over. 90-2588: While an employee mixed copper sulfate, some powder got inside the glove causing the irritation to his right forearm. The resulting dermatitis was subsequently complicated by an infection. 93-1839: A worker applied copper sulfate granules to canal water by a piece of equipment he called the "sandblaster"-that air blasts the material on the canal water. After copper sulfate dust landed on his neck and chest, he developed large pruritic, erythematous patches on the neck and chest.

(continues)

308

CHAPTER 11




Sensitivity

# of cases

Compound

CAS #


copper triethanolamine complex

68027-59-6 82027-59-6

3.5% formulation minimal irritant

Nonsensitizer

o

cupric oxide

1317-39-1

97% technical minimal irritant

Buehler method nonsensitizer

o

cuprous oxide

1317-38-0

85.7% category III irritant

Buehler method nonsensitizer

o

Anilazine

101-05-3

No data


2

carboxin

5234-68-4

Technical material is nonirritant

Mixture carboxin and other compoundsnonsensitizers

0

chloroneb

2675-77-6

30% formulation with 3.5% metalaxyl category III irritant

Negative test on mixture

0

fIusilazole

85509-19-9

60.7% formulation category III irritant

Buehler test on 20% formulation negative

0

fosetyl-al

39148-24-8


Buehler test on 80% formulation negative

0

imazaIiI

3554-44-0

Technical material 98.5% causes transient irritation

Buehler test on 13.5% formulation negative; technical material is labeled as sensitizer

0

iprodione

36734-19-7

50% material nonirritant

Nonsensitizer in the Buehler assay

methylene bis(thiocyanate)

6317- 18-6

10% material corrosive


mycIobutanil

88671-89-0

84.5% technical minimal irritant

Nonsensitizer

Pentachloronitrobenzene (PCNB)

82-68-8

No data

No data

0

sulfur

7704-34-89

98% formulation category IV

Nonsensitizer by Buehler method

28

TCMTB (thiocyanomethylthiobenzothiazole)

21564-17-0

30% formulation corrosive

10% material labeled as sensitizer

0

Case examples

Miscellaneous compounds 92-732: A mixer-loader for an aerial application developed rash on exposed skin areas while dumping wettable anilazine powder into mix tank. At examination, he had a generalized rash on face, neck, and arms thought to be allergic in nature.

84-897: A worker sprayed iprodione on grapes and developed a rash 3 days afterwards. He also worked in a thinning crew after app. No other crew members developed a similar rash. 0 92-1379: A worker overfilled a spray tank and spilled pesticide solution (mycIobutanil + adjuvant) on his feet. He rinsed his feet and shoes off with water, but did not remove his shoes. Examination showed pruritic dermatitis, scaling, and crusting of the bottoms of both feet.

82-1211: A worker's coveralls became complete while he applied sulfur and he developed dermatitis. 83-1252: An employee tore pants while applying sulfur to cotton, and got a rash in the area of the tear. 85-2263: A worker developed a rash on body, especially on neck, while dusting with sulfur. 87-174: A worker complained of rash after mixing, loading and applying Kolospray@ and had a 2 year history of sensitivity to the formulation.

(continues)


309




# of cases

Sensitivity

Case examples

Compound

CAS #

thiabendazole

148-79-8

Transient irritation with 98.5% formulation

Tests on mixtures only

triadimefon

4312173-43-3


95% technical material is a sensitizer in the Buehler test

vincIozolin

50471-44-8

Transient irritation


0

1,3 dichloropropene

542-75-6

Corrosive


11

aluminum phosphide

20859-73-8

No data

No data

84-2184: A rash developed on torso after application of Phostoxin® under tarp for rice in warehouse. There was no history of direct exposure.

dazomet

533-74-4

Technical material is a minimal irritant


90-2448: Applying dusty granular form of pesticide, a worker developed a rash at the belt-line as weII as front of legs, abdomen, and arms.

ethylene oxide

75-21-8

No data required as minimal dermal contact expected

Some products labeled as sensitizers

5

fosthiozate (2325)

not specified

Nonirritant

Technical material is sensitizer in the maximization test

0

metam-sodium

137-42-8

Corrosive

? Animal skin sensitizer

34

84-20: A worker handled vapam while wearing leather boots; the material soaked through the boots and irritated his feet. 85-1697: After making an application, he noticed pressure remaining in the hosel; the material splashed on him and skin began to bum.

methyl bromide

74-83-9

Corrosive

No data

59

82-32: Methyl bromide leaked into employee's boots when he changed cylinders on a tractor. His feet became inflamed, ulcerated, and infected over several days before employer noticed problem and sent the worker for treatment.

diquat

85-00-7

Irritant

No data

15

86-1498: An applicator wet his shoes with diquat and did not change them. He developed a rash on the top of one foot. By the time he saw a doctor, 8 days later, his foot had become infected.

paraquat

1910-42-5 50-2

Irritant


17

83-480: A gust of wind blew material onto arms which had previous abrasions and his condition was aggravated by contact with paraquat.

0

82-1412: While applying BayJeton® to grapes, the wind blew spray back on to the applicator. He suffered a reported aIIergic reaction.

Fumigants 88-2091: An employee was uncIogging one of the tubes through which the telone is injected on his spray rig when material spiIIed on his foot. He continued to work and the next day he had blistering and sweIIing offoot.

87-2720: A hospital worker stuck her hand in a sterilizer, before it had aerated to get rid of the ethylene oxide and suffered a chemical bum on her hand.

Herbicides Bipiridyls

Chloracetanilide alachlor

15972-60-8

metolachlor

51218-45-2

Sens per precaution

85% formulation category 3 irritant

Mixed formulations negative in Buehler test

84·537: An applicator developed a fine rash on trunk, arms, and legs two years in a row after handling alachlor. Condition was reported as suspected aIIergic dermatitis. 0

(continues)

310

CHAPTER 11



Compound

CAS #




# of cases

Sensitivity

Case examples

Nitroaniline compounds

benefin

1861-40-1

Technical material cat 2 irritant

Mixtures with other herbicides labeled as sensitizer

ethalfluralin

55283-68-6

Technical material cat 2 irritant


oryzalin

19044-88-3


75% formulation labeled as sensitizer

5

84-51: The wind blew material into an applicator's face, resulting in a rash on neck and face. 84-272: An applicator wiped his face with a wet glove, and developed a rash immediately.

pendimethalin

40487-42-1

2.5% formulation minimal irritant

Nonsensitizer

2

82-467: A mixerlloader/applicator developed a rash while applying pendiemethalin. Examination showed a rash on both arms up to elbow.

trifluralin

578064

80% formulation caused minimal, transient irritation


7

93-340: An applicator was loading his tractor with trifluralin and some of the material leaked out and was blown onto his face. He failed to wash the exposed area right away and developed developed an itching, burning, red rash on face.

bensulide

741-58-2

Nonirritant

Inadequate

glyphosate

1071-83-6

Minimal irritant

Nonsensitizer

70

sulfosate

81591-81-3


Sensitizer

0

tribufos

78-48-8

70% formulation labeled as corrosive; technical material category III

No data

0

94-75-7



0

dicamba

1918-00-9

Minimal irritant

Nonsensitizer

0

MCPP

7085-19-01

Minimal irritant

Nonsensitizer

0

fenoxaprop ethyl

66441-23-4

Irritant

Inadequate data

0

dithiopyr

97886-45-8

91.5% technical

12.7% material sensitizer in Buehler test

0

imazethapyr

81335-77-5

90% formulation category 4 in Draize

22.9% formulation negative in Buehler study

0

triclopyr

5721-4069-1

61.6% formulation causes transient irritation in the Draize test

Buehler study on 17% formulation was negative. However, the 44% formulation is labeled as sensitizer

2

0

90-1832: An applicator disconnected a filter valve, was sprayed in the face with herbicide, and developed pruritus.

Organophosphates 83-1770: A hose ruptured and sprayed his arm. A rash developed, reported to cause 20 days lost-work-time. 83-220: The wind blew spray mist onto a worker's forehead while he was applying a glyphosate formulation. He experienced a rash and itching at the site of contact that lasted for for several days. 83-917: As a worker was treating vineyard weeds with a glyphosate formulation, a hose burst on his backpack sprayer, covering his back with the material and he subsequently developed a rash at the site of contact.

Phenoxy herbicides

2,4-D

Pyridine derivatives

88-2832: The worker was spraying weeds along the roadside when a big rig drove by causing a shift in the wind direction and the spray blew back in her face. She developed mild erythema on the face.

(continues)


311


Compound

CAS #




# of cases

Sensitivity

Case examples

Thiocarbamates and carbamates

molinate

2212-67-1

thiobencarb

15% granular flake formulation was nonirritating in Draize test, but caused transient irritation when moistened prior to application

10% formulation was a nonsensitizer in the Buehler assay

84% material caused very minimal and transient irritation in Draize test

84% material is a nonsensitizer in Buehler test

87-937: An employee was loading Ordram® bags into the bucket of the loader truck for an aerial application. Some of the material got down his protective clothing, contacting his legs and feet and he subsequently developed a rash in the corresponding areas.

Triazines

atrazine

1912-24-9

No data

No data

0

cyanazine

21725-46-2

Technical material caused transient erythema in the Draize test

A Buehler study on a mixture of cyanzine and metolachlor was negative

0

prometryn

7287-19-6

Technical nonirritant in Draize

45% material minimal reaction in Buehler test

0

simazine

122-34-9


6.3% simazine negative in max. test

314-40-9

80% formulation category 4 Draize

Nonsensitizer in Buehler test 40% form, mixed with 40% diuron

0

83-2394: Mixlloading material, apparently urinated during operation, depositing material on penis. Developed rash.

Urea herbicides

bromacil

chlorsulfuron

64902-72-3

75% minimal irritant

No data

0

diuron

330-54-1

81 % formulation category III irritant


0

thidiazuron

51707-55-2

12% formulation with 6% diuron corrosive

Ginstar® sensitizer in Buehler test

0

dichlobenil

1194-65-6

Technical material category 4

0.55% root control sensitizer in maximization test

0

flumetsulam

98967-40-9

Mixture with metolachlor category 2 irritant

Mixture labeled as sensitizer

0

isoxaben

82558-50-7

Nonirritant

Nonsensitizer in Buehler assay

0

sethoxydim

74051-80-2

13% formulation category 11 irritant

13% formulation nonsensitizer in maximization test

polymerized pinene

None given

Summary of study describes material as slightly irritating

Nonsensitizer in human repeated insult patch test

0

Stephan C-65

Not given

Irritant

Nonsensitizer

0

Miscellaneous

88-1253: A worker was pumping up sprayer when leaky gasket caused solution to spray in face and he developed a dermatitis characterized as a chemical bum.

Adjuvants

312

CHAPTER 11


Sodium hypochlorite

Emmett et al., 1989; Foussereau et al., 1984; Mathias et al., 1983; Menne, 1991; Menne et al., 1991; Pilger et al., 1986; Thormann, 1982). In the pesticide handler database, the isothiazolin compounds appear to be relatively frequent causes of irritant skin reactions or frank chemical bums following accidental direct contact (e.g., cases 89-1312 and 90-552). The data are limited, however, because the individual isothiazolin compounds involved in each case are not identified.

JN~ Ld Octhilinone

li

N~N

O~9~OH Cyanuric acid

o

6

N

",,","""';=1;00" hexahydro-1,3,5-triethyl-s-triazine

N -

Cl

OH

Na-OCI

Kat~on DP, LX and other trade names

©IN

11.2.1.4 Phenolic Compounds Promexal X50 preservative

1,2-benzisothiazol-3(2H)-one Proxel

11.2.1.2 Quarternary Amines The quartemary amines have a capacity akin to isothiazolin compounds to cause both irritation and sensitization in animal test models. They are also often used in mixtures and are grouped together in the handler database. The 107 cases associated with quartemary amines accounted for 11.8% of the total number reported. Typical cases (88-1893) occurred on direct contact.

A review of the DPR label database showed 128 separate phenol compounds with DPR chemical codes, including sodium phenate salts of several phenol derivatives. O-phenyl phenol is a typical compound (Table 11.1), identified as an irritant in the Draize test, but no sensitization test was available for review. Its sodium salt has been identified as a cause of contact urticaria (Tuer et al., 1986). P-tert-butyl phenol is notable for being identified as a human sensitizer and as an occasional cause of leukoderma (Mancuso et al., 1996; O'Malley et al., 1988).

OH

~CH3)2-N-(C8HlO)~ + Cl Dioctyl dimethyl ammonium chloride

ortho-phenylphenol (CH3 h - N Cl

~.5H3 HO~CH3 CH 3 p-tert-butyl phenol

tXHx+2 x> 14

Alkyl dimethyl benzyl ammonium chloride

QCH3)2-N-(ClOH12~ +CI didecyl dimethyl ammonium chloride

11.2.1.3 Chlorine Compounds and Triazine Chlorine Stabilizers The animal data on sodium hypochlorite show irritant effects (Table 11.1), and sensitization has been reported in the form of contact urticaria (Hostynek et al., 1989). Nevertheless, most sodium hypochlorite-containing antimicrobials are not labeled as sensitizers. The cases reported in the handler database occurred principally in end users of sanitizers and disinfectants (e.g., 87-1468). Cyanuric acid was associated with several cases in the handler database. The other chlorine stabilizer hexahydro-l ,3,5-triethyl-s-triazine ([7779-27 -3]) was corrosive in the Draize test and a weak sensitizer in the Buehler assay.

11.2.1.5 Other Antimicrobial Compounds Among the miscellaneous antimicrobial compounds, only iodine was associated with a case in the handler database. Nevertheless, many of the compounds, (2,2-dibromo-3-nitrilpropionamide), 2-(hydroxymethy lamino )-ethanol, 2-bromo-2nitro-1,3-propanediol, and sodium pyrithione, appear to be corrosive or severely irritant. Bronopol has long been recognized as a sensitizer (Camarasa, 1986; Frosch et al., 1990), related to its capacity for releasing formaldehyde (Kranke et al., 1996); 1,2-dibromo-2,4-dicyanobutane (Tosti et a!., 1995; Vigan et a!., 1996) and iodine (Ancona et al., 1985; Erdmann et al., 1999), have also been associated with cases of contact sensitization. Organotin compounds have been reported as irritants (Gammeltoft, 1978). Allergenic effects of tributyl tin compounds generally are probably not significant, but tributyltin methacrylate is an equivocal sensitizer in the Buehler test (Table 11.1).


o 11

CHBr2-C-NH-CN

313

dler database appeared to be instances of mild irritation (e.g., 85-343), consistent with the transient irritation occurring in animal tests of technical material.

2,2-dibromo-3-nitril-propionamide OH 1

N-CH 3 1

CH 2-CH 2-OH

2-(hydroxymethylamino)-ethanol Br 1

HO-C-C-C-OH 1

N02 2-Bromo-2-nitropropane-1,3-diol Bronopol

Naled (Dibrom®) is labeled as a corrosive and also as a sensitizer (Table 11.1), presumably because of the cases of contact sensitivity (Edmundson and Davies, 1967) and irritation (Mick et aI., 1970) reported in the public domain literature. Cases of irritant dermatitis similar to those described by Mick were also found in the handler database (88-2330 and 88-942).

Sodium 1-hydroxypyridine-2-thione Omadine

1,2-Dibromo-2,4-dicyanobutane

11.2.2 INSECTICIDES AND INSECT REPELLANTS 11.2.2.1 Organophosphates The organophosphates are generally thought to cause minimal irritation (Rycroft, 1977) and are nonsensitizers in animal models (Table 11.1). Nevertheless, some organophosphates, such as parathion and malathion, have previously been reported as causes of contact sensitization or other dermatoses. Details of the animal studies and pesticide applicator cases are discussed for compounds which have been possibly associated with significant skin effects.

Concentrated forms of chlorpyrifos (e.g., 40.7% formulation of Lorsban) are irritants in animal studies, but all of the numerous formulated products containing chlorpyrifos are negative in the animal sensitization studies conducted by the Buehler method. Paradoxically, the technical formulation with 98% chlorpyrifos causes less irritation than the 40.7% formulation. The 11 cases associated with chlorpyrifos in the pesticide han-

Although a report by Milby identified malathion as a sensitizer, animal sensitization studies submitted for pesticide registration did not employ adequate negative controls to verify that the dermal reactions observed were due to sensitization rather than irritation (Milby and Epstein, 1964). Nevertheless, the irritation studies conducted, per se, showed that technical malathion causes minimal irritation. Cases associated with malathion in the handler database are consistent with mild, transient irritation. Although diazinon appears to be minimally irritant, in animal tests, cases in the handler database are consistent with an irritant mechanism (87-2537). No sensitization studies on diazinon were reviewed, but most of the diazinon products reviewed were labeled as sensitizers, as required in the V.S. EPA's diazinon reregistration standard. A case report from Australia identified an isomer of diazinon as a cause of porphyria cutanea tarda (Collins et aI., 1982).

314

CHAPTER 11


11.2.2.2 Carbamates None of the carbamate insecticides appeared to be markedly irritant or to be sensitizers in animal test models. One carbaryl formulation containing pyrethrins was found to be a sensitizer, but none of the formulations containing only carbaryl produced similar responses. Cases associated with carbamates in the handler database appeared to be principally transient irritation following direct contact (e.g., 88-297, 84-512). The transient irritation reported in the applicator cases may be due to nonpesticidal ingredients: Draize tests on carbaryl, methiocarb, methomyl, and propoxur did not show even transient irritation. Cases of dermatitis associated with application of carbamates are also reported in the public domain scientific literature (Bruynzeel, 1991; Vandekar, 1965).

o 0-8-NH-CH 3

00 Carbaryl

CH3-C

a category 11 irritant in the Draize test, but the concentration in ready-to-use formulations is less than 1%. The only dermal sensitization studies reported involve mixtures with pyrethroids. Cases reported in the handler database also had simultaneous exposure to pyrethrins or pyrethroids.

N-octyl-bicycloheptene dicarboximide (NOBD) is also used as a pyrethrin synergist in hundreds of formulations. The only irritation study available for review involved a mixture with diethyl toluamide, isochromyl cinchonerate. This mixture caused only transient irritation and was scored as category IV. There is no study available on the capacity of NOBD to cause sensitization. Cases reported to the handler database all involved mixtures with pyrethrins and other compounds.

0

=N-O-~-NH-CH3

~-CH3

Methomyl

11.2.2.3 Pyrethrins

n-octylbicyclJ"heptenedicarboximide

The California Pesticide Label Database shows more than 1000 pyrethrin formulations registered (2/98). The cases associated with pyrethrins reflect their broad-scale use, nearly always in mixtures with piperonyl butoxide, and often with pyrethroids, carbamates, organophosphates, and other materials. Many formulations containing pyrethrins are positive in the Buehler test, and this is reflected in the product labeling. Dermal irritation studies with technical pyrethrins (57% concentration) show only transient erythema, disappearing by 72 hours. Irritation suffered by users (Table 11.1) may be due to petroleum distillates or piperonyl butoxide (92% formulation is a grade 11 irritant in the Draize irritation test) contained in formulated products, as well as the pyrethrinlpiperonyl butoxide combination. C~

Jc-8-o--Q "'" C~

'C=CH CH!

C~ Pyrethrin I

0

C~CH=CHCH=C~

C~ 0

11.2.2.4 Pyrethrin Synergists Piperonyl butoxide is a synergist used with formulations of pyrethrins and pyrethroids, but is not chemically related to either group. The technical formulation (92%) was classified as

11.2.2.5 Synthetic Pyrethroids Effects of synthetic pyrethroids on the sodium channels of cutaneous nerve endings may cause paresthesias at levels of exposure that do not provoke visible erythema (Lisi, 1992). The standard Draize irritation study may be a poor means for evaluating such purely symptomatic endpoints. An alternative animal test developed by Cagen evaluates the sensory effect of pyrethroids through observations of grooming behavior focused on the site(s) of applied test material. The behavioral test demonstrated direct effects on grooming behavior for 4 hours after pyrethroid application and increased response to other chemical irritants (oil of mustard in the test model) for 24 hours after application (Cagen et aI., 1984).

Type I Pyrethroids These contain two cisltrans isomeric sites and may have as many as four isomers with ability to stimulate cutaneous nerves in the human epidermis (Flannigan and Tucker, 1985; Flanniganet aI., 1985a, b; Gammon, 1985; Gammon and Casida, 1983; Tucker et aI., 1984). Allethrin Technical d-allethrin is a cis-trans mixture that shows minimal irritation in the Draize test; a Buehler sensitization study was negative on a dilute end-use product (a mixture

11.2 Review of Use Categories of allethrin, cypermethrin, piperonyl butoxide, and petroleum distillates). Case 94-138 in the handler database (a formulation of allethrin and piperonyl butoxide) involved irritation on direct accidental contact.

315

Type 11 Pyrethroids These contain as many as three isomeric sites and most contain a cyano group attached near the ester linkage. They are relatively more potent systemic toxins than the type I pyrethroids and cause a greater degree of paresthesia in experimental studies on human volunteers. Cyfluthrin and Cyhalothrin Cyfluthrin and cyhalothrin both caused minimal irritation in the Draize test and were nonsensitizers in the Buehler assay. Neither was associated with cases in the handler database.

Permethrin Technical permethrin shows minimal irritation in the Draize test, with only very slight erythema persisting 72 hours after initial application. Several formulations appeared to be nonsensitizers in the Buehler assay, but a mixture of 1% permethrin and 1% piperonyl butoxide showed mild sensitization reaction on the rechallenge portion of the assay. Case 92-1381 in the handler database, associated with permethrin, appeared to be due to cumulative irritation.

Phenothrin and Tetramethrin A mixture of 5% phenothrin and 5% tetramethrin proved corrosive in Draize tests, according to the product labeling. Nevertheless, technical phenothrin caused no irritation in the same assay. Sensitization tests were negative except for a mixture of phenothrin and tetramethrin, which showed marginal reaction on rechallenge in the Buehler assay. No cases were included in the handler database.

Resmethrin Technical resmethrin caused mild persistent irritation in the Draize assay. A 3% formulation was also a sensitizer in the Buehler assay. No cases were included in the handler database.

--cr°

C=Ck-0 cH{ 1 C-o-CHz CfiJ

CfiJ

C~ Cl

~

CN

~CH-C~2cH-t-o-6H~

~~ 0

CH 3/ 'CH 3

0

A-cyhalothrin

Cypermethrin Technical cypermethrin causes minimal irritation in the Draize test. The 40% wettable powder is labeled as a sensitizer, but the Buehler assay on the 18% formulation of cypermethrin showed no evidence of sensitization. A case of dermatitis in the handler database was associated with accidental transfer of cypermethrin from the hands to the genitalia (88-2388).

o

CN

0

:~CH-C~~CH-~-O-CH~ ~ CHI 'CH3 Cypermethrin

Fenvalerate The 24% formulation is an irritant (category 11) in the Draize test and also a sensitizer in the Buehler assay. Case 86-1191 in the handler database was described as a bum in a patient who spilled fenvalerate on himself and did not promptly decontaminate his clothing.

Tetramethrin

CfiJ,

Cyfluthrin

CHz

~

Resmethrin

Fenvalerate

11.2.2.6 Organochlorines

-1

Thiophanate methyl

11.2.3.3 Thiocarbamates The thiocarbamate group of fungicides structurally resembles the rubber accelerator disulfiram (Antabuse®, tetraethylthiuram disulfide-CAS # 97-77-8), a common sensitizer present in both the European and the North American standard patch test series (Adams and Fischer, 1990). The prototype thiocarbamate fungicide, thiram (thiuram) is simply the methyl analog of disulfiram, and experimentally (Freundt and Netz, 1977) has a similar effect on the metabolism of alcohol. Technical formulations of thiram do not cause irritation in the Draize test, but the thiram case reported to the handler database was consistent with an irritant mechanism (84-1488). Most thiram formulations are labeled as potential sensitizers, because of reported cases of sensitization in the clinical literature (Cronin, 1980; Schultz and Hermann, 1958; Shelley, 1964).

Cases of allergic reactions to maneb documented with provocation (patch) testing have been reported in the clinical literature. Typical cases described from the Netherlands included two office workers who had purchased maneb spray to care for the plants in their office, and a 51-year-old woman who worked as an assistant in a flower shop (Nater et ai., 1979). Similar cases have been reported from the United States (Adams and Manchester, 1982), Italy (Peluso et ai., 1991), and Germany (Koch, 1996). A case reported to the handler database (Table 11.1) involved a possible allergic reaction after spraying a maneb-containing formulation of dithane® (84-811). Mancozeb is a polymer similar in structure to zineb and maneb, containing both zinc and manganese. The Draize irritation study reported to the registration database for the 80% formulation of mancozeb showed a minimal irritant reaction, and a combined formulation of mancozeb and thiophanate methyl was a sensitizer in the Buehler assay. A mancozeb case (88-1764) reported to the handler database was compatible with a simple irritant mechanism, but a recent case reported from Japan identified mancozeb as a cause of allergic contact dermatitis and photodermatitis (Higo et ai., 1996).

S

CH2-NH-~-S,-...----Mn/Zn

Thiram

t

The structure of ziram is very similar to that of thiram, but the compound contains a zinc atom between the two atoms of sulfur. It is also similar to zineb, which is a zinc/thiocarbamate polymer, and to the manganese/thiocarbamate polymer, maneb. A 96% formulation of ziram used as an industrial biocide was labeled as both an irritant and a sensitizer, but no animal data were available for review. Both ziram-associated cases (83-298, 84-518) reported to the handler database were consistent with an irritant mechanism.

CH3

~N-C-S-Z~S-C-~

CHi

~

~

"

CH3 CH

3

Ziram

Matsushita tested maneb and zineb experimentally with the guinea pig maximization procedure and found both compounds to be potent sensitizers with a high degree of mutual crossreactivity. Concentrations of 5% or more were found irritating

CH2-NH-C-S Mancozeb

\\ S

11.2.3.4 Copper Fungicides Copper compounds are used as both fungicides and antimicrobial agents. Copper has been identified as a sensitizer in the public domain literature based on human case reports (Rademaker, 1998; Verhagen, 1974), and most of the copper fungicides are labeled as potential sensitizers even where there are negative animal sensitization studies (e.g., copper hydroxide). Elemental copper, copper naphthenate, and cuprous oxide are irritant, but not corrosive, in the Draize test. Technical formulations of other compounds (copper ammonium carbonate, copper hydroxide, copper oxide, copper sulfide, and cupric oxide) cause only minimal irritation in animal tests. In the handler database, elemental copper, copper hydroxide, copper naphthenate, and copper sulfate were associated with cases of contact dermatitis consistent with dermal irritation (Table 11.1).


321

Copper naphthenate was also associated with one case (871724) described as a chemical bum.

11.2.3.5 Fungicides with Miscellaneous Structures Anilazine is a foliar and turf fungicide that has not been registered in California since 1990, but is currently being used elsewhere in the United States. Data on dermal irritation and sensitization from animal studies were not available for review, but the product has been reported as a human sensitizer in tomato harvesters (Schuman and Dobson, 1985; Schuman et aI., 1980) and in lawn care workers (Mathias, 1997). The sample case from the handler database described in Table 11.1 (92-732) was suspected to be caused by an allergic reaction, but a patch test was not carried out.

Carboxin is a systemic fungicide and seed protectant. It is a nonirritant in the Draize test. A mixture containing carboxin (15%), PCNB (15%), and metalaxyl (3.12%) was a nonsensitizer in the Buehler assay.

Chloroneb is a fungicide used for control of seedling diseases in beans, cotton, and soy beans. A formulation containing 30% chloroneb and 3.5% metalaxyl is a category III irritant in the Draize test. The same mixture was a nonsensitizer in the Buehler assay. There were no cases associated with either chloroneb or carboxin in the handler database. Cl

o

OCH 3

*

CH30

I

Chloroneb

Flusilazole is an organosilicon compound, formulated as dry granules and as an emulsifiable concentrate, used for control of ascomycetes and other fungi on cereals, fruits, and vegetables. A 60.7% formulation is an irritant in the Draize test, but a 20% formulation was a nonsensitizer in the Buehler assay. Fosetyl-aluminum is an aluminum salt of an organic acid, used as a systemic fungicide and bactericide. It is active against Oomycetes, Alternaria, and Penicillium on avocado, strawberries, and other crops. It is formulated as a water dispersable granule, as a liquid wettable powder, and as a liquid injectable.

The 80% formulation is a minimal irritant in the Draize test and a nonsensitizer in the Buehler assay. Imazalil (eni1conazole) is a systemic fungicide active against benzimidazole-resistant strains of fungi, formulated as an emulsifiable concentrate, as a water soluble powder, and as a soluble liquid. It is a nonirritant in the Draize test. Although the Buehler assay was negative on a 13.5% formulation, cases of sensitivity to the compound have been previously described in Europe (van Hecke and de Vos, 1983) and in Central America (Penagos, 1993). The technical material is labeled as a potential sensitizer.

There were no cases associated with ftusilazole, fosetylaluminum, or imazalil in the handler database. Iprodione is both a contact and a systemic fungicide used on a broad spectrum of crops. The 41.6% ftowable formulation was a nonsensitizer in the Buehler assay. The 50% formulation was a nonirritant in the Draize test; however, the iprodioneassociated case reported in the handler database (84-897) was compatible with an irritant mechanism following direct contact. CH(CH 3 h

NH\c

If ~CI

i f 'N----'\

Iprodione

~N o

0 Cl

Methylene bis(thiocyanate) is a reactive compound, used as a bactericide in water-cooling systems and pulp and paper mill operations. It is also used as a fungicide in wood preservation products. A 10% formulation was corrosive in the Draize test. The only sensitization study reviewed involved a mixture of methy lene bis( thiocyanate) (0.2%), chlorpyrifos (0.1 %), and TCMTB (0.2%), and did not show any evidence of an allergic

322

CHAPTER 11


response. Nevertheless, a 10% formulation used as an industrial preservative was labeled as a sensitizer.

Elemental Sulfur - transformation? products

S

--()g-

1/2S 2 0 3

",02,H 2 0 '~

NCS-CH 2 -SCN

Methylene bisthiocyanate

Myclobutanil is a systemic fungicide used on grapes and tree fruits. The dermal sensitization data reviewed were not sufficient to determine whether myclobutanil is a skin sensitizer. It caused no erythema or edema in the Draize test, but an accidental direct exposure involving myclobutanil and an adjuvant occluded against the skin without decontamination did prove capable of causing dermatitis (92-1379).

Pentachloronitrobenzene (PCNB) is a soil fungicide and seed treatment formulated as an emulsifiable concentrate, wettable powder, or flowable dust. Commonly treated crops include grain, cotton, cole crops, celery, strawberries, and ornamental flowers. No dermal irritation or sensitization data were available for review. Two reports in the public domain literature indicate that PCNB may be a human sensitizer. Cronin reported on a case seen at the St. John's Hospital in London (Cronin, 1980) and O'Malley described a positive patch test to PCNB as an incidental finding in a study of California nursery workers (O'Malley and Rodriguez, 1998a; O'Malley et aI., 1995). There were no reports involving PCNB in the handler database, as the cases identified in the cited study were not included in the California illness registry.

C I * C I Cl Cl

Cl

N02

S03

~

HS03 ,H 2 S04 _ _

1/2

H20

pH temperature ionic strength

TCMTB (thiocyanomethylthiobenzothiazole) is a compound used as both a fungicide and an antibacterial agent. It is contained in both wood preservation products and industrial preservatives. A 30% formulation was corrosive in the Draize test. A dilute wood preservative formulation (containing < 1% TCMTB) was a nonsensitizer in the Buehler assay, but a 30% formulation of TCMTB used as a bacteriocide, fungicide, and algicide was labeled as a sensitizer. There were no cases associated with TCMTB in the handler database.

ry--Nysr-

s~

~s TCMTB Busan

Thiabendazole is an agricultural fungicide with systemic activity against Fusarium, Penicillium, and other molds. It is formulated as a dust, a wettable powder, and a flowable dust. The 98.5% formulation is a minimal irritant in the Draize test, and no cases were associated with thiabendazole in the handler database. There were no sensitization studies done on thiabendazole per se, but the Buehler test was negative on a mixture of thiabendazole, captan, and pentachloronitrobenzene.

~---zJ N~N~ ~C') Thiabendazole S

PCNB

The large number of cases associated with elemental sulfur in California agriculture is striking and would seem to imply that sulfur is a potent skin irritant. Scattered cases have also been reported in applicators and field workers in the state of Washington. However, a 98% formulation caused only minimal irritation in the Draize test and was negative in the Buehler test. Other work has shown that a 25% concentration of wettable powder produced a 2+ irritant reaction on intradermal injection and was a sensitizer in the guinea pig maximization test (Matsushita et aI., 1977). Elemental sulfur has also been reported as a human allergen (Gregorczyk and Swieboda, 1968; O'Malley and Rodriguez, 1998b; O'Malley et aI., 1995; Schneider, 1978; Wilkinson, 1975). It is not clear which sulfur products (figure) may be responsible for its reported effects on the skin. The issue is explored more fully in the chapter on sulfur in this volume.

Triadimefon is a systemic fungicide for control of powdery mildew on cereals, deciduous fruit, and grapes. The 50% formulation is a minimal irritant in the Draize test. The 95% technical material is a sensitizer in the Buehler assay. The same test on a more dilute formulation was negative. Case 82-1412 in the handler database was a suspected allergic reaction to triadimefon.

Vinclozolin is a fungicide used for control of Botrytis, Sclerotinia, and Monilia on grapes, strawberries, and other crops.

11.2 Review of Use Categories Cl

It is a minimal irritant in the Draize test. It is labeled as a sen-

I

)0;(O~CH I

CH3

Cl

Vinclozolin

2

3

11.2.4 FUMIGANTS AND BIOCIDES Aluminum phosphide is a fumigant formulated as solid tablets that release phosphine gas on contact with air and water. It is used for both commodity fumigation and rodent control. Animal sensitization and irritation data were not available for review. Dermatitis cases have occurred in handlers following application (84-2184) and contact with partially spent dust. H20,02 AlP - - - PH3 PH3 ~

Cl

H-C-C=C I I ' H H H

sitizer, but no animal test data were available for review. The handler database did not contain any cases associated with the use of vinclozolin.

CI>rL)--o

/

323

1,3 Dichloropropene

Ethylene oxide is a commodity fumigant used in food processing and in hospital sterilization equipment. No dermal irritation or sensitization study was available for review; however, some ethylene oxide products are labeled as dermal sensitizers. Numerous case reports have described allergic contact dermatitis in hospital workers handling rubber products and other medical supplies sterilized with ethylene oxide (Alomar et aI., 1981; Alomar and Gimenez Camarasa, 1981; Fisher, 1988; Hanifin, 1971; Romaguera and GrimaIt, 1980; Romaguera et aI., 1977; Romaguera and Vilaplana, 1998; Taylor, 1977). The case described in the handler database (87-2720) involved a chemical bum following accidental direct contact with ethylene oxide gas. Fosthiozate is a compound that has limited use as a soil nematicide for root vegetable crops. It causes only minimal irritation in the Draize test, but the technical material is a sensitizer in the guinea pig maximization test. There were no cases associated with fosthiozate in the handler database.

+ AI(OHh

P04 + H20

Dazomet is a fumigant that releases irritant amines and mercaptans as it breaks down in soil. It is used as a preplanting treatment for ornamental beds and nurseries. The technical material causes minimal irritation in the Draize test, but most of the dazomet products reviewed (20-24% active ingredients) are all labeled as skin irritants. No dermal sensitization studies were available for review, but the compound has been reported as a sensitizer in the public literature (Black, 1973), and a series of cases of irritant contact dermatitis associated with dazomet has recently been reported from France (Gamier et aI., 1993). The latter cases were similar to a case reported to the handler database that occurred on contact of dazomet with skin covered with sweat (case 90-2448).

S==lSj CH3-N~N-CH3

Dazomet

Dichloropropene is a fumigant used principally for soil sterilization in the production of root vegetables and other crops. It is corrosive in the Draize tests. No animal sensitization studies were available for review, but the products registered in California are registered as sensitizers. A case reported by Nater and Gooskens describes an allergic reaction to a mixture of dichloropropane and dichloropropene, identifying dichloropropene as the most likely allergen (Nater and Gooskens, 1976). The 11 cases in the handler database (Table 11.1) appear consistent with a simple irritant mechanism (see case 88-2091).

Fosthiozate

Methyl bromide is a volatile fumigant used as a structural, soil, and commodity fumigant. It is corrosive in the Draize test, but no dermal sensitization data were available for review. The numerous cases in the handler database tended to involve prolonged occlusion of methyl bromide against the skin, as illustrated in Table 11.1 (82-32). Metam sodium, a soil fumigant and nematocide, is also effective against weeds and soil fungi. The reaction of metam with water produces methylisothiocyanate (MITC), carbon disulfide, hydrogen sulfide, and methyl amine. A 42% formulation was corrosive in the Draize test. In the Buehler assay, there was so much irritation present in the induction phase of the test that it was difficult to distinguish between irritant and allergic reaction. S

H2 0

CH3NH-~ Na+. 2 H20 _

"s

CH3N=C=S MITC

+ CH 3-NH 2 Methylamine

+ CS2

Carbon disulfide

+

CH 3N=C=O

MIC - up to 4% of the level of MITe

+ H2S Hydrogen disulflde

Cases of contact dermatitis associated with metam sodium have been reported in several jurisdictions around the world. In Germany, the cases stemmed from use of metam in the production of root vegetables (lung, 1975; Jung and Wolff, 1970a,

324

CHAPTER 11


1970b; Wolff and Jung, 1970). Cases of dennatitis were also reported from workers wading into the Sacramento river to clean up metam-sodium spilled into the river following a train derailment near Dunsmuir, California in July, 1991 (Koo et al., 1995). Between 1991 and 1995, one case of dennatitis associated with this compound was reported in a Washington applicator (O'Malley, 1997). There were 34 cases in the California handler database.

Acetochlor is not currently registered in California and consequently no animal sensitization or irritation data were available for review. However, the product infonnation in trade literature (Meister, 1995) indicates that acetochlor is a category 11 skin irri tan t. C H3

~2

.0

E ~1

15161718192021222324252627282930 Date during June, 1988

D

Crew 1

•

Crew2

III

Crew3

Figure 11.1 Dennatitis in three crews of nectarine harvesters-June, 1998. Reprinted with pennission from Han1ey and Belfus, State of the Art Reviews in Occupational Medicine.

tree fruit was lengthened to 21 days following the episode (O'Malley et aI., 1990). In August, 1995, the Department of Pesticide Regulation (DPR), Worker Health and Safety (WHS) Branch, received a report regarding an outbreak of dermatitis (sunburnlike erythema on the chest, neck, arms, and face) among crews of workers performing hand labor activities on a table grape ranch in northern Fresno County near Kerman, California. Of 202 workers (8 crews) lifting grape foliage over vine guide wires (turning cane), 65 (32.2%) sought treatment between August 9 and August 20. No rashes were reported among 54 workers (2 crews) working in the same vineyards pulling leaves. The incidence of workers seeking treatment varied from 3.7% in the crew apparently least affected to 87.5% in the crew with the largest number of reported cases. Because the crews left the Fresno area, it was not possible to determine whether the variation in reported cases was attributable to differences in the occurrence of rash or to differences in the likelihood of seeking treatment following onset of rash. Rash incidence/10 employed workers

10

8

6 4 2

o

o

10 20 30

40 50 60

70

Propargite-Residue Hours

Figure 11.2 Correlation between cumulative exposure to propargite and the occurrence of dennatitis.

330

CHAPTER 11


Crew Number

•• • • •

•

11.5 RESIDUE PROBLEMS IN OTHER JURISDICTIONS

35 -

30 ...,

~ 25

!::o 20 -

.! 15 -4

8 9

Total

910 -

Z

54

0-- T 8/8195

8/9/95

8/10/95

8/11/95

Date of Reported Onset

Figure 11.3

Crews affected by dermatitis and dates of reported onset.

The crews affected by dermatitis worked between August 7 and August 11 in blocks of Thompson seedless and Red Globe grapes (Fig. 11.3). Application records showed that these blocks had received treatments with propargite, Britz Buffer, glyphosate, methomyl, dichloronitroaniline, myclobutanil, Latron B 1956, gibberelin, triflumizole, elemental sulfur, and iprodione. At site 304, three dislodgeable foliar residue (DFR) samples were taken 8/15/95 by DPR Pesticide Use Enforcement (PUE)lFresno County Agricultural Commissioner (CAC). Single samples were also taken at site 301 and site 302. Propargite levels on 8/15/95 ranged from 0.37 to 0.66 ug/cm 2. Detectable residue was also found for elemental sulfur (0.16-1.1 !l-g/cm2), iprodione (non-detect [nd]-0.16 !l-g/cm2), myclobutani1 (nd0.16 !l-g/cm2), and dich10ronitroaniline (nd-0.045 !l-g/cm2). At site 301, the sulfur residue (1.1 !l-g/cm2) exceeded the residue of propargite (0.66 !l-g/cm2), but at other sites the sulfur residue was <

~ More acceptable - - - : - =

Magnitude of effect

\

\

Apply higher tiers or , mitigation

'"

Riskjudged" " , _ acceptable -

_ __

\\

Magnitude of effect

Figure 13.6 Presentation of exceedence probabilities (A) as a continuum of likehoods in an exceedence profile (B) and the use of these curves in decision making (C and D). Adapted from Solomon and Takacs, 2001.

and resting and other dormant stages are produced from which populations in the next season will develop. Similar mechanisms exist in environments with a dry season where ephemeral water bodies are subjected to drying out. Therefore, as many organisms in these regions undergo seasonal resets, a stressor return that occurs less frequently than once per season is likely to be tolerable from the viewpoint of the long-term productivity of the population and the sustainability of function in the ecosystem, especially if the effects are spatially restricted. Protection of longer-lived species without seasonal resets, such as some fish, birds, or mammals, may, however, require the consideration of return frequencies of several years or more. If a stressor is present nonuniformly in the environment, unexposed areas \\' ill act as refugia (metapopulations) for repopulation of potentially impacted areas. The relative size of the exposed and unexposed areas and their closeness is important, but this issue is particularly significant for assessing risks from pesticide use, where untreated fields, set-aside land, conservation headlands, crop rotations, and mixed farming practices guarantee that refugia will be present. Similarly, refugia exist in streams and rivers, and many organisms have resistant stages or propagules from which population recovery can occur. Thus, probabilistic risk assessments (and hazard quotients) are additionally conservative because they do not consider repopulation from unexposed refugia. The example of the more rapid than expected recovery of the biota in the Rhine River from an endosulfan spill illustrates this point (Friege, 1986).

13.4 UNCERTAINTY IN RISK ASSESSMENTS Uncertainty analysis is important in ecotoxicological risk assessment because it both identifies and, to the extent possible, quantifies the uncertainty in the entire process of problem formulation, analysis, and risk assessment. In addition, an assessment of uncertainty may allow identification of ways in which uncertainty can be reduced. Uncertainties in risk assessment have three sources: 1. Ignorance or imperfect knowledge of things that should be known is the first source of error. An example of this is the lack of prior knowledge that DDT would biomagnify in the food chain. 2. Systematic errors in the risk assessment process are those that may occur through computational mistakes or through incorrect instrumental calibration. These errors can be addressed through better quality control and quality assurance. 3. Nonsystematic errors are random or stochastic errors that result from the random nature of the system being assessed. These types of errors can be described and quantified, but cannot be avoided or corrected for. To the extent that probabilistic risk assessment uses distributions of data, errors of this type are incorporated in the assessment process. There is a reluctance in many risk assessors to admit that uncertainty exists in any decision; however, all scientific data have some uncertainty, as do all risk decisions made by humans. A description of sources of uncertainty is, in fact, helpful

References

to the risk manager because this allows identification of mechanisms by which additional certainty can be added to the risk assessment with the appropriate increase in comfort level in the result of the process on the part of the public and other users of the information.

13.5 RISK COMMUNICATION Once risk has been assessed, it will almost always be necessary to develop a risk communication strategy. This strategy may be needed for communication between the assessor and the manager or between the manager and politicians and the public. Communicating risk is not an easy task, especially if the result of the assessment is contrary to conventional wisdom or to the interests of certain stakeholder groups. Written risk communications must be designed to be understood by the lay public and this is a skill that few have. Verbal communication is even more difficult, especially if it is carried out in front of an audience where nonverbal communication can give mixed signals to the audience. Part of the difficulty with communicating risk assessment is that, by human definition, risk is assumed to be adverse. In the final analysis, in the ecosystem there is neither "good" nor "bad," and certainly nothing "adverse." An ecological change is labeled "adverse" by individuals or society and is basically a value judgment. Thus, to conduct a risk assessment means that someone has made a value judgment of which conditions will be defined as adverse. The public is often suspicious of the motives of those who communicate risks and may perceive conflicts of interest (Lackey, 1995). Risk communication is a form of persuasive communication that is designed to change behavior or what may be deeply held beliefs. Two of the keys to successful risk communication are (1) expressing the technical evaluation in a way that is meaningful to the audience (this can be achieved by using appropriate analogies to describe the risk assessment process and not using technical terminology that can be misunderstood) and (2) anticipating potential misunderstandings and dealing with them in a sympathetic way. In doing this, the communicator must recognize that groups who oppose a particular risk management strategy have a right to question a decision that affects them.

REFERENCES Ahlborg, U. G., Becking, G. c., Birnbaum, L. S., Brouwer, A., Derks, H. J. G., Feeley, M., Golor, G., Hanberg, A., Larsen, J. c., Liem, A. K. D., Safe, S. H., Schlatter, c., Waern, F., Younes, M., and Yrjanheikki, E. (1994). Toxic equivalence factors for dioxin-like PCBs. Report on a WHO-ECEH and IPCS consultation, December 1993. Chemosphere 28, 1049-1067. American Chemical Society Subcommittee on Environmental Analytical Chemistry (1980). Guidelines for data acquisition and data quality evaluation in environmental chemistry. Ana!' Chem. 52,2242-2249. ASTM (1991). "Annual Book of ASTM Standards, Section 11, Water and Environmental Technology." ASTM, Philadelphia. Baskin, Y. (1994). Ecosystem function ofbiodiversity. Bioscience 44, 657-660.

371

Bier, V. M. (1999). Challenges to the acceptance of probabilistic risk analysis. RiskAna!. 19,703-709. Burns, L. A. (1997). "Exposure Analysis Modeling System: User's Guide for EXAMS n, Version 2.97.5," Rep. 600/R-97/047. Office of Research and Development, U.S. EPA, Washington, DC. Cardwell, R. D., Parkhurst, B. R., Warren-Hicks, w., and Volosin, J. S. (1993). Aquatic ecological risk. Water Environ. Techno!. 5,47-51. Cardwell, R. D., Brancato, M. S., Toll, J., DeForest, D., and Tear, L. (1999). Aquatic ecological risks posed by tributyltin in United States surface waters: Pre-1989 to 1996 data. Environ. Toxieol. Chem. 18,567-577. Carrington, C. D. (1996). Logical probability and risk assessment. Human Eeo!. Risk Assess. 2, 62-78. Cunnane, C. (1978). Unbiased plotting positions. A review. 1. Hydro!. 37,205222. CWQG (1999). "Canadian Water Quality Guidelines (and updates)." Task Force on Water Quality Guidelines of the Canadian Council of Resource and Environment Ministers, Ottawa. ECOFRAM (1998). Ecological Committee on FIFRA Risk Assessment Methods. U.S. EPA, Washington, DC. ECOFRAM (1999). ECOFRAM Aquatic and Terrestrial Final Draft Reports, Vol. 1999. U.S. EPA, Washington, DC. Environment Canada (1997). "Environmental Assessments of Priority Substances under the Canadian Environmental Protection Act Guidance Manual Version 1.0," Rep. EPSI2ICC/3E. Environment Canada, Ottawa. Estes, J. A., Tinker, M. T., Williams, D., and Doak, D. F. (1998). Killer whale predation on sea otters linking oceanic and neashore ecosystems. Science 282, 473-476. European Union (1994). "Technical Guidance Documents in Support of the Risk Assessment of Existing Chemicals in Accordance with the Requirements of Council Directive," Rep. 931793IEEC (draft). European Union, Brussels. FAO (1989). "Revised Guidelines on Environmental Criteria for the Registration of Pesticides," Food and Agricultural Organization of the United Nations, Rome. Felton, J. C., Oomen, P. A., and Stephenson, J. H. (1986). Toxicity and hazard of pesticides to honeybees: Harmonization of test methods. Bee World 67, 114-124. Fletcher, J. S., Nellessen, J. E., and Pfleeger, T. E. (1994). Literature review and evaluation of the EPA food-chain (Kenaga) nomogram, an instrument for estimating pesticide residues on plants. Environ. Toxieo!. Chem. 13, 13831391. Friege, H. L. (1986). Monitoring of the River Rhine-Experience gathered from accidental events in 1986. In "Organic Micropollution in the Aquatic Environment. Proceedings of the 5th European Symposium," Rome, pp. 132-143. Ganzelmeier, H., Rautmann, D., Spangenberg, R., Streloke, M., Herrmann, M., Wenzelburger, H.-J., and Walter, H.-F. (1995). "Studies on the Spray Drift of Plant Protection Products." BIackwell Wissenschafts-Verlag, Berlin. Giddings, J. M., Biever, R. c., Annunziato, M. F., and Hosmer, A. J. (1996). Effects of diazinon on large outdoor pond microcosms. Environ. Toxieo!. Chem. 15,618-629. Giddings, J. M., Hall, L. W. J., and Solomon, K. R. (2000). An ecological risk assessment of diazinon in the Sacramento and San Joaquin River basins. Unpublished. Giddings, J. M., Solomon, K. R., and Maund, S. J. (2001). Probabilistic risk assessment of cotton pyrethroids: n. Aquatic mesocosm and field studies. Environ. Toxieol. Chem. 20, 660-668. Giesy, J. P., and Graney, R. L. (1989). Recent developments in and intercomparisons of acute and chronic bioassays and bioindicators. Hydrobiology 188/189,21-60. Giesy, J. P., Solomon, K. R., Coates, J. R., Dixon, K. R., Giddings, J. M., and Kenaga, E. E. (1999). Chlorpyrifos: Ecological risk assessment in North American aquatic environments. Rev. Environ. Contam. Toxieo!. 160, 1-129. Hall, L. W. J., and Anderson, R. D. (1995). The influence of salinity on the toxicity of various classes of chemicals to aquatic biota. Crit. Rev. Toxieol. 25,281-346.

372

CHAPTER 13

Ecotoxicological Risk Assessment

Hall, L. W. J., and Giddings, J. M. (2000). The need for multiple lines of evidence for predicting site-specific ecological effects. Human Eeo!. Risk Assess. 6, 679-710. Hall, L. W. J., Giddings, J. M., Solomon, K R., and Balcomb, R. (1999). An ecological risk assessment for the use of Irgarol 1051 as an algaecide for antifoulant paints. Crit. Rev. Toxieol. 29,367-437. Health Council of the Netherlands (1993). Ecotoxicological risk assessment and policy-making in the Netherlands-Dealing with uncertainties. Network 6/7,8-11. Hill, I. R, Heimbach, E, Leeuwangh, P., and Matthiessen, P., eds. (1994). "Freshwater Field Tests for Hazard Assessment of Chemicals." CRC Press, Boca Raton, FL. HC (1993). "A Strategy for the Virtual Elimination of Persistent Toxic Substances, Vo!. I. Report of the Virtual Elimination Task Force to the HC." International Joint Commission, Windsor, ON. James, E C. (1994). Society Actions. Joint EAS/AIBS review of President Clinton's plan for the management of forests in the Pacific Northwest. Bul!. Entomo!' Soc. Am. June, 69-75. Klaine, S. J., Cobb, G. P., Dickerson, R. L., Dixon, K R., Kendall, R J., Smith, E. E., and Solomon, K R (1996). An ecological risk assessment for the use of the biocide, dibromonitrilopropionamide (DBNPA) in industrial cooling systems. Environ. Toxieo!. Chem. 15,21-30. Kovacs, T. G., Martel, P. H., Voss, R H., Wrist, P. E., and Willes, R E (1993). Aquatic toxicity equivalency factors for chlorinated phenolic compounds present in pulp mill effluent. Environ. Toxieol. Chem. 12, 684-691. Lackey, R T. (1995). The future of ecological risk assessment. Human Eeo!. Risk Assess. 1, 339-343. Laudan, L. (1994). "The Book of Risks. Fascinating Facts About the Chances We Take Every Day." Wiley, New York. Lehtinen, K.-J., Axelsson, B., Kringstad, K, and Stromberg, L. (1991). Characterization of pulp mill effluents by the model ecosystem technique: SSVL investigations in the period 1982-1990. Nordie Pulp Paper Res. 1. 2,81-88. Liber, K, Kaushik, N. K, Solomon, K R., and Carey, J. H. (1992). Experimental designs for aquatic mesocosm studies: A comparison of the "ANOVA" and "regression" design for assessing the impact of tetrachlorophenol on zooplankton populations in limnocorals. Environ. Toxiea!. Chem. 11, 6177. Liber, K, Solomon, K R., Kaushik, N. K., and Carey, J. H. (1994). Impact of 2,3,4,6-tetrachlorophenol (DIATOX) on plankton communities in limnocorals. In "Aquatic Mesocosm Studies in Ecological Risk Assessment" (R. L. Graney, J. L. Kennedy, and J. H. Rogers, eds.), pp. 257-294. Lewis Publishers, Boca Raton, FL. Lynch, M. R, ed. (1995). "Procedures for Assessing the Environmental Fate and Ecotoxicology of Pesticides," pp. 1-54. SETAC Europe, Brussels. McBean, E. A., and Rovers, E A. (1992). Estimation of the probability of exceedences of a contaminant concentration. Ground Water MonU. Rev. 12, 115-119. Mullins, J. A, Carsel, R. E, Scarbrough, J. E., and Ivery, A M. (1993). "PRZM-2 A Model for Predicting Pesticide Fate in the Crop Root Zone and Unsaturated Soil Zones: Program and User's Manual for Release 2.0," Rep. 600/R-93/046. U.S. Environmental Protection Agency, Athens, GA NRC (1993). "Issues in Risk Assessment." National Academy Press, Washington,DC. OECD (1981). "OECD Guidelines for Testing of Chemicals." OECD, Paris. OECD (1984). "OECD Guidelines for Testing of Chemicals, Update of 1984." OECD, Paris. OECD (l992a). "GLP Consensus Document. Compliance of Laboratory Suppliers with GLP Principles. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 5," Rep. 49. OECD, Paris. OECD (l992b). "GLP Consensus Document. Quality Assurance and GLP. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 4," Rep. 48. OECD, Paris. OECD (I 992c). "GLP Consensus Document. The Application of GLP Principles to Field Studies. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 6," Rep. 50. OECD, Paris.

OECD (l992d). "Guidance for GLP Monitoring Authorities Guides for Compliance Monitoring Procedures for Good Laboratory Practice 2," Rep. 46. OECD, Paris. OECD (l992e). "Guidance for GLP Monitoring Authorities. Guidance for the Conduct of Laboratory Inspections and Study Audits 3," Rep. 47. OECD, Paris. OECD (1992f). "Guidance for GLP Monitoring Authorities. Guidance for the Conduct of Laboratory Inspections and Study Audits," Rep. 47. OECD, Paris. OECD (1992g). "The OECD Principles of Good Laboratory Practice. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 1," Rep. 45. OECD, Paris. OECD (l993a). "GLP Consensus Document. The Application of the GLP Principles to Short-Term Studies. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 7," Rep. 73. OECD, Paris. OECD (l993b). "GLP Consensus Document. The Role and Responsibilities of the Study Director in GLP Studies. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 8," Rep. 74. OECD, Paris. Okkerman, P. C., van der Plassche, E. J., Emans, H. J. B., and Canton, J. H. (1993). Validation of some extrapolation methods with toxicity data derived from multiple species experiments. Eeataxiea!' Environ. Safety 25, 341359. OMEE (1990). "The Ontario Ministry of the Environment Scoring System: A Scoring System for Assessing Environmental Contaminants," Hazardous Contaminants Branch, Ontario Ministry of the Environment, Toronto. Parker, R. (1999). "GENEEC." Environmental Fate and Effects Division, Office of Pesticide Programs, U.S. EPA, Washington, DC. Parkhurst, B. R, Warren-Hicks, W., Cardwell, RD., Volison, J., Etchison, T., Butcher, J. B., and Covington, S. M. (1996). "Aquatic Ecological Risk Assessment: A Multi-Tiered Approach to Risk Assessment," Rep. 91-AER-1. Water Environment Research Foundation, Alexandria, VA. Parrott, J. L., Hodson, P. v., Servos, M. R., Huestis, S. L., and Dixon, D. G. (1995). Relative potencies of polychlorinated dibenzo-p-dioxins and dibenzofurans for inducing mixed function oxygenase activity in rainbow trout. Environ. Taxieol. Chem. 14,1041-1050. Reinert, K H., Bartell, S. M., and Biddinger, G. R, eds. (1998). "Ecological Risk Assessment Decision-Support System: A Conceptual Design," pp. 1-98. SETAC Press, Pensacola, FL. Riley, D., ed. (1993). "Principles of Risk Assessment," Rep. 73. The Winand Staring Centre for Integrated Land, Soil and Water Research, Wageningen, Netherlands. Schwarz, R c., Schults, D. W., Ozretich, R. W., Lamberson, J. 0., Cole, E A, DeWitt, T. H., Redmond, M. S., and Ferraro, S. P. (1995). Sigma PAH: A model to predict the toxicity of polynuclear aromatic hydrocarbon mixtures in field-collected sediments. Environ. Taxieol. Chem. 14, 1977-1978. SETAC (1991). "Report of the Wintergreen Workshop on Microcosms." SETAC Foundation for Education, Pensacola, FL. SETAC (1994). "Pesticide Risk and Mitigation. Final Report of the Aquatic Risk Assessment and Mitigation Dialog Group." SETAC Foundation for Environmental Education, Pensacola, FL. Solomon, K R. (1996). Overview of recent developments in ecotoxicological risk assessment. Risk Ana!. 16,627-633. Solomon, K R (1999). Integrating environmental fate and effects information: the keys to ecotoxicological risk assessment for pesticides. In "Pesticide Chemistry and Bioscience: The Food-Environment Challenge" (G. T. Brooks and T. R. Roberts, eds.), pp. 313-326. Royal Society of Chemistry, London. Solomon, K. R., Baker, D. B., Richards, P., Dixon, K R., Klaine, S. J., La Point, T. w., Kendall, R. J., Giddings, J. M., Giesy, J. P., Hall, L. W. J., Weisskopf, c., and Williams, M. (1996). Ecological risk assessment of atrazine in North American surface waters. Environ. Taxieol. Chem. 15, 3176. Solomon, K R, and Chappel, M. J. (1998). Triazine herbicides: Ecological risk assessment in surface waters. In "Triazine Risk Assessment" (L. Ballantine, J. McFarland and D. Hackett, eds.), Vo!. 683, pp. 357-368. American Chemical Society, Washington, DC.

References

Solomon, K. R., Giddings, J. M., and Maund, S. J. (2001a). Probabilistic risk assessment of cotton pyrethroids: 1. Distributional analyses of laboratory aquatic toxicity. Environ. Toxico!. Chem. 20, 652-659. Solomon, K. R., Giesy, J. P., Kendall, R. J., Best, L. B., Coats, J. R., Dixon, K. R., Hooper, M. J., Kenaga, E. E., and McMurry, S. T. (200Ib). Chlorpyrifos: ecotoxicological risk assessment for birds and mammals in corn agroecosystems. Human. Eco!. Risk Assess. 7(2) (in press). Solomon, K. R., and Takacs, P. (2001). Probabilistic risk assessment using species sensitivity distributions. In "Species Sensitivity Distributions in Risk Assessment" (L. Postuma, T. Traas, and G. W. Suter, eds.), Chapter 15. CRC Press, Boca Raton, PL, USA (in press). Stephan, C. E., Mount, D. 1., Hansen, D. J., Gentile, J. H., Chaprnan, G. A., and Brungs, W. A. (1985). "Guidelines for Deriving Numerical National Water Quality Criteria for the Protection of Aquatic Organisms and their Uses," Rep. PB 85-227049. U.S. EPA ORD ERL, Dulutb MN. Stephenson, G. L., Kaushik, N. K., Solomon, K. R., Day, K. E., and Hamilton, P. (1986). Impact of methoxychlor on freshwater plankton communities in limnocorrals. Environ. Toxico!. Chem. 5,587-603. Suter, G., 11, Barnthouse, L. w., Bartell, S. M., Mill, T., Mackay, D., and Patterson, S. (1993). "Ecological Risk Assessment." Lewis Publishers, Boca Raton, FL. Teske, M. E., and Scott, R. L. (2000). AgDRIFT™: An update of the aerial spray model AGDISP. Tilman, D. (1996). Biodiversity: Population versus ecosystem stability. Ecology 77,350-363. Tilman, D., Wedlin, D., and Knops, J. (1996). Productivity and sustainability influenced by biodiversity in grassland ecosystems. Nature 379, 718-720. Urban, D. J., and Cook, N. J. (1986). "Standard Evaluation Procedure for Ecological Risk Assessment," Rep. 540/09-86/167, Hazard Evaluation Division, Office of Pesticide Programs, U.S. EPA, Washington, DC.

3'/3

U.S. EPA (1982). "Guidelines and Support Documents for Environmental Effects Testing." U.S. EPA, Washington, DC. U.S. EPA (1986). "Good Laboratory Practice Guidelines." U.S. EPA, Washington, DC. U.S. EPA (1992). "Framework for Ecological Risk Assessment," Rep. 630/R92/001. U.S. EPA, Washington, DC. U.S. EPA (l995a). Final water quality guidance for the Great Lakes system. Federal Register 15,366-15,425. U.S. EPA (l995b). "The Use of the Benchmark Dose Approach in Health Risk Assessment," Rep. 630/R-94/007. Risk Assessment Forum, U.S. EPA, Washington, DC. U.S. EPA (1998). "Guidelines for Ecological Risk Assessment," Rep. 630/R95/002F. Risk Assessment Forum, U.S. EP, Washington, DC. U.S. EPA (2001). ECOTOX Database System. U.S. Environmental Protection Agency, Washington, DC, http://www.epa.gov/ecotox/ Van den Brink, P. J., Van Wijngaarden, R. P. A., Lucassen, W. G. H., Broclc, T. C. M., and Leeuwangh, P. (1996). Effects of the insecticide Dursban® 4E (active ingredient chlorpyrifos) in outdoor experimental ditches: 11. Invertebrate community responses and recovery. Environ. Toxico!. Chem. 15, 1143-1153. Versteeg, D. J., Belanger, S. E., and Carr, G. J. (1999). Understanding singlespecies and model ecosystem sensitivity: Data-based comparison. Environ. Toxicol. Chem. 18, 1329-1346. Walker, B. (1992). Biodiversity and ecological redundancy. Conserv. Bio!. 6, 18-23. Walker, B. (1995). Conserving biological diversity through ecosystem resilience. Conserv. Bio!. 9,747-752. Wilson, R., and Crouch, E. A. C. (1987). Risk assessment and comparisons: An introduction. Science 236, 267-270.

CHAPTER

14 Developmental and Reproductive Toxicology of Pesticides Poomi Iyer California Environmental Protection Agency

14.1 INTRODUCTION Pesticides have been in use since the early days of modern agriculture and, as a class of chemicals, are well studied and subject to much regulation. There is growing concern about the safety of pesticides and how exposure (occupational/via food residues/contamination of air and water) may affect human health in general and reproductive outcome in particular. Attention to pesticide residues in food (fresh and processed), spurred by a demand for organically [as defined by the Organic Foods Act of California (1990)] grown produce, and reports on the levels of pesticides in the diets of infants and children (NAS, 1993) have led to the passage of federal regulations in the United States (EPA, 1996a). Exposure to pesticides is inherent in most agriculture-related occupations, and studies on pesticide use and pregnancy outcome generally focus on birth defects and the effects on the reproductive system. Endpoints investigated cover a broad range, including early and late fetal loss, alteration in gestational age at delivery, formation of terata (birth defects), infant/child morbidity and mortality, male/female sexual dysfunction, sperm abnormalities, amenorrhea, dysmenorrhea, and illness during pregnancy and parturition. Pesticides acting on the developing organism or on the reproductive system may produce adverse effects by one of several mechanisms. They may be direct-acting by being chemically reactive and (1) cause germ cell destruction (e.g., alkylating agents) or (2) exert their effects due to their structural similarity to endogenous molecules (e.g., hormone agonists/antagonists such as phytoestrogens). They could also act indirectly and interrupt reproduction (1) by metabolism to a direct-acting compound or reactive intermediate, (2) via endocrine alterations such as increased/decreased steroid clearance, and (3) by stimulating or inhibiting neuroendocrine responses at the level of the hypothalamus or pituitary. Developmental toxicants, through a direct- or indirect-acting mechanism, may result in embryolethality, frank malformations, or other undesirable sequelae such as growth retardation or functional alteration. Similarly, pesticides affecting reproHandbook of Pesticide Toxicology

Volume I. Principles

duction may act on selected stages targeting the prenatal stage, the prepubertal stage, or the adult, resulting in damage to the reproductive organs and/or impaired fertility. The potential of pesticides to adversely affect development is determined from studies conducted on animals to meet the regulations of the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA; EPA, 1982). Although this chapter is not intended to be encyclopedic in nature, the reproductive and developmental toxicity of a number of compounds regarded as pesticides will be discussed.

14.1.1 DEVELOPMENTAL TOXICITY Adverse effects on the conceptus, covering the period from conception to the completion of morphological structure and functional capability of the individual, are included in this category. After the malformations caused by the drug thalidomide in the early 1960s, the role of chemical exposure in the causation of birth defects has received much attention. Accordingly, following the pattern for drugs developed by the Food and Drug Administration (FDA) , animal studies using pesticides are conducted and findings from such studies are extrapolated to humans. In addition to malformations in fetuses, endpoints such as prenatal death, growth alterations, developmental variations, and maternal effects as well as postnatal development are also investigated. Postnatal neurodevelopment in animals and the field of behavioral changes are now gaining attention especially as alterations in these areas have been reported for several pesticides (Boyes et aI., 1997; Chernoff et aI., 1979a; Gray et aI., 1986; Ostby et aI., 1985; Sette et aI., 1989; Tilson et aI., 1988). Increased perinatal mortality has been reported in laboratory animals from excessive exposure to pesticides; in one case, the deaths may have been related to functional cardiac disorders (Grabowski and Daston, 1983). Also, postnatal exposure via lactation has resulted in the induction of cataracts in rat pups (Chernoff et aI., 1979b; Gaines and Kimbrough, 1970). Epidemiological studies have documented an association between spontaneous abortions and fetal deaths and maternal exposure (Barlow and Sullivan, 1982; Goulet and The-

375

Copyright © 200 I by Academic Press. All rights of reproduction in any fonn reserved.

376

CHAPTER 14

Developmental and Reproductive Toxicology of Pesticides

riault, 1991; Weinberg, 1993; Wi1son, 1979). The potential of pesticides to cause adverse effects on the developing individual has been demonstrated in laboratory animal studies with a range of effects at various stages of development. Of the numerous pesticides tested so far, about 43% have been documented to induce birth defects in experimental animals (J. L. Schardein, personal communication). Although this could be alarming, it must be remembered that all chemicals can interfere with some aspect of development if administered at a sufficiently high dose level at the appropriate time of development to certain species of animals. Development is a complex process, and the effects of a chemical depend on the time of exposure, the exposure level, and the extent of maternal effects. The nature of an insult during embryonic development is often less important than the developmental stage at which it occurs. This is because the steps in the sequence of tissue interactions during development are susceptible to disruption for a specific period of time. Typically, early exposure (i.e., during preimplantation and early postimplantation) results in fetal death, whereas exposure during the organogenesis period (3 weeks after conception through 2 months in humans) results in structural birth defects. However, there is evidence that exposure during the preimplantion period can also result in teratogenic effects (Rutledge et aI., 1992; Spielmann and Vogel, 1989). Pre- or postimplantation exposure of the developing conceptus to toxicants may also result in a "derailment" in the genetic control of development and the coordinated cascade of events that occur during normal development. Thus, developmental abnormalities may be induced by disrupting the coordinated expression of developmental genes involved in genomic imprinting, cell lineage specification, cell mixing and recognition, cell-cell interaction, cell migration and differentiation, and segmentation, depending on the time of exposure (Kimmel et aI., 1993). Exposure after the critical stage of organogenesis often results in growth retardation or other functional deficits. For regulatory purposes, hazard identification is based on the dose level at which an effect is noted, the observation of a dose-response relationship, and whether the adverse effect on the conceptus occurs at an exposure level below that which causes severe maternal toxicity. This is done partly to determine if the maternal effects are the underlying mechanism for the developmental effects noted. Testing for developmental toxicity therefore requires high doses even though humans may be exposed to low doses in practice. The details on testing protocols will be elaborated later on, suffice it to say at this point that the purpose for testing at high doses is to get an understanding of the mechanism of action of the chemical. The limitations of such testing, as in other toxicology studies, include the range of sensitivity within humans, extrapolation of effects observed at high doses to predict those likely to occur at low doses, as well as extrapolation from tests in animals to humans.

14.1.2 REPRODUCTIVE TOXICITY This category covers the adverse effects on the reproductive system, ranging from small decrements in reproductive abil-

ity in either male or female to a situation of overall infertility. It also includes effects on the reproductive organs irrespective of the influence on fertility in the affected individual. This is particularly relevant when animals are used as models for effects in humans because fertility in rodents is often difficult to disrupt and other indicators of reproductive function may be more sensitive. Hence, fertility cannot be used as the only tool to diagnose adverse effects. The effects of pesticides on reproduction may be acute or chronic and may be directed to a single sex. Also, pesticides affecting reproduction may act on selected stages targeting the prenatal stage, the prepubertal stage, or the adult. A few examples of reproductive toxicity by one or more of these mechanisms will be reviewed herein. Furthermore, due to unique anatomical and physiological characteristics in different species, the effects noted may differ. A pesticide causing reproductive toxic effects in one species may not be toxic in another and hence the relevance to humans needs to be examined. The impact on fertility also needs to be considered because the level of sperm production differs across species and a decrease in sperm count may not have the same impact in all species.

14.1.3 EPIDEMIOLOGY The degree to which pesticide exposure mayor may not be responsible for developmental problems in humans is not known. Available epidemiological data on the developmental toxicity of occupational and environmental pesticide exposure are limited in the sense that although a number of studies have some indications of elevated risk, the epidemiological evidence on the whole is unclear. The incidence of pesticide-related adverse reproductive/developmental outcomes has been extensively reviewed (Sever et aI., 1997). Elevated risk of limb anomalies (Lin et aI., 1994; Schwartz and LoGerfo, 1988; Schwartz et aI., 1986) has been associated with ecological exposure and occupational exposure; and orofacial clefts (Nurminen, 1995) have been related to maternal environmental exposure. In several countries, it has been noted that maternal agricultural occupation and pesticide exposure may be associated with elevated risk of spontaneous abortion and stillbirth (Goulet and Theriault, 1991; Heidam, 1984; Restrepo et al., 1990a, b; Rita et aI., 1987). Nonetheless, although reports from accidental exposure as well as occupational use have documented that pesticides can be incriminated in adverse reproductive outcomes, some studies have found no indication of reproductive hazards, presenting rather inconclusive results (Nurminen, 1995). Exposure to organochlorine and organophosphate pesticides in grape gardens in India resulted in higher abortion rates (almost sixfold) in 12 exposed couples compared to 15 nonexposed couples (Rita et aI., 1987). The compounds handled in this study included dichlorodiphenyltrichloroethane (DDT), lindane, quinalphos, dithane M45, metasystox, parathion, copper sulfate, dichlorovos, and dieldrin. Similarly, women working in vineyards in the Crimea also had higher rates of miscarriage after exposure to DDT, sulfur, methyl parathion, and copper sulfate (Nikitina, 1974). In China, women exposed

14.2 Mechanisms of Action

377

to chlorophenamidine (chlordimeform), dikishuang, and kitazin were found to be at increased risk of delivering stillbirths [relative risk (RR = 1.4-1.81)] as well as spontaneous abortions (RR = 1.90-4.00, depending on gravidity). The risks would have been even higher if previous adverse pregnancy outcomes had not been controlled for in the study (Weinberg, 1993). In rural California, second-trimester occupational exposure to pesticides was associated with an odds ratio (OR) of 4.8 in a casecontrol study of stillbirths and early neonatal deaths (Pastore et aI., 1995). Male and female farmers exposed to pesticides in central Sudan had a higher odds ratio for stillbirth (>500 g) in a case-control study: OR = 5.1; 95% confidence interval (Cl) = 1.4-9.6 (Taha and Gray, 1993).

in their site of action and most of these compounds are cytotoxic, carcinogenic, mutagenic, or developmentally toxic. They may also be toxic to the reproductive system and, in fact, the disruption of reproductive function could occur at doses lower than those that cause tumors. The classic example of such a mechanism is the case where the risk of sterility following many forms of cancer chemotherapy is considerably higher than the risk of second tumors (Kay and Mattison, 1985). Other direct-acting compounds are structurally similar to endogenous molecules, such as some organochlorines that may exert their effects through interaction with estrogen receptors. Organochlorines have been implicated in abnormal menses and impaired fertility (Mattison et al., 1983).

14.1.4 EXPOSURE

14.2.1.2 Indirect-Acting

Human malformations occur in roughly 5% of live births; therefore, to demonstrate an increase in the overall rate of malformation or incidence of a specific type of malformation from a documented exposure, a much larger population is required than if the background rate were zero (Fraser, 1977). The limitations often associated with epidemiological data, such as recall bias, lack of specificity, and use of surrogates for exposure, must be considered in evaluating findings that suggest an association with pesticide exposure. One explanation for these results is the specificity of the compounds involved in the exposure. Given that exposure is often categorized as either general pesticide use or agricultural setting, exposure to specific compounds is not evaluated. However, the effects of known classes of chemicals can be studied because a number of compounds have a common mechanism of action as reviewed in the next section. The adverse effects of pharmaceutical agents have been predicted from data on laboratory animals at exposures near maternally toxic levels (John son et aI., 1990). Much of the animal data on the reproductive and developmental effects of pesticides are generated for the purpose of pesticide registration under FIFRA and do not appear in the open literature. Hence, the amount of published information is limited. This chapter will attempt to fill that void by addressing several aspects in the area of developmental and reproductive toxicity of pesticides.

14.2 MECHANISMS OF ACTION Compounds used as pesticides have different mechanisms of action and these may be independent of the species targeted. Pesticides can therefore be studied by their mechanism of action. 14.2.1 BASIC MECHANISMS OF EXPOSURE 14.2.1.1 Direct-Acting

Pesticides that are direct-acting may exert their effect by being chemically reactive; these compounds may be nonspecific

Developmental/reproductive toxicants that are metabolized to either chemically reactive products or structures similar to endogenous molecules fall into this group. The embryo and fetus as well as both the ovary and the testis have been demonstrated to have microsomal monooxygenases, epoxide hydrases, and transferases responsible for metabolizing xenobiotics (Dixon and Lee, 1980; Heinrichs and Juchau, 1980; Mattison and Thorgeirsson, 1978, 1979; Pedersen et aI., 1985). The basic mechanisms outlined previously, along with timing, influence the various developmental effects that are observed. The concept that insult prior to the beginning of "organogenesis" results only in an "all (i.e., death) or none" effect is no longer considered accurate. Abnormal development subsequent to insult at preimplantation stages suggestive of early alterations in pattern formation has been reported for retinoic acid (Rutledge et aI., 1994). We are continuing to learn more regarding developmental stage-related sensitivities and this area of pattern formation and early alterations is of concern in the area of pesticide exposure. 14.2.2 TIMING OF EXPOSURE

Just as the time of exposure determines the developmental effects of a chemical, toxicity to the reproductive system also varies with the timing of exposure. Accordingly, reproductive toxicants can be classified as follows. 14.2.2.1 Prenatal Reproductive Toxicants

These are compounds that affect the developing reproductive system in utero, resulting in prenatal ovarian or testicular toxicity in humans and animals. These include the absence of or a considerable decrease in the number of primordial oocytes (e.g., primary or secondary amenorrhea). Thus, although it is possible that prenatal exposure could affect the oocyte, current study protocols are not designed to detect subtle changes that may occur. More frequent testing for toxicity to male reproductive processes is conducted because of the premise of male sensitivity and the ease of access to gametes and gonads. Furthermore, it is often presumed that the female gamete is better

378

CHAPTER 14


protected from mutagenic chemicals due to the probability that chemically induced deoxyribonucleic acid (DNA) damage in a primary oocyte is repaired prior to ovulation (Preston et al., 1995). Despite the differences between males and females in terms of anatomy and biological control mechanisms for reproduction, in the absence of data to the contrary, it should be assumed that both male and female gametes are equally sensitive to reproductive toxicants.

exposure because the developing animal may be extremely sensitive to toxicants during sex differentiation, and a number of these effects are difficult to detect until late in life. Ultimately, the concern for reproductive vulnerability is noted in its impact on pesticide regulation. This might translate into the cancellation of registration for a given compound or the chemical may be placed on a list, such as California's Proposition 65, aimed at controlling human exposure to compounds capable of adverse effects.

14.2.2.2 Prepubertal Reproductive Toxicants

The modulation of the hypothalamic-pituitary-ovarian axis is influenced by higher centers in the central nervous system. Both ovulation and ovarian hormone production require the interaction of components of this axis and the effects of specific compounds on any of these levels exert their influence on reproduction. The effects can be elucidated clinically by examining the impact on menstrual cyclicity in humans and estrus cyclicity in non primate animals. Although there are no data to link the increasing trend of early menarche with pesticide use, the area of estrogenic effects and their role on the onset of puberty is receiving attention (Thigpen et al., 1999). In the rat, studies on dams consuming diets containing high concentrations of estrogenic substances, with resultant exposure of their pups in utero and prior to weaning, suggest that the estrogenimprinting metabolism of the pups or future responses to other exogenous estrogenic substances may be altered (Lamartiniere et al., 1995). Thus, the effects of exposure may be noted in subsequent generations because a number of pesticides may have estrogenic potential. The prepubertal gonad may differ from the sexually mature gonad in its sensitivity to the toxic impact of pesticides and this is an endpoint that deserves examination. Contaminants of pesticides, such as 2,3,7,8-tetrachlorodibenzop-dioxin (TCDD), may, in fact, have such an effect, but the findings are not conclusive. There is increasing attention to the latent effects of pesticides on sexual differentiation in rodents. 14.2.2.3 Adult Reproductive Toxicants

The effects on the reproductive system may be observed in adults as well as in their progeny if exposure occurs over a long period of time. These are generally detected in the multi generation studies conducted in laboratory animals. Studies submitted for regulatory purposes (fenthion, oxydemeton methyl) have demonstrated effects such as increased epididymal vacuolation and other histopathological changes (California Department of Pesticide Regulation, 2000). Gender differences in response to chemical insult must be taken into consideration; for example, unlike the male where damage to the spermatocytes may only be a temporary problem, damage to oocytes is permanent. The new guidelines now require histopathology data on ovaries to detect changes that may be occurring due to prolonged pesticide exposure over various developmental periods. Additionally, recent studies have documented adult/pubertal alterations resulting from gestational and or neonatal exposures (Gray and Ke1ce, 1996). Hence, studies should include a comprehensive assessment of reproductive function after perinatal

14.3 REGULATORY ISSUES Outlined in this section are the many issues dealing with the use of pesticides and their regulation by state, federal, and international agencies. These include the conduct and interpretation of studies as well as the application of new findings and regulations. 14.3.1 HISTORY

The effects of thalidomide and the Kefauver-Harris Act in 1962 led the U.S. Food and Drug Administration (FDA) to strengthen drug testing. Currently, the U.S. government requires manufacturers to perform hazard assessments to determine the teratogenic potential of chemicals. The U.S. Environmental Protection Agency (EPA) published teratogenicity testing requirements in 1978 under FIFRA. Essentially, it mandated how testing was to be conducted and reported, which differed little from the FDA guidelines, except that exposure was to be initiated just before implantation and concluded the day before delivery (EPA, 1978). [Whereas the FDA requires three studies covering different segments of development, the EPA requires (1) a standard teratogenicity study with exposure during the main period of organogenesis and (2) a two-generation reproduction study.] In the early 1980s, the EPA specified the kinds of data and information required under FIFRA to support the registration of pesticides (EPA, 1982), reflecting guidelines proposed in 1978 (EPA, 1978, 1984). Similar regulations also went into effect through the EPA for chemicals under the Toxic Substances Control Act (TSCA); these were revised in 1985 (EPA, 1985). In 1986, the EPA published procedures to evaluate potential developmental toxicity associated with human exposure to environmental toxic ants (EPA, 1986). Also, a screening test for developmental neurotoxicity to include behavioral and neuropathology analyses was proposed (EPA, 1986; Francis, 1987) and finalized into test rules in 1988 and 1989 (EPA, 1988, 1989). Postnatal functional assessment has been recognized as an important part of developmental toxicity testing in the United States and is required in some cases. In other countries, requirements are in place for behavioral testing as a part of developmental toxicity testing (Barlow, 1985; EEC, 1983; Tanimura, 1985; WHO, 1986). In November 1986, voters in the state of California approved an initiative to address concerns about exposure to toxic chemicals. That initiative became the Safe Drinking Water and Toxic

14.3 Regulatory Issues Enforcement Act of 1986, better known as Proposition 65. This requires the governor to publish a list of chemicals that are known to the state to cause cancer, birth defects, or other reproductive harm. The chemicals that cause birth defects or other reproductive harm are called reproductive toxicants. The Proposition 65 list contains a wide range of chemicals, including dyes, solvents, pesticides, drugs, and food additives. If a pesticide is on the list, an employer must warn the employee if the exposure levels of the pesticide present a significant health risk; the employer may also choose to provide warning simply based on the presence of the chemical, even if the risk is not significant. In the case of worker exposure to pesticides, this warning is provided through the required hazard communication procedures, and, as an agricultural crop producer, the employer is also required to keep application-specific information on the pesticides used. A number of pesticides have been listed and subsequently withdrawn from registration for use in the state. A complete list of compounds may be accessed via the Internet at http://www.oehha.ca.gov/prop65/pdf/80400LSTA. pdf. Table 14.1 lists the currently registered active ingredients on the Proposition 65 list.

be detected in reproduction (two-generation) and developmental neurotoxicity studies. These guidelines provide information on the appropriate study design and methodology for the conduct of studies and may also be accessed via the EPA Web site at http//www.epa.gov/oppts_harmonizedl870_health_effects_test _guidelines/. Figure 14.1a and b describes the protocols for developmental and reproductive toxicity testing and notes the approximate dosing and breeding schedules. Rats: Day:

0

21 Rabbits: Day: 0

29

14.3.2 PRINCIPLES OF TESTING AND EVALUATION The Office of Prevention, Pesticides, and Toxic Substances (OPPTS) recently revised the 1982 Health Effects Test Guidelines (EPA, 1998). The OPPTS-harmonized guidelines have been developed for use in the testing of pesticides and toxic substances and the development of test data that must be submitted to the agency for review under federal regulations. The purpose of harmonizing these guidelines into a single set of OPPTS guidelines is to minimize variations among the testing procedures that must be performed to meet the data requirements of both the FIFRA (7 U.S.c. 136, et seq.), as amended by the Food Quality Protection Act (FQPA) (P.L. 104-170) and the Toxic Substances Control Act (TSCA) (15 u.s.c. 2601). The Organization of Economic Cooperation and Development (OECD) guidelines 414 and 416 and the OPPT guidelines under 40 CFR 798.4900 and 40 CFR 798.4700, OPP guidelines 833 and 83-4, provided the source material for developing these harmonized OPPTS test guidelines. Changes to the previous guidelines (EPA, 1984) were considered over a period of years and involved both industry and regulatory agencies and a period of public comment. An increase in the number of animals in the developmental toxicity study conducted in rabbits has been requested. The timing of exposure has also been extended from day 19 of gestation to the day before fetuses are examined. Endpoints previously not evaluated include the status of the ovaries, sperm/semen evaluation, and examination of vaginal smears for evaluation of estrous cycles and other effects resulting from endocrine disruption. In light of the data that the critical period for inducing abnormalities may extend to the postnatal period, for example, renal development (Couture, 1990), functional deficiencies and other postnatal effects are expected to

379

20 females per dose group (Evidence of sperm/plug in female or in bedding) Begin exposure (around implantation) generally on day 0 or 6 and continue until day before parturition C-section and examine fetuses (generally C-section is on gestation day 20 if mating is on gestation day 0) 20 females per dose group (Day of artificial insemination; natural mating can be used) Begin exposure (around implantation) generally gestation day 0 or 7 and continue until day before parturition C-section and examine fetuses

At a minimum, the test substance should be administered daily from around the time of implantation to the day before Cesarean section on the day prior to the expected day of parturition. Alternatively, if preliminary studies do not indicate a high potential for preimplantation loss, treatment may be extended to include the entire period of gestation, from fertilization to approximately 1 day prior to the expected day of parturition. It is preferred that the dams are exposed from the time of mating. The timing of implantation and expected delivery may vary with the strain. Figure 14.1a Protocol for developmental toxicity testing. Age of animals: (in weeks)

5-9

15-19

21-25 15-19

FO/Pl Start of study Exposure for 10 weeks in diet (or other route, based on the most likely human exposure scenario) Mating (conducted over 2 weeks or 3 estrous cycles; remating with a different partner is not undertaken) Gestation (approximately 3 weeks) Parturition -+ F1a Weaning (approximately 3 weeks) Growth (approximately 15 weeks) Mating Gestation (approximately 3 weeks) Parturition -+ F2a Weaning

In certain instances, such as poor reproductive performance in controls, or in the event of treatment-related alterations in litter size, the adults may be remated to produce an Fib or F2b litter. If production of a second litter is deemed necessary in either generation the dams should be remated approximately 1-2 weeks following weaning of the last F la or F2a litter. Figure 14.1b Approximate dosing and breeding schedule involved in a two-generation study of effects on the reproduction process in rats.

380

CHAPTER 14

Developmental and Reproductive Toxicology of Pesticides Table 14.1 Pesticides Known to the State to Cause Birth Defects or Reproductive Harm

Compound

CAS number

Altretamine

645056

August 20, 1999

Amitraz

33089611

March 30, 1999

Benomyl

17804352

July I, 1991

Bromacillithium salt

53404196

May 18, 1999

Bromoxynil

1689845

October I, 1990

Bromoxynil octanoate

1689992

May 18, 1999

Carbon disulfide

75150

July I, 1989

Chinomethionat (Oxythioquinox)

2439012

November 6,1998

Chlordecone (Kepone)

143500

January I, 1989 May 14, 1999

Arsenic (inorganic oxides)

Date listed

May I, 1997

May I, 1997

Cadmium

Chlorsulfuron

64902723

Cyanazine

21725462

April I, 1990

Cycloate

1134232

March 19, 1999

Cyhexatin

13121705

January I, 1989

2,4-D butyric acid

94826

June 18, 1999

2,4-DP (dichloroprop)

120365

April 27, 1999

1,2-Dibromo-3-chloropropane (DBCP)

96128

February 27, 1987

p'-DDT

789026

May 15, 1998

p, p'-DDT

50293

May 15, 1998

Dichlorophene

97234

April 27, 1999

Diclofop methyl

51338273

March 5, 1999

Dicumarol

66762

October I, 1992

Dinocap

39300453

April I, 1990

Dinoseb

88857

January I, 1989

Endrin

72208

May 15, 1998

Epichlorohydrin

106898

September I, 1996

Ethyl dipropylthiocarbamate

759944

April 27, 1999

Ethylene dibromide

106934

May 15, 1998

0,

Ethylene oxide

75218

February 27, 1987

Ethylene thiourea

96457

January I, 1993

Etretinate

54350480

July I, 1987

Fenoxaprop ethyl

66441234

March 26, 1999

Fluvalinate

69409945

November 6, 1998

Heptachlor

76448

August 20, 1999

Hexachlorobenzene

118741

January I, 1989

Hydramethylnon

67485294

March 5, 1999

Hydroxyurea

127071

May I, 1997

Linuron

330552

March 19, 1999

Mebendazole

31431397

February 27, 1987

Lead

August 20, 1999 July I, 1990

Mercury and mercury compounds Metham sodium

137428

May 15, 1998

Methotrexate

59052

January I, 1989

Methotrexate sodium

15475566

April I, 1990

Methyl bromide as a structural fumigant

74839

January I, 1993 July I, 1987

Methyl mercury

March 30, 1999

Metiram

9006422

Myclobutanil

88671890

April 16, 1999

Nabam

142596

March 30, 1999

Nicotine

54115

April I, 1990 (continues)

14.3 Regulatory Issues

381

Table 14.1 (continued) Compound

CAS number

Date listed

Oxadiazon

19666309

May 15, 1998

Oxydemeton methyl

301122

November 6,1998

Oxymetholone

434071

May 1, 1997 January 1, 1991

Polychlorinated biphenyls (as a contaminant) Potassium dimethyldithiocarbamate

128030

March 30, 1999

Propargite

2312358

June 15, 1999 January 29, 1999

Pyrimethamine

58140

Resmethrin

10453868

November 6, 1998

Sodium dimethyldithiocarbamate

128041

March 30, 1999

Sodium fiuoroacetate

62748

November 6,1998

Streptomycin sulfate

3810740

January 1, 1991

2,3,7,8-Tetrachlorodibenzo-para-dioxin

1746016

April 1, 1991

Thiophanate methyl

23564058

May 18, 1999

Triadimefon

43121433

March 30, 1999

Tributyltin methacrylate

2155706

December 1, 1999

Triforine

26644462,37273840

June 18, 1999

Vinclozolin

50471448

May 15, 1998

Warfarin

81812

July 1, 1987

(TCDD) (as a contaminant)

Updated November 26,1999.

14.3.3 CHOICE OF SPECIES IN TESTING The laboratory species typically used to test for developmental toxicity or for reproductive effects is the rat. Some strains are considered less suitable than others, and the rationale for the strain may vary with the compound and the effects it may cause in the species tested. The rabbit is the other species that is used as it is the one species (unlike the rat) that showed some signs for the compound thalidomide, the chemical that appeared safe in all other species tested. Additionally, because of the species specificity of teratogenic agents, the exact effects noted in laboratory animals are not necessarily those observed in humans. However, all proven human teratogens have parallel, but imperfect animal models. Determining which species is the most appropriate for extrapolation to humans for a given compound is difficult. Pesticides that involve food use are likely to have a higher potential exposure and are to be tested in two species as per FIFRA regulations. Among the species used for testing, the rat and mouse most successfully model the human reaction, but the rabbit is less likely than other species to give a falsepositive finding (EPA, 1991). The concomitant use of the rabbit with either the mouse or the rat is believed to enhance the predictive potential of the individual animal model (Schardein and Keller, 1989). Accordingly, the rat and rabbit are commonly used. Although no single species has clearly distinguished itself as being more advantageous in the detection of human teratogens over any other, it is concluded that safety decisions should be based on all reproductive and developmental toxicity data

in light of the agent's known pharmacokinetic, metabolic, and toxicologic profile.

14.3.4 CHOICE OF TESTING DOSES While reviewing studies submitted for regulatory purposes, a number of factors are taken into consideration. A compound may be embryolethal without being teratogenic; alternatively, it can also be both embryolethal and teratogenic. The teratogenic dose range, that is, the margin between the dose which will kill the fetus and that which does not have adverse effects on the fetus, is often very narrow. The wider the margin, the more potentially dangerous is the compound from a teratological perspective. Hence, it is recommended that the highest dose should only cause slight toxic effects on the pregnant animal (e.g., decreased body weight gain) such that a majority of the pregnancies reach term (Wilson, 1979). Choice of dosing regimens is critical to determining the potential of the compound to exert adverse effects. The mid-dose must not be much lower than the high dose as such a bracketing will result in a low no observed effect level (NOEL) and not provide data on the true nature of the chemical being evaluated. Although a low NOEL may appear to be more health protective than a higher NOEL, it is possible to miss the effects that the compound can cause at levels below the maximum tolerated dose. The choice of dose levels is critical to study design. Studies submitted with inappropriate doses are often unacceptable to regulatory agencies

382

CHAPTER 14


and the registrant has to either conduct a new study or provide justification to support the choice of doses employed. 14.3.5 INTERPRETING EFFECTS

Death of the conceptus may preclude expression of other major manifestations of developmental toxicity (i.e., structural abnormalities, altered growth, and functional deficit). Generally, the term teratogenicity refers to the observation of malformations, that is, permanent structural changes that may adversely affect survival, development, or function. Other developmental effects include variations, a term used to describe changes in the fetus that involve a divergence beyond the usual range of structural constitution that may not adversely affect survival or health. It is sometimes difficult to distinguish between malformations and variations because the responses constitute a continuum from normal to the extremely deviant. Other terms used are anomalies, deformations, and aberrations; however, they are not defined any better (EPA, 1991). To further confuse this already complex issue, these other terms may be used for either of these two categories, requiring a closer examination of the interpretation. Evaluating variations and interpreting their incidence is illustrated in the case of supernumerary ribs (SNRs). These are a common variant in some strains of mice used in standard teratology bioassays, and the increased incidence of SNRs may be induced by a wide variety of xenobiotics and/or general maternal stress. The significance of this defect in cross-species extrapolations has been problematic. In one study in mice, it was demonstrated that SNRs have a bimodal distribution composed of "rudimentary ribs" (RRs) with a mode of 0.3-0.4 mm and "extra ribs" (ERs) with a mode of 0.9-1.1 mm. The ERs and RRs were found to be morphologically distinct; the ERs were flat ended and distally joined by a cartilaginous portion, whereas the RRs were usually rounded distally and were without cartilaginous extensions. The 13th ribs were significantly longer in fetuses having SNRs than in those not having SNRs, whether treated or untreated. This relationship was present in all fetal ages examined and with both ER and RR groups, suggesting that SNRs are indicative of basic alterations in the development of the axial skeleton (Branch et aI., 1996). In the case of developmental toxicity, studies are reviewed taking into account the maternal effects observed. Developmental toxic effects in the presence of severe maternal toxicity are considered less severe than those observed in the absence of maternal effects. Generally, in order to determine whether or not the conceptus is uniquely susceptible, the developmental and maternal NOEL values are compared. The AID ratio (adult NOEL:developmental NOEL) has been advocated previously as an index of comparative teratogenic hazard (John son, 1981) and has been used to characterize the developmental effects of chemicals. The strategy of carefully characterizing the observed maternal toxicity at the individual level is also employed. To determine if the malformations observed were the result of maternal toxicity, two approaches may be adopted: consideration

of individual versus group mean data and examination of data during the specific period of gestation when the developmental malformations were likely to have occurred. Furthermore, because mere correlation of maternal toxicity with fetal effects does not imply causality (Chernoff et aI., 1987), maternal influence may not necessarily be the underlying mechanism of action. The severity of the effect on the fetus also needs to be considered; that is, the effect may be severe/life-threatening, whereas the maternal effects such as slight weight loss are minor or transient. Another confounding factor is that maternal effects may be reversible, whereas effects on the developing fetus may be permanent, underscoring the importance of characterization of the maternal effects. Examining the data in this manner leads to a more exacting interpretation of teratogenic potential. The use of a weight-of-evidence approach may help in verifying the apparent maternal toxicity at the individual level and in determining the developmental toxic potential of a chemical (lyer et aI., 1999). A weight-of-evidence approach should include the following: • Dose-response relationship • Supportive evidence in another species or related compounds • Closer scrutiny, focusing on individual data during the discrete time period(s) in which particular fetal malformations were most likely to have occurred In conducting the risk assessment, the developmental toxicity study serves as a surrogate for an acute toxicity study, based on the premise that the effects noted may be the result of a single exposure or an exposure over a short period. However, if the effects responsible for the NOAEL (no observed effect level)/RfD (reference dose) are known to result from multiple exposures, then use of the developmental toxicity study for acute effects would be inappropriate. Similar approaches are recommended for reproduction (twogeneration) and developmental neurotoxicity studies. Effects on lactation, acceptance of offspring, and sexual maturation are examined in the reproduction studies. Multigeneration studies are evaluated to determine if the effect noted is exacerbated in subsequent generations. Furthermore, spontaneous occurrence in control animals of stillborn pups and other developmental effects necessitate that evaluation of the data be subjected to rigorous statistical procedures. 14.3.6 STATISTICAL EVALUATION

One of the most important aspects of developmental toxicity analysis is that the litter is to be considered the experimental unit (EPA, 1991). Since it is the maternal unit that is exposed to the compound, the effects of the test substance on each fetus in a litter are related to the status of the animal bearing that fetus. Individual differences in maternal susceptibility can affect an entire litter, whereas others in the same dose group are unaffected. Hence, all fetuses in a single dose group are not

14.4 Toxicology Studies

equally at risk to the potential developmental effects of the test substance. Therefore, the accepted practice is to consider the litter as the experimental unit for developmental toxicity studies (Collins et aI., 1999; EPA, 1991; Gad and Weil, 1986; Gaylor, 1978). Evaluating the overall fetal effects helps further characterize the developmental toxic effects, but the litter is the preferred unit to evaluate the effects of the compound. Using model-fitting techniques and employing benchmark doses are encouraged if they are appropriate. Concurrent controls are the group of choice for comparison. Historical controls are recommended to serve as supportive evidence. 14.3.7 EXPOSURE ASSESSMENT

In evaluating the exposure for a given chemical, regulatory agencies take into consideration the amount used and the usage pattern (seasonality, etc.) and attempt to obtain the dose that reaches either the parent's germ cells or the developing conceptus. Although developmental toxicity is usually thought to be associated with maternal/embryonic exposure, there is increasing evidence for developmental effects due to male exposure (Colie, 1991; Sever, 1995). Agents associated with spontaneous abortions may also cause congenital malformations with the appropriate timing and dose; hence, the exposure pattern may determine the continuum of effects that might result. Additionally, it is thought that the steady accumulation of pesticides in the adipose tissues during a woman's lifetime may pose a risk, especially in the case of endocrine disrupting chemicals (Garcia-Rodriguez et aI., 1996). Issues associated with the importance of the timing and assessment of exposure during pregnancy have been discussed extensively (Hertz-Picciotto et aI., 1996). Knowledge of the time window of vulnerability has important implications for the assessment of risks. Along with the active ingredient, organic solvents are used extensively in pesticide formulations, and hence mixers and loaders of pesticides may be exposed to higher levels of both the active ingredient and the solvents/inerts. Effects caused by solvents could confound the issue and may impact exposure in an adverse manner. 14.3.8 IMPACT OF THE FOOD QUALITY PROTECTION ACT ON

DEVELOPMENTAL AND REPRODUCTIVE TOXIC EFFECTS OF PESTICIDES

In 1996, Congress passed landmark pesticide food safety legislation supported by the administration and a broad coalition of environmental, public health, agricultural, and industry groups. The bill was signed by the president on August 3, 1996, and the Food Quality Protection Act of 1996 became law (P.L. 104-170, formerly known as H.R. 1627). The EPA regulates pesticides under two major federal statutes. Under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA), the EPA registers pesticides for use in the United States and prescribes labeling and other regulatory requirements to prevent

383

unreasonable adverse effects on health or the environment. Under the Federal Food, Drug, and Cosmetic Act (FFDCA), the EPA establishes tolerances (maximum legally permissible levels) for pesticide residues in food. The Food Quality Protection Act (FQPA) amendments to the FFDCA direct the EPA to consider a number of factors in making risk assessments as part of the tolerance-setting procedure. Most of these provisions originated in recommendations from the National Academy of Sciences (NAS) 1993 report "Pesticides in the Diets of Infants and Children" and reflect concerns that children may be especially susceptible to pesticide exposure. Specifically, in setting a tolerance for pesticide residues in food, the FQPA directs the EPA to consider the following: use of an extra lO-fold safety factor to account for the susceptibility of children; the special susceptibility of children, including effects of in utero exposure; cumulative effects of exposure to the pesticide and substances having a common mode of action; aggregate exposure for all consumers (i.e., other routes, such as drinking water and home and garden applications); and potential for endocrine disrupting effects. Incorporating these factors into the tolerance-setting process poses significant challenges to the agency because there are many scientific uncertainties surrounding the use of these factors in risk assessment. To this end, specific areas are brought to the Scientific Advisory Panel (SAP) for review. Current practice is generally to use a WO-fold safety factor when the toxicity data are from animal studies and to apply extra factors of 3to lO-fold only when specific case-by-case evidence seems to warrant it. A presentation to the SAP, a position paper on the subject has been released (EPA, 1996b). The 1996 law represents a breakthrough, amending both major pesticide laws to establish a more consistent, protective regulatory scheme. It mandates a single, health-based standard for all pesticides in all foods; provides special protections for infants and children; expedites approval of safer pesticides; creates incentives for the development and maintenance of effective crop protection tools for U.S. farmers; and requires periodic reevaluation of pesticide registrations and tolerances to ensure that the scientific data supporting pesticide registrations will remain up to date in the future. Additional information may be accessed via the Internet at www.epa.gov/oppfeadl/fqpa/sciissue.htm.Awider perspective on the appropriateness of extra safety factors to protect children based on the use of one pesticide (chlorpyrifos) can be gleaned from other publications on the subject (Gibson et aI., 1999; Schardein and Scialli, 1999) including the risk characterization document for chlorpyrifos (California Department of Pesticide Regulation, 2000).

14.4 TOXICOLOGY STUDIES Pesticides can be broadly classified into different categories, based on the target of use. The major classes that will be discussed include herbicides, insecticides (including insect growth regulators), fungicides, and rodenticides. In addition, compounds that have pesticidal action (antiparasitic) are used as

384

CHAPTER 14 Developmental and Reproductive Toxicology of Pesticides

animal health products and will be discussed under a miscellaneous category. 14.4.1 HERBICIDES Included in this class of chemicals are the chlorphenoxy compounds, bipyridyls, dinitrophenols, triazines, substituted ureas, some of the carbamates, plant growth inhibitors, and amides. A number of herbicides have been studied for developmentally toxic effects and adverse effects on reproduction in animals (see Table 14.2). Chlorphenoxy Compounds Of the various herbicides used, a plethora of data is available on the phenoxy defoliants 2,4-dichlorophenoxyacetic acid (2,4-D) and 2,4,5-trichlorophenoxy ace acid (2,4,5-T) which have been used worldwide in forestry and agriculture. The phenoxy herbicide, Agent Orange, composed of equal parts of 2,4-D and 2,4,5-T, has received a lot of attention after its large-scale use in Vietnam by the U.S. military during the war years of 1962-1971. The contamination of these compounds with the dioxin TCDD in commercial preparations has further confounded the effects noted. TCDD is found to be extremely teratogenic in laboratory animals: It has a very low teratogenic minimal effective dose of 1-10 !l-g/kg in susceptible strains of mice (Peam, 1985), and effects such as hydronephrosis and cleft palate have been noted (Couture et aI., 1990). Results from Agent Orange exposure are largely inconclusive and more details, including the litigation history may be obtained in previous reports (Schardein, 1993; Schuck, 1986). The phenoxy herbicide 2,4,5-T has not been manufactured since 1983. 2,4-D One case of cephalic malformations and severe mental retardation was noted in an infant whose parents received prolonged exposure via the dermal route from forest spraying (Casey and Collie, 1984). Though no birth defects were found, an increase in spontaneous abortions and premature births was noted in a case-control study examining the effects of 2,4-D (Carmelli et aI., 1981). Possible adverse effects such as gestational and neonatallosses with NOELs of 20 mg/kg/day (rats) and 500 ppm (dogs) were noted in two studies (California Department of Pesticide Regulation, 2000). Markedly reduced gestational and neonatal survival was not accompanied by a commensurate degree of maternal toxicity. In the Ontario Farm Family Health Study exposure to phenoxy herbicides during the first trimester was generally not associated with increased risk of spontaneous abortion (Arbuckle et aI., 1999). However, the results suggest a possible role of preconception (possibly paternal) exposures to phenoxy herbicides in the risk of early spontaneous abortions. Of other phenoxy herbicides that have been studied, 4-chloro-2-methylphenoxyacetic acid ethyl ester caused (31 % incidence) cleft palate and anomalies of the heart and kidney in rats (Schardein, 1993). 4-chloro-o-toloxy acetic acid was also found to be teratogenic in mice and rats at high oral doses (Roll and Matthiaschk, 1983; Schardein, 1993).

Amitrole This nonselective postemergence herbicide is also an antithyroid agent. Although structural malformations were not noted when tested in animal studies, fetal thyroid lesions were observed in rats exposed to amitrole via drinking water (Schardein, 1993). Bromoxynil The developmental toxicity of the wide-spectrum herbicide bromoxynil (bromoxynil phenol; 3,5-dibromo-4-hydroxyphenyl cyanide) and its octanoate ester (2,6-dibromo-4-cyanophenyl octanoate) was evaluated in Sprague-Dawley rats and Swiss-Webster mice. The highest doses of both compounds increased the incidence of supernumerary ribs (SNRs) in the fetuses of treated rats, but did not induce other anomalies (Rogers et aI., 1991). In the teratogenicity study submitted to the California Department of Pesticide Regulation (CDPR), rats were dermally exposed to Buctril (containing 33.8% bromoxynil octanoate) diluted with water. Based on a dose-dependent increase in the incidence of extra thoracic ribs in the fetuses, at the 15 mg/kg/day level and above, the developmental NOEL was determined to be 10 mg/kg/day. Dinoseb Dinoseb (2-sec-butyl-4,6-dinitrophenol) has been shown to produce substantial spermatotoxicity after 1-5 doses in short-duration tests (Linder et aI., 1992). In mice at 17.7 mg/kg/day subcutaneous or intraperitoneal administration of dinoseb during organogenesis resulted in skeletal defects, cleft palate, hydrocephalus, and adrenal agenesis. Maternal toxicity, however, was noted at doses between 17.7-20 mg/kg/day (Gibson, 1973). Dinoseb has also been reported to produce a high incidence of dilated renal pelvis in the term rat fetus (McCormack et aI., 1980) as well as SNRs in mice (Kavlock et aI., 1985). Teratogenic effects such as an increased incidence of microphthalmia were also reported in the rat fed dinoseb in the diet (Giavini et aI., 1986). Eye defects and neural malformations were noted in the rabbit, leading to dinoseb's being banned by the EPA in 1986. Bipyridyl Compounds (Paraquat, Diquat) The herbicide paraquat has resulted in at least eight fetal deaths when taken during pregnancy as a result of maternal poisoning (Talbot et aI., 1988). However, no adverse effects were reported in animal developmental and reproductive toxicity studies submitted to the CDPR. On the other hand, for the herbicide diquat, adverse systemic effects were noted in the rat in parents and offspring (cataracts and eye pathologies in both sexes of Fo and Fl at more than 240 ppm; an increase of hypertrophy and hyperplasia of collecting duct epithelium and tubular dilatation in the renal papilla in both sexes of Fl at 240 ppm; Fl and F2 pups showed hydronephrosis at and above 240 ppm). In data submitted to the CDPR for the rat, the developmental NOEL and NOAEL were equal to 12 mg/kg/day with intrauterine growth retardation, as measured by decreased weight and delayed skeletal ossification, and hemorrhagic kidneys as the main effects observed (CD PR, 2000). Mice appear to be more sensitive to diquat than were rats. The NOEL in the mouse


385

Table 14.2 Developmental and Reproductive Toxicity Profile of Herbicidesa

Dose Chemical

Species

Toxicity profile

Acrolein

Rat

Reproduction

(mg/kg)b

3:pup l:parent

Alachlor (ethane

Rat

Comments

References

Decreased body

CDPR Database, 2000

weight in pups Heydens et ai., 1996

1000

sulfonate) Ametryn

Rat

Increased skeletal

50

Infuma et ai., 1987

variants

As cited in Schardein, 1993

Reproduction Rabbit Amitraz Amitrole

Rat

60 Reproduction

Mouse

Altered estrous cycles

Rat

Thyroid effects

Rabbit

Abortions, reduced weight gain

Infuma et aI., 1987

10.5 ppm

Pup mortality

CDPR Database, 2000

0.0004%

Via drinking


DPN #287,1994 water (3-G study) 4

DPN#0216

CDPR Database, 2000 (worksheet 033 45711)

Arsenic

Teratogenic: malformations

Arsenic acid

Mouse

Sodium arsenate

RatIMouse

Asu1am

Rat

CDPR Database, 2000 7.5

DPN #180 Parenterally (ip)

Reproduction: decreased

1000 ppm

CDPR Worksheet (360) 010

number of live births Atrazine

Rat

Fetal toxicity

As cited in Schardein, 1993 25257; 19

70

2-G study

Disruption of ovarian

Infuma et ai., 1987 Cooper et al., 1996

cycle, induced pseudopregnancy Reproduction Balagrin

Rabbit

Fetal toxicity

75

Infuma et ai., 1986

Mouse

Teratogenic

22


Rat

Teratogenic

1/50 LDsO 75

Mirkova, 1980 C1opyra1id

Rat

Delayed ossification Reproduction

Rabbit

Selectively toxic

Hayes et aI., 1984

(2-G study)

Dietz et aI., 1986

250

Cyanazine

Rat

Developmentally toxic

2,4-D

Mouse


Hayes et ai., 1984 Anophthalmia!

Lu et ai., 1982

microphthalmia 221


Teratogenic Rat

Fetal death

50

Teratogenic

2-G study

Reproduction 2,4-D + picloram

Hamster

Teratogenic

Mouse


20 0.2%

Teratogenic

Drinking water

Blakley et ai., 1989

route

2,4-D + 2,4,5-T

Rat

Behavioral effects

2,4-D butoxyethanol

Rat

Teratogenic

150

Rat

Teratogenic

150


Sheep

Teratogenic

3kg


50/125

Mohammad and St. Omer, 1988 As cited in Schardein, 1993

ester Rabbit 2,4-D butyiester

75

Sterility, fetal death 2,4-D diethylamine

Rat

Bifenox

Mouse

Teratogenic

Liberacki et aI., 1994

3%/10 ha pasture 0.5 LDSO 100

As cited in Schardein, 1993 Francis, 1986

(continues)

386

CHAPTER 14


Table 14.2 (continued) Dose Chemical

Species

Bromacil

Rat Rabbit

Bromoxynil

Toxicity profile

(mg/kg)b

Comments

References

250 ppm

Diet


250ppm

Diet

Mouse

Developmentally toxicC

96.4

Rat

Skeletal variation

15

Butiphos

Rat


Buturon

Mouse

Increased mortality

Chloramben

Rabbit

12.5

Rogers et aI., 1991 As cited in Schardein, 1993

100

Teratogenic Intrauterine growth

500

retardation

CDPR Database, 2000 (Worksheet 266 011 36993)

Chloridazon

Rat

Resorption

1/50 LDSO


4-Chloro-2-methyl

Rat

Teratogenic

100


phenoxyacetic acid ethyl ester 2-Chlorophenyl-4' -

Mouse

1000

Francis, 1990

Mouse

1000

Francis, 1990

3000

Tanaka et aI., 1997

nitropheny I ether 2-Chlorophenyl-4' nitrophenyl ether Chloroprophan

Mouse

2,4-D isooctyl ester

Rat

2,4-D

Rabbit

Developmentally toxic Teratogenic

750

Teratogenic

150


75


87.5


75


500

CDPR Database, 2000

isopropylamine 2,4-D propylene

Rat


glycol butyl ether ester 2,4-D

Rabbit

triisopropanolamine Dalapon

Rat

Daminozide

Rat

Diallate

Rabbit

Skeletal effects

(Worksheet 006 036526) 1000 10

Khera et aI., 1979b As cited in Schardein, 1993

2,4-DM

Rat

2,5-Dichlorophenyl-

Mouse

1000

Francis, 1990

Mouse

1000

Francis, 1990

Mouse

500

Francis, 1990

Mouse

1000

Francis, 1990

Mouse

400

Francis, 1990

400

Roll and Matthiaschk, 1983


3.4

4' -nitrophenyl ether 3,4 dichlorophenyl4' -nitrophenyl ether 2,6Dichlorophenyl-4' nitropheny I ether 2,5Dichlorophenyl-4' nitrophenyl ether 2,3Dichlorophenyl-4' nitropheny I ether Dichloroprop

Mouse

Teratogenic

Rat

Postnatal behavioral

5

Buschmann et aI., 1986

effects Dicotex

Rat

Dicuran

Rat

Teratogenic

20


5000 (continues)


38',


Dose Chemical

Species

Toxicity profile

Dinoseb

Rat

Teratogenic

(mg/kg)b

200ppm

Reproductive system

Comments

References

Diet, only

Giavini et aI., 1986

developmental


toxic by po route Diuron

Rat

500

Endothall

Rat

25

EPTC

Rat

1/20 LDso

Ethalfluralin

Rat

1000

Hamster

Teratogenic

Byrd et aI., 1990a Minta and Biernacki, 1981

20 40

Rabbit Rabbit

Increased resorptions,

30

CDPR Toxicology Summary,

delayed ossification Fiuoxypyrmethyl-


400

Rat Ethofumesate

1/ 2OLDSO

300

Rabbit Ethephon

As cited in Schardein, 1993 Trutter et aI., 1995

Rat

Skeletal variations

Rat

Reproductive system

1993 600

Carney et aI., 1995

hepty I ester Hexazinone

5000 ppm

Ioxynil octanoate

Mouse

100 ppm

Lenacil

Dog

500 ppm

Linuron

Diet; 3-G study

Rat

Reproductive system

Rat

Teratogenic

As cited in Schardein, 1993 Diet 3-G study

200

Reproductive system

Khera et aI., 1978 3-G study

100ppm 125 ppm

Rabbit Maleic hydrazide MCPA

Mecoprop

Embryotoxic

1/2 LDso

Teratogenic

1/2 LDso

Mouse

Teratogenic

200

Mouse

Teratogenic

400

Rat

Postnatal toxicity

Meturin

Rat

Molinate

Rat

Teratogenic: increased

Diet

Reproduction: sperm abnormalities, detached

As cited in Schardein, 1993 Khera et aI., 1979b As cited in Schardein, 1993 Roll and Matthiaschk, 1983 Roll and Matthiaschk, 1983 Buschmann et aI., 1986

13


1/10 LDsO 35

CDPR Toxicology Summary,

resorptions, intrauterine growth retardation Rat

As cited in Schardein, 1993 CDPR,1993

1600

Rat Rat

Kennedy and Kaplan, 1984

125

Rabbit

1998 DPN #228 5 ppm males

Jewell and Miller, 1998

20 ppm females

heads, ovarian interstitial tissue vacuolation 3-Monochlorophenyl-

Mouse

1000

Francis, 1990

4'-nitrophenyl ether Monolinuran

Mouse

Naphoxyacetic acid

Rat

Nitrofen

Mouse

Mortality

25

Teratogenic

25

Growth retardation


250

Henwood et aI., 1990

250

Nakao et aI., 1981

Teratogenic Rat

Teratogenic

121

dermal/oral

Costlow et aI., 1983

3-G & fertility Reproductive system Hamster

Teratogenic

studies 400

Gray et aI., 1985 (continues)

388

CHAPTER 14



Species

Toxicity profile

Oxadiazon

Rat

Teratogenic:

(mg/kg)b

Comments

References CDPR,2000

12

postimplantation loss incomplete ossification Rabbit

Teratogenic: resorptions,

60

constraint-related arthrogryposis Oxyfluorfen

Rat

Teratogenic: early

CDPR,2000

15

resorptions, decreased fetal weight, skeletal malformations and variations As cited in Schardein, 1993

Paraquat

Mouse

1,2,3,7,8-Penta-

Mouse

Teratogenic

4000

100

Bimbaum et aI., 1991

Mouse

Teratogenic

2400

Bimbaum et aI., 1991

Rat

Fetotoxicity

250

Rabbit

Fetotoxicity

20

bromodibenzofuran 1,3,4,7,8-Pentabromodibenzofuran Phosphinothricin Pichloram Pichloram

Ebert et aI., 1990 Breslin et aI., 1991

1000

Rat Rabbit

400

lohn-Greene et aI., 1985

Rabbit

500

Breslin et aI., 1994

1000

Breslin et aI., 1994

ethylhexyl ester Pichloram

Rabbit

Abortion

Prometryn

Rat

Teratogenic

Propach10r

Rat

Equivocally teratogenic

1/5 LDSO 1/5 LDSO 0.2 mgim 3

triisopropanolamine

Propazine

Rat

Decreased fetal weight

Simazine

Rat

Teratogenic

SLA 3992

Rat

Teratogenic

20

Rabbit

Teratogenic

20

2,4,5-T

Mouse

Teratogenic

15

Rat

Teratogenic

50

Mirkova and Ivanov, 1981 Machemer et aI., 1992 As cited in Schardein, 1993

40 Teratogenic

20 113

Sheep Primate

Inhalation route

3-G study

Reproductive system Rabbit Hamster


25

Growth retardation

40

Wilson, 1971 As cited in Schardein, 1993

Abortion 2,4,5-T butyl ester

Rat

Teratogenic

50

Mouse

Teratogenic

74

2,4,5-T phenol

Mouse

2,4,5-T propylene

Sheep

9 100

glycol butyl ether ester Tebuthiuton

Rat

Triallate

Rabbit

Trichloroacetic acid

Rat

1800 ppm

Diet

10 Developmentally toxic

330

Teratogenic

330

Smith et aI., 1988

(continues)


389


Dose Chemical

Species

Toxicity profile

Trichloropyr

Rat

Reproductive system,

(mg/kg)h

250

fetotoxicity (FI and Fz gen) Trichloropyr

Comments

References

reduced litter size

CDPR,2000

and pup weight

Rat


300

Breslin et aI., 1996a

Rat


300

Breslin et aI., 1996a

Resorption

250

Hanley et aI., 1987

butoxyethyl ester Trichloropyr triethylamine Tridiphane

Mouse

Teratogenic Rat

Skeletal variants

Hanley et aI., 1987

100

Reproductive system

2-G and repro

Rao et aI., 1986

studies Trifluralin

Triisopropanolamine

Mouse

Skeletal variation

Rat

Depressed fetal weight

1000

Rabbit


500

Rat

Beck, 1977

1000

Byrd and Markham, 1990 Breslin et aI., 1991

a Includes

plant growth regulators. bDose is oral unless stated otherwise; dose is the LOEL wherever effects were observed and the NOEL when there were no effects. C Developmental toxicity includes reduced fetal weight, increased embryo/fetal mortality, and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate the presence of malformations.

for both maternal toxicity (clinical signs, death) and developmental toxicity (skeletal anomalies, exencephaly, and umbilical hernia) was 1.0 mg/kg/day. However, the rabbit appeared to be the most sensitive laboratory animal to diquat in developmental toxicity studies. The NOEL for maternal toxicity (histopathological changes in the liver, intestine, and vasulature; mortality) was 3.0 mg/kg/day, but the developmental NOEL was below 1.0 mg/kg/day. Delayed ossification of the ventral tubercle of the cervical vertebrae was noted in all treatment groups compared to the controls. The incidence of fetal malformations was significantly greater in the low dose (1.0 mg/kg/day) and the high dose (10 mg/kg/day) compared to the controls; although the mid-dose lacked statistical significance, it appeared biologically significant (more than a two-fold increase over controls) and hence supportive of a treatment-related effect. Ethyl dipropylthiocarbamate (EPTC) The thiocarbamate class of pesticides has been shown to cause a wide range of effects. (EPTC) was determined to cause adverse developmental effects in the rat with a NOEL of 30 mg/kg/day, based on increased resorptions at levels below the maternal NOEL (100 mg/kg) (CDPR, 2000). Ethofumesate In a developmental tOXICIty study in rabbits, ethofumesate was determined to have adverse effects with a developmental NOEL of 30 mg/kg/day, based on increased resorptions and delayed ossification noted at the higher doses (300 and 3000 mg/kg/day) tested. Maternal toxic effects such as abortions and death were noted at the high

dose of 3000 mg/kg/day, resulting in a maternal NOEL of 300 mg/kg/day (CD PR, 2000). Molinate Results from several studies on the herbicide molinate have consistently demonstrated that exposure of male laboratory animals to the compound via the oral/inhalation route causes a decrease in fertility, abnormal sperm morphology, decreased epididymal sperm number, and/or testicular degeneration. Unexposed females mated to exposed males (rabbits/mice/rats) had significant (p < 0.05) preimplantation loss, possibly a result of the inability of the sperm to fertilize the ova. Female rats and mice exposed to molinate in the diet also exhibited significantly (p < 0.05) reduced litter sizes, along with histopathological abnormalities in the ovaries such as vacuolation and hypertrophy of the thecal/interstitial cells (CD PR, 2000). Recent data do suggest that humans are probably less sensitive and less likely than rats to experience the reproductive toxicity of molinate largely due to unequal rates of metabolism of molinate to molinate sulfoxide (Jewell and Miller, 1998). However, the relative degree of risk cannot be quantified at this time. In the rat, molinate demonstrated adverse effects such as increased resorptions and intrauterine growth prior to the onset of maternal toxicity with a developmental NOEL of 35 mg/kg/day. Nitrofen This pre- or postemergence herbicide induced a high incidence of diaphragmatic hernia and harderian gland alterations in mice fetuses subsequent to maternal oral exposure (Gray et al., 1983; Nakao et al., 1981). In rat studies, hydronephrosis and respiratory problems were noted (Costlow

390

CHAPTER 14


and Manson, 1980), whereas eye abnormalities were noted following percutaneous exposure to the dams (Francis and Metcalf, 1982). Exposure during only two gestational days altered the development of the para- and mesonephric ducts, resulting in renal malformations in females and agenesis of the vas, epididymis, and seminal vesicles in males (Gray et aI., 1985). The teratogenic activity of nitrofen has been attributed to alterations in maternal and fetal thyroid hormone status (Manson et aI., 1984). Triazines This class of compounds is heavily used throughout the world. They include herbicides such as atrazine, cyanazine, propazine, and simazine and the insecticide cyromazine. Eye defects such as anophthalmia, cryptophthalmia, microphthalmia, and cyclopia have been noted in animal studies for some of these compounds (CD PR, 2000). In a reproduction study in rats exposed to cyanazine, the significant toxicological finding was decreased pup viability: F1a pups at 250 ppm on day 21 and F2a pups at 150 and 250 ppm on day 4 (NOEL = 75 ppm). A possible adverse effect was indicated (pup NOEL < adult NOEL). In the rat teratogenicity study, the developmental toxicity NOEL was 5 mg/kg/day (increased number of fetuses and pups with microphthalmia or anophthalmia at 25 and 75 mg, decreased litter size and weight at 75 mg, increased total litter resorptions at 25 and 75 mg/kg, and decreased live litter size and survival to day 21 of lactation at 75 mg). Because developmental toxicity was seen at levels of cyanazine causing only slight maternal toxicity, the effects were considered adverse (CDPR, 2000; Iyer et al., 1999). Urea Herbicides Several urea herbicides induce genetic abnormalities in standard tests for genotoxic potential. They are generally the phenylureas and the effects of some of these compounds are detailed next. Diuron A widely used substituted urea herbicide, diuron induced wavy ribs at doses of 250 and 500 mg/kg/day (mid- and high dose) in rats. Ossification of the calvarium was delayed in fetuses of dams that received 125 mg/kg with the study yielding no NOEL (Khera et aI., 1979c). However, diuron did not produce any adverse effects for either reproduction or teratogenicity in studies submitted for registration (CDPR, 2000). Isoproturon Reproductive abnormalities, particularly those affecting sperm morphology and function, were noted in rats exposed to isoproturon (Behera and Bhunya, 1990). Maturational malformation of sperm and retarded spermatogenesis were also observed (Sarkar et aI., 1997). Linuron An antiandrogenic pesticide, linuron has been shown to induce a level of external effects consistent with its low affinity for the androgen receptor (AR), resulting in reduced anogenital distance, retained nipples, and a low incidence of hypospadias as well as malformed epididymides and testis atrophy (Gray et aI., 1999b). Additionally, linuron may produce Leydig cell tumors via an antiandrogenic mechanism where sustained

hypersecretion of luteinizing hormone (LH) appears to be responsible for the development of Leydig cell hyperplasia and adenomas (Cook et aI., 1993). Linuron may display several mechanisms of endocrine toxicity, one of which involves AR binding (Gray et aI., 1999b). Linuron produced malformations in the rat at 100 mg/kg/day but did not demonstrate teratogenic potential in the rabbit. Monolinuron, a related compound, has been shown to cause cleft palate in the mouse (Schardein, 1993). 14.4.2 INSECTICIDES

Included in this class of chemicals are the organophosphates, organochlorines, chlorinated cyclodienes, and carbamate esters. Many insecticides have demonstrated developmentally toxic effects and adverse effects on reproduction in animals (see Table 14.3). 14.4.2.1 Organochlorines, Including Chlorinated Cyclodier Aldrin Prenatal exposure to aldrin induced developmental changes (a decrease in the median effective time for incisor teeth eruption and an increase in the median effective time for testes descent) in rat pups and persistent behavioral alterations (the locomotor frequency of the experimental rats was higher than that of the controls at 21 and 90 days old) in adults after pregnant rats were subcutaneously treated with aldrin (1.0 mg/kg) or with its vehicle (0.9% NaCI solution plus Tween-80) from day 1 of pregnancy until delivery (Castro et aI., 1992). Aldrin may also have a direct inhibitory influence on gonadotrophin release and may exert a direct action on the testes (Chatterjee et aI., 1988). A review of the developmental toxicity of aldrin concludes that aldrin induces malformations (eye and digit defects and cleft palate) in mice and hamsters, a low frequency of malformations in rats, but was not teratogenic in dogs or swine. Aldrin is readily converted to dieldrin, and similar results were noted for dieldrin in these species as well as in the rabbit (Schardein, 1993). In the studies submitted for registration, possible adverse effects were noted in both the developmental toxicity studies (mouse and hamster at high doses) and the reproduction study in rats at the chronic toxicity dose range (CD PR, 2000). Amitraz In rats, adverse effects on reproduction were reduced litter size and substantial neonatal mortality at 200 ppm and slight to moderate neonatal mortality at 50 ppm, leading to a NOEL of 10.5 ppm. In the mouse, prolongation of the proestrus phase, a trend toward shortening of the diestrus phase, and depressed serum prolactin and progesterone levels with a NOEL of 25 ppm were observed (CD PR, 2000). In a developmental neurotoxicity study, rats were administered 20 mg/kg every third day and pups born were cross fostered. Open-field behavior (locomotion and rearing frequencies or immobility time) showed no significant differences, other than some transient delays (Palermo-Neto et aI., 1994). Postnatal exposure to


391

Table 14.3 Developmental and Reproductive Toxicity Profile of Insecticidesa Dose Chemical

Species

Toxicity profileb

Aldicarb

Rat

Acutely toxic; hence not

Rabbit Aldrin

Apholate Bendiocarb 1,3-Bis(carba-

considered teratogenic

(mg/kg)C 0.04

Teratogenic Reproduction

Hamster

Embryotoxic

50

Teratogenic

50

Rat


Rat

Reproduction

Risher et aI., 1987 As cited in Schardein, 1993

Mouse

Teratogenic

25 3-G study

Hodge et aI., 1967 As cited in Schardein, 1993

1


10 2

DPN#50094

100

Mouse

References

0.1

Rat

Sheep

Comments

CDPR Database, 2000 Schardein, 1993

moylthio)-2-N,N-

dimethy lamino propane 100

Rat Hamster Bromophos

Mouse

Carbaryl

Mouse


100 183

Nehez et aI., 1986

150 po or


5660 ppm (diet) Rat

Reproduction

Dog

Resorption, teratogenic

Pig

2000

2-G study

500

3-G study


6.25 30

Hamster

Fetal mortality

Rabbit

Teratogenic

150

Guinea

Teratogenic

300

125

pig Sheep

Teratogenic

Primate

Abortion

Cow Carbofuran

Mouse

250ppm

Diet

2 5.5

Liver histopathology

Rat

0.05 0.2

Adults

0.4

In utero, lactation

3 Rabbit Chlordane


50ppm

Dog

Pant et aI., 1995, 1997


0.5

Rat Mouse

Cell-mediated immune

Rat

(Temporary) tremors

80 g/kg

Days 7-17

Usami et aI., 1986

8

Throughout

Cranmer et aI., 1979

response

pregnancy 150-300 ppm

during and after

Ingle, 1952

gestation Chlordecone

Mouse

Reproductive failure

Rat

Fetotoxic

40ppm

Fertility study

10


Reproduction Chlordimeform

Rat

Postnatal behavioral deficit

Chlorfenvinphos

Rat

Ossification disorders

Huber, 1965 Canon and Kimbrough, 1979

100

~g

Diet


1/20 LDso

and carbaryl Chlormequat chloride

Rat Hamster

1000 ppm Teratogenic

Diet

100 (continues)

392

CHAPTER 14



Dose Toxicity profileb

Chemical

Species

Chlorpyrifos

Mouse

Fetotoxic

Rat

Reproduction

(mg/kg)C

Comments

References

2-0 study

Breslin et aI., 1996b

Deacon et aI., 1980

I

15

Ciafos

Rat

10kg

Coumaphos

Cow

28 g/45 kg

Crotoxyphos

Cow

Cyfiuthrin

Rat

As cited in Schardein, 1993 Topical route

BW 3.1 Reproduction: reduced pup viability; fetal malformations Rabbit

Postimplantation loss

Cypermethrin

Rat


DDT

Mouse

CDPR Database

50ppm 0.46mg/m 3

Toxicology profile DPN #50317

20

Shawky et aI., 1984

1/40 LDso

Teratogenic Reproduction

7ppm

diet


6-0 study

DEET

Rat

Reproduction

Rabbit


10

Rabbit

Incomplete ossification, other

30

CDPR Database, 1999

Mouse

86

Nehez et al., 1986

Mouse

85

Nehez et aI., 1986

1/50 LDso

3-0 study

skeletal effects Demethylbromophos sodium Demethylbromophostetramethylammonium Dialifor

Hamster

Teratogenic

100

Robens, 1970a

Diazinon

Rat


95-100

Campbell et aI., 1985 Dobbins, 1967

Reproduction: decreased pup survival, decreased ovary

CDPR Toxicology

< IOppm

Summary, 1999

weights Rabbit

0.25

Hamster Mouse


30 Reduced postnatal growtb

0.18

Spyker and Avery, 1977

6.6


Behavioral deficits Cow m-Dichlobenzene Dichlorvos

200

Rat Mouse

60

Rat

25

Rabbit

62


8.5

Pig Reproduction

6.2

Cow Dieldrin

Mouse

Teratogenic

6-0 study

Reproduction 6

Rat

3-0 study

Reproduction Hamster

Embryotoxic, teratogenic

sulfonamide

30

Rabbit

6

Sheep

25ppm 0.2

Dog N,N -diethyl-benzene-


15

Rat

Teratogenic

300

Rabbit

Resorption

25

Leland et al., 1992

(continues)


393


Dose Chemical

Species

Toxicity profileb

N ,N -Diethyl-n-

Rat

Reduced fetal weight

Rabbit

O,O-Dimethyl-S-

750

References Schoenig et aI., 1994 Wright et aI., 1992 Schoenig et aI., 1994

325

Pig Sheep

Dimethoate

Comments

Reproductively neurotoxic

toluamide Diflubenzuron

(mg/kg)C

100ppm

Fertility study

100ppm

Fertility study


Rat

Teratogenic

3

Khera et aI., 1979c

Cat

Teratogenic

12

Khera et aI., 1979a

Mouse

Reproduction

Mouse


5-G study

Budreau and Singh, 1973 As cited in Schardein, 1993

8

(2-acetylaminoethyl) dithiophosphate Empenthrin

Rat

Endosulfan

Rat

Delayed ossification

Endrin

Mouse

Teratogenic

2.5

As cited in Shepard, 1998

Hamster

Embryotoxic

5


Behavior deficits

5

Gray et aI., 1979

Teratogenic

0.75

Ethohexadiol

Rat

500

Developmental study

Fertility study

2 ml/kg

Kaneko et aI., 1992 Gupta et al., 1972

5

Occlusive


cutaneous Fenamiphos Fenbutatin oxide

Rat

3

Rabbit

2.5

Rat

Reproduction: decreased pup

Machemer et aI., 1992

250ppm

CDPR Database, 1999

weight gain in F 1 and F2 Fenitrothion

Mouse

DPN#214 80ppm

Rat

Postnatal behavioral deficits

Fensulfothion

Rabbit

Teratogenicity: malfonnations,

Fenthion

Rat

Diet

Lehotzky et aI., 1989

10 0.1

CDPR Worksheet 234-

incomlpete ossification Rat Fluvalinate

Rabbit

084054352, 1987 18

Astroff et aI., 1996

Epididymal cytoplasmic vacuolation Teratogenicity: malfonnations

CDPR Database, 1997 14ppm

2-G study

25

Fonnothion

Rabbit

Heptachlor

Rat

Heptachlor and

Rat

7ppm

Imidazolidinone

Rat

240

Isobenzan

Mouse

1 ppm

lOppm

(Worksheet 093 90460) CDPR Database, 1999

30 Cataracts in both generations

Benes et aI., 1973

As cited in Schardein, 1993 2-G study

Ruttkay-Nedecka et aI., 1972 Eisler, 1970

heptachlor epoxide As cited in Schardein, 1993 Diet

Isofenphos

Rat

Leptophos

Rat


Lindane

Mouse

Reduced fetal growth

Rat

Decreased fertility

0.5

Naishtein and Leibovich, 1971

Mortality and developmental

0.5


10 12.5 ppm

Mast et aI., 1985 Diet

Kanoh et aI., 1981 As cited in Schardein, 1993

delay 100ppm

Malathion

Rabbit

Inhibited development

40

Hamster

Inhibited development

20

Rat

Palmer et aI., 1978 As cited in Schardein, 1993

300 240

Rabbit

3-G study

Khera et aI., 1978 2-G study


100 (continues)

394

CHAPTER 14



Dose Chemical

Species

Toxicity profileb

Metam sodium

Rat

Teratogenicity: decreased fetal

(mg/kgY

Comments

CDPR Database, 1999

5

weights, severe malformations Rabbit

Postimplantation loss, severe

References

DPN#50150 5

defects (cleft palate, meningiocele) Methamidaphos

Rabbit Rat

Methomyl Methyl demeton

Rabbit Rat

Reproduction

Rat

Reproductive toxicity (altered

(oxydemeton

Methoxychlor

2-G study

CDPR Database, 2000

100 ppm

Diet

As cited in Schardein, 1993 As cited in Schardein, 1993 CDPR Database

and decreased fertility) Mouse

2.5 mg/gBW

Rat

2.5 mg/g BW

Rat

Fetopathic

Cow Methyl parathion

10ppm

ovarian, epididymal histology,

methyl) MethylISP

CDPR Database, 2000

2.5 Decreased pup weight gain

Rat


Wu et al., 1989 Khera et al., 1978

100 9.9


5

Frosch, 1990

Teratogenic Naled

Rat

Naphthylisothio-

Rat

Liver histopathology

100

Khera et al., 1979b

100


cyanate Nicotine

Mouse Rat


Resorption Teratogenic Developmentally toxic Postnatal behavioral deficits

0.008 I-lg/g BW 2

sc route

25mg

iv route Mini-osmotic pump

Hydrocephalus Toxic action in oocytes Rabbit Oxamyl

0.4

Antenatal

5

sc route

N-Octyl-

Reproduction

4 CDPR Database, 2000

10,000ppm

Rat

Vara and Kinnunen, 1951

1- and 3-G studies

Reproduction Rabbit

As cited Shepard, 1998

Kennedy, 1986

100

Rat


bicycloheptene dicarboximide Pentachlorophenol

Courtney et al., 1970a

75

Mouse Hamster

1.25

Rat

4

Embryotoxic

As cited in Schardein, 1993 Less toxic by po route


Welsh et aI., 1985 As cited in Schardein, 1993

Reproduction Phorate

1.94 mg/m 3

Rat

Phosalone Phosfolan

1/10 LDso Shen, 1983

Rat

Teratogenic

Rat

Embryotoxic

0.3

Teratogenic

0.3

Rabbit Photodieldrin

3-G study

Reproduction

Rat


3-G study

Reproduction Mouse

Inhalation route

Selectively toxic


35

Mouse

0.6

Rat

0.6

Chemoff etal., 1975

(continues)


395


Species

Photomirex

Rabbit

Toxicity profileb

(mglkg)C

Cataracts and reduced survival

Mouse

Teratogenic

20ppm

Reproduction


As cited in Schardein, 1993 study Ogata et aI., 1993

660


3,000

Rat Rabbit

References

10

Rat

in generational study Piperonyl butoxide

Comments


Potassium arsenate

Sheep

Propoxur

Rat

Neonatal CNS impairment

Ronnel

Rat


100 0.75

Rabbit

Teratogenic

Fox

Teratogenic Embryotoxic

1,000 ppm

Diet

2.5 Nafstad et aI., 1983 Berge and Nafstad, 1983

100

Khera et aI., 1981

2.5

Rotenone

Rat

Sarin

Rat

380 J.l-g/kg

Rabbit

15 J.l-glkg

Sodium arsenite

Hamster

25

Lu et aI., 1984 Bates and LaBorde, 1986 Teratogenic by iv route

Sodium selenite

Hamster

Soman

Rat

Teratogenic

Rabbit Sulfluramid

Rabbit

Spinosad

Rat

Hood and Harrison, 1982 WiIIhite, 1981

90

Ferm et aI., 1990

165 J.l-g/kg

Bates et aI., 1990

15 J.l-glkg Neonatal mortality

Stump et aI., 1997

0.3 200 10 (2-0)

Maternal NOEL

CDPR Database

= developmental NOEL

Rabbit 2,3,5,6-Tetra-

50

Rat

Zielke et aI., 1993

150

chloropyridine Thiometon

Rabbit

5


Toxaphene

Mouse

Teratogenic

15


Rat

Decreased skeletal ossification

35

Tribufos

Rat

Trichlorfan

Mouse

28

Rat

Teratogenic

400

Hamster

Teratogenic

400

Pig

Teratogenic

60

Trichloro-

Astroff et aI., 1996

300

Teratogenic


7.5

acetonitrile 1,2,3-

Rat

600

Black et aI., 1983

Trichlorobenzene 1,2,4-

Rat

300

Trichlorobenzene 1,3,5-

Rat

600

Trichlorobenzene Triphenyltin

Rat

hydroxide

20 Equivocally reduced fertility

100ppm

As cited in Schardein, 1993 Reproduction study

Valexon

Rat

Teratogenic Testicular toxin

a Includes

90 0.7

Fertility study

chemosterilants, repellants, and growth regulators. bDevelopmental toxicity includes reduced fetal weight, increased embryo/fetal mortality, and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate the presence of malformations. cDose is oral unless stated otherwise; is dose the LOEL wherever effects were observed and the NOEL when there were no effects.

396

CHAPTER 14


amitraz caused transient developmental and behavioral changes in the exposed offspring in a subsequent study (Palermo-Neto et aI., 1997). Results showed that the median effective time (ET50) for fur development, eye opening, testis descent, and onset of the startle response was increased in rats postnatally exposed to amitraz compared to those of the control. However, the age as incisor eruption, total unfolding of the external ears, vaginal and ear opening and the time taken to perform the grasping hind-limb reflex were not affected by amitraz exposure. Dibromochloropropane Dibromochloropropane (DBCP) is a brominated organochlorine that was used as a nematocide from the mid-1950s until its ban in the United States in the late 1970s (Whorton and Foliart, 1983). It was used in the United States, mostly in Hawaii and along the southern Atlantic and Pacific coasts, to protect citrus, grapes, peaches, pineapples, soybeans, and tomatoes. Of the pesticides studied to date, DBCP is the most toxic to the human male reproductive system. Torkelson et al. (1961) reported that DBCP caused testicular atrophy in rats, guinea pigs, and rabbits. Azoospermia and oligospermia were reported in DBCP production workers and these were linked with the length of time the persons worked with the chemical (Whorton et aI., 1977). Higher levels of follicle-stimulating hormone (FSH) were noted in a number of these individuals as well as in those who did not revert to normal spermatogenic levels long after exposure. Data from the animal studies revealed severe testicular insults, including degenerative changes in the semeniferous tubules, increase in Sertoli cells, reduction in the number of sperm, and increased abnormalities in the sperm cells. Epididymal (Kluwe, 1981), posttesticular effects (Kluwe et aI., 1983), and in vitro effects (Bartoov et al., 1987) were also noted. The mechanism of action was determined to be at the level of mitochondrial respiration, and DBCP (the parent compound) was demonstrated to inhibit carbohydrate metabolism at the reduced nicotinamide adenine dinucleotide (NADH) dehydrogenase step in the mitochondrial electron transport chain of rat sperm (Greenwell et aI., 1987). Workers on banana crops documented convincing evidence of increased spontaneous abortion in their family histories; follow-up studies among production workers in Israel showed that some recovered testicular function, but among their offspring, there was a predominance of females. Those who did not recover from azospermia were those with high levels of FSH (Goldsmith, 1997). No other studies have shown increased birth defects or increased infant mortality. Dichlorodiphenyltrichloroethane (DDT) Several investigators have examined the effects of DDT on both development and the reproductive system. DDT and its metabolite dichlorodiphenyldichloroethylene (DDE) have resulted in eggshell thinning in birds of different species (Porter and Wiemeyer, 1969). Thinner shells are associated with a higher disappearance and/or destruction of eggs, and it is this phenomenon, along with the huge public outcry subsequent to the publishing of Rachel Carson's book Silent Spring that led to

the ban on the use of DDT in the United States. The inconsistent effects of DDT and DDE on eggshell thickness might reflect differential sensitivities among species of birds. The thinning is thought to be similar to the way estrogen inhibits the formation of eggshells, although the calcium adenosinetriphosphatase (ATPase) inhibition by DDT demonstrated by Matsumura and Ghiasuddin (1979) may also be responsible. The interference (direct or indirect) with fertility and reproduction is thought, therefore, to be related to steroid metabolism and the inability of the bird to mobilize sufficient calcium to produce a strong eggshell to withstand the rigors of being buffeted around the nest; the resultant cracking allows the entry of bacteria, causing the developing embryo to die (Carson, 1962; Peakall, 1970). In male cockerels and rats, DDT (20% o,p'-DDT and 80% p,p'-DDT) reduced testicular size; in females o,p'-DDT administration yielded estrogenic effects such as edematous, blood-engorged uteri (Ecobichon and MacKenzie, 1974; Hayes, 1959). DDT has resulted in uterotrophic effects; o,p'-DDT has demonstrated increased weight in ovaries and uteri and an advanced time of vaginal opening (Gellert et aI., 1972). Other estrogenic effects such as increases in cellular progesterone receptors, tissue mass, DNA synthesis, and cell division have also been reported (Ireland et aI., 1980; Mason and Shulte, 1980; Nelson, 1974; Robison and Stancel, 1982). Kupfer and Bulger (1976) have shown that the o,p'isomer competes with estradiol for binding with estogen receptors in rat uterine cytosol. A review of studies conducted in Israel, India, and the Ukraine suggested that maternal and fetal tissue levels of DDT and metabolites were higher in fetal deaths than in other pregnancies; however, studies from Poland, Italy, and Florida observed no significant differences in the levels of organochlorine pesticides in samples from normal versus aborted pregnancies (Sever, 1988). DDT is known to be lipophilic and a potent bioaccumulator and can be detected in fatty tisues in the food chain long after its use. This and the fact that it is still widely used as an efficient agent in the control of mosquitoes causing malaria can result in marked long-term ecological impact. Methoxychlor Methoxychlor is a chlorinated hydrocarbon insecticide that has a much lower bioaccumulation potential than that of DDT. Early studies in pregnant rats demonstrated maternal and fetal toxicity (wavy ribs) at exposures of 200 and 400 mg/kg (Khera et al., 1978). In rats, methoxychlor is metabolized in vivo to 2,2-bis(p-hydroxyphenyl)-I, 1,1trichloroethane (HPTE), the active estrogenic form. It has direct estrogenic effects on the rat uterus and also inhibits the decidual cell response, which is an accepted model for implantation-associated effects. It has adverse effects on fertility, early pregnancy, and in utero development in females; in adult males, adverse effects such as altered social behavior following prenatal exposure to methoxychlor were noted (Cummings, 1997; Cummings and Gray, 1989). Recent work in mice concluded that neonatal exposure to methoxychlor at doses of 0.1, 0.5, and 1.0 mg/kg/day did not interfere with mating, but significant alterations were seen in initiating and/or


maintaining pregnancy. The deleterious effects on pregnancy may be due to the influence of neonatal methoxychlor treatments on the hypothalamic-pituitary-ovarian axis as well as on possible alteration of the uterine environment (Swartz and Eroschenko, 1998). The significance of this toxicity with respect to human health remains to be determined. In LongEvans hooded rats, methoxychlor affects the central nervous system (CNS), the epididymal sperm numbers, and the accessory sex glands and delays mating without significantly affecting the secretion of LH, prolactin, or testosterone. These data indicate that methoxychlor did not alter pituitary endocrine function in either an estrogenic or an antiandrogenic manner (200-400 mg/kg/day) and demonstrate a pronounced degree of target tissue selectivity (Gray et aI., 1999a). Chlordecone Chlordecone was sold as an insecticide and fungicide between 1958 and 1975. Better known by its trade name Kepone™, it was widely used and caused contamination of the lames River near the plant where it was manufactured in Virginia. Mirex, an insecticide that photodegraded to Kepone, has also been extensively studied and the toxicity of both compounds will be summarized here. Production workers exposed to chlordecone were noted to be oligospermic and had reduced sperm motility (Taylor et aI., 1978). Chlordecone primarily affects sperm motility and viability via mechanisms that are not completely understood. One study in CD-l mice documented that the pool of potentially ovulatory follicles was reduced subsequent to prolonged exposure to Kepone (Swartz and Mall, 1989). It is a potent inducer of the mixed-function oxidase system and may affect fertility by stimulating hepatic degradation of steroids. Although chlordecone (probably as the hydrate) has a binding affinity for estrogenic sites (Hammond et aI., 1979), other studies have concluded that the reproductive toxicity was not by a mimicry of estrogen (Cochran and Wiedow, 1984). Instead, it appears that chlordecone can act on the hypothalamic-pituitary axis (Hong et aI., 1985). Exposure to 50 and 75 mg/kg of chlordecone in the female rat before or after mating substantially reduced fertility (Uphouse, 1986). A review of the developmental toxicity of chlordecone exposure during gestation indicated fetotoxicity in mice and rats and some CNS impairment in fetuses in the rat (Schardein, 1993). Lindane (Hexachlorocyclohexane-HCH) Lindane, a nonaromatic chlorinated cyclic hydrocarbon, has shown some effects on the female reproductive system (Welch et aI., 1971), but, on the whole, the results are variable. In male rats fed 75 mg/kg/day of lindane for 90 days, testicular atrophy, degeneration of semeniferous tubules, and disruption of spermatogenesis were reported (Shivanandappa and Krishnakumari, 1983). Reduced epididymal sperm concentrations were also noted in other studies in rats given a single dose of 30 mg/kg/day (Dalsenter et aI., 1996). As with DDT, the estrogenic effects of lindane might occur via estrogen receptors, but studies in female Long-Evans rats dosed at 40 mg/kg/day for 7 days or in ovariectomized rats for 5 days did not show altered serum

397

estradiol concentrations or changes in the number of estrogen receptors. No changes in estrogen-dependent induction of progesterone in the hypothalamus, pituitary, or uterus were observed. Hence, it is thought that lindane may act via altering multiple processes such as the gamma-aminobutyric acid (GABA)-nergic system or via altered growth factors (Laws et aI., 1994). In reproduction studies, fertility was not reduced, but most of the F 1 pups died shortly after birth. Erratic estrous cycles were also noted with exposure to lindane (Gray et aI., 1988). The adverse liver effects in the rat reproduction study do not appear to be important to human safety because they were reversible after much higher exposures in the combined study (CDPR, 2000), and the kidney effects (apparent hydronephrosis) were species- and sex-specific. The reproductive NOEL was 20 ppm, based on reduced neonatal pup survival (largely due to total litter losses), slightly reduced pup growth rate, and slightly slower pup development such as delays in hair growth and tooth eruption (CD PR , 2000). In a suicidal poisoning, maternal ingestion of lindane resulted in the death of twin fetuses (Konje et aI., 1992). In addition to these and other endocrine effects, a dose-related increase in the incidence of fetuses with an extra 14th rib in CFY rats and an extra 13th rib in rabbits has been reported at 15 mg/kg/day but a lack of teratogenicity was determined (Palmer et aI., 1978). This was consistent with the negative teratogenicity results in mice, mutagenicity studies, and 3-generation rat reproduction studies. Regional changes in brain noradrenaline and serotonin levels have also been reported as developmental effects (Rivera et aI., 1991). Adverse effects, however, were not noted in the developmental toxicity studies submitted to the CDPR, and lindane and related HCH isomers are not listed as chemicals known to the state to cause reproductive toxicity under Proposition 65 or the Safe Drinking Water and Toxic Enforcement Act of 1986. 14.4.2.2 Organophosphates These compounds cause a combination of a reduction in brain acetylcholinesterase activity and altered reproductive behavior in a number of species. The reduced acetylcholinesterase has been associated with decreased egg production and serum LH and serum progesterone (Rattner et aI., 1982). The standard dominant lethal test in mice was negative for dichlorvos (Dean and Blair, 1976). Possible mechanisms of toxicity from studies on trichlorfon and parathion in the rat are thought to involve interference with steroid hormone binding to receptors in the liver, adrenal, uteri, and testes (Trajkovic et aI., 1981). In a case report, the organophosphate pesticide mercarbam crossed the placental barrier and caused the death of a 5-month fetus (Schardein, 1993). There has been some indication that organophosphates (OPs), in general, may affect the menstrual cycle and cause an early menopause in humans. Reproductive effects from exposures to mixtures of OPs have been documented by Mattison et al. (1983) and Nakazawa (1974) among women in agriculture. These effects included abnormal menstruation (e.g., hypermenorrhea, oligomenorrhea, amenorrhea) and early menopause. On the other hand, Willis

398

CHAPTER 14


et al. (1993) found no effects of pesticide exposures (including methyl parathion) on the pregnancy outcome among 535 women enrolled in a Southern California community clinic perinatal program.

Chlorpyrifos The developmental toxicity and reproductive toxicity of this widely used compound have been extensively studied. The studies submitted for registration under FIFRA did not show adverse developmental or reproductive effects. However, some recent studies on the role of cholinesterase in morphogenesis have used chlorpyrifos as the model compound and intimated an influence in learning disabilities among children who were exposed in utero or during the early postnatal period (Roy et aI., 1998). Chlorpyrifos, a phosphorothioate, undergoes oxidative desulfuration to form chlorpyrifos-oxon, which can then phosphorylate acetylcholinesterase, rendering it incapable of metabolizing the neurotransmitter acetylcholine to choline and acetate. The oxon, however, either may be detoxified by combining with carboxylesterase or may be hydrolyzed by oxonase to metabolites not capable of combining with acetylcholinesterase. It has been suggested that inhibition of DNA and protein synthesis may be direct noncholinergic effects of chlorpyrifos, but other mechanisms such as alterations in blood flow patterns may also be involved. In reviewing several studies as part of the risk characterization of the compound, it appears that newborn and juvenile rodents are more susceptible to the toxic effects of chlorpyrifos than adults. The increased susceptibility of young rats appeared not to be due to a difference in the affinity of the oxon for the acetylcholinesterase, but rather incomplete development of the enzyme systems that detoxify the oxon (CDPR, 2000). In vivo studies in rats have used newborns to explore the effects of chlorpyrifos on the ontogeny of the mammalian nervous system during the equivalent last trimester of human fetal development. However, sublethal concentrations were administered to postnatal day 1 rats via intraperitoneal or subcutaneous routes. In the human, in utero exposures are likely to be mitigated by the mother's metabolism and the availability of chlorpyrifos to the fetus may therefore not be comparable to newborn rodents being directly dosed. Aggregate exposure estimates, acute illness reports, and dermal exposure from surface wipes also point to low levels of chlorpyrifos (CDPR, 2000; Lewis et aI., 1994; Lu and Fenske, 1999). Higher exposures to children have been reported for oral nondietary and dermal exposure (Gurunathan et aI., 1998); the assumptions used, however, may not be appropriate. Thus, although the fetus (independent of maternal metabolism) may be theoretically more susceptible than the adult, the evidence that chlorpyrifos causes developmental neurotoxicity under physiologically relevant conditions is not compelling. The inhibition of cholinesterase activity (brain and liver) in pups was detected at doses nearly lethal to the rat dams (Tang et aI., 1999). Thus, taken together, the studies suggest that maternal effects will be observed prior to levels causing developmental toxicity. Chlorpyrifos has been banned in the United States for indoor application.

Diazinon This organophosphate insecticide has been tested extensively and yielded variable results for reproductive and developmental endpoints. Spyker and Avery (1977) exposed pregnant mice (9 mg/kg) and observed behavioral effects and functional impairments in overtly normal offspring along with neuropathology in the forebrain. The standard developmental toxicity studies in rats and rabbits did not demonstrate adverse effects (CD PR, 2000). In another review, malformations were not reported in hamsters and rabbits; renal, rib, and limb malformations and anomalies of the CNS and digits were noted in rats, and skull and teeth abnormalities were noted in puppies (Schardein, 1993). In both one-generation and two-generation reproduction studies, a variety of adverse effects were observed, including a decrease in the gestation index (number of litters with live offspring/number pregnant), a reduction in ovarian weights, and a prolonged gestation length. The reproduction NOEL was less than 10 ppm (1 mg/kg/day) or the lowest observed effect level (LOEL) was 10 ppm (CDPR, 2000). Behavioral effects were confirmed in neurotoxicity studies in rats (CD PR, 2000); hence, the developmental neurotoxicity of diazonon may need evaluation. Dimethoate Developmentally toxic effects were not noted in either the rat or the rabbit studies submitted under FIFRA (CDPR, 2000); however, rib defects in rats and polydactyly in cats were noted (Khera et aI., 1979c). In the mouse, variable results were noted with embryotoxicity without teratogenicity in earlier studies (Scheufler, 1975, 1976) and no adverse effects were found in another study (Courtney et aI., 1985). Reduced number of pregnant females and lowered pup weights and reduced litter sizes were noted in a second-generation reproduction study in rats (CDPR, 2000). Fenthion A reduction in fetal weights at 80 mg/kg was noted in a study in mice with an increase in malformations in 14.5% of the offspring (Budreau and Singh, 1973). In rats, however, a marginal increase in resorptions was noted at 18 mg/kg/day, demonstrating no other adverse developmental effects at lower doses and yielding a NOAEL of 4.2 mg/kg/day. Exposure to fenthion in the diet at 14 and 100 ppm in the reproduction study demonstrated epididymal cytoplasmic vacuolation (ECV) associated with decreased fertility, reduced survivability, and postnatal growth retardation resulting in a reproductive NOEL of2 ppm (CDPR, 2000). Parathion and Methyl Parathion Three multigeneration reproductive toxicity studies in rats have been submitted to regulatory agencies (CDPR, 2000), and decreased pup survival was consistently found in all three studies. A search of the literature revealed one study showing ovarian toxicity in rats (Dhondup and Kaliwal, 1997) and one study showing possible sperm abnormalities in mice (Mathew et aI., 1992). Testicular and reproductive effects have also been reported in avian species (Maitra and Sarkar, 1996; Solecki et aI., 1996). In the absence of clinical symptoms and behavior changes, reproductive effects such as reduction in the number of eggs laid (~ 20% reduction), egg

14.4 Toxicology Studies weight (~9% reduction), and eggshell thickness (7-10% reduction) were noted in the Japanese quail. Suppression of growth and ossification in both mice and rats were observed subsequent to methyl parathion exposure; in the mouse, high mortality and cleft palate were noted (Tanimura et aI., 1967). No other epidemiological data specific to methyl parathion are available. Malathion Malathion does not appear to cause adverse developmental or reproductive effects (CD PR, 2000). However, malathion administered to mice at 250 mg/kg (corresponding to 1/12 LDso) and examined 4, 14, 18, and 26 days after injection induced teratozoospermia. Sperm count at different time intervals was significantly increased compared to the controls, and there was a parallel increase in depletion of the semeniferous tubules; all germinal cell populations studied were affected by malathion, especially mice spermatid differentiation, which compromises mostly the flagella, perhaps due to an alkylating effect that disturbs the normal assembling of tail structural protein components (Contreras and Bustos-Obregon, 1999). No evidence for histopathological alteration or teratogenic anomalies in the fetuses was demonstrated, although placental transfer of malathion was indicated by the presence of the insecticide residues in fetuses from rats fed wheat material containing bound residues of malathion S-I,2-di(ethoxycarbonyl)-ethylO,O-dimethyl phosphorodithioate (Bitsi et al., 1994). The reproductive effects of the aerial spraying of the organophosphate insecticide malathion in California have been examined in a case-control study of spontaneous abortions «28 weeks) and stillbirths; relative risks were 1.21 (95% Cl = 0.94-1.52) for spontaneous abortions and 1.51 (95% Cl = 0.21-11.3) for stillbirths. A cohort of 7450 pregnancies identified through three Kaiser-Permanente facilities in the San Francisco Bay area, in relation to exposure to the pesticide malathion, applied aerially to control an infestation by the Mediterranean fruit fly, was examined for reproductive outcomes. No important association was found between malathion exposure and spontaneous abortion, intrauterine growth retardation, stillbirth, or most categories of congenital anomalies. Gastrointestinal anomalies noted were related to second-trimester exposure (OR = 2.6), based on 13 cases, and were not specific to any particular International Classification of Diseases code (Thomas et aI., 1992). Trichlorfon Trichlorfon, an organophosphate insecticide, has been associated with a cluster of babies born with Down's syndrome in Hungary (Czeizel et al., 1993). A case-control study and environmental investigations reported excessive use of trichlorfon at local fish farms. The high content of trichlorfon in the diet of pregnant women, including all of the mothers with affected offspring, along with an absence of known teratogenic factors such as familial inheritance and consanguinity, was supportive of the association. Trichlorfon or chlorophos, marketed under the brand name Dipterex, displayed embryotoxic and teratogenic effects in the Wistar rat after an oral dose of 80 mg/kg during a critical period of embryogenesis, but was negative at the low dose of 8 mg/kg (Martson and Voronina, 1976). A review of this insecticide and anthelmintic demonstrated a potent

399

developmental toxicity profile in laboratory animals such as the rat (oral and inhalation), mouse (oral), and hamster (oral), but teratogenicity was not noted if the exposure was via the intraperitoneal route (Schardein, 1993). Recent work from Norway characterized the teratogenic effects of trichlorfon on the guinea pig brain by determining the effective dose and sensitive period (Hjelde et al., 1998). Following oral or subcutaneous exposure, almost all regions of the brain were reduced in weight. The cerebellum was the most vulnerable region, but the medulla and hypothalamus were also greatly reduced in weight. While the mechanism behind the teratogenic effect is not known, alkylation of DNA or altered DNA repair may be involved. 14.4.2.3 Carbamates

Similar to organophosphates, carbamates result in the inhibition of cholinesterase and also exert an anesthetic effect. The dithiocarbamates have been used as fungicides and will be discussed under that category. Carbofuran Most studies with carbofuran have been negative for teratogenicity. However, one study in mice resulted in fine-structure abnormalities in mice (Schardein, 1993); in the FIFRA reproduction study, reduced body weight gain in adults and birthweights in offspring worsening to 15% by weaning were noted (CDPR, 2000). Decreased weights of the epididymides, seminal vesicles, ventral prostate, and coagulating glands were also noted in rats, along with decreased sperm motility, reduced epididymal sperm count, and increased morphological abnormalities in the head, neck, and tail regions of spermatozoa (Pant et al., 1995). Testicular and spermatotoxic effects were also noted at levels higher than 0.2 mg/kg in rats exposed to carbofuran in utero or via lactation (Pant et aI., 1997). Studies from Sri Lanka in rats have concluded that carbofuran administered orally 0.2, 0.4 and 0.8 mg/kg during early gestation is detrimental to pregnancy (enhanced preimplantation losses) and possibly harmful to neonatal development (Jayatunga et aI., 1998a). Similarly post-implantation losses were noted after exposure to carbofuran during mid-gestation (Jayatunga et al., 1998b). Bendiocarb This compound is a residual insecticide and appears to cause adverse effects on reproduction in the rat. A decrease in the number of pups and reduced survival at dose levels of 200 and 250 ppm resulted in a NOEL of 50 ppm (CD PR, 2000). Thiodicarb Adverse effects were documented in the rat reproduction study submitted for registration. Decreased pup weight gain at 100 ppm and above and decreased viability at 900 ppm were noted. The parental NOEL was 100 ppm (decreased Fo and Fl body weights); the reproductive LOEL was 100 ppm (reduced pup weight gain). A statistically based estimate of the NOEL provided an NEEL (no expected effect level) of 81 ppm in males and 80 ppm in females (CD PR, 2000).

400

CHAPTER 14


Zineb and Thiram The exposure of rats to zineb and thiram has documented an alteration (prolongation) of the estrous cycle in association with a reduction in ovarian and uterine weights (Ghizelea and Czeranschi, 1973). Direct gonadal effects have been noted in mouse and rabbit oocyctes, resulting in inhibition of oocyte meiotic maturation and prevention of germinal vesicle breakdown. In the mouse oocyte exposed to isopropyl-N-phenylcarbamate, the formation of nuclear condensates (macromolecules) has also been observed by Crozet and Szollosi (1979). Carbaryl Carbaryl (N -methyl-I-naphthyl carbamate; Sevin®) is a carbamate insecticide that is widely used, resulting in exposure during its use as well as via consumption of treated food. It is not metabolized to an active intermediate; the parent compound itself is thought to be the active agent. Carbaryl acts via inhibition of acetylcholinesterase by carbamylation of the active-site serine residue. Adverse effects on rodent spermatogenesis at 0.4-5 mg/kg ip (intraperitoneal) or orally have been reported (Weil et al., 1972), but several studies do not support this finding. Hence, although the data suggest that carbaryl does not produce a testicular effect similar to DBCP, human personnel may be affected as demonstrated by some studies (Wyrobek et al., 1981). The teratogenicity of carbaryl has been reviewed extensively, but the results can best be summarized as equivocal (Schardein, 1993). Data are available in a number of species; early studies in the rat reported terata but subsequent studies were negative. Studies in mice were variable, and the eye defects seen previously were not observed in later studies. In rabbits, the data are contradictory: Omphalocele and skeletal variations were noted at the high dose; however, malformations were not seen in the other study. A recent study in rabbits submitted to the CDPR documented agenesis of the gall bladder at the high dose in two fetuses of different litters, along with reduction in the size of the gallbladder and a missing bile duct, presenting the case for a continuum of effects (CD PR, 2000). Sheep that were fed carbaryl in the diet demonstrated heart defects; in the dog, where pregnant females were dosed with 6.25-50 mg/kg/day in the diet, malformations included abdominothoracic fissures, intestinal agenesis and displacement, brachygnathia, failure of skeletal formation, anurous (no tail), and superfluous phalanges. In addition to these malformations that were seen in several puppies, resorption was noted in 21 of 181 pups (11.6% fetal incidence; 21.1 % litter incidence). However in primates, minature swine, and cattle, exposure to carbaryl during gestation resulted in no malformations, although abortions were noted in the primate study (Schardein, 1993). 14.4.2.4 Pyrethroids

Another class of insecticide that is widely used is the pyrethroid group of compounds. These are considered to be relatively safe and are perceived to be innocuous because of the origin of the natural pyrethrins from the chrysanthemum family of plants. Earlier studies on the metabolism and toxicity of synthetic pyrethroids (fenothrin, furamethrin, proparthrin, resmethrin,

tetramethrin, and allemethrin) indicate that neither the cis- nor the trans-isomer of chrysanthemumate is teratogenic in rats, mice, and/or rabbits (Miyamoto, 1976). Toxicity studies with decamethrin, a synthetic pyrethroid, found no evidence of teratogenic activity in rats or mice at dose levels that produced marked maternal toxicity (Kavlock et al., 1979). However, numerous studies on the genetic toxicity potential of this group of compounds (cypermethrin and deltamethrin) have demonstrated a wide range of effects, including mitotic/chromosomal abnormalities and the induction of sister chromatid exchanges (Chauhan et al., 1997). In the studies submitted in support of registration for deltamethrin, no significant developmental toxicity was reported in rats; delayed ossification was noted in the high dose in rabbits, aIong with maternal effects (CDPR, 2000). However, a 5% deltamethrin formulation in Wistar-derived albino rats resulted in dose-dependent early embryonic death, retardation of fetal growth, hypoplasia of the lungs, and dilation of the renal pelvis with no skeletal abnormalities (Abd El-Khalik et al., 1993). Significant increases over respective controls were evident for chromosome aberrations, micronuclei, and sperm abnormalities (Bhunya and Pati, 1990). A number of effects of exposure to pyrethroids during early development have been described in rats and mice. Cypermethrin caused an apparent increase in blood-brain barrier permeability in 1O-day-old rat pups after a single or repeated doses of about 15% of the LDso but had no effect on the adult barrier (Gupta et al., 1999). A slightly lower daily dose of 4% of the LDso over postnatal days 10-16 caused an increase in renal D 1 receptor density in rats, which persisted at least until day 90 (Cantalamessa et al., 1998). Similarly low doses of bioallethrin to mice over postnatal days 10-16 decreased muscarinic receptor density in adult mouse neocortex and produced lasting changes in adult behavior (TaIts et al., 1998). These very interesting results have, however, proved impossible to reproduce in other laboratories using only slightly different protocols (unpublished results; D. E. Ray, personal communication), and their applicability to humans is at present uncertain. Similar results with DDT have been reproduced in rats as well as mice (P. Eriksson, personal communication) and in a second strain of mice (unpublished results; D. E. Ray, personal communication). The pyrethroids are also capable of producing gross effects on brain maturation and morphology, but only if given at dose levels that cause reduced body weight in the offspring (Patro et al., 1997), probably via a nonspecific developmental delay due to undernutrition. 14.4.2.5 Insect Growth Regnlators Methoprene A review of the developmental effects of methoprene indicated that a high incidence of multiple malformations was induced in mice, but not in rats (Schardein, 1993). In 1995, middle-school students reported (on the Internet) a high incidence of malformed frogs from a southern Minnesota farm pond. Consequently, increased rates of congenital anomalies in regions of Minnesota associated with pesticide use have heightened awareness of the possible effects of pesticides


(Garry et aI., 1996). Another group has implicated agricultural contaminants in the hindlimb deformities in frogs from a number of ponds in Quebec (Ouellet et aI., 1997). Additionally, some degradation products of the insect growth regulator S-methoprene have been reported to alter early frog embryo development in the laboratory (La Clair et aI., 1998). However, confirmation of these effects in mammlian species is lacking. The standard teratogenic studies conducted under FIFRA requirements for methoprene do not demonstrate similar results. Furthermore, it is not known if this compound, a juvenile growth hormone agonist, is used in quantities high enough to be a cause for concern. Recent findings linking the limb defects in frogs to a trematode parasite have shifted suspicion away from methoprene, but in the interest of providing the reader with the putative effects of this compound, the aforementioned information is provided. Diflubenzuron Other insect growth regulators such as diflubenzuron (Dimilin, TH 6040; N-[[(4-chlorophenyl)amino]carbonyl]-2,6-difluorobenzamide) have been tested in male and female layer-breed chickens from 1 day of age through a laying cycle at levels of 1, 2.5, 25, and 250 ppm in the feed. Feeding diflubenzuron at levels up to 250 ppm did not affect the characteristics measured such as egg production, egg weight, eggshell weight, fertility, hatchability, and effects on the progeny (Kubena, 1982). 14.4.3 FUNGICIDES Most fungicides tend to produce positive results in standard in vitro microbial mutagenicity tests. This is because the microorganisms used in such test systems are similar to the fungi. However, given the predictive possibility of the mutagenicity tests for teratogenic and carcinogenic potential, there is mounting concern about the developmental toxicity of these compounds. Several fungicides have documented developmental toxicity and details are given later in text (also see Table 14.4). There is evidence to suggest that fluconazole, a bis-triazole antifungal agent, exhibits dose-dependent teratogenic effects; however, it appears to be safe at lower doses (150 mg/day). Ketoconazole, flucytosine, and griseofulvin have been shown to be teratogenic and/or embryotoxic in animals. Iodides have been associated with congenital goiter and should not be used during pregnancy. Benomyl In reproduction studies in the rat, a reduction in epididymal sperm counts in pubertal animals was observed. Postpubertal animals showed a wide variation in susceptibility of sperm counts. Histological exams of testicular tissue showed an increased incidence of diffuse hypo-spermatocytogenesis in pubertal and postpubertal males (Carter et aI., 1984). In the 3000- and 1O,000-ppm males, lower sperm counts were noted. In addition, testicular atrophy and degeneration (4/30 and 29/30 in PI and 9/30 and 21125 in FI 3000- and 1O,OOO-ppm groups respectively) and oligospermia in the epididymides (unilateral and bilateral with 1130 at 3000 ppm and 26/30 at 10,000 ppm

401

in PI, 9/30 and 20125 in FI respectively) were observed. For the reproduction study, the NOEL was 500 ppm in males and 3000 ppm in females (decreased body weights). The NOEL for developmental toxicity was 31.2 mg/kg/day (dose-related reduction in fetal weight, hydrocephaly, microphthalmia, fused ribs, fused vertebrae, and decreased ossification in tail and in vertebral centra) in rats. Findings at the highest dose tested of 125 mg/kg/day included full litter resorptions in 6 of 11 surviving pregnant dams, enlarged lateral ventricles, enlarged renal pelves, and delayed ossification (more widespread than at 62.5 mg/kg/day). Fetotoxicity and teratogenicity findings in the absence of obvious maternal toxicity indicated possible adverse effects. MBC, a metabolite of benomyl, appeared to cause significant effects (postimplantation loss) in rabbits at the midand high-dose levels and resulted in a developmental NOEL of 10 mg/kg/day versus a maternal NOEL of 20 mg/kg/day. However, a teratology study in rabbits exposed to benomyl that was acceptable to the CDPR did not demonstrate possible adverse effects (CDPR, 2000). Dinocap Technical-grade dinocap, a complex-mixture fungicide, has been demonstrated to be teratogenic in the CD-l mouse, causing cleft palate, a dose-related increase in supernumerary ribs, a low frequency of exencephaly, umbilical hernia at high doses, otolith defects, weight deficits in fetuses at term, increased neonatal mortality, abnormal swimming behavior, and torticollis (Rogers et aI., 1986). Neither of the purified isomers, 2,4-dinitro-6-(1-methylheptyl)phenyl crotonate and 2,6-dinitro-4-( 1-methylheptyl)phenyl crotonate, exhibited any developmental toxicity when administered under identical conditions (Rogers et aI., 1987). Similar developmental defects were not seen in the rat and hamster (Rogers et aI., 1988). Hexachlorobenzene Hexachlorobenzene (HCB) is a preemergent fungicide and is ubiquitous in the environment. It has been isolated in the repoductive tract in several species, including humans (Jarrell et aI., 1998; Trapp et aI., 1984). Although HCB was not mutagenic in microbial test systems and was negative in dominant lethal mutation tests, it did cause terata in mice (renal and palate malformations) and in rats (increased incidence of the 14th rib). HCB was also found to be particularly toxic to the developing perinatal animal, transplacentally and via the milk, causing enlarged kidneys, hydronephrosis, hepatomegaly, and possible effects on the immune system (Ecobichon, 1996). It is commonly present in fat because of its lipophilicity and tendency to bioaccumulate (Mes et aI., 1982). The adverse reproductive effects of HCB have been reported in rats, minks, ferrets, and monkeys (Bleavins et aI., 1984; Iatropoulos et aI., 1976). These effects include decreased fertility, fecundity, and impaired cyclicity. In a recent study in Germany, HCB concentration correlated with maternal age (r = 0.249, p < 0.01), with 2.7-fold higher serum levels in offspring of 40-year-old as compared with 20-year-old women, concluding that the neonatal burden depends on maternal age and duration of pregnancy. This reflected the increase in body accumulation with these substances during human life as well as a continuous

402

CHAPTER 14


Table 14.4 Developmental and Reproductive Toxicity Profile of Fungicides Dose Chemical

Species

Toxicity profile"

Alkyldithiocarbamic

Rat

Teratogenic

(mg/kg)b

Comments

References As cited in Schardein, 1993

1/20 LDso

acid Benomyl

Mouse


Kavlock et aI., 1982

100

Teratogenic

Selectively toxic

Rat

Fetal mortality

200

Teratogenic

200

Reproduction

Teratogenic po; negative

Kavlock et aI., 1982 As cited in Schardein, 1993

via diet 3-G study

Bis(tri-N -butyltin) oxide

Rabbit

Visceral variations

Mouse

DevelopmentaIIy toxic

23.4

Teratogenic

11.7

Rat Bitertanol

Rat

Captafol

Rat


10

Teratogenic

10

Teratogenic

1/10 LDso 500

Primate

Carbendazim

200 2000

Teratogenic

Primate

37.5 75

Hamster

Teratogenic

Dog

Teratogenic

Rat

Vergieva, 1990 As cited in Schardein, 1993

100

Rat Rabbit

Crofton et aI., 1989

25 Teratogenic

Mouse Captan


ISO

Rabbit Hamster

Munley and Hurtt, 1996

180

Two species

300

Embryotoxic

100

Teratogenic

200

Robens, 1974 As cited in Schardein, 1993

160

Rabbit Rat

1.5 J.lg/100 g

Rabbit

47J.lg/animal

sc route

Janardhan et aI., 1984

iv route


BW Cupric acetate

Rat


0.01

Teratogenic

0.01

Toxic to testis Cycloheximide

Mouse

Teratogenic

Rabbit


Rat Cymoxanil

Rabbit

As cited in CDFA, 1987

30


0.05

As cited in CDFA, 1987

400


Teratogenic: early resorptions, skeletal variations

CDPR Database, 1999

Dazomet

Rat

Dinocap

Mouse


12

Selectively

Rat


100

By oral and

Gray et aI., 1996

toxic Rogers et aI., 1988

dermal routes Rabbit

Fetotoxic

48

By oral and

Costlow et aI., 1986

dermal routes Hamster Ethy lenebisisothiocyanate sulfide


Rat

12.5 200

Mouse Functional alterations

Rogers et aI., 1988 Chernoff et aI., 1979b

30

(continues)


403


Dose Chemical

Species

Ethy lthiuram

Rat

Toxicity profilea

(mglkg)b

Comments

References As cited in Schardein, 1993

60

monosulfide Ferbam

Mouse Rat

300 Teratogenic

150


Reproduction Flusilazole

Rat

Folpet

Mouse

Teratogenic

lOO

Rat

500

Rabbit Hamster


80 Teratogenic

Primate Hexachlorobenzene

Vergieva, 1990

1/5 LDso

500 Two species

Mouse

Teratogenic

Rat


100 10


DevelopmentaIIy neurotoxic Reproduction Rabbit Imazalil

Mouse

Teratogenic: increased

Rat

Teratogenic: resorptions,

Goldey and Taylor, 1992 10


10

CDPR database, 1999 DPN#413

skeletal defects NOEL =40

reduced fetal weights Reproduction: decreased

Pup NOEL = 20

litters and litter size Imidazolidinethione

Mouse Rat

800 DevelopmentaIIy toxic

10

Teratogenic

20

Resorption

10

Hamster

Teratogenic

270

Cat

Teratogenic

As cited in Schardein, 1993 Selectively toxic

Rabbit

Guinea

5 100

pig Isoprothiolane

Mouse

Mancozeb

Mouse Rat

DevelopmentaIIy toxic Teratogenic

Rabbit Maneb

Mouse

lOO 1330 1330 80

Increase in variations

375


Altered behavior Rat

Teratogenic

480

Reproduction Metiram

Rat

Teratogenic: decreased live

80

CDPR Database, 1999

56

CostIow et aI., 1983

litter size Ochthilinone

Rat Rabbit

Embryotoxic

DPN #217 1.5

Phenoxyacetic acid

Mouse

900

Phenylphenol

Mouse

2100

Rat


Rabbit Polycarbacin

Rat

Embryotoxic


150


250

Zablotny et aI., 1992

610


Teratogenic (continues)

404

CHAPTER 14



Species

Toxicity profilea

(mg/kg)b

Propamocarb

Rat

Teratogenic: skeletal

0.31 mllkg

Comments

CDPR Database, 1999

changes

DPN#50308

Propineb

Rat

Teratogenic

1000

Quintozene

Mouse

Teratogenic

500

Rat

500

Sodium phenylphenol

Mouse

400

Terrazole

Rabbit

Increased resorptions,

References


IS

CDPR Database, 1999

30


malformations 2,3,4,6-Tetra

Rat

Delayed ossification

Mouse

Retarded growth

chlorophenol Thiophanate ethyl

200

Reproduction Thiram

Triadimefon

Rabbit

3-G study

Mortality

0.01 LDso

Teratogenic

0.01 LDSO

Hamster

Teratogenic

Rat

Reduced fertility, viability,

250 1800ppm

CDPR Database, 1999

and lactation indices Rat

Decreased body weight gain,

lOO

cleft palate Rabbit

Malformations, variations,

50

delayed ossification 2,4,5-trichlorophenol

Mouse

Tridemorph

Mouse Rat

Teratogenic

900


245

Merkle et aI., 1984


60.2

Teratogenic

60.2

Selectively toxic

Vinclozolin

Rat

Reproductive malformations

100

Typical of

Gray et aI., 1994

endocrine disruption Zineb

Rat

Ziram

Rat

Teratogenic

2 g/kg 250

As cited in Schardein, 1993 Nakaura et aI., 1984

a Developmental toxicity includes reduced fetal weight, increased embryo/fetal mortality, and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate the presence of malformations. bDose is oral unless stated otherwise; is dose the LOEL wherever effects were observed and the NOEL when there were no effects.

transplacental transfer from mother to fetus during pregnancy (Lackmann et aI., 1999).

decrease fertility. Although it is a reproductive toxin, it does not appear to be teratogenic.

Ethylene Dibromide Ethylene dibromide (EDB; 1,2-dibromoethane), primarily a scavenger of lead compounds in gasoline, has also been used extensively as a fumigant for its chemical and biocidal properties as a soil sterilant and a spot fumigant or control agent in grain milling machinery, in the grain itself, and in fruit and vegetable infestations. In addition to its tumor-causing capabilities in rats and mice, it has been documented to cause changes in sperm morphology in bulls (Amir and Vo1cani, 1965). Spermatids appear to be the target for this compound and it has been shown to affect spermatogenesis in rat bulls and rams and to affect fertility in fowl (Alexeeff et aI., 1990). Human studies indicate that EDB may harm sperm and

Ketoconazole Ketoconazole, an imidazole antifungal agent, can compromise early pregnancy and may affect P450 enzymes of the mammalian steroidogenic system and inhibit progesterone synthesis in the ovary (Cummings et aI., 1997). It is a potential antiandrogenic agent and has displayed antihormonal activities, apparently by inhibiting ovarian hormone synthesis, resulting in delayed delivery and whole litter loss (Gray et aI., 1999a). Maneb Maneb produced fetal hydrocephalus in litters of rats receiving 480 mg/kg/day (Chemoff et aI., 1979a). In FIFRA studies, adverse developmental effects apparently occurred


because of contamination of maneb with ethylene thiourea (CDPR, 2000). The teratogenicity of a commercial formulation of the fungicide maneb (Maneb 80, containing 80% manganese ethylenebisdithiocarbamate and 20% inert ingredients) was evaluated in chick embryos. It was found to be teratogenic at all concentrations tested (0.5, 1.5, 4.5, and 13.5 g/l maneb aqueous solutions for 30 s), producing mainly unilateral lower limb deformities. No adverse effects on development were noticed after exposure to the inert ingredients (Maci and Arias, 1987). Metam Sodium Adverse effects on the reproductive system were not observed in a rat study; however, histopathology in the nasal cavity demonstrated the following: Bowman's duct hypertrophy with loss of alveolar cells, disorganization/degeneration/atrophy of olfactory epithelium, hyperplasia of olfactory epithelium, and dilatation of ducts of Bowman's glands at the high dose in females of both Fo and Fl generations (0.1 mg/ml in the drinking water). Developmental effects were noted in several studies and included severe malformations (meningocele, anophthalmia, hydrocephaly) at 60 mg/kg in the rat. The rat developmental NOEL was 5 mg/kg, based on a decrease in fetal weights, numerous skeletal developmental delays at 20 mg/kg, and higher levels and delayed ossification in hand and foot bones at 60 mg/kg. In the rabbit, postimplantation loss, early intrauterine deaths, and total litter resorptions were increased at 60 mg/kg with a developmental NOEL of 5 mg/kg/day (based on a decrease in mean live litter size, mean litter and fetal weights, and proportion of males/females at 60 mg/kg). Increases in severe defects (cleft palate and meningocele) at 60 mg/kg and skeletal variations at 20 mg/kg/day and above were also noted. In another study of Himalayan rabbits, embryotoxicity in the form of a statistically significant dose-related increase in dead implantations per pregnant animal at the mid- and high doses was noted. At the high dose, two fetuses with a neural tube closure defect (meningocele+spina bifida) were observed. This study was also acceptable under FIFRA guidelines: The developmental NOEL was 10 mg/kg/day, based on the increase in dead implantations at 30 mg/kg/day; the maternal NOEL was 30 mg/kg/day, based on decreased food consumption at 100 mg/kg/day (CD PR, 2000). Methyl Thiophanate The fungicide methyl thiophanate, widely used to control some of the most common fungal diseases in crops, is metabolized in animals into benzimidazole compounds, including the reproductive toxic ant carbendazim. However, standard toxicological tests did not indicate that methyl thiophanate may cause testicular toxicity and/or embryotoxicity, which are typical effects of many benzimidazoles. In the B6C3F1 mouse, in spite of the high doses administered, none of the testicular parameters examined (sperm head count, specific enzyme activities, histopathology on days 335 postdosing) showed significant alterations as compared to the controls at any time postdosing. Pregnant CD rat dams administered orally the limit dose of 650 mgkg- 1 body weight

405

day-l during the preimplantation (gestational day or GD 2-5) or peri-implantation (GD 6-9) phase showed maternal toxicity, with only marginal reductions of the growth of embryos and adnexa (Traina et aI., 1998). Earlier studies submitted by Atochem North America, Inc. (1985) for methyl thiophanate showed for the rat a teratogenic NOEL of 2500 ppm (125 mg/kg/day), based on the high dose tested (HDT), with a maternal NOEL of 250 ppm (12.5 mg/kg/day), an LEL of 1200 ppm (60 mg/kg/day), and a fetotoxic NOEL of 2500 ppm at HDT (EPA-IRIS database). For mice [study by Pennwalt Corp. (1970a)], the 1000mg/kg/day dose caused a decreased number of implantations; other details were unavailable because fetal examinations did not appear to include soft-tissue examinations. A threegeneration reproduction study in the rat [study by Pennwalt Corp. (1970b)] demonstrated a reproductive NOEL of 160 ppm (8 mg/kg/day) and an LEL of 640 ppm (32 mg/kg/day; HDT), based on reduced litter weights (EPA, IRIS Database). Pentachlorophenol Pentachlorophenol (PCP) is used primarily as a wood preservative. It has been shown to be fetotoxic and teratogenic during early gestation. Commerical PCP is contaminated with chlorinated dioxins and dibenzofurans, tetrachlorophenols, and hydroxychlorodiphenyl ethers (Williams, 1982), which can exert their own effects. Additionally, PCP was reported to be a contaminant in commercial creosote preparations used in wood preservation and may have contributed to its early fetotoxicity. Pentachlorophenol was not teratogenic in rats (Schwetz et aI., 1974). In studies submitted for registration, PCP was found to have adverse effects in the rat developmental toxicity study due to fetal resorptions, decreased fetal weights, ossification delays, and malformations, which findings cannot be assured to result strictly from maternal toxicity (CDPR, 2000). The maternal NOEL was 30 mg/kg/day (body weight and food consumption decrements) and the developmental NOEL was 30 mg/kg/day (increased fetal resorptions, decreased fetal weights, a modest incidence of malformations such as gastroschisis, hydrocephaly, and diaphragmatic hernia judged to be treatment-related, although not statistically significant; significant increase in incidence of dilated pelves delayed ossification in several areas and increased mean numbers of thoracic vertebrae and associated increased incidence of 14th ribs). Another study found a possible adverse effect in that a relatively low developmental toxicity NOEL was observed in the absence of maternal toxicity. A maternal NOEL of 200 ppm (13 mg/kg/day), based on reduced weight gain, clinical signs such as ringed eye, and possible vaginal hemorrhaging, and a developmental NOEL of 60 ppm (4 mg/kg/day), based on reduced fetal weights, misshapen centra, and a possible treatment-related increase in resorptions (significant increase in females with more than two resorptions), were noted (FDA, 1987), indicating a possible adverse effect. Terrazole Decreased live litter size, fetal weight and pup survival (24 h); increased resorptions; and malformations were noted at 45 mg/kg/day in rabbits. The developmental NOEL

406

CHAPTER 14 Developmental and Reproductive Toxicology of Pesticides

was 15 mg/kg/day; Adverse effects were indicated; even though the NOELs for maternal toxicity and developmental toxicity were equivalent, the developmental effects were quite marked (total resorptions equaled 31 in the high dose compared to 7 in the control; 24 h survival of 80% in the high dose compared to 99% in the control). The increased incidence of malformations in the high dose included tail defects, underdeveloped hind limbs, and crossed hind legs (CDPR, 2000).

VincIozoIin This fungicide has the unique claim of being a compound that results in abnormal rodent sex differentiation following exposure during critical stages of life. Effects such as hypospadias, ectopic testes, vaginal pouches, agenesis of the ventral prostate, and nipple retention in male rats were commonly observed (Gray et aI., 1994). In the FIFRA reproduction study, failure of F] males to acquire normal anatomical and functional male characteristics, marked retardation in neonatal growth and survival at dose levels not commensurately toxic to adults, and lenticular degeneration were the principal possible adverse effects. In the FIFRA developmental toxicity study in rats, decreased anogenital distance in males, a finding that was repeated in all the studies conducted and interpreted as feminization of male fetuses, was observed. Similar findings were not noted in the mouse or rabbit (CD PR, 2000). 14.4.4 RODENTICIDES Reviews on the effects of warfarin exposure indicate an uncommon, but strikingly similar pattern of congenital anomalies in children born to women exposed to the compound. The syndrome consists of nasal hypoplasia, stipled epiphyses and growth, retinal-optic atrophy and central nervous system anomalies (Friedman and Polifka, 1994; Schardein, 1993). Although warfarin is used as a rodenticide, it may also be administered to women with heart valve prosthesis. Ginsburg and Barron (1994) recommended not giving warfarin to women between 6 and 12 weeks of gestation.

also noted for commonly used compounds such as piperazine and ivermectin exposures.

Benzamidazole Family of Compounds Variable, but teratogenic potential has been noted for the benzamidazoles as a group. As reviewed by Schardein (1993), skeletal abnormalities were reported in sheep and rats exposed to parbendazole; however, teratogenicity was not reported at comparable or higher doses in hamsters, rabbits, cattle, and swine. Cambendazole was noted to have induced multiple defects in rats and sheep; flubendazole was found to be developmentally toxic in rats, producing multiple malformations; and mebendazole induced malformations in rats with up to 100% incidence in a dosedependent manner but was not teratogenic in rabbits even at high doses. Oxyfenbendazole induced multiple abnormalities in rats and sheep (Schardein, 1993) and swine (Morgan, 1982). Parbendazole has a safety index of over 30 times the recommended dose in healthy animals, but may be teratogenic at doses only slightly higher than the recommended one. It was parbendazole that first alerted scientists to the embryotoxicity of benzamidazoles (Manger, 1991). Thiabendazole is used as a veterinary anthelmintic and fungicide and has variable teratogenic effects in animals. Teratogenicity has been noted in some laboratory animal species (mice and rats), but other studies and reviews have claimed it to be relatively safe (Manger, 1991; Schardein, 1993). In acceptable FIFRA studies submitted for registration, adverse effects were not noted (CD PR, 2000). Another benzamidazole, benomyl (as reported earlier in this chapter), has also demonstrated fetotoxicity and teratogenicity findings in the absence of obvious maternal toxicity, indicating possible adverse effects.

See Tables 14.5 and 14.6 for the toxicity profiles of some miscellaneous pesticides and animal health pesticides.

Antimalarials Chloroquine and its congeners, which are inhibitors of dihydrofolate reductase, and primaquine are known to exert teratogenic effects, but because they are under the category of prescription drugs, the likelihood of exposure during pregnancy is low. Defects noted after quinine exposure include deafness due to auditory nerve hypoplasia, optidisc problems, limb anomalies, and visceral malformations as well as fetal deaths (Schardein, 1993). Rates of spontaneous abortion and birth defects were comparable in pregnant women taking mefloquine, compared with chloroquine-proguanil, or pyrimethamine-sulfadoxine prophylaxis in the first trimester of pregnancy (Phillips-Howard and Wood, 1996). Teratogenic effects for mefloquine were observed in animals but data from humans are lacking (Vanhauwere et aI., 1998).

Anthelmintics Several case reports have been published associating anthelmintic drugs with the induction of birth defects in humans. Ectromelia in infants whose mothers were treated with a tin-based tenifuge; multiple malformations (brain, jaw, ear, limb, and heart defects) subsequent to the intake of mebendazole during the first month of pregnancy; and the incidence of spina bifida and renal anomalies along with hydrocephalus due to quinacrine administration in the first trimester are noted in a review of the data (Schardein, 1993). Negative reports are

ImidacIoprid Imidacloprid is widely used against fleas in dogs and cats and also as an insecticide for use on soil, seed, or foliar treatment in rice, cereal, vegetables, cotton, and turf to control ricehoppers, thrips, termites, turf and soil insects, and some beetle species. In a rat developmental toxicity study, a high percentage of male fetuses and increased incidence of wavy ribs were noted at 94.1 mg/kg/day, indicating a possible adverse effect. The maternal NOEL was 25.9 mg/kg/d, based on decreased body weight gain and reduced food consumption of

14.4.5 ANIMAL HEALTH PRODUCTS, FUMIGANTS, AND MISCELLANEOUS PESTICIDES


407

Table 14.5 Developmental and Reproductive Toxicity Profile of Miscellaneous Pesticidesa Dose Chemical

Species

Toxicity profileb

Acrylonitrile

Rat

Teratogenic

(mg/kg)"

Comments


25

Testicular toxin Arsenic trioxide

WiIIhite, 1981 10

Rat

References

Teratogenic by

Stump et aI., 1998a---{;

ip route Reproduction Benzenesulfonic acid

Mouse

Embryo/fetal mortality

Four generations As cited in Schardein, 1993

5.5

hydrazide Diet

Morita et aI., 1981

Benzylbenzoate

Rat

1%

Biphenyl

Rat

500


Busan 77

Rabbit

125

Drake et aI., 1990

Chlorofebrifugine

Rat

9.3 (po) or


6ppm (diet) Chloropicrin

Rat


3.5 ppm

Inhalation

York et aI., 1994

route Rabbit


2ppm

Inhalation route

Chlorosil

Rat

100

CycIonite

Rat

50

Boikova et aI., 1981 Minor et aI., 1982

Reproduction Rabbit Dikurin

Rat

Diphenylamine

Rat

20 20 Renal lesions

1.5%

Reproduction Ethylene oxide

Mouse

Shepelskaya, 1988 Diet


2-G study

DevelopmentaIly toxic

150

iv route: also

Teratogenic

150

findings by

As cited in Schardein, 1993 RutJedge and Generoso, 1989

inhalation route Rat


10ppm

Reproduction Rabbit

Embryotoxicity

Gliftor

Mouse

Reduced fertility

Guanylthiourea

Rat

Teratogenic

Methyl bromide

Mouse


route 9

iv route

300

Kimmel et aI., 1982 As cited in Schardein, 1993

33 250

Rat

70 Testicular toxicity in Agenesis of gall bladder

Inhalation


route

reproductive studies Rabbit

Inhalation

2-G studies 70

Inhalation

Hardin et aI., 198 I

route Methylisothiozolinone

Rat

ScialIi et aI., 1995

15 Reproduction

Rabbit

N-methyl-N-(l-

Mouse

naphtyl)-fluoro

1.5 Fetal growth retardation

0.1 ppm


Drudge et aI., 1983 As cited in Schardein, 1993 Reproductively toxic in man

Diethylcarbamazine

Rat

100

Rabbit

200

2 x use

Dog

level Diethylcarbamazine and

13.2/10

Dog

Fenbendazole

Rat

Flubendazole

Rat

Over two

Rat

120 Teratogenic

40 40

Neonatal mortality and

::::0.4


decreased pup growth

Rodwell et aI., 1987

18/15

Dog


Muitigeneration


study 100 flglkg

Primate Ivermectin and pyrantel


generations generations

oxibendazole

Ivermectin

Over two

As cited in Schardein, 1993 Whorton and Foliart, 1988

In infants One-generation study

Mebendazole

Rat

Embryotoxic

10

Teratogenic

10


Rabbit

40


Naftalofos

Rat

15


Netobimin

Rat

Nitroxynil

Sheep

Teratogenic

71 34

(continues)


409

Table 14.6 (continued) Dose (mg/kg)C

IChemical

Species

Toxicity profileb

Oxfendazole

Rat

Teratogenic

16

Sheep

Teratogenic

23

Oxibendazole

14

13.5

Morgan, 1982

Mouse

30


Rat

149 Embryotoxic

10

Teratogenic

10

Abortion

60

Cow Pig Sheep

10 100

Hamster

90 Teratogenic

60

Mouse

400

Rat

400

Piperazine

Pig

PW 16

Rat

Pyrante1

Rat

400 15,000 Skeletal variations Reproductive effects

Rabbit

As cited in Schardein, 1993 3-0 study

Reproduction Rabbit

Ziborov et aI., 1982

440


440 3,000

Miscarriage

Horse Sodium arsenate


Pig

Rabbit

Permethrin

References

Cow

Rat Parbendazole

Comments

1,000 12.5

Mouse

120

ip/iv doses

Hood et aI., 1982

teratogenic in other species Terrazo1e

Rabbit

Increased resorptions and

15

CDPR Database, 1999

malformations Thiabendazole

Mouse

Teratogenic

700

Rat

Teratogenic

500

Sheep Tribendimin

Rat

Triclabendazole

Rat


200 200 Reduced fetal growth

100

a Includes

acaracides and anthelmintics. bDevelopmental toxicity includes reduced fetal weight, increased embryo/fetal mortality, and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate the presence of malformations. cDose is oral unless stated otherwise; dose is the LOEL wherever effects were observed or the NOEL when there were no effects.

the 94.1-mg/kg/day treatment group; the developmental NOEL was 25.9 mg/kg/day, based on increased incidence of wavy ribs in the fetuses of the 94.1-mg/kg/day treatment group (CDPR, 2000).

Methyl Bromide This gas has been used extensively as a fumigant to combat nematodes in strawberries and tomatoes. Alternatives to its use are needed due to its ozone-depleting properties and it is slated for replacement as per the Montreal protocol. Chemically, it is an alkylating agent and capable of neurotoxicity. The compound appears to demonstrate extreme differences among species, dogs being unable to tolerate doses

severalfold lower than those in the rat. Genetic polymorphism for the metabolism of this compound has also been noted. Exposure to methyl bromide in a two-generation reproduction study in Sprague-Dawley rats by inhalation affected fertility (the fertility index decreased from 90.9% in the controls to less than 68% in the 30- and 90-ppm groups) and decreased the body weights of parental and reduced the growth of neonatal rats. Pregnant animals were only exposed 5 days/week (for a total of 14-15 days) during their pregnancy and the pups were not directly exposed until after weaning on postnatal day 28. The parental NOAEL was 3 ppm (reduced fertility). The progeny NOAEL was 3 ppm, based on decreased pup body

410

CHAPTER 14


weight and reduced organ weights, including reduced F 1 brain weight/reduced width of the cerebral cortex. Data submitted for registration purposes were found to be marginally acceptable, but did not conclusively demonstrate the absence of neurotoxic potential. The developmental study in New Zealand white rabbits demonstrated maternal toxicity at 80 ppm (311 mg/m 3 ) such as reduced body weight and weight gain and clinical signs of central nervous system toxicity. Fetal effects that were not statistically significant but quite rare (i.e., considered biologically significant) were noted. These included omphalocele, hemorrhaging with or without hydrops, and retroesophageal right subclavian artery. Also gall bladder agenesis, fused sternebrae, and decreased fetal body weight were statistically significant at 80 ppm resulting, in a NOAEL for maternal toxicity and developmental effects of 40 ppm (155 mg/m3 ). In rats, a NOEL of 20 ppm for a developmental toxicity study based on delayed skeletal ossification and a maternal NOEL of 70 ppm or more were noted, but the study did not test at a high enough dose level. Reports of other studies via oral exposure in rats and another strain of rabbit demonstrated microphthalmia in rats and some skeletal malformations in the rabbit though not in a doseresponsive pattern (Kaneda et aI., 1998). The reports did not meet FIFRA specifications and historical negative-control data for the rabbit strain employed (Kbl:JW) are not generally available in the open literature. The oral route provides accurate dosage, but because it is more likely to be metabolized prior to reaching the brain than the inhalation route, it may be argued that a higher dosage may be needed to compare the oral route with the inhalation route. Although the pharmacokinetics of transplacental transfer of methyl bromide gas is not available because methyl bromide is known to have neurotoxic potential, and human exposure is most likely via inhalation or skin, inhalation may be the preferred route to detect neurotoxic damage. Hence, adverse effects to development were not observed in the oral studies, but the inhalation studies did demonstrate adverse effects in both reproduction and developmental toxicity studies.

than that for developmental toxicity, and one-quarter of these have been documented to be reproductive toxicants to humans and laboratory animals. There is greater concordance between laboratory animal models and humans for adverse effects on fertility than in the area of developmental effects; for example, male reproductive toxic ants acting on the testes in laboratory animals have the same site of action in humans (Schwetz, 1994). Developmental toxicity in animals, however, does not translate to the same kind of developmental defects in humans (Kimmel et aI., 1993; Schardein, 1989). This lack of concordance, noted for both drugs and other chemicals of commerce, has led to the interpretation that some adverse developmental effect in an animal study is potentially predictive of some adverse developmental effect in humans. Further confounding the issue are accounts that around 6070% of pesticides registered for use have not been adequately tested at the laboratory or clinical level (Mott and Snyder, 1987) and that 25% of the compounds exported from the United States are, in fact, banned or unregistered in the United States (Schardein, 1993). In evaluating global exposure patterns, the data submitted to regulatory agencies become more valuable. Regulatory agencies in the European Community are moving to reduce the number of experimental animals that are being sacrificed for studies. Although such a trend is helping to reduce unnecessary wastage of experimental animals, in vivo data submitted to agencies in the United States often serve as critical studies for a specific compound. Regulations in Japan (Ministry for Agriculture, Forests and Fisheries) have some similarities to U.S. reguations and so duplication of studies can occur, but the benefits of such studies serve to reduce the likelihood of a thalidomide disaster. The aim of testing and regulation is thus to minimize the liability to the manufacturers of chemicals, while assuring that the public will be exposed to a safe set of pesticides. This chapter discussed the majority of pesticides that have been used and reviewed the available data on their reproductive and developmental toxicity. Details for a specific compound or new active ingredient may be obtained from the databases that are accessible to the public at state and federal agencies or from publications.

14.5 CONCLUSIONS Over 4500 chemical tests have been reported using the FDA Segment 11 protocol and more than one-third have come up as positive for developmental toxicity (J. L. Schardein, personal communication). Approximately 25-30 chemicals or families of chemicals are considered human teratogens, based either on a Segment 11 or FIFRA study or on other animal models that have been optimized to detect a toxic effect (Schwetz, 1994). The pilot range-finding studies, in addition to helping to narrow the dose level, have also served to be predictive of the definitive study. Although much is known about the mode of action of some chemicals, the complexity of development suggests that there may be multiple mechanisms of interference with normal development. These mechanisms are not known even for known human teratogens. Similarly, hundreds of chemicals have been tested using protocols for reproductive toxicity; lO-fold less

REFERENCES Abd El-Khalik, M. M., Hanafy, M. S., and Abdel-Aziz, M. 1. (1993). Studies on the teratogenic effects of deltamethrin in rats. Deutsche Tierarztliche Wochenschrift 100, 142-143. Agarwal, R. c., Kumar, S., and Mehrotra, N. K. (1996). Micronucleus induction by diuron in mouse bone marrow. Toxicol. Lett. 89, 1-4. Alexeeff, G. v., Kilgore, W W, and Li, M. Y. (1990). Ethylene dibromide: Toxicology and risk assessment. Rev. Environ. Contam. Toxicol. 112, 49122. Al-Hachim, G. M., and Fink, G. B. (1968). Effect of DDT or parathion on conditioned avoidance response from DDT or parathion treated mothers. Psychopharmacologia 12,424-427. Amir D., and Volcani, R. (1965). Effect of dietary ethylene dibromide on bull semen. Nature (London) 206, 99-100. An der Lan, H. (1969). Possibilities of damage of progeny due to pesticides in warmblooded animals. Zentralbl. Bakeriol. Parasitenk. Abt. 1 Orig. 210, 234-240.

References

Andrews, J. E., and Courtney, K D. (1976). Inter- and intralitter variation of hexachlorobenzene (HCB) deposition in fetuses. Toxicol. Appl. Pharmacol. 37,128. Angerhofer, R. A., and Weeks, M. H. (1981). "Effect of Dermal Applications of N, N -Diethyl-m-toluamide (m-DET) on the Embryonic Development of Rabbits." USA EHA 75-51-00347; AD-A094778, 22 pp. Anonymous (1970). Another herbicide on the blacklist. Nature 226,309-311. Anonymous (1980). Report to the minister of health of an investigation into allegations of an association between human congenital defects and 2,4,5-T spraying in and around Te Kuiti. N. Z. Med. J. 91,314-315. Arbuckle, T. E., Savitz, D. A., Mery, L. S., and Curtis, K M. (1999). Exposure to phenoxy herbicides and the risk of spontaneous abortion. Epidemiology 10(6),752-760. Aschengrau, A., and Monson, R. R. (1989). Paternal military service in Vietnam and risk of spontaneous abortion. 1. Occup. Med. 31, 618-623. Astroff, A. B., Sangha, G. K, and Thyssen, J. H. (1996). The relationship between organophosphate-induced maternal cholinesterase inhibition and embryo/fetal effects in the Sprague-Dawley rat. Toxicologist 30, 191. Astroff, A. B., Young, A. D., Freshwater, K 1., Sangha, G. K., and Thyssen, J. H. (1998). Developmental and reproductive toxicity of KBR3023, a new insect repellent, in the Sprague-Dawley rat. Toxicologist 42, 255. Bage, G., Cekanova, E., and Larsson, K S. (1973). Teratogenic and embryotoxic effects of the herbicides di- and trichlorophenoxyacetic acids (2,4-D and 2,4,5-T). Acta Pharmacol. Toxicol. (Copenhagen) 32, 408-416. Baker, S. R., and Wilkinson, e. E (eds.) (1990). "Effects of Pesticides on Human Health." Advances in Modern Environmental Toxicology, Vo!. 18, Princeton Sci. Pub., Princeton, NJ. Barlow, S. M. (1985). United Kingdom: regulatory attitudes toward behavioural teratology testing. Neurobehav. Toxicol. Terato/. 7, 643-646. Barlow, S. M., and SuIIivan, E M. (1982). "Reproductive Hazards of Industrial Chemicals." Academic Press, London. Bartoov, B., Ventura, v., Ailenberg, M., Potashnik, G., and Mayevsky, A. (1987). Effects of dibromochloropropane and ethylene dibromide on biochemical events and ultramorphology of ejaculated ram spermatozoa in vitro. Am. J. Ind. Med. 11,647-658. Bates, H. K, and LaBorde, J. B. (1986). Developmental toxicity evaluation of soman in New Zealand white (NZW) rabbits. Teratology 33, Bates, H. K, and LaBorde, J. B. (1987). Developmental toxicity evaluation of soman in CD rats. Toxicologist 7, 174. Baxley, M. N., Hood, R. D., Vedel, G. C., Harrison, W. P., and Szczech, G. M. (1981). Prenatal toxicity of orally administered sodium arsenite in mice. Bull. Environ. Contam. Toxicol. 26, 749-756. Beck, S. L. (1977). Postnatal detection of prenatal exposure to herbicides in mice, using normally occurring variations in skeletal development. Teratology 1515A. Behera, B. e., and Bhunya, S. P. (1990). Genotoxic effect of isoproturon (herbicide) as revealed by three in vivo mutagenic bioassays. Indian J. Exp. BioI. 28,862-867. BeIIini, J. (1986). "High Tech Holocaust." Graham Tarrant, London. Benes, v., Sram, R. J., and Tuscany, R. (1973). Testing of mutagenicity of fenitrothione. Mutat. Res. 21, 23-24. Berge, G. N. and Nafstad, I. (1983). Teratogenicity and embryotoxicity of orally administered fench10rphos in blue foxes. Acta Vet. Scand. 24,99-112. Bhatnager, P., and Soni, I. (1988). Evaluation of the teratogenic potential of phosphamidon in mice by gavage. Toxicol. Lett. 42, 101-107. Bhunya, S. P., and Pati P. e. (1990). Effect of deltamethrin, a synthetic pyrethroid, on the induction of chromosome aberrations, micronuclei and sperm abnormalities in mice. Mutagenesis 5, 229-232. Bieyl, D. W. (1980). Embryotoxicity and teratogenicity of phosmet in mice. Arch. Exp. Veterinaermed. 34,791-795. Birnbaum, L. S. (1998). Developmental effects of dioxin. In "Reproductive and Developmental Toxicology" (K S. Korach, ed.), pp. 87-112. Dekker, New York. Bimbaum, L. S., Morrissey, R. E., and Harris, M. W. (1991). Teratogenic effects of 2,3,7,8-tetrabromodibenzo-p-dioxin and three polybrominated dibenzofurans in C57BU6N mice. Toxicol. App/. Pharmacol. 107, 141-152.

ne.

411

Bitsi, G. A., Singh, K., Khan, S. U., Akhtar, M. H., Kacew, S., and White, N. D. (1994). Fate of wheat bound malathion residues in rats during gestation. Chemosphere 29, 451-455. Black, W. D., VaIIi, V. E. 0., Ruddick, J. A., and ViIIeneuve, D. C. (1983). The toxicity of three trichlorobenzene isomers in pregnant rats. Toxicologist 3, 30. BlakIey, P. M., Kim, J. S., and Firneisz, G. D. (1989). Effects of paternal subacute exposure to Tordon 202c on fetal growth and development in CD-1 mice. Teratology 39, 237-241. Bleavins, M. R., Aulerich, R. J., and Ringer, R. K (1984). Effects of chronic dietary hexachlorobenzene exposure on the reproductive performance and survivability of mink and European ferrets. Arch. Environ. Contam. Toxico/. 13, 357-365. Blekherman, N. A., and Iiyina, V. I. (1973). Changes of ovary function in women in contact with organochlorine compounds. Pediatriya 52, 57-59. Boikova, V. v., Golikow, S. N., Korkhov, V. v., and Mots, M. N. (1981). Reproductive studies with a new cholinolytic chlorosil in rats. Farmakol. Toksikol. 44,83-95. Bolliger, e. T, van Zijl, P., and Louw, J. A. (1992). Multiple organ failure with the adult respiratory distress syndrome in homicidal arsenic poisoning. Respiration 59, 57-61. Borzsonyi, M., Pinter, A., SuIjan, A., and Torok, G. (1978). Carcinogenic effect of a guanidine pesticide administered with sodium nitrite on adult mice and on the offspring after prenatal exposure. Cancer Lett. 5, 107-113. Boucard, M., Beaulation, L. S., Mestres, R., and Allieu, M. (1970). Experimental study of teratogenesis: Influence of the timing and duration of the treatment. Therapie 25, 907-913. Bower, C., and Stanley, E J. (1980). Herbicides and cleft lip and palate. Lancet 2,1247. Boyes, W. K, Dourson, M. L., Patterson, J., TiIson, H. A., Sette, W. E, MacPhail, R. e., Li, A. A., and O'Donoghue, J. L. (1997). EPA's neurotoxicity risk assessment guidelines. Fundam. Appl. Toxicol. 40,175-184. Branch, S., Rogers, J. M., Brownie, C. E, and Chernoff, N. (1996) Supernumerary lumbar rib: Manifestation of basic alteration in embryonic development of ribs. J. Appl. Toxicol. 16, 115-119. Breslin, W. J., Schroeder, R. E., and HanIey, T R. (1991). Developmental toxicity of picloram potassium (K) and triisopropanolamine (TIPA) salts in the rat. Toxicologist 11, 74. Breslin, W. J., Vedula, U., Zablotny, C. L., and Stebbins, K E. (1994). Developmental toxicity studies with picIoram triisopropanolamine salt and picIoram 2-ethylhexyl ester in the rabbit. Toxicologist 14, 162. Breslin, W. J., BiIIington, R., and Jones, K (1996a). Evaluation of the developmental toxicity of tricIopyr triethylamine salt (TTEA) and tricIopyr butoxyethyl ester (TBEE) in rats. Teratology 53, 106. Breslin, W. J., Liberacki, A. B., Dittenber, D. A., and Quast, J. E (I 996b). Evaluation of the developmental and reproductive toxicity of chlorpyrifos in the rat. Fundam. Appl. Toxicol. 29, 119-130. Briggs, S. A., and the staff of the Rachel Carson Council (1992). "Basic Guide to Pesticides: Their Characteristics and Hazards." Hemisphere, Washington, De. Brogan, W. E, Brogan, e. E., and Dadd, J. T (1980). Herbicides and cleft lip and palate. Lancet 2, 597. Budreau, C. H., and Singh, R. P. (1973). Teratogenicity and embryotoxicity of demeton and fenthion in CF #1 mouse embryos. Toxicol. Appl. Pharmacol. 24, 324-332. Buist, R. (1986). "Food Chemical Sensitivity: What It Is and How to Cope with It." Avery, Garden City, NY. Bus, J. S., Preache, M. M., Cagen, S. Z., Posner, H. S., EIiason, B. C., Sharp, e. w., and Gibson, J. E. (1975). Fetal toxicity and distribution of paraquat and diquat in mice and rats. Toxicol. App/. Pharmacol. 33, 450460. Buschmann, J., Clausing, P., Salecki, E., Fischer, B., and Peetz, U. (1986). Comparative prenatal toxicity of phenoxyaIcanic herbicides: Effects on postnatal development end behavior in rats. Teratology 33, IIA-12A. Byrd, R. A., and Markham, J. K (1990). Developmental toxicity of dinitroanilines. I. Trifluralin. Teratology 41, 542-543.

412

CHAPTER 14


Byrd, R. A, Adams, E. R., Robinson, K., and Markham, J. K. (l990a). Developmental toxicity of dinitroanilines. n. Ethalfluralin. Teratology 41, 542. Byrd, R. A., Jordan, W. H., and Markham, J. K. (l990b). Developmental toxicity of dinitroanilines. rn. Oryzalin. Teratology 41, 542. California Department of Pesticide Regulation (CD PR) (2000). Database-Available Toxicology Summaries and "CDPR Risk Characterization Documents." http://www.cdpr.ca.gov/. "California Organic Foods Act of 1990," California Food and Agriculture Code 4600. Cam, C. (1960). Une nouvelle dermatose epidemique des enfants. Ann. Dermatol. Syphiligr. 87, 393-397. Cam, C., and Nigogosyan, G. (1963). Acquired toxic porphyria cutonia tarda due to hexachlorobenzene. J. Am. Med. Assoc. 183,88-91. Campbell, et al. (1985). Canon, S. B., and Kimbrough, R. D. (1979). Short-term chlordecone toxicity in rats including effects on reproduction, pathological organ changes and their reversibility. Toxicol. Appl. Pharmacol. 47, 469-476. Cantalamessa, E, Barili, P., Cavagna, R., Sabbatini, M., Tenore, G., and Amenta, E N. A. (1998). Influence of neonatal treatment with the pyrethroid insecticide cypermethrin on the development of dopamine receptors in the rat kidney. Mech. Aging Dev. 103, 165-178. Carmelli, D., Hoflherr, J., Tomsic, J., and Morgan, R. W. (1981). A case-control study of the relationship between exposure to 2,4-D and spontaneous abortions in humans. SRI International. Camey, E. W., Schroeder, R., and Breslin, W. J. (1995). Developmental toxicity study in rats with fluroxyopur methylheptyl ester. Teratology 51, 180. Carson, R. (1962). "Silent Spring." Houghton Mifflin, Boston. Carter, S. D., Hein, J. E, Rehnberg, G. L., and Laskey, J. W. (1984). Effect of benomyl on the reproductive development of male rats. J. Toxicol. Environ. Health 13, 53-68. Casey, P. H., and Collie, W. R. (1984). Severe mental retardation and multiple congenital anomalies of uncertain cause after extreme parental exposure to 2,4-D. J. Pediatr. 104,313-315. Castro, V. L., Bernardi, M. M., and Palermo-Neto, J. (1992). Evaluation ofprenatal aldrin intoxication in rats. Arch. Toxicol. 66, 149-152. CDFA (1987). Chatterjee, S., Ray, A., Ghosh, S., Bhattacharya, K., Pakrashi, A., and Deb, C. (1988). Effect of aldrin on spermatogenesis, plasma gonadotrophins and testosterone, and testicular testosterone in the rat. J. Endocrinol. 119, 7581. Chauhan, L. K., Agarwal, D. K., and Sundaraman, V. (1997). In vivo induction of sister chromatid exchange in mouse bone marrow following oral exposure to commercial formulations of alpha-cyano pyrethroids. Toxicol. Lett. 93,153-157. Chernoff, N., Kavlock, R. J., Beyer, P. E., and Miller, D. (1987). The potential relationship of maternal toxicity, general stress, and fetal outcome. Teratogen. Carcinogen. Mutagen. 7, 241-253. Chernoff, N., Kavlock, R. J., Kathrein, J. R., Dunn, J. M., and Haseman, J. K. (1975). Prenatal effects of dieldrin and photodieldrin in mice and rats. Toxicol. Appl. Pharmacol. 31,302-308. Chernoff, N., Kavlock, R. J., Rogers, E. H., Carver, B. D., and Murray, S. (1979a). Perinatal toxicity of maneb, ethylene thiourea, and ethylenebisisothiocyanate sulfide in rodents. J. Toxicol. Environ. Health 5,821-834. Chernoff, N., Linder, R. E., Scotti, T. M., Rogers, E. H., Carver, B. D., and Kavlock, R. J. (1979b). Fetotoxicity and cataractogenicity of mirex in rats and mice with notes on kepone. Environ. Res. 18,257-269. Cochran, R. c., and Wiedow, M. A. (1984). Chlordecone lacks estrogenic properties in the male rat. Toxicol. Appl. Pharmacol. 76, 519-525. Colie, C. E (1991). Male-mediated teratogenesis. Reprod. Toxicol. 7,3-9. Collins, T. EX., Sprando, R. L., Shackleford, M. E., Hansen, D. K., and Welsh, J. J. (1999). Food and drug Administration proposed testing guidelines for developmental toxicity studies. Regul. Toxicol. Pharmacol. 30, 39-44. Constable, J. D., and Hatch, M. C. (1984). Agent Orange and birth defects. N. Engl. J. Med. 310, 653-654. Contreras, H. R., and Bustos-Obregon, E. (1999). Morphological alterations in mouse testis by a single dose of malathion. J. Exp. Zool. 284, 355-359.

Cook, J. C., Mullin, L. S., Frame, S. R., and Biegel, L. B. (1993). Investigation of a mechanism for Leydig cell tumorigenesis by linuron in rats. Toxicol. Appl. Pharmacol. 119, 195-204. Cooper, C. L., Ozoktay, S., Tafreshi, M., and Alexander, L. I. (1983). Anencephaly: Agent Orange implications? J. Natl. Med. Assoc. 75, 93-94. Cooper, R. L., Stoker, T. E., Goldman, J. M., Parrish, M. B., and Tyrey, L. (1996). Effect of atrazine on ovarian function in the rat. Reprod. Toxicol. 10,257-264. Costlow, R. D., and Manson, J. M. (1980). Herbicide-induced hydronephrosis and respiratory distress: Effects of in utero exposure to nitrofen (2,4dichloro-4' -nitrophenyl ether). Toxicol. Appl. Pharmacol. (Abstract). Costlow, R. D., Hayes, A w., Moss, J. N., Smith, J. M., Rodwell, D., and Weatherholz, W. (1983). The effects of gestational exposure of rats and rabbits to Kathon biocide. Toxicologist 3, 17. Costlow, R. D., Lutz, M. E, Kone, W. w., Hurt, S. S., and O'Hara, G. P. (1986). Dinocap: Developmental toxicity studies in rabbits. Toxicologist 6, 85. Courtney, K. D. (1979). Hexachlorobenzene (HCB): A review. Environ. Res. 20, 225-226. Courtney, K. D., and Andrews, J. E. (1980). Extra ribs indicate fetotoxicity and maldevelopment. Teratology 21, 35A. Courtney, K. D., and Moore, J. A (1971). Teratology studies with 2,4,5trichlorophenoxy acetic acid and 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. Appl. Pharmacol. 20, 396-403. Courtney, K. D., Andrews, J. E., and Ebron, M. T. (1977). Teratology study of pentachlorobenzene in mice: No teratogenic effect at 50 or 100 mg. J. R. Chem. Soc. Med. Sci. Uhr. Compend. 5,587. Courtney, K. D., Andrews, J. E., Springer, J., and Dalley, L. (1985). Teratogenic evaluation of pesticides baygon, carbofuran, dimethoate and EPN. J. Environ. Sci. Health, B 20, 373-406. Courtney, K. D., Gaylor, D. w., Hogan, M. D., and Falk, H. L. (1970a). Teratogenic evaluation of pesticides: A large-scale screening study. Teratology 3, 199. Courtney, K. D., Gaylor, D. w., Hogan, M. D., and Falk, H. L. (l970b). "Record of the Hearing on 2,4,5-T before the Subcommittee on Energy, Natural Resources and the Environment of the Senate Committee on Commerce, April 15," p. 225. Courtney, K. D., Gaylor, D. W., Hogan, M. D., Falk, H. L., Bates, R. R., and Mitchell, L. (l970c). Teratogenic evaluation of 2,4,5-T. Science 168, 864866. Couture, L. A. (1990). 2,3,7,8-Tetrachlorodibenzo-p-dioxin Induced Hydronephrosis: Characterization of the Peak Period of Sensitivity for Placentally- and Lactationally-Induced Renal Lesion, and Assessment of Persistence. Ph.D. Dissertation, University of North Carolina, Chapel Hill. Couture, L. A, Abbott, B. D., and Birnbaum, L. S. (1990). A critical review of the developmental toxicity and teratogenicity of2,3,7,8-tetrachlorodibenzop-dioxin: Recent advances toward understanding the mechanism. Teratology 42, 619-627. Craig, G. R., and Ogilvie, D. M. (1974). Alterations of T-maze performance in mice exposed to DDT during pregnancy and lactation. Environ. Physiol. Biochem. 4, 189-199. Cranmer, J. S., Avery, D. L., and Bamett, J. B. (1979). Altered immune competence of offspring exposed during development to the chlorinated hydrocarbon pesticide chlordane. Teratology 19, 23A. Crofton, K. M., Dean, K. E, Boncek, V. M., Rosen, M. B., Sheets, L. P., Chernoff, N., and Reiter, L. W. (1989). Prenatal or postnatal exposure to bis(tri-n-butyltin)oxide in the rat: Postnatal evaluation ofteratology and behavior. Toxicol. Appl. Pharmacol. 97, 113-123. Crozet, N., and Szollosi, D. (1979). The effects of isopropyl-N-phenyl carbamate on meiotic maturation of mammalian oocytes. Ann. BioI. Anim. Biochim. Biophys. 19, 1131-1140. Cummings, A. M. (1997). Methoxychlor as a model for environmental estrogens. Crit. Rev. Toxicol. 27, 367-379. Cummings, A M., and Gray, L. E. (1989). Antifertility effect of methoxychlor in female rats: Dose- and time-dependent blockade of pregnancy. Toxicol. Appl. Pharmacol. 97, 454-462. Cummings, A. M., Ebron-McCoy, M. T., Rogers, J. M., Barbee, B. D., and Harris, S. T. (1992). Developmental effects of methyl benzimidazole car-

References

bamate following exposure during early pregnancy. Fundam. Appl. Toxicol. 18, 288-293. Cummings A M., Hedge, J. L., and Laskey, J. (1997). Ketoconazole impairs early pregnancy and the decidual cell response via alterations in ovarian function. Fundam. Appl. Toxicol. 40, 238-246. Cutting, R. T., Phuoc, T. H., Ballo, J. M., Benenson, M. W., and Evans, C. H. (1970). "Congenital Malformations, Hydatiform Moles, and Stillbirths in the Republic of Vietnam 1960-1969." U.S. Gov. Printing Office, Washington, DC. Czeizel, A. E., Elek, C., Gundy, S., Metneki, J., Nemes, E., Reis, A, Sperling, K, Timar, L., Tusnady, G., and Viragh, Z. (1993). 341, 539-542. Dalsenter, P. R., Faqi, AS., Webb, J., Merker, H. J., and Chahoud, 1. (1996). Reproductive toxicity and tissue concentrations of lindane in adult male rats. Hum. Exp. Toxicol. 15,406-410. Dan, B. B. (1984). Vietnam and birth defects. J. Am. Med. Assoc. 252, 936-937. Deacon, M. M., Murray, J. S., Piiney, M. K, Rao, K. S., Dittenber, D. A., Hanley, T. R., and John, J. A. (1980). Embryotoxicity and fetotoxicity of orally administered chlorpyrifos in mice. Toxicol. Appl. Pharmacal. 54, 31-40. Dean, B. J., and Blair, D. (1976). Dominant lethal assay in female mice after oral dosing with dichlorvos or exposure to atmospheres containing dichlorvos. Mutat. Res. 40, 67-72. De-Kun, L., Qi-Dong, Z., Xing-Bo, Q., Rong-Min, S., Xi-Lan, Z., He-Jian, C., Cheng-Su, W, JianPing, H., Chao, Q., Shou-Zhen, X., and Xue-Qi, G. (1986). An epidemiological study on the effect of N,N'-methylenebis(2amino-l,3,4-thiadiazole) (MATDA) on outcomes of pregnancy. Teratology 33,289-297. Delatour, P., Debroye, J., Lorgue, G., and Courtot, D. (1977). Experimental embryotoxicity of oxyfenbendazole in the rat and sheep. Reel. Med. Vet. 153, 639-645. Dhondup, P., and Kaliwal, B. B. (1997). Inhibition of ovarian compensatory hypertrophy by the administration of methyl parathion in hemicastrated albino rats. Reprod. Toxicol. 11, 77-84. Dickson, D. (1979). Herbicide claimed responsible for birth defects. Nature 282,220. Dietz et al. (1986). Dixon, R. L., and Lee, 1. P. (1980). Pharmacokinetic and adaptation factors involved in testicular toxicity. Fed. Prac. 39, 66-72. Dobbins, P. K (1967). Organic phosphate insecticides as teratogens in the rat. J. Fla. Med. Assoc. 54,452-456. Doherty, P. A., Ferm, V. H., and Smith, R. P. (1982). Congenital malformations induced by infusion of sodium cyanide in the golden hamster. Toxicol. Appl. Pharmacal. 64, 456-464. Donovan, J. W., Adena, M. A., Rose, G., and Batistutta, D. (1983). "Case Control Study of Congenital Anomalies and Vietnam Service." Aust. Gov. Pub. Serv., Canberra. Donovan, J. W., MacLennan, R., and Adena, M. (1984). Vietnam service and the risk of congenital anomalies: A case-control study. Med. J. Aust. 140, 394-397. Doroshina, M. V. (1983). Acute toxicity, embryotoxicity and teratogenicity of dimezole and benomyl. Tr. Vses. Inst. Gel'mintol. Jm. K. I. Skryabina 26, 41-46. Dougherty, W. J., Coulston, E, and Goldberg, L. (1973). Nonteratogenicity of 2,4,5-trichlorophenoxyacetic acid in monkeys (Macaca mulatta). Toxicol. Appl. Pharmacal. 25, 442. Drake, K D., Adam, G. P., McKenzie, J. J., and Rodweli, D. E (1990). Teratology study of busan 77 in rabbits. Toxicologist 10, 40. Drake, K D., Helmhout, S. L., Bonner, G. L., Adam, G. P., Michlewicz, K. G., and Rodwell, D. E (1989). A teratological evaluation of potassium dimethylthiocarbamate in rats and rabbits. Toxicologist 9, 273. Drudge et a!. (1983). Durham, W E, and Williams, C. H. (1972). Mutagenic, teratogenic, and carcinogenic properties of pesticides. Annu. Rev. Entomol. 17,123-148. Dzierzawski, A., and Minta, M. (1979). Embryotoxic effects of chlorfenvinphos and bromfenvinphos in laboratory animals. Bull. Vet. Inst. Pulawy 23, 3241. Earl, E L., Milier, E., and VanLoon, E. J. (1973). Reproductive, teratogenic, and neonatal effects of some pesticides and related compounds in beagle dogs

413

and miniature swine. In "Pesticides and the Environment: A Continuing Controversy," pp. 253-266. InterAmerican Conference on Toxicology and Occupational Medicine, Miami. Ebert, E., Leist, K. H., and Mayer, D. (1990). Summary of safety evaluation toxicity studies of glufosinate ammonium. Food Chem. Toxicol. 28, 339349. Ecobichon, D. J. (1996). Toxic effects of pesticides. In "Casarett and Doull's Toxicology: The Basic Science of Poisons" (C. Klassen, ed.), 677. Macmillan, New York. Ecobichon, D. J., and MacKenzie, D. O. (1974). The uterotropic activity of commercially and isomerically pure chlorobiphenyls in the rat. Res. Commun. Chem. Pathol. Pharmacal. 9, 85-95. Eibs, H. G., Spielmann, H., and Hagele, M. (1982). Teratogenic effects of cyproterone acetate and medroxyprogesterone treatment during the pre- and postimplantation period of mouse embryos. Teratology 25, 27-36. Eisler, M. (1968). Heptachlor: Toxicology and safety evaluation. Ind. Med. Surg. 37, 840-844. Eisler (1970). Elina, V. A. (1974). Effect of products of organochlorine herbicide production on specific functions of the female body. In "Gigiena Truda Sostoyanie Spetificheskikh Functs. Rabot Nefakbim Khim. Pros.-sti." (R. A Makysheva, ed.). Sverdl. Nauchno-issled. Inst. Okhr. Materin. Mladenchestva Minzdrava, Sverdlovsk, USSR, pp. 187-190. Elkington, J. (1985). "The Poisoned Womb: Human Reproduction in a Polluted World." Penguin, Harmondsworth, UK El-Zalabani, 1. M., Soliman, A A., Osman, A. 1., Wagih, 1. M., and Bassiouni, B. A. (1979). Effect of organophosphorus insecticides on pregnant rabbits. Bull. Alexandria Fac. Med. 15, 113-118. Emerson, J. L., Thompson, D. J., Strebing, R. J., Gerbig, C. G., and Robinson, V. B. (1971). Teratogenic studies on 2,4,5-trichlorophenoxyacetic acid in the rat and rabbit. Food Cosmet. Toxicol. 9, 395-404. Emsley, J. (1994). "The Consumer's Good Chemical Guide: A Jargon-free Guide to the Chemicals of Everyday Life." FreemanJSpektrum, New York. Erickson, J. D., Mulinare, J., McClain, P. W, Fitch, T. G., James, L. M., McClearn, A B., and Adams, M. J. (1984). Vietnam veterans' risks for fathering babies with birth defects. J. Am. Med Assoc. 252, 903-912. European Economic Community (EEC) (1983). Reproduction studies: Notes for guidance concerning the application of Chapter I(C) and (D) of part 2 of the Annex to Directive 75/318fEEC, with a view to granting a marketing authorization for a new drug. J. Eur. Community L332, 20. Everest, L. (1985). "Behind the Poison Cloud: Union Carbide's Bhopal Massacre." Banner Press, Chicago. Fagin, D., Lavelle, M., and the Center for Public Integrity (1996). "Toxic Deception: How the Chemical Industry Manipulates Science, Bends the Law, and Endangers Your Health." Carol Pub. Group, Secaucus, NJ. Ferm, V. H., Hanlon, D. P., Willhite, C. C., Choy, W. N., and Book, S. A (1990). Embryotoxicity and dose-response relationships of selenium in hamsters. Reprad. Toxicol. 4, 183-190. Field, B., and Kerr, C. (1979). Herbicide use and incidence of neural-tube defects. Lancet 1, 1341-1342. Field, B., and Kerr, C. (1988). Reproductive behavior and consistent patterns of abnormality in offspring of Vietnam veterans. J. Med. Genet. 25,819-826. Fish, S. A. (1966). Organophosphorus cholinesterase inhibitors and fetal development. Am. J. Obstet. Gynecol. 96, 1148-1154. Food and Drug Administration (FDA) (1987). Teratogenic potential of purified pentachlorophenol and pentachloroanisole in subchronically exposed Sprague-Dawley rats. Food Chem. Toxicol. 25, 163-172. Francis, B. M. (1986). Teratogenicity of bifenox and nitrofen in rodents. Enviran. Sci. Health, B 21,303-317. Francis, B. M. (1990). Relative teratogenicity of nitrofen analogs in mice: Unchlorinated, monochlorinated, and dichlorinated phenyl ethers. Teratology 41, 443-451. Francis, B. M., and Metcalf, R. L. (1982). Percutaneous teratogenicity of nitrofen. Teratology 25, 41A Francis, E. Z. (1987). "Developmental Neurotoxicology Guideline for the Triethylene Glycol Ether Series." Epidemiology and Surveillance Division, Public Health Service, Rockville, MD.

414

CHAPTER 14


Fraser, F. C. (1977). Relation of animal studies to the problem in man. In "Handbook of Teratology" (1. G. Wilson and F. C. Fraser, eds.), Vol. I, Chap. 3. Plenum, New York. Friedman, J. M. (1984). Does Agent Orange cause birth defects? Teratology 29, 193-221. Friedman, J. M., and Polifka, J. E. (1994). "Teratogenic Effects of Drugs: A Resource for Clinicians (TERIS)," p. 658. Johns Hopkins Press, Baltimore. Frosch, L. (1990). Prenatal toxicology of Wolfatox 80 in rats. Teratology 42, 26A. Funazaki, Z. (1971). Herbicides and deformities in Vietnam. Jpn. J. Public Health Nurse 27, 54-55. Fytizas-Danielidou, R. (1971). Effects of pesticides on the reproduction of white rats. I. Lebaycide. Meded. Fac. Landbouwwet. Rijksuniv. Gent. 36, 1146-1150. Gad, S. C., and Weil, C. S. (1986). Data analysis applications in toxicology. In "Statistics and Experimental Design for Toxicologists" (S. c. Gad and C. S. Weil, eds.), pp. 147-175. Telford, Caldwell, NJ. Gaines, T. B., and Kimbrough, R. D. (1966). The sterilizing, carcinogenic, and teratogenic effects of metepa in rats. Bull. World Health Org. 34, 317-320. Gaines, T. B., and Kimbrough, R. D. (1970). Oral toxicity of mirex in adults and suckling rats. Arch. Environ. Health 21, 7-14. Gale, T. F., and Ferm, V. H. (1973). Effects of the herbicides 2,4,5-T and pyrazon on embryogenesis in the hamster. Anat. Rec. 175,503. Galston, A. W. (1970). Herbicides, no margin of safety. Science 167,237. Garcia-Rodriguez, J., Garcia-Martin, M., Nogueras-Ocana, M., et al. (1996). Exposure to pesticides and cryptorchism: Geographical evidence of a possible association. Environ Health Perspect. 104, 1090-1095. Garry V. F., Schreinnemachers, D., Harkins, M. E., and Griffith, J. (1996). Pesticide applicators, biocides, and birth defects in rural Minnesota. Environ. Health Perspect. 104, 394-399. Gaylor, D. W. (1978). Methods and concepts of biometrics applied to teratology. In "Handbook of Teratology" (1. G. Wilson and F. C. Fraser, eds.), pp. 429-444. Plenum, New York. Gellert, R. J., Heinrichs, W. L., and Swerdloff, R. S. (1972). DDT homologues: Estrogen like effects on the vagina, uterus and pituitary of the rat. Endocrinology 9,1095-1100. Generoso, W. M., Rutledge, J. C., Cain, K. T., Hughes, L. A., and Braden, P. W. (1987). Exposure of female mice to ethylene oxide within hours of mating leads to fetal malformation and death. Mutat. Res. 176,269-274. Generoso, W. M., Rutledge, J. c., Cain, K. T., Hughes, L. A., and Downing, D. J. (1988). Mutagen-induced fetal anomalies and death following treatment of females within hours after mating. Mutat. Res. 199, 175-181. Ghizelea, G., and Czeranschi, L. (1973). Influence of the carbamate pesticides zineb and thiram on the estrous cycle and reproductive capacity of white rats. Environ. Res. 16, 123-130. Giavini, E., Broccia, M. L., Prati, M., and Vismara, C. (1986). Induction of teratogenic effects in the rat fetuses with dinoseb. Teratology 33, 19A. Giavini, E., Prati, M., and Vismara, C. (1980). Effects oftriphenyltin acetate on pregnancy in the rat. Bull. Environ. Contam. Toxicol. 24,936-939. Giavini, E., Prati, M., and Vismara, C. (1982). Rabbit teratology study with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Environ. Res. 27, 74-78. Gibson, J. E. (1973). Teratology studies in mice with 2-sec-butyl-4,6dinitrophenol (dinoseb). Food Cosmet. Toxicol. 11,31-43. Gibson, J. E., and Becker, B. A. (1970). Placental transfer, embryotoxicity, and teratogenicity of thallium sulfate in normal and potassium-deficient rats. Toxico!. App!. Pharmaco!' 16, 120-132. Gibson, J. E., Chen, W. L., Peterson, R. K. (1999). How to determine if an additional 10x safety factor is needed for chemicals: A case study with chlorpyrifos. Toxicol. Sci. 48, 117-122. Ginsburg, J. S., and Barron, W. M. (1994). Pregnancy and prosthetic heart valves. Lancet 344, 1170-1171. Gleiberman, S. E. (1980). Study of remote results of the use of insect repellents. IV. Study of the embryotoxic properties of N -benzolpiperidine and carboxide. Med. Parazitol. 49, 64-67. Gleiberman, S. E., Volkova, A. P., Nikolaev, G. M., and Zhukova, E. v. (1975). Embryotoxic properties of the repellent diethyltoluamide. Farmako!. Toksiko!. 38,202-205.

Gofmekler, V. A., and Tabakova, S. A. (1970). Action of chlorophos on the embryogenesis of rats. Farmakol. Toksiko!. 33,735-737. Goldey, E. S., and Taylor, D. H. (1992). Developmental neurotoxicity following premating maternal exposure to hexachlorobenzene in rats. Neurotoxico!' Terato!' 14, 15-21. Goldman, L. R., Smith, D. F., Neutra, R. R., Saunders, L. D., Pond, E. M., Stratton, J., Wailer, K., Jackson, R. J., and Kizer, K. W. (1985). Pesticide food poisoning from contaminated watermelons in California, 1985. Arch. Environ. Health 45, 229-236. Goldsmith, J. R. (1997). Dibromochloropropane: Epidemiological findings and current questions. Ann. N. Y. Acad. Sci. 26, 300-306. Goncharuk, G. A. (1968). Effect of mercury organic pesticides. Mercuran and mercurohexane on the progeny of white rats. Gig. Sanit. 33, 111-113. Goncharuk, G. A. (1971). Effect of organomercury pesticides on the generative function and offspring of rats. Gig. Sanit. 36, 32-35. Gordon, J. E., and Shy, C. M. (1981). Agricultural chemical use and congenital cleft lip and/or palate. Arch. Environ. Health 36, 213-220. Gough, M. (1986). "Dioxin, Agent Orange: The Facts." Plenum, New York. Goulet, L., and Theriault, G. (1991). Stillbirth and chemical exposure of pregnant workers. Scand. J. Work Environ. Health 17, 25-31 Grabowski, C. T., and Daston, G. P. (1983). Functional teratology of the cardiovascular and other organ systems. In "Issues and Reviews in Teratology" (H. Kalter, ed.), Vol. I, pp. 285-308. Plenum, New York. Gray, L. E., Jr., and Kelce, W. R. (1996). Latent effects of pesticides and toxic substances on sexual differentiation of rodents. Toxico!. Ind. Health 12, 515-531. Gray, L. E., and Ostby, J. S. (1995). In utero 2,3,7,8-tetrachlorodibenzo-pdioxin (TCDD) alters reproductive morphology and function in female rat offspring. Toxico!. App!. Pharmaco!' 133, 285-294. Gray, L. E., Ferrell, J., and Ostby, J. (1985). Prenatal exposure to nitrofen causes anomalous development of para- and mesonephric duct derivatives in the hamster. Toxicologist 5, 183. Gray, L. E., Kavlock, R., Chernoff, N., Lawton, D., and Gray, J. (1979). The effects of endrin administration during gestation on the behavior of the golden hamster. Toxico!. App!. Pharmaco!' 48, A200. Gray, L. E., Kavlock, R. J., Chernoff, N., Ostby, J., and Ferrell, J. (1983). Postnatal developmental alterations following prenatal exposure to the herbicide 2,4-dichlorophenyl-p-nitrophenyl ether: A dose response evaluation in the mouse. Toxico!. App!. Pharmacol. 67, 1-14. Gray, L. E., Jr., Ostby, J., Cooper, R. L., and Kelce, W. R. (1999a). The estrogenic and antiandrogenic pesticide methoxychlor alters the reproductive tract and behavior without affecting pituitary size or LH and prolactin secretion in male rats. Toxico!. Ind. Health IS, 37-47. Gray, L. E., Ostby, J. S., and Kelce, W. R. (1994). Developmental effects of an environmental antiandrogen: The fungicide vinclozin alters sex differentiation of the male rat. Toxicol. App!. Pharmaco!' 125,46-52. Gray, L. E., Ostby, J., Sigmon, R., Ferrell, J., Rehnberg, G., Linder, R., Cooper, R., Goldman, J., and Laskey, J. (1988). The development of a protocol to assess reproductive effects of toxicants in the rat. Reprod. Toxicol. 2,281-287. Gray, L. E., Jr., Rogers, J. M., Kavlock, R. J., Ostby, J. S., Ferrell, J. M., and Gray, K. L. (1986). Prenatal exposure to the fungicide dinocap causes behavioral torticollis, ballooning and cleft palate in mice, but not rats or hamsters. Teratogen. Carcinogen. Mutagen. 6, 33-43. Gray et al. (1996). Gray, L. E., Jr., Wolf, c., Lambright, c., Mann, P., Price, M., Cooper, R. L., and Ostby, J. (1999b). Administration of potentially anti androgenic pesticides (procymidone, linuron, iprodione, chlozolinate, p,p'-DDE, and ketoconazole) and toxic substances (dibutyl- and diethylhexyl phthalate, PCB 169, and ethane imethanesulphonate) during sexual differentiation produces diverse profiles of reproductive malformations in the male rat. Toxico!. Ind. Health 15,94-118. Greenwell, A., Tomaszewski, K. E., and Melnick, R. L. (1987). A biochemical basis for 1,2-dibromo-3-chloropropane (DBCP)-induced male infertility: Inhibition of sperm mitochondrial electron transport activity. Toxico!. App!. Pharmacol. 91,274-280.

References

Grether, J. K., Harris, J. A., Neutra, R., and Kizer, K. W. (1987). Exposure to aerial malathion application and the occurrence of congenital anomalies and low birthweight. Am. J. Public Health 77,1009-1010. Gianessi, L. P., and Puffer, C. (1990). Herbicide use in the United States: National summary report. In "Resources for the Future." Quality of the Environment Division, Washington, DC. Guittin, P., Trouiller, G., and Derrien, J. (1987). Postnatal behavioral toxicity in rats following prenatal exposure to an organophosphate. Teratology 36, 25A. Gupta, A., Agarwal, R., and Shukla, G. S., (1999). Functional impairment of blood-brain barrier following pesticide exposure during early development in rats. Hum. Exp. Toxicol. 18, 174-179. Gupta, P. K., Chandra, S. v., and Saxena, D. K. (1978). Teratogenic and embryotoxic effects of endosulfan in rats. Acta Pharmacol. Toxicol. (Copenhagen) 42, 150--152. Gurunathan, S., Robson, M., Freeman, N., Buckley, B, Roy, A., Meyer, R., Buckowsi, J., and Lioy, P. J. (1998). Accumulation of chlorpyrifos on residential surfaces and toys accessible to children. Environ. Health Perspect. 106,9-16. Gzhegotskii, M. 1., and Shtabskii, B. M. (1968). Chronic poisoning due to the herbicide celatox and its effect on the offspring. Vrach. Delo 11, 121-122. Hackel, C. (1983). Effect of methoprene on the embryonic development ofWistar rats. Rev. Bras. Genet. 6,639-647. Hall, J. G., Pallister, P. D., Clarren, S. K., Beckwith, J. B., Wiglesworth, F. w., Fraser, F. C., Cho, S., Benke, P. J., and Reed, S. D. (1980). Congenital hypothalamic hamartoblastoma, hypopituitarism, imperforate anus, and postaxial polydactyly-A new syndrome? 1. Clinical, causal and pathogenetic considerations. Am. J. Med. Genet. 7,47-74. Hall, P., and Selinger, B. (1981). Australian herbicide usage and congenital abnormalities. Chem. Aust. 48, 131-132. Hammond, B., Katzenellenbogen, B. S., Krauthammer, N., and McConnell, J. (1979). Estrogenic activity of the insecticide chlordecone (Kepone) and interaction with uterine estrogen receptors. Proc. Natl. Acad. Sci. U.S.A. 76, 6641-6645. Hanify, J. A., Metcalf, P., Nobbs, C. L., and Worsley, K. L. (1981). Aerial spraying of 2,4,5-T and human birth malformations: An epidemiological investigation. Science 212, 349-351. Hanley, T. R., Calhoven, L. L., Kociba, R. L., and Greene, J. A. (1989). The effects of inhalation exposure to sulfuryl fluoride on fetal development in rats and rabbits. Fundam. Appl. Toxicol. 13, 79-86. Hanley, T. R., Carney, E. w., and Johnson, E. M. (1998). 3,5,6-Trichloro-2pyridinol: Developmental toxicity studies in rats and rabbits. Toxicologist 42,255. Hanley, T. R., John-Greene, J. A., Hayes, W. c., and Rao, K. S. (1987). Embryotoxicity and fetotoxicity of orally administered tridiphane in mice and rats. Fundam. Appl. Toxicol. 8, 179-187. Hanley, T. R., Thompson, D. J., Palmer, A. K., Beliles, R. P., and Schwetz, B. A. (1984). Teratology and reproduction studies with triclopyr in the rat and rabbit. Fundam. Appl. Toxicol. 4, 872-882. Hardin, B. D., Bond, G. P., Sikov, M. R., Andrew, F. D., Beliles, R. P., and Niemeier, R. W. (1981). Testing of selected workplace chemicals for teratogenic potential. Scand. J. Work Environ. Health 7, 66-75. Harris, S. J., Bond, G. P., and Niemeier, R. W. (1979). The effects of 2-nitropropane, naphthalene, and hexachlorobutadiene on fetal rat development. Toxicol. Appl. Pharmacol. 48, A35. Harrison, W. P., and Hood, R. D. (1981). Prenatal effects following exposure of hamsters to sodium arsenite by oral or intraperitoneal routes. Teratology 23,40A. Harrison, W. P., Frazier, J. c., Mazzanti, E. M., and Hood, R. D. (1980). Teratogenicity of disodium methanarsonate and sodium dimethylarsinate (sodium cacodylate) in mice. Teratology 21, 43A. Hart, M. M., Adamson, R. H., and Fabro, S. (1971). Prematurity and intrauterine growth retardation induced by DDT in the rabbit. Arch. Int. Pharmacodyn. Ther. 192,286-290. Hassoun, E., d' Argy, R., and Dencker, L. (1983). Teratogenicity of TCDD and its congeners. Toxicologist 2, 67.

415

Hassoun, E., d' Argy, R., and Dencker, L. (1984a). Teratogenicity of 2,3,7,8tetrachlorodibenzofuran in the mouse. J. Toxicol. Environ. Health 14, 337352. Hassoun, E., d' Argy, R., Dencker, L., and Wahistrom, B. (l984b). Teratological studies on the TCDD congener 3,3',4,4'-tetrachloroazoxybenzene in sensitive and nonsensitive mouse strains-Evidence for direct effect on embryonic tissues. Arch. Toxicol. 55, 20--26. Hatch, M. C. (1985). Reproductive effects of herbicide exposure in Vietnam. Teratogen. Carcinogen. Mutagen. 5,231-250. Hayes, W. C., Hanley, T. R., Gushow, T. S., Johnson, K. A., and John, J. A. (1985). Teratogenic potential of inhaled dichlorobenzenes in rats and rabbits. Fundam. Appl. Toxicol. 5, 190--202. Hayes, W. c., Smith, F. A., John, J. A., and Rao, K. S. (1984). Teratologic evaluation of 3,6-dichloropicolinic acid in rats and rabbits. Fundam. Appl. Toxicol. 4,91-97. Hayes, W. J., Jr. (1959). The pharmacology and toxicology of DDT. In "The Insecticide DDT and Its Importance" (P. Muller, ed.), Vol. 2, pp. 9-247. Birkhauser, Basel. Heidam, L. Z. (1984). Spontaneous abortions among dental assistants, factory workers, painters, and gardening workers: A follow up study. J. Epidemiol. Community Health 38, 149-155. Heinrichs, W. L., and Juchau, M. R. (1980). Extrahepatic drug metabolism: The 62 gonads. In "Extrahepatic Metabolism of Drugs and Other Foreign Compounds" (T. E. Gram, ed.), pp. 319-332. SP Medical and Scientific Books, New York. Hemminki, K., Mutanen, P., and Niemi, M.-L. (1983). Spontaneous abortion in hospital workers who used chemical sterilizing equipment during pregnancy. Br. Med. J. 286, 1976-1977. Hemminki, K., Mutanen, P., Saloniemi, 1., Niemi, M.-L., and Vainio, H. (1982). Spontaneous abortions in hospital staff engaged in sterilizing instruments with chemical agents. Br. Med. J. 285, 1461-1462. Henwood, S., Mellon, K., and Osimitz, T. (1990). Teratology study with (2-naphthoxy) acetic acid in rats. Toxicologist 10,39. Herbst, M., and Bodenstein, G. (1972). Toxicology of lindane. In "Lindane" (E. Ulman, ed.), pp. 23-78. Verlag, Freiburg. Hertz-Picciotto, 1., Pastore, L. M., and Beaumont, J. J. (1996). Timing and patterns of exposures during pregnancy and their implications for study methods. Am. J. Epidemiol. 15,597-607. Heydens, W. F., Siglin, J. c., Holson, J. F., and Stegeman, S. D. (1996). Subchronic, developmental, and genetic toxicology studies with the ethane sulfonate metabolite of alachlor. Fundam. Appl. Toxico!. 33, 173-181. Hicks, S. P. (1952). Some effects of ionizing radiation and metabolic inhibition on developing mammalian nervous system. J. Pediatr. 40,489-513. Hjelde, T., Mehl, A., Schanke, T. M., and Fonnum, F. (1998). Teratogenic effects of trichlorfon (Metrifonate) ont he guinea-pig brain: Determination of the effective dose and the sensitive period. Neurochem. Int. 32, 469-477. Hoberman, A. M. (1978). Ultrastructural study of liver tissue from mice prenatally exposed to the cholinesterase inhibitor carbofuran. Teratology 17, 41A. Hodge, et al. (1967). Hoffman, D. J., Rattner, B. A., Sileo, L., Docherty, D., and Kubick, T. J. (1987). Embryotoxicity, teratogenicity, and aryl hydrocarbon hydroxylase activity in Forster's terns on Green Bay, Lake Michigan. Environ. Res. 42, 176-184. Hong, J. S., Hudson, P. M., Yoshikawa, K., Ali, S. F., and Mason, G. A. (1985). Effects of chlordecone administration on brain and pituitary peptide systems. Neurotoxicology 6, 167-182. Hood, R. D., and Harrison, W. P. (1982). Effects of prenatal arsenite exposure in the hamster. Bull. Environ. Contam. Toxicol. 29,671-678. Hood, R. D., Harrison, W. P., and Vedel, G. C. (1982). Evaluation of arsenic metabolites for prenatal effects in the hamster. Bull. Environ. Contam. Toxicol. 29,679-687. Huber, J. J. (1965). Some physiological effects of the insecticide kepone in the laboratory mouse. Toxicol. Appl. Pharmacol. 17,516-524. Iatropoulos, M. J., Hobson, w., Knauf, M. V., and Adams, H. P. (1976). Morphological effects of hexachlorobenzene toxicity in female rhesus monkeys. Toxicol. App!. Pharmacol. 37, 433-444.

416

CHAPTER 14


Ihrig, M. M., Sholat, S. L., and Baynes, C. (1998). A hospital-based casecontrol study of stillbirths and environmental exposure to arsenic using an atmosphere dispersion model linked to a geographical information system. Epidemiology 9, 290-294. Ilina, B. I. (1980). Effect of pesticides on development of gynecological diseases in field-crop growers. Pediatr. Akush. Ginekol. 42, 48-50. Infurna, R, Levy, B., Arthur, A., and Traina, V. (1986). Teratological evaluations of atrazine technical, a triazine herbicide, in rats and rabbits. Toxicologist 6, 92. Infurna, R, Yau, E., Traina, B., Wetzel, L., and Stevens, J. (1987). Teratological evaluations of ametryn technical, a triazine herbicide, in rats and rabbits. Toxicologist 7, 174. Ingle, L. (1952). Chronic oral toxicity of chlordane to rats. Arch. Ind. Hyg. Occup. Med. 6,357-367. Ireland, J. S., Mukku, V. R., Robison, A. K, and Stancel, G. M. (1980). Stimulation of uterine deoxyribonucleic acid synthesis by 1,1,I-trichloro-2-(pchlorophenyl)-2-(0-chlorophenyl)ethane (o,p' -DDT). Biochem. Pharmacol. 29, 1469-1474. Iskhakov, A. 1., and Magrupova, N. Kh. (1976). Experimental data on the embryotoxic effect ofrogor. Probl. Gig. Organ. Zdravookhr. Uzb. 5,83-85. Ivashin, V. M., Bandazher, Y. 1., Obozny, N. D., and Zakharchen, R. G. (1989). Development of addiction to carbophos in the offspring of rats at administration during pregnancy. Farmakol. Toksikol. 52, 87-90. Iyer, P., Gammon, D., Gee, J., and Pfeifer, K (1999). Characterization of maternal influence on teratogenicity: An Assessment of developmental toxicity studies for the herbicide cyanazine. Reg. Toxico!. Pharmacol. 29, 88-95. Janardhan, A., Sattur, P. B., and Sisodia, P. (1984). Teratogenicity of methyl benzimidazolecarbamate in rats and rabbits. Bull. Environ. Contam. Toxicol. 33,257-263. Jarrell, J., Gocmen, A, Foster, W., Brant, R, Chan, S., and Sevcik, M. (1998). Evaluation of reproductive outcomes in women inadvertently exposed to hexachlorobenzene in southeastern Turkey in the 1950s. Reprod. Toxico!. 12, 469-476. Jayatunga, Y. N. A, Dangalle, C. D., and Ratnasooriya (1998a). Hazardous effects of carbofuran on pregnancy outcome of rats. Medical Science Research 26, 33-37. Jayatunga, Y. N. A, Dangalle, C. D., and Ratnasooriya (1998b). Effects of midterm exposure to carbofuran on pregnancy outcome of rats. Medical Science Research 26, 679-683. Jewell, W. T, and Miller, M. G. (1998). Identification of a carboxylesterase as the major protein bound by molinate. Toxicol. Appl. Pharmacol. 149, 226234. John-Greene, J. A., Oulette, J. H., Jeffries, T. K, Johnson, K A, and Rao, K. S. (1985). Teratological evaluation of picloram potassium salt in rabbits. Food Chem. Toxicol. 23,753-756. Johnson, E. M. (1981). Screening for teratogenic hazards: Nature of the problems. Annu. Rev. Pharmacol. Toxicol. 21, 417-429. Johnson, E. M., Grabowski, C. J., Jensch, R. P., Juchau, M. R., Kimmel, C. A., Lamb, J., Manson, J. M., Marks, TA., Scott, W. J., Sheehan, D., Staples, R E., and Sullivan, F. M. (1990). Developmental Effects. In "Effects of Pesticides on Human Health" (S. R. Baker and C. F. Wilkinson, eds.), Advances in Modem Environmental Toxicology, Vol. 18, pp. 392-438. Princeton Sci. Pub., Princeton, NJ. Johnston, C. D., Woodard, G., and Cronin, M. T L. (1968). Safety evaluation of Botran (2,6-dichloro-4-nitroaniline) in laboratory animals. Toxicol. App!. Pharmacol. 12,314-315. Kagen, Y. S., Fudel-Ossipova, S. L., Khaikima, B. J., Kuzminskaya, U. A., and Kouton, S. D. (1969). On the problem of the harmful effect of DDT and its mechanism of action. Residue Rev. 27,43-79. Kamata, K (1983). Effect of 1,3,6,8-tetrachlorodibenzo-p-dioxin on the rat fetus. Oyo Yakuri 25,713-718. Kaneda, M., Hojo, H., Teramoto, S. and Maita, K (1998). Oral teratogenicity studies of methyl bromide in rats and rabbits. Food Chem. Toxicol. 36,421427. Kaneko, H., Kawaguchi, S., Misaki, Y., Koyama, y', Nakayama, A., Kawasaki, H., Hirohashi, A, Yoshitake, A., and Yamada, H. (1992). Mam-

malian toxicity of empenthrin (Vaporthrin, S-2852F). J. Toxicol. Sci. 17, 313-334. Kanoh, S., Ema, M., and Hon, Y. (1981). Studies on the toxicity of insecticides and food additives in the pregnant rats. I. Fetal toxicity of o-methyl-o(4-bromo-2,5-dichlorophenyl)phenyl thiophosphonate. Oyo Yakun 22,373380. Kasymova, R A. (1976). Experimental and clinical data on the embryotoxic effect ofbutiphos. Probl. Gig. Organ. Zdravookhr. Uzb. 5, 101-103. Katoh, M., Cacheiro, N. L., Cornett, C. V., Cain, K T, Rutledge, J. c., and Generoso, W. M. (1989). Fetal anomalies produced subsequent to treatment of zygotes with ethylene oxide or ethyl methanesulfonate are not likely due to the usual genetic causes. Mutat. Res. 210,337-344. Kavlock, R, Chernoff, N., Baron, R, Linder, R., Rogers, E., and Carver, B. (1979). Toxicity studies with decamethrin, a synthetic pyrethroid insecticide. J. Environ. Patho!. Toxicol. 2,751-766. Kavlock, R, Chernoff, N., Baron, R., Linder, R, Rogers, E., Carver, B., Dilley, J., and Simmon, V. (1979). Toxicity studies with decamethrin, a synthetic pyrethroid insecticide. J. Environ. Patho!. Toxicol. 2, 751-765. Kavlock, R. J., Chernoff, N., Gray, L. E., Gray, J. A, and Whitehouse, D. (1982). Teratogenic effects of benomyl in the Wistar rat and CD-I mouse, with emphasis on the route of administration. Toxico!. Appl. Pharmaco!' 62, 44--54. Kavlock, R. J., Chernoff, N., and Rogers, E. H. (1985). The effect of acute maternal toxicity on fetal development in the mouse. Teratogen. Carcinogen. Mutagen. 5, 3-13. Kawamura, S., Kato, T, Matsuo, M., Sasaki, M., Katsuda, Y., Hoberman, A M., and Yasuda, M. (1995). Species difference in developmental toxicity of an N -phenylimide herbicide between rats and rabbits and sensitive period of the toxicity to rat embryos. Congenital Anom. 35,123-132. Kay, H. H., and Mattison, D. R (1985). How radiation and chemotherapy affect gonadal function. Contemp. Obstet. Gynecol. 26, 109-127. Kennedy, G. L. (1986). Chronic toxicity, reproductive, and teratogenic studies with oxamyl. Fundam. Appl. Toxico!. 7, 106-118. Kennedy, G. L., and Kaplan, A M. (1984). Chronic toxicity, reproductive, and teratogenic studies of hexazinone. Fundam. Appl. Toxicol. 4, 960-971. Kennedy, G. L., Fancher, O. E., and Calandra, J. C. (1975). Nonteratogenicity of captan in beagle. Teratology 11, 223-226. Khadzhitodorova, E., and Andreev, A (1984). Effect of the organophosphate insecticide chloracetophon on skeletal ossification and the development of internal organs in white rat fetuses. Eksp. Med. Moffo!. 23,201-205. Khamidov, D. Kh., Vdovina, S. K., Sagatova, G. A, Muchnik, S. E., and Mirakhmedov, A. K. (1986). The transfer offluometuron (cotoran) through the placenta in the late stages of pregnancy. Uzb. Bio!. Zh. 2, 57-58. Khera, K S. (1973). Ethylenethiourea: Teratogenicity study in rats and rabbits. Teratology 7, 243-252. Khera, K S. (1979). Teratogenicity evaluation of commercial formulation of dimethoate (Cygon 4F) in the cat and rat. Toxicol. App!. Pharmacol. 48, A34. Khera, K. S. (1987). Ethylenethiourea: A review of teratogenicity and distribution studies and an assessment of reproduction risk. Crit. Rev. Toxico!. 18, 129-141. Khera, K S., and Iverson, F. (1978). Toxicity of ethylenethiourea in the pregnant cat following oral administration at low dosages. Toxicol. Appl. Pharmacol. 45,290-291. Khera, K S., and Whitta, L. L. (1968). Embryopathic effects of diquat and paraquat in the rat. Ind. Med. Surg. 37, 553. Khera, K S., Huston, B. L., and McKinley, W. P. (1971). Pre- and postnatal studies on 2,4,5-T, 2,4-D, and derivatives in Wistar rats. Toxicol. Appl. Pharmacol. 19,369-370. Khera, K S., Whalen, c., and Angers, G. (1981). Teratogenicity study on pyrethrins and rotenone (of natural origin) and ronnel in pregnant rats. Teratology 23, 45A-46A. Khera, K S., Whalen, C., and Trivett, G. (1978). Teratogenicity studies on linuron, malathion, and methoxychlor in rats. Toxicol. Appl. Pharmacol. 45, 435-444. Khera, K. S., Whalen, c., Trivett, G., and Angers, G. (1979a). Assessment of the teratogenic potential of biphenyl, ethoxyquin, piperonyl butoxide,

References

diuron, thiabendazole, phosalone, and lindane in rats. Toxieol. Appl. Pharmacal. 48, A33. Khera, K. S., Whalen, C., Trivett, G., and Angers, G. (1979b). Teratological assessment of maleic hydrazide and daminozide, and formulations of ethoxyquin, thiabendazole and naled in rats. J. Environ. Sci. Health, B 14, 563-577. Khera, K. S., Whalen, C., Trivett, G., and Angers, G. (1979c). Teratogenicity studies on pesticidal formulations of dimethoate, diuron and lindane in rats. Bull. Environ. Contam. Toxicol. 22, 522-529. Kimmel, C. A., Generoso, W. M., Thomas, R. D., and Bakshi, K. S. (1993). A new frontier in understanding the mechanisms of developmental abnormalities. Toxicol. Appl. Pharmacal. 119, 159-165. Kimmel, C. A., LaBorde, J. B., Jones-Price, C., Ledoux, T. A., and Marks, T. A. (1982). Fetal development in New Zealand white (NZW) rabbits treated iv with ethylene oxide during pregnancy. Toxicologist 2, 70. King, C. T. G., Horigan, E. A., and W?lk, A. L. (1971). Screening of the herbicides 2,4,5-T and 2,4-D for cleft palate production. Teratology 4, 233. King, S. (1981). Peropal-A new organotin miticide. Tin Its Uses 128, 12-14. Kitselman, C. H., Danm, P. A., and Borgmann, A. R. (1950). Toxicologic studies of aldrin (compound 118) on large animals. Am. J. Vet. Res. 11,378-381. Klotz, D. M., Arnold, S. F., and McLachlan, J. A. (1997). Inhibition of 17 betaestradiol and progesterone activity in human breast and endometrial cancer cells by carbamate insecticides. Life Sci. 60, 1467-1475. Kluth, D., Kangah, R., Reich, P., Tenbrinck, R., Tibboel, D., and Lambrecht, W. (1990). Nitrofen-induced diaphragmatic hernia in rats-An animal model. J. Pediatr. Surg. 25, 850--854. Kluwe, W. M. (1981). Acute toxicity of 1,2-dibromo-3-chloropropane in the F344 male rat. 11. Development and repair of the renal, epididymal, testicular and hepatic lesions. Toxicol. Appl. Pharmacal. 59, 84-95. Kluwe, W. M., Lamb, J. c., IV, Greenwell, A., and Harrington F. W. (1983). 1,2-dibromo-3-chloropropane (DBCP)-induced infertility in male rats mediated by post-testicular effect. Toxicol. Appl. Pharmacal. 71, 294-298. Klys, M., Kosun, J., Pach, J., and Kamenczak, A. (1989). Carbofuran poisoning of pregnant woman and fetus per ingestion. 1. Forensic Sci. 34, 1413-1416. Kolb-Meyers, V. (ed.) (1988). "Teratogens: Chemicals Which Cause Birth Defects." Elsevier, New York. Konje, J. c., Otolorin, E. 0., Sotunmbi, P. T., and Ladipo, O. A. (1992). Insecticide poisoning in pregnancy: A case report. J. Reprod. Med. 37, 992-994. Korshunova, E. P. (1988). Hygienic evaluation of embryotoxic effects of carbophos and formaldehyde mixture separated and combined with elevated temperature and air humidity. Gig. Sanit. 10, 13-15. Kronevi, T., and Backstrom, L. (1977). Kongenital tremor (Skaksjuka) hos grist. Sartryck Svensk Veterinartidning 21, 837-841. Krowke, R., Franz, G., and Neubert, D. (1989). Embryotoxicity: Is the TCDDinduced embryotoxicity in rats due to maternal toxicity? Chemosphere 18, 291-298. Kubena, L. F. (1982). The influence of diflubenzuron on several reproductive characteristics in male and female-layer-breed chickens. Poult. Sci. 61, 268271. Kunstater, P. (1982). "A Study of Herbicides and Birth Defects in the Republic of Vietnam: An Analysis of Hospital Records." Natl. Acad. Sci.-Natl. Acad. Press, Washington, DC. Kupfer, D., and Bulger, W. H. (1976). Studies on the mechanisms of estrogenic actions of o,p' -DDT: Interactions with the estrogen receptor. Pestic. Biochem. Physiol. 6, 461-470. Kvitnitskaya, V. A., and Kolesnichenko, T. S. (1971). Transplacental blastomogenic action of zineb on the progeny of mice. Vopr. Pitan. 30, 49-50. Lackmann, G. M., Angerer, J., Salzberger, U., and Tollner, U. (1999). Influence of maternal age and duration of pregnancy on serum concentrations of polychlorinated biphenyls and hexachlorobenzene in full-term neonates. BioI. Neonate 76, 214-219. La Clair, J., Bantle, J. A., and Dumont, J. (1998). Photoproducts and metabolites of a common insect growth regulator produce developmental deformities in Xenopus. Environ. Sci. Toxicol.32, 1453-1461. Lamartiniere, C. A., Moore, J., Holland, M., and Bames, S. (1995). Neonatal genistein chemo prevents mammary cancer. Proc. Soc. Exp. BioI. Med. 208, 120--123.

417

Laporte, J. R. (1977). Effects of dioxin exposure. Lancet 1, 1049-1050. Lappe, M. (1991). "Chemical Deception: The Toxic Threat to Health and the Environment." Sierra Club Books, San Francisco. Lathrop, G. D., Wolfe, W. H., Albanese, R. A., and Maynahan, P. M. (1984). "Project Ranch Hand 11. An Epidemiological Investigation of Health Effects in Air Force Personnel Following Exposure to Herbicides." Aerospace Medical Division, San Antonio. LaVecchio, F. A., Pashayan, H. M., and Singer, W. (1983). Agent Orange and birth defects. N. Engl. J. Med. 308,719-720. Laws, S. c., Carey, S. A., Hart, D. w., and Cooper, R. L. (1994). Lindane does not alter the estrogen receptor or the estrogen-dependent induction of progesterone receptors in sexually immature or ovariectomized adult rats. Toxicology 92, 127-142. Lehotzky, K., Szeberenyi, M. J., and Kiss, A. (1989). Behavioral consequences of prenatal exposure to the organophosphate insecticide Sumithion. Neurotoxicol. Teratol. 11,321-324. Leland, et al. (1992). Leland, T. M., Mendelson, G. F., Steinberg, M., and Weeks, M. (1972). Studies on the prenatal toxicity and teratogenicity of N, N -diethyl benzene sulfonamide in rats. Toxieol. Appl. Pharmacal. 23, 376-384. Le Marchand, L., Kolonel, L. N., Siegel, B. Z., and Dendle, W. H. (1986). Trends in birth defects for a Hawaiian population exposed to heptachlor and for the United States. Environ. Health 41, 145-148. Lewis, R. G., Fortmann, R. C., and Camann, D. E. (1994). Evaluation of methods for monitoring the potential exposure of small children to pesticides in the residential environment. Arch. Environ. Contam. Toxicol. 26, 37-46. Li, X., Johnson, D. c., and Rozman, K. K. (1995). Reproductive effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in female rats: Ovulation, hormonal regulation, and possible mechanisms. Toxicol. Appl. Pharmacal. 133,327. Liberacki, A. B., Zablotny, C. L., Yano, B. L., and Breslin, W. J. (1994). Developmental toxicity studies on a series of 2,4-D salts and esters in rabbits. Toxicologist 14, 162. Lin, S., Marshall, E. G., and Davidson, G. K. (1994). Potential parental exposure to pesticides and limb reduction defects. Scand. J. Work Environ. Health 20, 166-179. Linder, R. E., Strader, L. F., Slott, V. L., and Suarez, J. D. (1992). Endpoints of spermatotoxicity in the rat after short duration exposures to fourteen reproductive toxicants. Reprod. Toxicol. 6,491-505. Lindhout, D., and Hageman, G. (1987). Amyoplasia congenita like condition and maternal malathion exposure. Teratology 36, 7-9. Linedecker, C. (1982). "Kerry: Agent Orange and an American Family." St. Martin's, New York. Lingh, w., Younes, M., Pfeil, R., and Solecki, R. (1992). Scientific basis for risk assessment (reproductive toxicity) for pesticides as practiced by the Bundesgesundheitsamt (BGA). In "Risk Assessment of Prenatally-Induced Adverse Health Effects" (D. Neubert, R. J. Kavlock, H.-J. Merker, and J. Klein, eds.), pp. 127-139. Springer-Verlag, Berlin. Lipson, A. (1983). Herbicides and teratogenesis. Med. J. Aust. 2,367-368. Lipson, A., and Gaffey, W. R. (1983). Agent Orange and birth defects. N. Engl. J. Med. 309,491-492. Llorens, J., Crofton, K. M., Tilson, H. A., Ali, S. F., and Mundy, W. R. (1993). Characterization of disulfoton-induced behavioral and neurochemical effects following repeated exposure. Fundam. Appl. Toxicol. 20, 163-169. Lovre, S. c., McCreesh, A. H., and Weeks, M. H. (1977). Safety evaluation of insect repellent cyclohexamethylene carbamide. Toxico!. Appl. Pharmacal. 41,132. Lowry, R. B., and Alien, A. B. (1977). Herbicides and spina bifida. Can. Med. Assoc. J. 117,580. Lu, c., and Fenske, R. A. (1999). Dermal transfer of chlorpyrifos residues from residential surfaces: Comparison of hand press, hand drag, wipe, and polyurethane foam roller measurements after broadcast and aerosol pesticide applications. Environ. Health Perspect. 107,463-467. Lu, C. c., Tang, B. S., and Chai, E. Y. (1982). Teratogenicity evaluations of technical Bladex in Fischer-344 rats. Teratology 25, 59A-60A.

418

CHAPTER 14


Lu, M.-H., and Kennedy, G. L. (1986). Teratogenic evaluation ofmancozeb in the rat following inhalation exposure. Toxicol. Appl. Pharmacol. 84, 355368. Lu, M. H., Filler, R., Bates, H. K., LaBorde, J. B., Bazare, J., Gaylor, D. w., and Kimmel, C. A. (1984). Teratogenicity evaluation of Sarin in rats. Teratology 29, 45A. Lugo, G., Cassady, G., and Palmisano, P. (1969). Acute maternal arsenic intoxication with neonatal death. Am. 1. Disabled Child 117, 328-330. Machemer, L., Schmidt, U., and Holzum, B. (1992). Specific and non-specific developmental effects. In "Risk Assessment of Prenatally-Induced Adverse Health Effects" (D. Neubert, R. J. Kavlock, H.-J. Merker, and J. Klein, eds.), pp. 85-100. Springer-Verlag, Berlin. Maci, R., and Arias, E. (1987). Teratogenic effects of the fungicide maneb on chick embryos. Ecotoxicol. Environ. Saf 13, 169-173. Maitra, S. K., and Sarkar, R. (1996). Influence of methyl parathion on gametogenic and acetylcholinesterase activity in the testis of whitethroated munia (Lonchura malabarica). Arch. Environ. Contam. Toxicol. 30, 384--389. Makita, T., Hashimoto, Y., and Noguchi, T. (1973). Mutagenic, cytogenetic and teratogenic studies on thiophanate-methyl. Toxicol. Appl. Pharmacol. 24, 206--215. MakIetsova, N. Y. (1979). Characteristics of the course of pregnancy, childbirth, and the period after birth in female workers in contact with the pesticide zineb. Pediatr. Akush. Ginekol. 41, 45-46. MakIetsova, N. Y., and Lanovoi, 1. D. (1981). Status of gynecological morbidity of women with occupational contact with the pesticide zineb. Pediatr. Akush. Ginekol. 43, 60--61. Manger, B. R. (1991). In "Veterinary and Applied Pharmacology and Therapeutics" (G. C. Brander, D. M. Pugh, R. J. Bywater, and W. L. Jenkins, eds.), p. 531. Bailliere Tindall, London. Manson, J. M., Brown, T. J., and Baldwin, D. M. (1984). Teratogenicity of nitrofen (2,4-dichloro-4' -nitrodiphenyl ether) and thyroid function in the rat. Toxicologist 4, 166. Marinova, G., Osmankova, D., Dermendzhieva, L., Khadzhikolev, L., Chakurova, 0., and Kaneva, Y. (1973). Professional injuries: Pesticides and their effects on the reproductive functions of women working with pesticides. Pediatr. Akush. Ginekol. 12, 138-140. Martson, L. v., and Shepelskaya, N. R. (1980). Reproductive function in animals to polychlorocamphene. Gig. Sanit. 45, 14-16. Martson, L.v., and Voronina, V. M. (1976). Experimental study of the effect of a series of phosphororganic pesticides (Dipterex and Imidan) on embryogenesis. Environ. Health Perspect. 13, 121-124. Mason, R. R., and Shulte, G. J. (1980). Estrogen-like effects of o,p'-DDTon the progesterone receptor of rat uterine cytosol. Res. Commun. Chem. Pathol. Pharmacol. 29, 281-290. Mast, T. J., Bracco, C. A., Rowland, J. R., and Hendrickx, A. G. (1985). Oftanol exposure during organogenesis in rat: Serum cholinesterase depression and pregnancy outcome. Teratology 31, 47 A. Mathew, G., Vijayalaxmi, K. K., and Abdul Rahiman, M. (1992) Methyl parathion-induced sperm shape abnormalities in mouse. Mutat. Res. 280, 169-173. Matsumura, F., and Ghiasuddin, S. M. (1979). Characteristics ofDDT-sensitive Ca-ATPase in the axonic membrane. In "Neurotoxicology of Insecticides and Pheromones" (T. Narahashi, ed.), pp. 245-257. Plenum, New York. Mattison, D. R., and Thorgeirsson, S. S. (1978). Gonadal aryl hydrocarbon hydroxylase in rats and mice. Cancer Res. 39, 1368-1373. Mattison, D. R., and Thorgeirsson, S. S. (1979). Ovarian aryl hydrocarbon hydroxylase activity and primordial oocyte toxicity of polyclic aromatic hydrocarbon in mice. Cancer Res. 39, 3471-3475. Mattison, D. R., Bogumil, R. J., Chapin, R., Hatch, M. Hendrickx, A., JarreIl, J., LaBarbera, A. L., Schrader, S. M., and Selevan, S. (1990). Reproductive toxicity of pesticides. In "Effects of Pesticides on Human Health" (R. S. Baker and C. F. WiIkinson, eds.). Advances in Modem Environmental Toxicology, Vol. 18, pp. 299, 313, 338. Princeton Sci. Pub., Princeton, NJ. Mattison, D. R., Nightingale, M. S., and Shiromizu, K. (1983). Effects of toxic substances on female reproduction. Environ. Health Perspect. 48, 43-52.

McBride, W. G., and Machin, M. (1989). Placental transfer and teratogenic potential of malathion in the rabbit. Teratology 40, 260. McCormack, K. M., Abuelgasim, A., Sanger, V. L., and Hook, J. B. (1980). Postnatal morphology and functional capacity of the kidney following prenatal treatment with dinoseb in rats. 1. Toxicol. Environ. Health 6, 633-643. McLachlan, J. A., and Dixon, R. L. (1972). Gonadal function in mice exposed prenatally to p, p'-DDT. Toxicol. Appl. Pharmacol. 22, 327. McLaughlin, J., Reynolds, E. F., Lamar, J. K., and Marliac, J. P. (1969). Teratology studies in rabbits with captan, folpet, and thalidomide. Toxicol. Appl. Pharmacal. 14,641. McNulty, W. P. (1984). Fetotoxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) to rhesus monkeys (Macaca mulaua). Am. 1. Primatol. 641-47. McParland, P. J., and McCracker, R. M. (1973). Benzene hexachloride poisoning in cattle. Vet. Rec. 93, 369-371. Merkle, J., Schulz, v., and Gelbke, H. P. (1984). An embryotoxicity study of the fungicide tridemorph and its commercial formulation calixin. Teratology 29,259-269. Mes, J. D., Davies, D. J., and Thurton, D. (1982). Polychlorinated biphenyl and other chlorinated hydrocarbon residues in adipose tissue of Canadians. Bull. Environ. Contam. Toxicol. 28, 97-104. Mestitizova, M. (1967). On reproduction studies and the occurrence of cataracts in rats after long-term feeding of the insecticide heptachlor. Experientia 23, 42-43. Minor, J. L., Short, R. D., Van Goethem, D. L., Wong, L. C. K., and Dacre, J. C. (1982). Mutagenic and reproductive studies of hexahydro-I,3,5-trinitro1,3,5-triazine (RDX) in rats and rabbits. Toxicologist 2, 34--35. Minta, M., and Biernacki, B. (1981). Embryotoxicity and teratogenicity of ethephon, a chemical regulator of biological processes in plants. Med. Weter. 37, 153-156. Mirkhamidova, P., Mirakhmedov, A. K., Sagatova, G. A., Isakova, A. v., and Khamidov, D. K. (1981). Effect of butiphos on the structure and function of the liver in rabbit embryos. Uzb. BioI. Zh. 5,45-47. Mirkova, E. (1980). Embryotoxic and teratogenic effect of the herbicide balagrin. Khig. Zdraveopaz. 23,214--219. Mirkova, E., and Ivanov, 1. (1981). Embryotoxic effect of triazine herbicide polyzin 50. Probl. Khig. 6, 36-43. Miyamoto, J. (1976). Degradation, metabolism and toxicity of synthetic pyrethroids. Environ. Health Perspect. 14, 15-28. Mohammad, F. K., and St. Omer, V. E. V. (1988). Behavioral and neurochemical alterations in rats prenatally exposed to 2,4-dichlorophenoxyacetate (2,4-D) and 2,4,5-trichlorophenoxyacetate (2,4,5-T) mixture. Teratology 37, 515. Moore, J. A., and Courtney, K. D. (1971). Teratology studies with the trichlorophenoxyacid herbicides, 2,4,5-T and silvex. Teratology 4, 236. Morato, G. S., Lemos, T., and Takahashi, R. N. (1989). Acute exposure to maneb alters some behavioral functions in the mouse. Neurotoxicol. Teratol. 11,421-425. Morgan, D. W. (1982). Toxicity study of oxyfendazole in pregnant sows. Vet. Rec. 111, 161-163. Morita, S., Yamada, A., Ohgaki, S., Noda, T., and Taniguchi, S. (1981). Safety evaluation of chemicals for use in household products. n. Teratological studies on benzyl benzoate and 2-(morpholinothio)benzothiazole in rats. Annu. Rep. Osaka City Inst. Public Health Environ. Sci. 43,90-97. Mott, L., and Snyder, K. (1987). "Pesticide Alert: A Guide to Pesticides in Fruits and Vegetables." Sierra Club Books, San Francisco. Munley, S. M., and Hurtt, M. E. (1996). Developmental toxicity study of benomy I in rabbits. Toxicologist 30, 192. Murphy, S. D. (1975). Pesticides. In "Toxicology: The Basic Science of Poisons" (L. J. Casarett and J. Doull, eds.), pp. 408-453. MacmiIIan, New York. Muscarella, D., Dennett, D., Babich, J. G., and Noden, D. (1982). The effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on fetal development in the ferret. Toxicologist 2, 73. Nafstad, 1., Berge, G., Sannes, E., and Lyngset, A. (1983). Teratogenic effects of the organophosphorus compound fenchlorphos in rabbits. Acta Vet. Scand. 24, 295-304.

References

Nagymajtenyi, L., Selypes, A., and Berencsi, G. (1985). Chromosomal aberrations and fetotoxic effects of atmospheric arsenic exposure in mice. J. Appl. Toxicol. 5,61-63. Naishtein S. 1., and Leibovich, D. L. (1971). Effect of small doses of DDT, gamma-hexachlorocyclohexane and their mixtures on sexual function and embryogenesis in rats. Gig. San it. 36, 19-22. Nakanishi, M., Hamada, Y., and Izaki, K. (1970). Insecticides. G. Toxicological studies on a new pyrethroid, kikuthrin. Bochakogaku 35, 113-116. Nakao, Y., and Ueki, R. (1987). Congenital diaphragmatic hernia induced by nitrofen in mice and rats: Characteristics as animal model and pathogenetic relationship between diaphragmatic hernia and lung hypoplasia. Congenital Anom. 27, 397-417. Nakao, Y., Ueki, R., Tada, T., Iritani, L., and Kishimoto, H. (1981). Experimental animal model of congenital diaphragmatic hernia induced chemically. Teratology 24, 11A. Nakaura, S., Tanaka, S., Kawashima, K., Takanaka, A., and Omori, Y. (1984). Effects of zinc diethyidithiocarbamate on prenatal and postnatal developments in rats. Bull. Natl. Inst. Hyg. Sci. (Tokyo) 102, 55-61. Nakazawa, T. (1974). Chronic organophosphorous intoxication in women. J. Jpn. Assoc. Rural Med. 22, 756-758. National Academy of Sciences (NAS). (1993). "Pesticides in the Diets of Infants and Children." Natl. Acad. Sci., Washington, DC. Neeper-Bradley, T. L., Fisher, L. C., Butler, B. L., and Ballantyne, B. (1991). Developmental toxicity evaluation of2-ethyl-I,3-hexanedioI (EHD) administered cutaneously to CD-(Sprague-Dawley) rats. Toxicologist 11,341. Nehaz, M., Paldy, A., Selypes, A., Scheufler, H., Berencsi, G., and Freye, H. A. (198 I). The teratogenic and mutagenic effects of dinitro-o-cresolcontaining herbicide on the laboratory mouse. Ecotoxicol. Environ. Sa! 5, 38-44. Nehez, M., Huszta, E., Mazzag, H., Scheufler, G., Fischer, G. W., and Desi, L. (1986). Cytogenetic and embryotoxic effects of bromophos and demethylbromophos. Regul. Toxicol. Pharmacol. 6,416. Nelson, C. J., Holson, J. F., Green H. G., and Gaylor, D. W. (1979). Retrospective study of the relationship between agricultural use of 2,4,5-T and cleft palate occurrence in Arkansas. Teratology 19,377-384. Nelson, J. A. (1974). Effects of dichlorodiphenyltrichloroethane (DDT) analogs and polychlorinated biphenyl (PCB) mixtures on 17-beta-(3H)estradiol binding to rat uterine receptor. Biochem. Pharmacol. 15,447-451. Neubert, D., Biankenburg, G., Chahoud, 1., Franz, G., Herken, R., Kastner, M., Klug, S., Kroger, J., Krowke, R., Lewandowski, C., Merker, H.-J., Schulz, T., and Stahimann, R. (1986). Results of in vivo and in vitro studies for assessing prenatal toxicity. Environ. Health Perspect. 70, 89-103. Neubert, D., Krowke, R., Chahoud, 1., and Franz, G. (1987). Studies on the reproductive toxicity of 2,3,7,8-TCDD in rodents and non-human primates. Teratology 35, 66A. Nikitina, Y. 1. (1974). Course of labor and puerperium in the vineyard workers and milkmaids in Crimea. Gig. Tr. Pro! Zabol. 18, 17-20. Noda, K., Numata, H., Hirabayashi, M., and Endo, L. (1972). Influence of pesticides on embryos. On influence of organophosphoric pesticides. Oyo Yakuri 6,667-672. Nora, J. J., Nora, A. H., Sommerville, R. J., Hill, R. M., and McNamara, D. G. (1967). Maternal exposure to potential teratogens. J. Am. Med. Assoc. 202, 1065-1069. Norris, R. (ed.) (1982). "Pills, Pesticides and Profits: The International Trade in Toxic Substances." North River Press, Croton-on-Hudson, NY. Nowak, w., Lotocki, w., Stasiewicz, A., and Badurski, J. (1971). Dieldrin poisoning during pregnancy. Pol. Tyg. Lek. 26,958-959. Nurminen, T. (1995). Maternal pesticide exposure and pregnancy. J. Occup. Environ. Med. 37, 935-940. O'Donoghue, J. L. (1997). EPA's neurotoxicity risk assessment guidelines. Fundam. Appl. Toxicol.40, 175-184. Ogata, A., Ando, H., Kubo, Y., Sasaki, M., and Suzuki, K. (1993). Teratogenicity ofpiperonyl butoxide in ICR mice. Teratology 48, 529. Ogi, D., and Hamada, A. (1965). Case reports on fetal deaths and malformations of extremities probably related to insecticide poisoning. 1. Jpn. Obstet. Gynecol. Soc. 17,569.

419

O'Leary, J. A., Davis, J. E., and Feldman, M. (1970). Spontaneous abortion and human pesticide residues of DDT and DDE. Am. J. Obstet. Gynecol. 108, 1291-1292. OIler, W. L., Cairns, T., Bowman, M. C., and Fishbein, L. (1980). A toxicological risk assessment procedure-A proposal for a surveillance index for hazardous chemicals. Arch. Environ. Contam. Toxicol. 9, 483-490. Olson, J. R., McGarrigle, B. P., Tonucci, D. A., Schlecter, A., and Eichelberger, H. (1990). Developmental toxicity of 2,3,7,8-TCDD in the rat and hamster. Chemosphere 20, 1117-1123. Onnis, A., and Grella, P. (1984). "The Biochemical Effects of Drugs in Pregnancy." Vols 1 and 2. Halsted, New York. Orlova, N. v., and Zhalbe, E. P. (1968). Experimental data concerning the problem of permissible amounts of sevin in foodstuffs. Vopr. Pitan. 27, 49-55. Ostby, J. S., Gray, L. E., Kavlock, R. J., and Ferrell, J. M. (1985). The postnatal effects of prenatal exposure to low doses of nitrogen (2,4-dichlorophenylp-nitrophenyl ether) in Sprague-Dawley rats. Toxicology 34, 285-297. Ouellet, M., Bonin, J., Rodrigue, J., Degranges, J. L., and Lair, S. J., (1997). Hindlimb deformities (ectromelia, ectrodactyly) in free-living anurans from agriculatural habitats. J. Wildlife Diseases 33, 95-104. Palermo-Neto, J., FIorio, J. C., and Sakate, M. (1994). Developmental and behavioral effects of prenatal amitraz in rats. Neurotoxicol. Terato!' 16,65-70. Palermo-Neto, J., Sakate, M., and FIorio, J. C. (1997). Developmental and behavioral effects of postnatal amitraz exposure in rats. Braz. J. Med. BioI. Res. 8, 989-997. Palmer, A. K., Bottomley, A. M., Worden, A. N., et al. (1978). Effect of lindane on pregnancy in the rabbit and rat. Toxicology 9, 239-247. Pan, X., Wang, J., and Wu, Z. (1993). A prospective study on the relationship between environmental exposure to pesticides and adverse pregnancy outcomes. Chin. Environ. Sci. 4,91-93. Pant, N., Prasad, A. K., Srivastava, S. c., Shankar, R., and Srivastava, S. P. (1995). Effect of oral administration of carbofuran on male reproductive system of rat. Hum. Exp. Toxico!. 14, 889-894. Pant, N., Shankar, R., and Srivastava, S. P. (1997). In utero and lactational exposure of carbofuran to rats: Effect on testes and sperm. Hum. Exp. Toxicol. 16, 267-272. Pastore, L., Hertz-Piccioto, 1., and Beaumont, J. (1995). A case-control study of still births in relation to residential and occupational exposures. Am. J. Epidemio!. 141, S73. Patro, N., Mishra, S. K., Chattopadhyay, M., and Patro, 1. K., (1997). Neurotoxicological effects of deltamethrin on the postnatal development of cerebellum of rat. J. Biosci. 22, 117-130. Peakall, D. B. (1970). Pesticides and reproduction in birds. Sci. Am. 222, 72-78. Peam, J. H. (1985). Herbicides and congenital malformations: A review for the paediatrician. Aust. Paediatr. J. 21, 237-242. Pedersen, R. A., Meneses, J., Spindle, A., Wu, K., and Galloway, S. M. (1985). Cytochrome P-450 metabolic activity in embryonic and extraembryonic tissue lineages of mouse embryos. Proc. Natl. Acad. Sci. 82, 3311-3315. Pestic. Toxic Chem. News (1989). Peters, H. A. (1976). Hexachlorobenzene poisoning in Turkey. Fed. Proc. 35, 2400-2403. Petrova-Vergieva, T., and Ivanova-Chemishanska, L. (1971). Teratogenic effect of zinc ethylenebis(dithiocarbamate) (zineb) in rats. Eksp. Med. Moiol. 10, 226-230. Phillips-Howard, P. A., and Wood, D. (1996). The safety of antimalarial drugs in pregnancy. Drug Sa! 14, 131-145. Pim, L. R. (1981). "The Invisible Additives: Environmental Contaminants in Our Food." Doubleday, Garden City, NY. Pimental, D., and Levitan, L. (1986). Pesticides: Amounts applied and amounts reaching pests. BioScience 36, 86-91. Porter, S. D., and Wiemeyer, R. D. (1969). Dieldrin and DDT: Effects on sparrow hawk eggshells and reproduction. Science 165, 199-200. Potashnik, G., and Abelovich, D. (1985). Chromosomal analysis and health status of children conceived to men during or following dibromochloropropane-induced spermatogenic suppression. Andrologia 17, 291296.

420

CHAPTER 14


Preston, R J" Fennell, T. R., Leber, A. P., Sielken, R. L. Jr., and Swenberg, J. A. (1995). Reconsideration of the genetic risk assessment for ethylene oxide exposures. Environ. Mo!. Mutagen. 26, 189-202. Rao et a!. (1986). Rattner, B. A., Sileo, L., and Scanes, C. G. (1982). Hormonal responses and tolerance to cold of female quail Colinus-virginianus following parathion ingestion. Pestie. Bioehem. Physio!. 18, 132-138. Regenstein, L. (1982). "America the Poisoned." Acropolis Books, Washington, DC. Regulation 2092/911EEC. http://europa.eu.intlcommlsg/consolid/enl391r2092/ artf.htm. Restrepo, M., Minoz, N., Day, N. E., Parra, J. E., Deromero, L., and Xuan, N. D. (l990a). Prevalence of adverse reproductive outcomes in a population occupationally exposed in Colombia. Scand. J. Work Environ. Health 16, 232-238. Restrepo, M., Munoz, N., Day, N., Parra, J. E., Hernandez, c., Blettner, M., and Giraldo, A. (1990b). Birth defects among children born to a population occupationally exposed to pesticides in Colombia. Seand. J. Work Environ. Health 16, 239-246. Risher, J. E, Mink, E L., and Stara, J. E (1987). The toxicologic effects of the carbamate insecticide aldicarb in mammals: A review. Environ. Health Perspeet. 72, 267-281. Rita, P., Reddy, P. P., and Reddy, S. v. (1987). Monitoring of workers occupationally exposed to pesticides in grape gardens of Andhra Pradesh. Environ. Res. 44, 1-5. Rivera, S., Sanfelieu, c., Sunol, C., et a!. (1991). Regional effects on the cerebral concentration of noradrenaline, serotonin and dopamine in suckling rats after a single dose of lindane. Toxicology 69, 43-54. Roan, C. c., Matanoski, G. E., Mcilnay, C. Q., Olds, K. 1., Pylant, E, Trout, J. R, and Wheeler, P. (1984). Spontaneous abortions, stillbirths, and birth defects in families of agricultural pilots. Arch. Environ. Health 39, 56-60. Robens (1970a). Robens, J. E. (1974). Teratogenesis. In "Current Veterinary Therapy. V. Small Animal Practice" (R. W. Kirk, ed.), pp. 152-154. Saunders, Philadelphia. Robison, A. K., and Stancel, G. M. (1982). The estrogenic activity of DDT: Correlation of estrogenic effect with nuclear level of estrogen receptor. Life Sci. 31, 2479-2484. Rodwell, D. E., Nemec, M. D., Tasker, E. J., Murphy, J. M., and Simpson, J. E. (1987). Measuring the effects of a heartwormlhookworm preventative on canine reproduction. Vet. Med. pp. 438-447. Rogers, J. M., Barbee, B., Burkhead, L. M., Rushin, E. A., and Kavlock, R J. (1988). The mouse teratogen dinocap has lower AID ratios and is not teratogenic in the rat and hamster. Teratology 37, 553-559. Rogers, J. M., Carver, B., Gray, L. E., Gray, J. A., and Kavlock, R J. (1986). Teratogenic effects of the fungicide dinocap in the mouse. Toxicologist 6, 91. Rogers J. M., Francis, B. M., Barbee, B. D., and Chernoff, N., (1991). Developmental toxicity of bromoxynil in mice and rats. Fundam. App!. Toxieol. 17,442-447. Rogers, J. M., Francis, B. M., and Chernoff, N. (1990). Developmental toxicity assessment of bromoxynil in rats and mice. Toxicologist 10, 30. Rogers, J. M., Gray, L. E., Carver, B. D., and Kaviock, R. J. (1987). The developmental toxicity of dinocap in the mouse is not due to two isomers of the major active ingredients. Teratology 35, 62A. Roll, R. (1971a). Teratologic studies with thiram (TMTD) [tetramethylthiuram disulfidel on two strains of mice. Arch. Toxieo!. (Berlin) 27, 173-186. Roll, R (1971b). Untersuchungen uber die teratogenic Wirkung von 2,4,5-T bei Mausen. Food Cosmet. Toxieo!. 9,671-676. Roll, R., and Matthiaschk, G. (1983). Comparative studies on the embryotoxicity of 2-methyl-4-chlorophenoxyacetic acid, mecoprop and dichlorprop in NMRI mice. Arzneim. Forseh. 33, 1479-1483. Rosenstein, L., Bruce, A., Rogers, N., and Lawrence, S. (1977). Neurotoxicity of Kepone in perinatal rats following in utero exposure. Toxieol. App!. Pharmaeo!' 41, 142-143. Roy, T. S., Andrews, J. E., Seidler, E J., and Slotkin, T. A. (1998). Chlorpyrifos elicits mitotic abnormalities and apoptosis in neuroepithelium of cultured rat embryos. Teratology 58, 62-68.

Ruddick, J. A., Black, W. D., Villeneuve, D. c., and Valli, V. F. (1983). A teratological evaluation following oral administration of trichloro- and dichlorobenzene isomers to the rat. Teratology 27, 73A-74A. Ruddick, J. A., Newsome, W. H., and Nash, L. (1976). Correlation of teratogenicity and molecular structure: Ethylenethiourea and related compounds. Teratology 13, 263-266. Rutledge, J. c., and Generoso, W. M. (1989). Fetal pathology produced by ethylene oxide treatment of the murine zygote. Teratology 39, 563-572. Rutledge, J. c., Generoso, W. M., Shourbaji, A., Cain, K. T., Gans, M., and Oliva, J. (1992). Developmental anomalies derived from exposure of zygotes and first-cleavage embryos to mutagens. Mutat. Res. 296,167-177. Rutledge, J. c., Shourbaji, A. G., Hughes, L. A., Polifka, J. E., Cruz, Y. P., Bishop, J. B., and Generoso, W. M. (1994). Limb and lower-body duplications induced by retinoic acid in mice. Proe. Nat!. Aead. Sei. U.S.A. 91, 5436-5440. Ruttkay-Nedecka, J., Cerey, K., and Rosival, L. (1972). Evaluation of the chronic toxic effect of heptachlor. "Kongr. Chem. Pol.'nohospod," 2nd, p. 2. Saegusa, T., Naito, Y., and Narama, L. (1987). Teratogenicity study of subcutaneously administered d-d- T80-prallethrin (S4068SF) in rabbits. Oyo Yakuri 234,319-325. Saillenfait, A. M., Bonnet, P., Guenier, J. P., and Deceaurrer, J. (1989). Inhalation teratology study on hexachloro-l,3-butadiene in rats. Toxieo!. Lett. 47, 235-240. Saillenfait, A. M., Gallissot, F., Bonnet, P., and Protois, J. C. (1996). Developmental toxicity of inhaled ethylene oxide in rats following short-duration exposure. Fundam. Appl. Toxieo!. 34,223-227. Saito, M., Kumagai, Y., and Narama, L. (1987). A teratological evaluation following subcutaneous administration of metoxadiazone (S-21074) to rats. Oyo Yakuri 34, 147-162. Saito, R, Teramoto, S., and Shirasu, T. (1980).2 generation reproduction studies in rats with 4,5,6,7-tetrachlorophthalide fthalide. 1. Pestie. Sci. 5, 357362. Sare, W. M., and Forbes, P. 1. (1972). Possible dysmorphogenic effects of an agricultural chemical: 2,4,5-T. N. Z. Med. J. 75, 37-38. Sare, W. M., and Forbes, P. 1. (1977). The herbicide 2,4,5-T and its possible dysmorphogenic effects. N. Z. Med. J. 85,439. Sarkar, S. N., Majumdar, A. C., and Chattopadhayay, S. K. (1997). Effect of isoproturon on the male reproductive system: Clinical, histological and histoenzymological studies in rats. Indian J. Exp. Bioi. 35, 133-138. Schaefer, c., and Peters, P. W. J. (1992). Intrauterine diethyltoluamide exposure and fetal outcome. Reprod. Toxieol. 6, 175-176. Schantz, S. L., and Bowman, R. E. (1989). Learning in monkeys exposed prenatally to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Neurotoxieo!' Terato!. 11, 13-19. Schardein, J. L., (1993). "Chemically Induced Birth Defects." Dekker, New York. Schardein, J. L. (1998). Schardein, J. L., and Keller, K. A. (1989). Crit. Rev. Toxieol. Schardein J. L., and Scialli, A. R (1999). The legislation of toxicologic safety factors: The Food Quality Protection Act with chlorpyrifos as a test case. Reprod. Toxieol. 13, 1-14. Scheufler, H. (1975). Effect ofrelatively high doses of dimethoate and trichlorfon on the embryogenesis of laboratory mice. Bio!. Rundseh. 13, 238-240. Scheufler, H. (1976). Experimental testing of chemical agents for embryotoxicity, teratogenicity and mutagenicity-ontogenic reactions of the laboratory mouse to these injections and their evaluation-A critical analysis method. Bio!. Rundseh. 14,227-229. Schmidt, R. (1979). Prenatal toxic effect of hexamethylphosphoric acid triamide (HMPT). z. Gesamte Hyg. 25, 662-664. Schoenig, G. P., Neeper-Bradley, T. L., Fisher, L. c., and Hartnagel, R E. (1994). Teratologic evaluations of N, N-diethyl-m-toluamide (DEET) in rats and rabbits. Fundam. App!. Toxieol. 23, 63-68. Schuck, P. H. (1986). "Agent Orange on Trial: Mass Toxic Disasters in the Courts." Belknap, Cambridge, MA. Schwartz, D. A., and LoGerfo, J. P. (1988). Congenital limb reduction defects in the agricultural setting. Am. J. Public. Health 78, 654-658.

References

Schwartz, D. A., Newsum, L. A., Markowitz, and Heifetz, R. (1986). Parental occupation and birth outcome in an agricultural community. Scand. J. Work Environ. Health 12,51-54. Schwetz, B. A. (1994). "Chemical Industty Institute of Toxicology Activities," Vol. 14, p. 2. Chem. Ind. of Toxicol., Research Triangle Park, Ne. Schwetz, B. A., Keeler, P. A., and Gehring, P. J. (1974). The effect of purified and commercial grade pentachlorophenol on rat embtyonal and fetal development. Toxicol. Appl. Pharmacol. 28, 151-161. Scialli, et al. (1995). Sendrowski, e., Jadczak, J., and Karolski, K. (1977). Case of sirenomelia. Wia~Lek.

30,893-895.

Sesline, D. H., and Jackson, R. J. (1994). The effects of prenatal exposure to pesticides. In "Prenatal Exposure to Toxicants: Developmental Consecqences" (H. L. Needleman and D. Bellinger, eds.), pp. 233-248. Johns Hopkins Press, Baltimore. Sette, W. F. (1989). Adoption of new guidelines and data requirements for more extensive neurotoxicity testing under FIFRA. Toxicol. Ind. Health 5, 181194. Setterberg, F., and Shavelson, L. (1993). "Toxic Nation: The Fight to Save Our Communities from Chemical Contamination." Wiley, New York. Sever, L. E. (1995). Male-mediated developmental toxicity. Epidemiology 6, 573-574. Sever, L. E. (1988). "The State of the Art and Current Issues Regarding Reproductive Outcomes Potentially Associated with Environmental Exposures: Reduced Fertility, Reproductive Wastage, Congenital Malformations, and Birthweight." Report 600/8-89/103, Work Group on Reproductive and Developmental Epidemiology: Research Issues, U.S. Environmental Protection Agency, Washington, CD. Sever, L. E. (1997). Sever, L. E., ArbuckIe, T. E., and Sweeney, A. (1997). Occup. Environ. Med. Shalette, M. L., Cotes, N., and Goldsmith, E. D. (1963). Effects of 3-amino1,2,4-triazole treatment during pregnancy on the development and structure of the thyroid of the fetal rat. Anat. Rec. 145, 284. Shawky, A. S. H., Gomaa, E. A., Bakty, H., Kadty, A. M., and Sherif, R. (1984). Mutagenicity and teratogenicity of the synthetic pyrethroid insecticide cypermethrin in albino rats. Genetics 107,598. Shen, S. Y. (1983). Toxicity of cyolane. Chung Hua Yu I Hsueh Tsa ChIh 17, 216-218. Shepard (1986). Shepard (1988). Shepard (1998). Shepelskaya, N. R. (1988). Gonadotoxic effect of dikurin under intragastric administration by probes and with food. Gig. Sanit. 11,78-79. Sherman, J. D. (1995). Chlorpyrifos (Dursban)-associated birth defects: A proposed syndrome. Int. J. Occupat. Med. Toxicol. 4. Sherman, J. D. (1997). Dursban revisited: Birth defects, U.S. Environmental Protection Agency, and Centers for Disease Control. Arch. Environ. Health 52,332-333. Shivanandappa, T., and Krishnakumari, M. K. (1983). Hexachlorocyclohexaneinduced testicular dysfunction in rats. Acta Pharmacol. Toxicol. (Copenhagen) 52, 12-17. Short, R. D., Minor, J. L., Lee, e.-e., Chernoff, N., and Baron, R. L. (1980). Developmental toxicity of Guthion in rats and mice. Arch. Toxicol. 43, 177186. Short, R. D., Minor, J. L., Winston, J. M., Seifter, J., and Lee, C. C. (1978). Inhalation of ethylene dibromide during gestation by rats and mice. Toxicol. Appl. Pharmacol. 46, 173-182. Short, R. D., Russel, J. Q., Minor, J. L., and Lee, C.-e. (1976). Developmental toxicity of ferric dimethyidithiocarbamate and bis(dimethylthiocarbamoyl) disulfide in rats and mice. Toxieol. Appl. Pharmaeol. 35, 83-94. Smalley, H. E., O'Hara, P. J., Bridges, C. H., and Radeleff, R. D. (1969). The effects of chronic carbaryl administration on the neuromuscular system of swine. Toxicol. Appl. Pharmacol. 14,409-419. Smith, A. H., Fisher, D. 0., Pearce, N., and Chapman, e. J. (1982). Congenital defects and miscarriages among New Zealand 2,4,5-T sprayers. Arch. Environ. Health 37, 197-200.

421

Smith, F. A., Schwetz, B. A., and Nitschke, K. D. (1976). Teratogenicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in CF-l mice. Toxieol. Appl. Pharmacol. 38,517-523. Smith, M. K., Randall, J. L., and Stober, J. A. (1988). Developmental effects of trichloroacetic acid in Long-Evans rats. Teratology 37, 495. Smith, R. F. and Goldman, L. (1983). Behavioral effects of prenatal exposure to ethylene dibromide. Neurobehav. Toxieol. Teratol. 5, 579-586. Solecki, R., Faqi, A. S., Pfeil, R., and Hilbig, V. (1996). Effects of methyl parathion on reproduction in the Japanese quail. Bull. Environ. Contam. Toxieol. 57, 902-908. Spagnolo, A., Bianchi, F., Calabro, A., Calzolari, E., Clementi, M., Mastroiacovo, P., Meli, P., Petrelli, G., and Tenconi, R. (1994). Anophthalmia and benomyl in Italy: A multicenter study based on 940,615 newborns. Reprod. Toxicol. 8, 397-403. Sparschu, G. L., Dunn, F. L., Lisowe, R. w., and Rowe, V. K. (1971). Study of the effects of high levels of 2,4,5-trichlorophenoxyacetic acid on foetal development in the rat. Food Cosmet. Toxicol. 9, 527-530. Sparschu, G. L., Dunn, F. L., and Rowe, V. K. (1970). Teratogenic study of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat. Toxieol. Appl. Pharmacol. 17,317-318. Spasovski, M., Khristeva, v., Pervov, K., Kirhov, v., Dtyanovska, T., Panova, Z., Bobev, G., Gincheva, D., and Ivanova, S. (1980). Health state of the workers in the production of ethylene and ethylene oxide. Khig. Zdraveopaz. 23,41-47. Spielmann, H., and Vogel, R. (1989). Unique role of studies on preimplantation embtyos to understand mechanisms of embtyotoxicity in early pregnancy. Crit. Rev. Toxicol. 20, 51-64 Spyker, J. M., and Avety, D. L. (1977). Neurobehavioral effects of prenatal exposure to the organophosphate diazinon in mice. J. Toxicol. Environ. Health 3, 989-1002. Stellman, S. D. (1988). Health and reproductive outcome: American Legionnaires in relation to combat experience in Vietnam: Associated and contributing factors. Environ. Res. 47, 150. Stellman, S., and Stellman, J. (1980). Health problems among 535 Vietnam veterans potentially exposed to herbicides. Am. J. Epidemiol. 112, 444. Sterling, T. D. (1971). Difficulty of evaluating the toxicity and teratogenicity of 2,4,5-T from existing animal experiments. Science 174, 1358-1359. Sternberg, S. S. (1979). The carcinogenesis, mutagenesis and teratogenesis of insecticides: Review of studies in animals and man. Pharmacol. Ther. 6, 147-166. Stevens, K. M. (1981). Agent Orange toxicity: A quantitative perspective. Hum. Toxieol. 1,31-39. Stump, D. G., Clevidence, K. J., Knapp, J. F., Holson, J. F., and Farr, e. H. (1998a). An oral developmental toxicity study of arsenic trioxide in rats. Teratology 57, 216-217. Stump, D. G., Fleeman, T. L., Nemec, M. D., Holson, J. F., and Farr, e. H. (1998b). Evaluation of the teratogenicity of sodium arsenate and arsenic trioxide following single oral or intraperitoneal administration in rats. Teratology 57, 217. Stump, D. G., Nemec, M. D., Holson, J. F., Piccirillo, V. J., and Mares, J. T. (1997). Study of the effects of sulfluramid on pre- and postnatal development, maturation and fertility in the rabbit. Toxicologist 36, 357. Stump, D. G., Ulrich, e. E., Holson, J. F., and Farr, C. H. (1998c). An inhalation developmental toxicity study of arsenic trioxide in rats. Teratology 57, 216. Sullivan, F. M., and Barlow, S. M. (1979). Congenital malformation and other reproductive hazards from environmental chemicals. Proe. R. Soc. Lond. 205,91-110. Swartz, W. J., and Eroschenko, V. P. (1998). Neonatal exposure to technical methoxychlor alters pregnancy outcome in female mice. Reprod. Toxicol. 12, 565-573. Swartz, w.J., and Mall, G. M. (1989). Chlordecone-induced follicular toxicity in mouse ovaries. Reprod. Toxicol. 3, 203-206. Swentzel, K. e., Angerhofer, R. A., Haight, E. A., McCreesh, A. H., and Weeks, M. H. (1978). Safety evaluation of the synthetic pyrethroid insecticide, resmethrin, as a clothing impregnant. Toxieol. Appl. Pharmacol. 45, 243. Taha, T. E., and Gray, R. H. (1993). Agricultural pesticide exposure and perinatal mortality in central Sudan. Bull. World Health Org. 71, 317-321.

422

CHAPTER 14


Talbot, A. R., Fu, C. c., and Hsieh, M. F. (1988). Paraquat intoxication during pregnancy: A report of9 cases. Vet. Hum. Toxieo!. 30,12-17; Erratum. Vet. Hum. Toxieo!. 30, 245. Talts, U., Fredriksson, A., and Eriksson, P. (1998). Changes in behavior and muscarinic receptor density after neonatal and adult exposure to bioallethrin. Neurobio!. Aging 19, 545-552. Tanaka, T., Fujitani, T., Takahashi, 0., Oishi, S., and Yoneyama, M. (1997). Developmental toxicity of chlorpropham in mice. Reprod. Toxieol. 11,697701. Tang, J., Carr, R. L., and Chambers, J. E. (1999). Changes in rat brain cholinesterase activity and muscarinic receptor density during and after repeated oral exposure to chlorpyrifos in early postnatal development. Toxieo!' Sci. 51, 265-272. Tanimura, T. (1985). Guidelines for developmental toxicity testing of chemicals in Japan. Neurobehav. Toxieo!. Terato!' 7, 647. Tanimura, T., Katsuya, T., and Nishimura, H. (1967). Embryotoxicity of acute exposure to methyl parathion in rats and mice. Arch. Environ. Health 15, 609-613. Taylor, J. R., Selhorst, J. B., Houff, A. S., and Martinez, A. J. (1978). Chlordecone intoxication in man. 1. Clinical observations. Neurology 28, 626-630. Taylor, P. (1986). "Practical Teratology," p. 2. Academic Press, San Diego. Teramoto, S., Shingu, A., Kaneda, M., and Saito, R. (1978). Teratogenicity studies with ethylenethiourea in rats, mice and hamsters. Congenital Anom. 18, 11-17. TerHaar, G. (1980). An investigation of possible sterility and health effects from exposure to ethylene dibromide. Banbury Rep. 5, 167-188. Thigpen, J. E., Setchell, K D. R., Goelz, M. F., and Forsythe, D. B. (1999). The phytoestrogenic content of rodent diets. Environ. Health Perspeet. 107, AI82-AI83. Thomas, D., Goldhaber, M., and Petitti, D. (1990). Reproductive outcomes in women exposed to malathion. Am. J. Epidemio!. 132,794-795. Thomas, D. c., Petitti, D. B., Goldhaber, M., Swan, S. H., Rappaport, E. B., and Hertz-Picciotto, 1. (1992). Reproductive outcomes in relation to malathion spraying in the San Francisco Bay area, 1981-1982. Epidemiology 3, 3239. Thomas, H. F. (1980). 2,4,5-T use and congenital malformation rates in Hungary. Lancet 2, 214-215. Thomas, H. F. and Czeizel, A. (1982). Safe as 2,4,5-T? Nature 295,276. Thompson, D. J., Emerson, J. L., Strebling, R. J., Gerbig, C. G., and Robinson, V. B. (1972). Teratology and postnatal studies on 4-amino-2,5,6trichloropicolinic acid (picloram) in the rat. Food Cosmet. Toxieol. 10, 797-803. Tilson, H. A. (1998). Developmental neurotoxicology of endocrine disruptors and pesticides: Identification of information gaps and research needs. Environ. Health Perspeet. 106,807-811. Torkelson, T. R., Sadek, S. R., Rowe, V. K, Dodama, J. K, Anderson, H. H., Loquvam, G. S., and Hine, C. H. (1961). Toxicologic investigation of 1,2dibromo-3-chloropropane. Toxieo!. App!. Pharmaeol. 3, 545-559. Traina, M. E., Fazzi, P., Macri, c., Ricciardi, c., Stazi, A. V., Urbani, E., and Mantovani, A. (1998). In vivo studies on possible adverse effects on reproduction of the fungicide methyl thiophanate. J. App!. Toxieol. 18,241-248. Trajkovic, D., Matic, G., Radojcic, M., and Petrovic, J. (1981). Effects of trichlorphon and parathion on steroid hormones binding to sytosol receptors of different target tissues ofrats. Acta Bio!. Med. Exp. 6, 71-76. Trapp, M., Bauloh, V., Bonnet, H. G., and Herschen, W. (1984). Ferti!. Steri!. 42,146-148. Trost, C. (1984). "Elements of Risk: The Chemical Industry and Its Threat to America." Times Books, New York. Trutter, J. A., Arce, G. T., Piccirillo, V. J., Wakefield, A. E., and Robertson, D. B. (1995). Rat developmental toxicity study with disodium salt of Endothall. Teratology 51, 200. Tsaregorodtseva, G. N., and Talanov, G. A. (1973). Embryotoxic and teratogenic effect of chlorophos, TCM-33, sevin and dicresyl on white rats. Tr. Vses. Nauehno-issled. Inst. Vet. Sanit. 47, 150--155. Turck, P. A., Eason, C. T., and Wickstrom, M. (1998). Assessment of the developmental toxicity of sodium monofiuoroacetate (1080) in rats. Toxicologist 42,258-259.

Tyrkiel, E. (1978). Effect of o-isopropoxyphenyl-N -methylcarbamate (Propoxur) on embryonal development of mice. Roez. Panstw. Zak!. Hig. 29, 655-664. Ueki, R., Nakao, Y., Nishida, T., Nakao, Y., and Wakabayashi, T. (1990). Lung hypoplasia in developing mice and rats induced by maternal exposure to nitrofen. Congenital Anom. 30, 133-143. Umpierre, C. C. (1981). Embryolethal and teratogenic effects of sodium arsenite in rats. Teratology 23, 66A. Unsworth, B., Hennen, S., Krishnakumaran, A., Ting, P., and Hoffman, N. (1974). Teratogenic evaluation of terpenoid derivatives. Life Sei. 15, 16491655. Uphouse, L. (1986). Single injection with chlordecone reduces behavioral receptivity and fertility of adult rats. Neurobehav. Toxieo!. Terato!' 8, 121126. Uphouse, L., and Williams, J. (1989). Sexual behavior of intact female rats after treatment with 0, p'-DDTor p, p'-DDT. Reprod. Toxieo!. 3,33-41. Usarni, M., Kawashima, K, Nakaura, S., Yamaguchi, M., Tanaka, S., Takanaka, A., and Omori, Y. (1986). Effect of chlordane on prenatal development ofrats. Eisei Shikenjo Hokoku 104, 68-73. U.S. Environmental Protection Agency (EPA). IRIS Database http://www.epa.gov/ngispgm3/iris/subst/0336.htm. U.S. Environmental Protection Agency (EPA). (1978). Proposed guidelines for registering pesticides in the U.S.: Hazard evaluation: Humans and domestic animals. Fed. Regist. 43, 37336-37403. U.S. Environmental Protection Agency (EPA). (1982). "Pesticide Assessment Guidelines (FIFRA). Subdivision F. Hazard Evaluation: Humans and Domestic Animals." EPA Report 540/9-82-025, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA). (1984). 40 CFR Part 158. Data requirements for pesticide registration: Final rule. Fed. Regis. 49, 4285642905. U.S. Environmental Protection Agency (EPA). (1985). "Standard Evaluation Procedure. Teratology Studies." EPA Report 540/9-85-018, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA). (1986). Developmental neurotoxicity screen. Fed. Regist. 51, 17890--17894. U.S. Environmental Protection Agency (EPA). (1988). Diethylene glycol butyl ether and diethylene glycol ether acetate: Final test rule. Fed. Regist. 53, 5932-5953. U.S. Environmental Protection Agency (EPA). (1989). Triethylene glycol monomethyl ether: Final test rule. Fed. Regist. 54, 13472-13477. U.S. Environmental Protection Agency (EPA). (1991). Guidelines for developmental toxicity risk assessment. Fed. Regist. 56, 63788-63826. U.S. Environmental Protection Agency (EPA). (1996a). "The Food Quality Protection Act of 1996." H.R. 1627, P.L. 104-170. U.S. Environmental Protection Agency (EPA). (1996b). "Is an Additional Uncertainty Factor Necessary and Appropriate to Assess Pre- and Post-Natal Developmental and Reproductive Effects in Infants and Children Exposed to Pesticides throught Chronic Dietary Exposure?" U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA). (1998). "Health Effects Test Guidelines OPPTS 870.3700. Prenatal Developmental Toxicity Study" EPA 12-98-207. U.S. EPA, Washington, DC. Uzoukwu, M., and Sleight, S. D. (1972a). Dieldrin toxicosis: Fetotoxicosis, tissue concentrations, and microscopic and ultrastructural changes in guinea pigs. Am. J. Vet. Res. 33,579-583. Uzoukwu, M., and Sleight, S. D. (1972b). Effects of dieldrin in pregnant sows. J. Am. Vet. Med. Assoc. 160, 1641-1643. van den Bosch, R. (1989). "The Pesticide Conspiracy." Univ. of California Press, Berkeley. Vanhauwere, B., Maradit, H., and Kerr, L. (1998). Post-marketing surveillance of prophylactic mefioquine (Lariam) use in pregnancy. Am. J. Trop. Med. Hyg.58,17-21. Van Strum, C. (1983). "A Bitter Fog: Herbicides and Human Rights." Sierra Club Books, San Francisco. Vara and Kinnunen (1951).

References

Vashakidze, V. 1. (1965). Some questions of the harmful action of sevin on the reproductive function of experimental animals. Soobshch. Akad. Nauk Gruz. SSR 39, 471-474. Veis, V. P. (1970). Some data on the status of the sexual sphere in women who have been in contact with organochlorine compounds. Pediatr. Akush. Ginekol. 32,48-49. Vergieva, T. (1984). Experimental study of the teratogenicity and embryotoxicity of endodan. Probl. Khig. 988-995. Vergieva, T. (1985). Behavioral teratology-Results achieved and perspectives of development. J. Hyg. Epidemiol. Microbiol. Immunol. 29, 121-127. Vergieva, T. (1990). Triazoles teratogenicity in rats. Teratology 42, 27A-28A. Voloshina, L. T. (1985). Embryotropic effect offenuron. Vopr. Onkol. 31, 103105. Vos, J. G., and Moore, J. A. (1974). Suppression of cellular immunity in rats and mice by maternal treatment with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Int. Arch. Allergy Appl. Immunol. 47, 777-794. Wargo, J. (1996). "Our Children's Toxic Legacy: How Science and Law Fail to Protect Us from Pesticides." Yale Univ. Press, New Haven, CT. Weber, H., Harris, M. W., Haseman, J., and Birnbaum, L. S. (1985). Teratogenic potency of TCDD, TCDF and TCDD-TCDF combinations in C57BU6N mice. Toxicol. Lett. 26, 159-167. Weil, C. S., Woodside, M. D., Carpenter, C. P., and Smyth, H. F., Jr. (1972). Current status of tests of carbaryl for reproductive and teratogenic effect. Toxicol. Appl. Pharmacol. 23, 351-364. Weinberg, C. R. (1993). Toward a clearer definition of confounding. Am. J. Epidemiol. 137, 1-8. Weir, D., and Schapiro, M. (1981). "Circle of Poison: Pesficides and People in a Hungry World." Institute for Food and Development Policy, San Francisco. Welch, R. M., Levin, w., Kuntzman, Jacobson, M., and Conney, A. H. (1971). Effect of halogenated hydrocarbon insecticides on the metabolism and uterotropic action of estrogens in rats and mice. Toxicol. Appl. Pharmacol. 19, 234--246. Welsh, J. J., Collins, T. F. X., Black, T. N., Graham, S. L., and 0' Donnell, M. W. (1985). Teratogenic potential of purified PCP and PCA in subchronically exposed Sprague-Dawley rats. J. Am. Coli. Toxicol. 4, 143-144A. Wheeler, L., and Strother, A. (1974). Placental transfer, excretion, and disposition of [' 4C] Zectron and [' 4C] Mesurol in maternal and fetal rat tissues. Toxicol. Appl. Pharmacol. 30, 163-174. Whiteford, R. B. (1977). Deformed puppies. Vet. Rec. 100, lI8. Whiteside, T. (1979). "The Pendulum and the Toxic Cloud: The Course of Dioxin Contamination." Yale Univ. Press, New Haven, CT. Whorton, M. D., and Foliart, D. E. (1983). Mutagenicity, carcinogenicity and reproductive effects of dibromochloropropane (DBCP). Mutat. Res. 123, 13-30. Whorton and Foliart (1988). Whorton, D., Krauss, R. M., Marshall, S., and Milby, T. (1977). Infertility in male pesticide workers. Lancet 2, 1259-1261. Willhite, C. C. (1981). Arsenic-induced axial skeletal (dysraphic) disorders. Exp. Mol. Pathol. 34, 145-158. Willhite, C. c., Ferm, V. H., and Smith, R. P. (1981). Teratogenic effects of aliphatic nitriles. Teratology 23,317-323. Willhite, C. c., Ferm, V. H., Marin-Padilla, M., and Smith, R. P. (1980). Developmental malformations induced by metabolically-liberated cyanide. Toxicol. Appl. Pharmacol. A88.

423

Wilkinson, C. F. (1990). In "Effects of Pesticides on Human Health" (S. R. Baker and C. F. Wilkinson, eds.). Advances in Modern Environmental Toxicology, Vo!. 18, p. 25. Princeton Sci. Pub., Princeton, NJ. Williams, P. L. (1982). Pentachlorophenol, an assessment of the occupational hazard. Am. Ind. Hyg. Assoc. J. 43, 799--810. Willis, W. 0., de Peyster, A., Molgaard, C. A., Walker, C., and MacKendrick, T. (1993). Pregnancy outcome among women exposed to pesticides through work or residence in an agricultural area. J. Occup. Med. 35, 943-949. Wilson, J. G. (1971). "Report on the Treatment of Pregnant Rhesus Monkeys with 2,4,5-T Acid." Report submitted to Swedish National Poisons and Pesticides Board, Stockholm. Wilson, J. G. (1979). The evolution of teratological tests. Teratology 20(2), 205-211. Wilson, J. G., Fradkin, R., and Schumacher, H. J. (1970). Influence of drug pretreatment on the effectiveness of known teratogenic agents. Teratology 3,210--211.

Wong, 0., Utidjian, H. M. D., and Karten, V. S. (1979). Retrospective evaluation of reproductive performance of workers exposed to ethylene dibromide. J. Occup. Med. 21, 98-102. World Health Organization (WHO) (1986). "Draft Guideline for the Assessment of Drugs and Other Chemicals for Behavioural Teratogenicity." WHO Regional Office for Europe, Copenhagen. Wright, D. M., Hardin, B. D., Goad, P. w., and Chrislip, D. W. (1992). Reproductive and developmental toxicity of N, N -diethyl-m-toluamide in rats. Fundam. Appl. Toxicol. 19, 33-42. Wu, H., Zhiwei, L., Xu, H., Ruikun, S., Ma, T., Shi, N., Siu, R., and Liu, Y. (1989). Toxicological studies on the organophosphorus insecticide methylISP. J. Tongii. Med. Univ. 9, 58-64. Wyrobek, A. J., Watchmaker, G., Gordon, L., Wong, K., Moore, D., n, and Whorton, D. (1981). Sperm shape abnormalities in carbaryl-exposed employees. Environ. Health Perspect. 40, 255-265. Yakubova, Z. N., Shamova, N. A., Miftakhova, F. A., and Shilova, L. F. (1976). Gynecological disorders in workers engaged in ethylene oxide production. Kazan. Med. Zh.57,558-560. Yamamoto, H., Kuchli, M., and Hayano, T. (1970). Effect of prothrin (D-1201), a pyrethroidal insecticide, on mouse and fetuses. Oyo Yakun 4, 779-787. York, R. G., Butala, J. H., Ulrich, C. E., and Schardein, J. L. (1994). Inhalation developmental toxicity studies of chloropicrin in rats and rabbits. Teratology 49,419. Zablotny, C. L., Breslin, W. J., and Kociba, R. J. (1992). Developmental toxicity of orthophenylphenol (OPP) in New Zealand white rabbits. Toxicologist 12, 103. Ziborov, N. A., Malakhova, E. L., and Veselova, T. P. (1982). Evaluation of the effect of sodium fluorosilicate on swine reproductive function. Byoll. Vses. Inst. Gel'mintol. 32, 33-36. Zielke, G. J., Yano, B. L., and Breslin, W. J. (1993). 2,3,5,6Tetrachloropyridine: Combined repeat dose and reproductive/developmental toxicity screen in Sprague-Dawley rats. Toxicologist 13,77.

Zurawski, J. M., and Kelly, E. A. (1997). Pregnancy outcome after maternal poisoning with brodifacoum, a long-acting warfarin-like rodenticide. Obstet. Gynecol. 90, 672-673.

CHAPTER

15 Worker Exposure: Methods and Techniques Graham Chester Syngenta

15.1 INTRODUCTION

15.2.1 DERMAL EXPOSURE

Pesticides are biologically active compounds, which may pose a health risk to agricultural workers during or after their use. Operators involved in handling, dispensing, and applying pesticides and postapplication crop reentry workers will be exposed to these compounds through different routes and to varying extents. It is essential both for stewardship and regulatory approval purposes that the possible health risk associated with this exposure is assessed using quantitative information on the toxicological hazard and the amount of exposure. This chapter presents methods by which exposure can be determined. It is not intended to be a complete review of the literature or to provide detailed guidance on how to measure operator exposure or conduct a field exposure study but rather a summary of current "state of the art" principles and methodology involved in measuring exposure of agricultural workers to pesticides. The processes by which such exposure data are used and interpreted are dealt with elsewhere in this book.

In practice, two measurements or estimations are usually made for all work activities associated with the use of pesticides:

1. Potential dermal exposure-the total amount of pesticide coming into contact with the protective clothing, work clothing, and skin. 2. Actual dermal exposure-the amount of pesticide coming into contact with the bare (uncovered) skin and the fraction transferring through protective and work clothing or via seams to the underlying skin, which is therefore available for percutaneous absorption. The biological availability or absorption of a pesticide via the dermal route of exposure is a property of the formulated product and the diluted material and is a separate subject in its own right. Given the significance of the dermal route, precise determinations of percutaneous absorption are key components of the overall assessment of the absorbed dose of the pesticide for risk assessment. 15.2.2 EXPOSURE BY INHALATION

15.2 ROUTES OF EXPOSURE Agricultural workers involved in the use of pesticides and postapplication crop reentry activities may be exposed to pesticides via the skin, by inhalation, or by accidental oral ingestion. Exposures via the first two routes are usually determined separately, but little attention is paid to oral ingestion because it is difficult to estimate or measure. Exposure is usually greatest by the dermal route, although inhalation can be an important route for pesticides that have significant vapor pressures, are applied in confined spaces, or have an application technique which generates a significant proportion of respirable or inhalable particles. Handbook of Pesticide Toxicology Volume 1. Principles

Fundamentally, as far as possible health effects are concerned and setting aside volatile pesticides for the moment, the only spray droplets or particles that pose a potential risk comprise the so-called inhalable or inspirable fraction, which is the mass fraction of airborne particulate capable of entering the respiratory tract via the nose and the mouth, so providing a source of absorption into the body, either from direct inhalation or from subsequent oral absorption. This is considered to be the most important indicator of potential inhalation exposure (ACGIH, 1985; Vincent and Mark, 1987). The inhalable fraction depends on the speed and direction of the air movement, on the rate of breathing, and on other factors. For sampling purposes, inhalable particles can be considered to have a mass median diameter of ~1 00 Jl.m diameter or less. The respirable fraction is the mass

425

Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved..

426

CHAPTER 15

Worker Exposure: Methods and Techniques

fraction of inhaled particles, which penetrates to the unciliated airways. For sampling purposes, respirable particles can be considered to have a diameter between 0 and 15 J.l.m (ISO, 1995). In risk assessment it is common to assume that volatile airborne pesticides are completely retained and absorbed via the respiratory tract, unless there are specific data to demonstrate otherwise. A reasonable default value is 50% retention and absorption of vapors as adopted by the Worker Health and Safety Branch of the California EPA Department of Pesticide Regulation (Thongsinthusak et aI., 1993). Inhalation exposure is usually a small fraction of the total exposure and can, in some cases, be ignored, for example, the mixing and loading of liquid formulations, particularly if a closed loading system is involved. Conditions under which exposure by the inhalation route becomes important usually involve the use of volatile pesticides or of dusts, fumigants, and sprays, especially in enclosed spaces. It should, however, be borne in mind that a higher proportion (up to 75%; Ross et aI., 2000) of the inhaled dose may be retained systemically, compared with the proportion absorbed after dermal exposure, which could be as low as 1% or less of the available dermal dose. 15.2.3 ORAL EXPOSURE

Some of the larger airborne particulates may be trapped in the mouth or nasal passages and subjected to oral ingestion. Some of the exposure, which is measured as inhalation, may indeed be trapped and absorbed in this way. No serious attempts have been made to measure separately the amount of exposure by this route because of the obvious difficulties involved. Biological monitoring takes into account all routes of absorption, but it is usually unable to distinguish between their relative contributions.

15.3 PREVIOUS REVIEWS AND GUIDANCE ON METHODOLOGY Durham and Wolfe (1962), Wolfe (1976), and Davis (1980) were responsible for the earliest reviews of methodology. These reviews and particularly the methodology of Durham and Wolfe (1962) were used to develop the World Health Organization (WHO) standard protocol for the measurement of exposure (WHO, 1975, 1982). The first WHO protocol (1975) advocated the Durham and Wolfe "patch" method to estimate dermal exposure and included a reference to biological monitoring through the use of cholinesterase activity measurement for organophosphorus insecticides. The revised protocol of 1982 proposed an alternative method for the measurement of dermal exposure, the "whole body" sampling technique, and gave an overview of biological monitoring as a means of measuring absorption arising from all routes of exposure. This revised protocol was used to develop the U.S. National Agricultural Chemicals Association (NACA) guidelines for mixerloader-applicator exposure studies (Mull and McCarthy, 1986).

These guidelines placed primary emphasis on the use of the Durham and Wolfe (1962) patch methodology rather than the whole body technique. NACA also published guidelines for conducting field biological monitoring studies as a means of measuring the absorption of pesticides (NACA, 1985). Both the United States Environmental Protection Agency (U.S. EPA, 1987) and NACA guidelines contained detailed reviews of the advantages and limitations of the methods for the measurement of exposure to and absorption of pesticides. The International Group of National Associations of Manufacturers of Agrochemical Products (GIFAP), now known as the Global Crop Protection Federation, published a position paper on general aspects of monitoring studies for the assessment of worker exposure to pesticides (GIFAP, 1990). These guidelines were intended to inform the nonspecialist of the various approaches to exposure/absorption evaluation and their significance. Curry and Iyengar (1992) reviewed and compared the currently available guidelines (published and unpublished) for the evaluation of exposure to individuals using pesticides and those exposed to residues in indoor and outdoor environments. Harmonized exposure and biological monitoring guidelines were proposed in a workshop held in the Netherlands (Chester, 1993; Henderson et aI., 1993). The guidelines agreed on at this workshop were further discussed at a Health CanadaINorth Atlantic Treaty Organization-sponsored "Workshop on Methods of Pesticide Exposure Assessment" held in Canada in 1993, resulting in a draft guidance document to be submitted to the Organization for Economic Co-operation and Development (OECD). A designated peer review group established at the workshop revised the guidance document, which was submitted to the OECD and published as an OECD Guidance Document (OECD,1997). The U.S. EPA meanwhile published a series of test guidelines in 1996 describing their preferred approaches to passive dosimetry and biological monitoring for exposure during indoor and outdoor occupational and residential use of pesticides (U.S. EPA, 1996). The guidance in these OECD and U.S. EPA documents represents the most up to date, harmonized approaches to the assessment of exposure to pesticides and the reader is referred to them for more detailed information.

15.4 DESIGN OF AGRICULTURAL WORKER EXPOSURE STUDIES The purpose of a worker exposure and/or biological monitoring study is to generate data for use in a risk assessment. Good study design is therefore a key consideration in ensuring that relevant and useful exposure data are obtained. In deciding on the basic methodological approach, reference can be made to the tiered approach to exposure and risk evaluation to determine if passive dosimetry will suffice or whether the use of biological monitoring is warranted to give the most accurate

15.4 Design of Agricultural Worker Exposure Studies

determination of the dose absorbed by the worker (Henderson et aI., 1993). Agricultural worker exposure studies can be regarded as being of two types: pre- and reregistration studies and postregistration surveillance studies (OECD, 1997). Studies of the first type involve the test subjects complying fully with the requirements of the product label, in particular, use of protective clothing and equipment, application rates, and clean-up procedures. By ensuring compliance, certain constraints are imposed on the activities of the workers that may influence the amount and variability of their exposures. Studies done according to these criteria are also appropriate for inclusion in generic exposure data bases since they will have been conducted according to a set of standardized principles. Studies of the second type are done primarily in support of product stewardship and postregistration evaluation of actual pesticide use conditions and practices. Adverse effects on the health of workers might be reported, or there might be a possible need to study the extent of compliance with product label precautions and recommendations. Therefore, the study design should take into account the need to measure exposure under the actual conditions of use and would be free of the constraints imposed on pre- or reregistration studies. Other factors that may influence the sampling strategy include possible concern about specific work activities during pesticide use, including the use of nonstandard application equipment. Passive dosimetry enables separate measurements of the respective contributions of these activities to the total exposure. Identification of the differences in the magnitude of exposure attributable to these activities permits the use of different regulatory proposals for reduction of exposure to acceptable levels. An example is the recommended use of additional protective equipment during procedures with greater potential for exposure, for example, handling and mixing of the concentrated formulation. In designing a study consideration should be given to whether passive dosimetry and biological monitoring should be conducted concurrently. Both may be justified to provide data for inclusion in generic data bases, to examine the relationship between exposure and absorption, and to provide a second measurement should one fail. The V.S. EPA, in their latest test guidelines, however, is strongly opposed to this idea (U.S. EPA, 1996). Their concern stems from the reasonable perception that the dermal passive dosimeters, in particular, intercept pesticide as it comes into contact with the worker, thus interfering with the process of dermal contamination and absorption and reducing the biological monitoring. However, if clothing dosimeters are representative of what workers normally use then this disadvantage can be overcome. This matter is discussed further in Section 15.6. Concerning the number of measurements of exposure and/or absorbed dose in the field study, as a general guide, OECD (1997) proposes a minimum of 10 different subjects. The V.S. EPA (1996) requires a minimum of 15 replicated measurements of exposure, not necessarily in different test subjects. Factors to consider in making the decision on the number of subjects are:

• • • •

427

the likely end use of the data; the nature of any identified toxicological endpoint; the required level of statistical confidence; and the overall manageability of the study.

Where feasible, subjects should be randomly selected from the worker population. It is recommended that a sufficient number of measurements be made in different locations to cover the range of use procedures, conditions, and application equipment for which exposure data are required. Variability in exposure can be addressed by increasing the number of subjects rather than repeatedly monitoring the same individuals. Variability between workers is typically greater than that within the same worker (Kromhout et aI., 1993; Rappaport, 1991). In addition, as many sites as possible should be included rather than having subjects use the same equipment under the same conditions. This applies particularly if location is believed to have a significant impact on the variability of the exposure measurements. Certain types of pesticide application procedures render management of the study difficult, such as those involving aerial application. In such cases, repeated monitoring of the same individuals is a possible option, although the limitations of such a choice should be recognized. Should biological monitoring be necessary, a further limitation is introduced in that interindividual variation in metabolism of the pesticide would be less well evaluated. The duration of the study will be prolonged owing to the possible need to collect urine samples over several days for each individual. The consequence is that repeat monitoring cannot commence until urine collection is complete. These difficulties should be considered when using biological monitoring in such circumstances. Ideally, the duration of a single measurement of exposure or absorbed dose should be representative of the typical working day, so that all the work activities that contribute to total exposure, such as equipment repair and clean up, are assessed. This criterion applies particularly to studies involving biological monitoring in which the skin is the predominant, if not only, route of exposure and absorption. For many pesticides that are not well absorbed dermally, absorbed dose data should not be linearly extrapolated on the basis of time or amount of active ingredient used in the same way as passive dosimetry exposure data. The percutaneous absorption of a pesticide depends on the rate at which it is absorbed, the area of skin contaminated, and the duration of skin contact. The rate will increase up to the maximum steady state rate. Once this is achieved, further increases in deposition of pesticide on the skin will have no further impact on the absorption process as it is saturated. Certainly, an increase in deposition can result in increased absorption up to maximum rate, but not in direct proportion (Chester, 1988). In studies involving passive dosimetry and volatile or unstable pesticides, a shorter monitoring period should be considered. This would be based on a consideration of the physicalchemical properties of the compound. The choice of monitoring duration should account for the possibility of dosimeter saturation. Ideally, a single set of dosimeters should be used per

428

CHAPTER 15


worker; however, a change of dosimeters during the workday may be necessary if different tasks are to be monitored separately. This can, however, be difficult to manage from the standpoint of practicality. The choice of use pattern (including application equipment) should account for factors such as whether it is the predominant one for the product or a minor one for which no generic data are available to the investigator or regulatory authority to enable a risk assessment. In pre- and reregistration studies the product should be used in the study at a representative recommended rate of application and on the likely maximum area of crop treatable in a working day under local conditions. It should also be applied in accordance with all the label recommendations for use. These principles also apply to studies involving postapplication crop reentry in which exposure should be determined after the shortest permissible reentry period, if known. In postregistration surveillance studies these criteria should not be enforced; this ensures maximum representativeness of actual use conditions and exposure variability. Where the product label recommends the use of protective clothing and/or equipment, these items should be provided to the subjects in pre- and reregistration studies by the supervisory team to ensure standardization. This will benefit the scientific interpretation of the data, because the variable of differing standards of protection by different types and conditions of protective equipment will have been removed. Inclusion of criteria such as these ensures exposure and risk assessment for the product in accordance with the label recommendations for regulatory use. However, efforts should be made to ensure that the recommended protective clothing and equipment are practical and realistic to use under local conditions. In postregistration surveillance studies, the study team should not mandate use of protective clothing and equipment, although ethical or legal viewpoints on product label recommendations in this respect should be considered.

15.5 TEST SUBJECTS Agricultural workers should be the test subjects in a field study, rather than inexperienced volunteers. If this is not possible, use of nonprofessional personnel may have to be considered, provided that they are given the requisite training in the handling and use of the pesticide and equipment. The disadvantage of this choice is that the subjects would not necessarily be representative of the worker population. Males or females may be considered for inclusion in passive dosimetry studies. However, in studies involving biological monitoring, the decision is more difficult owing to the potential impact on the interpretation of the metabolite excretion data due to possible sex differences in metabolism and kinetics. If it is likely that the product will be used by both sexes then it is important to know if there are differences. All subjects should be asked to provide written, informed consent to participate in a study after they are provided with the requisite information on the pesticide. Potential subjects must

be informed that they are free to withdraw from the study at any time. Subjects should be screened for any preexisting medical conditions that may be affected by use of the pesticide, depending upon its toxicological profile. Depending upon the circumstances and local custom, it may be appropriate to provide the subjects with information on their individual results.

15.6 METHODS FOR MEASURING EXPOSURE (PASSIVE DOSIMETRY) 15.6.1 PATCH METHOD FOR DERMAL EXPOSURE

In this method, the potential contamination of the workers' skin and clothing is measured using a variable number of absorbent cloth or paper patches, attached to body regions inside and outside clothing. The surface area covered by the patches represents less than lO% of the total body surface area. After a defined or measured period of exposure, the patches are removed and analyzed for pesticide content. The quantity of a pesticide on a patch of known area is then related to the area of limb or other body part on the assumption that deposition is uniform over the body parts. Body part surface areas can be obtained from standard reference texts and exposure guidance documents, the most recent of which are those of the V.S. EPA (1996,1999) and the OECD (1997). The assumption of uniform deposition is perhaps the principal disadvantage of the patch technique. This is illustrated by the extrapolation of the value given by half the limit of quantification to the total body part; this may give a substantial underor overestimate of exposure. The principal disadvantage can be mitigated to a certain extent by increasing the number of patches located on body parts likely to receive significant exposure. Individual body part exposure values are then added to give a total potential exposure expressed in mg/hour, mg/day, or mg/kg product handled or applied. In the WHO protocol fewer patches are recommended, representing only 3% of the body area. In both cases standard body part surface areas are used in correcting individual patch values (WHO, 1982; U.S. EPA, 1996). According to the WHO protocol, only the pesticide contacting the normally unclothed area of skin, for example, head, neck, hands, and forearms, is used to calculate actual exposure. In temperate climates about lO% of the total body area is normally unprotected during use of pesticides and other agricultural activities. Normal work clothing, such as cotton trousers or shirts, is absorbent and may retain and allow penetration of a proportion of the pesticide contamination. Therefore, it is still necessary to estimate exposure to the covered areas of the body. Consequently the amount of pesticide penetration is often measured using patches attached beneath the clothing. Without such measurements, an estimate of penetration of normal clothing must be used. However, penetration of clothing by pesticides is a highly variable process, which is influenced by factors such as the type of formulation (liquid or solid) and the amount or volume of deposition on the clothing, dampness of the clothing, pesticide vapor pressure,

15.6 Methods for Measuring Exposure (Passive Dosimetry)

429

the location of deposition (e.g., seams), and the type of fabric. Further uncertainties are introduced by the method of sampling for clothing penetration. The patch method may give significant under- or overestimates of exposure, depending on whether the patches have captured the nonuniform, random deposition of concentrate splashes or spray droplets. This limitation applies equally to crop reentry procedures, such as harvesting, where contact with the crop is not a uniform process. Despite the readily apparent limitations, the patch method remains a useful method for exposure evaluation.

acceptability applicable to all pesticides, the surrogate compound should not significantly alter the physical properties of the formulation or spray mixture. The key question that determines the utility of a tracer or dye is whether it affects, is retained by, or penetrates clothing similar to or in parallel with the pesticide of interest. The technique can only be used in exposure studies and not in biological monitoring (although it can be used with concurrent biological monitoring of a pesticide). Its greatest utility probably lies in substituting for pesticides that are particularly unstable during the sampling and analytical phases.

15.6.2 USE OF FLUORESCENT TRACERS AND VISIBLE DYES: QUANTIFICATION BY ANALYSIS OR VIDEO IMAGING

15.6.3 WHOLE BODY METHOD

Dermal exposure can be quantified directly by measuring deposition of fluorescent materials or visible dyes on the clothing and/or skin. The fluorescent tracer or dye can be substituted and extracted from passive dosimeters and analyzed in the same way as the pesticide. By adjustment for concentration differences, an estimate of exposure to the pesticide can be obtained, similar to extrapolation from one monitored pesticide to another used in the same way. A recent technical advance has been the development of a video imaging/fluorescent tracer technique (Fenske et aI., 1986a, b; Fenske, 1990). This method involves the incorporation of a fluorescent tracer in a pesticide formulation and subsequent visual and quantitative analysis using a video imaging technique. It reveals non uniform patterns of exposure that escape detection by the patch method. It also demonstrates that exposure can occur beneath protective clothing. An important advantage of this technique is that the skin serves as a collection medium rather than dosimeter patches or clothing. The main limitations of this method are the assumptions that the relative transfer of the tracer and the pesticide in the field and their permeation of the clothing are equivalent. However, these assumptions are analogous to those involving use of generic exposure databases. That is, the exposure to a pesticide measured under a given set of conditions is assumed to represent the exposure associated with a second pesticide under the same conditions. Ancillary studies can assess possible differences in the relative transfer of the tracer and the pesticide. The techniques are particularly useful for the training of operators by demonstrating the extent of their contamination, thus enabling modification of their working practices to reduce exposure. Roff (1994) developed the technique a stage further by using a dodecahedrallighting system to illuminate the contaminated worker. Measurement errors inherent in the Fenske technique caused by body surface morphology are apparently reduced. This rather large piece of equipment has been used under field conditions. If a tracer compound or visible dye is chosen, its performance and suitability as a surrogate should be validated before the field study. Apart from the usual criteria of quality control

The whole body method came into use during the late 1970s1 early 1980s (WHO, 1982; Abbot et aI., 1987). The method involves the use of clothing, usually two layers of cotton or cottonlblend material that act as the pesticide collection media. The outer layer of clothing should be representative of what the workers might wear under normal circumstances. The inner layer, usually a set of combination "long johns," represents the skin. This method overcame one of the inherent problems of the patch method, i.e., the assumption of uniformity of pesticide deposition on the skin and clothing. Exposure of the head is assessed by use of a hood or hat, preferably made of the same material. A face wipe technique can also be used, in which the skin of the face and anterior and posterior neck is wiped with cotton swabs containing a suitable detergent to remove the pesticide contaminant. Any additional protective clothing and equipment recommended for the product under study are worn over the sampling clothing, thus enabling an evaluation of their protection efficiency. The use of the whole body method overcomes the perceived problem of non uniformity of deposition. Furthermore, extrapolation from small target areas to larger body regions is not necessary. For these reasons, the method is believed to give a more accurate estimate of potential and actual dermal exposure. The whole body method can be adapted for concurrent use with biological monitoring by use of work clothing as dermal dosimeters, which represents what the workers would normally wear under the prevailing conditions. This is contrary to the view expressed by the U.S. EPA (1996). Whereas the patch and standard whole body methods place sampling media between the pesticide and the clothing or skin thus acting as a barrier interfering with the normal process of skin contamination and percutaneous absorption-the U.S. EPA's main concern-the advantage of this method is that the capture, retention, and penetration properties of the normal work clothing are mimicked as closely as possible. It is important, therefore, to have an understanding of the range of normal work clothing worn by the worker population under study. Clothing for sampling should be selected cautiously, using the minimum that might be worn under the prevailing conditions. Therefore, the assessment of residual clothing contamination and transfer to the skin

430

CHAPTER 15


beneath the clothing is as realistic as possible. This method is particularly relevant for North European and North American temperate countries where the typical work clothing consists of a "T" shirt, long sleeved shirt, socks and long trousers, and/or coveralls. Actual exposure of the skin beneath the clothing can be estimated by detemining the ratio of outer to inner clothing penetration or transfer of the pesticide. An obvious limitation is that the permeation and transfer properties of the outer and inner clothing are assumed to be the same. For analytical considerations, it may be necessary to use noncolored, white materials such as cotton or cotton/polyester mixtures. As in the standard whole body method the clothing is sectioned into individual body parts and analyzed separately to determine the regional distribution of total potential and actual exposure. Few attempts have been made to validate methods for the monitoring of dermal exposure, for example, by using biological monitoring to compare the derivations of absorbed dose. Until this is done, all methods should be viewed as providing only an approximate estimate of the dermal exposure. Dermal exposure methods will still be needed, as biological monitoring cannot be applied to all pesticides. Exposure method development is therefore a continuing need. 15.6.4 HAND EXPOSURE

Measurement of hand exposure is one of the most important aspects of a study to monitor worker exposure. The contribution of the hands to total exposure is well documented and was originally recognized in the seminal works of Batchelor and Walker (1954) and Durham and Wolfe (1962). The U.S. EPA (1996) reviewed the literature on studies that had included hand exposure measurements and concluded that its contribution to total exposure ranged from around 40 to 98%, depending upon the application method. The methods for measuring hand exposure include using lightweight absorbent gloves or sections cut from gloves and swabbing or rinsing the hands in various solvents, for example, 95% ethanol (U.S. EPA, 1996). Mild detergent solutions can be used in the hand-wash technique, for example, Aerosol OT. All these methods have advantages and limitations and it is difficult to evaluate the accuracy of any procedure. This is the reason for one of the major issues with developing generic hand exposure data bases-lack of standardization in measurement technique. It is feasible to evaluate the efficiency of recovery of a pesticide for a hand-wash or hand-rinse method under laboratory conditions using human subjects. This has been investigated by Fenske and Lu (1994) for the insecticide, chlorpyrifos, using a solvent hand-rinse method. Their findings suggest that exposure to pesticides such as chlorpyrifos that are well adsorbed to the skin cannot be estimated accurately by the hand-rinse method. However, Geno et al. (1996) showed for chlorpyrifos and pyrethrin that hand wipes with isopropanol removed in excess of 90% of the material applied to hands. Fenske et al. (1999) compared the hand exposures of orchard apple thinners

to azinphos-methyl using three methods: glove, hand-wash, and wipe. Hand exposure estimates derived from the three methods differed significantly. Based upon the hand-wash measurements and the laboratory recovery/efficiency study, it was concluded that the glove method gave a 2.4-fold overestimate whereas the wipe method gave a lO-fold underestimate. The authors also concluded that methods should be validated and standardized to enable the development of more accurate hand exposure estimates. The generally held view is that the use of gloves results in a significant overestimation of total dermal exposure, owing to the retention of more of the pesticide than would otherwise be retained by the skin. Gloves also contain foreign materials such as sizing, which may be coextracted with the pesticide. At low levels of contamination this may cause analytical difficulties. However, glove contamination with dirt and grease arising from the worker's activities are a more likely cause of analytical problems. Washing the hands with a solvent such as ethanol might cause skin damage or disrupt skin barrier function and enhance percutaneous absorption of the pesticide. In the many studies that have used this technique there has been little evidence that these effects occurred. In studies involving concurrent passive dosimetry and biological monitoring, a hand-washing procedure involving standardized detergent and water can be used. This procedure is identical to that described earlier as the hand-wash method, except that the measurement is taken only when the workers would normally wash their hands. The basis is that when the total absorbed dose is determined with biological monitoring there should be no interference with the normal process of dermal contamination and percutaneous absorption. Therefore the use of gloves or a solvent washing or rinsing technique is inappropriate, because these methods would retain pesticide otherwise available for absorption or potentially disrupt the barrier function of the skin. Inevitably there is a loss of standardization of the intervals at which samples are taken. However, it does give some information on the extent of hand exposure that might be of value in overall data interpretation. 15.6.5 INHALATION EXPOSURE

Exposure by inhalation is usually a minor route of absorption in comparison with the dermal route. There are exceptions to this, for example, when dusts, fine aerosols, and fumigants are applied or when materials are applied indoors. The U.S. EPA (1996) reviewed several exposure studies and found that the inhalation route contributed negligible amounts (nondetectable) to about 9% of total exposure. In most cases, the contribution was less than 1%. The extent of the contribution depends upon the method of application, whether used outdoors or indoors and on factors such as the volatility of the pesticide. Significantly, for pesticides that are poorly absorbed via the skin, the inhalation route can become the most important route of absorption.

15.7 Methods to Measure the Absorbed Dose

431

The important reviews of methodology for field monitor- particulates and vapor can be achieved by mounting the filter ing of airborne pesticides were provided by Van Dyk and sampling head in front of the vapor trap in a "sampling train." Visweswariah (1975) and Lewis (1976). The former reviewed This train allows retention of any vapor stripped off the filter the sampling media available for collection of pesticides but on the resin. The material on the filter can be analyzed both with particular emphasis on static environmental sampling gravimetrically and/or chemically and an estimate made of the rather than personal sampling. pesticide content of the particulate sample. If use of such a A personal air sampling method is the most appropriate for sampling train is needed, laboratory validation of the sampling the determination of potential inhalation exposure of workers. efficacy, particularly of the adsorbent resin, is necessary owing Several techniques are available such as gauze pads in place of to the possibility of stripping material from the resin by the relfilters in respirators for agricultural use, pioneered by Durham atively high flow rate of 2 lImin. The measurement of the inhalable fraction with use of adsorand Wolfe (1962), midget impingers, solid adsorbents, and filter cassettes attached to battery-powered personal sampling bent resins for the vapor phase is recommended as the method pumps. A personal sampling technique involving sampling de- of first choice. vices located in the breathing zone and sampling pumps is preferred for reasons of practicability and representativeness. 15.7 METHODS TO MEASURE THE Breathing rates for the calculation of inhalation exposure from airborne concentration data can be obtained from standard refABSORBED DOSE erence texts such as the D.S. EPA's Exposure Factors Handbook (1999). 15.7.1 BIOLOGICAL MONITORING The advantage of using the modified respirator is that the subject produces the airflow so that breathing rate and total vol- Biological monitoring of pesticide workers was first used as a ume of air inhaled do not have to be estimated. However, the means of assessing health effects or modification of biochemgauze pads must be capable of trapping the pesticide efficiently. ical parameters as a consequence of exposure to organophosThe respirator must also fit properly to the face. Perhaps the phorus compounds by measurement of plasma cholinesterase main disadvantage of this technique is that the subject must levels (for example, Peoples and Knaak, 1982). This type of aswear the respirator for the duration of the monitoring period, sessment, used in the chemical industry for many years, can be which ideally should be a complete working day, which may termed biological effect monitoring and must be distinguished cause discomfort. from the type of monitoring that determines the absorption of Midget impingers, which traditionally have used ethylene chemicals by measuring the chemical or its metabolites in body glycol as the trapping medium, have a long history of use in fluids, usually urine, blood, or exhaled breath. measuring agricultural worker inhalation exposure. The techAnalysis of body fluids and excreta, usually urine, for parent nique suffers from the major drawbacks of spillage of the trap- compound or metabolites can provide both a qualitative and a ping liquid and inefficient trapping and retention of some pes- quantitative measurement of absorbed dose for pesticides that ticides. Microimpingers have been developed, which overcome lend themselves to this form of monitoring. The technique has the first of these drawbacks. a distinct advantage over passive dosimetry because it evaluates Personal air samplers allow the use of respirators or dust actual, rather then potential, absorption. It integrates absorption masks for protection, if required by the product label. How- from all routes of exposure: dermal, inhalation, and primary and ever, they do not measure the true exposure of workers wearing secondary oral ingestion. However, it is difficult to differentiate respiratory protection. The choice of sampling medium is de- the contributions to the absorbed dose from different aspects termined by the nature of the pesticide. A filter cassette or of the work procedures or to distinguish between the relative sampling head should be used for spray particulates and a solid contributions of the different routes of exposure to the total adsorbent material for volatile compounds. absorbed dose. Van Heemstra-Lequin and Van Sittert (1986), The inhalable fraction (all material capable of being drawn Wang et al. (1989), and Henderson et al. (1993) reviewed biinto the nose and mouth) is the most biologically relevant frac- ological monitoring in the context of pesticides. The OECD tion to measure. An example of a suitable device is the Institute (1997) provided detailed guidance on how to conduct biologof Occupational Medicine personal sampling head designed ical monitoring studies. specifically to collect this fraction (Vincent and Mark, 1987). Early biological monitoring studies on pesticides were able For use of this device, a sampling flow rate of 2 lImin is a spe- to demonstrate absorption without quantifying the amount of cific requirement. There is now a commercially available device absorption. For example, Durham and Wolfe (1962) measured manufactured by SKC that collects, as it separate fractions, the p-nitrophenol, a metabolite of parathion, in the urine of workinhalable and respirable components. ers; Swan (1969) measured paraquat in the urine of backpack Examples of suitable adsorbent materials for some volatile spray operators. The absorbed dose of a pesticide can only compounds are activated charcoal and Tenax and XAD-2 resins be quantified accurately if the metabolism and pharmacokimounted in stainless steel or glass tubes. The choice of ma- netics of the compound are understood, ideally from human terial should be determined by analytical retention (trapping studies. Woollen (1993) reviewed the specific requirements for efficiency) and extractability studies. Concurrent sampling for this type of biological monitoring study. Studies that have met

432

CHAPTER 15


these requirements include those on the herbicide fluazifopbutyl (Chester and Hart, 1986) and 2,4-dichlorophenoxyacetic acid amine (Grover et aI., 1986; Ritter and Franklin, 1989). Data from human dosing studies facilitate the design of a field sampling strategy and secondly define the body fluid matrix of choice. Ideally, urine is the matrix of choice as its collection is noninvasive and the collection of 24-hour urine samples is practicable. Complete 24-hour urine collections are essential, and this can be checked in a number of ways, for example, by measuring the concentration of creatinine. Substantially incomplete collections are readily apparent and these samples are either excluded or an allowance is made, for example, by use of a correction factor based upon the average daily urine volume for the individual concerned (Woollen, 1993). Specific gravity and osmalarity are alternative means of checking for completeness of urine collection (Allesio et aI., 1985). Unstable or highly volatile pesticides are not good candidates for passive dosimetry, despite the efforts to accurately assess field, storage, and transit losses. Biological monitoring should be considered for these pesticides and may be the only means of obtaining adequate quantitative data from which the absorbed dose can be derived. An example of the successful use of biomonitoring to estimate exposure occurring primarily by the inhalation route is urinary monitoring for a metabolite of the fumigant 1,3-dichloropropene (Osterloh et al., 1989; Van Welie et aI., 1991). If a human metabolism study were impracticable, then animal metabolism data might be used, if metabolism and excretion kinetics are similar in several animal species, then it could be assumed that humans will metabolize and excrete the compound in a similar manner. This carries a degree of uncertainty. There are examples, described in detail by Woollen (1993), that demonstrate the limitations ofthis approach. Apart from interspecies differences in metabolism, there is the possibility of dose-dependent differences, which might necessitate metabolism studies in animals and humans at doses similar to the anticipated worker exposures. Pesticides that are extensively metabolized to a large number of metabolites are not good candidates for biological monitoring. The absorbed dose cannot be determined accurately using data on a minor metabolite, particularly if there is wide interindividual variation in the proportion of the parent compound excreted as this metabolite. However, a minor metabolite might provide some useful information as a "biological indicator" in the absence of more abundant metabolites. Overall, it can be concluded that if the requisite human metabolism data are available for a pesticide, biological monitoring provides the most accurate means of estimating the absorbed dose for quantitative risk assessment.

REFERENCES Abbott, 1. M., Bonsall, J. L., Chester, G., Hart, T. B., and Turnbull, G. J. (1987). Worker exposure to a herbicide applied with ground sprayers in the United Kingdom. Am. Ind. Hyg. Assoc. 1. 48,167-175.

Alessio, L., Bolin, A., Dell'Orto, A., Toffoletto, E, and Ghezzio, I. (1985). Reliability of urinary creatinine as a parameter used to adjust values of urinary biological indicators. Int. Arch. Occup. Environ. Health 55, 99-106. American Conference of Government Industrial Hygienists (ACGIH)fTechnical Committee on Air Sampling Procedure (1985). "Particle Size Selective Sampling in the Workplace." ACGIH, Cincinnati. Batchelor, G. S., and Walker, K. C. (1954). Health hazards involved in the use of parathion in fruit orchards of North Central Washington. Am. Med. Assoc. Arch. Ind. Hyg. 10,522-529. Chester, G. (1988). Pesticide applicator exposure-towards a predictive model for the assessment of hazard. Aspects Appl. Bio!. 18,331-343. Chester, G. (1993). Evaluation of agricultural worker exposure to, and absorption of pesticides. Ann. Occup. Hyg. 37, 509-523. Chester, G., and Hart, T. B. (1986). Biological monitoring of a herbicide applied through backpack and vehicle sprayers. Toxicol. Lett. 33, 137-149. Curry, P., and Iyengar, S. (1992). Comparison of exposure assessment guidelines for pesticides. Rev. Environ. Contam. Toxico!. 129, 79-93. Davis, J. E. (1980). Minimizing occupational exposure to pesticides: personal monitoring. Residue Rev. 75, 35-50. Durham, W. E, and Wolfe, H.T. (1962). Measurement of the exposure of workers to pesticides. Bull. WHO 26, 75-91. Fenske, R. A. (1990). Non-uniform dermal deposition patterns during occupational exposure to pesticides. Arch. Environ. Contam. Toxicol. 19,332-337. Fenske, R. A., Leffingwell, J. T., and Spear, R. C. (1986a). A video imaging technique for assessing dermal exposure-I. Instrument design and testing. Am. Ind. Hyg. Assoc. 1. 47, 764-770. Fenske, R. A., and Lu, C. (1994). Determination of handwash removal efficiency; incomplete removal of the pesticide chlorpyrifos from skin by standard handwash techniques. Am. Ind. Hyg. Assoc. 1. 55, 425-432. Fenske, R. A., Simcox, N. J., Camp, J. E., and Hines, C. J. (1999). Comparison of three methods for assessment of hand exposure to azinphos-methyl (Guthion) during apple thinning. App. Occup. Environ. Hyg. 14,618-623. Fenske, R. A., Wong, S. M., Leffingwell, J. T., and Spear, R. C. (1986b). A video imaging technique for assessing dermal exposure-H. Fluorescent tracer testing. Am. Ind. Hyg. Assoc. 1. 47, 771-775. Geno, P. w., Camann, D. E., Harding, H. J., Villabos, K., and Lewis, R. G. (1996). Handwipe sampling and analysis procedure for the measurement of dermal contact with pesticides. Arch. Environ. Contam. Toxicol. 30, 132138. Groupement International des Associations NationaIes de Fabricants de Produits Agrochemiques (GIFAP). (1990). "Monitoring Studies in the Assessment of Field Worker Exposure to Pesticides," Technical Monograph No. 14. GIFAP, Brussels. Grover, R., Franklin, C. A., Muir, N. I., Cessna, A. J., and Riedel, D. (1986). Dermal exposure and urinary metabolite excretion in farmers repeatedly exposed to 2,4-D amine. Toxicol. Lett. 33, 73-83. Henderson, P. Th., Brouwer, D. H., Opdam, J. J. G., Stevenson, H., and Stouten, J. Th. J. (1993). Proceedings of workshop on: risk assessment for worker exposure to agricultural pesticides. Ann. Occup. Hyg. 37, 499-507. International Organization for Standardization (ISO). (1995). "Air QualityParticle Size Fraction Definitions for Health-Related Sampling." ISO 7708: 1995(E). Kromhout, H., Symanski, E., and Rappaport, S. M. (1993). A comprehensive evaluation of within and between worker components of occupational exposure to chemical agents. Ann. Occup. Hyg. 37, 253-270. Lewis, R. G. (1976). Sampling and analysis of airborne pesticides. In "Air Pollution from Pesticides and Agricultural Processes" (R. E. Lee, Jr., ed.). CRC Press, Cleveland. Mull, R., and McCarthy, J. E (1986). Guidelines for conducting mixer-Ioaderapplicator studies. Vet. Hum. Toxicol. 28, 328-336. National Agricultural Chemicals Association (NACA). (1985). "Guidelines for Conducting Biological Monitoring-Applicator Exposure Studies." NACA, Washington, DC.

References

Osterloh, J. D., Wang, R., Schneider, F., and Maddy, K. (1989). Biological monitoring of dichloropropene: air concentrations, urinary metabolite, and renal enzyme excretion. Arch. Environ. Health 44, 207-213. OECD (1997). Guidance Document for the Conduct of Studies of Occupational Exposure to Pesticides During Agricultural Application. OECD Environmental Health and Safety Publications Series on Testing and Assessment No. 9. Environment Directorate, OECD Paris. Peoples, S. A., and Knaak, J. (1982). Monitoring pesticide blood cholinesterase and analysing blood and urine for pesticides and their metabolites. In "Pesticide Residues and Exposure" (J. R. Plimmer, ed.), ACS Symposium Series, Vol. 182, pp. 41-57. Am. Chem. Soc., Washington, DC. Rappaport, S. M. (1991). Assessment of long-term exposures to toxic substances in air. Ann. Occup. Hyg. 35, 61-121. Ritter, L., and Franklin, C. A. (1989). Use of biological monitoring in the regulatory process. In "Biological Monitoring for Pesticide Exposure" (R. G. M. Wang, C. A. Franklin, R. C. Honeycutt, and J. C. Reinert, eds.), ACS Symposium Series, Vol. 382, pp. 354-367. Am. Chem. Soc., Washington, DC. Roff, M. W. (1994). A novel lighting system for the measurement of dermal exposure using a fluorescent dye and an image processor. Ann. Occup. Hyg. 38,903-919. Ross, J. H., Driver, J. H., Cochran, R. C., Thongsinthusak, T., and Krieger, R. 1. (2000). Could pesticide toxicology studies be more relevant to occupational risk assessment? 1. Occup. Hyg. 45(Suppll), 5-17. Swan, A. A. B. (1969). Exposure of spray workers to paraquat. Br. 1. Ind. Med. 26,322-329. Thongsinthusak, T., Ross, J. H., and Meinders, D. (1993). "Guidance for the Preparation of Human Pesticide Exposure Documents, HS-1612, May 4." California Environmental Protection Agency, Worker Health and Safety Branch. U.S. Environmental Protection Agency. (1987). Pesticide Assessment Guidelines, Subdivision U, Applicator Exposure Monitoring. U.S. EPA, Washington, DC.

433

U.S. Environmental Protection Agency. (1996). Occupational and Residential Exposure Test Guidelines, OPPTS 875.1000, EPA 712-C-96-261. U.S. EPA, Washington, DC. U.S. Environmental Protection Agency. (1999). Exposure Factors Handbook, EPN600/C-99/001, February. U.S. EPA, Office of Research and Development, Washington, DC. Van Dyk, L. P., and Visweswariah, K. (1975). Pesticides in air: sampling methods. Residue Rev. 55, 91-134. Van Heemstra-Lequin, E. A. H., and Van Sittert, N. J., eds. (1986). Biological monitoring of workers manufacturing, formulating and applying pesticides. Toxicol. Lett. 33, 1-236. Van Welie, R. T., van Duyn, P., Brouwer, D. H., van Hemmen, J. J., Brouwer, E. J., and Vermuelen, N. P. (1991). Inhalation exposure to 1,3dichloropropene in the Dutch flower-bulb culture. Part n. Biological monitoring by measurement of urinary excretion of two mercapturic acid metabolites. Arch. Environ. Contam. Toxicol. 20, 6-12. Vincent, J. H., and Mark, D. (1987). Comparison of criteria for defining inspirable aerosol and the development of appropriate samplers. Am. Ind. Hyg. Assoc. 1. 48, 451-457. Wang, R. G. M., Franklin, C. A., Honeycutt, R. c., and Reinert, J. C. (1989). "Biological Monitoring for Pesticide Exposure: Measurement, Estimation and Risk Reduction," ACS Symposium Series, Vol. 382. Am. Chem. Soc., Washington, DC. Wolfe, H. R. (1976). Field exposure to airborne pesticides. In: "Air Pollution from Pesticides and Agricultural Processes" (R. E. Lee, Jr. ed.). CRC Press, Cleveland, Ohio. Woollen, B. H. (1993) Biological monitoring for pesticide absorption. Ann. Occup. Hyg. 37, 525-540. World Health Organization. (1975). "Survey of Exposure to Organophosphorus Pesticides in Agriculture," Standard Protocol, VBCI75.9. WHO, Geneva. World Health Organization. (1982). "Field Surveys of Exposure to Pesticides," Standard Protocol, VBC/82.1. WHO, Geneva.

CHAPTER

16 Residential Exposure Assessment: An Overview Jeffrey H. Driver, John H. Ross, Muhilan D. Pandian infoscientific.com & risksciences.net

Jeffrey B. Evans D.S. Environmental Protection Agency, Office of Pesticide Programs

Gary K. Whitmyre risksciences, LLC

16.1 INTRODUCTION Following the use of products in and around the home, postapplication chemical exposures to consumers may occur in a variety of microenvironments that correspond to the daily activities in which adults and children engage. These activity patterns may place individuals in contact with a variety of chemicals including pesticides (e.g., dislodgeable foliar residue exposures during gardening, lawn chemical exposures after reentry onto treated turf; and chemical emissions from treated surfaces inside the residence). To understand the potential health significance of these exposures it is necessary to characterize their sources and estimate their magnitude. In response to these needs, efforts have been undertaken to develop methodologies for quantifying pesticide and other chemical exposures in soil, air, food, and water (Cal-EPA, 1994; McKone, 1991, 1993; Ott, 1985; Thompson et aI., 1984; Vaccaro et al., 1996; Wallace, 1987, 1989, 1990, 1991; Wall ace et aI., 1982, 1984, 1985, 1986, 1987a, b, c, 1988, 1989, 1991a, b). The U.S. Environmental Protection Agency (U.S. EPA), for example, in response to the the Food Quality Protection Act of 1996 (FQPA; http://www.epa.gov/docs/oppfeadsllfqpa),has been revising exposure monitoring guidelines that emphasize nonoccupational, residential exposure to pesticides; these guidelines are referred to as "Series 875, Occupational and Residential Exposure Test Guidelines Group B: Post-Application Monitoring Test Guidelines" (http://www.epa.gov/docs/opptsfrs/OPPTS_ Harmonized/87 5_Occupational_and_ResidentiaLExposure_ TesCGuidelines; Whitford et al., 1999). The series 875 guidelines provide information and protocols relevant for persons required to submit postapplication exposure data under 40 CFR 158.390 (Fig. 16.1). Generally, these data are required under the Handbook of Pesticide Toxicology Volume 1. Principles

Federal Insecticide, Fungicide and Rodenticide Act (FIFRA) when certain toxicity and/or exposure criteria have been met for a given pesticide (Driver and Wilkinson, 1996). Although at a low level relative to occupational exposures, the major source of chemical exposures for the general population, appears to result from the use of products in and around the home (Hill et aI., 1995; U.S. EPA, 1999a; Whitmore et aI., 1994). For example, the National Academy of Sciences Committee on Urban Pest Management noted that 5000 health-related incidents involving pesticides were reported as occurring in homes in the United States from 1966 to 1979 (NRC, 1980). More recent data regarding pesticide healthrelated incidents can be obtained from the American Association of Poison Control Centers (http://www.aapcc.org), the U.S. EPA (http://www.epa.gov/pesticides), or state regulatory agencies, (e.g., California Department of Pesticide Programs (http://www.cdpr.ca.gov). It is perhaps not surprising, therefore, that the potential health risks associated with exposure to chemicals such as pesticides occurring in and around the home (in air and from surfaces) and from other sources (e.g., consumer products, combustion appliances, and outdoors) are being evaluated much more carefully now than in the past (Driver and Wilkinson, 1996). Pesticides, of course, are just one of many types of chemicals to which humans are exposed in the home. During the past decade and a half, a number of studies, most notably the Total Exposure Assessment Methodology (TEAM) studies sponsored by the EPA, have demonstrated that, for a variety of contaminants, indoor air and other residential pathways are often a more significant source of exposure than corresponding outdoor pathways (Curry et aI., 1994; Furtaw et aI., 1993; Pellizzari et aI.,

435


436

CHAPTER 16

Residential Exposure Assessment: An Overview

Dissipation Studies Dislodgeable Foliar Residue (DFR) Dissipation Study Soil Residue Dissipation (SDR) Study Indoor Surface Residue (ISR) Dissipation Study

SOURCES

Measurements ofHuman Exposure Dermal Exposure (passive dosimetry) Inhalation Exposure Biological Monitoring Other Relevant Data Descriptions of Human Activity Data Data Reporting and Calculations Detailed Product Use Information

ACTIVITIESI PATHWAYSI ROUTES

ApplicatIon

Post-Application

//j

Figure 16.1 U.S. EPNOPPTS Series 875, Group B: description of required studies (adapted from ILSI, 1998). RECEPTORS

1987,1993; Thomas et aI., 1993; Wallace, 1993). Indeed, some of the studies found that indoor concentrations of some chemicals were higher than those outdoors and raised serious questions about the relative contribution of indoor sources to the total exposure (Dockery and Spengler, 1981; Melia et al. , 1978 ; Ott, 1985; Spengler et al., 1983). The assessment of potential indoor exposures has also been recognized by industry as a key component in the overall risk evaluation for consumer products (Hakkinen et al., 1991; Hakkinen, 1993). Several studies of potential indoor air exposures to chemicals have been reported to confirm the safety of various consumer products (Gibson et al., 1991; Hendricks, 1970; Wooley et aI., 1990).

16.2 OVERVIEW OF GENERAL ISSUES The residential environment should be considered in very dynamic terms. Chemicals that are released into or otherwise enter the residential environment tend to partition into various compartments, either through direct dispersion in indoor air or through adsorption onto surfaces that serve as "sinks" from which material can subsequently be released into the air (Ross et aI., 1990, 1991). The amount of a given chemical in each compartment is depleted over time by various mechanisms, including air exchange with the outdoors or with other rooms of the house, and by chemical or physical transformation and/or degradation. Simulation of the behavior of pesticides (and other chemicals) in residential environments can be modeled using the principles of fugacity (Matoba, 1996). Fugacity employs the unit of pressure (Pa), which refers to the external force of a chemical escaping from one compartment or medium to another. For example, there is evidence that particulate contaminants, whether generated inside the residence or brought in from outdoors are adsorbed to surfaces and are later resuspended and recycled within the house after a disturbance (e.g., walking on floors and rugs, sweeping and dusting, and vacuuming ; Roberts et al., 1992). A simplistic depiction of the relationship between potential sources and exposure pathways in the context of a residential exposure assessment is illustrated in Fig. 16.2. Thus, the residence can be considered an exposure unit containing multiple compartments with which the human

Figure 16.2 1998).

Shematic diagram for nondietary exposure (adapted from ILSr,

receptor can contact. Figure 16.3 illustrates the components of a residential exposure model and Fig. 16.4 illustrates the decision logic associated with construction of a residential exposure assessment. Although inhalation exposure and indoor air quality have received the most attention to date, there are a number of noninhalation exposure pathways that are likely to be of equal or greater importance for human residential exposures to pesticides and other chemicals. These include potential dermal exposure to dislodgeable chemical residues from surfaces such as floors and carpets or from hard surfaces resulting from the use of formulations for cleaning and disinfection and potential ingestion of surface contaminants resulting from hand/objectto-mouth activity, particularly in infants and toddlers. Several studies and/or reviews provide examples of noninhalation residential exposures and the complexities involved (Calvin, 1992; CTFA, 1983; Driver et aI., 1989; Eberhart, 1994; ECETOC, SOURCE CHARACTERISTICS ~ PRODUCT USE PATTERNS & PLAUSIBLE SCENARIOS

~

HUMAN ACTIVITY PATTERNS POPULATION DEMOGRAPHIC:-:::-& STRATIFICATION

NON-DIETARY EXPOSURE MODEL HUMAN EXPOSURE FACTORS _ _ PHYSICO-CHEMICALPROPERTIES RESIDENTIAL BUILDING

~ ~

FACTOR~

TEMPORAL & SPATIAL DOMAINS UNCERTAINTY ANALYSIS MODEL VALIDATION

Figure 16.3 Residential exposure assessment: key components (adapted from ILSI, 1998).

16.3 Lessons Learned from Key Studies Hazard Identification and Toxicity Endpoint Selection (e.g., toxicological benchmarks; time-to-effect; exposure/dose metric selection; relevant subpopulations, etc.)

1

Evaluation of Consumer and Commercial Product Use Information (e.g., application methods, frequencies, locations, rates, formulation types, active ingredient concentrations, post-application activities, consumerlprofessional survey data, market share data, etc.)

1

Evaluation of Exposure Monitoring Data (for actual pesticide and/or relevant surrogate chemicals; data limitations; uncertainty & variability; data quality objectives)

1

Determination of Relevant (Plausible) Aggregate Exposure Scenarios and Related Routes and Pathways of Exposure

~ Development and Implementation of Screening-Level Deterministic and/or Stochastic

Exposure~ Model Validation & Refinement

Figure 16.4 Decision logic and model development/validation process (adapted from ILSI, 1998).

1994; Harris and Solomon, 1992; Harris et aI., 1992; Turnbull and Rodricks, 1989; Vermiere et aI., 1993).

16.3 LESSONS LEARNED FROM KEY STUDIES Pesticides applied in and around homes by both professional applicators and consumers are used in different ways for different purposes: (1) indoor uses (e.g., floor sprays or foggers for fleas) and outdoor uses (e.g., treatment of wasp nests and ant mounds; use of antimicrobial products in swimming pools); (2) turf uses (e.g., granular applications for control of soildwelling insect pests, preemergent and postemergent herbicide sprays) and ornamental uses (e.g., foliar sprays for shrubs); (3) home garden uses (e.g., fungicide dusts for tomatoes); and (4) structural pest control uses (e.g., structural treatment or insecticidal soil barriers to protect against termite invasion). The vast majority of V.S. households use pesticides (Whitmore et aI., 1992) and these uses undoubtedly present many opportunities for exposure during intended, label-directed use, misuse, and accidents (Whitmyre et aI., 1996). Other sources of indoor exposure to pesticides for the general population may be from ambient air, food, water, ambient particles and indoor house dust (Jenkins et al., 1992; Pellizzari et aI., 1993; Wallace, 1991,1993; Whitmore et aI., 1994). Residential pesticide monitoring studies have included general surveys of many different pesticides and measurements of air and surface concentrations of specific pesticides after applications of termiticides, crack and crevice or baseboard treatments, total release aerosols or foggers, broadcast applications, and hand-held sprays (Fenske et aI., 1990; Racke and Leslie, 1993; Whitmore et aI., 1994). These studies typically

437

demonstrate that, after pesticide use, measurable though relatively low-level residues exist in homes and that indoor exposures are often higher than outdoor exposures. Although, in most cases, such exposures are associated with negligible health risks (Whitmore et aI., 1994), potential residential exposures to infants and children continue to be the subject of debate and scientific investigation (Berteau et aI., 1989; Byrne et aI., 1998; Gibson et aI., 1998; NRC, 1993; Ross et aI., 1990, 1991; Vaccaro et aI., 1996; Zweiner and Ginsburg, 1988). However, pesticide biomonitoring of both adults and children demonstrate that absorbed doses from all sources range from fractions of micrograms to single digit micro grams per kilogram of body weight (Adgate et aI., 1998; Hill et aI., 1995; Krieger et aI., 2000, 2001; Vaccaro et aI., 1996). The Nonoccupational Pesticide Exposure Study involving about 250 residents of Jacksonville, Florida and Springfield, Massachusetts, clearly demonstrated measurable levels of indoor exposure. Participants in this study carried personal monitors for 24 hours that provided indoor and outdoor measurements of 32 common household pesticides and structural termiticides (Immerman and Schaum, 1990). The indoor air concentrations of these materials exceeded the outdoor air levels by factors even larger than those measured in the earlier TEAM studies on volatile organic compounds (VOCs). It was hypothesized that, with termiticides and other chemicals used outside the home, some material entered the house on soil particles in addition to infiltration in air from treated areas beneath and around the house. It was also noted that the use of "walkoff" rugs in hallways and the practice of removing shoes on entering the house would help to reduce indoor exposure levels (Roberts et aI., 1992). In the TEAM studies mentioned earlier, median personal concentrations of VOCs in V.S. residences were found to be 2-5 times outdoor levels and maximum personal concentrations were 5-70 times the highest outdoor levels (Wallace, 1993). The variability in indoor personal exposures probably reflects differences in human activity patterns that bring individuals into contact with chemicals indoors and suggests the importance of specific sources of residential exposures that may not be available to all individuals. Smokers, for example, had benzene exposures 6-10 times higher than those of nonsmokers, individuals wearing freshly dry cleaned clothes had significantly higher exposures to tetrachloroethylene (Wallace et aI., 1991 c), and persons using mothballs and solid deodorizers in the residence had greatly elevated exposures to p-dichlorobenzene relative to those of nonusers (Wallace, 1993). The most recent TEAM study, known as PTEAM, focused on measuring personal exposures to inhalable particles (PMIO) of approximately 200 Riverside, California, residents using specially designed indoor sampling devices. A major finding from this study was that personal exposures to particles in the daytime were 50% greater than either indoor or outdoor concentrations (Wallace, 1993). It has been hypothesized that these data suggest that individuals are exposed to a "personal cloud" of particles as they go about their daily activities (Wallace, 1993). Resuspension of household dust through walking in the resi-

438

CHAPTER 16


dence or by vacuuming, cooking, or sharing a home with a smoker lead to significant particle exposures. The recent Valdez Air Health Study conducted in Valdez, Alaska (Wallace, 1993) generally supports the findings of the TEAM studies in terms of the greater importance of individual indoor exposure sources than outdoor sources. In the Valdez study, mean personal concentrations of benzene were roughly 3-4 times higher than outdoor levels, despite the presence of a significant outdoor source of benzene in the community (i.e., a petroleum storage and loading terminal).

16.4 GUIDANCE FOR RESIDENTIAL POSTAPPLICATION EXPOSURE ASSESSMENT METHODS AND DATA SOURCES FOR EXPOSURE FACTORS The most recent effort to develop guidance for residential exposure assessment methods was initiated by the EPA's Office of Pesticide Programs in the Standard Operating Procedures for Residential Exposure Assessment (U.S. EPA, 1999a). The passage of the FQPA mandated the EPA to immediately begin routinely addressing nondietary and non-occupational pesticide exposures for the general population. These are exposures that can occur in a residential setting (or other areas frequented by the general population) and that do not occur as part of the diet or as a result of participation in occupational practices. These exposures may include breathing vapors while inside a treated home, exposures to children playing on a treated lawn, or exposures attributable to the mouthing behaviors of infants and children. Before passage of the FQPA, the Agency addressed these kinds of exposures on a case-by-case basis, typically in the "special review" process. The intent of the EPA SOPs is to provide a means for consistently calculating single pathway, screening level exposures and not to provide guidance on other related topics such as aggregate (multisource to a given pesticide) or cumulative (multi source to two or more pesticides with a presumed common mode of action) exposure assessment. These SOPs are the backbone of the Agency's current approach for completing initial tier (screening-level) residential exposure assessments. However, the state-of-the-science continues to evolve since the release of the original document in 1997 and the emphasis of industry, as well as of academia and others, has clearly focused on the scientific and policy issues raised by the implementation of the FQPA and the use of the first-generation SOPs. Thus, revisions to the SOPs are ongoing to reflect the development of scientific information and the development of refined methods for estimation of potential residential exposures to adults and children. Additional guidance for dermal exposure assessment methods and dermal permeability coefficients for some organic chemicals are contained in the EPA's dermal exposure assessment guidance document (U.S. EPA, 1992). Given that skin surface area and body weight are closely correlated, total skin surface area to body weight ratios for use in residential exposure assessments have been recommended (Phillips et aI.,

1993). Another excellent source for methodology and data relevant to consumer product exposure assessments is ECETOC (1994). A number of relevant data sources exist for key variables or factors used in performing residential exposure assessment. Data useful in estimating human exposures (e.g., distributions of body weights and skin surface areas, inhalation rates, and residential occupancy periods) can be obtained from the American Industrial Health Council's Exposure Factors Sourcebook l (AIHC, 1995) and the EPA's Exposure Factors Handbook (U.S. EPA, 1999b), which has recently been updated. Residential "environmental factors" such as air exchange rates have been summarized by Pandian et al. (1993). Human time-activity data in the United States were summarized by the EPA (U.S. EPA, 1991) and compiled in the THERdbASE software (Pandian and Furtaw, 1995), which is available on the Internet at http://www.therd.com. Multiple data sources for time-activity data have been included in EPA's Consolidated Human Activity Database, which is planned for future release via the Internet.

16.5 RESEARCH NEEDS Given that the potential for postapplication exposures largely exists because of product use in and around the home, the need to develop and validate models for prediction of multipathway, multiroute exposures and absorbed dose is evident. Historically, efforts have focused on indoor air and associated inhalation exposures. Jayjock and Hawkins (1993), for example, have explored the complementary roles of indoor air mode ling and data development in improving the level of confidence in estimates of indoor inhalation exposures. More recently, dermal and incidental ingestion exposures have been the focus of monitoring and modeling efforts (e.g., the Outdoor Residential Exposure Task Force, the Non-Dietary Exposure Task Force, the OP Case Study Group, and Residential Exposure Joint Venture (Zartarian and Leckie, 1998; Zartarian et aI., 2000). Multipathway, multiroute modeling efforts for pesticides include the Residential Exposure Assessment Model (REAM), the Stochastic Human Exposure and Dose Simulation model (SHEDS); the Cumulative and Aggregate Risk Evaluation System (CARES), and LifeLine (U.S. EPA, 1999a). The use of real-world data to validate residential exposure models is critical to developing estimates that are more representative than worst-case estimates typically obtained from unvalidated modeling approaches (Whitmyre et aI., 1992a, b). Other research activities related to residential exposure assessment currently being sponsored by the EPA include the National Human Exposure Assessment Survey. In addition, the EPA has recently concluded a cooperative agreement, referred to as the Residential Exposure Assessment Project (REAP) with the Society for Risk Analysis and the International Society of Exposure Analysis, to develop a reference textbook describing 1For the latest version of the Exposure Factors Sourcebook, contact the Update Coordinator, American Industrial Health Council, Suite 760, 2001 Pennsylvania Avenue, N.W., Washington, DC 20006-1807; phone: 202-833-2131.

References

relevant methodologies, data sources, and research needs for residential exposure assessment. The REAP will complement other EPA initiatives, such as the development of the series 875 guidelines and will facilitate a sharing of information and other resources between the EPA, other federal and state agencies, industry, academia, and other interested parties. Residential exposures to pesticides and other chemicals are estimated by means of either monitoring and/or predictive modeling but, unfortunately, little or no guidance is available for those attempting such estimates. Key areas requiring attention include: • Characterization of temporal product use patterns (particularly the likelihood of co-occurrence of more than one product use event) and associated demographic and postapplication activity information relevant to occupants of homes using products. • Source characterization, including emission rates, surface deposition, tranferability to human clothing and skin, and physicochemical factors driving fate and transport processes. • The complex interaction over time of environmental media residue concentrations with humans resulting from variable time-activity patterns that determine subsequent residential exposures (inhalation, dermal, and incidental ingestion). • Identification of the fundamental principles, concepts, and methods for conducting multipathway/multiroute residential exposure assessments, including unique pathways such as incidental dermal exposure to dislodgeable pesticide residues from treated lawns, incidental ingestion of contaminated soil particles during gardening, hand-to-mouth transfer by infants and children, and dermal exposure to dislodge able pesticide residues from carpets and other treated surfaces, and incidental ingestion of postapplication residues in food. • Characterization of key human exposure factors (ranges and distributions of factors such as age-specific inhalation rates, product use patterns, and human time-activity data) and residential building factors (distributional data on housing stock type, number and size of rooms, air exchange rates, source emission rates, and sink effects, i.e., adsorption! desorption from various surfaces in the home) that influence residential exposure and dosimetry. • Continued development and validation of methods for measuring and mode ling indoor chemical fate processes (e.g., volatilization from surfaces and dislodgeable residue kinetics), chemical concentrations in complex matricies (such as house dust) and human intake (e.g., incidental ingestion, inhalation, and dermal exposure). • Development and validation of methods for extrapolating from short-term monitoring data to long-term exposure scenarios and for extrapolation of adult monitoring data to children. • Continued development and application of methods for quantifying uncertainty and variability (e.g., Monte Carlo methods) in residential exposure (and risk) estimates.

439

• The development and use of effective methods for comparing and communicating residential exposure and risk estimates to risk managers and the general public.

REFERENCES Adgate, J., Quackenboss, J., Needham, L., Pellizari, P., Lioy, P., Shubat, P., and Sexton, K. (1998). Comparison of urban versus rural pesticide exposure in Minnesota children. In "Annual Conference of International Society for Environmental Epidemiology (ISEE) and International Conference for Society of Exposure Analysis (ISEA)," July 1998, Vol. 9, No. 4, Suppl., Abstract 920. American Industrial Health Council (AIHC). (1995). "Exposure Factors Sourcebook." AIHC, Washington, DC. Berteau, P. E., Knaak, J. B., Mengle, D. C., and Schreider, J. B. (1989). Insecticide absorption from indoor surfaces. In "Biological Monitoring for Pesticide Exposure" (R. G. Wang, C. A. Frankiin, R. C. Honeycutt, and J. C. Reinert, eds.), ACS Symposium Series, Vol. 382, pp. 315-326. Am. Chem. Soc., Washington, DC. Byrne, S. L., Shurdut, B. A., and Saunders, D. G. (1998). Potential chlorpyrifos exposure to residents following standard crack and crevice treatment. Env. Health Perspect. 106,725-731. California Environmental Protection Agency (Cal-EPA). (1994). CalTOX™, a Multimedia Total Exposure Model for Hazardous-Waste Sites. Spreadsheet user's guide, version 1.5. NTIS Publication No. PB95-100467. Office of Scientific Affairs, Dep. of Toxic Substances Control, Cal-EPA, Sacramento. Calvin, G. (1992). Risk management case history-detergents. In "Risk Management of Chemicals" (M. L. Richards, ed.). Royal Society of Chemistry, United Kingdom. Cosmetic, Toiletry and Fragrance Association, Inc. (CTFA). (1983). "Summary of the Results of Surveys of the Amount and Frequency of Use of Cosmetic Products by Women." Report prepared by ENVIRON Corp. CTFA, Washington, DC. Curry, K. K., Brookman, D. J., Whitmyre, G. K., Driver, J. H., Hackman, R. J., Hakkinen, P. J., and Ginevan, M. E. (1994). Personal exposures to toluene during use of nail lacquers in residences: description of the results of a preliminary study. J. Expos. Anal. Environ. Epidemiol. 4, 443-456. Dockery, D. W., and Spengler, J. D. (1981). Indoor-outdoor relationships of respirable sulfates and particles. Atmos. Environ. 15,335-343. Driver, J. H., Konz, J. J., and Whitmyre, G. K.. (1989). Soil adherence to human skin. Bull. Environ. Contam. Toxicol. 43, 814-820. Driver, J. H., and Wilkinson, C. F. (1996). Pesticides and human health: science, regulation and public perception. In "Risk Assessment and Management Handbook for Environmental, Health and Safety Professionals" (R. V. Kalluro, S. M. Bartell, R. M. Pitblado, and R. S. Stricoff, eds.). McGraw-Hill, New York. Eberhart, D. C. (1994). Current activities in assessing human exposures to lawn chemicals. In "Workshop on Residential Exposure Assessment, Annual Meeting of the International Society for Exposure Analysis and the International Society for Environmental Epidemiology," Research Triangle Park, NC. European Centre for Ecotoxicology and Toxicology of Chemicals (ECETOC). (1994). Tech. Rep. 58, Assessment of Non-Occupational Exposure Chemicals, Brussels. Fenske, R. A., Black, K. G., Elkner, K. P., Lee, c., Methner, M. M., and Soto, R. (1990). Potential exposure and health risks of infants following indoor residential pesticide applications. Am. J. Publ. Health 80, 689-693. Furtaw, E. J., Pandian, M. D., and Behar, J. V. (1993). Human exposure in residences to benzene vapors from attached garages. In "Proceedings of International Conference: Indoor Air '93," Helsinki, Finland. Gibson, J. E., Peterson, R. K. D., and Shurdut, B. A. (1998). Human exposure and risk from indoor use of chlorpyrifos. Environmental Health Perspectives 106,303-306. Gibson, W. S., Keller, F. R., Foltz, D. J., and Harvey, G. J. (1991). Diethylene glycol monobutyl ether concentrations in room air from application of

440

CHAPTER 16


cleaner formulations to hard surfaces. 1. Expos. Anal. Environ. Epidemiol. 1,369-383. Hakkinen, P. J. (1993). Cleaning and laundry products: human exposure assessments. In "Handbook of Hazardous Materials," pp. 145-151. Hakkinen, P. J., Kelling, e. K, and Callender, J. e. (1991). Exposure assessment of consumer products: human body weights and total body surface areas to use, and sources of data for specific products. Vet. Hum. Toxicol. 33,61-65. Harris, S. A., and Solomon, K R (1992). Human exposure to 2,4-D following controlled activities on recently-sprayed turf. 1. Environ. Sci. Health B27(1),9-22. Harris, S. A., Solomon, K R., and Stephenson, G. R. (1992). Exposure of homeowners and bystanders to 2,4-dichlorophenoxyacetic acid (2,4-D). 1. Environ. Sci. Health B27(1), 23-38. Hendricks, M. G. (1970). Measurement of enzyme laundry product dust levels and characteristics in consumer use. 1. Am. Oil. Chem. Soc. 47,207-211. HilI, R H., Jr., Head, S. L., Baker, S., Gregg, M., Shealy, D. B., Bailey, S.L., WilIiams, e. e., Sampson, E. J., and Needham, L. L. (1995). Pesticide residues in urine of adults living in the United States: reference range concentrations. Environ. Res. 71, 99-108. Immerman, W. W., and Schaum, J. L. (1990). "Nonoccupational Pesticide Exposure Study (NOPES)." NTIS Publication No. PB90-152224, 256 p. Research Triangle Park, NC: Atmospheric Research and Exposure Assessment Laboratory, U.S. Environmental Protection Agency. Report prepared by Research Triangle Institute, Research Triangle Park, NC. International Life Sciences Institute (ILSI) (1998). (S. S. Olin, ed.). "Aggregate Exposure Assessment." ILSI Risk Sciences Institute Workshop Report, Washington, De. Jayjock, M. A and Hawkins, N. e. (1993). A proposal for improving the role of exposure modeling in risk assessment. Am. Ind. Hyg. Assoc. 1. 54, 733-741. Jenkins, P. L., PhilIips, T. H., Mulberg, E. J., and Hui, S. P. (1992). Activity patterns of Californians: use of and proximity to indoor pollutant sources. Atmos. Environ. 26, 2141-2148. Krieger, R I., Bernard, C. E., Dinoff, T. M., Fell, L., Osimitz, T. G., Ross, J. H., and Thongsinthusak, T. (2000). Biomonitoring and whole body cotton dosimetry to estimate potential human dermal exposure to semivolatile chemicals. 1. Expos. Anal. Environ. Epidemiol. 10,50-57. Krieger, R. I., Bernard, e. E., Dinoff, T. M., Ross, J. H., and WiIIiams, R L. (2001). Biomonitoring of persons exposed to insecticides used in residences. Ann. Occup. Hyg. 45(Suppl 1),5143-5153. Matoba, Y. (1996). Simulation of indoor behavior of insecticides applied by various methods. In "SP World," No. 24. Sumitomo Chemical Co., Osaka, Japan. McKone, T. E. (1991). Human exposure to chemicals from multiple media and through multiple pathways: research overview and comments. Risk Anal. 11,5-10. McKone, T. E. (1993). Understanding and modeling multipathway exposures in the home. In "Reference House Workshop II: Residential Exposure Assessment for the '90s." Society for Risk Analysis, 1993 Annual Conference, Savannah, GA. Melia, R J. W, F1orey, e. duV., Darby, S. e., Palmes, E. D., and Goldstein, B. D. (1978). Differences in N02 levels in kitchens with gas or electric cookers. Atmos. Environ. 12, 1379-1381. National Research Council (NRC). (1980). Committee on Urban Pest Management. Nat. Acad. Press, Washington, De. National Research Council (NRC). (1993). "Pesticides in the Diets of Infants and Children. Committee on Pesticides in the Diets of Infants and Children," Board on Agriculture and Board on Environmental Studies and Toxicology, Commission on Life Sciences, Nat. Acad. Press, Washington, DC. Ott, W. R. (1985). Total human exposure: an emerging science focuses on humans as receptors of environmental pollution. Environ. Sci. Technol. 19, 880. Pandian, M. D., Ott, W R, and Behar, J. V. (1993). Residential air exchange rates for use in indoor air and exposure mode ling studies. 1. Expos. Anal. Environ. Epidemiol. 3,407-416. Pandian, M. D. and Furtaw, E. J. (1995). ''THERdbASE: Total Human Exposure Relational Database and Advanced Simulation Environment." Harry Reid

Center for Environmental Studies, University of Nevada at Las Vegas, Las Vegas. Developed under contract to the U.S. EPA, Office of Research and Development, Environmental Monitoring Systems Laboratory, Las Vegas. Pellizzari, E. D., Hartwell, T. D., Perritt, R. L., Sparacino, e. M., Sheldon, L. S., Whitmore, R W., and Wallace, L. A. (1987). Comparison of indoor and outdoor residential levels of volatile organic chemicals in five U.S. geographic areas. Environ. Int. 12,619-623. PeIIizzari, E. D., Thomas, K W, Clayton, C. A, Whitmore, R W, Shores, R e., Zelon, H. S., and Peritt, R L. (1993). "Particle Total Exposure Assessment Methodology (PTEAM): Riverside, California Pilot Study," Vo!. I, EPAJ600/SR-93/050. U.S. EPA, Research Triangle Park, Ne. PhilIips, L. J., Fares, R. J., and Schweer, L. G. (1993). Distributions of total skin surface area to body weight ratios for use in dermal exposure assessments. 1. Expos. Anal. Environ. Epidemiol. 3, 331-338. Racke, K D., and Leslie, A. R (eds.). (1993). "ACS Symposium Series 522. Pesticides in Urban Environments: Fate and Significance." 203rd National Meeting of the American Chemical Society, San Francisco, California, April 5-10, 1992. Published by the American Chemical Society, Washington, DC, ISBN 0-8412-2627-X, 378 p. Roberts, J. W., Budd, W. T., Ruby, M. G., Camann, D. E., Fortmann, R e., Lewis, R. G., Wallace, L. A, and Spittler, T.M. (1992). Human exposure to pollutants in the floor dust of homes and offices. 1. Expos. Anal. Environ. Epidemiol. 1 (Supp!. 1),127-146. Ross, J. H., Fong, H. R., Thongsinthusak, T., Margetich, S., and Krieger, R (1991). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: using the CDFA roller method. Interim report II. Chemosphere, 22, 975-984. Ross, J., Thongsinthusak, T., Fong, H. R, Margetich, S., and Krieger, R (1990). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: an interim report. Chemosphere, 20, 349-360. Spengler, J. D., Duffy, C. P., Letz, R., Tibbets, T. W, and Ferris, B. G. Jr. (1983). Nitrogen dioxide inside and outside 137 homes and implications for ambient air quality standards and health effects research. Environ. Sci. Technol. 17(3), 164--168. Thomas, K W, PeIIizzari, E. D., Clayton, C. A., Perritt, R. L., Dietz, RN., Goodrich, R. W, Nelson, W e., and Wallace, L. A (1993). Temporal variability of benzene exposures for residents in several New Jersey homes with attached garages or tobacco smoke. 1. Expos. Anal. Environ. Epidemiol. 3, 49-73. Thompson, D. G., Stephenson, G.R, and Sears, M. K (1984). Persistence, distribution, and dislodgeable residues of 2,4-D following its application to turfgrass. Pestic. Sci. 15,353-360. Turnbull, D., and Rodricks, J. V. (1989). A comprehensive risk assessment of DEHP as a component of baby pacifiers, teethers and toys. In "The Risk Assessment of Environmental and Human Health Hazards: A Textbook of Case Studies" (D. J. Paustenbach, ed.). WiIey, New York. U.S. Environmental Protection Agency (U.S. EPA). (1991). "Time Spent in Activities, Locations, and Microenvironments: A California-National Comparison," USEPA Pub!. No. 600/4-91/006. Office of Research and Development, Environmental Monitoring Systems Laboratory, Las Vegas. U.S. Environmental Protection Agency (U.S. EPA). (1992). "Dermal Exposure Assessment: Principles and Applications," Washington, DC: USEPA Pub!. No. 600/8-91-011. Exposure Assessment Group, Office of Health and Environmental Assessment, Office of Research and Development, Washington, De. U.S. Environmental Protection Agency (U.S. EPA). (1999a). "Overview of Issues Related to the Standard Operating Procedures for Residential Exposure Assessment." Presented to the EPA Science Advisory Panel U.S. EPA, OPP, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1999b). "Exposure Factors Handbook;' USEPA Pub!. No. EPAJ600/C-99/001. National Center for Environmental Assessment, Cincinnatti. Vaccaro, J. R., Nolan, R J., Murphy, P. G., and Berbrich, D. B. (1996). "The Use of Unique Study Design to Estimate Exposure of Adults and Children to Surface and Airborne Chemicals," STP 1287, pp. 166-183. Am. Soc. for Testing and Materials, West Conshohocken, PA.

References

Vermiere, T.G., van der Poel, P., van de Laar, R. T. H., and Roelfzema, H. (1993). Estimation of consumer exposures to chemicals: applications of simple models. Sei. Total Environ. 136, 155-176. WaIIace, L. A. (1987). "The TEAM Study: Summary and Analysis, Vo!. I," EPA 600/6-87/002a. V.S. Environmental Protection Agency, Office of Research and Development, Nat. Tech. Information Service, Springfield, VA. WaIIace, L. A. (1989). The exposure of the general population to benzene. Cell Bio!. Toxieol. 5, 297-314. WaIIace, L. A. (1990). Major sources of exposure to benzene and other volatile organic compounds. Risk Ana!' 10, 59-64. WaIIace, L. A. (1991). Comparison of risks from outdoor and indoor exposure to toxic chemicals. Environ. Health Perspeet. 95, 7-13. WaIIace, L. (1993). A decade of studies of human exposure: what have we learned? Risk Anal. 13, 135-139. Wallace, L. A., Zweidinger, R., Erickson, M., Cooper, S., Whitaker, D., and Pellizzari, E. (1982). Monitoring individual exposure: measurement of volatile organic compounds in breathing-zone air, drinking water, and exhaled breath. Environ Internat. 8, 269-282. WaIIace, L. A., PeIlizzari, E., HartweII, T., Rosenzweig, R., Erickson, M., Sparacino, c., and ZeIon, H. (1984). Personal exposure to volatile organic compounds: I. Direct measurement in breathing-zone air, drinking water, food, and exhaled breath. Environ. Res. 35, 293-319. WaIIace, L. A., PeIlizzari, E., HartweII, T., Sparacino, c., Sheldon, L., and Zelon, H. (1985). Personal exposures, indoor-outdoor relationships and breath levels of toxic air agents measured for 355 persons in New Jersey. Atmos. Environ. 19, 1651-1661. WaIIace, L. A., PeIIizzari, E., HartweII, T., Whitmore, R., Sparacino, c., and Zelon, H. (1986). Total exposure assessment methodology (TEAM) study: personal exposures, indoor-outdoor relationships, and breath levels of volatile organic compounds in New Jersey. Environ. Int. 12,369-387. WaIIace, L. A., Pellizari, E. D., HartweII, T. D., Sparacino, c., Whitmore, R., Sheldon, L., Zelon, H., and Perrit, R. (l987a). The TEAM study: personal exposures to toxic substances in air, drinking water, and breath of 400 residents of New Jersey, North Carolina, and North Dakota. Environ. Res. 43, 290-307. WaIIace, L. A., Pellizari, E., Hartwell, T., Perritt, K., and Ziegenfus, R. (1987b). Exposures to benzene and other volatile organic compounds from active and passive smoking. Areh. Environ. Health 42, 272-279. WaIIace, L. A., PeIlizari, E., Leaderer, B., Hartwell, T., Perritt, R., Zelon, H., and Sheldon, L. (I 987c). Emissions of volatile organic compounds from building materials and consumer products. Atmos. Environ. 21,385-393. WaIIace, L. A., Pellizzari, E. D., HartweII, T. D., Whitmore, R., Perritt, R., and Sheldon, L. (1988). The California TEAM study: breath concentrations and personal exposures to 26 volatile compounds in air and drinking water of 188 residents of Los Angeles, Antioch, and Pittsburgh, CA. Atmos. Environ. 22,2141-2163.

441

WaIIace, L. A., PeIIizzari, E. D., Hartwell, T. D., Davis, v., Michael, L. C., and Whitmore, R. W. (1989). The influence of personal activities on exposure to volatile organic compounds. Environ. Res. 50, 37-55. WaIIace, L. A., Nelson, W. c., Ziegenfus, R., and PeIlizzari, E. (l991a). The Los Angeles TEAM study: personal exposures, indoor-outdoor air concentrations, and breath concentrations of 25 volatile organic compounds. J. Expos. Ana!. Environ. Epidemiol. 1(2),37-72. WaIIace, L. A., PeIIizzari, E., and Wendel, C. (l99Ib). Total volatile organic concentrations in 2700 personal, indoor, and outdoor air samples collected in the VSEPA TEAM studies. Indoor Air 4, 465-477. WaIIace, L. A., Pellizzari, E., Sheldon, L., Hartwell, T., Perritt, R., and Zelon, H. (1991c). Exposures of dry cleaning workers to tetrachloroethylene and other volatile organic compounds: Measurements in air, water, breath, blood, and urine. In "Annual Meeting of the International Society for Exposure Analysis and Environmental Epidemiology," Atlanta. Whitford, E, Kronenberg, J., Lunchick, c., Driver, J., TomerIin, R., Wolt, J., Spencer, H., Winter, c., and Whitmyre, G. (1999). "Pesticides and Human Health Risk Assessment: Policies, Processes and Procedures," Purdue Pesticide Programs, Publication PPP-48. Purdue Vniv. Cooperative Extension Service, West Lafayette, IN. Whitmore, R. W, KeIIy, J. E., and Reading, P. L. (1992). "National Home and Garden Pesticide Vse Survey: Final Report," NTIS PB92-174739. V.S. EPA, Office of Pesticide Programs and Toxic Substances, Washington, DC. Whitmore, R. W., Immerman, E W., Camann, D. E., Bond, A. E., Lewis, R. G., and Schaum, J. L. (1994). Non-occupational exposures to pesticides forresidents of two V.S. cities. Areh. Environ. Contam. Toxieol. 26,47-59. Whitmyre, G. K., Driver, J. H., Ginevan, M. E., Tardiff, R. G., and Baker, S. R. (1992a). Human exposure assessment. I: understanding the uncertainties. Toxieo!. Ind. Health 8, 297-320. Whitmyre, G. K., Driver, J. H., Ginevan, M. E., Tardiff, R. G., and Baker, S. R. (l992b). Human exposure assessment. 11: quantifying and reducing the uncertainties. Toxieo!. Ind. Health 8, 321-342. Whitmyre, G. K., Driver, J. H., and Hakkinen, P. J. (1996). Assessment of residential exposures to chemicals. In "Fundamentals of Risk Analysis and Risk Management" (V. Molak, ed.), pp. 125-141. CRC Lewis Publishers, Boca Raton, Florida. Wooley, J., Nazaroff, W W, and Hodgson, A. T. (1990). Release of ethanol to the atmosphere during use of consumer cleaning products. 1. Air Waste Management Assoe. 40, 1114-1120. Zartarian, V. G., and Leckie, J. O. (1998). Dermal exposure: the missing link. Environ. Sei. Teehnol. March I, 134-137. Zartarian, V. G., Ozkaynak, H., Burke, J. M., ZufaII, M. J., Rigas, M. L., and Furtaw, E. J. (2000). Amodeling framework for estimating residential exposure to and dose of chlorpyrifos via dermal residue contact and non-dietary ingestion. Environ. Health Perspeet. 108,505-514. Zweiner, R. J., and Ginsburg, C. M. (1988). Organophosphate and carbamate poisoning in infants and children. Pediatries 81,121-126.

CHAPTER

17 Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure Barbara J. Peters en , Susan H. Youngren, and Cassi L. Walls Novigen Sciences

Exposure assessment methodology is rapidly changing to meet the demands of consumers, regulators, and researchers. Passage of the Food Quality Protection Act (FQPA) in 1996 challenged the discipline of risk assessment to develop methods that are scientifically defensible and that provide estimates to meet the legal requirements. New methods that allow quantitative estimation of exposure from different sources and from mUltiple chemicals are now needed. Additionally, richer data sets also need to be incorporated into the new methodologies to reflect realistic exposure patterns. The current methods that use default assumptions or simplistic methodologies generate "worst case" exposure and risk estimates that are extreme. These simplified approaches do not permit any discrimination between factors that are likely to reflect risks that really exist and those that are truly hypothetical. Researchers in academic and industrial positions are generating extensive exposure data sets. However, these data need to be understood. For example, the results of biomonitoring studies need to be interpreted in relation to potential exposures and risks. Specifically, environmental residues are being measured with increasingly sensitive analytical methods that can now detect contaminants at much lower levels. These analyses produce results that require statistical treatment of variations in levels of detection and sampling variability in new ways. Fortunately, computer technology also has improved so that it is possible to conduct analyses relatively rapidly that would have been almost impossible even 10 years ago. This chapter focuses on methodology to estimate dietary exposures and provides options that range from "worst case" screening assessments to much more refined and accurate assessments. Separate sections introduce the relatively new and much more complicated techniques that are under consideration for conducting aggregate and cumulative assessments. Handbook of Pesticide Toxicology Volume 1. Principles

An aggregate assessment estimates exposure to one chemical via multiple pathways including inhalation, dermal, and oral routes. A cumulative assessment estimates exposure to multiple chemicals via multiple pathways. The methods for estimating cumulative exposure are presented from the point that it has been determined that the chemicals cause toxicity by a common mechanism. Figure 17.1 provides a schematic overview of an aggregate and cumulative assessment process.

17.1 OVERVIEW There are several features that are required in any exposure assessment. Examples of common features include the ability to provide estimates of exposure that are statistically representative of the population to be evaluated, the ability to consider the impact of various scenarios and data, and the ability to provide documentation that is suitable for regulatory decisions and/or to support peer review. Typically the analyst also requires estimates for a range of user-specified subpopulations as well as for the overall population. In the case of aggregate or cumulative exposure, it is also necessary to be able to estimate exposure to single or multiple compounds, respectively, for relevant time periods. The goal of aggregate and cumulative exposure assessments is to characterize the exposure of the population of concern (e.g., adults, toddlers), and to identify the variability and uncertainty associated with that exposure. The exposures are characterized by estimating the level of chemical uptake via ingestion, inhalation, and/or dermal absorption of the substance over various time periods. The typical time periods over which exposures may be evaluated include daily/acute, short-term (1-7 days), intermediate-

443


444

CHAPTER 17

Modeling Dietary Exposure Non-Dietary Assessment

Dietary Assessment USDA CS F[]

Pesticide Use Patterns

Other Residue Adjustments

Figure 17.1

Components of an aggregate and cumulative risk assessment.

term, and chronic (up to 1 year) time periods. These time periods are based on the toxicity profile of the chemical. In addition, the exposure assessment also can be useful to identify the potential importance of a specific route relative to other pathways of exposure. In these types of assessments, exposure can result from residues in food, residues in or around the residence, from residues in parks/schools/towns, and/or residues from occupational uses of the pesticide. Therefore, it is desirable to be able to estimate the contribution of each route to exposure. That is, the method should identify the proportion of exposure that can result from oral, dermal, or inhalation, or a combination of these routes. In many cases, exposures from more than one source need to be considered. For example, oral exposure can arise as a result of residues in the diet or from other pathways such as toddler hand-to-mouth activity. The methods should allow the user to aggregate exposures as appropriate for the scenarios under consideration. Aggregation may be relevant to one chemical contained in one product that has multiple routes of exposure. For example, a compound might be used exclusively in a lawn product that is applied

by the homeowner. In this case, the person applying the compound may be exposed by dermal and inhalation routes while applying the compound and using the lawn after the treatment. Aggregation might also be relevant to one chemical contained in multiple products and with multiple exposure routes. For example, if a chemical is used on crops and also as a termiticide, possible routes of exposure are oral (from residues on food), dermal (from contact with treated surfaces in the home), and inhalation (from residues in the air around the home). For cumulative exposure assessments, the methods need to simulate situations that encompass exposure to more than one chemical, with multiple uses or sources and multiple exposure routes (Fig. 17.2). Cumulative assessments should be limited to chemicals that have a similar mode of action toxicologically. This requires that the relative toxicity of the various chemicals included in the analysis be quantitatively specified. It is also necessary that the exposure methods account for temporal and spatial considerations. The temporal aspect should consider when exposures occur simultaneously. For example, the method should account for overlap of exposures based on product usage information and chemical degradation. The spatial or regional aspect should be included in the exposure methodology. For example, the types of contaminants encountered in a home in Florida may be very different from those found in a home in northern Maine. In summary, exposure estimates must be able to assess exposures that are specific to both time and location.

17.2 EXPOSURE MODELS 17.2.1 GENERAL EXPOSURE MODEL Exposure assessment models can be designed for a specific route or for a specific source of exposure. Routes of exposure - Dietary

80

""-Use B

70 01

~

01 ::J

e

--Use C

x

........ TOTAL

60 50 40

::J

!/)

0 c.. 30

> shoulder > buttocks > abdomen; for occluded skin, the order is back > abdomen > buttocks > shoulder. Wester et al. (1994) also demonstrated that pyrethrin absorption through the human forearm is less than the predicted absorption in the human scalp. This anatomical difference is somewhat consistent with lindane absorption through the forearm (18%), forehead (34%), and palm (34%) of rhesus monkeys (Moody and Ritter, 1989). This anatomical range for lindane is similar to that for dermal absorption of DEET (diethyl-m-toluamide) in rhesus monkeys (Moody et aI., 1989). There is also significant data to suggest that dermal absorption of permethrin, aminocarb, DEET, and fenitrothion in monkey foreheads is twice that in monkey forearms (Moody and Franklin, 1987; Moody et aI., 1987; Sidon et aI., 1988). However Moody et al. (1990, 1992) demonstrated that there is no difference between the absorption of acid and amine forms of 2,4-D in rhesus monkey forearm and forehead and forearm and palm regions. The palmar absorption data conflict with the accepted dogma that absorption through palmar skin should theoretically be less than that in forearm skin because of the thickness of the stratum corneum in palmar skin (Maibach et aI., 1971). It is proposed that because of the hydrophilic nature of 2,4-D-amine, absorption can occur through polar routes such as eccrine glands, which are more frequent in the palmar skin than in the forearm skin. This anatomical difference does not explain the discrepancy with lindane, which is more lipophilic than 2,4-D and least likely to be absorbed via a polar route. Despite a threefold range in follicle area in the marmoset, no differences in absorption rates of paraquat, mannitol, water, and ethanol were observed between different body sites (Scott et aI., 1991). However, among the different species examined in this

520

CHAPTER 22

Pesticide Disposition: Dermal Absorption

study, there was an 80-times range in follicle area, which correlated with observed differences in the rates of mannitol and paraquat absorption. The authors concluded that this correlation was only possible with relatively slowly absorbed test penetrants such as paraquat and mannitol. Further work is needed to determine the extent to which the unique anatomical features at different body sites play a role in absorption and penetration of both lipophilic and hydrophilic pesticides. 22.4.2 AGE-RELATED DIFFERENCES

Very little information is available on age-related absorption of pesticides. The generally accepted theory is that dermal bioavailability should be greater in the young than in adults because of the underdeveloped skin barrier in the young. Differences in dermal absorption between young and adult rat skin has been demonstrated for 11 out of 14 pesticides (Shah et aI., 1987a). Dermal absorption of four pesticides (atrazine, carbaryl, chlorpyrifos, and hexachloro-biphenyl) was greater in the young, whereas dermal absorption of seven pesticides [carbofuran, captan, dinoseb, disodium methanearsonate (DSMA), monosodium methanearsonate (MSMA), nicotine, and parathion] was greater in adults. In vitro studies with human skin further demonstrated that triclocarban absorption was greater in newborn foreskin (2.5%) than in adult abdominal skin (0.23%; Wester and Maibach, 1985). The latter comparison is confounded by anatomical site differences. Another study clearly indicated greater penetration of carbofuran in young rats as measured by both in vivo and in vitro methods (Shah et aI., 1987b). Carbofuran absorption in young rats was 34% for flow-through systems, 12% for static systems, and 34% for in vivo systems, whereas in older rats, absorption was 11.2, 8.8, and 11.8%, respectively. Although there were some differences between the systems used, there are clear agerelated differences for absorption of this carbamate insecticide. Lack of proper mixing and lipophilicity of carbofuran probably explain why absorption was low in the static cell system. In another in vitro study, carbofuran absorption through human newborn foreskin was 78% in 24 hours (Shehata-Karam et aI., 1988), which is greater than in previously described rodent models. This is not surprising because a low dose was used and dose-dependent effects on carbofuran absorption have been reported (Shah et al., 1987a, b). Dermal absorption studies with dinoseb exposure in rats have, however, demonstrated the opposite relationship, with absorption being greater in adults than in the young (Hall et al., 1992). In adult rats, 85% was absorbed with a half-life of 4.9 hours in one phase of the biphasic absorption process and the remaining 15% with a half-life of 34 hours in the second phase. In young rats, 59.7% was absorbed with half-life of 5.4 hours in the first phase and the remaining 40% with a half-life of 298 hours. Dermal absorption of captan was similar in adult (9.8%) and young (8.4%) rats in vivo after a 72-hour exposure (Fisher et al., 1992). Penetration also increased as applied dose decreased. However, in vitro studies utilizing static cells and flow-though diffusion cells provide conflicting results.

The static system gave penetration values twice those obtained for in vivo, while the flow-through diffusion system overestimated in vivo absorption by 43% in young and underestimated in vivo absorption in adults by 19%. This underscores the need to use a validated model prior to making such comparisons. 22.4.3 VARIATION BETWEEN AND WITHIN SPECIES

The use of laboratory animal models can sometimes lead to overprediction of pesticide absorption in human skin. Differences in permeability properties between human skin and those of laboratory animal skin can account for this overprediction. Furthermore, inherent structural differences in skin biology (e.g., skin thickness, sebaceous secretions) make speciesspecies extrapolation of dermal absorption data very difficult. These differences in epidermal and dermal anatomy and physiology have been well documented, although the basic architecture of skin is similar in all terrestrial mammals (Monteiro-Riviere, 1991). It is plausible that a high density of hair follicles attenuates the thickness of the interfollicular epidermis, which may promote absorption. Additionally, species differences in stratum corneum lipid composition may be the overriding factor in determining the rate and extent of absorption. Because of these most glaring differences that distinguish rodent skin from human skin, skin from rodents may not always serve as a suitable model. However, pig skin is functionally and structurally similar to that of human skin (Monteiro-Riviere, 1991) and, therefore, percutaneous absorption of toxicants through pig skin should mimic absorption through human skin. Recent studies have demonstrated that the range of percutaneous absorption of carbaryl, lindane, malathion, and parathion in pig skin in vivo (Carver and Riviere, 1989) or in vitro (Chang et aI., 1994b) is similar to that observed in humans (Feldmann and Maibach, 1974). However, this was not the case when the permeability of hydrophilic chemicals (mannitol, water, and paraquat) and lipophilic chemicals (carbaryl, aldrin, and fluazifop-butyl) in pig ear skin was compared with human abdominal skin and rat dorsal skin (Dick and Scott, 1992). This study demonstrated that for hydrophilic chemicals, pig ear skin and rat skin overestimated permeability in human skin. Although the permeability was generally higher in animal skin than in human skin for the lipophilic chemicals, the permeability of carbaryl in human and pig skin was almost identical. The permeability of lipophilic chemicals in pig skin correlated better with data from human skin compared to the permeability of hydrophilic chemicals. Bartek et al. (1972) demonstrated good agreement between human in vivo and pig in vivo dermal absorption data for lipophilic chemicals, such as butter yellow, DDT, haloprogin, lindane, and testosterone, and hydrophilic chemicals, such as caffeine, acetylcysteine, and cortisone. Further studies with nonhuman primates observed that lindane absorption from rhesus monkey forearm (18%) was about twice that for the ventral forearm in humans (9.3%) (Feldmann and Maibach, 1974) and for the ventral abdomen in pigs (7.7%) (Chang et aI., 1994b).

22.4 Factors that Affect Dermal Absorption

Finally, note that these species comparative data do not prove any specific anatomical route of penetration, only that there is a species difference. In summary, the rate of absorption of most chemicals can be ranked in the decreasing order rabbits > rodents > weanling pig

= rhesus monkey::::: human.

22.4.4 PHYSICOCHEMICAL PROPERTIES AND DOSE EFFECTS 22.4.4.1 Dose Effects Limited information is available in the literature regarding the effects of dose or multiple dose exposure on percutaneous absorption of pesticides. Most studies have demonstrated that absorption efficiency tends to decrease significantly as topical dose increases in simple mixtures. This makes predicting pesticide bioavailability very difficult. This decrease in absorption efficiency has been demonstrated with drugs (e.g., cortisone, testosterone, benzoic acid) in human and rhesus monkey skin (Scheuplein and Ross, 1974; Wester and Maibach, 1976), with lindane in humans (Maibach and Feldmann, 1974), and with parathion in pigs in vitro (Chang and Riviere, 1991). Other studies by Shah et al. (1987a) demonstrated that as the topical dose of 14 pesticides increased, the median penetration value decreased in both young and adult rats in vivo. Of the 14 pesticides tested, a highly significant effect of dose on skin permeability was observed with carbaryl, captan, folpet, and polychlorobiphenyl (PCB). More specifically, skin penetration expressed as a percentage of recovered carbaryl dose ranged from 30-37,12-20, and 4-5% in skin dosed with 31-37, 108, and 539-l-l-g/cm2 of carbaryl, respectively. Insecticides with skin penetration values higher than those for carbaryl were permethrin (16-57%), chlorpyrifos (59-90%), and parathion (58-82%). The range of skin penetration for these other three pesticides was associated with application of low, medium, and high doses. Note that this dose-dependent effect on absorption has not been demonstrated in human skin treated with parathion (Maibach and Feldmann, 1974) nor in rats treated with methylated arsenicals (Shah et aI., 1987a) and dinoseb (Hall et aI., 1992). Repeated or multiple exposure to malathion (15 I-l-g!cm2) for 15 days does not appear to alter absorption in guinea pigs (Bucks et aI., 1985; Courtheoux et aI., 1986) or in humans (Wester et aI., 1983). Therefore, malathion, does not appear to induce changes in barrier function as observed with multiple exposure to drugs such as hydrocortisone and salicylic acid (Roberts and Harlock, 1978; Wester et at., 1980b). However, be aware that some experimental protocols or real world occupational scenarios involve daily soap and water treatment prior to daily topical doses that can significantly decrease barrier function and can increase absorption (Wester and Maibach, 1989). One human in vivo 5-day study demonstrated significant dermal absorption (14%) of 2,4-D-amine, despite rigorous washing at 24 hours posttreatment (Moody et aI., 1992). However, only 0.5% absorption occurred after washing at 1 minute posttreatment.

521

22.4.4.2 Pesticide Chemistry Physicochemical factors, such as molecular weight and structure, lipophilicity, pKa, ionization, solubility, partition coefficients, and diffusivity, can influence the dermal absorption of various classes of pesticides. In addition, the penetration of acidic and basic pesticides is influenced by the skin surface, which is weakly acidic (pH 4.2-5.6). Although several of these factors (e.g., molecular weight and partition coefficients) have been used to predict absorption of various drug classes (Bunge and Cleek, 1995; Cleek and Bunge, 1993; Potts and Guy, 1992), this approach has not been applied to pesticides. It is expected that these physicochemical factors influence the rate and extent of absorption for various classes of pesticides. Paraquat and diquat are hydrophilic pesticides that exist as fixed charged cations and remain dissociated at all pH values. Very little paraquat or diquat is, therefore, expected to be absorbed by skin, although percutaneous absorption of paraquat has resulted in systemic effects and deaths in humans (Smith, 1988). Dermal absorption studies in human volunteers demonstrated 0.29, 0.23, and 0.29% absorption in the leg, forearm, and forearm, respectively (Wester et aI., 1984). One in vivo study in rats, demonstrated that paraquat absorption (3.5%) was greater in rats because an occlusive dressing was used and also because of differences in skin thickness between species (Chui et al., 1988). Topical application of 3-,24-, and 200-mg doses of paraquat to IPPSFs for 8 hours resulted in penetration (skin and perfusate) of 0.91, 1.09, and 0.50%, respectively (Srikrishna et aI., 1992). These absorption data are comparable to the human in vivo data. Despite the limited amounts absorbed, they were sufficient to cause morphological and biochemical changes in the IPPSFs. Other studies have determined that the in vitro permeability constants for paraquat in various animal species (rat, hairless rat, nude rat, mouse, hairless mouse, rabbit, guinea pig) are 40-1600 times greater than for humans (Walker et aI., 1983). Like paraquat, very little diquat is absorbed (0.3%) in the human forearm in vivo (Maibach and Feldmann, 1974). Diquat absorption increased to 1.4% with occlusion and to 3.8% with damaged skin. Data from these in vivo and IPPSF studies suggest that paraquat- or diquatinduced dermatotoxicity is a highly probable mechanism, a priori, for dermal absorption of these hydrophilic and charged pesticides. The various forms of organic arsenicals used as herbicides include dimethylarsinic acid (DMA) and the sodium salts of methylarsinic acid, MS MA and DSMA. Because of the carcinogenic potential of these herbicides, there have been no human dermal absorption studies. Many of the commercial MSMA and DMSA formulations contain various surfactants [U.S. Environmental Protection Agency, (EPA), 1975], that alter skin barrier properties and may promote permeation of the arsenicals. Skin penetration of MSMA and DSMA in mice can range from 2 to 22 and 1.2 to 15% of dose, respectively, within 72 hours (Hall et al., 1989; Shah et aI., 1987a). Surprisingly, absorption was greater in adult mice than young mice for both arsenicals and there was no dose effect for either of these arsenicals. In similar studies, a constant fraction of the dose (12.4%)

522

CHAPTER 22


penetrated mice skin within 24 hours in vitro from aqueous vehicles over an wide dosage range (Rahman and Hughes, 1994). Of this amount, only 4% were absorbed into the receptor fluid. In the same study, short-term exposure (1 hour) resulted in 0.63% absorption in skin and 0.03% absorption in receptor fluid. For the remaining 23-hour perfusion, 0.54% of the dose remained in the skin, whereas 0.13% was still absorbed. Although the pH levels of MSMA and DSMA were at least 1.18-1.7 units higher than the pKa of these arsenica1s, indicating 93-98% of the arsenicals were ionized (mono- and dibasic forms), ionization had no effect on absorption though mouse skin. Topical application of aqueous solutions (20, 100, and 250 !-LI) of 10 !-Lg ofDMA in mice skin in vitro resulted in 5.1625.22% dose in receptor fluid and 2-16% dose in skin tissue within 24 hours (Hughes et al., 1995). This study also determined that lag times were about 1 hour, but within 4 hours, 50% of the total (24 hour) cumulative dose detected in the receptor fluid had penetrated the skin at doses ranging from 10 to 500 !-Lg. As demonstrated in previous studies with MSMA and DSMA, short-term exposure to DMA (1 hour) resulted in about 1% absorption, and the percentage of dose in perfusate and skin was unaffected by changes in applied dose or pH of the dosing solution. Many of the chlorinated hydrocarbon insecticides have a known tendency to accumulate in adipose tissue. Because of the high lipophilicity of these compounds compared to organophosphates, partitioning of absorbed compound from dermis to subcutaneous adipose tissue ultimately affects absorption flux profile and tissue distribution. For example, parathion and lindane have similar molecular weights, but the difference in partition coefficients is a very plausible explanation for why measured and predicted absorption of lindane was demonstrated experimentally to be less than that for parathion (Chang et aI., 1994b). In a 24-hour in vivo study in rabbits, greater quantities of DDT and dieldrin were retained on the skin surface, skin, and adipose tissue compared to carbaryl, parathion, and malathion, which are less lipophilic than the organoch10rines (Shah and Guthrie, 1977). Not surprisingly, blood concentration was significantly lower at 8 and 24 hours for the organochlorines. Carbaryl is the most studied of all carbamate pesticides and despite its low toxicity, appears to penetrate human and animal skin more readily than most other pesticides. This increased permeability compared to most other pesticides is most likely associated with its unique physicochemical characteristics. Almost complete penetration of carbaryl was observed when carbaryl was dissolved in acetone and applied to the forearm and jaw angle of six human volunteers (Feldmann and Maibach, 1974; Maibach et al., 1971). The data from these studies demonstrated that 74% of the 24-hour applied dose (4 !-Lg/cm2) was excreted in urine over 5 days. Utilizing deconvolution analysis of the same human data (skin to urine and blood to urine data), cumulative absorption over 5 days was estimated to be 63% of the applied dose, with 45% of this occurring 8 hours after onset of penetration (Fisher et aI.,

1985). When this analysis was performed with 11 other pesticides (e.g., parathion, aldrin, diquat), more of the carbaryl dose was absorbed within 120 hours compared to other pesticides (0.3-20%). Only carbaryl had a lag time (3.5 hours), which was followed by rapid absorption. About 50% of the 120-hour total absorption of parathion and dieldrin occurred in the first 4 hours. Therefore, although the absorption rate appeared to be less (due to the 3.5-hour lag time) with carbaryl compared to parathion in humans, the extent of absorption during 120 hours was greater for carbaryl. This 1ag period may be unique to carbaryl, although it can be formulation dependent or related to body clearance. Although dermal absorption usually follows Fick's law of diffusion, other studies have demonstrated that dermal absorption for carbaryl as well as dinoseb is a biphasic process. This phenomenon is probably related to the physicochemical properties of the penetrant (Hall et aI., 1992; Shah and Guthrie, 1983). 22.4.5 PESTICIDE FORMULATION AND MIXTURES

Insecticide efficacy, the stability of active ingredients, and programmed release of active ingredients from the vehicle/device are the most important characteristics controlled for when pesticides are formulated (Krenek and Rohde, 1988). EPA registration does not always require percutaneous absorption studies. For this reason, more efficacy data than dermal pharmacokinetic data are available in the literature. Furthermore, most of the available pesticide absorption data pertain to binary mixtures (pesticide + vehicle). Technical grade formulations are, however, complex mixtures of formulation additives and, therefore, risk assessment based on data from exposure to binary mixtures may be inappropriate. Pesticides are usually formulated to contain active and inactive or inert ingredients. The latter component(s) can enhance the rate and extent of absorption or slow the release of the active ingredient and thus reduce the rate and extent of absorption (Walters and Roberts, 1993). These "inert" ingredients are often classified as adjuvants, surfactants, preservatives, solvents, diluents, thickeners, and stabilizers. These pesticide additives were first covered by the Food and Drug Administration and now are covered by EPA regulation 40 CFR 180.1001 and also TSCA and FIFRA (Seaman, 1990). This increasing list of inerts as well as the prohibitive cost to obtain 40 CRF 180.1001 clearance of new inerts strongly support the need to evaluate the influence of current and novel inerts on the toxicology and dermal absorption of active ingredients in pesticide formulations. Several studies have demonstrated the penetration enhancing ability of acetone compared to water, ethanol, or other vehicles commonly used in dermal absorption studies. Early work by O'Brien and Dannelley (1965) showed that in comparison with benzene and corn oil, acetone was best at enhancing carbaryl absorption. More recent studies also have demonstrated the enhancing effect of acetone compared with other solvent systems on the absorption of carbaryl, p-nitrophenol, and 2,4-D (Baynes and Riviere, 1998; Brooks and Riviere, 1995; Moody et aI., 1992).

22.4 Factors that Affect Dennal Absorption

However, other studies have demonstrated that commercial formulations are more effective than acetone in enhancing pesticide absorption. Methyl parathion absorption in vitro in human skin at 24 hours was 1.3% in acetone, but was significantly increased to 5.2% in a commercial formulation (Sartorelli et aI., 1997). Likewise, in vivo dermal exposure studies of lindane in humans resulted in approximately 60% with a white spirits formulation and 5% with an acetone vehicle (Dick et aI., 1997a, b). In these latter experiments, more of the lindane dose (79%) remained on the skin surface at 6 hours with acetone than with the white spirits formulation (10.5%), and significant levels of lindane accumulated in the stratum corneum with white spirits (30%) and with acetone (14.3%) at 6 hours. These findings strongly suggest that the white spirits formulation enhanced lindane penetration with respect to acetone vehicle. The in vitro studies with human skin also demonstrated a similar pattern although only 18 and 0.3% of the dose was absorbed into the perfusate at 6 hours for the white spirits formulation and the acetone vehicle, respectively. Topical application of 1% commerciallotion of lindane in vitro in human and guinea pig skin resulted in absorption levels as high as 71.72 and 35.31 %, respectively, at 48-hour exposure (Franz et aI., 1996). Dermal absorption of alachlor as an emulsifiable concentrate and microencapsulated formulation was demonstrated to be 8.5 and 3.8%, respectively, in rhesus monkeys after a 12-hourexposure (Kronenberg et aI., 1988). About 88% of the systemically absorbed dose were excreted in urine within 48 hours. However, the differences between these two formulations were not statistically significant. Although dilution of either of these formulations (1 :29) slightly enhanced alachlor absorption, these effects were surprisingly not statistically significant. One in vitro study with human skin demonstrated similar absorption data (0.54%) after an 8-hour exposure and peak fluxes within 3-5 hours postapplication (Bucks et aI., 1989a). However, a significant effect of formulation dilution with water was observed in this study, even though the same mass of alachlor was applied to skin. Not surprisingly, a greater fraction of alachlor was present on the skin surface and skin tissue than in the receptor fluid, and the high capacity for stratum corneum binding demonstrated in this study is not unique for related chlorinated aromatic chemicals. Data from several studies have suggested that the insect repellent, DEET, enhances transdermal delivery of drugs and toxicants (Moody et aI., 1987; Windheuser et aI., 1982). Some studies have demonstrated that DEET can act as a transdermal accelerant of2,4-D-amine (Moody et aI., 1992). Recent studies in our laboratory have, however, determined that DEET blocked permethrin absorption and inhibited carbaryl absorption in acetone, but not in dimethyl sulfoxide DMSO mixtures (Baynes et al., 1997; Baynes and Riviere, 1998). The insecticide synergist, piperonyl butoxide, was also shown to enhance carbaryl absorption (Baynes and Riviere, 1998). These diffusion studies further demonstrated that piperonyl butoxide does not enhance absorption through inert latex membranes, but does so in porcine skin sections. This observation suggests that some chemical-biological interaction or other mechanisms (e.g., ir-

523

ritation) may occur in skin to enhance the absorption of pesticides. An expected, but important finding in these carbaryl experiments was that increased dilution of the carbaryl formulation with water, especially in the presence of the surfactant, sodium lauryl sulfate (SLS), enhanced carbaryl absorption. The penetration enhancing effect of SLS also was observed with parathion (Qiao et al., 1996). In addition to the formulation additives, these agrochemicals may contain isomers, homo logs, or breakdown products that form after synthesis and/or formulation and during storage (Chambers and Dorough, 1994). Although these impurities can potentially alter the toxicity and toxicokinetics of the pesticide, many toxicology and dermal absorption studies have ignored these impurities and used the pure rather than the technical grade pesticide. There is evidence that technical grade malathion can be more lethal (eightfold difference) in rats than the purified form (Umetsu et aI., 1977). Other studies have demonstrated that organophosphates such as malathion and fenitrothion can potentiate the toxicity of the carbamate insecticide carbaryl (Takahashi et aI., 1987). Previous metabolism studies in the IPPSF (Carver et aI., 1990) demonstrated a significant first-pass metabolism of parathion to p-nitrophenol and paraoxon, and that these metabolites may be present simultaneously during absorption of parathion. Environmental exposure to parathion is never to pure parathion because spontaneous degradation occurs during storage. When mixtures of parathion and its metabolites were dosed and then assayed for parathion and its two metabolites across pig skin in vitro, significant interactions were detected. In general, the nontoxic metabolites p-nitrophenol and I-naphthol can significantly enhance the absorption of the parent compounds, parathion and carbaryl, respectively (Baynes and Riviere, 1998; Chang et aI., 1994a). Surprisingly, p-nitrophenol did not enhance the absorption of paraoxon; this toxic metabolite of parathion and parathion appears to decrease the absorption of p-nitrophenol and paraoxon. In other related absorption studies, pretreatment with 3% fenvalerate decreased subsequent absorption of parathion, increased subsequent lindane absorption, and had no effect on subsequent fenvalerate or carbaryl absorption (Chang et aI., 1995). These results underscore the chemical specificity of these interactions and reinforce the concept that the percutaneous absorption of a mixture cannot be predicted from individual component studies. These data suggest that other mechanisms in addition to vehicle and surfactant effects must be operating simultaneously; hence further investigation is required. The data reinforce the concept that the permeability of a mixture cannot be predicted from individual component studies. Many of the mechanisms of pesticide mixtures interactions are not well understood and are not easy to model, although a biophysical model for parathion was attempted (Williams et aI., 1996). It should also be recognized that it is more often the formulation additives and other environmental factors rather than the active ingredient that compromise the skin barrier and eventually enhance pesticide absorption. There is epidemiological evidence that agricultural pesticides can cause dermatoses (Abrams et aI., 1991;

524

CHAPTER 22


Cellini and Offidani, 1994; Guo et al., 1996) and there is experimental evidence that UV irradiation can enhance skin reactions to topical agricultural chemical treatment (Kimura et al., 1998). In the latter study, significant reactions were observed for several herbicides. Maibach and Feldmann (1974) demonstrated that dermal absorption of pesticides such as parathion, azodrin, and diquat occurs more readily (ninefold) through damaged skin than through normal skin. It is, therefore, plausible to assume that the formulation additive can inflict local reversible or irreversible "damage" to the skin structure and physiology, and that it is these interactions that modulate dermal absorption of most pesticides. 22.4.6 ENVIRONMENTAL FACTORS 22.4.6.1 Temperature

Changes in ambient air temperature can alter lipid fluidity in the intercellular lipid domain of the stratum corneum. This alteration in the intercellular pathway can theoretically alter pesticide penetration through the stratum corneum. Previous in vivo studies have demonstrated that increased percutaneous absorption of a cholinesterase inhibitor (VX) was a function of skin temperature (Craig et aI., 1977). In humans topically exposed to parathion at different ambient temperatures (11, 25, and 40°C), the urinary excretion of the metabolite p-nitrophenol paralleled the increase in ambient temperature (Hayes et al., 1964). Sev~ral in vitro experiments with pig skin also demonstrated that increasing air temperature from 37 to 42°C significantly increased parathion absorption (Chang and Riviere, 1991). Increased ambient temperatures can also increase the evaporation of volatile pesticides from the skin, thereby reducing the topical dose available for absorption. Increasing air flow over the skin increases evaporative loss and significantly decreases dermal residues in the upper skin layer of pigs for DDT, malathion, parathion, and DEET (Reifenrath et al., 1991). Wester et al. (1992a) demonstrated that isophenfos concentrations on the human skin surface in vivo was less than 1% dose at 24 hours and that evaporation from the skin surface during absorption reduced the dose available for penetration and absorption. Finally, it should be recognized that skin surface conditions in vitro are more easily controlled than in vivo, and data from in vitro studies can significantly underestimate evaporation in vivo. 22.4.6.2 Humidity and Occlusion

Skin hydration can be increased by occlusion, with high relative humidity or immersion conditions (e.g., swimming or bathing). Although previously it was assumed that hydration changes only affect dermal absorption of polar compounds, there is significant data that suggest that at high relative humidity, this hydration effect becomes more important for nonpolar molecules such as pesticides and is most likely secondary to an increase in diffusivity of the penetrating molecule (Behl et aI., 1980). Under relative humidity conditions greater than 80%, parathion

absorption was significantly increased in pig skin in vitro by as much as 2-3 times the value under standard conditions of 60% relative humidity (Chang and Riviere, 1991, 1993). The practical application of occlusion is when pesticides get into and under the clothing of workers and form the ideal reservoir for penetration and absorption into the skin. Occlusion can change dermal absorption by various mechanisms, among which are reducing loss of evaporation from the skin surface, enhanced skin hydration, changes in cutaneous metabolism, dermal irritation, and altered cutaneous blood circulation (e.g., vasodilation). Occlusion can increase hydration of the stratum corneum from as little as 5-15% to as much as 50% (Bucks et aI., 1989b), thereby modulating the absorption profile for the pesticide. One in vivo study with pigs (Qiao et aI., 1997) demonstrated that occlusion significantly enhanced pentachlorophenol (PCP) absorption from 29.1 to 100.72% dose and changed the shape of the absorption profile in blood and plasma. The study also suggested that occlusion changed the local metabolism of PCP and as a result, the 14C partitioning between plasma and red blood cells. Occlusion was also kinetically related to modification of cutaneous biotransformation of topical parathion (Qiao and Riviere, 1995). Occlusion enhanced the cutaneous metabolism of parathion to paraoxon and to p-nitrophenol as well as the percutaneous absorption and penetration of both parathion and p-nitrophenol. Occlusion also reduced parathion and p-nitrophenollevels in the skin, but increased p-nitrophenol and p-nitrophenol-glucuronide in the blood. Other in vivo studies (Qiao et al., 1993) showed that dermal occlusion significantly enhanced the rate and extent of parathion absorption in pigs in the abdomen (43.94 vs. 7.47%), buttocks (48.47 vs. 15.60%), back (48.82 vs. 25.00%), and shoulder (29.28 vs. 17.41 %). Although significant anatomical site differences were observed with non occluded skin, these site differences were concealed with occluded skin. In vitro studies with parathion also demonstrated that occlusion increased absorption from 0.46-7.69 to 1.04-17.46% at doses ranging from 4 to 400 ).lg/cm2 (Chang and Riviere, 1993). Pesticides can be transferred from cotton fabric into and through human skin as demonstrated in several studies (Snodgrass, 1992; Wester et aI., 1996a), but it should be recognized that these studies were under occlusive conditions. Dermal absorption of malathion was 3.92% with ethanol wet fabric and 0.6% with 2-day-treated cotton sheets (Wester et aI., 1996a). However, malathion absorption was increased to 7.34% when the 2-day-treatedldried cotton fabric was wetted with aqueous ethanol. In the same study, absorption of glyphosphate was 1.42% in water solution, 0.74% when applied as wet cotton sheets, and 0.08% when applied as 2-day-treatedldried cotton sheets. Absorption increased to 0.36% when the 2-daytreatedldried cotton sheets were wetted with water to simulate sweating and wet conditions. Military uniforms are impregnated with permethrin as a defense against nuisance and disease-bearing insects. Application of fabric impregnated with permethrin to the backs of rabbits resulted in a 3.2% migration to the skin surface with 2% of the impregnant being absorbed

22.4 Factors that Affect Dennal Absorption

and 1.2% remaining on the skin surface after 7 days of continuous skin contact (Snodgrass, 1992). The implications of these interactions, especially for agricultural workers during pesticide application in humid climates or for military personnel under combat conditions in the desert, should not be underestimated.

22.4.6.3 Soil Pesticide adsorption to soil can alter the amount of pesticide available for dermal absorption. It should also be recognized that exposure conditions such as exposure time, pesticide concentration, soil load, and soil characteristics are important variables that can theoretically influence absorption (Bunge and Parks, 1997). Soil adherence to skin, for instance, can vary from 10- 3 to 102 mg/cm 2 and has been shown to be activity dependent (Kissel et al., 1996). Predicting dermal absorption of pesticides from contaminated soils is, therefore, not a simple process and becomes problematic because there are very few studies that have addressed many of these issues. For several pesticides (e.g., pep, 2,4-D, chlordane), percutaneous absorption in acetone vehicle appears to be slightly less or not significantly different from absorption from soil. However, for several other pesticides (e.g., DDT, organic arsenicals), soil appears to reduce percutaneous absorption of the pesticide. The interactions between soil and several of these pesticides are subsequently described in more detail, but note that in vitro skin models are, in general, not very predictive of in vivo absorption when exploring these interactions (Wester and Maibach, 1998). Although DDT is no longer widely used in the United States, residues in soil are still detectable and human contact with contaminated soil can result in DDT exposure. One study demonstrated that in vivo absorption of DDT in rhesus monkeys was significantly less from soil (3.3% dose) than from acetone vehicle (18.9%; Wester et aI., 1990). The absorption of DDT in acetone in rhesus monkey is not significantly different from DDT absorption in man (10.4% dose; Feldmann and Maibach, 1974). In vivo absorption from acetone or soil was not similar to in vitro absorption ( r(Xi/PhH)

+Pr

+ l]p L

L ar(o) + 8 L

aO(o, m,

Es(o)] + c.

p) (17)

L IT covers all substituents, Xi, on the benzene ring. The second term on the right-hand side takes care of the three components of the "forward" effect. Lao, Lar, and L Es are made for substituents at positions indicated in parentheses relative to individual hydrogen-bonding substituents. The L sign outside the brackets means to sum up the forward effects on every hydrogen-bonding substituent. Each of the a, P, Pr, 8, and c values is calculated by regression analysis (Nakagawa et aI., 1992). The preceding procedure, counting substituent effects toward every hydrogen-bonding substituent "forwardly" and "multiply," has been nicely used to analyze some 200 log P values of multi substituted benzenes (n = 210, s = 0.118, r = 0.994) with some corrections for intramolecular hydrogen bonding and a buttressing effect of vicinally located substituents on the solvation of hydrogen-bonding substituents (Nakagawa et aI., 1992). Recently, a procedure similar to that

described previously has been applied to analyze and predict the log P values of substituted pyridines and diazines, in which the fused N atom in the ring is dealt with as being a substituent (Yamagami et aI., 1995). The procedure works well with certain approximations, but indicates that further elaboration is required for heteroaromatic compounds. The log P value of a number of zwitterionized di- to pentapeptides at the isoelectric point has been analyzed with the use of free-energy related parameters under defined conditions (Akamatsu and Fujita, 1992). The IT parameter of the side chain of the amino acid units is that for aliphatic substituents (Iwasa et aI., 1965; Leo et aI., 1971). Along with the term for the steric effect of the side chain in terms of a variation of Es, correction terms for polar side chains interacting with the backbone -CONH - structure and the ,B-turn formation are included (Akamatsu and Fujita, 1992). A similar set of aliphatic parameters is used in the analyses of the log P values of primary, secondary, and tertiary amines and quaternary ammonium ions pairing with picrate (Takayama et aI., 1985). 29.3.5 COMPUTER-AIDED PROCEDURES To make a more comprehensive procedure that covers a wide variety of compounds as aliphatic, alicyclic, (hetero )aromatic, and various combinations, it should be computerized with an access to a large database of reliably measured log P values. Various "rules" governing the structural contribution to log P, and correction terms for various intramolecular interaction features specific to each series should be incorporated into the program software, and calculated values should be immediately compared with the measured values in the database whenever available. The software could be constructed so as to recognize structural features of compounds when it is input into the computer and to output the calculated log P value "automatically" after data processing according to the rules and corrections covering any types of structural features. There are quite a few programs based on this type of concept (Sangster, 1997b). One of them, the "fragmental method," was first proposed by Nys and Rekker (1973). They have been developing software for their own fragmental method (Rekker and Mannhold, 1992). With a procedure different from that used in the Rekker method, Leo and Hansch developed another fragmental method, the software of which is called CLOGP (Hansch and Leo, 1979, 1995; Leo, 1991,1993; Leo et aI., 1975). In this section, the CLOGP procedure is briefly described. In principle, the fragmental method is based on a "correlation" equation such as the following to elucidate the measured log P value: (18) Here, a is the number of occurrences of molecular fragment f of type n, and b is the number of occurrences of correction factor F of type m. The fragmental hydrophobicity index f differs from the IT value for a certain substituent X. For instance, the log P of chlorobenzene is conceptually expressible by either of

29.3 Nonexperimental Estimations of log P

the following equations, although the f(C6HS) value is algorithmically dealt with as being divisible into smaller fragments in the CLOGP: log P(PhCI) = log P(PhH) log P(PhCI) = f(C6HS)

+ n(CI/PhH),

+ f(CI).

(19) (20)

Because log P(PhH) in Eq. 19 is expressible as the sum of f(C6HS) and f(H), the relationship between the nand f values can be represented as n(CI/PhH)

=

f(CI, aromatic) - f(H, aromatic).

(21)

Table 29.2 lists the f values of some representative substituents as well as the F values of typical correction features (Leo, 1998). As indicated, an "isolating carbon" (IC) atom is defined as the one not doubly or triply bonded to a heteroatom. An IC can be bonded to heteroatom inside an aromatic ring, and one IC can be multiply bonded to another. The IC and hydrogens attached to it (ICHs) are considered hydrophobic fragments. All atoms and groups of covalently bonded atoms, which are left after removing ICs and ICHs, are considered polar fragments. The polar fragments do not contain ICs, but each is connected to ICs with one or more bonds. The f value is assigned first to each fragment. Depending on the bond environment, such as aliphatic and aromatic as well as benzyl and vinyl, the f value of polar monovalent fragments is found to vary, with the aromatic value highest and the aliphatic value is lowest. Vinyl and benzyl values are intermediate. This difference is probably due to variations in the degree of delocalization of electron pairs of the polar fragment. The correction factors, F, for six types of fragment interactions are noted in the aliphatic systems. When halogens, X, and hydrogen-bonding fragments, Y, are located either geminally or vicinally in such arrangements as X -C- X, X -CC- X, X-C-Y, X-CC-Y, Y-C-Y, and Y-CC-Y, the F value for "proximity polar" interactions is assigned to each of them as positive corrections with detailed rules depending on their structural features. For aromatic systems, electronic interactions among substituents modifying their hydrogen-bonding capability, such as those described in the preceding section, are incorporated after modifications/simplifications. The "ordinary" electronic effect of a certain substituent designated by the 0'°(0, rn, p) in Eqs. 16 and 17 is simplified/approximated here as being expressible as a single a value, some of which are listed in Table 29.2. The p value of substituents/fragments, which is not easily accessible directly by the analyses according to such correlation equation as Eqs. 13-17, is estimated/calculated so that each of the 0'/ P combinations (or pyaX products) makes up the difference between the measured value and the simple fragment sum. The p values of some fragments are also shown in Table 29.2. Provisions are also made for multiply substituted compounds, including heteroaromatic systems. Thus, the susceptibility p of each substituent to multiple interaction is not additive but attenuated starting with the greatest 0'/ P combination of fragments (substituents) regardless of whether they are present as substituents or as fragments fused in a heteroaromatic ring.

659

For ortho disubstitutions, besides the "ordinary" pO' interaction, a negative correction factor is assigned when the effect is regarded as a twisting of one of the substituents out of the ring plane, whereas a positive correction factor is defined for the internal hydrogen bond formation. In addition, the CLOGP program uses correction factors for the bond flexibility in aliphatic systems. For chain compounds, the total correction is made by a negative unit F value multiplied by the bond number connecting (or outside of) fragments minus unity (not counting those to H). For alicyclic compounds, the bond flexibility correction is made by a less negative factor. The branching structures in the alkane chain at ICs and at polar fragments are also considered as negative correction factors. Table 29.3 shows the CLOGP components of methomyl, an oxime carbamate insecticide (Leo, 1998). The polar fragment, including the amide H, is enclosed by the broken line, and three CH3 groups, each including an IC and three ICHs, connected to the polar fragment, are circled in the structural formula of methomyl. The CLOGP estimation is in a good agreement with that measured by Drabek and Bachmann (1983). Table 29.4 is that for the more complicated triazine herbicide, terbutryn (Leo, 1998). The six polar fragments numbered from 1 to 6 are shaded circles, whereas three aromatic ring ICs and seven aliphatic ICs with respective ICHs (five CH3, one CH2, and a quaternary C) are just circled. In Table 29.4, the f value of secondary (2°) amino (-1.030) and thio (0.030) fragments is that for the aromatic substituents. The f value of aromatic ring IC (0.130) is defined as being lower than that of aliphatic IC (0.195). In this molecule, the correction for the bond flexibility is made for each of the "side chains." The NH-tertiary butyl substituent is considered double-branched, whereas the NH-ethyl is a straight chain. Electronic interactions are considered between the three substituents and the three - N = fragments fused in the triazine ring as mentioned before. In terbutryn, the electronic interactions are assumed to be "faded with use" in terms of the ay px product. For instance, each of the fused "aza" nitrogens and secondary amino groups (X) undergoes the electronic effect ay of other fragments according to the susceptibility Px and its fading factor. The situation is illustrated in the lower half of Table 29.4. The ortho correction factor of 0.400 is empirically assigned because the measured log P value of 2-methylthiopyridine (Yamagami and Fujita, 1995) deviates positively from the CLOGP similar to this value. The value of 3.129 here for terbutryn is calculated using the newest version of the program (Leo, 1998) and differs from that in earlier publications, 3.73 (Hansch and Leo, 1995; Leo, 1991), but it is closer to the more recently measured value of 3.38 (Liu and Qian, 1995). With the preceding types of improvements for the estimation of the log P values of complicated compounds, the latest version of the CLOGP program seems to work very well. The correction factors are physicochemically as well as empirically reasonable enough. The elaboration in the treatment of the electronic interactions among aromatic substituents to assume the fading effect has been observed to also work well for the calculation of the log P of such candidate azole fungicides as shown

660

CHAPTER 29

Hydrophobicity as a Key Parameter of Environmental Toxicology

in Table 29.5 with the five-membered heteroaromatic system (Kataoka et al., 1989). Although a number of correction factors are required, this is a matter of course if one can understand that there are various types of interactions between solutes and solvents as well as between substructures within single molecules that can affect the log P value. Further elaboration is needed because a number of factors have yet to be taken into account (Leo, 1998). It is recommended not to evaluate the computer-aided system in terms of its computational efficiency/simplicity but to judge it in terms of the physical organic background of the program.

29.4 PHYSICOCHEMICAL SIGNIFICANCE OF log P IN ENVIRONMENTAL QUANTITATIVE STRUCTUREACTIVITY RELATIONSHIPS 29.4.1 BEHAVIOR IN SOIL

The behavior of pesticides in soil is governed by adsorption, movement, vaporization, and degradation (Arnold and Briggs, 1990). The adsorption of organic chemicals to soil and sediment plays a very important role in their transport and mobility in the environment. The adsorption is an exothermic process, and a strong adsorption is observed in a low temperature range. The soil adsorption coefficient, Kd, in a soil/slurry-water system is expressed by the ratio of the amount of compounds adsorbed in soil (!-lg/g soil) to the concentration in water (J.Lg/ml) at an equilibrium state under defined conditions. The soil adsorption of organic un-ionizable pesticides has been shown to be expressible by the hydrophobic/hydrophilic balance parameterized by the partition between organic solvent and water or the chromatographic retention (Hance, 1967). Thus, for a certain soil sample, the log Kd value is observed to be almost linearly correlated with the log P value for a related series of compounds. For instance, Uchida and Kasai (1980), using a local soil sample in Japan, obtained the following linear relationship between the log Kd values of a rice blast fungicidal isoprothiolane (VI: RI = R2 = i-Pr) and related compounds and their log P values: log Kd

= 0.41 (±0.06) log P

n = 12,

s = 0.150,

= 0.980.

, "-8 0N

(22)

Uchida et al. (1982b) also found that the log Kd values calculated from Eq. 22 for structurally unrelated buprofezin (VII), an insect growth regulator, and flutoranil (VIII), a rice sheath blight fungicide, agree well with experimentally estimated Kd values using the same soil sample as before. The basic mechanism of the soil adsorption of organic compounds from the water phase had been recognized earlier to be that involved in the distribution between soil organic matter and water (Goring, 1962). Thus, the "soil organic-carbon adsorption coefficient" or Koe should be normalized by the proportion of

N)=N-tett_Su

VI

VII

I CF~

CX;° N'V0'~

I

iso-Pr

~

VIIJ

organic components in soil samples, Foe, as determined in a separate experiment and defined by

= Kd/ Foe.

Koe

(23)

The definition assumes that pesticide adsorption by soils is entirely due to organic matter, even though the organic matter is a complex mixture of carbon, hydrogen, and nitrogen compounds, which acts as a nonpolar film coating soil particles. The Koe value is relatively constant for a particular compound among soil samples from different origins. Briggs (1981 b) reported that the Koe value of herbicidal phenylureas measured with English soils agrees well with that measured with Australian soils. Dzombak and Luthy (1984) showed that the Koe value of chlorobenzenes hardly depends on soil samples having various fractional organic matter ranging from 0.15 to 33%. Oliver (1987) determined Koe values for chlorinated hydrocarbons, including polychlorinated biphenyls with suspended sediments, collected from various districts in the United States, yielding log Koe

= O. 76 log P + 1.66,

n = 19,

s = 0.27,

r =

(24) 0.93.

These relationships for a number of individual series of compounds were comprehensively examined by Sabljic et al. (1995). They derived the following equation for a number of organic compounds belonging to such classes as acetanilides, carbamates, phenylureas, phosphates, triazines, triazoles, and uracils, including practically used agrochemicals: log Koe = 0.47 log P

- 0.40(±0.24),

r

0)-- ,iso-Pr

n = 216,

s

+ 1.09,

= 0.425,

(25) r =

0.826.

In Eq. 25, log Koe varies from 0 to 5 and log P from -I to 8. They proposed that this equation can apply to the prediction of log Koe values of nonmeasured agrochemicals irrespective of origin of the soil samples. In fact, Eq. 25 is very similar to the following equation, which was derived for phenylurea-type herbicides independently (Liu and Qian, 1995) using soil samples from China: log Koe n =9,

= 0.59(±0.15) log P + 1.18(±0.43), s = 0.152, r = 0.961.

(26)

29.4 Physicochemical Significance of log P in Environmental Quantitative Structure-Activity Relationships

The mobility of pesticides in soil is important in governing their persistency as well as downward movement to pollute the ground water and lateral movement to pollute surface water. Mobility can be estimated using soil column chromatography or soil thin-layer chromatography. Uchida and Kasai (1980) studied the mobility of isoprothiolane and its analogs (VI) using a soil column chromatography and expressed it as log f.l. , where f.l. is the ratio of the volume of soil packed in the column to the volume of the aqueous phase required to elute the solute i.e., f.l. = Vsoil/ Veluent. They derived the following equation: log f.l.

= -1.44(±0.17) 10gKd - 0.45(±0.33),

n = 11,

s

= 0.193,

r

= 0.986.

(27)

Equation 27, together with Eq. 22, indicates that the greater the log P, the lower the mobility of the compounds in soil. Briggs (1973) obtained similar results for herbicidal phenylureas with soil thin-layer chromatography. Their mobility in tenns of the RM (Boyce and Milborrow, 1965) nonnalized by the content of organic matter was related to the log Koc value irrespective of the source of soils. Helling (1971) measured the mobility of ionic/ionizable herbicides such as dicamba, 2,4-dichlorophenoxyacetic acid (2,4-D), fenac, picloram, and diquat by thin-layer chromatography with 14 kinds of soil samples. The mobility of ionic!ionizable compounds was not simply related to the log Koc value. Their soil adsorption mechanism is not a simple hydrophobic partitioning into the soil organic matter from the aqueous phase, but includes various interactions. According to Wauchope et al. (1992), these interactions include (a) binding of cations to negatively charged sites on clay surfaces (a very strong interaction), (b) binding of anions to soil anion-exchange sites (a very weak interaction), and (c) specific chemical binding mechanisms such as the phosphate-fixation-like binding of glyphosate and the arsenicals to soil metal oxides. In many cases, anionic and cationic pesticides, which give very low and very high Kd values, respectively, have no reported soil adsorption values, probably because the extreme values involved are difficult to measure. The vapor losses of volatile pesticides from soil to air depend primarily on the air/soil distribution constant, KAS. Vaporization from water is conveniently estimated by the Henry's law constant, KAW (air/water). In the soil/water/air system, the soil adsorption lowers the concentration of pesticides in the water phase and, in turn, in air. The air/wet soil distribution could be approximated by the ratio of KAW and a soil/water distribution constant such as Koc. The KAS value is thus a function of water solubility, vapor pressure, and soil adsorption. According to Arnold and Briggs (1990), the KAS constant is roughly expressible as a function of boiling point and log P value under defined conditions for such factors as the air/soil ratio, distribution of the pesticide in the soil sample, and climate. To a first approximation, pesticides in soil exhibit an exponential degradation according to the first-order kinetics. The degradation half-life, dTso, can be estimated from the reciprocal of the first-order rate constant. Degradation in soil occurrs,

661

however, as a combination of mechanistically complex processes. It generally includes abiotic and biotic processes. The abiotic degradation is due to chemical reactions such as hydrolysis, photolysis, air oxidation, and others. Depending on the structural feature, pesticides could suffer from various types of reactions. No common parameter such as log P alone is capable of describing a variety of reactivities. The reaction rate can be evaluated if experimentally established model systems are available. For biotic degradation of organic compounds, there have been quite a few efforts to establish the QSAR model correlation equations. With the use of acclimated mixed microbial cultures, Babeu and Vaishnav (1987) measured the 5-day BOD (biological oxygen demand in mmol/mmol chemical) of a wide variety of organic compounds, including aIcohols, acids, esters, ketones, and aromatics. They examined the correlation of 10g(BOD) with various physicochemical parameters for 45 compounds. Their analysis, showing that the 10g(BOD) values fit well a correlation equation with quadratic tenns of log(theoretical BOD), seems to be reasonable (n = 45, r = 0.862) among others. There would be an optimum theoretical BOD value for compounds to be most biodegradable. Molecules in which the number of carbon atoms is high necessarily have a high theoretical BOD value. They are often highly hydrophobic. The highly hydrophobic compounds, being trapped by cell membrane lipids, would not be easily incorporated into microbial cells. In fact, using the BOD data of Babeu and Vaishnav, Zakarya et al. (1993) fonnulated a correlation equation showing that the experimental BOD values are parabolically related to log P and linearly related to molecular volume (n = 43, s = 0.575, r = 0.906). In contrast to the preceding studies, Dearden and Nicholson (1986) discussed the significance of the electronic structure of the molecule. The 5-day BOD value for various types of compounds, including amines, phenols, aldehydes, acids, halogenated hydrocarbons, and amino acids, supplied by the U.S. Environmental Protection Agency, Duluth, Minnesota, was shown to be highly dependent on a new electronic parameter that is expressible as the difference in the modulus of atomic charge across a key bond in the molecule, which could be attacked by microbials (n = 79, s = 3.459, r = 0.993). In spite of the fact that the biodegradability is estimated under simplified conditions without soil, the QSAR model building needs much improvement. In reality, assignment of a single dTso value to each pesticide is impossible under field conditions. Although it could be related to dTso values from model systems, it is also highly sensitive to the type of soil with varying mineralogy, carbon and water content, and pH, and the distribution and activity of soil microbials as well as climate. Thus, it is almost impossible to build a comprehensive single QSAR model for the degradation of a variety of pesticide classes. Careful studies should be made under defined conditions perhaps on a series-to-series basis.

662

CHAPTER 29 Hydrophobicity as a Key Parameter of Environmental Toxicology

29.4.2 BIOACCUMULATION

Processes involved in the accumulation of environmental chemicals in various organisms through aquatic phases are generally classified into two types: bioconcentration and biomagnification (Connell, 1988). Bioconcentration is the accumulation of chemicals dissolved in water in fish and aquatic organisms through the gills and body surface directly. The bioconcentration factor (BCF) is defined as the ratio of the concentration of a chemical in an aquatic organism to that in the aqueous phase under steady-state conditions. The measurement of the BCF has been made with the average concentration of the chemical in the whole body absorbed through the gills, skin, and digestive tract of fish of small to moderate size reared in a sublethal aqueous concentration. Sometimes, the BCF is estimated as that for the lipid content of fishes. Following the work of Neely et al. (1974) showing that the log value of the BCF of nonpolar compounds can be correlated with their log P value, a number of examples have been accumulated. Mackay (1982) critically reviewed the BCF values and proposed that the fundamental process observed in bioconcentration is such that P and BCF are proportional; that is, the slope of the log BCF versus log P correlation should "theoretically" be close to unity. In the earlier publications (Neely et al., 1974; Veith et aI., 1979), the slope was often lower than unity, even when dealing with nonpolar compounds. Mackay suggested that the lower slope was the result of overestimation (by calculation) of the log P of compounds of high molecular weight. After omitting compounds, the log P value of which are above 6 or are unreliable, compounds that are ionizable, and compounds that can act as surfactants, Mackay (1982) proposed the following equation for 43 compounds. These are mostly nonpolar such as lindane, DDT, and (polyhalogenated) aromatic hydrocarbons with log P values ranging from unity to 6. 10gBCF -log P n

= 43,

r

= -1.32(±0.25),

= 0.974.

(28)

Equation 28 is valid as far as inert compounds having log P values lower than 6 are concerned. It also indicates that variations in fish species, with which BCF is measured, are insignificant in the general relationship between log P and log BCF values. The lipid content of the fish used in developing Eq. 28 has been estimated as being 5% (Connell and Hawker, 1988) and does not vary significantly among fish species. There are quite a few examples conforming to Eq. 28. Oliver and Niimi (1983) determined the BCF of chlorobenzenes in rainbow traut (Salmo gairdneri) as log BCF = 1.022(±0.057) log P - 0.632, (29) n =

11,

r = 0.993.

In this and other correlation equations (Connell, 1988; Davies and Dobbs, 1984; Isnard and Lambert, 1988; Oliver and Niimi, 1985; Opperhuizen et aI., 1985), the slope of the log P term

is indeed close to unity and the intercept ranges from -0.5 to -1.3. Considerable deviations from the linear relationship represented by Eqs. 28 and 29 have been observed, however, for highly hydrophobic compounds with a log P value > 6 (Bruggeman et aI., 1984; Opperhuizen et aI., 1985). Reduced membrane permeation (Opperhuizen et aI., 1985), lowered lipid solubility (Banerjee and Baughman, 1991), and other possible reasonings have been proposed. To simulate this situation better, Bintein and Devillers (1993) proposed the following equation using the bilinear model developed by Kubinyi (1977) for a number of compounds, roughly one-third of which belongs to pesticides: 10gBCF

= 0.91OlogP -1.97510g(6.8 x 10-7 P n = 154,

r

= 0.950,

+ 1) s

(30)

0.784,

= 0.347,

F

= 463.5.

In Eq. 30, F is the ratio of regression and residual variances. The compounds were selected so that they are mostly inert and cover a wide range of log P values, ranging from unity to 9. In the second term on the right-hand side, conventionally written as b[log(,8 P + 1)], the ,8 value is supposed to correspond to the volume ratio between lipid and aqueous phases involved in the entire system for the manifestation of biological "activity." For small P values, (,8 P + 1) is close to unity, so 10g(,8 P + 1) is 0, and Eq. 30 takes the form of Eqs. 28 and 29. For large P values, (,8 P + 1) is almost equal to ,8 P, so that 10g(,8 P + 1) is linear with log P. The value of -log,8 nearly corresponds to the log P value where the log BCF is maximum. Biphasic functions with linear ascending and descending sides and a rounded apical part are represented by this model. In Eq. 30, the positive slope for the ascending side is 0.910, and the negative descending slope is (0.9lO - 1.975 = -1.065). Among the compounds included in Eq. 30, 24 compounds have a log P higher than 6; that is, they are covering a part of the apical region and the descending phase. In measuring the BCF values of compounds having a large log P value, one needs to carefully set the test period. Oliver and Niimi (1983) showed that 120 days are needed to attain the equilibrated steady state for polychlorobenzenes in fish. Hawker and Connell (1985a, b) suggested that half a year may be required to obtain the steady state for compounds with a log P of around 6 and about 10 years for those with a log P of about 8, and formulated a QSAR similar to Eqs. 28 and 29 for the BCF value under nonequilibrium conditions with a certain exposure period. Devillers et al. (1996) compared in detail the versatility of the bilinear model expressed by Eq. 30 with that of linear, quadratic, and polynominal correlation models published from different organizations. They selected 342 log BCF values for 181 compounds, which are mostly inert. Some values were independently measured in duplicate for a single compound. The selection criteria for these log BCF values required that the BCF data are obtained only after a steady state was established and


that, if one or more values appeared out of line in a publication, all the data contained in that publication are not used. They concluded that the bilinear model represented by Eq. 30 is among the best in predictive performance in terms of the root mean square (rms) value for residuals between log BCF values experimentally measured and calculated from model equations. Banerjee and Baughman (1991) proposed another model for the nonlinearity of log BCF versus log P. They tried to rationalize the breakdown of the linear relationship not only for the highly hydrophobic compounds, but also for many azo dyes (multifunctionalized azobenzenes), the log P value of which is below 6 (mostly between 3.5 and 4.5) (Anliker and Moser, 1987). Their model considered the fact that large compounds of low lipid solubility such as azo dyes and polyhaloaromatic hydrocarbons, have lower than expected BCF values because of the difficulty in cavity formation in lipids. Thus, with an approximation in which the lipid solubility of compounds (generally unavailable) is replaced with the solubility in octanol, Soct (M), they proposed the following equation for a set of compounds. These include inert pesticides, aromatic and aliphatic (halogeno)hydrocarbons, and polar but mostly nonionized azo dyes, the log P values of which ranged from 1.5 to 8.3. log BCF

= -1.13 + 1.02 log P +0.84 log Soct + 0.0004(mp - 25), n

= 36,

(31)

r = 0.95.

In Eq. 31, mp is the melting point (in 0c), which was intended to allow octanol solubilities for both liquids and solids to be included in a single equation (Valvani and Yalkowsky, 1980). For liquids, mp is regarded as 25 to remove the entire term. For small compounds, the log Soct and (mp - 25) terms in Eq. 31 tend to be constant and Eq. 31 takes the same form as Eq. 29. Most of the compounds included in Eq. 31 are solids. Without the log Soct and (mp - 25) terms, the correlation was much poorer (r = 0.73). In the region of log P > 6, even though the log P increases, the BCF value could decrease because of a decrease in the lipid (octanol) solubility, resulting in the descending phase in the log BCF/ log P relationship. The previous relationships seem to be in accord with the low fish toxicity of an insecticide of nonester-type pyrethroids, silaftuofen (IX) (Sieburth et aI., 1990). The log P value of this compound has been estimated as being about 10 (Okimoto et aI., 1994). Its uptake in fish should well be very low.

663

1989; Oliver and Niimi, 1985; Opperhuizen and Voors, 1987). Such deviations from Mackay's postulate (Mackay, 1982) have been attributed to relatively high rates of biotransformation (de Bruijn and Hermens, 1991; Opperhuizen and Voors, 1987; Southworth et aI., 1980). Uchida et al. (1982a) measured the BCF value of isoprothiolane and its analogs (VI) with the killifish (Orizias latipes) and proposed the following: 10gBCF = 0.65(±0.17)logP -1.17(±0.62),

s

n =9,

= 0.197,

r

= 0.962,

F

(32)

= 85.5.

The reason that the coefficient of the log P term is smaller than unity is attributable to the fact that the test compounds have two hydrolyzable ester groups. Uchida et al. also measured the rate of disappearance of the compounds in the entire system. The rate followed approximately the zero-order kinetics, and the rate constant, logk, tended to increase with decreasing bulkiness of the ester substituents in terms of the STERIMOL width (Verloop, 1983) and with increasing log P value. This was taken to indicate that the hydrolytic degradation would occur under conditions "reacting" with fish. The "apparent" log BCF term in Eq. 32 should be compensated for the degradation effect corresponding to the log k value so that the size of the log P term under the "real" conditions would be closer to unity. de Bruijn et al. (1993) examined the effect ofbiotransformation of a set of organophosphorus insecticides on bioconcentration. The compounds belonged to O,O-dimethyl-O-phenyl phosphorothioates in which various substituents such as CN, N02, SMe, and halogens are located at the 2-, 4-, and 5positions of the benzene ring either singly or multiply. Their log BCF values were measured using guppies, from which the following equation was derived: 10gBCF = 0.80(±0.12)logP +0.45, (33) n =

12,

s = 0.35,

r = 0.910.

One of the most important metabolic pathways of dimethyl phosphorothioates has been shown to be the demethylation of one of the two methyl groups by glutathione (Fukami and Shishido, 1966). Thus, the rate of demethylation under pseudofirst-order conditions, k (min- l mg protein-I), was measured using a glutathione S-methyltransferase preparation. Introduction of the log k term into Eq. 33 yielded the following: log BCF = 0.94(±0.08) log P -0.63(±0.14) logk - 2.31,

n = 12,

For less inert compounds, including pesticides in which certain reactive/vulnerable functions are required for their biological activity, the slope has been observed to be significantly lower than unity. Moreover, the correlation is often of a lower statistical quality (de Wolf et aI., 1992; Niimi et aI.,

s = 0.21,

(34) r =

0.971.

The improvement in the correlation quality is significant and the slope of the log P term is close to unity in Eq. 34. The preceding two examples indicate that, if an appropriate term for the vulnerability is incorporated, then Mackay's postulate is valid even for the unstable series of pesticides. Mackay's postulate has also been observed for bioconcentration in mollusks, daphnias, and aquatic microbes (Baughman and Paris, 1981; Geyer et al., 1982; Hawker and Connell, 1986).

664

CHAPTER 29


The structure-activity relationships for biomagnification are not as well established as those for bioconcentration. A number of mechanisms, which are not well understood, are involved in the entire process of biomagnification, which occurs through the food chain. However, it is highly probable that the biomagnification is also able to be significantly related to the octanol/water partition coefficient (Esser, 1986). Davies and Dobbs (1984) indicated that the uptake of chemicals from both food and water results in tissue concentrations comparable to those resulting from water alone. It should be mentioned that the uptake of pesticides from food is far less important than the uptake from water and that only a part of the residue present in the lower level biota is transferred to the higher level of the food chain (Ellgehausen et aI., 1980).

(halogen-substituted) a1cohols, esters, ketones, phthalates, substituted benzenes, and several pesticides (mostly rather stable herbicides). For inert organic compounds, the log P value of which is very high (>5-6), the pLC50 value has been observed to be lower than expected by such linear correlations as Eqs. 35 and 36, probably because of the limit in the water solubility (solubility cutoff) to be partitioned into the fish body required for the symptom of toxicity. This situation was confinned by Veith et al. (1983), who formulated a bilinear correlation of excellent quality such as the following for the 4-day toxicity to the fathead minnow of five classes of unreactive compounds similar to those included in Eq. 36: pLC 50 (M) = 0.94 log P

-0.94log(6.8 x 10-5 P

29.4.3 AQUATIC TOXICITY

n = 65,

The aquatic toxicity of "simple" organic compounds has been recognized as being closely related to their lipophilicity (hydrophobicity). Overton and Meyer independently proposed the "lipoid theory of narcosis" about 100 years ago (Lipnick, 1989), narcosis being considered as a toxic effect that can be lethal. The modern fonnulation of the aquatic toxicity in terms of the QSAR was, however, first published by Hansch and Dunn (1972). They found that the earlier toxicity data of various sets of homologous a1cohols or miscellaneous inert compounds, in terms of the minimum narcotic and the minimum lethal concentrations, can mostly be explained by a single parameter log P with a slope close to unity. Subsequently, a number of hypotheses of aquatic toxicology were combined with the QSAR concept. Thus, K6nemann (1981) investigated the QSAR of "environmental pollutants" of structurally heterogeneous organic compounds such as methyl- and chlorobenzenes, aliphatic chlorohydrocarbons, and a1cohols, among others. For the 14-day LCso (M) value in the guppy (Poecilia reticulata), the following equation was derived: pLCso n =

= 0.871 log P + 1.13,

50,

r

= 0.988,

(35) s = 0.237.

In this set of compounds, the pLC50 (M) value varies from 6.15 for pentachlorobenzene (log P = 5.69) to 0.24 for diethyleneglycol (log P = -1.30). The selection of compounds included in Eq. 35 was intended so that they are rather stable and unreactive (inert) as well as un-ionizable. Thus, pesticides that are "reactive" to "specific" targets in vivo are not included. With a larger (4 day) toxicity data set against juvenile fathead minnows (Pimephales promelus) from the V.S. Environmental Protection Agency, Duluth, Minnesota, the following equation was derived (McCarty et aI., 1992): pLC50 (M) = 0.90(±0.04) log P n = 150,

r = 0.959.

+ 1.29(±0.12), (36)

The compounds included in Eq. 36 are inert and cover the range of log P similar to that in Eq. 35. They are rather stable, belonging to halogenated aliphatic hydrocarbons, ethers,

+ 1) + 1.25,

(37)

r = 0.999.

Although the detailed mechanism of the acute fish toxicity is not completely clear, the observed symptoms caused by inert and unreactive compounds strongly suggest a mechanism categorized as general narcosis (Lipnick, 1990). In Eqs. 3537, the coefficient of the log P term is close to unity and the intercepts are almost equivalent to one another irrespective of the test organism and the range of test compounds. A number of similar correlations have been accumulated for sets of inert compounds (Cronin and Dearden, 1995; Ikemoto et aI., 1992; Lipnick, 1990). Thus, the general narcotic (analgesic) potency is dependent only on the overall hydrophobicity and not on the specificity of the chemical structure. Because every compound is supposed to exert at least this type of narcotic activity, the QSAR correlations similar to Eqs. 35-37 are considered to predict the minimum toxicity or the "baseline toxicity" (Lipnick, 1990) of any organic compound to aquatic biota, unless the compound is biodegradable and therefore less toxic. The toxic effects of mixtures of compounds from a single group in terms of the mechanism (type) of toxicity should be concentration additive. Concentration addition means that the LC50 of a mixture is observed at the sum of concentrations, c, of individual compounds (as the fraction of their own LC50) being 1.0 (L c /LC50 = 1.0). The compounds included in Eq. 35 have been shown to be concentration additive, that is, to exert their effect by an equivalent mechanism (K6nemann, 1980). Veith and Broderius (1987) noted that a number of compounds that appear to produce narcosis are significantly more toxic than the baseline toxicity. They are more polar and often have weakly acidic and/or hydrogen-bond donor groups, such as substituted phenols, mostly existing as the nonionized fonn under experimental conditions, and substituted anilines. This structurally heterogeneous set of compounds has been examined by the concentration additivity test, showing that these phenols and anilines indeed belong to a group exerting a narcotic syndrome [Type 11 (or polar) narcotic syndrome] differing from that of the unreactive inert compounds [Type I (or nonpolar) narcotic syndrome]. For phenols and anilines substituted both singly and multiply, by alkyl, alkoxy, halogen, N02, the


following equation was formulated with the 4-day LCso value against juvenile fathead minnows: pLCso(M) = 0.65(±0.07) log P n =39,

r

+ 2.29(±0.22),

= 0.95.

(38)

Prior to this, Saarikoski and Viluksela (1982) analyzed the 4-day LCso value of variously substituted phenols against guppies (P. reticulata) at pH 6, 7, and 8. The substitution pattern of phenols was carefully selected so that the collinearity between the log P and 1'1 log Ka values is as low as possible. 1'1 log Ka, the difference between the log Ka and the reference log Ka of unsubstituted phenol, was used as a parameter for the electron-withdrawing effect of substituents. They also corrected for the effect of ionization on the LCso value in terms of an "effective" concentration of the neutral form defined empirically (Saarikoski and Viluksela, 1981). For phenols substituted with Me, Cl, t-Bu, OMe, OH, and N02 groups singly or multiply, Saarikoski and Viluksela derived the following equations: pLCso (M; corrected for pH 6) = 0.67 log P + 0.191'1 log Ka n =

19,

r = 0.985,

+ 2.33, s

(39)

= 0.17,

pLCSO (M; corrected for pH 7)

= 0.71 log P + 0.191'1 log Ka + 2.23, n = 19,

r = 0.978,

(40)

s = 0.21.

Because Eqs. 39 and 40 are practically identical, they can serve to predict the toxicity at any pH from 6 to 8. It is interesting to note that, for phenols, the 1'1 log Ka value of which is so low that the effects of ionization on the LCso and electron withdrawal of substituents are not great, Eqs. 39 and 40 are very similar to Eq. 38. For phenols with more acidic hydrogen than anilines, the electron-withdrawing effect of substituents on the OH, which may hydrogen-bond with possible taget site(s), seems to be accounted for explicitely in Eqs. 39 and 40. Equations 38-40 probably reflect the feature of Type 11 (or polar) narcotics, in which the slope of the log P term is lower and the intercept is higher than those of the nonpolar narcotics. The higher intercept may result from the hydrogen-bond donor group in phenols and anilines enhancing the interaction with the target site(s). The lower slope could mean that there is an increasing shift toward nonpolar narcosis with increasing hydrophobicity. For nonpolar narcosis, the slope is higher. However, the toxity of non polar narcotics is lower than that of polar narcotics in the region of low log P values. Thus, the shift of the narcosis type could make the slope higher. A possible reason for this may be related to differences in distribution so that compounds with a high log P value will show an increasing tendency to accumulate in the lipid phase (Cronin and Dearden, 1995). In Eq. 38, such phenols as Cls- and 2,4-(N02h-derivatives are not included because these compounds do not share the concentration additivity with those included. In Eqs. 39 and 40, phenols with 2,5- and 2,4-(N02h-substitution patterns are not included because of their much higher toxicity than that

665

predicted. The structural characteristics of these outliers are shared by the pesticidal phenols mostly acting as uncouplers with the mitochondrial oxidative phosphorylation. They are neither Type I nor Type II narcotics, but exert a toxicity about lO-fold higher than that predicted by the Type II correlation equations. A number of compounds with biologically and/or chemically "reactive" sites have toxicities considerably greater than those predicted for either nonpolar or polar narcosis. Various pesticides with specific structural features and/or specific functional moieties to inteact with respective target site(s) are such compounds not categorized into narcotics. These compounds are often defined to show "reactive toxicity" (Hermens, 1990). Ikemoto et al. (1992) examined the excess toxicity exerted by various pesticides against killifish (Oryzias lapipes) over the baseline toxicity which was formulated for nonpolar inert compounds such as homologous alkanols, chlorobenzenes, and alkylbenzenes. Photosynthesis-disrupting herbicides, such as diuron, and chitin-synthesis-disrupting larvicides, such as diflubenzuron (IV: Xl = X2 = F, Y = Cl) and buprofezin (VII) were almost on the line for the nonpolar narcotic compounds. Their targets do not exist in fish. Such neurotoxic insecticides as DDT, dieldrin, fenvalerate, and lindane are more toxic than the baseline by 1.2-2.5 log units, and the respiration-inhibiting rotenone is higher than 3 log units more toxic. The neurotoxicity and respiratory toxicity are probably common between fish and insects. Besides the specific biochemical mechanisms exerted by pesticides, a variety of chemical reactivity mechanisms such as electro- and nucleophilic, redox, and free-radical processes are thought to be involved in the interactions of various types of toxicants with biological systems. Especially important are toxicants expected to work as electrophiles which can react with nucleophilic groups such as NH2, OH, and SH in deoxyribonucleic acid (DNA) and proteins. Hermens (1989) classified the reaction mechanisms of nucleophilic groups in biological systems with electrophilic toxic ants into (a) nucleophilic displacement reaction, (b) addition to carbon-oxygen double bonds, and (c) addition to activated carbon-carbon double bonds (the Michael-type addition). He also surveyed various molecular substructures present in possible toxicants where these types of reactions might be responsible for their unwanted activity (Hermens, 1990). In this type of reactive toxicity, one should not expect simple relationships between toxicity and hydrophobicity even when accompanied by electronic parameters such as those represented by Eqs. 35-40. For instance, Hermens et al. (1985) derived the following equation for the 14-day toxicity to guppies of 15 alkyl, alkenyl, acylmethyl, and benzyl halides: pLCso(M) = 0.224 log P - 1.3210g(2484 + k- l ) + 10.05, n

= 15,

r

= 0.956,

(41)

s = 0.39.

The pLCso value was only very poorly correlated with the log P value alone (r = 0.41), but the correlation was much improved

666

CHAPTER 29


when the pseudo-first-order rate constant, k (in min- 1), with a model nucleophile (4-nitrobenzylpyridine) was included. The second term on the right-hand side of Eq. 41 means that the variation in the toxicity is biphasic with respect to log k. For compounds with very low reactivity, that is, for those with very small k or very large k- 1 values, the value of 2484 within the parenthess can be neglected relative to the k- 1 value. Therefore, the value of the negatively signed second term, being approximated by + 1.3210g k, initially increases nearly linearly with increasing logk, but, in the region above logk > -3.4 (k > 1/2484), follows a plateau-like pattern. Examples of the use of experimentally estimated reactivity indices have been reviewed by Hermens (1990), who also analyzed the 14-day LCso values of epoxides and aldehydes and formulated correlations somewhat similar to Eqs. 38 and 41, respectively. In the reactive toxicants, the term "reactive" encompasses a wide spectrum of chemical processes as mentioned previously. Further classification must be made for hard (charge controlled) and soft (orbital controlled) interactions (Comporti, 1989) because relevant molecular descriptors are different between these interactions. Soft electrophiles cover one of the largest groups of anthropogenic toxic ants exceeded only by narcotics. Soft electrophilic interactions of toxicants with biomolecules are regulated by the ability of toxicants to accept electron density represented by the "nucleophilic" or electron-acceptor superdelocalizability, SN (the term "nucleophilic" is from the side of "biological" reactants), through the orbital interactions. The SN value on the activated unsaturated carbon atoms and the frontier charge of the lowest unoccupied molecular orbital, f(LUMO), along with the log P value, are important in predicting the toxicity of such soft (pro )electrophiles (Mekenyan and Veith, 1993). The "proelectrophilic" mechanism of the toxicity refers to compounds that are not direct-acting electrophiles but are metabolized to electrophiles such as primary and secondary allyl and propargyl alcohols (Lipnick et aI., 1985). Veith and Mekenyan (1993) examined to extend this approach to a large set of n -electron systems such as aromatic compounds, including variously substituted hydrocarbons, phenols and ani lines, and unsaturated aliphatic compounds, including alkenols and alkynols. These were selected to represent a wide variation in n bonding and a variety of modes of toxic actions. The compounds include polar and nonpolar narcotics, uncouplers with oxidative phophorylation, and "reactive" electrophiles and proelectrophiles. Veith and Mekenyan derived the following equation for the 4-day LCso value of 114 compounds against fathead minnow: pLCso(M) = -1.49(±0.53) + 0.56(±0.04) log P +13.7(±1.7)SN(av.), n

= 114,

r

= 0.90,

s

= 0.43,

(42)

toxicity of proelectrophiles, when the SN(av.) value is calculated for their metabolites. The authors (Mekenyan and Veith, 1993) further showed that the soft electrophiles included in Eq. 42 can be clustered according to their SN values by defining "isoelectrophilic windows" along the toxicity response plane. Nonpolar narcotics are located in the lowest SN region where toxicity varies almost only with hydrophobicity. Polar narcotics are more toxic than nonpolar narcotics at similar values of log P and the toxicity increase can be illustrated by higher SN values (by stronger electronic interactions with cellular soft nucleophiles). Highly reactive soft electrophiles, which have dissociable protons, act as uncouplers. Electrophiles without dissociable protons indicate symptoms of reactive toxicity consistent with covalent bonding. Perhaps because of their high specificity, the aquatic toxicology of pesticides has not been studied intensively in terms of the QSAR. There have been studies for sets of miscellaneous pesticides, but their toxicity is usually dealt with as that of QSAR outliers with higher potency than narcotic toxicants (Ikemoto et aI., 1992). Exceptions are the organophosphorus insecticides (Hermens et aI., 1987; Schiiiirmann, 1992) where a series of O,O-dimethyl phenylphosphorothioates are substituted at the 2-, 4-, and 5-positions on the benzene ring by various groups such as H, CN, N02, Me, halogens, and SMe. The 14-day LCso toward guppies were measured and analyzed for 12 analogs to give the following (Verhaar et aI., 1994): pLCso(M) = 0.38(±0.12) log P - 27.9(±7.78)p(PO) + 11.0(±5.43), n = 12,

r

s = 0.371,

F

= 12.9.

In Eq. 43, p(PO) is the bond order between the central phosphorus atom and the phenoxy oxygen calculated by MOPAC 6.0 (Stewart, 1990). This descriptor was believed to contain information about the strength of the corresponding bond reflecting the nature of the phenoxide as the leaving group and thus to have relevance for the phosphorylation step of acetylcholinesterase inhibition. The negative sign of this term suggests that the lower the PO bond order, the more easily the phosphorylation of the serine OH of acetylcholinesterase occurs, leading to the higher toxicity. The preceding study originally analyzed the LCso data with the use of experimentally estimated reactivity indices such as the rate constant k with the model nucleophile, 4nitrobenzylpyridine (Hermens et aI., 1987). The following equation was fotmulated for 10 analogs (two compounds were added later to formulate Eq. 43): pLCso(M)

= 0.34(±0.1O)

L:n

+0.76(±0.19) logk - 3.57,

F = 238.7.

SN(av.) is the superdelocalizability averaged for the values assigned to atoms included in the conjugated n -electron system. It is expected to represent the global electron-acceptor character of the molecule. The log P and SN (av.) parameters are orthogonal for these compounds. Equation 42 accurately predicts the

= 0.861,

(43)

= 0.29. same set of compounds, the use of L CT n

= 10,

r = 0.912,

(44)

s

instead of With the log k yielded a poorer correlation. The significant contribution to toxicity by the demethylation rate, k, can be taken as a possible (but not definite) hint that the in vivo demethylation may be

References

involved in the acute fish toxicity of this series of compounds (Schtitirmann, 1992). There are an increasing number of aquatic toxicity QSAR publications other than those described previously in which quantum-chemical descriptors are used to illustrate the electronic mechanism involved (Cronin et aI., 1995; Schultz et al., 1995). The examples described in this chapter are those in which the QSAR procedure is successuful in building models to elucidate the acute aquatic toxicity of the rather limited number of compounds included in the individual sets. The correlation equations could predict the toxicity of nontested compounds with related structures. However, there are a vast number of miscellaneous compounds the environmental behavior of which should be evaluated. Not only because of time constraints, but also because of limited financial resources, it is impossible to test all of the existing chemicals experimentally. For commodity chemicals, pharmaceuticals, and pesticides, it is better to predict adverse (environmental) toxicological behaviors before synthesis. In many of the preceding examples, the mode (or type) of aquatic toxicological action was more or less preestablished or categorized without much difficulty. For a large number of compounds, ranging from (non)polar narcotics, to reactive toxicants, and to those acting with specific biochemical mechanism, the ab initio identification of the mode of action becomes very difficult. Thus, attempts have been made to "identify" the mode of toxic activity according to substructures that are expected to participate in certain specific types of reactivity (Karabunarliev et aI., 1996a, b; Mekenyan and Veith, 1994; Verhaar et aI., 1992). The procedure is computerized with incorporation of the substructure search algorithm into expert systems. For sets of compounds searched for from the large data set, quantum-chemical descriptors are used to analyze the toxicity data quantitatively along with such log P values estimated with use of the CLOGP method (Leo, 1993).

ACKNOWLEDGMENTS The authors would like to dedicate this chapter to Professor Corwin Hansch of Pomona College, Clare mont, California, on his 80th birthday. They would like to express their sincere thanks to Dr. Albert Leo of Pomona College for his invaluable discussions as well as for his careful review of the manuscript.

REFERENCES Akamatsu, M., and Fujita, T. (1992). Quantitative analyses of hydrophobicity of di- to pentapeptides having unionizable side chains with substituent and structural parameters. J. Pharm. Sci. 81,164-174. Akamatsu, M., Yoshida, Y., Nakamura, H., Asao, M., Iwamura, H., and Fujita, T. (1989). Hydrophobicity of di- and tripeptides having unionizable side chains and correlation with substituent and structural parameters. Quant. Struct.-Act. Re/at. 8, 195-203. Anliker, R, and Moser, P. (1987). The limits of bioaccumulation of organic pigments in fish: Their relation to the partition coefficient and the solubility in water and octanol. Ecotoxico!' Environ. Sa! 13,43-52.

667

Amold, D. J., and Briggs, G. G. (1990). Fate of pesticides in soil: P.redictive and practical aspects. In "Environmental Fate of Pesticides" (D. H. Hutson and T. R Roberts, eds.), pp. 101-122. Wiley, New York. Avdeef, A. (1991). Fast simultaneous determination of log P and pKa by potentiometry: para-Alkoxyphenol series (methoxy to pentoxy). In "QSAR: Rational Approaches to the Design of Bioactive Compounds" (C. Silipo and V. Vittoria, eds.), pp. 119-122. Elsevier, Amsterdam. Avdeef, A. (1992). pH-metric log P. 1. Difference plots for determining ionpair octanol-water partition coefficients of multiprotic substances. Quant. Struct.-Act. Relat. 11,510--517. Babeu, L., and Vaishnav, D. D. (1987). Prediction of biodegradability for selected organic chemicals. J. Ind. Microbiol. 2, 107-115. BaneIjee, S., and Baughman, G. L. (1991). Bioconcentration factors of lipid solubility. Environ. Sci. Technol. 25, 536-539. Baughman, G. L., and Paris, D. F. (1981). Microbial bioconcentration of organic pollutants from aquatic systems: A critical review. Crit. Rev. MicrobioI. 205-227. Bintein, S., and Devillers, J. (1993). Non-linear dependence of fish bioconcentration on n-octanollwater partition coefficient. SAR QSAR Environ. Res. 1, 29-39. Boyce, C. B. c., and Milborrow, B. V. (1965). A simple assessment of partition data for correlating structure and biological activity using thin-layer chromatography. Nature 208, 537-539. Braumann, T. (1986). Determination of hydrophobic parameters by reversalphase liquid chromatography: Theory, experimental techniques, and application in studies on quantitative structure-activity relationships. 1. Chromatogr. 373, 191-225. Briggs, G. G. (1969). Molecular structure of herbicides and their sorption by soils. Nature 223, 1288. Briggs, G. G. (1973). A simple relationship between soil adsorption of organic chemicals and their octanollwater partition coefficients. In "Proceedings of the Seventh British Insecticide and Fungicide Conference," pp. 83-86. Briggs, G. G. (l981a). Theoretical and experimental relationships between soil adsorption, octanol-water partition coefficients, water solubilities, bioconcentration factors, and the parachor. J. Agric. Food Chem. 29, 1050-1059. Briggs, G. G. (1981 b). Adsorption of pesticides by some Australian soils. Aust. J. Soil Res. 19,61-68. Brooke, D. N., Dobbs, A. J., and Williams, N. (1986). Octanol:water partition coefficient (P): Measurement, estimation, and interpretation, particularly for chemicals with P > 105. Ecotoxicol. Environ. Sa! 11, 251-260. Bruggeman, W. A., Opperhuizen, A., Wijbenga, A., and Hutzinger, O. (1984). Bioaccumulation of super-lipophilic chemicals in fish. Toxico!. Environ. Chem. 7, 173-189. Chamberlain, K., Evans, A. A., and Bromilow, R H. (1996). 1-0ctanol/water partition coefficient (Kow) and pKa for ionizable pesticides measured by a pH-metric method. Pestic. Sci. 47, 265-271. Charton, M. (1981). Electrical effect substituent constants for correlation analysis. In "Progress in Physical Organic Chemistry" (R W. Taft, ed.), Vol. 13, pp. 119-251. Wiley, New York. Clarke, F. H. (1984). Ionization constants by curve fitting: Application to the determination of partition coefficients. J. Pharm. Sci. 73, 226-230. Comporti, M. (1989). Three models offree radical-induced cell injury. Chem.Bioi. Interact. 72, I-56. Connell, D. W. (1988). Bioaccumulation behavior of persistent organic chemicals with aquatic organisms. In "Reviews of Environmental Contamination and Toxicology" (G. W. Ware, ed.), Vol. 102, pp. 117-154. Springer-Verlag, New York. ConnelI, D. W., and Hawker, D. W. (1988). Use of polynomial expressions to describe the bioconcentration of hydrophobic chemicals by fish. Ecotoxico!' Environ. Sa! 16, 242-257. Cronin, M. T. D., and Dearden, J. C. (1995). QSAR in toxicology.!. Prediction of aquatic toxicity. Quant. Struct.-Act. Re/at. 14, 1-7. Cronin, M. T. D., Bryant, S. E., Dearden, J. C., and Schultz, T. W. (1995). Quantitative structure-activity study of the toxicity of benzonitriles to the ciliate Tetrahymena pyriformis. SAR QSAR Environ. Res. 3,1-13. Davies, RP., and Dobbs, A. J. (1984). The prediction of bioconcentration in fish. Water Res. 18, 1253-1262.

668

CHAPTER 29


Dearden, J. c., and Bresnen, G. M. (1988). The measurement of partition coefficients. Quant. Struct.-Act. Relat. 7, 133-144. Dearden, J. C., and Nicholson, R. M. (1986). The prediction of biodegradability by the use of quantitative structure-activity relationships: Correlation of biological oxygen demand with atomic charge difference. Pestic. Sci. 17, 305-310. de Bruijn, J., and Hermens, J. (1990). Relationships between octanol/water partition coefficients and total molecular surface area and total molecular volume of hydrophobic organic chemicals. Quant. Struct.-Act. Relat. 9,1121. de Bruijn, J., and Hermens, J. (1991). Uptake and elimination kinetics of organophosphorus pesticides in the guppy (Poecilia reticulata): Correlations with the octanol/water partition coefficient. Environ. Toxicol. Chem. 10,791-804. de Bruijn, J., Busser, E, Seinen, W., and Hermens, J. (1989). Determination of octanol/water partition coefficients for hydrophobic organic chemicals with the "slow-stirring" method. Environ. Toxicol. Chem. 8,499-512. de Bruijn, J., Seinen, W., and Hermens, J. (1993). Biotransformation of organophosphorus compounds by rainbow trout (Oncorhynchus mykiss) liver in relation to bioconcentration. Environ. Toxicol. Chem. 12, 10411050. Devillers, J., Bintein, S., and Domine, D. (1996). Comparison of BCF models based on log P. Chemosphere 33,1047-1065. de Wolf, w., de Bruijn, J. H. M., Seinen, w., and Hermens, J. L. M. (1992). Influence of biotransformation on the relationship between bioconcentration factors and octanol-water partition coefficients. Environ. Sci. Technol. 26, 1197-1201. Drabek, J., and Bachmann, E (1983). Proinsecticides: Structure-activity relationship in carbamoyl sulfenyl N -methylcarbamates. In "Pesticide Chemistry: Human Welfare and the Environment" (J. Miyamoto and P. C. Kearney, eds.), Vol. 1, pp. 271-277. Pergamon, Oxford. Dzombak, D. A., and Luthy, R. G. (1984). Estimating adsorption of polycyclic aromatic hydrocarbons on soils. Soil Sci. 137,292-308. Ellgehausen, H., Guth, J. A., and Esser, H. O. (1980). Factors determining the bioaccumulation potential of pesticides in the individual compartments of aquatic food chains. Ecotoxicol. Environ. Sa! 4, 134-157. Esser, H. O. (1986). A review of the correlation between physicochemical properties and bioaccumulation. Pestic. Sci. 17,265-276. Fujita, T. (1983). Substituent effects in the partition coefficient of disubstituted benzenes: Bidirectional Hammett-type interactions. In "Progress in Physical Organic Chemistry" (R. W. Taft, ed.), Vol. 14, pp. 75-113. Wiley, New York. Fujita, T., and Nishioka, T. (1976). The analysis of the ortho effect. In "Progress in Physical Organic Chemistry" (R. W. Taft, ed.), Vol. 12, pp. 49-89. Wiley, New York. Fujita, T., Iwasa, J., and Hansch, C. (1964). A new substituent constant, IT, derived from partition coefficient. J. Am. Chem. Soc. 86, 5175-5180. Fujita, T., Nishioka, T., and Nakajima, M. (1977). Hydrogen-bonding parameter and its significance in quantitative structure-activity studies. J. Med. Chem. 20, 1071-1081. Fukami, J., and Shishido, T. (1966). Nature of a soluble, glutathione-dependent enzyme system active in cleavage of methyl parathion to demethyl parathion. 1. Econ. Entomol. 59, 1338-1346. Geyer, H., Sheehan, D., Kotzias, D., Freitag, D., and Korte, E (1982). Prediction of ecotoxicological behavior of chemicals: Relationship between physicochemical properties and bioaccumulation of organic chemicals in the mussel. Chemosphere 11, 1121-1134. Goring, C. A. I. (1962). Control of nitrification by 2-chloro-6(trichloromethyl)pyridine. Soil Sci. 93, 211-218. Hance, R. J. (1967). Relationship between partition data and the adsorption of some herbicides by soils. Nature 214, 630-631. Hansch, C, and Dunn, W. J. (1972). Linear relationships between lipophilic character and biological activity of drugs. J. Pharm. Sci. 61, 1-19. Hansch, c., and Fujita, T. (1964). Rho-sigma-pi analysis, a method for the correlation of biological activity and chemical structure. J. Am. Chem. Soc. 86, 1616-1626.

Hansch, C., and Fujita, T. (1995). "Classical and Three-Dimensional QSAR in Agrochemistry," ACS Symposium Series No. 606. Am. Chem. Soc., Washington, DC. Hansch, c., and Leo, A. (1979). "Substituent Constants for Correlation Analysis in Chemistry and Biology." Wiley, New York. Hansch, c., and Leo, A. (1995). "Exploring QSAR: Fundamentals and Applications in Chemistry and Biology." Am. Chem. Soc., Washington, DC. Hansch, C., Leo, A., and Hoekman, D. (1995). "Exploring QSAR: Hydrophobic, Electronic, and Steric Constants." Am. Chem. Soc., Washington, DC. Hansch, c., Quinlan, J. E., and Lawrence, G. L. (1968). The linear free-energy relationship between partition coefficients and the aqueous solubility of organic liquids. J. Org. Chem. 33, 347-350. Hansch, c., Vittoria, A., Silipo, C., and Jow, P. Y. C. (1975). Partition coefficients and the structure-activity relationship of the anesthetic gases. J. Med. Chem. 18,546-548. Hawker, D. W., and Connell, D. W. (1985a). Relationships between partition coefficient, uptake rate constant, clearance rate constant and the time to equilibrium for bioconcentration. Chemosphere 14, 1205-1219. Hawker, D. W., and Connell, D. W. (1985b). Prediction of bioconcentration factors under non-equilibrium conditions. Chemosphere 14,1835-1843. Hawker, D. W., and Connell, D. W. (1986). Bioconcentration oflipophilic compounds by some aquatic organisms. Ecotoxicol. Environ. Sa! 11, 184-197. HeIling, C. S. (1971). Pesticide mobility in soils. Ill. Influence of soil properties. Soil Sci. Soc. Am. Proc. 35, 743-748. Hermens, J. L. M. (1989). Quantitative structure-activity relationships of environmental pollutants. In "Handbook of Environmental Chemistry" (0. Hutzinger, ed.), pp. 111-162. Springer-Verlag, Berlin. Hermens, J. L. M. (1990). Electrophiles and acute toxicity to fish. Environ. Health Perspect. 87, 219-225. Hermens, J., Busser, E, Leeuwanch, P., and Musch, A. (1985). Quantitative correlation studies between the acute lethal toxicity of 15 organic halides to the guppy (Poecillia reticulata) and chemical reactivity towards 4-nitrobenzylpyridine. Toxicol. Environ. Chem. 9, 219-236. Hermens, J., de Bruijn, J., Pauly, J., and Seinen, W. (1987). QSAR studies for fish toxicity data of organophosphorus compounds and other classes of reactive organic compounds. In "QSAR in Environmental Toxicology-II" (K. L. E. Kaiser, ed.), pp. 135-152. Reidel, Dordrecht. Ikemoto, Y., Motoba, K., Suzuki, T., and Uchida, M. (1992). Quantitative structure-activity relationships of nonspecific and specific toxicants in several organism species. Environ. Toxicol. Chem. 11,931-939. Isnard, P., and Lambert, S. (1988). Estimating bioconcentration factors from octanol-water partition coefficient and aqueous solubility. Chemosphere 17, 21-34. Iwasa, J., Fujita, T., and Hansch, C. (1965). Substituent constants for aliphatic functions obtained from partition coefficients. J. Med. Chem. 8, 150-153. Karabunarliev, S., Mekenyan, O. G., Karcher, W., Russom, C. L., and Bradbury, S. P. (1996a). Quantum-chemical descriptors for estimating the acute toxicity of electrophiles to the fathead minnow (Pimephales promelas): An analysis based on molecular mechanisms. Quant. Struct.-Act. Relat. 15, 302-310. Karabunarliev, S., Mekenyan, O. G., Karcher, W., Russom, C. L., and Bradbury, S. P. (1996b). Quantum-chemical descriptors for estimating the acute toxicity of substituted benzenes to the guppy (Poecilia reticulata) and fathead minnow (Pimephales promelas). Quant. Struct.-Act. Relat. 15,311-320. Karger, B. L., Gant, J. R., Hartkopf, A., and Weiner, P. H. (1976). Hydrophobic effects in reversed-phase liquid chromatography. J. Chromatogr. 128, 6578. Kataoka, T., Hayase, Y., Hatta, T., Hayashi, Y., Murabayashi, A., Makisumi, Y., and Fujita, T. (1989). Quantitative structure-activity study of fungicidal I-substituted cis-2-(I H -1 ,2,4-triazol-I-yl)cycloalkanols. Pestic. Biochem. Physiol. 34, 228-239. Kaufman, J. J., Semo, N. M., and Koski, W. S. (1975). Microelectrometric titration measurement of the pKa's and partition and drug distribution coefficients of narcotics and narcotic antagonists and their pH and temperature dependence. J. Med. Chem. 18,647-655. Keller, C. (1993). "Grundlagen der Radiochemie," 3rd ed., p. 247. Otto Sal1e Verlag, Frankfurt am Main; Japanese Translation (1993). "Houshakagaku

References

no Kiso" (T. Kishikawa, translator), pp. 304-305. Gendai Kagakusha, Tokyo. Konemann, H. (1980). Structure-activity relationships and additivity in fish toxicities of environmenal pollutants. Eeotoxieol. Environ. Sa! 4, 415-421. Konemann, H. (1981). Quantitative structure-activity relationships in fish toxicity studies. 1. Relationship for 50 industrial pollutants. Toxicology 19, 223-228. Kubinyi, H. (1977). Quantitative structure-activity relationships. 7. The bilinear model, a new model for nonlinear dependence of biological activity on hydrophobic character. J. Med. Chem. 20, 625-629. Kutter, E., and Hansch, C (1969). Steric parameters in drug design. Monoamine oxidase inhibitors and antihistamines. J. Med. Chem. 12,647652. Leo, A. (1991). Hydrophobic parameter: Measurement and calculation. In "Methods in Enzymology" (J. J. Langone, ed.), Vol. 202, pp. 544-591. Academic Press, San Diego. Leo, A. (1993). Calculating log Poet from structures. Chem. Rev. 93, 12811306. Leo, A. (1998). Private communication to T Fujita. The calculation was made with the use of the CLOGP Desktop Version 2.50 for Mac and PCWindows, available at Biobyte Corp., Claremont, CA. Leo, A., Hansch, C, and Elkins, D. (1971). Partition coefficients and their uses. Chem. Rev. 71, 525-616. Leo, A., Jow, P. Y. C, Silipo, C, and Hansch, C (1975). Calculation of hydrophobic constant (log P) from 7r and f constants. 1. Med. Chem. 18, 865-868. Lipnick, R. L. (1989). Narcosis, electrophile and proelectrophile toxicity mechanisms: Application of SAR and QSAR. Environ. Toxieol. Chem. 8, 1-12. Lipnick, R. L. (1990). Narcosis: Fundamental and baseline toxicity mechanism for nonelectrolyte organic chemicals. In "Practical Applications of Quantitative Structure-Activity Relationships (QSAR) in Environmental Chemistry and Toxicology" (W. Karcher and J. DeviIIers, eds.), pp. 281293. Kluwer Academic, Dordrecht. Lipnick, R. L., Johnson, D. E., Gilford, J. H., Bickings, C K., and Newsome, L. D. (1985). Comparison of fish toxicity screening data for 55 alcohols with the quantitative structure-activity relationship predictions of minimum toxicity for nonreactive nonelectrolyte organic compounds. Environ. Toxieol. Chem. 4,281-296. Liu, J., and Qian, C. (1995). Hydrophobic coefficients of s-triazine and phenylurea herbicides. Chemosphere 31, 3951-3959. Mackay, D. (1982). Correlation ofbioconcentration factors. Environ. Sci. Teehno!. 16,274-278. Masutani, T, Seyama, 1., Narahashi, T, and Iwasa, J. (1981). Structure-activity relationship for grayanotoxin derivatives in frog skeletal muscle. J. Pharmacal. Exp. Ther. 217, 812-819. Matsuda, K., Hamada, M., Nishimura, K., and Fujita, T. (1989). Quantitative structure-activity studies of pyrethroids. 17. Physicochemical substituent effects of substituted benzyl esters of "kadethric" acid on symptomatic and neurophysiological activities. Pestie. Bioehem. Physiol. 35, 300-314. McCarty, L. S., Mackay, D., Smith, A. D., Ozbum, G. w., and Dixon, D. G. (1992). Residue-based interpretation of toxicity and bioconcentration QSARs from aquatic bioassays: Neutral narcotic organics. Environ. Toxico!. Chem. 11,917-930. Mekenyan, O. G., and Veith, G. D. (1993). Relationships between descriptors for hydrophobicity and soft electrophilicity in predicting toxicity. SAR QSAR Environ. Res. 1, 335-344. Mekenyan, O. G., and Veith, G. D. (1994). The electronic factor in QSAR: MO-parameters, competing interactions, reactivity and toxicity. SAR QSAR Environ. Res. 2, 129-143. Menn, J. J., and Borkovec, A. B. (1989). Insect neuropeptides: Potential new insect control agents. J. Agrie. Food Chem. 37, 271-278. Mitsutake, K., Iwamura, H., Shimizu, R., and Fujita, T (1986). Quantitative structure-activity relationship of photosystem 11 inhibitors in chloroplasts and its link to herbicidal action. 1. Agrie. Food Chem. 34, 725-732. Miyake, K., and Terada, H. (1982). Determination of partition coefficients of very hydrophobic compounds by high-performance liquid chromatography on glyceryl-coated controlled-pore glass. J. Chromatogr. 240, 9-20.

669

Nakagawa, Y., Izumi, K., Oikawa, N., Sotomatsu, T, Shigemura, M., and Fujita, T. (1992). Analysis and prediction of hydrophobicity parameters of substituted acetanilides, benzamides and related aromatic compounds. Environ. Toxieol. Chem. 11,901-916. Nakagawa, S., Okajima, N., Kitahaba, T, Nishimura, K., Fujita, T., and Nakajima, M. (1982). Quantitative structure-activity studies of substituted benzyl chrysanthemates. 1. Correlation between symptomatic and neurophysiological activities against American cockroaches. Pestie. Bioehem. Physio!. 17,243-258. Neely, W. B., Branson, D. R., and Blau, G. E. (1974). Partition coefficient to measure bioconcentration potential of organic chemicals in fish. Environ. Sci. Teehno!. 8, 1113-1115. Niimi, A. J., Lee, H. B., and Kissoon, G. P. (1989). OctanoIlwater partition coefficients and bioconcentration factors of chloronitrobenzenes in rainbow traut (Salmo gairdneri). Environ. Toxieo!. Chem. 8, 817-823. Nishikawa, S. (1989). "Structure-Activity Relationships of Cytokinin-Active Compounds." Ph.D. Thesis, Kyoto University, Kyoto, Japan. Nishimura, K., and Fujita, T (1983). Quantitative structure-activity relationships of DDT and its related compounds. J. Pestie. Sci. 8, 69-80. Nishimura, K., Ohoka, M., and Fujita, T (1987). Quantitative structure-activity studies of pyrethroids. 10. Physicochemical substituent effects of substituted benzyl pyrethrates on symptomatic and neurophysiological activities. Pestie. Bioehem. Physiol. 28, 257-270. Noble, A. (1993). Partition coefficients (n-octanol-water) for pesticides. J. Chromatogr. 642, 3-14. Nys, G. G., and Rekker, R. E (1973). Statistical analysis of a series of partition coefficients with special reference to the predictability of folding of drug molecules: The introduction of hydrophobic fragmental constants (f-values). Chim. Ther. 8, 521-535. Oikawa, N., Nakagawa, Y., Nishimura, K., Ueno, T, and Fujita, T (1994). Quantitative structure-activity studies of insect growth regulators. 10. Substituent effects on larvicidal activity of I-tert-butyl-I-(2-chlorobenzoyl)2-(substituted benzoyl)hydrazines against Chilo suppressalis and design synthesis of potent derivatives. Pestie. Bioehem. Physiol. 48, 135-144. Okimoto, H., Nishimura, K., Matsuda, K., Hamada, M., Ueno, T, and Fujita, T. (1994). Neurophysiological effects and quantitative structure-activity analyses of insecticidal silaneophanes. Pestie. Bioehem. Physiol. 49, 83-93. Oliver, B. G. (1987). Partitioning relationships for chlorinated organics between water and particulates in the St. Clair, Detroit, and Niagara rivers. In "QSAR in Environmental Toxicology-I1" (K. L. E. Kaiser, ed.), pp. 251-260. Reidel, Dordrecht. Oliver, B. G., and Niimi, A. J. (1983). Bioconcentration of chlorobenzenes from water by rainbow trout: Correlation with partition coefficients and environmental residues. Environ. Sei. Technol. 17,287-291. Oliver, B. G., and Niimi, A. J. (1985). Bioconcentration factors of some halogenated organics for rainbow trout: Limitations in their use for prediction of environmental residues. Environ. Sei. Teehno!. 19,842-849. Opperhuizen, A., Velde, E. W., Gobas, E A. P. C, L1em, D. A. K., and Steen, J. M. D. (1985). Relationship between bioconcentration in fish and steric factors of hydrophobic chemicals. Chemosphere 14, 1871-1876. Opperhuizen, A., and Voars, P. L. (1987). Uptake and elimination of polychlorinated aromatic ethers by fish: Chloroanisoles. Chemosphere 16, 953-962. Organization for Economic Co-operation and Development (OECD) (1981). "Guidlines for Testing of Chemicals." No. 107, Organization for Economic Co-operation and Development, Bureau for Information, Paris. Organization for Economic Co-operation and Development (OECD) (1989). "Guidlines for Testing of Chemicals." No. 117, Organization for Economic Co-operation and Development, Bureau for Information, Paris. Rekker, R. E, and Mannhold, R. (1992). "Calculation of Drug Lipophilicity." VCH, Weinheim. Saarikoski, J., and Viluksela, M. (1981). Influence of pH on the toxicity of substituted phenols to fish. Arch. Environ. Contam. Toxieo!. 10,747-753. Saarikoski, J., and Viluksela, M. (1982). Relation between physicochemical properties of phenols and their toxicity and accumulation in fish. Eeotoxico!. Environ. Sa! 6, 501-512.

670

CHAPTER 29


Sabljic, A., Guesten, H., Verhaar, H., and Hennens, J. (1995). QSAR modelling of soil sorption: Improvements and systematics of log Koc vs. log Kow correlations. Chemosphere 31, 4489-4514. Sangster, J. (1997a). "Octanol-Water Partition Coefficients: Fundamentals and Physical Chemistry," pp. 57-112. Wiley, Chichester. Sangster, J. (1997b). "Octanol-Water Partition Coefficients: Fundamentals and Physical Chemistry," pp. 113-156. Wiley, Chichester. Schultz, T. w., Sinks, G. D., and Hunter, R. S. (1995). Structure-toxicity relationships for alkanones and alkenones. SAR QSAR Environ. Res. 3, 27-36. Schiiiinnann, G. (1992). Ecotoxicology and structure-activity studies of organophosphorus compounds. In "Rational Approaches to Structure, Activity, and Ecotoxicology of Agrochemicals" (W. Draber and T. Fujita, eds.), pp. 485-541. CRC Press, Boca Raton, FL. Shimizu, R., Iwamura, H., and Fujita, T. (1988). Quantitative structure-activity relationships of photosystem II inhibitory anilides and triazines: Topological aspects of their binding to the active site. J. Agric. Food Chem. 36, 1276-1283. Sieburth, S. M., Manly, C. J., and Gammon, D. W. (1990). Organosilane insecticides. I: Biological and physical effects of isosteric replacement of silicon for carbon in etofenprox and MTI-800. Pestic. Sci. 28, 289-307. Smith, R. N., Hansch, c., and Ames, M. M. (1975). Selection of a reference partitioning system for drug design work. J. Pharm. Sci. 64, 599-609. Sotomatsu, T., Nakagawa, Y., and Fujita, T. (1987). Quantitative structureactivity studies of benzoylphenylurea larvicides. 4. Benzoyl ortho substituent effects and molecular confonnation. Pestic. Biochem. Physiol. 27, 156-164. Sotomatsu, T., Shigemura, M., Murata, Y., and Fujita, T. (1993). OctanoIlwater partition coefficient of ortho-substituted aromatic solutes. J. Pharm. Sci. 82, 776-781. Southworth, G. R., Keffer, C. c., and Beauchamp, J. J. (1980). Potential and realized bioconcentration. A comparison of observed and predicted bioconcentration of azaarenes in the fathead minnow (Pimephales promelas). Environ. Sci. Technol. 14, 1529-1531. Stewart, J. J. P. (1990). "MOPAC Manual," Version 6.0. Quantum Chemistry Program Exchange, Indiana University, Bloomington, Indiana. Taft, R. W. (1956). Separation of polar, steric, and resonance effects in reactivity. In "Steric Effects in Organic Chemistry" (M. S. Newman, ed.), pp. 597-603. Wiley, New York. Takahashi, J., Kirino, 0., Takayama, c., and Kamoshita, K. (1988). Studies on fungicidal activity of N -pheny lcarbamates. IV. Detennination of the hydrophobicity of N-phenylcarbamates by high-perfonnance liquid chromatography. J. Chromatogr. 436, 316-322. Takayama, c., Akamatsu, M., and Fujita, T. (1985). Effects of structure on l-octanoIlwater partitioning behavior of aliphatic amines and ammonium ions. Quant. Struct.-Act. Relat. 4, 149-160. Terada, H. (1986). Detennination of log Poet by high-perfonnance liquid chromatography, and its application in the study of quantitative structureactivity relationships. Quant. Struct.-Act. Relat. 5, 81-88. Terada, H., Kitagawa, K., Yoshikawa, Y., and Kametani, F. (1981). Partition and ion-pair partition of 2,4-dinitrophenol, an uncoupler of oxidative phosphorylation. Chem. Pharm. Bull. 29, 7-14. Uchida, M., and Kasai, T. (1980). Adsorption and mobility of fungicidal dialkyl dithiolanylidenemalonates and their analogs in soil. J. Pestic. Sci. 5, 553558. Uchida, M., Funayama, S., and Sugimoto, T. (1982a). Bioconcentration of fungicidal dialkyl dithiolanylidenemalonates in Orizias latipes L. J. Pestic. Sci. 7,181-186. Uchida, M., Kurihara, N., Fujita, T., and Nakajima, M. (1974). Inhibitory effects of BHC isomers on Na+ -K+ -ATPase, yeast growth, and nerve conduction. Pestic. Biochem. Physiol. 4, 260-265.

Uchida, M., Nishizawa, H., and Suzuki, T. (1982b). Hydrophobicity of buprofezin and fIutoranil in relation to their soil adsorption and mobility in rice plants. J. Pestic. Sci. 7, 397-400. Valvani, S. c., and Yalkowsky, S. H. (1980). Solubility and partitioning in drug design. In "Physical Chemical Properties of Drugs" (S. H. Yalkowsky, A. A. Sinkula, and S. C. Valvani, eds.), pp. 201-229. Dekker, New York. Veith, G. D., and Broderius, S. J. (1987). Structure-toxicity relationships for industrial chemicals causing type (ll) narcosis syndrome. In "QSAR in Environmental Toxicology-II" (K. L. E. Kaiser, ed.), pp. 385-391. Reidel, Dordrecht. Veith, D. v., and Mekenyan, O. G. (1993). A QSAR approach for estimating the aquatic toxicity of soft electrophiles [QSAR for soft electrophiles J. Quant. Struct.-Act. Relat. 12,349-356. Veith, G. D., Call, D. J., and Brooke, L. T. (1983). Structure-toxicity relationships for the fathead minnow, Pimephales promelas: Narcotic industrial chemicals. Can. J. Fish. Aquat. Sci. 40, 743-748. Veith, G. D., DeFoe, D. L., and Bergstedt, B. V. (1979). Measuring and estimating the bioconcentration factor of chemicals in fish. J. Fish. Res. Board Can. 36, 1040-1048. Verhaar, H. J. M., Eriksson, L., Sjoestroem, M., Schiiiinnann, G., Seinen, W., and Hennens, J. L. M. (1994). Quant. Struct.-Act. Relat. 13, 133-143. Verhaar, H. J. M., van Leeuwen, C. J., and Hennens, L. M. (1992). Classifying environmental pollutants. I. Structure-activity relationships for prediction of aquatic toxicity. Chemosphere 25, 471-491. Verloop, A. (1983). The STERIMOL approach: Further development of the method and new application. In "Pesticide Chemistry: Human Welfare and the Environment" (J. Miyamoto and P. C. Kearney, eds.), Vol. I, pp. 339344. Pergamon, Oxford. Wang, P. S., and Lien, E. J. (1980). Effects of different buffer species on partition coefficients of drugs used in quantitative structure-activity relationships. J. Pharm. Sci. 69, 662-668. Wauchope, R. D., ButtIer, T. M., Homsby, A. G., Augustijn-Beckers, P. W. M., and Burt, J. P. (1992). The SCS/ARS/CES pesticide properties database for environmental decision-making. In "Reviews of Environmental Contamination and Toxicology" (G. W. Ware, ed.), Vol. 123, pp. 1-25. SpringerVerlag, New York. Yamagami, c., and Fujita, T. (1995). Hydrophobicity parameter of heteroaromatic compounds derived from various partitioning systems. In "Classical and Three-Dimensional QSAR in Agrochemistry" (C. Hansch and T. Fujita, eds.), ACS Symposium Series 606, pp. 36-47. Am. Chem. Soc., Washington, DC. Yamagami, c., Ogura, T., and Takao, N. (1990). Hydrophobicity parameters detennined by reversed-phase liquid chromatography. I. Relationship between capacity factors and octanol-water partition coefficients for monosubstituted pyrazines and the related pyridines. J. Chromatogr. 514, 123-136. Yamagami, c., Takao, N., and Fujita, T. (1995). Analysis and prediction of 1octanoIlwater partition coefficients of substituted diazines with substituent and structural parameters. In "QSAR and Drug Design-New Development and Applications" (T. Fujita, ed.), pp. 153-183. Elsevier, Amsterdam. Yamagami, c., Yokota, M., and Takao, N. (1994). Hydrophobicity parameters detennined by reversed-phase liquid chromatography. IX. Relationship between capacity factor and octanol-water partition coefficient of monosubstituted pyrimidines. Chem. Pharm. Bull. 42, 907-912. Yukawa, Y., and Tsuno, Y. (1959). Resonance effect in Hammett relationship. II. Sigma constants in electrophilic reactions and their intercorrelation. Bull. Chem. Soc. Jpn. 32, 965-971. Zakarya, D., Belkhadir, M., and Fk:ih-Tetouani, S. (1993). Quantitative structure-biodegradability relationships (QSBRs) using modified autocorrelation method (MAM). SAR QSAR Environ. Res. 1,21-27.

CHAPTER

30 Modern Approaches to Analysis of Pesticide Residues in Foods and the Environment Luis O. Ruzo PTRL West, Inc.

Thomas Class PTRLEurope

30.1 INTRODUCTION The capability to determine the actual residues of pesticides, and increasingly of their metabolites, is at the core of the complex and sometimes bewildering regulatory processes in all industrialized nations. Thus, in the United States we are currently experiencing the replacement of laws based on long-term toxicological effects, such as the Delaney Clause, with those based on quantitative determinations that lead to a different type of risk assessment, as is the Food Quality Protection Act of 1996 (FQPA). To understand the benefits and limitations of modem analytical techniques, it is necessary to examine the regulatory requirements that they aim to satisfy. Residue chemistry data are used by regulatory agencies to estimate the exposure of the general population, as well as discrete subpopulations, to pesticide residues in food and water and for setting and enforcing tolerances for such residues in food crops or animal feed. These data are also used to monitor environmental contamination of soil, air, and water and thus to determine adverse effects that may arise from transport of residues between these compartments. The tolerance values are a key component of the regulatory equation as they represent the amounts of pesticiderelated materials legally allowed to be present in a given matrix. Thus, tolerances are legally enforceable limits that are currently under scrutiny for implementation of the FQPA. The passage of this law by the U.S. Congress in 1996 signaled a fundamental change in the way exposure to pesticides is evaluated by introducing the concept of aggregate exposure to several compounds of a given chemical class and mode of action. These expanded requirements will necessitate the development of new methodHandbook of Pesticide Toxicology Volume 1. Principles

ologies for analysis, specifically to address the lower limits of quantitation (LOQs) (Ragsdale, 1998). A similar situation is observed in the European Community (EC). Here (re-)registration of pesticides and enforcement of tolerances or maximum residue limits (MRLs) requires that existing multiresidue methods be assessed for their applicability toward the determination of active substances and their toxicologically relevant metabolites.

30.2 METHOD VALIDATION Because this chapter will deal primarily with analytical techniques now in use for developing methodologies, it is worth reviewing briefly the criteria that enforcement agencies use to evaluate method performance. Methods used by government laboratories are generally developed by the pesticide registrants. A method is considered acceptable upon validation. This process may entail the examination of various parameters but always includes the establishment of a limit of quantitation (LOQ) and a limit of detection (LOD). The former is usually set at the lowest matrix fortification level for which acceptable, quantifiable recoveries of the analyte(s) are obtained.

30.2.1 CONFIRMATION OF THE ANALYTICAL METHOD This entails the fortification of untreated (control) matrices (crops, soil, water, etc.) with varying concentrations of analyte. Thus, the processed sample is fortified at two or three levels in duplicate or triplicate (United States) or in five replicates (EC). The fortified samples and corresponding untreated controls are

671


672

CHAPTER 30

Modern Approaches to Analysis of Pesticide Residues

then subjected to extraction, cleanup, and chromatographic separation and quantitation. The ultimate recovery of analytes must be in the 70-120% (70-110% in Europe) range based on the initial concentration, with repeatability demonstrated by relative standard deviations of Q 20%. The lowest concentration for which this is achieved during validation is generally considered the limit of quantitation or determination (LOQ) for the method. The limit of detection (LOD) may be any value below the LOQ at which an analyte signal is clearly distinguished from background signals present or absent in matrix extracts. At present, methods aim at an LOQ of 0.01 ppm (mgikg) for most foodstuffs, whereas much lower values are considered desirable for water analyses, where the LOQ Q 0.05 I-Lg/I (ppb) is required, or for compounds of known toxicological significance. Important aspects to be considered in the conduct of validation studies are provided by lenke (1996a-c).

30.2.2 INDEPENDENT LABORATORY VALIDATION

For a method to be used in the enforcement of tolerances or MRLs, it must be rugged and straightforward. Therefore, the U.S. Environmental Protection Agency (EPA) and the EC require that such methods be validated independently by laboratories that are not familiar with the procedures or analytes (EC Directorate General for Agriculture, 1998; EPA, 1996). Because the equipment necessary for the enforcement method must be generally accessible and affordable for the enforcement laboratories, some of the most advanced techniques (such as MSIMS, or in the European Community LC-MS) are still not acceptable to regulatory agencies. Measures to demonstrate the validity of results obtained by enforcement methods include round-robin testing using identical samples with fortified or incurred residues, which allow assessment of the reproducibility of commonly employed enforcement methods.

30.2.3 METHOD RADIOVALIDATION (EPA,1996)

A stringent test for an analytical method is its reproducibility when applied to incurred residues as opposed to an externally fortified matrix. Plant and animal metabolism studies utilizing 14C-Iabeled pesticides generate matrices containing incurred residues that can be readily quantified with radiochemical methodology, which is quite different (and simpler) than that generally developed for an analytical residue method. In order for a method to be considered fully validated, the results obtained when the "cold" method is applied to matrices containing 14C-Iabeled incurred residues must agree closely with the results arising from quantitation of the radiocarbon conducted by radiochemical methods such as liquid chromatography and liquid scintillation counting.

30.3 DEVELOPMENT OF THE ANALYTICAL METHOD 30.3.1 MULTIRESIDUE METHODS

Multiresidue, multiclass methods are generally the most cost effective overall approach for pesticide analysis in foods, soil, and water. Regulatory enforcement methods routinely deal with multiresidue determinations under fairly standardized conditions. For example, the widely used DFG S19 method (Thier and Zeumer, 1987) establishes conditions for extraction and quantitation of organochlorine, organophosphorous, and nitrogencontaining pesticides (typically 80-100 compounds per method) in crop plants under standardized conditions. The original method (Specht et al., 1995) uses an acetone/ethyl acetate!cyclohexane mixture for extraction and partition, thus replacing dichloromethane and introducing a "one-beaker" extraction and partition procedure. In addition to use with watery plant matrices, the DFG S19 method has also been successfully employed for dry plant matrices (straw, hops, tobacco, herbal teas), for oily crops (oil seed rape, sunflower seeds, nuts, etc.), for animal matrices (milk, whole egg, muscle, fat), and for soil. Thus, a single solvent (acetone) with fixed amounts of added water (in a ratio of 2: 1 solvent water) or acetonitrile (for oily matrices) is used for extraction of compounds with a wide range of polarity. Cleanup involves gel permeation chromatography (GPC) and fractionation on a small silica gel column followed by separationlquantitation with capillary gas chromatography equipped with selective detectors. The Netherlands' Inspectorate for Health Protection has recently published in its 6th edition (1996) an extended collection of multiresidue methods for a multitude of pesticides: The Netherlands' multiresidue method 1 (MRM-1) covers all pesticides that can be analyzed by capillary gas chromatography using selective detectors [electron-capture detector (ECD), nitrogen-phosphorus detector (NPD), flame photometric detector (FPD)] and increasingly full-scan ion trap mass spectrometric detection. Three submethods described in The Netherlands' MRM-2 use high-performance liquid chromatography (HPLC) and postcolumn derivatization with fluorescence detection for N-methylcarbamate (including metabolites) and phenylurea pesticides or precolumn switching employing a precolumn packed with internal surface reversed-phase material for chlorophenoxy herbicides. The latter class of compounds is also analyzed more selectively by The Netherlands' MRM3 after derivatization with pentafluorobenzyl bromide (PFBBr) as esters. Derivatives of aromatic amines are covered by The Netherlands' MRM-4 (two submethods) and by the German DFG S6 and S6-A methods, all of which use alkaline hydrolysis and steam distillation of the amines, followed by various derivatization procedures, and gas chromatography. The Luke method is a multiresidue method currently employed in the United States for enforcement of tolerances and import tolerances. Its cleanup includes mainly solid-phase extraction (SPE) cartridges of various selectivities with either gas or liquid chromatography of pesticides in separate fractions.

30.3 Development of the Analytical Method

However, because of the current emphasis on metabolites of toxicological concern, target methods are generally developed with an individual or closely related compounds in mind. Frequently, this involves simultaneous determination of the parent pesticide and its metabolites, which are generally of greater polarity (because they typically arise from oxidation or hydrolytic cleavage reactions). 30.3.2 EXTRACTION Generally, extraction methodology must be developed such that nearly quantitative recovery of target analytes is obtained. At this stage of method development, the use of radiolabeled standards is invaluable because it allows for rapid determination of the percentage extractability by direct liquid scintillation counting (LSC). In later steps, the radiotracer is useful in determining efficiencies for each step of the proposed method. Traditional solvent extraction must take into account the chosen solvent's water miscibility, ease of solvent concentration/removal, safety, and disposal costs. DFG S 19 uses water/acetone (100 ml1200 ml) or acetonitrile/acetone for oily matrices and allows a sample size ranging from 109 of very dry material (straw, herbs) to 100 g for watery matrices. The Netherlands' multi methods use mainly acetone or ethyl acetate for extraction; other conventional extraction systems include methanol or acetonitrile. Some methods use combined extractionlhydrolysis steps either for deconjugation or to form common moiety products, which may be separated from the extraction mixture by steam distillation (e.g., The Netherlands' MRM-4 and the DFG S6 for derivatives of aromatic amines). Several new extraction approaches are being developed such as supercritical fluid extraction (SFE) and pressurized liquid extraction (also known as accelerated solvent extraction, ASE). Typically, SFE gives very clean extracts with somewhat low recoveries and ASE gives good recoveries but the samples are exposed to high temperatures and pressures and require more extensive cleanup to remove co-extracted matrix components. The latter techniques allow only the extraction of relatively small amounts of samples (e.g., 5-10 g), which requires an increased emphasis on sample preprocessing and homogenization (to assure homogeneity) if market sample size amounts to 5-10 kg. SFE offers potential advantages for removing trace levels of target agrochemical ana1ytes from various matrices. Of particular interest are the enhanced extraction rates obtained due to the high diffusivity of critical fluids. In addition to rapid extraction, improved penetration of the matrix with subsequent high recovery of bound residues is feasible (Hawthorne et aI., 1992; Lira, 1988; Lopez-Avila and Dodhiwala, 1990). However, the success of commonly used fluid systems such as carbon dioxide or polar-modified carbon dioxide in binary mixtures is limited to, at best, moderately polar analytes. Because many factors can influence SFE efficiency (Erstfeld and Chen, 1998; Fahmy et aI., 1993; Snyder et aI., 1992, 1993), including pressure, temperature, fluid flow rate, extraction time, and modifier, considerable effort must be invested

673

during method development before choosing SFE as the extraction technique. However, there is evidence that at least in some cases (as with chlorothalonil) it compares favorably to traditional approaches such as Soxhlet extraction (Erstfeld and Chen, 1998). Thorough descriptions of the instrumentation utilized by SFE are available in the literature (McNally et aI., 1992; Riekkola et aI., 1992). Excellent short reviews of SFE applications are those of Bond (1994) and King (1989). Supercritical fluid extraction has been reported to be successful in multiresidue analysis with pyrethroids (Argauer et al., 1997), organophosphates (Skopec et aI., 1993), and other compound classes (Jones, 1996, 1997; King et aI., 1993). In fact, even the hydrophobic avermectins are amenable to SFE techniques (Brooks and Uden, 1995). There is general agreement among researchers that although the scope of applicability for SFE may be limited, great advantages are provided in selected cases by the cleaner extracts obtained. Accelerated solvent extraction (ASE) is rapidly gaining acceptance as an alternative extraction approach with positive results in magnitude of residue, multiresidue, soil dissipation, and animal health studies (Stanek and Keller, 1998). ASE is based on the use of a variety of solvents under elevated temperatures and pressure. The higher temperatures involved accelerate the kinetics (as with Soxhlet) and the elevated pressure keeps the solvent in the liquid phase. The technique is amenable to automation (Ezzell, 1998). Traditionally, compounds that are difficult to extract from environmental matrices have been subjected to reflux conditions such as Soxhlet extraction. This approach is wasteful in time and solvent use. In fact, ASE compares favorably with Soxhlet and supercritical fluid extraction (David and Seiber, 1996; Frost et aI., 1997; Lou et aI., 1997). ASE is most effective with thermally stable low- or medium-polarity substrates. Extraction results are often better than those obtained with traditional methods (Conte et aI., 1997; Ezzell et aI., 1995). Sample preparation is quite simple (Richter et aI., 1996), involving grinding and mixing of the soil followed by air drying or admixture with drying agents (e.g., sodium su1fate). 30.3.3 CLEANUP OF EXTRACTS The primary extraction methodology can result in significant cleanup of the sample by separating the ana1yte from the bulk of interfering matrix components. However, primary extraction methods are designed to decrease sample bulk rather than to achieve complete purification. Therefore, additional steps are generally needed. The degree of purification ranges from none to very little with binary solvent systems (e.g., aqueous acetonitrile or methanol) to significant with solvent systems that allow homogeneous partition such as in the DFG S19 method, which uses a sequential system consisting of NaCl, water, acetone, ethyl acetate, and cyclohexane and results in rather clean extracts for watery crops, or with SFE or ASE, as discussed previously. Therefore, purification regimes must generally be instituted subsequent to the initial analyte separation.

674

CHAPTER 30 Modern Approaches to Analysis of Pesticide Residues

Traditional approaches to extract cleanup usually involve liquid/liquid partition. The most common technique focuses on differences in solubility (and polarity) of the matrix constituents. Thus, it is common to find methods in which (a) fatty components in matrices are removed by partition with nonpolar solvents (acetonitrilelhexane fat cleanup, extraction of aqueous extracts, or residues with dichloromethane) or (b) acidic or basic analytes are converted to water-soluble salts and the aqueous phase extracted with organic solvents to remove matrix components. Gel permeation chromatography (GPC) is an established cleanup technique that separates (size exclusion) high-molecular-weight compounds such as proteins, fats, and sugars from relatively low molecular weight compounds such as pesticides in animal and plant matrices. Typically, GPC involves injection of large-volume solutions of analyte ('"'-'5 ml) via autosampler onto a low-pressure glass column containing polystyrene divinylbenzene beads conditioned with the appropriate solvent. Fractions of eluate are then collected and concentrated prior to further chromatographic analysis. A recent modification (Chambers, 1998) involves using a high-pressure steel column, lower injection volumes, and collection of smaller fractions, thus minimizing solvent use and concentration time. GPC is an integral part of multiresidue methods (Thier and Zeumer, 1987). A more efficient, but time-consuming method involves liquid/solid partition. For several decades, silica gel and florisil and size exclusion supports (gel permeation) have been used in liquid chromatography (LC). Thin-layer chromatography is far less effective. However, the advent of reverse-phase (RP) systems and high-efficiency SPE (and microextraction) cartridges has revolutionized the approaches to analyte purification. Whereas normal-phase (silica-based) supports could handle only organic solvents of medium to low polarity, reverse-phase systems can extract hydrophobic pesticide residues directly from aqueous solutions without involving significant amounts of organic solvents. The first report on applications of solid-phase extraction (SPE) with reverse-phase supports (Belardi and Pawliszyn, 1989) involved chemically modified fused-silica fibers. There are now an increasing variety of other solid supports relying on ion exchange, size exclusion, and other physicochemical properties (Zhang et aI., 1994). The technique has rapidly advanced, especially for the analysis of pesticide traces in systems such as surface and groundwater (Balinova, 1993; Eisert and Levsen, 1995a; Field et aI., 1997; Hatrik et aI., 1994; Moore et aI., 1995) and procymidone in wine (Vrruty et aI., 1997). SPE has also become an integral part of multiresidue methods (Analytical Methods for Pesticide Research, 1996; Barnabas et aI., 1995; Nouri et aI., 1995). A great advantage of SPE is that in many cases similar HPLC column supports are available (C8, C18, aminopropyl), which can be used to predict the chromatographic behavior of the analyte relative to potential interferences. Solid-phase extraction Empore disks are an alternative to SPE cartridges. These disks may eliminate certain cleanup procedures and further reduce the use of organic solvents. The

disks may be extracted with small amounts of solvents directly in the autosampler vial (Field and Monahan, 1995, 1996), thus allowing for increased method automation. In general, SPE methodology is increasingly being incorporated into on-line systems (Marce et aI., 1995; Maris et aI., 1985; Nielen et aI., 1987) with an emphasis on polar pesticides such as diuron and bromacil in water samples (Parrilla et aI., 1993; Sancho et aI., 1997; Sennert et aI., 1995).

30.3.4 SEPARATION AND QUANTITATION 30.3.4.1 Gas Chromatography and Mass Spectrometry For the past 40 years, gas chromatography (GC) has been the most widely utilized technique to analyze pesticide residues. Advances in chromatography, in particular, capillary column technology, have provided an increasing variety of thermally stable stationary phases, thus improving selectivity. The development of highly specific carbon, phosphorus, sulfur, and nitrogen detectors based on flame ionization and photometry (NPD, FPD) and on electron capture (ECD) culminated with the direct coupling of capillary GC columns with decreased carrier gas flow to mass spectrometers, in particular, those equipped with quadrupole analyzers (March, 1997). Thus, selectivity and sensitivity were improved tremendously. Large numbers of samples could thus be screened cheaply and efficiently once automation was introduced for sample injection. In typical applications, such a GC-MS system can provide quantitation of 20-40 samples overnight, including full- or selected-ion spectra on several components. In fact, GC-MS is already being used extensively for the V.S. Department of Agriculture Pesticide Data program, as exemplified by the routine simultaneous determination of diphenyl amine, o-phenylphenol, and propargite in apples (Yu et aI., 1997). Other significant developments in GC-MS included the following: 1. The use of high-resolution mass spectrometry (HRMS) with magnetic sector instruments, which led to limits of detection in the femtogram range, especially for analytes that contain heteroatoms with significant mass defects (e.g., chlorine) and that are prone to give simple spectra on negative chemical ionization (NCI). 2. The introduction of the ion trap mass spectrometer (ITD), which resulted in better sensitivities in the full-scan mode, thus providing improved identification power in combination with nontarget multiresidue enforcement methods. 3. The introduction of multiple MS capabilities either by means of a row of quadrupoles (triple-quads), providing MSIMS in space; or by the use of ion trap mass spectrometers, allowing MS n (multiple MS experiments) in time.

30.3 Development of the Analytical Method The MSIMS or MS n techniques result quite often in better selectivity and thus in improved sensitivity compared to the single-quadrupole or early ion trap techniques, whereas the price for the instruments is still much lower than that for the high-resolution MS instruments. Thus, the versatility of these instruments, especially of the ion trap mass spectrometers equipped either with an external ionization source (which allows negative and positive chemical ionization) or with internal ionization (which seems to give better ion yields on electron impact and positive chemical ionization), has resulted in a tremendous improvement in the limit of detection for many analytes that can be chromatographed by capillary Gc. Multiple MS thus serves as an additional "cleanup" technique, allowing for very low detection limits of analytes in matrix. For example, nanogram per liter detection of several pesticides in aquatic matrices has been reported using tandem GC-MS (Boyd-Boland et aI., 1996; Rossi et aI., 1997; Steen et aI., 1997). That GC-MSIMS is rapidly becoming the preferred technique for pesticide analysis (Feigel, 1997) is due in part to the added capability for confirmation of identity, which allows for elimination of the false positives often detected in GC-MS analysis (de Cruz et aI., 1996; Schachterle and Feigel, 1996). Furthermore, the possibility of conducting two or more MS n experiments in the search for ions not present in co-eluting interferences greatly simplifies the extraction and purification process. GC-MS has proven useful in multiresidue methods, sometimes dealing with mixtures containing in excess of 100 compounds, in a very cost-effective manner (Fillion et aI., 1995; Liao et aI., 1991). In accordance with present trends toward automation of pesticide analysis, it may be expected that in-line techniques coupling solid-phase microextraction of pesticide mixtures with GC-MS (Eisert and Levsen, 1995b) will be increasingly reported. 30.3.4.2 Liquid Chromatography and Mass Spectrometry

A major shortcoming of GC-MS techniques is their inability to analyze samples of low volatility, high polarity, or thermal instability. Because regulatory requirements focused only on the parent pesticide, GC-MS could accomplish the required goals for the great majority of commercial products. However, in the past decade the need for quantitative data on degradation products and metabolites has increased and now the complete expression for "toxic residues" routinely includes one or more metabolites in addition to the pesticide itself. A multitude of derivatization techniques (Lunn and Hallwig, 1998) is available to convert hydroxy, carboxy, amino, and other polar derivatives to entities amenable to GC-MS. The derivatization approach is expensive and often unreliable in view of the variety of matrices and co-extractives involved. Furthermore, the analytes of interest are often obtained in aqueous fractions incompatible with the majority of available derivatizing agents. Liquid chromatography (LC) can successfully handle polar, nonvolatile compounds but, as a residue technique, it has

675

developed more slowly than GC, and, mainly due to the lack of element-specific detectors, LC has not reached the wide applicability of gas chromatographic methods. Although detectors utilizing ultraviolet absorbance, fluorescence (with/without postcolumn derivatization), electrochemical properties, and refractive index have proven useful, the advent of combined liquid chromatography and mass spectrometry (LC-MS) has provided renewed and strong interest in LC as a residue analytical tool. The union of the two analytical techniques in LC-MS combines an instrument that operates in the condensed phase with one that operates at reduced pressures. Thus, for LC, it is the mobile phase and flow rate that affects separation of analytes; for MS, it is the ionization mode and factors affecting the production and transport of gas-phase ions into the analyzer that are important. In the 1980s, several approaches evolved toward finding a practical application for LC-MS: First were the moving belt interface and direct liquid introduction, which were later replaced by particle beam and thermospray technologies (Cairns and Siegmund, 1990). These methods have been applied to a great variety of analytical problems involving pesticides, their metabolites, and conjugates (Brown, 1990). In particular, it is worth noting that LC-MS proved itself as a viable technique with the successful analysis of thermally labile sulfonylurea herbicides (Shalaby and George, 1990). These compounds are used at very low application rates so the use of thermospray techniques established the increased sensitivity of LC-MS as compared to other methods. In spite of interface improvements such as thermospray, complications in the use of LC-MS were common and based on the introduction of a fluid stream into a vacuum system. At present, the best results are being obtained with ion sources developed over the past decade that operate at or near atmospheric pressure. Atmospheric pressure chemical ionization (APCI) relies on nebulization of the solvent stream followed by thermal evaporation. Thus, the mixture is ionized in the vapor phase and the reactant ions are formed from the components present in the LC eluent. Chemical ionization utilizes solvent molecules as the "reagent gas." APCI can handle high flow rates and high electrolyte concentrations. The second widely utilized atmospheric pressure technique is electrospray ionization (ESI). Here the sample has to be in an ionized form in solution. Neutral samples can be converted to ions by adjustment of pH or by addition of electrolytes (such as ammonium acetate) to form ion-molecule complexes. The influence of the species utilized for ion-pair processes in LC is quite important as evidenced in the analysis of diquat and paraquat (Startin et aI., 1998). The sample solution is then dispersed into an electrically charged aerosol. The target ions are separated from the droplet interface by the action of an electric field at the surface of the charged droplet, which, upon partial solvent removal, develops a substantially smaller cross section. The gas-phase ions are then transported to the mass analyzer. Because electrospray ionization can be accomplished at low temperatures, this tech-

676

CHAPTER 30

Modem Approaches to Analysis of Pesticide Residues

nique is ideal for thermally labile compounds. ESI generally works best at low flow rates and with lower concentrations of electrolytes. Both APCI and ESI are quite sensitive to eluent composition and to matrix effects. Thus, extensive method development is required in the early stages of analysis to optimize sensitivity, selectivity, and compatibility with matrix components and eluent solvents. In spite of these drawbacks, APCI and ESI are rapidly becoming the LC-MS techniques of choice. As with GC-MS, tandem (MSIMS), or MSn, methods for LC have proliferated (Gilbert et aI., 1995). The great advantage of the MSIMS technique using triple-quadrupole instruments (MSIMS-in-space) and of the MS n (-in-time) technique using ion trap technology lies in the gain of selectivity using the "multiple-reaction mode" (MRM). In the case of APCI, protonated quasi-molecular parent ions are selected either by passing the first quadrupole or by eliminating all other ions in the ion trap, and then fragmented further by collision-induced dissociation in the middle quadrupole or in the trap by applying energy. The daughter ion fragments formed are then filtered by the third quadrupole or by the ion trap and yield highly selective MS signals. The majority of pesticides and metabolites can be detected by LC-MSIMS in concentrations greater than 10 ng/ml with high selectivity, allowing much abbreviated cleanup procedures. However, although the current view is that MSIMS techniques can largely eliminate cleanup steps, the issues and strategies that have historically been important in sample preparation and analysis remain fundamentally important to reliable LC-MSIMS. The "dilute and shoot" approach is valid in some cases where analyte concentrations are high and matrix components do not co-elute or otherwise interfere with ionization. Contamination of HPLC column and MS source by matrix components, however, may cause variation and drift of the LC-MS signals and thus hamper unattended automated analysis of crude extracts. This can be avoided to a certain extent by introducing precolumn-switching techniques or automated on-line SPE cleanup steps. An obvious approach to bypass matrix effects is to prepare and analyze the quantitation standards in solutions containing matrix. However, the EC and U.S. guidelines diverge on this point. Whereas in Europe the use of standard in-matrix calibrants is encouraged, the EPA does not generally allow it. As with all mass spectrometric techniques, the use of isotopically labeled internal standards results in an improved reliability and ruggedness of methods. This approach is very well established in pharmaceutical analysis, but nonradioactivelabeled tracer compounds are available for only a limited number of pesticides and metabolites. Use of internal 13C_ or 2H-labeled standards is of particular usefulness when extraction techniques result in partial losses of analytes, as in the case with chlorinated anilines (Hurlbut et aI., 1998), which react with matrix components, and of pentachlorophenol (Gremaud and Turesky, 1997). Analytical methods using GC or LC-MSIMS and MS n techniques are per se target methods (i.e., tailored to detect with

high selectivity and sensItIvIty one or few analytes). Thus, the use of MSIMS has limitations for enforcement methods that should ideally cover a multitude of pesticides and relevant metabolites. This is especially true for LC-MS due to the limited separation power of HPLC in comparison to the much higher peak capacity obtained by capillary Gc. This disadvantage can be overcome by two approaches, as follows. First, if several rugged automated short LC-Msn methods analyze sample extracts after a general cleanup procedure consecutively for several groups or classes of compounds, this would allow unattended screening of one extract for many analytes. As sample extraction and cleanup are time-consuming steps, this approach allows increased sample throughput for enforcement purposes. Second, a technique called data-dependent full-scan MSIMS can be implemented with ion trap mass spectrometers. This application first screens in the full-scan mode the mass spectral information for compounds eluting from the HPLC column. As soon as ions formed from a potential analyte are detected in a defined retention time window, the ion trap switches to a predetermined MSIMS method for improved selectivity, thus providing additional information and confirmation. A combination of these two approaches could very well result in the future in reliable and affordable multiresidue methods for pesticides and relevant metabolites not covered by GC-based multimethods. 30.3.5 IMMUNOASSAY TECHNIQUES

The use of immunoassays (lAs) as pesticide residue analytical methods represents a radical departure from the more conventional chromatographic approaches. Immunoassays utilize antibodies that have been prepared in animals (commonly rabbits, mice, or sheep) to a particular pesticide or family of pesticides. These molecules are too small to elicit immune responses by themselves, but, upon coupling of a chemical analog of the pesticide to a carrier (usually a protein), the "conjugate" may evoke production of antibodies. For the method to be successful, these antibodies must be able to bind selectively to the free pesticide. The key steps in development of an antibody test are as follows (Gee et aI., 1995; Harris et aI., 1998): 1. Synthesis of a pesticide (or derivative), coupled to a suitable carrier protein for immunization. 2. Immunization of rabbits, mice, and/or other species; preparation and purification of antibodies. 3. Development of initial immunoassay using pesticide standards; checking assay sensitivity and specificity. 4. Assessment of assay performance with water and soil matrices in laboratory-spiked and field samples. 5. Formatting of methods as prototype kits, stabilization and stability trials on components and prototypes. 6. Field trials of kits and training workshops. The advantages associated with lA include low detection limits and high analyte selectivity. Because sample preparation

30.3 Development of the Analytical Method

is minimal, the methods allow for high throughput, thus increasing cost effectiveness. These advantages have been extensively reviewed in the literature (Harris et aI., 1995, 1998). A key limitation in the development of lA methodology is the longer time required when compared to traditional instrumental methods. Specifically, the selection of the target analyte analog (hapten) is critical for the production of the high-affinity antibodies required for high selectivity and sensitivity. The functional group used for protein coupling should not mask the key structural feature(s) of the target analytes so that the immune system of the host animal can recognize it. Hapten design and synthesis have been extensively reviewed (Goodrow et aI., 1990, 1995). The coupling (conjugation) of the hapten with a carrier protein must result in a product that is of adequate solubility and stability under the reaction conditions and that contains the appropriate functional groups (Brinckley, 1992; Erlanger, 1980). The unreacted products are then separated from the conjugate by dialysis, gel filtration, or other methods. Animal immunization to obtain polyclonal or monoclonal antibodies is conducted under conditions that enhance the immune response (Harlow and Lane, 1988; Tijssen, 1985). The affinity of the antibodies for the immunizing hapten is then evaluated to determine antibody titers against it. Inhibition experiments are then conducted to determine the potential for each target analyte toward inhibition of binding between the hapten and the antibody. Validation of the lA method is conducted after matrix effects and the influence of pH, salts, solvents, and other components are identified. As with other animal methods, validation involves determination of fortified recoveries in the appropriate matrices. Schneider et at. (1995) detail troubleshooting procedures that may be considered during method development and validation. Numerous examples of successful lA methods are reported for organophosphates, pyrethroids, triazines, urea herbicides, and other compound classes (Gee et aI., 1995; Harris et aI., 1998). For the analysis of specific agrochemicals, commercially available immunoassay kits are generally more cost effective than traditional instrumental analysis. This factor makes the continuing development of lA methods attractive, in particular, for use in monitoring studies and for pesticide analysis in developing countries. However, substantial purification of analyte from matrix is required before lA methods can be applied. 30.3.6 CAPILLARY ELECTROPHORESIS Electrophoresis refers to the migration of electrically charged species when dissolved in an electrolyte through which an electric current is passed. Capillary electrophoresis (CE) combines a variety of modem analytical techniques with a wide range of applications, including the analysis of biopolymers [such as deoxyribonucleic acid (DNA), proteins, peptides], natural products, pharmaceuticals and drugs, and fine chemicals, including agrochemicals. CE makes use of various separation modes with distinct ranges of application, such as the following:

677

1. Free-solution capillary electrophoresis (FSCE), which is mainly used for the separation of ions based on differences in their charge-to-mass ratios, whereby the pH controls the dissociation and protonation of functional groups on the analytes. 2. Micellar electrokinetic capillary chromatography (MECC or MEKC), which uses relatively high levels of ionic surfactants forming micelles. The separation of neutral analytes is based on the hydrophobic interaction with the micelles that migrate in the capillary. The use of chiral cyclodextrins provides a relatively cost effective and powerful method for enantioselective separations. 3. Capillary electrochromatography (CEC) is a fusion of liquid chromatography and capillary electrophoresis, where the capillary is packed with stationary phase similar to those used in liquid chromatography, and the flow of the mobile phase is caused by the electroosmotic flow (EOF) between the electrodes. The selectivity of the separation depends on partition between the stationary and mobile phases. Capillary electrophoresis has to be considered for the analysis of polar and charged analytes and thus follows the trend in pesticide chemistry to use more hydrophilic water-soluble active substances and for the requirement to include polar metabolites into the residue definition. For routine residue analysis, however, CE is generally only considered when conventional chromatographic approaches such as capillary GC, HPLC, and LC-MS fail to provide straightforward solutions, for the following reasons. First, expertise and instrumentation in GC, HPLC, and LC-MS are more readily available in residue research, contract, and enforcement laboratories, and the pressure to use CE for residue analysis only exists when the other techniques do not provide rational and cost-effective methods. Second, there are several aspects of CE that do not facilitate its use in residue analysis:

1. Whereas CE provides excellent sensitivity for many applications, the limitation in sample size, which is in the nanoliter range, results in insufficient overall sensitivity. Injection size can be increased by the use of a process called "stacking" where the analytes are concentrated in a sample zone prior to the chromatographic separation, thus reducing the starting peak width. Miniaturization of extraction and cleanup techniques may provide a means to obtain decreased final extract volumes. However, in residue analysis, there is a limit to decreasing the original sample size. 2. The most frequently used detection methods in CE are ultraviolet (UV) absorbance or UV diode array. Very short path lengths, however, again limit the sensitivity obtained by CE. On the other hand, the use of a mass spectrometric detector (quadrupole, triple-quadrupole, ion trap, or time-of-flight mass spectrometer, MSIMS techniques), predominantly with ESI sources, provides good sensitivity and high selectivity. The

678

CHAPTER 30


presence of electrolytes and surfactants in the caSe of MECC, however, limits applications and overall performance. In summary, with the current trend away from highly lipophilic pesticides of great environmental persistence toward more polar active substances and the necessity to detect polar metabolites, there is an increasing number of potential applications in pesticide residue analysis for CE.

30.4 SUMMARY Regulatory requirements that address pesticide concentrations in foods and environmental matrices have significantly accelerated the development of faster, less costly, and more sensitive and specific analytical techniques. Extraction techniques utilizing modem approaches such as supercritical fluid and accelerated solvent extractions allow for use of smaller sample sizes and solvent volumes. Cleanup of extracts can now be accomplished with commercial chromatography products and the process can be automated to address large sample numbers. The most dramatic improvements have taken place in the analytical instrumentation and automation options available. In particular, the development of multiple ionization techniques and secondary ion production in mass spectrometry have advanced detection limits and improved selectivity. Developments in immunoassay-based analysis promise to provide low detection limits and high analyte selectivity coupled with relatively low cost.

REFERENCES Argauer, A. l., Eller, K. 1., Pfeil, R. M., and Brown, R. l. (1997). Detennining ten synthetic pyrethroids in lettuce and ground meat by using ion-trap mass spectrometry and electron capture gas chromatography. J. Agric. Food Chem. 45, 180-184. Balinova, A. (1993). Solid-phase extraction followed by high-perfonnance liquid chromatographic analysis for monitoring herbicides in drinking water. J. Chromatogr. 643,203-207. Barnabas, I. l., Dean, l. R., Fowlis, 1. A., and Owen, S. P. (1995). Automated detennination of s-triazine herbicides using solid-phase microextraction. J. Chromatogr. 705, 305-312. Belardi, R. P., and Pawliszyn, l. B. (1989). The application of chemically modified fused silica fibers in the extraction of organics from water matrix samples and their rapid transfer to capillary columns. Water Pollut. Res. J. Can. 24, 179-189. Bond, P. M. (1994). Supercritical fluid chromatography-Finding a place in the chromatographer's arsenal? Pestic. Sci. 41, 369-374. Boyd-Boland, A. A., Magdic, S., and Pawliszyn, l. (1996). Simultaneous detennination of 60 pesticides in water using solid phase microextraction and gas chromatography-mass spectrometry. Analyst 121, 929-938. Brinkley, M. (1992). A brief survey of methods for preparing protein conjugates with dyes, haptens and cross-linking reagents. Biocon}. Chem. 3, 2. Brooks, M. W., and Uden, P. C. (1995). Detennination of abamectin from soil and animal tissue by SFE and fluorescence detection. Pes tic. Sci. 43, 141146. Brown, M. A. (1990) (ed.) "Liquid ChromatographylMass Spectrometry;' ACS Symposium Series 420. Am. Chem. Soc., Washington, DC.

Cairns, T., and Siegmund, E. G. (1990). The development ofliquid chromatography/mass spectrometry. In "Liquid ChromatographylMass Spectrometry" (M. A. Brown, ed.), pp. 1-15, ACS Symposium Series 420. Am. Chem. Soc., Washington, DC. Chambers, J. (1998). Miniaturized GPC for the effective cleanup of crop extracts containing pesticide residues. In "Ninth International Congress of Pesticide Chemistry (IUPAC)," Vol. 2, 7 A-046. Conte, E., Milani, R., Morali, G., and Aballe, F. (1997). J. Chromatogr., A 765, 121-125. David, M. D., and Seiber, l. N. (1996). Comparison of extraction techniques, including supercritical fluid, high-pressure solvent, and Soxhlet, for organophosphorus hydraulic fluids from soil. Anal. Chem. 68, 3038-3044. de Cruz, I., Lacroix, G., Mougin, C., and Grolleau, G. (1996). Residues of chlorinated pesticides in eggs of the gray heron: Contribution of capillary gas chromatography ion-trap spectrometry. J. High Resolut. Chromatogr. 19, 62--64. EC Directorate General for Agriculture (1998). "Guidance Document on Residue Analytical Methods." 8064NIJ97-rev 4 15/12/98. Eisert, R., and Levsen, K. (1995a). Detennination of organophosphorus, triazine and 2,6-dinitroaniline pesticides in aqueous samples via solidphase microextraction (SPME) and gas chromatography with nitrogenphosphorus detection. Fresenius' J. Anal. Chem. 351, 555-562. Eisert, R., and Levsen, K. (1995b). Detennination of pesticides in aqueous samples by solid-phase microextraction in-line coupled to gas chromatographymass spectrometry. J. Am. Soc. Mass Spectrom. 6, 1119-1130. ErIanger, B. F. (1980). The preparation of antigenic hapten-carrier conjugates: A survey. Methods Enzymol. 70, 85. Erstfeld, K. M., and Chen, C. Y. (1998). Comparison of supercritical fluid and Soxhlet extraction of chlorothalonil from cranberry bog soils. J. Agric. Food Chem. 46, 499-503. Ezzell, J. (1998). The use of SW-846 Method 3545 for automated extraction of environmental samples. Am. Environ. Lab. 24-25. Ezzell, J., Richter, B., Felix, D., Black, S., and MeikIe, J. (1995). A comparison of accelerated solvent extraction with conventional solvent extraction for organophosphorus pesticides and herbicides. LC-GC 390-398. Fahmy, T. M., Paulaitis, M. E., Johnson, D. M., and McNally, M. E. P. (1993). Modifier effects in the supercritical fluid extraction of solutes from clay, soil, and plant materials. Anal. Chem. 65, 1462-1469. Feigel, C. (1997). GCIMSIMS: Method of choice for pesticide residue analysis. Inside Laboratory Management 34-35. Field, J. A., and Monahan, K. (1995). In-vial derivatization and Empore disk elution for the quantitative detennination of the carboxylic acid metabolites of dacthal in groundwater. Anal. Chem. 34, 3357-3362. Field, J. A., and Monahan, K. (1996). Chlorinated acid herbicides analysis in water by strong anion-exchange disk extraction and in-vial elution and derivatization. J. Chromatogr., A 741, 85-90. Field, J. A., Reed, R. L., Sawyer, T. E., and Martinez, M. (1997). Diuron and its metabolites in surface and ground water by solid phase extraction and in-vial elution. J. Agric. Food Chem. 45, 3897-3902. Fillion, J., Hindle, R., Lacroix, M., and Selwyn, J. (1995). Multiresidue detennination of pesticides in fruits and vegetables by gas chromatographymass-selective detection and liquid chromatography with fluorescence detection. J.-Assoc. Off. Anal. Chem. 78, 1252-1266. Frost, S., Dean, J., Evans, K., Harradine, K., Cary, C., and Comber, M. (1997). Extraction of hexaconazole from weathered soils: A comparison between Soxhlet extraction, microwave assisted extraction, supercritical fluid extraction and accelerated solvent extraction. Anal. Chem. 69,2171-2180. Gee, S. J., Hammock, B. D., and Skenitt, J. H. (1995). Diagnostics for plant agrochemicals-A meeting of chemistry and immunoassay. In "New Diagnostics in Crop Sciences" (1. H. Skenitt, and R. Appels, eds.), pp. 243-276. CAB, International, Oxford. Gilbert, J. D., Olah, T. v., and McLaughlin, D. A. (1995). Biochemical and biotechnological applications of electrospray ionization mass spectrometry. In "ACS Symposium Series 619" (A. P. Snyder, ed.), pp. 330-350. Am. Chem. Soc., Washington, DC.

References

Goodrow, M. H., Harrison, R. 0., and Hammock, B. D. (1990). Hapten synthesis, antibody development, and competitive inhibition enzyme immunoassay for s-triazine herbicides. J. Agric. Food Chem. 38, 990. Goodrow, M. H., Sanborn, J. R., Stoutamire, D. W., Gee, S. J., and Hammock, B. D. (1995). Strategies for immunoassay hapten design. In "Immunoanalysis of Agrochemicals: Emerging Technologies" (J. O. Nelson, A. E. Karu, and R. B. Wong, eds.), pp. 119-126, ACS Symposium Series 586. Am. Chem. Soc., Washington, DC. Gremaud, E., and Turesky, R. J. (1997). Rapid analytical methods to measure pentachlorophenol in wood. J. Agric. Food Chem. 95, 1229-1233. Harlow, E., and Lane, D. (1988). "Antibodies: A Laboratory Manual." Cold Spring Harbor Laboratory Press, Cold Spring Harbor, NY. Harris, A. S., Lucas, A. D., Kriimer, P. M., Marco, M. P., Gee, S. J., and Hammock, B. D. (1995). Use of immunoassay for the detection of urinary biomarkers of exposure. In "New Frontiers in Agrochemical Immunoassay" (D. A. Kurtz, J. H. Skerritt, and L. Stanker, eds.), p. 217. Assoc. Official Anal. Chem., Arlington, VA. Harris, A. S., Wengatz, I., Wortberg, M., Kreissig, S. B., Gee, S. J., and Hammock, B. D. (1998). Development and application of immunoassays for biological and environmental monitoring. In "Multiple Stresses in Ecosystems" (J. J. Cech, B. W. Wilson, and D. G. Crosby, eds.), pp. 135-153. Lewis Publishers, CRC Press. Hatdk, S., Lehotay, J., and Tekel, J. (1994). Simultaneous HPLC determination of phenylurea herbicides and their corresponding anilines in water after online preconcentration. J. High Resolut. Chromatogr. 17,756-758. Hawthorne, S. B., Miller, D. J., and Langefeld, J. J. (1992). Advances in analytical supercritical fluid extraction (SFE). In "Hyphenated Techniques in Supercritical Fluid Chromatography and Extraction" (K. Jinno, ed.), Chap. 12. Elsevier, Amsterdam. Hurlbut, D. B., Johnston, J. J., Daniel, S. R., and Tawara, J. (1998). Gas chromatography/mass spectrometry method for the quantitation of 3-chloro-ptoluidine hydrochloride in birds using a deuterated surrogate. J. Agric. Food Chem. 46,4610-4615. Inspectorate for Health Protection (1996). "Analytical Methods for Pesticide Research," 6th ed. Inspectorate for Health Protection, Rijswijk, The Netherlands. Jenke, D. R. (l996a). Chromatographic method validation: A review of current practices and procedures. I. General concepts and guidelines. J. Liq. Chromatogr. Re/at. Technol. 19,719-736. Jenke, D. R. (1996b). Chromatographic method validation: A review of current practices and procedures. Il. Guidelines for primary validation parameters. J. Liq. Chromatogr. Relat. Technol. 19,737-757. Jenke, D. R. (1996c). Chromatographic method validation: A review of current practices and procedures. Ill. Ruggedness, re-validation, and system suitability. J. Liq. Chromatogr. Re/at. Technol. 19, 1873-1891. Jones, F. W. (1996). Multi-residue analysis of pesticides in wool wax and lanolin using gel permeation and gas chromatography. J. Agric. Food Chem. 44, 3197-3201. Jones, E W. (1997). SFE as a cleanup technique for gas chromatographic analysis of pesticides in wool wax. J. Agric. Food Chem. 45, 2569-2572. King, J. W. (1989). Fundamentals and applications of supercritical fluid extraction in chromatographic science. J. Chromatogr. Sci. 27, 355-364. King, J. w., Hopper, M. L., Lutchefeld, R. G., Taylor, S. L., and Orton, W. L. (1993). Optimization of experimental conditions for the supercritical carbon dioxide extraction of pesticides residues from grains. J.-Assoc. Off. Anal. Chem. Int. 76,857-864. Liao, W., Joe, T., and Cusick, W. G. (1991). Multi-residue screening method for fresh fruits and vegetables with gas chromatographic/mass spectrometric detection. J.-Assoc. Off. Anal. Chem. 78, 554-565. Lira, C. T. (1988). Physical chemistry of supercritical fluids. In "Supercritical Fluid Extraction and Chromatography" (B. A. Charpentier and R. S. Michael, eds.), Chap. 1. Am. Chem. Soc., Washington, DC. Lopez-Avila, v., and Dodhiwala, N. S. (1990). Supercritical fluid extraction and its application to environmental analysis. J. Chromatogr. Sci. 28, 468-476. Lou, X., Janssen, H., and Cramers, C. (1997). Parameters affecting the accelerated solvent extraction of polymeric samples. Anal. Chem. 69,1598-1603.

679

Lunn, G., and Hallwig, L. C. (1998). "Handbook of Derivatization Reactions for HPLC." Wiley, New York. Marce, R. M., Prosen, H., Crespo, c., Calull, M., Borrull, E, and Brinkman, U. A. Th. (1995). On-line trace enrichment of polar pesticides in environmental waters by reversed-phase liquid chromatography-diode array detection-particle beam mass spectrometry. J. Chromatogr., A 696, 63-74. March, R. E. (1997). An introduction to quadrupole ion trap mass spectrometry. J. Mass Spectrom. 32,351-369. Maris, E A., Geerdink, R. B., Frei, R. w., and Brinkman, U. A. Th. (1985). On-line trace enrichment for improved sensitivity in liquid chromatography with direct liquid introduction mass spectrometric detection. J. Liq. Chromatogr. 323, 113-120. McNally, M. E. P., Deardorff, C. M., and Fahmy, T. M. (1992). Supercritical fluid extraction-new directions and understandings. In "Supercritical Fluid Technology" (E Bright and M. E. McNally, eds.), Chap. 12. Am. Chem. Soc., Washington, DC. Moore, K. M., Jones, S. R., and James, C. (1995). Multi-residue analytical method for diuron and carbamate pesticides in water using solid-phase extraction and liquid chromatography-mass spectrometry. Water Res. 29, 1225-1230. Nielen, M. W. E, Valk, A. J., Frei, R. w., Brinkman, U. A. Th., Mussche, Ph., De Nijs, R., Ooms, B., and Smink, W. (1987). Fully automated sample handling system for liquid chromatography based on pre-column technology and automated cartridge exchange. J. Chromatogr. 393, 69-83. Nouri, B., Fouillet, B., Toussiant, G., Chambon, P., and Chambon, R. (1995). High-performance liquid chromatography with diode array detection for the determination of pesticides in water using automated solid-phase extractions. Analyst 120, 1133-1136. Parrilla, P., Martinez-Vidal, J. L., and Fernandez-Alba, A. R. (1993). Optimization of the separation, isolation and recovery of selected pesticides in water samples by solid-phase extraction and HPLC photodiode array detection. J. Liq. Chromatogr. 16,4019-4029. Ragsdale, N. N. (1998). Impact of US Food Quality Protection Act on pesticide research priorities. In "Ninth International Congress of Pesticide Chemistry (IUPAC)," Vol. 2, 8A-006. Richter, B., Jones, B., Ezzell, J., and Porter, N. (1996). Accelerated solvent extraction: A technique for sample preparation. Anal. Chem. 68, 1033-1039. Riekkola, M. L., Manninen, P., and Hartonen, K. (1992). SFE, SFE/GC and SFE/SFC: Instrumentation and applications. In "Hyphenated Techniques in Supercritical Fluid Chromatography and Extraction" (K. Jinno, ed.), Chap. 14. Elsevier, Amsterdam. Rossi, D., Hoffman, K., Janiczek-Dolphin, N., Bockbrader, H., and Parker, T. (1997). Tandem-in-time mass spectrometry as a quantitative bioanalytical tool. Anal. Chem. 69,4519-4523. Sancho, J. v., Hidalgo, c., and Hernandez, E (1997). Direct determination of bromacil and diuron residues in environmental water samples by coupled-column liquid chromatography and large-volume injection. J. Chromatogr., A 761, 322-326. Schachterle, S., and Feigel, C. (1996). Pesticide residue analysis in fresh produce by gas chromatography-tandem mass spectrometry. J. Chromatogr., A 754,411-422. Schneider, P., Gee, S. J., Kreissig, S. B., Harris, A. S., Kriimer, P., Marco, M. P., Lucas, A. D., and Hammock, B. D. (1995). Troubleshooting during the development and use of immunoassays for environmental monitoring. In "New Frontiers in Agrochemical Immunoassay" (D. A. Kurtz, J. H. Skerritt, and L. Stanker, eds.), pp. 103-130. Assoc. Official Anal. Chem., Arlington, VA. Sennert, S., Volmer, D., Levsen, K., and Wiinsch, G. (1995). Multiresidue analysis of polar pesticides in surface and drinking water by on-line enrichment and thermospray LC-MS. Fresenius' J. Anal. Chem. 351, 642-649. Shalaby, L. M., and George, S. W. (1990). Multiresidue analysis of thermally labile sulfonylurea herbicides in crops by liquid chromatography/mass spectrometry. In "Liquid ChromatographylMass Spectrometry" (M. A. Brown, ed.), pp. 75-92, ACS Symposium Series 420. Am. Chem. Soc., Washington, DC. Skopec, Z. v., Clark, R., Harvey, P. M. A., and Wells, R. J. (1993). Analysis of organophosphorus pesticides in rice by supercritical fluid extraction and

680

CHAPTER 30


quantitation using an atomic emission detector. J. Chromatogr. Sci. 31,445449. Snyder, J. L., Grob, R. L., McNaIIy, M. E., and Oostdyk, T. S. (1992). Comparison of supercritical fluid extraction with classical sonication and Soxhlet extractions of selected pesticides. Anal. Chem. 64, 1940-1946. Snyder, J. L., Grob, R. L., McNaIIy, M. E., and Oostdyk, T. S. (1993). The effect of instrumental parameters and soil matrix on the recovery of organochlorine and organophosphate pesticides from soil using supercritical fluid extraction. J. Chromatogr. Sci. 31, 183-191. Specht w., Pelz, S., and Gilsbach, W. (1995). Gas-chromatographic determination of pesticide residues after cleanup by gel-permeation chromatography and mini-silica gel chromatography. Fresenius' J. Anal. Chem. 353, 183190. Stanek, M., and KeIIer, G. (1998). Determination of pesticide residues using accelerated solvent extraction. In "Ninth International Congress of Pesticide Chemistry (IUPAC)," Vol. 2, 7A-034. Startin, J. R., Hird, S. J., Jones, A., and HilI, A. R. C. (1998). Analysis of residues of paraquat and diquat in plant and animal tissues by LC-MS. In "Ninh International Congress of Pesticide Chemistry Book of Abstracts," Vol. 2, 7A-007.

Steen, R. J. C. A., Freriks, I. L., Cofino, W. P., and Brinkman, U. A. Th. (1997). Large-volume injection gas chromatography-ion trap tandem mass spectrometry for the determination of pesticides in the marine environment at low ng/llevel. Anal. Chem. Acta 353, 153-163. Thier, H. P., and Zeumer, H. (1987). "Manual of Pesticide Residue Analysis." DFG Pestic. Comm., Weinheim. Tijssen, P. (1985). "Laboratory Techniques in Biochemistry and Molecular Biology, Volume 15: Practice and Theory of Enzyme Immunoassays." Elsevier, Amsterdam. Urruty, L., Montury, M., Braci, M., Fournier, J., and Dournel, J. M. (1997). Comparison of two recent solventless methods for determination of procymidone residues in wine. J. Agric. Food Chem. 45,1519-1522. U.S. Environmental Protection Agency (EPA) (1996). "Residue Chemistry Test Guidelines," OPPTS 860 Series. Yu, L., Schoen, R., Dunkin, A., Firman, M., and Cushman, H. (1997). Rapid identification and quantitation of diphenylamino, o-phenylphenol and propargite residues on apples by GC-MS. J. Agric. Food Chem. 45, 748-752. Zhang, Z., Yang, M. J., and Pawliszyn, 1. (1994). Solid-phase microextraction. Anal. Chem. 66, 844-853.

CHAPTER

31 Risk Assessment and Risk Management: The Regulatory Process* Penelope A. Fenner-Crisp U.S. Environmental Protection Agency

31.1 INTRODUCTION In the United States, primary authority for pesticide regulation resides with the US. Environmental Protection Agency (EPA) under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) and the Federal Food, Drug, and Cosmetic Act (FFDCA). Under FIFRA, EPA registers pesticides for use. This law also authorizes the Agency to prescribe the conditions of use of pesticide products. Under FFDCA, EPA establishes maximum allowable levels of pesticide residues ("tolerances") in foods and animal feeds. These tolerances are enforced by the Food and Drug Administration (FDA) of the Department of Health and Human Services (HHS) for most foods and by the U.S. Department of Agriculture (USDA) for meat, poultry, and some egg products.

31.2 HISTORICAL BACKGROUND OF PESTICIDE REGULATION IN THE UNITED STATES Regulation of pesticides at the federal level has been in place for nearly a century. Each time the law has been amended, the number and nature of the directives have been expanded and embellished. Some, but not all, of these changes will be described. Only those areas of the law(s) that remain in effect under current legislation are discussed. The first legislation passed was the Federal Insecticide Act of 1910 (FIA, 1910). The provisions of this act, essentially only a labeling statute, were limited to the prohibition of the manufacture of any insecticide or fungicide that was "adulterated or misbranded." No requirement for registration and no establishment of standards of safety were included at that time. Congress *This document reflects the opinions only of the author and does not necessarily represent official policy of the U.S. Environmental Protection Agency. Handbook of Pesticide Toxicology Volume 1. Principles

passed the first version of the Federal Insecticide, Fungicide, and Rodenticide Act in 1947, adding the requirement of registration by the Secretary of Agriculture before sale or distribution in interstate or foreign commerce, but without providing the US. Department of Agriculture the power to deny or cancel a registration if the registration did not comply with the provisions of the law (FIFRA, 1947) A 1964 amendment to FIFRA did provide USDA with the authority to deny or rescind a registration and to issue an immediate suspension of registration if necessary to prevent an imminent hazard to human health (FIFRA,1964). Nearly half a century after Congress passed the first federal pesticide regulatory legislation, it amended the Federal Food, Drug, and Cosmetic Act to require the Food and Drug Administration to establish maximum acceptable levels ("tolerances") for pesticide residues in foods and animal feeds (FFDCA, 1954). This requirement applied only to raw agricultural commodities. Four years later, Congress once again amended FFDCA to include a requirement for a tolerance in a processed food, but only if the pesticide residue in that processed food was expected to exceed the tolerance level in the related raw agricultural commodity (FFDCA, 1958). In 1970, under President Nixon's government Reorganization Plan No. 3, the primary federal authority for the regulation of pesticides was transferred from USDA and FDA to the newly created EPA (Nixon, 1970). Between 1970 and 1990, Congress amended FIFRA six times, each time adding to, or enhancing, the Agency's existing responsibilities. One of the provisions added in 1972 was that a pesticide could be registered only if it did not cause "unreasonable adverse effects" on human health or the environment [Federal Environmental Pesticide Control Act of 1972 (FEPCA, 1972)]. The 1972 revisions also established the requirement for reregistration of all existing pesticides within a five-year time frame. When this was not accomplished by 1978, Congress relaxed the timelines [Federal Pesticide Act of 1978 (FPA, 1978)]. Ten years later,

681

Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved.

682

CHAPTER 31

Risk Assessment and Risk Management

following continued delays in completing the reregistration process, Congress once again established specific timetables in a five-phase program for all active ingredients registered before November 1, 1984, but this time with targeted funding to support the reregistration activity (FIFRA, 1988). Completion of the reregistration process is estimated for 2006. Other important provisions in the FIFRA amendments of 1972 included authorization (1) to require federal registration of pesticides sold within states, (2) to classify pesticides in general and/or restricted use categories (based upon their inherent acute toxicity) and otherwise regulate their usage (e.g., specify the maximum allowable application rate and frequency), and (3) to register establishments that make pesticide products and require them to maintain records, while also being able to inspect these producers, as well as those establishments which hold pesticides for sale, for compliance with the applicable provisions of FIFRA. Amendments to FIFRA in 1975 included the requirement for EPA to notify the Secretary of Agriculture in advance of issuing proposals for regulations or to cancel or otherwise change the registration status of a pesticide, and to consider the impact on agriculture when cancellation actions are being considered (FIFRA, 1975). In addition, the seven-member FIFRA Scientific Advisory Panel (SAP) was established to "comment as to the impact on health and the environment" of proposed cancellation actions and regulations. In the 25 years since this provision was added to FIFRA, the SAP has been consulted on many scientific issues reflecting a much broader range of EPA's pesticide regulatory activities, for example, the proposed classification of human cancer potential of a pesticide, the design of a testing protocol, and the risk assessment methodologies developed to address aggregate and cumulative exposure and risk assessment. Congress once again substantially revised FIFRA in 1978 (FIFRA, 1978). Provisions included granting data submitters 10 years of exclusive use of their data on new active ingredients while transferring the responsibility for managing the issue of data compensation from the Agency to outside arbitrators. The 1978 revisions also removed most trade secret protection for health and safety data, an early example of public rightto-know. Other changes included the granting of conditional registration authority which allows EPA to approve proposed uses before the full set of supporting data are submitted and reviewed; the establishment of a procedure for interim administrative review (a process known as Special Review) if "a validated test or other significant evidence raised prudent concerns of unreasonable adverse risk to man or to the environment;" a prohibition against disclosure of data to foreign or multinational pesticide producers; and a requirement for the recognition of the distinction between agricultural and nonagricultural pesticides when processing registration petitions and in setting registration standards and guidelines. Among the modifications to FIFRA in 1980 was the provision that the Scientific Advisory Panel could create its own subpanels, resolving the limitation in scope of expertise that the smaller, seven-member permanent panel may have on any

specific issue (FIFRA, 1980). This allowed expansion, of the panel's capabilities to provide more substantive expert scientific peer review of the increasingly diverse and complex issues brought to it by the Agency. The 1980 provisions also required EPA to request SAP comment "as to the impact on health and the environment" on proposed suspension actions (in addition to the requirements for consultation on proposed cancellations) and to set up a process for peer review "with respect to the design, protocols, and conduct of major scientific studies conducted under this Act by the Environmental Protection Agency or by any other Federal agency, any State or political subdivision thereof, or any institution or individual under grant, contract, or cooperative agreement from or with the Environmental Protection Agency" and of "the results of any such scientific studies relied upon by the Administrator with respect to any actions the Administrator may take relating to the change in classification, suspension, or cancellation of a pesticide." In 1988, in addition to prescribing the multi phase reregistration program, Congress made other changes to FIFRA, including creation of a "fast-track" registration process for enduse products for "me-too" registrants ("me-too" registrants are those who seek to register products similar to an already registered pesticide product); the necessity of gaining Congressional approval to indemnify registrants holding suspended or cancelled products; and a series of provisions related to the storage, disposal, and transport of suspended or cancelled pesticides and pesticide containers (FIFRA, 1988). Record keeping requirements were expanded to include all registrants and applicants for registration, in addition to those previously required of producers. The Food, Agriculture, Conservation, and Trade Act of 1990 (FACTA, 1990) added requirements for certified pesticide applicators to maintain records of their use of restricted-use chemicals, prohibited registrants of minor-use pesticides from submitting field trial data from geographic areas where the chemical would not be used, and provided discretion to the Administrator to reduce or waive registration fees if the cost would "significantly reduce the availability of the pesticide." In addition, there were several new requirements related to voluntary cancellation of minor-use pesticides, including a provision for public notice and comment upon a registrant's application for voluntary cancellation.

31.3 CURRENT STATE OF PESTICIDE REGULATION IN THE UNITED STATES-THE FOOD QUALITY PROTECTION ACT OF AUGUST 3, 1996 Amendments to FIFRA and FFDCA in 1996 brought both incremental and broad, sweeping changes to the legal foundation for pesticide regulation in the United States [The Food Quality Protection Act of 1996 (FQPA, 1996)]. The Food Quality Protection Act represents the outcome of long and complex deliberations to resolve inconsistencies in the previous legislation, both within and between the two statutes.

31.3 Current State of Pesticide Regulation in the United States

683

FIFRA §4 specifies that tolerances and exemptions from tolerances must be reassessed as part of reregistration to determine whether they meet the requirements of the FFDCA. These reassessments must be made as soon as EPA has sufficient information to assess dietary risk, but no later than when it makes product reregistration decisions. These determinations for existing tolerances and exemptions and the need for any additional tolerances or exemptions must be published in the Federal Register and appropriate regulatory action under FIFRA and/or FFDCA begun promptly.

economic incentive to support the costs of registration or reregistration. In addition, the minor use pesticide must play a significant role in managing pest resistance or in an integrated pest management (IPM) program. There also must be a lack of efficacious alternatives or the alternatives must pose a greater risk to human health or the environment than does the pesticide under evaluation. Many minor-use crops are fruits and vegetables, which are significant components of the human diet. The law provides additional incentives for the development and maintenance of minor use registrations in a number of ways: extension of time to generate residue data and for exclusive use of these data; greater flexibility to waive data requirements; the option to waive some or all of the fees usually charged to support and maintain registration; and expedited review of minor-use applications by the Agency. None of these provisions would apply, however, if the minor use is determined to pose unreasonable risks or if the lack of data would significantly delay EPA decisions. FQPA establishes a USDA revolving grant program and a program for support of public health pesticides to be implemented jointly by the Public Health Service of HHS and EPA. By virtue of instituting this program, the federal government bears the cost of developing the required data to support the registration and reregistration of the public health use, as it does for minor-use pesticides used on agricultural crops under the USDA Inter-Regional Project Number 4 (IR-4) program. A public health pesticide is defined in FIFRA §2(nn) as "any minor use pesticide product registered and used predominantly in public health programs for vector control or for other recognized health protection uses, including the prevention of viruses, bacteria, or other microorganisms (other than viruses, bacteria, or other microorganisms on or in living man or other animal) that pose a threat to public health." A "vector" is defined in FIFRA §2(00) as "any organism capable of transmitting human discomfort or injury, including mosquitoes, flies, fleas, cockroaches, or other insects and ticks, mites, or rats." Perhaps the most notable example of a public health pesticide is DDT, s~ill used in some parts of the world, but not the United States, for the control of mosquitoes bearing the malaria vector.

31.3.1.4 Registration Renewal

31.3.1.6 Antimicrobial Pesticides

The requirement for periodic registration review was introduced. Formal procedures are to be established, with the aim of updating a pesticide's registration eligibility at least once every 15 years. The goal of this requirement is to ensure that all pesticides continue to meet up-to-date standards for safety testing and the protection of human health and the environment.

FIFRA §2(mm) defines an antimicrobial pesticide as one which is intended to "disinfect, sanitize, reduce, or mitigate growth or development of microbiological organisms" or "protect inanimate objects, industrial processes or systems, surfaces, water, or other chemical substances from contamination, fouling, or deterioration caused by bacteria, viruses, fungi, protozoa, algae, or slime" and, in this use, is exempt from a tolerance under FFDCA §408. Wood preservatives, antifouling paints, agricultural fungicides, aquatic herbicides, and liquid chemical sterilants intended for use on critical or semicritical medical devices as defined under FFDCA are not included within the definition. FQPA contains special provisions for antimicrobial pesticides, essentially removing them from the shadow of pesticides intended for agricultural and other nonfood uses and prompting

31.3.1 FIFRA-KEY CHANGES AND ADDITIONS 31.3.1.1 Emergency Snspension EPA may now suspend a pesticide registration in an emergency situation without simultaneously issuing a notice of intent to cancel, a change from the previous requirement for simultaneous action. A notice of intent to cancel must be issued within 90 days or the suspension will automatically expire. Determination of imminent hazard (to human health or the environment) constitutes grounds for suspension. This process is invoked to prevent unacceptable risks from occurring during the time required to cancel or otherwise modify the registration of a pesticide. Any action to suspend, cancel or modify the registration of a pesticide under FIFRA must be accompanied by a similar and simultaneous action on any associated tolerances under FFDCA. 31.3.1.2 FIFRA Scientific Advisory Panel A Science Review Board to consist of 60 scientists was established to be available to the permanent panel to assist in the scientific peer reviews conducted by the panel. Formation of the board complements the earlier modification to FIFRA which allowed the panel to create its own subpanels as needed. 31.3.1.3 Tolerance Reevaluation as Part of Reregistration

31.3.1.5 Protections for Minor-Use Pesticides, Including Public Health Pesticides A minor use is defined as one in which the pesticide is used on an animal, on a commercial agricultural crop for which the total U.S. acreage is less than 300,000 acres, or for the protection of public health, but does not, on its own, provide sufficient

684

CHAPTER 31


more focused attention on facilitating more timely registration decisions. The law requires the identification, evaluation, and implementation of reforms to the registration process for this class of pesticides to reduce review periods, providing explicit goals in number of days depending upon the action requested (i.e., a new use of an already-registered active ingredient; a new product; "me-too's"; and amendments to existing uses). A separate administrative unit has been established within the Office of Pesticide Programs that deals only with the registration, reregistration, and Special Review processes for antimicrobial pesticides. A separate section in 40 CFR 158 describes the data requirements necessary to support registration or registration for these products. 31.3.1.7 Reduced Risk or "Safer" Pesticides In the early 1990s, the Office of Pesticide Programs set up a system by which reduced risk or "safer" pesticides would be given priority attention in the registration process. The most current guidelines governing expedited review of conventional and biological pesticides were issued in 1997 [U.S. Environmental Protection Agency (U.S. EPA, 1997b)]. FQPA provided the statutory mandate for continuing this expedited consideration of applications for pesticides which meet one or more of the criteria for a reduced risk pesticide. A pesticide qualifies for expedited review as a reduced risk pesticide if its use "may reasonably be expected to accomplish 1 or more of the following:" (1) reduce the risks to human health; (2) reduce the risks to nontarget organisms; (3) reduce the potential for contamination of ground water, surface water, or other valued environmental resources; and (4) broaden the adoption of IPM strategies, or make them more available or effective [FIFRA §3(c)(10)(B)]. 31.3.1.8 Data Collection The keystone of FQPA is the inclusion of special provisions for infants and children. Title III ofFQPA addresses data collection activities to assure the health of infants and children. It states that USDA, in cooperation with FDA and/or EPA, "shall coordinate the development and implementation of survey procedures to ensure that adequate data on food consumption patterns of infants and children are collected"; shall ensure there will be improved data collection on occurrence of pesticide residues in foods, particularly those most likely consumed by infants and children; and shall evaluate the current status of pesticide usage information and move to improve usage information gathering activities. Information in all three of these areas is critical to the conduct of credible and accurate estimates of risk from exposure to pesticide residues in the diet. 31.3.2 FFDCA-KEY CHANGES AND ADDITIONS The most significant changes in pesticide regulation resulting from the passage of FQPA impact the tolerance-setting process described in FFDCA. Definitional and process changes

were mandated and the factors that are to be considered when conducting risk assessments and making risk management decisions were expanded. 31.3.2.1 The Delaney Clause Until FQPA was passed, pesticide residues in processed foods were considered to be "food additives" regulated under FFDCA §409. If a pesticide residue was expected to exceed the level which was allowed under a §408 tolerance for the raw agricultural commodity, it became necessary to establish a separate food additive regulation for the processed food under §409. However, the Delaney clause in §409 prohibits the establishment of food additive regulations for any substance "if it is found to induce cancer when ingested by man or animal, or if it is found, after tests which are appropriate for the evaluation of the safety of food additives, to induce cancer in man or animal. ... " Under the new law, pesticide residues are excluded from the definition of "food additive." Thus, the Delaney clause is no longer applicable to pesticide residues in processed foods. All pesticide residues, whether in raw or processed foods, are regulated only under FFDCA §408, which does not contain the prohibition against setting tolerances for carcinogens. 31.3.2.2 Definition of "Safe" Under FFDCA §408(b )(2)(A), the standard for establishing a tolerance is based on whether the tolerance is "safe." To be "safe" means that there is "a reasonable certainty that no harm will result from aggregate exposure to the pesticide chemical residue, including all anticipated dietary exposures and all other exposures for which there is reliable information." This definition is consistent with the standard applied historically to nonpesticide food additives and color additives by the Food and Drug Administration. For threshold effects (i.e., those effects for which a level can be identified that would not be expected to cause or contribute to any adverse human health consequences), the safety standard is satisfied if the aggregate exposure is lower than the no-effect level by "an ample margin of safety." Traditional regulatory policy states that, as a default, exposure estimated to be at or below a level WO-fold lower than the critical no-effect level identified in the animal toxicology database would meet the safety standard. For nonthreshold effects (i.e., those for which no no-effect level can be identified), a pesticide will satisfy the safety standard if the increased lifetime risk, expressed as a probability, is "negligible." Traditionally, "negligible" has been defined as being no greater than a one-in-a-million excess lifetime risk for nonoccupational exposures. 31.3.2.3 Aggregate Exposure Aggregate exposure is defined as that which occurs from all food uses for the pesticide, as well as from exposure that occurs from all nonoccupational sources. This would include exposures from drinking water, nonfood pesticidal uses (e.g., lawn and garden use or indoor residential, school, or public building

31.3 Current State of Pesticide Regulation in the United States

applications) and those exposures that may result from nonpesticidal uses (e.g., as a human pharmaceutical or a hazardous waste site contaminant). Principles for conducting an aggregate exposure assessment are being developed by EPA (U.S. EPA, 1999d). Among the factors that must be taken into account when establishing, modifying, leaving in effect, or revoking a tolerance or an exemption from a tolerance is the risk that may ensue from the aggregate exposure to the pesticide under evaluation. A tolerance represents a single pesticide-use combination. That is, one tolerance would be needed if Chemical X were to be used on potatoes. A separate tolerance would be required if Chemical X also were to be used on lettuce. Therefore, when making a decision with regard to anyone use, EPA must consider the exposure and risk that would occur not only as a consequence of that particular use, but also all other existing food and nonfood uses. In other words, can this new use be added to the existing "risk cup" for Chemical X? 31.3.2.4 Common Mechanism of Toxicity and Cumulative Risk Assessment Another factor that must be taken into account when establishing, modifying, leaving in effect, or revoking a tolerance or an exemption from a tolerance is the cumulative effects of the pesticide under evaluation and other substances with which it may share a common-mechanism of toxicity. "Other substances" are not just other pesticides, but may be chemicals such as drugs, commodity chemicals, or environmental contaminants. When a common mechanism finding is made, then a cumulative risk assessment is to be conducted. The first step is to determine the need for a cumulative risk assessment. This is done by conducting a hazard assessment for that pesticide. In the course of performing this assessment, information would come to light to suggest that this pesticide may share a common mechanism with at least one other pesticide. This process continues until all likely pesticide candidates are identified. After it is concluded that two or more pesticides are candidates for a common mechanism group, an effort is made to determine if reliable information exists to suggest that any non pesticides also may share the same mechanism of toxicity. EPA has developed criteria by which to judge whether substances share a common mechanism of toxicity (V.S. EPA, 1999a). At the conclusion of the process to determine those substances which actually do share a common mechanism of toxicity, those remaining substances are subjected to cumulative risk assessment. It is possible that the final cumulative risk assessment may not include all of the substances that constituted the original common mechanism group. Modifications to group membership would be informed by the results of the individual aggregate exposure assessments that also would be conducted on each original candidate for the common mechanism group. The nature, magnitude, and timing of exposure to each substance in the aggregate and how the exposures (or their biological consequences) to individual substances overlap become critical factors in determining the final group to be included in the cumulative risk assessment. Guidance is being developed by EPA for the conduct of cumulative

685

risk assessments. Early articulations of principles are presented in several Agency documents (V.S. EPA, 1999a, b, 2000a, b). The first group of substances to be subjected to a cumulative risk assessment are the cholinesterase-inhibiting organophosphorus insecticides. 31.3.2.5 Special Considerations for Infants and Children When making tolerance decisions, EPA also must implement several new requirements related to assuring the safety of infants and children. The Agency must assess (aggregate) risk based upon available information about: (1) dietary consumption patterns that are likely to yield disproportionately higher exposures or risks; (2) special susceptibilities to pesticides, including neurological differences between infants and children and adults, and the effects of in utero exposure to pesticide chemicals; and (3) the cumulative effects of the pesticide residues and other substances that have a common mechanism of toxicity. The "reasonable certainty of no harm" safety standard must be ensured and a specific safety determination for infants and children must be made. The provision that has prompted the most controversy and has had the greatest impact upon the risk assessment and regulatory decision-making process under FQPA is the obligatory application of an additional safety factor. FFDCA §408(b )(2)(C) states that "in the case of threshold effects ... an additional tenfold margin of safety for the pesticide chemical residue and other sources of exposure shall be applied for infants and children to take into account potential pre- and postnatal toxicity and completeness of data with respect to exposure and toxicity to infants and children." A different margin of safety may be used only if, on the basis of reliable data, such a margin will be safe. It should noted that any different margin of safety could be greater or lesser than the default 10 x . Since the passage of FQPA, EPA has developed a series of policy guidance documents, representing the evolution of its approach to implementing the "FQPA Safety Factor" provision of the law. The most current thinking on this topic can be found in the draft document entitled The Office of Pesticide Programs' Policy on Determination of the Appropriate FQPA Safety Factor(s)for Use in the Tolerance-Setting Process (V.S. EPA, 1999b). This document describes when FQPA safety factor decisions are needed; what the FQPA lOx safety factor is "in addition to"; how to judge the completeness of the toxicology and exposure databases; when a database uncertainty factor greater than I x is applied; how to determine, and account for, the degree of concern for pre- and postnatal toxicity; and the process for determination of the appropriate FQPA safety factor(s). Earlier articulations of principles are presented in several Agency documents (V.S. EPA, 1996, 1998a, b). 31.3.2.6 Consumer Right-to-Know FFDCA §408(0) states that EPA shall publish, and provide to large grocery stores, a publication which describes the risks and benefits of pesticide residues on food purchased in those

686

CHAPTER 31


stores by consumers, a listing of those pesticides for which the limited benefits-based tolerances have been issued, and recommendations for consumers on how they can reduce dietary exposures to pesticides in a manner consistent with maintaining a healthy diet. EPA developed and distributed a brochure, Pesticides and Food, to 30,000 grocery stores during the winter of 1999. Copies also went to public health officials, libraries, and the medical community. Over 4 million copies are in circulation. The brochure also can be found on EPA's Office of Pesticide Programs' Web site (www.epa.gov/pesticides/food). 31.3.2.7 Estrogenic Substances Screening Program FFDCA §408(p) states the EPA shall "develop a screening program, using appropriate validated test systems and other scientifically relevant information, to determine whether certain substances may have an effect in humans that is similar to an effect produced by a naturally occurring estrogen, or such other endocrine effect as the Administrator may designate." The Agency was given two years to develop the screening program, another year to implement it, and four years to report on its findings. The Agency established an advisory committee to assist it with the development of the program. In August 1998, the committee released its recommendations, most of which were adopted by the Agency. Implementation of the program began soon thereafter, with the focus being on the standardization and validation of the proposed components of the screening and testing batteries that make up the screening program. Details on the program and the Report to Congress (August 2000) can be found on the Web site of EPA's Office of Science Coordination and Policy (www.epa.gov/scipoly/oscpendo/index.htm).

remaining post-1984 chemicals; biopesticides; and the rest of the food-use inert ingredients. EPA published a notice in the Federal Register on August 4, 1997, outlining its plans for carrying out the tolerance reassessment process and identifying the individual substances in each of the three priority groups (U.S. EPA,1997a).

31.4 CURRENT REGULATORY PROCESS The registration of pesticides in the United States is bound by a structure defined by congressional legislation, as interpreted in formal regulations and other less-formal articulations of policy and practice. Proposals for changes and final changes to the regulations are published in the Federal Register. All final, formal regulations also can be found in the Code of Federal Regulations (40 CFR Parts 150-189). The CFR is updated annually to reflect any changes in the regulations that may have been finalized during the year. Daily issues of the Federal Register and the current CFR can be found on the Web site of the U.S. Government Printing Office (http://www.gpo.gov/). Frequently, statements of policy related to pesticide regulation are published by EPA's Office of Pesticide Programs (OPP) as Pesticide Registration (PR) Notices. These PR Notices as well as other documents articulating OPP's regulatory and risk assessment policies and practices can be found on OPP's Web site (http://www.epa.gov/pesticides/). Because regulatory approaches and practices are continually evolving as the state-ofthe-science and its interpretation mature, prospective pesticide registrants are strongly encouraged to meet with pesticide officials before proceeding with data generation and submission of petitions for registration or before executing changes in the registration status of their products.

31.3.2.8 Tolerance Reassessment FQPA required EPA to reevaluate all tolerances and exemptions in effect on the day before enactment of the act. A schedule was imposed that was to assure that 33% of such tolerances and exemptions were reviewed within 3 years of enactment; a second 33% within 6 years; and the remaining number within 10 years. Priority was to be given to those tolerances or exemptions that appeared to pose the greatest risk to public health (i.e., review the "worst first"). EPA divided all chemicals into three groups, with Group 1 containing the pesticides that appeared to pose the greatest risks. Group 1 is made up of several subgroups: organophosphorus compounds (OPs); carbamates; pesticides which had previously been characterized as probable human carcinogens (Groups Bland B2) according to EPA's classification scheme published in its cancer risk assessment guidelines in 1986; high-hazard food-use inert ingredients; and any chemicals that exceed their reference dose (RID) by unacceptable levels. Priority Group 2 contains those pesticides characterized as possible human carcinogens (Group C) and all reregistration chemicals which remained unfinished in 1996 (i.e., of those registered before 1984). Priority Group 3 contains all remaining pre-FQPA chemicals for which reregistration eligibility decisions already had been made by August 1996; all

31.4.1 REGISTRATION Registering a pesticide product for use in the United States under the regulations of the Federal Insecticide Fungicide and Rodenticide Act is equivalent to acquiring a federal license to sell or distribute a product in commerce. To do so without EPA approval is a federal crime. In theory, all pesticide products destined for use in the United States must be registered. A pesticide product is generally made up of more than one constituent. It may include one or more "active" ingredient(s) and one or more "inert" ingredient(s). A pesticide active ingredient is defined as "(1) any substance or mixture of substances intended for preventing, destroying, repelling, or mitigating any pest; (2) any substance or mixture of substances intended for use as a plant regulator, defoliant, or dessicant, and (3) any nitrogen stabilizer" [FIFRA §2(u)]. An "inert ingredient" is any substance or group of similar substances, other than the active ingredient, which is intentionally included in a pesticide product. Both the active ingredient(s) and the formulation(s) which constitute the product(s) are subject to registration requirements. If the intended use of the product includes application to agricultural

31.4 Current Regulatory Process

commodities destined for human or animal consumption, a tolerance or exemption from a tolerance also must be granted under FFDCA §408. New animal drugs, animal feeds containing a new animal drug, and liquid chemical sterilants for use on critical or semicritical medical devices are excluded from the definition of "pesticide." Although FIFRA §3 requires the registration of all materials which meet the definition of "pesticide" (i.e., either an active or inert ingredient) this section along with several other sections of the law provides for exemptions. For instance, exemptions may be granted if the material is being transferred from one site of an establishment to another when both are operated by the same producer, if it is regulated by another federal agency (e.g., certain biological control agents and human drugs), if it is "of a character which is unnecessary to be subject to this Act" [FIFRA §25(b)], or if it is in the preregistration status of having been granted an experimental-use permit under FIFRA §5 or an emergency use under FIFRA § 18. Approval of a registration is dependent upon the successful fulfillment of a series of data requirements, among other factors. The number and types of studies to be conducted vary with the intrinsic chemistry, anticipated inherent toxicity, and proposed use pattern of the pesticide. Pesticides of conventional chemistry proposed for use on agricultural commodities generally require the greatest amount of information, whereas nonfooduse conventional chemicals, antimicrobials, and biopesticides such as microbials and biochemicals generally require less. Part 158 of 40 CFR presents the regulatory roadmap specifying the types and amounts of data and other information needed by EPA to decide whether to approve an application for a new or amended registration or reregistration under FIFRA §3, for an experimental-use permit under FIFRA §5, or for a emergency exemption under FIFRA § 18. The data requirements specified in this part cover the areas of product chemistry, toxicology for human health and terrestrial mammals, wildlife and aquatic toxicology, nontarget insects (e.g., honey bees), environmental fate, aerial drift evaluation, reentry protection (primarily in the occupational setting), plant protection, product performance, residue chemistry (for food uses), and biochemical and microbial pesticides. Some of these kinds of data are always required for the evaluation of some or all types of products. Other kinds of data are required only under certain conditions; that is, if the product's proposed pattern of use, the results of earlier studies, or other circumstances warrant the development of such data. For example, the acute delayed neurotoxicity study in the hen is required only if the pesticide is an organophosphate or a metabolite thereof and causes inhibition of acetylcholinesterase or is structurally related to a substance that is known to cause delayed neurotoxicity. Another example would be the requirement to develop residue chemistry data only if the pesticide is proposed for use on food crops. Some, but not all, subparts of Part 158 have been modified since their original promulgation in 1984. Other data requirements have been added without the benefit of formal promulgation of regulations. This has occurred because the state-of-the-science has evolved and matured in many areas in

687

the intervening years. The Office of Pesticide Programs has modified many of its data requirements, communicating these changes via Pesticide Regulation Notices and other written materials. Thus, although the current Part 158 establishes data requirements that are applicable to various general use patterns, some unique aspect of a proposed use and/or the possibility of modification of the original data requirements for any use, general or unique, argues strongly for consultation between the prospective registrant and the Agency before beginning any data generation or information development. 31.4.2 REREGISTRATION AND REGISTRATION RENEWAL

Four times, over approximately 25 years (1972, 1978, 1988, 1996), Congress acted to require EPA to update the registration status of existing pesticide products. The most proscriptive directive was introduced in the amendments to FIFRA in 1988. The reregistration scheme articulated at that time remains in place, now made more complex by the requirement to reassess all tolerances in place at the time FIFRA and FFDCA were amended in the Food Quality Protection Act in 1996. The reregistration directives in the 1972 and 1978 amendments to FIFRA covered those pesticides registered up to that point. The accelerated reregistration program under FIFRA 1988 (FIFRA §4) encompassed all pesticide active ingredients initially registered before November 1, 1984. At that time, there were approximately 1150 active ingredients and over 20,000 product formulations registered in the United States. Because many of these 1150 active ingredients were related to one other (e.g., different salts of the same substance, such as sodium and calcium hypochlorite), they were organized into about 600 "cases" or groups of related pesticide active ingredients. These 600 cases were divided into four lists: List A-List A, which contains most of the food use pesticides, consists of the 194 chemical cases (or 350 individual active ingredients) for which EPA had issued registration standards prior to FIFRA 1988. Each registration standard document summarized the data available for a pesticide, called in any additional studies needed for reregistration, and required necessary product labeling changes. Lists B, C, and D-The remaining pesticides requiring reregistration, and for which no registration standard had been developed in previous reregistration attempts, were divided into three lists based on their potential for human exposure and other factors, with List B containing pesticides of greater concern and List D pesticides of less concern. Some of the classification criteria included the potential or known occurrence of residues in food or drinking water, significance of outstanding data requirements, potential for worker exposure, Special Review or restricted-use status, and unintended adverse effects on animals and plants.

FIFRA 1988 established a reregistration process consisting of five phases, with time frames and responsibilities for both

688

CHAPTER 31


EPA and the pesticide producers or registrants. The pesticides on Lists B, C, and D went through all five phases. Because EPA had already substantially reviewed them under the Registration Standards program, the List A pesticides moved directly to Phase 5.

Phase 1, list active ingredients-As required, EPA published Lists A, B, C, and D within 10 months ofFIFRA 1988 (by October 24, 1989) and asked registrants of these pesticides whether they intended to seek reregistration. Phase 2, declare intent and identify studies-Phase 2 required registrants to notify EPA whether or not they intended to reregister their products; to identify and commit to providing necessary new studies; and to pay the first installment of the reregistration fee. During this phase, EPA issued guidance to registrants for preparing their Phase 2 and Phase 3 responses. Phase 2 activities were completed in 1990. Nearly 250 cases, which included nearly half of the existing products, did not proceed past this phase, as the registrants chose not to support their continued registrations. Phase 3, summarize studies-During Phase 3, following EPA guidance, registrants were required to submit summaries and reformat acceptable studies, "flag" studies indicating adverse effects, recommit to satisfying all applicable data requirements, and pay the final installment of the reregistration fee. Phase 3 ended in October 1990. Phase 4, EPA review and data call-in-During Phase 4, EPA reviewed all Phase 2 and Phase 3 submissions and required registrants to meet any unfulfilled data requirements within four years. Phase 4 was completed in 1993. Phase 5, reregistration decisions-In this final phase, which remains ongoing, EPA reviews all the studies that have been submitted and decides whether or not the active ingredient(s) and the pesticide products containing the active ingredient(s) are eligible for reregistration-whether the data base is substantially complete, and whether or not the pesticide causes unreasonable adverse effects to humans or the environment when used according to product labeling. EPA also considers whether the pesticide meets the new safety standard of the FQPA and conducts tolerance reassessment for those pesticides which have food uses. The results of the Agency's review are presented in a Reregistration Eligibility Decision (RED) document. Products containing the pesticide active ingredient are reregistered after certain product-specific data and revised labeling are submitted and approved. All the active ingredients in a pesticide product must be eligible before the product is considered to be reregistered. Because of the FQPA requirement that pesticides sharing a common mechanism of toxicity with other substances (with or without pesticidal uses of their own) shall be evaluated for inclusion in a cumulative risk assessment with these other substances, the Agency has been issuing Interim REDs (IREDs) for some pesticides, reflecting its judgment concerning reregistration eligibility based solely on the individual pesticide's aggregate risk assessment, but reserving judgment on full reregistration eligibility until the cumulative risk assessment process is completed.

The timetable for reregistration that had been established in response to the congressional directives in FIFRA 1988 required modification following the passage of FQPA. Among the new dimensions added in FQPA was the requirement to reassess, within a lO-year time frame, all previously granted tolerances and exemptions from the requirement for a tolerance against the new safety standard that FQPA had established. This directive applied to tolerances (and exemptions) for all food-use pesticides, without regard to the date of their original registration. The consequences of this directive were, among others, that those food-use pesticides for which REDs already had been completed under the FIFRA 1988 reregistration process needed to be revisited for tolerance reassessment. Some of these also would need to be considered for inclusion in a cumulative risk assessment. Because EPA is using the reregistration program to accomplish tolerance reassessment, the timetable for completion of reregistration has been extended to encompass the 10-year time frame for tolerance reassessment, 1996-2006. At the end of 2000, reregistration was about 75% complete. Persons interested in obtaining details on the status of individual cases are referred to EPA's Office of Pesticide Programs Web site, where status reports and other materials on reregistration can be found (http://www.epa.gov/pesticides/reregistration. htm). FQPA also required EPA to establish a new registration review program ("registration renewal"). This new program obligates EPA to review every registered pesticide on a IS-year cycle. This new program would include all pesticides registered since November 1, 1984, as well as those that had been through earlier reregistration processes. Implementation of such a program would assure that pesticides are being reviewed periodically and updated to meet current scientific and regulatory standards. 31.4.3 SPECIAL REVIEW EPA not only has the authority to register pesticides, but also to cancel, suspend, or modify the registration of any pesticide or use of such pesticide that the Agency has determined to have the potential to "cause unreasonable adverse effects on the environment" [FIFRA §6(b )]. FIFRA §(2)(bb) states that, in making a final judgment whether to cancel or modify the conditions of registration of a pesticide or any of its uses, the Agency must weigh the potential for adverse effects against the costs and benefits ("economic, social, and environmental") derived from the use(s) of the pesticide. Dietary risks must be judged against the "reasonable certainty of no harm" safety standard under Section 408 of FFDCA. Risks related to use of public health pesticides are weighed against the health risks such as those from the diseases transmitted by the vector to be controlled by the pesticide. The formal procedure for conducting the necessary regulatory assessment under FIFRA §6 is commonly known as Special Review. Regulations governing the Special Review process are articulated in 40 CFR Part 154. The formal Special Review

References

process which includes a FIFRA §6 cancellation or suspension hearing is resource-intensive and time-consuming. In practice, the Agency has more often used less formal procedures to achieve the same goal of reducing the potential risks to acceptable limits. In more recent times, the reregistration process has been the principal mechanism for negotiating changes in the registration status of pesticides. It is more efficient and, whereas the formal Special Review process generally is a dialogue only between the Agency and the registrant(s) (and, sometimes, the hearing judge), with user groups free to submit benefits and other economic information to USDA, the less formal procedure encourages and supports a more active role for user groups and other interested stakeholders, such as public interest groups and the public health community. Persons interested in obtaining details on the status of chemicals in Special Review are referred to EPA's Office of Pesticide Programs Web site, where its report is available (http://www.epa.gov/docs/SpeciaIReview/srOOstatus.pdf).This report is updated annually. Risk reduction measures taken as a result of assessments conducted during the reregistration process are detailed in individual REDs and IREDs and summarized periodically in status reports. These can be found on EPA's Office of Pesticide Programs' Web site, where status reports and other materials on reregistration can be found http://www.epa.gov/pesticides/reregistration.htm). Portions of this Web page are updated annually at a minimum.

31.5 WEB SITES Government Printing Office Code of Federal Regulations and Federal Register (http://www.gpo.gov/) U.S. Environmental Protection Agency Estrogenic Substances Screening Program (http://www.epa.gov/scipoly/oscpendo/index.htm) Pesticides and Food brochure (http://www.epa. gov/pesticides/food) Reregistration (http://www.epa.gov/pesticides/reregistration.htm) Special Review (http://www.epa. gov/docs/SpecialReview/srOOstatus. pdf) Statements of policy (http://www.epa.gov/pesticidesl)

REFERENCES Code of Federal Regulations (1990). 40 CFR Parts 150-189. Protection of the environment. Subchapter E-Pesticide Programs. FACTA (1990). Food, Agriculture, Conservation, and Trade Act of 1990, Public Law No. 101-624, secs. 1491-1496, 104 Stat. 3359. FEPCA (1972). Federal Environmental Pesticide Control Act of 1972, Public Law 92-516,86 Stat. 973. FFDCA (1954). Federal Food, Drug, and Cosmetic Act of 1954 Miller Amendment to FFDCA §408, Public Law No. 518, 68 Stat. 511.

689

FFDCA (1958). Federal Food, Drug, and Cosmetic Act of 1958 Food Additives Amendment to FFDCA §409, Public Law 85-929, 72 Stat. 1785. FIA (1910). Federal Insecticide Act of 1910, Chap. 191,36 Stat. 331. FIFRA (1947). Federal Insecticide, Fungicide, and Rodenticide Act of 1947, Public Law No. 80-104,61 Stat. 163. FIFRA (1964). Federal Insecticide, Fungicide, and Rodenticide Act of 1964, Public Law 88-305,78 Stat. 190. FIFRA (1975). Federal Insecticide, Fungicide, and Rodenticide Act of 1975, Public Law No. 94-140, 89 Stat. 751. FIFRA (1980). Federal Insecticide, Fungicide, and Rodenticide Act of 1980, Public Law No. 96-539, 94 Stat. 3194. FIFRA (1988). Federal Insecticide, Fungicide, and Rodenticide Act Amendments of 1988, Public Law No. 100-532, 102 Stat. 2654. FPA (1978). Federal Pesticide Act of 1978, Public Law No. 95-396, 92 Stat. 819. FPA (1996). Food Quality Protection Act of 1996, amending the Federal Insecticide, Fungicide, and Rodenticide Act and the Federal Food, Drug, and Cosmetic Act, Public Law No. 104-170, 11 Stat. 1513. Nixon, R. M., President (1970). Reorganization Plan No. 3 of 1970. 40 CFR pt. I and Fed. Reg. 35, 15623. U.S. Environmental Protection Agency (1996). "Is an Additional Uncertainty Factor Necessary and Appropriate to Assess Pre- and Postnatal Developmental and Reproductive Effects in Infants and Children Exposed to Pesticide through Chronic Dietary Exposure?" Presented to the FIFRA Scientific Advisory Panel, October 1996. U.S. Environmental Protection Agency (1997a). Raw and processed food schedule for pesticide tolerance reassessment. Fed. Reg. 62(149), 4201942030, August 4, 1997. U.S. Environmental Protection Agency (1997b). "Guidelines for Expedited Review of Conventional Pesticides under the Reduced-Risk Initiative and for Biological Pesticides." Pesticide Registration Notice 97-3, September 4, 1997. U.S. Environmental Protection Agency (1998a). "Use of IOx Safety Factor to Address Special Sensitivity of Infants and Children to Pesticides." Presented to the FIFRA Scientific Advisory Panel, March 1998. U.S. Environmental Protection Agency (1998b). "Standard Operating Procedures of the Health Effects Division's FQPA Safety Factor Committee." Presented to the FIFRA Scientific Advisory Panel, December 1998. U.S. Environmental Protection Agency (I 999a). "Guidance for Identifying Pesticide Chemicals and Other Substances Which Have a Common Mechanism of Toxicity." February 1999. U.S. Environmental Protection Agency (l999b). "The Office of Pesticide Programs' Policy on Detennination of the Appropriate FQPA Safety Factor(s) for Use in the Tolerance-Setting process." May 1999. U.S. Environmental Protection Agency (1999c). "Proposed Guidance on Cumulative Risk Assessment of Pesticide Chemicals That Have a Common Mechanism of Toxicity: Issues Pertaining to Hazard and Dose Response Assessment." Presented to the FIFRA Scientific Advisory Panel, September 1999. U.S. Environmental Protection Agency (l999d). "Chapter 4: Exposure Assessment and Characterization and Chapter 6: Estimation and Characterization of Cumulative Risk. Proposed Guidance on Cumulative Risk Assessment of Pesticide Chemicals That Have a Common Mechanism of Toxicity." Presented to the FIFRA Scientific Advisory Panel, December 1999. U.S. Environmental Protection Agency (2000a). "End Point Selection and Detennination of Relative Potency in Cumulative Hazard Assessment: A Pilot Study of Organophosphorus Pesticide Chemicals." Presented to the FIFRA Scientific Advisory Panel, September 2000. U.S. Environmental Protection Agency (2000b). "Cumulative Risk: A Case Study of the Estimation of Risk from 24 Organophosphate Pesticides." Presented to the FIFRA Scientific Advisory Panel, December 2000.

CHAPTER

32 Risk Assessment for Acute Exposure to Pesticides Roger C. Cochran* Department of Pesticide Regulation, California Environmental Protection Agency

32.1 INTRODUCTION

32.2 TOXICOLOGICAL DATA

Regulatory agencies have tended to focus their assessments of pesticide risk on the potential for toxicological effects to arise from repetitive, long-term usage of the chemicals (Barnes and Dourson, 1988; WHO, 1978). This emphasis on developing reference doses (RIDs) for the potential effects of chronic exposure to pesticides may have gained impetus from public concerns about cancer or the impact of pesticides on the environment (Carson, 1962; NRC, 1987). Consequently, much of the toxicological database required under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) examines the effects of repetitive, subchronic or chronic dosing by the oral route. From a public health perspective, however, the recorded human illnesses attributed to acute exposures to pesticides may be of greater significance than those connected with potential chronic exposures (Mehler et aI., 1992). Risk assessment, as defined by the National Research Council (1983), consists of five components: (1) Hazard identification encompasses examination of the toxic effects of the chemical; (2) dose-response assessment evaluates the dose level of the chemical necessary to cause manifestation of toxic effects; (3) exposure assessment estimates the amount of the chemical that people are likely to absorb; (4) risk characterization predicts the likelihood that people, exposed to the chemical to the degree estimated, will become ill; and (5) risk appraisal examines the strengths and weaknesses of the estimates of the various toxicological and exposure parameters and expresses the degree of confidence in the projected risks. Acute exposure, here, refers to human encounters with pesticides in the course of one day or less. A pesticide, as defined by the U.S. Environmental Protection Agency (EPA), is any chemical, or mixture of chemicals, intended to be used in preventing, destroying, repelling, or mitigating any pest (Federal Register, 1998).

The first two of the five components needed for risk assessment require an extensive knowledge of the toxicological effects of a chemical. Under FIFRA, the toxicological database for a pesticide is defined by the guideline requirements (EPA, 1984). This database includes acute lethality studies (oral, dermal, and inhalation), subchronic toxicity studies (90-day oral, inhalation, and dermal toxicity; 21128-day dermal toxicity; developmental toxicity; reproductive toxicity), chronic toxicity studies (I-year nonrodent toxicity; oncogenicity; and combined chronic toxicity/oncogenicity), and neurotoxicity studies (neurotoxicity screening battery; 90-day neurotoxicity; developmental neurotoxicity) (Federal Register, 1998). Only a few of these study types contain data that can be used to explore the toxicological effects from a single day's (acute) exposure to a pesticide. Acute lethality studies, for example, use a range of single doses to elicit toxic effects. However, these studies are designed to set toxicity categories for labeling information (EPA, 1998a). Virtually all of the older acute lethality studies, regardless of the route of exposure (oral, dermal, or inhalation), generally do not have data on nonlethal, systemic effects that occurred at less than lethal dosages. The single-dose, neurotoxicity screening battery is currently being required only for those pesticides designed to be neurotoxins (e.g., organophosphates, carbamates, and pyrethroids) (EPA, 1998b). Consequently, data from this test, which includes components of histopathology, tissue and blood chemistry, as well as clinical signs and performance testing, are not available for most pesticides. Thus, data on acute effects for most pesticides have to be teased out of repetitive dosing studies. Subchronic, reproductive, and chronic toxicity studies may have data concerning clinical signs that appear within 1-2 days at the beginning of the studies. All other data on potential systemic toxicity in these study types are obtained at the end of the study period and cannot be attributed to acute toxicity. Developmental toxicity studies provide an exception. Because develop-

*The opinions expressed in this chapter represent the views of the author and do not necessarily reflect the views and policies of the Department of Pesticide Regulation. The mention of trade names or commercial products does not constitute endorsement or recommendation for use. Handbook of Pesticide Toxicology Volume 1. Principles

691


692

CHAPTER 32

Risk Assessment for Acute Exposure to Pesticides

mental toxicity may be manifested as the result of a single dose (EPA, 1991; Ogata et aI., 1984; Schardein, 1985), it is assumed, in the absence of data to the contrary, that the observed developmental effects are elicited from a single dose. This assumption mayor may not be valid. Nonetheless, developmental toxicity studies (see Chapter 16) tend to be a major source of critical no observed effect levels (NOELs) for conducting risk assessments on potential acute exposures to nonneurotoxic pesticides. Although a developmental endpoint for exposure to toxins is only relevant in women of child-bearing age, the assumption that all other population subgroups are as sensitive results in margins of safety (MOSs) that protect the health of these other subgroups for other endpoints that may occur at higher dosages. The MOS is defined as the ratio of the critical NOEL to the estimated exposure. Published research studies may also provide sufficient data for dose-response assessment. These studies, however, tend to be designed to clarify the mechanism of action of a specific type of pesticide toxicity. Nonetheless, the peer-reviewed reports sometimes describe a range of concentrations used to elicit an effect from a single dose. Such a study may provide the basis for a regulatory NOEL, particularly in the case of experiments with human subjects. The main drawbacks to these published studies are (l) the lack of individual animal data because they are typically not reported or archived and, thus, (2) the need to rely on the author's interpretation of the results. Under FIFRA, pharmacokinetic data are sought to obtain information on how a pesticide is absorbed, distributed, biotransformed, and excreted, as well as to aid in understanding the mechanism of toxicity (EPA, 1998c). Information may also be obtained about potential tissue-specific accumulation and induction of biotransformation. Most of the pharmacokinetic data are derived from studies using the oral route of exposure. Dermal pharmacokinetic studies tend to consider solely dermal penetration and/or absorption. Pharmacokinetic studies on the inhalation of pesticides are comparatively rare, seemingly limited to fumigants. Pharmacokinetic data can have a profound effect on the dose-response assessment for a pesticide. The estimated absorbed dose of a pesticide necessary to cause toxic effects may be modified downward if there is evidence of less than 100% absorption through the route used in the dose-response assesment. Information regarding bioavailability via the oral route is useful as many pesticides and their metabolites are excreted in variable amounts in the feces. Estimations of absorbed dosages from inhalation toxicity studies rely on default assumptions concerning breathing rates, tidal volumes, and chemical retention and absorption to estimate absorbed dosages (Raabe, 1986, 1988; Zielhuis and van der Kreek, 1979). Such estimates, when derived from whole-body inhalation studies, can be confounded by the fact that rats exposed to dusts or chemical vapors via whole body absorb 5-8 times more material than rats exposed via nose only (Blair et aI., 1974; Hext, 1991; Iwasaki et aI., 1987; Jaskot and Costa, 1994; Landry et aI., 1986; Langard and Nordhagen, 1980; Ty1 et al., 1995; Wolff et al., 1982). The additional absorption noted in whole-body inhalation exposure

studies appears to be due to an unquantifiable oral component, possibly from grooming behavior (Cochran et aI., 1997). Even nose-only inhalation toxicity studies may have a significant oral component due to grooming activity (Hext, 1991).

32.3 EXPOSURE DATA The second, and equally important, half of the risk assessment equation is the estimate of human exposure. The chief source of exposure to pesticides through the oral route is from the diet (see Chapter 19; Cochran et aI., 1994). There can also be an oral contribution from hand-to-mouth activity in adults and children or pica in children (Binder et aI., 1986; Calabrese and Stanek, 1992; Calabrese et aI., 1989, 1991; Carlisle, 1992; Clausing et aI., 1987; EPA, 1996). Pica in children, however, appears to be highly unusual, as only a single instance of intentional imbibing of dirt was reported out of more than 200 children whose soil ingestion from hand-to-mouth activity was documented in the preceding publications. Currently, dietary exposures are estimated by most governmental agencies through a process that combines data on dietary consumption with data on pesticide residues measured on food (Cochran et aI., 1995a; FAOIWHO, 1988, 1997). Dietary consumption data are generally derived from government surveys (Cochran et aI., 1995a; FAOIWHO, 1997; Trichopoulou, 1994; USDA, 1989-1991). Data for potential pesticide residues associated with EPA or European Union (EU) label-approved direct food uses, as well as information about possible secondary residues in animal tissues, are also necessary for estimating human dietary exposures. These data are derived from governmental monitoring programs (CD PR, 1997; FAOIWHO, 1999; USDA, 1996). However, dietary exposure to pesticides is only a fraction of the total human exposure experience. Much of human occupational (persons engaged in the process of pesticide application) or nonoccupational (other than dietary) exposure to pesticides results from the handling of pesticides or other activity patterns that place people in contact with the pesticides. In general, most of the occupational and nonoccupational exposure to pesticides is through the dermal and/or inhalation routes (Ross et aI., 1992; Wolfe, 1976). Exposure estimates for these scenarios are based on environmental monitoring, passive dosimetry, or biological monitoring of individuals involved in the active handling of pesticides or engaged in activities in areas treated with those pesticides (Bonasall, 1985; Lavy and Mattice, 1986). An extensive, detailed discussion of the techniques used for estimating occupational and nonoccupational exposures may be found in Chapter 21. Environmental monitoring involves measurements of pesticide concentrations in the ambient air and on surfaces. The translation of measured air concentrations into an estimated absorbed dose for humans requires assumptions on respiratory frequency, volume, and absorption of the pesticide (EPA, 1996; Raabe, 1986, 1988; Zielhuis and van der Kreek, 1979). Estimation of human dermal exposure from surface concentrations of

32.4 Examples

pesticides in the environment relies on the precision of various generic transfer factors (EPA, 1996; Pandian et aI., 1999). Passive dosimetry gauges air concentrations in the breathing zone and measures dermal concentrations of pesticides through the use of hand washes, dermal patches, and/or articles of clothing (Wolfe, 1976). The same assumptions for inhalation are used with air concentrations of pesticides measured in the breathing zone as were used for those detected in the ambient air. Concentrations of pesticides extracted from monitoring patches attached to the skin are assumed to be representative of chemical concentrations over a specified body surface area (Wolfe, 1976). A single value, based on submitted, chemical-specific studies (a default of 100% has been used if specific data were not available), serves as the basis for estimating the absorbed dose (EPA, 1992a). It is known that the percentage of pesticide absorbed through the skin varies inversely with the concentration of the chemical (Wester and Maibach, 1976). However, at the present time, there are no scientific models available that examine the effect of multiple concentrations of pesticides on the skin, separated spatially and/or chronologically, on the absorbed daily dosage (Wester and Maibach, 1993). Biological monitoring provides an estimate of the aggregate exposure to a pesticide from all routes. Unfortunately, very few biomonitoring studies have been conducted for more than a handful of pesticides. Chemical-specific, human stoichiometric data are essential to the process of estimating absorbed dosages from excreted pesticide metabolites. Consequently, the principal limiting factor seems to be the lack of human pharmacokinetic data on most pesticides. Chemical-specific information is preferred for exposure data from either environmental monitoring or passive dosimetry. Surrogate exposure data (from pesticides with similar chemical and physical properties, as well as similar preparation and application practices) and generic databases, such as the Pesticide Handlers Exposure Database (PHED, 1995), are used as substitutes. The use of surrogate exposure data increases the level of uncertainty in exposure estimates. Differences in volatility between the chemical under consideration and surrogate chemicals may affect air concentrations in an unquantifiable manner. Likewise, differences in chemical properties could affect transfer factors, clothing penetration, and dermal adsorption. Differences in application rates cause assumptions to be made on the relationship between the amount of chemical handled and the amount of exposure through all routes. The principal difficulty associated with the use of PHED to estimate exposure data is that the data subsets, which are combined by the program to form work categories, are not homogeneous (van Hemmen, 1992). For example, one source of variability is that each of those studies has a different minimum detection level for the analytical method. It should be noted that the detection of dermal exposure to the body regions is not standardized. Some studies observe exposure to only selected body regions, such as the hands, arms, and face, with other body regions considered 100% protected from exposure by work clothing. Other studies have more extensive dermal measurements.

693

Consequently, the subsets derived from the database for dermal exposure have different numbers of observations for each of the body regions. Finally, the PHED database is predicated on the relationship between the amount of pesticide handled and the degree of occupational exposure. Yet, for example, within the data set used to estimate exposures for groundboom applications without the presence of a cab, there is no correlation between the amount of pesticides being used and the amount of dermal or inhalation exposures that workers receive. The net effect of this lack of correlation between exposure and the amount of chemical used is an inability to predict, with accuracy, what exposures any worker will receive in a given work category. When PHED is used for estimating potential acute (single day) occupational exposures, the only data point that can be provided is the average exposure value. Because the variability in each of the data subsets in a given category is unrelated to that in any of the other data subsets, it is not possible to estimate the overall variability in exposure. Yet, in a given study, there is variability in worker exposure. Even though individuals confine their activities to label-approved personal protective equipment and labor practices, they do not receive the same exposure. Depending on the shape of the distribution curve, the average exposure value may represent the maximum potential exposure of as few as 50% of the workers. The amount of exposure for the other workers, who also follow label requirements, could be greater-though the magnitude of the exposure cannot be calculated.

32.4 EXAMPLES Thus far, we have examined the nature of the toxicological and exposure databases used in generating a risk assessment. How the data fit together can best be explored through critical examination of some examples of completed risk assessments for acute exposure to pesticides. The examples provided are risk assessments conducted for the California Department of Pesticide Regulation (CDPR). 32.4.1 ETHOPROP The first case study is the risk characterization document (RCD) for ethoprop (Cochran et aI., 1995b). Ethoprop (O-ethyl-S,Sdipropyl phosphorodithioate) is an organophosphate pesticide used as an insecticide, nematicide, and fungicide (suppression of white mold on peanuts) on food and nonfood crops. The oral LDso for ethoprop was 61 mglkg in male rats and 33 mglkg in female rats. Examination of the toxicological database indicated that rabbits were the most sensitive laboratory species to ethoprop exposure, with a dermal LDso of 24 mglkg. Clinical signs of acute toxicity were characteristic of cholinesterase inhibition and included diarrhea, excessive urination, lacrimation, tremors, and convulsions. As the principal route of exposure for most pesticide applicators using ethoprop was through the skin, it would have been

694

CHAPTER 32


preferable to use the dose-response data of adverse effects observed in short-term dermal toxicity studies as the basis for calculating margins of safety for workers with short-term exposure to ethoprop. However, none of the submitted or published data established a single-dose dermal NOEL for clinical signs in rats or rabbits using technical-grade ethoprop. In rats, a single dermal dose of a formulation, Mocap® 6EC, resulted in cholinergic signs (salivation, irregular respiration, prostration, and morbidity), with a NOEL of 160 mg formulationikg. Dermal exposure of rabbits to Mocap 6EC resulted in clinical signs (ataxia, depression, dilation of pupils, excessive salivation, and loss of the righting reflex) with a NOEL of 8.7 mg active ingredientlkg. However, examination of the database suggested that ethoprop, diluted in formulations, had more toxicity through the dermal route than technical-grade ethoprop. This suggested that the inert ingredients in the formulation had facilitated the passage of ethoprop through the skin. As the inert ingredients in formulations are frequently changed, none of the short-term dermal toxicity studies were considered appropriate as the basis for assessing the risk of short-term occupational exposures to ethoprop. Instead of a dermal NOEL, the absorbed dose from an oral NOEL was used to estimate margins of safety from short-term exposure to ethoprop. Short-term oral NOELs were derived from developmental studies in rats and rabbits and from a single-dose neurotoxicity study in rats. Ethoprop did not produce developmental malformations in rats. Fetal toxicity in rat studies was manifested as decreased fetal weight. In rats, the lowest observed effect level (LOEL) for maternal toxicity [cholinergic signs-soft stools (8/25 animals) and anogenital staining (3/25 animals)] was 18 mg/kg-day with a NOEL of 9 mg/kg-day. The effects were manifested after 2 days of dosing. A lower NOEL for cholinergic signs (1.6 mg/kg-day) was reported in an earlier rat developmental study. However, the happenstance of dose selection appeared to determine this NOEL. Considering the two studies together, the NOEL (9 mg/kg-day) from the later study was not precluded as a possible NOEL for the earlier study as well. The single-dose LOEL for cholinergic signs, reduced motor activity, and reduced scores on the functional observational battery was 25 mg/kg with a NOEL of 5 mg/kg in both male and female rats. Again, 9 mg/kg was not precluded as the actual NOEL. Developmental toxicity was not observed in rabbits at any dose. Maternal toxicity in rabbits, characterized by signs of cholinesterase inhibition [soft stools (2/8 animals), anogenital staining (2/8 animals), and death (1/8 animals)], was observed by day 2 at 5.0 mg/kg-day, with a NOEL of 2.0 mg/kg-day. A lower NOEL in rabbits, 0.125 mg/kg for decrement in maternal weight gain (14%), was noted in an earlier study. However, the endpoint (decrement in weight gain) required 12 days of dosing to be manifested. Consequently, this NOEL could not be used to assess health risks associated with potential single-dose exposures to ethoprop. The oral NOEL (2 mg/kg-day) for maternal toxicity in rabbits (cholinergic signs and death) was used to assess the health risks from potential short-term exposures to ethoprop.

Exposure estimates for the various occupational categories were based on monitoring data from ethoprop exposures and calculations from monitoring data for surrogate active ingredients (diazinon, turbofos) with similar application rates and chemical properties. These estimates were based on 8-h workdays during the application season and assumed 100% dermal absorption, as no dermal absorption data were available. This health protective assumption probably overstated the exposure, as reported in vivo human dermal absorption for five other organophosphate pesticides ranged from 8 to 46% (Wester and Maibach, 1985, 1993). Uptake of ethoprop via the inhalation route was assumed to involve 50% retention by the lungs and 100% absorption (Raabe, 1986, 1988). The average daily dosage (ADD), actually the geometric means of exposure, ranged from 0.2 J.lg/kg-day for irrigators to 139 J.lg/kg-day for incorporators (workers incorporating the applied ethoprop into the soil) working with the EC formulation. The use of geometric mean values underestimates potential short-term exposures of populations of workers (EPA, 1992b). Consequently, the 95th percentile [geometric mean x (standard deviation)1.645] of short-term worker exposure was also examined. The 95th percentiles of short-term exposure for loader/applicator/incorporators working with the 5G and lOG formulations were 71 and 45 J.lg/kg-day, respectively. All but two of the tolerances for ethoprop are for "negligible residues," as the EPA does not expect that any residues of ethoprop will be found on raw agricultural commodities (EPA, 1988a). Nonetheless, to be health protective, CDPR has a policy of conducting dietary risk assessments if tolerances exist for a pesticide on edible commodities. The CDPR surveillance programs from 1987 to 1991 indicated that ethoprop levels in raw agricultural commodities (RACs) were nondetectable. The minimum detection limit (MDL) was 0.05 ppm. Crops monitored in this survey between 1989 and 1990 were cabbage and potatoes, where ethoprop was mostly used. Field studies indicated that ethoprop residues on registered crops were less than 0.02 ppm. Examination of the FDA program for fiscal year (FY) 1985FY 1990 revealed only two values. These were 0.680 ppm in strawberries (1987) and 0.140 ppm in apples (1989). The mean theoretical (acute) daily dietary exposure for all population subgroups ranged from 0.02 to 0.08 J.lg/kg-day, with children (1-6 years of age) having the highest theoretical exposure. Although theoretical acute dietary exposure was combined with acute occupational exposure in the RCD, the dietary contribution was negligible. Consequently, it is not considered here. The MOS for exposure to ethoprop was calculated as the ratio of an oral NOEL, established in laboratory animal studies, to the potential exposure dosage (greater than 95% through the dermal route) estimated for the human population. The MOSs for potential acute exposure, based on an oral NOEL of 2.0 mg/kg-day for cholinergic signs and death in rabbits, ranged from 14 for incorporators using the EC formulation to 10,000 for the irrigators (Table 32.1). If the 95th percentile of shortterm exposure were considered for workers using the 5G and lOG formulations, the MOSs would be 29 and 40, respectively.

32.4 Examples Table 32.1 Margins of Safety for Potential Acute (Daily) Exposures to Ethoprop Acute

Work task

Mosa

EC formulation MixerlIoader/applicator

32 14

Incorporator

10,000

Irrigator 50 formulation Loader/applicator/incorporator

400

100 formulation Loader/applicator/incorporator

425

aBased on an NOEL of 2.0 mg/kg-day for cholinergic signs and death in a rabbit study MOS = NOEL(2000 ~g/kg-day) ADD

In the absence of scientific evidence to the contrary, effects reported in laboratory studies are expected to occur in humans at similar dosages. When the NOEL is from a laboratory animal study, a MOS of 100 is generally considered adequate for protection against potential acute toxicity of a chemical. This uncertainty factor assumes that humans are 10 times more sensitive to the acute effects of a toxin than are laboratory animals and that the difference in susceptibility to the toxicity of a compound within the human population spans only an order of magnitude (Davidson et aI., 1986; Dourson and Stara, 1983, 1985; EPA, 1986). If the critical NOEL is derived from a human study, a different number, 10, is used, incorporating the single uncertainty factor for human variability. After the RCD for cthoprop was released, the manufacturer dropped production of the EC formulation and changed the method of handling the granular formulations in order to reduce the estimated exposure. Certain generic exposure reduction factors were dictated by the addition of personal protective equipment and procedures (EPA, 1997; Thongsinthusak et al., 1993). These changes in handling procedures of the granular formulations theoretically resulted in a substantial decrease in the estimated exposure to ethoprop and a concomitant increase in the estimated MOSs to more than 100 for the 95th percentile of worker exposure. 32.4.2 MEVINPHOS Mevinphos (2-carbomethoxy-l-methyl-vinyl dimethyl phosphate) is an organophosphate insecticide used to control aphids, mites, grasshoppers, cutworms, leafhoppers, caterpillars, and many other insects on a broad range of field, forage, vegetable, and fruit crops. This highly toxic pesticide (rat oral LDso "-'2 mg/kg) had a large number of human illness reports associated with its use (Cochran et al., 1996). Examination of the toxicological database for mevinphos indicated that the principal adverse effects (cholinergic signs) were associated with inhibition of acetylcholinesterase activity. Both plasma and red

695

blood cell cholinesterase activities were significantly reduced compared to controls in several acute studies. The EPA, in its Guidelines for Neurotoxicity Risk Assessment, lists alteration in the degradation of neurotransmitters as a possible adverse effect because it can lead to unwanted changes in the function of the nervous system (EPA, 1998b). However, the guidelines do not specify the level of inhibition of brain acetylcholinesterase activity that constitutes an adverse effect. Even statistically significant brain cholinesterase inhibition caused by organophosphorous insecticides may not lead to cholinergic signs in laboratory animals (Bushnell et aI., 1993, 1994; Chanda and Pope, 1996; Stanton et al., 1994). Organophosphorous insecticide poisoning in humans may lead not only to cholinergic signs but also symptoms, for example, headaches (Ellenhom et aI., 1997), which cannot be ascertained in laboratory animals. Consequently, statistically significant levels of inhibition of brain acetylcholinesterase activity in laboratory animals may be used as a surrogate for this manifestation of an impaired nervous system function, which is detectable only in humans (EPA, 1998b; JMPR, 1998). The depression of plasma or red blood cell cholinesterase activity is generally considered an indication of exposure to a neurotoxic substance, rather than an adverse effect in itself (Carlock et aI., 1999; EPA, 1988b, 1990, 1993, 1998b; JMPR, 1998). The biological significance of inhibition of blood cholinesterase has remained controversial (Carlock et aI., 1999; Chen et aI., 1999; EPA, 1988b, 1990, 1993; JMPR, 1998; Lotti, 1995). The cholinesterase in the blood does not appear to act as a "sink" to reduce paraoxon toxicity, as that same toxicity was not potentiated by prior immunological reduction of blood acetylcholinesterase activity (Padilla et aI., 1992). Consequently, these parameters were not used to characterize the potential risk to humans. As in the case of ethoprop, the principal route of exposure for most mixerlloaders associated with mevinphos use was through the skin. No single-dose dermal toxicity studies were available. However, the toxicological database for mevinphos indicated that the inhibition of brain cholinesterase activity following continuous dosing was likely the result of the last dosage, rather than a cumulative effect of the repeated dosages. Thus, the NOEL (1 mg/kg-day) for significant (p < 0.05) brain cholinesterase inhibition (>50%) in a 21-day dermal toxicity study in rabbits might be considered a single-dose NOEL. The single-dose oral NOEL for neurotoxicity (clinical signs, sensorimotor alterations, reduced neuromuscular performance, and inhibition of brain cholinesterase activity) in the rat was 0.1 mg/kg. Cholinergic signs (loose stool and undescribed other effects) were noted in a study of 20 human subjects following oral ingestion of capsules with mevinphos up to an estimated dosage of 0.036 mg/kg (Rider et aI., 1975). Unfortunately, individual data from that study were unavailable, and a NOEL could not be established. In a subsequent study, a different group of investigators (Verberk, 1977; Verberk and Salle, 1977) built upon the earlier study. A NOEL for cholinergic signs in eight humans was established as 0.025 mg/kg or greater (only one dose level and no cholinergic effects) (Verberk, 1977; Ver-

696

CHAPTER 32


berk and Salle, 1977). Although the study ran for 28 days, no cholinergic signs were reported at any time. Thus, the NOEL for cholinergic signs and symptoms (0.025 mg/kg) in humans applies to both acute (single dose) and subchronic (28 days) time periods. The actual I-day NOEL for cholinergic signs may be higher, as a toxicological endpoint caused by a single dose of a chemical can generally be achieved by a lower repetitive dose (Klaassen and Eaton, 1991). For the short-term time frame, this human I-day oral NOEL was approximately the same as the oral NOELs for cholinergic signs from studies on dogs, rats, and rabbits. Because of the uncertainty over the rabbit dermal NOEL, the human oral NOEL (0.025 mg/kg) for cholinergic signs was used in the ReD to calculate margins of safety for potential short-term occupational and acute dietary exposures. The studies and data forming the basis for estimating worker exposure were based on both passive dosimetry data and calculations from foliar residue data. The mean exposure values used for the risk assessment are shown in Table 32.2. The exposure estimates for pilots of fixed-wing aircraft and helicopters, loaders for helicopters, and fiaggers were obtained from studies conducted in Monterey and Imperial counties (Maddy et aI., 1981, 1982). In those studies, the mean absorbed dosage was calculated from the amount of mevinphos collected by cotton gauze and cloth patches placed on the bodies of workers. Approximately 80% of the patches had no detectable levels of mevinphos. Accordingly, each of those patches was assigned a default value equivalent to 50% of the MDL. The measured

or theoretical levels of residue on patches from specific body regions were then multiplied by the total surface area of that zone to obtain an estimate of the dermal exposure. The estimates for each of the zones were then summed for total body exposure. Inhalation exposure was estimated from measured air concentrations of mevinphos in the breathing zone. The 95th percentile of short-term worker exposure ranged from 1.1 f.Lg/kg-day for helicopter pilots to 33.8 f.Lg/kg-day for helicopter mixerlloaders. These values are representative of the maximum acute occupational exposures that workers might be expected to encounter. Potential acute dietary exposure to mevinphos for alllabeled uses, based on the 95th percentile of user-day exposure for all population subgroups, ranged from 1.0 to 3.3 f.Lglkg-day. The potential dietary exposure of the population subgroup of males, aged 20 and over, was chosen as a surrogate sUbpopulation for the purposes of estimating combined occupational and potential dietary exposures. The choice was based on two factors: (1) occupational exposures were derived using passive dosimetry of agricultural workers from this population subgroup, and (2) the dietary exposure values are approximately the same as those of any other population subgroup that might contribute to the agricultural workforce. The potential acute dietary exposure of this population subgroup was 1.3 f.Lglkg-day. This value was added to the mean estimated occupational exposures. However, the theoretical combined acute exposures were probably overestimates of the actual exposures. It was unlikely that the agricultural workers engaged in activities associated with mevinphos

Table 32.2 Estimates of Acute Occupational Exposure to Mevinphos and the Respective Acute MOSs

Work task

ADDa

Acute

Combinedc

Combined

(~g/kg-day)

MOSb

ADD

acute MOS

Helicopters Mixerlloaders

2.4

10

3.7

7

Pilots

0.5

50

1.8

14

Fixed-wing aircraft Mixerlloaders

1.6

16

2.9

9

Pilots

0.5

50

1.8

14

Flaggers

0.04

625

1.34

19

Ground application Mixerlloader/applicators

3.8

7

5.1

5

3.2

8

4.5

5

(open cab) Mixerlloader/applicators (closed cab) Harvesters Field workers (vegetables) Field workers (fruits) Field workers (grapes)

neg.-1.5 0.8-11 0.4-1.1

17-50

1.3-2.8

2-31

2.1-12.3

2-12

25-63

1.7-2.4

11-15

9-14

aThe geometric mean of the estimated absorbed daily dosage (ADD) for helicopter, fixed-wing aircraft, and ground application. For harvesters, the values represent the range of exposures for workers harvesting different commodities. The dermal absorption was 16.8%. Inhalation retention and inhalation absorption were 50% and 100%, respectively, assuming a 75.9 kg body weight for workers. bMOS based on an NOEL of 0.025 mg/kg for cholinergic signs in human studies. cCombined occupational and dietary exposures.

32.4 Examples

use would also be in the 95th percentile of dietary exposure to RACs, each with the maximum measured level of mevinphos residues. The MOSs for mean acute occupational exposures, based on the NOEL of 25 ~g/kg for human cholinergic signs, ranged from 2 (apple harvesters) to 625 (flaggers in enclosed vehicles). If the 95th percentile of short-term exposures were considered for each of the job categories, the MOSs would range from less than 1 (mixerlloaders involved in helicopter applications) to 23 (helicopter pilots). Combining potential acute dietary exposure with mean occupational exposures caused a substantial drop in the MOS for all job categories. The combined MOSs ranged from 2 (apple harvesters) to 19 (flaggers in closed cabs). Arguably, it might have been preferable to use the dermal NOEL of 1 mg/kg for inhibition of brain cholinesterase activity from the 2l-day rabbit study. The specificity of the route of exposure can affect the time course of systemic absorption as well as the chemical nature of the toxin. If this NOEL (l mg/kg) were applied to the calculated mean dermal exposures, the MOSs would then be 69, 100,64, and 76 for mixerlloaders associated with helicopter, fixed-wing aircraft, open-cab ground applications, and closed-cab ground applications, respectively. If the 95th percentile of short-term exposure were considered for these workers, the MOSs would then be 5, 26, 20, and 30, respectively. These margins of safety remain less than the value (100, for a critical NOEL from a laboratory animal study) conventionally recommended to protect people from the toxic effects of a chemical. The greatest uncertainty was associated with the exposure assessment. As more than 70% of the dermal patches analyzed in the occupational exposure studies involving mevinphos and a surrogate pesticide contained nondetectable levels of residues, the accuracy of the occupational exposure estimates were questioned, too. One alternative would have been to use the PHED to estimate worker exposures. However, PHED mean exposure values were approximately the same as those used. A second alternative would have been a biological monitoring study, in which the absorbed dose would have been estimated from urinary metabolites of the parent compound. However, no acceptable study of this type was available. The personal protective equipment and clothing already required for mevinphos handlers was close to the maximum level permitted in California's climate (CCR, 1989). Consequently, it did not appear possible to mitigate the estimated excessive exposures. Before any regulatory action was taken, the manufacturer voluntarily withdrew the registration of mevinphos (EPA, 1994). 32.4.3 PROPOXUR

Propoxur [2(l-methylethoxy)phenol methyl carbamate] is a carbamate insecticide used alone or in combination with other insecticides in interior crack-and-crevice treatments, room foggers, flea and tick sprays, flea and tick collars, ant and cockroach traps, insecticide tapes, ant and cockroach sprays, wasp, bee, and hornet sprays, and flea and tick dips for pets. It does

697

not have any food uses in the United States, but it is approved for crop use in the EU. As was the case with ethoprop, the principal route of exposure for most pesticide applicators using propoxur is through the skin. Consequently, the dose-response data of adverse effects observed in short-term dermal toxicity studies were the initial choice as the basis for calculating margins of safety for workers with short-term exposure to propoxur (Cochran et aI., 1997). A single dermal dose of 2000 mg/kg caused clinical signs (fasciculations, decreased motor activity, hyperreactivity) in rabbits. The single-dose dermal NOEL for clinical signs was 1000 mg/kg in both rabbits and rats. These single-dose dermal NOELs for clinical signs were considerably greater than the oral NOELs in the same species. In a rabbit developmental study, the maternal NOEL (l day) was 10 mg/kg-day, based on cholinergic signs and death at 30 mg/kg. In the rat, the LOEL (30 min) for cholinergic signs (convulsions, reduced motility, apathy, bristling coat) from a single oral dose was 25 mg/kg with a NOEL of 5 mg/kg. The LOEL for maternal toxicity (cholinergic signs) in a rat developmental study was 9 mg/kgday, with a I-day NOEL of 3 mg/kg-day. A single oral dose of 5 mg/kg resulted in cholinergic signs (muscle fasciculations) in dogs, but a dose of 4 mg/kg did not produce any signs. In a single-oral-dose neurotoxicity study, the LOEL for cholinergic signs (excessive chewing and reclining posture) and significant brain cholinesterase inhibition was 2 mg/kg-day. Differences in the effective dose were due to the slower and reduced percentage of dermal absorption compared to oral absorption. Despite the fact that dermal dosing is more germane to human exposure scenarios, the dermal NOELs, 1000 mg/kg for clinical signs in rats and rabbits, were not used as the basis for assessing the risks from acute exposure to propoxur. Nor were oral NOELs for clinical signs in laboratory animals used as the basis for risk characterization. The toxicological basis for characterizing the risk from acute exposure to propoxur was an oral NOEL from a human study. The human oral NOEL was used because (1) the use of human dose-response data eliminates the uncertainty associated with extrapolating to humans from laboratory animal studies and (2) the quality of the dermal toxicity studies was not comparable to the clinical observations in the human study. In the human study, volunteers (number unstated) were reported to have exhibited cholinergic signs (stomach discomfort, blurred vision, moderate facial redness, and sweating) after a single bolus oral dose of 0.36 mg/kg. Doses of 0.2 mg/kg administered every half hour for up to 2 1/2 h (a total of 1 mg/kg) produced no cholinergic signs. Thus, the 30-min NOEL for cholinergic signs in humans following a single bolus dose was 0.2 mg/kg. In the same study, red blood cell cholinesterase activity was depressed about 2% after the first dose and 10% after a total of 5 doses. This indicated a cumulative inhibitory effect on cholinesterase activity by multiple doses of propoxur. The NOELs for clinical signs (specified amounts-up to 1 mg/kgfor specific lengths of time-up to 2 1/2 h) were used to evaluate the health risks from potential acute exposures of different durations to propoxur.

698

CHAPTER 32


The studies and data, which formed the basis for estimating worker exposure, were based on passive dosimetry using patches attached to clothing. Ethanol hand washes were collected to assess hand exposure, and air levels were monitored with personal pumps. Patches with nondetectable levels of propoxur were given default values equal to 50% of the MDL. The monitored activities in one location took 1.8 h to complete. This was defined as one cycle. The geometric mean exposure values used for the risk assessment are shown in Table 32.3. The 95th percentiles of the absorbed cycle dosage for the respective work tasks were as follows: aerosol (1 %) applicator, 2.30 I-lg/kg-cycle; bait (2%) applicator, 0.61 I-lg/kg-cycle; spray (0.95%) applicator, 0.32 I-lg/kg-cycle; and spray (70WP) applicator, 8.0 I-lg/kg-cycle. Nonoccupational exposures to office workers and home residents may occur through dermal contact with treated surfaces and, to a lesser extent, via inhalation of pesticide vapors. The potential passive exposures of residents to propoxur after crackand-crevice treatment of a home were based on studies submitted by the registrant. The data were derived from wipe samples in various rooms of the home. Analysis of the samples indicated a log-normal distribution of surface residues throughout the house. Air concentrations in the home were more or less constant. It was assumed that infants (6-9 months old) had a body weight of 7.5 kg with 0.45 m2 of surface area-50% of which could be exposed to pesticides. Their breathing rate was

Table 32.3 Mean Acute Exposures to Propoxur and Their Respective MOSs Absorbed cycle dosage Activity

(iJ.g/kg-cycle )a

Acute MOSb

Aerosol (1%) applicator (N = 32)

0.95

842

Bait (2%) applicator (N = 32)

0.19

4210

Spray (0.95%) applicator (N = 32)

0.16

5000

Spray (70WP) applicator (N = 16)

1.47

544

Passive exposure Infant (6-9 months)

1.46

548

Adolescent (12 years)

0.22

3643

Adult

0.37

2000

Active exposure Dog groomer (N = 15)

10.3

97

aGeometric mean of one application-assumes that workers' body weights were 76 kg, dermal penetration was 0.351 %Jh, and respiratory uptake was 50%. The monitored activities in one location took 1.8 h to complete, which was rounded off to 2 h. This is defined as one cycle. For people living in homes treated with crack-and-crevice treatments of propoxur for insect control, the exposures were calculated using default assumptions. It was assumed that infants (6-9 months) had a body weight of 7.5 kg with 0.45 m 2 of surface area-50% of which could be exposed to pesticides. Their breathing rate was 0.5 m3Jh with 100% absorption. For children (12 years), it was assumed they had a body weight of 40.5 kg with 1.37 m2 of body surface area and a breathing rate of 0.9 m3Jh. For adults, the assumptions were 76 kg body weight, 2.0 m2 surface area, and 1 m3Jh breathing rate. b Acute MOSs were based on a 2-h human NOEL of 800 iJ.g/kg for cholinergic signs.

0.5 m 3Jh with 100% absorption through the inhalation route. For children (12 years old), it was assumed that they had a body weight of 40.5 kg with 1.37 m2 of body surface area and a breathing rate of 0.9 m3Jh. For adults, the assumptions were 76 kg body weight, 2.0 m2 surface area, and 1 m 3Jh breathing rate. The geometric means of passive, nonoccupational exposures ranged from 2-h absorbed dosages of 0.22 to 1.4 I-lg/kgday (Table 32.3). Infants, 6-9 months of age, had the highest potential exposure. The study used to estimate exposure to a pet owner for flea control on two dogs involved a biomonitoring study of 15 professional dog groomers who sprayed an average of 20 dogs during an 8-h workday. A metabolite of propoxur, 2-isopropoxyphenol, was measured in urine samples collected from the study's participants. The estimated 24-h absorbed dose of propoxur was based on the amount of the metabolite in the urine. The mean absorbed dose, normalized for two dogs, is presented in Table 32.3. It was assumed that the dogs could be sprayed by a nonprofessional in a period of 2 h. The 95th percentile of the absorbed dosage for an adult engaged in spraying two dogs per day for ticks was 77.1 I-lg/kg-day. The MOSs for mean acute exposure to propoxur, based on the human NOEL of 0.8 mg/kg for a 2-h period, ranged from 97 to 5000 (Table 32.3). The MOSs for the 95th percentile of the absorbed cycle dosages ranged from 100 (applicators handling 70WP) to 2500 for spray applicators using 0.95% formulation. The MOS for the 95th percentile of the absorbed cycle dosage for dog owner/groomers was 13. The acute NOEL for propoxur was based on human oral exposure leading to cholinergic signs. Even though a single bolus dose of 0.36 mg/kg produced short-lasting stomach discomfort, blurred vision, moderate facial redness, and sweating, five oral doses of 0.2 mg/kg at 30-min intervals over a period of 2 1/2 h did not cause cholinergic signs. This indicated that carbamylation of cholinesterase, caused by bolus oral doses of propoxur, was rapidly reversed in the human body (Ellenhom et aI., 1997). However, the preponderance of occupational or nonoccupational acute exposure to propoxur was through the dermal route (approximately 99% in most instances). As absorption of propoxur via the dermal route is generally slower than absorption from the gut, decarbamylation, body metabolism, and clearance probably limit the effects of acute dermal exposure to propoxur. Consequently, the margins of safety under actual exposure conditions are probably greater than indicated.

32.4.4 DIQUAT DIBROMIDE

Diquat dibromide (6,7 -dihydrodipyrido-[ 1,2-a :2', l' -c]pyrazinediium ion) is a contact herbicide that damages plant tissues quickly, causing plants to appear frostbitten because of cell membrane destruction. It also reduces plant photosynthetic activity. This nonselective contact herbicide is used for desiccation of potato vines and seed crops, control of sugarcane flowering; and industrial and aquatic weed control. As in the preceding examples, most of the exposures to diquat dibromide involve dermal absorption (Cochran et aI.,

32.4 Examples

1994). Consequently, a short-term dermal NOEL would have been desirable as the critical NOEL for assessing the risks of acute human exposure to diquat. Unfortunately, no such singledose dermal studies were available in the CDPR database or from a search of the open literature. A subchronic dermal exposure study on the effects of diquat dibromide on rats indicated systemic effects (death) began after 6 days of repetitive dosing. In rabbits, ulceration of the gastric mucosa, degeneration of the convoluted tubules in the kidneys, areas of hemorrhage in the thymus, and congestion of the lungs and lung blood vessels accompanied by death were observed at 12.5 mg/kg-day after 3 days of repetitive dermal dosing with diquat dichloride. However, deficiencies in the dosing regime, and the difference in the chemical identity of the test material, precluded the use of this study as the basis for regulating short-term exposure to diquat dibromide. The toxicological basis for assessing the risks associated with potential short-term exposure to diquat, therefore, was identified in oral dosing studies. A single oral dose of diquat dibromide at 75 mg/kg caused clinical signs (diarrhea and stained nose) in female rats. The NOEL for clinical signs was 25 mg/kg. In a rat developmental study, the developmental NOEL (delayed ossification) was 12 mg/kg-day, whereas the maternal NOEL (decrement in body weight gain) was 4 mg/kg-day for exposure to diquat by gavage. Mice appeared to be more sensitive to diquat than were rats. The mouse NOELs for maternal toxicity (clinical signs, death) and developmental toxicity (skeletal anomalies, exencephaly, and umbilical hernia) were both 1.0 mg/kg-day. The rabbit appeared to be the most sensitive laboratory animal to diquat in developmental toxicity studies. The maternal NOEL was 3.0 mg/kg-day (histopathological changes in the liver, intestine, and vasculature; mortality), but there was no developmental NOEL. The ossification of the ventral tubercle of the cervical vertebrae was delayed significantly in all treatment groups compared to controls. In addition, the incidence of fetal malformations was significantly greater in the low-dose (1 mg/kg-day) and high-dose (10 mg/kg-day) groups compared to the controls. The incidence at the mid-dose (more than a twofold increase over controls) lacked statistical significance, but may have represented a biologically significant finding, supportive of a treatment-related effect. This hypothesis was consistent with the suggested common mechanism (interference with cell migration) for the observed anomalies across the different treatment groups. Although the fetal malformations in the low-dose group (1 mg/kg-day) and the high dose group (10 mg/kg-day) were quantitatively different from the controls, the malformations were not qualitatively different from either the concomitant or the historical controls. Nonetheless, the possibility that diquat caused a significant (p < 0.05) increase in the number of malformations found at the low dose (1 mg/kgday) could not be ignored. As there was no NOEL, an estimated no effect level (ENEL) was calculated. Because the magnitude (incidence) of the effect was small, and the slope of the dose response was fairly shallow, an uncertainty factor of 3 was used to derive an ENEL of 0.33 mg/kg-day. As absorption of diquat

699

across the gut in a rat pharmacokinetic study was approximately 10%, the ENEL was divided by a factor of 10 to reflect the absorbed dosage that would be expected to have no toxic effect. This adjusted ENEL, 0.033 mg/kg-day, based on the observations of delayed ossification and fetal malformations, was used to calculate margins of safety for potential short-term exposure to diquat. It was assumed that the developmental toxicity observed in pregnant rabbits could occur as the result of a single dose. Further, it was assumed that absorption of oral dosages by the rabbit and human would be limited to the same degree as in the rat. The data that formed the basis for estimating worker exposure were based on monitoring studies for diquat and calculations from studies involving a surrogate active ingredient (paraquat) with similar application rates and chemical properties. Dermal absorption constituted the principal route of exposure. Monitoring data for both aquatic use and ground spraying indicated that less than 1% of the total exposure came through the inhalation route. Exposure data from aerial application of the surrogate herbicide, paraquat, indicated that pilots and flaggers could be exposed as much through the inhalation route as through the dermal route. The exposure estimates used for the risk assessment are shown in Table 32.4. Potential short-term exposures ranged from 0.2 !-lg/kg-day for mixers and applicators injecting diquat into aquatic environments to 106 !-lg/kg-day for ground applicators driving tractors with no cabs and a normal ground clearance. The potential short-term exposure to drift was 0.5 !-lg/kg-day at 50 and 0.01 !-lg/kg-dayat 1600 m. Potential short-term exposures (4 h) for adult males swimming in treated water 24 h after application ranged from 0.2 to 1.3 !-lg/kg-day. Only mean exposure values were available; potential upperbound exposures would likely have been greater. MOSs for short-term occupational exposures, based on the adjusted ENEL of 33 !-lg/kg for developmental toxicity and maternal clinical signs, ranged from less than 1 (ground applicators) to 165 (aquatic mixers and injectors). The MOSs for potential short-term exposure of swimmers to diquat dibromide ranged from 26 (theoretical water concentration) to 165 (measured water concentration). If the developmental ENEL of 0.033 mg/kg-day had not been used as the basis for calculating the MOSs for acute exposure, the next best oral NOEL was 1 mg/kg-day for clinical signs and death in the mouse (LOEL = 2 mg/kg-day) from a developmental study. This NOEL would also have been adjusted to 0.1 mg/kg-day because of the 10% oral absorption. Using the adjusted NOEL of 0.1 mg/kg-day for clinical signs and death in the mouse, the MOSs for mean acute occupational exposures would have ranged from less than 1 to 500, and the MOSs for mean acute nonoccupational exposures would have ranged from 77 to 1000. Consequently, the conclusions would not have changed. Margins of safety for some exposures would have remained less than 100. A mitigation process was initiated to ascertain whether changing the manner in which diquat was applied would reduce exposure. Additional exposure studies and a new acute dermal

700

CHAPTER 32


Table 32.4 Potential Mean Absorbed Daily Dosages and Margins of Safety from Exposures to Diquat Dibromide

Activity

ADD

Short-term

(j.lg/kg-day)

Mosa

Aquatic Mixer (injection)

0.2

165

Applicator (injection)

0.2

165

Applicator (handgun)

3.6

9

Boat driver (handgun)

0.9

37

Aerial Mixerlloader

7.8

4

Pilot

0.3

III

FJagger

8.1

4

Ground application Applicator-normal clearance, cab

106

human. For thiocarbamate cleavage, the rank order was rat > dog > mouse » rabbit > monkey (no data were available for humans). Wickramaratne et al. (1998) argued that the metabolic differences, combined with the unique role of high-density lipoproteins in cholesterol mobilization in rodents (which is inhibited by metabolites of molinate) as opposed to other mammals which rely on low-density lipoproteins (whose esterase, acetyl-CoA, is not inhibited by molinate metabolites) as their primary source of cholesterol, suggest that the rodent data on testicular toxicity is not relevant to humans. 34.3.3 DIBROMOCHLOROPROPANE The nemacide dibromochloropropane (DBCP) reached notoriety in 1977 when it was discovered that workers in a pesticide manufacturing plant had become oligospermic. Only 7 of 26 DBCP-exposed workers had normal sperm counts 11 years later (Lahdetie, 1995). Although the effects of DBCP on the adult testes appear to be related to direct effects on spermatogenic cells, most likely through damage to DNA (Lag et aI., 1989), there is some evidence that the mode of action on the developing fetus might have a endocrine basis. Determining which is a cause and which is an effect in such situations, however, is complicated. Warren et al. (1988) exposed pregnant rats to 25 mg/kg DBCP on days 14.5-19.5, 16.5-19.5, or 18.5-19.5 of gestation. Treatment for 6 days reduced intratesticular testosterone concentrations on day 20.5 by 50%. In adulthood, all exposure durations reduced male body weights, whereas testes weights were reduced 75% by 2 days of exposure and 90% following 4 or 6 days of exposure. In the brain, the volume of the normally sexually dimorphic nucleus of the preoptic area in males treated for 6 days in utero was not different from that of control females, and these males all displayed female lordosis behavior. A few of the males treated for 4 or 6 days had testes lacking seminiferous tubules.

737

34.3.4 TRIAZOLES Inhibition of aromatase activity, the enzyme which converts testosterone to estrogen, in the ovarian granulosa cell by the antifungal triazole 1, I-di -( 4-ftuorophenyl)-2-(1 ,2,4-triazoll-yl)-ethanol (RI51885) has been linked to a blockade of ovulation, particularly when given during diestrus I or 11 of the estrous cycle (Middleton et aI., 1986; Milne et aI., 1987). Doses as low as 5 mg/kg by gavage completely suppressed ovulation. When given at midday of diestrus 11, there were no effects on serum LH, FSH, or progesterone until the afternoon of proestrus. However, plasma estradiollevels were reduced by nearly 50% within 12 hours of treatment and remained low for an additional 12 hours. The agent was not directly uterotrophic in the ovariectomized mature rat, but doses of 25 mg/kg were able to reduce estradiol-stimulated uterine weight increases. 34.3.5 TRIBUTYLTIN One of the clearest examples of pesticide-induced populationlevel effects in wildlife via alterations of the endocrine system is that of altered reproductive development and subsequent population declines in marine snails brought about by exposure to tributyltin, an antifouling agent (Ankley and Giesy, 1998). In sensitive species, tributyltin has been shown to inhibit the metabolism of testosterone to estrogen by aromatase, with subsequent development of male reproductive organs in females, condition referred to as "imposex." Marine snails appear to be especially sensitive to this mechanism of action. Effective concentrations are in the range of nanograms per milliliter and population declines are a global observation. Despite the dramatic nature of the effect in marine snails, tributyltin-induced inhibition of aromatase has not been reported in higher organisms, including mammals. However, it is toxic to the mammalian immune system by a nonsteroid-receptor mode of action.

34.4 HYPOTHALAMIC-PITUITARY FEEDBACK LOOPS 34.4.1 FEMALE NEUROENDOCRINE REGULATION 34.4.1.1 Dithiocarbamates Dithiocarbamates are a broad chemical class including fungicides such as the ethylenbisdithiocarbamates, metam sodium and thiram. They are also metal chelating agents and are known to inhibit the synthesis of neurotransmitters, particularly norepinephrine, via chelation of the copper-containing portion of the enzyme dopamine-,B-hydroxylase. Norepinephrine plays a critical role in the release of gonadotropin-releasing hormones (GnRH) from the hypothalamus. During a short time period (between 1400 and 1600 hours on the day of proestrus), the sequential feedback of estrogen and then progesterone stimulates

738

CHAPTER 34

Pesticides as Endocrine-Disrupting Chemicals

the activity of a-adrenergic neurons which induces a phasic release of GnRH. This, in turn, triggers the anterior pituitary to release a surge of LH. Concentrations of LH rapidly rise in serum from levels below 1 ng/ml to 5-10 ng/ml and ovulation is induced. There is a minimum concentration for this surge to be effective in inducing ovulation in spontaneous ovu1ators such as the rat and the human. Because of the critical timing of events, and the multiple steps which are susceptible to disruption, analysis of the control of ovulation has proven a particularly useful tool in understanding neuroendocrine toxicology. Ovariectomized, estrogen-primed female rats given a single injection of 50 or 100 mg/kg thiram at 1100 hours lacked the expected elevation in LH measured over the succeeding 4 hours. A dose level of 25 mg/kg blocked the surge in some animals, and attenuated it in others (Stoker et aI., 1993). Administration of 50 mg/kg to intact females on the afternoon of vaginal proestrus delayed ovulation by 1 day. When mated the following evening, there were no differences in the proportion of sperm-positive females compared to controls mated the previous evening, but there was a significant decrease in embryo viability between gestation days 7 and 11, indicating that ovulation delayed for 24 hours was deleterious to the ripening oocytes (Stoker et aI., 1996). Doses of 50 and 100 mg/kg metam sodium induced nearly identical effects, and the surge effect was reversed by the a-adrenergic agonist c1onidine. In addition, anterior and posterior hypothalamic norepinephrine levels fell 3 hours after injection and were accompanied by a rise in dopamine (Go1dman et aI., 1994). Using sodium dimethyldithiocarbamate and comparing systemic versus intrabursal injections, Goldman et al. (1997) went on to demonstrate that, although the effect on hypothalamic catecholamine synthesis may underlie the ovulatory blockade, there is also a local ovarian response that is independent of ovarian norepinephrine concentrations. Thus, there appear to be multiple, albeit uncharacterized, modes of action by which the dithiocarbamates may disrupt the regulation of ovulation. 34.4.1.2 Formamidines The formamidines are a class of insecticides which include amitraz and chlordimeform, the latter of which is no longer marketed in the United States due to its carcinogenic potential. Their mode of insecticidal activity is based on mimicking the insect neurotransmitter octopamine, but they are also capable of binding to and inhibiting a-adrenergic receptors in mammals. As seen with the dithiocarbamates, interference with norepinephrine action in the hypothalamus can lead to significant alterations in female reproductive function. For instance, both amitraz and chlordimeform block the LH surge (Cooper et aI., 1994), which is mediated in part by norepinephrine. 34.4.1.3 Atrazine Concern for the endocrine-disrupting effects of atrazine, a chlorotriazine herbicide, arose following the observation of increased incidence of mammary tumors in a chronic bioassay in female Sprague-Dawley (SD) rats exposed to 400 ppm atrazine

in the diet for 104 weeks. These tumors also appeared in control females, but occurred earlier in the treated females. No other tumors were present in the treated Sprague-Dawley female rats, nor in male Sprague-Dawley rats or male and female Fischer 344 rats (Stevens et aI., 1994; Thakur et aI., 1998). The finding of an earlier onset of mammary tumors led to an investigation into the estrogenicity of atrazine, but under equilibrium conditions, atrazine was not able to compete with estradiol for binding to rat uterine estrogen receptors. A weak competition was noted if the cytoso1s were preincubated at 25°C prior to incubation with the tracer (Tennant et aI., 1994a). Somewhat conflicting results have been seen in other studies. Daily exposure of adult Fischer rats to 120 mg/kg for 7 days resulted in fewer treated females displaying normal estrous cycles, and the number of days in diestrus increased significantly. Fertility was reduced in females during the first week after exposure, but pregnancy outcome was not affected in those that became inseminated (Simic et aI., 1994). However, treatment of adult, ovariectomized SD rats with up to 300 mg/kg atrazine by oral gavage for 3 days did not result in an increase in uterine weight, nor were there increases in uterine progesterone levels, suggesting the lack of an estrogenic potential. Indeed, when estradiol (2 !J.g/kg subcutaneously) was given in conjunction with 300 mg/kg or orally administered atrazine, there was a weak inhibition (~25%) of the uterotrophic response (Tennant et al., 1994b). In a similar study, immature female SD rats were dosed with 0, 50, 150, or 300 mg/kg atrazine by gavage for 3 days. Uterine weight was not increased, but decreases in uterine progesterone receptors and peroxidase activities were noted; however, when combined with estradiol, antiestrogenic effects of atrazine including decreases in uterine progesterone receptor binding and uterine peroxidase was not noted on the uterus (Connor et aI., 1996). In this same study, atrazine did not affect basal or estradiol-induced MCF-7 cell proliferation, nor did it display agonist or antagonist action against estradiolinduced 1uciferase activity in MCF-7 cells transfected with a Ga14-regulated human estrogen receptor chimera. To further evaluate effects on reproductive function, female Long Evans (LE) and SD rats that had been screened for regular 4-day estrous cycles, received 0, 75, 150, or 300 mg/kg/day atrazine by gavage for 21 days. In both strains, atrazine disrupted the regular 4-day estrous cycles. For the LE rats, all dose levels were effective, whereas SD rats required a higher dose (150 mg/kg/day) for a longer time for this effect to appear. The increased time spent in vaginal diestrus was associated with elevated serum progesterone and low estradiol concentrations, indicative of a repetitive pseudopregnant condition. This hormonal condition was not considered by the authors to be conducive to the development of mammary tumors, although there was some indication of prolonged estrous at the lowest dose tested (Cooper et aI., 1996). The strain difference noted in the premature onset of mammary tumors (insensitive Fischer 344 rats versus sensitive SD rats) has been attributed to differences in the normal aging of the reproductive tract in these strains (Eldridge et aI., 1994; Stevens et aI., 1994; summarized in Chapin et aI., 1996). Re-

34.4 Hypothalamic-Pituitary Feedback Loops

productive cycling in the female SD rat begins to decline in animals less than a year of age, presumably due to the loss of sensitivity of adrenergic neurons in the hypothalamus that control GnRH release to the pituitary. This loss of stimulation reduces FSH and LH release, and ultimately ovulatory failure. In turn, the ovaries contain many follicles but no corpora lutea. In contrast, adrenergic neurons of female Fischer 344 rats do not seem to lose their sensitivity to estrogen stimulation, and regular cycling is maintained for a much longer time. Also in contrast to the onset of persistent estrus, reproductive aging in the Fischer 344 is believed to be due to an inability to control daily prolactin surges, a prolonged activity of the corpora lutea (i.e., repetitive pseudopregnancy), and a higher level of progesterone release. Hence, the endocrine milieu of the aging SD rat, but not the Fischer 344 rat, favors development of mammary tumors and helps explain the difference in incidence of spontaneous tumors as females of these strains age. How atrazine accelerates the neuroendocrine aging of the reproductive axis in the SD rat, however, has not been determined. Although the induction of mammary tumors by atrazine may not be relevant to humans (International Agency for Research on Cancer, 1999), the action of atrazine on the hypothalamus may be of some significance to human health. 34.4.2 THYROID TUMORS Endocrine disruption of the pituitary-thyroid axis is a relatively well understood process by which endogenous chemicals induce thyroid follicular cell neoplasia. The physiological regulation of thyroid cell growth and function involves a complex interactive network of trophic factors that are mediated by a number of second messenger systems (Hard, 1998). TSH is the main growth factor for follicular cells, with insulin-like growth factor 1 (IGF-l), epidermal growth factor (EGF), basic fibroblast growth factor (bFGF), and transforming growth factor f3 (TGF-f3) also involved in various ways. Activation of TSH receptors stimulates G protein-dependent rises in cAMP and phospholipase C, with resulting consequences of iodine uptake and release, thyroid peroxidase (TPO) generation, thyroid hormone synthesis and release, and thyroid cell growth and division. Relative to metabolism, T4 is secreted by the thyroid, but must be converted to T3 via either Type I 5' -diodinase in the liver or Type 11 5' -diodenase in the brain, pituitary, and brown adipose tissue. There are three main carrier proteins for thyroid hormones, thyroxine binding protein (65%), transthyretin (20%), and albumin (10%); only about 5% of the hormone is unbound (in the rat, thyroxine binding protein is absent during most of adult life). Further metabolism occurs in the liver, intestines, and kidneys and involves inactivation of biological activity by conjugation with glucuronic acid or sulfate. Whether by reduced synthesis due to inhibition of TPO, reduced peripheral de-iodination, or by elevated turnover via induction of conjugating enzymes, sustained release of TSH in response to decreased circulating levels of thyroid hormones is intimately involved in thyroid gland neoplasia. This suggests that nonlinear thyroid cancer dose-response considerations can be applied

739

to chemicals that reduce thyroid hormone levels, increase TSH and thyroid cell division, and are judged to lack mutagenic activity (Hill et aI., 1998). Although much of thyroid gland physiology is similar across experimental animals and humans, there are, as noted above, some important differences that may reduce the sensitivity of humans relative to rodents (Hard, 1998). Interestingly, childhood radiation is the only known exogenous risk for thyroid gland carcinogenesis in humans. A total of 240 pesticides have received an in-depth review for potential carcinogenicity by the US EPA, with evidence of induction of thyroid follicular tumors in appropriate chronic tests present for (Hurley, 1998). Thyroid tumors were second only to liver tumors in the frequency with which they were observed. Three pesticides (amitrole, ethylene thiourea, and mancozeb) induced a high incidence (>0.48) at relatively low daily doses (3.5-30.9 mg/kg/day). All but 2 of the 24 pesticides induced thyroid tumors only in rats; none induced tumors only in mice; 8 induced tumors only in males; none were positive in females only. Sixteen induced tumors in at least one other site, the most frequent being the liver; other sites of preponderance being the glandular stomach, mammary gland, parathyroid, pancreas islet cells, testis, and thyroid C cell. Based on gene mutation and chromosomal aberration tests, a direct mutational mode of action appears possible for only 3 of the 24 pesticides: acetochlor, ethylene thiourea, and etridiazole. None of the pesticides has had a complete evaluation of all potential sites of antithyroid action (inhibition of iodide uptake, inhibition of thyroid peroxidase, damage to thyroid follicular cells, inhibition of thyroid hormone release, inhibition of 5' -monodeiodenase activity, and enhancement of metabolism and excretion by the liver), but 12 have sufficient information to infer a mode of action. Three (amitrole, ethylene thiourea, and mancozeb) inhibit thyroid peroxidase; four inhibit the iodide pump (amitrole, ethiozin, ethylene thiourea and pentachloronitrobenzene), nine stimulate thyroid hormone metabolism and excretion (acetochlor, clofentezine, fenbuconazole, fipronil, pendimethrin, pentachloronitrobenzene, prodiamine, pyrimethanil, and thiazopyr). 34.4.3 LEYDIG CELL TUMORS Leydig cells, which are situated in the interstitial space between seminiferous tubules in the testes, serve as the primary source for synthesis of androgens in mammals. As in the thyroid gland, alterations in endocrine feedback loops in the hypothalamic-pituitary-gonadal axis have been linked to Leydig cell hyperplasia-and, ultimately, neoplasia (reviewed in Cook et aI., 1999). The primary trophic influence over Leydig cells is exerted by pulsatile release of LH by the anterior pituitary following stimulation by GnRH originating in the preoptic and medial basal areas of the hypothalamus. In the Leydig cell, LH activates adenyl cyclase via a G protein, which initiates a cascade of events leading to increased steroidogenesis. Testosterone completes the feedback loop by providing inhibitory signals to the hypothalamus and pituitary to lower LH secretion. Also similar to observations regarding the thyroid

740

CHAPTER 34


gland, a number of other trophic factors affecting the Leydig cells have been identified, including IGF-l, TGF-a, TGF-,B, bFGF, interleukin I (IL-l), inhibin, and activin. Because Leydig cell hyperplasia is frequently noticed in areas proximate to seminiferous tubule damage, a local endocrine imbalance has been assumed to play a key role, but in nearly all cases a sustained elevation in serum LH is a key mediator of the response. Almost all Leydig cell tumors are benign adenomas, and the distinguishing feature between hyperplasia and tumors is the size of the nodule. There are species differences in the susceptibility to Leydig cell hyperplasia produced by exposure to pesticides. For example, Murakami et al. (1995) studied the effect of procymidone, an antiandrogenic dicarboximide fungicide, in inducing testicular interstitial cell tumors in rats and mice. Male SpragueDawley rats and ICR mice were exposed via the diet to 7006000 ppm and 1000-10,000 ppm, respectively, for 3 months. In the rat, serum and intratesticular testosterone and serum and pituitary LH levels were increased in treated animals throughout the exposure. The endocrine effects were more pronounced at 4 versus 13 weeks. Interstitial cells were also hypersensitive to hCG in vitro at 2, 4, and 13 weeks, as evidenced by the rate of testosterone release. In contrast, although serum and pituitary levels of LH were elevated after 4 weeks of exposure, no significant changes in testosterone were detected in mice, either in vivo or in vitro. For a reason not yet understood, rat Leydig cells appear more sensitive to chronic stimulation by LH than the mouse cells, an effect which correlates with observations of tumors. As noted by Cook et al. (1999), Leydig cell tumors are exceptionally rare in human populations.

34.5 MISCELLANEOUS OR UNKNOWN MODES OF ACTION 34.5.1 NITROFEN Nitrofen (2,4-dichlorophenyl-p-nitrophenyl ether) is a preemergent herbicide removed from the U.S. market in the early 1980s due to concerns over its developmental toxicity. Administration during gestation to both rats and mice induces a constellation of developmental alterations that in many respects resemble those of altered thyroid hormone function. Consistent with the similarity in effects, nitrofen also bears structural similarity to thyroid hormone. These effects include delayed maturation of the lungs and pulmonary surfactant, small or missing Harderian glands, growth deficits, delayed opening of the eyes, and altered neurobehavioral development. In addition, malformations of the cardiovascular system and urogenital track and diaphragmatic hernias have been observed (Gray et aI., 1982; Ostby et aI., 1985). Developmental toxicity was observed at doses as low as 4.17 mg/kg/day given to rats on gestation days 8-16. In subsequent studies to examine whether nitrofen interferes with the hypothalamic-thyroid axis, adult mice were given 500 or 1000 mglkglday for 3 days (Gray and Kavlock, 1983). Serum thyroxine was reduced by 60% in the high-dose

group and by 20% at the low dose, but no changes in serum T3 levels were found. In this study, there was no effect on body weight, but liver weights were slightly increased. In support of the role of altered thyroid function induced by nitrofen on the developing organism, Manson et al. (1984) reported that a single dose of 250 mg/kg administered to pregnant rats on gestation day 11 significantly depressed TSH. The decrease was most evident at 6 and 24 hours after dosing, but were still present at term. Maternal serum T4 levels levels were likewise depressed at 8 and 24 hours after treatment, but were normal at term. No effects were seen in T3. Although confounded by the effects ot thyroidectomy, administration of T4 (4 Ilg per 100 g body weight) with 25 mg/kg nitrofen (days 9-11 of gestation) resulted in a 70% reduction in malformations, with the heart most protected and the kidney least protected. Competitive radioimmunoassay binding studies indicated that the 4-hydroxy-2,5-dichloro-4'-aminophenyl ether metabolite could displace T3. Collectively these results indicate that some of the nitrofen-induced developmental toxicity is potentially mediated by alterations in the thyroid status of the maternal organism and her fetus.

34.6 IMPACT ON TESTING GUIDELINES Assessment of the potential developmental and reproductive risks of environmental contaminants is generally determined through application of testing guidelines that are established by regulatory agencies such as the U.S. EPA or by international coordinating bodies such as the Organization of Economic Cooperation and Development (OECD). In 1991, the U.S. EPA began a process to update its developmental and reproductive testing guidelines to ensure that they incorporated contemporary scientific methodology. Traditionally, these tests had been very apical in nature; that is, they relied on end points which were diagnostic of adverse biological outcomes, but did not provide clarification of potential modes of action, target organs, or most sensitive life stage or gender. For example, the multi generation reproductive test guidelines require groups of animals (generally rats) to be exposed to the test chemical beginning shortly after weaning and continuing until production of the second generation (thus concluding with examination of offspring of animals exposed from fertilization and through reproduction). In the past, the primary end points that were evaluated in such tests included fertility (are the animals capable of reproducing?), fecundity (how many offspring are produced?), and growth of the offspring. A large body of evidence suggests, especially for endocrine-disrupting chemicals, that these end points are neither very sensitive to reproductive disturbance nor indicative of the underlying biological effect. Both these issues raise concern regarding the suitability of previously issued test guidelines to satisfactorily detect and characterize reproductive hazard. As part of this effort to revise the testing guidelines, particular emphasis was placed on improving the ability to detect the action of chemicals that may act via the endocrine system to

34.6 Impact on Testing Guidelines

741

Table 34.4 Summary of Relative Sensitivity of End Points in the Traditional Multigenerative Reproduction Study (Fertility, Fecundity, and Somatic Growth) Compared with Alternate Measures of Reproductive Function or Capacity for Chemicals that Work through the Estrogen, Androgen, and Ah Receptorsa Weak estrogenb Female

Male

Fertility

+

+

Fecundity

+

Male

Growth ofF1

++

Physiology

Male

Female + (TTP)

+ -(AGD)

-(AGD)

+++ (YO)

++ (PS)

+ (YO)

++(EC)

+ + + (EO)

+++(AGD) + (PS)

Female

++

Sex differentiation Puberty

Ah agonistd

Antiandrogen C

-(AGD)

Gamete number

++ (CSC)

+ (ESC)

+ + + (ESC)

Accessory sex gland weights

++

+++

+

Gonad weight Pituitary hormones

+ +++(PL)

Steroid hormones

-(isT)

+

Malformations

-(T) + (HS)

++(YT)

QThe lowest dose level at which a statistically significant effect was observed is noted by symbols (the lower the dose within a chemical, the number of "+"s, see footnotes b-d for chemical, exposure duration, and dose levels). Reprinted with permission from Kavlock (1999). bMethoxychlor, GDI5-PD21; dose levels: -, 200 mg/kg/day; +, 100 mg/kg/day; ++,50 mg/kg/day; + + +, 25 mg/kg (Gray et aI., I988a, 1989). Fertility and fecundity data are for the parental generation. cYinc!ozolin, GDI4-PD3; dose levels: -, 100-200 mglkg/day; +,50 mg/kg/day; + + + 12.5-25 mg/kg/day; + + +, 3-6 mg/kg/day (Gray et aI., 1994; Ostby et aI., 1997). dDioxin, GD 15; dose levels; -, I j.lg/kg; +,0.8 j.lg/kg; ++, 0.2 j.lg/kg; + + +,0.05 j.lglkg (Gray et aI., 1995, 1997b, c; Gray and Ostby, 1995). Abbreviations: AGD, anogenital distance; CSC, caudal sperm count; EC, estrous cyC!icity; EO, age at eye opening; ESC, ejaculated sperm count; GD, gestation day; HS, hypopsadias; isT, in vitro stimulated testosterone release from testes; PD, postnatal day; PL, prolactin; PS, age at preputial gland separation; YO, age at vaginal opening; YT, vaginal thread; T, serum testosterone; TTP , time to pregnancy.

perturb reproduction. Data from multigeneration studies (summarized in Table 34.4), obtained from the same laboratory so that comparison across end points and chemicals is relatively straightforward, indicate that, for chemicals that act via the estrogen receptor (e.g., methoxychlor), the androgen receptor (e.g., vinclozolin), or the Ah receptor (e.g., dioxin), the traditional end points of fertility, fecundity, and growth do tend to pick up effects, but confirm that there are a lack of sensitivity and a poor ability to characterize the overall impact. For example, one of the most sensitive indicators of developmental exposure to an estrogen is accelerated puberty in the female (Gray et aI., 1988a), whereas diminished anogenital distance and accessory sex gland weights are most sensitive to developmental exposure to an anti-androgen (Gray et aI., 1994), and decreased ejaculated sperm counts are the most sensitive to chemicals that act via the Ah receptor (Gray and Ostby, 1995). None of these effects would have been identified by the traditional multigeneration reproductive test. The effects of methoxychlor on the growth of the parental males present an interesting example of where additional information can help interpret the data (Gray et aI., 1988b). In this instance, the impairment of growth might be considered a manifestation of systemic toxicity if it were not known that estrogens will decrease appetite in males and hence result in decreased food consumption and body growth. How the emerging finding (Maness et aI., 1998) that methoxychlor interacts with both the estrogen receptor and the androgen (as an agonist and antago-

nist, respectively), remains to be reconciled with the pattern of effects observed in vivo, where the estrogenic effects appear to prevail. The newly harmonized multigeneration reproductive testing guidelines (U.S. EPA, 1997b) now include a number of end points to monitor reproductive performance and health. These include assessments of the following: female estrous cyclicity; sperm parameters (total number, percentage progressively motile, and sperm morphology in both the parental and FI generations); the age at puberty in the FI generation (vaginal opening in the female, preputial separation in the males); an expanded list of organs for pathology, gravimetric analysis, and/or histopathology to identify and characterize effects at the target organ; some triggered end points including anogenital distance in the F2 generation and primordial follicu1ar counts in the parental and FI generations. For the new prenatal developmental toxicity test guidelines (U.S. EPA, 1997b), one important modification related to the improved detection of endocrine disruptors was the expansion of the period of dosing from the end of organogenesis (i.e., palatal closure) to the end of pregnancy in order to include the developmental period of urogenital differentiation. Collectively these modifications of the test guidelines should markedly improve the characterization of endocrine-mediated effects during reproduction and development.

742

CHAPTER 34


34.7 IMPACT OF FOOD QUALITY PROTECTION ACT Over the past several years, two environmental laws enacted by the United States Congress specifically require that pesticides and other chemicals found in or on food or in drinking-water sources be tested for their potential to cause "estrogenic or other endocrine effects in humans." The Food Quality Protection Act of 1996 (FQPA) and the Safe Drinking Water Act Amendments of 1996 (SDWA) require the U.S. EPA to, within 2 years of enactment, develop a screening program using appropriate, valid test systems to determine whether substances may have estrogenic or other endocrine effects in humans. The screening program must undergo a public comment period and peer review and be implemented within 3 years. The laws require that the manufacturers, registrants, or importers of the pesticides and other substances conduct the testing according to the program the U.S. EPA develops. At joint workshops, cosponsored by U.S. EPA, the Chemical Manufacturers Association, and the World Wildlife Fund (Ankley et aI., 1998; DeVito et aI., 1999; Gray et aI., 1997a), a number of assays potentially suitable for assessing endocrine-disrupting chemicals (EDCs), particularly those for detecting (anti-)estrogenic, (anti-)androgenic, and (anti-)thyroidogenic effects, were identified and critiqued. Based upon input from these and other workshops and its own deliberations, the Endocrine Disruptor Screening and Testing Advisory Committee (EDSTAC), an advisory committee to the U.S. EPA on implementation of the FQPA and SDWA, has recommended a battery of assays for both screening and testing potential EDCs that will be used to address the mandates of the FQPA and SDWA (U.S. EPA, 1998). The assays are intended to detect potential interaction both with the sex steroids (estrogen and testosterone) and with thyroid hormone function, and they include assessment of both potential human health effects and effects in wildlife. To help prioritize chemicals for screening and testing, the EDSTAC recommended a high-throughput screening (HTPS) cellular-based, receptor-mediated gene transcription assay for chemicals which act as either agonists or antagonists for estrogen, androgen, or thyroid receptors. It has been estimated that perhaps 15,000 chemicals would be evaluated in the HTPS. The EDSTAC recommendation for the "Tier 1" screening (TIS) battery includes three in vitro assays and five in vivo assays. The in vitro assays in TIS include an estrogen receptor binding or transcriptional activation assay; an androgen receptor binding or transcriptional activation assay; and a steroidogenesis assay using minced testis. The five in vivo screens recommended include the rodent 3-day uterotrophic assay, a rodent 20-day pubertal female assay for effects on thyroid function, a male rodent 5- to 7-day Hershberger assay, a frog metamorphosis assay for thyroid effects, and a fish gonadal recrudescence assay. It is estimated that perhaps as many as 1500 chemicals would enter the TIS, and positive chemicals would move into a second level (T2T), where more defined toxicological responses would be characterized. In the U.S. EPAs Reports to Congress in 2000 (U.S. EPA, 2000) it stated that it is proceeding on two fronts to imple-

ment the screening and testing program. The first element is establishing a method for setting priorities for screening. For commercial chemicals and environmental contaminants other than pesticides, this method will include use of a database and software that EPA is developing. Prioritization of pesticidal active ingredients will be done in conjunction with a review of existing data on health and environmental effects. In the second element, it is ensuring that the Tier 1 and 2 assays are scientifically valid. Validation includes developing protocols to conduct specific assays, evaluating their effectiveness, and ensuring that the assays can be performed reliably and consistently in different laboratories. This effort is being conducted in liaison with the Interagency Coordinating Committee for the Validation of Alternative Methods (ICCVAM) and is following ICCVAM principles. In the Report to Congress, EPA estimated that the Tier 1 screens would be validated by the end of 2002, and the Tier 2 by 2004. Updates on progress of the screening and testing program can be found at: http.//www.epa.gov/ endocrine/.

ACKNOWLEDGMENTS Without the most helpful assistance of many members of the Reproductive Toxicology Division, including Janice Brown, Jerome Goldman, Susan Laws, Earl Gray, and Ralph Cooper, this effort would not have been possible.

REFERENCES Adami, H.-O., Lipworth, L., Titus-Ernstoff, L., Hsieh, c.-c., Hanberg, A., Ahlborg, U., Baron, J., and Trichopoulos, D. (1995). Organochlorine compounds and estrogen-related cancers in women. Cancer Causes and Control 6,551-566. Ahlborg, U. G., Lipworth, L., Titus-Ernstoff, L., Hsieh, C.-C., Hanberg, A., Baron, J., Trichopoulos, D., and Adami, H.-O. (1995). Organochlorine compounds in relation to breast cancer, endometrial cancer, and endometriosis: An assessment of the biological and epidemiological evidence. Crit. Rev. Toxieo!. 25,463-531. Ankley, G. T., and Giesy, J. P. (1998). Endocrine disruptors in wildlife: A weight-of-evidence perspective. In "Principles and Processes for Evaluating Endocrine Disruption in Wildlife" (R. Kendall, R. Dickerson, J. Giesy, and W. Suk, eds.), pp. 349-367. SETAC Press, Pensacola, FL. Ankley, G. T., Johnson, R. D., Detenbeck, N. E., Bradbury, S. P., Toth, G., and Folmar, L. C. (1997). Development of a research strategy for assessing the ecological risk of endocrine disruptors. Rev. Toxieol. 1, 231-267. Ankley, G., Mihaich, E., Stahl, R., Tillitt, D., Colborn, T., McMaster, S., Miller, R., Bantle, J., Campbell, P., Denslow, N., Dickerson, R., Folmar, L., Fry, M., Giesy, J., Gray, L. E., Guiney, P., Hutchinson, T., Kennedy, S., Kramer, v., LeBlanc, G., Mayes, M., Nimrod, A., Patino, R., Peterson, R., Purdy, R., Ringer, R., Thomas, P., Touart, L., Van Der Kraak, G., and Zacharewski, T. (1998). Overview of a workshop on screening methods for detecting potential (anti-) estrogenic/androgenic chemicals in wildlife. Environ. Toxieol. Chem. 17(1),68-87. Arctic Monitoring and Assessment Programme (1997). "Arctic Pollution Issues: A State of the Arctic Environment Report." AMAP, Oslo. Bal, H. (1984). Effect of methoxychlor on reproductive systems of the rat. Proe. Soc. Exp. Bioi. Med. 176, 187-196. Beard, A. P., and Rawlings, N. C. (1998). Reproductive effects in mink (Mustela vison) exposed to pesticides lindane, carbofuran and pentachlorophenol in a multigeneration study. 1. Reprod. Fertil. 113,95-104.

References

Beard, A. P., and Rawlings, N. C. (1999). Endocrine and reproductive function in ewes exposed to organohlorine pesticides lindane and pentachlorophenol. J. Environ. Toxieol. Health 56, 23-46. Bigsby, R. M., Caperell-Grant, A., Madhukar, B. V. (1997). Xenobiotics released from fat during fasting produce estrogenic effects in ovariectomized mice. Cancer Res. 57(5), 865-869. Bitman, J., Cecil, H., Harris, S. J., and Fries, G. F. (1968). Estrogenic activity of o,p'-DDT in the mammalian uterus and avian oviduct. Science 162, 371. Bulger, W. H., Muccitelli, R. M., and Kupfer, D. (1978). Studies on the in vivo and in vitro estrogenic activities of methoxychlor and its metabolites. Role of hepatic mono-oxygenase in methoxychlor activation. Biochem. Phannacol. 27, 2417-2423. Buttar, H. S., Moffatt, J. H., and Bura, C. (1989). Pregnancy outcome in ketoconazole-treated rats and mice. Teratology 39, 444. Cannon, S. B., and Kimbrough, R. D. (1979). Short-term chlordecone toxicity in rats including effects on reproduction, pathological organ changes, and their reversibility. Toxieo!. App!. Phannaeo!. 47, 469-476. Chapin, R. E., Harris, M. w., Davis, B. J., Ward, S. M., Wilson, R. E., Mauney, M. A, Lockhart, A. C., Smialowicz, R. J., Moser, V. c., Burka, L. T., and Collins, B. J. (1997). The effects of perinatal/juvenile methoxychlor exposure on adult rat nervous, immune, and reproductive system function. Fundam. App!. Toxieo!. 40, 138-157. Chapin, R. E., Stevens, J. T., Hughes, C. L., Kelce, W. R., Hess, R. A., and Daston, G. P. (1996). Symposium overview: Endocrine modulation of reproduction. Fundam. App!. Toxieo!. 29,1-17. Clement, J. G., and Okey, A. B. (1974). Reproduction in female rats born to DDT-treated parents. Bull. Environ. Contam. Toxieo!. 12,373-377. Cochran, R. c., Formoli, T. A, Pfeifer, K F., and Aldous, C. N. (1997). Characterization of risks associated with the use of molinate. Regu!. Toxieo!. Phannaeo!. 25, 146-157. Cochran, R. c., and Wiedow, M. A. (1984). Chlordecone lacks estrogenic properties in the male rat. Toxieol. App!. Phannaeol. 76,519-525. Colborn, T., and Clements, C. eds. (1992). "Chemically-Induced Alterations in Sexual and Functional Development: The WildlifelHuman Connection," Advances in Modern Environmental Toxicology, Vol. 21. Princeton Sci. Publ., Princeton, NJ. Como, J. A., and Dismukes, W. E. (1994). Oral azole drugs as systemic antifungal therapy. New England J. Med. 330(4), 263-272. Connor, K, Howell, J., Chen, I., Liu, H., Berhane, K., Sciarretta, c., Safe, S., and Zacharewski, T. (1996). Failure of chloro-s-triazine derived compounds to induce estrogen receptor-mediated responses in vivo and in vitro. Fundam. App!. Toxieo!. 30, 93-101. Cook, J. c., Foster, P. M. D., Hardesty, J. F., Klinefelter, G. K, and Sharpe, R. M. (1999). A review of Leydig cell hyperplasia in rodents and their relevance to man. Crit. Rev. Toxieol. 29, 169-261. Cook, J. C., Mullins, L. S., Frame, S. R., and Biege1, L. B. (1993). Investigation of a mechanism for Leydig cell tumorigenesis by linuron in rats. Toxieo!. App!. Pharmaeo!' 119(2), 195-204. Cooper, J. R., Vodicnik, M. J., and Gordon, J. H. (1985). Effects of perinatal kepone exposure on sexual differentiation of the brain. Neurotoxieology 6, 183-190. Cooper, R. L., and Kav10ck, R. J. (1997). Endocrine disruptors and reproductive development: A weight of the evidence overview. Endocrinology 152, 159166. Cooper, R. L., Barrett, M. A, Goldman, J. M., Rehnberg, G. L., McElroy, W. K, and Stoker, T. E. (1994). Pregnancy alterations following xenobiotic-induced delays in ovulation in the rat. Fundam. App!. Pharmacol. 22, 474-480. Cooper, R. L., Chadwick, R. w., Rehnberg, G. L., Goldman, J. M., Booth, K c., Hein, J. F., and McElroy, W. K (1989). Effect of lindane on hormonal control of reproductive function in the female rat. Toxieol. App!. Pharmaeol. 99,384. Cooper, R. L., Stoker, T. E., Goldman, J. M., Parrish, M. B., and Tyrey, L. (1996). Effect of atrazine on ovarian function in the rat. Reprod. Toxieol. 10(4),257-264. Cummings, A. M. (1997). Methoxychlor as a model for environmental estrogens. Cri!. Rev. Toxieol. 27, 367-379.

743

Cummings, A. M., Hedge, J. L., and Laskey, J. (1997). Ketoconazole impairs early pregnancy and the decidual cell response via alterations in ovarian function. Fundam. Appl. Toxieol. 40, 238-246. Davis, D. L., Bradlow, H. L., Wo1ff, M., Woodruff, T., Hoe1, D. G., and AntonCu1ver, A. (1993). Medical hypothesis: Xenoestrogens as preventable causes of breast cancer. Environ. Health Perspeet. 101, 372-377. DeCoster, R., Caers, I., Halterman, c., and Debroye, M. (1985). Effect of a single administration of ketoconazole on total and physiologically free plasma testosterone and 17fi-oestradiol levels in healthy male volunteers. Eur. J. Clin. Phannaeo!. 29, 489. DeVito, M., Biegel, L., Brouwer, A., Brown, S., Brucker-Davis, F., Oliver, A., Christensen, R., Colborn, T., Cooke, P., Crissman, J., Crofton, K, Doerge, D., Gray, E., Hauser, P., Hurley, P., Kohn, M., Lazar, J., McMaster, S., McCLain, M., McConnell, E., Meier, C., Miller, R., Tietge, J., and Tyl, R. (1999). Screening methods for thyroid hormone disruptors. Environ. Health Perspeet. 107(5),407-415. Eldridge, J. c., Fleenor-Heyser, D. G., Extrom, P. c., Wetzel, L. T., Breckenridge, C. B., Gillis, J. H., Luempert, L. G., Ill, and Stevens, J. T. (1994). Short-term effects of chlorotriazines on estrous in female Sprague-Dawley and Fischer 344 rats. J. Toxieo!. Environ. Health 43, 155-167. Ellis, M. K, Richardson, A. G., Foster, J. R., Smith, F. M., Widdowson, P. S., Farnworth, M. J., Moore, R. B., Pitts, M. R., and Wickramaratne, G. A. (1998). The reproductive toxicity of molinate and metabolites to the male rat: Effects of testosterone and sperm morphology. Toxieo!. App!. Phannaco!. 151,22-32. Eroschenko, V. P. (1981). Estrogenic activity of the insecticide chlordecone in the reproductive tract of birds and mammals. 1. Toxieo!. Environ. Health 8, 731-742. Eroschenko, V. P., and Mouse, M. A (1979). Neonatal administration of insecticide chlordecone and its effects on the development of the reproductive tract in the female mouse. Toxieo!. App/. Phannaeo!. 49,151-159. European Commission (1996). "European Workshop on the Impact of Endocrine Disruptors on Human Health and Wildlife. Report of the Proceedings." Report EUR 17549, Directorate-General XII, Weybridge, U.K European Commission (1998). "Endocrine Disruptors Research in the EU. Report of a Meeting." Report EUR 18345, European Commission, DirectorateGeneral XIII, Brussels, 1998. European Commission (1999). "CS TEE Opinion on Human and Wildlife Health Effects of Endocrine Disrupting Chemicals, with Emphasis on Wildlife and Ecotoxicology Test Methods." European Commission, Directorate General XXIV, Brussels, 1999. Feyk, L. A., and Giesy, J. P. (1998). Xenobiotic modulation of endocrine function in birds. In "Principles and Processes for Evaluating Endocrine Disruption in Wildlife" (R. Kendall, R. Dickerson, J. Giesy, and W. Suk, eds.), pp. 121-140. SETAC Press, Pensacola, PL. Fisher, A. L., Keasling, H. H., and Schueler, F. W. (1952). Estrogenic action of some DDT analogues. Proe. Soc. Exp. Bio!. Med. 81, 439-441. Fromtling, R. A. (1988). Overview of medically important antifungal azole derivatives. Clin. Mierobio!' Rev. 1, 187-217. Gellert, R. J. (1978). Kepone, mirex, dieldrin, and aldrin: Estrogenic activity and the induction of persistent vaginal estrus and anovulation in rats following neonatal treatment. Environ. Res. 16(1-3), 131-138. Gellert, R. J., Heinrichs, W. L., and Swerdloff, R. (1974). Effects of neonatally administered DDT homologues on reproductive function in male and female rats. Neuroendocrinology 16, 84-94. Goldman, J. M., Parrish, M. B., Cooper, R. L., and McElroy, W. K (1997). Blockade of ovulation in the rat by systemic and ovarian intrabursal administration of the fungicide sodium dimethyldithiocarbamate. Reprod. Toxieol. 11, 185-190. Goldman, J. M., Stoker, T. E., Cooper, R. L., McElroy, W. K, and Hein, J. (1994). Blockade of ovulation in the rat by the fungicide sodium-Nmethyldithiocarbamate: Relationship between effects on the luteinizing surge and alterations in hypothalamic catecholamines. Neurotoxieo!' Terato/. 16,257-268. Good, E. E., Ware, G. w., and Miller, D. F. (1965). Effects of insecticides on reproduction in the laboratory mouse. I. Kepone. J. Econ. Entomo!' 58, 754757.

744

CHAPTER 34


Gray, L. E., Jr. (1982). Neonatal chlordecone exposure alters behavioral sex differentiation in female hamsters. Neurotaxiealagy 2, 67-80. Gray, L. E., and Kavlock, R. J. (1983). The effects of the herbicide 2,4(dichlorophenyl)-p-nitrophenyl ether (NIT) on serum thyroid hormones in adult female mice. Taxieol. Lett. 15,23J. Gray, L. E., and Ostby, J. S. (1995). In utero 2,3,7,8-tetrach10rodibenzo-pdioxin (TCDD) alters reproductive morphology and function in female rat offspring. Taxiea!. App!. Pharmacal. 133, 285-294. Gray, L. E., Kavlock, R J., Chernoff, N., Ferrell, J., McLamb, J., and Ostby, J. (1982). Prenatal exposure to the herbicide 2,4-(dichlorophenyl)-pnitrophenyl ether destroys the rodent Harderian gland. Science 215(4530), 293-294. Gray, L. E., Kelce, W. R, Monosson, E., Ostby, J. S., and Birnbaum, L. (1995). Exposure to TCDD during development permanently alters reproductive function in male LE rats and hamsters: Reduced ejaculated and epididymal sperm numbers and sex accessory gland weights in offspring with normal androgenic status. Taxiea!. App!. Pharmacal. 131(1), 108-118. Gray, L. E., Kelce, W. R., Weise, T., Tyl, R, Gaido, K., Cook, J., Klinefelter, G., Desaulniers, D., Wilson, E., Zacharewski, T., Wailer, c., Foster, P., Laskey, J., Reel, J., Giesy, J., Laws, S., MacLachlan, J., Breslin, W., Cooper, R, Di Giulio, R., Johnson, R, Purdy, R, Mihaich, E., Safe, S., Sonnenschein, C., Welshons, W., Miller, R., McMaster, S., and Colborn, T. (1997a). Endocrine screening methods workshop report: Detection of estrogenic and androgenic hormonal and antihormonal activity for chemicals that act via receptor or steroidogenic enzyme mechanisms. Reprod. Taxieol. 11(5),719-750. Gray, L. E., Ostby, J., Ferrell, J., and Goldman, J. (1988a). Methoxychlorinduced alterations of estrogen-dependent running wheel activity, the reproductive tract and pituitary function in the female rat. Taxieol. App!. Pharmacal. 96, 525-540. Gray, L. E., Ostby, J., Sigmon, R., Ferrell, J., Rehnberg, G., Linder, R, Cooper, R, Goldman, J., and Laskey, J. (l988b). The development of a protocol to assess reproductive effects of toxicants in the rat. Reprad. Taxiea!. 2, 281287. Gray, L. E., Jr., Ostby, J., Ferrell, J., Rehnberg, G., Linder, R, Cooper, R, Goldman, J., Slott, v., and Laskey, J. (1989). A dose-response analysis of methoxychlor-induced alterations of reproductive development and function in the rat. Fundam. Appl. Taxieal. 12(1), 92-lO8. Gray, L. E., Ostby, J. S., and Kelce, W. R (1994). Developmental effects of an environmental antiandrogen: The fungicide vinclozolin alters sex differentiation of the male rat. Taxiea!. App!. Pharmacal. 129,46-52. Gray, L. E., Ostby, J. S., and W. R., Kelce, W. R (1997b). A dose-response analysis of the reproductive effects of a single gestational dose of 2,3,7,8tetrachloro-p-dioxin (TCDD) in male Long Evans Hooded rat offspring. Taxiea!. App!. Pharmacal. 146, 11-20. Gray, L. E., Wolf, c., Mann, P., and Ostby, J. S. (1997c). In utero exposure to low doses of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) alters reproductive development of female Long Evans Hooded rat offspring. Taxiea!. App!. Pharmacal. 146, 237-244. Gray, L. E., Jr., Ostby, J., Cooper, R L., and Kelce, W. R (l999a). The estrogenic and antiandrogenic pesticide methoxychlor alters the reproductive tract and behavior without affecting pituitary size or LH and prolactin secretion in male rats. Taxiea!. Ind. Health 15(1-2), 37-47. Gray, L. E., Jr., Ostby, J., Monosson, E., and Kelce, W. R (1999b). Environmental antiandrogens: Low doses of the fungicide vinclozolin alter sexual differentiation of the male rat. Taxieol. Ind. Health 15(1-2), 48-.D4. Gray, L. E., Wolf, c., Lambright, c., Mann, P., Price, M., Cooper, R., and Ostby, J. (1999c). Administration of potentially anti-androgenic pesticides (procymidone, linuron, iprodione, chlozolinate, p,p'-DDE and ketoconazole) and toxic substances (dibutyl- and diethylhexylphthalate, PCB 169, and ethane dimethane sulphonate) during sexual differentiation produce diverse profiles of reproductive malformations in the male rat. Taxiea!. Ind. Health 15, 94-118. Griffin, J. E., and Ojeda, S. R., eds. (1988). "Textbook of Endocrine Physiology." Oxford Univ. Press, New York. Guzelian, P. S. (1982). Comparative toxicology of chlordecone (kepone) in humans and experimental animals. Ann. Rev. Pharmaca!' Taxieal. 22, 89-113.

Hadley, M. E. ed. (1996). "Endocrinology," 4th ed. Prentice Hall, Upper Saddle River, NJ. Hard, G. C. (1998). Recent developments in the investigation of thyroid regulation and thyroid carcinogenesis. Environ. Health Perspeet. 106(8),427-436. Heinrichs, W. L., Gellert, R J., Bakke, J. L., and Lawrence, N. L. (1971). DDT administered to neonatal rats induces persistent estrous syndrome. Science 147,306. Herbst, A. L., Kurman, R. J., and Scully, R E. (1972). Vaginal and cervical abnormalities after exposure to stilbestrol in utero. Obstet. Gyneeal. 40(3), 287-298. Hill, R. N., Crisp, T. M., Hurley, P. M., Rosenthal, S. L., and Singh, D. V. (1998). Risk assessment of thyroid follicular cell tumors. Enviran. Health Perspeet. 106(8),447-457. Hosokawa, S., Murakami, M., Ineyama, M., Yamada, T., Yoshitake, A., Yamada, H., and Miyamato, J. (1993). The affinity of procymidone to androgen receptor in rats and mice. J. Taxiea!. Sei. 18, 83-93. Hoyer, A. P., Grandjean, P., Jorgensen, T., Brock, J. W., and Hartvig, H. B. (1998). Organochlorine exposure and the risk of breast cancer. Lancet 352, 1816-1820. Huber, J. J. (1965). Some physiological effects of the insecticide kepone in the laboratory mouse. Taxieal. App!. Pharmacal. 7,516-524. Huff, J. E., and Gerstner, H. B. (1978). Kepone: A literature summary. J. Environ. Pathology Taxiea!. 1,377-395. Hurley, P. M. (1998). Mode of carcinogenic action of pesticides inducing thyroid follicular cell tumors in rodents. Environ. Health Perspect. 106(8), 437-445. Institute for Environment and Health (1995). "IEH Assessment on Environmental Oestrogens: Consequences to Human Health and Wildlife." Institute for Environment and Health, Leicester. International Agency for Research on Cancer (1999). "Some Agents which Targer Specific Organs in Rodent Bioassays," IARC Monograph on the Carcinogenic Risk of Chemicals to Humans, Vo!. 73. IARC Press, Lyon, France. International Programme on Chemical Safety (1998). "Report ofIPCS/OECD Scoping Meeting on Endocrine Disruptors, March 1998," IPCSIEDC2/03 World Health Organization, Geneva, Switzerland. Available from http://endocrine.ei.jrc.it. Japan Chemical Industry Association (1997). "A Study on Hormone-like (Hormone Mimic) Effects of Exogenous Substances." Japan Chemical Industry Association, Japan Chemical Industry Ecology-Toxicology and Information Center. Japan Environment Agency (1998). "Strategic Programs on Environmental Endocrine Disruptors." Jewell, W. T., and Miller, M. G. (1998). Identification of a carboxyesterase as the major protein bound by molinate. Taxiea!. App!. Pharmacal. 149, 256264. Jewell, W. T., Hess, R. A., and Miller, M. G. (1998). Testicular toxicity of molinate in the rat: Metabolic activation via sulfoxidation. Taxieol. App!. Pharmacal. 149, 159-166. Johnson, D. C. (1996). Estradiol-chlordecone (kepone) interactions: Additive effect of combinations for uterotrophic and embryo implantation functions. Taxiea!. Left. 89, 57-.D4. Joshi, S. c., Jain, G. C., and Lata, M. (1994). Effects of ketoconazole (an imidazole antifungal agent) on the fertility and reproductive function of male mice. Acta Eurapaea Fertilitatis 25(1), 55-58. Kavlock, R J. (1999). Overview of endocrine design in research activity in the United States. Chemasphere 39, 1227-1236. Kavlock, R. J., and Ankley, G. T. (1996). A perspective on the risk assessment process for endocrine-disruptive effects on wildlife and human health. Risk Anal. 16(6),731-739. Kavlock, R J., Daston, G. P., DeRosa, c., Fenner-Crisp, P., Gray, L. E., Kaattari, S., Lucier, G., Luster, M., Mac, M. J., Maczka, C., Miller, R., Moore, J., Rolland, R., Scot!, G., Sheehan, D. M., Sinks, T., and Tilson, H. A. (1996). Research needs for the risk assessment of health and environmental effects of endocrine disruptors: A report of the US EPA sponsored workshop. Environ. Health Perspeet. 104(Supp!. 4), 715-740.

References Kelce, W. R., Stone, C. R., Laws, S. c., Gray, L. E., Kempapainen, J. A., and Wilson, E. M. (1995). Persistent DDT metabolite p,p'-DDE is a potent androgen receptor antagonist. Nature 375,581-585. Kelce, W. R., Monosson, E., Gamcsik, M. P., Laws, S. c., and Gray, L. E. (1994). Environmental hormone disruptors: Evidence that vinclozolin developmental toxicity is mediated by antiandrogenic metabolites. Toxico!. App!. Pharmacol. 126,276-285. Kendall, R. J., Brouwer, A., and Giesy, J. P. (1998). A risk based field and laboratory approach to assess endocrine disruption in wildlife. In "Principles and Processes for Evaluating Endocrine Disruption in Wildlife" R. Dickerson, J. Giesy, and W. Suk, eds.), pp. 1-16. SETAC Press, Pensacola, FL. Khera, K. S., Whalen, C., and Trivett, G. (1978). Teratogenicity studies on linuron, malathion and methoxychlor. Toxico!. App!. Pharmacol. 45, 435. Korach, K. S., and McLachlan, J. A. (1985). The role of the estrogen receptor in diethylstilbestrol toxicity. Arch. Toxico!. Suppl. 8, 33-42. Krieger, N., Wolff, M. S., Hiatt, R. A., Rivera, M., Vogelman, J., and Orentreich, N. (1994). Breast cancer and serum organochlorines: A prospective study among white, black and Asian women. J. Nat!. Cancer Inst. 86,589-599. Kuiper, G. G. J. M., Lemmen, J. G., Carlsson, B., Corton, J. C., Safe, S. H., van der Saag, P. T., van der Bureg, B., and Gustafsson, J.-A. (1998). Interaction of estrogenic chemicals and phytoestrogens with estrogen receptor. Endocrinology 139,4252-4263. Kupfer, D., and Bulger, W. H. (1987). Metabolic activation of pesticides with proestrogenic activity. Fed. Proc. 46, 1864-1869. Lag, M., Soderlund, J., Brunborg, G., Dahl, J. E., Holme, J. A., Omichinski, J. G., Nelson, S., and Drybing, E. (1989). Species differences in testicular necrosis and DNA damage, distribution and metabolism of 1,2-dibromo-3chloropropane. Toxicology 58, 133-144. Lahdetie, J. (1995). Occupation- and exposure-related studies on human sperm. J. Occupational Exposure Med. 36, 922-930. Larson, P. S., Egle, J. L., Jr., Hennigar, G. R., Lane, R. W., and Borzelleca, J. F. (1979). Acute, subchronic, and chronic toxicity of chlordecone. Toxico!. App!. Pharmacol. 48,29-41. Laws, S. C., Carey, S. A., Hart, D. w., and Cooper, R. C. (1994). Lindane does not alter the estrogen receptor or the estrogen-dependent induction of progesterone receptors in sexually immature or ovariectomized adult rats. Toxicology 92, 127-142. Linder, R. E., Scotti, T. M., McElroy, K., Laskey, J. W., Strader, L. F., and Powell, K. (1983). Spermatotoxicity and tissue accumulation of chlordecone (kepone) in male rats. 1. Toxico!. Environ. Health 12,183-192. Linder, R. E., Strader, L. F., Slott, V. L., and Suarez, J. D. (1992). Endpoints of spermatotoxicity in the rat after short duration exposures to fourteen reproductive toxicants. Reprod. Toxicol. 6(6),491-505. Maness, S., McDonnell, D. P., and Gaido, K. W. (1998). Inhibition of androgen receptor---{\ependent transcriptional activity by DDT isomers and methoxychlor in HepG2 human hepatoma cells. Toxicol. App!. Pharmacol. 151, 135-142. Manson, J. M., Brown, T., and Baldwin, D. M. (1984). Teratogenicity of nitrofen (2,4-dichloro-4' -nitrophenyl ether) and its effects on thyroid function in the rat. Toxicol. Appl. Pharmacol. 73, 323-335. McFarland, L. Z., and Lacy, P. B. (1969). Physiologic and endocrinologic effects of the insecticide kepone in the Japanese quail. Toxico!. App!. Pharmacol. 15(2),441-449. McLachlan, J. A., ed. (1980). Estrogens in the Environment. Elsevier, North Holland, New York. McLachlan, J. A., ed. (1985). Estrogens in the Environment n. Influences on Development. Elsevier, New York. McLachlan, J. A., and Korach, K. S. (1995). Symposium on estrogens and the environment, Ill. Envir. Health Perspect. 103(Suppl. 7), 3-4. McLachlan, J. A., and Newbold, R. R. (1987). Estrogens and development. Environ. Health Perspect. 75, 25-27. Middleton, M. c., Milne, C. M., Moreland, S., and Hasmall, R. L. (1986). Ovulation in rats is delayed by a substituted triazole. Toxicol. App!. Pharmacol. 83,230-239. Milne, C. M., Hasmall, R. L., Russell, A., Watson, S. c., Vaughn, Z., and Middleton, M. C. (1987). Reduced estradiol production by a substituted triazole results in delayed ovulation in rats. Toxicol. Appl. Pharmacol. 90, 427-435.

745

Minor, J. L., Knapp, H. F., Stuart, B.O., Killinger, J. M., Zwicker, G. M., and Freudenthal, R. I. (1984). Evaluation of male rat fertility following inhalation exposure to ordram. Toxicologist 4, 80. Monosson, E., Kelce, W. R., Lambright, c., Ostby, J., and Gray, L. E., Jr. (1999). Peripubertal exposure to the antiandrogenic fungicide, vinclozolin, delays puberty, inhibits the development of androgen-dependent tissues, and alters androgen receptor function in the male rat. Toxicol. Ind. Health 15(1-2),65-79. Murakami, M., Hosokawa, S., Yamada, T., Harakawa, M., Ito, M., Koyama, Y., Kimura, J., Yoshitake, A., and Yamata, H. (1995). Species-specific mechanism in rat Leydig cell tumorigenesis by procymidone. Toxico!. Appl. Pharmacol. 131, 244-252. National Research Council (NRC) (1999). "Hormonally Active Agents in the Environment." National Academy Press, Washington, DC. Nelson, J. A., Struck, R. F., and James, R. (1978). Estrogenic activities of chlorinated hydrocarbons. J. Toxico!. Environ. Health 4, 325-339. Organization of Economic Cooperation and Development (1997). "Appraisal of Test Methods for Sex Hormone Disrupting Chemicals." Draft detailed review paper. Environment Directorate, OECD, Paris. Available at www.oecd.org/ehs. O'Shea, T., Reeves, R. R., and Long, A. K. eds. (1999). "Marine Contaminants and Persistent Ocean Contaminants: Proceedings of the Marine Mammal Commission Workshop," Keystone Colorado, 12-15 October 1998. Marine Mammal Commission, 4340 East-West Highway, Room 905, Bethesda, MD 20814. Ostby, J. S., Gray, L. E., Kavlock, R. J., and Ferrell, J. M. (1985). The postnatal effects of prenatal exposure to low doses of nitrofen (2,4) in SpragueDawley rats. Toxicology 34, 285-297. Ostby, J., Kelce, w., Lambright, c., Wolf, c., Mann, P., and Gray, L. E. (1999). The fungicide procymidone alters sexual differentiation in the male rat by acting as an androgen-receptor antagonist in vivo and in vitro. Toxico!. Ind. Health 15,80-93. Palmitar, R. D., and MulvihiII, E. R. (1978). Estrogenic activity of the insecticide Kepone on the chicken oviduct. Science 201, 356-358. Pont, A., WiIIiams, P. L., Loose, D. S., Feldman, D. S., Reitz, R. E., Bochra, c., and Stevens, D. A. (1982a). Ketoconazole inhibits adrenal steroid synthesis. Ann. Intern. Med. 97, 370-372. Pont, A. P., WiIIiams, P. L., Azhar, S., Reitz, R. E., Bochra, c., Smith, E. R., and Stevens, D. A. (1982b). Ketoconazole blocks testosterone synthesis. Arch. Intern. Med. 142, 2137-2140. Reiter, L. W., DeRosa, C., Kavlock, R. J., Lucier, G., Mac, M. J., Melillo, J., Melnick, R. L., Sinks, T., and Walton, B. T. (1998). The V.S. federal framework for research on endocrine disruptors and an analysis of research programs supported during fiscal year 1996. Environ. Health Perspect. 106(3), 105-113. Ronis, M. J., Barger, T. M., Gandy, J., Bell, L. M., and Green, K. (1995). Antiandrogenic effects of perinatal cypermethrin exposure in the developing rat. Neurotoxicology 16, 673. Rosencrans, J. A., Squibb, R. E., Johnson, J. H., Tilson, H. A., and Hong, J. S. (1985). Effect of neonatal chlordecone exposure on pituitary-adrenal function in adult Fischer 344 rats. Neurobehavioral Toxico!. Teratol. 7, 33-37. Schwartz, W. J., and Corkem, M. (1992). Effects of methoxychlor treatment of pregnant mice on female offspring of the treated and subsequent pregnancies. Reprod. Toxicol. 6, 431-437. Simic, B. S., Kniewald, J., and Kniewald, Z. (1994). Effect of atrazine on reproductive performance in the rat. J. App!. Toxico!. 14(6),401-404. Sonino, N. (1987). The use of ketoconazole as an inhibitor of steroid production. New England J. Med. 317, 812-818. Squibb, R. E., and Tilson, H. A. (1982). Effects of gestational and perinatal exposure to chlordecone (Kepone) on the neurobehavioral development of Fischer 344 rats. Neurotoxicology 3, 17-26. Steinmetz, R., Young, P. C. M., Caperall-Grant, A., Gize, E. A., Madhukar, B. v., Ben-Jonathan, N., and Bigsby, R. M. (1996). Novel estrogenic action of the pesticide residue ,B-hexachlorocyclohexane in human breast cancer cells. Cancer Res. 56, 5403-5409. Stevens, J. T., Breckinridge, C. B., Wetzel, L. T., Gillis, J. H., Luempert, L. G., III, and Eldridge, J. C. (1994). Hypothesis for mammary tu-

'/46

CHAPTER 34


morigenesis in Sprague-Dawley rats exposed to certain triazine herbicides. 1. Toxieo!. Environ. Health 43, 139-153. Stoker, T. E., Cooper, R. L., Goldman, J. M., and Andrews, J. E. (1996). Characterization of pregnancy outcome following thiram-induced ovulatory delay in the female rat. Neurotoxieo!' Teratol. 18,277-282. Stoker, T. E., Goldman, J. M., and Cooper, R. L. (1993). The dithiocarbamate fungicide thiram disrupts the honnonal control of ovulation in the female rat. Reprod. Toxieol. 7, 21I-218. Swartz, W. J., Eroschenko, V. P., and Schutzman, R. L. (1988). Ovulatory response of chlordecone (kepone)-exposed mice to exogenous gonadotropins. Toxicology 51,147-153. Swedish Environmental Protection Agency (1998). "Endocrine Disrupting Substances-Impainnent of Reproduction and Development" (P.-E. OIsson, B. Borg, B. Brunstrom, H. Hakansson, and E. Klasson-Wehler, eds.). Elanders Gotab, Stockholm. Tattersfield, L., Matthiessen, P., CampbeIl, P., Grandy, N., and Lange, R., eds. (1997). "Workshop on Endocrine Modulators and Wildlife: Assessment and Testing." SETAC-Europe, Brussels. Tennant, M. K., HilI, D. S., Eldridge, J. c., Wetzel, L. T., Breckenridge, C. B., and Stevens, J. T. (1994a). Chloro-s-triazine antagonism of estrogen action: Limited interaction with estrogen receptor binding. 1. Toxieo!. Environ. Health 43,197-211. Tennant, M. K., HilI, D. S., Eldridge, J. c., Wetzel, L. T., Breckenridge, C. B., and Stevens, J. T. (l994b). Possible anti-estrogenic properties of chloro-striazines in rat uterus. 1. Toxieo!. Environ. Health 43, 183-196. Thakur, A. K., Wetzel, L. T., VoeIker, R. w., and Wakefield, A. E. (1998). Results of a two-year oncogenicity study in the Fischer 344 rats with atrazine. In "Triazine Herbicides Risk Assessment." (L. G. Ballentine, J. E. McFarland, and D. S. Hackett, eds.), ACS Symposium Series, Vo!. 683, pp. 384398. Am. Chem. Soc., Washington, DC. Toppari J., Larsen, J. c., Christiansen, P., Giwercman, A., Grandjean, P., Guiliette, L. J., Jr., Jegou, B., Jensen, T. K., Jouannet, P., Keiding, N., Leffers, H., McLachlan, J. A., Meyer, 0., Muller, J., Rajpert-De Meyts, E., Scheike, T., Sharpe, R., Sumpter, J., Skakkebaek, N. E. (1996). Male reproductive health and environmental xenoestrogens. Environ. Health Perspeet. 104(Supp!. 4), 741-803.

Umweltbundesamt (1995). "Endocrinically Active Chemicals in the Environment," Texte Series 3, Vo!. 36, Expert Round, 9-10 March 1995, Berlin, Gennany. Uphouse, L. (I 985a). Effects of chlordecone on neuroendocrine function of female rats. Neurotoxieology 6, 191-210. Uphouse, L. (1985b). Single injection with chlordecone reduced behavioral receptivity and fertility of adult rats. Neurobehavioral Toxieol. Teratol. 8, 121-126. Uphouse, L., and WilIiams, J. (1989). Diestrous treatment with lindane disrupts the female reproductive cycle. Toxieo!. Lett. 48, 21. U.S. Environmental Protection Agency (U.S. EPA) (l997a). "Special Report on Environmental Endocrine Disruption: An Effects Assessment and Analysis." EPAl6301R-012, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (l997b). Toxic Substances Control Act test guidelines; final rule, Federal Register 62(158), 4381943861 (August 15). U.S. Environmental Protection Agency (US. EPA) (1998). "Endocrine Disruptor Screening and Testing Advisory Committee (EDSTAC) Final Report," Vo!. 1. U.S. Environmental Protection Agency (U.S. EPA) (2000). Endocrine Disruptor Screening Program-Report to Congress, August 2000. WaIler, D. P., Martin, A., Vickery, B. H., and Zaneveld, L. J. (1990). The effect ofketoconazole on fertility of male rats. Contraception 41(4), 41I-417. Warren, D. w., Ahmad, N., and Rudeen, P. K. (1988). The effects of fetal exposure to 1,2-dibromo-3-chloropropane on adult male reproductive function. Bio!. Reprod. 39, 707-716. Wickramaratne, G. A., Foster, J. R., Ellis, M. K., and Tomenson, J. A. (1998). Molinate: Rodent reproductive toxicity and its relevance to humans-a review. Regul. Toxieo!. Pharmaeol. 27, 1l2-1I8. You, L., Casanova, M., Archibeque-Engle, S., Sar, M., Fan, L. Q., and Heck, H. A. (1998). Impaired male sexual development in perinatal Sprague-Dawley and Long-Evans hooded rats exposed in utero and lactationally to p,p'-DDE. Toxieol. Sci. 45(2),162-173.

CHAPTER

35 Genetic Toxicity of Pesticides David A. Eastmond and Sharada Balakrishnan University of California, Riverside

35.1 INTRODUCTION Pesticides are biologically active compounds selected and used for their toxic properties. In many cases, these agents are highly specific in their pesticidal effects, acting on a unique molecular target or affecting a narrow range of organisms. Other pesticides can affect a much broader range of targets and organisms, including humans. As a result, there exist ongoing concerns about the health effects of pesticide exposure in humans. These concerns have been heightened by pesticide-related poisoning episodes that have occurred over the past 50 years such as those involving hexachlorobenzene (Schmid, 1960), methylmercury (Bakir et aI., 1973), malathion (Baker et aI., 1978), dibromochloropropane (Whorton et aI., 1979), aldicarb (Green et aI., 1987), and methylparathion (Rehner et aI., 2000). In addition to acute effects, substantial concerns exist about chronic effects such as cancer and heritable diseases that might stem from pesticide exposure. An association between pesticide exposure and cancer has been suspected for more than 40 years following reports of the occurrence of elevated levels of skin and lung cancer in European farmers using arsenical insecticides in grape production (Jungmann, 1966; Roth, 1958; Thiers et aI., 1967). In a few cases, the association between pesticide exposure and cancer has been confirmed (Blair and Zahm, 1995; IARC, 1987a, 1994; IOM, 1999; Zahm et al., 1997). However, in many cases, these concerns remain unsubstantiated either due to an underlying lack of an association or because of the difficulties in conducting epidemiological studies in these exposed populations. Even where associations have been seen or suspected, identifying the specific agent responsible has been difficult for many reasons, including poorly defined and variable exposure levels, exposure to multiple pesticides as well as other potentially carcinogenic agents, long latency periods, small study populations, and other confounding factors. As additional limitations, human epidemiological studies are costly and can only take place following exposure-an approach that is not considered protective of public health. Because of these difficulties, regulatory agencies and other organizations have turned to chronic animal bioassays, short-term tests, and other relevant data, in addition to human epidemiological studies, to evaluate the carcinogenicity and potential carcinogenicity of Handbook of Pesticide Toxicology Volume 1. Principles

pesticides and other agents. These agencies, such as the International Agency for Research on Cancer (IARC),* the D.S. Environmental Protection Agency (EPA), and the National Toxicology Program (NTP) have adopted a weight-of-the-evidence approach to make decisions on the carcinogenicity of an agent. For example, after reviewing the human, animal, and relevant biological data for one class of pesticides, IARC concluded that the spraying and application of nonarsenical insecticides entail exposures that are probably carcinogenic to humans (IARC, 1991). To date, a relatively small number of pesticides ( < 10) have been recognized by one or more of these organizations as human carcinogens (Goldman, 1998; IARC, 2000; NTP, 2000). It should be noted that, in these cases, the primary evidence has come, not from agricultural uses, but from studies of exposed workers manufacturing the agent for other industrial uses or, in the case of inorganic arsenic, from therapeutic and industrial uses as well as environmental exposures (IARC, 1987a). For instance, most of the evidence for the carcinogenicity of agents such as inorganic arsenic, benzene, cadmium, and chromium(VI), which historically were used as pesticides,t as well as agents currently registered for use such as ethylene oxide and coal tar, have been obtained from studies involving nonagricultural uses. In some cases, it is believed that the agent responsible for the toxic effects seen in the pesticideexposed individuals is a contaminant or an "inert" ingredient in the pesticide formulation rather than the active ingredient itself. For example, many of the adverse effects proposed as *The International Agency for Research on Cancer, part of the World Health Organization, produces authoritative evaluations of the carcinogenic risks of chemicals and other agents to humans. Following a critical review of both human and animal studies, IARC classifies agents as exhibiting sufficient evidence of carcinogenicity, limited evidence of carcinogenicity, inadequate evidence of carcinogenicity, or evidence suggesting a lack of carcinogenicity in animals or humans. As a final step, the IARC considers the entire body of evidence, including mechanistic information, to reach an overall evaluation of the carcinogenicity of the agent to humans. Other regulatory agencies such as the U.S. Environmental Protection Agency and the U.S. National Toxicology Program use similar approaches to evaluate the carcinogenicity of chemical agents. t Some forms of arsenic are still registered in the United States for use under severely restricted conditions. Arsenicals continue to be used as insecticides and wood preservatives in other countries (Zahm et aI., 1997).

747


748

CHAPTER 35

Genetic Toxicity of Pesticides

being associated with the chlorophenoxyacetic acid herbicides are believed to be due to contamination by low levels of2,3,7,8tetrachlorodibenzo-para-dioxin (TCDD), a potent animal and human carcinogen (IARC, 1997; IOM, 1999, 2000). Furthermore, it is conceivable that other cancers such as the leukemias and non-Hodgkin's lymphomas that have been attributed to pesticide exposure may in part be due to the use of benzene and other solvents as ingredients in the formulation products (Blair and Zahm, 1995; Petrelli et ai., 1993). As indicated previously, chronic testing in animals is also used by regulatory agencies to evaluate the carcinogenic effects of chemical agents. Animal bioassays have been conducted for a considerable number of individual pesticides and a significant number have been reported as tumorigenic in one or more animal tissues. According to Zahm and Ward (1998), of the 51 pesticides evaluated prior to 1990 by the U.S. National Cancer Institute and the National Toxicology Program, 24 exhibited carcinogenicity in chronic animal bioassays. These authors further reported that, as of 1997, the IARC had classified 26 pesticides as having sufficient evidence of carcinogenicity in animals and 19 as having limited evidence. However, because of their cost, lengthy duration, and concern that the results may not be directly relevant to humans, these bioassay results are considered as less than ideal and are usually evaluated in conjunction with additional types of biological information. In addition to human and chronic animal studies, regulatory agencies often rely on other relevant biological data to assist in the evaluation of carcinogenicity. These other data may include information on preneoplastic lesions, tumor pathology, genetic and related effects, structure-activity relationships, metabolism and pharmacokinetics, physicochemical parameters, and mechanisms of action (IARC, 199ge). In particular, short-term tests evaluating the genetic toxicity of the agent are often relied on in the decision-making process. The development and interpretation of these short-term tests have stimulated the development of the field of genetic toxicology. As a subspecialty of toxicology, genetic toxicology is concerned with the adverse effects of chemicals and other physical agents on the deoxyribonucleic acid (DNA) and other genetic components of living organisms. The primary focus of this discipline is to identify the agents and mechanisms involved in the formation of mutations-heritable genetic alterations in cells. When broadly defined, mutagenesis encompasses the induction of DNA damage as well as all types of genetic alterations, ranging from a single nucleotide change in the DNA sequence to large-scale changes in chromosome structure and number. The recognition that cancer is fundamentally a genetic disease, combined with the close association that has been seen between mutagenicity and carcinogenicity, has led to the use of mutation and genotoxicity assays as screens to identify agents likely to be carcinogenic or cause other genetic diseases. Over the past 30 years, a large number of short-term tests have been developed as screening tools to identify genotoxic and mutagenic chemicals. These short-term tests may employ bacteria, yeast, plants, insects, isolated mammalian cells, or whole animals and can be performed for a fraction of the cost and time required for a long-

term cancer bioassay. In addition, a number of these assays have been modified for use in detecting genetic alterations occurring in human populations exposed to genotoxic and carcinogenic agents. The objective of this chapter is to provide an overview of the methods of genotoxicity testing and their application to identifying pesticides capable of inducing genetic damage. The initial section will focus on the most common short-term tests that are employed for detecting the genotoxicity of pesticides in model systems and the use of these assays to detect genetic alterations in exposed humans. This will then be followed by an overview of the results of genotoxicity studies that have been performed on individual agents and studies of genetic damage in pesticideexposed workers. The last section will briefly address the value and interpretation of this information in the safety evaluation and risk assessment process.

35.2 GENOTOXICITY TESTS Hundreds of short-term tests have been developed to screen chemicals for potential mutagenic and carcinogenic effects. These assays measure effects ranging from DNA adduct formation to mutations induced in transgenic animals. A listing of representative short-term tests as well as a brief description of how these effects are measured is presented in Tables 35.1 and 35.2. Each of these genotoxicity assays has its own unique characteristics and measures only a subset of the possible heritable alterations involved in cancer and other genetic diseases. As a result, combinations of short-term tests are often used to increase the likelihood of detecting genotoxic effects. Over the years, requirements for genotoxicity testing have been established in the United States and other developed nations, and agents being proposed as new pesticides must undergo testing prior to registration. Using the EPA requirements as an example, the initial test battery includes (1) a gene mutation assay in bacteria, typically the Salmonella typhimurium reverse mutation assay; (2) one of several gene-inactivating (forward) mutation assays using mammalian cells in culture; and (3) an in vivo assay for chromosomal effects in mammalian bone marrow cells using either metaphase analysis for structural aberrations or the micronucleus assay (Auletta et aI., 1993; Dearfield et aI., 1991). Depending on the results of the initial battery, as well as other relevant information, additional shortterm testing or a chronic animal bioassay may be required. As a general principle, agencies such as the EPA place greater weight on tests conducted in eukaryotes than in prokaryotes and in mammalian species rather than in submammalian species when conducting a hazard evaluation of a chemical (Auletta et aI., 1993). For heritable noncancer risks, the results from studies in germ cells are accorded more weight than those obtained using somatic cells. Because of their prominent role in the testing of pesticides, the principal required assays will be described in more detail. For more detailed reviews of these and other short-term genotoxicity tests, the reader is referred to more

35.2 Genotoxicity Tests

749

Table 35.1 Representative Short-Term Tests for Genotoxicity

Type of test

Specific test

DNA adduct formation

Covalent binding of radiolabeled chemicals 32p-postlabeling of adducts Measurement of oxygen radical-derived adducts Immunological detection of adducts

DNA damage in microorganisms

Pol A test rec test Mitotic recombination, mitotic crossing over, or mitotic gene conversion in yeast (D3, D4, D5, or D7 assays)

DNA damage in mammalian cells

Unscheduled DNA synthesis (UDS) Single-cell gel electrophoresis (Comet) assay Sister chromatid exchange (SCE)

Gene mutation in bacteria and fungi

Salmonella microsome reversion assay (Ames test) WP2 assay Yeast "forward" and "reverse" assays Miscellaneous

Gene mutation in higher systems

HPRT, TK, and Na/K-ATPase assays in vitro Sex-linked recessive lethal assay Tradescantia or maize waxy locus plant tests HPRT assay in vivo Mutation in lac IlIac Z-bearing transgenic animals

Chromosomal effects in isolated cell systems

In vitro cytogenetics assays In vitro micronucleus test Aneuploidy assays

Chromosomal effects in whole organisms

In vivo cytogenetics Micronucleus test Nondisjunction assay Heritable translocation assay Dominant lethal assay Alterations in germ cells

Oncogenic transformation

Transformation assays (clonal or focus)

Modified from U.S. EPA (1979).

comprehensive sources (IARC, 1980; IPCS, 1985; Rice et aI., 1999b). 35.2.1 BACTERIAL REVERSE MUTATION ASSAY

The bacterial reverse mutation assay uses specially engineered amino acid-requiring strains of Salmonella typhimurium (S. typhimurium) or, less frequently, Escherichia coli (E. coli) to detect point mutations, which involve the substitution, deletion, or insertion of one or a few DNA base pairs (IPCS, 1985; D.S. EPA, 1998c). The widely used Salmonella assay was developed by Ames, McCann, and Yamasaki and is commonly known as the Ames test (Ames et aI., 1975; McCann and Ames, 1976). The basis for the assay is as follows: Following exposure

to a mutagenic chemical, mutations are detected that reverse existing gene-inactivating mutations present in the Salmonella test strains, thereby restoring the ability of the bacteria to synthesize the essential amino acid. The bacteria carrying the reverse mutations (called revertants) are detected by their ability to grow in the absence of the amino acid required by the parental test strain. Many of the test strains have also been engineered to increase the sensitivity of the assay. These enhancements include a modification of the cell wall to be more permeable to lipophilic chemicals, inactivation of a gene involved in DNA excision repair, and addition of another gene coding for an error-prone DNA repair gene. In addition, the assay is conducted in the presence and absence of a mammalian metabolic system to increase the sensitivity of the assay to chemicals requiring metabolic activation for genotoxicity. Most

750

CHAPTER 35


Table 35.2 The Measurement of Genotoxic Effects in Short-Term Tests DNA binding (32 P-postIabeIing, 8-0H-dG, and others)

DNA damage in bacteria (Pol A test, rec test)

The covalent binding of a chemical to DNA is used as a measure of its

Two strains of bacteria are used that are identical except in their

reactivity and potential for genotoxicity. The DNA adducts can be

ability to repair DNA damage; one strain can repair damage

detected and quantitated by using a radiolabeled chemical,

whereas the other cannot. Both strains are exposed to the test

by labeling the adducted nucleotide after formation, through

substance, and the extent to which cells are killed is measured

an analysis for specific adducts, or by immunological techniques.

for each. If the repair-deficient strain has a greater degree of cell killing, DNA damage is assumed to have occurred.

DNA damage in yeast (mitotic recombination, mitotic gene conversion, or

Chromosomal effects in isolated cells or whole organisms (cytogenetic

mitotic crossing over)

assays or in vitro micronucleus assay)

Special strains of yeast cells are used to test for these effects.

Treated cells (or cells from treated organisms) are stained and

When the cells change color from white to either pink or red, DNA

then examined under the microscope for various chromosomal

damaging potential is indicated.

abnormalities. Lost, broken, or misarranged chromosomes or the formation of micronuclei indicate genotoxicity

Gene mutation in bacteria or fungi (Ames test, WP2 assay, yeast assays, and others)

Oncogenic transformation (transformation assays) When certain types of mammalian cells are treated in vitro

Special strains of bacteria are used that cannot grow without

with carcinogens, they undergo cancer-like transformation.

a nutritional supplement. Certain types of mutations will permit

If these cells are injected into appropriate experimental animals,

these bacteria to grow in unsupplemented media. By treating the

tumors will appear. Most frequently, transformed cells are

cells and then seeing if they can grow in unsupplemented media,

distinguished by their unusual growth patterns in culture,

mutagenicity can be measured. Distinguishing mutated bacteria from

such as abnormal piling-up and disorientation of cells.

nonmutated bacteria is not necessary using this procedure, because only mutant cells can grow and form visible colonies. DNA damage in mammalian cells (unscheduled DNA synthesis, sister chromatid exchange, and single-cell gel electrophoresis 'Comet' assay)

Micronucleus test in vivo Animals are treated with a chemical, and their red blood cells are removed, stained, and examined under the microscope. If small

Abnormal distribution of a dna marker indicates whether DNA damage

fragments of the genetic material (micronuclei) are observed,

has occurred. Microscopic examination, photographic measurements,

chromosomal damage is indicated. Normal red blood

and computerized image analysis systems are used to detect the DNA

cells will not contain any genetic material or fragments of

damage.

genetic material.

Gene mutation in mammalian cells or plants Mammalian cells (HPRT, TK, and Na/K-ATPase assays) In these

Drosophila melanogaster (sex-linked recessive lethal test for mutations: nondisjunction and heritable translocation assays for chromosomal effects)

systems, mutations that confer resistance to a poison are measured. Cells

Drosophila have a variety of "marker" traits that can be used

are first treated with a test chemical and then exposed to the poison.

to signal whether gene mutations or chromosome disturbances

Because only mutant cells can survive and grow, mutagenicity can be

have occurred. Specially "marked" male or female flies are treated

measured simply by observing the extent of growth in the poisonous

with a substance, mated, and then their offspring are observed to see

Plant cells: environment. (Tradescantia and maize waxy locus)

if they have certain specific features, such as unusual eye color or shape.

Mutations in these plants are detected by looking for color changes in the stamen hairs (Tradescantia) or pollen grains (maize). Modified from U.S. EPA (1979).

commonly, the metabolic system is a cofactor-supplemented post-mitochondrial fraction (S9) prepared from the livers of rats treated with enzyme-inducing agents such as Arochlor 1254. By using these specially engineered and rapidly growing bacteria, chemically induced mutations occurring at low frequencies (1 x 10-6 ) in tens to hundreds of millions of bacteria can be rapidly and inexpensively detected. However, the targets for reverse mutations in the test strains are very small in relationship to the bacterial genome and only a narrow range of point mutations can be detected. To increase the range of point mutations that can be detected, regula-

tory guidelines recommend that five different test strains of bacteria be used (U.S. EPA, 1998c). This assay has been shown to be moderately to highly efficient at predicting carcinogenicity with predictive values generally ranging from 0.5 to 0.8, depending on the study and the characteristics of the chemicals being tested (Brusick, 1987). However, due to fundamental differences between prokaryotic and eukaryotic organisms, these assays are not able to detect mutations induced by some types of chemical agents, for example, topoisomerase inhibitors and nucleoside analogs (U.S. EPA, 1998c).

35.3 Genotoxicity Testing of Pesticides

35.2.2 IN VITRO MUTATION ASSAY IN MAMMALIAN CELLS The in vitro mammalian cell gene mutation assay can be used to detect gene-inactivating mutations induced by chemicals (U.S. EPA, 1998d). Mouse, Chinese hamster, or human cell lines are exposed to the test chemical and mutations occurring in endogenous genes such as thymidine kinase (TK), hypoxanthineguanine phosphoribosyl transferase (HPRT), and a transgene of xanthine-guanine phosphoribosyl transferase (XPRT) are measured. Using the assay for mutations in thymidine kinase as an example, cells with a mutation converting the TK heterozygote (T K+I-) to cells lacking a functional TK allele (T K-I-) are resistant to the cytotoxic effects of the nucleotide analog, trifluorothymidine (TFr). Thymidine-proficient cells are sensitive to TFr, which inhibits cellular metabolism and halts cell division. As a result, mutant cells are able to proliferate in the presence of TFr, whereas normal cells that contain the functional TK allele are unable to grow. Similarly, cells deficient in HPRT or XPRT are selected based on their resistance to 6thioguanine or 8-azaguanine, respectively. In these assays, cells are exposed to the test chemical both in the presence and in the absence of metabolic activation for a suitable period of time, then subcultured to allow phenotypic expression prior to mutant selection with the toxic nucleotide analog. Mutant frequency is then determined, after an appropriate incubation period, by seeding known numbers of cells in a medium containing the selection agent to detect mutant cells and in a medium without the selection agent to determine the cloning efficiency. The principal advantage of this assay is that it allows rare mutations occurring in mammalian cells to be detected simply and relatively inexpensively. Moreover, because this assay measures gene-inactivating mutations in eukaryotic cells, it is capable of detecting a much broader range of mutagenic events (i.e., large deletions, recombination, etc.) than a bacterial mutation assay.

35.2.3 IN VIVO CYTOGENETIC ASSAY The in vivo chromosome aberration assay is used for the detection of structural chromosome aberrations induced by test chemicals in the bone marrow of mammals, typically rodents (U.S. EPA, 1998e). In this assay, animals are administered the test substance by an appropriate route of exposure and are sacrificed at selected times (typically 12-36 h) after treatment. Prior to sacrifice, the animals are treated with a spindledisrupting agent to arrest rapidly dividing bone marrow cells in the metaphase stage of the cell cycle. Chromosome preparations are made from the bone marrow cells and, following staining, the metaphase cells are analyzed for structural damage to the chromosomes. Some information on changes in chromosome number (aneuploidy and polyploidy) can also be obtained. Because a chromosome break can occur within most, if not all, DNA sequences throughout the genome, this assay is believed to be highly sensitive at detecting agents inducing doublestranded breaks in DNA. In previous studies, it has been shown

751

that, when tested, most human cancer-causing agents induce increased levels of chromosome aberrations in the bone marrow of rodents (Ashby and Paton, 1993). Moreover, this assay is thought to be particularly valuable in that chromosomal alterations are the underlying cause of many genetic diseases and play an important role in carcinogenesis. The in vivo aberration assay is considered particularly useful for assessing mutagenic hazards in that it allows normal in vivo metabolism, toxicokinetics (absorption, distribution, and excretion), and DNA repair processes to occur (Auletta et aI., 1993). 35.2.4 MICRONUCLEUS ASSAY The micronucleus assay is similar to the in vivo aberration assay in that both measure chromosome alterations in treated mammals and, according to most regulatory guidelines, either can be used in the initial testing (Auletta et aI., 1993; Dearfield et aI., 1991). The micronucleus assay detects chromosome breakage and loss occurring following chemical treatment. Although micronuclei can be formed in any dividing tissue of any species following treatment, for regulatory purposes the assay is almost always conducted in the bone marrow or, less frequently, the peripheral blood erythrocytes of rodents (U.S. EPA, 1998f). As a bone marrow erythroblast develops into a newly formed ribonucleic acid (RNA)-containing (polychromatic) erythrocyte, the main nucleus is extruded. In a damaged cell, the micronucleus that has been formed remains behind in the anucleate cytoplasm. Using a stain such as acridine orange that differentially stains RNA and DNA, the DNA-containing micronucleus can easily be visualized in the cytoplasm of the newly formed RNA-containing erythrocytes. An increase in the frequency of micronuclei following treatment with a test chemical indicates that an increase in chromosome damage has occurred. The assay can be performed in one of two ways: with a single dose followed by two or more sampling times or with two or more sequential doses followed by a single harvest. As with the in vivo aberration assay, this in vivo assay allows normal metabolism, toxicokinetics, and DNA repair to occur. In addition, many human and animal carcinogens when tested have shown positive results in this assay (Ashby and Paton, 1993).

35.3 GENOTOXICITY TESTING OF PESTICIDES As indicated previously, genotoxicity testing is required for the registration of new pesticides in the United States and most developed nations. Testing has also been performed for many of the pesticides that were registered prior to the current testing requirements. It should be noted, however, that often the results of these tests are considered proprietary and are not published in the public domain. Published genotoxicity test results for many pesticides and other agents evaluated by the EPA and IARC are available in both graphical and tabular forms in the Genetic Activity Profile (GAP)

752

CHAPTER 35


database (Waters et aI., 1991, 1999). This program can be downloaded without charge at www.epa.gov/gapdb. Two valuable sources of the summary results of unpublished tests on pesticides are the toxicological summaries compiled by the California Department of Pesticide Regulation (available at www.cdpr.ca.gov/docs/toxsums/toxsumlist.htm) and the toxicological evaluations performed as part of the joint meeting of the Food and Agricultural Organization panel of experts on pesticide residues in food and the environment [see FAOIWHO (1999) for a recent example]. A representative listing of specific pesticides, along with their activity in various genotoxicity tests and evaluations for carcinogenicity, is shown in Table 35.3. As is evident from the table, a variety of patterns of responses can be seen. Some agents are clearly genotoxic and carcinogenic, whereas others have shown activity in the genotoxicity assays without showing an increase in tumors in the cancer bioassays. Other pesticides have primarily exhibited negative results in short-term genotoxicity assays but have shown increases in tumors in chronic animal testing. Other agents have demonstrated no genotoxic or carcinogenic effects in in vitro or in vivo studies. Finally, many agents have given mixed or equivocal responses in genotoxicity or carcinogenicity tests. Interpretation of this latter pattern of responses is particularly challenging due to the likelihood of false positive results when many short-term assays are conducted or assays are performed under conditions (high concentrations, increased osmolality, pH, oxygen tension, etc.) that may differ significantly from those likely to be encountered in vivo. In addition, the pathological evaluation of many different tissues and organs also increases the likelihood of false positives in a chronic animal cancer bioassay. For illustration, examples of each of the preceding patterns of response will be presented.

35.4 PATTERNS OF RESPONSE 35.4.1 PESTICIDES EXHIBITING BOTH GENOTOXICITY AND CARCINOGENICITY 35.4.1.1 Ethylene Oxide

Ethy lene oxide, or epoxy ethane, is an insecticidal fumigant used for stored food products, bedding, carpets, and clothing (Gehring et aI., 1991). It is also used to sterilize heat-sensitive medical devices and as an intermediate in the synthesis of other chemicals, particularly ethylene glycol (Dellarco et aI., 1990). Structurally, it is a reactive chemical that exerts its cytotoxic effects by alkylating a broad range of critical cellular macromolecules such as DNA and proteins (Dellarco et aI., 1990; Gehring et aI., 1991). Given its ability to alkylate DNA, it is not surprising that it exhibits genotoxic effects in most genotoxicity assays. Ethylene oxide has been shown to be mutagenic in bacterial and mammalian cells,

to increase chromosome aberrations and micronuclei in the bone marrow of rodents, and to exhibit positive responses in a series of other genotoxicity assays (Dellarco et aI., 1990; IARC, 1994). In reviewing the evidence, the IARC has concluded that ethylene oxide is both an animal and a human carcinogen (IARC, 1994). In addition to affecting somatic cells, ethylene oxide is also an established germ cell mutagen, which has been shown to induce dominant lethal mutations and translocations in rodents (Dellarco et aI., 1990). Ethylene oxide is one of the few agents for which heritable risks to humans have been evaluated (Rhomberg et aI., 1990). 35.4.1.2 Ethylene Dibromide

Ethylene dibromide, EDB or 1,2-dibromoethane, has been used as a fumigant for stored grain, fruits, and vegetables (Gehring et al., 1991). It has also been used as a soil treatment for nematodes and as a scavenger in tetraethyl lead-containing gasoline. Ethylene dibromide is metabolic ally activated through both microsomal- and glutathione transferase-dependent pathways to form reactive DNA and protein-binding metabolites (Gehring et aI., 1991). EDB has been shown to be mutagenic in bacteria and mammalian cells, to bind to DNA, to induce DNA strand breakage, and to increase unscheduled DNA synthesis (U.S. EPA, 1997). Although genotoxic in the majority of in vitro tests and in vivo assays for DNA breakage, EDB has shown largely negative results in in vivo assays of chromosome damage and dominant lethal mutations (IARC, 1999b). These somewhat differing results may reflect the target organ specificity, as well as the types of DNA damage induced by this agent. EBD has been shown to exhibit carcinogenic effects in multiple animal species. However, the evidence for carcinogenic effects in humans is considered inadequate (IARC, 1999b; U.S. EPA, 1997). Both the IARC and the EPA consider EDB to be a probable human carcinogen. Similar patterns can be seen for other pesticides or "inert" ingredients such as chromium(VI), arsenic, formaldehyde, benzene, and creosote. In each of these cases, the agent is carcinogenic in either animals and/or humans and is positive in most genetic toxicity assays. t The lack of activity in a few short-term tests suggests that the agent acts through a specific genotoxic mechanism, that target organspecific effects or metabolism may be occurring, or that the genotoxicity result is in error (i.e., a false negative). Based on the strongly positive results observed, most of these types of agents have been banned for use as pesticides or are registered for use under highly restricted conditions. +Benzene and arsenic have consistently exhibited negative results in gene mutation assays but have been positive for the induction of chromosomal alterations in vivo. The critical genetic alterations in the carcinogenicity of these agents appear to be chromosomal in nature.

35.4 Patterns of Response

753

Table 35.3 Short·Term Genotoxicity Results and Evaluation of Carcinogenic Risk for Selected Pesticides Chromosomal aberrations/micronuclei

Mutation

Pesticide

Salmonella

Mammalian

(Ames test)

cells

in vitro

in vivo

lARC classification Human

Animal

carcinogenicity

carcinogenicity

Inorganic metals

+a +a, _lh

+la. +Ib

SE

LE

+

eb, ea

SE

SE

+

+a, +ah

+b, +a

SE

SE

_a, eb

NR

NR

ab

ND

lE

+Iah

+1'

ND

lE

Arsenic compounds Cadmium chloride Chromium(Vl)

+

compounds Carbamates

Propoxur* +a, +lh

Carbaryl Aldicarb

+, +lh

Chlorinated hydrocarbon insecticides

0

0

lE

SE

-1

0 ea,_lah

0 +a, (+)ah

lE

SE

lE

SE

0

+lah

-lb, +Ia

lE

LE

NR

NR

(+)

+a

0

lE

lE

0 +lah

+a, _ b

ND

lE

ND

lE

0

+a, +lb _la

NR

NR

0

-la

ND

lE

+a, -lah

lE

SE

0

lE

ND

ND

ESL

Chlordane Heptachlor DDT

Aldrin Endosulfan* Endrin Pyrethroids

Deltamethrin Fenvalerate

0 -I

Cypermethrin Permethrin Organophosphate insecticides

Dichlorvos

+

+1

Parathion Methyl parathion Malathion

0

Diazinon

e

Chlorpyrifos* lsazofos

+

+ -I

-I

-I

0

0

0

+a, +bh, +Iah, +Ib ea, -Iah

e _a, -lb b

ND

lE

NR

NR

NR

NR

NR

NR

ND

SE

Fungicides

Captafol Pentachlorophenol Thiram

+

e

Ziram

+

0

ortho-Phenylphenol*

+1

Chlorothalonil Hexachlorobenzene*

0 +

+1

lE

SE

lE

lE

+Ia, +Ib, +Iah

ND

LE

-, +Ib a

lE

LE

lE

SE

0

lE

SE

a

ND

SE

+Ia, +lah

+Ib

lE

SE

_Ib

lE

SE

-la _Iah

+b, +Ia

ND

LE

-1' _Ib

ND

LE

lE

LE

+a

lE

LE

-1', _Iah, _Ib, +Ibh _a, + Ib, _Ibh

1,4-Dichlorobenzene* Propylene oxide

+1' +1',+lah ea ea _ah

0 +Iah +b, +Ia

Herbicides

Atrazine

-I

Monuron

e

Picloram

0 -I

Simazine Trifluralin

0 eah

(continues)

754

CHAPTER 35 Genetic Toxicity of Pesticides

Table 35.3 (continued) Mutation

Pesticide

Salmonella

Mammalian

(Ames test)

cells

MCPA 2,4-D

0

Chromosomal aberrations/micronuclei in vitro

in vivo

rARC classification Human

Animal

carcinogenicity

carcinogenicity

+1 ab

-1', _b,_ ah ab b

lE

ND

LE

lE lE

+1

0

0

lE

Bentazon*

-I

0

-lb

NR

NR

2,4,5-T

0

0

-lb,+1'

LE

lE

Amitrole

e

-1'

lE

SE

+

+a, +Ib, _lab

_a, _lb

lE

SE

+Ih

LE

SE

lE

SE

Methyl chloride

+

Fumigants and nematocides Acrylonitrile

+

Ethylene oxide*

+

+

Ethylene dibromide*

+

+

+a, +Ib

+a, +b -la, _b

+

+a, +ah

_ab, _a, _b, +bhf

LE

SE

+ e

ea

+b, -1"

lE

LE

+a

SE

0

-lab, eb

+b, +1"' - a -

lE

Carbon tetrachloride

lE

SE

Tetrachloroethylene

0

0

-lab

lE

LE

1,3 Dichloropropene

e

lE

SE

Formaldehyde

+

Methyl bromide*

+

DBCP

+

+

+

Solvents and others Xylene

-I", -lab

lE

lE

Benzene

+a

+a, +b

SE

SE

Piperonyl butoxide

0

0

ND

lE

a, chromosomal aberrations; b, micronucleus; h, human cells; s, spermatogonia; (+), weakly positive; e, equivocal/inconclusive; 0, no test results were located; + I, positive in one study; -I, negative in one study; +, positive in more than one study or the majority of studies; -, negative in more than one study or the majority of studies; f, micronucleus formation was positive in buccal mucosal cells in humans whereas it was negative in peripheral blood Iymphocytes, possibly due to the high reactivity of formaldehyde at the primary site of exposure; lE, inadequate evidence for carcinogenicity; LE, limited evidence for carcinogenicity; SE, sufficient evidence for carcinogenicity; ESL, evidence suggesting lack of carcinogenicity; ND, no adequate data were available; NR, not reviewed by the rARe. *See text for additional details. This table was compiled primarily from five sources: (1) the rARC Monographs on the Evaluation of Carcinogenic Risks in Humans; (2) the Environmental Health Criteria series published by the International Programme on Chemical Safety; (3) the toxicological evaluations performed as part of the Joint meeting of the FAO panel of experts on pesticide residues in food and the environment; (4) the Genetic Activity Profile (GAP) database generated jointly by the EPA and rARC; and (5) the toxicological data review summaries prepared by the California Department of Pesticide Regulation.

35.4.2 PESTICIDES EXHIBITING GENOTOXICITY WITH LIMITED OR NO EVIDENCE OF CARCINOGENICITY 35.4.2.1 Methyl Bromide Methyl bromide, or bromomethane, has been widely used as a fumigant for control of insects, nematodes, fungi, and weeds (Gehring et aI., 1991; IPCS, 1985). Although methyl bromide has been shown to react with both DNA and proteins, its mechanism for toxicity remains to be elucidated (IPCS, 1995). Methyl bromide has been shown to be genotoxic in most short-term genotoxicity tests (IARC, 1999c; IPCS, 1995): It induced mutations in bacteria and mammalian cells, increased the incidence of micronuclei in vivo in mouse and rat bone marrow erythro-

cytes, and was shown to bind covalently to the DNA in several rat and mouse organs. In contrast, methyl bromide has produced mixed, largely negative responses in chronic animal bioassays. In a short 13week study in which methyl bromide was administered by oral gavage, it was reported to produce squamous cell carcinomas of the forestomach (IARC, 1999c; D.S. EPA, 1990). However, this result was questioned by other investigators, and, upon reexamination of histological slides, a group of National Toxicology Program pathologists concluded that the lesions were hyperplasia and inflammation rather than neoplasia (U.S. EPA, 1990). In inhalation studies, the most relevant route of human exposure, methyl bromide was reported to be largely negative, although there was some limited evidence for tumorigenicity in various tissues. According to the IARC, no significant increase in tumors was observed in two inhalation studies in mice and one in rats (IARC, 1999c). In another rat inhalation study, a signif-


icant increase in pituitary gland adenomas was seen in males treated at the highest dose. However, a detailed examination of two of the inhalation studies described as negative led some reviewers to suggest that methyl bromide was capable of inducing tumors in some tissues (CDPR, 1999b). Based on its evaluation of the literature, the IARC concluded that there is limited evidence in experimental animals for the carcinogenicity of this agent (lARC, 1999c), whereas the EPA considered the data inadequate to reach any conclusion (U.S. EPA, 1990). Both the IARC and the EPA stated that there is inadequate evidence to make conclusions about the carcinogenicity of methyl bromide in humans (IARC, 1999c; U.S. EPA, 1990). As indicated previously, methyl bromide is genotoxic in most in vitro and in vivo assays. Although there is some evidence for the carcinogenicity of methyl bromide, it has not exhibited consistent carcinogenic effects in most studies. The reason for the discrepancy between the short-tenn tests and the animal bioassay results is not clear. These negative test results could be false negatives, reflecting inadequacies of the animal bioassays. However, several bioassays have been conducted with similar results and no carcinogenic effects were seen even in the comprehensive mouse bioassay conducted by the National Toxicology Program (NTP, 1992). Alternatively, methyl bromide may alkyl ate DNA in vivo at sites that are readily repaired or lead directly to celllethality rather than heritable mutations. 35.4.3 PESTICIDES EXHIBITING CARCINOGENICITY WITHOUT APPRECIABLE GENOTOXICITY 35.4.3.1 Propoxur

Propoxur, or Baygon, is an important carbamate insecticide used primarily against household insects and pests of domestic animals. It is considered among the top 10 most widely used home and garden pesticides in the United States (Grossman, 1995). Similar to other carbamate insecticides, propoxur inhibits acetylcholinesterase, an enzyme involved in neurotransmission, producing neurotoxic effects in insects and nontarget organisms. Propoxur has yielded negative results in the majority of short-term genotoxicity tests that have been conducted (FAOIWHO, 1990). It was negative in bacterial and mammalian mutation assays and in bacterial DNA repair assays. It was reported to be negative in most chromosome aberration and micronucleus assays in vitro and in vivo (FAOIWHO, 1990), although two positive studies have recently been published (Agrawal and Mehrotra, 1997; Wei et aI., 1997). In chronic studies, no evidence of carcinogenic effects was seen in mice treated with propoxur for 24 months or hamsters treated for 53 weeks (FAOIWHO, 1990). However, in a series of studies conducted in rats, highly significant increases in hyperplasia and bladder tumors were seen at high doses of propoxur. No hyperplastic effects in the bladder were seen in short-tenn studies employing mice, dogs, or monkeys, whereas effects

755

were seen in short-tenn studies in the bladders of SpragueDawley rats. Dietary studies have indicated that, in addition to high doses, the urothelial effects of propoxur are dependent on high urinary pH. The carcinogenic effects of propoxur in rats have been proposed to be due to chronic mitogenic stimulation of propoxur or a metabolite on the urothelium rather than from a direct genotoxic or mutagenic effect (Cohen et aI., 1994). 35.4.3.2 Hexachlorobenzene

Historically, hexachlorobenzene (HCB) was commonly used as a seed treatment for prevention of fungal growth on crops such as wheat, barley, oats, and rye (IPCS, 1997). Concern for human health and the environment resulted in its discontinued use as a pesticide in many countries during the 1970s. Hexachlorobenzene is currently found as an unintentional by-product in several high-volume chlorinated solvents (carbon tetrachloride, trichloroethylene, and perchloroethylene) and in various pesticides, including pentachloronitrobenzene, chlorothalonil, dimethyl 2,3,5,6-tetrachlorotereohthalate (DCPA), picloram, and pentachlorophenol (ATSDR, 1997). In general, studies investigating the genotoxicity of HCB have indicated that it exhibits weak or no genotoxic activity (Brusick, 1986; Gorski et aI., 1986; IPCS, 1997). In most studies, hexachlorobenzene exhibited no detectable mutagenic activity in Salmonella either with or without microsomal activation. No increase in structural chromosome aberrations was seen in Chinese hamster lung cells (Ishidate et aI., 1988). Canonero and associates evaluated the in vitro genotoxicity of HCB in primary cultures of rat and human hepatocytes (Canonero et aI., 1997). An induction of micronuclei but not DNA strand breaks was seen in rat hepatocytes treated with hexachlorobenzene. In the studies with human hepatocytes, the authors reported that hexachlorobenzene induced a weak but significant increase in the frequency of both DNA breaks and micronuclei. Low levels of DNA binding were seen following the in vivo treatment of rats with hexachlorobenzene (Gopalaswamy and Nair, 1992). Additionally, no increase in sister chromatid exchanges (SCEs) in the bone marrow of male mice or DNA fragmentation in the liver of rats was observed in hexachlorobenzene-treated animals (Gorski et aI., 1986). Hexachlorobenzene also failed to induce dominant lethal mutations in male rats (Simon et aI., 1979). In contrast with the largely negative results in the genotoxicity studies, hexachlorobenzene exhibited carcinogenic effects in a series of animal studies, increasing the incidence of tumors in rats, hamsters, and mice (lARC, 1987b; IPCS, 1997; U.S. EPA, 1985, 1996). Increased tumor fonnation was seen in the liver and kidney as well as the adrenal, parathyroid, and thyroid glands of the treated animals. To date, the mechanisms underlying carcinogenesis in these organs remain unclear. Several theories have been proposed to explain the basis for certain tumors induced by hexachlorobenzene. For example, it has been proposed that the liver tumors occur as a secondary effect resulting from chronic toxicity to this organ (Carthew and Smith, 1994). It has been postulated that the male kidney tumors were due

756

CHAPTER 35


to an accumulation of the male rat-specific protein alpha 2uglobulin in the proximal renal tubular cells, resulting in a sustained cell proliferation and eventually neoplasia in this organ (Bouthillier et aI., 1991). Finally, others have proposed that the thyroid tumors were the result of a chronic stimulation of cell proliferation in the thyroid gland due to a chronic imbalance in thyroid hormones resulting from an induction of glucuronosyl transferases by hexachlorobenzene (DFG, 1998). All of these theories indicate that hexachlorobenzene exerts its carcinogenic effects through indirect or "nongenotoxic" mechanisms. Assuming that these mechanisms are correct, the difference in the observed genotoxicity and carcinogenicity results would be expected. Following a review of the data, the IARC and the EPA have determined that there was sufficient evidence to conclude that HCB induces cancer in laboratory animals (lARC, 1987b; U.S. EPA, 1996). The evidence in humans is inadequate to draw definite conclusions. However, for regulatory purposes, the EPA considers hexachlorobenzene to be a probable human carcinogen (U.S. EPA, 1996), whereas the IARC considers it to be a possible human carcinogen (IARC, 1987b).

35.4.4 NONGENOTOXIC AGENTS WITHOUT EVIDENCE OF CARCINOGENICITY 35.4.4.1 Endosulfan Endosulfan, or thiodan, is a chlorinated insecticide used on a wide variety of food and non-food crops, including grapes, cantaloupes, lettuce, tomatoes, alfalfa, and cotton. Although a few positive responses have been reported in short-term tests (Smith, 1991), endosulfan is generally viewed by regulatory bodies as being nongenotoxic (CDFA, 1988; FAOIWHO, 1999). Endosulfan has primarily exhibited negative results in both bacterial and mammalian cell gene mutation assays. It was also negative in inducing chromosome aberrations or micronuclei in vitro as well as in vivo. In addition, it has been reported to be negative in other genotoxicity assays. Endosulfan did not exhibit carcinogenic effects in chronic bioassays conducted using mice or rats (FAOIWHO, 1999). Epidemiological studies of cancer in humans have not been conducted.

35.4.4.2 Chlorpyrifos Chlorpyrifos, or Dursban, is a broad-spectrum organophosphate insecticide with widespread usage on food commodities, turf, and ornamental plants. It has been commonly used indoors and for structural pest control. It is one of the most widely used pesticides in the United States and has been one of the top five insecticides used in residential settings (U.S. EPA, 1999). In common with other organophosphate insecticides, upon bioactivation, chlorpyrifos inhibits acetylcholinesterase, an enzyme involved in neurotransmission, producing neurotoxic effects in insects and nontarget organisms. Consequently, genotoxic effects would not be expected nor are they seen (CDPR, 1999a; U.S. EPA, 1999). Chlorpyrifos did not induce gene mutations in either bacterial and mammalian systems, although it was reported to induce slight increases in genetic alterations in yeast

as well as DNA damage in bacteria. No increase in chromosome aberrations was seen in an in vitro study using rat lymphocytes or in two in vivo studies evaluating micronuclei in the mouse bone marrow. It was ineffective at inducing unscheduled DNA synthesis in isolated rat hepatocytes. Chlorpyrifos was evaluated for carcinogenic potential in both rats and mice with no evidence of carcinogenicity (CDPR, 1999a; U.S. EPA, 1999).

35.4.4.3 Bentazon Bentazon, 3-( 1-methylethy 1)-1 H -2,1 ,3-benzothiadiazin-4(3 H)one-2,2-dioxide, is a herbicide used in agriculture for control of broadleaf weeds in crops such as soy beans, rice, corn, peanuts, and lima beans (U.S. EPA, 1998g). As summarized from EPA reports (U.S. EPA, 1998b, g), Bentazon is not chemically reactive and no highly reactive species have been identified during its metabolism. Bentazon was negative in bacterial mutation assays, in a mammalian cell assay, in the unscheduled DNA synthesis assay, and in the mouse micronucleus assay in vivo. In chronic animal bioassays, no increases in tumors were seen in the rat. A slight dose-related increase in hepatocellular tumors was seen in the mouse studies. However, upon reexamination, it was concluded that the incidence did not differ significantly from the controls. In its evaluation of the toxicity of Bentazon, the EPA concluded that bentazon was essentially noncarcinogenic in animals and was not likely to cause cancer in humans.

35.4.5 PESTICIDES EXHIBITING MIXED RESULTS IN GENOTOXICITY OR CANCER TESTS

35.4.5.1 ortho- Phenyl phenol ortho-Phenylphenol (OPP) and its sodium salt, sodium o-phenylphenate (SOPP), are broad-spectrum fungicides and disinfectants with widespread agricultural, industrial, and domestic usage. OPP has historically been among the most widely used home and garden pesticides (Grossman, 1995). Investigations into the genotoxic effects of SOPP and OPP have indicated that these compounds are inactive or weakly active in bacterial mutation assays (NTP, 1986). Some evidence for the mutagenicity of OPP has been seen in mammalian cell assays. A weak increase in mutations was seen at the TK locus in treated CHO cells (NTP, 1986), whereas a strong increase in ouabain-resistant mutants was reported to occur in an ultraviolet-sensitive human Rsa cell line following treatment with OPP (Suzuki et aI., 1985). Negative results were also observed when measuring unscheduled DNA synthesis in rat hepatocytes following exposure to SOPP (Reitz et aI., 1983). In cytogenetic studies, several reports indicate that OPP and its metabolite phenylhydroquinone have induced sister chromatid exchanges and structural chromosomal aberrations in CHO cells in the presence of exogenous metabolic activation (NTP, 1986; Tayama and Nakagawa, 1991; Tayama et al., 1989; Tayama-Nawai et al., 1984), whereas others have reported negative or ambiguous results (Ishidate, 1988;


NTP, 1986). Phenylhydroquinone was also shown to induce chromosome-containing micronuclei upon prostaglandin[H] synthase-mediated activation in V79 cells (Lambert and Eastmond,1994). Following the in vivo administration of radiolabeled OPP and SOPP to male F344 rats, no increases in the covalent binding of these compounds to rat bladder DNA were observed using either liquid scintillation counting (Reitz et al., 1983) or a highly sensitive accelerator mass spectrometric technique (Kwok and Eastmond, 1997). Binding to bladder proteins was seen in both studies. Contradictory results have been reported for DNA binding using the 32p postlabeling technique with one group reporting detectable OPP-derived adducts (Ushiyama et aI., 1992) whereas another, focusing on adduct formation in the target urothelial cells, reported negative results (Smith et aI., 1998). A modest increase in DNA breakage in the bladder was detected in rats (Morimoto et aI., 1989) and mice (Sasaki et aI., 1997) following treatment with OPP or SOPP. In addition, a significant increase in micronucleated bladder cells was reported in rats administered a high dose of OPP in the diet (Tadi-Uppala et aI., 1996). OPP and SOPP have been tested for carcinogenicity in both mice and rats by administration in the diet. Increases in bladder tumors were seen in multiple rat studies following treatment with OPP and SOPP (CDPR, 1997; IARC, 1999d). SOPP appears to be more potent and consistent in inducing carcinogenic effects, and it has been proposed that urinary pH plays an important role in the bioactivation and carcinogenesis of these compounds (Fujii et aI., 1987; Kwok and Eastmond, 1997). The effects appear to be specific to the rat as little evidence of carcinogenicity was observed in chronically treated mice (IARC, 1999d), and bladder toxicity was not seen in short-term studies in mice, guinea pigs, hamsters, and dogs (Co see et aI., 1992; Hasegawa et aI., 1990). Upon review of the data, the IARC concluded that OPP was not classifiable as to its carcinogenicity to humans and that SOPP was possibly carcinogenic to humans (IARC, 1999d). The mechanisms underlying the carcinogenic effects of OPP remain to be fully elucidated. It has been proposed that OPP acts as a bladder carcinogen in rats by inducing cytotoxicity and hyperplasia without directly binding to DNA (Smith et aI., 1998). In this case, the genotoxicity may be indirect, occurring through the formation of oxygen radicals, through an enhancement of spontaneous mutations, or through an interaction with protein targets (Appel, 2000; Kwok and Eastmond, 1997). The inconsistent results seen in the short-term tests may, in part, be a reflection of this indirect mechanism of genotoxicity. 35.4.5.2 1,4-Dichlorobenzene l,4-Dichlorobenzene, or para-dichlorobenzene (p-DCB), is commonly used to control moths, molds, and mildew, and as a bathroom deodorizer. p-DCB is also used as an intermediate in the synthesis of polyphenylene sulfide (PPS) resin. The genotoxicity of p-DCB has been investigated with mixed, largely negative results (IARC, 1999a). p-DCB was not mutagenic in bacteria or mammalian cells in vitro but did exhibit

757

some evidence of DNA damage and mutagenicity in yeast. pDCB produced mixed results in in vitro cytogenetic assays with both positive and negative reports for micronuclei and sister chromatid exchanges. It was negative in inducing DNA strand breaks and chromosome aberrations in vitro. p-DCB failed to exhibit genotoxic effects in vivo, exhibiting negative responses in unscheduled DNA synthesis, in the chromosome aberration assay, in the dominant lethal assay, and in the in vivo micronucleus assay. It was reported as positive in one DNA strand breakage assay and in one in vivo micronucleus assay. p-DCB bound to DNA in the liver, lung, and kidney of mice but not in that of male rats (IARC, 1999a). It also induced DNA damage in the liver and spleen but not in the kidney, lung, or bone marrow of mice. The IARC stated that no conclusion could be drawn from the few data on genotoxicity in vivo (IARC, 1999a). In contrast to the negative genotoxicity results, p-DCB induced carcinogenic effects in both rats and mice. Following oral administration, p-DCB increased the incidence of liver tumors in male and female mice as well as the incidence of renal carcinomas in male rats (IARC, 1999a). In evaluating the significance of these tumors, the IARC concluded that the evidence did not support a mechanism of renal cell tumor formation that involved a direct interaction between p-DCB or its metabolites with DNA. The male kidney tumors induced by p-DCB were due to an accumulation of the male rat-specific protein alpha 2u-globulin in the proximal renal tubular cells that eventually resulted in neoplasia in this organ. This mechanism is widely accepted as not being relevant to humans (U.S. EPA, 1991; IARC, 1999a; Rice et al., 1999a). However, the IARC Working Group had more concern for the liver tumors that were seen at a high incidence in the male and female mice. Because p-DCB was reported to cause DNA damage in the liver and spleen of mice and bound weakly to DNA, the tumors in the liver were thought to be potentially relevant to humans. IARC concluded that p-DCB was an animal carcinogen and possibly carcinogenic to humans (lARC, 1999a). As illustrated in the preceding examples, different patterns of genotoxic and carcinogenic effects can be seen in short-term tests and in animal bioassays. In many instances, the outcome of the studies and interpretation of their relevance to humans is relatively straightforward, indicating that these agents pose or do not pose significant carcinogenic risks to humans. However, in other cases, the interpretation of the results can be quite challenging. In almost all cases, scientists and regulators rely on a weight-of-the-evidence approach, where the number, consistency, and quality of the studies is combined with mechanistic, structure-activity, and other information to reach conclusions about the genotoxicity and likely human carcinogenicity of the agent. In addition to the short-term tests and animal results, information about the genotoxic effects of the pesticide in humans can contribute significantly to the risk assessment process.

758

CHAPTER 35


35.5 HUMAN BIOMONITORING To identify pesticides and other agents capable of inducing genotoxicity in humans and to identify groups at elevated risk for cancer or other genetic diseases, biological markers of exposure and effect have been developed to measure genetic changes in exposed humans (Albertini and Hayes, 1997; Albertini et aI., 2000; Sorsa et aI., 1992; Tucker et aI., 1997; Wild and Pisani, 1997). These biomarkers range from early premutagenic lesions such as covalent adducts between the chemical and DNA to heritable mutations in endogenous genes such as HPRT. Although these studies have primarily been conducted using somatic cells, a few have been performed using germ cells in which chromosomal changes in human sperm have been monitored. Among the most commonly used biomarkers is the measurement of structural and numerical alterations in lymphocyte chromosomes. In this assay, the frequencies of chromosome changes occurring in metaphase preparations of stimulated peripheral blood lymphocytes from individuals in an exposed group are measured and compared with those of an appropriate control. Increased frequencies of genetic alterations are believed to indicate that an exposure has occurred that is biologically significant and mechanistically related to cancer and other genetic diseases (Sorsa et aI., 1992). Consistent with this, recent studies have shown that individuals with elevated frequencies of structural chromosomal aberrations in their peripheral blood lymphocytes are at increased risk for the development of cancer (Bonassi et aI., 1995; Hagmar et aI., 1994, 1998). It should be noted that for one frequently measured endpoint, sister chromatid exchanges, such an association was not seen (Hagmar et aI., 1994,1998). A considerable number of studies have been conducted using various biomarkers to measure genetic alterations in the cells of pesticide-exposed workers. A list of genetic biomarker studies obtained primarily from a search of MED LINE (and references cited therein) is shown in Table 35.4. Although these studies represent only a fraction of the studies that have been conducted, the results and patterns of response are probably representative of those commonly seen in pesticide biomonitoring studies. As can be seen from the table, numerous reports from many countries have been published on genotoxic effects in pesticideexposed workers. In many of these, higher frequencies of genotoxic effects have been seen in the exposed workers. However, most studies have been conducted on agricultural workers who have been exposed to many different pesticides. As a result, it is difficult to identify the actual genotoxic agent involved. For example, in the studies conducted in southeast India by Rupa and associates, the cotton field applicators reported having used 11 different pesticides in the period preceding the study (Rupa et aI., 1989b). Even in cases where the exposed workers were exposed primarily to a single pesticide, the reported outcome may be influenced by other confounding factors such as tobacco smoking, age, exposure to solvents, inert ingredients, etc.

As is also apparent from the table, studies have been performed for only a small portion of the thousands of pesticides currently being used. In addition to the paucity of information on most pesticides, interpreting the results of biomonitoring studies such as these and their significance for workers exposed at lower levels or the general public exposed at much lower levels can be difficult. For example, ethylene oxide has exhibited positive responses in the majority of biomonitoring studies and endpoints measured. This is consistent with the known reactivity of this agent and its results in the short-term genotoxicity tests. This would indicate that, at high exposure levels, ethylene oxide poses a genotoxic and carcinogenic risk. However, these studies provide little information about the risk at lower exposure levels, requiring an extrapolation of risk to be made from high exposures to lower exposures. In contrast, negative results were seen in two biomonitoring studies of ethylene dibromide, an agent that yielded positive results in most short-term tests and was carcinogenic in animals. Although one might interpret these results as indicating that EDB is not genotoxic in humans, this conclusion could easily be in error. In this case, the negative results could simply be due to a combination of low exposures and a limited sample size. Quantitative measures of pesticide exposure are infrequently performed in these types of studies. Other results, such as those reported for dichlorodiphenyltrichoroethane (DDT), dimethoate, deltamethrin, and cypermethrin, are also challenging to interpret. Based on the short-term test results, one would not expect these agents to be genotoxic. For example, DDT was negative in 133 out of the143 short-term genotoxicity tests listed in the EPAJIARC Genetic Activity Profile database. This suggests that the positive results seen in these types of biomonitoring studies might be due to other factors such as solvent exposure, tobacco use, etc. (Petrelli et aI., 1993) or may simply be false positives. However, the overall number of positive studies in Table 35.4 far exceeds a reasonable estimate of false positives and indicates that pesticide exposure is frequently associated with genotoxic effects in exposed workers. Although the majority of biomonitoriing studies have been conducted using somatic cells, a small number of studies have been conducted to measure effects in germ cells. Interestingly, positive effects have been reported in three of the four studies conducted to date. Significant increases in aneuploid sperm were seen in agricultural workers exposed to 1,2-dibromo-3chloropropane (DBCP) (Kapp et aI., 1979), in Chinese factory workers exposed to organophosphates (Padungtod et aI., 1999) and in Indian applicators and sprayers exposed to a variety of pesticides (predominantly organophosphate insecticides) (Rupa et aI., 1997). Moreover, increases in breakage/exchanges affecting the lcen-lql2 region of chromosome 1 were also detected in the sperm of the Indian cotton field workers (Rupa et aI., 1997). Notably in earlier studies by Rupa and associates, the Indian group of applicators involved in the sperm and lymphocyte aberration studies had previously been reported to exhibit significant decreases in reproductive performance (fertility, pregnancy loss, and birth anomalies) (Rupa et aI., 1991b). These initial reports indicate that exposure to certain pesticides can

35.5 Human Biomonitoring

759

Table 35.4 Summary of Results of Genotoxicity Studies of Pesticide-Exposed Workers Study group

Location

Pesticide

Endpoint

Result

Reference Linnainmaa, 1983

Foliage sprayers

Finland

2,4-D and MCPA

SCE

Negative

Workers

United States

DBCP

Sperm aneuploidy

Positive

Kapp et aI., 1979

Workers in insecticide plants

Brazil

DDT

Cs aberrationsa

Positive

Rabello et al., 1975

Sprayers

Syria

Deltamethrin and

Cs aberrations

Positive

Mohammad et aI., 1995

SCE

Positive

Larripa et aI., 1983 Steenland et aI., 1986

cypermethrin Brazil

Dimethoate

Papaya workers

Hawaii

Ethylene dibromide

SCE

Negative

Papaya workers

Hawaii

Ethylene dibromide

Cs aberrations

Negative

Steenland et aI., 1986

Pesticide sprayers

United States

Ethylene dibromide

SCE

Negative


Accidental exposure of firefighters

Pesticide sprayers

United States

Ethylene dibromide

Cs aberrations

Negative


Factory workers

Sweden

Ethylene oxide

SCE

Negative

Hogstedt et aI., 1983

Factory workers

Sweden

Ethylene oxide

Cs aberrations

Positive


Factory workers

Sweden

Ethylene oxide

Micronuclei

Negative

Hogstedt et al., 1983

Factory workers

Sweden

Ethylene oxide

MicronucIeib

Positive


Sanitary workers

Italy

Ethylene oxide

Cs aberrations

Positive

Sarto et aI., 1984

Sanitary workers

Italy

Ethylene oxide

SCE

Positive

Sarto et aI., 1984

Sterilizer operators

United States

Ethylene oxide

Cs aberrations

NegativeC

Galloway et aI., 1986

Malathion workers

United States

Malathion

Micronuclei

Negative

Titenko-Holland et aI., 1997

Workers (production)

Czechoslovakia

Mancozeb-

Cs aberrations

Positive

lablonicbi et al., 1989

SCE

Positive

lablonicka et aI., 1989

Calvert et aI., 1998

containing fungicide Novozir Mn80 Workers (production)

Czechoslovakia

Mancozebcontaining fungicide Novozir Mn80

Fumigation workers

United States

Methyl bromide

HPRT mutations

Negative

Fumigation workers

United States

Methyl bromide

Micronucleid

Equivocal

Calvert et aI., 1998

Pesticide plant workers

Brazil

Methyl parathion

Cs aberrations

Negative

de Cassia Stocco et aI., 1982

Pesticide-preparing workers

Hungary

Monochlorinated benzene

HPRT mutation

Negative

Major et aI., 1992

Patients (attempted suicide or

Hungary

Organophosphates

Cs aberrationse

Positive

van Bao et aI., 1974

exposed during work) Fumigant applicators

United States

Phosphine

Cs aberrations

Positive

Garry et aI., 1989

Fumigant applicators

United States

Phosphine

SCE

Negative

Garry et aI., 1989

Pesticide applicators

United States

Phosphine

Cs rearrangements

Positive

Garry et aI., 1992

Pesticide sprayers

Hungary

Pyrethroids

Cs aberrations!

Positive

Nehez et aI., 1988

Workers (fitters, packers,

Former Soviet

Zineb

Cs aberrations

Positive

Pilinskaya, 1974

Ziram

Cs aberrations

Positive

Pilinskaya, 1970 Yoderetal.,1973

truck drivers) Store workers and packers

Union Former Soviet Union


United States

Herbicides

Cs aberrations

Positive


United States

Insecticides

Cs aberrations

Positive

Yoder et aI., 1973

Farmers

Denmark

Fungicides

Aneuploid sperm

Negative

Harkonen et aI., 1999


United States

Pesticides

Cs aberrations

Positive

Yoder et aI., 1973

Sprayers

New Zealand

Pesticides

SCE

Negative

Crossen and Morgan, 1978

Pesticide workers

Sweden

Pesticides

Cs aberrations

Negative


Exposed workers

Hungary

Pesticides

Cs aberrations

Positive

Nehez et aI., 1981

Agricultural workers

Former Soviet

Pesticides

Cs aberrations

Positive

Volnjanskaya, 198 I

Union (continues)

760

CHAPTER 35


Table 35.4 (continued) Study group

Location

Pesticide

Endpoint

Result

Reference

Floriculturists

Argentina

Pesticides

SCE

Positive

Dulout et aI., 1985

Floriculturists

Argentina

Pesticides

Cs aberrationsg

Negative

Dulout et aI., 1985

Greenhouse pesticide sprayers

Hungary

Pesticides

Cs aberrations!

Positive

Desi et aI., 1986

Ornamental plant breeders

Argentina

Pesticides

Cs aberrations

Negative

Dulout et aI., 1987

Pesticide workers

Mexico

Pesticides

Cs aberrations

Positive

Gayon et aI., 1987

Mixers and field sprayers

Hungary

Pesticides

Cs aberrations

Positive

PaIdy et aI., 1987

Pesticide sprayers in vineyards

India

Pesticides

Cs aberrations

Positive

Rita et aI., 1987

Pesticide sprayers

Hungary

Pesticides

Cs aberrations!

Positive

Nehez et aI., 1988

Vegetable garden workers

India

Pesticides

Cs aberrations

Positive

Rupa et aI., 1988

Fumigant applicators

United States

Pesticides

Cs aberrations

Positive

Garry et aI., 1989

Pesticide sprayers

India

Pesticides

Cs aberrations

Positive

Rupa et aI., 1989b

Pesticide mixers and sprayers

India

Pesticides

Cs aberrations

Positive

Rupa et aI., 1989a


Canada

Pesticides

Micronuclei h

Positive

San et aI., 1989


Spain

Pesticides

SCE

Negative

Carbonell et aI., 1990

Workers in flower industry

Italy

Pesticides

Cs aberrations

Positive

De Ferrari et aI., 1991

Workers in flower industry

Italy

Pesticides

SCE

Positive

De Ferrari et aI., 1991

Cotton field workers

India

Pesticides

Cs aberrations

Positive

Rupa et aI., 1991a


India

Pesticides

SCE

Positive

Rupa et aI., 1991c


United States

Pesticides

Cs rearrangements

Positive

Garry et aI., 1992

Workers in plastic greenhouses

Greece

Pesticides

Cs aberrations

Positive

Kourakis et aI., 1992

Pesticide workers

Mexico

Pesticides

SCE

Negative

G6mez-Arroyo et aI., 1992

Floriculturists

Italy

Pesticides

Micronuclei

Positive

Bolognesi et aI., 1993a

Floriculturists

Italy

Pesticides

Micronuclei

Positive

Bolognesi et al.. 1993b


Spain

Pesticides

Cs aberrations

Positive

Carbonell et al.. 1993


Spain

Pesticides

SCE

Negative


Pesticide packers

Egypt

Pesticides

Cs aberrations

Positive

Anwar,1994

Pesticide packers

Egypt

Pesticides

SCE

Negative

Anwar,1994

Farm workers

Spain

Pesticides

Cs aberrations

Positive


Greenhouse sprayers

Scandinavia

Pesticides

SCE

Negative

Lander and Ronne, 1995


India

Pesticides

Cs aberrations i

Positive

Rupa et aI., 1995

Dealers and controllers

Syria

Pesticides

Cs aberrations

Positive

Mohammad et aI., 1995

Farmers

Colombia

Pesticides

SCE

Negative

Hoyos et aI., 1996

Farmers

Colombia

Pesticides

Cs aberrations

Negative

Hoyos et aI., 1996

Pesticide sprayers

Greece

Pesticides

SCE

Negative

Kourakis et al., 1996

Pesticide sprayers

Hungary

Pesticides

Cs aberrations

Positive

Nehez and Desi, 1996

Farmers

Italy

Pesticides

SCE

Negative

Pasquini et aI., 1996

Farmers

Italy

Pesticides

Micronuclei

Positive

Pasquini et aI., 1996

Greenhouse floriculturists

Italy

Pesticides

DNA adducts

Positive

Peluso et aI., 1996


Italy

Pesticides

Cs aberrations

Negative

Scarpato et aI., 1996


Italy

Pesticides

Micronuclei

Negative


SCE

Negative


Positive

Joksic et aI., 1997


Italy

Pesticides

Cs

aberrations j

Vineyard growers

Yugoslavia

Pesticides

Pesticide sprayers

France

Pesticides

DNA damage

Increase

Lebailly et aI., 1998

Mixers and applicators

India

Pesticides

Cs aberrations in

Positive

Rupa et aI., 1997

sperm i (continues)

35.6 Genotoxicity and Risk Assessment

761

Table 35.4 (continued) Pesticide

Endpoint

Chile

Pesticides

Micronuclei

Negative

Venegas et aI., 1998

Italy

Pesticides

Micronuclei

Positive

Falck et aI., 1999

Pesticide industry workers

India

Pesticides

SCE

Positive

Padrnavathi et aI., 2000

Farm workers

Canada

Pesticides

Micronuclei

Equivocal

Davies et aI., 1998

Greenhouse workers

Spain

Pesticides

Micronuclei

Negative

Lucero et aI., 2000

Factory workers

China

Pesticides

Aneuploid spenn

Positive

Padungtod et aI., 1999

Study group

Location

Pesticide sprayers Greenhouse workers

Result

Reference

Cs aberrations refer to both chromosome and/or chromatid aberrations. Micronuclei in bone marrow cells. C Negative at low and moderate exposures but positive at high exposures. d Micronuclei in oropharyngeal cells. e Breaks; unstable and stable chromosomal aberrations. f Numerical chromosomal aberrations. g Structural chromosomal aberrations, but exchanges showed a statistically significant increase in exposed over controls.

a

b

Micronuclei in exfoliated urothelial cells. Affecting the lcen-Iql2 region. j Unstable chromosomal aberrations during prespraying period. All studies were perfonned on peripheral blood Iymphocytes unless otherwise noted.

h i

induce chromosome alterations in the sperm of the exposed workers and may contribute to decreased reproductive performance of the workers.

35.6 GENOTOXICITY AND RISK ASSESSMENT As described previously, short-term tests for genotoxicity are required by regulatory agencies for pesticide registration and play an important role in the safety evaluation and risk assessment process. For the few agents that have been evaluated for heritable risks, genotoxicity assays, particularly those assessing heritable effects in germ cells, have played a critical role. Historically, in cancer risk assessment, the short-term test results and human biomonitoring studies have been used to alert agencies and the public to pesticides with potential cancercausing properties as well as to provide valuable supplemental information for the positive or negative results seen in animal bioassays. In recently implemented or proposed regulatory strategies, genotoxicity information plays an increasingly important role in the risk assessment process. DNA reactivity and mechanisms of genotoxicity are being used to provide in sights into an agent's mode of action and, as a result, may play a pivotal role in determining whether linear or nonlinear (apparent threshold) models will be used for extrapolation from high animal doses to lower exposure levels. In the EPA approach, genotoxic effects may also be modeled as precursor events to provide the basis for the selection of a certain extrapolation procedure (Wiltse and Dellarco, 1996). The use of mechanistic or mode-of-action information plays an important role in the cancer risk assessment guidelines proposed by the EPA (Wiltse and Dellarco, 1996) as well as in those implemented by other national and international regulatory groups such as the IARC

(IARC, 199ge) and the German Commission for the Investigation of Health Hazards of Chemical Compounds in the Work Area of the Deutsche Forschungsgemeinschaft (Neumann et aI., 1998). The evaluation of ethylene oxide provides a recent example of the contribution of genotoxicity data to the cancer risk assessment process. Upon reviewing the literature on the carcinogenicity of ethylene oxide in humans and animals, the IARC Working Group concluded that there was limited evidence for the carcinogenicity of ethylene oxide in humans but sufficient evidence in animals (IARC, 1994). However, in its overall evaluation, the IARC Working Group concluded that ethylene oxide is carcinogenic to humans. In making this conclusion, the IARC took into consideration evidence that "ethylene oxide is a directly acting alkylating agent that: (i) induces a sensitive, persistent dose-related increase in the frequency of chromosomal aberrations and sister chromatid exchanges in the peripheral lymphocytes and micronuclei in the bone-marrow cells of exposed workers; (ii) has been associated with malignancies of the lymphatic and haematopoietic system in both humans and experimental animals; (iii) induces a dose-related increase in the frequency of haemoglobin adducts in exposed humans and dose-related increases in the numbers of adducts in both DNA and haemoglobin in exposed rodents; (iv) induces gene mutations and heritable translocations in germ cells of exposed rodents; and, (v) is a powerful mutagen and clastogen at all phylogenetic levels." In a similar fashion with differing conclusions, the IARC recently evaluated data on the relevance of rodent tumors of the urinary bladder, renal cortex, mammary gland, and thyroid gland induced by agents such as atrazine, chlorothalonil, OPP, p-DCB, and saccharin and their relevance to carcinogenic risk in humans (Rice et aI., 1999a). In a number of cases, the lack of genotoxicity exhibited by these agents or their metabolites

762

CHAPTER 35


played an important role in its conclusions that the mechanisms by which agents such as atrazine and saccharin induced cancer in rodents were not relevant to humans (Rice et aI., 1999a). Other governmental groups have reached similar conclusions (NTP, 2000; U.S. EPA, 1991, 1998a). It should be emphasized that, in all cases, critical evaluation should be used in the interpretation and application of short-term test results in the risk assessment process. Given the large number of tests that can be performed in different cells or strains and at multiple dose levels, positive results should be expected in some tests by random chance alone. As a result, reproducibility and consistency become particularly important in evaluating genotoxicity test results. Short-term tests can also be performed in vitro or in vivo under conditions that will produce positive test results but that are unlikely to pose significant genotoxic risks to humans. For example, there is increasing recognition that positive responses in the in vitro chromosome aberration assay can be caused by mechanisms such as endonuclease activation that are not likely to occur at lower doses (Galloway, 2000; Scott et aI., 1991). These tend to occur more frequently at high test concentrations under conditions in which high osmolality, extremes of pH, or excessive cytotoxicity are seen. Similarly, genotoxic effects may occur at concentrations in vitro that most likely would not occur in vivo as other types of toxic effects such as neurotoxicity would be dose limiting. A comparison of in vitro concentrations or in vivo animal plasma concentrations with expected plasma levels in humans under conditions of normal (and above normal) usage can assist in the interpretation of the test data. Conversely, negative results in short-term genotoxicity tests should not be given undue weight as they do not exclude the possibility that an effect occurred in tissues that were not examined, that inadequate bioactivation was used, that the test was improperly conducted, or that the agent induces another type of genetic damage (IARC, 199ge; Proctor et aI., 1986). Additionally, negative results in these assays cannot be considered to rule out the carcinogenicity of agents that act through other mechanisms (e.g., receptor-mediated effects, cellular toxicity with regenerative proliferation, or peroxisome proliferation) (IARC, 199ge). By using a weight-of-evidence approach to evaluate the data, the likelihood of error (both false positives and false negatives) can be minimized. In a similar fashion, to confidently use human biomonitoring studies to evaluate risk, one should ensure that the biomarker of interest was sufficiently sensitive to detect changes at the exposure levels of interest, that the number of exposed and control individuals in the study was adequate, that an acceptable number of measurements was collected, and that major confounding variables were controlled. In addition, the identification of the specific pesticide and information on the exposure levels, although frequently difficult to obtain, can add significantly to the evaluation. Although it is uncommon for all of the preceding conditions to be fulfilled, results of human biomonitoring studies, when the specific agent is known, can play a valuable role in the risk assessment process (see the previous example for ethylene oxide).

In conclusion, a significant number of pesticides have exhibited genotoxic effects in short-term genotoxicity assays and may pose significant risks to humans. Consistent with this, chromosomal alterations have been seen in many studies monitoring genotoxic effects in pesticide-exposed workers. However, these studies often involve exposures to multiple pesticides and potential confounding factors and at levels much higher than those experienced by the general pUblic. The ongoing challenge for researchers, regulators, and those interested in environmental health is to effectively use genotoxicity data to distinguish noncarcinogenic and nonmutagenic pesticides from those capable of inducing cancer and heritable mutations in humans, to determine which of the latter pose significant risks at human exposure levels, and to identify safe methods and levels for the use of these agents or to eliminate their usage altogether.

REFERENCES Agency for Toxic Substances and Disease Registry (ATSDR) (1997). "Toxicological Profile of Hexachlorobenzene-Update." PB/97/12 I 065/AS, Agency for Toxic Substances and Disease Registry. Agrawal, R. c., and Mehrotra, N. (1997). Assessment of mutagenic potential of propoxur and its modulation by indole-3-carbinol. Food Chem. ToxicoZ. 35, IOS1-IOS4. Albertini, R. J., and Hayes, R. B. (1997). Somatic cell mutations in cancer epidemiology. IARC Sci. Pub!. 142, 159-IS4. Albertini, R. J., Anderson, D., Douglas, G. R., Hagmar, L., Hemminki, K, Merlo, E, Natarajan, A. T., Norppa, H., Shuker, D. E., Tice, R., Waters, M. D., and Aitio, A. (2000). IPCS guidelines for the monitoring of genotoxic effects of carcinogens in humans. Mutat. Res. 463, 111-172. Ames, B. N., McCann, J., and Yamasaki, E. (1975). Methods for detecting carcinogens and mutagens with the Salmonella/mammalian-microsome mutagenicity test. Mutat. Res. 31, 347-364. Anwar, W. A. (1994). Assessment of cytogenetic changes in human populations at risk in Egypt. Mutat. Res. 313, IS3-191. Appel, K E. (2000). The carcinogenicity of the biocide ortho-phenylphenol. Arch. Toxicol. 74,61-71. Ashby, J., and Paton, D. (1993). The influence of chemical structure on the extent and sites of carcinogenesis for 522 rodent carcinogens and 55 different human carcinogen exposures. Mutat. Res. 286, 3-74. Auletta, A. E., Dearfield, K L., and Cimino, M. C. (1993). Mutagenicity test schemes and guidelines: U.S. EPA Office of Pollution Prevention and Toxics and Office of Pesticide Programs. Environ. Mol. Mutagen. 21, 3S-45; Discussion 46-57. Baker, E. L., Jr., Warren, M., Zack, M., Dobbin, R. D., Miles, J. w., Miller, S., Aldennan, L., and Teeters, W. R. (l97S). Epidemic malathion poisoning in Pakistan malaria workers. Lancet 1, 31-34. Bakir, F., Damluji, S. E, Amin-Zaki, L., Murtadha, M., Khalidi, A., al-Rawi, N. Y., Tikriti, S., Dahahir, H. I., Clarkson, T. w., Smith, J. C., and Doherty, R. A. (1973). Methylmercury poisoning in Iraq. Science 181, 230-241. Blair, A., and Zahm, S. H. (1995). Agricultural exposures and cancer. Environ. Health Perspect. 103, 205-20S. Bolognesi, c., Parrini, M., Bonassi, S., Ianello, G., and Salanitto, A. (l993a). Cytogenetic analysis of a human population occupationally exposed to pesticides. Mutat. Res. 285, 239-249. Bolognesi, c., Parrini, M., Reggiardo, G., Merlo, E, and Bonassi, S. (l993b). Biomonitoring of workers exposed to pesticides. Int. Arch. Occup. Environ. Health 65, SIS5-SIS7. Bonassi, S., Abbondandolo, A., Camurri, L., Dal Pra, L., De Ferrari, M., Degrassi, E, Fomi, A., Lamberti, L., Lando, c., Padovani, P., Sbrana, I., Vecchio, D., and Puntoni, R. (1995). Are chromosome aberrations in circulating Iymphocytes predictive of future cancer onset in humans? Cancer Genet. Cytogenet. 79, 133-135.

References

Bouthillier, L., Greselin, E., Brodeur, J., Viau, c., and Charbonneau, M. (1991). Male rat specific nephrotoxicity resulting from subchronic administration of hexachlorobenzene. Toxicol. Appl. Phannacol. 110,315-326. Brusick, D. (1987). "Principles of Genetic Toxicology." Plenum, New York. Brusick, D. J. (1986). Genotoxicity of hexachlorobenzene and other chlorinated benzenes. IARC Sci. Publ. 77, 393-397. California Department of Food and Agriculture (CDFA) (1988). "Summary of Toxicology Data: Endosulfan." California Department of Food and Agriculture, Sacramento. Available at http://www.cdpr.ca. gov/docs/toxsums/pdfs/259. pdf. California Department of Pesticide Regulation (CDPR) (1997). "Summary of Toxicology Data: o-Phenylphenol." Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento. Available at http://www.cdpr.ca.gov/docs/toxsums/pdfs/448.pdf. California Department of Pesticide Regulation (CDPR) (I 999a). "Summary of Toxicology Data: Chlorpyrifos." Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento. Available at http://www.cdpr.ca.gov/docs/toxsums/pdfs/385.pdf. California Department of Pesticide Regulation (CDPR) (1999b). "Summary of Toxicology Data: Methyl Bromide." Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento. Available at http://www.cdpr.ca.gov/docs/toxsums/pdfs/385.pdf. Calvert, G. M., Talaska, G., Mueller, C. A., Ammenheuser, M. M., Au, W. W., Fajen, J. M., Fleming, L. E., Briggle, T., and Ward, E. (1998). Genotoxicity in workers exposed to methyl bromide. Mutat. Res. 417, 115-128. Canonero, R., Campart, G. B., Mattioli, E, Robbiano, L., and Martelli, A. (1997). Testing of p-dichlorobenzene and hexachlorobenzene for their ability to induce DNA damage and micronucleus formation in primary cultures of rat and human hepatocytes. Mutagenesis 12, 35-39. Carbonell, E., Puig, M., Xamena, N., Creus, A., and Marcos, R. (1990). Sister chromatid exchange in Iymphocytes of agricultural workers exposed to pesticides. Mutagenesis 5,403-405. Carbonell, E., Valbuena, A., Xamena, N., Creus, A., and Marcos, R. (1995). Temporary variations in chromosomal aberrations in a group of agricultural workers exposed to pesticides. Mutat. Res. 344, 127-134. Carbonell, E., Xamena, N., Creus, N., and Marcos, R. (1993). Cytogenetic biomonitoring in a Spanish group of agricultural workers exposed to pesticides. Mutagenesis 8, 511-517. Carthew, P., and Smith, A. G. (1994). Pathological mechanisms of hepatic tumour formation in rats exposed chronically to dietary hexachlorobenzene. 1. Appl. Toxicol. 14,447-452. Cohen, S. M., Cano, M., Johnson, L. S., St. John, M. K., Asamoto, M., Garland, E. M., Thyssen, J. H., Sangha, G. K., and van Goethem, D. L. (1994). Mitogenic effects of propoxur on male rat bladder urothelium. Carcinogenesis 15,2593-2597. Cosee, P. E, Stebbins, K. E., Stott, W. T., Johnson, K. A., and Atkin, L. (1992). ortho-Phenylphenol: Short term/chronic oral toxicity studies in beagle dogs. Toxicologist 12, 121. Crossen, P. E., and Morgan, W. E (1978). Cytogenetic studies of pesticide and herbicide sprayers. N. Z. Med. 1. 88, 192-195. Davies, H. W., Kennedy, S. M., Teschke, K., and Quintana, J. P. E. (1998). Cytogenetic analysis of South Asian berry pickers in British Columbia using the micronucleus assay in peripherallymphocytes. Mutat. Res. 416, 101-113. Dearfield, K. L., Auletta, A. E., Cimino, M. c., and Moore, M. M. (1991). Considerations in the U.S. Environmental Protection Agency's testing approach for mutagenicity [see Comments]. Mutat. Res. 258, 259-283. de Cassia Stocco, R., Becak, W., Gaeta, R., and Rabello-Gay, M. N. (1982). Cytogenetic study of workers exposed to methyl-parathion. Mutat. Res. 103, 71-76. De Ferrari, M., Artuso, M., Bonassi, S., Bonatti, S., Cavalieri, Z., Pescatore, D., Marchini, E., Pisano, v., and Abbondandolo, A. (1991). Cytogenetic biomonitoring of an Italian population exposed to pesticides: Chromosome aberration and sister-chromatid exchange analysis in peripheral blood Iymphocytes. Mutat. Res. 260, 105-113. Dellarco, V. L., Generoso, W. M., Sega, G. A., Fowle, J. R. D., and JacobsonKram, D. (1990). Review of the mutagenicity of ethylene oxide. Environ. Mol. Mutagen. 16, 85-103.

763

Desi, 1., Palotas, M., Vetro, G., Csolle, 1., Nehez, M., Zimanyi, M., Ferke, M., Huszta, E., and Nagymajtenyi, L. (1986). Biological monitoring and health surveillance of a group of greenhouse pesticide sprayers. Toxicol. Left. 33, 91-105. Deutsch Forschungsgemeinschaft (DFG) (1998). "Hexachlorbenzol: Documentation for the MAK Value." Deutsch Forschungsgemeinschaft. Dulout, EN., Pastori, M. c., Gonzalez Cid, M., Matos, E., von Guradze, H. N., Maderna, C. R., Loria, D., Albiano, D., and Sobel, N. (1987). Cytogenetic analysis in plant breeders. Mutat. Res. 189,381-386. Dulout, EN., Pastori, M. C., Olivero, O. A., Gonzalez Cid, M., Loria, D., Matos, E., Sobel, N., de Bujan, E. C., and Albiano, N. (1985). Sisterchromatid exchanges and chromosomal aberrations in a population exposed to pesticides. Mutat. Res. 143,237-244. European Commission (1999). "CSTEE Opinion on Human and Wildlife Health Effects of Endocrine Disrupting Chemicals, with Emphasis on Wildlife and Ecotoxicology Test Methods." European Commission, Directorate General XXIV, Brussels, 1999. U.S. Environmental Protection Agency (EPA) (1979). "Short-Term Tests for Carcinogens, Mutagens and Other Genotoxic Agents." EPA-625/9-79-003, U.S. Environmental Protection Agency, Research Triangle Park, NC. U.S. Environmental Protection Agency (EPA) (1985). "Health Assessment Document for Chlorinated Benzenes-Final Report." EPN600/8-84/015F, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1990). "BromomethaneIntegrated Risk Information System (IRIS) Record." U.S. Environmental Protection Agency, Washington, DC. Available at http://www.epa. gov/iris/substl0015.htm#II. U.S. Environmental Protection Agency (EPA) (1991). "Alpha2u-Globulin: Association with Chemically-induced Renal Toxicity and Neoplasia in the Male Rat." EPN625/3-91/019F, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1996). "HexachlorobenzeneIntegrated Risk Information System (IRIS) Record." U.S. Environmental Protection Agency, Washington, DC. Available at http://www.epa. gov/iris/substl0374.htm. U.S. Environmental Protection Agency (EPA) (1997). "1,2-DibromoethaneIntegrated Risk Information System (IRIS) Record." U.S. Environmental Protection Agency, Washington, DC. Available at http://www.epa. gov/iris/substl0361.htm. U .S. Environmental Protection Agency (EPA) (1998a). "Assessment of Thyroid Follicular Cell Tumors." EPN6301R-97/002, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (I 998b). "BentazonIntegrated Risk Information System (IRIS) Record." U.S. Environmental Protection Agency, Washington, DC. Available at http://www.epa.gov/iris/substlOI34.htm. U.S. Environmental Protection Agency (EPA) (I 998c). "Health Effects Test Guidelines OPPTS 870.5100 Bacterial Reverse Mutation Test." EPA 712C-98-247, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (I 998d). "Health Effects Test Guidelines OPPTS 870.5300 In Vitro Mammalian Cell Gene Mutation Test." EPA 712-C-98-221, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998e). "Health Effects Test Guidelines OPPTS 870.5385 Mammalian Bone Marrow Chromosome Aberration Test." EPA 712-C-98-225, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (I 998f). "Health Effects Test Guidelines OPPTS 870.5395 Mammalian Erythrocyte Micronucleus Test." EPA 712-C-98-226, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998g). "Toxicological Review of Bentazon (CAS 25057-89-0): In Support of Summary Information on the Integrated Risk Information System (IRIS)." U.S. Environmental Protection Agency, Washington, DC. Available at http://www.epa.gov/iris/toxreviews/0134-tr.pdf.

764

CHAPTER 35


U.S. Environmental Protection Agency (EPA) (1999). "Chlorpyrifos: HED Preliminary Risk Assessment for the Reregistration Eligibility Decision (RED) Document." U.S. Environmental Protection Agency, Washington, DC. Available at http://www.epa.gov/pesticides/op!chlorpyrifos.htm. Falck, G. c., Hirvonen, A., Scarpato, R, Saarikoski, S. T., Migliore, L., and Norppa, H. (1999). Micronuclei in blood Iymphocytes and genetic polymorphism for GSTMl, GSTTl and NAT2 in pesticide-exposed greenhouse workers. Mutat. Res. 441, 225-237. Food and Agricultural OrganizationIWorld Health Organization (FAOIWHO) (1990). Propoxur. In "Pesticide Residues in Food-1989: Toxicological Evaluations," pp. 183-214. WHO, Rome. Food and Agricultural OrganizationIWorld Health Organization (FAOIWHO) (1999). Endosulfan. In "Pesticide Residues in Food-1998: Toxicological Evaluations," pp. 127-158. WHO, Geneva. Fujii, T., Nakamura, K., and Hiraga, K. (1987). Effects of pH on the carcinogenicity of o-phenylphenol and sodium o-phenylphenate in the rat urinary bladder. Food Chem. Toxicol. 25,359-362. Galloway, S. M. (2000). Cytotoxicity and chromosome aberrations in vitro: Experience in industry and the case for an upper limit on toxicity in the aberration assay. Environ. Mol. Mutagen. 35, 191-201. Galloway, S. M., Berry, P. K., Nichols, W. w., Wolman, S. R., Soper, K. A., Stolley, P. D., and Archer, P. (1986). Chromosome aberrations in individuals occupationally exposed to ethylene oxide, and in a large control population. Mutat. Res. 170,55-74. Garry, V. F., Danzl, T. J., Tarone, R, Griffith, J., Cervenka, J., Krueger, L., Whorton, E. B., Jr., and Nelson, R L. (1992). Chromosome rearrangements in fumigant appliers: Possible relationship to non-Hodgkin's lymphoma risk. Cancer Epidemiology, Biomarkers and Prevention 1, 287-291. Garry, V. F., Griffith, J., Danzl, T. J., Nelson, R L., Whorton, E. B., Krueger, L. A., and Cervenka, J. (1989). Human genotoxicity: Pesticide applicators and phosphine. Science 246, 251-255. Gayon, N. Z., Gayon, M. C., and Angulo, M. C. (1987). Clastogenic chromosomal aberrations in a population of workers exposed to different pesticides. Salud Publica Mex. 29, 506-511. Gehring, P. J., Nolan, R J., Watanabe, P. G., and Schumann, A. M. (1991). Solvents, fumigants and related compounds. In "Handbook of Pesticide Toxicology" (w. J. J. Hayes and E. R J. Laws, eds.), pp. 637-730. Academic Press, San Diego. Goldman, L. R. (1998). Chemicals and children's environment: What we don't know about risks. Environ. Health Perspect. 106, 875-880. G6mez-Arroyo, S., Noriega-Aldana, N., Osorio, A., Galicia, F., Ling, S., and Villalobos-Pietrini, R (1992). Sister-chromatid exchange analysis in a rural population of Mexico exposed to pesticides. Mutat. Res. 281,173-179. Gopalaswamy, U. v., and Nair, C. K. (1992). DNA binding and mutagenicity of lindane and its metabolites. Bull. Environ. Contam. Toxicol. 49,300--305. Gorski, T., Gorska, E., Gorecka, D., and Sikora, M. (1986). Hexachlorobenzene is non-genotoxic in short-term tests. IARC Sci. Publ. 399-401. Green, M. A., Heumann, M. A., Wehr, H. M., Foster, L. R, Williams, L. P., Jr., Polder, J. A., Morgan, C. L., Wagner, S. L., Wanke, L. A., and Witt, J. M. (1987). An outbreak of watermelon-borne pesticide toxicity. Am. J. Public Health 77, 1431-1434. Grossman, J. (1995). What's hiding under the sink: Dangers of household pesticides [News]. Environ. Health Perspect. 103, 550-554. Hagmar, L., Bonassi, S., Stromberg, U., Brogger, A., Knudsen, L. E., Norppa, H., and Reuterwall, C. (1998). Chromosomal aberrations in lymphocytes predict human cancer: A report from the European Study Group on Cytogenetic Biomarkers and Health (ESCH). Cancer Res. 58, 4117-4121. Hagmar, L., Brogger, A., Hansteen, 1. L., Heim, S., Hogstedt, B., Knudsen, L., Lambert, B., Linnainmaa, K., Mitelman, F., Nordenson, 1., et al. (1994). Cancer risk in humans predicted by increased levels of chromosomal aberrations in Iymphocytes: Nordic study group on the health risk of chromosome damage. Cancer Res. 54,2919-2922. Hiirkanen, K., Viitanen, T., Larsen, S. B., Bonde, J. P., and Lahdetie, J. (1999). Aneuploidy in sperm and exposure to fungicides and lifestyle factors. ASCLEPIOS. A European Concerted Action on Occupational Hazards to Male Reproductive Capability. Environ. Mol. Mutagen. 34, 39-46.

Hasegawa, R, Takahashi, S., Asamoto, M., Shirai, T., and Fukushima, S. (1990). Species differences in sodium o-phenylphenate induction of urinary bladder lesions. Cancer Lett. 50, 87-91. Hagstedt, B., Gullberg, B., Hedner, K., Kolnig, A. M., Mitelman, F., Skerfving, F., and Widegren, B. (1983). Chromosome aberrations and micronuclei in bone marrow cells and peripheral blood lymphocytes in humans exposed to ethylene oxide. Hereditas 98, 105-113. Hagstedt, B., Kolnig, A. M., Mitelman, F., and Skerfving, S. (1980). Cytogenetic study of pesticides in agricultural work. Hereditas 92, 177-178. Hoyos, L. S., Carvajal, S., Solano, L., Rodriguez, J., Orozco, L., Lopez, J., and Au, W. W. (1996). Cytogenetic monitoring of farmers exposed to pesticides in Colombia. Environ. Health Perspect. 104,535-538. Institute of Medicine (IOM) (1999). "Veterans and Agent Orange: Update 1998." Natl. Acad. Press, Washington, DC. Institute of Medicine (IOM) (2000). "Veterans and Agent Orange: HerbicideIDioxin Exposure and Type 2 Diabetes." Natl. Acad. Press, Washington, DC. International Agency for Research on Cancer (IARC) (1980). "Long-Term and Short-Term Screening Assays for Carcinogens: A Critical Appraisal." International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (1987a). Arsenic and arsenic compounds. In "Overall Evaluations of Carcinogenicity: An Updating of IARC Monographs Volumes 1-42," pp. 100--106. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (1987b). Hexachlorobenzene (Group 2B). In "IARC Monographs of the Evaluation of Carcinogenic Risks to Humans: Overall Evaluations of Carcinogenicity: An Updating of IARC Monographs Volumes 1-42," pp. 219-220. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (1991). Occupational exposures in insecticide application, and some pesticides. In "IARC Monographs on the Evaluation of Carcinogenic Risks to Humans," Vol. 53, pp. 45-92. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (1994). Ethylene oxide. In "IARC Monographs on the Evaluation of Carcinogenic Risks to Humans," Vol. 60, pp. 73-159. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (1997). Polychlorinated dibenzo-para-dioxins. In "IARC Monographs on the Evaluation of Carcinogenic Risks to Humans," Vol. 69, pp. 33-344. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (1999a). Dichlorobenzenes. In "IARC Monographs on the Evaluation of Carcinogenic Risks to Humans," Vol. 73, pp. 223-276. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (1999b). Ethylene dibromide. In "IARC Monographs on the Evaluation of Carcinogenic Risks to Humans," Vol. 71, pp. 641-669. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (1999c). Methyl bromide. In "IARC Monographs on the Evaluation of Carcinogenic Risks to Humans," Vol. 71, pp. 721-735. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (1999d). o-Phenylphenol. In "IARC Monographs on the Evaluation of Carcinogenic Risks to Man," Vol. 73, pp. 451-480. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (199ge). Preamble. In "IARC Monographs," pp. 9-31. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (2000). "IARC Monographs on the Evaluation of Carcinogenic Risks to Humans." Available at http://193.51.164.11lmonoevaUgrlist.html. International Programme of Chemical Safety (IPCS) (1985). "Guide to ShortTerm Tests for Detecting Mutagenic and Carcinogenic Chemicals," Environmental Health Criteria 51. International Programme of Chemical Safety, WHO, Geneva.

References

International Programme of Chemical Safety (IPCS) (1995). "Methyl Bromide," Environmental Health Criteria 166. International Programme of Chemical Safety, WHO, Geneva. International Programme of Chemical Safety (IPCS) (1997). "Hexachlorobenzene," Environmental Health Criteria 195. International Programme of Chemical Safety, WHO, Geneva. Ishidate, M. J. (1988). "Data Book of Chromosomal Aberration Test In Vitro." Elsevier, Amsterdam. Ishidate, M., Jr., Hamois, M. C., and Sofuni, T. (1988). A comparative analysis of data on the clastogenicity of 951 chemical substances tested in mammalian cell cultures. Mutat. Res. 195, 151-213. Jablonicka, A., Polakova, H., Karelova, J., and Vargova, M. (1989). Analysis of chromosome aberrations and sister-chromatid exchanges in peripheral blood lymphocytes of workers with occupational exposure to the mancozebcontaining fungicide Novozir Mn80. Mutat. Res. 224,143-146. Joksic, G., Vidakovic, A., and Spasojevic-Tisma, V. (1997). Cytogenetic monitoring of pesticide sprayers. Environ. Res. 75, 113-118. Jungmann, G. (1966). Arsenic cancer in vintagers. Landarzt 42, 1244-1247. Kapp, K W, Jr., Picciano, D. J., and Jacobson, e. B. (1979). Y-chromosomal nondisjunction in dibromochloropropane-exposed workmen. Mutat. Res. 64,47-51. Kourakis, A., Mouratidou, M., Barbouti, A., and Dimikiotou, M. (1996). Cytogenetic effects of occupational exposure in the peripheral blood lymphocytes of pesticide sprayers. Carcinogenesis (Oxford) 17, 99-101. Kourakis, A., Mouratidou, M., Kokkinos, G., Barbouti, A., Kotsis, A, Mourelatos, D., and Dozi-Vassiliades, 1. (1992). Frequencies of chromosomal aberrations in pesticide sprayers working in plastic green houses. Mutat. Res. 279, 145-148. Kwok, E. S., and Eastmond, D. A (1997). Effects of pH on nonenzymatic oxidation of phenylhydroquinone: Potential role in urinary bladder carcinogenesis induced by o-phenylphenol in Fischer 344 rats. Chem. Res. Toxicol. 10,742-749. Lambert, A e., and Eastmond, D. A. (1994). Genotoxic effects of the o-phenylphenol metabolites phenylhydroquinone and phenylbenzoquinone in V79 cells. Mutat. Res. 322, 243-256. Lander, F., and Ronne, M. (1995). Frequency of sister chromatid exchange and hematological effects in pesticide-exposed greenhouse sprayers. Scand. 1. Work Environ. Health 21, 283-288. Larripa, 1., Matos, E., Vinuesa, M., and Salum, S. (1983). Sister chromatid exchanges in a human population accidentally exposed to an organophosphorus pesticide. Rev. Brasi!. Genet. 6, 719-727. Lebailly, P., Vigreux, e., Lechevrel, C., Ledemeney, D., Godard, T., Sichel, F., LeTaler, J. Y, Henry-Amar, M., and Gauduchon, P. (1998). DNA damage in mononuclear leukocytes of farmers measured using the alkaline comet assay: Discussion of critical parameters and evaluation of seasonal variations in relation to pesticide exposure. Cancer Epidemiology, Biomarkers and Prevention 7, 917-927. Linnainmaa, K (1983). Sister chromatid exchanges among workers occupationally exposed to phenoxy acid herbicides 2,4-D and MCPA. Teratogen. Carcinogen. Mutagen. 3, 269-279. Lucero, L., Pastor, S., Suarez, S., Durban, K, Gomez, e., Parron, T., Creus, A, and Marcos, R. (2000). Cytogenetic biomonitoring of Spanish greenhouse workers exposed to pesticides: Micronuclei analysis in peripheral blood Iymphocytes and buccal epithelial cells. Mutat. Res. 464, 255-262. Major, J., Kemeny, G., and Tompa, A (1992). Genotoxic effects of occupational exposure in the peripheral blood Iymphocytes of pesticide preparing workers in Hungary. Acta Medica Hungarica 49, 79-90. McCann, J., and Ames, B. N. (1976). Detection of carcinogens as mutagens in the Salmonella/microsome test: Assay of 300 chemicals: Discussion. Proc. Nat!. Acad. Sci. U.S.A. 73, 950-954. Mohammad, 0., Walid, A. A., and Ghada, K (1995). Chromosomal aberrations in human lymphocytes from two groups of workers occupationally exposed to pesticides in Syria. Environ. Res. 70, 24-29. Morimoto, K., Sato, M., Fukuoka, M., Hasegawa, K, Takahashi, T., Tsuchiya, T., Tanaka, A., Takahashi, A., and Hayashi, Y (1989). Correlation between the DNA damage in urinary bladder epithelium and the urinary 2-phenyl-

765

1,4-benzoquinone levels from F344 rats fed sodium o-phenylphenate in the diet. Carcinogenesis 10, 1823-1827. National Research Council (1999). "Hormonally Active Agents in the Environment." National Academy Press, Washington, De. National Toxicology Program (NTP) (1986). "Toxicology and Carcinogenesis Studies of ortho-Phenylphenol (CAS 90-43-7) Alone and with 7,12Dimethylbenz(a)anthracene (CAS 57-97-6) in Swiss CD-l Mice (Dermal Studies)." National Toxicology Program, Public Health Service, National Institutes of Health, U.S. Department of Health and Human Services, Research Triangle Park, Ne. National Toxicology Program (NTP) (1992). "NTP Technical Report on the Toxicology and Carcinogenesis Studies of Methyl Bromide (CAS 74-839) in B6C3Fl Mice (Inhalation Studies)." National Toxicology Program, Public Health Service, National Institutes of Health, U.S. Department of Health and Human Services, Research Triangle Park, Ne. National Toxicology Program (NTP) (2000). "Ninth Report on Carcinogens." National Toxicology Program, U.S. Department of Health and Human Services, Research Triangle Park, NC. Nehez, M., and Desi, 1. (1996). A genetic toxicological study of pesticide workers. Int. 1. Environ. Health Res. 6, 201-208. Nehez, M., Berencsi, G., Paldy, A., Czeizel, A, Szentesi, 1., Levay, K, Maurer, J., and Nagy, E. (1981). Data on the chromosome examinations of workers exposed to pesticides. Regu!. Toxico!. Pharmacol. 1, 116-122. Nehez, M., Boros, P., Ferke, A., Mohos, J., Palotas, M., Vetr6, G., Zimanyi, M., and Desi, 1. (1988). Cytogenetic examination of people working with agrochemicals in the southern region of Hungary. Regu!. Toxico!. Pharmacal. 8, 37-44. Neumann, H. G., Thielmann, H. W., Filser, J. G., Gelbke, H. P., Greim, H., Kappus, H., Norpoth, K H., Reuter, U., Vamvakas, S., Wardenbach, P., and Wichmann, H. E. (1998). Changes in the classification of carcinogenic chemicals in the work area (Section III of the German list of MAK and BAT values). 1. Cancer Res. Clin. Oncol. 124,661-669. Padmavathi, P., Aruna Prabhavathi, P., and Reddy, P. P. (2000). Frequencies of SCEs in peripheral blood Iymphocytes of pesticide workers. Bull. Environ. Contam. Toxicol. 64, 155-160. Padungtod, e., Hassold, T. J., Millie, E., Ryan, L. M., Savitz, D. A, Christiani, D. e., and Xu, X. (1999). Sperm aneuploidy among Chinese pesticide factory workers: Scoring by the FISH method. Am. 1. Ind. Med. 36, 230-238. Paldy, A, Puskas, N., Vincze, K, and Hadhazi, M. (1987). Cytogenetic studies on rural populations exposed to pesticides. Mutat. Res. 187, 127-132. Pasquini, R., Scassellati-Sforzolini, G., Angeli, G., Fatigoni, e., Monarca, S., Beneventi, L., DiGiulio, A. M., and Bauleo, F. A (1996). Cytogenetic biomonitoring of pesticide-exposed farmers in central Italy. 1. Environ. Patho!' Toxico!. Oncol. 15,29-39. Peluso, M., Merlo, F., Munnia, A., Bolognesi, C., Puntoni, K, and Parodi, S. (1996). (32)P-postlabeling detection of DNA adducts in peripheral white blood cells of greenhouse floriculturists from western Liguria, Italy. Cancer Epidemiology, Biomarkers and Prevention 5, 361-369. Petrelli, G., Siepi, G., Miligi, L., and Vineis, P. (1993). Solvents in pesticides. Scand.l. Work Environ. Health 19, 63-65. Pilinskaya, M. A. (1970). Chromosome aberrations in the persons contacted with ziram. Genetika 6, 157-163. Pilinskaya, M. A (1974). Results of cytogenetic examinations of people having professional contact with the fungicide zineb. Genetika 10, 140-146. Proctor, B. L., Gaulden, M. E., and Dowd, M. A. (1986). Reactivity and fate of benzene and formaldehyde in culture medium with and without fetal calf serum: Relevance to in vitro mutagenicity testing. Mutat. Res. 160, 259266. Rabello, M. N., Dealmeida, W F., Pigati, P., Ungaro, M. T., Murata, T., Perira, e. A, and Becak, W. (1975). Cytogenetic study on individuals occupationally exposed to DDT. Mutat. Res. 28, 449-454. Rehner, T. A, Kolbo, J. K, Trump, R., Smith, e., and Reid, D. (2000). Depression among victims of south Mississippi's methyl parathion disaster. Health and Social Work 25, 33. Reitz, K H., Fox, T. K, Quast, J. F., Hermann, E. A, and Watanabe, P. G. (1983). Molecular mechanisms involved in the toxicity of orthophenylphenol and its sodium salt. Chem.-Bio!. Interact. 43, 99-119.

766

CHAPTER 35


Rhomberg, L., Dellarco, V. L., Siegel-Scott, c., Dearfield, K. L., and JacobsonKram, D. (1990). Quantitative estimation of the genetic risk associated with the induction of heritable translocations at Iow-dose exposure: Ethylene oxide as an example. Environ. Mol. Mutagen. 16, 104-125. Rice, J. M., Baan, R. A., Blettner, M., Genevois-Charmeau, C., Grosse, Y., McGregor, D. B., Partensky, C., and Wilbourn, J. D. (1999a). Rodent tumors of urinary bladder, renal cortex, and thyroid gland in IARC Monographs evaluations of carcinogenic risk to humans. Toxicol. Sci. 49, 166-171. Rice, J. M., Venitt, S., and McGregor, D. (1999b). The use of short- and medium-term tests for carcinogens and data on genetic effects in carcinogenic hazard evaluation. fARC Sci. Publ. 536. Rita, P., Reddy, P. P., and Reddy, S. v. (1987). Monitoring of workers occupationally exposed to pesticides in grape gardens of Andhra Pradesh. Environ. Res. 44, 1-5. Roth, R. (1958). Concerning bronchial cancers in vine-growers injured by arsenic. Virchows Arch. 331, 119-137. Rupa, D. S., Eastmond, D. A., and Reddy, P. P. (1997). Detection of chromosomal alterations in the sperm of pesticide-exposed workers using fluorescence in situ hybridization (FISH). Environ. Mol. Mutagen. 29, 44. Rupa, D. S., Hasegawa, L., and Eastmond, D. A. (1995). Detection of chromosomal breakage in the Icen-Iql2 region of interphase human lymphocytes using multicolor fluorescence in situ hybridization with tandem DNA probes. Cancer Res. 55, 640-645. Rupa, D. S., Reddy, P. P., and Reddi, O. S. (1989a). Analysis of sister-chromatid exchanges, cell kinetics and mitotic index in lymphocytes of smoking pesticide sprayers. Mutat. Res. 223, 253-258. Rupa, D. S., Reddy, P. P., and Reddi, O. S. (1989b). Frequencies of chromosomal aberrations in smokers exposed to pesticides in cotton fields. Mutat. Res. 222, 37-41. Rupa, D. S., Reddy, P. P., and Reddi, O. S. (199Ia). Clastogenic effect of pesticides in peripheral lymphocytes of cotton-field workers. Mutat. Res. 261, 177-180. Rupa, D. S., Reddy, P. P., and Reddi, O. S. (199Ib). Reproductive performance in population exposed to pesticides in cotton fields in India. Environ. Res. 55, 123-128. Rupa, D. S., Reddy, P. P., Sreemannarayana, K., and Reddi, O. S. (199Ic). Frequency of sister chromatid exchange in peripheral lymphocytes of male pesticide applicators. Environ. Mol. Mutagen. 18, 136-138. Rupa, D. S., Rita, P., Reddy, P. P., and Reddi, O. S. (1988). Screening of chromosomal aberrations and sister chromatid exchanges in peripheral lymphocytes of vegetable garden workers. Hum. Taxieol. 7, 333-336. San, R. C. H., Rosin, M. P., See, R. H., Dunn, B. P., and Stich, H. F. (1989). Use of urine for monitoring human exposure for genotoxic agents. fn "Biological Monitoring for Pesticide Exposure-Measurement, Estimation, and Risk Reduction" (R. G. M. Wang, C. A. Franklin, R. C. Honeycutt, and J. C. Reinert, eds.), pp. 98-116. Am. Chem. Soc., Washington, DC. Sarto, F., Cominato, 1., Pinton, A. M., Brovedani, P. G., Faccioli, C. M., Bianchi, v., and Levis, A. G. (1984). Cytogenetic damage in workers exposed to ethylene oxide. Mutat. Res. 138, 185-195. Sasaki, Y. F., Saga, A., Akasaka, M., Yoshida, K., Nishidate, E., Su, Y. Q., Matsusaka, N., and Tsuda, S. (1997). fn vivo genotoxicity of orthophenylphenol, biphenyl, and thiabendazole detected in multiple mouse organs by the alkaline single cell gel electrophoresis assay. Mutat. Res. 395, 189-198. Scarpato, R., Migliore, L., Hirvonen, A., Falck, G., and Norppa, H. (1996). Cytogenetic monitoring of occupational exposure to pesticides: Characterization of GSTMl, GSTTl, and NAT2 genotypes. Environ. Mol. Mutagen. 27,263-269. Schmid, R. (1960). Cutaneous porphyria in Turkey. N. Engl. J. Med. 263, 397398. Scott, D., Galloway, S. M., Marshall, R. R., Ishidate, M., Jr., Brusick, D., Ashby, J., and Myhr, B. C. (1991). International Commission for Protection against Environmental Mutagens and Carcinogens. Genotoxicity under extreme culture conditions. A report from ICPEMC Task Group 9. Mutat. Res. 257, 147-205.

Simon, G. S., Tardiff, R. G., and Borzelleca, J. F. (1979). Failure of hexachlorobenzene to induce dominant lethal mutations in the rat. Toxicol. Appl. Pharmacol. 47, 415-419. Smith, A. G. (1991). Chlorinated hydrocarbon insecticides. fn "Handbook of Pesticide Toxicology" (w. J. J. Hayes and E. R. J. Laws, eds.), pp. 731915. Academic Press, San Diego. Smith, R. A., Christenson, W. R., Bartels, M. J., Arnold, L. L., St. John, M. K., Cano, M., Garland, E. M., Lake, S. G., Wahle, B. S., McNett, D. A., and Cohen, S. M. (1998). Urinary physiologic and chemical metabolic effects on the urothelial cytotoxicity and potential DNA adducts of o-phenylphenol in male rats. Toxicol. Appl. Pharmacol. 150,402-413. Sorsa, M., Wilbourn, J., and Vainio, H. (1992). Human cytogenetic damage as a predictor of cancer risk. fARC Sci. Publ. 543-554. Steenland, K., Carrano, A., Clapp, D., Ratcliffe, J., Ashworth, L., and Meinhardt, T. (1985). Cytogenetic studies in humans after short-term exposure to ethylene dibromide. 1. Occup. Med. 27, 729-732. Steenland, K., Carrano, A., Ratcliffe, J., Clapp, D., Ashworth, L., and Meinhardt, T. (1986). A cytogenetic study of papaya workers exposed to ethylene dibromide. Mutat. Res. 170, 151-160. Suzuki, H., Suzuki, N., Sasaki, M., and Hiraga, K. (1985). Orthophenylphenol mutagenicity in a human cell strain. Mutal. Res. 156, 123-127. Tadi-Uppala, P., Hasegawa, L., Rupa, D. S., and Eastmond, D. A. (1996). Detection of micronuclei and cell proliferation in the rat bladder induced by the fungicides o-phenylphenol and sodium orthophenylphenate. fn "Proceedings of the American Association for Cancer Research Annual Meeting," Vo!. 37, pp. 127-128. Tayama, S., and Nakagawa, Y. (1991). Sulfhydryl compounds inhibit the cytoand geno-toxicity of o-phenylphenol metaboIites in CHO-Kl cells. Mutat. Res. 259, 1-12. Tayama, S., Kamiya, N., and Nakagawa, Y. (1989). Genotoxic effects of o-phenylphenol metabolites in CHO-KI cells. Mutat. Res. 223, 23-33. Tayama-Nawai, S., Yoshida, S., Nakao, T., and Hiraga, K. (1984). Induction of chromosome aberrations and sister-chromatid exchanges in CHO-KI cells by o-phenylpheno!. Mutat. Res. 141,95-99. Thiers, H., Colomb, D., Moulin, G., and Colin, L. (1967). Cutaneous arsenical cancer in viticultivators in Beaujolais. Ann. Dermatol. Syphiligr. (Paris) 94, 133-158. Titenko-Holland, N., Windham, G., Kolachana, P., Reinisch, F., Parvatham, S., Osorio, A. M., and Smith, M. T. (1997). Genotoxicity of malathion in human lymphocytes assessed using the micronucleus assay in vitro and in vivo: A study of malathion-exposed workers. Mutat. Res. 388, 85-95. Tucker, J. D., Eastmond, D. A., and Littlefield, L. G. (1997). Cytogenetic endpoints as biological dosimeters and predictors of risk in epidemiological studies. fARC Sci. Publ. 142, 185-200. Ushiyama, K., Nagai, F., Nakagawa, A., and Kano, 1. (1992). DNA adduct formation by o-phenylphenol metabolite in vivo and in vitro. Carcinogenesis 13, 1469-1473. van Bao, T., Szabo, 1., Ruzicska, P., and Czeizel, A. (1974). Chromosome aberrations in patients suffering acute organic phosphate insecticide intoxication. Humangenetik 24, 33-57. Venegas, W., Zapata, 1., and Marcos, R. (1998). Micronuclei analysis in lymphocytes of pesticide sprayers from Concepcion, Chile. Teratogen. Carc'inogen. Mutagen. 18, 123-129. Volnjanskaya, V. (1981). Level of chromosome aberrations in agricultural workers. Gigiena Truda i Professional'nye Zabolevaniia 12,47-48. Waters, M. D., Stack, H. F., Garrett, N. E., and Jackson, M. A. (1991). The Genetic Activity Profile database. Environ. Health Perspect. 96, 41-45. Waters, M. D., Stack, H. F., and Jackson, M. A. (1999). Short-term tests for defining mutagenic carcinogens. fARC Sci. Publ. 499-536. Wei, L. Y., Chao, J. S., and Hong, C. C. (1997). Assessment of the ability of propoxur, methomyl, and aldicarb, three carbamate insecticides, to induce micronuclei in vitro in cultured Chinese hamster ovary cells and in vivo in BALB!c mice. Environ. Mol. Mutagen. 29, 386-393. Whorton, D., MiIby, T. H., Krauss, R. M., and Stubbs, H. A. (1979). Testicular function in DBCP exposed pesticide workers. J. Occup. Med. 21, 161-166. Wild, C. P., and Pisani, P. (1997). Carcinogen-DNA and carcinogen-protein adducts in molecular epidemiology. fARC Sci. Publ. 143-158.

References

Wiltse, J., and Dellarco, V. L. (1996). U.S. Environmental Protection Agency guidelines for carcinogen risk assessment: Past and future. Mutat. Res. 365, 3-15. Yoder, J., Watson, M., and Benson, W. W. (1973). Lymphocyte chromosome analysis of agricultural workers during extensive occupational exposure to pesticides. Mutat. Res. 21, 335-340.

767

Zahm, S. H., and Ward, M. H. (1998). Pesticides and childhood cancer. Environ. Health Perspect. 106(suppl. 3), 893-908. Zahm, S. H., Ward, M. H., and Blair, A. (1997). Pesticides and cancer. Occup. Med. 12,269-289.

CHAPTER

36 Immunotoxicity of Pesticides Kathleen E. Rodgers University of Southern California

36.1 INTRODUCTION

36.1.2 PESTICIDES

36.1.1 IMMUNE SYSTEM The immune system, the system by which foreign invaders are controlled or eliminated, contains two components: innate and acquired immunity. Innate immunity provides protection in an nonspecific manner and does not have a lag time before protection is conferred. Innate immunity consists of barriers, such as skin and mucous membranes, neutrophils, cells of the macrophage lineage, and natural killer cells. The last three elements are involved in the release of inflammatory mediators, such as enzymes and bioactive lipids, cytokines, which can stimulate cellular function or act to eliminate the invader directly, and reactive oxygen and nitrogen intermediate, which act together to inhibit cellular proliferation or are cytotoxic. The specific immune response requires a lag time for the generation of the response and has memory that allows a more intense, rapid, and specific response upon reexposure. There are two arms of the specific immune response, cellular immunity and humoral immunity. The effectors of cellular immunity are cytotoxic T cells that kill virally infected cells or tumor cells by direct contact. The effectors of humoral immunity are antibodies, which are generated by B cells and eliminate antigens by the formation of immune complexes, by complement fixation, or by enhancement of phagocytosis (opsonization) or antibody-directed cellular cytotoxity. The generation of cellular and humoral immune responses is regulated by suppressor T cells and helper T cells. Subsets of helper T cells have been defined which elaborate different cytokines, express different receptors, and are instrumental in the support of either cellular or humoral immunity depending upon the cell type. The optimal functioning of the immune system is under exquisite regulation that allows elimination of the foreign invader without a great deal of detriment to the host. Immune disorders arise when this balance is disrupted. Handbook of Pesticide Toxicology Volume 1. Principles

Pesticides are xenobiotics which are by definition biocidal and are designed to be selective to the species to be killed through metabolism or targeting of a site that is specific to the target organism. However, this is not accomplished for most pesticides and the safety of the pesticide is based on the quantity of chemical used and the method of application. Therefore, there are reasons for unique concern regarding the toxicity of pesticides and consideration as to the dose of a pesticide to which a test animal is exposed. For most pesticides, a marker for poisoning has been established based on the most sensitive parameter of toxicity in laboratory animals that is readily measured in humans. Although exposure to a dose of compound sufficient to cause acute poisoning may have an effect on the immune system, many studies have examined the effects of pesticides on the immune system, with particular emphasis on the administration of a nontoxic dose of compound. By administration of a nontoxic dose of pesticide, one can avoid the complications of the effects of stress and toxicities to other physiologic systems on the function of the immune system. The influence of pesticides on the immune response in humans has largely been ignored. The immune-mediated complications of pesticide exposure to humans that are most often noted are allergic reactions, especially contact dermatitis. This is probably due to the fact that an allergic reaction is readily observed and can be attributed to the causative agent through measurement of immune components specific for this agent. Although a great deal of data are available from animal studies which show that pesticides are immunosuppressive, there is no good evidence at this time for immune suppression in the general popUlation as a result of environmental exposure. This may be due to the lack of well-designed longitudinal studies to examine this potential effect of pesticides. To fully delineate the effects of pesticides on the generation of immune responses, sensitive tests (i.e., those that would measure an alteration in a functional immune response and not simply basal immune function) with low interassay, day-to-day, and person-to person variability should be conducted on persons occupationally ex-

769


770

CHAPTER 36

Immunotoxicity of Pesticides

posed to well-known levels and compositions of pesticides or those exposed to high levels after accidental exposures. These studies should be longitudinal to allow the determination of the kinetics of potential recovery and to eliminate the possibility that initial observations of variations from the norm were simply due to intrinsic variabilities within the measurements. The roadblocks to conducting such experiments include (1) lack of appropriate control groups, (2) inherent day-to-day and personto-person variability in commonly used assays of immune responsiveness, and (3) lack of knowledge of the physiologic significance of alterations in these parameters.

36.2 ORGANOCHLORINE PESTICIDES Organochlorine insecticides include many classes of compounds, such as chlorinated ethane derivatives (DDT), cyclodienes (chlordane, aldrin, heptachlor), and hexachlorocyclohexanes (lindane). These insecticides were widely used from the mid-1940s to the mid-1960s in the control of insects in agriculture, soil, and structures. This class of compound is less acutely toxic than anticholinesterase pesticides, but has a greater potential for chronic toxicity. 36.2.1 DDT The effects of dichlorodiphenyltrichloroethane (DDT) on the immune system were studied in several species using a variety of functional parameters. In most of these studies described here, DDT was given repeatedly to mimic the persistence of the chemical in the environment. Administration of 100 ppm orally of DDT to chickens for 40 days led to a decrease in the weight of lymphoid organs, including the spleen, thymus, and bursa of Fabricus (Subba Rao and Glick, 1977). In this same study, the humoral immune response to sheep red blood cells (SRBCs), a T-cell-dependent antigen, was also studied, but no effect on this parameter was noted. In another study, the humoral immune response to bovine serum albumin (BSA) after oral exposure to 100-400 ppm DDT orally for 5 weeks was shown to be decreased (Glick, 1974). One study examined the effects of DDT on the immune system of the guinea pig following a single intraperitoneal administration of 15 mg/kg DDT. Following this treatment, the antidiphtheria hemagglutinin titer was unchanged, but anaphylactic shock was decreased (Gablicks et aI., 1973). Studies also showed that the level of protein in the diet affected the ability of exposure to DDT to result in immunosuppression with greater than 3% protein being protective (Banerjee et aI., 1995). Further, combined exposure to DDT and stress will enhance the observed immunosuppression (Banerjee et aI., 1997). Two studies examined the effects of DDT on the cellular and humoral immune response in rats. Administration of 200 ppm DDT orally for 35 days decreased the response to ovalbumin (Wassermann et aI., 1969). In contrast, administration of 40 mg/kg/day DDT orally for 60 days increased both the humoral immune response and the delayed-type hypersensitivity

(DTH) response to BSA (Luki et aI., 1973). Administration of 200 ppm DDT or DDT metabolites (DDE and DDA) in their diet for 5 weeks suppressed the generation of both humoral immune and cell-mediated responses to ovalbumin (Banerjee et aI., 1996). DDT (0.25 mg/kg/day orally for 1 month) had no effect on the phagocytosis of bacteria by polymorphonuclear neutrophils (PMN) (Crocker et aI., 1969). Administration of up to 150 ppm DDT orally to rabbits for 4 weeks had no effect on the humoral immune response to SRBC. Only the high dose of 150 ppm affected the DTH response to tuberculin in this study (Andre et al., 1983). Administration of 20 or 200 pm DDT orally for 1 month decreased anaphylactic shock symptoms (Gablicks et aI., 1975). Only a few studies have examined the effects of DDT on the human immune response. In vitro exposure of human peripheral blood mononuclear cells (PBMC) or PMN had no effect on the mitogenic response to phytohemagglutinin, but slightly decreased the chemotactic response (Lee et aI., 1979). On the other hand, occupational exposure to DDT is thought to depress the PMN function as measured by chemotaxis, nitro blue tetrazolium (NBT) reduction test, and phagocytosis (Hermanowicz et aI., 1982). In addition, there was an increase in the incidence of infections in the occupationally exposed group. On the other hand, no correlation was found between DDT blood levels and the ability to respond to diphtheria immunization in children (Costa and Schvartsman, 1977). Others have suggested that DDT may induce allergic contact dermatitis in humans (Vanat and Vanat, 1971). In summary, DDT has been shown to increase, have no effect on, or decrease the immune system depending upon the dose, route, timing of administration, species, and antigen. 36.2.2 CHLORDANE Most studies of the effects of chlordane on the immune system were performed in mice. Adult mice are relatively resistant to the immunotoxic effects of chlordane. The generation of a humoral immune response was unaffected by administration of 0.1-8 mg/kg/day for 14 days. In this same study, the generation of some cell-mediated immune responses studied was elevated by this treatment regime (John son et aI., 1986). On the other hand, in vitro exposure of murine splenocytes from adult mice to chlordane resulted in a dose-dependent suppression of immune response at doses that reduced cell viability in culture (Johnson et aI., 1987). In contrast, the murine immune system was found to be most sensitive to the immunotoxic effects of chlordane when exposure occurred in utero. BALB/C mice were exposed in utero to 0.16 or 8 mg/kg/day chlordane given to the dam throughout gestation. These treatments did not affect the generation of a humoral immune response to SRBC, but the high dose did inhibit the DTH response to oxazolone (Menna et aI., 1985; Spyker-Crammer et aI., 1982). In studies using a similar treatment regime, alterations were found in the following immune parameters at one or more of the doses administered: natu-

36.2 Organochlorine Pesticides

ral killer (NK) activity, bone marrow hematopoiesis [as measured by colony-forming units-granulocyte macrophage (CFUGM) and colony-forming units-stem cell (CFU-S)], fetal liver hematopoiesis (as measured by CFU-GM and CFU-S), and the DTH response to influenza (Barnett et al., 1985a, b, 1990a, b; Blaylock et al., 1990; Chuang et al., 1992). Some of these alterations occurred up to 200 days after birth. No effects on the generation of a cytotoxic T-lymphocyte (CTL) response or the proliferative responses to T- and B-cell mitogens were observed (Barnett et al., 1985a, b). Further studies have shown that in utero exposure to chlordane led to enhanced peritoneal macrophage function similar to that observed with inflammatory stimuli (i.e., decreased 5' nucleotidase and transferrin receptor expression and protein synthetic pattern similar to inflammatory macrophages) (Theus et al., 1992). In this study, no additional effects on macrophage function were observed in chlordane-treated mice upon stimulation with thioglycollate. In vitro exposure of guinea pig PMN to chlordane resulted in the stimulation of many membrane-related events, including respiratory burst, membrane potential, calcium mobilization, and the release of membrane-bound calcium (Suzald et al., 1988). In humans, there has been a tentative association made between exposure to chlordane and blood dyscrasias (AMA Council on Drugs, 1962; Furie and Trubowitz, 1976; Infante etal., 1976;Stieglitzetal., 1967). In summary, the immune system was most sensitive to chlordane-induced immunotoxicity when exposure occurred in utero. Alterations also occurred following in vitro exposure of white blood cells to chlordane.

36.2.3 DIELDRIN The humoral immune response to bacterial antigen was decreased by oral administration of 50 ppm dieldrin to rabbits (Wassermann et al., 1969). In addition, administration of 2040 ppm dieldrin to ducks decreased the resistance to hepatitis virus (Friend and Trainer, 1974). Others have studied the effects of dieldrin on the murine immune system. In this model, dieldrin was administered in corn oil by intraperitoneal injection for up to 16 days. Antiviral resistance to the mouse hepatitis virus 3 (MHV3) was decreased after exposure to dieldrin (Krzystyniak et al., 1985, 1986). Dieldrin acted to increase host susceptibility to virus through suppression of the humoral immune responses and alterations in cell-mediated immunity (Bernier et aI., 1987; Hugo et aI., 1988a, b; Krzystyniak et aI., 1985). However, studies showed that the use of a pathogenic antigen, which could itself suppress the immune response, in the process of examining the immunotoxic effects of dieldrin did not contribute to the observed immune suppression (Fournier et aI., 1988). The authors suggest that the alterations in cell-mediated immunity may occur at the level of antigen recognition rather than proliferation of the lymphocytes in response to stimulation (Hugo et aI., 1988a, b).

771

Alterations in macrophage function that may contribute to the observed reduction in host resistance to MHV3 were observed (Krzystyniak et aI., 1986, 1987). Specifically, antigen presentation and phagocytosis were altered by exposure to dieldrin (Bernier et aI., 1988; Kaminski et aI., 1982; Krzystyniak et aI., 1985, 1989). In these studies, it was shown that, after exposure to dieldrin, the quantity of antigen taken up, cell associated, and released by macrophages was suppressed. These alterations in macrophage function and the decrease in the ability to generate a humoral immune response may occur through alteration in membrane integrity and fluidity (Alberts et aI., 1965; Antunes-Madeira and Madiera, 1979).

36.2.4 HEPTACHLOR Administration of 1 ppm of heptachlor to chickens for 3-8 weeks decreased the weight of the bursa of Fabricus (lymphoid organ in the chicken) (Rodica and Stefania, 1973). In addition, administration of heptachlor to rats decreased the levels of serum gamma globulins (Klimova, 1970). In vitro exposure of guinea pig PMNs to heptachlor and heptachlor epoxide induced the generation of superoxide anion, altered the membrane potential, induced calcium mobilization, and induced the release of membrane-bound calcium (Suzald et aI., 1988).

36.2.5 LINDANE Very few studies of the effect of lindane on the immune response have been performed. Administration of 150 ppm lindane per day orally for 1 month to mice did not alter the levels of IgA, IgG 1, IgG2a, and IgM in the serum or the ability to generate a humoral immune response to SRBC (Andre et aI., 1983). However, the serum levels of IgG2b were elevated by this exposure. In the rabbit, administration of 3 mg/kg/day or more lindane for 5 weeks decreased the humoral immune response to bacterial antigen (Desi et aI., 1978; Kaliser, 1968). In this study, the no-observable-adverse-effect level (NOAEL) for the immunotoxic effects of lindane was shown to be 1.5 mg/kg/day. In vitro exposure of human PBMC to 0.1-0.3 mM lindane inhibited the mitogenic response to a T-cell mitogen (Roux et aI., 1979). Further studies were conducted on the immunotoxicity of ,B-hexachlorocyclohexane (HCH), an isomeric contaminant of lindane, in mice. Mice were given up to 300 ppm HCH in their diet for 30 days. At 300 ppm, but not 100 ppm, suppression of the ability of splenocytes to proliferate in response to mitogen to generate a CTL response and NK activity was suppressed (Cornacoff et aI., 1988). As with chlordane, there is circumstantial evidence for the association of lindane with aplastic anemia and agranulocytosis, as well as other blood dyscrasias, in humans. However, the incidence of this type of toxicity associated with lindane exposure is very low (Gewin, 1959; Jedlicka et al., 1958; Loge, 1965; Mastromattco, 1964; Samuels and Milby, 1971; Sianchez-Madel et aI., 1963; West, 1967).

772

CHAPTER 36


36.2.6 MIREX Administration of 100 ppm mirex orally to chickens for 40 days decreased the weight of the thymus and the spleen, but increased the weight of the bursa of Fabricus (Subba Rao and Glick, 1977). However, administration of at least 500 ppm mirex orally for 5 weeks was required to inhibit the generation of a humoral immune response to BSA (Roux et al., 1979). 36.2.7 TOXAPHENE One study was conducted to examine the effects of toxaphene on the immune system. In this study, administration of 100200 ppm either to adult animals or in utero decreased the humoral immune response to BSA and phagocytosis by peritoneal macrophages, but had no effect on the DTH response to tuberculin antigen (AlIen et aI., 1983). 36.2.8 ENVIRONMENTAL EXPOSURE Recent studies have evaluated the effect of living in a contaminated environment on the immune function of marine mammals. Recent mass mortality among several populations has lead to the hypothesis of increased susceptibility to viral and opportunistic infections. In these studies, harbor seals were fed fish from either contaminated or noncontaminated sources over 2.5 years. Seals fed contaminated fish were suppressed in both humoral and cell-mediated immune responses (De Swart et aI., 1996; Ross et aI., 1995. Further studies in free-ranging dolphins showed that an inhibition of proliferative responses was correlated with concentrations of organochlorines in the blood (Lahvis et aI., 1995). 36.2.9 SUMMARY In summary, the effects of a number of organochlorine pesticides on the immune system have been examined. For the most part, these compounds suppressed the generation of immune responses and reduced the resistance of the test animal to infections. Several laboratories have shown that these compounds alter membrane fluidity and modulate membranemediated events. In fact, those who suggest a mechanism of action for the immunotoxic effects of organochlorine pesticides refer to this as a possibility.

36.3 ANTICHOLINESTERASES 36.3.1 ORGANOPHOSPHATES ESTERS Organophosphate pesticides are widely used compounds in agriculture, by consumers, and in public health situations due to their relatively low toxicity, rapid removal from the environment, and lack of bioaccumulation. However, organophosphates have the dubious distinction of being the pesticides most

often implicated in poisonings due to pesticides (parathion contributing in large part to this). Organophosphates of low mammalian toxicity, such as malathion, are used in situations such as structural treatment for mosquito eradication, spraying of tobacco plants, and aerial spraying of urban populations for eradication of fruit flies. Therefore, there is potential for both high exposure to organophosphate pesticides during occupational exposure and low-level exposure of large segments of the general population. A great deal of information is available with regards to the effects of organophosphates on the immune system. 36.3.1.1 Parathion The effects of parathion on the immune system have been extensively studied. Wiltrout et al. (1978) showed that subacute administration of parathion (2.2-22.3 mg/kg/day) to mice blocked the generation of a humoral immune response (Wiltrout et aI, 1978). In addition, Dandliker et al. (1985) demonstrated that parathion was able to suppress both cellular and humoral immunity (Dandliker et aI., 1985). Alternatively, one study showed a decrease in the lymphoid organ weight with no change in the humoral immune response following repeated exposure to parathion. Others showed that parathion suppressed the humoral immune response following acute, subacute, and in vitro exposure (Bartholomew et aI., 1984; Casale et aI., 1983; Duggan et al., 1984). Peroral dosing of parathion to mice with a cytomegalovirus infection elevated mortality (Raise, 1983). The proliferative response of human lymphocytes to mitogens was suppressed following in vitro exposure to paraoxon (Waterhouse and Tourney, 1984). Paraoxon and two structurally related compounds inhibited the production of interleukin 2 by rat splenocytes (Pruett and Chambers, 1988). The in vitro exposure (1-125 J.-Lg/ml) of splenocytes to parathion and methyl parathion blocked the generation of a cell-mediated immune response (Rodgers et aI., 1986b). Methyl parathion (up to 3 mg/kg/day) increased the virulence of Salmonella typhimurium infection in rabbits (Fan, 1981; Fan et aI., 1984). In another study, methyl parathion administered over 4 weeks (up to 1/10 LDso) did not affect the generation of humoral or cellular immune responses in rabbits (De si et aI., 1978). Administration of a single high dose or repeated lower doses of methyl parathion to mice elevated the humoral immune response with no effect on DTH reaction (Institoris et al., 1992). Administration of 0.22-0.44 mglkg methyl parathion over three generations was studied. Alterations in immune function were detected at a dose of 0.29 mg/kg methyl parathion, but the parameters altered varied between generations (Institoris et aI., 1995). In vitro exposure of human PBMC to methyl parathion did not affect the proliferative response to mitogen, but decreased the chemotactic response (Lee et aI., 1979). Studies were also conducted as to the effects of parathion on hematopoiesis. Gallichio et al. Gallichio et al. showed that oral administration of parathion for 14 days, at a dose that did not affect the body weight or generate cholinergic symptoms, altered

36.3 Anticholinesterases the bone-marrow-derived stem cell colonies for up to 2 weeks after the last dose of parathion (Gallichio et aI., 1987a). In vitro exposure of human bone marrow cells to paraoxon or malaoxon significantly depressed the in vitro generation of colonies of burst-forming units-erythroid (BFU-E), colony-forming uniterythroid (CFU-E), and CFU-GM (Gallichio et aI., 1987b). These studies are difficult to correlate due to the differences in exposure route, immune parameters studied, and the species studied. In general, however, parathion was shown to be immunosuppressive, but the mechanism of these effects is unknown. 36.3.1.2 Malathion There have been several studies on the effects of malathion on the immune response. Repeated exposure to malathion results in allergic responses in man, guinea pigs, rabbits, rats, and mice (Centeno et aI., 1970; Cushman and Street, 1983; Hazelton, 1992; Magnusson and Kligman, 1987; Milby and Epstein, 1964; Vijay et aI., 1978). In contrast, a DTH response to malathion (up to 100 Il-g) was not observed in mice and guinea pigs (Cushman and Street, 1983; Kynoch and Smith, 1992). Administration of low doses of malathion for prolonged periods results in a decrease in the humoral immune response. For example, low doses of malathion (up to 50 mg/kg) given for 5-6 weeks to rabbits significantly lowered the humoral immune response to bacterial antigen (Desi et aI., 1986). In addition, a cholinergic dose of malathion suppressed the generation of a humoral immune response, whereas mUltiple low doses did not (Casale et aI., 1983). In contrast, exposure of mice to malathion dip (2 or 8%) or chickens to 400-1600 ppm malathion in their diet did not alter the generation of humoral immunity (Relford et aI., 1989; Varshneya et aI., 1988). Administration of high, noncholinergic doses of malathion (up to 715 mg/kg) to mice elevated the generation of a humoral immune response and proliferative responses to mitogen. Acute or subacute administration of malathion did not affect the generation of a CTL response to allogeneic tumor (Rodgers et al., 1986c). In vitro exposure of human PBMC or mouse splenocytes to malathion suppressed the proliferative response to mitogens and the generation of hydrogen peroxide (Rodgers et aI., 1986a; Rodgers and Ellefson, 1990a). In mouse splenocytes, the generation of CTL responses was also blocked by in vitro exposure to malathion (up to l25Il-g/kg). Further studies indicated that, when malathion was coincubated with liver enzymes to allow metabolism, the metabolites of malathion were no longer able to block the generation of a CTL response or the proliferative response to mitogens (Rodgers et aI., 1985a, 1986b; Rodgers and Ellefson, 1990a). In contrast, in vitro exposure of murine peritoneal cells or human PBMC to metabolized malathion elevated the respiratory burst activity of these cells (Rodgers and Ellefson, 1990a). Further studies were conducted to determine the mechanism of action of malathion on the immune system. Cell separation and reconstitution experiments after acute administration of high, noncholinergic doses of malathion (up to

773

715 mg/kg) showed that the macrophages were the cell type affected by malathion (Rodgers and Ellefson, 1990a). Further studies showed that acute administration of malathion elevated the respiratory burst of peritoneal leukocytes. The lowest observable adverse effect level (LOAEL) and NOAEL for the effect of acute administration of malathion on the respiratory burst of peritoneal leukocytes were shown to be 0.25 and 0.1 mg/kg malathion, respectively (Rodgers and Ellefson, 1992). Further studies in animals administered malathion for 14 or 90 days showed systemic degranulation of basophilic cells and macrophage activation (Rodgers and Xiong, 1997a, b). Microscopic examination of the peritoneal cells showed that peritoneal mast cells were degranulated within 4 hr after malathion administration. In addition, the percentage of peritoneal phagocytes ingesting mast cell granules and the number of granules ingested per cell were elevated. Further, the systemic release of mast cell mediators, ,B-hexosaminidase and histamine, was observed after oral administration of malathion (Rodgers and Ellefson, 1992; Rodgers and Xiong, 1998). This exposure to mast cell products may elevate macrophage function. Studies involving mast cell-deficient mice showed that the presence of mast cells was necessary for an elevation in macrophage and immune function after malathion was administered (Rodgers, 1997). More recently, it was shown that the release of mast cell mediators, both inflammatory mediators and histamines, contributes to alterations in macrophage function that occur after oral administration of malathion (Rodgers and Xiong, 1996, 1997c). In vitro exposure of rat basophilic leukemia (RBL-l) cells to paraoxon, but not parathion, and malathion caused degranulation of these cells (Rodgers and Ellefson, 1992). Further studies on purified metabolites of malathion showed that degranulation of both normal basophilic cells (rat and human) and the RBL-l tumor cell line occurred after in vitro exposure to not only the potent inhibitors of anticholinesterase, malaoxon and isomalathion, but also the nonneurotoxic metabolite, dicarboxylic acid, of malathion (Xiong and Rodgers, 1997). This may explain why immunologic effects were observed at doses two to three orders of magnitude lower than the noncholinergic dose of malathion. These data, together with the report in the literature that diisopropyl fluorophosphate (DFP) and soman cause mast cell degranulation in an IgE-independent manner, suggest that malathion or a metabolite of malathion may act to elevate the immune response through inhibition of a cell surface-serine esterase on mast cells, subsequent degranulation of mast cells, and exposure of macrophages to mast cell products (Kazimierczak et aI., 1984). Recent studies have shown that weekly administration of relatively high doses of malathion (33-100 mg/kg) will accelerate the onset of autoimmune disease in mice predisposed to SLE, but not in their littermates which do not have a gene for accelerated onset of autoimmune disease (Rodgers, 1997). 36.3.1.3 Effects of Impurities in Organophosphate Pesticides The effects of impurities in organophosphate pesticides (malathion, acephate, and fenitrothion) on the immune system have

774

CHAPTER 36


been studied. The maJonty of the studies were conducted with O,O,S-trimethyl phosphorothioate (OOS-TMP). Following acute, nontoxic doses of OOS-TMP (up to 10 mg/kg), the generation of both cell-mediated and humoral immunity was blocked following in vivo and in vitro exposure to antigen (Devens et aI., 1985; Rodgers et aI., 1985a, 1986b). However, OOS-TMP did not affect the proliferative response to mitogens, but did elevate the production of interleukin 2 (IL-2) (Rodgers et aI., 1986a). More recent studies have shown that 0,0,0trimethyl phosphorothioate (OOO-TMP) (up to 40 mg/kg) elevated humoral and cell-mediated immune responses and protected against OOS-TMP-induced immune suppression when coadministered with OOS-TMP (Rodgers et aI., 1989a). In addition, exposure to low levels of OOS-TMP protected against an immunosuppressive dose of OOS-TMP. Subchronic (14-day) exposure to OOS-TMP increased the humoral and cell-mediated immune responses, mitogenic responses, and IL-1 production (Rodgers et aI., 1985b). The suppression of the immune response following acute administration of OOS-TMP (up to 10 mg/kg) was dose and time dependent, and macrophages were shown to be the cell type most affected by OOS-TMP (Devens et aI., 1985; Rodgers et aI., 1985c). Macrophages from OOS-TMP-treated mice were shown to (1) be larger in size, (2) have increased nonspecific esterase activity, (3) be less effective at antigen presentation, (4) have increased phagocytic activity, (5) secrete increased levels oflL-l, (6) have decreased la and F4/80 expression, (7) release suppressive factors, (8) have increased respiratory burst activity, and (9) secrete increased levels of neutral proteases, plasminogen activator, elastase, and collagenase (Rodgers and Ellefson, 1988a, b, 1990b; Rodgers et aI., 1985c, d, 1987b). These effects were transient and macrophage function was comparable to controls within 7 days (at which time immune function was similar to controls) (Rodgers and Ellefson, 1988b, 1990b). OOS-TMP also caused thymic atrophy (Devens et al., 1985). More recent investigations have shown a reduction in the number of cells expressing T-cell markers in the thymus (Rodgers et aI., 1987a-c). Studies are ongoing to determine the identity of mouse thymus cells which are targeted following acute administration of OOS-TMP. In summary, acute in vivo administration of OOS-TMP was immunosuppressive while stimulating macrophage function; repeated exposures stimulated cellular and humoral immune responses. In vivo exposure to 0 ,S,S,-trimethyl phosphorodithioate enhanced or suppressed the generation of cell-mediated or humoral immune response at nontoxic or toxic (assessed by suppression of plasma cholinesterase) doses, respectively, following in vivo or in vitro stimulation with antigen (Rodgers et aI., 1987c, 1988b). In vivo exposure to OSS-TMP also elevated proliferative responses to mitogens, but suppressed IL-2 production (Rodgers et aI., 1988b). Fourteen-day exposure to OSS-TMP (20 or 40 mg/kg) elevated or suppressed (60 or 80 mg/kg), depending upon the dose, the generation of immune responses (Rodgers et aI., 1989b). Further studies showed that OSS-TMP altered both T- and B-lymphocyte function (Thomas

and Imamura, 1986). In vitro exposure to OSS-TMP enhanced or suppressed immune function, depending upon the OSS-TMP concentration and the in vitro metabolism system used (Rodgers et aI., 1988b; Thomas and Imamura, 1986). OSS-TMP also inhibited the cytolytic function of cloned murine and human CTL, but only if present during the time when the cell to be lysed was being recognized (as measured by conjugation) by the CTL (Rodgers et aI., 1988a). OSS-TMP was immunostimulatory at noncholinergic doses and immunosuppressive at cholinergic doses (similar to that described previously for malathion). In vitro exposure to OSS-TMP suppressed humoral and cellmediated immune function. Finally, one study showed that O,O-dimethyl, S-ethyl phosphorothioate, a synthetic analog of the impurities described previously, blocked cell-mediated and humoral immune response after in vitro exposure through impairment of lymphocyte function (Thomas et aI., 1986). These studies show that the impurities found in technical malathion can modulate immune function. The cell type affected, the duration of the effect, and the immune parameter modulated varied from compound to compound and were related to the duration of exposure. 36.3.1.4 Other Organophosphate Pesticides The effects of many organophosphates have been assessed on at least one facet of the immune system. Carbophenothion and crufomate suppressed the proliferation of human lymphocytes in response to mitogen (Park and Lee, 1978). Acute administration of dichlorvos slightly decreased splenic weight of mice, but did not affect the generation of a humoral immune response (Cas ale et aI., 1983). In addition, chronic, low-level exposure to dichlorvos (up to 1/10 LD50) suppressed the generation of serum antibodies following vaccination of rabbits with Salmonella typhimurium and the generation of cell-mediated immunity following a tuberculin vaccination (De si et aI., 1978, 1979). Acute administrations of cholinergic doses of dichlorvos (up to LD50) resulted in mobilization of bone marrow cells and suppression of cellular and humoral immune responses. The mediator (corticosterone or acetylcholine) associated with these alterations varied with the parameter measured (Zabrodski, 1993). In vitro treatment of rabbit PMN by diisopropylethyl phosphate or triisopropyl phosphate reduced locomotion of leukocytes (Woodin and Harris, 1973). Oral, acute administration of DFP to guinea pigs enhanced the serum complement and hemolysin activity and the generation of a humoral immune response, but suppressed lysozyme activity. Alternatively, repeated administration of DFP suppressed complement, hemolysin, and lysozyme activities, and the generation of a humoral immune response (Lis and Mierzejewski, 1980). Intraperitoneal injection of dimethoate reduced the thymic and splenic weight of treated mice and blocked the generation of a humoral immune response (Tiefenbach and Lange, 1980). Further administration of 7.04-14.1 mg/kg dimethoate over three generations to rats affected immune function. However, the parameter affected varied from generation to generation (lnstitoris

36.3 Anticholinesterases

et aI., 1995). Administration of fenchlorphos to chickens for 3-8 weeks increased the weight of the bursa of Fabricus (Rodica and Stefania, 1973). In vitro exposure of murine splenocytes to fenthion (up to 125 I-lg/ml) blocked their ability to generate a cell-mediated immune response (Rodgers et aI., 1986b). In contrast, topical administration of fenthion to newborn mice with encephalomyocarditis virus infection did not alter mortality (Crocker et aI., 1974). Leptophos (up to 500 pm) administered orally for 12 weeks did not affect the generation of the humoral immune response in the mouse (Koller et aI., 1976). Demeton-O-methyl decreased the generation of a humoral immune response in the rat when a single high dose was administered (Nikolayev et al., 1972). Monocrotophos (up to 8 mg/kg), given intraperitoneally one time per week for 6 weeks, modulated several hematological parameters, including an increase in clotting time, white blood cell count, splenic cellularity, and the percentage of large lymphocytes, neutrophils and basophils (Gupta et aI., 1982). In vitro exposure of human basophils to soman led to an IgEindependent release of histamine (Meier et aI., 1985). Esa et al. (1988) showed that in vitro exposure of human mononuclear cells to triphenyl phosphine oxide and tetra-o-cresyl piperazinyl diphosphoramidate caused suppression of antigen-specific proliferation. In addition, treatment of human monocytes with triphenyl phosphine oxide, tetra-o-cresyl piperazinyl diphosphoamidate, triphenyl phosphate, and triphenyl thiophosphate significantly inhibited their ability to present antigen to immune T cells. Exposure of mice to tris-(2,3-dichloropropyl) phosphate decreased the proliferative response of splenocytes to mitogens and increased the incidence of tumors after challenge, but did not alter splenic or thymic weight, hematological parameters, DTH, serum Ig levels, humoral immune responses to T-cell-dependent and -independent antigens, and mortality following Listeria monocytogenes infection (Luster et aI., 1981). Triphenyl phosphate caused an allergic reaction and suppressed the immune system by subchronic administration (Carlsen et aI., 1986; Hinton et aI., 1987).

36.3.1.5 Effects of Organophosphates Esters on Humans Based on Epidemiology Some epidemiology studies have indicated that organophosphates may have an effect on the human immune system. Exposure to organophosphates has been shown to cause allergic reactions (3- to 4-month exposure), a decrease in rosette-forming T cells and an increase in B cells, a decrease in leukocyte phagocytic activity, and an increased susceptibility to colds and subjective health complaints (Bellin and Chow, 1974; Hermanowicz and Kossman, 1984; Kanezaki et aI., 1973; Katsenovich et al., 1981; Malenkii, 1978; Sawinsky and Durst, 1973; Zaninovic, 1977). One study has shown a decrease in monocyte esterase activity in workers occupationally exposed to an organophosphate compound (Lee and Waters, 1977). Occupational exposure to organophosphate pesticide decreased PMN chemotaxis and adhesion, but increased NBT reduction. In these studies, there may be a suggestion of immune modula-

775

tion, but the extent and mechanism of these effects are difficult to ascertain at this time. 36.3.1.6 Summary Most of these studies show that a variety of organophosphate pesticides reduce immune function in a variety of species. Although the site of action has not been identified at a molecular level, most investigators agree that organophosphate compounds probably act through inhibition of serine esterases. 36.3.2 CARBAMATES Carbamates, like organophosphate pesticides, act through inhibition of anticholinesterase and the symptoms of toxicity are similar to those observed with organophosphates. However, most of these compounds have low dermal toxicity, unlike most organophosphates. Carbamates are not broad spectrum pesticides, but, like organophosphates, they are relatively nonpersistent in the environment. Therefore, these compounds are widely used to eradicate the species for which they are indicated. 36.3.2.1 Carbaryl Carbaryl increased the serum level of IGGI and IgG2b without affecting the other Ig classes following oral exposure for 1 month (Andre et aI., 1983). In addition, administration of carbaryl (up to 10 I-lg/day) to quail for 5 days lowered their resistance to the protozoan parasite Histomonas meleagrides (Zeakes et aI., 1987). The humoral and cellular immune response of rabbits to antigen was unchanged following oral carbaryl (up to 150 ppm) for 4 weeks (Street and Sharma, 1975). Carbaryl suppressed a humoral immune response at very high doses (50% LDso) (Wiltrout et al., 1978). A study was conducted more recently in which the effect of carbaryl, administered by three different routes, was examined. The humoral immune response was suppressed after inhalation of carbaryl (up to 335 mg/kg), but not after oral or dermal exposure (Ladics et aI., 1994). In vitro exposure of splenocytes to carbaryl blocked their ability to generate a cellular immune response (Rodgers et aI., 1986b). Carbaryl suppressed the expression of complement activity in human serum when added to the assay (Casale et aI., 1989). Carbaryl and some of its metabolites inhibited the proliferation of interleukin-2-dependent T cells (Bavari et aI., 1989). In vitro exposure of large granular lymphocytes to carbaryl was also shown to inhibit the proliferation and the induction of NK activity in response to interleukin 2 (Casale et aI., 1990; Street and Sharma, 1975). These authors, as do others as discussed previously, suggest that carbaryl and other compounds that inhibit serine esterases act through this mechanism to modulate the generation and expression of immune responses. 36.3.2.2 Carbofuran Carbofuran did not affect the generation of a humoral immune response, but significantly suppressed cellular immunity when

776

CHAPTER 36


given over a 4-week period (Street and Sharma, 1975). Carbofuran (up to 0.6 mg/kg) also decreased the humoral immune response to neutral and pathogenic antigens and increased the cytolysis of macrophages by virus (Fournier et aI., 1988). However, carbofuran did not affect the generation of a DTH response in vivo or a cell-mediated immune response in vitro. On the other hand, carbofuran suppressed the expression of complement activity when added to human serum during the assay (Casale et aI., 1989). Carbofuran (up to 0.25 mg/kg), given intraperitoneally over 6 weeks, modulated several hematological parameters (Gupta et aI., 1982). 36.3.2.3 Aminocarb

Aminocarb decreased humoral immune response to neutral and pathogenic antigens and increased the cytolysis of macrophages by virus (Fournier et aI., 1988). Aminocarb did not decrease the resistance of mice to Salmonella typhimurium and MHV3 (Krzystyniak et aI., 1989). Aminocarb suppressed the generation of serum antibody titer to MHV3, but when given orally, enhanced the humoral immune response to a nonpathogenic antigen (Fournier et aI., 1986). Aminocarb did not affect the generation of a cell-mediated immune response or the ability of macrophages to process antigen (Fournier et aI., 1988). A more recent study compared the immunotoxic potential of aminocarb given by four different routes. Oral and dermal administration of relatively low doses of aminocarb elevated a humoral immune response. Intraperitoneal administration suppressed a humoral immune response whereas inhalation had no effect (Bernier et aI., 1995). 36.3.2.4 Aldicarb

Studies by Fiore et al. (1986) showed that women drinking well water containing detectable levels of aldicarb exhibited increased percentages and absolute numbers of CD8-positive peripheral blood lymphocytes. As a result of this increase in CD8+ cells, there was a decrease in the CD4/CD8 ratio in these women. In contrast, additional studies of exposed persons showed no effect on immune function or increase in clinical illness (Hong, 1991). Animal studies were conducted to further examine this possible alteration in immune function. Olson et al. (1987) and Shirazi et al. (1990) showed that low levels of aldicarb decrease the humoral immune response, but Thomas and co-workers (Thomas et aI., 1987; Thomas and Ratajczak, 1988) showed that low levels of aldicarb (up to 1000 ppb) did not affect the generation of cellular and humoral immune responses or the resistance of the host to infection or tumor challenge. Further studies were done on mice that received 1-100 ppb aldicarb in their drinking water for 34 days (Thomas et aI., 1990). In this study, no alterations were found in the percentage and absolute number of T cells, B cells, or T-cell subpopulations in the spleen (Thomas et aI., 1990). In addition, administration of aldicarb did not alter splenic NK activity or the ability of splenocytes to generate an allogeneic CTL response. A more recent study confirmed that chronic (90-day)

administration of 0.1-10 ppb aldicarb in drinking water did not affect any immune parameters measured (Hajoui et aI., 1992). Others have studied the effects of administration of aldicarb on macrophage and T-cell function. In these studies, aldicarb (up to 1000 ppb) in corn oil administered one time by intraperitoneal injection. Aldicarb administered intraperitoneally decreased the ability of macrophages to lyse tumors in an antibody-dependent cell-mediated cytotoxicity assay, but did not alter NK activity (Selvan et aI., 1989). In addition, aldicarb treatment suppressed the generation of a syngeneic mixed lymphocyte reaction by selectively decreasing the stimulatory activity of the macrophages without directly affecting autoreactive T cells (Dean et aI., 1990a, b). Further studies showed that the proliferation of splenocytes from treated mice to Con a and anti-CD3 antibodies was decreased after aldicarb exposure. Cell separation and reconstitution experiments showed that alteration in macrophage function, specifically decreases in IL-I production, may be responsible for this decrease in T-cell proliferation after stimulation (Dean et aI., 1990a, b). 36.3.2.5 Ethyl and Methyl Carbamate

Ethyl carbamate inhibited humoral immune response to T-celldependent and -independent antigens (Haran-Ghera and Peled, 1967; Luster et aI., 1982; Malmgren et aI., 1952; Parmiani, 1970; Parmiani et aI., 1969). However, ethyl carbamate did not affect or only slightly affected cell-mediated immunity (DiMarco et aI., 1972; Lappe and Steinmuller, 1970; Luster et aI., 1982; Parmiani, 1970). Ethyl carbamate did not affect the resistance of mice to encephalomyocarditis virus infection, but increased the incidence of induced leukemia (Chieco-Bianchi et aI., 1963). Administration of ethyl carbamate (200-400 mg/kg) for 14 days to mice reduced splenic and thymic weight and increased splenic myelopoiesis. Preinduction of P450 liver enzymes with phenobarbital resulted in an increase in the immunosuppression observed after administration of ethyl carbamate (Jeong et al., 1995). Macrophage phagocytic and bacteriocidal functions were unaffected, but the release of cytostatic factors from macrophages was elevated. Bone marrow myelopoietic function and splenic NK activity were suppressed after exposure to ethyl carbamate (up to 0.8 mg/kg) (Gupta et aI., 1982; Luster et aI., 1983). In contrast, 14-day exposure to methyl carbamate did not affect splenic or thymic weight, cell-mediated or humoral immunity, mitogenic responses, macrophage function, bone marrow function, or NK activity (Luster et aI., 1982). Perinatal exposure of mice to ethyl carbamate resulted in the induction of tumors in adults. A study of immune function of the pups after in utero exposure to ethyl carbamate (up to 1000 mg/kg) on days 7-16 of gestation or of neonates after exposure on postpartum days 5-14 (with a total of 1-2 mg/g ethyl carbamate given) was conducted. Postnatal exposure suppressed NK activity only. However, prenatal exposure increased leukocyte counts and suppressed the generation of a humoral immune response (Luebke et aI., 1986).

36.5 Herbicides 36.3.2.6 Summary The effect of carbamates on the immune system has been studied in a variety of systems. These compounds were shown either to not affect or to suppress the generation of immune responses depending upon the dose, route, and timing of exposure.

36.4 PYRETHROIDS Pyrethrum is a naturally occurring pesticide extracted from the chrysanthemum flower. Over the last several years, several pyrethroid pesticides, based upon the structure of pyrethrum, have been developed. This group of compounds is very active, has a high insect/mammal toxicity ratio, and does not persist in the environment. Therefore, these compounds are widely used. Very little information, however, is available regarding the effects of this class of compounds on the immune system. Initial studies showed that after acute and subchronic administration of cypermethrin (up to 1/2 LDso), a synthetic pyrethroid, there was an early dose-dependent suppression of the generation of a humoral immune response to Salmonella typhimurium in rabbits at doses that do not cause other toxicologic symptoms (De si et aI., 1986). In rats, the generation of a humoral immune response to SRBC and ovalbumin was suppressed by cypermethrin administration (up to 40 mg/kg) (Desi et aI., 1986). Further studies were conducted in mice and goats. Administration of cypermethrin (up to 50 mg/kg) intraperitoneally for 26 days (mice) or dermal exposure (up to 41.6 mg/kg) for 30 days (goats) resulted in a decrease in the DTH response to 2,4dinitrofluorobenzene. In addition, the generation of a humoral immune response was suppressed in this study (Tamang et aI., 1988). However, administration of up to 12 mg/kg cypermethrin for 28 days had no effect on immune function (Madsen et aI., 1996). Studies were conducted on the effects of exposure to deltamethrin for 10-30 days on the generation of immune responses in mice and rats (Kowalczyk-Bronisz et aI., 1990; Madsen et aI., 1996). In these studies, the authors observed little effect on the immune system. In contrast, oral administration of 6 mg/kg deltamethrin for 84 days or 15 mg/kg for 14 days resulted in suppression of humoral and cellular immune reponses (Lukowicz-Ratajczak and Krechniak, 1992). Exposure of mice to a newer pyrethroid pesticide, Supercypermethrin Forte, resulted in inhibition of a humoral immune response only after administrations of doses that resulted in mortality in some mice (Siroki et aI., 1994). Exposure of mouse splenocytes to allethrin, cypermethrin, fenpropathrin, and pennethrin in vitro decreased the proliferative response to mitogen (Stelzer and Gordon, 1984). These few studies suggest that cypermethrin may inhibit the immune system in a variety of species, but these effects are not observed for all compounds.

36.5 HERBICIDES Herbicides are chemicals used for the destruction of unwanted foliage. Other than the effects of dioxin, a contaminant in some

777

herbicides, on the immune system, very few studies have been done to examine the effects of herbicides on the immune system. Oral administration of 100 mg/kg/day atrazin, a triazine herbicide, for 3 days to rats decreased the number of white blood cells, but did not affect lymphoid organ weight or serum immunoglobulin levels. Oral administration of diuron, a substituted urea herbicide, for 3 weeks to rats increased the weight of lymphoid organs (Vos and Krajnc, 1983). One study conducted in rats showed that mecoprop, a phenoxy acid herbicide, altered the structure of the spleen and the thymus and altered the number of blood lymphocytes and granulocytes (Moeller and Solecki, 1989). The authors suggest that these alterations may be the result of chemically induced stress. Administration of 20-320 mg/kg/day Ordram for 12 days had no consistent effect on organ weights, natural killer activity, proliferative responses, DTH responses, and humoral immune response (Smialowicz et aI., 1985). Studies have also been conducted on the effects of propanil, a postemergence herbicide used in rice and wheat production. Propanil was given through intraperitoneal injection. Acute administration of this compound increased splenic weight and cellularity and suppressed the generation of humoral and cell-mediated immune responses at doses of 50 or 400 mg/kg, respectively (Barnett and Gandy, 1989). In addition, administration of 50-200 mg/kg propanil resulted in reduction in the number of myeloid and erythroid progenitor cells (Blyler et aI., 1994). Analysis of thymocyte subpopulations after propanil (100 mg/kg) administration showed a decrease in the number of single and double positive thymocytes (with no effect on splenic or lymph node populations) (Zhao et al., 1995). Further studies were conducted on the major metabolite of propanil, 3,4-dichloroaniline (DCA), on the immune system (Barnett et aI., 1992). Again, the compound was administered intraperitoneally. DCA, like propanil, increased the splenic weight and cellularity. In addition, DCA suppressed the generation of a humoral immune response to both T-celldependent and -independent antigens. Both propanil and DCA inhibited NK activity, but neither chemical affected the generation of a CTL response. These studies suggest that this herbicide suppressed selected immune responses, but not all immune cell types are affected. In one study, the effect of phenoxy herbicide exposure on immune function of exposed farmers was evaluated. This study showed a reduction in CD4 and CD8 cells, and proliferative responses were decreased to 12 days after exposure. The level of circulating cells, but not the proliferative responses, was normal by 70-90 days after exposure (Faustini et al., 1996). In studies of commercial herbicides, administration of 2,4dichlorophenoxyacetic acid (2,4-D) or Round Up for 26-28 days (twice weekly) did not affect immune function (Blakley, 1997; Blakley et aI., 1998). However, exposure to Tordon resulted in inhibition of humoral immunity at all exposure levels (Blakley, 1997).

778

CHAPTER 36

Irnmunotoxicity of Pesticides

36.6 SUMMARY Many pesticides that are widely used and have great potential for occupational and public exposure have received only a cursory examination with regards to their immunotoxic potential. Many of the studies that are available were done during the early period of immunotoxicology and many reports do not state whether or not any other toxic signs were observed. In addition, very few compounds have received a thorough examination under well-defined treatment conditions. That is, many of the studies were done in multiple species, through various routes of administration, and using a variety of assays of immune function. For many of the compounds that have received extensive study, the site of action has been determined at the cellular level, and for some at the biochemical level, but the molecular site of action has not been determined for any of the pesticides discussed in this chapter.

36.7 FUTURE STUDIES Future studies should include a comprehensive examination of the immunotoxic potential for at least one compound in each class of pesticides. These studies should consider the route by which humans are exposed to a compound and an attempt should be made to mimic the human situation. In addition, the metabolic capacity for the compound of the test animal and humans should be considered when the test animal is being selected. Once the immunotoxic potential and parameters under which immune suppression or enhancement occur are established, the cellular, biochemical, and molecular site of action should be established. The establishment of a mechanism of action will allow (1) determination if this site is comparable between the test animal and the human immune system, (2) analysis of the potential for immunotoxicity for other compounds within the class by examination of structure-activity relationships at the site of action rather than the intact immune system, and (3) further dissection of the immune system through the use of toxic chemicals as biochemical tools. The biomarkers that should be used to evaluate toxity in human studies should be discerned from animal studies. For example, exposure of mice to malathion caused the release of histamine from basophilic cells resulting in a transient increase in the peripheral blood of exposed animals. The applicability of this biomarker to human populations could be assessed by measurement of workers acutely exposed to the agent. For example, the level of histamine in the blood of persons occupationally exposed could be assessed at baseline (e.g., after beginning of work after a long weekend) and then midday and at the end of the workday. By comparing baseline levels of histamine to those in blood taken during the workday, the utility of this parameter in human populations can be assessed. The difficulties with biomarkers of the immune system are the vast difference in immune function between people and within a person due to

changes other than environmental exposures. Until this variability can be controlled, accounted for, or understood, assessment of changes in immune function as a result of incidental exposure will be fraught with difficulty.

REFERENCES Alberts, R. w., Rodriques de Lores, Amaiz, G., and De Robertis, E. E. (1965). Sodium potassium-activated ATPase and potassium-activated p-nitrophenyl phosphatase: A comparison of the subcellular localizations in the rat brain. Proc. Natl. Acad. Sci. U.S.A. 53, 557-564. Alien, A. L., Koller, L. D., and Pollock, G. A. (1983). Effect of toxaphene exposure on immune responses in mice. 1. Toxicol Environ. Health 11,6169. AMA Council on Drugs (1962). Registry on blood dyscrasias. A. Am. Med. Assoc.179,888-890. Andre, E, Gillon, E, Andre, c., Lafont, S., and Jourdan, G. (1983). Pesticidecontaining diet augments anti-sheep red blood cell non reaginic antibody responses in mice but may prolong murine infection with Giardia Muris. Environ. Res. 32, 145-150. Antunes-Madeira, M. c., and Madiera, V. M. C. (1979). Interaction of insecticides with lipid membranes. Biochim. Biophys. Acta 550,384-392. BaneJjee, B. D., Koner, B. C., and Ray, A. (1997). Influence of stress on DDTinduced humoral immune responsiveness in mice. Environ. Res. 74(1), 4377. Banerjee, B. D., Ray, A., and Pasa, S. T. (1996). A comparative evaluation of immunotoxicity of DDT and its metabolities in rats. Indian 1. Exp. Bioi. 34(6),517-522. BaneJjee, B. D., Saha, S., Mohapatra, T. K., and Ray, A. (1995). Influence of dietary protein on DDT-induced immune responsiveness in rats. Indian 1. Exp. Bio!. 33(10), 739-744. Bamett, J. B., and Gandy, J. (1989). The effect of acute propanil exposure on the immune response ofC57BV6V6 mice. Fundam. Appl. Toxicol. 12,757764. Bamett, J. B., Blaylock, B. L., Gandy, J., Menna, J. H., Denton, R., and Soderberg, L. S. E (1990a). Long-term alteration of adult bone marrow colony formation by prenatal chlordane exposure. Fundam. Appl. Toxicol. 14, 688695. Bamett, J. B., Blaylock, B. L., Gandy, J., Menna, J. H., Denton, R., and Soderberg, L. S. F. (1990b). Alteration offetalliver colony formation by prenatal chlordane exposure. Fundam. Appl. Toxicol. 15, 820-822. Bamett, J. B., Gandy, J., Wilboum, D., and Theus, S. A. (1992). Comparison of the immunotoxicity of propanil and its metabolite 3,4,-dichloroaniline in C57BV6 mice. Fundam. Appl. Toxico!. 18, 628-631. Bamett, J. B., Holcomb, D., Menna, J. H., and Soderberg, L. S. (1985a). The effect of prenatal chlordane exposure on specific anti-influenza cell-mediated immunity. Toxicol. Lelt. 25, 229-238. Bamett, J. B., Soderberg, L. S., and Menna, J. H. (1985b). The effect of prenatal chlordane exposure on the delayed hypersensitivity response of BALB/C mice. Toxicol. Lelt. 25, 173-183. Bartholomew, P. M., Casale, G. P., and Duggan, W. J. (1984). Effect of repeated parathion exposure on the primary Igm response and bone marrow stem cells in C57BV6 mice. Toxicologist 4, 159. Bavari, S., Casale, G. P., Gold, R. E., and Vitzhum, E. (1991) Modulation of interleukin-2-driven proliferation of human large granular Iymphocytes by carbaryl an anticholinesterase insecticide. Fundam. Appl. Toxicol. 17,6174. Bavari, S., Duszynski, c., and Casale, G. P. (1989). Effects of carbaryl and its metabolites (alpha naphthol alpha-naphthyl-betaglucuronide and alpha naphthyl sulfate) on cell cycle traverse by CTLL-2 cells. Toxicologist 9, 203. Bellin, J. S., and Chow, I. (1974). Biochemical effects of chronic low-level exposure to pesticides. Res. Commun. Chem. Pathol. Pharmacal. 9, 325337.

References

Bemier, J., Foumier, M., Blais, Y., Lombardi, P., Chevalier, G., and Krzystyniak, and K (1988). Immunotoxicity of aminocarb. L comparative stndies of sublethal exposure to aminocarb and dieldrin in mice. Pestic. Biochem. Physio!. 30, 238-250. Bemier, J., Girard, D., Krzystyniak, K, Chevalier, G., Trottier, B., Nadeau, D., Rola-Pleszczynski, M., and Foumier, M. (1995). Toxicology 99(1-2), 135146. Bemier, J., Hngo, P., Krzystyniak K, and Fonmier, M. (1987). Suppression of humoral immunity in inbred mice to dieldrin. Toxicol. Letl. 35,231-240. Blakley, B. R. (1997). Effect of roundup and tordon 202c herbicides on antibody production in mice. Vet. Hum. Toxicol. 39(4), 204-206. Blakley, B. R., Yole, M. J., Brousseau, P., Boermans, H., and Foumier, M. (1998). Effect of 2,4-dicholorophenoxyacetic acid triflura1in and triallate herbicides on immune function. Vet. Hum. Toxicol. 40(1), 5-10. Blaylock, B. L., Soderberg, L. S. E, Gandy, J., Menna, J. R, Denton, R., and Bamett, J. B. (1990). Cytotoxic T-Iymphocyte and NK responses in mice treated prenatally with chlordane. ToxicoL Letl. 51,41-49. Blyler, G., Landreth, K S., Lillis, T., Schafer, R., Theus, S. A., Gandy, J., and Bamett, J. B. (1994). Selective myelotoxicity of propaniL Fundam. AppL Toxicol. 22(4), 505-510. Carlsen, L., Andersen, K E., and Egsgaard, R (1986). Triphenyl phosphate allergy from spectacle frames. Contact Dermatitis 15, 274-277. Casale, G. P., Bavari, S., and Connolly, J. J. (1989). Inhibition of human serum by diisopropyl fluorophosphate (DFP) and selected anticholinesterase insecticides. Fundam. AppL Toxico!. 12,460-468. Casale, G. P., Bavari, S., Gold, R. E., and Vitzthum, E. E (1990). Suppression of interleukin-2 enhancement of human natural killer cell activity by carbaryl Toxicologist 10, 220. Casale, G. P., Cohen, S. D., and Di Capua, R. A (1983). The effects of organophosphate-induced cholinergic stimulation on the antibody response to sheep erythrocytes in inbred mice. Toxicol. AppL PharmacoL 68, 198205. Centeno, E. R., Johnson, W. J., and Sehon, A H. (1970). Antibodies to two common pesticides DDT and malathion. Int. Arch. Allergy AppL Immunol. 37,1-13. Chieco-Bianchi, L., Fiore-Donati, L., and De Benedict, G. (1963). Influence of urethan on susceptibility to leukemia induced by graffi virus in mice. Nature 198, 292-293. Chuang, L. E, Liu, Y., Killam, K, and Chuang, R. Y. (1992). Modulation by the insecticides heptachlor and chlordane of the cell-mediated immune proliferative responses ofrhesus monkeys. In Vivo 6, 29-32. Conradt, P., Mueller, W. E, Loose, L., Klein, E, Coulston, E, and Korte, E (1979). Incorporation of [3H] uridine into RNA under influence of dieldrin and polychlorinated biphenyls. EcotoxicoL Environ. Sa! 3, 10-17. Comacoff, J. B., Lauer, L. D., House, R. V, Tucker, A N., Thurmond, L. M., Vos, J. G., Working, P. K, and Dean, J. H. (1988). Evaluation of the immunotoxicity of beta-hexachlorocyclohexane (beta-HCH). Fundam. AppL Toxicol. 11,293-299. Costa, M. C. L., and Schvartsman, S. (1977). Antibody titers and blood levels of DDT after diphteric immunization in children. Acta PharmacoL ToxicoL 41,249. Crocker et aL (1969). Crocker, J. E S., Bozee K R., and Ozere, R. L. (1974). Insecticide and viral infection as a cause of fatty visceral changes and encephalopathy in the mouse. Lancet 2, 22-24. Crocker, 1. E S., Ozere, R. L., Safe, H. S., Digout, S. c., Rozee, K. R., and Hutzinger, O. (1976). Lethal interaction of ubiquitous insecticide carriers with virus. Science 192, 1251. Cushman, J. R., and Street, J. C. (1983). Allergic hypersensitivity to the insecticide malathion in BALB/C mice. ToxicoL AppL Pharmacol. 70, 29. Dandliker, W. B., Hides, A. N., and Levinson, S. A. (1980). Effects of pesticides on the immune response. USNITS PbRep PB80-8i] 532, 14. Dandliker, W. B. et al. (1985). De Swart, R. L., Ross, P. S., Vos, J. G., and Osterhause, A. D. (1996). Impaired immunity in harbour seals (Phoca vitulina) exposed to bioaccumulated environmental contaminants: Review of a long-term feeding study. Environ. Health Perspect. 104(4),823-828.

779

Dean, T. N., Kakkanaiah, V N., Nagarkatti, M., and Nagarkatti, P. S. (1990a). Immunosuppression by aidicaib of T cell responses to antigen-specific and polyclonal stimuli results from defective IL-1 production by macrophages. ToxicoL AppL PharmacoL 106,408-417. Dean, T. N., Selvan, R. S., Misra, H. P., Nagarkatti, M., and Nagarkatti, P. S. (1990b). Aldicarb treatment inhibits stimulatory activity of macrophages without affecting the T-cell responses in the syngeneic mixed lymphocyte reaction. Int. J. ImmunopharmacoL 12,337-348. Desi L, Dobronyi, L, and Varga, L. (1986). Immuno-neuro- and general toxicologic animal studies on a synthetic pyrethroid: Cypermethrin. EcotoxicoL Environ. Sa! 12, 220. Desi, L, Varga, L., and Farkas, L (1978). Studies on the immunosuppressive effect of organochlorine and organophosphoric pesticides in subacute experiments. J. Hyg. EpidemioL MicrobioL ImmunoL 22, 115-122. Desi, L, Varga, L., and Farkas, L (1979). The effects of DDVP an organophosphorus pesticide on the humoral and cell-mediated immunity of rabbits. Arch. ToxicoL 4, 171-174. Devens, B. H., Grayson, M. H., Imamura, T., and Rodgers, K. E. (1985). 0,0 ,S-Trimethyl phosphorothioate mediated effects on immunocompetence. Pestic. Biochem. PhysioL 24, 251-259. DiMarco, A. T., Franceshi, c., and Prodi, G. (1972). Selective thymus-derived cell enriched in the rat spleen as a result of immunodepression by urethan. Cancer Res. 32, 1569-1573. Duggan, Q. J., Casale, G. P., and Cohen, S. D. (1984). Paraoxon induced suppression of the in vitro response of murine spleen cells to sheep red blood cells. Toxicologist 4, 159. Esa, A R, WaIT, G. A., and Newcombe, D. S. (1988). Immunotoxicity of organophosphorus compounds. modulation of cell-mediated immune responses by inhibition of monocyte accessory functions. Clin. ImmunoL Immunopatho!. 49,41-52. Fan, A M. M. (1981). Effects of pesticides on immune competency: influence of methyl parathion and carbofuran on immunologic response Salmonella typhimurium. Diss. Abstr. Int. B 41, 2962. Fan, A., Street, J. c., and Nelson, R. M. (1984). Immunosuppression in mice administered methyl parathion and carbofuran by diet. ToxicoL AppL Pharmacol. 45, 235. Faustini, A, Settimi, L., Pacifici, R., Fano, V, Zuccaro, P., and Forastiere, E (1996). Immunological changes among farmers exposed to phenoxy herbicides: Preliminary observations. Occup. Environ. Med. 53(9), 583-585. Fiore, M. c., Anderson, H. A., Hong, R., Golubjatnikov, R., Seiser, J. E., Nordstrom, D., Hanrahan, L., and Belluck, D. (1986). Chronic exposure to aldicarb-contaminated ground water and human immune function. Environ. Res. 41, 633-645. Foumier, M., Bemier, 1., Flipo, D., and Krzystyniak, K (1986). Evaluation of pesticides effects on humoral response to sheep erythrocytes and mouse hepatitis virus 3 by immunosorbent analysis. Pestic. Biochem. PhysioL 26, 353-359. Foumier, M., Chevalier, G., Nadeau, D., Trotter, B., and Krzystyniak, K (1988). Virus-pesticide interactions with murine cellular immunity after sublethal exposure to dieldrin and aminocarb. 1. Toxicol. Environ. Health 1, 103-118. Friend, M., and Trainer, D. O. (1974). Experimental dieldrin-Duck hepatitis virus interaction studies. J. Wild. Manage. 38,896-901. Furie, B., and Trubowitz, S. (1976). Insecticides and blood dyscrasia. Chlordane exposure and self-limited refractory megoblastic anemia. J. Am. Med. Assoc. 235, 1720-1722. Gablicks, J., AI-Zubaidy, T., and Askain, E. (1975). DDT and immunological responses. Ill. Reduced anaphylaxis and mast cell populations in rats fed DDT. Arch. Environ. Health 30, 81-84. Gablicks, J., Askari, E. M., and Yolen, N. (1973). DDT and immunological responses. I. Serum antibodies and anaphylactic shock in guinea pigs. Arch. Environ. Health 26, 305-309. Gallichio, VS., Casale, G. P., Bartholomew, P. M., and Watts, T. D. (1987a). Altered colony forming activities of bone marrow hematopoietic stem cells in mice following short-term in vivo exposure to parathion. Int. J. Cell Cloning 15,231.

780

CHAPTER 36


Gallichio, V. S., Casale, G. P., and Watts, T. D. (1987b). Inhibition of human bone marrow-derived stem cell colony formation (CFU-E, BFU-E and CFU-GM) following in vitro exposure to organophosphates. Exp. Hemato!' 15,1099. Gewin, H. M. (1959). Benzene hydrochloride and aplastic anemia. J. Am. Med. Assoe. 171, 1624. Glick, B. (1974). Antibody-mediated immunity in the presence of mirex and DDT. Poult. Sei. 53, 1476-1485. Gupta, M., Bagchi, G., and Bandyopadhyay, S. (1982). Hematological changes produced in mice by nuvacron or furadan. Toxicology 25, 255-260. Gupta, M. et a!. (1983). Hajoui, 0., Filipo, D., Mansour, S., Foumier, M., and Krzystyniak, K. (1992). Immunotoxicity of subchronic versus chronic exposure to aldicarb in mice. Int. J. Immunopharmaeol. 14(7), 1203-1211. Haran-Ghera, N., and Peled, A. (1967). The mechanism of radiation action in leukemogenesis-isolation of a leukemogenic filtrable agent from tissues of irradiated and normal C57B mice. Br. J. Cancer 21, 730--737. Hazelton (1992). Dermal sensitization study with AC 6,601 (malathion) 57% EC in guinea pigs. In Health Risk Assessment of Aerial Applications of Malathion Bait. Hazelton Laboratories, Madison, W1. Hermanowicz, A., and Kossman, S. (1984). Neutrophil function and infectious disease in workers occupationally exposed to phosphoorganic pesticides: Role of mononuclear-derived chemotactic factor for neutrophils. Clin. Immunol. Immunopatho!. 33, 13-22. Hermanowicz, A., Nawarska, Z., Borys, D., and Naslankiewicz, A. (1982). The neutrophil function and infectious disease in workers occupationally exposed to organochlorine insecticides. Int. Arch. Oeeup. Environ. Health 50, 329-340. Hinton, D. M., Jessop, J. J., Amold, A., Albert, R. H., and Hines, F. A. (1987). Evaluation of immunotoxicity in subchronic feeding study of triphenyl phosphate. Toxieo!. Ind. Health 3, 71-89. Hong, R (1991). Effects of environmental toxins on lymphocyte function: Studies in rhesus and man. Ann. Allergy 66(6),474-480. Hugo, P., Bemier, J., Krzystyniak, K, and Foumier, M. (1988a). Transient inhibition of mixed lymphocyte reactivity by dieldrin in mice. Toxieol. Lett. 41,1-9. Hugo, P., Bemier, J., Krzystyniak, K., Potworowski, E. F., and Foumier, M. (1988b). Abrogation of graft-versus-host reaction by dieldrin in mice. Toxieo!' Lett. 41, 11-22. Infante et a!. (1976). Institoris, L., Siroki, 0., and Desi, 1. (1995). Immunotoxicity study of repeated small doses of dimethoate and methylparathion administered to rats over three generations. Hum. Exp. Toxieo!. 14(11),879-883. Institoris, L., Siroki, 0., Toth, S., and Desi, 1. (1992). Immunotoxic effects of MPT-IP containing 60% methylparathion in mice. Hum. Exp. Toxieol. 11(1), 11-16. Jedlicka, V. L., Hefmanska, Z., Smfd, A., and Kouba, A. (1958). Paramyeloblastic leukemia appearing simultaneously in two blood cousins after simultaneous contact with gamma hexane (hexachlorocyclohexane). Acta Med. Seand 161,447-451. Jeong, T. C, Cha, S. w., Park, J. 1., Ha, CS., Han, S. S., and Roh, J. K (1995). Role of metabolism in ethyl carbamate induced suppression of antibody response to sheep erythrocytes in female BALB/c mice. Int. J. Immunopharmaeol. 17(12), 1035-1044. Johnson, KW., Holsapple, M. P., and Munson, A. E. (1986). An immunotoxicological evaluation of gamma-chlordane. Fundam. App!. Toxieo!. 6, 317-326. Johnson, K w., Kaminski, N. E., and Munson, A. E. (1987). Direct suppression of cultured spleen cell responses by chlordane and the basis for differential effects on in vivo and in vitro immunocompetence. J. Toxieol. Environ. Health 22, 497-515. Kaliser, L. A. (1968). An in vitro and in vivo study of the effect of DDT on the phagocytic activity of rat white blood cells. Toxieo!. App!. Pharmaeo!' 13, 353-357. Kaminski, N. E., Roberts, J. E, and Guthrie, E. E (1982). The effect of DDT and dieldrin on rat peritoneal macrophages. Pestie. Bioehem. Physiol. 17, 191-195.

Kanezaki, H., Sera, T., Inoue, Y., and Takahashi, T. (1973). On the health disturbance of the inhabitants around a pesticide factor in Araki Area in Kurume City. J. Jpn. Assoe. Rural Med. 22, 198. Katsenovich, L. A., Ruzybakiev, R. M., and Fedorina, L. A. (1981). T and B immunity in patients with pesticide poisoning. Gig. Tr. Pro! Zabo!. 4, 1719. Kazimierczak, W., Muir, H. L., Macglashan, D. W., and Lichtenstein, L. M. (1984). An antigen activated DFP-inhibitable enzyme controls basophil desensitization. J. Immuno!. 132, 399. Klimova, J. (1970). Soderzhomie SH-gruppv syvortkekrovi i pechemi kryzpriotravlerii geptakholorom. Gig. Tr. Pro! Zabo!. 14, 56. Koller, L. D., Exon, J. H., and Roan, J. G. (1976). Immunological surveillance and toxicity in mice exposed to the organophosphate pesticide leptophos. Environ. Res. 12, 238-242. Kowalczyk-Bronisz, S. H., Gieldanowski, J., and Bubak, B. (1990). Immunological profile of animals exposed to pesticide-deltamethrin. Arch. Immuno!. Theor. Exp. Warsz. 38, 229-238. Krzystyniak, K, Bemier, J., Hugo, P., and Foumier, M. (1986). Suppression of MHV3 virus-activated macrophages by dieldrin. Bioehem. Pharmaeol. 15, 2577-2586. Krzystyniak, K, Flipo, D., Mansour, S., and Foumier, M. (1989). Suppression of avidin processing and presentation by mouse macrophages after sublethal exposure to dieldrin. Immunopharmaeology 18, 157-166. Krzystyniak, K, Hugo, P., Flipo, D., and Foumier, M. (1985). Increased susceptibility to mouse hepatitis virus 3 of peritoneal macrophages exposed to dieldrin. Toxieo!. Appl. Pharmaeol. 80, 397-408. Krzystyniak, K., Trottier, B., Jolicoeur, P., and Foumier, P. (1987). Macrophage functional activities versus cellular parameters upon sublethal pesticide exposure in mice. Mo!. Toxieol. 1,247-259. Kynoch, S. R, and Smith, P. A. (1992). Delayed contact hypersensitivity in the guinea pig with malathion (Fytanow) technical. In "Health Risk Assessment of Aerial Application of Malathion Bait." Huntington Research Center Huntington, Cambridgeshire, England. Ladics, G. S., Smith, C, Heaps, K, and Loveless, S. E. (1994). Evaluation of the humoral immune response of CD rats following a 2-week exposure to the pesticide carbaryl by the oral, dermal or inhalation routes. J. Toxieo!. Environ. Health 42(2), 143-156. Lahvis, G. P., Wells, R. S., Kuehl, D. w., Stewart, J. L., Rhinehart, H. L., and Via, C S. (1995). decreased Iymphocyte responses in free-ranging bottlenose dolphins (Tursiops Truneatus) are associated with increased concentrations of PCBs and DDT in peripheral blood. Environ. Health Perspeet. 193(Suppl. 4), 67-72. Lappe, M. A., and Steinmuller, D. S. (1970). Depression of weak allograft immunity in the mouse of neonatal or adult exposure to urethan. Cancer Res. 30,674-678. Lee, M. J., and Waters, H. C (1977). Inhibition of monocyte esterase activity by organophosphorus insecticides. Blood 50, 947-951. Lee, T. P., Moscati, R, and Park, B. H. (1979). Effects of pesticides on human leukocyte functions. Res. Conunun!. Chem. Patho!. Pharmaeo!' 23, 597609. Lis, T., and Mierzejewski, T. (1980). Inhibition of immune responsivity by diisopropyl phosphorofluoridate. Arch. Toxieo!. 4, 151-155. Loge, J. P. (1965). Aplastic anemia following exposure to benzene hexachloride (lindane). J. Am. Med. Assoe. 193, 110--114. Loose, L. D. (1982). Macrophage induction of T-suppressor cells in pesticideexposed and protozoan-infected mice. Environ. Health Perspeet. 43, 89. Loose, L. D., Silkworth, J. B., Charbonneau, T., and Blumenstock, F. (1981). Environmental chemical-induced macrophage dysfunction. Environ. Health Perspeet. 39,79-91. Luebke, R w., Riddle, M. M., Rogers, R R, Rowe, D. G., Gamer, R J., and Smialowicz, R. J. (1986). Immune function in adult C57B1I6 mice following exposure to urethan pre- or postnatally. J. Immunopharmaeo!. 8,243-257. Luki, M. L., Popeskovic, L., and Jankovic, B. D. (1973). Potentiation of immune responsiveness in rats treated with DDT. Fed. Proe. 32,1037, A4615. Lukowicz-Ratajczak, J., and Krechniak, J. (1992). Effects of deltamethrin on the immune system in mice. Environ. Res. 59(2), 467-475.

References

Luster, M. 1., Dean, J. H., Boorman, G. A., Archer, D. L., Lauer, L., Lawson, L. D., Moore, J. A., and Wilson, R. E. (1981). The effects of ortho phenyl-phenol tris (2,3-dichloropropyl) phosphate and cyclophosphaniide on the immune system and host susceptibility of mice following subchronic exposure. Toxicol. Appl. Pharmacol. 58,252-261. Luster, M. 1., Dean, J. H., Boorman, G. A., Dieter, M. P., and Hayes, H. T. (1982). Immune functions in methyl and ethyl carbamate treated mice. Clin. Exp. Immunol. 50, 223-230. Luster, M. I. et al. (1983). Madsen, C., Claesson, M. H., and Ropke, C. (1996). Immunotoxicity of the pyrethroid insecticides deltametrin and alpha-cypermetrin. Toxicology 107(3),219-227. Magnusson, B., and Kligman, A. M. (1987). "Allergic Contact Dermatitis in the Guinea Pig. Identification of Contact Allergens." Thomas Springfield, IL. Malenkii, V. P. (1978). Clinical immunological characteristics of chronic pneumonia in farm equipment operators. Vrach. Delo 6, 53. Malmgren, R. A., Bennison, B. E., and Mckinley, T. W. (1952). Reduced antibody titres in mice treated with carcinogenic and cancer chemotherapeutic agents. Proc. Soc. Exp. BioI. Med. 79,484-488. Mastromattco, E. (1964). Hematological disorder following exposure to insecticides. Can. Med. Assoc. 1. 90, 1166. Meier, H. L., Gross, C. L., Papenneister, B., Kagey-Sobotka, A., and Kilduff, J. E. (1985). Histamine release by esterase inhibitors. The regulation of histamine release from human leukocytes of allergic and non-allergic individuals by the serine esterase inhibitors diisopropyl fluorophosphate and pinacolyl methyl phosphorofluoridate. Int. Arch. Allergy Appl. Immunol. 77, 218-221. Menna, J. H., Barnett, J. B., and Soderberg, L. S. (1985). Influenza type a virus infection of mice exposed in utero to chlordane; survival and antibody studies. Toxicol. Lett. 24,45-52. Milby, T. H., and Epstein, W. L. (1964). Allergic contact sensitivity to malathion. Arch. Environ. Health 9, 434-437. Milby, T. H., and Samuels, A. J. (1971). Human exposure to lindane: Clinical hematological and biochemical effects. 1. Occup. Med. 13, 256-258. Moeller, T., and Solecki, R. (1989). Pathomorphologic and hematologic studies of the immunotoxicity of the phenoxyalkane acid mecropropin rats. Z. Gesamte Hyg. 35, 258-260. Nikolayev, A. 1., Ponomareva, L. A., and Geller, L. S. (1972). Immunosuppressive actions of some pesticides. Farmakol. Toksikol. (Moscow) 35, 352. Olson, L., Erickson, B., and Hinsdale, R. (1987). Aldicarb immunomodulation in mice. An inverse dose-response to parts per billion levels in drinking water. Arch. Environ. Contam. Toxicol. 16,433-439. Park, B. H., and Lee, T. H. (1978). Effects of pesticides on human leukocyte functions. In "Proc. 4th FDA Science Symp." (1. Masher, ed.), Annapolis, MD,p.273. Parmiani, G. (1970). Immune depressive effect of urethan on the homograft response in mice. Int. 1. Cancer 5, 260. Parmiani, G., Colnaghi, M. 1., and Della Porta, G. (1969). Immunodepressive and leukemogenic effects of urethan in Ch3 and SWR mice. Proc. Soc. Exp. BioI. Med. 130, 828. Pruett, S. B., and Chambers, J. B. (1988). Effects of paraoxon p-nitrophenol phenyl saligenin cyclic phosphate and phenol on the rat interleukin 2 system. Toxicol. Lett. 40, 11. Raise, B. T. (1983). Role of adaptive immune defense mechanisms. In "Immunobiology of Herpes Simplex Virus Infection" (B. T. Rouse and C. Lopez, eds.), pp. 69-73. CRC Press, Boca Raton, FL. Relford, R. L., Ainsworth, A. J., and Harkness, J. E. (1989). Effects of a commercial malathion dip preparation on the cellular and humoral response of BALB/c mice. Lab. Anim. Sci. 39,56-59. Rodgers, K. E. (1997). Effects of oral administration of malathion on the course of disease in MRL-Lpr. 1. Autoimmunity 10, 367-373. Rodgers, K. E., and Ellefson, D. D. (1988a). Cytofluorometric changes in macrophage cell surface markers following exposure to O,O,S-trimethyl phosphorothioate. FASEB 1.47(3),1617.

781

Rodgers, K. E., and Ellefson, D. D. (1988b). Effects of acute administration of 0,0,S-trimethy1 phosphorothioate on the respiratory burst and phagocytic activity of splenic and peritonealleukocytes. Agents Actions 24, 152-160. Rodgers, K. E., and Ellefson, D. D. (1990a). Modulation of respiratory burst activity and mitogenic response of human peripheral blood mononuclear cells and murine sp1enocytes by malathion. Fundam. Appl. Toxicol. 14,309-317. Rodgers, K. E., and Ellefson, D. D. (1990b). Modulation of macrophage secretion of proteases and protease inhibitory activity by acute administration of O,O,S-trimethyl phosphorothioate. Agents Actions 29, 277-285. Rodgers, K. E., and Ellefson, D. D. (1992). Mechanism of modulation of murine peritoneal cell function and mast cell degranulation by low doses of malathion. Agents Actions 35, 57-63. Rodgers, K. E., and Xiong, S. (1996). Contribution of mast cell mediators and histamine to alterations in macrophage function after malathion administration. Fundam. Appl. Toxicol. 33,100-108. Rodgers, K. E., and Xiong, S. (1997a). Effect of administration of malathion for 90 days on macrophage function and mast cell degranulation. Toxicol. Lett. 93, 73-82. Rodgers, K. E., and Xiong, S. (1997b). Effect of administration of malathion for 14 days on macrophage function and mast cell degranulation. Fundam. Appl. Toxicol. 37, 95-99. Rodgers, K. E., and Xiong, S. (1997c). Contribution of inflammatory mast cell-mediators to alterations in macrophage function after malathion administration. Int. 1. Immunopharmacol. 19, 149-156. Rodgers, K. E., and Xiong, S. (1998). Effects of acute administration of malathion by oral and dermal routes on serum histamine levels. Int. 1. Immunopharmacol. xxx, xxx-xxx. Rodgers, K. E., St Amand, K., and Xiong, S. (1996). Effects of Malathion on the humoral immunity and macrophage function in mast cell deficient mice. Fundamental and Applied Toxicology 31, 252. Rodgers, K. E., Ellefson, D. D., and Ware, C. F. (1987a). Cytofluorometric analysis of thymic and splenic lymphoid populations following acute administration of 0,0 ,S-trimethyl phosphorothioate. Toxicologist 7, 232. Rodgers, K. E., Grayson, M. H., Imamura, T., and Devens, B. H. (1985a). In vitro effects of malathion and O,O,S-trimethyl phosphorothioate on cytotoxic T lymphocyte responses. Pestic. Biochem. Physiol. 24, 260-266. Rodgers, K. E., Grayson, M. H., and Ware, C. F. (1988a). Inhibition of cytotoxic T-Iymphocyte and natural killer cell-mediated lysis by O,S,S-trimethyl phosphorodithioate is at an early postrecognition step. 1. Immunol. 140, 564-570. Rodgers, K. E., Haviland, D. L., and Ware, C. F. (1989a). Protection from 0,0 ,S-trimethyl phosphorothioate induced immune suppression. Immunopharmacology 17, 131-140. Rodgers, K. E., Imamura, T., and Devens, B. H. (1985b). Effects of subchronic treatment with O,O,S-trimethyl phosphorothioate on cellular and humoral immune response systems. Toxicol. Appl. Pharmacol. 81,310-318. Rodgers, K. E., Imamura, T., and Devens, B. H. (1985c). Investigations into the mechanism of immunosuppression caused by acute treatment with O,O,Strimethyl phosphorothioate. I. Characterization of the immune cell population affected. Immunopharmacology 10,171-180. Rodgers, K. E., Imamura, T., and Devens, B. H. (1985d). Investigations into the mechanism of immunosuppression caused by acute treatment with O,O,Strimethyl phosphorothioate. n. Effect on the ability of murine macrophages to present antigen. Immunopharmacology 10, 181-189. Rodgers, K. E., Imamura, T., and Devens, B. H. (1986a). Organophosphorus pesticide immunotoxicity: Effects of 0,0 ,S-trimethyl phosphorothioate on cellular and humoral immune response systems. Immunopharmacology 12, 193-202. Rodgers, K. E., Imamura, T., and Devens, B. H. (1987b). Investigations into the mechanism of immunosuppression caused by acute treatment with O,O,Strimethyl phosphorothioate: Generation of suppressive macrophages from treated animals. Toxicol. Appl. Pharmacol. 88,270-281. Rodgers, K. E., Leung, N., Imamura, T., and Devens, B. H. (1986b). Rapid in vitro screening assay for immunotoxic effects of organophosphorus and carbamate insecticides on the generation of cytotoxic T-Iymphocyte responses. Pestic. Biochem. Physiol. 26,292-301.

782

CHAPTER 36


Rodgers, K. E., Leung, N., Ware, C. F., Devens, B. H., and Imamura, T. (1986c). Lack of immunosuppressive effects of acute and subacute administration of malathion on murine cellular and humoral immune responses. Pestic. Biochem. Physio!. 25,358-365. Rodgers, K. E., Leung, N., Ware, C. F., and Imamura, T. (1987c). Effects of O,S,S-trimethyl phosphorodithioate on immune function" Toxicology 43, 201-216. Rodgers, K. E., Leung, N., and Ware, C. F. (1988b). Effects of acute administration of 0 ,S,S-trimethyl phosphorodithioate on the generation of cellular and humoral immune response following in vitro stimulation. Toxicology 51,241-253. Rodgers, K. E., Stem, M. L., and Ware, C. F. (1989b). Effects of subacute administration of O,S,S-trimethyl phosphorodithioate on cellular and humoral immune response parameters. Toxicology 54, 183-195. Rodica G., and Stefania, M. (1973). Effects of some insecticides on the bursa of Fabricus in chickens. Arch. Exp. Vet. Med. 27, 723-728. Ross, P. S., De Swat, R. L., Reijnders, P. l., Van Loveren, H., Vos, l. G., and Osterhause, A. D. (1995). Contaminant related suppression of delayed type hypersensitivity and antibody responses in harbor seals fed herring from the baltic sea. Environ. Health Perspect. 102(2), 162-167. Roux, F., Treich, 1., Brun, C., Desoize, B., and Fournier, E. (1979). Effect of lindane on human Iymphocyte response to phytohemagglutinin. Biochem. Pharmaco!' 28,2419-2426. Samuels, A. l., and Milby, T. H. (1971). Human exposure to lindane: Clinical hematological and biochemical effects. J. Occup. Med. 13, 147. Sianchez-Madel L., Castanedo l. P., and Garcia-Rojas F. (1963). Insecticides and aplastic anemia. N. Engl J. Med. 269, 1365-1367. Sawinsky, A., and Durst, l. (1973). The effect of pesticides on the activity of phagocytes. Z. Gesamte Hyg. 19, 863-865. Selvan, R. S., Dean, T. N., Misra, H. P., Nagarkatti, P. S., and Nagarkatti, M. (1989). Aldicarb suppresses macrophage but not natural killer (NK) cellmediated cytotoxicity of tumor cells," Bull. Environ. Contam. Toxicol. 43, 676-682. Shirazi, M. A., Erickson, B. l., Hinsdill, R. D., and Wyman, l.A. (1990). An analysis of risk from exposure to aldicrab using immune response of nonuniform populations of mice. Arch. Environ. Contam. Toxicol. 19(3), 447-456. Siroki, 0., Institoris, L., Tatar, E., and Desi, 1. (1994). Immunotoxicological investigation of SCMF a new pyrethroid pesticide in mice. Hum. Exp. Toxicol. 13(5), 337-343. Smialowicz, R. l., Luebke, R. W., Rogers, R. R., Riddle, M. M., and Rowe, D. G. (1985). Evaluation of immune function in mice exposed to ordram. Toxicology 37(3-4), 307-314. Spyker-Crammer, l. M., Bamett, l. B., Avery, D. L., and Cranmer, M. F. (1982). Immunoteratology of chlordane: Cell-mediated and humoral immune responses in adult mice exposed in utero. Toxicol. Appl. Pharmaco!' 62, 402-408. Stelzer, l., and Gordon, M. A. (1984). Effects of pyrethroids on lymphocyte mitogenic responsiveness. Res. Commum. Chem. Patho!. Pharmacol. 46, 137-150. Stieglitz, R., Stobbe, H., and Scheiettman, W (1967). Knochenmarkschaden beruflicher einwirkung des insektizids gamma hexachlorocyclohexan. Acta Hemato!' 36,337. Street, l. C., and Sharma, R. P. (1975). Alteration in induced cellular and humoral immune responses by pesticides and chemicals of environmental concern: Quantitative studies of immunosuppression by DDT, Arochlor 1254 carbaryl carbofuran, and methyl parathion. Toxicol. App!. Pharmaco!' 32,587-602. Subba Rao, D. S. v., and Glick, B. (1977). Pesticide effects on the immune response and metabolic activity of chicken Iymphocytes. Proc. Soc. Exp. Bio!. Med. 154,27-29. Suzald, E., Inoue, B., Okimasu, E., Ogata, M., and Utsumi, K. (1988). Stimulative effect of chlordane on the various functions of the guinea pig leukocytes. Toxico!. App!. Pharmaco!' 93, 137-145.

Tamang, R. K., lha, G. l., Gupta, M. K., Chauhan, H. v., and Tiwary, B. K. (1988). In vivo immunosuppression by synthetic pyrethroid (cypermethrin) pesticide in mice and goats. Vet. Immuno!. Immunopatho!. 19,299-305. Theus, S. A., Lau, K. A., Tabor, D. R., Soderberg, L. S. F., and Bamett, l. B. (1992). In vivo prenatal chlordane exposure induces development of endogenous inflammatory macrophages. J. Leukocyte BioI. 51, 366--372. Thomas, 1. K., and Imamura, T. (1986). Modulation of cellular and humoral response by O,S,S-trimethyl phosphorothioate an impurity in commercial malathion. Toxicology 39, 1-12. Thomas, 1. K., Koizumi, A., and Imamura, T. (1986). Suppressive effect of O,O-dimethyl,5-ethyl phosphorothioate on immune response. J. Toxico!. Environ. Health 19, 465-476. Thomas, P., and Ratajczak, H. (1988). Assessment of carbamate pesticide immunotoxicity. Toxicol. Ind. Health 4,381-390. Thomas, P., Ratajczak, H., Demetral, D., Hagen, K., and Baron, R. (1990). Aldicarb immunotoxicity: Functional analysis of cell-mediated immunity and quantitation of Iymphocyte subpopulations. Fundam Appl. Toxico!. 15, 221-230. Thomas, P., Ratajczak, H., Eisenberg, W, Furedi-Machacek, M., Ketels, K. v., and Barbera, P. W. (1987). Evaluation of host resistance and immunity in mice exposed to the carbamate pesticide Aldicarb. Fundam. Appl. Toxico!. 9,82-89. Tiefenbach, B., and Lange, P. (1980). Studies on the action of dimethoate on the immune system. Arch. Toxico!. 4, 167-170. Vanat, S. v., and Vanat, 1. M. (1971). Contributions to the toxic-allergic reaction induced by DDT. Klin. Med. 49, 126--127. Varshneya, C., Bahga, H. S., and Sharma, L. D. (1988). Effect of insecticides on humoral immune response in cockerels. Br. Vet. J. 144,610--612. Vijay, H. M., Mendoza, C. E., and Lavergne, G. (1978). Production ofhomocytotropic antibodies (Ige) to malathion in the rat. Toxicol. App!. Pharmacol. 44, 137-142. Vos, l. G., and Krajnc, E. 1. (1983). Immunotoxicity of pesticides. In "Developments in the Science and Practice of Toxicology," 229-240. Elsevier, Amsterdam. Wassermann, M., Wassermann, D., Gershon, Z., and Zellennayer, L. (1969). Effect of organochlorine pesticides on body defense system. Ann. N. Y. Aead. Sci. 160, 393-401. Waterhouse, l., and Tourney, T. (1984). The effects of organophosphorus and carbamate cholinesterase inhibitors on in vitro immune responses. Toxicologist 4, 159. West, 1. (1967). Lindane and hematological reactions. Arch. Environ. Health 15,97-101. Wiltrout, R. W, Ercegovich, C. D., and Ceglowski, W S. (1978). Humoral immunity in mice following oral administration of selected pesticides. Bull. Environ. Contam. Toxieo!. 20,423-431. Woodin, A. M., and Harris, A. (1973). The inhibition of locomotion of the polymorphonuclear leukocyte by organophosphorus compounds. Exp. Cell Res. 77,41-46. Xiong, S., and Rodgers, K. E. (1997). Effects of malathion metabolites on degranulation of and mediator release by human and rat basophilic cells. J. Toxicol. Environ. Health 51(2), 101-117. Zabrodski, P. F. (1993). The mechanisms of the immunotropic effects of organophosphorus compounds. Bull. Eksp. BioI. Med. 116(8), 181-183. Zaninovic, M. (1977). Agranulocytosis caused by exposure to insecticides. Arh. Hig. Rada Toksikol. 28, 43. Zeakes, S. l., Hanson, M. F., and Robel, R. l. (1987). Increased susceptibility of bob whites (Colinus virginianus) to Histomonas meleagriditis after exposure to sevin insecticides. Avian Dis. 25, 981-987. Zhao, W, Schafer, R., Cuff, C. F., Gandy, l., and Bamett, l. B. (1995). Changes in primary and secondary lymphoid organ T-cell subpopulations resulting from acute in vivo exposure to propanil. J. Toxico!. Environ. Health 46(2), 171-181.

CHAPTER

37 Sensitive Population Groups Richard J. J ackson, Carol H. Rubin, and Michael McGeehin National Center for Environmental Health, Centers for Disease Control and Prevention

37.1 INTRODUCTION

37.1.1 EXPOSURE AMONG SENSITIVE POPULATION GROUPS

Exposure to pesticides does not affect all humans uniformly. Age, sex, genetic make-up, health status, and previous or concurrent exposures influence individual sensitivity. These parameters are interrelated and may combine to influence both qualitative and quantitative differences in sensitivity. Thus sensitivity and susceptibility are inextricably related. Defined population groups may be more susceptible to the toxic effects of pesticide exposure because of a greater inherent sensitivity and also because certain characteristics of the subpopulation may result in greater exposure. For example, young children are usually physiologically more sensitive than adults to a given pesticide exposure level (Mortensen et aI., 1996; Pope et aI., 1991). At the same time, the activities of young children (e.g., crawling on floor and engaging in hand-to-mouth behaviors) increase the likelihood of exposure in a household setting. This chapter will consider the question of sensitivity with regard to realistic exposure scenarios. Pesticides are often included in the list of chemicals that elicit adverse health outcomes in people identified as multiply chemically sensitive. Such people react to a wide variety of chemicals at exposure levels that are usually tolerated by the general population. This is a complex and controversial condition that is well-profiled in the literature (Cone and SuIt, 1992; Ziem and McTamney, 1997) and beyond the scope of this chapter. Sensitivity to pesticides is a multifaceted emerging issue. As new pesticide formulations become available, and as the number of mixtures of pesticides that are on the market increases, it is likely that parameters of sensitivity and susceptibility will modify. The following discussion attempts to summarize the observed effects of pesticides on humans and also the suspected effects, based upon animal models. Some of the categories of sensitivity addressed in this chapter (e.g., genetically imposed sensitivity) are presented elsewhere in this Handbook in greater depth and in a framework that goes beyond issues of sensitive populations. Handbook of Pesticide Toxicology Volume 1. Principles

Hill et al. (1995) measured 12 urinary metabolite pesticide residues, reflecting exposure to more than 30 different pesticides, in a sample of 1000 adults from the Third National Health and Nutrition Examination Survey, 1988-1994. Six of the pesticide residues were detectable in more than half of the population sampled (Hill et al., 1995). Para-nitrophenol (p-NP), the residue representing exposure to methyl parathion, was detected in 41 % of the samples. However, exposure rates are substantially higher among certain sensitive population groups than they are in the total popUlation. This was demonstrated in 1995 during an assessment of exposure to methyl parathion which had been illegally applied indoors. The urinary metabolite levels that define actual human exposure varied by age and sex (Fig. 37.1). People spending more time in the home (e.g., infants, the elderly, and the unemployed) had increased exposure potential and demonstrably elevated p-NP metabolite levels (Esteban et aI., 1996). At the same time, the reasons that kept many of these people at home (e.g., being of pre-school age, being pregnant, or having a chronic disease) also defined physiologically sensitive population groups.

37.2 QUALITATIVE AND QUANTITATIVE ASPECTS OF SENSITIVITY Pharmakokinetic differences in absorption, distribution, metabolism, and excretion are the basis for most subpopulation differences in pesticide sensitivity. This is particularly apparent among infants and children. 37.2.1 ABSORPTION Dermal absorption and gastrointestinal tract (GIT) absorption depend upon ratios of surface area to body weight, as well as on the characteristics of the absorptive surface. Children are more susceptible to dermal pesticide exposure because they have a greater surface area relative to their weight. The ratio of surface

783


784

CHAPTER 37

Sensitive Population Groups

400 350 300 250 200 150 100 50

o

Age Group in Years Figure 37.1 Distribution of creatinine-adjusted para-nitrophenol (the urinary metabolite marker for methyl parathion) by sex and age among 306 persons exposed to the indoor application of methyl parathion in Lorain County, Ohio (1994).

area to weight of a newborn may be up to 2.5 times that of an adult. Thus similar exposure levels lead to significantly higher doses in infants than in adults (Rasmussen, 1979). Dermal characteristics may also increase sensitivity. A fullterm infant is born with a completely developed stratum corneum, the main barrier to percutaneous absorption (Lester, 1983). Human and animal studies of the antimicrobial hexachlorophene have shown no significant difference in percutaneous absorption between adults and full-term infants (Plueckhahn, 1973). Solomon and Fahrner (1977) found no difference in y-benzene hexachloride levels in the brain after topical application to newborn and adult guinea pigs. Similarly, Wester et al. (1977) reported no difference in percutaneous absorption of testosterone between newborn and adult Rhesus monkeys. In contrast, preterm infants appear to have increased percutaneous absorption for various chemicals (Shuman et aI., 1975). Serum levels of hexachlorophene have been higher in preterm than full-term infants bathed with the chemical (Greaves et aI., 1975; Tyra1a et aI., 1977). Dermal exposure of preterm infants to hexachlorophene has resulted in severe toxicity manifested by apnea, convulsions, and coma (Shuman et aI., 1975). Percutaneous absorptive changes associated with aging are not well-defined (Roskos et aI., 1986). In a study of percutaneous absorption of 14 pesticides in young and adult rats, 11 of the pesticides exhibited significant age- and dose-dependent differences in skin penetration. Among the pesticides that exhibited significant age-dependent differences in absorption, the highest young/adult penetration ratio was 1.53 (Hall et aI., 1988). However, this study did not include aged or senile rats. 37.2.1.1 GIT Absorption The extent to which pesticides are absorbed after oral exposure is determined by gastric acidity, emptying time, and intestinal motility (Morselli et aI. , 1980). Full-term neonates develop stomach acidity within 24 hours of birth, whereas their gastrointestinal motility is irregular and unpredictable (Morselli, 1976). Although there is little information on the rates of gastrointestinal absorption of pesticides in infants and children,

several studies have been done on the gastrointestinal absorption of pharmaceuticals in neonates. Morselli (1976) found that newborns absorb some pharmaceuticals (e.g., digoxin) the same as adults, some (e.g., ampicillin) to a greater extent, and some (e.g., phenobarbitol, chloramphenicol) to a lesser extent. The complexity of the gastrointestinal system makes it difficult to determine the effect of maturation on the absorption of pesticides (Warner, 1986). Hoffmann (1982) reviewed a number of studies on the gastrointestinal absorption of exogenous chemicals in experimental animals at various stages of development. He concluded that drugs which cross the intestinal epithelium by passive diffusion are absorbed from the gastrointestinal tract of immature animals at a higher rate than from the tract of adult animals. The higher absorption rate did not, however, result in higher blood levels because of a greater distribution volume in the neonate. 37.2.1.2 Transplacental Absorption Although adverse outcomes associated with fetal exposure to pesticides is not well defined, the passage of pesticides through the placenta is documented (Kreuzer et aI., 1997). In vitro placental perfusion by parathion resulted in significant transfer and 50% acetylcholinesterase depression (Benjaminov et aI., 1992). 37.2.2 DISTRffiUTION Once absorbed, pesticides are distributed within the plasma to various tissues and organs. Distribution varies according to the size of the individual, blood flow to the tissue, the pH of the body fluids, the distribution of body water in intracellular and extracellular compartments, and the extent of protein binding of the pesticide (Warner, 1986). These factors may compete with or complement each other. Increased sensitivity occurs when imbalances in distribution lead to a higher pesticide concentration in the target organ. An investigation of the passage of paraquat through the blood-brain barrier showed a high rate of entry into the brain of neonatal rats than into the brain of adult and elderly rats (Widdowson et aI. , 1996). A similar study reported higher paraquat concentrations among both very young

37.3 Age-Related Differences in Sensitivity and very old rats (Corasaniti et aI., 1991). Although there is little research directly relating to pesticide distribution, reviews of pharmaceutical distribution describe potential sensitivities among infants and children (Kearns and Reed, 1989).

785

resulting in high concentrations of the drug and its metabolites persisting in the blood. The decreased renal function of newborns, combined with their slower biometabolism, increases the likelihood that they will have pharmakokinetically based pesticide sensitivity.

37.2.3 METABOLISM

A major determinant of the toxicological effect of a pesticide is the manner in which it is handled in the body. Although the liver is the primary organ for xenobiotic metabolism, the kidney, intestines, lungs, and skin are also capable ofbiotransforming certain compounds, and alterations in these organs may lead to increased pesticide sensitivity (Reed and Besunder, 1989). Many such alterations are age-dependent. Neonates do not have full metabolic capacity; metabolic maturity is achieved at about 6 months of age in full-term infants and more slowly in premature infants (Warner, 1986). At birth, the metabolic deficiencies of the neonate include decreased hydroxylation and plasma enterase activity, a decreased amount of cytochrome P450, and a deficient glucorodination process (Done, 1964; Morselli, 1976; Morselli et aI., 1980; Warner, 1986). Metabolic activity may be further reduced by pathological conditions such as respiratory distress, cardiac insufficiency, hyperbilirubinemia, and low dietary intake (Morselli et aI., 1980). Once children's metabolic mechanisms are mature, their metabolic activity rises rapidly until they are about 3 years old. From 3 years to puberty, their metabolic activity slowly declines to adult levels and then further slows as they age (Warner, 1986). Neonates' metabolism of certain exogenous chemicals has also been found to be qualitatively different from that of adults. For example, in preterm and full-term infants, theophylline undergoes N-methylation to caffeine, whereas the opposite process occurs in adults (Reed and Besunder, 1989). Miller et al. (1976) described similar age-related differences in the metabolism of acetaminophen. Young rats store lindane more readily and for longer periods than do adult rats (Solomon and Fahrner, 1977). 37.2.4 EXCRETION

The kidney is the primary route of elimination for most pesticides. Although the kidneys of newborns have a full complement of glomeruli, their tubular size and mass are less than those of adults (Reed and Besunder, 1989). Adult kidney function is achieved at about one year of age (Morselli, 1976). Only a handful of studies have been done on the actual renal clearance of drugs and their metabolites from children and infants. The clearance of penicillin in premature infants was only 17% that in older children when corrected for surface area (Barnett et aI., 1949). Studies of chloramphenicol concentrations in the blood of neonates showed an inverse correlation between the age ofthe infant and the half-life of the drug (Reed and Besunder, 1989). The conjugation of chloramphenicol with glucoronic acid and renal excretion was lower among neonates,

37.3 AGE-RELATED DIFFERENCES IN SENSITIVITY As summarized previously and discussed at length elsewhere (Thomas, 1995), age-related sensitivity to pesticide exposure is a function of differences in size, in the maturity of biochemical and physiological functions in major body systems, and in body composition (proportions of water, fat, protein, and mineral content) (Forbes, 1987). In addition to having increased physiological sensitivity to pesticides, children are also at greater risk from pesticides because they have more opportunity for exposure (Go1dman, 1995; NRC, 1993; Thomas, 1995). 37.3.1 AGE-RELATED DIFFERENCES IN EXPOSURE

Sources and routes of exposure for children include ingestion (food and water), inhalation, and dermal contact inside and outside the home, and take-home-toxin exposure (Rogan, 1980). In Missouri, results of a telephone-interview survey of 238 families showed that nearly all families (97.8%) used pesticides at least one time per year in the home, garden, orchard, or yard; two-thirds used pesticides more than five times per year. More than 80% of families used pesticides during a pregnancy, and 70% used pesticides during the first 6 months of a child's life (Davis et aI., 1992). Other studies have quantitatively confirmed pesticide residue levels in home environments (Bradman et aI., 1997; Gurunathan et aI., 1998). The Gurunathan study showed that 2 weeks after certified applicators sprayed a single application of chlorpyrifos in apartment rooms, the pesticide continued to accumulate on children's toys and hard surfaces. Dietary exposure to pesticides varies both qualitatively and quantitatively with age. Infants and children consume more calories of food per unit of body weight than do adults, but they also consume far fewer types of foods. Consequently, infants and young children may consume much more of certain foods suspected of having elevated pesticide residues (e.g., apples), especially processed foods (e.g., apple juice, apple sauce). The younger the child, the less diverse the foods that he or she consumes (Thomas, 1995). 37.3.2 AGE-RELATED DIFFERENCES IN TOXICITY 37.3.2.1 Fetuses, Infants, and Young Children

Children are less able than adults to metabolize and excrete toxic substances (Drew et aI., 1983; Finhorn, 1982; Hudson

786

CHAPTER 37


et aI., 1972; Kalland, 1982; Widdowson and Dickerson, 1960). Their rapidly developing organ systems, especially the central nervous system, are highly susceptible to chemical interference. Exposures during brief but critical periods early in development can permanently alter the structure or function of an organ system (Swenberg and Fedtke, 1992; Vesell, 1982). Fetal sensitivity to pesticides may lead to unique transgenerational manifestations. In a study of cancer incidence among the progeny of male rats exposed to ethy lnitrosourea before mating, Tomatis et al. (1981) found that the fertility rate was lower and the preweaning mortality rate was higher in the experimental group than in the control group. Survival rates after weaning, as well as the total incidence of tumors, were similar in the progeny of treated males and of controls. However, neurogenic tumors were more frequent among progeny of rats in the experimental group (p = 0.08). These findings suggest that the toxic effects of mutagens can be transmitted through the male germ line. In an earlier study of the occurrence of tumors in first-, second-, and third-generation descendants of rats exposed to N-nitrosomethylurea during pregnancy, Tomatis et al. (1975) confirmed previous observations that exposure to a carcinogen during prenatal life may cause a genetic-chemical interaction leading to an increased cancer risk, which may persist for more than one generation. Other rat studies found that vinclozolin, a fungicide for fruits and vegetables, possessed antiandrogenic activity, causing feminization of male fetuses, sterility of adult males, and other developmental variations (Gray et aI., 1994; Ke1ce et aI., 1994). Pesticides can also differentially affect the fetus. In a review of the potential effects of chemical carcinogens during pregnancy and the perinatal period, Rice (1979) found that the fetal rat is up to several orders of magnitude more sensitive than the adult to certain carcinogens. Most such agents are direct-acting and independent of metabolism. Rice further reported that the fetus may be less vulnerable than the adult to those substances requiring enzyme-mediated metabolic conversion to chemically reactive derivatives to effect carcinogenesis. Newborns may exhibit the most extreme quantitative differences in sensitivity to chemicals because they are the group most anatomically and physiologically different from adults (Calabrese, 1986; Rodier, 1980). For example, infants may be at higher risk than adults of experiencing serious side effects from organophosphate poisoning because their nervous systems are not fully developed and because their ability to detoxify such agents is also generally lower than that of adults (Fuortes, 1993). Research suggests that neonatal and weanling animals have heightened sensitivity to organophosphates (Mendoza and Shields, 1977). Feeding methyl parathion to pregnant rats led to behavioral changes in offspring without visual signs of maternal toxicity (Gupta et aI., 1985). Lu et al. (1965) found that the LD 50 for malathion in rats increased from 124 mg/kg/day in the newborn to 386 mg/kg/day in the preweanling and to 925 mg/kg/day in the adult. On the other hand, Moretto et al. (1991) reported resistance to organophosphate-induced delayed polyneuropathy among young chickens. He attributed this lack

of sensitivity to a more efficient repair mechanism in developing chicks compared with that in hens. In addition, Pope et al. (1991) reported that cholinesterase activity recovers faster in neonates than in adults. Human Research Evidence of age-related sensitivity among human populations is generally based on ecological or caseseries reports (Garcfa-Rodrfguez et aI., 1996; Garry et aI., 1996; Kristensen et aI., 1997). An investigation of a cluster of cases of congenital abnormalities in a Hungarian village in 1989-1990 by Czeizel et al. (1993) showed that 11 of 15 (73 %) babies born live had congenital abnormalities; of these 11, four had Down's syndrome. The study identified a strong association between the occurrence of Down's syndrome and maternal consumption of fish containing elevated (100 mg/kg) levels of trichlorfon. In a statewide survey of 856 Iowa municipal drinking-water supplies in 1986-1987, Munger et al. (1997) concluded that newborn singletons in communities with herbicide-contaminated wells had elevated rates of intrauterine growth retardation. In a study exploring the role of chlorinated hydrocarbon pesticides in causing spontaneous abortions and premature labor, Saxena et al. (1980) found considerably higher amounts of organochlorine pesticide residues in the circulating blood and placental tissue of women undergoing spontaneous abortion or premature labor compared with the amount of these residues in women in full-term labor. Sharpe et al. (1995) reported a significantly increased risk for Wilms' tumor among Brazilian children whose parent(s) were occupationally exposed to pesticides during pregnancy. Currently, we lack sufficient evidence to define the adverse health effects of pesticide exposure on fetuses; more rigorous research is needed (Garcfa, 1998). 37.3.2.2 Adult and Aged Populations

Most occupational exposures to pesticide occur in healthy adults, and most safety regulations are standardized based upon this population. Scant research has been done that directly addresses pesticide sensitivity in geriatric populations. Although older people usually have higher levels of persistent chlorinated pesticides, these levels have not been associated with increased sensitivity (Sim et al., 1998). Similarly, aging of exposure pathways (e.g., changes in skin permeability) or metabolic pathways (e.g., reduction in hepatic biotransformation capabilities) has not been related to changes in pesticide sensitivity (Roskos et aI., 1986).

37.4 SEX-RELATED SENSITIVITY Sex-specific pesticide sensitivity, as well as sex-specific health effects, may occur after exposure to pesticides both during developmental and reproductive-age periods. Sex-based pharmacokinetic differences that may predispose a person to a sensitivity to pesticides have been described primarily in animal models. Sex-specific health effects of pesticides most often act

37.4 Sex-Related Sensitivity

through the mechanism of disrupting normal hormonal or endocrine relationships. Potential adverse outcomes include decreased sperm count, reduced sperm quality, infertility, premature menopause, or altered sexual behavior (Kavlock et aI., 1996). 37.4.1 ANIMAL STUDIES Research in animal models, primarily in rats, has suggested that males and females may be differentially sensitive to the toxicity of pesticides (Gaines, 1960, 1969). This difference is often attributed to a sex-specific variation in liver microsomal enzyme activity (Snawder and Chambers, 1991). The complexities and variations in pesticide metabolism are discussed in detail elsewhere in this text. Overall, sex does not appear to be a major modifier of the effects of pesticides. For example, chickens administered a single oral dose of an organophospate all experienced a delayed neurotoxicity but the ataxia was significantly more pronounced earlier in males than in females (Odom et al., 1992). However, the ultimate toxic effect of pesticide exposure was the same for both males and females despite variations in enzymatic activity that determined the temporal progression of the reaction. Animal studies of adverse birth outcomes have been used to evaluate the effect of pesticide exposure on paternally mediated birth defects (Anderson et aI., 1996). Potential mechanisms include direct germ-cell effects and indirect effects through transfer of chemicals to the mother via seminal fluid. Olshan and Faustman (1993) reviewed the experimental evidence for malemediated effects on offspring due to a variety of physical and chemical exposures and concluded that more basic animal and human research was necessary to determine the public health relevance of this exposure route. 37.4.2 HUMAN STUDIES: MALES Pesticides can adversely affect spermatogenesis, as documented by the potentially irreversible aspermia after occupational exposure to dibromochloropropane(DBCP) (Uihdetie, 1995; Whorton et aI., 1977). There has been concern that global increases in pesticide use may be causing a worldwide decline in the concentration and quality of sperm (Giwercman et aI., 1993). Empirical results have been conflicting, however. For example, two French studies that used different methods and different populations found varying results. Auger et al. (1995) analyzed sperm from 1351 healthy men and showed a decline in sperm quality among men in Paris from 1973 through 1992. However, Bujan et al. (1996) collected sperm samples from 302 healthy men in Toulouse and found no change in sperm concentrations over a similar period, 1977-1992. In a retrospective study of Danish farmers, Larsen et al. (1998) found no effect of pesticide exposure on male fecundity. Although there have been several meta analyses of more than 60 sperm health studies, the specific relationship between pesticides and male reproductive status remains undefined (Becker and Berhane, 1997).

787

37.4.3 HUMAN STUDIES: FEMALES Sex-specific effects in women may manifest as infertility, endometriosis, or breast cancer. Smith et al. (1997) reported an increased risk for medically diagnosed infertility among women occupationally exposed to pesticides. Lebel et al. (1998) compared plasma concentrations of 11 chlorinated pesticides in 156 women in a case-control study. Neither crude geometric mean concentrations nor crude or adjusted means of the sum of chlordanes, nor the sum of dichlorodiphenyltrichloroethanes (DDEs) differed between case and control subjects. There was no significant linear trend in the adjusted odds ratios for endometriosis as organochlorine concentrations increased. Extensive reviews of the literature on organochlorines have concluded that there is not enough evidence to either support or reject the hypothesis that certain organochlorine compounds [such as DDT, DDE, polychlorinated biphenyls (PCBs), and tetrachloro-p-dioxin (TCDD)] increase the risk for breast, endometrial, or other human cancers (Adami et aI., 1995; Ahlborg et aI., 1995). Two other recent literature reviews concluded that DDE levels are not consistently elevated in women with breast cancer and that occupationally exposed women do not have an increased incidence of breast cancer (Safe and Zacharewski, 1997; Safe, 1997). However, individual epidemiological studies have not shown consistent results. Wolff et al. (1993) reported higher mean (P = 0.31) DDE levels for 58 female case subjects compared with 171 matched control subjects. After adjustment, researchers found that the DDE level was associated with a fourfold increase in relative risk for breast cancer. In a nested case-control study of 150 case and 150 control subjects, Krieger et al. (1994) found no differences (mean difference, 0.2 parts per billion; 95% confidence interval, -6.7, 7.2) between serum levels of DDE in case as compared with control subjects. However, DDE levels were higher among case subjects who were black than among control subjects who were black (mean difference, 5.7 parts per billion; 95% confidence interval, -3.3,14.8). Organochlorinelevels, in general, were significantly higher among black and Asian women than among white women. H0yer et al. (1998) reported that dieldrin level was associated with a significantly increased dose-related risk (adjusted odds ratio, 2.05; 95% confidence interval, 1.17,3.57) for breast cancer among 240 case subjects and 477 control subjects who were originally enrolled in the Copenhagen City Heart Study. B-Hexachlorocyclohexane increased the risk for breast cancer slightly but not significantly (adjusted odds ratio, 1.36; 95% confidence interval, 0.79, 3.57). There were no overall associations between the risk for breast cancer and DDT or DDE. Moysich et al. (1998) found no significant association between DDE exposure (odds ratio, 1.34; 95% confidence interval, 0.71, 2.55) and postmenopausal breast cancer among women enrolled in New York from 1986 to 1991. Hunter et al. (1997) reported that the median level of DDE was lower (P = 0.14) among 226 case subjects than among their matched pairs. Van't Veer et al. (1997) measured DDE in adipose tissue aspirated from the buttocks of women in several European countries and

788

CHAPTER 37


found 9.2% (P = 0.36) lower age-adjusted DDE concentrations among women with breast cancer compared with the concentrations found in control subjects. Pregnant or soon-to-be pregnant women represent a special category of people with a sensitivity to pesticides. Reports of pregnancy loss or of adverse birth outcomes associated with maternal pesticide exposure support the potential for risk (Sever et al., 1997). However, study results have been ambiguous and defined parameters of risk have not been identified (Nurminen, 1995; Savitz et al., 1989).

37.5 GENETIC PREDISPOSITION Genetic polymorphisms (i. e., genes with several variants, or alleles, the rarest of which occurs in at least 1% of the population) are important factors in determining an organism's sensitivity to environmental hazards (Barrett et al., 1997). Genetic sensitivity may influence the activation, detoxification, and cellular uptake of a pesticide (Table 37.1). Polymorphisms in genes that code for metabolic enzymes vary widely within a population and appear to play a primary role in altering sensitivity to environmental exposures (Garte, 1998; Takahashi et al., 1998; Wolff and Weston, 1997).

37.5.1 POLYMORPHISMS IN METABOLIC ENZYMES Pesticides are detoxified in the body through a series of reactions that ultimately biotransform them into water-soluble com-

pounds that are readily eliminated in the urine or feces. Several of the metaboloc enzymes that catalyze these reactions are known to be polymorphic. To date, the best known of these are the cytochrome P450 enzymes, the glutathione S-transferases (GSTs), and Paraoxonase (PONl). Although it is likely that polymorphisms in these enzymes influence human sensitivity to pesticide toxicity, research in this area is just beginning (Costa, 1996; Hirvonen, 1995; Hodgson and Levi, 1996).

37.5.1.1 Cytochrome P450 Enzymes The influence of cytochrome P450 enzyme polymorphisms has probably been best studied in patients with heart disease who demonstrate variable sensitivity to the anticoagulant effects of warfarin. Three allelic variants of the cytochrome P450 2C9 gene (CYP2C9) are known to affect the rate of oxidation of warfarin in humans (Yamazaki et al., 1998). One of the heterozygote forms was found to metabolize (S)-warfarin, but not (R)-warfarin, at a slower rate than wild-type CYP2C9. (S)Warfarin has a greater anticoagulant potency and, therefore, the CYP2C9 polymorphism may account for part of the variability observed between patients treated with therapeutic doses of warfarin (Takahashi et al., 1998). Pesticide exposure and family history have both been identified as risk factors for Parkinson's disease, and several studies have explored possible gene-toxin interactions in the etiology of this disease. In one study, subjects who reported pesticide exposure and who also had the polymorphism that metabolizes debrisoquine (CYP2D6 29B+) were three times more likely to have Parkinson's disease with dementia than Parkinson's with-

Table 37.1 Genetic Polymorphisms Potentially Involved in Modulating Human Susceptibility to Pesticide Toxicity Polymorphism

Effect of atypical phenotype

Pesticide class studied

Effect on toxicity

Reduced metabolism of xenobiotics

Coumarin rodenticides

Increased (Pope et aI., 1991)


Pesticides, in general

Increased (Hubble et aI., 1998)

Reduced metabolism of several

Organophosphates

Increased (Li et aI., 1993;

Fumigants

Increased (Ploemen et aI., 1995;

Metabolic enzymes Cytochrome P450 2C9 (CYP2C9) Cytochrome P450 2D6 (CYP2D6) Paraoxonase (PON1)

Mackness et aI., 1997)

organophosphates (OP) Glutathione S -transferase ()


Thier et aI., 1996)

(GSTTl) Glutathione S -transferase

7r


Pesticides, in general

Increased (Menegon et aI., 1998)

Reduced binding to OPs, which

Organophosphates

Increased (Fontoura-da-Silva and

(GSTP1) Target molecules Butyrylcholinesterase

increases availability of OP to bind

Chautard-Freire-Maia, 1996)

to acetylcholinesterase GABA-gated chloride ion channel Sodium ion channel

Reduced binding of pesticide to ion

Avermectins, cyclodienes

channels on neurons Reduced binding of pesticide to ion channels on neurons

Resistance to toxic effects seen in insects (Ffrench-Constant et aI., 1993)

DDT, pyrethroids

Resistance to toxic effects seen in insects (Knipple et aI., 1994)

37.5 Genetic Predisposition out dementia. Neither pesticide exposure nor the CYP2D6 polymorphism alone was found to be associated with increased risk for Parkinson's disease with dementia (Hubble et aI., 1998).

37.5.1.2 Glutathioue S-Transferases Several classes of GSTs are known to be polymorphic (e.g., GSTMl, GSTTl, GSTPl) and may influence sensitivity to pesticide toxicity. For instance, genomic loci coding for GSTs in the housefly have been implicated in mediating insecticide resistance (Zhou and Syvanen, 1997). Epidemiologic studies have investigated the potential health impact of interactions between GST polymorphisms and pesticide exposures. In a study of Parkinson's disease, patients with idiopathic Parkinson's disease were compared to healthy control subjects. Although pesticide exposure and family history were found to be risk factors for Parkinson's, none of the GST classes was independently associated with illness. However, when the analysis was restricted to participants who reported exposure to pesticides, the distribution of GSTPl genotypes varied significantly between patients with Parkinson's disease and healthy control subjects. This suggests that GSTPl polymorphisms might influence susceptibility to Parkinson's disease after pesticide exposure (Menegon et aI., 1998).

37.5.1.3 Paraoxonase Serum paraoxonase (PONl) is important in the detoxification of the organophosphate (OP) insecticides parathion, diazinon, and chlorpyrifos. Evidence in animals suggests that PONl protects them from poisoning by the OPs it metabolizes. Injection of purified PONl was shown to protect rodents against acute OP toxicity (Li et aI., 1993). In a recent study, PONl-deficient mice were much more sensitive than their wild-type littermates to the toxic effects of chlorpyrifos oxon, the activated form of chlorpyrifos, and to chlorpyrifos itself (Shih et aI., 1998). Studies in humans are just beginning. One recent study compared the PONl genotypes of patients with sporadic idiopathic Parkinson's disease to the genotypes of healthy control subjects. A specific allele of the PON 1 gene was found significantly more often in patients with Parkinson's disease than healthy control subjects. The authors suggested that PONl might influence susceptibility to Parkinson's disease by modulating the effect of the environmental neurotoxins, such as OP pesticides, that are metabolized by PONl (Kondo and Yamamoto, 1998).

37.5.2 POLYMORPHISMS IN TARGET AND TRANSPORT MOLECULES Genetic variability also occurs in molecules that bind to pesticides in the body. Polymorphisms in these genes influence sensitivity to pesticide toxicity by modulating the binding affinity between the pesticide and molecule. Target site (cholinesterases, ion-channel receptors) and transport molecules (serum albumin, P-glycoproteins) have been found

'/89

to exhibit genetic variation in various species and can affect individual sensitivity to pesticides. The cholinesterases (ChE) are the target sites for OP and carbamate pesticides. All vertebrates have two distinct ChEs, acetylcholinesterase (AcChE) and butyrylcholinesterase (BuChE), that hydrolyze the neurotransmitter acetylcholine. OPs bind irreversibly to ChEs, thus effectively inhibiting their ability to metabolize acetylcholine. Evidence suggests that BuChE acts as a scavenger, binding OPs and other toxins before they can bind to and inhibit AcChE. At least 20 polymorphisms in the BuChE gene have been found, most of which are functional and alter the activity of BuChE. The most common BuChE variant occurs in less than 5% of Europeans and Americans and in up to 11 % of people in other populations; therefore, genetic polymorphisms in BuChE may be important contributors to the observed variability in individual susceptibility to OPs and carbamates (Schwarz et aI., 1995). A recent epidemiologic study investigated the association between BuChE polymorphisms and AcChE activity in pesticide-exposed farmers. The farmers were classified as mildly poisoned if their RBC-AcChE activity levels less than 87.5%. The authors reported that the atypical BuChE phenotype was found significantly more often in farmers considered to be mildly poisoned than in the farmers with normal RBC-AcChE activity (::0::87.5%). These data suggest that different BuChE genetic variants offer differential protection against AcChE suppression by the OP and carbamate pesticides (Fontoura-da-Silva and and Chautard-Freire-Maia, 1996). Albumin also binds many compounds in the blood, therefore rendering them unavailable to bind to other molecules or to enter cells. Natural mutants of human serum albumin have different binding affinities for warfarin and may, therefore, affect sensitivity to the effects of warfarin (Vestberg et aI., 1992). Certain pesticides, such as 2,4-D, are bound almost exclusively to serum albumin in humans (Rosso et aI., 1998). Whether or not genetic variation in the binding affinity of serum albumin affects individual sensitivity to these pesticides has not been determined. P-glycoproteins are membrane-transport molecules that remove avermectins from cells in nematodes and mice, and are known to influence sensitivity to avermectin-induced neurotoxicity (Umbenhauer et aI., 1997; XU et aI., 1998). To investigate the role of P-glycoprotein in the transport of pesticides in humans, an experiment was conducted using murine melanoma cells transfected with the human MDRl gene, which codes for P-glycoprotein. The researchers found that a number of pesticides, including ivermectin and several organophosphate and organochlorine pesticides, are capable of binding to human P-glycoprotein. However, none of the pesticides, except endosulfan, was transported out of the cells (Bain and LeBlanc, 1996). Therefore, the role of P-glycoprotein, and consequently the MDRl gene, in altering susceptibility to pesticides in humans remains unclear (Umbenhauer et aI., 1997).

790

CHAPTER 37


37.6 HEALTH STATUS Compromised health status (e.g., malnutrition or immunosuppression) or preexisting disease (e.g., skin disease or seizure disorders) may increase sensitivity to pesticide exposure (Table 37.2). Often interrelationships between these factors are not precisely known. A case report by Solomon et al. (1995) described a neurotoxic reaction to a pesticide in an adult HIVseropositive patient. In this case, the routine application of 1% lindane for microscopically confirmed scabies precipitated seizures and encephalopathy. The authors attribute the adverse outcome to several independent factors, including diminished seizure threshold associated with HIV status, concurrent administration of chlorpromazine, and epidermal barrier dysfunction. Under normal circumstances, the skin acts as a barrier, preventing pesticides and other toxic ants from entering the circulation and causing systemic effects. However, Feldmann and Maibach (1974) applied 12 radiolabeled pesticides and herbicides to the forearm of human subjects and demonstrated that, even under normal circumstances, all of the chemicals tested were absorbed into the systemic circulation. Percutaneous absorption is a complicated process (Wester and Maibach, 1983), involving a series of steps that may be compromised by dermatitis, skin hydration status, or age. For example, increased trancutaneous absorption ascribed to congenital ichthyosiform erythroderma led to nausea and convulsions in a 1-year-old boy after a single application of 1% lindane cream (Friedman, 1987). Similarly, both Lange et al. (1981) and Tenenbein (1991) reported higher lindane blood levels among patients with generalized scabetic dermatitis than among treated patients whose skin was intact. Absorption is also facilitated by the hydration status of skin. Overly hydrated (e.g., recently bathed) skin facilitates systemic absorption. This observation is reported in the literature as a problem when bathing occurs before initiating topical treatment for mite or lice infestations. Overhydration can also potentially occur during the bathing or dipping of pets, especially if the pet owner or animal technician has skin excoriations and is treating

several pets within a short time. Cutaneous absorption is further enhanced if the pesticide is in a lipid vehicle (Solomon and Fahmer,1977). Disease processes that, like many pesticides, target the cholinergic system, may enhance pesticide sensitivity. Alzheimer's disease and several other neurodegenerative disorders (e.g., Parkinson's disease, Huntington's disease, and amyotrophic lateral sclerosis) involve pathological changes in the amount of available AcChE and BuChE (Rakonczay and Brimijoin, 1988). Increased senile plaque formation with decreased AcChE activity and increased circulating BuChE is seen in patients with Alzheimer's disease, and the severity of their dementia parallels the decline in available choline acetyltransferase in the brain (Perry et aI., 1978). These relationships between disease processes and pesticide sensitivity are complex and not well-defined. For example, anticholinergic drugs, such as carbamates and OPs, may cause symptoms similar to those observed in people with degenerative disease. At the same time, recent evidence suggests that use of selective anticholinergic agents may be therapeutic, especially for patients with Alzheimer's disease (Knapp et aI., 1994).

37.7 PREVIOUS OR CONCURRENT EXPOSURES THAT ALTER SENSITIVITY Individual sensitivity to pesticide exposure may be enhanced or diminished by simultaneous exposure to other chemicals. Such synergistic or antagonistic effects are best described in animal and plant models where research is conducted to maximize pesticide efficacy (Bemard and Philogene, 1993). Less is known about human exposure (Krishnan and Brodeur, 1994). In the workplace, simultaneous exposure to different classes of chemicals (e.g., pesticides and solvents) is often inevitable, and personal protective equipment is usually used to avoid interactive effects. Less understood, but increasingly likely, are

Table 37.2 Examples of Preexisting Conditions or Diseases that May Influence Pesticide Metabolism and Pesticide Health Effects Condition

Effect

Reference

Alzheimer's disease

Decreases available AcChE

Schwarz et al. (1995)

Parkinson's disease

Symptoms worsen with anti-ChE exposures

Ott and Lannon (1992)

Down's syndrome

Acetylcholinesterase enzyme deficiency

Percy et al. (1993)

Loiasis

Ivermectin may cause encephalopathy when patient has high

Gardon et al. (1997)

microfilaraemia Asthma

Potential for fatal asthma following pyrethrin inhalation

Wax and Hoffman (1994)

Carbamates enhance lung dysfunction in asthmatics

Senthilselvan et al. (1992)

Infectious hepatitis

Decreased plasma cholinesterase

Wagner (1995)

HIV

Decreased seizure threshold

Solomon et al. (1995)

Excoriated skin

Increases pesticide absorption

Ginsburg et al. (1977)

Increased absorption; seizures after lindane application

Friedman (1987)

Overly hydrated skin Ichthyosiform erythroderma

37.7 Previous or Concurrent Exposures that Alter Sensitivity

the pesticide interactions that may occur in the home due to complex product formulations and multiple product use. In a non-occupational setting the variety of pesticide exposures is usually not well-defined. For example, an analysis of moth- and mite-proofed household products, including vacuum cleaner bags, consistently found mixtures of pyrethroid, synergists, and (DEET) (Utunomiya et aI., 1997). In addition, low-level dietary exposure can occur through consumption of pesticide residues on fruits and vegetables (Melnyk et aI., 1997; Ratner et aI., 1983). Individuals with chronic multiple routes of exposure may be more sensitive to an acute pesticide exposure. This exposure may be further complicated if there is also a simultaneous exposure to nonpesticide cholinesteraseinhibiting chemicals. Such chemicals may potentiate the effect of organophosphate and carbamate exposure, thus potentially further enhancing sensitivity. 37.7.1 SYNERGISM Synergy is the interaction of two or more agents so that their combined effect is greater than the sum of their individual efforts. The result may be pesticide reactions or poisonings among population groups not previously identified as sensitive. A classic example of synergism is illustrated by the interaction of facilitatory toxins (e.g., snake venoms) that increase the release of acetylcholine, and anticholinesterase pesticides that inhibit the destruction of acetylcholine (Harvey and Karlsson, 1982). Similarly, simultaneous or concurrent use of organophospate pesticides may lead to intensified reactions if one pesticide interferes with the physiological detoxification of the other (Cohen and Murphy, 1974). Such an additive effect can also occur when, ostensibly, a single pesticide is being used. Baker et al. (1978) describe the enhanced toxicity of improperly stored malathion used by Pakistani applicators. The same synergistic toxic effect was observed in Belgium after people were exposed to pure malathion that had been stored for more than 5 years. The liquid pesticide had converted to a synergistic, and highly toxic, mixture of malathion and isomalathion (Dive et aI., 1994). Recent research suggests that pesticides frequently used indoors may synergistically interact with the materials that are in house dust (e.g., organic compounds and metals). Kang and Fang (1997) determined that several polycyclic aromatic hydrocarbons (PAHs) commonly found in house dust increased the potency of chlorpyrifos to inhibit AcChE by as much as 85% in vitro. However, there is a fine line between synergism and antagonism, and the combined effect of exposure to different chemicals is still not well-defined. The sequence of exposure may result in disparate outcomes. For example, exposure to the serine esterase inhibitor phenylmethylsulfonyl fluoride (PMSF) before OP exposure can protect a person against organophosphorusinduced delayed neurotoxicity (OPIDN). However, PMSF administration after OP exposure will exacerbate OPIDN (Pope and Padilla, 1990). Research in sheep has suggested that defined

791

periods must elapse between applications of different pesticides to avoid otherwise toxic pesticide interactions (Mohammad and St. Omer, 1985). It is likely that pesticide exposure occurs in combination with exposures to other agents, such as solvents (Petrelli et aI., 1993). Table 37.3 briefly lists results of animal and human studies which show that such combined exposures may alter sensitivity. People who are sensitive to solvents may be at increased risk for illness when they are exposed to low levels of pesticides. 37.7.1.1 Poverty and Pesticides Poverty may increase the potential for exposure to pesticides and also may be the underlying reason for nutritional deficits that heighten sensitivity. Poor people are more likely than those who are not poor to have higher residential exposures from heavy spraying for severe pest infestations in substandard housing (Moses et aI., 1993). Spraying by unlicensed applicators also increases the likelihood that inappropriate strengths or illegal formulations are being used (Esteban et aI., 1996). Often, itinerant and migrant workers receive additional exposures while working in fields recently sprayed with pesticides. The pesticides or pesticide residues are often carried into the home on work clothing. Research has also shown that less affluent people are more likely than affluent people to have higher levels of chlorinated pesticides stored in their bodies (Davies et aI., 1972). Other studies have shown that black people have higher chlorinated pesticide levels than white people (Finklea et aI., 1972; Rogan et aI., 1986). People living in poverty may be more sensitive to pesticide exposure because of nutritional factors such as low body fat, micronutrient imbalance, or protein deficiency. Although scant research has been done among human populations, multiple rat studies suggest that starvation and crowding increase the toxic effects of pesticides (Hayes and Laws, 1991). The combined effect of multiple environmental exposures that interact with pesticides (e.g., lead and tobacco smoke) is not known. 37.7.2 POTENTIATION Potentiation occurs when one exogenous chemical enhances or increases the effect of another. For instance, several overthe-counter and prescription medications potentiate the anticoagulant effect of warfarin. Examples of these potentiators include levamisole (Wehbe and Warth, 1996), ginseng (Janetzky and Morreale, 1997), fluoxetine (Dent and Orrock, 1997), fluoroquinolones (Jolson et aI., 1991), and Danshen (Yu et aI., 1997). People who are exposed to warfarin and who are concurrently exposed to any of the drugs just listed may experience adverse health effects because of potentiation. Although more commonly used therapeutically, warfarin is also an anticoagulant rodenticide. The 1997 Annual Report of the American Association of Poison Control Centers Toxic Exposure Surveillance System categorized 93.5% of the human poison exposures reported by poison control centers in the United States in

792

CHAPTER 37


Table 37.3 Examples of Interaction between Environmental Exposures and Pesticide Exposure Exposure

Example of effect

Reference

Solvents

Lipid solvents increase dermal

Solomon and Fahrner (1977)

absorption (animal) Poverty

Low body fat or starvation increases susceptibility to acute intoxication from organochlorines and organophosphates (animal) Seizures following routine use of lindane in malnourished

Clarke and Clarke (1975), Iyaniwura (1990) Pramanik and Hansen (1979)

child; increases topical effects of y-benzene hexachloride Selenium deficiency Heavy metals

Potentiates paraquat-induced liquid peroxidation of lung

Glass et al. (1985)

tissue Dithiocarbamates increase movement of lead across

Oskarsson and Lind (1983)

blood-brain barrier (animal) Alcohol

Potentiates endosulfan hepatotoxicity (animal)

Singh and Pandey (1991)

Increases methyl parathion-induced chromosomal

Kumar et al. (1993)

aberrations (human) Antagonism of the acute toxicity of parathion (animal)

O'Shaughnesy and Sultatos (1995)

Smoking

Increases chromosomal abberations (human)

Rupa et al. (1989),

Chloroform and

Dithiocarbamates decrease bioactivation and decrease

Gopinath and Ford (1975)

Scarpato et al. (1996, 1997) carbon tetrachloride

toxic effects of chloroform and carbon tetrachloride (animal) Kepone increases hepatotoxicity of chloroform (animal)

1997 (Litovitz et aI., 1998). Of the 2,192,088 human exposures reported, 14,795 cases were associated with exposure to anticoagulants; 90% of the anticoagulant poisonings were among children under six years of age. Thus it appears that, when anticoagulant poisonings occur, they are often among young children. 37.7.3 NONPESTICIDE CHEMICALS THAT INHIBIT CHOLINESTERASE People can also exhibit sensitivity to pesticides when nonpesticide ChE-inhibiting chemicals alter the effect of OPs and carbamates and vice-versa. For example, Ware et al. (1990) describe a prolonged response to the neuromuscular blockade effects of succinylcholine in a patient whose exposure to an organophosphate pesticide had occurred within the week that the exposure occurred. OP and carbamate pesticides inhibit AcChE and interfere with nerve conduction. When a person is exposed to these pesticides, limited amounts of circulating BuChE (or pseudocholinesterase) may act as toxin scavengers. This action potentially prevents the pesticides from interacting with AcChE and disrupting nerve conduction (Neville et aI., 1990a, b; Schwarz et aI., 1995). However, such a potentially protective effect may be lost if the BuChE is reacting to synthetic or naturally occurring chemicals that can also act as ChE substrates. If sufficient quantities of these exogenous drugs are present, then a person may be more susceptible to lower levels of OP and carbamate insecticides than he or she would be otherwise. Nonpesticide

Hewitt et al. (1986)

chemicals that inhibit cholinesterase may also potentiate the effect of pesticide exposure. This sensitivity is similar to the sensitivity imposed by the genetic BuChE allele discussed in Section 40.4. Intuitively, if BuChE is an important first line of defense against low-level pesticide exposure, then people with decreased levels of BuChE will be at greater risk of poisoning from cholinesterase inhibitors than people whose levels are within normal limits (Anton, 1988; Cregler, 1989; Devenyi, 1989). Nonetheless, the opposite effect was observed when the corollary of this scenario was tested. Rats were pretreated with an OP (tetraisopropyl pyrophosphoramide) and then given a low toxic dose of cocaine (a BuChE substrate). Contrary to expectations, significantly more fatalities occurred among the rats that did not receive OP pretreatment than among those that did (Kambam et al., 1992a). Results of a similar experiment with pigs showed that inhibition of BuChE accelerated the metabolism of cocaine through an alternative route (Kambam et aI., 1992b). Although interaction clearly occurs when people are simultaneously exposed to different cholinesterase-binding chemicals, the direction of such an interaction and the manifestation of adverse health effects is not well-demonstrated; further research is necessary to understand these relationships. Rats that were pretreated with monoclonal antibodies to rat AcChE and then given a dose of an OP did demonstrate reduced AcChE activity but did not exhibit neurobehavioral changes (Padilla et aI., 1992). To further complicate matters, other drugs (e.g., tricyclic antidepressants) may act as anticholinergic blocking agents (Bal et aI., 1990). These drugs act like atropine on all

37.8 Implications for Public Health

793

Table 37.4 Naturally Occurring and Synthetic Cholinesterase Substrates Type of chemical

References

Examples

Analgesics

Aspirin, acetaminophen

Valentino et al. (1981)

Narcotics

Cocaine, procaine, heroin

Isenschmid et al. (1989),

Fungal antibiotics

Puromycin

Hersh (1981)

Gatley (1991) Onchidal

Mollusc secretion

Abrahamson et al. (1989)

Reptile polypeptides

FascicuIin

Karlsson et al. (1985)

Metals

Aluminum, scandium, yttrium

Marquis and Lerrick (1982)

Neuromuscular relaxants

Succinylcholine

Neville et al. (1990a)

Physostigmine derivatives

Neostigmine, demecarium, pyridostigmine

Shaw et al. (1985) Coleman et al. (1987)

muscarinic sites and may mask the effects of pesticide exposure (Marrs, 1993). Although this action essentially protects a person from all but nicotinic and central-nervous-system effects, it increases the likelihood of misdiagnosis of pesticide poisoning. Exogenous cholinesterase inhibitors may also have the effect of increasing the overall effective dose. Low-level exposure to OPs may result in a person exhibiting neurological symptoms when other cholinesterase inhibitors are also in the blood. Health care providers must be aware of the variety of environmental and recreational chemicals (Table 37.4) that are cholinesterase inhibitors because, even if they do not magnify or mask clinical signs of pesticide poisoning, their presence may affect treatment decisions and treatment efficacy.

37.8 IMPLICATIONS FOR PUBLIC HEALTH Historically, risk assessment models for pesticide toxicity have been formulated to reflect the pharmacokinetic patterns of the adult male. Pesticide regulations defining permissible levels of exposure have been based upon these risk assessments and have been primarily concerned with occupational exposures that are most relevant to a healthy working adult male. It is only recently that exposure levels in children, parous women, and other sensitive population groups have been acknowledged. In 1988, Congress requested the National Academy of Sciences (NAS) to evaluate the V.S. Environmental Protection Agency's (EPA) existing risk assessment methods to determine if they were adequately addressing the exposure and the potential risk that pesticides may pose to infants and children. That same year, the Children's Health Protection Advisory Committee (which had been formed to advise, consult with, and make recommendations to the EPA regarding reevaluation of existing EPA regulations to better protect children's health) reported that selected pesticides (triazines, OPs, and carbamates) were among the top five priority issues (Reigart, 1988). In 1993, the NAS report, Pesticides in the Diet of Infants and Children, recommended that future studies should compare metabolism and toxicity in adult and immature animals and that carcinogenicity

research should consider in utero exposure. EPA responded by revising its residue guidelines (Fenner-Crisp, 1995). In 1996, the Food Quality Protection Act (FQPA) was passed. It sought to provide added protection against pesticide risk, especially for infants and children, by setting a lO-year schedule for EPA to reevaluate 10,000 existing tolerances for pesticide residues on food (Goldman, 1998). In the absence of reliable data on children's toxicity, the FQPA directed EPA to use an extra lO-fold safety factor in its risk assessments to ensure that the greater sensitivity of infants and children would be addressed. 37.8.1 INFORMATION GAPS AND RESEARCH NEEDS Despite legislative progress in defining risk, most risk assessment models continue to rely upon animal data which may not accurately represent human exposure or human health outcomes. For example, based upon 2-year low-dose rat studies, proteinuria was identified as the most sensitive toxic end point for chlordecone exposure; however, even in high-dose occupational exposures, proteinuria has not been observed as a human health outcome (Guzelian, 1992). Similarly, animal models may not be able to adequately identify the effect of pesticide exposure on endocrine function during crucial periods of neurological development (Tilson, 1998). As discussed throughout this chapter, sensitive population groups experience numerous routes of exposure to an increasing variety of pesticide formulations. Risk assessment should reflect such exposure realities. Measuring pesticide metabolites in human biological specimens is the most accurate method of defining human exposure. Studies that actually quantify human exposure can more accurately associate adverse health effects with specific pesticides. Traditional risk assessment has used high-dose exposures (e.g., animal models or accidental high-dose occupational exposures) to predict low-dose health outcomes. This scenario may not be relevant to exposures that impact the endocrine system, especially during periods of rapid development (Sheehan and

794

CHAPTER 37


vom Saal, 1997). Both the FQPA and the Safe Drinking Water Act require EPA to develop a screening program to test whether exogenous chemicals affect humans in the same way as a naturally occurring estrogen. Research activities must be focused on those populations (e.g., fetuses) most vulnerable to exogenous estrogens. Responsible use of pesticides and rational pesticide regulations must consider sensitive human population groups. Further research that includes assessment of all exposure routes, human biomonitoring of exposure, and the timing of exposure relative to development stages is necessary.

ACKNOWLEDGMENT The authors acknowledge Kim Blindauer, Amanda Niskar, Luke Naeher, and Anyana Banerjee for contributing both time and expertise to the preparation of this chapter.

REFERENCES Abrahamson, S. N., Zoran, R, Manker, D., Faulkner, J. D., and Taylor, P. (1989). Onchidal: A naturally occurring irreversible inhibitor of acetylcholinesterase with a novel mechanism of action. Mol. Pharmacol. 36, 349354. Adami, H., Lipworth, L., Titus-Emstoff, L., Hsieh, C., Hanberk, A., Ahlborg, u., Baron, J., and Trichopoulos, D. (1995). Organochlorine compounds and estrogen-related cancers in women. Cancer Causes Control 6, 551-566. Ahlborg, U. G., Lipworth, L., Titus-Emstoff, L., Chung-Cheng, H., Hanberg, A., Baron, J., Richopoulos, D., and Adami, H.-O. (1995). Organochlorine compounds in relation to breast cancer, endometrial cancer and endometriosis: An assessment of the biological and epidemiological evidence. Crit. Rev. Toxicol. 25, 463-53\. Anderson, D., Edwards, A. J., Brinkworth, M. H., and Hughes, J. A. (1996). Male-mediated FI effects in mice exposed to 1,3-butadiene. Toxicology 113, 120-127. Anton, A. H. (1988). Unexpected cocaine-induced fatalities: A possible cause (letter). Drug Intelligence Clin. Pharm. 22,914. Auger, J., Kunstmann, J. M., Czyglik, F., and Jouannet, P. (1995). Decline in semen quality among fertile men in Paris during the past 20 years. New England J. Med. 332,281-285. Bemard, c., and Philogene, B. (1993). Insecticide synergists: Role, importance, and perspectives. J. Toxicol. Environmen. Health 38, 199-223. Bain, L. J., and LeBlanc, G. A. (1996). Interaction of structurally diverse pesticides with the human MDRl gene product p-glycoprotein. Toxicol. Appl. Pharmacol. 141,288-298. Baker, E. L., Jr., Warren, M., Zack, M., Dobbin, R. D., Miles, J. w., Miller, S., Alderman, L., and Teeters, W. R (1978). Epidemic malathion poisoning in Pakistan malaria workers. Lancet 1, 31-34. Baldessarini, R J. (1985). "Chemotherapy in Psychiatry" (revised and enlargen ed.). Harvard Univ. Press, Cambridge, MA. Baldessarini, R J., Marsh, E. R, Kula, N. S., Zong, R S., Gao, Y. G., and Neumeyer, J. L. (1990). Effects of isomers of hydroxyaporphines on dopamine metabolism in rat brain regions. Biochem. Pharmacol. 40,417423. Bamett, H. L., McNamara, H., Shultz, S., and Tompsett, R (1949). Renal clearances of sodium penicillin G, procaine penicillin G, and inulin in infants and children. Pediatrics 418-422. Barrett, C. J., Vainio, H., Eakall, D., and Goldstein, B. D. (1997). 12th meeting of the scientific group on methodologies for the safety evaluation of chemicals: Susceptibility to environmental hazards. Environ. Health Perspect. 105(Suppl. 4), 699-737.

Becker, S., and Berhane, K. (1997). A meta-analysis of 61 sperm count studies revisited. Fertility and Sterility 67, 1103-1108. Benjaminov, 0., Hoffer, E., and Taitelman, U. (1992). Parathion transfer and acetylcholinesterase activity in an in-vitro prefused term human placenta. Vet. Hum. Toxieol. 34, 10-12. Bradman, M. A., Hamly, M. A., Draper, W., Seidel, S., Teran, S., Wakeham, D., and Neutra, R (1997). Pesticide exposures to children from California's central valley: Results of a pilot study. J. Exposure Anal. Environ. Epidemiol. 7,217-234. Bujan, L., Mansat, A., Pontonnier, F., and Mieusset, R (1996). Time series analysis of sperm cocentration in fertile men in Toulouse, France between 1977 and 1992. Br. Med. J. 312, 471-472. Calabrese, E. J. (1986). "Age and Susceptibility in Toxic Substances," Environmental Sience and Technology Series. Wiley-Interscience, New York. Clarke, E. G. c., and Clarke, M. L. (1975). "Veterinary Toxicology," (1st ed.). Williams & Wilkins, Baltimore. Cohen, S. D., and Murphy, S. D. (1974). A simplified bioassay for organophosphate detoxification and interactions. Toxicol. and Appl. Pharmacol. 27, 537-550. Coleman, B. A., Michel, L., and Oswald, R (1987). Interaction of a benzomorphan opiate with acetylcholinesterase and the nicotinic acetylcholine receptor. Mol. Pharmacol. 32, 456-462. Cone, J. E., and SuIt, T. A. (1992). Acquired intolerance to solvents following pesticide/solvent exposure in a building: A new group of workers at risk for multiple chemical sensitivities? Toxicol. Ind. Health 8, 29-39. Corasaniti, M. T., Defilippo, R., Rodino, P., Nappi, G., and Nistico, G. (1991). Evidence that paraquat is able to cross the blood-brain barrier to a different extent in rats of various age. Funct. Neurol. 6, 385-39\. Costa, L. G. (1996). Biomarker research in neurotoxicology: The role of mechanistic studies to bridge the gap between the laboratory and epidemiological investigations. Environ. Health Perspect. 104(Suppl. 1),55-67. Cregler, L. L. (1989). Protracted elimination of cocaine metabolites (letter). Am. J. Med. 86, 632-633. Czeizel, A. E., Elek, c., Gundy, S., Metneki, J., Nemes, E., Reis, A., Sperling, K., Timar, L., Tusnady, G., and Vinigh, Z. (1993). Environmental trichlorform and cluster of congenital abnormalities. Lancet 341, 539-542. Davies, J. E., Edmundson, W. E, Rafonell, A., Cassady, J. c., and Morgade, C. (1972). The role of social class in human pesticide pollution. Am. J. Epidemiol. 96,334-341. Davis, R J., Brownson, R c., and Garcia, R (1992). Family pesticide use in the home, garden, orchard, and yard. Arch. Environ. Contam. Toxicol. 22, 262-266. Dent, L. A., and Orrock, M. W. (1997). Warfarin-fluoxetine and diazepamfluoxetine interaction. Pharmacotherapy 17, 170-172. Devenyi, P. (1989). Cocaine implications and pseudocholinesterase (letter). Ann. Internal Med. 110,167-168. Dive, A., Mahieu, P., Van Binst, R, Hassoun, A., Lison, D., De Bisschop, H., Nemery, B., and Lauwerys, R. (1994). Unusial manifestations after malathion poisoning. Hum. Exp. Toxicol. 13,271-274. Done, A. K. (1964). Developmental pharmacology. Clin. Pharmacol. Ther. 5, 432-479. Drew, T. R, Boorman, G. A., Haseman, J. K., McConnell, E. E., Busey, W. M., and Moore, J. A. (1983). The effect of age and exposure duration on cancer induction by a known carcinogen in rats, mice, and hamsters. Toxicol. Appl. Pharmacol. 68, 120-130. Esteban, E., Rubin, c., Hill, R., Olson, D., and Pearce, K. (1996). Association between indoor residential contamination with methyl parathion and urinary para-nitrophenol. J. Exposure Anal. Environ. Epidemiol. 6, 375-387. Feldmann, R J., and Maibach, H. 1. (1974). Percutaneous penetration of some pesticides and herbicides in man. Toxicol. Appl. Pharmacol. 28, 126-132. Fenner-Crisp, P. A. (1995). Pesticides-the NAS report: How can the recommendations be implemented. Environ. Health Perspect. 106, 159-162. Ffrench-Constant, RH., Steichen, J. c., Rocheleau, T. A., Aronstein, K., and Roush, R T. (1993). A single-amino acid substitution in a gammaaminobutyric acid subtype A receptor locus is associated with cyclodiene insecticide resistance in Drosophila populations. Proe. Natl. Aead. Sei. U.S.A. 90, 1957-1961.

References

Finhom, L. (1982). Oncodevelopmental biology and medicine. J. Int. Soc. Oncodev. BioI. Med. 4, 219-229. Finklea, J., Priester, L. E., Crason, J. P., Hauser, T., Hinners, T., and Hammer, D.1. (1972). Polychlorinated biphenyl residues in human plasma expose a major urban pollution problem. Am. J. Public Health 62, 645-651. Fontoura-da-Silva, S. E., and Chautard-Freire-Maia, E. A. (1996). Butyrylcholinesterase variants (BCHE and CHE2 loci) associated with erythrocyte acetylcholinesterase inhibition in farmers exposed to pesticides. Hum. Heredity 46, 142-147. Forbes, G. B. (1987). "Human Body Composition: Growth, Aging, Nutrition, and Activity." Springer-Verlag, New York. Friedman, S. J. (1987). Lindane neurotoxic reaction in nonbullous congenital ichthyosifonn erythrodenna. Arch. Dermatol. 123, 1056-1058. Fuortes, L. J. (1993). Cholinesterase-inhibiting pesticide toxicity. In "Case Studies in Environmental Medicine" (S. Wagner, ed.). U.S. Department of Health & Human Services, Agency for Toxic Substances and Disease Registry, Atlanta. Gaines, T. B. (1960). The acute toxicity of pesticides to rats. Toxicol. Appl. Pharmacol. 2,88-99. Gaines, T. B. (1969). Acute toxicity of pesticides. Toxicol. Appl. Pharmacol. 14,515-534. Garcfa, A. M. (1998). Occupational exposure to pesticides and congenital malfonnations: A review of mechanisms, methods, and results. Am. J. Ind. Med. 33, 232-240. Garcfa-Rodriguez, J., Garcfa-Martfn, M., Nogueras-Ocafia, M., Luna-delCastillo, J., Garcfa, M. E., Olea, N., and Lardelli-Claret, P. (1996). Exposure to pesticides and cryptorchidism: Geographical evidence of a possible association. Environ. Health Perspect. 104, 1090-1095. Gardon, J., Gardon-Wendel, N., Ngangue, D., Kamgno, J., and Chippaux, J.-P. (1997). Serious reactions after mass treatment of onchocerciasis with ivermectin in an area endemic for Laa loa infection. Lancet 350, 18-22. Garry, V. F., Schreinemachers, D., Harkins, M. E., and Griffith, J. (1996). Pesticide appliers, biocides, and birth defects in rural Minnesota. Environ. Health Perspect. 104, 394-399. Garte, S. (1998). The role of ethnicity in cancer susceptibility gene polymorphisms: The example ofCYPIAI. Carcinogenesis 19, 1329-1332. Gatley, S. J. (1991). Activities of the enantiomers of cocaine and some related compounds as substrates and inhibitors of plasma butyry lcholinesterase. Biochem. Pharmacol. 41, 1249-1254. Ginsburg, C. M., Lowry, w., and Reisch, J. S. (1977). Absorption of lindane (gamma benzene hexachlorine) in infants and children. J. Pediatrics 91, 998-1000. Giwercman, A., Carlsen, E., Keiding, N., and Skakkeb<ek, N. E. (1993). Evidence for increasing incidence of abnonnalities of the human testis: A review. Environ. Health Perspect. 101(Supp!. 3),65-71. Glass, M., Sutherland, M. w., Fonnan, H. J., and Fisher, A. B. (1985). Selenium deficiency potentiates paraquat-induced lipid peroxidation in isolated perfused rat lung. J. Appl. Physiol. 59, 619-622. Goldman, L. R. (1995). Children-unique and vulnerable. Environmental risks facing children and recommendations for response. Environ. Health Perspect. 103(Supp!. 6), 13-18. Goldman, L. R. (1998). Chemicals and children's environment: What we don't know about risks. Environ. Health Perspect. 106(Supp!. 3), 875-880. Gopinath, c., and Ford, J. H. (1975). The role of microsomal hydroxylases in the modification of chlorofonn hepatotoxicity in rats. Br. J. Exp. Pathol. 56, 412-422. Gray, E. L., Ostby, J. S., and Kelce, W. R. (1994). Developmental effects of an environmental antiandrogen: The fungicide vinclozolin alters sex differentiation of the male rat. Toxicol. Appl. Pharmacol. 129,46-52. Greaves, S. J., Ferry, D. G., McQueen, E. G., Malcolm, D. S., and Buckfield, P. M. (1975). Serial hexachlorophene blood levels in the premature infant. New Zealand Med. J. 81,334-336. Gupta, R. C., Rech, R. H., Lovell, K. L., Welsch, F., and Thomburg, J. E. (1985). Brain cholinergic, behavioral, and morphological development in rats exposed in utero to methylparathion. Toxicol. Appl. Pharmacol. 77,405-413.

795

Gurunathan, S., Robson, M., Freeman, N., Buckley, B., Roy, A., Meyer, R., and Bukowski, J. (1998). Accumulation of chlorpyrifos on residential surfaces and toys accessible to children. Environ. Health Perspect. 106,9-16. Guzelian, P. S. (1992). The clinical toxicology of chlorodecone as an example of toxicology risk assesment for man. Toxicol. Lett. 64-65. Hall, L. L., Henry, L. F., Martha, R. S., Robert, J. M., Neil, c., and Shah, P. V. (1988). Dose response of skin absoprtion in young and adult rats. In "Perfonnance of Protective Clothing: Second Symposium, ASTM STP 989" (S. Z. Mansdorf, R. Sager, and A. P. Nielsen, eds.), pp. 177-194. Philadelphia. Harvey, A. L., and Karlsson, E. (1982). Protease inhibitor homologues from mamba venoms: Facilitation of acetylchole release and interactions with prejunctional blocking toxins. Br. J. Pharmac. 77, 153-161. Hayes, W. J., and Laws, E. R., eds. (1991). "Handbook of Pesticide Toxico!." Vo!. I. Academic Press, San Diego. Hersh, L. B. (1981). Inhibition of aminopeptidase and acetylcholinesterase by puromycin and puromycin analogs. J. Neurochem. 36, 1594-1596. Hewitt, A. L., Caille, G., and Plaa, G. L. (1986). Temporal relationships between biotransfonnation, detoxification, and chlordecone potentiation of chlorofonn-induced hepatotoxicity. Can. J. Physiol. Pharmacol. 64,477482. Hill, R. H., Jr., Head, S. L., Baker, S., Gregg, M., Shealy, D. B., Bailey, S. L., Williams, C. c., Sampson, E. J., and Needham, L. L. (1995). Pesticide residues in urine of adults living in the United States: Reference range concentrations. Environ. Res. 71,99-108. Hirvonen, A. (1995). Genetic factors in individual responses to environmental exposures. J. Occupational Environ. Med. 37, 37-43. Hodgson, E., and Levi, P. E. (1996). Pesticides: An important but underused model for the environmental health sciences. Environ. Health Perspect. 104(Supp!. 1),97-106. Hoffmann, H. (1982). Absorption of drugs and other xenobiotics during development in experimental animals. Pharmacol. Ther. 16, 247-260. H0yer, A. P., Grandjean, P., J0gensen, T., Brock, J. W., and Hartvig, H. B. (1998). Organochlorine exposure and risk of breast cancer. Lancet 352, 1816-1820. Hubble, J. P., Kurth, J. H., Glatt, S. L., Kurth, M. C., Schellenberg, G. D., Hassanein, R. E. S., Liebennan, A., and Koller, W. C. (1998). Gene-toxin interaction as a putative risk factor for Parkinson's disease with dementia. Neuroepidemiology 17,96-104. Hudson, R. H., Richard, K. T., and Haegele, M. A. (1972). Effect of age on sensitivity: Acute oral toxicity of 14 pesticides to mallard ducks of several ages. Toxicol. Appl. Pharmacol. 22, 556-561. Hunter, D. J., Hankinson, S. E., Laden, F., Colditz, G. A., Manson, J. E., Willett, W. c., Speizer, F. E., and Wolff, M. S. (1997). Plasma organochlorine levels and the risk of breast cancer. New England J. Med. 337, 12531308. Isenschmid, D. S., Levine, B. S., and Caplan, Y. H. (1989). A comprehensive study of the stability of cocaine and its metabolites. J. Anal. Toxicol. 13, 250-256. Iyaniwura, T. T. (1990). Mammalian toxicity and combined exposure to pesticides. Vet. Hum. Toxicol. 32, 58-62. Janetzky, K., and Morreale, A. P. (1997). Probable interaction between warfarin and ginseng. Am. J. Health System Pharm. 54, 692-693. Jolson, H. M., Tanner, A. L., Green, L., and Grasela, T. H. (1991). Adverse reaction reporting of interaction between warfarin and fluoroquinolones. Arch. Internal Med. 151, 1003-1004. Kalland, T. (1982). Long-tenn effects on the immune system of an early exposure to diethylstilbestero!. In "Banbury Report: Environmental Factors in Human Growth and Development" (V. R. Hunt, M. K. Smith, and D. Worth, eds.). Cold Spring Harbor Laboratory Press, Cold Spring Harbor, NY. Kambam, J., Mets, B., Hickman, R. M., Janicki, P., James, M. F. M., Fuller, B., and Kirsch, R. E. (I 992a). The effects of inhibition of plasma cholinesterase and hepatic microsomal enzyme activity on cocaine, benzoylecgonine, ecgonine methyl ester, and narcocaine blood levels in pigs. J. Lab. CUn. Med. 120, 323-328.

796

CHAPTER 37


Kambam, J. R, Naukam, R., and Bennan, L. M. (1992b). Inhibition of pseudocholinesterase activity protects from cocaine-induced cardiorespiratory toxicity in rats. J. Lab. Clin. Med. 119,553-556. Kang, J., and Fang, H. (1997). Polycyclic aromatic hydrocarbons inhibit the activity of acetylcholinesterase purified from electric eel. Biochem. Biophys. Res. Commun. 238, 367-369. Karlsson, E., Mbugua, P. M., and Rodrfguez-Ithurralde. (1985). Anticholinesterase toxins. Pharmacal. Ther. 30, 259-276. Kavlock, R. J., Daston, G. P., DeRosa, C., Fenner-Crisp, P., Gray, E. L., Kaattari, S., Lucier, G., Luster, M., Mac, M. J., Maczka, C. M. R, Miller, R, Moore, J., Rolland, R, Scott, G., Sheehan, D. M., Sinks, T, and Tilson, H. A (1996). Research needs for the risk assessment of health and environmental effects of endocrine disruptors: A report of the U.S. EPAsponsored workshop. Environ. Health Perspect. 104(Suppl. 4), 1-36. Kearns, L. G., and Reed, M. D. (1989). Clinical phannacokinetics in infants and children: A reappraisal. Clin. Pharmacokinetics 17,29-67. Kelce, W. R, Monosson, E., Gamcsik, M. P., laws, S. c., and Gray, L. E. (1994). Environmental honnone disruptors: Evidence that vinclozolin developmental toxicity is mediated by antiandrogenic metabolites. Toxicol. Appl. Pharmacal. 126, 276-285. Knapp, M. J., Knopman, D. S., Solomon, P. R., Pendlebury, W. W., Davis, C. S., and Gracon, S. I. (1994). A 30-week randomized controlled trial of a high dose tacrine in patients with Alzheimer's disease. J. Am. Med. Assoc. 271, 985-991. Knipple, D. C., Doyle, K E., Marsella-Herrick, P. A, and Sodurland, D. M. (1994). Tight genetic linkage between the kdr insecticide resistance trait and a voltage-sensitive sodium channel gene in the house fly. Proc. Natl. Acad. Sci. U. S. A. 91, 2483-2487. Kondo, I., and Yamamoto, M. (1998). Genetic polymorphisms of paraoxonase 1 (PON1) and susceptibility to Parkinson's disease. Brain Res. 806, 271-273. Kreuzer, P. E., Csanady, G. A, Baur, c., Papke, K. 0., Greim, H., and Filser, J. G. (1997). 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and congeners in infants. A toxicokinetic model of human lifetime body burden by TCDD with special emphasis on its uptake by nutrition. Arch. Toxicol. 71,383-400. Krieger, N., Wolff, M. S., Hiatt, R. A., Rivera, M., Vogelman, J., and Ostreich, N. (1994). Breast cancer and serum organochlorines: A prospective study among white, black, and Asian women. J. Natl. Cancer Inst. 86, 589599. Krishnan, K, and Brodeur, J. (1994). Toxic interactions among environmental pollutants: Corroborating laboratory observations with human experience. Environ. Health Perspect. 102(Suppl. 9), 11-17. Kristensen, P., Irgens, L. M., Anderson, A., Bye, A. S., and Sundheim, L. (1997). Birth defects among offspring of Norwegian fanners, 1967-1991. Epidemiology 8, 537-544. Kumar, S. K, Ankathil, R, and Devi, K S. (1993). Chromosomal aberrations induced by methyl parathion in human peripherallymphocytes of alcoholics and smokers. Hum. Exp. Toxicol. 12, 285-288. Lahdetie, J. (1995). Occupation- and exposure-related studies on human spenn. J. Occupational Environ. Med. 37,922-929. Lange, M., Nitzche, K, and Zesch, A (1981). Percutaneous absorption of lindane in healthy volunteers and scabies patients. Arch. Dermatol. Res. 271, 387-399. Larsen, S. B., Joffe, M., Bonde, J. P., and ASCLEPIOS study group. (1998). Time to pregnancy and exposure to pesticides in Danish fanners. Occupational Environ. Med. 55, 278-283. Lebel, G., Dodin, S., Ayotte, P., Marcoux, S., Ferron, L. A, and Dewailly, E. (1998). Organochlorine exposure and the risk of endometriosis. Fertility Sterility 69, 221-228. Lester, R S. (1983). Topical fonnulatory for the pediatrician. Pediatric Clinician North Am. 30, 749-765. Li, W., Costa, L. G., and Furlong, C. E. (1993). Serum paraoxonase status: A major factor in detennining resistance to organophosphates. J. Toxicol. Environ. Health 40, 337-346. Litovitz, T L., Klein-Schwartz, W., Dyer, K S., Shannon, M., Lee, S., and Powers, M. (1998). 1997 Annual Report of the American Association of Poison Control Centers Toxic Exposure Surveillance System. Am. J. Emergency Med. 16,443-497.

Lu, E C., Jessup, D. c., and Lavallee, A (1965). Toxicity of pesticides in young versus adult rats. Food Cosmetic. Toxicol. 3,591-596. Mackness, M., Mackness, M. I., Arrol, S., Turki, W., and Durrington, P. N. (1997). Effect of the molecular polymorphisms of human paraoxonase (PON1) on the rate of hydrolysis of paraoxon. Br. J. Pharmacal. 122,265268. Marquis, J. K, and Lerrick, A. J. (1982). Noncompetitive inhibition by aluminum, scandium and yttrium of acetylcholinesterase from Electrophorous electricus. Biochem. Pharmacal. 31, 1437-1440. Marrs, T C. (1993). Organophosphate poisoning. Pharmac. Ther. 58,51-66. Melnyk, L. J., Berry, M. R., and Sheldon, L. S. (1997). Dietary exposure from pesticide application on fanns in the agricultural health pilot study. J. Exposure Anal. Environ. Epidemiol. 7,61-80. Mendoza, C. E., and Shields, J. B. (1977). Effects on esterases and comparison of 150 and LD50 values of malathion in suckling rats. Bull. Environ. Contam. Toxicol. 17,9-15. Menegon, A., Board, P. G., Blackburn, A. C., Mellick, G. D., and Le Couteur, D. G. (1998). Parkinson's disease, pesticides, and glutathione transferase polymorphisms. Lancet 352, 1344-1346. Miller, R. P., Roberts, R. J., and Fischer, L. J. (1976). Acetominophen elimination kinetics in neonates, children, and adults. Clin. Pharmacal. Ther. 19, 284-294. Mohammad, E K, and SI. Omer, V. V. (1985). Toxicity and interaction of topical organophosphate insecticide dichlorvoscrotoxyphos and phenothiazine anthelmintic in sheep previously exposed to drugs. Vet. Hum. Toxicol. 27, 181-184. Moretto, A., Capodicasa, M., and Lotti, M. (1991). Age sensitivity to organophosphate-induced delayed polyneuropathy: Biochemical and toxicological studies in developing chicks. Biochem. Pharmacal. 41, 14971504. Morselli, P. L. (1976). Clinical phannacokinetics in neonates. Clin. Pharmacokinetics 1,81-98. Morselli, P. L., Morselli, R E, and Bossi, L. (1980). Clinical phannacokinetics in newborns and infants: Age related diferences and therapeutic implications. Clin. Pharmacokinetics 5, 485-527. Mortensen, S. R, Chanda, S. M., Hooper, M. J., and Padilla, S. (1996). Maturational differences in chlorpyrifos-oxonase activity may contribute to agerelated sensitivity to chlorpyrifos. J. Biochem. Toxicol. 11,279-287. Moses, M., Johnson, E. S., Anger, W. K, Burse, V. W., Horstman, S. W., Jackson, R. J., Lewis, R. G., Maddy, K T, McConnell, R, Meggs, W. J., and Zahm, S. H. (1993). Environmental equity and pesticide exposure. Toxicol. Ind. Health 9, 914-959. Moysich, K B., Ambrosone, C. B., Vena, J. E., Shields, P. G., Mendola, P., Kostyniak, P., Greizerstein, H., Graham, S., Marshall, J. R, Schistennan, E. E, and Freudenheim, J. L. (1998). Environmental organochlorine exposure and postmenopausal breast cancer risk. Cancer Epidemio!. Biomarkers Prevention 7, 181-188. Munger, R, Isacson, P., Hu, S., Burns, I., Hnason, J., Lynch, C. E, Cherryholmes, K, Van Dorpe, P., and Hausler, W. J. (1997). Intrauterine growth retardation in Iowa communities with herbicide-contaminated drinking water supplies. Environ. Health Perspect. 105,308-314. Neville, L. E, Gnatt, A, Loewenstein, Y, and Soreq, H. (1990). Aspartate70 to glycine substitution confers resistance to naturally ocurring and synthetic anionic-site ligands on in-ovo produced human butyry lcholinesterase. J. Neurosci. Res. 27, 452-460. Neville, L. E, Gnatt, A, Padan, R., Seidman, S., and Soreq, H. (1990b). Anionic site interactions in human butyrylcholinesterase disrupted by two single point mutations. J. Bio!. Chem. 265, 20735-20738. NRC (1993). "Pesticides in the Diets of Infants and Children." National Academy Press, Washington, DC. Nunninen, T (1995). Maternal pesticide exposure and pregnancy outcome. J. Occupational Environ. Med. 37,935-940. Odom, A, Gross, B. W., and Ehrich, M. (1992). Role of socialization, stress and sex of chickens in response to anesthesia and in response to an organosphophate neurotoxicant. Vet. Hum. Toxicol. 34, 134-137. Olshan, A E, and Faustman, E. M. (1993). Male-mediated developmental toxicity. Ann. Rev. Public Health 14, 159-181.

References

O'Shaughnesy, J, A., and Sultatos, L. G. (1995). Interaction of ethanol and the organophophorous insecticide parathion. Biochem. Pharmacol. 50, 19251932. Oskarsson, A., and Lind, B. (1983). Increased lead levels in brain after longterm treatment with lead and dithiocarbamate or thiuran derivatives in rats. Acta Pharmacol. Toxicol. 56,309-315. Ott, B. R., and Lannon, M. C. (1992). Exacerbation of Parkinsonism by Tacrine. Clin. Neuropharmacol. 4,322-325. Padilla, S., Moser, V. c., Pope, C. N., and Brimijon, W S. (1992). Paraoxon toxicity is not potentiated by prior reduction in blood acety!cholinesterase. Toxicol. Appl. Pharmacol. 117, 110-115. Percy, M. E., Markovic, V. D., Dalton, A. J., McLachlan, D. R. C., Berg, J. M., Rusk, A. C. M., Somerville, M. J., Chodakowski, B., and Andrews, D. E (1993). Age-associated chromosome 21 loss in Down syndrome: Possible relevance to mosaicism and Alzheimer disease. Am. 1. Med. Genetics 45, 584-588. Perry, E. K., Tomlinson, B. E., Blessed, G., Bergmann, K., Gibson, P. H., and Perry, R. H. (1978). Correlation of cholinergic abnormalities with senile plaques and mental test scores in senile dementia. Br. Med. 1. 2,1457-1459. Petrelli, G., Siepi, G., Miligi, L., and Vineis, P. (1993). Solvents in pesticides. Scand. J. Work Environ. Health 19, 63-65. Ploemen, I., Wormhoudt, L. W., van Ommen, B., Commandeur, J. N. M., Vermeulen, N. P. E., and van Bladeren, P. J. (1995). Polymorphism in the glutathione conjugation activity of human erythrocytes towards ethylene dibromide and 1,2-epoxy-3-(p-nitrophenoxy)-propane. Biochim. Biophys. Acta 1243,469-476. Plueckhahn, V. D. (1973). Infant antiseptic skin care and hexachlorophene. Med. J. Australia 1, 93-100. Pope, C. N., and Padilla, S. (1990). Potentiation of organophosphorous-induced delayed neurotoxicity by phenylmethylsulfonyl fluoride. 1. Toxicol. Environ. Health 31, 261-273. Pope, C. N., Chakraborti, T. K., Chapman, M. L., Farrar, J. D., and Arthun, D. (1991). Comparison of in vivo cholinesterase inhibition in neonatal and adult rats by three organophosphorothioate insecticides. Toxicology 68, 5161. Pramanik, A. K., and Hansen, R. C. (1979). Transcutaneous gamma benzene hexachloride absorption and toxicity in infants and children. Arch. Dermatol. 115, 1224-1225. Rakonczay, Z., and Brimijoin, S. (1988). Biochemistry and pathophysiology of the molecular forms of cholinesterases (Review). Sub-Cellular Biochem. 12,335-378. Rasmussen, J. E. (1979). Percutaneous absorption in children. In "Year Book of Dermatology" (R. L. Dobson, ed.), pp. 15-38. Chicago. Ratner, D., Baruch, 0., and Karola, V. (1983). Chronic dietary anticholinesterase poisoning. Israel J. Med. Sci. 19,810-814. Reed, M. D., and Besunder, J. B. (1989). Developmental pharmacology: Ontogenic basis of drug disposition. Clin. Pharmacol. 36, 1053-1074. Reigart, R. (1988). "Report of the Children's Health Protection Advisory Committee to the V.S. Environmental Protection Agency Regarding the Selection of Five Regulations for Re-Evaluation." Children's Health Protection Advisory Committee. Rice, J. M. (1979). Perinatal period and pregnancy: Intervals of high risk for chemical carcinogens. Environ. Health Perspect. 29, 23-27. Rodier, P. M. (1980). Chronology of neuron development: Animal studies and their clinical implications. Dev. Med. Child Neurol. 22, 525-545. Rogan, W. J. (1980). The exposure and routes of childhood chemical exposures. J. Pediatrics 97,861-865. Rogan, W J., Gladen, B. c., McKinney, J. D., Carreras, N., Hardy, P., Thullen, J., Tingelstad, J., and Tully, M. (1986). Polychlorinated biphenyls (PCBs) and dichlorodiphenyl dichloroethene (DDE) in human milk: Effects of materials factors and previous lactation. Am. J. Public Health 76, 172177. Roskos, K. V., Guy, R. H., and Maibach, H.1. (1986). Percutaneous absorption in the aged. Dermatol. Clin. 4,455-465. Rosso, S., Gonzalez, M., Bagatolli, L., Duffard, R. 0., and Fidelio, G. D. (1998). Evidence of strong interaction of 2,4-dichlorophenoxyacetic acide herbicide with human serum albumin. Life Sci. 63,2343-2351.

797

Rupa, D. S., Reddy, P. P., and Reddi, O. S. (1989). Frequencies of chromosomal aberrations in smokers exposed to pesticides in cotton fields. Mutation Res. 222,37-41. Safe, S. H. (1997). Is there an association between exposure to environmental estrogens and breast cancer? Environ. Health Perspect. 105(Suppl. 3), 675678. Safe, S. H., and Zacharewski, T. (1997). Organochlorine exposure and risk for breast cancer (Review). Prog. Clin. BioI. Res. 396, 133-145. (63 refs.) Savitz, D. A., Whelan, E. A., and Kleckner, R. C. (1989). Self-reported exposure to pesticides and radiation related to pregnancy outcome-results from the National Natality and Fetal Mortality Surveys. Public Health. Rep. 104, 473-477. Saxena, M. C., Siddiqui, M. K. J., Bhargava, A. K., Seth, T. D., Krishnamurti, C. R., and Kutty, D. (1980). Role of chlorinatedhydrocarbon pesticides in abortions and premature labour. Toxicology 17. Scarpato, R., Hirvonen, A., Migliore, L., Falck, G., and Norppa, H. (1997). Influence of GSTM I and GSTTl polymorphisms on the frequency of chromosome aberrations in Iymphocytes of smokers and pesticide-exposed greenhouse workers. Mutation Res. 389,227-235. Scarpato, R., Migliore, L., Hirvonen, A., and Fa!ck, G. N. H. (1996). Cytogenic monitoring of occupational exposure to pesticides: Characterization of GSTMI, GSTTl, and NAT2 genotypes. Environ. Mol. Mutagenesis 27, 263-269. Schwarz, M., Glick, D., Loewenstein, Y., and Soreq, H. (1995). Engineering of human cholinesterases explains and predicts diverse consequences of administration of various drugs and poisons. Pharmac. Ther. 67,283-322. Senthilselvan, A., Mcduffie, H. H., and Dosman, J. A. (1992). Association of asthma with use of pesticides: Results of a cross-sectional survey of farmers. Am. Rev. Respiratory Disease 146, 884-887. Sever, L. E., Arbuckle, T., and Sweeney, A. (1997). Reproductive and developmental effects of occupational pesticide exposure: The epidemiologic evidence. Occupational Med.: State 01 the Art Rev. 12,305-325. Sharpe, C. R., Franco, E. L., Camargo, B. de., Lopes, L. E, Barreto, J. H., Johnsson, R. R., and Mauad, M. A. (1995). Parental exposures to pesticides and risk ofWilms' tumor in Brazil. Am. 1. Epidemiol. 141,210-217. Shaw, K. P., Aracava, Y., Akaike, A., Daly, J. W, Rickett, D. L., and Albuquerque, E. X. (1985). The reversible cholinesterase inhibitor physostigmine has channel-blocking and agonist effects on the acetylcholine receptor-ion channel complex. Mol. Pharmacol. 28,527-538. Sheehan, D. M., and vom Saal, E S. (1997). Low dose effects of endocrine disruptors-A challenge for risk assessment. Risk Policy Rep. (Sept. 19), 31-36. Shih, D. M., Gu, L., Yu-Rong, X., Navab, M., Li, W H. S., Castellani, L. W, Furlong, C. E., Costa, L. G., Fogelman, A. M., and Lusis Aldons, J. (1998). Mice lacking serum paraoxonase are susceptible to organophosphate toxicity and atherosclerosis. Nature 394, 284-287. Shuman, R. M., Leech, R. W, and Alvord, E. C. (1975). Neurotoxicity of hexachlorophene in the human: A clinicopathologic study of 248 children. Pediatrics 54, 689-695. Sim, M., Forbes, A., McNeil, J., and Roberts, G. (1998). Termite control and other determinants of high burdens of cyclodiene insectides. Arch. Environ. Health 53, 114-121. Singh, S. K., and Pandey, R. S. (1991). Ethanol potentiates in vivo hepatotoxicity of endosulfan in adult male rats. Indian J. Exp. Bioi. 29, 1035-1038. Smith, E. M., Hammonds-Ehlers, M., Clark, M. K., and Fuortes, L. (1997). Occupational exposures and risk of female infertility. J. Occupational Environ. Med. 39, 138-147. Snawder, J. E., and Chambers, J. E. (1991). Sex differences in the induction of hepatic microsomal metabolism of parathion by phenobarbital and ft-naphthoflavone. FASEB J. 5, AI564. Solomon, B. A., Haut, S. R., Carr, E. M., and Shalita, A. R. (1995). Neurotoxic reaction to Lindane in an HIV-seropositive patient. J. Family Practice 40, 291-296. Solomon, L. M., and Fahmer, L. (1977). Gamma benzene hexachloride toxicity. Arch. Dermatol. 113,353-357. Swenberg, J. A., and Fedtke, N. (1992). Age-related differences in DNA adduct formation and carcinogenesis of vinyl chloride in rats. In "Similarities and

798

CHAPTER 37


Differences Between Children and Adults: Implications for Risk Assesment" (P. S. Guzelian, C. J. Henry, and S. S. Olin, eds.), pp. 163-171. International Life Sciences Institute, Washington, DC. Takahashi, H., Kashima, T., Nomizo, Y., Muramoto, N., Shimizu, T., Nasu, K, Kubota, T., Kimura, S., and Echizen, H. (1998). Metabolism of warfarin enantiomers in Japanese patients with heart disease having different CYP2C9 and CYP2C19 genotypes. CUn. Pharmacol. Ther. 63,519-528. Tenenbein, M. (1991). Seizures after Lindane therapy. 1. Am. Geriatrics Soc.

39,394-395. Thier, R., Pemble, S. E., Kramer, H., Taylor, J. B., Guengerich, P. F., and Ketterer, B. (1996). Human glutathioe S-transferase TI-I enhances mutagenicity of 1,2-dibromoethane, dibromomethane and 1,2,3,4-diepoxybutane in Salmonella typhimurium. Carcinogenesis 17, 163-166. Thomas, R. D. (1995). Age-specific carcinogenesis: Environmental exposure and susceptibility. Environ. Health Perspect. 103(Suppl. 6), 45-48. Tilson, H. A. (1998). Developmental neurotoxicology of endocrine disruptors and pesticides: Identification of information gaps and research needs. Environ. Health Perspect. 106,807-811. Tomatis, L., Cabral, 1. R. P., Likhachev, A. 1., and Ponomarkrov, V. (1981). Increased cancer incidence in the progeny of male rats exposed to ethylnitrosourea before mating. Int. J. Cancer 28, 475-478. Tomatis, L., Hilfrich, J., and Turusov, V. (1975). The occurrence of tumors in Fl, F2 and F3 descendants of BD rats exposed to N -nitrosomethylurea during pregnancy. Int. J. Cancer 15, 385-390. Tyrala, E. E., Hillman, L. S., Hillman, R. E., and Dodson, W. E. (1977). Clinical pharmacology of hexachlorophene in newborn infants. J. Pediatrics 91, 481-486. Umbenhauer, D. R., Lankas, G. R., Pippert, T. R., Wise, D. L., Cartwright, M. E., Hall, S. J., and Beare, C. M. (1997). Identification of a p-glycoproteindeficient subpopulation in the CF-l mouse strain using a restriction fragment length polymorphism. Toxicol. Appl. Pharmacol. 146, 88-94. Utunomiya, A., Hasegawa, K, and Mori, Y. (1997). Analysis of pyrethroid pesticides, synergists and repellent in moth/mite-proofed household products and their mutagenicity. Jpn. J. Toxicol. Environ. Health 43, 366-375. Valentino, R. J., Lockridge, 0., Eckerson, H. w., and La Du, B. N. (1981). Prediction of drug sensitivity in individuals with atypical serum cholinesterase based on in vitro biochemical studies. Biochem. Pharmacol. 30, 1643-1649. van't Veer, P., Lobbezoo, 1. E., Martfn-Moreno, Guallar, E., G6mez-Aracena, J., Kardinaal, A. F. M., Kohlmeier, L., Martin, B. c., Strain, J. J., Thamm, M., van Zoonen, P., Baumann, B. A., Huttunen, J. K, and Kock, F. J. (1997). DDT (dicophane) and postmenopausal breast cancer in Europe: Casecontrol study. Br. Med. J. 315, 81-85. Vesell, E. S. (1982). Dynamically interacting genetic and environmental factors that affect the reponse of developing individuals to toxicants. In "Banbury Report: Environmental Factors in Human Growth and Development" (V. R. Hunt, M. K. Smith, and D. Worth, eds.). Cold Spring Harbor Laboratory Press, Cold Spring Harbor, NY. Vestberg, K., Galliano, M., Minchiotti, L., and Kragh-Hansen, U. (1992). Highaffinity binding of warfarin, salicylate and diazepam to natural mutants of

human serum albumin modified in the c-terminal end. Biochem. Phamacol. 44, 151-1521. Wagner, S. L. (1995). Pitfalls in the laboratory diagnosis of pesticide intoxication. J. AOC Int. 78, 1-3. Ware, M. R., Frost, M. L., Berger, J. J., Stewart, R. B., and DeVane, L. C. (1990). Electroconvulsive therapy complicated by insecticide ingestion. J. Clin. Psychopharmacol. 10,72-73. Warner, A. (1986). Drug use in the neonate: Interrelationships of pharmacokinetics, toxicity, and biochemical maturity. Clin. Chem. 32, 721-727. Wax, P. M., and Hoffman, R. S. (1994). Fatality associated with inhalation of a pyrethrin shampoo. J. Toxicol.-Clinical Toxicol. 32, 457-460. Wehbe, T. w., and Warth, J. A. (1998). A case of bleeding requiring hospitalization that was likely caused by an interaction between warfarin and levamisole. Clinical Pharmacol. Ther. 59, 360--362. Wester, R. C., and Maibach, H. 1. (1983). Cutaneous pharmacokinetics: 10 steps to percutaneous absorption. Drug Metab. Rev. 14, 169-205. Wester, R. C., Noonan, P. K, Cole, M. P., and Maibach, H. 1. (1977). Percutaneous absorption of testosterone in the newborn rhesus monkey: Comparison to the adult. Pediatrics 11, 737-739. Whorton, D., Krauss, R. M., Marshall, S., and Milby, T. H. (1977). Infertility in male pesticide workers. Lancet (Dec. 17), 1259-1261. Widdowson, E. M., and Dickerson, J. W. T. (1960). The effect of growth and function on the chemical composition of soft tissues. Biochem. J. 77, 30-43. Widdowson, P. S., Farnworth, M. J., and Simpson, M. G. (1996). Influence of age on the passage of paraquat through the blood-brain barrier in rats: A distribution and pathological examination. Hum. Exp. Toxicol. 15,231236. Wolff, M. S., and Weston, A. (1997). Breast cancer risk and environmental exposures. Environ. Health Perspect. 105(Suppl. 4), 891-896. Wolff, M. S., Toniolo, P. G., Lee, E. w., Rivera, M., and Dubin, N. (1993). Blood levels of organochlorine residues and risk of breast cancer. J. Natl. Cancer Inst. 85, 648-652. Xu, M., Molento, M. B. w., Ribeiro, P., and Beech, R. P. R. (1998). Ivermectin resistance in nematodes may be caused by alteration of P-glycoprotein homolog. Mol. Biochem. Parasitol. 91, 327-335. Yamazaki, H. N. K C. K, Ozawa, N., Kawai, T., Suzuki, Y., Goldstein, J. A., Guengerich, P., and Shimada, T. (1998). Comparative studies on the catalytic roles of cytochrome P450 2C9 and its Cys- and Leu-variants in the oxidation of warfarin, flurbiprofen, and diclofenac by human liver microsomes. Biochem. Pharmacol. 56,243-251. Yu, C. M., Juliana, C. N., and Sanderson, J. E. (1997). Chinese herbs and warfarin potentiation by 'Danshen.' J. Internal Med. 241, 337-339. Zhou, Z. H., and Syvanen, M. (1997). A complex glutathione transferase gene family in the housefly Musca demestica. Mol. General Genetics 256, 187194. Ziem, G., and McTamney, J. (1997). Profile of patients with chemical injury and sensitivity. Environ. Health Perspect. 105(Suppl. 2),417-435.

CHAPTER

38 Pesticide Residues in Food and Cancer Risk: A Critical Analysis Lois Swirsky Gold, Thomas H. Slone, Bruce N. Ames, and Neela B. Manley University of California, Berkeley Ernest Orlando Lawrence Berkeley National Laboratory

38.1 INTRODUCTION Possible cancer hazards from pesticide residues in food have been much discussed and hotly debated in the scientific literature, the popular press, the political arena, and the courts. Consumer opinion surveys indicate that much of the U.S. public believes that pesticide residues in food are a serious cancer hazard (Opinion Research Corporation, 1990). In contrast, epidemiologic studies indicate that the major preventable risk factors for cancer are smoking, dietary imbalances, endogenous hormones, and inflammation (e.g., from chronic infections). Other important factors include intense sun exposure, lack of physical activity, and excess alcohol consumption (Ames et aI., 1995). The types of cancer deaths that have decreased since 1950 are primarily stomach, cervical, uterine, and colorectal. Overall cancer death rates in the United States (excluding lung cancer) have declined 19% since 1950 (Ries et aI., 2000). The types that have increased are primarily lung cancer [87% is due to smoking, as are 31 % of all cancer deaths in the United States (American Cancer Society, 2000)], melanoma (probably due to sunburns), and non-Hodgkin's lymphoma. If lung cancer is included, mortality rates have increased over time, but recently have declined (Ries et aI., 2000). Thus, epidemiological studies do not support the idea that synthetic pesticide residues are important for human cancer. Although some epidemiologic studies find an association between cancer and low levels of some industrial pollutants, the studies often have weak or inconsistent results, rely on ecological correlations or indirect exposure assessments, use small sample sizes, and do not control for confounding factors such as composition of the diet, which is a potentially important confounding factor. Outside the workplace, the levels of exposure to synthetic pollutants or pesticide residues are low and rarely Handbook of Pesticide Toxicology Volume 1. Principles

seem toxicologically plausible as a causal factor when compared to the wide variety of naturally occurring chemicals to which all people are exposed (Ames et aI., 1987, 1990a; Gold et aI., 1992). Whereas public perceptions tend to identify chemicals as being only synthetic and only synthetic chemicals as being toxic, every natural chemical is also toxic at some dose, and the vast proportion of chemicals to which humans are exposed are naturally occurring (see Section 38.2). There is, however, a paradox in the public concern about possible cancer hazards from pesticide residues in food and the lack of public understanding of the substantial evidence indicating that high consumption of the foods that contain pesticide residues-fruits and vegetables-has a protective effect against many types of cancer. A review of about 200 epidemiological studies reported a consistent association between low consumption of fruits and vegetables and cancer incidence at many target sites (Block et aI., 1992; Hill et aI., 1994; Steinmetz and Potter, 1991). The quarter of the population with the lowest dietary intake of fruits and vegetables has roughly twice the cancer rate for many types of cancer (lung, larynx, oral cavity, esophagus, stomach, colon and rectum, bladder, pancreas, cervix, and ovary) compared to the quarter with the highest consumption of those foods. The protective effect of consuming fruits and vegetables is weaker and less consistent for hormonally related cancers, such as breast and prostate. Studies suggest that inadequate intake of many micronutrients in these foods may be radiation mimics and are important in the carcinogenic effect (Ames, 2001). Despite the substantial evidence of the importance of fruits and vegetables in prevention, half the American public did not identify fruit and vegetable consumption as a protective factor against cancer (U.S. National Cancer Institute, 1996). Consumption surveys, moreover, indicate that 80% of children and adolescents in the United States (Krebs-Smith et

799

Copyright © 2001 by Academic Press. All rights of reproduction in any fonn reserved.

800

CHAPTER 38

Pesticide Residues in Food and Cancer Risk: A Critical Analysis

aI., 1996) and 68% of adults (Krebs-Smith et aI., 1995) did not consume the intake of fruits and vegetables recommended by the National Cancer Institute (NCI) and the National Research Council: five servings per day. One important consequence of inadequate consumption of fruits and vegetables is low intake of some micronutrients. For example, folic acid is one of the most common vitamin deficiencies in people who consume few dietary fruits and vegetables; folate deficiency causes chromosome breaks in humans by a mechanism that mimics radiation (Ames, 2001; BIount et aI., 1997). Approximately 10% of the V.S. population (Senti and Pilch, 1985) had a lower folate level than that at which chromosome breaks occur (BIount et aI., 1997). Folate supplementation above the recommended daily allowance (RDA) minimized chromosome breakage (Fenech et al., 1998).

Given the lack of epidemiological evidence to link dietary synthetic pesticide residues to human cancer, and taking into account public concerns about pesticide residues as possible cancer hazards, public policy with respect to pesticides has relied on the results of high-dose, rodent cancer tests as the major source of information for assessing potential cancer risks to humans. This chapter examines critically the assumptions, methodology, results, and implications of cancer risk assessments of pesticide residues in the diet. Our analyses are based on results in our Carcinogenic Potency Database (CPDB) (Gold et aI., 1997b, 1999; http://potency.berkeley.edu), which provide the necessary data to examine the published literature of chronic animal cancer tests; the CPDB includes results of 5620 experiments on 1372 chemicals. Specifically, the following are addressed in the section indicated: Section 38.2. Human exposure to synthetic pesticide residues it the diet compared to the broader and greater exposure to natural chemicals in the diet Section 38.3. Cancer risk assessment methodology, including the use of animal data from high-dose bioassays in which half the chemicals tested are carcinogenic Section 38.4. Increased cell division as an important hypothesis for the high positivity rate in rodent bioassays and implications for risk assessment Section 38.5. Providing a broad perspective on possible cancer hazards from a variety of exposures to rodent carcinogens, including pesticide residues, by ranking on the HERP (human exposure/rodent potency) index Section 38.6. Analysis of possible reasons for the wide disparities in published risk estimates for pesticide residues in the diet Section 38.7. Identification and ranking of exposures in the V.S. diet to naturally occurring chemicals that have not been tested for carcinogenicity, using an index that takes into account the acutely toxic dose of a chemical (LDso) and average consumption in the V.S. diet Section 38.8. Summary of carcinogenicity results on 193 active ingredients in commercial pesticides.

38.2 HUMAN EXPOSURES TO NATURAL AND SYNTHETIC CHEMICALS Current regulatory policy to reduce human cancer risks is based on the idea that chemicals that induce tumors in rodent cancer bioassays are potential human carcinogens. The chemicals selected for testing in rodents, however, are primarily synthetic (Gold et aI., 1997a, b, c, 1998, 1999). The enormous background of human exposures to natural chemicals has not been systematically examined. This has led to an imbalance in both data and perception about possible carcinogenic hazards to humans from chemical exposures. The regulatory process does not take into account (1) that natural chemicals make up the vast bulk of chemicals to which humans are exposed; (2) that the toxicology of synthetic and natural toxins is not fundamentally different; (3) that about half of the chemicals tested, whether natural or synthetic, are carcinogens when tested using current experimental protocols; (4) that testing for carcinogenicity at near-toxic doses in rodents does not provide enough information to predict the excess number of human cancers that might occur at low-dose exposures; and (5) that testing at the maximum tolerated dose (MTD) frequently can cause chronic cell killing and consequent cell replacement (a risk factor for cancer that can be limited to high doses) and that ignoring this effect in risk assessment can greatly exaggerate risks. We estimate that about 99.9% of the chemicals that humans ingest are naturally occurring. The amounts of synthetic pesticide residues in plant foods are low in comparison to the amount of natural pesticides produced by plants themselves (Ames et aI., 1990a, b; Gold et aI., 1997a). Of all dietary pesticides that Americans eat, 99.99% are natural: They are the chemicals produced by plants to defend themselves against fungi, insects, and other animal predators. Each plant produces a different array of such chemicals (Ames et aI., 1990a, b). We estimate that the daily average V.S. exposure to natural pesticides in the diet is about 1500 mg and to burnt material from cooking is about 2000 mg (Ames et aI., 1990b). In comparison, the total daily exposure to all synthetic pesticide residues combined is about 0.09 mg based on the sum of residues reported by the V.S. Food and Drug Administration (FDA) in its study of the 200 synthetic pesticide residues thought to be of greatest concern (Gunderson, 1988; V.S. Food and Drug Administration, 1993a). Humans ingest roughly 5000-10,000 different natural pesticides and their breakdown products (Ames et aI., 1990a). Despite this enormously greater exposure to natural chemicals, among the chemicals tested in long-term bioassays in the CPDB, 77% (1050/1372) are synthetic (i.e., do not occur naturally) (Gold and Zeiger, 1997; Gold et aI., 1999). Concentrations of natural pesticides in plants are usually found at parts per thousand or million rather than parts per billion, which is the usual concentration of synthetic pesticide residues. Therefore, because humans are exposed to so many more natural than synthetic chemicals (by weight and by number), human exposure to natural rodent carcinogens, as defined

38.2 Human Exposures to Natural and Synthetic Chemicals

801

Table 38.1 Carcinogenicity Status of Natural Pesticides Tested in Rodentsa Carcinogensb : N =37

Acetaldehyde methylformylhydrazone, allyl isothiocyanate, arecoline·HCI, benzaldehyde, benzyl acetate, caffeic acid, capsaicin, catechol, clivorine, coumarin, crotonaldehyde, 3,4-dihydrocoumarin, estragole, ethyl acrylate, N2-y-glutamyl-p-hydrazinobenzoic acid, hexanal methylformylhydrazine, p-hydrazinobenzoic acid·HCI, hydroquinone, I-hydroxyanthraquinone, lasiocarpine, d-limonene, 3-methoxycatechol, 8-methoxypsoralen, N -methyl-N -formylhydrazine, a-methylbenzyl alcohol, 3-methylbutanal methylformylhydrazone, 4-methyicatechol, methylhydrazine, monocrotaline, pentanal methylformylhydrazone, petasitenine, quercetin, reserpine, safrole, senkirkine, sesamol, symphytine

Noncarcinogens: N =34

Atropine, benzyl aicohol, benzyl isothiocyanate, benzyl thiocyanate, biphenyl, d-carvone, codeine, deserpidine, disodium glycyrrhizinate, ephedrine sulfate, epigallocatechin, eucalyptol, eugenol, gaIlic acid, geranyl acetate, J'i-N-[y-I(+)-glutamyI1-4hydroxymethylphenylhydrazine, glycyrrhetinic acid, p-hydrazinobenzoic acid, isosafrole, kaempferol, dl-menthol, nicotine, norharman, phenethyl isothiocyanate, pilocarpine, piperidine, protocatechuic acid, rotenone, rutin sulfate, sodium benzoate, tannic acid, I-trans-8 9-tetrahydrocannabinol, turmeric oleoresin, vinblastine

aPungal toxins are not included. bThese rodent carcinogens occur in absinthe, allspice, anise, apple, apricot, banana, basil, beet, black pepper, broccoli, Brussels sprouts, cabbage, cantaloupe, caraway, cardamom, carrot, cauliflower, celery, cherries, chili pepper, chocolate, cinnamon, cloves, coffee, coliard greens, comfrey herb tea, coriander, corn, currants, dill, eggplant, endive, fennel, garlic, grapefruit, grapes, guava, honey, honeydew melon, horseradish, kale, lemon, lentils, lettuce, licorice, lime, mace, mango, marjoram, mint, mushrooms, mustard, nutmeg, onion, orange, paprika, parsley, parsnip, peach, pear, peas, pineapple, plum, potato, radish, raspberries, rhubarb, rosemary, rutabaga, sage, savory, sesame seeds, soybean, star anise, tarragon, tea, thyme, tomato, turmeric, and turnip.

by high-dose rodent tests, is ubiquitous (Ames et al., 1990b). It is probable that almost every fruit and vegetable in the supermarket contains natural pesticides that are rodent carcinogens. Even though only a tiny proportion of natural pesticides have been tested for carcinogenicity, 37 of 71 that have been tested are rodent carcinogens that are present in the common foods listed in Table 38.1. Humans also ingest numerous natural chemicals that are produced as by-products of cooking food. For example, more than 1000 chemicals have been identified in roasted coffee, many of which are produced by roasting (Clarke and Macrae, 1988; Nijssen et al., 1996). Only 30 have been tested for carcinogenicity according to the most recent results in our CPDB, and 21 of these are positive in at least one test (Table 38.2), totaling at least 10 mg of rodent carcinogens per cup of coffee (Clarke and Macrae, 1988; Fujita et aI., 1985; Kikugawa et aI., 1989; Nijssen et aI., 1996). Among the rodent carcinogens in coffee are the plant pesticides caffeic acid (present at 1800 ppm; C1arke and Macrae, 1988) and catechol (present at 100 ppm; Rahn and K6nig, 1978; Tressl et aI., 1978). Two other plant pesticides in coffee, chlorogenic acid and neochlorogenic acid (present at 21,600 and 11,600 ppm, respectively; Clarke and Macrae,

1988) are metabolized to caffeic acid and catechol but have not been tested for carcinogenicity. Chlorogenic acid and caffeic acid are mutagenic (Ariza et aI., 1988; Fung et aI., 1988; Hanham et aI., 1983) and clastogenic (Ishidate et aI., 1988; Stich et aI., 1981). Another plant pesticide in coffee, d-limonene, is carcinogenic but the only tumors induced were in male rat kidney, by a mechanism involving accumulation of c¥2u-globulin and increased cell division in the kidney, which would not be predictive of a carcinogenic hazard to humans (Dietrich and Swenberg, 1991; Rice et al., 1999). Some other rodent carcinogens in coffee are products of cooking, for example, furfural and benzo(a)pyrene. The point here is not to indicate that rodent data necessarily implicate coffee as a risk factor for human cancer, but rather to illustrate that there is an enormous background of chemicals in the diet that are natural and that have not been a focus of carcinogenicity testing. A diet free of naturally occurring chemicals that are carcinogens in high-dose rodent tests is impossible. It is often assumed that because natural chemicals are part of human evolutionary history, whereas synthetic chemicals are recent, the mechanisms that have evolved in animals to cope with the toxicity of natural chemicals will fail to protect against

Table 38.2 Carcinogenicity Status of Natural Chemicals in Roasted Coffee Positive: N = 21

Acetaldehyde, benzaldehyde, benzene, benzofuran, benzo(a)pyrene, caffeic acid, catechol, 1,2,5,6-dibenzanthracene, ethanol, ethylbenzene, formaldehyde, furan, furfural, hydrogen peroxide, hydroquinone, isoprene, limonene, 4-methyicatechol, styrene, toluene, xylene

Not positive:

Acrolein, biphenyl, chOline, eugenol, nicotinamide, nicotinic acid, phenol, piperidine

N=8 Uncertain:

Caffeine

Yet to test:

~ 1000

chemicals

802

CHAPTER 38


synthetic chemicals, including synthetic pesticides (Ames et aI., 1987). This assumption is flawed for several reasons (Ames et aI., 1990b, 1996; Gold et aI., 1997a, b, c):

1. Humans have many natural defenses that buffer against normal exposures to toxins (Ames et aI., 1990b) and these are usually general, rather than tailored for each specific chemical. Thus, they work against both natural and synthetic chemicals. Examples of general defenses include the continuous shedding of cells exposed to toxins-the surface layers of the mouth, esophagus, stomach, intestine, colon, skin, and lungs are discarded every few days; deoxyribonucleic acid (DNA) repair enzymes, which repair DNA that was damaged from many different sources; and detoxification enzymes of the liver and other organs, which generally target classes of chemicals rather than individual chemicals. That human defenses are usually general, rather than specific for each chemical, makes good evolutionary sense. The reason that predators of plants evolved general defenses is presumably to be prepared to counter a diverse and ever-changing array of plant toxins in an evolving world; if a herbivore had defenses against only a specific set of toxins, it would be at great disadvantage in obtaining new food when favored foods became scarce or evolved new chemical defenses. 2. Various natural toxins, which have been present throughout vertebrate evolutionary history, nevertheless cause cancer in vertebrates (Ames et aI., 1990b; Gold et aI., 1997b, 1999; Vainio et aI., 1995). Mold toxins, such as aflatoxin, have been shown to cause cancer in rodents, monkeys, humans, and other species. Many of the common elements, despite their presence throughout evolution, are carcinogenic to humans at high doses (e.g., the salts of cadmium, beryllium, nickel, chromium, and arsenic). Furthermore, epidemiological studies from various parts of the world indicate that certain natural chemicals in food may be carcinogenic risks to humans; for example, the chewing of betel nut with tobacco is associated with oral cancer. Among the agents identified as human carcinogens by the International Agency for Research in Cancer (IARC) 62% (37/60) occur naturally: 16 are natural chemicals, 11 are mixtures of natural chemicals, and 10 are infectious agents (IARC, 1971-1999; Vainio et aI., 1995). Thus, the idea that a chemical is "safe" because it is natural, is not correct. 3. Humans have not had time to evolve a "toxic harmony" with all of their dietary plants. The human diet has changed markedly in the last few thousand years. Indeed, very few of the plants that humans eat today (e.g., coffee, cocoa, tea, potatoes, tomatoes, corn, avocados, mangos, olives and kiwi fruit) would have been present in a hunter-gatherer's diet. Natural selection works far too slowly for humans to have evolved specific resistance to the food toxins in these newly introduced plants. 4. Some early synthetic pesticides were lipophilic organochlorines that persist in nature and bioaccumulate in adipose tissue, for example, dichlorophenyltrichloroethane (DDT), aldrin, and dieldrin (DDT is discussed in

Section 38.5). This ability to bioaccumulate is often seen as a hazardous property of synthetic pesticides; however, such bioconcentration and persistence are properties of relatively few synthetic pesticides. Moreover, many thousands of chlorinated chemicals are produced in nature (Gribble, 1996). Natural pesticides also can bioconcentrate if they are fat soluble. Potatoes, for example, were introduced into the worldwide food supply a few hundred years ago; potatoes contain solanine and chaconine, which are fat-soluble, neurotoxic, natural pesticides that can be detected in the blood of all potato-eaters. High levels of these potato glycoalkaloids have been shown to cause reproductive abnormalities in rodents (Ames et aI., 1990b; Morris and Lee, 1984). 5. Because no plot of land is free from attack by insects, plants need chemical defenses-either natural or synthetic-to survive pest attack. Thus, there is a trade-off between naturally-occurring pesticides and synthetic pesticides. One consequence of efforts to reduce pesticide use is that some plant breeders develop plants to be more insect resistant by making them higher in natural pesticides. A recent case illustrates the potential hazards of this approach to pest control: When a major grower introduced a new variety of highly insect-resistant celery into commerce, people who handled the celery developed rashes when they were subsequently exposed to sunlight. Some detective work found that the pest-resistant celery contained 6200 parts per billion (ppb) of carcinogenic (and mutagenic) psoralens instead ofthe 800 ppb present in common celery (Beier and Nigg, 1994; Berkley et aI., 1986; Seligman et aI., 1987).

38.3 THE HIGH CARCINOGENICITY RATE AMONG CHEMICALS TESTED IN CHRONIC ANIMAL CANCER TESTS Because the toxicology of natural and synthetic chemicals is similar, one expects, and finds, a similar positivity rate for carcinogenicity among synthetic and natural chemicals that have been tested in rodent bioassays. Among chemicals tested in rats and mice in the CPDB, about half the natural chemicals are positive, and about half of all chemicals tested are positive. This high positivity rate in rodent carcinogenesis bioassays is consistent for many data sets (Table 38.3): Among chemicals tested in rats and mice, 59% (350/590) are positive in at least one experiment, 60% of synthetic chemicals (2711451), and 57% of naturally occurring chemicals (79/139). Among chemicals tested in at least one species, 52% of natural pesticides (37171) are positive, 61 % of fungal toxins (14123), and 70% of the naturally occurring chemicals in roasted coffee (21130) (Table 38.2). Among commercial pesticides reviewed by the EPA (U.S. Environmental Protection Agency, 1998), the positivity rate is 41 % (79/193); this rate is similar among commercial pesticides that also occur naturally and those that are only synthetic, as well as between commercial pesticides that have been canceled and those still in use. (See Section 38.8 for detailed summary results

38.3 The High Carcinogenicity Rate Among Chemicals Tested in Chronic Animal Cancer Tests Table 38.3 Proportion of Chemicals Evaluated as Carcinogenic Chemicals tested in both rats and micea Chemicals in the CPDB Naturally occurring chemicals in the CPDB Synthetic chemicals in the CPDB

350/590 (59%)

79/139 (57%) 2711451 (60%)

Chemicals tested in rats and/or micea Chemicals in the CPDB Natural pesticides in the CPDB

70211348 (52%) 37171 (52%)

Mold toxins in the CPDB

14123 (61%)

Chemicals in roasted coffee in the CPDB

2l/30 (70%)

Commercial pesticides in the CPDB

79/193 (41 %)

Physicians' Desk Reference (PDR): Drugs with reported cancer tests b FDA database of drug submissionsc

117/241 (49%) 125/282 (44%)

aFrom the Carcinogenic Potency Database (Gold et aI., 1997c, 1999). bDavies and Monro (1995). CContrera et al. (1997). 140 drugs are in both the FDA and the PDR databases.

of carcinogenicity tests on the 193 commercial pesticides in the CPDB, including results on the positivity of each chemical, its carcinogenic potency, and target organs of carcinogenesis.) Because the results of high-dose rodent tests are routinely used to identify a chemical as a possible cancer hazard to humans, it is important to try to understand how representative the 50% positivity rate might be of all untested chemicals. If half of all chemicals (both natural and synthetic) to which humans are exposed were positive if tested, then the utility of a test to identify a chemical as a "potential human carcinogen" because it increases tumor incidence in a rodent bioassay would be questionable. To determine the true proportion of rodent carcinogens among chemicals would require a comparison of a random group of synthetic chemicals to a random group of natural chemicals. Such an analysis has not been done. It has been argued that the high positivity rate is due to selecting more suspicious chemicals to test for carcinogenicity. For example, chemicals may be selected that are structurally similar to known carcinogens or genotoxins. That is a likely bias because cancer testing is both expensive and time consuming, making it prudent to test suspicious compounds. On the other hand, chemicals are selected for testing for many reasons, including the extent of human exposure, level of production, and scientific questions about carcinogenesis. Among chemicals tested in both rats and mice, chemicals that are mutagenic in Salmonella are carcinogenic in rodent bioassays more frequently than nonmutagens: 80% of mutagens are positive (1761219) compared to 50% (1351271) of nonmutagens. Thus, if testing is based on suspicion of carcinogenicity, then more mutagens should be selected than nonmutagens; however, of the chemicals tested in both species, 55% (271/490) are not mutagenic. This suggests that prediction of positivity is often not the basis for selecting a chemical to test. Another argument against selection bias is the high positivity rate for drugs (Ta-

803

ble 38.3), because drug development tends to favor chemicals that are not mutagens or suspected carcinogens. In the Physicians' Desk Reference (PDR), however, 49% (1171241) of the drugs that report results of animal cancer tests are carcinogenic (Davies and Monro, 1995) (Table 38.3). Moreover, while some chemical classes are more often carcinogenic in rodent bioassays than others (e.g., nitroso compounds, aromatic amines, nitroaromatics, and chlorinated compounds), prediction is still imperfect. For example, a prospective prediction exercise was conducted by several experts in 1990 in advance of the 2-year National Toxicology Program bioassays. There was wide disagreement among the experts on which chemicals would be carcinogenic when tested, and the level of accuracy varied by expert, thus indicating that predictive knowledge is uncertain (Omenn et aI., 1995). One large series of mouse experiments by Innes et al. (1969) has frequently been cited (U.S. National Cancer Institute, 1984) as evidence that the true proportion of rodent carcinogens is actually low among tested substances (Table 38.4). In the Innes study, 119 synthetic pesticides and industrial chemicals were tested, and only 11 (9%) were evaluated as carcinogenic. Our analysis indicates that those early experiments lacked power to detect an effect because they were conducted only in mice (not in rats), they included only 18 animals in a group (compared with the standard protocol of 50), the animals were tested for only 18 months (compared with the standard 24 months), and the Innes dose was usually lower than the highest dose in subsequent mouse tests if the same chemical was tested again (Gold and Zeiger, 1997; Gold et aI., 1999; Innes et al., 1969). To assess whether the low positivity rate in the Innes study was due to the lack of power in the design of the experiments, we used results in our CPDB to examine subsequent bioassays on the Innes chemicals that had not been evaluated as positive (results and chemical names are reported in Table 38.4). Among the 34 chemicals that were not positive in the Innes study and were subsequently retested with more standard protocols, 17 had a subsequent positive evaluation of carcinogenicity (50%), which is similar to the proportion among all chemicals in the CPDB (Table 38.4). Of the 17 new positives, 7 were carcinogenic in mice and 14 in rats. Innes et al. had recommended further evaluation of some chemicals that had inconclusive results in their study. If those were the chemicals subsequently retested, then one might argue that they would be the most likely to be positive. Our analysis does not support that view, however. We found that the positivity rate among the chemicals that the Innes study said needed further evaluation was 7 of 16 (44%) when retested, compared to 10 of 18 (56%) among the chemicals that Innes evaluated as negative. Our analysis thus supports the idea that the low positivity rate in the Innes study resulted from lack of power. Because many of the chemicals tested by Innes et al. were synthetic pesticides, we reexamined the question of what proportion of synthetic pesticides are carcinogenic (as shown in Table 38.3) by excluding the pesticides tested only in the Innes series. The Innes studies had little effect on the positivity rate: Table 38.3 indicates that of all commercial pesticides in the

804

CHAPTER 38


Table 38.4 Results of Subsequent Tests on Chemicals (Primarily Pesticides) not Found Carcinogenic by rnnes et al. (1969) Percentage carcinogenic when retested Retested chemicals All retested

Mice

Rats

Either mice or rats

7/26 (27%)

14/34 (41 %)

17134 (50%)

rnnes: not carcinogenic

3/10 (30%)

9/18 (50%)

10/18 (56%)

rnnes: needs further evaluation

4/16 (25%)

5/16 (31 %)

7116 (44%)

Of 119 chemicals tested by rnnes et al., 11 (9%) were evaluated as positive by rnnes et al. Carcinogenic when retested: atrazine (R), azobenzene* (R), captan (M, R), carbaryl (R), 3-(p-chlorophenyl)-I,I-dimethylurea* (R), p,p'-DDD* (M), folpet (M), manganese ethylenebisthiocarbamate (R), 2-mercaptobenzothiazole (R), N -nitrosodiphenylamine* (R), 2,3,4,5,6-pentachlorophenol (M, R), o-phenylphenol (R), piperonyl butoxide* (M, R), piperonyl sulfoxide* (M),2,4,6-trichlorophenol* (M, R), zinc dimethyldithiocarbamate (R), zinc ethylenebisthiocarbamate (R). Not carcinogenic when retested: (2-chloroethyl)trimethylammonium chloride*, calcium cyanamide*, diphenyl-p-phenylenediamine, endosulfan, p, p'ethyl-DDD*, ethyl tellurac*, isopropyl-N -(3-chlorophenyl) carbamate, lead dimethyldithiocarbamate*, maleic hydrazide, mexacarbate*, monochloroacetic acid, phenyl-,B-naphthylamine*, rotenone, sodium diethyldithiocarbamate trihydrate*, tetraethylthiuram disulfide*, tetramethy Ithiuram disulfide, 2,4,5trichlorophenoxyacetic acid. (M), positive in mice when retested; (R), positive in rats when retested; *, rnnes et al. stated that further testing was needed.

CPDB, 41 % 791193 are rodent carcinogens; when the analysis is repeated by excluding those Innes tests, 47% (77/165) are carcinogens.

38.4 THE IMPORTANCE OF CELL DIVISION IN MUTAGENESIS AND CARCINOGENESIS What might explain the high proportion of chemicals that are carcinogenic when tested in rodent cancer bioassays (Table 38.3)? In standard cancer tests, rodents are given a chronic, near-toxic dose: the maximum tolerated dose (MTD). Evidence is accumulating that cell division caused by the high dose itself, rather than the chemical per se, contributes to cancer in such tests (Ames and Gold, 1990; Ames et aI., 1993a; Butterworth and Bogdanffy, 1999; Cohen, 1998; Cunningham, 1996; Cunningham and Matthews, 1991; Cunningham et aI., 1991; Heddle, 1998). High doses can cause chronic wounding of tissues, cell death, and consequent chronic cell division of neighboring cells, which is a risk factor for cancer (Ames and Gold, 1990; Gold et aI., 1998). Each time a cell divides, there is some probability that a mutation will occur, and thus increased cell division increases the risk of cancer. At the low levels of pesticide residues to which humans are usually exposed, such increased cell division does not occur. The process of mutagenesis and carcinogenesis is complicated because many factors are involved, for example, DNA lesions, DNA repair, cell division, clonal instability, apoptosis, and p53 (a cell cycle gene that is mutated in half of human tumors) (Christensen et aI., 1999; Hill et aI., 1999). The normal endogenous level of oxidative DNA lesions in somatic cells is appreciable (Helbock et aI., 1998). In addition, tissues injured by high doses of chemicals have an inflammatory immune response involving activation of white cells in response to cell death (Adachi et aI., 1995; Czaja et aI., 1994; Gunawardhana et aI., 1993; Laskin and Pendino, 1995; Roberts and Kimber, 1999). Activated white cells release mutagenic oxidants (including peroxynitrite, hypochlorite, and

H202). Therefore, the very low levels of synthetic pesticide residues to which humans are exposed may pose no or only minimal cancer risks. It seems likely that a high proportion of all chemicals, whether synthetic or natural, might be "carcinogens" if administered in the standard rodent bioassay at the MTD, primarily due to the effects of high doses on cell division and DNA damage (Ames and Gold, 1990; Ames et aI., 1993a; Butterworth et aI., 1995; Cunning ham, 1996; Cunningham and Matthews, 1991; Cunningham et aI., 1991). For nonmutagens, cell division at the MTD can increase carcinogenicity; for mutagens, there can be a synergistic effect between DNA damage and cell division at high doses. Ad libitum feeding in the standard bioassay can also contribute to the high positivity rate (Hart et aI., 1995). In calorie-restricted mice, cell division rates are markedly lower in several tissues than in ad libitum-fed mice (Lok et aI., 1990). In dosed animals, food restriction decreased tumor incidence at all three sites that were evaluated as target sites (pancreas and bladder in male rats, liver in male mice), and none of those sites was evaluated as target sites after 2 or 3 years (U.S. National Toxicology Program, 1997). In standard National Cancer Institute (NCI)INational Toxicology Program (NTP) bioassays, for both control and dosed animals, food restriction improves survival and at the same time decreases tumor incidence at many sites compared to ad libitum-feeding. Without additional data on how a chemical causes cancer, the interpretation of a positive result in a rodent bioassay is highly uncertain. Although cell division is not measured in routine cancer tests, many studies on rodent carcinogenicity show a correlation between cell division at the MTD and cancer (Cunningham et aI., 1995; Gold et aI., 1998; Hayward et aI., 1995). Extensive reviews of bioassay results document that chronic cell division can induce cancer (Ames and Gold, 1990; Ames et aI., 1993b; Cohen, 1995b; Cohen and Ellwein, 1991; Cohen and Lawson, 1995; Counts and Goodman, 1995; Gold et aI., 1997b). A large epidemiological literature reviewed by PrestonMartin et al. (1990, 1995) indicates that increased cell division by hormones and other agents can increase human cancer.

38.4 The Importance of Cell Division in Mutagenesis and Carcinogenesis

Several of our findings in large-scale analyses of the results of animal cancer tests (Gold et aI., 1993) are consistent with the idea that cell division increases the carcinogenic effect in high-dose bioassays, including the high proportion of chemicals that are positive; the high proportion of rodent carcinogens that are not mutagenic; and the fact that mutagens, which can both damage DNA and increase cell division at high doses, are more likely than nonmutagens to be positive, to induce tumors in both rats and mice, and to induce tumors at multiple sites (Gold et aI., 1993, 1998). Analyses of the limited data on dose response in bioassays are consistent with the idea that cell division from cell killing and cell replacement is important. Among rodent bioassays with two doses and a control group, about half the sites evaluated as target sites are statistically significant at the MTD but not at half the MTD (p < 0.05). The proportions are similar for mutagens (44%, 148/334) and nonmutagens (47%, 76/163) (Gold and Zeiger, 1997; Gold et aI., 1999), suggesting that cell division at the MTD may be important for the carcinogenic response of mutagens as well as nonmutagens that are rodent carcinogens. To the extent that increases in tumor incidence in rodent studies are due to the secondary effects of inducing cell division at the MTD, then any chemical is a likely rodent carcinogen, and carcinogenic effects can be limited to high doses. Linearity of the dose-response relationship also seems less likely than has been assumed because of the inducibility of numerous defense enzymes that deal with exogenous chemicals as groups (e.g., oxidants, electrophiles) and thus protect humans against natural and synthetic chemicals, including potentially mutagenic reactive chemicals (Ames et aI., 1990b; Luckey, 1999; Munday and Munday, 1999; Trosko, 1998). Thus, true risks at the low doses of most exposures to the general population are likely to be much lower than what would be predicted by the linear model that has been the default in U.S. regulatory risk assessment. The true risk might often be O. Agencies that evaluate potential cancer risks to humans are moving to take mechanism and nonlinearity into account. The U.S. Environmental Protection Agency (EPA) recently proposed new cancer risk assessment guidelines (U.S. Environmental Protection Agency, 1996a) that emphasize a more flexible approach to risk assessment and call for the use of more biological information in the weight-of-evidence evaluation of carcinogenicity for a given chemical and in the dose-response assessment. The proposed changes take into account the issues that were discussed previously. The new EPA guidelines recognize the dose dependence of many toxicokinetic and metabolic processes and the importance of understanding cancer mechanisms for a chemical. The guidelines use nonlinear approaches to low-dose extrapolation if warranted by mechanistic data and a possible threshold of dose below which effects will not occur (National Research Council, 1994; U.S. Environmental Protection Agency, 1996a). In addition, toxicological results for cancer and noncancer endpoints could be incorporated together in the risk assessment process. Also consistent with the results discussed previously, are the recent IARC consensus criteria for evaluations of carcino-

805

genicity in rodent studies, which take into account that an agent can cause cancer in laboratory animals through a mechanism that does not operate in humans (Rice et aI., 1999). The tumors in such cases involve persistent hyperplasia in cell types from which the tumors arise. These include urinary bladder carcinomas associated with certain urinary precipitates, thyroid follicular-cell tumors associated with altered thyroidstimulating hormone (TSH), and cortical tumors of the kidney that arise only in male rats in association with nephropathy that is due to CV2u urinary globulin. Historically, in U.S. regulatory policy, the "virtually safe dose," corresponding to a maximum, hypothetical risk of one cancer in a million, has routinely been estimated from results of carcinogenesis bioassays using a linear model, which assumes that there are no unique effects of high doses. To the extent that carcinogenicity in rodent bioassays is due to the effects of high doses for the nonmutagens, and a synergistic effect of cell division at high doses with DNA damage for the mutagens, this model overestimates risk (Butterworth and Bogdanffy, 1999; Gaylor and Gold, 1998). We have discussed validity problems associated with the use of the limited data from animal cancer tests for human risk assessment (Bemstein et aI., 1985; Gold et aI., 1998). Standard practice in regulatory risk assessment for a given rodent carcinogen has been to extrapolate from the high doses of rodent bioassays to the low doses of most human exposures by multiplying carcinogenic potency in rodents by human exposure. Strikingly, however, due to the relatively narrow range of doses in 2-year rodent bioassays and the limited range of statistically significant tumor incidence rates, the various measures of potency obtained from 2-year bioassays, such as the EPA value, the TDso, and the lower confidence limit on the TDlO (LTDlO), are constrained to a relatively narrow range of values about the MTD, in the absence of 100% tumor incidence at the target site, which rarely occurs (Bemstein et aI., 1985; Freedman et aI., 1993; Gaylor and Gold, 1995, 1998; Gold et aI., 1997b). For example, the dose usually estimated by regulatory agencies to give one cancer in a million can be approximated simply by using the MTD as a surrogate for carcinogenic potency. The "virtually safe dose" (VSD) can be approximated from the MTD. Gaylor and Gold (1995) used the ratio MTDITDso and and TDso found by Krewski et al. the relationship between (1993) to estimate the VSD. The VSD was approximated by the MTD1740,000 for rodent carcinogens tested in the bioassay program of the NCIINTP. The MTD1740,000 was within a factor of 10 of the VSD for 96% of carcinogens. This is similar to the finding that in near-replicate experiments of the same chemical, potency estimates vary by a factor of 4 around a median value (Gold et aI., 1987a; Gold et aI., 1989; Gaylor et aI., 1993). Using the benchmark dose approach proposed in the EPA carcinogen guidelines, risk estimation is similarly constrained by bioassay design. A simple, quick, and relatively precise determination of the LTDlO can be obtained by the MTD divided by 7 (Gaylor and Gold, 1998). Both linear extrapolation and the use of safety or uncertainty factors proportionately reduce

q;

q;

806

CHAPTER 38


a tumor dose in a similar manner. The difference in the regulatory "safe dose," if any, for the two approaches depends on the magnitude of uncertainty factors selected. Using the benchmark dose approach of the proposed carcinogen risk assessment guidelines, the dose estimated from the LTDlO divided, for example, by a WOO-fold uncertainty factor, is similar to the dose of an estimated risk of less than 10-4 using a linear model. This dose is 100 times higher than the VSD corresponding to an estimated risk of less than 10-6 . Thus, whether the procedure involves a benchmark dose or a linearized model, cancer risk estimation is constrained by the bioassay design.

38.5 THE HERP RANKING OF POSSIBLE CARCINOGENIC HAZARDS Given the lack of epidemiological data to link pesticide residues to human cancer, as well as the limitations of cancer bioassays for estimating risks to humans at low exposure levels, the high positivity rate in bioassays, and the ubiquitous human exposures to naturally occurring chemicals in the normal diet that are rodent carcinogens (Tables 38.1-38.3), how can bioassay data best be used if our goal is to evaluate potential carcinogenic hazards to humans from pesticide residues in the diet? In several papers, we have emphasized the importance of setting research and regulatory priorities by gaining a broad perspective about the vast number of chemicals to which humans are exposed. A comparison of potential hazards can be helpful in efforts to communicate to the public what might be important factors in cancer prevention and when selecting chemicals for chronic bioassay, mechanistic, or epidemiologic studies (Ames et aI., 1987, 1990b; Gold and Zeiger, 1997; Gold et aI., 1992). There is a need to identify what might be the important cancer hazards among the ubiquitous exposures to rodent carcinogens in everyday life. One reasonable strategy for setting priorities is to use a rough index to compare and rank possible carcinogenic hazards from a wide variety of chemical exposures to rodent carcinogens at levels that humans receive, and then to focus on those that rank highest in possible hazard (Ames et aI., 1987; Gold et aI., 1992, 1994a). Ranking is thus a critical first step. Although one cannot say whether the ranked chemical exposures are likely to be of major or minor importance in human cancer, it is not prudent to focus attention on the possible hazards at the bottom of a ranking if, using the same methodology to identify a hazard, there are numerous common human exposures with much greater possible hazards. Our analyses are based on the HERP (human exposure/rodent potency) index, which indicates what percentage of the rodent carcinogenic dose (TDso in mg/kg/day) a human receives from a given average daily exposure for a lifetime (mg/kg/day). TDso values in our CPDB span a 10 million-fold range across chemicals (Gold et aI., 1997c). Human exposures to rodent carcinogens range enormously as well, from historically high workplace exposures in some occupations or pharmaceutical dosages to very low exposures from residues of synthetic chemicals in food or water.

The rank order of possible hazards for the given exposure estimates will be similar for the HERP ranking, for a ranking of regulatory "risk estimates" based on a linear model, or for a ranking based on TDlO, since all 3 methods are proportional to the dose. Overall, our analyses have shown that synthetic pesticide residues rank low in possible carcinogenic hazards compared to many common exposures. HERP values for some historically high exposures in the workplace and some pharmaceuticals rank high, and there is an enormous background of naturally occurring rodent carcinogens in typical portions or average consumption of common foods. This result casts doubt on the relative importance of low-dose exposures to residues of synthetic chemicals such as pesticides (Ames et aI., 1987; Gold et aI., 1992, 1994a). A committee of the National Research Council recently reached similar conclusions about natural versus synthetic chemicals in the diet and called for further research on natural chemicals (National Research Council, 1996). (See Section 38.7 for further work on natural chemicals. ) The HERP ranking in Table 38.5 is for average U.S. exposures to all rodent carcinogens in the CPDB for which concentration data and average exposure or consumption data were both available, and for which known exposure could be chronic for a lifetime. For pharmaceuticals the doses are recommended doses; for the workplace, they are past industry or occupation averages. The 87 exposures in the ranking (Table 38.5) are ordered by possible carcinogenic hazard (HERP), and natural chemicals in the diet are reported in boldface. Our early HERP rankings were for typical dietary exposures (Ames et aI., 1987; Gold et aI., 1992), and results are similar. Several HERP values make convenient reference points for interpreting Table 38.5. The median HERP value is 0.0025%, and the background HERP for the average chloroform level in a liter of U.S. tap water is 0.0003%. A HERP of 0.00001 % is approximately equal to a regulatory VSD risk of 10-6 based on the linearized multi-stage model (Gold et aI., 1992). Using the benchmark dose approach recommended in the new EPA guidelines with the LTDlO as the point of departure (POD), linear extrapolation would produce a similar estimate of risk at 10-6 and hence a similar HERP value (Gaylor and Gold, 1998), if information on the carcinogenic mode of action for a chemical supports a nonlinear dose-response curve. The EPA guidelines call for a margin-of-exposure approach with the LTDlO as the POD. Based on that approach, the reference dose using a safety or uncertainty factor of 1000 (i.e., LDlO/1000) would be equivalent to a HERP value of 0.001 %, which is similar to a risk of 10-4 based on a linear model. If the dose-response relationship is judged to be nonlinear, then the cancer risk estimate will depend on the number and magnitude of safety factors used in the assessment. The HERP ranking maximizes possible hazards to synthetic chemicals because it includes historically high exposure values that are now much lower [e.g., DDT, saccharin, butylated hydroxyanisole (BHA), and some occupational exposures]. Additionally, the values for dietary pesticide residues are averages in the total diet, whereas for most natural chemicals the ex-

38.5 The HERP Ranking of Possible Carcinogenic Hazards

807

Table 38.5 Ranking Possible Carcinogenic Hazards from Average V.S. Exposures to Rodent Carcinogens Possible hazard:

Potency TDso

HERP

Human dose of

(mg/kg/day)a

(%)

Average daily V.S. exposure

rodent carcinogen

Rats

Mice

Exposure references

140

EDB: production workers (high

Ethylene dibromide, 150 mg

1.52

(7.45)

Ott et al. (1980), Ramsey et al. (1978)

exposure) (before 1977) 17

Clofibrate

Clofibrate, 2 g

169

14

Phenobarbital, 1 sleeping pill

Phenobarbital, 60 mg

(+)

6.09

Havel and Kane (1982) AMA(l983)

6.8

1,3-Butadiene: rubber industry workers

1,3-Butadiene, 66.0 mg

(261)

13.9

Matanoski et al. (1993)

6.2

Comfrey-pepsin tablets, 9 daily

Comfrey root, 2.7 g

626

Tetrachloroethylene, 433 mg

101

(126)

Andrasik and Cloutet (1990)

Formaldehyde, 6.1 mg

2.19

(43.9)

Siegal et al. (1983)

Acrylonitrile, 405 !-lg

16.9

Trichloroethylene, 1.02 g

668

(1580)

(1978-1986) Hirono et al. (1978), Culvenor et al. (1980)

(no longer recommended) 6.1

Tetrachloroethylene: dry cleaners with dry-to-dry units (1980-1990)

4.0

Formaldehyde: production workers (1979)

2.4

Acrylonitrile: production workers

Blair et al. (1998)

(1960-1986) 2.2

Trichloroethylene: vapor degreasing

Page and Arthur (1978)

(before 1977) 2.1

Beer, 257 g

Ethyl alcohol, 13.1 ml

9110

(-)

Stofberg and Grundschober (1987)

1.4

Mobile home air (14 h/day)

Formaldehyde, 2.2 mg

2.19

(43.9)

Connor et al. (1985)

1.3

Comfrey-pepsin tablets, 9 daily

Symphytine, 1.8 mg

1.91

Methylene chloride, 471 mg

724

(1100)

CONSAD (1990)

(-)

Stofberg and Grundschober (1987) McCann et al. (1987)

Hirono et al. (1978), Culvenor et al. (1980)


Methylene chloride: workers, industry average (l940s-1980s)

0.5

Wine,28.0g

Ethyl alcohol, 3.36 ml

9110

0.5

Dehydroepiandrosterone (DHEA)

DHEA supplement, 25 mg

68.1

0.4

Conventional home air (14 h/day)

Formaldehyde, 598 !-lg

2.19

(43.9)

0.2

Omeprazole

Omeprazole, 20 mg

199

(-)

0.2

Fluvastatin

Fluvastatin, 20 mg

125

0.1

Coffee, 13.3 g

Caffeic acid, 23.9 mg

297

(4900)

Stofberg and Grundschober (1987),

0.1

d-Limonene in food

d-Limonene, 15.5 mg

204

(-)


0.04

Bread, 67.6 g

Ethyl Alcohol 243 mg

9110

(-)


0.04

Lettuce, 14.9 g


297

(4900)

TAS (1989), Herrmann (1978)

PDR (1998) PDR (1998) Clarke and Macrae (1988)

Wolm et al. (1974) 0.03

Safrole in spices

Safrole, 1.2 mg

(441)

51.3

Hall et al. (1989)

0.03

Orange juice, 138 g

d-Limonene, 4.28 mg

204

(-)

TAS (1989), Schreier et al. (1979)

0.03

Comfrey herb tea, 1 cup (1.5 g root)

Symphytine, 38 J.l g

1.91

Culvenor et al. (1980)


Tomato, 88.7 g


297

(4900)

TAS (1989), Schmidtlein and Herrmann (l975a)

0.03

Pepper, black, 446 mg

d-Limonene, 3.57 mg

204

(-)


0.02

Coffee, 13.3 g

Catechol, 1.33 mg

88.8

(244)


0.02

Furfural in food

Furfural, 2.72 mg

(683)

197


0.02

Mushroom (Agaricus bisporus) 2.55 g

Mixture of hydrazines, etc.

20,300


Hasselstrom et al. (1957) Tressl et al. (1978), Rahn and Konig (1978)

(whole mushroom)

Toth and Erickson (1986), Matsumoto et al. (1991) (continues)

808

CHAPTER 38


Table 38.5 (continued) Possible hazard:

Potency TDSO

HERP

Human dose of

(%)

Average daily V.S. exposure

0.02

Apple, 32.0 g

0.02

Coffee, 13.3 g

0.01

BHA: daily U.S. avg (1975)

0.01

Beer (before 1979), 257 g

rodent carcinogen

(mg/kg/day)a Rats

Mice

Exposure references


297

(4900)

EPA (1989a), Mosel and Herrmann (1974)

Furfural, 2.09 mg

(683)

197


BHA,4.6mg

606

(5530)

FDA (1991b)

Dimethylnitrosamine, 726 ng

0.0959

(0.189)

Stofberg and Grundschober (1987), Fazio et al. (1980), Preussmann and Eisenbrand (1984)

0.008

Aflatoxin: daily U.S. avg (1984-1989)

Aflatoxin, 18 ng

0.007

Cinnamon, 21.9 mg

Coumarin, 65.0 I!g

13.9

(103)

Poole and Poole (1994)

0.006

Coffee, 13.3 g

Hydroquinone, 333 I!g

82.8

(225)


0.0032

(+)

FDA (1992b)

Tressl et al. (1978), Heinrich and Baltes (1987) 0.005

Saccharin: daily V.S. avg (1977)

Saccharin, 7 mg

2140

(-)

0.005

Carrot, 12.1 g

Aniline, 624 I!g

194b

(-)

TAS (1989), Neurath et al. (1977)

0.004

Potato, 54.9 g

Caffeic acid, 867 I!g

297

(4900)

TAS (1989), Schmidtlein and Herrmann

0.004

Celery, 7.95 g


297

(4900)

ERS (1994), Stohr and Herrmann (1975)

NRC (1979)

(1975c) 0.004

White bread, 67.6 g

Furfural, 500 I!g

(683)

197


0.003

d-Limonene

Food additive, 475 !J.g

204

(-)

Clydesdale (1997)

0.003

Nutmeg, 27.4 mg

d-Limonene,466 !J.g

204

(-)


0.003

Conventional home air (14 h/day)

Benzene, 155 !J.g

(169)

77.5

McCann et al. (1987)

0.002

Coffee, 13.3 g

4-Methylcatechol, 433 I!g

248

Bejnarowicz and Kirch (1963) Stofberg and Grundschober (1987), Heinrich and Baltes (1987), IARC (1991) 0.002

Carrot, 12.1 g


297

(4900)

TAS (1989), Stohr and Herrmann (1975)

0.002

Ethylene thiourea: daily V.S. avg (1990)

Ethylene thiourea, 9.51 !J.g

7.9

(23.5)

EPA (1991a)

0.002

BHA: daily V.S. avg (1987)

BHA,700 !J.g

606

(5530)

FDA (1991b)

0.002 0.001

DDT: daily V.S. avg (before 1972 ban)d

DDT, 13.8!J.g

(84.7)

Plum,2.00g


297

12.8 (4900)

ERS (1995), Mosel and Herrmann (1974)

0.001

Pear, 3.29 g


297

(4900)


0.001

[VDMH: daily V.S. avg (1988)]

[VDMH, 2.82 !J.g (from Alar)]

(-)

3.96

EPA (1989a)

0.0009

Brown mustard, 68.4 mg

Allyl isothiocyanate, 62.9 I!g

96

(-)


0.0008

DDE: daily V.S. avg (before 1972 ban)d

DDE,6.91 !J.g

(-)

12.5

Duggan and Comeliussen (1972)

0.0007

TCDD: daily V.S. avg (1994)

TCDD, 12.0 pg

0.0000235

(0.000156)

EPA (1994b)

0.0006

Bacon, 11.5 g

Diethylnitrosamine, 11.5 ng

0.0266

(+)


0.0006

Mushroom (Agaricus bisporus) 2.55 g

Glutamyl-p-hydrazinobenzoate,

277


0.0005

Bacon, 11.5 g

Dimethylnitrosamine, 34.5 ng

0.0959

(0.189)


0.0004

Bacon, 11.5 g

N -Nitrosopyrrolidine, 196 ng

(0.799)

0.679


0.0004

EDB: daily V.S. avg (before 1984 ban)d

EDB,420ng

1.52

(7.45)

EPA (1984b)

0.0004

Tap water, 11iter (1987-1992)

Bromodichloromethane, 13 !J.g

(72.5)

47.7

AWWA(1993)

0.0003

Mango, 1.22 g

d-Limonene, 48.8 I!g

204

(-)

ERS (1995), Engel and Tress1 (1983)

Duggan and Come1iussen (1972)

Mosel and Herrmann (1974)

Car1son et al. (1987)

Sen et al. (1979) Chauhan et al. (1985)

107 I!g

Sen et al. (1979) Tricker and Preussmann (1991)

(continues)

38.5 The HERP Ranking of Possible Carcinogenic Hazards

809

Table 38.5 (continued) Possible hazard: HERP

Potency TDso (mg/kg/day)Q

Human dose of Average daily V.S. exposure

rodent carcinogen

Rats

0.0003

Beer, 257 g

Furfural, 39.9 IJg

(683)

197


0.0003

Tap water, 1 liter (1987-1992)

Chloroform, 17 J,lg

(262)

90.3

AWWA(1993) Gloria et at. (1997)

(%)

Mice

Exposure references

0.0003

Beer (1994-1995), 257 g

Dimethylnitrosamine, 18 ng

0.0959

(0.189)

0.0003

Carbaryl: daily V.S. avg (1990)

Carbaryl, 2.6 J,lg

14.1

(-)

FDA (1991a)

0.0002

Celery, 7.95 g

8-Methoxypsoralen,4.86IJg

32.4

(-)

ERS (1994), Beier et at. (1983)

(-)

0.0002

Toxaphene: daily V.S. avg (1990)d

Toxaphene, 595 ng

0.00009

Mushroom (Agaricus bisporus),

p-Hydrazinobenzoate, 28 IJg

0.00008

PCBs: daily V.S. avg (1984-1986)

5.57

FDA (l991a)

454b


(9.58)

Gunderson (1995) FDA (199Ia)

Chauhan et at. (1985)

2.55 g PCBs, 98 ng

1.74

0.00008

DDEIDDT: daily V.S. avg (l990)d

DDE,659ng

(-)

12.5

0.00007

Parsnip, 54.0 mg

8-Methoxypsoralen, 1.57 IJg

32.4

(-)

VFFVA (1989), Ivie et at. (1981)

0.00007

Toast, 67.6 g

Vrethane, 811 ng

(41.3)

16.9


0.00006

Hamburger, pan fried, 85 g

PhIP, 176 ng

4.22b

(28.6 b )

TAS (1989), Knize et at. (1994)

0.00006

Furfural

Food additive, 7.77 J,lg

(683)

0.00005

Estragole in spices

Estragole, 1.99 IJg

0.00005

Parsley, fresh, 324 mg

8-Methoxypsoralen, 1.17 IJg

0.00005

Estragole

Food additive, 1.78 J,lg

0.00003


MeIQx, 38.1 ng

0.00002

Dicofol: daily V.S. avg (1990)

Dicofol, 544 ng

0.00001

Beer, 257 g

U rethane, 115 ng

0.000006


0.000005

Hexachlorobenzene: daily V.S. avg

Canas et at. (1989) 197

Clydesdale (1997)

51.8


(-)

VFFVA (1989), Chaudhary et at. (1986)

51.8

Clydesdale (1997)

1.66

(24.3)

TAS (1989), Knize et at. (1994)

(-)

32.9

FDA (l99Ia)

(41.3)

16.9


IQ,6.38ng

1.65 b

(19.6)

TAS (1989), Knize et at. (1994)

Hexachlorobenzene, 14 ng

3.86

(65.1)

FDA (l991a) FDA (l991a)

32.4

Canas et al. (1989)

(1990) 0.000001

Lindane: daily V.S. avg (1990)

Lindane, 32 ng

(-)

30.7

0.0000004

PCNB: daily V.S. avg (1990)

PCNB (Quintozene), 19.2 ng

(-)

71.1

FDA (l991a)

0.0000001

Chlorobenzilate: daily V.S. avg (l989)d

Chlorobenzilate, 6.4 ng

(-)

93.9

FDA (l99Ia)

0.00000008

Captan: daily V.S. avg (1990)

Captan, 115 ng

2080

(2110)

FDA(l991a)

0.00000001

Folpet: daily V.S. avg (1990)

Folpet, 12.8 ng

(-)

1550

FDA (l991a)

; (2) an upper bound estimate of hypothetical, lifetime daily human exposure, TMRC; and (3) an upper bound estimate of excess cancer risk over a lifetime, calculated as potency x exposure. We obtained data from the EPA for 19 of the 26 chemicals discussed by the NRC (Quest et aI., 1993; U.S. Environmental Protection Agency, 1984a, 1985-1988, 1985a, 1985b, 1986b, 1987b, 1988b, 1989b, 1989c, 1999a). We were not able to identify the animal data used in the NRC report for cryomazine, diclofop methyl, ethalfluralin, ethylene thiourea, o-phenylpheno1, pronamide, and terbutryn. To verify that we had identified the correct rodent results, we attempted to replivalue for each of the 19 pesticides to define cate the EPA the data set for our comparison of risk estimates. The Tox-Risk program (Crump & Assoc.) was used to calculate as the 95% upper confidence limit on the linear term in the LMS, which theoretically represents the slope of the dose-response curve in the low-dose region. If it was not clear which target site had been used by the EPA, we calculated more than one and used in our subsequent comparison of potency estimates whichever value. If the EPA memorandata best reproduced the EPA dum for a chemical stated that the was the geometric mean of two or more experiments, we used the same method. The bioassay data that most accurately reproduced the EPA for each chemical are given in Table 38.6. Superscripts indicate the EPA weight-of-evidence classification given in the NRC report, followed by subsequent reevaluations of the classification. Using the data in Table 38.6 with the Tox-Risk program, overall there was good reproducibility in potency estimation (Table 38.7). We were able to reproduce the EPA qr value for

q;

q;

q;

q;

q;

q;

38.6 Pesticide Residues in Food: Investigation of Disparities in Cancer Risk Estimates Table 38.7 Reproducibility of the EPA

qf Values Reported by the NRC

Pesticide

EPA qi reported by NRC (mg/kg/day)-I

Chlorothalonil

2.4 x 10- 2

Asulam

2.0 x 10- 2 3.4 x 10- 2

Oryzalin Permethrin Chlordimeform Fosetyl Al Captafol Oxadiazon Cypermethrin Folpet Linuron Captan Alachlor Acephate Benomyl Metolachlor Glyphosate Parathion Azinphosmethyl

817

3.0 x 10- 2 9.4 x 10- 1 4.3 x 10- 3

Recalculated qi

Recalculated q

(mg/kg/day)-I

EPAqi

f/

1.3 x 10- 2 1.4 x 10- 2

O.S 0.7

2.S x 10- 2 2.0 x 10- 2 7.2 x 10- 1 3.7 x 10- 3

0.7 0.7 0.8

2.S x 10- 2 1.3 x 10- 1

2.4 x 10- 2

0.9 1.0

1.3 x 10- 1

1.0

1.9 x 10- 2 3.S x 10- 3

2.1 x 10- 2 3.8 x 10- 3

1.1

3.3 x 10- 1 2.3 x 10- 3

3.7 x 10- 1 3.4 x 10- 3

1.1

6.0 x 10- 2 6.9 x 10- 3 2.1 x 10- 3 2.1 x 10- 3

9.S x 1.3 x 4.6 x 8.7 x

1.6

S.9 x 10- 5 1.8 x 10- 3 I.S x 10- 7

10- 2 10- 2 10- 3 10- 3 4.8 x 10-4 1.3 x 10°

7.3 x 10- 1

1.1

I.S 1.9 2.2 4.1 6.1 720 4,900,000

Recalculated qj uses the bioassay data in Table 38.6 and a linearized multistage model.

15 chemicals within a factor of 2.2, and for 17 within a factor of 6. The median ratio of the q{ reported by the NRC to the recalculated q{ is 1.1. We could not approximate the q{ for parathion or azinphosmethyl. The q{ published in the NRC report for azinphosmethyl appears to be an error (W. Bumam, Office of Pesticide Programs, EPA, personal communication). We concluded that the data set of 15 chemicals with a q{ reproducibility within a factor of 2.2 would be used in the comparison of risk estimates. The four-chemicals for which we could not reproduce the q{ within a factor of 2.2 have all been reevaluated by the EPA since the NRC report: Azinphosmethyl and glyphosate are considered to have evidence of noncarcinogenicity to humans (i.e., superscript E in Table 38.6) (Bumam, 2000); a margin-of-exposure approach is recommended for metolachlor (MOE in Table 38.6); and parathion is classified as having limited evidence without a q{ value (Cnq) in Table 38.6. 38.6.2 COMPARISON OF POTENCY ESTIMATES: q~ AND TDso

Using the incidence data identified as those used by the EPA (Table 38.6), we estimated the TDso, that is, the dose rate (in mg/kg body weight/day) that is estimated to reduce by 50% the proportion of tumor-free animals at the end of a standard lifespan (Peto et aI., 1984; Sawyer et aI., 1984). The TDso does not involve extrapolation to low dose. It is inversely related to the slope (Peto et aI., 1984; Sawyer et aI., 1984; see Section 38.8 for details), and a comparison with q{ can be made by

using In(2)jTDso. An adjustment for rodent-to-human extrapolation, such as a surface area or other allometric correction factor, is usually applied to the q{ for regulatory purposes. For comparison purposes, the TDso was adjusted by the same interspecies scaling factor that was used by the EPA for q{, that is, (body weight)2/3, a factor of approximately 5.5 for rats and 13.0 for mice. The two potency estimates were then compared by computing the ratio q{ j(ln(2)jTDso). The dose calculation and standardization methods used for the TDso calculation in this chapter follow the EPA methods, some of which differ from the standard methodology used to estimate TDso in the CPDB. 38.6.3 COMPARISON OF HUMAN EXPOSURE ESTIMATES

The risk estimates in the NRC report (National Research Council, 1987) differed from those in the HERP ranking for dietary residues of synthetic pesticides (Section 38.5). The NRC reported upper bound estimates of daily human exposure (i.e., the EPA TMRC). In contrast, the HERP values in Table 38.5 used the daily exposure estimates from the FDA Total Diet Study (TDS). Thirteen pesticides discussed in the NRC report were measured in the TDS, and we compared the exposure estimates from the two sources for these 13. We used results from the TDS for the years 1984-1986 (Gunderson, 1995; U.S. Food and Drug Administration, 1988), which are the closest to the time of the NRC report.

818

CHAPTER 38


The EPA TMRC is a theoretical maximum estimate for potential human dietary exposure to synthetic pesticides. Pesticides registered for food crop use in the United States must first be granted tolerances under the Federal Food, Drug and Cosmetic Act (FFDCA). Tolerances are the maximum, legally allowable residues of the pesticide, or its active ingredient, on raw agricultural commodities and in processed foods. A tolerance is typically set for each pesticide for each crop use (e.g., corn, barley, wheat) based on field trials. The manufacturer conducts these trials, using varying rates of application under diverse environmental conditions, to determine both the minimum application rate needed to be effective against pest targets and the duration of time before harvest when it has to be applied (these are the rates specified on the pesticide label). Residue measurements are made on various parts of the crop at several time intervals after application to determine the rate of decline in residues of the pesticide active ingredient, its metabolites, and/or degradation products. The maximum measured residue is then used to establish the tolerance. Each crop use of a pesticide can have a different tolerance. Thus, the tolerance value is an upper bound estimate of total pesticide residue on a crop in the field, rather than in the marketplace or in table-ready foods. To obtain the TMRC, the tolerance value is multiplied by the mean U.S. food consumption estimate for each food item on which the pesticide is legally permitted, and exposures are combined for all such foods. The EPA, in calculating the TMRC, generally assumed that (l) each pesticide is used on all (100%) acres for each crop that the pesticide is permitted to be used on and (2) residues are present at the tolerance level (the highest allowable level in the field) in every food for which the pesticide is permitted. The National Food Consumption Survey conducted by the U.S. Department of Agriculture (USDA) is used for average food consumption estimates. Thus, the TMRC represents the hypothetical maximum exposure for a given pesticide (in mg/kg body weight/day) using field trial residue data. In contrast, the FDA Total Diet Study (TDS) measures detectable levels of pesticide residues as they are consumed, using a market basket survey of eight age-gender groups (Gartrell et aI., 1986; Gunderson, 1988, 1995; U.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a). Market baskets of foods are collected 4 times per year, once from each of four geographic regions of the United States. Each market basket consists of 234 identical foods purchased from local supermarkets in three cities in each geographic area. The foods are selected to represent the diet of the U.S. population, prepared table-ready, homogenized together and then analyzed for pesticide residues, including some metabolites and impurities (Gartrell et aI., 1986; Gunderson, 1988, 1995; U.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a). The levels of pesticide residues that are found are used in conjunction with the same USDA food consumption data used by the EPA in the TMRC in order to estimate the average dietary intake of pesticide residues in (mg/kg body weight/day) (Yess et aI., 1993). The TDS has been conducted annually by the FDA since 1961 (U.S. Food and Drug Administration, 1990), initiated primarily

in response to public concern about radionuclides in foods that might result from atmospheric nuclear testing. It is important to note that the TDS is distinct from FDA regulatory monitoring programs whose primary purpose is to ascertain that residues on crops at the "farm gate" or in the marketplace do not exceed maximum allowable levels and do not result from illegal pesticide use on crops for which the pesticide is not registered. FDA regulatory monitoring is designed only to make certain that regulations for pesticide use and application are followed, whereas the TDS is designed to provide an estimate of average daily dietary intake of pesticide residues in foods. Analytical methods for the TDS have been modified over time to permit measurement at concentrations 5 to 10 times lower than those used in FDA regulatory or incidence level monitoring. Generally, these methods can detect residues at 1 ppb (Gartrell et aI., 1986; Gunderson, 1988, 1995; U.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a). 38.6.4 COMPARISON OF RISK ESTIMATES

Of the chemicals for which we were able to reproduce the EPA

qt reported by the NRC, 10 were measured in the FDA Total Diet Study, and these were used to compare risk estimates based on different exposure assessments. Our analyses of the sources of variation in cancer risk estimates for dietary synthetic pesticides are presented in Tables 38.8-38.10. A comparison of the variation in potency estimates to the variation in exposure estimates is given in Table 38.8. Table 38.9 reports hypothetical dietary exposure estimates from the NRC report, i.e., the TMRC and measured residues in the FDA TDS. In Table 38.10, risk estimates based on the TMRC are compared to risk estimates based on the TDS, using in both cases the EPA as reported by the NRC. Because of missing data or NRC results that could not be reproduced, not all chemicals are included in every table; we have used all chemicals for which appropriate data were available. TDso values were calculated from the same dose and incidence data in Table 38.6 that were used to recalculate and these TDso values are reported in Table 38.6. Table 38.8 values for the 19 compares TDso values to recalculated chemicals, using the ratio qj /(In(2)/TDso). The qj and TDso values are within a factor of 2 of each other for 10 chemicals, and within a factor of 3 for 18 chemicals. These small differences in potency estimates are within the range of differences in potency estimates from near-replicate tests where the same chemical is tested in the same sex, strain and species of test animal (Gold et aI., 1987a, 1989, 1998; Gaylor et aI., 1993). Differences in potency values are larger only for azinphosmethyl, by a factor of 6.1; there was no statistically significant increase in tumor incidence for azinphosmethyl. In contrast to the similarity of potency estimation between In(2) /TDso and ,there is enormous variation in dietary exposure estimates for synthetic pesticides between the EPA TMRC values and the FDA average dietary residues in foods prepared as consumed (Tables 38.8 and 38.9). For 5 pesticides (alachlor,

qt

qt,

qt

qt

38.6 Pesticide Residues in Food: Investigation of Disparities in Cancer Risk Estimates

819

Table 38.8 Comparison of Variation in Measures of Potency and Exposure Pesticides included in

Ratio of potency:

the TDS (FDA)

recalculated

PermethrinCq AcephateNA(Cnq)

1.5

579

0.7

1,130

ParathionCq(Cnq)

2.6

6,300

Azinphosmethy lD(E) FolpetB2

6.1

7,530

q[ /(In(2)/TDso)

0.8

Ratio of exposure: EPAlFDA

9,650"

LinuronCq(Cnq) Captan B2

2.5 1.7

16,900

ChlorothalonilNA (B2)(MOE)

1.9

99,100

Alachlo~2(MOE)

0.9

Captafol B2 CypermethrinCq(Cnq) Oxadiazon B2 (Cq)

2.2

11,600

1.2 3.0

Pesticides not measured in the TDS (FDA) AsulamNA(Cnq)

2.5

NAc

BenomylCq

2.2

NAC

Chlordimeform B2 Fosetyl AlCq(Cnq)(Unclassified)

1.7

NAC

1.8

NAC

GlyphosateCq(E)

2.5

NAC

MetolachlorCq(Cnq)(MOE) Oryzalin Cq

1.8

NAC

2.5

NAC

aFolpet was not detected by the FDA in 1984-1986. This value is for 1987. bThe FDA did not detect any residues; therefore, no ratio could be calculated. CNot applicable because not measured by the FDA. Asulam had no food uses.

Table 38.9 Dietary Exposure Estimates in 1986 by the EPA and the FDA for Pesticides Measured in the FDA Total Diet Studya Daily intake (I-lglkg/day) Pesticide

EPA TMRC (1986)

FDA TDS (1984-1986)

PermethrinCq Captan B2 FolpetB2

14.0 206 92.6 5.41 11.3 8.19 4.65 9.91 0.408 23.8 0.197 0.0938 0.486c

0.0242 0.0122

AcephateNA(Cnq) AzinphosmethyID(E) ParathionCq(Cnq) LinuronCq(Cnq) ChlorothaloniINA (B2)(MOE) Alachlo~2(MOE)

Captafol B2 CypermethrinCq(Cnq) Oxadiazon B2 (Cq) PronamideCq (B2)

0.0096 0.0048 0.0015 0.0013 0.0004 0.0001

NDb NDb NDb NDb NDb

aFDA dietary estimates are for 60--65-year-old females for 1984-1986 (Gunderson, 1995). Because of the agricultural usage of these chemicals and the prominence of fruits and vegetables in the diet of older Americans, the residues are slightly higher than for other adult age groups. bNot detected at limit of quantification (~I ppb). cDid not appear in Tables 38.1 and 38.3 because no bioassay data were available.

captafol, cypermethrin, oxadiazon and pronamide), FDA found no residues at the 1 ppb limit of quantification (Gartrell et aI., 1986; Gunderson, 1988, 1995; D.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a; Yess et aI., 1993). Among chemicals detected by FDA, the TDS estimates were lower than the TMRC estimates by a factor of 99,100 for chlorothalonil, 16,900 for captan, 11,600 for linuron, and 9,650 for folpet (Table 38.8). For 4 other chemicals, the TDS estimates ranged from 579 to 7,530 times lower than TMRC. For the pesticides that EPA classified as having the strongest evidence of carcinogenicity in animal studies (B2), the differences in exposure estimates for EPA vs. FDA are particularly large (Table 38.8). Examination of FDA pesticide residue data collected over a period of 14 years (Gartrell et aI., 1986; Gunderson, 1988, 1995; D.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a) indicates that dietary exposure to pesticide residues has not changed markedly over time. Thus, the large differences in exposure estimates between EPA and FDA cannot be explained simply by changes in pesticide use patterns. In standard regulatory risk assessment, an estimate of the lifetime excess cancer risk is obtained by multiplying q~ by human exposure; the true risk, however, may be zero, as the 1986 EPA cancer risk assessment guidelines indicated (D.S. Environmental Protection Agency, 1986a). A comparison of the risk estimates obtained by multiplying the q~ in the NRC report by

820

CHAPTER 38


Table 38.10 Comparison of Cancer Risk Estimates Based on Different Exposure Measures: TMRC Versus TDsa Cancer risk reported by NRC

Cancer risk based on TDS

Pesticideb

based on TMRC (EPA)

(FDA)

LinuronCq(Cnq)

1.5 x 10- 3

1.3 x 10- 7

Captafol B2

5.9 x 10-4 4.7 x 10-4

0

CaptanB2 PermethrinCq FolpetB2 Chlorothalonil NA (B2)(MOE) AcephateNA(Cnq)

2.8 x 10- 8 7.3 x 10- 7

4.2 x 10-4 3.2 x 10-4

3.4 x 10- 8

2.4 x 10-4

2.4 x 10- 9 3.3 x 10- 8

OxadiazonB2 (Cq)

3.7 x 10- 5 2.4 x 10-5 1.2 x 10- 5

0 0

CypermethrinCq(Cnq)

3.7 x 10- 6

0

AlachlorB 2 (MOE)

Each risk> I x 10- 6

qt

Each risk < I x 10- 6

qt

aRisk estimates use values in the NRC report for pesticides with reproducible values (see Table 38.2, column 1). EPA risks are reported in the NRC book Regulating Pesticides in Food (1987). bThree chemicals measured in the Total Diet Study (Table 38.4) are excluded: For parathion and azinphosmethyl, the values could not be reproduced; for pronamide, we were unable to obtain bioassay results.

qt

TMRC vs. TDS exposure values is presented in Table 38.10. The risks based on TMRC are also reported by NRC, and range from 10-3 to 10-6 . In contrast, risk estimates using TDS are all lower than 10-6 . There are no risk estimates in Table 38.10 for the chemicals that FDA did not detect, i.e., if there is no exposure, there is no risk. Even if the undetected chemicals are considered to be present in minute quantities, below the limit of quantification, risk estimates for these undetected chemicals would be negligible, i.e., less than 10-6 . Thus, for synthetic pesticide residues in the diet, large discrepancies in cancer risk estimates are due to differences in exposure estimates rather than to differences in carcinogenic potency values estimated by different methods from rodent bioassay data. The high risk estimates reported by NRC in 1987 were overestimates based on EPA human exposure assessments which assumed that dietary residues were at tolerance levels. For example, the TDS did not detect any residues in table-ready foods for 4 pesticides that were evaluated in the NRC report as greater than 10-6 risks (Table 38.10). 38.6.5 USE OF EXPOSURE ASSESSMENTS IN RISK ASSESSMENT The results of our analysis emphasize the importance of exposure assessment in risk estimation for synthetic pesticide residues in the diet. Both the TDS of FDA and the TMRC of EPA link estimates of food consumption patterns for groups of individuals with an estimate of pesticide concentrations in food. Since FDA and EPA use the same USDA consumption surveys to estimate dietary patterns, food consumption is not a source of variation in their exposure estimates. However, the methods of estimating the concentrations of pesticide residues in food differed markedly. The FDA measured actual residues in food

items that are bought at the market and prepared as typically eaten; the EPA used a theoretical construct, based on worst case assumptions for the maximally exposed individual and maximally allowable levels, to estimate residues that could legally occur on a given food crop at the farm gate or in the marketplace. The EPA assumption that every pesticide registered for use on a food commodity is used on every crop is another source of overestimation of exposure (Winter, 1992). In California, for example, 54 insecticides were registered for use on tomatoes in 1986; however, the maximum number of insecticides used by any tomato grower was 5, 52% of tomato growers used 2 or fewer insecticides, and 31 % used none at all (Chaisson et al., 1989). Similar findings are reported for herbicides and fungicides. FDA monitoring programs have been criticized for not measuring enough pesticides or sampling enough food items, for aggregating foods under a single representative core food (e.g., apple pie to represent all types of fruit pies), and for statistical design and sampling. In several other independent studies, however, frequency of detection and residue concentrations have also been consistently low, for example, residue data from FOODCONTAM, a national database for state surveys on pesticide and other residues in foods (Minyard and Roberts, 1991). McCarthy (1991) collected residue data on 16 pesticides for 50 crops at the "farm gate"; although all crops had been treated with the label rates of pesticide application, 93% of 134 samples had concentrations below half the tolerance. Post-harvest treatment of crops, such as removing husks or outer leaves, shelling, peeling, and washing, all reduce residue levels still further (Yess et aI., 1993), as does processing. Eilrich (1991) measured residue levels on four produce crops "from the farm gate to the table" for a fungicide whose active ingredient is

38.6 Pesticide Residues in Food: Investigation of Disparities in Cancer Risk Estimates chlorothalonil and found that dietary residues were similar to those reported by the FDA. Analyses by Nigg et al. (1990) and Winter (1992) of residue data from the California Department of Food and Agriculture confirm the FDA regulatory monitoring findings. Most crops have no detectable residues; crop residues that are found are small fractions of tolerance values. Thus, tolerances are poor indicators of human exposure, a function for which they were not designed. Although it is possible that a small percentage of people who obtain food crops close to the farm gate may have higher incidental dietary exposures, these concentrations are very unlikely to persist over time and would still be substantially lower than the TMRC values. In the TDS, approximately 264 pesticides, metabolites, and impurities are analyzed; only 51 had detectable residues, and only 3 were present in more than 10% of the sample foods (V.S. Food and Drug Administration, 1991a). These findings are similar to those obtained from the TDS during the 10 previous years (Gartrell et aI., 1986; Gunderson, 1988; V.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a) and to those from surveys on pesticides of special interest. Even if exposure estimates based on the TDS were underestimates by an order of magnitude, the potential risks estimated using the EPA qi would still be low. The use of the TMRC as an estimate of human dietary exposure in quantitative cancer risk assessment is not justified, from either a scientific or a public policy perspective, because this measure often grossly exaggerates actual consumer exposure. The TMRC uses tolerances as surrogates for concentration in foods and therefore, by definition, the TMRC is not representative of the level likely to reach the consumer (Chaisson et aI., 1989). It does not take into account percentage of crop treated, actual pesticide application practices, chemical degradation from farm gate to table, and cooking or other processing. Some subsequent EPA exposure estimates have used "anticipated residues" instead of TMRC, which are calculated using tolerances and processing factors, using tolerances and percentage of crop treated, using field trial data, or using monitoring data. The anticipated residue also tends to be an overestimate because it is based on the average residue observed from maximum allowable pesticide application of a pesticide during field trials. Actual pesticide use is not always at the maximum level; hence, actual residues tend to be lower than the anticipated level (Chaisson et aI., 1989). For example, the EPA subsequently used anticipated residues to evaluate linuron and reported that less than 1% of the crop of barley, oats, and rye was treated. Despite this finding, for risk assessment purposes the EPA assumed that 100% of the crop was treated. The linuron comparison indicates how anticipated residues can be an overestimate: The TMRC in the NRC report was 4.65 f..Lg/kg/day; the anticipated residue reported by EPA was 0.185 f..Lg/kg/day (U.S. Environmental Protection Agency, 1995b); the TDS value was 0.0004 f..Lg/kg/day (Gunderson, 1995). Recent developments by government agencies have responded to the need for better quality information on exposure assessment of dietary residues. In response to the need identi-

821

fied by the National Academy of Sciences (NAS) for a standardized exposure database while developing the report Pesticides in the Diets of Infants and Children, the EPA has begun a National Pesticide Residue Database (NPRD) that collects data from the FDA, the VSDA, and private and commercial sources (http://www.epa.gov/pesticides/nprd). A mUltiagency effort, the Pesticide Data Program (PDP), is providing more information on actual exposure to dietary residues, food consumption, and pesticide usage (V.S. Environmental Protection Agency, 1999a). The PDP was established by the VSDA in 1991 to monitor pesticide residues in fresh and processed fruits and vegetables at terminal markets or distribution centers. Sampling procedures are designed to measure residues close to the time of consumption. Since 1994, the PDP testing protocol has included several foods in addition to fresh produce, such as canned and frozen fruits and vegetables and milk. The PDP is a critical component of the Food Quality Protection Act of 1996, and hence focuses on commodities that are consumed by infants and children (http://www. ams.usda.gov/science/pdp/what.htm). In 1998, PDP produce samples originated from 40 states and 25 foreign countries (V.S. Department of Agriculture, 2000). The PDP is currently used by the EPA to support its dietary risk assessment process [e.g., Eiden (1999)] and by the FDA to refine sampling for enforcement of tolerances. Given that exposure assessments for pesticide residues are available from the FDA TDS for about 38 years, it might be reasonable to compare those assessments to the new PDP assessments. A more complete characterization of exposures has been undertaken for some chemicals using biomarkers of exposure or distributions of exposure factors. Monte Carlo methods and other variance propagation techniques have been used to characterize the interindividual variability in exposures within a population and the uncertainty in exposure estimates (McKone, 1997). 38.6.6 USE OF TOXICOLOGICAL DATA IN RISK ASSESSMENT

Throughout this chapter, we have presented data indicating the limitations of tumor incidence results from rodent cancer tests in efforts to estimate human risk at low exposures. Our analysis of differences in risk estimates for dietary pesticide residues indicated that carcinogenic potency values were similar for In(2)jTDso and qi and therefore did not contribute substantially to the disparities in risk estimation. Similarity in potency estimates is expected: Bernstein et al. (1985) showed that carcinogenic potency values from standard bioassays are restricted to an approximately 32-fold range surrounding the maximum dose tested, in the absence of 100% tumor incidence. Estimates of carcinogenic potency derived from statistical models are highly correlated with one another because they are all highly correlated with the MTD, regardless of whether the estimate is based on the one-stage, multi stage, or Weibull model (Krewski et aI., 1990). This constraint on potency estimation

822

CHAPTER 38


contrasts with the enormous extrapolation that is required from the MTD in bioassays to the usual human exposure levels of pesticide residues, often hundreds of thousands of times lower than the MTD. One implication of the boundedness of potency estimation based on tumor incidence data is that for a given exposure estimate, the risk estimate can be approximated from the MTD in the bioassay without conducting an experiment. We have shown that the VSD at 10-6 (Gaylor and Gold, 1995) and the risk estimate based on the LTDlO, whether using safety factors for a nonlinear dose-response relationship or a linear model, can all be approximated from the MTD within a factor of 10 of the estimate that would be obtained from tumor incidence data in standard bioassays (Gaylor and Gold, 1998) (see Section 38.4). Adequate qualitative evaluation of the weight of evidence for carcinogenicity of a chemical and quantitative extrapolation from high to low dose requires more information for a chemical, about pharmacokinetics, mechanism of action, cell division, induction of defense and repair systems, and species differences. The new EPA guidelines and recent evaluations of several chemicals by the EPA, recognize the importance of such additional information (U.S. Environmental Protection Agency, 1996a). The proposed EPA guidelines permit the use of nonlinear approaches to low-dose extrapolation if warranted by mechanistic data. In recent years, the EPA has reevaluated several of the weight-of-evidence classifications for pesticides in the NRC report. This is consistent with the recommendation in the proposed EPA cancer risk assessment guidelines, which calls for use of available toxicological data in a characterization of the weight of evidence. Several pesticides are no longer considered appropriate for quantitative risk estimation (see the superscripts in parentheses in Table 38.6). Of the 19 pesticides from the NRC report for which we obtained bioassay data from EPA, only 11 are currently considered by the EPA as appropriate for quantitative risk estimation (Table 38.6 superscripts). This contrasts with the NRC report evaluation that the risks for 16 of the 19 were greater than 10-6 . For example, 1inuron had the highest risk estimate of all pesticides in the NRC analysis. It was subsequently reclassified by the EPA as inappropriate for quantitative risk assessment based on biological considerations: The testicular tumors in rats were late forming and benign and were a relatively common tumor type; the hepatocellular tumors in mice were benign and only in the highest dose group; and there was no evidence of mutagenic activity (U.S. Environmental Protection Agency, 1988a, 1999b). For evaluation of the mode of action of a given chemical using the new EPA risk assessment guidelines, information other than bioassay data can be developed and included in the assessment of weight of evidence and whether the dose-response relationship is likely to be nonlinear; for example, pharmacokinetic data on absorption, distribution, and metabolism can be used to predict target organ concentrations and then compared in different species. Other relevant results can be obtained from studies of cell division at and below the carcinogenic dose or from receptor-binding assays. New animal models with genetic alterations that are designed to make an animal resemble the

human more closely or to make the animal more sensitive to a given response can complement or take the place of long-term cancer tests, for example, transgenic mouse models that use unique phenotypic properties such as the p53 gene-deficient model or receptor-binding assays (Blaauboer et aI., 1998). Critical evaluation and validation of these new methodologies and increasing use of fundamental toxicological research will improve the regulatory evaluation of potential human risk. Although the proposed guidelines offer some incentive to generate mechanistic data on a chemical, for most chemicals no such data will be available, and the default procedure will continue to be used. If bioassay data are to be used in risk assessment, it is desirable to facilitate generation of mechanistic data on the chemicals of interest (Clayson and Iverson, 1996), including chemicals for which past risk assessments have resulted in regulation.

38.7 RANKING POSSIBLE TOXIC HAZARDS FROM NATURALLY OCCURRING CHEMICALS IN THE DIET Because naturally occurring chemicals in the diet have not been a focus of cancer research, it seems reasonable to investigate some of them further as possible hazards because they often occur at high concentrations in common foods. Only a small proportion of the many chemicals to which humans are exposed will ever be investigated, and there is at least some toxicological plausibility that high-dose exposures may be important. Moreover, the proportion positive in rodent cancer tests is similar for natural and synthetic chemicals, about 50% (see Section 38.3), and the proportion positive among natural plant pesticides is also similar (Table 38.3). Therefore, one would expect many of the untested natural chemicals to be rodent carcinogens. In order to identify and prioritize untested dietary chemicals that might be a hazard to humans if they were to be identified as rodent carcinogens, we have used an index, HERT, which is analogous to HERP (see Section 38.5). HERT is the ratio of human exposure/rodent toxicity (LDso) in mg/kg/day expressed as a percentage, whereas HERP is the ratio of human exposure/rodent carcinogenic potency (in mg/kg/day) expressed as a percentage. HERT uses readily available LDso values rather than the TDso values from animal cancer tests that are used in HERP. This approach to prioritizing untested chemicals makes assessment of human exposure levels critical at the outset. The validity of the HERT approach is supported by three analyses: First, we have found that for the exposures to rodent carcinogens for which we have calculated HERP values (Gold et aI., 1992), the rankings by HERP and HERT are highly correlated (Spearman rank order correlation = 0.89). Second, we have shown that without conducting a 2-year bioassay the regulatory VSD can be approximated by dividing the MTD by 740,000 (Gaylor and Gold, 1995; and Section 38.4). Because the MTD is not known for all chemicals and the MTD and LDso are both measures of toxicity, acute toxicity (LDso) can reasonably be used as a surrogate for chronic toxicity (MTD).

38.7 Ranking Possible Toxic Hazards from Naturally Occurring Chemicals in the Diet

Third, LDso and carcinogenic potency are correlated (Travis et aI., 1990; Zeise et aI., 1984); therefore, HERT is a reasonable surrogate index for HERP because it simply replaces TDso with LDso. We have calculated HERT values using LDso values as a measure of toxicity and human exposure estimates based on the available data on concentrations of untested natural chemicals in commonly consumed foods and average consumption of those foods in the U.S. diet. Literature searches identified the most commonly consumed foods (Stofberg and Grundschober, 1987; Technical Assessment Systems 1989; United Fresh Fruit and Vegetable Association, 1989) and concentrations of chemicals in those foods (Nijssen et aI., 1996; U.S. National Institute for Occupational Safety and Health, 1999). We considered any chemical with available data on rodent LDso that had a published concentration of ~ 10 ppm in a common food and for which estimates of average U.S. consumption of that food were available. The natural pesticides among the chemicals in the HERT table (Table 38.11) are marked with an asterisk. Among the set of 121 HERT values (Table 38.11), the HERT ranged across 6 orders of magnitude. The median HERT value for average dietary exposures is 0.007%. It might be reasonable to investigate further the chemicals in the diet that rank highest on the HERT index and that have not been adequately tested in chronic carcinogenicity bioassays in rats and mice. We have nominated to the NTP the chemicals with the highest HERT values as candidates for carcinogenicity testing. These include solanine and chaconine, the main alkaloids in potatoes, which are cholinesterase inhibitors that can be detected in the blood of almost all people (Ames, 1983, 1984; Harvey et aI., 1985); chlorogenic acid, a precursor of caffeic acid; and caffeine, for which no adequate standard lifetime study has been conducted in mice. In rats, cancer tests of caffeine have been negative, but one study that was inadequate because of early mortality, showed an increase in pituitary adenomas (Yamagami et aI., 1983). How would the synthetic pesticides that are rodent carcinogens and that are included in the HERP ranking (Table 38.5) compare to the natural chemicals that have not been tested for carcinogenicity (Table 38.11) if they too were ranked on HERT? We calculated HERT using LDso values for the synthetic pesticide residues that are rodent carcinogens in the HERP table and found that they rank low in HERT compared to the naturally occurring chemicals in Table 38.11; 88% (1071121) of the HERT values for the natural chemicals in Table 38.11 rank higher in possible toxic hazard HERT than any HERT value for any synthetic pesticide that is a rodent carcinogen in the HERP table (Table 38.5). The highest HERT for the synthetic pesticides would be for DDT in 1970 before the ban (0.00004%), which is more than lOO-fold lower than the median HERT for the natural chemicals in the HERT table. Many interesting natural toxicants are ranked in common foods in the HERT table. Oxalic acid, a plant pesticide, which is one of the most frequent chemicals in the table, occurs widely in nature. It is usually present as the potassium or calcium salt and also occurs as the free acid (Hodgkinson, 1977). Oxalic acid

823

is reported in many foods in Table 38.11; the highest contributors to the average U.S. diet are coffee (HERT = 0.09%), carrot (0.08%), tea (0.02%), chocolate (0.01 %), and tomato (0.01 %). Excessive consumption of oxalate has been associated with urinary tract calculi and reduced absorption of calcium in humans (Beier and Nigg, 1994; Hodgkinson, 1977). Because of the high concentrations of natural pesticides in spices, we have reported the HERT values for average intake in Table 38.11, even though spices are not among the foods consumed in the greatest amounts by weight. The highest concentrations of chemicals in Table 38.11 are found in spices, which tend to have higher concentrations of fewer chemicals (Nijssen et aI., 1996). (Concentrations can be derived from Table 38.11 by the ratio of the average consumption of the chemical and the average consumption of the food.) High concentrations of natural pesticides in spices include those for menthone in peppermint oil (243,000 ppm), y-terpinene in lemon oil (85,100 ppm), citral in lemon oil (75,000 ppm) piperine in black pepper (47,100 ppm), and geranial in lemon juice (14,400 ppm) and lemon oil (11,300 ppm). Natural pesticides in spices have antibacterial and antifungal activities (Billing and Sherman, 1998) whose potency varies by spice. A recent study of recipes in 36 countries examined the hypothesis that spices are used to inhibit or kill food spoilage microorganisms. Results indicate that as mean annual temperature increases in a geographical area (and therefore so does spoilage potential), there is an increase in number of spices used and use of the spices that have greatest antimicrobial effectiveness. The authors argue that spices are used to enhance food flavor, but, ultimately, are continued in use because they help to eliminate pathogens and therefore contribute to health, reproductive success, and longevity (Billing and Sherman, 1998). Cyanogenesis, the ability to release hydrogen cyanide, is widespread in plants, including several foods, of which the most widely eaten globally are cassava and lima bean (Poulton, 1983). Cassava is consumed widely throughout the tropics and is a dietary staple for over 300 million people (Bokanga et aI., 1994). There are few effective means of removing the cyanogenic glycosides that produce hydrogen cyanide (HCN), and cooking is generally not effective (Bokanga et aI., 1994; Poulton, 1983). For lima beans in Table 38.6, the HERT is 0.01 %. Ground flaxseed, a dietary supplement (http://www.heintzmanfarms.comJ; Gruenwald et aI., 1998), contains about 500 ppm hydrogen cyanide glycosides. The HCN in flaxseed appear to be inactivated in the digestive tract of primates (Mazza and Oomah, 1995). The increasing popUlarity of herbal supplements in the United States raises concerns about possible adverse effects from high doses or drug interactions (Saxe, 1987). Because the recommended doses of herbal supplements are close to the toxic dose and because about half of natural chemicals are rodent carcinogens in standard animal cancer tests, it is likely that many dietary supplements from plants will be rodent carcinogens that would rank high in possible carcinogenic hazard (HERP) if they were tested for carcinogenicity. Whereas pharmaceuticals are federally regulated for purity, identifica-

824

CHAPTER 38


Table 38.11 Ranking Possible Toxic Hazards to Naturally Occurring Chemicals in Food on the HERT (Human ExposurelRodent Toxicity) Index Possible hazard: HERT

Average human

LDSO (mg/kg)

(%)

Average daily consumption of food

consumption of chemical

Rats

Mice

Exposure references

4.3

Coffee, 500 ml (13.3 g)

*Caffeine, 381 mg

(192)

127

Stofberg and Grundschober (1987), Macaulay et al. (1984), IARC (1991)

0.3

Tea, 60.2 ml (903 mg)

*Caffeine, 29.4 mg

(192)

127

Stofberg and Grundschober (1987), Martinek and Wolman (1955), Wolman (1955), Lee (1973), Groisser (1978), Bunker and McWilliams (1979), Galasko et al. (1989), IARC (1991)

0.3

Potato, 54.9 g

*a-Chaconine, 4.10 mg

(84P)

19P

TAS (1989), Bushway and Ponnampalam (1981), Takagi et al. (1990)

0.2

Cola, 174 ml

*Caffeine, 20.8 mg

(192)

127

EPA (1996b), Bunker and McWilliams (1979), Galasko et al. (1989)

0.1

Coffee, 500 ml

*Chlorogenic acid, 274 mg

4000P

Stofberg and Grundschober (1987), Baltes (1977), IARC (1991)

0.09

Coffee, 500 ml

*Oxalic acid, 25.2 mg

382

Stofberg and Grundschober (1987), Kasidas and Rose (1980), IARC (1991), Vernot et al. (1977)

0.09

Black pepper, 446 mg

*Piperine, 21.0 mg

(514)

0.08

Carrot, boiled, 12.1 g


382

0.08

Chocolate (cocoa solids) 3.34 g

*Theobromine, 48.8 mg

(1265)

0.05

Lemon juice, 1.33 ml

*Geranial, 19.2 mg

500

EPA (1996b), Mussinan et al. (1981)

0.05

Coffee, 500 ml

*Trigonelline, 176 mg

5000

Stofberg and Grundschober (1987), Clinton (1986), IARC (1991)

0.03


*Caffeine, 2.30 mg

(192)

0.02

Tea, 60.2 ml


382

0.G2

Isoamyl alcohol: U.S. avg (mostly

Isoamyl alcohol, 18.4 mg

1300

330

Stofberg and Grundschober (1987) TAS (1989), Zarembski and Hodgkinson (1962), Vernot et al. (1977)

837

127

Stofberg and Grundschober (1987), IARC (1991)

Stofberg and Grundschober (1987), Zoumas et al. (1980) Stofberg and Grundschober (1987), Zarembski and Hodgkinson (1962), Kasidas and Rose (1980), IARC (1991), Vernot et al. (1977) Stofberg and Grundschober (1987)

beer, alcoholic beverages) 0.01

Beer, 257 m1


1300

Stofberg and Grundschober (1987), Arkima (1968)

0.01



382

Stofberg and Grundschober (1987), Kasidas and Rose (1980), Vernot et al. (1977)

0.01

Tomato, 88.7 g


382

Stofberg and Grundschober (1987), Zarembski and Hodgkinson (1962), Kasidas and Rose (1980), Vernot et al. (1977)

0.01

Coffee, 500 ml

2-Furancarboxylic acid, 821 J.!g

0.01

Lima beans, 559 mg

Hydrogen cyanide, 28.5 J.!g

0.01

Potato chips, 5.2 g

*a-Chaconine, 136 J.!g"

0.01

Sweet potato, 7.67 g

*Ipomeamarone, 336 J.!g

0.009

Potato, 54.9 g

*a-Solanine, 3.68 mg

590

0.008

Isobutyl alcohol: U.S. avg

Isobutyl alcohol, 14.1 mg

2460

0.008

Hexanoic acid: U.S. avg

Hexanoic acid, 15.8 mg

3000

0.007

Phenethyl alcohol: U.S. avg

Phenethyl alcohol, 8.28 mg

1790

0.007

Carrot, 12.1 g

*Carotatoxin, 460 J.!g

0.006

Ethyl acetate: U.S. avg

Ethyl acetate, 16.5 mg

lOOP

(84P)

Stofberg and Grundschober (1987), Tressl et al. (1978), IARC (1991), Kitamura et al. (1978)

3.7

EPA (1996b), Viehoever (1940), Montgomery (1964)

19P

Stofberg and Grundschober (1987), Friedman and Dao (1990)

50

Stofberg and Grundschober (1987), Coxon et al. (1975) TAS (1989), Bushway and Ponnampalam (1981), Takagi et al. (1990) Stofberg and Grundschober (1987)

(5000)


1001

Crosby and Aharonson (1967), Wulf et al. (1978)

4100


(beer, grapes, wine)

(5620)


(mostly alcoholic beverages)

(continues)


825


Possible hazard:

HERT

Average human

(%)



LDsO (mglkg) Rats Mice

0.005

Celery, 7.95 g


382

0.005

Coffee, 500 m1

*3·Methylcatechol, 203 flg

0.005

Potato, 54.9 g


Exposure references ERS (1994), Zarembski and Hodgkinson (1962), Vernot et al. (1977)

56V 382

Stofberg and Grundschober (1987), Heinrich and Baltes (1987), IARC (1991) TAS (1989), Zarembski and Hodgkinson (1962), Vernot et al. (1977)

0.004

Beer, 257 ml

Phenethy1 alcohol, 5.46 mg

1790


0.004

Corn, 33.8 g


382

Stofberg and Grundschober (1987), Kohman (1939), Vernot et al. (1977)

0.004

Corn, 33.8 g

Methylamine, 906 flg

0.004

Peppermint oil, 5.48 mg

*Menthone, 1.33 mg

0.004

White bread, 67.6 g

Propionaldehyde, 2.09 mg

0.004

Beer, 257 ml

Isobutyl alcohol, 6.40 mg

2460

0.003

Tomato, 88.7 g

Methyl alcohol, 13.4 mg

5628

0.003

Wine, 28.0 ml


1300

0.003

Coffee, 500 ml

Pyrogallol, 555 flg

0.003

Apple, 32.0 g

*Oxalic acid, 704 flg

382

EPA (1989a), Zarembski and Hodgkinson (1962), Kasidas and Rose (1980), Vernot et al. (1977)

0.003

Butyl alcohol: V.S. avg

Butyl alcohol, 1.45 mg

790


500 (1410)

317

Stofberg and Grundschober (1987), Neurath et al. (1977)

800

Stofberg and Grundschober (1987), Lorenz and Maga (1972)

(7300)

TAS (1989), Nelson and Hoff (1969), Kazeniac and Hall (1970)

300

Stofberg and Grundschober (1987), Tressl et al. (1978), IARC (1991)



Stofberg and Grundschober (1987), Postel et al. (1972)

(mostly apple, beer) 317

TAS (1989), Neurath et at. (1977)

Lettuce, 14.9 g Beer, 257 ml

Methylamine, 567 flg

0.003

Propyl alcohol, 3.29 mg

1870

(6800)


0.002

Banana, 15.7 g

trans- 2-Hexenal, 1.19 mg

(780)

685

TAS (1989), Hultin and Proctor (1961)

0.002

Orange, 10.5 g


382

0.002

Wine, 28.0 ml

Ethyl lactate, 4.16 mg

(>5000)

0.002 0.002

Tomato, 88.7 g White bread, 67.6 g

* p-Coumaric acid, Butanal, 3.44 mg

2490

0.002

Tea, 60.2 ml

*Theobromine, 1.11 mg

0.002 0.002

Apple, 32.0 g Tomato, 88.7 g

0.002 0.002

0.003

TAS (1989), Zarembski and Hodgkinson (1962), Vernot et al. (1977) 2500

Stofberg and Grundschober (1987), Postel et al. (1972), Shinohara et at. (1979)

657P

TAS (1989), Schmidtlein and Herrmann (1975a) Stofberg and Grundschober (1987), Lorenz and Maga (1972), Smyth et al. (1951)

(1265)

837

*Epicatechin, 1.28 mg

1000P

Stofberg and Grundschober (1987), Blauch and Tarka (1983), Nagata and Sakai (1985), IARC (1991) EPA (1989a), Risch and Herrmann (1988)

Beer, 257 ml

*Tomatine, 621 flg Ethyl acetate, 4.42 mg

(5620)

500 4100

Lettuce, 14.9 g


382

0.001

Apple, 32.0 g

*p-Coumaric acid, 573 flg

0.001

Apple, 32.0 g

*Chlorogenic acid, 3.39 mg

4000P

0.001

Coffee, 500 ml

Maltol, 462 flg

(1410)

0.001

Coffee, 500 ml

Nonanoic acid, 188 flg

1.02 mg

TAS (1989), Eltayeb and Roddick (1984) Stofberg and Grundschober (1987), Rosculet and Rickard (1968) TAS (1989), Kasidas and Rose (1980), Vernot et al. (1977)

657P

EPA (1989a), Mosel and Herrmann (1974) EPA (1989a), Jurics (1967), Perez-Ilzarbe et al. (1991)

550

Stofberg and Grundschober (1987), Tressl et al. (1978), IARC (1991)

224V

Stofberg and Grundschober (1987), Kung et al. (1967), IARC (1991) (continues)

826

CHAPTER 38


Table 38.11 (continued) Possible hazard: HERT

Average human

LDSO (mg/kg)

(%)



Rats

0.001

5-Methylfurfural: U.S. avg

5-Methylfurfural, 1.71 mg

2200

Stotberg and Grundschober (1987)

*fi-Pinene, 3.28 mg

4700


Mice

Exposure references

(mostly coffee) 0.001

fi-Pinene: U.S. avg (mostly pepper, lemon oil, nutmeg)

0.001

Broccoli, 6.71 g

*Oxalic acid, 268 fig

382

ERS (1994), Kohman (1939), Vernot et al. (1977)

0.001

Strawberry, 4.38 g


382

Stotberg and Grundschober (1987), Zarembski and Hodgkinson (1962), Kasidas and Rose (1980), Vernot et al. (1977)

0.0009

Orange juice, 138 ml

Methyl alcohol, 3.48 mg

5628

0.0009

a-Pinene: U.S. avg (mostly pepper,

*a-Pinene, 2.25 mg

3700

(7300)

TAS (1989), Kirchner and Miller (1957), Tanner and Limacher (1984), Nisperos-Carriedo and Shaw (1990) Stotberg and Grundschober (1987)

nutmeg, lemon oil) 0.0009

White bread, 67.6 g

2-Butanone, 1.65 mg

2737

(4050)

Stotberg and Grundschober (1987), Lorenz and Maga (1972)

0.0008

Coffee, 500 ml

Pyridine, 5 I 9 fig

891

(1500)

Stotberg and Grundschober (1987), Silwar et al. (1987), IARC (1991)

0.0008

Acetone: U.S. avg

Acetone, 1.74 mg

(5800)

3000


(698)

316

Stotberg and Grundschober (1987), Neurath et al. (1977)

317

Stotberg and Grundschober (1987), Neurath et al. (1977)

(mostly tomato, bread, beer) 0.0008

Cucumber, pickled, 11.8 g

Dimethylamine, 182 fig

0.0008

Cabbage, raw, 12.9 g

Methylamine, 169 fig

0.0007

Tomato, 88.7 g


4000P

0.0007

Wine, 28.0 ml

Methyl alcohol, 2.84 ml

5628

0.0007

Coffee, 500 ml

2-Methylpyrazine, 894 fig

1800


0.0007

Coffee, 500 ml

2,6-Dimethylpyrazine, 432 fig

880


0.0007

Cabbage, raw, green, 12.9 g

*p-Coumaric acid, 303 fig

0.0006

Peach, 9.58 g


4000P

Stotberg and Grundschober (1987), Jurics (1967), MolIer and Herrmann (1983), Senter et al. (1989)

0.0006


*3-Carene, 2.00 mg

4800

Stotberg and Grundschober (1987), Pino et al. (1990)

0.0006

Cabbage, boiled, 12.9 g

*OxaIic acid, 155 fig

382

Stotberg and Grundschober (1987), Zarembski and Hodgkinson (1962), Vernot et al. (1977)

0.0006

Coffee, 500 ml

Butyric acid, 785 fig

2000

Stotberg and Grundschober (1987), Kung et al. (1967), IARC (1991)

0.0006

Coffee, 500 ml

2,5-Dimethylpyrazine, 399 fig

1020


0.0005

Coffee, 500 ml

5-Methylfurfural, 798 fig

2200


0.0005

Grapes, I I g


382

Stotberg and Grundschober (1987), Kohman (1939), Vernot et al. (1977)

0.0005

Grapes, I I g


4000P

Stotberg and Grundschober (1987), Jurics (1967)

0.0005


*fi-Pinene, 1.50 mg

4700

Stotberg and Grundschober (1987), Pino et al. (1990)

TAS (1989), Winter and Herrmann (1986) (7300)

657P

Stotberg and Grundschober (1987), Postel et al. (1972)

Stotberg and Grundschober (1987), Schmidtlein and Herrmann (1975b)

0.0004

Cucumber (raw flesh), 11.8 g


382

Stotberg and Grundschober (1987), Kasidas and Rose (1980), Vernot et al. (1977)

0.0004

Potato chips, 5.2 g

*a-Solanine, 179 fig

590

Stotberg and Grundschober (1987), Ahmed and MUlIer (1978)

0.0004

Coffee, 500 ml

Propanoic acid, 785 fig

2600

Stotberg and Grundschober (1987), Kung et al. (1967), IARC (1991)

(continues)


827

Table 38.11 (continued) Possible hazard: Average human

HERT

LDSO (mglkg)

(0/0)



Rats

0.0004

Peach, canned, 9.58 g

*Oxalic acid, 115 fl,g

382

0.0004

Lettuce, 14.9 g

Benzylamine, 172 fl,g

0.0004

Lemon juice, 1.33 ml

Octanal, 1.60 mg

5630

EPA (1996b), Mussinan et al. (1981)

0.0004

a-Phellandrene: D.S. avg

*a-Phellandrene, 1.59 mg

5700


Hexanal, 1.35 mg

4890

Mice

Exposure references Stofberg and Grundschober (1987), Zarembski and Hodgkinson (1962), Vernot et al. (1977)

600P

TAS (1989), Neurath et al. (1977)

(mostly pepper) 0.0004

White bread, 67.6 g

0.0004


*a-Pinene, 1.02 mg

3700

0.0004

Banana, 15.7 g

2-Pentanone, 424 fl,g

1600

0.0003

Grapes, 11 g

*Epicatechin, 243 fl,g

0.0003

Onion, raw, 14.2 g

Dipropyl trisulfide, 189 fl,g

0.0003

Coffee. 500 ml

2-Ethyl-3-methylpyrazine,

(8292)

Stofberg and Grundschober (1987), Lorenz and Maga (1972)

1600

TAS (1989), Hultin and Proctor (1961)

l000P

Stofberg and Grundschober (1987), Jurics (1967), Lee and Jaworski (1987)

800


Stofberg and Grundschober (1987), Pino et al. (1990)

Stofberg and Grundschober (1987), IARC (1991)

880

186 fl,g 0.0003

Pear, 3.29 g

*Chlorogenic acid, 823 fl,g

4000P

Stofberg and Grundschober (1987), J urics (1967)

0.0003

Carrot, 12.1 g


4000P

TAS (1989), Winter et al. (1987)

0.0003

Lemon oil, 8 mg

*y-Terpinene, 681 fl,g

3650

Stofberg and Grundschober (1987), Ikeda et al. (1962), Staroscik and Wilson (1982a, 1982b)

0.0003

Lemon oil, 8 mg

*Geranial, 90.4 fl,g

500

Stofberg and Grundschober (1987), Bernhard (1960), Staroscik and Wilson (1982a, 1982b)

0.0003

Lemon oil, 8 mg

*,B-Pinene, 832 fl,g

4700

Stofberg and Grundschober (1987), Ikeda et al. (1962), Staroscik and Wilson (1982a, 1982b)

0.0002

Broccoli (raw), 6.71 g

*p-Coumaric acid, 90.6 fl,g

0.0002

Lemon oil, 8 mg

*Citral, 600 fl,g

4960

0.0001

Isoamyl acetate: D.S. avg

Isoamyl acetate, 1.70 mg

16,600

Dimethyl sulfide, 324 fl,g

3300

657P

ERS (1994), Schmidtlein and Herrmann (l975b)

(6000)

Stofberg and Grundschober (1987), Giinther (1968) Stofberg and Grundschober (1987)

(mostly beer, banana) 0.0001

Corn, canned, 33.8 g

0.0001

Onions, green, cooked, 137 mg

*Oxalic acid, 31.5 fl,g

382

0.0001

Coffee, 500 ml

Hexanoic acid, 245 fl,g

3000

0.0001

Pear, 3.29 g

*Epicatechin, 80.9 fl,g

(3700)

Stofberg and Grundschober (1987), Williams et al. (1972), Buttery et al. (1994)

(5000)

Stofberg and Grundschober (1987), Kung et al. (1967), IARC (1991) Stofberg and Grundschober (1987), Mosel and Herrmann (1974), Risch and Herrmann (1988)

EPA (1996b), Kohman (1939), Vernot et al. (1977)

1000P

0.00007

Nutmeg, 27.4 mg

*Myristicin, 207 fl,g

4260

0.00006

Banana, 15.7 g

Methyl alcohol, 236 fl,g

5628

0.00005

Lemon oil, 8 mg

*a-Pinene, 139 fl,g

3700

0.00005

Banana, 15.7 g

Isoamyl acetate, 584 fl,g

16,600

TAS (1989), Tressl et al. (1970)

0.00005

Strawberry, 4.38 g


4000P

Stofberg and Grundschober (1987), Jurics (1967)

0.00004


*a-Phellandrene, 162 fl,g

5700

0.00002

Grapefruit juice, 3.29 ml

Methyl alcohol, 95.4 fl,g

5628

0.00002

Lemon oil, 8 mg

*a-Terpinene, 23.2 fl,g

1680

0.00001

Lemon oil, 8 mg

*a-Terpineol, 29.6 fl,g

Ehlers et al. (1998) (7300)

TAS (1989), Hultin and Proctor (1961) Stofberg and Grundschober (1987), Ikeda et al. (1962), Staroscik and Wilson (1982a, 1982b)

Stofberg and Grundschober (1987), Pino et al. (1990) (7300)

Stofberg and Grundschober (1987), Kirchner et al. (1953), Lund et al. (1981), Tanner and Limacher (1984), Pino et al. (1986) Stofberg and Grundschober (1987), Staroscik and Wilson (1982a, 1982b)

2830

Stofberg and Grundschober (1987), Staroscik and Wilson (1982a, 1982b) (continues)

828

CHAPTER 38


Table 38.11 (continued) Possible hazard: HERT (%)


Average human

LDso (mg/kg)


Rats

Mice

Exposure references

2830

Stofberg and Grundschober (1987), Pino et al. (1990)

0.00001


*a-Terpineol, 25.0 J.4g

0.00001

Garlic, blanched, 53.3 mg

Diallyl disulfide, 2.05 J.4g

260

EPA (1996b), Yu et al. (1989)

0.00001

Lemon oil, 8 mg

*Terpinolene, 29.6 J.4g

4390

Stofberg and Grundschober (1987), Staroscik and Wilson (1982a, 1982b)

0.000008


Diallyl trisulfide, 592 ng

0.000001


Diallyl sulfide, 2.28 J.4g

100 2980

EPA (1996b), Yu et al. (1989) EPA (1996b), Yu et al. (1989)

LDSO: Values are from the Registry of Toxic Effects of Chemical Substances (RTECS). Parentheses indicate the species with the higher (weaker) LDSO, which is not used in the HERT calculation. Daily human exposure: The average amount of the food consumed daily per person in the United States; when a chemical is listed rather than a food item, the value is the per person average in the total diet. All other calculations assume a daily dose for a lifetime. Possible hazard: The amount of chemical reported under "Human dose of chemical" is divided by 70 kg to give a mg/kg of human exposure. The HERT is this human dose (mg/kglday) as a percentage of the rodent LDso (mg/kg). An * preceding a chemical name indicates that the chemical is a natural pesticide. Abbreviations for LDso values: LO, LDLO; P, intraperitoneal injection; V, intravenous injection; J, injection (route not specified).

tion, and manufacturing procedures and additionally require evidence of efficacy and safety, dietary supplements are not. We found that several dietary supplements would rank high in the HERT table if we had included them by using the recommended dose and the LDso value for the extract: ginger extract (HERT = 0.8%), ginkgo leaf extract (HERT = 0.7%), ginseng extract (HERT = 0.7%), garlic extract (HERT = 0.1%), and valerian extract (HERT = 0.01%). These results argue for greater toxicological testing requirements and regulatory scrutiny of dietary supplements on the grounds that they may be carcinogens in rodents and that, if so, they are likely to rank high in possible carcinogenic hazard. Because these products lack requirements for toxicological testing, the NTP has established a research program on medicinal herbs and ingredients.

38.8 SUMMARY OF CARCINOGENICITY RESULTS IN THE CPDB ON ACTIVE INGREDIENTS OF COMMERCIAL PESTICIDES THAT HAVE BEEN EVALUATED BY THE U.S. EPA This section presents summary results on each of 193 commercial pesticide ingredients that are listed by the EPA in "Status of Pesticides in Registration, Reregistration, and Special Review," its Rainbow Report (U.S. Environmental Protection Agency, 1998) and that are also included in the CPDB. Results for pesticides that are negative for carcinogenicity in the CPDB are included. Approximately 1900 pesticides are listed in the Rainbow Report, but only 193 have published results of carcinogenicity experiments that meet the inclusion criteria of the CPDB (Gold and Zeiger, 1997). Table 38.12 provides a quick overview of the CPDB results on each pesticide, including the following information: the sex-species groups that have

been tested, the strongest level of evidence of carcinogenicity based on the opinion of the published author, carcinogenic potency (TDso), target organs in each species, and mutagenicity in Salmonella typhimurium. Carcinogenicity results for rats and mice are reported in Table 38.12, and in Table 38.13 for hamsters, monkeys, and dogs. For each pesticide, the details on each experiment are reported in the CPDB, published in the CRC Handbook of Carcinogenic Potency and Genotoxicity Databases (Gold et aI., 1997c) and in Environmental Health Perspectives (Gold et aI., 1999) as well as reported in http://potency.berkeley.edu. The following describe the data reported in Tables 38.12 and 38.13. Pesticides A chemical is considered a pesticide if it appears in the EPA "Status of Pesticides in Registration, Reregistration, and Special Review" (U.S. Environmental Protection Agency, 1998), the Rainbow Report. Included in the Rainbow Report is the status of pesticides that are undergoing pesticide reregistration, that have completed pesticide reregistration, that are under special review, or that are "new" (i.e., that have been registered since 1984). For 79 of the 193 commercial pesticides in the table, the active ingredient is no longer contained in any registered pesticide product; for these cases of voluntary or regulated cancellation, we indicate this fact by an asterisk next to the chemical name. If a commercial pesticide is also a chemical that occurs naturally, the chemical name is in boldface. Mutagenicity in Salmonella A chemical is classified as mutagenic in the Salmonella assay "+" if it was evaluated as either "mutagenic" or "weakly mutagenic" by Zeiger (1997) or as "positive" by the Gene-Tox Program (Auletta, personal communication; Kier et aI., 1986). All other chemicals evaluated for mutagenicity by these two sources are reported as "-." The symbol "." indicates that these sources did not provide an

38.8 Summary of Carcinogenicity Results in the Carcinogenic Potency Database

829

Table 38.12 Summary of Carcinogenicity Results in Rats and Mice in the Carcinogenic Potency Database on 193 Active Ingredients Commercial Pesticides that Have Been Evaluated by the V.S. Environmental Protection Agency Harmonic mean of Sal-

TDSO (mglkg/day)

Pesticide

CAS

moneIIa

Rat

Acrolein

107-02-8

Acrylonitrile*

107-13-1

+ +

16.9ffi .v

Mouse

Rat target sites Male

Female

ezy nrv orc smi

ezy mgl nas nrv

sto Aldicarb

116-06-3

Aldrin*

309-00-2

Allantoin*

97-59-6

Allyl isothiocyanate

57-06-7

3-Aminotriazoles

61-82-5

Anethole*

104-46-1

Anilazine*

101-05-3

Antimony potassium tartrate*

28300-74-5

Arsenate, sodium s

7631-89-2

Arsenious oxide

1327-53-3

Arsenite, sodium*

7784-46-5

Aspirin

50-78-2

Atrazine

1912-24-9

Azinphosmethy1

86-50-0

Benzaldehyde*

100-52-7

Benzene*

96 9.94ffi

thy

pit thy

B-

B-

Benzoic acid*

65-85-0

Benzyl aIcohol*

100-51-6

o-Benzyl-p-chlorophenol

120-32-1

Biphenyl*

92-52-4

Bis(tri-n-butyItin)oxide,

56-35-9

liv(B)

liv

liv

B-

B-

B-

B-

B-

B31. 7m

mgl

hmoute

ezy nas orc ski

ezy nas orc sto

+ 169m

1490ffi 77.5 m ,v

sto vsc 532-32-1

liv ubi

25.3 ffi

B-

71-43-2

Benzoate, sodium*

Female

orc smi sto

l.27ffi

+

Mouse target sites Male

sto

sto

ezy hag hmo

ezy hmo lun

lun pre

mglova

vsc

1350

kid

technical grade Boric acid

10043-35-3

tert-Butyl alcohol*

75-65-0

Butyl p-hydroxybenzoate*

94-26-8

p-tert-Butylphenol*

98-54-4

Cadmium chlorides *

10108-64-2

Calcium chloride*

10043-52-4

Capsaicin

404-86-4

Captan

133-06-2

Carbaryl

63-25-2

Carbon tetrachloride

56-23-5

Chloramben*

133-90-4

Chloranil*

118-75-2

Chlordane, technical grade*

57-74-9

Chlorinated trisodium phosphate

56802-99-4

Chlorine

7782-50-5

3-Chloro-p-toluidine

95-74-9

Chlorobenzilate*

510-15-6

(2-Chloroethyl) trimethyl-

999-81-5

64.6

21900

0.0114 ffi ,v

kid

thy

hmo 1un pro tes

lun

kid

ute

tba(B)

tba(B)

liv

Iiv mgl

167 ffi ,n

+ + +

2080m 14.1 2.29m ,n

2110m 150m

19i

19i

smi

smi

adrliv

adrliv

5230

liv

1.37m ,v

liv

liv

liv

liv

+ B93.9m ,v

B-

ammonium chloride

(continues)

830

CHAPTER 38


Table 38.12 (continued) Harmonic mean of SalPesticide

CAS

Chloroforms 3-(p-Chlorophenyl)-I,I-di-

monella

TDSO (mg/kg/day)

Rat target sites

Mouse target sites

Rat

Mouse

Male

Female

Male

Female

67-66-3

262 ffi

90.3 ffi

kid

liv

kid liv

liv

150-68-5

131

kid liv

2270ffi

kid

kid lun vsc

lun vsc

liv lun

lun

hrno liv lun

hmo liv lun

methylurea* Chloropicrin

76-06-2

Chlorothalonil

1897-45-6

Citric acid

77-92-9

Clonitralid

1420-04-8

Copper-8-hydroxyquinoline

10380-28-6

Coumaphos

56-72-4

Cyanamide, calcium*

156-62-7

Cyclohexanone*

108-94-1

Daminozide

1596-84-5

p,p'-DDD*

72-54-8

DDTs*

50-29-3

Deltamethrin

52918-63-5

Diallate*

2303-16-4

Diazinon

333-41-5

1,2-Dibromo-3-chloropropane*

96-12-8

+

I

kid

+ 2500n

1030ffi

84.7ffi

30.7 ffi 12.8ffi ,v

liv

liv

26.7 ffi

+ +

ute

0.259 ffi

2.72ffi

liv nas orc sto

adrmgl nas

lun nas sto

lun nas sto

lun sto vsc

eso lun mgl nas

orc sto 1,2-Dibromoethane*

106-93-4

3,5-Dichloro(N -I, l-dimethyl-2-

+

1.52ffi

23950-58-5

7.45 ffi ,V

nas per pit

liv lun mgl nas

sto vsc

pit sto vsc

119

sto sub vsc liv

propynyl)benzamide 2,3-Dichloro-1A-naphtho

117-80-6

quinone* 2,6-Dichloro-4-nitroaniline

99-30-9

1,2-Dichlorobenzene*

95-50-1

1A-Dichlorobenzene Dichlorodifluoromethane*

106-46-7 75-71-8

1,2-Dichloroethane*

107-06-2

a-(2,4-Dichlorophenoxy)propi-

120-36-5

+ 644

398 ffi

kid

+

8.04ffi

IOlffi

sto sub vsc

+

4.16

70.4ffi

hmo pan

mgl

liv

liv

lun

lun mgl ute

sto

sto

onic acid

204- Dichlorophenoxyacetic

acid

2,4-Dichlorophenoxyacetic

94-75-7 94-80-4

acid, n-butyl ester* 2,4-Dichlorophenoxyacetic

94-11-1

acid, isopropyl ester 3-(3A-Dichlorophenyl)-I,I-di-

330-54-1

methylurea Dichlorvos

62-73-7

Dicofol

115-32-2

32.9

liv

Dieldrins *

60-57-1

0.912ffi

liv

O,O-Diethyl-o-(3,5,6-trichloro-

2921-88-2

liv

2-pyridyl)phosphorothioate Dimethoate

60-51-5

Dimethoxane

828-00-2

+ +

716

hrno kid liv ski sub

(continues)


831


CAS

Dimethylarsinic acid

75-60-5

2,4-Dinitrophenol*

51-28-5

Dioxathion*

78-34-2

n-Dodecylguanidine acetate

2439-10-3

EDTA, trisodium salt

150-38-9

monella

TDSO (mg/kglday) Rat

Mouse


Female


Female

B-

B-

hag lun

haghmo lun

+

trihydrate* Endosulfan

115-29-7

Endrin*

72-20-8

Ethoxyquin

91-53-2

Ethyl alcohol

64-17-5

p, p'-Ethyl-DDD*

72-56-0

Ethylene glycol*

107-21-1

Ethylene oxide

75-21-8

Ethylenebisdithiocarbamate,

142-59-6

9110

adr liv pan pit

+ +

21.3 ffi ,v

63.7ffi

hmonrv per

nrv sto

mglute disodium di(2-Ethylhexyl)phthalate*

117-81-7

Eugenol

97-53-0

Fenaminosulf, formulated*

140-56-7

Fenthion

55-38-9

Fenvalerate

51630-58-1

Ferric dimethy ldithiocarbamate

14484-64-1

Fluometuron

2164-17-2

Fluoride, sodium

7681-49-4

Formaldehyde'

50-00-0

Fosetyl AI

39148-24-8

Furfural'*

98-01-1

Gibberellic acid

77-06-5

Glycerol a-monochlorohydrin*

96-24-2

Heptachlor

76-44-8

fJ-l,2,3,4,5,6-HexachlorocycIo-

625 ffi

894ffi

liv

Iiv

Iiv

2.19 m.v

43.9

hmonas

hmo nas

nas

liv

+

+

3660

+

683

ubi 197ffi

liv

liv

liv

+ Iiv

liv

319-85-7

1.21 m 27.8 m

liv

Iiv

58-89-9

30.7 m

liv

liv lun

hexane y -1 ,2,3,4,5,6-HexachlorocycIo-

hexane Hexachlorophene*

70-30-4

3-(Hexahydro-4,7 -methanoin-

2163-79-3

dan-5-yl)- I, I-dimethylurea* Hydrochloric acid

7647-01-0

Hydrogen peroxide

7722-84-1

8-Hydroxyquinoline*

148-24-3

Isopropyl-N -(3-chlorophenyl)

101-21-3

+ +

7540

smi

carbamate' Isopropyl-N -phenyl

122-42-9

carbamate' * Kepone*

143-50-0

Malathion

121-75-5

Maleic hydrazide

123-33-1

2.96

0.982m

liv

Iiv

liv

(continues)

832

CHAPTER 38



CAS

Manganese ethylenebisthiocar-

12427-38-2

2-Mercaptobenzothiazole*

149-30-4

2-Mercaptobenzothiazole, zinc

155-04-4

monella

TDSO (mg/kg/day) Rat

Mouse


Female

157

tba(B)

tba(B)

344m

adr hmo pan pre

adrpit


Female

bamate

Mercuric chloride*

7487-94-7

Methidathion

950-37-8

Methoxychlor

72-43-5

Methyl bromide

74-83-9

Methyl parathion

298-00-0

Methylene chloride*

75-09-2

Metronidazole*

443-48-1

Mexacarbate*

315-18-4

Mirex*

2385-85-5

Naphthalene

91-20-3

I-Naphthalene acetamide

86-86-2

I-Naphthalene acetic acid

86-87-3

Nickel (11) sulfate

10101-97-0

3.12

st~

6.04

+ + + +

724m.i 542ffi l.77ffi

liv

1100m,i

mgl

mgl

livlun

liv lun

506ffi

pit tes

livmgl

lun

hmolun

1.45 ffi

adr kid liv

hmoliv

liv

liv

163 i

lun

+

hexahydrate* Nicotine

54-11-5

Nitrate, sodium

7631-99-4

Nitrite, sodium'

7632-00-0

Nitrofen*

1836-75-5

Oleate, sodium*

143-19-1

Oxamyl

23135-22-0

Oxytetracycline.HCI

2058-46-0

Parathion

56-38-2

Pentachloronitrobenzene

82-68-8

2,3,4,5,6-Pentachlorophenol

87-86-5

+ +

167ffi 420

hmo(B) liv 115ffi

hmo(B) liv pan

liv vsc

71.1 24ffi

liv

liv adr liv

adr liv vsc

(Dowicide EC-7)

Phenol

108-95-2

Phenothiazine*

92-84-2

Phenylmercuric acetate*

62-38-4

o-Phenylphenate, sodium

132-27-4

o-Phenylphenol

90-43-7

Phosphamidon*

13171-21-6

PicIoram, technical grade

1918-02-1

Piperonyl butoxide in solvent

51-03-6

Piperonyl sulfoxide*

120-62-7

Polysorbate 80*

9005-65-6

Potassium bicarbonate

298-14-6

Propazine

139-40-2

Propyl N -ethyl-N -butylthiocar-

1114-71-2

+ +

545 ffi ,v

kid ubi

232

ubi

ubi

62.2 13,OOOffi

liv ubi

ubi

adr nas

mgl nas

bamate n-Propyl isome*

83-59-0

Propylene glycol*

57-55-6

1,2-Propylene oxide

75-56-9

FD & C red no. 3*

16423-68-0

Rotenone

83-79-4

+

74.4ffi ,V

912ffi

st~

nas

nas

(continues)


833


CAS

Safrole*

94-59-7

Simazine

122-34-9

Sodium bicarbonate

144-55-8

Sodium chloride

7647-14-5

Sodium chlorite

7758-19-2

Sodium dichromate

10588-01-9

Sodium hypochlorite

7681-52-9

Strobane*

8001-50-1

Sulfallate*

95-06-7

Telone II

542-75-6

2,4,5,4' -Tetrachlorodiphenyl

116-29-0

monella

TDso (mg/kg/day)

Rat target sites

Mouse target sites

Rat

Mouse

Male

Female

Male

Female

441ffi

5l.3 ffi ,v

liv

liv(B)

liv

liv

lun

+

4.M

+ +

26,lffi 94ffi

42.2ffi

sto

mgl

49.6

liv sto

sto

lOl ffi

126ffi

hmo kid

hmo

lun 0.884ffi

hmoliv mgl lun sto ubI

sulfone* Tetrachloroethylene*

127-18-4

Tetrachlorvinphos

961-11-5

Tetrakis(hydroxymethyl)phos-

55566-30-8

liv

liv

228

liv

5.57 ffi 584ffi

liv

liv

liv

liv

I 550ffi

smi sto

smi

liv

liv

tba

tba

phonium sulfate Tetramethylthiuram disulfide

137-26-8

Thiabendazole

148-79-8

Toxaphene*

8001-35-2

Trichloroacetic acid*

76-03-9

I, 1,1-Trichloroethane, technical

71-55-6

+ + +

grade* Trichlorofluoromethane*

75-69-4

N -(Trichloromethy lthio )phthal-

133-07-3

+

imide 2,4,6-Trichlorophenol*

88-06-2

2,4,5-Trichlorophenoxyacetic

93-76-5

405

lO70ffi

hmo

acid* Triethanolamine

102-71-6

Triethylene glycol

112-27-6

Trifluralin, technical grade

1582-09-8

Triphenyltin hydroxide

76-87-9

Urea*

57-13-6

Xylene mixture (60%

1330-20-7

lOOffi

+

330

liv lun sto

m-xylene, 9% o-xylene, 14% p-xylene, 17% ethylbenzene) FD & C yellow no.5

1934-21-0

Zinc dimethyldithiocarbamate

137-30-4

Zinc ethylenebisthiocarbamate*

12122-67-7

+

40.7 ffi

tba(B) thy

tba(B)

255

tba(B)

tba(B)

Abbreviations: " not tested; (B), data reported only for both sexes combined. Tissue codes: adr, adrenal gland; eso, esophagus; ezy, ear/Zymbal's gland; hag, harderian gland; hmo, hematopoietic system; kid, kidney; 19i, large intestine; liv, liver; lun, lung; mgl, mammary gland; nas, nasal cavity (includes tissues of the nose, nasal turbinates, paranasal sinuses, and trachea); nrv, nervous system; orc, oral cavity (includes tissues of the mouth, oropharynx, pharynx, and larynx); ova, ovary; pan, pancreas; per, peritoneal cavity; pit, pituitary gland; pre, preputial gland; pro, prostate; ski, skin; smi, small intestine; sto, stomach; sub, subcutaneous tissue; tba, all tumor bearing animals; tes, testes; thy, thyroid gland; ubI, urinary bladder; ute, uterus; vsc, vascular system. In a series of footnotes, we provide additional information about TDso values and test results in the CPDB. These are as follows: i, carcinogenic in rodents only by the inhalation route of administration; m, more than one positive test in the species in the CPDB; n, no results that were evaluated as positive by the published author for this species in the CPDB have statistically significant TDSO values (two-tailed p < 0.1); s, species other than rats or mice are reported for this chemical in Table 38.13; v, variation is greater than lO-fold among statistically significant (p < 0.1) TDso values from different positive experiments. Note: The commercial pesticides in boldface also occur naturally. 'Voluntary or regulated cancellations. The Active Ingredient Is No Longer Contained in Any Registered Pesticide Product.

834

CHAPTER 38


Table 38.13 Summary of Carcinogenicity Results in the Carcinogenic Potency Database in Other Species on 11 Commercial Pesticides Ingredients Evaluated by the V.S. Environmental Protection Agency Harmonic mean of Pesticide

CAS

Salmonella

TD50 (mg/kg/day)

Target sites

Hamsters 3-Aminotriazole

61-82-5

Cadmium chloride*

10108-64-2

DDT*

50-29-3

Dieldrin*

60-57-1

Formaldehyde

50-00-0

Furfural*

98-01-1

Isopropyl- N -(3-chlorophenyl)carbamate

101-21-3

Isopropyl-N -phenyl carbamate*

122-42-9

Nitrite, sodium

7632-00-0

+ + +

Cynomolgus monkeys DDT*

50-29-3

Rhesus monkeys Arsenate, sodium

7631-89-2

DDT*

50-29-3

Dogs Chloroform

67-66-3

The Commercial Pesticides in Boldface also Occur Naturally. 'Voluntary or regulated cancellations. The Active Ingredient Is No Longer Contained in Any Registered Pesticide Product. Abbreviations: ., not tested.

evaluation. For a chemical of interest, results for other genotoxicity tests are reported for some chemicals in the Genotoxicity Database (Zeiger, 1997). Carcinogenicity For each positive chemical in the CPDB, results are included on carcinogenic potency (by species) and target organ (by sex-species); if there are no positive results, then the symbol "-" appears. The classification of positivity in this summary table is based on a positive result in at least one experiment. There may be additional experiments on the same chemical that are negative in the CPDB, but this is not reflected in the table. An experiment is classified as positive or negative on the basis of the published author's opinion. A target site is classified as positive for NCIINTP if the evaluation in technical report was "carcinogenic" or "clear" or "some" evidence of carcinogenic activity ["c" or "p" on the plot of the CPDB (Gold et aI., 1997c, 1999)]. In the general literature, a site is classified as a target if the author of the published paper considered tumors to be induced by compound administration ("+" on the plot). In some cases authors do not clearly state their evaluation (blank in author's opinion in plot), and in some NCIINTP technical reports the evidence for carcinogenicity is considered "associated" or "equivocal"; these are not classified as positive. We use the author's opinion to determine positivity because it often takes into account more information than statistical significance alone, such as historical control rates for particular sites, survival and latency, and/or dose response. Generally, this

designation by author's opinion corresponds well with the results of statistical tests for the significance of the dose-response effect (two-tailed p < 0.01). For some chemicals, the only experiments in the CPDB for a species or a sex-species group were NCIINTP bioassays that were evaluated as inadequate, and we indicate these with an "I," in the potency and target organ fields. Carcinogenic Potency In the CPDB, a standardized quantitative measure of carcinogenic potency, the TD50, is estimated for each set of tumor incidence data. In a simplified way, the TD50 may be defined as follows: For a given target site(s), if there are no tumors in control animals, then the TD50 is that chronic dose rate (in mg/kg body weight/day) that would induce tumors in half the test animals at the end of a standard life span for the species. Because the tumor(s) of interest often does occur in control animals, the TD50 is more precisely defined as that dose rate (in mg/kg body weight/day) that, if administered chronically for the standard life span of the species, will halve the probability of remaining tumorless throughout that period. The TD50 is analogous to the LD50, and a low TD50 value indicates a potent carcinogen, whereas a high value indicates a weak one. The TD50 and the statistical procedures adopted for estimating it from experimental data have been described elsewhere (Gold et aI., 1997c; Peto et aI., 1984; Sawyer et aI., 1984). The range of TD50 across chemicals in the CPDB is at least lO millionfold for carcinogens in each sex of rat or mouse.

References

In Table 38.12, a carcinogenic potency value is reported for a chemical in each species with a positive evaluation of carcinogenicity in at least one test. If there is only one positive test on the chemical in the species, then the most potent TDso value from that test is reported. When more than one experiment is positive, in order to use all the available data, the reported potency value is a harmonic mean of the most potent TDso values from each positive experiment. We have shown that the harmonic mean is similar to the most potent TDso value for chemicals with more than one positive test (Gold et aI., 1989, 1997b). The harmonic mean (TH) is defined as TH =

ACKNOWLEDGMENTS We thank the many people who have worked on the analyses discussed in this chapter. Several collaborators were authors on work that has been updated in this chapter, including, Leslie Bernstein, David Freedman, David Gaylor, Bonnie R. Stem, Joseph P. Brown, Georganne Backrnan Garfinkel, Lars Rohrbach, and Estie Hudes. This work was supported through the University of California, Berkeley by National Institute of Environmental Health Sciences Center Grant ESOl896 (BNA and LSG), and by support for research in disease prevention from the Dean's Office of the College of Letters and Science (LSG); and by U.S. Department of Energy Grant DE-AC-03-76SF00098 through the E.O. Lawrence Berkeley National Laboratory (LSG).

1 --c--

1 n

;; L

i=l

1

REFERENCES

T,. I

To obtain the harmonic mean from each positive experiment, we select the lowest TDso value from among positively evaluated target sites with a statistically significant dose response (two-tailed p < 0.1). If no positive sites have a significant dose response, then we select the most potent (lowest TDso) from among positively evaluated sites with p :::: 0.1. When some experiments have positive significant results and others have only positive nonsignificant results, we discard the nonsignificant experimental results for the calculation of the harmonic mean. In some experiments, no TDso could be estimated because all dosed animals had the tumor of interest, and only summary data were available for animals with the tumor. For these cases, we use the 99% upper confidence limit of TDso as a replacement for the TDso. In a series of superscripts following the TDso value, we provide additional information about the carcinogenic potency and other test results in the CPDB. These are as follows: i = carcinogenic in rodents only by the inhalation route of administration. m = more than one positive test in the species in the CPDB. n

835

= no

results that were evaluated as carcinogenic by the published author for this species in the CPDB have statistically significant TDso values (two-tailed p < 0.1). s = species other than rats or mice are reported for this chemical in Table 38.13. v = variation is greater than lO-fold among statistically significant (two-tailed p < 0.1) TDso values from different positive experiments.

Target Sites Target sites are reported for each sex-species group with a positive result in the CPDB. Target sites are identified on the basis of an author's positive opinion for the particular site, in any experiment in the sex-species, using all results from both the general literature and the NCIINTP bioassays. Hence, if a chemical has two target sites listed in a sex-species, the results may represent two different experiments. Occasionally, the CPDB results are only for both sexes combined and this has been indicated with (B) next to the target site.

Adachi, Y., Moore, L. E., Bradford, B. U., Gao, w., and Thurman, R. G. (1995). Antibiotics prevent liver injury in rats following long-term exposure to ethanol. Gastroenterology 108, 218-224. Adamson, R. H., Shozo, T., Sugimura, T., and Thorgeirsson, U. P. (1994). Induction of hepatocellular carcinoma in nonhuman primates by the food mutagen 2-amino-3-methylimidazo[4,S-flquinoline. Environ. Health Perspect. 102, 190-193. Ahmed, S. S., and Miiller, K. (1978). Effect of wound-damages on the glycoalkaloid content in potato tubers and chips. Lebensm.-Wiss. Technol. 11, 144-146. American Cancer Society (2000). "Cancer Facts and Figures-2000." American Cancer Society, Atlanta. American Medical Association (AMA) Division of Drugs (1983). "AMA Drug Evaluations," 5th ed., pp. 201-202. AMA, Chicago. American Water Works Association (AWWA) (1993). "DisinfectantlDisinfection By-Products Database for the Negotiated Regulation." AWWA, Washington, DC. Ames, B. N. (1983). Dietary carcinogens and anti-carcinogens: Oxygen radicals and degenerative diseases. Science 221, 1256-1264. Ames, B. N. (1984). Cancer and diet. Science 224, 668-670, 757-760. Ames, B. N. (2001). DNA damage from micronutrient deficiency is likely to be a major cause of cancer. Mutat. Res. 475, 7-20. Ames, B. N., and Gold, L. S. (1990). Chemical carcinogenesis: Too many rodent carcinogens. Proc. Natl. Acad. Sci. U.S.A. 87, 7772-7776. Available at http://socrates.berkeley.edu/mutagenlpnasl.html. Ames, B. N., and Gold, L. S. (1991). Risk assessment of pesticides. Chem. Eng. News 69, 28-32, 48-49; Forum: 27-55. Ames, B. N., Gold, L. S., and Shigenaga, M. K. (1996). Cancer prevention, rodent high-dose cancer tests, and risk assessment. Risk Anal. 16,613-617. Available at http://potency.berkeley.edu/textlriskanaleditorial.html. Ames, B. N., Gold, L. S., and Willett, W. C. (1995). The causes and prevention of cancer. Proc. Natl. Acad. Sci. U.S.A. 92, S2S8-526S. Available at http://socrates.berkeley.edulmutagenlames. pnas3.html. Ames, B. N., Magaw, R., and Gold, L. S. (1987). Ranking possible carcinogenic hazards. Science 236, 271-280. Letters: 237,235 (1987); 237, 1283-1284 (1987); 237, 1399-1400 (1987); 238, 1633-1634 (1987); 240, 1043-1047 (1988). Ames, B. N., Profet, M., and Gold, L. S. (I 990a). Dietary pesticides (99.99% all natural). Proc. Natl. Acad. Sci. U.s.A. 87, 7777-7781. Available at http://socrates.berkeley.edulmutagenlpnas2.html. Ames, B. N., Profet, M., and Gold, L. S. (l990b). Nature's chemicals and synthetic chemicals: Comparative toxicology. Proc. Natl. Acad. Sci. U.S.A. 87, 7782-7786. Available at http://socrates.berkeley.edu/mutagenlpnas3.html. Ames, B. N., Shigenaga, M. K., and Gold, L. S. (1993a). DNA lesions, inducible DNA repair, and cell division: Three key factors in mutagenesis and carcinogenesis. Environ. Health Perspect. 101, 35-44. Ames, B. N., Shigenaga, M. K., and Hagen, T. M. (1993b). Oxidants, antioxidants, and the degenerative diseases of aging. Proc. Natl. Acad. Sci. U.s.A. 90,7915-7922.

836

CHAPTER 38


Andrasik, J., and Cloutet, D. (1990). Monitoring solvent vapors in drycleaning plants.Int. Fabricare Inst. Focus Dry Cleaning 14, 1-8. Ariza, R. R., Dorado, G., Barbanch, M., aud Pueyo, C. (1988). Study of the causes of direct-acting mutagenicity in coffee aud tea using the Ara test in Salmonella typhimurium. Mutat. Res. 201, 89-96. Arkima, V. (1968). Die quautitative gaschromatographische Bestimmung der hoheren aliphatischen und aromatischen Alkohole im Bier. Mschr. Brauerei 21,25-27. Baltes, W. (1977). Rosteffekte auf die Kaffeezusammensetzung. Colloque Scientijique International sur le Cafe 8, 85-96. Beier, R. C., aud Nigg, H. N. (1994). Toxicology of naturally occurring chemicals in food. In "Foodborne Disease Haudbook" (Y. H. Hui, J. R. Gorham, K. D. MurreIl, and D. O. CIiver, eds.), Vol. 3, pp. 1-186. Dekker, New York. Beier, R. c., Ivie, G. w., Oertli, E. H., aud Holt, D. L. (1983). HPLC aualysis of linear furocoumarins (psoralens) in healthy celery Apium graveolens. Food Chem. Toxicol. 21, 163-165. Bejnarowicz, E. A., and Kirch, E. R. (1963). Gas chromatographic aualysis of oil of nutmeg. J. Pharm. Sci. 52, 988-993. Berkley, S. F., Hightower, A. w., Beier, R. c., Fleming, D. w., Brokopp, C. D., Ivie, G. W., and Broome, C. V. (1986). Dermatitis in grocery workers associated with high natural concentrations of furanocoumarins in celery. Ann. Int. Med. 105,351-355. Bernhard, R. A. (1960). Analysis aud composition of oil of lemon by gas-liquid chromatography. J. Chromatogr. 3,471-476. Bernstein, L., Gold, L. S., Ames, B. N., Pike, M. C., aud Hoel, D. G. (1985). Some tautologous aspects of the comparison of carcinogenic potency in rats and mice. Fundam. Appl. Toxicol. 5, 79-86. Bernstein, L., Ross, R. K., aud Pike, M. C. (1990). Hormone levels in older women: A study of postmenopausal breast cancer patients aud healthy population controls. Br. J. Cancer 61, 298-302. Billing, J., and Shermau, P. W. (1998). Antimicrobial functions of spices: Why some like it hot. Q. Rev. BioI. 73,3-49. Birnbaum, L. S. (1994). The mechauism of dioxin toxicity: Relationship to risk assessment. Environ. Health Perspect. 102, 157-167. Blaauboer, B. J., Balls, M., Barratt, M., Casati, S., Coecke, S., Mohamed, M. K., Moore, J., RaIl, D., Smith, K. R., Tennant, R., Schwetz, B. A., Stokes, W. S., and Younes, M. (1998). 13th meeting of the scientific group on methodologies for the safety evaluation of chemicals (SGOMSEC): Alternative testing methodologies and conceptual issues. Environ. Health Perspect. 106,413418. Blair, A., Stewart, P. A., Zaebst, D. D., Pottern, L., Zey, J. N., Bloom, T. F., Miller, B., Ward, E., and Lubin, J. (1998). Mortality of industrial workers exposed to acrylonitrile. Scand. J. Work Environ. Health 24, 25-41. Blauch, J. L., and Tarka, S. M., Jr. (1983). HPLC determination of caffeine aud theobromine in coffee, tea, and instant hot cocoa mixes. J. Food Sci. 48, 745-747,750. Block, G., Patterson, B., and Subar, A. (1992). Fruit, vegetables aud cancer prevention: A review of the epidemologic evidence. Nutr. Cancer 18, 1-29. Blount, B. C., Mack, M. M., Wehr, C. M., MacGregor, J. T., Hiatt, R. A., Wang, G., Wickramasinghe, S. N., Everson, R. B., aud Ames, B. N. (1997). Folate deficiency causes uracil misincorporation into human DNA and chromosome breakage: Implications for caucer aud neuronal damage. Proc. Natl. Acad. Sci. U.S.A. 94, 3290-3295. Bogen, K. T., aud Gold, L. S. (1997). Trichloroethylene cancer risk: Simplified calculation of PBPK-based MCLs for cytotoxic endpoints. Regul. Toxicol. Pharmacol. 25, 26-42. Bokanga, E., Essers, A. J. A., Poulter, N., Rosling, H., and Tewe, O. (eds.) (1994). International Workshop on Cassava Safety. International Society for Horticultural Science, Wageningen, Netherlauds. Acta Horticulturae 375. Botterweck, A. A. M., Verhagen, H., Goldbohm, R. A., Kleinjaus, J., and van den Brandt, P. A. (2000). Intake of butylated hydroxyanisole aud butylated hydroxy toluene and stomach cancer risk: Results from aualyses in the Netherlands cohort study. Food Chem. Toxicol. 38, 599-605. Bradfield, C. A., and Bjeldanes, L. F. (1987). Structure-activity relationships of dietary indoles: A proposed mechanism of action as modifiers of xenobiotic metabolism. 1. Toxicol. Environ. Health 21,311-323.

Bunker, M. L., and McWiIIiams, M. (1979). Caffeine content of common beverages. J. Am. Diet. Assoc. 74,28-32. Burdock, G. A. (2000). Dietary supplements and lessons to be learned from GRAS. Regul. Toxicol. Pharmacol. 31, 68-76. Burnam, W. L. (2000). "Office of Pesticide, Programs List of Chemicals Evaluated for Carcinogenic Potential." Memoraudum, Office of Prevention, Pesticides aud Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. Bushway, R. 1., and Ponnampalam, R. (1981). a-Chaconine and a-solauine content of potato products aud their stability during several modes of cooking. J. Agric. Food Chem. 29, 814-817. Butterworth, B. E., and Bogdanffy, M. S. (1999). A comprehensive approach for integration of toxicity aud caucer risk assessments. Regul. Toxicol. Pharmacol. 29,23-36. Butterworth, B. E., ConnoIly, R. B., and Morgan, K. T. (1995). A strategy forestablishing mode of action of chemical carcinogens as a guide for approaches to risk assessments. Cancer Lett. 93, 129-146. Buttery, R. G., Stern, D. J., aud Ling, L. C. (1994). Studies on flavor volatiles of some sweet corn products. J. Agric. Food Chem. 42, 791-795. California Environmental Protection Agency, Staudards and Criteria Work Group (1994). "California Cancer Potency Factors: Update." California Environmental Protection Agency, Sacramento. Canas, B. J., Havery, D. c., Robinson, L. R., SuIIivan, M. P., Joe, F. L., Jr., and Diachenko, G. W. (1989). Chemical contaminauts monitoring: Ethyl carbamate levels in selected fermented foods and beverages. J.-Assoc. Off. Anal. Chem. 72, 873-876. Carlson, D. G., Daxenbichler, M. E., VauEtten, C. H., Kwolek, W. F., and WiIIiams, P. H. (1987). Glucosinolates in crucifer vegetables: Broccoli, Brussels sprouts, cauliflower, coIlards, kale, mustard greens, and kohlrabi. J. Am. Soc. Hortic. Sci. 112, 173-178. Chaisson, C. F., Petersen, B. J., Eickhoff, J. C., aud Slesinski, R. S. (1989). "Pesticides in Our Food: Facts, Issues, Debates, Perceptions." Technical Assessment Systems, Washington, DC. Chaudhary, S. K., Ceska, 0., Tetu, C., Warrington, P. J., Ashwood-Smith, M. J., aud Poulton, G. A. (1986). Oxypeucedauin, a major furocoumarin in parsley, Petroselinum crispum. Planta Med. 6, 462-464. Chauhan, Y., Nagel, D., Gross, M., Cerny, R., aud Toth, B. (1985). Isolation of N2-[y-L-(+)-glutamyl]-4-carboxyphenylhydrazine in the cultivated mushroom Agaricus bisporus. J. Agric. Food Chem. 33, 817-820. Christensen, J. G., Goldsworthy, T. L., aud Cattley, R. C. (1999). Dysregulation of apoptosis by c-myc in transgenic hepatocytes aud effects of growth factors aud nongenotoxic carcinogens. Mol. Carcinogen. 25, 273-284. Clarke, R. J., and Macrae, R. (eds.) (1988). "Coffee," Vols. 1-3. Elsevier, New York. Clayson, D. B., and Iverson, F. (1996). Cancer risk assessment at the crossroads: The need to turn to a biological approach. Regul. Toxicol. Pharmacol. 24, 45-59. Clayson, D. B., Iverson, F., Nera, E. A., aud Lok, E. (1990). The significauce of induced forestomach tumors. Annu. Rev. Pharmacol. Toxicol. 30,441-463. Clinton, W. P. (1986). The chemistry of coffee. Colloque Scientijique International sur le Cafe 11, 87-92. Clydesdale, F. M. (ed.) (1997). "Food Additives: Toxicology, Regulation, and Properties." CRC Press, Boca Raton, FL. Cohen, S. M. (l995a). Role of urinary physiology and chemistry in bladder carcinogenesis. Food Chem. Toxicol. 33,715-730. Cohen, S. M. (1995b). Humau relevance of auimal carcinogenicity studies. Regul. Toxicol. Pharmacol. 21, 75-80. Cohen, S. M. (1998). Cell proliferation and carcinogenesis. Drug Metab. Rev. 30, 339-357. Cohen, S. M., aud EIlwein, L. B. (1991). Genetic errors, cell proliferation, and carcinogenesis. Cancer Res. 51, 6493-6505. Cohen, S. M., aud Lawson, T. A. (1995). Rodent bladder tumors do not always predict for humans. Cancer Lett. 93, 9-16. Connor, T. H., Theiss, J. C., Hauna, H. A., Monteith, D. K., and Matney, T. S. (1985). Genotoxicity of organic chemicals frequently found in the air of mobile homes. Toxicol. Lett. 25, 33-40.

References

CONSAD Research Corporation (1990). "Economic Analysis of OSHA's Proposed Standards for Methylene Chloride." Final Report, OSHA Docket H-71, CONS AD Research Corporation, Pittsburgh. Contrera, 1., lacobs, A., and DeGeorge, 1. (1997). Carcinogenicity testing and the evaluation of regulatory requirements for pharmaceuticals. Regul. Taxieal. Pharmacal. 25: 130--145. Counts, 1. L., and Goodman, 1. I. (1995). Principles underlying dose selection for, and extrapolation from the carcinogen bioassay: Dose influences mechanism. Regul. Taxieal. Pharmacal. 21, 418--421. Coxon, D. T, Curtis, R E, and Howard, B. (1975). Ipomeamarone, a toxic furanoterpenoid in sweet potatoes (Ipamea batatas) in the United Kingdom. Faad Casmet. Taxieol. 13, 87-90. Crosby, D. G., and Aharonson, N. (1967). The structure of carotatoxin, a natural toxicant from carrot. Tetrahedron 23, 465--472. Crump, K. S. (1984). An improved procedure for low-dose carcinogenic risk assessment from animal data. J. Enviran. Pathal. Taxieal. Oneal. 5, 339348. Crump & Assoc., Tox-Risk Computer Program, Version 3.0. Clement International Corp., Ruston, LA. Culvenor, C. C. 1., Clarke, M., Edgar, 1. A., Frahn, 1. L., lago, M. v., Peterson, 1. E., and Smith, L. W (1980). Structure and toxicity of the alkaloids of Russian comfrey (Symphytum x Uplandicum nyman), a medicinal herb and item of human diet. Experientia 36, 377-379. Cunningham, M. L. (1996). Role of increased DNA replication in the carcinogenic risk of nonmutagenic chemical carcinogens. Mutat. Res. 365, 59-69. Cunningham, M. L., and Matthews, H. B. (1991). Relationship of hepatocarcinogenicity and hepatocellular proliferation induced by mutagenic noncarcinogens vs. carcinogens. n. 1- vs. 2-nitropropane. Taxieol. Appl. Pharmacal. 110,505-513. Cunningham, M. L., Foley, 1., Maronpot, R R, and Matthews, H. B. (1991). Correlation of hepatocellular proliferation with hepatocarcinogenicity induced by the mutagenic noncarcinogen : carcinogen pair-2,6- and 2,4diaminotoluene. Taxical. Appl. Pharmacal. 107,562-567. Cunningham, M. L., Pippin, L. L., Anderson, N. L., and Wenk, M. L. (1995). The hepatocarcinogen methapyrilene but not the analog pyrilamine induces sustained hepatocellular replication and protein alterations in F344 rats in a 13-week feed study. Taxicol. Appl. Pharmacal. 131,216-223. Czaja, M. 1., Xu, 1., lu, Y., Alt, E., and Schmiedeberg, P. (1994). Lipopolysaccharide-neutralizing antibody reduces hepatocyte injury from acute hepatotoxin administration. Hepatalagy 19, 1282-1289. Dashwood, R. H. (1998). Indole-3-carbinol: Anticarcinogen or tumor promoter in Brassica vegetables? Chem.-Bial. Interact. 110, 1-5. Davies, T. S., and Momo, A. (1995). Marketed human pharmaceuticals reported to be tumorigenic in rodents. J. Am. Call. Taxieal. 14,90--107. Dietrich, D. R, and Swenberg, 1. A. (1991). The presence of £¥2u-globulin is necessary for d-limonene promotion of male rat kidney tumors. Cancer Res. 51,3512-3521. Duggan, R. E., and Corneliussen, P. E. (1972). Dietary intake of pesticide chemicals in the United States (Ill), lune 1968-April 1970. Pest. Manit. J. 5, 331-341. Economic Research Service (ERS) (1994). "Vegetables and Specialties Situation and Outlook Yearbook." U.S. Department of Agriculture, Washington, DC. Economic Research Service (ERS) (1995). "Fruit and Tree Nuts Situation and Outlook Yearbook." U.S. Department of Agriculture, Washington, DC. Ehlers, D., Kirchhoff, 1., Gerard, D., and Quirin, K. (1998). High-performance liquid chromatography analysis of nutmeg and mace oils produced by supercritical C02 extraction--comparison with steam-distilled oilscomparison of East Indian, West Indian and Papuan oils. Int. J. Faad Sci. Technal. 33,215-223. Eiden, C. (1999). "Human Health Risk Assessment: Azinphos-Methyl." U.S. Environmental Protection Agency, Washington, DC. Eilrich, G. L. (1991). Tracking the fate of residues from the farm gate to the table. In "Pesticide Residues and Food Safety: A Harvest of Viewpoints" (B. G. Tweedy, H. 1. Dishburger, L. G. Ballantine, 1. McCarthy, and 1. Murphy, eds.), pp. 202-212. Am. Chem. Soc., Washington, DC.

837

Eltayeb, E. A., and Roddick, 1. G. (1984). Changes in the alkaloid content of developing fruits of tomato (Lyeapersican esculentum Mill.). 1. Analyses of cultivars and mutants with different ripening characteristics. J. Exp. Bat. 35, 252-260. Engel, K. H., and Tressl, R (1983). Studies on the volatile components of two mango varieties. J. Agric. Faad Chem. 31,796-801. Fazio, T, Havery, D. c., and Howard, 1. W (1980). Determination of volatile N -nitrosamines in foodstuffs. I. A new clean-up technique for confirmation by GLC-MS. n. A continued survey of foods and beverages. In "NNitroso Compounds: Analysis, Formation and Occurrence" (E. A. Walker, L. Griciute, M. Castegnaro, and M. Borzsonyi, eds.), pp. 419-435. IARC Scientific Publication 31, International Agency for Research on Cancer, Lyon. Fenech, M., Aitken, C., and Rinaldi, 1. (1998). Folate, vitamin B12, homocysteine status and DNA damage in young Australian adults. Carcinagenesis 19, 1163-1171. Fernandez-Salguero, P. M., Hilbert, D. M., Rudikoff, S., Ward, 1. M., and Gonzalez, E 1. (1996). Aryl-hydrocarbon receptor-deficient mice are resistant to 2,3,7,8-tetrachlorodibenzo-p-dioxin-induced toxicity. Taxieol. Appl. Pharmacal. 140, 173-179. Freedman, D. A., Gold, L. S., and Slone, T. H. (1993). How tautological are inter-species correlations of carcinogenic potency? Risk Anal. 13, 265-272. Friedman, M., and Dao, L. (1990). Distribution of glycoalkaloids in potato plants and commercial potato products. J. Agrie. Faad Chem. 40,419-423. Fujita, Y., Wakabayashi, K., Nagao, M., and Sugimura, T (1985). Implication of hydrogen peroxide in the mutagenicity of coffee. Mutat. Res. 144, 227230. Fung, V. A., Cameron, T P., Hughes, T 1., Kirby, P. E., and Dunkel, V. C. (1988). Mutagenic activity of some coffee flavor ingredients. Mutat. Res. 204,219-228. Galasko, G. T. E, Furman, K. I., and Alberts, E. (1989). The caffeine content of non-alcoholic beverages. Faad Chem. Taxical. 27,49-51. Gartrell, M. 1., Craun, 1. c., Podrebarac, D. S., and Gunderson, E. L. (1986). Pesticides, selected elements, and other chemicals in adult total diet samples, October 1980--March 1982. J. Assac. Off. Anal. Chem. 69, 146-161. Gaylor, D. W, and Gold, L. S. (1995). Quick estimate of the regulatory virtually safe dose based on the maximum tolerated dose for rodent bioassays. Regul. Taxieol. Pharmacal. 22, 57-63. Gaylor, D. W., and Gold, L. S. (1998). Regulatory cancer risk assessment based an a quick estimate of a benchmark dose derived from the maximum tolerated dose. Regul. Taxicol. Pharmacal. 28, 222-225. Gaylor, D. W., Chen, 1. 1., and Sheehan, D. M. (1993). Uncertainty in cancer risk estimates. Risk Anal. 13, 149-154. Gloria, M. B. A., Barbour, 1. E, and Scanlan, R A. (1997). N-Nitrosodimethylamine in Brazilian, U.S. domestic, and U.S. imported beers. J. Agric. Faad Chem. 45,814-816. Gold, L. S., and Zeiger, E. (eds.) (1997). "Handbook of Carcinogenic Potency and Genotoxicity Databases." CRC Press, Boca Raton, PL. Gold, L. S., Garfinkel, G. B., and Slone, T. H. (1994a). Setting priorities among possible carcinogenic hazards in the workplace. In "Chemical Risk Assessment and Occupational Health: Current Applications, Limitations, and Future Prospects" (C. M. Smith, D. C. Christiani, and K. T. Kelsey, eds.), pp. 91-103. Auburn House, Westport, CT. Gold, L. S., Manley, N. B., Slone, T. H., and Rohrbach, L. (1999). Supplement to the Carcinogenic Potency Database (CPDB): Results of animal bioassays published in the general literature in 1993 to 1994 and by the National Toxicology Program in 1995 to 1996. Environ. Health Perspeet. 107,527-600. Available at http://ehpnet1.niehs.nih.gov/docs/1999/suppl-4/toc.html. Gold, L. S., Slone, T. H., and Ames, B. N. (1997a). Prioritization of possible carcinogenic hazards in food. In "Food Chemical Risk Analysis" (D. R Tennant, ed.), pp. 267-295. Chapman & Hall, London. Available at http://potency.berkeley.edultextlmaff.html. Gold, L. S., Slone, T H., and Ames, B. N. (l997b). Overview of analyses of the Carcinogenic Potency Database. In "Handbook of Carcinogenic Potency and Genotoxicity Databases" (L. S. Gold, and E. Zeiger, eds.), pp. 661685. CRC Press, Boca Raton, PL.

838

CHAPTER 38


Gold, L. S., Slone, T. H., and Ames, B. N. (1998). What do animal cancer tests tell us about human cancer risk? Overview of analyses of the Carcinogenic Potency Database. Drug Metab. Rev. 30, 359-404. Available at http://potency.berkeley.edultextldrugmetrev.html. Gold, L. S., Slone, T. H., Ames, B. N., Manley, N. B., Garfinkel, G. B., and Rohrbach, L. (1997c). Carcinogenic Potency Database. In "Handbook of Carcinogenic Potency and Genotoxicity Databases" (L. S. Gold, and E. Zeiger, eds.), pp. 1-605. CRC Press, Boca Raton, FL. Gold, L. S., Slone, T. H., and Bernstein, L. (1989). Summary of carcinogenic potency (TDso) and positivity for 492 rodent carcinogens in the Carcinogenic Potency Database. Environ. Health Perspect. 79,259-272. Gold, L. S., Slone, T. H., Manley, N. B., and Ames, B. N. (1994b). Heterocyclic amines formed by cooking food: Comparison of bioassay results with other chemicals in the Carcinogenic Potency Database. Cancer Let!. 83,21-29. Gold, L. S., Slone, T. H., Stem, B. R., and Bernstein, L. (1993). Comparison of target organs of carcinogenicity for mutagenic and non-mutagenic chemicals. Mutat. Res. 286, 75-100. Gold, L. S., Slone, T. H., Stem, B. R., Manley, N. B., and Ames, B. N. (1992). Rodent carcinogens: Setting priorities. Science 258, 261-265. Available at http://potency.berkeley.edultextlscience.html. Gold, L. S., Wright, c., Bernstein, L., and de Veciana, M. (1987a). Reproducibility of results in "near-replicate" carcinogenesis bioassays. J. Natl. Cancer Inst. 78, 1149-1158. Gold, L. S., Backman, G. M., Hooper, N. K, and Peto, R. (1987b). Ranking the potential carcinogenic hazards to workers from exposures to chemicals that are tumorigenic to rodents. Environ. Health Perspect. 76,211-219. Goodman, D. G., and Sauer, R. M. (1992). Hepatotoxicity and carcinogenicity in female Sprague-Dawley rats treated with 2,3,7,8-tetrachlorodibenzo-pdioxin (TCDD): A pathology working group reevaluation. Regul. Toxicol. Pharmacol. 15, 245-252. Goodman, J. I. (1994). A rational approach to risk assessment requires the use of biological information: An analysis of the National Toxicology Program (NTP), final report of the advisory review by the NTP Board of Scientific Counselors. Regul. Toxicol. Pharmacol. 19,51-59. Gribble, G. W. (1996). The diversity of natural organochlorines in living organisms. Pure Appl. Chem. 68, 1699-1712. Groisser, D. S. (1978). A study of caffeine in tea. I. A new spectrophotometric micro-method. Il. Concentration of caffeine in various strengths, brands, blends, and types of teas. Am. J. Clin. Nutr. 31, 1727-1731. Groopman, J. D., Zhu, J. Q., Donahue, P. R., Pikul, A., Zhang, L. S., Chen, J. S., and Wogan, G. N. (1992). Molecular dosimetry of urinary aflatoxin-DNA adducts in people living in Guangxi Autonomous Region, People's Republic of China. Cancer Res. 52, 45-52. Gruenwald, J., Brendler, T., and Jaenicke, C. (1998). "PDR for Herbal Medicines." Medical Economics Company, Montvale, NJ. Gunawardhana, L., Mobley, S. A., and Sipes, I. G. (1993). Modulation of 1,2dichlorobenzene hepatotoxicity in the Fischer-344 rat by a scavenger of superoxide anions and an inhibitor of Kupffer cells. Toxicol. Appl. Pharmacol. 119,205-213. Gunderson, E. L. (1988). Chemical contaminants monitoring: FDA Total Diet Study, April 1982-April 1984, dietary intakes of pesticides, selected elements, and other chemicals. J. Assoc. Off. Anal. Chem. 71, 1200-1209. Gunderson, E. L. (1992). "Alphabetical Listing of Organic Pesticide Residues and Industrial Chemicals Detected by the Total Diet Study." U.S. Environmental Protection Agency, Washington, DC. Gunderson, E. L. (1995). Dietary intakes of pesticides, selected elements, and other chemicals: FDA Total Diet Study, June 1984-April 1986. J.-Assoc. Off. Anal. Chem. 78,910-921. Giinther, H. (1968). Untersuchungen an Cintronenolen mit Hilfe der Gaschromatographie und der Infrarotspektroskopie. Dtsch. Lebensm.-Rundsch. 4, 104-111. Hall, R. L., Henry, S. H., Scheuplein, R. J., Dull, B. J., and Rulis, A. M. (1989). Comparison of the carcinogenic risks of naturally occurring and adventitious substances in food. In "Food Toxicology: A Perspective on the Relative Risks" (S. L. Taylor, and R. A. Scanlan, eds.), pp. 205-224. Dekker, New York.

Hanham, A. E, Dunn, B. P., and Stich, H. E (1983). Clastogenic activity of caffeic acid and its relationship to hydrogen peroxide generated during autooxidation. Mutat. Res. 116, 333-339. Hard, G. c., and Whysner, J. (1994). Risk assessment of d-limonene: An example of male rat-specific renal tumorigens. Crit. Rev. Toxicol. 24,231-254. Hart, R., Neumann, D., and Robertson, R. (1995). "Dietary Restriction: Implications for the Design and Interpretation of Toxicity and Carcinogenicity Studies." ILSI Press, Washington, DC. Harvey, M. H., Morris, B. A., McMillan, M., and Marks, V. (1985). Measurement of potato steroidal alkaloids in human serum and saliva by radioimmunoassay. Hum. Toxicol. 4,503-512. Hasselstrom, T., Hewitt, E. J., Konigsbacher, K. S., and Ritter, J. 1. (1957). Composition of volatile oil of black pepper. J. Agric. Food Chem. 5,53-55. Havel, R. J., and Kane, J. P. (1982). Therapy ofhyperlipidemic states. Ann. Rev. Med. 33,417. Hayashi, E, Tamura, H., Yamada, J., Kasai, H., and Suga, T. (1994). Characteristics of the hepatocarcinogenesis caused by dehydroepiandrosterone, a peroxisome proliferator, in maleF-344 rats. Carcinogenesis 15, 22152219. Hayward, J. J., Shane, B. S., Tindall, K R., and Cunningham, M. L. (1995). Differential in vivo mutagenicity of the carcinogen/non-carcinogen pair 2,4and 2,6-diaminotoluene. Carcinogenesis 16, 2429-2433. Hecht and Hoffmann (1998). Heddle, J. A. (1998). The role of proliferation in the origin of mutations in mammalian cells. Drug Metab. Rev. 30, 327-338. Heikes, D. L. (1994). SFE with GC and MS determination of safro le and related allylbenzenes in sassafras teas. J. Chromatogr. Sci. 32, 253-258. Heinrich, L., and Baltes, W. (1987). Uber die Bestimmung von Phenolen im Kaffeegetrank. Z. Lebensm. Unters.-Forsch. 185,362-365. Helbock, H. J., Beckman, K B., Shigenaga, M. K, WaIter, P. B., Woodall, A. A., Yeo, H. C., and Ames, B. N. (1998). DNA oxidation matters: The HPLC-electrochemical detection assay of 8-oxo-deoxyguanosine and 8-oxo-guanine. Proc. Natl. Acad. Sci. US.A. 95, 288-293. Herrmann, K (1978). Review on nonessential constituents of vegetables. Ill. Carrots, celery, parsnips, beets, spinach, lettuce, endives, chicory, rhubarb, and artichokes. Z. Lebensm. Unters.-Forsch. 167,262-273. Hill, L. L., Ouhtit, A., Loughlin, S. M., Kripke, M. L., Ananthaswamy, H. N., and Owen-Schaub, L. B. (1999). Fas ligand: A sensor for DNA damage critical in skin cancer etiology. Science 285, 898-900. Hill, M. J., Giacosa, A., and Caygill, C. P. J. (eds.) (1994). "Epidemiology of Diet and Cancer." ElIis Horwood, New York. Hirono, I., Mori, H., and Haga, M. (1978). Carcinogenic activity of Symphytum officinale. J. Natl. Cancer Inst. 61, 865-868. Hodgkinson, A. (1977). "Oxalic Acid in Biology and Medicine." Academic Press, New York. Hultin, H. 0., and Proctor, B. E. (1961). Changes in some volatile constituents of the banana during ripening, storage, and processing. Food Technol. 15,

440-444. Huxtable, R. (1995). Pyrrolizidine alkaloids: Fascinating plant poisons. Newsletter, Fall, pp. 1-3. Center for Toxicology, Southwest Environmental Health Sciences Center. Ikeda, R. M., Stanley, W. L., Rolle, L. A., and Vannier, S. H. (1962). Monoterpene hydrocarbon composition of citrus oils. J. Food Sci. 27,593-596. Innes, J. R. M., Ulland, B. M., Valerio, M. G., Petrucelli, L., Fishbein, L., Hart, E. R., Pallota, A. J., Bates, R. R., Falk, H. L., Gart, J. J., Klein, M., Mitchell, I., and Peters, J. (1969). Bioassay of pesticides and industrial chemicals for tumorigenicity in mice: A preliminary note. J Natl. Cancer Inst. 42, 1101-1114. International Agency for Research on Cancer (IARC) (1971-1999). IARC Monographs on the Evaluation of Carcinogenic Risk of Chemicals to Humans, Vols. 1-73. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (1991). "Coffee, Tea, Mate, Methylxanthines and Methylglyoxal." International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (IARC) (1993). "Some Naturally Occurring Substances: Food Items and Constituents, Heterocyclic Aromatic

References

Amines and Mycotoxins," IARC Monographs on the Evaluation of Carcinogenic Risk of Chemicals to Humans, Vo!. 56. International Agency for Research on Cancer, Lyon. International Agency for Research on Cancer (1997). "Polychlorinated Dibenzo-para-dioxins and Polychlorinated Dibenzofurans," IARC Monographs on the Evaluation of Carcinogenic Risk of Chemicals to Humans, Vo!. 69. International Agency for Research on Cancer, Lyon International Life Sciences Institute (1996). Occurrence and significance of ochratoxin A in food. ILSI Europe Workshop, January 10--12, 1996, Aixen-Provence, France. ILSI Europe Newsletter, February 1996, p. 3. Irene, S. R. (1995). "List of Chemicals Evaluated for Carcinogenic Potentia!." Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. Ishidate, M., Jr., Hamois, M. C., and Sofuni, T. (1988). A comparative analysis of data on the clastogenicity of 951 chemical substances tested in mammalian cell cultures. Mutat. Res. 195, 151-213. Ivie, G. W., Holt, D. L., and Ivey, M. (1981). Natural toxicants in human foods: Psoralens in raw and cooked parsnip root. Science 213, 909-910. Jurics, E. W. (1967). Zur Analaytik der in Friichten am haufigsten vorkommenden Hydroxyzimtsauren und Catechine. Ernahrungs-forschung 3, 427-433. Kasidas, G. P., and Rose, G. A. (1980). Oxalate content of some common foods: Determination by an enzymatic method. f. Hum. Nutr. 34, 255-266. Kazeniac, S. J., and Hall, R. M. (1970). Flavor chemistry of tomato volatiles. f. Food Sci. 35,519-530. Kelloff, G. J., Boone, C. W, Crowell, J. A., Steele, V. E., Lubet, R. A., Doody, L. A., Malone, W E, Hawk, E. T., and Sigman, C. C. (1996a). New agents for cancer chemoprevention. f. Cell Biochem. Supp!. 26, 1-28. Kelloff, G. J., Crowell, J. A., Hawk, E. T., Steele, V. E., Lubet, R. A., Boone, C. W., Covey, J. M., Doody, L. A., Omenn, G. S., Greenwald, P., Hong, W. K., Parkinson, D. R., Bagheri, D., Baxter, G. T., Blunden, M., Doeltz, M. K., Eisenhauer, K. M., Johnson, K., Knapp, G. G., Longfellow, G., Malone, W E, Nayfield, S. G., Seifried, H. E., Swall, L. M., and Sigman, C. C. (1996b). Strategy and planning for chemopreventive drug development: Clinical development plans n. 1. Cell Biochem. Supp!. 26, 54-71. Key, T., and Reeves, G. (1994). Organochlorines in the environment and breast cancer. Br. Med. f. 308, 1520--1521. Kier, L. E., Brusick, D. J., Auletta, A. E., Von Halle, E. S., Brown, M. M., Simmon, V. E, Dunkel, V., McCann, J., Mortelmans, K., Prival, M., Rao, T. K., and Ray, V. (1986). The Salmonella typhimuriumlmammalian microsomal assay: A report of the U.S. Environmental Protection Agency Gene-Tox Program. Mutat. Res. 168, 69-240. Kikugawa, K., Kato, T., and Takahashi, S. (1989). Possible presence of 2-amino-3,4-dimethylimidazo[4,5- flquinoline and other heterocyclic amine-like mutagens in roasted coffee beans. f. Agric. Food Chem. 37, 881-886. Kirchner, J. G., and Miller, J. M. (1957). Volatile water-soluble and oil constituents of Valencia orange juice. f. Agric. Food Chem. 5, 283-291. Kirchner, J. G., Miller, J. M., Rice, R. G., Keller, G. J., and Fox, M. M. (1953). Volatile water-soluble constituents of grapefruit juice. f. Agric. Food Chem. 1,510--512. Kitamura, S., Koga, H., Tatsumi, K., Yoshimura, H., and Horiuchi, T. (1978). Relationship between biological activities and enzymatic reduction of nitrofuran derivatives. f. Pharm. Dyn. 1, 15-21. Knize, M. G., Dolbeare, EA., Carroll, K. L., Moore, D. H., n, and Felton, J. S. (1994). Effect of cooking time and temperature on the heterocyclic amine content of fried beef patties. Food Chem. Toxico!. 32, 595--603. Kohman, E. E (1939). Oxalic acid in foods and its behavior and fate in the diet. f. Nutr. 18, 233-246. Krebs-Smith, S. M., Cook, A., Subar, A. E, Cleveland, L., and Friday, J. (1995). US adults' fruit and vegetable intakes, 1989 to 1991: A revised baseline for the Healthy People 2000 objective. Am. f. Public Health 85, 1623-1629. Krebs-Smith, S. M., Cook, A., Subar, A. E, Cleveland, L., Friday, J., and Kahle, L. L. (1996). Fruit and vegetable intakes of children and adolescents in the United States. Arch. Pediatr. Adolesc. Med. 150,81-86.

839

Krewski, D., Gaylor, D. W, Soms, A. P., and Szyszkowicz, M. (1993). An overview of the report-Correlation between Carcinogenic Potency and the Maximum Tolerated Dose: Implications for Risk Assessment. Risk Ana!. 13, 383-398. Krewski, D., Szyszkowicz, M., and Rosenkranz, H. (1990). Quantitative factors in chemical carcinogenesis: Variation in carcinogenic potency. Regu!. Toxico!. Pharmaco!' 12, 13-29. Kuiper-Goodman, T., and Scott, P. M. (1989). Risk assessment of the mycotoxin ochratoxin A. Biomed. Environ. Sci. 2, 179-248. Kung, J. T., McNaught, R. P., and Yeransian, J. A. (1967). Determining volatile acids in coffee beverages by NMR and gas chromatography. f. Food Sci. 32, 455-458. Laskin, D. L., and Pendino, K. J. (1995). Macrophages and inflammatory mediators in tissue injury. Annu. Rev. Pharmaco!' Toxico!. 35,655--677. Laskin, D. L., Robertson, EM., Pilaro, A. M., and Laskin, J. D. (1988). Activation of liver macrophages following phenobarbital treatment of rats. Hepatology 8, 1051-1055. Lee, C. Y., and Jaworski, A. (1987). Phenolic compounds in white grapes grown in New York. Am. f. Eno!. Vitic. 38,277-281. Lee, S. (1973). Better procedures for analyzing caffeine in tea could help establish standards for tea mixes. Tea Coffee Trade f. 144, 26-40. Lijinsky, W (1999). N-Nitroso compounds in the diet. Mutat. Res. 443, 129138. Lok, E., Scott, E W., Mongeau, R., Nera, E. A., Malcolm, S., and Clayson, D. B. (1990). Calorie restriction and cellular proliferation in various tissues of the female Swiss Webster mouse. Cancer Let!. 51,67-73. Lorenz, K., and Maga, J. (1972). Staling of white bread: Changes in carbonyl composition and GLC headspace profiles. f. Agric. Food Chem. 20,211213. Luckey, T. D. (1999). Nurture with ionizing radiation: A provocative hypothesis. Nutr. Cancer 34, 1-11. Lund, E. D., Kirkland, C. L., and Shaw, P. E. (1981). Methanol, ethanol, and acetaldehyde content of citrus products. f. Agric. Food Chem. 29, 361-366. Macaulay, T., Gallant, C. J., Hooper, S. N., and Chandler, R. E (1984). Caffeine content of herbal and fast-food beverages. f. Cancer Diet. Assoc. 45, 150156. Martinek, R. G., and Wolman, W. (1955). Xanthines, tannins, and sodium in coffee, tea, and cocoa. f. Am. Med. Assoc. 158, 1030--1031. Matanoski, G., Francis, M., Correa-Villasenor, A., Elliot, E., Santos-Brugoa, c., and Schwartz, L. (1993). Cancer epidemiology among styrenebutadiene rubber workers. IARC Sci. Pub!. 127, 363-374. Matsumoto, K., Ito, M., Yagyu, S., Ogino, H., and Hirono, 1. (1991). Carcinogenicity examination of Agaricus bisporus, edible mushroom, in rats. Cancer Let!. 58, 87-90. Mazza, G., and Oomah, B. D. (1995). Flaxseed, dietary fiber, and cyanogens. In "Flaxseed in Human Nutrition" (S. c. Cunnane, and L. U. Thompson, eds.), pp. 56--81. AOCS Press, Champaign, IL. McCann, J., Horn, L., Girman, J., and Nero, A. V. (1987). Potential risks from exposure to organic carcinogens in indoor air. In "Short-Term Bioassays in the Analysis of Complex Environmental Mixtures" (S. S. Sandhu, D. M. deMarini, M. J. Mass, M. M. Moore, and J. L. Mumford, eds.). Plenum, New York. McCarthy, J. E (1991). Average residues vs. tolerances: An overview of industry studies. In "Pesticide Residues and Food Safety" (B. G. Tweedy, H. J. Dishburger, L. G. Ballantine, J. McCarthy, and J. Murphy, eds.), pp. 182-191. Am. Chem. Soc., Washington, DC. McCormick, D. L., Rao, K. V. N., Johnson, W. D., Bowman-Gram, T. A., Steele, V. E., Lubet, R. A., and Kelloff, G. J. (1996). Exceptional chemopreventive activity of low-dose dehydroepiandrosterone in the rat mammary gland. Cancer Res. 56,1724-1726. McKone, T. E. (1997). Presidential Address: International Society of Exposure Analysis. f. Expo. Ana!. Environ. Epidemio!. 7, 403-408. Minyard, J. P., Jr., and Roberts, W E. (1991). FOODCONTAM: A state data resource on toxic chemicals in foods. In "Pesticide Residues and Food Safety" (B. G. Tweedy, H. J. Dishburger, L. G. Ballantine, J. McCarthy, and J. Murphy, eds.), pp. 151-161. Am. Chem. Soc., Washington, DC.

840

CHAPTER 38


Moller, B., and Herrmann, K. (1983). Quinic acid esters of hydroxycinnamic acids in stone and pome fruit. Phytochemistry 22, 477-481. Montgomery, R. D. (1964). Observations on the cyanide content and toxicity of tropical pulses. West Indian Med. 1. 13, 1-11. Morris, S. C., and Lee, T. H. (1984). The toxicity and teratogenicity of Solanaceae glycoalkaloids, particularly those of the potato (Solanum tuberosum): A review. Food Technol. Aust. 36, 118-124. Mosel, H. D., and Herrmann, K. (1974). The phenolics of fruits. Ill. The contents of catechins and hydroxycinnamic acids in pome and stone fruits. Z. Lebensm. Unters.-Forsch. 154,6-11. Munday, R., and Munday, C. M. (1999). Low doses of dially1 disu1fide, a compound derived from garlic, increase tissue activities of quinone reductase and glutathione transferase in the gastrointestinal tract of the rat. Nutr. Cancer 34, 42-48. Mussinan, C. J., Mookherjee, B. D., and Ma1colm, G. I. (1981). Isolation and identification of the volatile constituents of fresh lemon juice. In "Essential Oils" (B. D. Mookherjee, and C. J. Mussinan, eds.). Allured Publishing, Wheaton, IL. Nagata, T., and Sakai, S. (1985). Purine base pattern of Camellia irrawadiensis. Phytochemistry 24, 2271-2272. National Academy of Sciences (1970). "The Life Sciences: Recent Progress and Application to Human Affairs, the World of Biological Research, Requirement for the Future." Committee on Research in the Life Sciences, Washington, DC. National Academy of Sciences (1981). National Research Council (1979). "The 1977 Survey ofIndustry on the Use of Food Additives." Natl. Acad. Press, Washington, DC. National Research Council (1987). "Regulating Pesticides in Food: The Delaney Paradox." Natl. Acad. Press, Washington, DC. National Research Council (1994). "Science and Judgment in Risk Assessment." Committee on Risk Assessment of Hazardous Air Pollutants, Natl. Acad. Press, Washington, DC. National Research Council (1996). "Carcinogens and Anticarcinogens in the Human Diet: A Comparison of Naturally Occurring and Synthetic Substances." Natl. Acad. Press, Washington, DC. Nelson, P. E., and Hoff, J. E. (1969). Tomato volatiles: Effect of variety, processing and storage time. J. Food Sci. 34, 53-57. Neurath, G. B., Diinger, M., Pein, E G., Ambrosius, D., and Schreiber, O. (1977). Primary and secondary amines in the human environment. Food Cosmet. Toxicol. 15, 275-282. Nigg, H. N., Beier, R. c., Carter, 0., Chaisson, c., Franklin, c., Lavy, T., Lewis, R. G., Lombardo, P., McCarthy, J. E, Maddy, K. T., Moses, M., Norris, D., Peck, c., Skinner, R., and Tardiff, R. G. (1990). Exposure to pesticides. In "The Effects of Pesticides on Human Health" (S. R. Baker, and C. E Wilkinson, eds.), Vol. 18, pp. 35-104. Princeton Sci. Publ., Princeton, NJ. Nijssen, L. M., Visscher, C. A., Maarse, H., Willemsens, L. C., and Boelens, M. H. (eds.) (1996). "Volatile Compounds in Foods: Qualitative and Quantitative Data," 7th ed. TNO-CIVO Food Analysis Institute, Zeist, The Netherlands. Nisperos-Carriedo, M. 0., and Shaw, P. E. (1990). Comparison of volatile f1avor components in fresh and processed orange juices. 1. Agric. Food Chem. 38, 1048-1052. Oelkers, W. (1999). Dehydroepiandrosterone for adrenal insufficiency. N. Engl. 1. Med. 341, 1073-1074. Ogawa, K., Tsuda, H., Shirai, T., Ogiso, T., Wakabayashi, K., Dalgard, D. w., Thorgeirsson, U. P., Thorgeirsson, S. S., Adamson, R. H., and Sugimura, T. (1999). Lack of carcinogenicity of 2-amino-3,8-dimethylimidazo[4,5flquinoxaline (MeIQx) in cynomolgus monkeys. Ipn. 1. Cancer Res. 90, 622-628. Omenn, G. S., Stuebbe, S., and Lave, L. B. (1995). Predictions of rodent carcinogenicity testing results: Interpretation in light of the Lave-Omenn value-of-information model. Mol. Carcinogen. 14, 37-45. Opinion Research Corporation (1990). "Trends, Consumer Attitudes, and the Supermarket." Food Marketing Institute, Washington, DC.

Ott, M. G., Scharnweber, H. C., and Langner, R. R. (1980). Mortality experience of 161 employees exposed to ethylene dibromide in two production units. Br. 1. Ind. Med. 37, 163-168. Page, N. P., and Arthur, J. L. (1978). "Special Occupational Hazard Review of Trichloroethylene." National Institute for Occupational Safety and Health, Rockville, MD. Prakash, A. S., Pereira, T. N., Reilly, P. E. B., and Seawright, A. A. (1999). Pyrrolizidine alkaloids in human diet. Mutat. Res. 443, 53-57. Perez-Ilzarbe, J., Hernandez, T., and Estrella, I. (1991). Phenolic compounds in apples: Varietal differences Z. Lebensm. Unters.-Forsch. 192,551-554. Peto, R., Pike, M. C., Bernstein, L., Gold, L. S., and Ames, B. N. (1984). The TDSO: A proposed general convention for the numerical description of the carcinogenic potency of chemicals in chronic-exposure animal experiments. Environ. Health Perspect. 58, 1-8. Physicians' Desk Reference, 52nd ed. (1998). Medical Economics Company, Montvale, NJ. Pino, J., Rodriguez-Feo, G., Borges, P., and Rosado, A. (1990). Chemical and sensory properties of black pepper oil (Piper nigrum L.). Nahrung 34, 555560. Pino, J., Torricella, R., and Orsi, F. (1986). Correlation between sensory and gas-chromatographic measurements on grapefruit juice volatiles. Acta Alimentaria 15, 237-246. Pons, W. A., Jr. (1979). High pressure liquid chromatographic determination of aflatoxins in corn. l.-Assoc. Off. Anal. Chem. 62, 586-594. Poole, S. K., and Poole, C. E (1994). Thin-layer chromatographic method for the determination of the principal polar aromatic flavour compounds of the cinnamons of commerce. Analyst 119, 113-120. Postel, W., Drawert, E, and Adam, L. (1972). Gaschromatographische Bestimmung der Inhaltsstoffe von Garungsgetranken. Ill. Fliichtige Inhaltsstoffe des Weines. Chem. Mikrobiol. Technol. Lebensm. 1, 224-235. Poulton, J. E. (1983). Cyanogenic compounds in plants and their toxic effects. In "Handbook of Natural Toxins: Plant and Fungal Toxins" (R. E Keeler, and A. T. Tu, eds.), Vol. 1, pp. 117-157. Dekker, New York. Preston-Martin, S., Momoe, K., Lee, P. J., Bernstein, L., Kelsey, J., Henderson, S., Forrester, D., and Henderson, B. (1995). Spinal meningiomas in women in Los Angeles County: Investigation of an etiological hypothesis. Cancer Epidemiol. Biomarkers Prev. 4, 333-339. Preston-Martin, S., Pike, M. c., Ross, R. K., and Jones, P. A. (1990). Increased cell division as a cause of human cancer. Cancer Res. 50,7415-7421. Preussmann, R., and Eisenbrand, G. (1984). N-Nitroso carcinogens in the environment. In "Chemical Carcinogenesis" (C. E. Searle, ed.), 2nd ed., Vol. 2, pp. 829-868. Am. Chem. Soc., Washington, DC. Qian, G., Ross, R. K., Yu, M. C., Yuan, J., Henderson, B. E., Wogan, G. N., and Groopman, J. D. (1994). A follow-up study of urinary markers of aflatoxin exposure and liver cancer risk in Shanghai, People's Republic of China. Cancer Epidemiol. Biomarkers Prev. 3, 3-10. Quest, J. A., Fenner-Crisp, P. A., Burnam, W., Copley, M., Dearfield, K. L., Hamernik, K. L., Saunders, D. S., Whiting, R. J., and Engler, R. (1993). Evaluation of the carcinogenic potential of pesticides. 4. Chloroalkylthiodicarboximide compounds with fungicidal activity. Regul. Toxico!. Pharmacol. 17,19-34. Quest, J. A., Hamernik, K. L., Engler, R., Burnam, W. L., and FennerCrisp, P. A. (1991). Evaluation of the carcinogenic potential of pesticides. 3. Aliette. Regul. Toxicol. Pharmacol. 14,3-11. Rahn, w., and Konig, W. A. (1978). GCIMS investigations of the constituents in a diethyl ether extract of an acidified roast coffee infusion. 1. High Resolut. Chromatogr. Chromatogr. Commun. 1002,69-71. Ramsey, J. C., Park, C. N., Ott, M. G., and Gehring, P. J. (1978). Carcinogenic risk assessment: Ethylene dibromide. Toxicol. Appl. Pharmacol. 47,411414. Rao, M. S., Subbarao, v., Yeldandi, A. v., and Reddy, J. K. (I 992a). Hepatocarcinogenicity of dehydroepiandrosterone in the rat. Cancer Res. 52, 2977-2979. Rao, M. S., Subbarao, v., Yeldandi, A. v., and Reddy, J. K. (1992a). Inhibition of spontaneous testicular Leydig cell tumor development in F-344 rats by dehydroespiandrosterone. Cancer Lett. 65, 123-126.

References Rice, J. M., Baan, R. A., Blettner, M., Genevois-Charmeau, c., Grosse, Y., McGregor, D. B., Partensky, c., and Wilbourn, J. D. (1999). Rodent tumors of urinary bladder, renal cortex, and thyroid gland in IARC Monographs evaluations of carcinogenic risk to humans. Toxico!. Sci. 49, 166-17 I. Ries, L. A. G., Eisner, M. P., Kosary, C. L., Hankey, B. F., Miller, B. A., Clegg, L., and Edwards, B. K. (eds.) (2000). "SEER Cancer Statistics Review, 1973-1997." National Cancer Institute, Bethesda, MD. Risch, B., and Herrmann, K. (1988). Die Gehalte an HydroxyzintsaureVerbindungen und Catechinen in Kern- und Steinobst. Z. Lebensm. Unters.Forsch. 186, 225-230. Roberts, R. A., and Kimber, I. (1999). Cytokines in non-genotoxic hepatocarcinogenesis. Carcinogenesis 20, 1397-1401. Rosculet, G., and Rickard, M. (1968). Isolation and characterization of fiavor components in beer. Am. Soc. Brew. Chem. Proc., 203-213. Safe, S., Wang, F., Porter, w., Duan, R., and McDougal, A. (1998). Ah receptor agonists as endocrine disruptors: Antiestrogenic activity and mechanisms. Toxicol. Left. 102-103,343-347. Sawyer, c., Peto, R., Bernstein, L., and Pike, M. C. (1984). Calculation of carcinogenic potency from long-term animal carcinogenesis experiments. Biometrics 40, 27-40. Saxe, T. G. (1987). Toxicity of medicinal herbal preparations. Am. Fam. Physician 35,135-142. Schmidtlein, H., and Herrmann, K. (I 975a). Ober die Phenolsauren des Gemiises. H. Hydroxyzimtsauren und Hydroxybenzoesauren der Fruchtund Samengemiisearten. Z. Lebensm. Unters.-Forsch. 159,213-218. SchmidtIein, H., and Herrmann, K. (1975b). Uber die Phenolsauren des Gemiises. I. Hydroxyzimtsauren und Hydroxybenzoesauren der Kohlarten und anderer Cruciferen-Blatter. Z. Lebensm. Unters.-Forsch. 159, 139-148. SchmidtIein, H., and Herrmann, K. (l975c). Ober die Phenolsauren des Gemiises. IV. Hydroxyzimtsauren und Hydroxybenzosliuren weiterer Gemiisearten und der Kartoffeln. Z. Lebensm. Unters.-Forsch. 159, 255263. Schreier, P., Drawert, F., and Heindze, I. (1979). Uber die quantitative Zusammensetzung natiirlicher und technologish verlinderter pfianzIicher Aromen. Chem. Mikrobio!. Techno!. Lebensm. 6, 78-83. Schwartz, A. G., Hard, G. c., Pashko, L. L., Abou-Gharbia, M., and Swern, D. (1981). Dehydroespiandrosterone: An anti-obesity and anti-carcinogenic agent. Nutr. Cancer 3, 46-53. SeIigman, P. J., Mathias, C. G., O'Malley, M. A., Beier, R. c., Fehrs, L. J., Serrill, W. S., and Halperin W. E. (1987). Phytophotodermatitis from celery among grocery store workers. Arch. Dermato!' 123, 1478-1482. Sen, N. P., Seaman, S., and Miles, W. F. (1979). Volatile nitrosamines in various cured meat products: Effect of cooking and recent trends. 1. Agric. Food Chem. 27, 1354-1357. Senter, S. D., Robertson, J. A., and Meredith, F. I. (1989). Phenolic compounds of the mesocarp of Cresthaven peaches during storage and ripening. J. Food Sci.54, 1259-1260, 1268. Senti, F. R., and Pilch, S. M. (1985). Analysis of folate data from the second National Health and Nutrition Examination Survey (NHANES H). 1. Nutr. 115,1398-1402. Shinohara, T., Shimazu, Y., and Watanabe, M. (1979). Dosage de I'acetoi'ne du lactate d'ethyle dans les vins par chromatographie en phase gazeuse, et etude de leur formation dans les vins. Agric. Bio!. Chem. 43, 2569-2577. Siegal, D. M., Frankos, V. H., and Schneiderman, M. (1983). Formaldehyde risk assessment for occupationally exposed workers. Regul. Toxicol. Pharmacol. 3,355-371. Silwar, R., Kamperschroer, H., and Tressl, R. (1987). GaschromatographischRostkaffeearomasmassenspektrometrische Untersuchungen des Quantitative Bestimmung wasserdampffiiictiger Armoastoffe. Chem. Mikrobiol. Technol. Lebensm. 10, 176-187. Smyth, H. F., Jr., Carpenter, C. P., and Weil, C. S. (1951). Rangefinding toxicity data: List IV. Arch. Ind. Hyg. Occup. Med. 4, II9-122. Snyderwine, E. G., Turesky, R. J., Turteltaub, K. w., Davis, C. D., Sadrieh, N., Schut, H. A. J., Nagao, M., Sugimura, T., Thorgeirsson, U. P., Adamson, R. H., and Snorri, S. (1997). Metabolism of food-derived heterocyciic amines in nonhuman primates. Mutat. Res. 376, 203-210.

841

Staroscik, J. A., and Wilson, A. A. (1982a). Quantitative analysis of coldpressed lemon oil by glass capillary gas chromatography. J. Agric. Food Chem. 30,507-509. Staroscik, J. A., and Wilson, A. A. (l982b). Seasonal and regional variation in the quantitative composition of cold-pressed lemon oil from California and Arizona. J. Agric. Food Chem. 30, 835-837. Steinmetz, K. A., and Potter, J. D. (1991). Vegetables, fruit, and cancer. I. Epidemiology. Cancer Causes Control 2, 325-357. Stich, H. F., Rosin, M. P., Wu, C. H., and Powrie, W. D. (1981). A comparative genotoxicity study of chlorogenic acid (3-0caffeoylquinic acid). Mutat. Res. 90, 201-212. Stotberg, J., and Grundschober, F. (1987). Consumption ratio and food predominance of fiavoring materials: Second cumulative series. Perfum. Flavor. 12, 27-56. StOhr, H., and Herrmann, K. (1975). Uber die Phenolsauren des Gemiises: HI. Hydroxyzimtsauren und Hydroxybenzoesauren des Wurzelgemiises. Z. Lebensm. Unters.-Forsch. 159,219-224. Stoll, B. A. (1999). Dietary supplements of dehydroepiandrosterone in relation to breast cancer risk. Eur. J. Clin. Nutr. 53, 771-775. Takagi, K., Toyoda, M., Fujiyama, Y., and Saito, Y. (1990). Effect of cooking on the contents of a-chaconine and a-solanine in potatoes. J. Food Hyg. Soc. Jpn. 31,67-73. Takayama, S., Sieber, S. M., Adamson, R. H., Thorgeirsson, U. P., Dalgard, D. w., Amold, L. L., Cano, M., EkIund, S., and Cohen, S. M. (1998). Long-term feeding of sodium saccharin to nonhuman primates: Implications for urinary tract cancer. J. Natl. Cancer Inst. 90, 19-25. Takayama, S., Sieber, S. M., Dalgard, D. W., Thorgeirsson, U. P., and Adamson, R. H. (1999). Effects of long-term oral administration of DDT on nonhuman primates. J. Cancer Res. Clin. Oncol. 125,219-225. Tanner, H., and Limacher, H. (1984). Direktbestimmung von Methanol, Ethanol und Acetaldehyd. Fliissiges Obst 51, 182-184. Technical Assessment Systems (TAS) (1989). Exposure I Software Package. Technical Assessment Systems, Washington, DC. Theranaturals (2000). Theranaturals BC Caps. Available at http://www. theranaturals.com. Thorgeirsson, U. P., Dalgard, D. W., Reeves, J., and Adamson, R. H. (1994). Tumor incidence in a chemical carcinogenesis study in nonhuman primates. Regul. Toxicol. Pharmacol. 19, 130-151. Tomatis, L., and Bartsch, H. (1990). The contribution of experimental studies to risk assessment of carcinogenic agents in humans. Exp. Pathol. 40, 251266. Toth, B., and Erickson, J. (1986). Cancer induction in mice by feeding of the uncooked cultivated mushroom of commerce Agaricus bisporus. Cancer Res. 46, 4007-4011. Travis, C. c., Richter Pack, S. A., Saulsbury, A. w., and Yambert, M. W. (1990). Prediction of carcinogenic potency from toxicological data. Mutat. Res. 241,21-36. Tressl, R., Bahri, D., Kiippler, H., and Jensen, A. (1978). Diphenole und Caramelkomponenten in Rostkaffees verschiedener Sorten. n. z. Lebensm. Unters.-Forsch. 167, I II-II4. Tressl, R., Drawert, F., Heimann, W., and Emberger, R. (1970). Uber die Biogenese von Aromastoffen bei Pfianzen und Friicten. VI. Mitteilung: Ester, Alkohole, Carbonylverbindungen und Phenolather des Bananenaromas. Z. Lebensm. Unters.-Forsch. 142,313-321. Tricker, A. R., and Preussmann, R. (1991). Carcinogenic N-nitrosamines in the diet: Occurrence, formation, mechanisms and carcinogenic potential. Mutat. Res. 259,277-289. Trosko, J. E. (1998). Hierarchical and cybernetic nature of biologic systems and their relevance to homeostatic adaptation to low-level exposures to oxidative stress-inducing agents. Environ. Health Perspect. 106,331-339. United Fresh Fruit and Vegetable Association (UFFVA) (1989). "Supply Guide: Monthly Availability of Fresh Fruit and Vegetables." United Fresh Fruit and Vegetable Association, Alexandria, VA. U.S. Department of Agriculture (2000). "Pesticide Data Program, Annual Summary Calendar Year 1998." Agricultural Marketing Service, Washington, DC.

842

CHAPTER 38


U.S. Environmental Protection Agency (l984a). "Review of Rat and Mouse Data for the Carcinogenicity of Linuron." Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (l984b). "Ethylene Dibromide (EDB) Scientific Support and Decision Document for Grain and Grain Milling Fumigation Uses." Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, De. U.S. Environmental Protection Agency (1985-1988). "Peer Review of [... ]." Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, De. [Alachlor 1986; Asulam 1988; Captafol1987; Captan 1986; Cypermethrin 1988; Folpet 1986; Metolachlor 1985; Oryzalin 1985; Oxadiazon 1986]. U.S. Environmental Protection Agency (1985a). "Chlordimeform Risk Assessment." Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, De. U.S. Environmental Protection Agency (1985b). "Consensus Review of Acephate." Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (l986a). Guidelines for carcinogen risk assessment. Fed. Reg. 51, 33992-34003. U.S. Environmental Protection Agency (l986b). "Abbreviated Peer Review Meeting on Guthion." Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (l987a). "Peer Review of Chlorothalonil." Office of Pesticides and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. Review found in Health Effect Division Document 007718. U.S. Environmental Protection Agency (1987b). "Chlorothalonil-Rat Study, Qualitative and Quantitative Risk Assessment." Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (I 988a). Regulation of pesticides in food: Addressing the Delaney paradox policy statement; notice. Fed. Reg. 53,41112-41123. U.S. Environmental Protection Agency (1988b). "Permethrin-Quantitative Risk Assessment, Two Year Chronic/Oncogenicity Mouse (Females) Study." Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (1989a). "Daminozide Special Review. Technical Support Document-Preliminary Determination to Cancel the Food Uses of Daminozide." Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, De. U.S. Environmental Protection Agency (l989b). "MBC (INE-965)Qualitative and Quantitative Risk Assessment, CD-l Mouse Study (ReEvaluation)." Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (1989c). "Second Peer Review of Parathion." Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (1991a). "Report of the EPA Peer Review Workshop on Alpha2u-globulin: Association with Renal Toxicity and Neoplasia in the Male Rat." U.S. Environmental Protection Agency, Washington, De. U.S. Environmental Protection Agency (l991b). "EBDCIETU Special Review. DRES Dietary ExposurelRisk Estimates." Memo from R. Griffin to K. Martin, U.S. Environmental Protection Agency, Washington, De. U.S. Environmental Protection Agency (1992). Ethylene bisdithiocarbamates (EBDCs); notice of intent to cancel; conclusion of special review. Fed. Reg. 57,7484--7530. U.S. Environmental Protection Agency (I 994a). "Estimating Exposure to Dioxin-Like Compounds (Review Draft)." U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (l994b). "Health Assessment Document for 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and Related Compounds." U.S. Environmental Protection Agency, Washington, De.

U.S. Environmental Protection Agency (1995a). "Re-evaluating Dioxin: Science Advisory Board's Review of EPA's Reassessment of Dioxin and Dioxin-Like Compounds." U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (1995b). "Reregistration Eligibility Decision (RED): Linuron." U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (1996a). Proposed Guidelines for Carcinogen Risk Assessment. Fed. Reg. 61, 17960-18011. U.S. Environmental Protection Agency (1996b). "Exposure Factors Handbook." Draft Report, U.S. Environmental Protection Agency, Washington, De. U.S. Environmental Protection Agency (1998). "Status of Pesticides in Registration, Reregistration, and Special Review." U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (l999a). Pesticides: Science policy issues related to the food quality protection act. Fed. Reg. 64, 162-166. U.S. Environmental Protection Agency (1999b). "Integrated Risk Information System (IRIS)." Office of Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH. U.S. Environmental Protection Agency (2000). "Exposure and Human Health Reassessment of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and Related Compounds." Draft Final, U.S. Environmental Protection Agency, Washington, De. U.S. Food and Drug Administration (1960). Refusal to extend effective date of statute for certain specified food additives. Fed. Reg. 25, 12412. U.S. Food and Drug Administration (1988). FDA Pesticide Program: Residues in foods 1987. J.-Assoc. Off. Anal. Chem. 71, 156A-174A. U.S. Food and Drug Administration (1989). FDA Pesticide Program: Residues in foods 1988. J.-Assoc. Off. Anal. Chem. 72, 133A-152A. U.S. Food and Drug Administration (1990). FDA Pesticide Program: Residues in foods 1989. J.-Assoc. Off. Anal. Chem. 73, 127A-146A. U.S. Food and Drug Administration (1991a). FDA Pesticide Program: Residues in foods 1990. J.-Assoc. Off. Anal. Chem. 74, 121A-141A. U.S. Food and Drug Administration (1991b). "Butylatedhydroxyanisole (BHA) Intake." Memo from Food and Color Additives Section to L. Lin, U.S. Food and Drug Administration, Washington, De. U.S. Food and Drug Administration (1992a). FDA Pesticide Program-Residue monitoring 1991. J.-Assoc. Off. Anal. Chem. 75, 136A-158A. U.S. Food and Drug Administration (1992b). "Exposure to Aflatoxins." U.S. Food and Drug Administration, Washington, DC. U.S. Food and Drug Administration (1993a). Food and Drug Administration Pesticide Program: Residue monitoring 1992. J.-Assoc. Off. Anal. Chem. 76,127A-148A. U .S. Food and Drug Administration (1993b). "Assessment of Carcinogenic Upper Bound Lifetime Risk Resulting from Aflatoxins in Consumer Peanut and Corn Products." Report of the Quantitative Risk Assessment Committee, U.S. Food and Drug Administration, Washington, De. U.S. National Cancer Institute (1984). Everything doesn't cause cancer: But how can we tell which things cause cancer and which ones don't? NIH Publication 84-2039, U.S. National Cancer Institute, Bethesda, MD. U.S. National Cancer Institute (1996). Why eat five? J. Natl. Cancer Inst. 88, 1314. U.S. National Institute for Occupational Safety and Health (1999). "Chemical Abstracts" (CA Search). DIALOG, Knight-Ridder Information, Mountain View,CA. U.S. National Toxicology Program (1997). "Effect of Dietary Restriction on Toxicology and Carcinogenesis Studies in F344IN Rats and B6C3F 1 Mice." U.S. National Toxicology Program, Research Triangle Park, NC. U.S. National Toxicology Program (2000a). "Ninth Report on Carcinogens." U.S. National Toxicology Program, Research Triangle Park, NC. U.S. National Toxicology Program (2000b). "Background Information Indole3-carbinol (BC) 700-06-1." U.S. National Toxicology Program, Research Triangle Park, Ne. U.S. National Toxicology Program (2001). Addendum to Ninth Report on Carcinogens: 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD); DIOXIN CAS No. 1746-01-6. Research Triangle Park, NC. NTP.

References

Vainio, H., Wilbourn, l. D., Sasco, A. l., Partensky, C., Gaudin, N., Heseltine, E., and Eragne, 1. (1995). Identification des facteurs cancerogenes pour l'homme dans les Monographies du CIRC. Bull. Cancer 82, 339-348. van Vollenhoven, R. F. (2000). Dehydroepiandrosterone in systemic lupus erythematosus. Rheum. Dis. Clin. North. Am. 26, 349-362. Vernot, E. H., MacEwen, l. D., Haun, C. C., and Kinkead, E. R. (1977). Acute toxicity and skin corrosion data for some organic and inorganic compounds and aqueous solutions. Toxicol. Appl. Pharmacol. 42,417-423. Viehoever, A. (1940). Edible and poisonous beans of the lima type (Phaseolus lunatus L.): A comparative study, including other similar beans. Thai Sci. Bull. 2, 1-99. Ward, l. M., Peters, l. M., Perella, C. M., and Gonzalez, F. l. (1998). Receptor and nonreceptor-mediated organ-specific toxicity of di(2ethylhexyl)phthalate (DEHP) in peroxisome proliferatoractivated receptor a-null mice. Toxicol. Pathol. 26, 240-246. Wilker, C., lohnson, L., and Safe, S. (1996). Effects of developmental exposure to indole-3-carbinol or 2,3,7,8-tetrachlorodibenzo- p-dioxin on reproductive potential of male rat offspring. Toxicol. Appl. Pharmacol. 141,68-75. Williams, M. P., Hoff, l. E., and Nelson, P. E. (1972). A precise method for the determination of dimethyl sulfide in processed foods. 1. Food Sci. 37, 408-410. Winter, C. (1992). Pesticide tolerances and their relevance as safety standards. Regul. Toxicol. Pharmacal. 15, 137-150. Winter, M., and Herrmann, K. (1986). Esters and glucosides of hydroxycinnamic acids in vegetables. 1. Agric. Food Chem. 34,616-620. Winter, M., Brandl, W., and Herrmann, K. (1987). Bestimmung von Hydroxyzimtsaure-Derivaten in Gemiise. Z. Lebensm. Unters.-Forsch. 184, 11-16. Wolm, G., Tunger, L., and Rothe, M. (1974). Verkurzte Teigbereitung bei Weirbrot und ihr Einfluauf das Aroma. 2. Mitt. Alkanole als Aroma-Index. Nahrung 18, 157-164. Wolman, W. (1955). Instant and decaffeinated coffee. 1. Am. Med. Assoc. 159, 250. Woodyatt, N. l., Lambe, K. G., Myers, K. A., Tugwood, l. D., and Rovert, R. A. (1999). The peroxisome proliferator (PP) response element upstream of the

843

human acyl CoA oxidase gene is inactive among a sample human population: Significance for species differences in response to PPs. Carcinogenesis 20,369-372. World Health Organization (1993). "Polychlorinated Biphenyls and Terphenyls." Environmental Health Criteria 140, WHO, Geneva. Wulf, L. W, Nagel, C. W, and Branen, A. L. (1978). Analysis of myristicin and falcarinol in carrots by high-pressure liquid chromatography. 1. Agric. Food Chem. 26, 1390-1393. Wu-Williams, A. H., Zeise, L., and Thomas, D. (1992). Risk assessment for aflatoxin B]: A modeling approach. RiskAnal. 12,559-567. Yamagami, T., Handa, H., luji, T., Munemitsu, H., Aoki, M., and Kato, Y. (1983). Rat pituitary adenoma and hyperplasia induced by caffeine administration. Surg. Neural. 20, 323-331. Yess, N. l., Gunderson, E. L., and Roy, R. R. (1993). U.S. Food and Drug Administration monitoring of pesticide residues in infant foods and adult foods eaten by infants/children. l.-Assoc. Off. Anal. Chem. 76,492-507. Yu, M. c., Tong, M. l., Govindarajan, S., and Henderson, B. E. (1991). Nonviral risk factors for hepatocellular carcinoma in a low-risk population, the nonAsians of Los Angeles County, California. 1. Natl. Cancer Inst. 83, 18201826. Yu, T., Wu, C., and Liou, Y. (1989). Effects of pH adjustment and subsequent heat treatment on the formation of volatile compounds of garlic. 1. Food Sci. 54,632-635. Zarembski, P. M., and Hodgkinson, A. (1962). The oxalic acid content of English diets. Br. 1. Nutr. 16,627-634. Zeiger, E. (1997). Genotoxicity Database. In "Handbook of Carcinogenic Potency and Genotoxicity Databases" (L. S. Gold, and E., Zeiger, eds.), pp. 687-729. CRC Press, Boca Raton, FL. Zeise, L., Wilson, R., and Crouch, E. (1984). Use of acute toxicity to estimate carcinogenic risk. Risk Anal. 4, 187-199. Zoumas, B. L., Kreiser, W R., and Martin, R. A. (1980). Theobromine and caffeine content of chocolate products. 1. Food Sci. 45,314-316.

CHAPTER

39 Perceptions of Pesticides as Risks to Human Health Paul Slovic Decision Research, Inc.

39.1 INTRODUCTION Public perceptions of risk have been studied systematically for more than 20 years, within the United States and in other countries. Throughout that time period, the use of pesticides has been perceived as one of the most risky activities pursued by human societies. It is difficult to pinpoint the origins of these perceptions but certainly Rachel Carson's book, Silent Spring, first published in 1962, has played an important role. In Carson's story, pesticides are singled out as among the most potent "elixirs of death." Referring to synthetic pesticides, Carson observed, "They have immense power not merely to poison but to enter the most vital processes of the body and change them in sinister and often deadly ways" (Carson, 1962, p. 25). Ironically, as will be shown, the heavy reliance on testing of chemicals with animals and the quantitative risk assessment that has developed during this same time period may have reinforced and maintained the public's fears of pesticides and other chemicals.

39.2 RISK-PERCEPTION STUDIES One of the first quantitative studies of risk perception took place in the United States in the late 1970s and early 1980s. This research showed that perceptions of risk can be described in terms of numerous characteristics or dimensions. Figure 39.1, for example, presents a spatial display of hazards within a perceptual space derived from more than 40,000 individual judgments. The factors in this space reflect the degree to which a hazard is perceived to be known or understood (vertical dimension) and the degree to which it evokes perceptions of dread, uncontrollability, and catastrophe (horizontal dimension). Research has also demonstrated that social response to risk is closely related to the position of a hazard within this space. The further to the right a hazard appears, the higher its perceived risk, the more people want to see its current risks reduced, and Handbook of Pesticide Toxicology Volume 1. Principles

the more they want to see strict regulation employed to achieve reduced risk. Media coverage appears to be most extensive and intense when something goes wrong in the upper right-hand quadrant of the space, in an activity whose risks are seen as being poorly understood, evoking dread, and potentially leading to a catastrophe. In this light, it is interesting to see that pesticides fall in the problematic upper right quadrant of the space, reflecting the fact that respondents in this study characterized them as poorly known or understood, delayed in effect, relatively new, uncontrollable, evoking dread, catastrophic, fatal, inequitable, and posing high risk to future generations. Their location in this space is not too distant from activities related to the use of nuclear power. Whereas public judgments of risk seem closely related to the characteristics that define the space in Fig. 39.1, expert's judgments of risk are not closely related to any of these various risk characteristics. Instead, experts appear to see riskiness as synonymous with expected annual mortality. As a result, many conflicts in society over "risk" may result from experts and laypeople having different definitions of the concept. In this light, it is not surprising that expert recitations of "risk statistics" often do little to change people's attitudes and perceptions. In addition to constructing spatial displays such as that in Fig. 39.1, research has compared perceptions of risk and benefit from a large number of activities and technologies. It is particularly instructive to compare perceptions of various radiation and chemical technologies. Nuclear power has a very high perceived risk and low perceived benefit, whereas diagnostic X-rays have the opposite pattern (low perceived risk, high perceived benefit). A parallel finding occurs with chemicals. Nonmedical sources of exposure to chemicals (e.g., pesticides, food additives, alcohol, cigarettes) are seen as very low benefit and high in risk; chemicals used in medicine (e.g., prescription drugs, antibiotics, vaccines) are generally seen as high in benefit and low in risk, despite the fact that they can be very toxic substances.

845


846

CHAPTER 39

Perceptions of Pesticides as Risks to Human Health

Factor 2 Unknown risk

Laetrile. Microwave Ovens.

• DNA Technology

Water Fluoridation. Saccharin. .Nitrates Water Chlorination.·Hexachlorophene Coal Tar Hairdyes. Polyvinyl. · • Chloride Ora I Con tracep tIves. Diagnostic Valium • • IUD • Darvon

X-Rays Antibiotics. .Rubber Mfg.

Auto Lead. • Lead Paint

• Caffeine .Aspirin

.Electric Fields .DES .Nitrogen Fertilizers

.SST

• Radioactive Waste .Cadmium Usage .2,4,5-T Mirex • Trichloroethylene • Nuclear Reactor Accidents .Uranium Mining • Pesticides • Nuclear Weapons • Asbestos • PCBs Fallout Insulation • lOOT • Satellite Crashes Mercury FisSil Fuels Coal Burning (Pollution)

-r__~~~--~~--~~~~--~-------------------Factor1

______________._v_a_~~in_es____________ • Skateboards

Smoking (Disease). • Power Mowers • Snow mobiles Trampolines.

• Tractors • Alcohol • Chainsaws

. . • Elevators Home SWlmmln9 Po?ls • • Electric WIr & Appl ( Downhill SkIIng. . Recreational Boating • • Smoking Motorcycles Electric WIr & Appl (Shock). ·Bicycles

•

.Auto Exhaust (CO) • LNG Storage & .D-CON Transport I M" (D' ) oa Inlng Isease • Large Dams • SkyScraper Fires

.C

Dread risk

• Nerve Gas Accidents

Nuclear Weapons (War) • Underwater • Construction .Coal Mining Accidents • Sport Parachutes • General Aviation • High Construction • Railroad Collisions lcohol • Commercial Aviation ccidents • Auto Racing Auto A~idents • Handguns • Dynamite

1Factor 21

Controllable Not Dread Not Global Catastrophic Consequences Not Fatal Equitable Individual Low Risk to Future Generations Easily Reduced Risk Decreasing Voluntary

Not Observable Unknown to Those Exposed Effect Delayed New Risk Risk Unknown to Science

Observable Known to those Exposed Effect Immediate Old Risk Risks Known to Science

Uncontrollable Dread Global Catastrophic Consequences Fatal Not Equitable Catastrophic High Risk to Future Generations Not Easily Reduced Risk Increasing Involuntary

1

Factor 1 1

Figure 39.1 Location of 81 hazards in a two-factor space derived from the relationships among 15 risk characteristics. Each factor is made up of a combination of characteristics, as indicated by the lower diagram. [Slovic, P., Fischhoff, B., and Lichtenstein, S. (1985). Characterizing perceived risk. In "Perilous Progress: Technology as Hazard" (R. W. Kates, C. Hohenemser, and J. X. Kasperson, eds.), pp. 91-123. Westview, Boulder, CO.]

Research throughout the 1980s and 1990s has continued to show high levels of concern regarding the risks from the use of pesticides. Figures 39.2 and 39.3 present data from parallel national surveys in the United States and France conducted in 1992. The U.S. survey shows significant concerns regarding

pesticides in food, which were seen as close in risk to drinking alcohol, motor vehicle accidents, and nuclear power plants and higher in risk than two rather significant hazards, bacteria in food and storms and floods. The picture in France was similar (Figure 39.3). There was a somewhat greater frequency of

39.2 Risk-Perception Studies

847

Cigarette smoking Street drugs AIDS Nuclear waste Stress Chemical pollution Ozone depletion Suntanning Drinking alcohol Motor vehicle accidents Pesticides in food Nuclear power plants Blood transfusions Outdoor air quality Climate change Bacteria in food High-voltage power lines Food irradiation Coal/oil power plants Genet engr bacteria Radon in home Storms & floods VDTs Commercial air travel Medical X-rays 00%

•

Figure 39.2

20%

High risk

D

40%

60%

Moderate ~ Slight risk L::::::.:::J risk

100%

80%

' D ;

,Almost no risk

D

Don't know

Perceived health risks to American public: 1992 national survey.

high-risk responses to pesticides in food in France than in the United States. Similar results were obtained in a 1993 national survey of the Canadian public (Figure 39.4), with more than 30% of Canadians judging pesticides in food as high risk and more than 70% judging them as high or moderate risk. Pesticides in food were rated as higher in risk than drinking alcohol, nuclear power plants, asbestos, and bacteria in food. When the same survey was given to members of the Society of Toxicology of Canada (see Fig. 39.5), pesticides in food were judged far lower in risk (less than 10% judged them as high risk; only about 20% judged them high or moderate risk). Contrary to the public's opinions, the toxicologists judged bacteria in food to be riskier than pesticides in food. Members of the British Toxicology Society (BTS) surveyed by Slovic et al. (1997) also judged pesticides in food to pose rather small risks (see Fig. 39.6). The most recent data that my colleagues and I have collected come from a national survey in the United States in 1997. Pes-

ticides were seen as posing high risk to the American public by 26% of the respondents and posing high or moderate risk by 69%. Pesticides were rated almost as risky as stored nuclear waste, motor vehicles, nuclear power plants, and natural disasters (see Fig. 39.7). In another segment of that survey, respondents judged risks to individuals to be almost or as great from pesticides as from handguns and violent crime. 39.2.1 SOCIAL, CULTURAL, AND POLITICAL INFLUENCES ON RISK PERCEPTION

The data shown in Figs. 39.1-39.7 reflect only the tip of the iceberg. Below the surface are rumblings of quite complex social, cultural, and political forces shaping the observed ratings of risks from pesticides and other hazards. Recent studies have shown that such factors as gender, race, political worldview, affiliation, emotion, and trust are strongly correlated with risk

848

CHAPTER 39


Nuclear waste AIDS Street drugs Cigarette smoking Chemical pollution Motor vehicle accidents Ozone depletion Drinking alcohol Stress Pesticides in food Genet engr bacteria Bacteria in food Suntanning Blood transfusions Nuclear power plants Food irradiation Climate change Outdoor air quality Storms & floods High-voltage power lines VDTs Medical X-rays Coal/oil-burning plants Commercial air travel Radon in home 00% High nsk

Figure 39.3

20%

40%

Moderate risk

60% Almost no nsk

80%

c

100% Don·t know

Perceived health risks to French public: 1992 national survey.

judgments. Equally important is that these factors influence the judgments of experts as well as the judgments of laypersons. For example, gender is strongly related to risk judgments and attitudes. Several dozen studies have documented the finding that men tend to judge risks as smaller and less problematic than do women (Brody, 1984; Carney, 1971; DeJoy, 1992; Gutteling and Wiegman, 1993; Gwartney-Gibbs and Lach, 1991; Pillisuk and Acredolo, 1988; Sj6berg and Drottz-Sj6berg, 1993; Slovic et aI., 1989, 1993; Spigner et aI. , 1993; Steger and Witt, 1989; Stem et aI. , 1993). A number of hypotheses have been put forward to explain sex differences in risk perception. One approach has been to focus on biological and social factors. Women have been characterized as more concerned about human health and safety because they are socialized to nurture and maintain life (Steger and Witt, 1989). They have been characterized as physically more vulnerable to violence, such as rape, and this may sensitize them to other risks (Baumer, 1978; Riger et al., 1978). The combination of biology and social experience

has been put forward as the source of a "different voice" that is distinct to women (Gilligan, 1982; Merchant, 1980). A lack of knowledge and familiarity with science and technology has also been suggested as a basis for these differences, particularly with regard to nuclear and chemical hazards. Women have been discouraged from studying science and there are relatively few women scientists and engineers (Alper, 1993). However, Barke, Jenkins-Smith, and Slovic have found that female physical scientists judge risks from nuclear technologies to be higher than do male physical scientists (Gilligan, 1982; Merchant, 1980). Similar results with scientists were obtained by Slovic et al. (1997), who found that female members of the BTS were far more likely than male toxicologists to judge societal risks, including pesticides, as moderate or high. Certainly female scientists in these studies cannot be accused of lacking knowledge and technological literacy. Some other factors must influence their decisions. Hints about the origin of these sex differences come from a study by Flynn et al., in which 1512 Americans were asked, for

39.2 Risk-Perception Studies

Cigarette smoking Ozone depletion Breast implants Street drugs Stress Chemical pollution Crime and violence Suntanning AIDS Motor vehicle accidents Nuclear waste Alcohol and pregnancy PCBs or dioxin Pesticides in food Food additives Drinking alcohol Nuclear power plants Climate change on prescription med. Asbestos Waste incinerators Malnutrition Hi-volt power lines Food irradiation Prescription drugs Genet engr bacteria Outdoor air quality Bacteria in food Molds in food Mercury in fillings Tap water Medical X-rays Indoor air quality VDTs Contraceptives Heart pacemakers Bottled water Contact lenses

-~

I

849

---• • •• • • I

I

~

••

I

I

I

I

I I

--•

00 %

I _ ~~h

-• ---I

• I

•

I

I

I

20%

60%

40%

~

Moderate risk

Slight nsk

I

80%

I Almost no risk

100%

Don't know -----l

Figure 39.4 Health risks to the Canadian public: 1992 Health and Welfare Canada survey (N = 1506). [Slovic, P., Malmfors, T., Krewski, D., Mertz, C. K., Neil, N., and Bartlett, S. (1995). Intuitive toxicology. n. Expert and lay judgments of chemical risks in Canada. RiskAnal. 15,661-675.]

each of 25 hazard items, to indicate whether the hazard posed (1) little or no risk, (2) slight risk, (3) moderate risk, or (4) high risk to society (Flynn et aI., 1994). Figure 39.8 shows the difference in the percentage of males and females who rated a hazard as a "high risk." All differences are to the right of the 0% mark, indicating that the percentage of high-risk responses was greater for women on every item (Flynn et aI., 1994). Perhaps the most striking result from this study is shown in Fig. 39.9, which presents the mean risk ratings separately for white males, white females, nonwhite males, and nonwhite

females. Across the 25 hazards, white males produced riskperception ratings that were consistently much lower than the means of the other three groups (Flynn et aI., 1994). When the data underlying Fig. 39.9 were examined more closely, Flynn et al. observed that not all white males perceived risks as low. The "white-male effect" appeared to be caused by about 30% of the white-male sample who judged risks to be extremely low (Flynn et aI., 1994). The remaining white males were not much different from the other subgroups with regard to perceived risk.

850

CHAPTER 39


Cigarette smoking Motor vehicle accidents Stress Alcohol and pregnancy Crime and violence Suntanning Breast implants Street drugs AIDS Ozone depletion Drinking alcohol Asbestos Molds in food Nuclear waste PCBs and dioxins Bacteria in.food Chemical pollution Malnutrition Nuclear power reactors Medical X -rays Climate changes Pesticides in food Indoor air pollution Rx drugs Waste incinerators Contraceptives Non-Rx drugs Power lines Outdoor air quality Food additives Tap water Contact lenses Genet engr bacteria Pacemakers VDTs Mercury in fillings Food irradiation Bottled water

_.

•

I

I I

I I

•

I I I I

I

I _ _ _ _ _ _ _ _ • • • •

p

I I I

I

•• ••

I

I

I

I

I

I

00%

--•

I

•

20%

I

40%

60%

, ~oderate- S~ght fISk

-

fisk

D

80%

Almost no risk

-

100%

Don't know

Figure 39.5 Health risks to the Canadian public: 1993 survey of the Society of Toxicology of Canada (N = 150). [Slovic, P., Malmfors, T., Krewski, D., Mertz, C. K. , Neil, N., and Bartlett, S. (1995). Intuitive toxicology.

n. Expert and lay judgments of chemical risks in Canada. Risk Anal.

Further analyses showed that the subgroup of white males who perceive risks to be quite Iow can be characterized by very great trust in institutions and authorities and by anti-egalitarian attitudes, including a disinclination toward giving decisionmaking power to citizens in areas of risk management. The results of this study raise new questions. What does it mean for the explanations of gender differences when we see that the sizable differences between white males and white females do not exist for non white males and nonwhite females? Why do a substantial percentage of white

15,661-675.)

males see the world as so much less risky than everyone else sees it? Obviously, the salience of biology is reduced by these data on risk perception and race. Biological factors should apply to nonwhite men and women as well as to white men and women. The present data thus move us away from biology and toward sociopolitical explanations. Perhaps white males see less risk in the world because they create, manage, control, and benefit from many of the major technologies and activities. Perhaps women and non white men see the world as more dangerous

39.3 Intuitive Toxicology: Expert and Lay Judgments of Chemical Risks

851

Cigarette smoking Asbestos Dioxins Motor vehicle traffic Suntanning Nuclear waste Crime and violence

liii~~~~~~~i~~~;liiliiiiiii~~

Alcoholic beverages Environmental tobacco smoke

1:=:=======:;;;;;;;;

Chemicals in the workplace Depletion of the ozone layer Burning fossil fuels Nuclear power reactors Breast implants Radon in homes Outdoor air pollution Waste incinerators Indoor air pollution Prescription drugs Pesticides in food

i~~~iiii!iiiiii~!~~~~~~~~~~~1

Non-prescription drugs Contraceptive pills Medical X-rays Electric and magnetic fields Food irradiation Mercury in dental fillings Food additives Tap water 00%

20%

40%

~ ~oderate. S~ght ~ r~k r~k

60 %

D

80%

Almost no risk

.

100%

Don't know

Figure 39.6 Health risks to the average exposed citizen of your country: 1994 British Toxicology Society (N = 312). [Slovic, P. , Malmfors, T., Mertz, C. K., Neil , N. , and Purchase, 1. F. H. (1997). Evaluating chemical risks: Results of the British Toxicology Society. Hum. Exp. Toxicol. 16,289-304.]

because in many ways they are more vulnerable, because they benefit less from many of its technologies and institutions, and because they have less power and control over what happens in their communities and their lives. Although the survey conducted by Flynn et al. was not designed to test these alternative explanations, the race and gender differences in perceptions and attitudes point toward the role of power, status, alienation, trust, perceived government responsiveness, and other sociopolitical factors in determining perception and acceptance of risks. According to this view, the problem of risk conflict and contro-

versy goes beyond science. It is deeply rooted in the social and political fabric of our society.

39.3 INTUITIVE TOXICOLOGY: EXPERT AND LAY JUDGMENTS OF CHEMICAL RISKS The preceding sections have described general attitudes and perceptions regarding pesticides and other hazards. In parallel

852

CHAPTER 39


Multiple sexual partners Street drugs 2nd-hand cigarette smoke Stored nuclear waste Motor vehicles Chemical manufacturing Nuclear power plants Natural disasters Pesticides Blood transfusions Lead in dust or paint Coal/oil-burning p. plants Tap water Airplane travel Radon in homes Electromagnetic fields Cellular phones Vaccines Asteroids 00%

• I

Figure 39.7 Survey.]

20%

High nsk

D

40%

Moderate fisk

[SJ Slight fisk

60%

U

80%

Almost no fisk

100%

~ Don·t ~

know

Perceived health risks to American public: U.S. population as a whole. [1997 National Risk

with this work, another stream of research has focused on chemical risks and attempted to go beneath the surface differences between experts and laypersons to document and understand the causes of these different views. This research has centered around a concept labeled intuitive toxicology (Kraus et aI., 1992; Slovic et aI., 1995, 1997). Research on intuitive toxicology was motivated by the premise that different assumptions, conceptions, and values underlie much of the discrepancy between expert and lay views of chemical risks. Research attempted to address this issue by exploring the cognitive models, assumptions, and inference methods that comprise laypeople's "intuitive toxicological theories" and comparing these theories with the cognitive models, assumptions, and inference methods of scientists working in the field of toxicology. The work began by identifying several fundamental principles and judgmental components within the science of risk

assessment. Questions were developed based on these fundamentals in order to determine the extent to which laypeople and experts share the same beliefs and conceptual framework. Questions addressed the following topics: (a) dose-response sensitivity; (b) trust in animal and bacterial studies; (c) attitudes toward chemicals; (d) attitudes toward reducing chemical risks; (e) conceptions of toxicity, including the toxicity of natural versus synthetic substances and the toxicity of prescription drugs versus chemicals in general; and (f) interpretation of evidence regarding cause-effect relationships between exposure to chemicals and human health. Questions on these topics were incorporated into a survey designed for both experts and the public. Each question was designed, whenever possible, according to a guiding hypothesis about how experts and "lay toxicologists" might respond. For example, a key principle in toxicology is the fact that "the dose makes the poison." Any substance can cause a toxic effect if the


Stress Suntanning Nuclear Waste Nuclear Power Plants Ozone Depletion AIDS Drinking Alcohol Hi-Volt Power Lines Street Drugs Motor Vehicle Accidents Blood Transfusions Chemical Pollution Pesticides in Food Bacteria in Food Cigarette Smoking Storms & Floods Radon in Home Climate Change Food Irradiation Outdoor Air Quality Coal/Oil Burning Plants Genet Engr Bacteria Medical X-Rays Commercial Air Travel VDTs

853

45.3 34.3 52.4 25.9 38.8 54.7 34.7 15.8 55.6 32.9 25.1 41.6 32.0 18.7 57.9 11.5 12.2 22.9 18.0

-10%

-5%

0%

5%

10%

15%

20%

25%

Percent difference in high risk Figure 39.8 Perceived health risks to American public by gender: difference between males and females. Note: Base percentage equals male high-risk response. Percentage difference is female high-risk response minus male high-risk response. [F1ynn, J., Slovic, P., and Mertz, C. K. (1994). Gender, race, and perception of environmental health risks. Risk Anal. 14, 1101-1108.]

dose is great enough. Thus, experts were predicted to be quite sensitive to considerations of exposure and dose when responding to questions on this topic. In contrast, the often-observed concerns of the public regarding very small exposures or doses of chemicals led to the hypothesis that the public would have more of an "all-or-none" view of toxicity and would be rather insensitive to concentration, dose, and exposure (thus equating any exposure with harm). Because the science of toxicology and the discipline of risk assessment rely so heavily upon animal studies, experts were predicted to have a more favorable view than laypersons regarding the value of such studies. The prediction that laypersons lack sensitivity to dose-response considerations and thus fear even small exposures to toxic or carcinogenic substances led to the prediction that they would exhibit far more negative attitudes toward chemicals than experts. This last prediction was confirmed dramatically in the studies. The members of the public who responded to these surveys associated exposure to chemicals to a remarkable extent with danger, cancer, and death, consistent with the general opinions described in Figs. 39.1-39.7 for pesticides and other chemicals. Specifically, studies of intuitive toxicology on national populations in the United States, Canada, and France have found that about 70% of the public believe that "if a person is exposed to a chemical that can cause cancer, then that person will probably

get cancer some day" (Kraus et al., 1992; Krewski et al., 1995). About 75% of the respondents in these surveys agreed that "If even a tiny amount of a cancer-producing substance was found in my tap water, I wouldn't drink it." More than 50% agreed that "There is no safe level of exposure to a cancer-causing chemical." The concern that any exposure to a carcinogen, no matter how small, is likely to cause cancer is linked to a desire to avoid chemicals and reduce the risks of exposure to them at any cost. About 75% of the public surveyed agreed that "I try hard to avoid contact with chemicals and chemical products in my daily life." About 62% agreed that "It can never be too expensive to reduce the risks from chemicals." Responses of toxicologists were not at all in agreement with these views. Of particular importance in this research is the finding, as predicted, that the public is much less sensitive than the experts to considerations of dose and exposure. Although the public recognizes the importance of these factors in some domains (e.g., prescription drugs), they generally tend to view chemicals as either safe or dangerous and they appear to equate even small exposures to toxic or carcinogenic chemicals with almost certain harm. This orientation was found to be associated with high levels of concern regarding chemicals, including very small residues of chemicals on food, and a desire to reduce chemical risks at any cost.

854

CHAPTER 39


---+- White Male

-

-+- - White Female

____ Non-White Male - ..... - Non-White Female I _ ______

._ ~.-------.J

Cigarette Smoking Street Drugs AIDS Stress Chemical Pollution Nuclear Waste Motor Vehicle Accidents Drinking Alcohol Suntanning Ozone Depletion Pesticides in Food Outdoor Air Quality Blood Transfusions Coal/Oil Burning Plants Climate Change

'a. I

Bacteria in Food Nuclear Power Plants Food Irradiation Storms & Floods Genet Engr Bacteria Radon in Home Hi-Volt Power Lines VDTs Medical X-Rays Commercial Air Travel

L-_ _ _ _ _ _ _ _ _ _ _ _ _ _ _ _ _ _ _---'

2

3

4

Slight Risk

Moderate Risk

High Risk

Figure 39.9 Mean risk-perception ratings by race and gender. [Flynn, J., Slovic, P., and Mertz, C. K. (1994). Gender, race, and perception of environmental health risks. Risk Anal. 14, 1104.]

Views on the validity of animal studies have been found to be more complex than expected. Consider two survey items that have been studied repeatedly. One is statement SI: "Would you agree or disagree that the way an animal reacts to a chemical is a reliable predictor of how a human would react to it?" The second statement, S2, is a little more specific: "If a scientific study produces evidence that a chemical causes cancer in animals, then we can be reasonably sure that the chemical will cause cancer in humans." When members of the American and Canadian public responded to these items, they showed moderate agreement with SI; about half the people agreed and half disagreed that animal tests were reliable predictors of human reactions to chemicals. However, in response to S2, which stated that the animal study found evidence of cancer, there was a jump in agreement to about 70% among both male and female respondents (see Fig. 39.10). The important point about the pattern of response is that agreement was higher on the second item. What happens when toxicologists are asked about these two statements? Figure 39.11 shows that toxicologists in the United States and toxicologists in the United Kingdom responded sim-

ilarly to the public on the first statement but differently on the second (Kraus et aI., 1992). They exhibited the same rather middling level of agreement with the general statement about

Percent agree 70% 60% 50% 40% 30% 20% 10% 00% Men

Women

Figure 39.10 Agreement among members of the V.S. public with statements SI and S2. [Kraus, N. N., Malmfors, T., and Slovic, P. (1992). Intuitive toxicology: Expert and lay judgments of chemical risks. Risk Anal. 12,215-232.]


Percent agree 90% 75% 60% 45% 30% 15% 00% U.S. Public N=262

U.S Toxicologists N= 170

U.K. Toxicologists N=312

Figure 39.11 Agreement among the public and toxicologists with statements SI and S2. [Slovic, P. (1997). Trust, emotion, sex, politics, and science: Surveying the risk-assessment battlefield. In "Environment, Ethics, and Behavior" (M. H. Bazerman, D. M. Messick, A. E. Tenbrunsel, and K. A. Wade-Benzoni, eds.), p. 299. New Lexington, San Francisco.)

animal studies as predictors of human health effects.! However, when these studies were said to find evidence of carcinogenicity in animals, the toxicologists were less likely to agree that the results could be extrapolated to humans. Thus, findings that lead toxicologists to be less willing to generalize to humans lead the public to see the chemical as more dangerous for humans. Figure 39.12 presents the responses for S! and S2 among men and women toxicologists in the United Kingdom (208 men and 92 women). Here, we see another interesting finding. The men agree less on the second statement than on the first, but the women agree more, just like the general public. Among toxicologists, women are more willing than men to say that one can generalize to humans from positive carcinogenicity findings in animals.

Percent agree 90% 75% 60% 45% 30% 15% 00% Men (-23, n = 208)

Women (+11 . n=92)

Figure 39.12 Agreement among men and women toxicologists in the United Kingdom with statements SI and S2. [Slovic, P. (1997). Trust, emotion, sex, politics, and science: Surveying the risk-assessment battlefield. In "Environment, Ethics, and Behavior" (M. H. Bazerman, D. M. Messick, A. E. Tenbrunsel, and K. A. Wade-Benzoni, eds.), pp. 299. New Lexington, San Francisco.) 1This is a rather surprising result, given the heavy reliance on animal studies in toxicology.

855

These studies of intuitive toxicology have yielded a number of intriguing findings that likely pertain to views about pesticides. One is the low percentage of agreement that animal studies can predict human health effects. Another is that toxicologists show even less confidence in equating human cancers with studies that find cancer in animals resulting from chemical exposure. The public, on the other hand, has high confidence in animal studies that find cancer. These studies also help us understand why the public has come to fear pesticides and other chemicals so greatly. As regulators have sought to develop more effective ways to meet public demands for a safer and healthier environment, they have come to rely heavily on quantitative risk assessment based on animal tests. Such tests often find evidence of cancer at high dose levels. Many scientists are skeptical of such evidence, on the grounds that high doses overwhelm the animals' defense mechanisms and produce cancers that would not occur in humans under normal conditions of exposure. This skepticism is seen in the high percentage of toxicologists who lack confidence in evidence for carcinogenicity derived from animal studies. The public, on the other hand, exhibits a high degree of confidence in positive findings from animal studies. Thus, the large number of animal studies performed over the years may have done a better job of scaring the public than of informing science about chemical carcinogenesis. Another contributing factor is that interpretation of the animal data has been based on a linear model that assumes that there is no level of exposure to a carcinogen that is without some degree of risk. Multiplying even very small probabilities of contracting cancer across large numbers of exposed individuals will likely project at least some number of deaths. This frightens people. Using upper 95% confidence bounds in the linear extrapolation makes the scenario even more frightening. Thus, the many people who believe there is no safe level of exposure to a carcinogen may have learned this from hearing about the linearity assumption or seeing risk estimates projected from a linear model. Psychological and anthropological research also helps us understand the nature of the public's fear of exposure to toxic substances that are said (by scientists using a linear model) to be toxic at all levels. For example, Frazer (1959) and Mauss (1972) describe a belief, widespread in many cultures, that things that have been in contact with each other may influence each other through transfer of some of their properties via an "essence." Thus, "once in contact, always in contact," even if that contact (exposure) is brief. Rozin et al. (1986) show that this belief system, which they refer to as a "law of contagion," is common in our present culture. The implication of these notions is that even a minute amount of a toxic substance in one's food (e.g., a pesticide residue) will be seen as imparting toxicity to the food; any amount of a carcinogenic substance will impart carcinogenicity, etc. The "essence of harm" that is contagious is typically referred to as contamination. Being contaminated clearly has an all-or-nothing quality to it-like being alive or pregnant. When a young child drops a sucker on the floor, the brief contact with "dirt" may be seen as contaminating the candy, causing the par-

856

CHAPTER 39


ent to throw it away rather than washing it off and returning it to the child's mouth. This all-or-nothing quality, irrespective of the degree of exposure, is evident in the observation by Erikson (1990) that "To be exposed to radiation or other toxins ... is to be contaminated in some deep and lasting way, to feel dirtied, tainted, corrupted" (p. 122). A contagion or contamination model is much more likely to hold in a world in which scientists use linear extrapolation to estimate risks than in a world that recognizes the beneficial effects of chemicals at low doses. We do not, for example, view ourselves as being "contaminated" by exposures to prescription drugs. Another relevant psychological tendency is to confound perception of risk with perception of benefit. If an activity or substance conveys some benefit upon us, we are likely to perceive it as less risky (Alhakami and Slovic, 1994) and more acceptable (Starr, 1969).

39.4 CONCLUSION In this brief overview, an attempt has been made to show the depth and complexity of the public's concerns regarding the risks from pesticides and other chemicals. These concerns transcend national borders and seem to have held remarkably constant for several decades, in spite of views of many toxicologists and other scientists that are quite the opposite from public views. Fortunately, the research described here will help us understand why public attitudes are the way they are and why they are so resistant to change. These attitudes are a complex product of human psychology and culture interacting with complex and idiosyncratic sciences such as toxicology, epidemiology, and risk assessment. One thing is clear. Risk communication efforts conducted by public relations specialists cannot turn public views around and may, in fact, exacerbate them. Some investigators have taken the limitations of risk science, the difficulty of creating and maintaining trust, and the subjective nature of risk judgments as signs pointing to the need for a radically different approach to dealing with conflicts regarding pesticides and other chemical products. This "new" approach focuses on introducing more public participation into both risk assessment and risk decision making to make the decision process more democratic, improve the relevance and quality of technical analysis, and increase the legitimacy and public acceptance of the resulting decisions. Work by scholars and practitioners in Europe and North America has begun to lay the foundation for improved methods of public participation within deliberative decision processes that include negotiation, mediation, oversight committees, and other forms of public involvement (English, 1992; Kunreuther et aI., 1993; National Research Council, 1996; Renn et aI., 1991, 1995). Those who are concerned about promoting rational decisions about the use of pesticides would be well advised to give careful consideration to this approach.

REFERENCES Alhakami, A. S., and Slovic, P. (1994). A psychological study of the inverse relationship between perceived risk and perceived benefit. Risk Anal. 14, 1085-1096. Alper, J. (1993). The pipeline is leaking women all the way along. Science 260, 409-411. Baumer, T L. (1978). Research on fear of crime in the United States. Victimology 3, 254-264. Brody, C. J. (1984). Differences by sex in support for nuclear power. Social Forces 63, 209-228. Carney, R. E. (1971). Attitudes toward risk. In "Risk Taking Behavior: Concepts, Methods, and Applications to Smoking and Drug Abuse" (R. E. Carney, ed.). Thomas, Springfield, IL. Carson, R. (1962). "Silent Spring." Fawcett, New York. DeJoy, D. (1992). An examination of gender differences in traffic accident risk perception. Accident Analysis and Prevention 24, 237-246. English, M. R. (1992). "Siting Low-Level Radioactive Waste Disposal Facilities: The Public Policy Dilemma." Quorum, New York. Erikson, K. (1990). Toxic reckoning: Business faces a new kind of fear. Harvard Business Rev. 68, 118-126. Flynn, J., Slovic, P., and Mertz, C. K. (1994). Gender, race, and perception of environmental health risks. Risk Anal. 14, 1101-1108. Frazer, J. G. (1959). "The New Golden Bough: A Study in Magic and Religion." Macmillan Co., New York (original work published in 1890). Gilligan, C. (1982). "In a Different Voice: Psychological Theory and Women's Development." Harvard Univ. Press, Cambridge, MA. Gutteling, J. M., and Wiegman, O. (1993). Gender-specific reactions to environmental hazards in the Netherlands. Sex Roles 28,433-447. Gwartney-Gibbs, P. A., and Lach, D. H. (1991). Sex differences in attitudes toward nuclear war. 1. Peace Res. 28, 161-174. Kraus, N. N., Malmfors, T, and Slovic, P. (1992). Intuitive toxicology: Expert and lay judgments of chemical risks. Risk Anal. 12,215-232. Krewski, D., Slovic, P., Bartlett, S., Flynn, J., and Mertz, C. K. (1995). Health risk perception in Canada. n. Worldviews, attitudes and opinions. Human and Ecological Risk Assessment 1, 231-248. Kunreuther, H., Fitzgerald, K., and Aarts, T. D. (1993). Siting noxious facilities: A test of the facility siting credo. Risk Anal. 13, 301-318. Mauss, M. (1972). "A General Theory of Magic." Norton, New York (original work published in 1902). Merchant, C. (1980). "The Death of Nature: Women, Ecology, and the Scientific Revolution." Harper & Row, New York. National Research Council, Committee on Risk Characterization (1996). "Understanding Risk: Informing Decisions in a Democratic Society" (P. C. Stem and H. V. Fineberg, eds.). Natl. Acad. Press, Washington, DC. Pillisuk, M., and Acredolo, C. (1988). Fear of technological hazards: One concern or many? Social Behavior 3, 17-24. Renn, 0., Webler, T., and Johnson, B. B. (1991). Public participation in hazard management: The use of citizen panels in the U.S. Risk-Issues in Health and Safety 2, 197-226. Renn, 0., Webler, T, and Wiedemann, P. (1995). "Fairness and Competence in Citizen Participation." Kluwer Academic, Dordrecht. Riger, S., Gordon, M. T, and LeBailly, R. (1978). Women's fear of crime: From blaming to restricting the victim. Victimology 3, 274-284. Rozin, P., Millman, L., and Nemeroff, C. (1986). Operation of the laws of sympathetic magic in disgust and other domains. J. Personality and Social Psychology 50, 703-712. Sjoberg, L., and Drottz-Sjoberg, B. M. (1993). "Attitudes Toward Nuclear Waste." Rhizikon Research Report 12, Center for Risk Research, Stockholm School of Economics, Stockholm. Slovic, P. (1997). Trust, emotion, sex, politics, and science: Surveying the risk-assessment battlefield. In "Environment, Ethics, and Behavior" (M. H. Bazerman, D. M. Messick, A. E. Tenbrunsel, and K. A. Wade-Benzoni, eds.), pp. 277-313. New Lexington, San Francisco. Slovic, P., Fischhoff, B., and Lichtenstein, S. (1985). Characterizing perceived risk. In "Perilous Progress: Technology as Hazard" (R. W. Kates, C. Hohenemser, and J. X. Kasperson, eds.), pp. 91-123. Westview, Boulder, CO.

References

Slovic, P., Flynn, l., Mertz, C. K., and Mullican, L. (1993). "Health Risk Perception in Canada." Report 93-EHD-170, Department of National Health and Welfare, Ottawa. Slovic, P., Kraus, N. N., Lappe, H., Letzel, H., and Malmfors, T. (1989). Risk perception of prescription drugs: Report on a survey in Sweden. Pharm. Med. 4, 43-65. Slovic, P., Malmfors, T., Krewski, D., Mertz, C. K., Neil, N., and Bartlett, S. (1995). Intuitive toxicology. n. Expert and lay judgments of chemical risks in Canada. RiskAnal. 15,661-675. Slovic, P., Malmfors, T., Mertz, C. K., Neil, N., and Purchase, 1. F. H. (1997). Evaluating chemical risks: Results of a survey of the British Toxicology Society. Hum. Exp. Toxicol. 16,289-304.

857

Spigner, C., Hawkins, W., and Loren, W. (1993). Gender differences in perception of risk associated with alcohol and drug use among college students. Women and Health 20, 87-97. Starr, C. (1969). Social benefit versus technological risk. Science 165, 12321238. Steger, M. A. E., and Witt, S. L. (1989). Gender differences in environmental orientations: A comparison of publics and activists in Canada and the V.S. Western Political Quarterly 42,627-649. Stem, P. c., Dietz, T., and Kalof, L. (1993). Value orientations, gender, and environmental concern. Environ. Behav. 25, 322-348.

CHAPTER

40 Mammalian Toxicity of Microbial Pest Control Agents Andrew L. Rubin Department of Pesticide Regulations, California Environmental Protection Agency

40.1 INTRODUCTION Concern for the safety of agricultural workers and the public, as well as for the integrity of ecosystems, has fueled an interest in the use of microbes as pest control agents (Siegel and Shadduck, 1992). Following the registration in the United States of Bacillus popillae as an insecticide in 1948, the list of federally registered microbial pest control agents (MPCAs) has grown to include not only bacteria, but also fungi, yeasts, viruses, and protozoa. As of August 2000, there were 245-250 products listing approximately 60 different MPCAs as active ingredients under federal registration [R. Torla, U.S. Environmental Protection Agency (EPA), personal communication]. In California, 91 products containing 33 MPCAs were under active registration as of September 2000 [Department of Pesticide Regulation (DPR) registration database]. Although occupying only a small fraction of the current pesticide market, the number of pounds of MPCA active ingredients applied agriculturally nearly tripled in California between 1990 and 1999 (DPR, 2000) (Table 40.1). Bacillus thuringiensis (Bt) products overwhelmingly dominated this segment of the market, with much of the increase over the 1990-1999 period due to the use of B. thuringiensis o-endotoxins encapsulated in killed Pseudomonas jluorescens. Nonetheless, the use of MPCAs occupying a much smaller market fraction, notably Agrobacterium radiobacter, Ampelomyces quisqualis, Beauveria bassiana, Candida oleophila, Gliocladium virens, Lagenidium giganteum, Metarhizium anisopliae, Myrothecium verrucaria (killed), Pseudomonas jluorescens, and Streptomyces griseoviridis, also rose. This is evident in the greater than 12-fold increase in combined sales of products containing these ingredients in California between 1991 and 1998, reflecting both agricultural and nonagricultural uses (DPR sales database). Greater human exposure to these organisms under both occupational and nonoccupational scenarios is a reasonable expectation. Handbook of Pesticide Toxicology Volume 1. Principles

In this chapter, the current regulatory system in the United States for assessing the potential for toxicity, infectivity, and pathogenicity of MPCAs in humans is reviewed. In addition, toxicologic overviews for several prominent or proposed MPCAs are provided. These overviews are directed primarily at toxicity and pathogenicity issues arising from exposure to viable microbial organisms, though individual microbial toxins are considered in some cases. It is hoped that the reader will gain an appreciation for the unique problems confronting regulators as they assess the possibility of human health impacts resulting from the use of MPCAs.

40.2 TOXICITY TESTING REQUIREMENTS FOR MPCAS By 1981, with publication by the World Health Organization (WHO) of an approach to the safety testing of MPCAs (WHO, 1981), it was clear that differences between conventional chemicals and MPCAs required the development of a separate MPCA toxicity testing scheme. Shadduck (1983) outlined the premises upon which conventional chemical testing was (and is) based and provided reasons why these premises were not applicable to MPCAs. Siegel (1997) rearticulated Shadduck's arguments in the following manner. First, high doses of conventional chemicals generate biological effects which are expressed either as overt toxicity or as cellular or organ system responses designed to detoxify and excrete the chemical. With MPCAs it is often impossible to generate such effects without first killing the host by suffocation or by circulatory or gastrointestinal blockage. Second, metabolic and excretion pathways are often predictive of conventional chemical toxicity. In contrast MPCAs are not known to be degraded or genetically altered during passage through the host. Third, persistence and accumulation of chemicals within host organisms necessitates long-term testing for chronic effects. In general, MPCAs have not been shown to colonize mammals nor to produce chronic

859


860

CHAPTER 40

Mammalian Toxicity of MPCAs

Table 40.1 Total Pounds of MPCA Active Ingredients Applied per Year in California, 1990-1999 1 Year Active ingredient

1990

1991

1992

1993

1994

Agrobacterium radiobacter

0.10

0.00

0.41

2.20

3.84

Ampelomyces quisqualis

1995

1996

1997

1998

1999 6.44

6.03

14.0

27.8

19.8

0.42

2.59

9.05

39.7

17.9

1300

4890

2210

Bacillus sphaericus, serotype H-5A5B, strain 2362 Bacillus subtilis GB03

0.00

0.03

0.03

62,400

72,000

85,900

65,300

257

15,600

12,500

12,800

34,500

73,700

47,800

36,400

Beauveria bassiana, strain GHA

0.53

573

1240

915

Candida oleophila, isolate 1-182

0.00

305

103

55.0

Bacillus thuringiensis2

51,400

50,700

53,300

60,900

67,800

80,600

Bacillus thuringiensis, spp. kurstaki, genetically engineered

Encapsulated 8-endotoxin of Bacillus thuringiensis in killed Psuedomonas jluorescens 3

35.0

1820

7960

14,300

Codling moth granulosis virus

14,500

320

Gliocladium virens, GL-21 (spores)

15.3

144

156

104

86.1

Lagenidium giganteum (California strain)

86.9

151

0.10

134

859

499

Metarhizium anisopliae, var. anisopliae, strain ESFl

1.03

0.75

0.20

3.12

36.8

10.5

1100

8500

18,800

0.00

0.00

0.00

3640

3660

2100

Myrothecium verrucaria, dried fermentation solids and solubles Nosema [ocustae, spores

0.00

0.01

0.01

0.05

Pseudomonas jluorescens, strain A506

0.16

206

3040

Pseudomonas syringae, strain ESC-ll

34.00

Pseudomonas syringae, strain ESC-IO Streptomyces griseoviridis, strain K61

0.14

21.3

Trichoderma harzianum, Rifai strain KRL-AG2

Total pounds applied

51,400

50,700

55,100

68,900

82,200

95,800

15.4

0.01

0.05

1.42

1.74

4.90

1.90

64.5

39.2

60.3

122

100,000

169,000

166,000

139,000

1Data are from the Pesticide Use Report (DPR, 2000). The figures represent the amounts applied predominantly under agricultural conditions. Where no figure appears, there was no reported usage. Where "0.00" appears, it is the result of rounding off very small reported usages. 2The figures for B. thuringiensis include the following subspecies and strains: subspecies aizawai (GC-91 protein), subspecies aizawai (serotype H-7), subspecies aizawai [strain SD-1372, lepidopteran active toxin(s)], subspecies israelensis (serotype H-14), subspecies kurstaki (serotype 3A,3B), subspecies kurstaki (strain EG 2348), subspecies kurstaki (strain EG 2371), subspecies kurstaki (strain SA-l 1), subspecies kurstaki (strain BMP 123), subspecies kurstaki (strain HD-l), and subspecies san diego. 3The figures for encapsulated 8-endotoxin of B. thuringiensis in killed P. jluorescens include endotoxins from subspecies kurstaki and san diego.

effects. (Two caveats should be noted here. Some MPCAs are capable of persisting within mammals for longer than a few days without mUltiplying. This necessitates careful examination of their host clearance pattern, which would allow persistence to be distinguished from active infection (Siegel and Shadduck, 1990a). Also, viral agents targeted at mammalian pests present unique problems due to their mammalian host ranges (see the discussion of rabbit hemorrhagic disease virus, Section 43.3.3.2». Fourth, structure-activity relationships, which are often applicable to conventional chemicals, are not relevant to MPCAs.

The WHO testing scheme for determining the toxicity of MPCAs was based on four principles (WHO, 1981): (1) MPCAs pose "inherently different" risks to humans than conventional pesticides; (2) findings of minimal or no toxicity in laboratory testing ("negative results") are likely; (3) tiered testing, wherein negative results at one level preclude testing at higher levels, is appropriate; and (4) testing protocols should maximize the possibility of generating adverse effects in the host organism. This approach allowed for a more expedient registration process than that in effect for conventional chemicals. For the great majority of agents, which show negative results under the short-term

40.2 Toxicity Testing Requirements for MPCAs

Tier I requirements, longer term and more expensive studies were avoided. In 1982, the United States Environmental Protection Agency published its "Pesticide Assessment Guidelines, Subdivision M: Guidelines for Testing Biorational Pesticides" (U.S. EPA, 1982). This document, which was revised in 1989, incorporated WHO's philosophy and testing schemes into a system that remains in force today in the United States. A brief discussion of the MPCA testing requirements is presented in the following paragraphs. More complete treatments have appeared elsewhere (Betz et aI., 1990; McClintock, 1999; McClintock et aI., 1995; Siegel, 1997; Siegel and Shadduck, 1992). Few, if any, MPCA candidates have been carried beyond Tier I testing. 40.2.1 TIER I

Under Subdivision M, Tier I, three acute systemic exposure routes, oral, pulmonary, and intravenous (intraperitoneal for larger microbes), are required for unformulated MPCAs or for the technical grade active ingredient. Because aerosolization of viable microorganisms is problematic, intratracheal dosing is often used as a surrogate for pulmonary exposure by the inhalation route. Both the intravenous-intraperitoneal and the intratracheal routes are more invasive than those routes generally required for conventional chemicals, fulfilling the WHO principle of maximizing the likelihood of adverse effects. For each of these tests, between 107 and 108 colony-forming units (cfu), or the highest obtainable dose, is administered to mice or rats. The animals are monitored over a 4-week period for mortality, clinical signs of morbidity, body weight, gross pathology, and microbial clearance. Clearance is a measure of the ability of the host to remove invading microorganisms over time; as such, it is an indication of the presence or absence of an active infection (McClintock et aI., 1995). It is usually assessed by culturing the MPCA from homogenates of various organ systems, blood, and excretory products at established time intervals after dosing. Colony-forming units are enumerated and a pattern of clearance established. Although complete clearance can be demonstrated within a few days for most MPCAs (a result not unexpected because these organisms are rarely adapted for life under mammalian body conditions), clearance can take 50 days or more for persistent organisms. In such cases it is considered sufficient to demonstrate a clearance pattern and to show that the organism does not produce an active infection which can colonize and multiply within the host. In contrast to the first three routes, dermal toxicity test guidelines recommend use of the manufacturing use and formulated end use products, applied at 2 g per kilogram of body weight. This allows for an evaluation of the potential for local irritation, as well as that for systemic toxicity by this exposure route. As is often the case with conventional pesticidal products, it is the formulation components, not the active ingredients, which drive irritation reactions. In addition, there is another, though perhaps minor, consideration. Although it is unlikely that the vast majority of MPCA candidates would penetrate the dermal

861

barrier when applied as aqueous pastes or suspensions, such an event may occur when formulation components with high dermal permeability are also present. Tier I also requires testing for ocular irritation. As for dermal irritation, the formulation components are likely to play a primary role in this process; hence the requirement that the manufacturing and end use products be tested. Clearance determinations are not required for either dermal or ocular exposures. Nonetheless, putative MPCAs could conceive establish themselves at least temporarily in the eyes. Ocular applications of B. sphaericus and B. thuringiensis subsp. israelensis in the rabbit eye led to detections at that site for as long as 8 weeks postdosing (Siegel and Shadduck, 1990a). Cell culture tests are required only in the case of viral pest control agents; other classes of MPCAs are not likely to initiate infections of individual cells. A number of tests using both primary mammalian cell cultures and established mammalian cell lines are necessary to evaluate the potential toxicity and infectivity of the form of the virus considered to be most infective in susceptible cell cultures or in whole organisms (e.g., insects). These include a plating efficiency test, an infectivity evaluation, and a cell morphologic transformation assay. The latter assay, done specifically in Syrian hamster embryo cells, would be extended to an examination of viral tumorigenicity in hamsters if the MPCA proved capable of morphologically transforming cells in culture. Details of the specific cell culture tests are provided in McClintock (1999). Finally, Tier I requires testing for hypersensitivity if common use practices lead to repeated dermal or inhalation exposure. However, this requirement is generally waived because injection hypersensitivity studies using MPCAs are expected to be positive, whereas topical exposure studies are expected to be negative. Reports of hypersensitivity incidents occurring during manufacture or testing of MPCA products are nonetheless required (McClintock, 1999). 40.2.2 TIER 11

When acute toxicity is observed in the absence of pathogenicity and infectivity by the oral, dermal, or pulmonary exposure route, an acute LDso study is undertaken. Those exposure routes that produced toxicity in Tier I testing are used to establish the median lethal dose and slope after a 14-day postdose observation period. A subchronic study is required as well; the test article is administered daily for at least 90 days at a dose of at least 108 cfu/animal/day. The animals are monitored throughout for toxicity and pathogenicity-infectivity. In addition, organs, tissues, and body fluids are assayed for the presence of the microorganism. 40.2.3 TIER III

If significant toxicity in Tier 11 studies is observed, or if an ability to overcome natural host barriers to infection is detected, Tier III studies may be necessary for the registration process

862

CHAPTER 40


to be continued. Studies designed to detect effects on reproduction, fertility, oncogenicity, immunodeficiency, and primate infectivity may be called for based on the specific signs noted in Tiers I and 11. However, it is unlikely that organisms which show signs of significant toxicity or mammalian infectivity, defined as the ability to multiply within the host, will be carried beyond Tier I testing. Organisms targeted against mammalian pests clearly present a conundrum to this testing scheme, though it appears that none have yet been considered for registration in the United States.

40.3 TOXICITY OF INDIVIDUAL MPCAS The following discussion focuses on specific toxicity issues pertaining to selected prominent microbes which either are in use or have been considered for use as MPCAs. 40.3.1 BACTERIA 40.3.1.1 Bacillus Thuringiensis Bacillus thuringiensis is the most well-known and widely used of all pesticidal microbes. This gram-positive, spore-forming, facultative soil saprophyte was first registered for use in the United States in 1961 (U.S. EPA, 1998), after having been granted a temporary tolerance exemption for use in food and forage crops in 1958 (Fisher and Rosner, 1959). Bacillus thuringiensis subspecies have shown specificity against various orders of insects, including dipterans (B. thuringiensis subsp. israelensis), lepidopterans (B. thuringiensis subspp. kurstaki and aizawai), and coleopterans (B. thuringiensis subsp. tenebrionis). Recent B. thuringiensis isolates are active against nematodes, mites, and protozoa, as well as against other insect orders (Schnepf et aI., 1998). The entomopathogenic activity is primarily based on production during the stationary growth phase of a parasporal protein crystal. The crystal is composed of "Cry" (for "crystal") and, at least in B. thuringiensis subspp. israelensis and morrisoni (among currently commercially relevant B. thuringiensis subspecies) , "Cyt" (for "cytolytic") proteins. Knowledge of the identity, specificity, and structure of these "o-endotoxins" has expanded enormously over the past two decades. The coding sequences are known for over 100 of the relevant genes (Schnepf et aI., 1998). The Cry protoxin is activated by solubilization and proteolytic cleaveage under the alkaline gut conditions prevalent in susceptible insects. The resultant protein causes larval death by receptor-mediated lysis of the midgut epithelium (McClintock et aI., 1995; Schnepf et aI., 1998). Most Cry genes code for proteins in the 65-138 kDa range, with size at least partially dependent on the strain pathotype (Beegle and Yamamoto, 1992; Drobniewski, 1994). Differences in insect toxicity may be a function of different Cry solubilities in the insect gut, in addition to different inherent characteristics such as receptor affinity (Schnepf et aI., 1998). In some cases the spores can contribute to the insecticidal ac-

tivity of the parasporal crystal proteins, though how this comes about is unclear (Beegle and Yamamoto, 1992). The Cyt toxins are hemo1ytic and cytolytic proteins, with protoxin molecular weights in the 25-28 kDa range (Drobniewski, 1994). Cyt proteins do not exhibit sequence homology with Cry proteins (Hofte and Whiteley, 1989). They appear to disrupt insect cell membranes through detergent-like effects (Butko et al., 1997) and/or through the formation of cation-selective channels (Drobniewski, 1994). Thomas and Ellar (1983) found that intravenous injection of solubilized parasporal crystal proteins from B. thuringiensis subsp. israelensis was toxic to mice, in contrast to the lack of toxicity upon injection of a similar preparation from B. thuringiensis subsp. kurstaki. This was probably due to the presence of a 28-kDa Cyt protein in the former preparation. The toxicity of the isolated 28-kDa protein from B. thuringiensis subsp. israelensis was subsequently verified by intraperitoneal injection into mice (Mayes et aI., 1989). Interestingly, neither the solubilized kurstaki nor israelensis preparations provoked a toxic response in mice by the oral route (Thomas and Ellar, 1983). Several other B. thuringiensis molecules deserve mention. The ,B-exotoxin (thuringiensin), a heat-tolerant nonproteinaceous compound which is toxic to houseflies, mammals, and other nontarget organisms, has been demonstrated in B. thuringiensis subsp. thuringiensis and in one B. thuringiensis subsp. aizawai strain (U.S. EPA, 1998). ,B-Exotoxin works by inhibiting DNA-dependent RNA polymerase (Beegle and Yamamoto, 1992). For purposes of registration it is necessary to demonstrate the absence of ,B-exotoxin in B. thuringiensis formulations (McClintock et aI., 1995). In addition, a proteinaceous, heat-labile, insecticidal a-exotoxin with a molecular weight in the 45-50 kDa range has been identified (Beegle and Yamamoto, 1992). a-Exotoxin has properties similar to the 50-kDa enterotoxin of B. cereus (see Section 40.3.1.2). Finally, insecticidal activity can be enhanced by the expression of other proteinaceous toxins, among them phospholipases, proteases, chitinases, and secreted vegetative insecticidal proteins (Schnepf et aI., 1998). Extensive toxicologic testing of intact commercial B. thuringiensis strains has not resulted in appreciable toxicity, pathogenicity, or infectivity (U.S. EPA, 1998). In one early study, ingestion by humans of 3 x 109 B. thuringiensis spores/day (the subspecies was not identified) for 5 days, or by rats of 2 x 10 12 spores/kg produced no toxicity (Fisher and Rosner, 1959). Exceptions may occur when noncommercial subspecies, unusual exposure routes, or, as indicated previously, the use of isolated toxins as opposed to whole organisms are examined. Intracerebral exposure of laboratory rats is lethal if sufficient numbers of organisms are injected (Siegel and Shadduck, 1990b), though human exposure by this route is unlikely. Warren et al. (1984) reported an incident in which local and lymphatic inflammation requiring antibiotic therapy occurred when a laboratory worker sustained an accidental injection with spent medium containing B. thuringiensis subsp. israelensis and Acenitobacter calcoaceticus var. anitratus. In this case, A. calcoaceticus, a skin-dwelling bacterium, could have provided the proteases

40.3 Toxicity of Individual MPCAs

necessary to activate the protoxin, either by releasing them into the spent medium or into extracellular fluids at the injection site. It was, nonetheless, unclear to what extent the pathology was due to intoxication and to what extent it was due to bacterial persistence or infection with an accompanying inflammatory response from the host. Isolated health concerns pertaining to noncommercial B. thuringiensis strains have occasionally surfaced. Bacillus thuringiensis subsp. konkukian (serotype H34) was detected in the wounds of a French soldier injured by a land mine explosion (Hernandez et aI., 1998). The ability of B. thuringiensis subsp. konkukian (serotype H34) to cause tissue damage was demonstrated by cutaneous application of bacterial suspensions isolated from the wounds to normal and immunosuppressed mice. Inflammatory lesions developed in all mice treated with 107 cfu. These healed spontaneously in the normal animals, but progressed in the immunosuppressed animals. In a study reported in the Russian literature, B. thuringiensis subsp. galleriae was shown to cause syndromes in humans similar to those found in B. cereus-related food poisoning (Pivovarov et aI., 1977). Bacillus thuringiensis with cytotoxic characteristics similar to enterotoxin-producing B. cereus (see Section 43.3.1.2) was isolated from stools in a gastroenteritis outbreak occurring in a Canadian chronic care institution (Jackson et aI., 1995). Although the B. thuringiensis subspecies was not identified, the expression of B. cereus traits is not surprising, as some consider B. thuringiensis and B. cereus to be variants of the same species (Schnepf et aI., 1998). The production of Bacillus diarrheal enterotoxin was demonstrated in various commercial preparations of B. thuringiensis by Damgaard (1995), though the levels were generally low compared to those found in a reference culture of B. cereus that had been isolated from a food poisoning outbreak. It was noted, however, that a role for B. thuringiensis in food poisoning may be underestimated due to the requirement for a special staining technique to differentiate B. thuringiensis from the more conventionally assayed B. cereus. Under field conditions, reports of clinically significant symptoms in humans have been rare considering the length of time that B. thuringiensis has been in use as a pesticide. In one case, a farmer developed a corneal ulcer containing B. thuringiensis after being splashed in the eye with a commercial B. thuringiensis product (Samples and Buettner, 1983). A survey of farm workers exposed to commercial B. thuringiensis subsp. kurstaki sprays failed to identify clinical syndromes in the eye, respiratory tract, or skin (Bernstein et aI., 1999). However, positive skin-prick allergy tests and induction of IgG and IgE antibodies were documented in some exposed individuals, suggesting that allergenicity could result from repetitive exposure. Exposure of immunocompromised individuals may pose a special challenge. This was noted in an epidemiologic study conducted in an area of Oregon that had undergone spraying with B. thuringiensis subsp. kurstaki (Green et aI., 1990). Bacterial cultures from patients undergoing routine exams revealed 55 that were B. thuringiensis-positive. Bacillus thuringiensis was ruled out as a specific pathogen in 52 of those patients, but was not ruled out in the remaining 3 cases. Although all of these

863

latter infections may have been opportunistic, having occurred in people with established medical conditions, exacerbation of preexisting symptoms or development of new symptoms as a result of B. thuringiensis exposure was not ruled out. Despite the documented absence of mammalian toxicity in in vivo testing by most economically important B. thuringiensis strains, the ability of these bacteria to persist for extended periods within mammals after injection or intratracheal administration (McClintock et aI., 1995; Siegel and Shadduck, 1990b) has occasioned concern, particularly in light of the close phylogenetic relationship between B. thuringiensis and other medically significant Bacillus species such as B. cereus, B. anthracis, and B. sphaericus. For example, changes in the microenvironment could lead to increases in production of B. cereuslike enterotoxin. Siegel and Shadduck (1990a, 1992) discussed at length their operational definition of infectivity, which requires demonstration of multiplication of the microorganism within the host and disruption of functional or structural homeostasis. This contrasts with their definition of persistence, in which an organism is present either in a multiplying or quiescent state, but does not disrupt the host. Bacillus thuringiensis generally shows a consistent, though in some cases prolonged, clearance pattern (McClintock et aI., 1995). However, in one study using B. thuringiensis subsp. israelensis, splenic bacterial counts following intraperitoneal injection into mice showed no tendency to decline even after 80 days (Siegel and Shadduck, 1990a). This was interpreted as evidence that multiplication had occurred within the host, though there was no sign of toxicity. Combined with the record of safe use, a conclusion that persistence in animals does not translate into toxicity is warranted. 40.3.1.2 Bacillus Cereus Bacillus cereus is a gram-positive, catalase-positive, rod-shaped saprophyte that is closely related to, if not con specific with, B. thuringiensis. Bacillus cereus does not produce the parasporal inclusion body which is so important to the entomopathogenic activity of B. thuringiensis. One B. cereus strain, BP01, is currently registered in the United States as a plant growth regulator. Bacillus cereus is interesting from a medical standpoint because it is implicated in pathologies of the lung, ear, eye, gall bladder, and urinary tract and as an opportunistic invader in trauma or disease cases (Drobniewski, 1994; Goepfert et aI., 1972). It also is reported to have been an opportunistic pathogen in several cancer patients (Banerjee et aI., 1988) and is an epidemiologically significant food-based pathogen. Two food poisoning syndromes, a diarrheal type, associated with consumption of a diversity of food types, and an emetic type, most commonly associated with rice and pasta consumption, are caused by B. cereus (Drobniewski, 1993; Logan and Turnbull, 1999). A heat-labile enterotoxin complex, consisting of two or three protein components with molecular weights of 38-57 kDa, appears to be responsible for the diarrheal symptoms. A heatstable 10-kDa peptide is associated with the emetic symptoms.

864

CHAPTER 40


Serotyping has identified the B. cereus strains most likely to produce the two syndromes. Other factors, notably hemolysins and phospholipases, are involved in the establishment of local and systemic infections (Drobniewski, 1993). In view of the medical and epidemiologic significance of B. cereus, vigilance with respect to the dissemination and application of pesticidal products containing this organism is warranted, as well as strict attention to proper strain identification. Demonstration of an inability to produce enterotoxins or emetic toxins also may be useful in the assessment of potential risks posed by proposed commercial B. cereus strains. 40.3.1.3 Bacillus Sphaericus

Several isolates of B. sphaericus, a crystalliferous spore-former similar to B. thuringiensis, have shown promise as mosquito larvicides (Saik et aI., 1990), making them potentially useful for controlling tropical diseases such as malaria and filariasis (Murthy, 1997). However, B. sphaericus has provoked human health concerns. It was implicated in several cases of meningitis (AlIen and Wilkinson, 1969; Siegel and Shadduck, 1990b), in the formation of a lung pseudotumor (Isaacson et aI., 1976), and as an opportunistic pathogen in a cancer patient (Banerjee et aI., 1988). Interestingly, B. sphaericus isolates from the meningitis patients and from the patient with the pseudotumor did not prove pathogenic in rabbits and mice exposed intraperitoneally or intravenously, or in mice after intracerebral exposure (Siegel and Shadduck, 1990b), raising questions about the applicability of animal testing to human pathology in this case. It is worth noting, however, that B. sphaericus is a species complex, divisible into six groups based on DNA homology, with larvicidal strains clustering in homology group IIA (Krych et aI., 1980). Even among entomocidal strains, the particular B. sphaericus strain designation, exposure route, and host animal strain continue to be relevant when analyzing the toxicity of these organisms in mammalian systems. A more detailed examination of two particular studies may serve to illustrate this point. Shadduck et al. (1980) investigated the pathogenicity in mice, rats, and rabbits of three entomocidal B. sphaericus strains, SSII-1, 1404-9, and 1593-4, after exposure by various routes. Neither death nor systemic illness resulted under any exposure scenario. Conjunctival instillation into rabbits of up to 109 infectious units (iu) of B. sphaericus caused local lesions which were severe at the higher doses. Less severe lesions developed after instillation of autoclaved bacteria, suggesting that a portion of the pathologic response was attributable to the presence of heat-stable foreign material. Generally mild brain lesions occurred in some rats upon intracerebral injection of all three strains. In strain SSII -1 this occurred at a dose as low as 1.2 x 106 iu per rat. Intracerebral hemorrhage was commonly noted in mice upon intracerebral injection of ~3 x 108 iu of any of the three strains, though rabbits similarly treated with 1593-4 did not react in this way. Subcutaneous injection of strain 1404-9 resulted in an abscess in one of five mice exposed to the highest dose, 6.7 x 109 iu per animal. Intraperitoneal injection into rats with 3.2 x 108 iu of strain 1404-9 or

4.7 X 108 iu of strain 1593-4, respectively, were without effect. Viable bacilli were detected in eye cultures 10-14 days after conjunctival instillation into rabbits of as few as 1.2 x 103 iu of strain SSII-1. Persistence in the eye was also detected with injections of 109 iu of strain 1404-9 and 108 iu of strain 1593-4, though lower doses were not tested. Similarly, bacilli were detected in brain cultures 10-14 days after intracerebral injections of as few as 108 iu per rat of strain 1593-4, 1.2 x 108 iu of strain SSII -1, and 3.2 x 108 iu of strain 1404-9. Lower concentrations were not tested for strains 1404-9 or SSII-1. These results indicate that the bacterium can persist in mammals, at least under certain conditions. Whether or not bacterial replication occurred was not explored with respect to the ocular exposure route. However, a pattern of cerebral clearance was established after intracerebral injection of 5.5 x 105 iu of strain 1593-4 into rats. Detections of greater than 600 iu per 100 mg wet brain tissue were noted on day 3 postinjection. No more than 30 iu per 100 mg tissue were observed at various times between days 5 and 12. The brains were considered sterile by day 14. In a follow-up study, conjunctival instillation into rabbits of 4.48 x 108 cfu 1 B. sphaericus strain 2362 did not cause local toxicity (Siegel and Shadduck, 1990a), contrasting with the results of the earlier study using other strains. Nonetheless, culturable bacilli were recovered 8 weeks after treatment (longer recovery times were not examined). A clear pattern of splenic clearance was established after intraperitoneal injection of 1.2 x 107 cfu into mice, though some bacilli (165 cfu per gram of spleen) were still present at study termination on day 67. Intraperitoneal injection of 8 x 108 cfu resulted in the death of 42 of 49 mice within 24 hours. Interestingly, injection of 3.8 x 108 cfu per animal autoclaved strain 2362 resulted in the deaths of 3/6 mice between 24 and 48 hours postinjection. These results were surprising in light of the observation of no toxicity by the intraperitoneal route of B. sphaericus strains SSII-1, 1404-9, and 1593-4 in the earlier study (Shadduck et aI., 1980). Production of a soluble extrabacterial toxin was ruled out as a cause because passage through cellulose acetate filters removed the toxicity. The authors speculated that the strain of mouse (outbred CD-I) could have been particularly sensitive to the effects of strain 2362, citing a study from a different laboratory showing no effect of this strain in Swiss mice. Alternatively, they considered that the particular culture conditions of the bacterium, which they did not know, could have generated a more lethal bacterial isolate. In any case, they noted that these were extremely high doses which are not likely to be relevant to human exposures in the field. The possibility that B. sphaericus has unique pathogenic properties (e.g., it may be infective only in health-compromised individuals) which are not evident in conventional animal testlThe designation "colony forming units" (cfu) was used in the later study (Siegel and Shadduck, 1990a) to signify that the bacterial titers were quantitated by culturing appropriate dilutions of the inocula in agar and counting the resultant bacterial colonies at a later time. "Infectious units" (iu) was used in the earlier study (Shadduck et aI., 1980) because titers were quantitated by serial tube dilution; determination of infectious units was by turbidity in brain infusion broth.


ing cannot yet be excluded. Certainly the direct impacts of entomocidal B. sphaericus strains on human populations should be monitored until sufficient evidence for their safety is obtained under conditions of actual pesticidal use. 40.3.1.4 Burkholderia Cepacia

A gram-negative, nutritionally versatile, and highly antibioticresistant bacterium, B. cepacia has stimulated economic interest due to its ability to inhibit soil-borne plant pathogens and to degrade hydrocarbons associated with sites of environmental contamination (Parke, 1998). It also causes serious opportunistic infections in humans suffering from chronic granulomatous disease and cystic fibrosis (CF) (Butler et al., 1995). As many as 40% of patients in some CF centers develop B. cepacia infections, with 35% of those patients exhibiting "cepacia syndrome" characterized by grave pulmonary pathogenesis, bacteremia, and death (Holmes et al., 1998). Other figures are perhaps less alarming, but serious nonetheless. LiPuma (1998) cited respiratory culture results from 1996 showing that 3.6% of cystic fibrosis patients showed evidence of infection. Of those infected, 20% "succumb to a rapidly progressive necrotizing pneumonia." Burkholderia cepacia also is implicated in nosocomial infections of non-CF patients, as well as in the "foot rot" syndrome experienced by soldiers in swampy terrain (Holmes et al., 1998). The primary mode of transmission to CF patients appears to be from other CF patients, though transmission from non-CF patients is also possible. Social isolation measures have been necessary in some circumstances (WaIters and Smith, 1993), but these have been accompanied by poor psychosocial outcomes (Butler et al., 1995; LiPuma, 1998). Burkholderia cepacia strains are divided among five genomovars, collectively known as the B. cepacia complex. According to the analysis of Vandamme et al. (1997), genomovars 11 and III are the best represented among CF patients. However, they caution that a systematic study of genomovar distribution has not yet been done, nor is the relative significance for cepacia syndrome of the various strains yet understood. Representatives of all genomovars have been detected in CF patients (Vandamme et al., 1997). Attempts have been made to identify markers associated with epidemic transmission of B. cepacia among CF patients. Such markers could simplify the risk assessment process for proposed B. cepacia pesticidal strains. Cable pili, peritrichous appendages that facilitate binding to CF mucin and airway epithelial cells (Goldstein et al., 1995; Sajjan and Forstner, 1993), may comprise one marker phenotype. These structures were identified in an epidemic strain transmitted among CF patients in clinics in Toronto and Edinburgh (Sun et al., 1995). Another candidate, a l.4-kb DNA fragment known as the "B. cepacia epidemic strain marker" (BCESM), was identified in seven epidemic strains of the bacteria, but was not present in nonepidemic strains (Mahenthiralingam et al., 1997). Unfortunately, the presence of neither BCESM nor of cable pili is considered a certain indicator of transmission potential. The cable pilus gene was detected in only one of the seven epidemic strains

865

examined in the Mahenthiralingam study. And as pointed out by LiPuma (1998), the presence of nonepidemic strains in the respiratory tracts of CF patients demonstrates their colonizing ability even as they lack BCESM. In addition to these markers, other virulence factors and epidemic strain-associated markers have been considered. At this point, it appears that no single marker will provide absolute predictive ability for clinically important B. cepacia strains. However, they may provide some initial screening capability when considering potential pesticidal strains. George et al. (1991) investigated the effect on CD-1 mice of intranasal instillation of 5.3 x 108 cfu B. cepacia strain AC1100. Ruffled fur, weight loss, and inactivity were noted during the first two days following treatment, but recovery was evident thereafter. Increased lung weights, apparent between days 2 and 14 (study termination), were attributed to macrophage influx and endotoxin-mediated edema. Declining numbers of B. cepacia were evident through day 7 in the lungs and through day 2 in the nasal cavity. Burkholderia cepacia was also present in the gastrointestinal tract for the first two days after treatment. This was attributed to mucociliary evacuation from the lung to the mouth and thence to the stomach, with bacterial survival afforded by the mucus coating acquired in the respiratory system. In a later study (George et al., 1999), endotoxinresistant C3H1HeJ mice were subjected to intranasal instillation with B. cepacia strain ATCC 25416. Lethality was observed at as low as 2.2 x 108 cfu per mouse, with an approximate LDso of 7 x 108 cfu per mouse. Although this appears inconsistent with the relative lack of effect in mice seen at a similar dose in the 1991 study, the considerable splay evident in the mortality data should be recognized, as well as the fact that different bacterial and mouse strains were used. Other differences from the earlier study were evident. In the 1999 study, no changes in lung weights were detected through 14 days posttreatment, possibly reflecting the endotoxin-resistant status of the mouse strain used in that study or the lower bacterial dose applied (~107 cfu per mouse vs. 5.3 x 108 cfu per mouse in the earlier study). Also in contrast to the 1991 study, B. cepacia appeared to be stably established in the lungs, small intestine, cecum, and large intestine through study termination on day 14. Moreover, B. cepacia was cultured from the liver and spleen at 3 hours, and from the mesenteric lymph nodes through 10 days. These data imply that, under some circumstances, B. cepacia could gain a more tenacious hold in a mammalian system. Subdivision M testing of B. cepacia isolate M36, submitted for registration as an active ingredient to the State of California, indicated that clearance had occurred through the feces within 7 days of oral dosing of rats with 2.85 x 108 cfu. However, severe fibrous adhesions between pleural surfaces of the thoracic cavity, lungs, and pericardial sac were noted in 1/3 males and grey lung coloring in 2/3 females (DPR, 1994a). Intratracheal dosing of 1.9 x 108 cfu per rat resulted in pulmonary clearance by day 22, with lung discoloration evident through day 8 (DPR, 1994b). Infectivity of this strain by all routes of exposure was discounted because there was no evidence of multi-

866

CHAPTER 40 Mammalian Toxicity of MPCAs

plication. Significantly, use of M36 has been eliminated due to the presence of BCESM. Use of other pesticidal strains currently are restricted to agricultural applications such as seed treatment and drip irrigation to minimize exposure to susceptible populations (D. Gurian-Sherman, U.S. EPA, personal communication). Finally, because of its large and highly adaptable genome, there is concern that pathogenic strains of B. cepacia could be generated through gene transfer or recombination if large numbers of putative nonpathogenic organisms are artificially introduced into the environment (Holmes et aI., 1998). For this reason, cystic fibrosis advocacy groups have expressed serious misgivings about the registration of B. cepacia products before adequate testing and assurance of nontransformability is available (PTCN, 1997).

40.3.2 FUNGI

40.3.2.1 Metarhizium Anisopliae In use since the late 1800s, M. anisopliae is a Deuteromycete (Fungi imperfecta) used in the United States largely for the control of cockroaches, though it also is effective against other orthopterans and against coleopterans. It has a wide geographic distribution, existing as an insect or nematode parasite, or in various soils, sediments, spoil heaps, and other environments (Domsch et aI., 1980). Death of the host insect results when contact with conidia, the environmentally stable asexual spore stage, leads to infection. This is followed by enzyme-mediated exoskeletal degradation, mycelial development, and sporation (Ward et aI., 1998). Massing of conidia in affected insects lends a characteristic green color, hence the name "green muscardine" for the insect disease (Ferron, 1981). Insecticidal activity also may reside in a family of cyclodepsipeptides known as destruxins, 15 of which were identified in M. anisopliae as of 1989 (Gupta et aI., 1989), as well as in other toxic substances (Domsch et aI., 1980). Standardized laboratory studies have not demonstrated toxicity or infectivity of M. anisopliae in rats, mice, or rabbits, though persistence without multiplication was reported (Siegel and Shadduck, 1990b). Allergenicity was demonstrated in mice following intraperitoneal sensitization and intratracheal challenge with crude protein extracts of mycelia and conidia (Ward et aI., 1998). Although applicability of the mouse model to humans exposed solely via inhalation was not experimentally addressed, evidence for allergy under occupational situations has been reported (Kaufman and Bellas, 1996). A single case of keratomycosis, an increasing problem with fungi in general due to increased use of antibacterial drugs, immunosuppressants, and corticosteroids (Ishibashi et aI., 1986), was reported in an 18-year old man (Cepero de Garcia et al., 1997). A single case of hyphomycotic rhinitis was reported in a cat (Muir et aI., 1998). Disseminated infection with severe morbidity was reported in an immunocompromised child (Burgner et aI., 1998), highlighting the need for great caution where exposures of sick individuals are possible.

40.3.2.2 Beauveria Bassiana Beauveria bassiana, a Deuteromycete long known for its entomopathogenic properties, causes an insect disease known as white muscardine. The organism produces a number of cyclodepsipeptides such as beauvericin which may account for at least part of its insect toxicity (Miller et al., 1983). Beauveria has been used as a medicant in Japan for over a millenium (Ignoffo, 1973). Allergic responses have been reported in humans following inhalation of spore preparations, though repeated handling of cultures did not reveal adverse effects in another study (Ignoffo, 1973). A study from China noted hypersensitivity-like pulmonary reactions in mice and rats after a single exposure to B. bassiana. However, the low room temperatures may have constituted a significant stress to the animals (Song et al., 1989, cited in Semalulu et al., 1992). Russian investigators have reported the LDso to be greater than 1.1 x 1010 and greater than 2.2 x 1010 fungal cells in albino rats exposed intragastrically and intraperitoneally, respectively, and greater than 4 x 1010 fungal cells in rabbits exposed intravenously (Mel'nikova and Murza, 1980). Beauveria bassiana has been implicated in at least two cases of keratomycosis, though both patients had long histories of antibiotic and corticosteroid use (Ishibashi et al., 1986). A study in the French literature isolated organisms from the genus Beauveria from an individual with bronchopneumonia (Freour et al., 1966, cited in Semalulu et al., 1992). Direct inoculation into rabbit corneas of B. bassiana isolated from a patient with keratitis resulted in inflammation, corneal ulcers, corneal haze, injection of the iris, and sparse-to-moderate fungal growth in the cornea, though the severity was less than that seen in parallel eyes treated with Candida albicans and tended to resolve itself with time (Ishibashi et al., 1986). Injection of B. bassiana into the quadriceps muscles of CD-1 mice led to focal muscle necrosis, edema, and inflammation, with the severity of the responses dependent on the number of organisms injected (Semalulu et aI., 1992). Muscle regeneration was visible by 7 days. Viable spores capable of initiating colonies in artificial media were not detected after 3 days. Thus it is unlikely that this organism, which does not grow at temperatures above 32 QC, can infect or colonize mammals under normal circumstances.

40.3.2.3 Gliocladium Wrens Gliocladium virens is a common, soil-dwelling saprophyte that is useful in controlling Pythium ultimum and Rhizoctonia solani, organisms that cause damping-off disease in greenhouses. Such antifungal activity may be due partially to production of gliotoxin, a relatively nonselective antibiotic that is also immunosuppressive and moderately toxic to mammals'(Lumsden and WaIter, 1995). However, studies using G. virens isolate GL-21 did not reveal unusual toxicity, though conjunctival irritation for at least 72 hours in the rabbit eye irritation study and death by capillary obstruction in the rat intravenous study were noted (DPR, 1993). The potential for G. virens-induced allergenicity is not known.


40.3.2.4 Lagenidium Giganteum The ability of this aquatic saprophyte to parasitize, and eventually kill, mosquito larvae demonstrates its potential usefulness in mosquito control programs. Conventional toxicity testing with an organism as large as L. giganteum, which can produce cells greater than 200 ).tm in length, was impossible because intratracheal instillation of as few as 1.16 x 105 oospores per rat resulted in the prompt death of many of the treated animals from acute pneumonia, airway obstruction, or severe inflammation (Siegel and Shadduck, 1987). A similar result was obtained upon intravenous injection of 1.78 x 106 cfu into mice, where embolism killed several animals within 36 hours of treatment (Kerwin et al., 1990). Lowering the numbers either of active or autoclave-inactivated organisms still resulted in tissue damage after intratracheal or intraperitoneal exposures (Siegel and Shadduck, 1987). In all likelihood, these lesions represented local inflammatory responses to large amounts of foreign biological material. Although there was histologic evidence for persistence, multiplication within the mammalian hosts did not occur. Oospore treatment of rat skin or rabbit eyes did not result in irritation. The potential for allergenicity remains untested. 40.3.3 VIRUSES Interest in the use of viruses as pest control agents is based on their promise as specific vectors, either in native form or as genetically engineered constructs designed as delivery agents for pesticides or as immunocontraceptives for mammalian wildlife control. Establishment of the range of target species susceptible to infection by a given virus is perhaps the major issue in the assessment of risks associated with use of that virus as a pest control agent. For obvious reasons, human health concerns are magnified when the virus in question is infective in, or pathogenic toward, mammalian species. 40.3.3.1 Baculoviruses Baculoviruses are double-stranded DNA viruses that are highly specific pathogens of insects and other arthropods (Huber, 1995). Two major groups of baculoviruses, the nucleopolyhedrosis viruses (NPVs) and the granuloviruses (GVs), are enveloped by proteinaceous occlusion bodies that protect the virions from adverse environmental impacts (Saik et al., 1990). Infection in susceptible insects occurs when the occlusion bodies are ingested from leaf surfaces and dissolved under the alkaline conditions of the insect gut (Black et aI., 1997). Death of the host results from the multiplication of freed virions within bodily tissues (Huber, 1995). Based on studies both in humans and in laboratory animals, human safety concerns appear to be minimal. Dietary consumption over a 5-day period of 6 x 109 Helicoverpa zea NPV polyhedra was without effect (Heimpel and Buchanan, 1967), as was occupational exposure to the same organism over a 26month period in a virus production facility (Huang et aI., 1977). Exposure of mice and guinea pigs to H. zea NPV inclusion

86"

bodies, virus rods, and polyhedral protein by several routes (inhalation, oral gavage, and intradermal, intraperitoneal, intracerebral, and intravenous injection) revealed no effects (Ignoffo and Heimpel, 1965). Similar negative results were obtained using the Trichoplusia ni NPV (Heimpel, 1966). No unusual responses were detected in rats following acute subcutaneous injection of 1.2 x 109 polyhedral inclusion bodies (PIB) of H. zea NPV into neonates or acute intravenous injection of the same quantity into adults (Bames et aI., 1970). In addition, no effects were identified with dietary exposures of 90 days or 2 years using feed preparations containing viral loads between 6 x 107 and 6 x 109 PIB per 100 g. Allergenic responses were not detected in guinea pigs after 3 weeks of inhalation exposure at 1 hour/day, 5 days/week, to H. zea NPV free viral rods (obtained from 3 x lOll PIB/day) or intact PIB (3 x lOll/day), or after dermal exposure to 1.5 x 1011 PIB for 5 days followed by intradermal injections of 1.2 x 108 PIB in each of four bodily sites (Meinecke et aI., 1970). Pigs force-fed with Mamestra brassicae larval NPV at a dose level of 5 x 107 polyhedra per gram of body weight showed a slight, transitory rise in body temperature (Doller et aI., 1983). There was, however, no evidence for lymphatic involvement, no effect on leukocyte counts, and no evidence for viral replication or organ infection within the hosts. In one accounting published in 1965, over 26 baculoviruses had been tested in 10 mammalian species without indication of toxicity (Ignoffo and Heimpel, 1965). Indeed, baculoviruses do not appear capable of multiplying within mammalian hosts (Black et aI., 1997). Finally, the ubiquity of baculoviruses in the human food supply attests to their harmlessness (Black et al., 1997; Heimpel et aI., 1973). 40.3.3.2 Rabbit Hemorrhagic Disease Virus Rabbit hemorrhagic disease virus (RHDV) is a single-stranded RNA virus belonging to the calicivirus family. It is linked to a syndrome of necrotising hepatitis, hemorrhage, and death in European rabbits (Oryctolagus cuniculus) that has become known as rabbit hemorrhagic disease. This disease was first recognized in Angora rabbits exported from Germany to China in 1984. It subsequently spread from China to Korea, Europe (including the British Isles), Mexico, Israel, and North Africa (Nowotny et aI., 1998). Efficacy of the virus is not completely understood and may depend on such host factors as rabbit density, viral resistance, and age distribution, as well as nonhost factors, including the presence or absence of vertebrate or invertebrate vectors and the quality of the virus preparations used (Parkes et aI., 1999). Nonetheless, RHDV has been purposefully spread in Australia, where introduction of European rabbits in the 19th century left the island continent with a monumental and ongoing rabbit infestation. The rabbits' prodigious appetite, burrowing practices, and reproductive capacity have wreaked ecological and agricultural devastation, with serious consequences for the Australian economy. New Zealand has also considered using RHDV to control its own rabbit infestation, though importation was prohibited in 1997 due to uncertainties surrounding the epidemiology and efficacy of the virus

868

CHAPTER 40


[Ministry of Agriculture and Forestry (MAF), 1997]. Despite this ruling, the virus became established in New Zealand by virtue of illegal importation and release (Sissons and Grieve, 1998). Subsequent, deliberate spread by farmers, although initially of unclear legality, eventually was legalized through an amendment to the law. Recent regulations provide for the importation of the virus, as long as it is done with appropriate permits (J. Parkes, Landcare Research, New Zealand, personal communication). Following government-sanctioned tests of the efficacy of rabbit-to-rabbit transmission on an uninhabited Australian island in 1995, RHDV jumped to the mainland by an unknown mechanism. This led to the approval of hundreds of deliberate releases on the mainland, and brought into focus the issue of human safety in the affected areas. The wide host ranges of other calicivirus types, along with the apparent virulence of some caliciviruses in humans [representatives of four of five calicivirus categories are considered to be human pathogens (Smith et al., 1998)], and the high mutation rates characteristic of RNA viruses in general, militate against premature judgements on the issue of possible direct human impacts. Claims of high RHDV specificity toward European rabbits are based primarily on serologic analysis of blood from various experimentally infected species [Bureau of Resourse Sciences (BRS), 1996]. These studies have been criticized for applying an inappropriately high standard for a positive judgement of infection. It has been argued that the data actually support the opposite conclusion, that many species do indeed show evidence of infection (Smith, 1998). Ultimately, when dealing with serologic data, rigorous tests will be required to distinguish active infection from simple exposure. In any case, regardless of controversies over the extent of the RHDV host range, or whether there is a credible risk to humans, it seems clear that the rapid spread of the disease puts rabbit populations at risk in areas where eradication of those rabbits may not be desired. Where the possibility of human infectivity or illness resulting from contact with infected rabbits has been examined directly, divergent conclusions have been drawn, even when they are based on the same data set. Mead et at. (1996), in a report to the Australian government, as well as Carman et at. (1998) in the follow-up report in the open literature, found no evidence for infection or for symptoms of disease in over 250 people, many of whom had high exposure to the virus through direct handling of diseased rabbits. Smith et at. (1998), using the same data but categorizing putative exposure levels differently, saw correlations between exposure and incidence of a number of pathologies, including flu or fever, diarrhea or gastroenteritis, neurologic symptoms, rashes, and hepatitis. Matson (1998) also differed from Mead in his analysis of the human serologic data provided in the original Mead report, seeing plausible evidence for infection in some people from South Australia who had handled infected rabbits. Because of these conflicting data and their interpretations, clarification of the issue of RHDV specificity and safety to humans will depend on the rigorous analysis of well-planned future studies, both of an epidemiologic and experimental nature.

40.4 GENERAL CONCLUSIONS The promise of microbial pest control agents resides in their host organism specificity and in their relatively benign ecosystem and human health impacts. The requirements for toxicity testing of these agents in the United States under Subdivision M of the Federal Insecticide, Fungicide, and Rodenticide Act are designed to provide a fast and efficacious means to identify problematic MPCAs, while moving the rest toward registration. In general, this approach appears to have functioned well. Nonetheless, as is clearly recognized in those guidelines, MPCAs pose unique challenges to humans. They are, after all, alive, and thus carry at least a theoretical potential for adaptation and survival in novel microenvironments. It is not inconceivable that hazards to humans posed by living organisms could be missed in animal studies. For example, it might be argued that particular attention must be paid to the welfare of sensitive human subpopulations such as those who are diseased or immunocompromised, or who might be allergic to specific microorganisms. Attention to strain type is also very important, particularly when the microbe in question belongs to a medically significant species or genus. In the final analysis, continued monitoring of the health effects of MPCAs under conditions of actual pesticidal manufacture and use is warranted to ensure the safety of this interesting and viable approach to pest control.

ACKNOWLEDGMENTS For helpful discussions and comments on the manuscript, the author thanks Drs. Derek Gammon, Doug Gurian-Sherman, J. Thomas McClintock, John Parkes, Joel Siegel, and Alvin Smith. For help gaining access to the Pesticide Use Report of the Department of Pesticide Regulation (State of California), the author thanks Dr. Larry Wilhoit.

REFERENCES Alien, B. T., and Wilkinson, H. A., III (1969). A case of meningitis and generalized Shwartzman reaction caused by Bacillus sphaericus. Hopkins Med. 1. 125,8-13. Banerjee, C, Bustamante, C L, Wharton, K, Talley, E., and Wade, J. C (1988). Bacillus infections in patients with cancer. Arch. Intern. Med. 148, 17691774. Bames, KW., Meinecke, C E, McLane, W. C, and Rehnborg, C S. (1970). Long-term feeding and other toxicity-pathogenicity studies on rats using a commercial preparation of the nuclear-polyhedrosis virus of Heliothis zea. 1. Invert. Pathol. 16, 112-115. Beegle, C C, and Yamamoto, T. (1992). Invitation paper (C P. Alexander Fund); History of Bacillus thuringiensis berliner research and development. Can. Entomologist 124, 587-616. Bernstein, L L, Bernstein, J. A., Miller, M., Tierzieva, S., Bernstein, D. L, Lummus, Z., Selgrade, M. K., Doerfler, D. L, and Seligy, V. L (1999). Immune responses in farm workers after exposure to Bacillus thuringiensis pesticides. Environ. Health Perspect. 107, 575-582. Betz, E S., Forsyth, S. E, and Stewart, W. E. (1990). Registration requirements and safety considerations for microbial pest control agents in North America. In "Safety of Microbial Insecticides" (M. Laird, L A. Lacey, and E. W. Davidson, eds.), pp. 3-10. CRC Press, Boca Raton, EL.

References

Black, B. c', Brennan, LA., Dierks, P. M., and Gard, I. E. (1997). Commercialization of baculoviral insecticides. In "The Baculoviruses" (L K. Miller, ed.), pp. 341-387. Plenum Press, New York. Bureau of Resourse Sciences (BRS) (1996). "Rabbit Calicivirus Disease: A Report under the Biological Control Act of 1984." Australian Bureau of Resource Sciences, Canberra. Burgner, D., Eagles, D., Burgess, M., Procopis, P., Rogers, M., Muir, D., Pritchard, R., Hocking, A., and Priest, M. (1998). Disseminated invasive infection due to Metarhizium anisopliae in an immunocompromised child. 1. Clin. Microbiol. 36(4), 1146-1150. Butko, P. E, Huang, E, Pusztai-Carey, M., and Surewicz, W. K. K. (1997). Interaction of the 8-endotoxin CytA from Bacillus thuringiensis var. israelensis with lipid membranes. Biochemistry 36, 12862-12868. Butler, L L, Doherty, C. J., Hughes, J. E., Nelsonh, J. w., and Govan, J. R. W. (1995). Burkholderia cepacia and cystic fibrosis: Do natural environments present a potential hazard? 1. Clin. Microbiol. 33, 1001-1004. Carman, J. A., Garner, M. G., Catton, M. G., Thomas, S., Westbury, H. A., Cannon, R. M., Collins, G. J., and Tribe, I. G. (1998). Viral haemorrhagic disease of rabbits and human health. Epidemiol. Infection 121,409-418. Cepero de Garcia, M. C., Arboleda, M. L, Barraquer, E, and Grose, E. (1997). Fungal keratitis caused by Metarhizium anisopliae var. anisopliae. 1. Med. Vet. Mycol. 35, 361-363. Damgaard, P. H. (1995). Diarrhoeal enterotoxin production by strains of Bacillus thuringiensis isolated from commercial Bacillus thuringiensis-based insecticides. FEMS Immunol. Med. Microbiol. 12, 245-250. Department of Pesticide Regulation (DPR) (1993). Data package summary and recommendation sheet, Gliocladium virens GL-21 dated 12/16/93. Doc. D51957>RA>FOO 11688>POO003>S931216, California Department of Pesticide Regulation, Sacramento. Department of Pesticide Regulation (DPR) (1994a). Review of the study: "Acute oral toxicity/pathogenicity study of SMP-l in rats" dated 2121190. Record 122641, Vo!. 51990-002, California Department of Pesticide Regulation, Sacramento. Department of Pesticide Regulation (DPR) (1994b). Review of the study: "Acute pulmonary toxicity/pathogenicity study of SMP-l in rats" dated 2121190. Record 122643, Vo!. 51990-002, California Department of Pesticide Regulation, Sacramento. Department of Pesticide Regulation (DPR) (2000). "Pesticide Use Report." Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento. Doller, G., Groner, A., and Straub, O. C, (1983). Safety evaluation of nuclear polyhedrosis virus replication in pigs. App!. Environ. Microbiol. 45, 12291233. Domsch, K. H., Gams, W., and Anderson, T.-H. (1980). Metarrhizium; Sorok 1883. In "Compendium of Soil Fungi;' Vo!. 1, pp. 413-415. Drobniewski, E A. (1993). Bacillus cereus and related species. Clin. Microbiol. Rev. 6, 324-338. Drobniewski, E A. (1994). The safety of Bacillus species as insect vector control agents. 1. Appl. Bacteriol. 76, 101-109. Ferron, P. (1981). Pest control by the fungi Beauveria and Metarhizium. In "Microbial Control of Pests and Plant Diseases" (H. D. Burges, ed.), pp. 465482. Academic Press, London. Fisher, R., and Rosner, L (1959). Toxicology of the microbial insecticide, Thuricide. 1. Agric. Food Chem. 7,686-688. Freour, P., Lahourcade, M., and Chomy, P. (1966). Une mycose nouvelle: Etude clinique et mycologique d'une localisation pulmonaire de Beauveria. Societe Medicale des H8pitaux de Paris 117, 197-200. George, S. E., Kohan, M. J., Whitehouse, D. A., Creason, J. P., Kawanishi, C. Y., Sherwood, R. L, and Claxton, L D. (1991). Distribution, clearance, and mortality of environmental pseudomonads in mice upon intranasal exposure. Appl. Environ. Microbiol. 57,2420-2425. George, S. E., Nelson, G. M., Kohan, M. J., Brooks, L R., and Boyd, C. (1999). Colonization and clearance of environmental microbial agents upon intranasal exposure of strain C3H1HeJ mice. 1. Toxico!. Environ. Health (part A) 56, 419-431. Goepfert, J. M., Spira, W. M., and Kim, H. U. (1972). Bacillus cereus: Food poisoning organism. A review. 1. Milk Food Technol. 35,213-227.

869

Goldstein, R., Sun, L, Jiang, R.-Z., Sajjan, U., Forstner, J. E, and Campanelli, C, (1995). Structurally variant classes of pilus appendage fibers coexpressed from Burkholderia (Pseudomonas) cepacia. 1. Bacteriol. 177, 1039-1052. Green, M., Heumann, M., Sokolow, R., Foster, L R., Bryant, R., and Skeels, M. (1990). Public health implications of the microbial pesticide Bacillus thuringiensis: An epidemiological study, Oregon, 1985-86. Am. 1. Publ. Health 80, 848-852. Gupta, S., Roberts, D. w., and Renwick, J. A. A. (1989). Insecticidal cyclodepsipeptides from Metarhizium anisopliae. 1. Chem. Soc. Perkin Trans. I, 2347-2357. Heimpel, A. M. (1966). Exposure of white mice and guinea pigs to the nuclearpolyhedrosis virus of the cabbage looper, Trichoplusia ni. 1. Invert. Pathol. 8,98-102. Heimpel, A. M., and Buchanan, L K. (1967). Human feeding tests using a nuclear-polyhedrosis virus of Heliothis zea. 1. Invert. Pathol. 9,55-57. Heimpel, A. M., Thomas, E. D., Adams, J.R., and Smith, L J. (1973). The presence of nuclear polyhedrosis virus of Trichoplusia ni on cabbage from the market shelf. Environ. Entomol. 2,72-75. Hernandez, E., Ramisse, E, Ducoureau, J.-P., Cruel, T., and Cavallo, J.-D. (1998). Bacillus thuringiensis subsp. konkukian (serotype H34) superinfection: Case report and experimental evidence of pathogenicity in immunosuppressed mice. 1. Clin. Microbiol. 36,2138-2139. Hofte, H., and Whiteley, H. R. (1989). Insecticidal crystal proteins of Bacillus thuringiensis. Microbiol. Rev. 53,242-255. Holmes, A., Govan, J., and Goldstein, R. (1998). Agricultural use of Burkholderia (Pseudomonas) cepacia: A threat to human health? Emerging Infectious Diseases 4, 221-227. Huang, H., Ignoffo, c'M., and Shapiro, M. (1977). Physical and clinical examinations of personnel involved in production of the insect virus, Baculovirus heliothis.l. Kansas Entomol. Soc. 50, 200-202. Huber, J. (1995). Opportunities with baculoviruses. In "Biological Control: Benefits and Risks" (H. M. T. Hokkanen, and J. M. Lynch, eds.), pp. 201206. Cambridge Univ. Press, Cambridge, U.K. Ignoffo, C. M. (1973). Effects of entomopathogens on vertebrates. Ann. N. Y. Acad. Sci. 217, 141-172. Ignoffo, C. M., and Heimpel, A. M. (1965). The nuc1ear-polyhedrosis virus of Heliothis zea (Boddie) and Heliothis virescens (Fabricius). V. Toxicitypathogenicity of virus to white mice and guinea pigs. 1. Invert. Pathol. 7, 329-340. Isaacson, P., Jacobs, P. H., MacKenzie, A. M. R., and Mathews, A. W. (1976). Pseudotumour of the lung caused by infection with Bacillus sphaericus. 1. Clin. Pathol. 29, 806-811. Ishibashi, Y., Kaufman, H. E., Ichinoe, M., and Kagawa, S. (1986). The pathogenicity of Beauveria bassiana in the rabbit cornea. Mykosen 30(3), 115-126. Jackson, S. G., Goodbrand, R. B., Ahmed, R., and Kasatiya, S. (1995). Bacillus cereus and Bacillus thuringiensis isolated in a gastroenteritis outbreak investigation. Letl. Appl. Microbiol. 21, 103-105. Kaufman, G., and Bellas, T. (1996). Occupational allergy to Metarhizium. Allergy Asthma Proc. 17, 116. Kerwin, J. L, Dritz, D.A., and Washino, R. K. (1990). Confirmation of the saftey of Lagenidium giganteum (Oomycetes: Lagenidiales) to mammals. 1. Econ. Entomol. 83(2), 374-376. Krych, V. K., Johnson, J. L, and Yousten, A. A. (1980). Deoxyribonucleic acid homologies among strains of Bacillus sphaericus. Intern. Systematic Bacteriol. 30, 476-484. LiPuma, J. J. (1998). Burkholderia cepacia. Management issues and new insights. Clinics Chest Med. 19,473-486. Logan, N. A., and Turnbull, P. C, B. (1999). Bacillus and recently derived genera. In "Manual of Clinical Microbiology" (P. R. Murray, E. J. Baron, M. A. Pfaller, E C, Tenover, and R. H. Yolken, eds.), pp. 357-369. ASM Press. Lumsden, R. D., and Waiter, J. E (1995). Development of the biocontrol fungus Gliocladium virens: Risk assessment and approval for horticultural use. In "Biological Control: Benefits and Risks" (H. M. T. Hokkanen, and 1. M. Lynch, eds.), pp. 263-269. Cambridge Univ. Press, Cambridge, UK.

870

CHAPTER 40


Ministry of Agriculture and Forestry (MAF) (1997). MAF says "no" to RCD as rabbit control in New Zealand. Media release, Ministry of Agriculture and Forestry. Available at http://www.maf.govt.nzIMAFnetlpress/020797rcd. htm. Mahenthiralingam, E., Simpson, D. A., and Speert, S. P. (1997). Identification and characterization of a novel DNA marker associated with epidemic Burkholderia cepacia strains recovered from patients with cystic fibrosis. 1. CZin. Microbio!' 35,808-816. Matson, D. O. (1998). Re-analysis of serologic data from the Australian study of human health risks of infection by rabbit haemorrhagic disease virus. In "Rabbit Control, RCD: Dilemmas & Implications, Proceedings of the Rabbit Control, RCD: Dilemmas and Implications Conference," Wellington, New Zealand, 30--31 March 1998 (compiled by B. D. W. Jarvis), pp. 62--66. New Zealand Association of Scientists, supported by The Royal Society of New Zealand. Mayes, M. E., Held, G. A., Lau, c., Seely, J. C., Roe, R. M., Dauterman, W. C., and Kawanishi, C. Y. (1989). Characterization of the mammalian toxicity of the crystal polypeptides of Bacillus thuringiensis subsp. israelensis. Fundam. App!. Toxico!. 13,310--322. McClintock, J. T. (1999). The federal registration process in requirements for the United States. In "Biopesticides: Use and Delivery" (F. R. Hall, and J. J. Menn, eds.), Methods in Biotechnology, Vo!. 5, pp. 415~1. Humana Press, Totowa, NJ. McClintock, J. T., Schaffer, C.R., and Sjoblad, R. D. (1995). A comparative review of the mammalian toxicity of Bacillus thuringiensis-based pesticides. Pestic. Sci. 45,95-105. Mead, C., Kaldor, J., Canton, M., Gamer, G., Crerar, S., and Thomas, S. (1996). "Rabbit Calicivirus and Human Health. Report of the Rabbit Calicivirus Human Health Study Group." Department of Primary Industries and Energy, Australian Government, Canberra. Meinecke, C. F., McLane, W. c., and Rehnborg, C. S. (1970). Inhalation and dermal allergenicity studies of a nuciear-polyhedrosis virus of Heliothis zea in guinea pigs. 1. Invert. Patho!. 15,207-210. Mel'nikova, E. A., and Murza, V. I. (1980). Investigation of the safety of industrial strains of microorganisms and microbial insecticides. 1. Hyg. Epidemio!. Microbiol. Immuno!. 24,425-431. Miller, L. K., Lingg, AJ., and Bulla, L.A., Jr. (1983). Bacterial, viral, and fungal insecticides. Science 219, 715-721. Muir, D., Martin, P., Kendall, K., and Malik, R. (1998). Invasive hyphomycotic rhinitis in a cat due to Metarhizium anisopliae. Med. Myco!. 36,51-54. Murthy, P. S. R. (1997). Mucous membrane irritancy study of Bacillus sphaericus (1593) and B. thuringiensis (H-14) formulations in rabbit. Bio!. Memoirs 23, 11-13. Nowotny, N., Ros Bascunana, C., Ballagi-Pordany, A., Belak, S., GaviewWiden, D., and Uhlen, M. (1998). Phylogeny and variability of rabbit haemorrhagic disease virus, and the present situation of rabbit haemorrhagic disease in Europe. In "Rabbit Control, RCD: Dilemmas & Implications, Proceedings of the Rabbit Control, RCD: Dilemmas and Implications Conference," Wellington, New Zealand, 30--31 March 1998 (compiled by B. D. W. Jarvis), pp. 47-51. New Zealand Association of Scientists, supported by The Royal Society of New Zealand. Parke, J. L. (1998). Burkholderia cepacia: Friend or foe? APSnet feature (American Phytopathological Society), Oct. 1-31, 1988. Available at http:// www.scisoc.org/featurelBurkholderiaCepaciaITop.htrn!. Parkes, J. P., Norbury, G. L., and Heyward, R. P. (1999). Has rabbit haemorrhagic disease worked in New Zealand? Proc. New Zealand Soc. Animal Production 59, 245-249. PTCN (1997). Pseudomonas cepacia use opposed by cystic fibrosis groups. Pestic. Toxic Chem. News Apr. 2,17-18. Pivovarov, Y. P., Ivashina, S.A., and Padalkin, V. P. (1977). Hygienic assessment of investigation of insecticide preparations. Gig. Sanit. 8, 95-98. Ray, D. E. (1991). Pesticides derived from plants and other organisms. In "Handbook of Pesticide Toxicology" (w. 1. Hayes, Jr., and E. R. Laws, Jr., eds.), pp. 585-636. Academic Press, San Diego.

Saik, J. E., Lacey, L. A., and Lacey, C. M. (1990). Safety of microbial insecticides to vertebrates---Domestic animals and wildlife. In "Safety of Microbial Insecticides" (M. Laird, L. A. Lace, and E. W. Davidson, eds.), pp. 115132. CRC Press, Boca Raton, PL. Sajjan, U., and Forstner, J. (1993). Role of a 22-kilodalton pilin protein in binding of Pseudomonas cepacia to buccal epithelial cells. Infect. Immun. 61, 3156--3163. Samples, J. R., and Buettner, H. (1983). Corneal ulcer caused by a biologic insecticide (Bacillus thuringiensis). Am. 1. Ophthalmo!. 95,258-260. Schnepf, E., Crickmore, N., van Rie, J., Lerecius, D., Baum, J., Feitelson, J., Zeigler, D. R., and Dean, D. H. (1998). Bacillus thuringiensis and its pesticidal crystal proteins. Microbio!' Mol. BioI. Rev. 62, 775-806. Semalulu, S. S., MacPherson, J. M., Scheifer, H. B., and Khachatourians, G. G. (1992). Pathogenicity of Beauveria bassiana in mice. 1. Vet. Med. B 39, 81-90. Shadduck, J. A. (1983). Some considerations on the safety evaluation of nonviral microbial pesticides. Bull. World Health Organization 61,117-128. Shadduck, J. A., Singer, S., and Lause, S. (1980). Lack of mammalian pathogenicity of entomocidal isolates of Bacillus sphaericus. Environ. Entomol. 9,403-407. Siegel, J. P. (1997). Testing the pathogenicity and infectivity of entomopathogens to mammals. In "Manual of Techniques in Insect Pathology" (L. A. Lacey, ed.), pp. 325-336. Academic Press, San Diego. Siegel, J. P., and Shadduck, J. A. (1987). Safety of the entomopathogenic fungus Lagenidium giganteum (Oomycetes: Lagenidiales) to mammals. 1. Econ. Entomol. 80(5),994--997. Siegel, J. P., and Shadduck, J. A. (1990a). Clearance of Bacillus sphaericus and Bacillus thuringiensis ssp. israelensis from mammals. 1. Econ. Entomol. 83,347-355. Siegel, J. P., and Shadduck, J. A. (1990b). Safety of microbial insecticides to vertebrates-humans. In "Safety of Microbial Insecticides" (M. Laird, L. A. Lacey, and E. W. Davidson, eds.), pp. 101-113. CRC Press, Boca Raton, PL. Siegel, J. P., and Shadduck, J. A. (1992). Testing the effects of microbial pest control agents on mammals. In "Microbial Ecology. Principles, Methods, and Applications" (M. A. Levin, R. J. Seidler, and M. Rogul, eds.), pp. 745759. McGraw-Hill, New York. Sissons, c., and Grieve, J. (1998). Introduction. In "Rabbit Control, RCD: Dilemmas and Implications, Proceedings of the Rabbit Control, RCD: Dilemmas and Implications Conference," Wellington, New Zealand, 30--31 March 1998 (compiled by B. D. W. Jarvis), pp. v-vi. New Zealand Association of Scientists, supported by The Royal Society of New Zealand. Smith, A. W. (1998). Calicivirus models of emerging and zoonotic diseases. In "Rabbit Control, RCD: Dilemmas and Implications, Proceedings of the Rabbit Control, RCD: Dilemmas and Implications Conference," Wellington, New Zealand, 30--31 March 1998 (compiled by B. D. W. Jarvis), pp. 37-42. New Zealand Association of Scientists, supported by The Royal Society of New Zealand. Smith, A. w., Skilling, D. E., Cherry, N., Mead, J. H., and Matson, D. O. (1998). Calicivirus emergence from ocean reservoirs: Zoonotic and interspecies movements. Emerging Infectious Diseases 4, 13-20. Song, J. v., Houng, S. M., Lin, G. H., Lou, C. Z., Hou, M. S., Tzan, L. F., and Chang, O. Z. (1989). Experimental study of farmers' lung-like lesions caused by Beauveria bassiana. Chung-hua-Ping-Li-Hsuch-Tse-Chin 18, 111-114. Sun, L., Jiang, R.-Z., Steinbach, S., et al. (1995). The emergence of a highly transmissible lineage of cbl+ Pseudomonas (Burkholderia) cepacia causing CF centre epidemics in North America and Britain. Nature Med. 1,661. Thomas, W. E., and Ellar, D. J. (1983). Bacillus thuringiensis var. israelensis crystal o-endotoxin: Effects on insect and mammalian cells in vitro and in vivo. 1. Cell Sci. 60, 181-197. U.S. Environmental Protection Agency (U.S. EPA) (1982). "Pesticide Assessment Guidelines. Subdivision M, Biorational Pesticides." Office of Pesticide and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (1998). "Reregistration Eligibility Decision (RED). Bacillus thuringiensis." Prevention, Pesticides,

References

and Toxic Substances, D.S. Environmental Protection Agency, Washington, DC. Vandamme, P., Holmes, B., Vancanneyt, M., Coenye, T., Hoste, B., Coopman, R., Revets, H., Lauwers, S., Gillis, M., Kersters, K., and Govan, J. R. W. (1997). Occurrence of multiple genomovars of Burkholderia cepacia in cystic fibrosis patients and proposal of Burkholderia multivorans sp. novo Int. J. Systematic Bacteriol. 47, 1188-1200. Walters, S., and Smith, E. G. (1993). Pseudomonas cepacia in cystic fibrosis: Transmissibility and its implications. Lancet 342, 3-4.

871

Ward, M. D. w., Sailstad, D. M., and Selgrade, M. K. (1998). Allergic responses to the biopesticide Metarhizium anisopliae in Balb/c mice. Toxicol. Sci. 45, 195-203. Warren, R. E., Rubenstein, D., Ellar, D. J., Kramer, 1. M., and Gilbert, R. J. (1984). Bacillus thuringiensis var. israelensis: Protoxin activation and safety. Lancet 1, 678-679. World Health Organization (WHO) (1981). Mammalian safety of microbial agents for vector control: A WHO memorandum. Bull. World Health Organization 59, 857-863.

CHAPTER

41 The Influence of Age on Pesticide Toxicity Carey Pope Oklahoma State University

41.1 GENERAL CONCEPTS IN DIFFERENTIAL SENSITIVITY TO PESTICIDES Age-related differences in sensitivity to pesticides can occur for a wide variety of reasons. Toxicokinetic differences among different age groups can contribute to differential sensitivity, with differences in biotransformation often being a major factor. In other instances, toxicodynamic differences may exist which lead to age-related differences in sensitivity. For example, during development and maturation, a critical time of exposure or "window of opportunity" during which a developmental process occurs may impart selective sensitivity. At the other end of the spectrum, changes associated with aging may alter sensitivity to pesticide toxicity. Moreover, the relative contributions of toxicokinetic and toxicodynamic factors in age-related sensitivity may differ markedly among the various classes of pesticides, and even among members of the same class of toxicants. Exposures to pesticides are often age-related, based on age-specific behaviors, diets, or other factors. Thus, the nature of age-related differences in sensitivity to pesticides is complex, and broadbased generalities are typically unjustified. See Table 41.1. With the common routes of exposure (i.e., oral, dermal, and inhalation), a pesticide must first be absorbed before systemic toxicity can occur. In a comparative study of 14 different pesticides, 11 of these exhibited age-related differences in percutaneous absorption (Shah et al., 1987). Interestingly, four of the 14 showed greater absorption in young (33 days old), while seven of the 14 showed more extensive absorption in adults (82 days old). Moreover, even within the same class of pesticide (e.g., the organophosphorus toxicants parathion and chlorpyrifos), no clear age-related pattern of dermal absorption was evident, that is, chlorpyrifos showed greater absorption in young while parathion showed greater absorption in older animals. Hall and co-workers (1992) reported that dermal absorption of the dinitrophenol pesticide dinoseb was lower (about 20%) in 33-day-old rats compared to adults (82 days of age). Thus, Handbook of Pesticide Toxicology Volume 1. Principles

differences in rates or extent of absorption can contribute to differential sensitivity among age groups. Once absorbed, differences in tissue distribution or rates of elimination between age groups can contribute to differential sensitivity. Older animals (and people) typically have a higher fat content than younger individuals, which can have an important effect on distribution, accumulation, and storage of highly lipophilic pesticides, for example, organochlorines. Deichmann (1972) reported that DDT was eliminated from the body most efficiently in neonates and less so in older rats, at least partially because of differences in partitioning of the pesticide into fatty tissues. Changes in biotransformation during maturation and aging can often contribute to age-related differences in sensitivity. Immature and very old animals generally have lower biotransformation capacities, for example, lower levels of CYP450 (Benke and Murphy, 1975; Mehendale, 1980; Wynne et al., 1987). If a pesticide is activated by CYP450 to a more toxic metabolite, lower levels of CYP450 could be associated with lower sensitivity to that pesticide. In contrast, pesticides which are inactivated by monooxygenases may be relatively more toxic in groups with lower CYP450 levels. Lower activities of phase 11 reactions in neonatal or aged animals may also increase sensitivity to certain pesticides (Borghoff et al., 1988; Das et al., 1981; Egaas et al., 1995). Because of the complexity of pathways and the multiplicity of reactions generally involved in xenobiotic metabolism, however, differences in individual metabolic processes between age groups have to be considered in context to appreciate the net consequences of biotransformation on age-related toxicity. For example, while young rats exhibit lower rates of CYP450-mediated activation of the organophosphorus pesticide parathion to its active metabolite, paraoxon, lesser capacity in neonates for detoxification of paraoxon appears to be a prominent difference contributing to higher sensitivity (Benke and Murphy, 1975). Toxicodynamic differences can also contribute to age-related sensitivity. The ability to restore function following toxic ant exposure may be higher in some age groups than in others. For

873


874

CHAPTER 41

Age and Pesticide Toxicity

Table 41.1 General Factors Contributing to Age-Related Differences in Sensitivity to Pesticides Toxicokinetic

Toxicodynamic

Exposure-based

Differences in absorption

Age-related expression of target molecules

Age-related behaviors

Differences in distribution/elimination Differences in biotransformation

or sensitive processes

Age-related diets

Differential capacities to recover from or adapt to toxicant insult

example, young rats challenged with hepatotoxic ants recover much better than older animals, apparently because of more rapid and robust synthesis of new cells following the initial tissue damage (Dalu and Mehendale, 1996). More rapid recovery of acetylcholinesterase (AChE) activity in younger animals (Chakraborti et aI., 1993; Moser, 1999; Pope et al., 1991; Pope and Liu, 1997) and slower recovery in aged animals (Michalek et aI., 1990) following acute exposure to an organophosphorus anticholinesterase may make those age groups differentially sensitive to subsequent anticholinesterase exposures. Thus age-related differences in sensitivity to pesticides can be influenced by multiple toxicokinetic and toxicodynamic factors. An important consideration in the differential sensitivity to pesticides is the time available for toxicity to develop. Children have a longer time to live than adults; thus, if pesticide exposure leads to the development of some form of delayed toxicity, for example, tumor formation, a child has more time for this adverse effect to be exhibited. Conversely, older individuals have experienced a longer time to accumulate residues or damage from long-term exposures. As noted before, the critical time-dependent nature of developmental stages is also an important consideration in age-related differences in response to pesticides. The endogenous metabolite bilirubin, for example, induces encephalopathy in the developing nervous system only at certain early timepoints when the blood-brain barrier is deficient (Lee et aI., 1995; Wennberg et aI., 1993). Another factor of particular importance to age-related differences in susceptibility is differential exposures. Age-related behaviors may contribute to differential exposure and sensitivity. For example, young children tend to sample the environment by taste. If the opportunity arises for oral "sampling" of a pesticide container, the young child may be much more susceptible to toxicity based on a greater likelihood of such exposure. In general, young children tend to be more exploratory and inquisitive than adults, which can sometimes lead to contact with inappropriately stored chemicals. Many lipophilic xenobiotics concentrate in breast milk; thus, breastfed infants may be preferentially exposed to such toxicants (Mussalo-Rauhamaa et aI., 1984; Schildkraut et aI., 1999). Young children eat more in proportion to their body size and they tend to eat more frequently than adults. When pesticide residues are consumed with the food, the relative frequency of exposure can be important if recovery takes longer than the time between exposures. Toddlers are also in contact with the floor more than are adults. With a higher surface-area-to-body-weight ratio, dermal contact may

Differences in time available for cumulative exposures and/or expression of toxicity

be more extensive than in adults. When pesticide residues fall to the floor after household applications, or become associated with carpeting or furniture, there is a higher probability of direct dermal contact in children playing on those surfaces (Fenske et aI., 1990; Lu and Fenske, 1999). Conversely, adults can be exposed to chemicals in the workplace, an exposure possibility which is generally missing in young children and older adults. Obviously, there are many reasons why exposures to pesticides can be age-related. The role of differential exposure in age-related sensitivity to pesticides is a critical issue and is discussed in more detail in Chapter 16. It is apparent, however, that age-related differences in sensitivity to pesticides can be caused by differences in inherent sensitivity to the pesticide, differences in exposure, or both. Clearly, multiple factors can contribute to differential susceptibility to pesticides throughout life. Risk assessment for pesticides relies heavily on data generated from animal studies. In fact, the United States Environmental Protection Agency recently limited the use of human data in the pesticide registration process (EPA, 1998a). The use of rodent animal models to estimate age-related differences in sensitivity in humans has some inherent problems, however, in particular when modeling the effects of early postnatal exposures. Developmentally, the maturational states of experimental animals and humans at parturition and perinatal periods can be quite different (Romijn et al., 1991). If neonatal rodents are more sensitive than adults to a particular pesticide, but only briefly during the early postnatal period, they may not represent a valid model for children because of the species differences in maturation relative to the timing of exposure. The comparative development, maturation, and aging of organ systems between man and experimental animals must be kept in mind when extrapolating age-related differences in sensitivity from animal models.

41.2 CHILDREN'S HEALTH AND REGULATION OF PESTICIDES IN THE UNITED STATES Ideally, regulatory policies governing the use of pesticides should be conservative enough to allow for protection of all members of the population. With noncarcinogenic toxicants, an uncertainty factor of ten is generally incorporated into the risk assessment process for such purposes (Bames and Dourson,

41.3 Age-Related Differences in Sensitivity to Pesticides

1988), assuming that variability in sensitivity to a particular agent within subpopulations is no greater than an order of magnitude. For this to be true and for all members of the population to be protected, all possible extrinsic and intrinsic modifiers of toxicity, for example, nutrition, disease, physiological stressors, or genetic polymorphisms, must together contribute to less than a ten-fold variation in sensitivity in the entire population. One intrinsic modifier of toxicity, age, has received considerable attention in recent years. In particular, the relative sensitivity of developing infants and children to pesticides has been the focus of concern (Bearer, 1995; Fenner-Crisp, 1995; Garrettson, 1997; Goldman, 1995; Little, 1995; Tilson, 1998). In 1988, the National Academy of Sciences (NAS) initiated a concerted effort to evaluate pesticide exposures in infants and children and to determine if the health of children was adequately addressed in the regulation of pesticides. The Committee on Pesticides in the Diets of Infants and Children, composed of scientists from industry, government, and academia, was established within the National Research Council of NAS in 1988 to evaluate the relative sensitivity of infants and children to pesticides. The conclusions eventually reached by this select committee had far-reaching consequences (see later). In 1989, public attention in the United States was focused by media coverage of a report from the Natural Resources Defense Council (NRDC) entitled Intolerable Risk: Pesticides in our Children's Food (NRDC, 1989) on the possibility that children were being exposed to excessive levels of pesticide residues in food products. The executive summary of this report begins "Our nation's children are being harmed by the very fruits and vegetables we tell them will make them grow up healthy and strong." While the basis of many claims in the NRDC report may have been inaccurate (Wilkinson and Ginevan, 1989), the public attention raised by this report had a significant impact; that is, it strengthened the commitment to ensure that children's health was adequately considered in the risk assessment of pesticides. Four years later, the National Academy of Sciences published the report Pesticides in the Diets of Infants and Children (NAS, 1993), which detailed conclusions from the NRC committee with the same name. Major findings of this committee included (i) that both quantitative and qualitative differences in toxicity of pesticides can occur between children and adults, but quantitative differences are usually less than a factor of ten, (ii) that infants and adults differ quantitatively and qualitatively in the types of pesticide exposures in the diet, a factor generally more important than differences in inherent sensitivity, (iii) that assessment of pesticide exposures should consider dietary as well as nondietary sources, and (iv) that "in the absence of data to the contrary, there should be a presumption of greater toxicity to infants and children" (NAS, 1993). The findings from this committee provided impetus for federal legislation addressing pesticide regulation, in particular regarding potential problems with differential exposure and sensitivity in children. In 1996, the Food Quality Protection Act (FQPA), containing sections relating to the protection of infants and children from pesticide exposures, was passed into law. The FQPA

875

amended the Federal Insecticide, Fungicide and Rodenticide Act and the Federal Food, Drug and Cosmetic Act (FFDCA). Section 408(b)(2)(C) of FFDCA states that with "threshold" adverse effects, "an additional tenfold margin of safety for the pesticide chemical residue ... shall be applied for infants and children to take into account potential pre- and post-natal toxicity and completeness of the data with respect to exposure and toxicity to infants and children." Further, this section of FFDCA states that the "Administrator may use a different margin of safety for the pesticide chemical residue only if, on the basis of reliable data, such margin will be safe for infants and children." In October of 1995, the United States Environmental Protection Agency (EPA) announced that it would explicitly evaluate risks to infants and children in all regulatory actions, and in April of 1997, Executive Order 13045 directed Federal agencies to identify and assess environmental health and safety risks to children (EPA, 1998b). Thus, the default position of EPA in pesticide regulatory decisions was to use an additional 10 x uncertainty factor (the FQPA factor) for threshold effects to ensure the protection of infants and children from pesticide toxicity. The EPA Office of Pesticide Programs proposal included, however, the possibility of either removing or reducing the magnitude of the FQPA factor if "reliable data" were available that suggested infants and children would be adequately protected under those conditions (EPA, 1999). Thus, risk assessment procedures for pesticides registered with the U.S. EPA now incorporate an additional uncertainty factor for infants and children unless sufficient data indicates that young are not at higher risk. The conclusions from the NAS report (NAS, 1993) regarding risks to infants and children were based on two parameters, that is, differences in sensitivity and differences in exposure. The following is a brief summary of evidence pertaining to age-related differences in response to pesticides. It should be noted that while the recent focus of concern in the United States has been on the possibly higher susceptibility of infants and children, because of the demographics of societal aging, elderly individuals and their relative susceptibility to pesticides may become a more important issue in the future (Overstreet, 2000). Alterations in cholinergic neurotransmission with aging and associated neurological disorders such as Alzheimer's disease may be particularly important in contributing to differential sensitivity to the cholinesteraseinhibiting agents and with other pesticides which may alter cholinergic neurotransmission.

41.3 AGE-RELATED DIFFERENCES IN SENSITIVITY TO PESTICIDES It is apparent that, as with other types of xenobiotics (Done, 1964; Goldenthal, 1971), there is no consistent effect of age on acute sensitivity to pesticides across all classes or even within a class of compounds. There are various factors that could contribute to differential toxicity, whether one compares different age groups, different species, different sexes, different strains,

876

CHAPTER 41


Table 41.2 Studies Reporting Age-Related Differences in Sensitivity with the Major Classes of Pesticides Type of study: Pesticide class Organophosphorus

relative sensitivity

References

Animal study:

Brodeur and DuBois (1963); Gagne and Brodeur (1972);

Immature more

Benke and Murphy (1975); Mendoza (1976); Long et al. (1986);

sensitive than adults

Pope et al. (1991); Atterberry et al. (1997); Moser and Padilla (1998); Karantb and Pope (2000)

Animal study: Adults

Lu et al. (1965); Chakraborti et al. (1993); Pope and Liu (1997);

more sensitive than

lohnson and Barnes (1970); Moretto et al. (1991);

immature

Peraica et al. (1993); Pope et al. (1992, 1993); Harp et al. (1997)

Animal study: Aged

Veronesi et al. (1990); Karantb and Pope (2000)

adults more sensitive than young adults Human study:

Diggory et al. (1977)

Children more sensitive than adults Organochlorines

Animal study:

Eriksson (1997); Samanta and Chainy (1997);

Immature more

linna et al. (1989)

sensitive than adults Animal study: Adults

Lu et al. (1965); Kiran and Varma (1988)

more sensitive than immature Carbamates

Animal study:

Moser (1999) (based on lethality)

Immature more sensitive than adults Animal study: Aged

Knisely and Hamm (1989); Takahashi et al. (1991)

adults more sensitive than young adults Human study:

Lifshitz et al. (1997) (depending on endpoint)

Children more sensitive tban adults Pyrethroids

Animal study:

Cantalamessa (1993); Sheets et al. (1994)

Immature more sensitive than adults

or any other factor. These contributing factors will be examined in more detail with specific examples of pesticides potentially capable of eliciting age-related effects. See Table 41.2. 41.3.1 ORGANOPHOSPHORUS PESTICIDES

Organophosphorus pesticides (OPs) elicit toxicity through inhibition of AChE (Mileson et aI., 1998). Age-related differences in sensitivity to OPs have been reported in many experimental studies (Benke and Murphy, 1975; Brodeur and DuBois, 1963; Gagne and Brodeur, 1972; Gaines and Linder, 1986; Mendoza, 1976; Moser and Padilla, 1998; Pope et aI., 1991). In general (but not always), neonatal animals are more sensitive to the acute toxicity ofOPs. Lu and co-workers (1965) reported a maturational decrease in sensitivity to malathion among newborn,

14- to 16-day-old, and adult rats. Mendoza (1976) reported that 1-day-old rats were about nine times more sensitive to lethality from acute exposure to malathion. Mortality in newborn pigs following dermal application of chlorpyrifos (2.5% aerosol) was markedly higher when exposure occurred within the first three hours of life than at 30-36 hours after birth, suggesting a rapid change in sensitivity in the first days following parturition (Long et aI., 1986). Pope and co-workers (1991) reported that 7-day-old rats were 2-9 times more sensitive than adult (90 days of age) rats to the acute toxicity of methyl parathion, parathion, and chlorpyrifos. In contrast, methamidophos appears to elicit little age-related toxicity (Moser, 1999; Padilla et aI., 2000). Several factors could contribute to age-related differences in response to acute OP exposures. Gagne and Brodeur (1972)


investigated potential metabolic factors in the higher sensitivity of weanling rats to parathion and concluded that limited detoxification of parathion and its metabolite paraoxon were at least partially responsible. Later, Benke and Murphy (1975) evaluated metabolic contributions to age-related differences in sensitivity to parathion and methyl parathion. When metabolic rates were compared to LDso values in different age groups, high correlations were noted between lethality and liver and plasma A-esterase activity, oxon dealkylation and dearylation, and binding to "noncritical tissue constituents" in liver and plasma. They concluded that more robust metabolic inactivation of the active oxons of these two pesticides in more mature animals was primarily responsible for decrease in sensitivity with age. Atterberry and co-workers (1997) compared the toxicity and biotransformation of parathion and chlorpyrifos in neonatal and adult rats and concluded that differences in liver carboxylesterase activity and CYP450-dependent dearylation were important in differential age-related sensitivity to these pesticides. Moser and colleagues (1998) concluded that differences in liver carboxylesterase and A-esterase activities formed the basis for age-related differences in sensitivity to acute chlorpyrifos exposures. Other studies have indicated that maturational differences in the capacity for detoxification of organophosphates by A-esterases and carboxylesterases may contribute to higher sensitivity to these pesticides in immature animals (Costa et aI., 1990; Li et aI., 1993, 1995; Maxwell, 1992; Pond et aI., 1995). Thus, considerable evidence suggests that immature animals are more sensitive to the acute toxicity of several OP pesticides because of limited detoxification of either the parent compound or its active metabolite. Young children also appear to be more sensitive to acute toxicity from OP exposure. In a case of parathion-contaminated food in Jamaica, the highest incidence of lethality was in children less than five years of age (Diggory et aI., 1977). Differences in metabolic capacities between very young children and older children or adults may also be primary determinants in age-related sensitivity to acute OP exposures. Augustinsson and Barr (1963) showed that serum arylesterase (A-esterase) activity was very low in newborn children but increased steadily during the first six months of life. Ecobichon and Stephens (1973) reported that plasma cholinesterase and A-esterase activities increased dramatically in children during the first year of life, after which no further increases occurred. Any active anticholinesterases in the blood of very young children would therefore be less likely to bind to nontarget cholinesterases or to be hydrolyzed by A-esterases; thus, more inhibitor would be available to reach target tissues. As detoxification of active OP anticholinesterases is thought to be a prominent factor in age-related sensitivity (Atterberry et aI., 1997; Benke and Murphy, 1975; Mortensen et aI., 1996; Moser and Padilla, 1998), these studies suggest that dramatically higher acute sensitivity in children may only exist in the very young (:sI year of age), however, when these detoxification processes appear most limited. In addition to metabolic differences that may contribute to age-related sensitivity to OP pesticides, some toxicody-

877

namic differences among age groups could also be important. Organophosphorus and carbamate pesticides are toxic by virtue of their ability to inhibit AChE (Fukuto, 1990). Species differences in sensitivity of AChE to inhibition by some OP anticholinesterases have been reported (Kemp and Wallace, 1990). Thus, there could be a degree of selective toxicity among age groups based on the molecular interaction between the toxic ant and its "receptor," AChE. Several studies have reported, however, that AChE sensitivity to the inhibitors is not a contributing factor to age-related differences in sensitivity (Atterberry et aI., 1997; Benke and Murphy, 1975; Mortensen et aI., 1998). Upon extensive inhibition of AChE in the nervous system, the neurotransmitter acetylcholine accumulates in synapses, causing excessive stimulation of cholinergic receptors on postsynaptic cells, leading to cholinergic toxicity. It is known that feedback inhibition of acetylcholine release can occur through activation of muscarinic acetylcholine receptors located on presynaptic terminals (Allgaier et aI., 1993; Vickroy and Cadman, 1989; Weiler et aI., 1989). Activation of these presynaptic muscarinic receptors diminishes further acetylcholine release and thereby may reduce the excessive stimulation of postsynaptic cholinergic receptors following extensive AChE inhibition. Pedata and co-workers (1983) reported that muscarinic autoreceptor function was absent in 7-day-old rat brain but viable in brain from 21-day-old animals. Thus, with extensive AChE inhibition, very young rats do not have an adaptive mechanism which limits further neurotransmitter release in times of excess (e.g., when AChE is inhibited). Pedata and co-workers (1983) and Meyer and Crews (1984) reported that evoked acetylcholine release was lower in tissues from both neonatal and aged brain compared to animals 1-6 months of age. Differences in the amount of acetylcholine released upon stimulation between the age groups may therefore contribute to differences in response to AChE inhibitors. The function of muscarinic autoreceptors appears markedly reduced with aging in some rat brain regions (Araujo et aI., 1990). A deficit or lack of feedback inhibition of acetylcholine release in some age groups may limit their adaptation to synaptic AChE inhibition and contribute to higher sensitivity (Pope, 1999). Differences in acute sensitivity to OP anticholinesterases between neonatal and adult rats may therefore have both a toxicokinetic and a toxicodynamic basis. It should be stressed, however, that the studies cited above generally used lethality as the endpoint for estimating age-related sensitivity. By definition, dosages at or near those causing death would have to be considered "high"-level exposures. Less information is available regarding age-related differences in sensitivity to lower levels of exposure. While of prominent importance with acute, high-level exposures where detoxification systems may be saturated, differential detoxification capacities in neonatal and adult animals may have lesser importance when repeated, lower-level exposures occur. With lower, nonlethal dosages, less AChE activity would be inhibited, with lesser signs of cholinergic toxicity. At even lower dosages, some degree of AChE inhibition could occur in the absence of any overt toxicity (Nostrandt et aI., 1997).

878

CHAPTER 41


Under these conditions, feedback inhibition of acetylcholine release (or lack of that adaptive mechanism in neonatal animals) would have little consequence. Thus, with acute dosages of pesticide high enough to cause some level of AChE inhibition but with no alteration of cholinergic neurotransmission, two factors which apparently make younger animals more sensitive (lower detoxification capabilities, lesser adaptive regulation of neurotransmitter release) may have no relevance. With such repeated lower-level exposures, however, another toxicodynamic factor (i.e., recovery of AChE activity following inhibition) may play a more prominent role. As mentioned before, AChE activity following OP exposure recovers much faster in neonatal tissues (Pope et ai., 1991) and much slower in aged animals (Micha1ek et al., 1990) than in adult tissues. While neonatal rats were more sensitive to single, high dosages of chlorpyrifos, adults exhibited more extensive changes in cholinergic neurochemical markers (i.e., AChE inhibition, muscarinic receptor binding) following repeated, intermittent dosing (40 mg/kg, every four days for a total of four exposures) (Chakraborti et al., 1993; Pope and Liu, 1997). Apparently, while young rats are more sensitive to the acute effects of chlorpyrifos, they can recover much faster than adults from the biochemical insult. When exposures are sufficiently separated in time, neonatal animals can regain AChE activity faster and avoid cumulative inhibition with repeated exposures. In contrast, in particular with OPs such as chlorpyrifos which produce long-term inhibition of AChE, activity recovers more slowly in adult tissues, allowing accumulative inhibition with subsequent exposures. Thus, one could argue that with acute chlorpyrifos dosing, young animals are more sensitive than adults, but with repeated dosing, age-related sensitivity is reversed. Clearly, the nature of the exposures (acute vs. repeated, high level vs. low level) can influence age-related differences in sensitivity to these toxicants. Few studies have evaluated the relative sensitivity of aged animals to organophosphates. Acetylcholinesterase activity in some brain regions (e.g., hippocampus, cortex) but not others (e.g., pons-medulla) of rats declines with aging (Bisso et ai., 1991; Meneguz et aI., 1992). As mentioned before, recovery of AChE activity as well as muscarinic receptor binding following repeated organophosphate exposures was impaired in aging brain, in particular in the cerebral cortex (Michalek et ai., 1990). Age-related differences in baseline activity of cholinergic neurochemical processes or their adaptive responses to pesticide exposure could therefore influence sensitivity to some anticholinesterases. Veronesi and co-workers (1990) evaluated the effects of chronic fenthion exposure (25 mg/kg, three times a week for 10 months) in either young (2 months old) or aged (12 months old) rats. Using this dosing treatment schedule, chronic (10 months) fenthion exposures initiated in young rats produced gliosis and necrosis in the dentate gyrus and CA!, CA3, and sometimes CA2 regions of the hippocampus. Aged rats treated with the same regimen of fenthion exhibited similar degrees of hippocampal degeneration earlier during the progression of exposure, that is, by 2 months, and much more extensive pathol-

ogy than noted in the younger animals when evaluated following 10 months of exposure. These studies suggest that persistent acetycholinesterase inhibition by fenthion can produce neuropathological changes in the rat hippocampus and that aged rats are more sensitive than younger rats to such effects. Karanth and Pope (2000) compared acute sensitivity to chlorpyrifos and parathion in neonatal (7 days old), juvenile (21 days old), adult (90 days old), and aged (24 months old) Sprague Dawley rats. Neonatal and juvenile rats were more sensitive than adults to both toxicants. Adult and aged rats were similar in sensitivity to chlorpyrifos, but aged animals were markedly more sensitive than adults to parathion. Moreover, plasma carboxylesterase activity among groups was highly correlated with acute sensitivity to parathion, further suggesting a toxicokinetic basis for the age-related differences in sensitivity to this pesticide. The above discussion pertains to differences in sensitivity among different age groups to the cholinergic toxicity of OP pesticides. Some recent reports suggest that OP pesticides may affect macromolecular synthesis and cell viability in the brain following early postnatal exposures, independent of AChE inhibition (Slotkin, 1999). Whitney and co-workers (1995) reported that DNA and protein synthesis could be affected by chlorpyrifos in a time-dependent and brain regional-dependent manner. When postnatal rats (11-14 days of age) were given chlorpyrifos (1 mg/kg/day), a delayed reduction in DNA concentration and content in forebrain was noted at 15-20 days of age (Campbell et aI., 1997). Reductions in cellular RNA concentration and content were also reported in the brainstem and fore brain following repeated postnatal chlorpyrifos exposures in rats (Johnson et aI., 1998). Song and co-workers (1997) reported that repeated postnatal exposures to chlorpyrifos in rats affected multiple components of the adenylyl cyclase cascade system (e.g., activity of adenylyl cyclase, G-protein function, expression of neurotransmitter receptors coupled to adenylyl cyclase). Moreover, changes in these processes were noted in the cerebellum, a brain region with only sparse cholinergic innervation. Other studies suggest that anticholinesterases may affect neuronal adhesion and neurite extension (Bigbee et ai., 1999; Dupree and Bigbee, 1994; Small et aI., 1995; Song et aI., 1998). Thus, OP pesticides may be capable of altering macromolecule synthesis, intracellular signaling, and neuronal adhesion/outgrowth in the developing brain, apparently independent of catalytic inhibition of AChE. Some organophosphorus toxicants can induce a delayed neuropathological response referred to as organophosphorusinduced delayed neurotoxicity (OPIDN) (Abou-Donia, 1981). This form of neurotoxicity is not associated with AChE inhibition but has been correlated with the inhibition of another enzyme in the nervous system called neurotoxic esterase (NTE) (Johnson, 1976, 1980). Individuals affected by this delayed neurotoxicity exhibit gait disturbances (incoordination and difficulties in walking) and sensory deficits (numbness and tingling, particularly in the fingers and toes), which mayor may not follow signs of toxicity characteristic of AChE inhibition. Degeneration of certain nerve tracts in both the central and the


peripheral nervous systems has been demonstrated in OPIDN. It has more recently been observed that some compounds [e.g., the common protease and NTE inhibitor phenylmethylsulfonyl fluoride (PMSF)], while not being capable of inducing delayed neurotoxicity, can potentiate or promote delayed neurotoxicity caused by an OP (Lotti et aI., 1991; Pope and Padilla, 1990; Pope et aI., 1993). The sequence of administration of the two compounds is of paramount importance; that is, for delayed neurotoxicity to be exacerbated, OP exposure must precede exposure to the potentiating agent. Young animals are resistant to delayed neurotoxicity (Johnson and Barnes, 1970; Moretto et aI., 1991). Before the age of about 6-7 weeks, chickens (the animal model of choice for studies of delayed neurotoxicity) are completely resistant to functional and morphological signs of OPIDN. From about 7-10 weeks of age, sensitivity develops, and at about 12-14 weeks of age, they become completely sensitive (Moretto et aI., 1991; Pope et aI., 1992). Studies have also examined the potentiation of OPIDN in young animals (Peraica et aI., 1993; Pope et aI., 1992). As stated above, five-week-old chickens are normally resistant to the clinical and morphological changes associated with delayed neurotoxicity. If OP exposure is followed by treatment with PMSF, however, overt delayed neurotoxicity can be demonstrated. On the other hand, clinical and morphological changes typical of OPIDN are generally not elicited in very young chickens (e.g., two weeks of age) regardless of the dose of the OP or whether a potentiating agent is given after the OP (Harp et aI., 1997). Just as the mechanism(s) underlying OPIDN itself has not been elucidated, the basis for age-related differences in sensitivity to delayed neurotoxicity remains unknown. In contrast to age-related sensitivity to acute toxicity from most OPs, however, young are less sensitive to the delayed neurotoxicity of OPs. 41.3.2 ORGANOCHLORINE INSECTICIDES At one time, organochlorines (OCs) constituted the highest-use pesticide class in the world. With increased awareness of ecological damage, global contamination, and insect resistance, the use ofOCs has decreased. The most well-known OC, DDT, has been extensively studied. In acute toxicity studies, DDT is actually less toxic to neonatal rats than to adults (Lu et aI., 1965). In this same study, dieldrin, another OC, was also reported to be less toxic in neonatal rats. Later studies have suggested that early neonatal exposure to DDT (0.5 mg/kg, po) can have long-lasting consequences (Eriksson et aI., 1984, 1993). Total cholinergic muscarinic receptor (eH]QNB) density was increased in cortex one week after DDT exposure in lO-day-old rats, but no effect was noted in hippocampus. Moreover, muscarinic receptor binding was still altered at 4 months of age following this single treatment with DDT, but at this time there was a reduction in binding density. Functional alterations (deficits in locomotor habituation) were also noted in rats four months after acute DDT exposure (Eriksson, 1997). Of particular interest in these studies

879

was the observation that neonatal (10 days old) rats treated with DDT (0.5 mg/kg) showed an increase in cortical muscarinic receptor binding one week after exposure, whereas adult rats treated similarly showed a decrease in receptor binding. Moreover, neither 3-day-old rats nor 19-day-old rats showed the same response (i.e., up-regUlation of muscarinic receptors) when treated similarly with DDT (Eriksson, 1997). Subsequent studies have shown that 1O-day-old mice treated with DDT and then challenged at five months of age with bioallethrin showed increased expression of the m4 subtype of muscarinic receptors in selected brain regions (cortex and striatum) (Talts et aI., 1998b). Thus, there appears to be a critical developmental window in which alteration of the cholinergic system can occur following early DDT exposure, and changes in muscarinic receptor density induced by DDT appear specific for the m4 subtype. While most OCs have been banned from use in the United States, their use continues in other countries. A few OCs are still used in the United States, for example, lindane (y-hexachlorocyclohexane). Lindane is still commonly prescribed for treatment of scabies and pediculosis. Rivera et al. (1990) reported that repeated, relatively low-level exposures to lindane (10 mg/kg/day for seven days) during postnatal week 1 or 2 induced transient changes in reflex behaviors (e.g., surface righting, cliff avoidance) and locomotor hyperactivity, in the absence of overt signs of toxicity. Serrano and co-workers (1990) reported that early postnatal lindane exposure reduced the level of myelin basic protein and 2',3' -cyclic nucleotide 3' -phosphodiesterase activity, an enzyme in high concentrations in myelin and myelin-forming cells, in a dose-dependent manner. Lindane exposure (either acute [20 mg/kg] or repeated [10 mg/kg/day for seven days]) in rats 15 days of age caused complex behavioral changes (improvement in passive avoidance behavior, alterations in locomotor activity) and apparent enhanced turnover of brain monoaminergic neurotransmitters (Rivera et aI., 1998). While these studies only evaluated toxicity in postnatally maturing animals, the endpoints evaluated and the changes noted suggested that higher sensitivity may exist in younger individuals. Samanta and Chainy (1997) reported that acute lindane exposure (50 mg/kg, i.p.) caused only minimal lipid peroxidation in liver of 30-day-old chickens but more extensive oxidative changes in 7 -day-old animals. Furthermore, superoxide dismutase was inhibited and glutathione levels were elevated by lindane in 7-day-old but not 30-day-old chickens. Thus, lindane can cause diverse age-related changes that generally target younger animals. Endosulfan is another OC that is still commonly used in the United States. Kiran and Varma (1988) studied the toxicity of endosulfan in different age groups of rats (12.5 mg/kg/day for four days beginning at 15,30,70, and 365 days of age). Hyperglycemia and glycogen depletion were most extensive in 365day-old animals and least affected in the youngest age group. Liver aldolase activity was also reduced more in older rats than in younger animals. In contrast, red blood cell Na+IK+ ATPase activity was inhibited more in the youngest age group. These

880

CHAPTER 41


results indicate complex age-related differences in response to endosulfan. Chlordecone is an organochlorine that causes hyperexcitability, tremors, incoordination, and other signs of neurotoxicity (Tilson and Mactutus, 1982). Several studies have evaluated the effects of early postnatal exposure to chlordecone. Tilson and co-workers (1982) reported that rats exposed acutely on postnatal day 4 to chlordecone had markedly altered responses during reversal of visually cued nose poke behavior when tested at about four months of age. Neonatal chlordecone exposure was also reported to alter passive avoidance performance (Mactutus et aI., 1982). Jinna and co-workers (1989) reported that chlordecone inhibited rat brain ATPases (Na+ /K+ ATPase, Ca++ ATPase) in an age-related manner, that is, neonatal enzyme activity was more sensitive to inhibition by chlordecone in vitro. Chlordecone has been shown to potentiate the hepatotoxicity of halogenated solvents, for example, carbon tetrachloride (Soni and Mehendale, 1998). Twenty and 45-day-old rats were resistant to chlordecone-enhanced hepatotoxicity relative to 60-day-old animals, however (Dalu and Mehendale, 1996). Dosages of chlordecone (10 ppm in the diet for 15 days) and carbon tetrachloride (0.1 ml/kg, i.p.) that caused 100% lethality in the adult rats caused 0% and 25% lethality in 20- and 45-dayold animals. It was concluded from these studies that the relative ability of the liver to recover from injury was the prominent factor underlying age-related differences in toxic outcome, with immature animals being more competent than adults at restoring tissue integrity and function. Thus, while these studies do not indicate age-related differences in sensitivity to chlordecone alone, they suggest that the modulation of solvent hepatotoxicity by chlordecone can occur in an age-related manner. Many of the OCs, for example, DDT, chlordecone, methoxychlor, chlordane, and endosulfan, have also been noted to interact directly with hormonal receptors (Tilson, 1998). The DDT analog, methoxychlor, is still used in the United States and has been shown to both alter sex-related hormones and reproductive function in rats treated postnatally (Chapin et aI., 1997). The endocrine-disrupting capacity of these agents could be cause for concern with early exposures (Cassidy et aI., 1994; Chapin et aI., 1997; Davis et aI., 1993). The reader is referred to Chapter 16 for more information on endocrine disruption by pesticides. 41.3.3 CARBAMATES

Knisely and Hamm (1989) investigated the comparative actions of physostigmine on nociception in different age groups of rats (3, 17, and 25 months of age). Tail-flick latencies were dose-dependently altered in all age groups by physostigmine, but more extensive increases in latency were noted in the 17- and 25-month-old animals with higher dosages, suggesting higher sensitivity in the aged animals to this carbamate anticholinesterase. Such changes could be an indication of upregulation of cholinergic receptors due to loss of cholinergic innervation with aging.

Takahashi and co-workers (1991) compared the motor, sensory, and thermoregulatory responses of young adults (3 months of age) and older adults (12 months of age) to carbaryl (10 or 50 mg/kg, p.o.). Carbaryl affected nociception primarily in the older animals. Hypothermia was also affected in an agerelated manner. Locomotor changes following carbamate exposure, however, were similar between the two age groups. Again, these data illustrate the potential for age-related differences in response to a pesticide when one endpoint is used and conversely, lack of age-related differences in sensitivity when based on another endpoint. Moser (1999) reported that aldicarb was about twice as toxic in preweanling rats compared to adults using the acute maximum tolerated dose as the endpoint of sensitivity. Interestingly, preweanling rats exhibited fewer signs of functional toxicity than older animals, in the presence of similar levels of brain and blood cholinesterase inhibition. Furthermore, the young rats were resistant to locomotor alterations noted in older animals following aldicarb administration. Lifshitz and co-workers (1997) retrospectively compared the clinical course of poisoning in children (1-8 years of age) and adults (17--41 years of age) following carbamate pesticide exposures. In all cases, blood serum cholinesterase inhibition was approximately the same (10-30% below the lower limit of normal). Interestingly, signs of coma/stupor and hypotonia were noted in 100% of the children but in none of the adults. While miosis was noted in 92% of the adults, this sign was only recorded in 55% of the children. Moreover, muscle fasciculations were observed in 83% of the adults and in only 6% of the children. While the relative level of AChE inhibition in the target tissues was unknown, these results suggest that children may respond differently than adults following acute anticholinesterase exposures producing relatively similar degrees of blood cholinesterase inhibition. 41.3.4 PYRETHROID INSECTICIDES

Eriksson and Nordberg (1990) studied the effects of early postnatal exposures to one of two different pyrethroid insecticides, bioallethrin (a type I pyrethroid) and deltamethrin (a type 11 pyrethroid), on cholinergic receptors in mouse brain. With lower levels of exposure, bioallethrin (0.72 mg/kg/day from postnatal day 10-16) reduced high-affinity muscarinic receptor binding in brain, whereas deltamethrin (0.71 mg/kg/day) increased high-affinity binding, both in the absence of overt signs of toxicity. Deltamethrin also increased cortical [3H]nicotine binding. Higher levels of repeated exposure (72 and 1.2 mg/kg/ day for bioallethrin and deltamethrin, respectively) caused overt toxicity (tremor, choreoathetosis), but only deltamethrin affected cholinergic receptor binding under these conditions. Early exposure to bioallethrin in mice (0.7 mg/kg/day from postnatal day 10-16) was also reported to increase sensitivity to bioallethrin when administered at 7 months of age, suggesting long-term changes in sensitivity following exposure during postnatal maturation (Talts et aI., 1998a). These studies, similar


to those with early postnatal exposures to DDT (Eriksson et aI., 1984), indicate that development of some components of the cholinergic system may be sensitive to alteration by early postnatal exposure to "noncholinergic" pesticides (i.e., pesticides not having a primary action on some aspect of cholinergic neurotransmission). Cantalamessa (1993) compared the acute toxicity and metabolism of cypermethrin and permethrin in neonatal and adult rats. With both pesticides, an age-related decrease in acute toxicity was noted. Cypermethrin and permethrin were 16.8 and 4.4 times more toxic (based on 24-hour oral LDso values) in 8-day-old animals compared to adults. Carboxylesterase inhibition (by tri-ortho-cresyl phosphate) in neonatal animals failed to alter acute toxicity, but lethality was increased in adults by this pretreatment, suggesting that neonatal animals may be more sensitive to acute toxicity of these pyrethroids at least partially because of incomplete development of this detoxification system. Sheets and co-workers (1994) evaluated the sensitivity of preweanling, wean ling, and adult rats to a wide dose range of deltamethrin. Younger rats (11 and 21 days of age) were markedly more sensitive than adults (72 days of age) to the acute lethality of deltamethrin (LDso: 11 days = 5.1 mg/kg; 21 days = 11 mg/kg; 72 days = 81 mg/kg, p.o.). In contrast, using acoustic startle response to evaluate functional toxicity of lower-level exposures, the ED50 was the same between Il-dayold and 72-day-old animals. Based on these studies, age-related differences in sensitivity to deltamethrin could be considered substantial (if based on acute lethality) or nonexistent (if based on the acoustic startle response). Clearly, the selection of the endpoint used to define sensitivity, as well as the exposure conditions, can qualitatively influence determination of age-related susceptibility to these pesticides. 41.3.5 MISCELLANEOUS PESTICIDES

Gaines and Linder (1986) examined the comparative acute sensitivity of weanling (4-6 weeks of age) and adult rats to 34 pesticides from different chemical classes. The immature rats were more sensitive to only four of those pesticides. Moreover, differences in acute sensitivity to pesticides were generally only two- to threefold in magnitude. One problem with this study, however, was the age of the younger animals used, that is, 4- to 6-week-old rats. Similar studies using less mature animals (e.g., 1- to 3-week-old rats) may have yielded different conclusions. Watkinson studied the cardiotoxicity of the formamidine pesticide chlordimeform. Weanling (22-30 days of age [Watkinson, 1985]) and aged (24 months of age [Watkinson, 1986]) rats were treated sequentially with 5, lO, 30, 60, and 120 mg/kg chlordimeform (i.v.) or vehicle, and mean arterial blood pressure and heart rate were monitored. While chlordimeform reduced heart rate and blood pressure in both age groups, the magnitude of the changes was greater in the aged animals. Arrythmias were also less pronounced in younger animals and required higher thresholds of chlordimeform. In addition, while

881

a single injection of chlordimeform (60 mg/kg, i.v.) was lethal to all aged rats tested, only 23% of the weanling rats died following this level of exposure. Lower sensitivity of young rats to the lethality of chlordimeform had been previously reported (Robinson and Smith, 1977). Thus, it appears that younger animals are less sensitive than aged rats to the toxicity of the formamidine insecticide, chlordimeform. Ivermectin is a broad-spectrum antiparasitic agent (Campbell and Benz, 1984). Relatively low-level exposure to ivermectin (4 mg/kg/day) during gestation (GD 6-20) and lactation (postnatal days 2-20) caused lOO% lethality in pups with no apparent toxicity in dams (Poul, 1988). When exposure was limited to gestation, only 22% lethality was noted in the offspring. Lower exposure levels (1 mg/kg/day) had no effect on survival but delayed some developmental endpoints, including cliff avoidance and locomotion. Lankas and co-workers (1989) reported that newborn rodents were particularly sensitive to the neurotoxicity of ivermectin. Following application of ivermectin to control an ectoparasite infestation, Skopets and coworkers (1996) noted evidence of higher sensitivity of young mice to ivermectin. While all adults tolerated the ivermectin exposures, preweanling mice developed seizures or tremors, and lethality was observed in some cases. Together, these data suggest that neonatal rodents are more sensitive than adults to the acute toxicity of ivermectin. As ivermectin is typically prevented from access to the central nervous system in adults (Lovell, 1990), incomplete blood-brain barrier formation appears to contribute to these age-related differences in sensitivity (Lankas et aI., 1989). Age-related differences in sensitivity were noted following acute dibromochloropropane exposure (250 mg/kg, s.c.) in 4- and 9-week-old rats (Saegusa, 1987). It was noted that the older animals exhibited a higher incidence of lethality, more extensive body weight reductions, and more extensive tissue damage in kidney, intestine, and testes. Dithiobiuret (DTB, thioimidodicarbonic diamide) was originally proposed as a rodenticide and is a prototypical motor neuron toxicant that produces a flaccid weakness following repeated exposures (Atchison et aI., 1982). Using failure of the rotorod test as an indication of neuromuscular toxicity, Atchison and co-workers (1982) studied the sensitivity of weanling (25 days old), juvenile (50 days old), and adult (80 days old) rats to DTB (1 mg/kg/day). In females, the mean time to onset of rotorod failure was about six days in wean ling rats, four days in juveniles, and only about three days in adults. Neither differences in total accumulation of DTB nor distribution appeared to contribute to the differences in DTB toxicity among the age groups. These data provide another example of higher sensitivity to neurotoxic ants in adults compared to younger animals. Using a series of immunotoxicity assays, Smialowicz and co-workers (1989) reported that preweanling rats (3-24 days of age) were somewhat more sensitive than adults to tributyltininduced immune alterations. In addition, natural killer cell activity was only affected in the neonatal animals. Furthermore, some immune responses were altered in lO-week-old animals

882

CHAPTER 41


treated prior to weaning, suggesting long-term changes in immune function could occur with early exposure to tributyltin. Children may be more sensitive to the insect repellent, DEET (diethyl m-toluamide) (Couch and Johnson, 1992). DEET is used safely by an estimated 200 million people each year around the world (Brown and Hebert, 1997), but severe neurological manifestations have occasionally been associated with its use (Osimitz and Murphy, 1997). Four boys (age 3-7 years) had seizures within 48 hours of applying DEET to the skin. Six young girls (ages 1.5-8 years) exhibited seizures, ataxia, and/or coma after dermally applying DEET, and three of those children later died. These types of neurological signs have been reported in adults following oral consumption of large amounts of DEET (Tenenbeim, 1987). Thus, while occurrence is rare, children may exhibit serious signs and symptoms of neurotoxicity and can die following dermal application of this widely used repellent. Because of the scarcity of data on absorption, metabolism, or elimination of DEET in children, it is unclear why children may be more sensitive to this compound (Garrettson, 1997).

41.4 CONCLUSIONS Changes in sensitivity to pesticides can occur throughout the life span from early postpartum to senescence. Recently, there has been considerable concern that children may be at higher risk than adults to pesticides. Enactment of the Food Quality Protection Act in 1996 was in response to this concern, and called for consideration of additional safety in the risk assessment of pesticides to protect infants and children. It is clear from review of both experimental and clinical data, however, that there is no hard-and-fast rule regarding age-related differences in sensitivity to pesticides. While neonates may be more sensitive to the acute toxicity of some pesticides, adults may be more sensitive to others. Even within a class of toxic ants, for example, within the organophosphorus anticholinesterases, examples of higher sensitivity in both age groups can be demonstrated. In fact, even when a single pesticide is considered, age-related differences in sensitivity may change qualitatively depending on the conditions of exposure (e.g., acute vs. repeated dosing, high- vs. low-level exposures) or the endpoint measured. While maturational differences in biotransformation capacity may be limiting in some cases, for example, with acute, high-level exposures where detoxification enzymes could become saturated, such metabolic differences may be of lesser importance with repeated, lower levels of exposure to the same pesticides. Similarly, differences in the ability to recover following pesticide exposure may be much more important when repeated exposures occur than following acute exposures. Storage and clearance of pesticides may also be more important with repeated, long-term exposures. Age-related sensitivity to pesticides should therefore be evaluated on a case-by-case basis, recognizing both the factors which influence age-related differences in response and the critical importance of appropriate endpoint selection for establishing differential sensitivity.

Relative sensitivity can be expressed in one of two ways; that is, a subpopulation exhibits differences in sensitivity to a particular form of toxicity or a subpopulation exhibits qualitatively different forms of toxicity to the same pesticide. Young animals may be more sensitive to the acute lethality of some pesticides (e.g., chlorpyrifos), but this does not necessarily mean that young animals will be more sensitive to the same pesticides when sensitivity is based on nonlethal endpoints of toxicity. Risk assessments are typically performed using a "critical" endpoint, generally the most sensitive end point to the toxicant in question derived from a series of toxicity studies. Thus, even if a pesticide causes a qualitatively different form of toxicity in different age groups, the risk assessment and estimation of tolerable exposure levels will not change unless this response occurs at dosages lower than those defining the critical endpoint. There will always be uncertainty in risk assessment. One factor which contributes to that uncertainty is age and its influence on the response to a particular toxic ant. If the critical endpoint for a particular pesticide is well established based on "reliable" data derived from studies across all age groups, the contribution of age-related differences in sensitivity to such uncertainty can be minimized. Knowledge of mechanisms which contribute to such age-related differences in response to pesticides will ultimately aid in the safer use of these chemicals.

REFERENCES Abou-Donia, M. B. (1981). Organophosphorus ester-induced delayed neurotoxicity. Annu. Rev. Pharmacal. Taxicol. 21, 511-548. Allgaier, c., Choi, B. K, and Hertting, G. (1993). Muscarine receptors regulating electrically evoked release of acetylcholine in hippocampus are linked to pertussis toxin-sensitive G proteins but not to adenylate cyclase. J. Neurachem. 61, 1043-1049. Araujo, D. M., Lapchak, P. A., Meaney, M. J., Collier, B., and Quirion, R (1990). Effects of aging on nicotinic and muscarinic autoreceptor function in the rat brain: Relationship to presynaptic cholinergic markers and binding sites. J. Neurasci. 10,3069-3078. Atchison, W. D., Dickins, J., and Peterson, R E. (1982). Age dependence of dithiobiuret neurotoxicity in male and female rats. Neurotaxicalagy 3, 233241. Atterberry, T. T., Bumett, W. T., and Chambers, J. E. (1997). Age-related differences in parathion and chlorpyrifos toxicity in male rats: Target and nontarget esterase sensitivity and cytochrome P450-mediated metabolism. Toxicol. Appl. Pharmacol. 147,411-418. Augustinsson, K-B., and Barr, M. (1963). Age variation in plasma arylesterase activity in children. Clin. Chem. Acta 8, 568-573. Bames, D. G., and Dourson, M. (1988). Reference dose (RID): Description and use in health risk assessments. Regul. Toxicol. Pharmacol. 8, 471-486. Bearer, C. F. (1995). Environmental health hazards: How children are different from adults. Future. Child 5, 11-26. Benke, G. M., and Murphy, S. D. (1975). The influence of age on the toxicity and metabolism of methyl parathion and parathion in male and female rats. Toxicol. Appl. Pharmacol. 31, 254-269. Bigbee, J. w., Sharma, K v., Gupta, J. J., and Dupree, J. L. (1999). Morphogenic role for acetylcholinesterase in axonal outgrowth during neural development. Environ. Health Perspect. 107 (Suppll), 81-87. Bisso, G. M., Briancesco, R, and Michalek, H. (1991). Size and charge isomers of acetylcholinesterase in the cerebral cortex of young and aged rats. Neurochem. Res. 16, 571-575.

References

Borghoff, S. J., Stefanski, S. A., and Birnbaum, L. S. (1988). The effect of age on the glucuronidation and toxicity of 4,4'-thiobis(6-t-butyl-m-cresol). Toxicol. Appl. Pharmacol. 92, 453-466. Brodeur, J., and DuBois, K. P. (1963). Comparison of the acute toxicity of anticholinesterase insecticides to weanling and adult male rats. Proc. Soc. Exp. BioI. Med. 114, 509-511. Brown, M., and Hebert, A. A. (1997). Insect repellents: An overview. J. Am. Acad. Dermatol. 36,243-249. Campbell, C. G., Seidler, F. J., and Slotkin, T. A. (1997). Chlorpyrifos interferes with cell development in rat brain regions. Brain Res. Bull. 43, 179-189. Campbell, W. c., and Benz, G. W. (1984). Ivermectin: A review of efficacy and safety. J. Vet. Pharmacol. Ther. 7,1-16. Cantalamessa, F. (1993). Acute toxicity of two pyrethroids, permethrin, and cypermethrin in neonatal and adult rats. Arch. Toxicol. 67,510-513. Cassidy, R. A., Vorhees, C. V., Minnema, D. J., and Hastings, L. (1994). The effects of chlordane exposure during pre- and postnatal periods at environmentally relevant levels on sex steroid-mediated behaviors and functions in the rat. Toxicol. Appl. Pharmacol. 126, 326-337. Chakraborti, T. K., Farrar, J. D., and Pope, C. N. (1993). Comparative neurochemical and neurobehavioral effects of repeated chlorpyrifos exposures in young and adult rats. Pharmacol. Biochem. Behav. 46, 219-224. Chapin, R. E., Harris, M. Davis, B. J., Ward, S. M., Wilson, R. E., Mauney, M. A., Lockhart, A. C., Smialowicz, R. J., Moser, V. c., Burka, L. T., and Collins, B. J. (1997). The effects of perinatal/juvenile methoxychlor exposure on adult rat nervous, immune, and reproductive system function. Fundam. Appl. Toxicol.40, 138-157. Costa, L. G., McDonald, B. E., Murphy, S. D., Omenn, G. S., Richter, R. J., Motulsky, A. G., and Furlong, C. E. (1990). Serum paraoxonase and its influence on paraoxon and chlorpyrifos-oxon toxicity in rats. Toxicol. Appl. Pharmacol. 103,66-76. Couch, P., and Johnson, C. E. (1992). Prevention of Lyme disease. Am. J. Hosp. Pharm.49,1164-1173. Dalu, A., and Mehendale, H. M. (1996). Efficient tissue repair underlies the resiliency of postnatally developing rats to chlordecone + CCl4 hepatotoxicity. Toxicology 111, 29-42. Das, M., Dixit, R., Seth, P. K., and Mukhtar, H. (1981). Glutathione-Stransferase activity in the brain: Species, sex, regional, and age differences. J. Neurochem. 36, 1439-1442. Davis, D. L., Bradlow, H. L., Wolff, M., Woodruff, T., Hoel, D. G., and AntonCulver, H. (1993). Medical hypothesis: Xenoestrogens as preventable causes of breast cancer. Environ. Health Perspect. 101, 372-377. Deichmann, W. B. (1972). Toxicology of DDT and related chlorinated hydrocarbon pesticides. 1. Occup. Med. 14, 285-292. Diggory, H. J., Landrigan, P. J., Latimer, K. P., Ellington, A. C, Kimbrough, R. D., Liddle, J. A., Cline, R. E., and Smrek, A. L. (1977). Fatal parathion poisoning caused by contamination of flour in international commerce. Am. J. Epidemiol. 106, 145-153. Done, A. K. (1964). Developmental pharmacology. Clin. Pharmacol. Ther.5, 432-479. Dupree, J. L., and Bigbee, J. W. (1994). Retardation of neuritic outgrowth and cytoskeletal changes accompany acetylcholinesterase inhibitor treatment in cultured rat dorsal root ganglion neurons. J. Neurosci. 39, 567-575. Ecobichon, D. J., and Stephens, D. S. (1973). Perinatal development of human blood esterases. Clin. Pharmacol. Ther. 14,41-47. Egaas, E., Falls, J. G., and Dauterman, W. C. (1995). A study of gender, strain and age differences in mouse liver glutathione-S-transferase. Comp. Biochem. Physiol. C. Pharmacol. Toxicol. Endocrinol. 110, 35-40. Environmental Protection Agency (EPA) (1998a). EPA statement on human testing. http://www.epa.gov/scipoly/sap/l998/december/epastmt.htm. Environmental Protection Agency (EPA) (1998b). Presentation for FIFRA Scientific Advisory Panel by Office of Pesticide Programs, Health Effects Division on FQPA Safety Factor for Infants and Children. http://www.epa.gov/scipoly/sap/1998/marchllOx.htm. Environmental Protection Agency (EPA) (1999). The office of pesticide program's policy on determination of the appropriate FQPA safety factor(s) for use in the tolerance setting process. http://www.epa.gov/scipoly/sap/1999/may/lOxpoli.pdf.

w.,

883

Eriksson, P. (1997). Developmental neurotoxicity of environmental agents in the neonate. Neurotoxicology 18,719-726. Eriksson, P., and Nordberg, A. (1990). Effects of two pyrethroids, bioallethrin and deltamethrin, on subpopulations of muscarinic and nicotinic receptors in the neonatal mouse brain. Toxicol. Appl. Pharmacol. 102, 456-463. Eriksson, P., Falkeborn, Y., Nordberg, A., and Slanina, P. (1984). Effects of DDT on muscarine- and nicotine-like binding sites in CNS of immature and adult mice. Toxicol. Lett. 22, 329-334. Eriksson, P., Johansson, U., Ahlbom, J., and Fredriksson, A. (1993). Neonatal exposure to DDT induces increased susceptibility to pyrethroid (bioallethrin) exposure at adult age-Changes in cholinergic muscarinic receptor and behavioural variables. Toxicology 77, 21-30. Fenner-Crisp, P. A. (1995). Pesticides-the NAS report: How can the recommendations be implemented? Environ. Health Perspect. 103 (Suppl 6), 159-162. Fenske, R. A., Black, K. G., Elkner, K. P., Lee, C L., Methner, M. M., and Soto, R. (1990). Potential exposure and health risks of infants following indoor residential pesticide applications. Am. J. Public Health 80, 689-693. Fukuto, T. R. (1990). Mechanism of action of organophosphorus and carbamate insecticides. Environ. Health Perspect. 87, 245-254. Gagne, J., and Brodeur, J. (1972). Metabolic studies on the mechanisms of increased susceptibility of weanling rats to parathion. Can. J. Physiol. Pharmacol. 50,902-915. Gaines, T. B., and Linder, R. E. (1986). Acute toxicity of pesticides in adult and weanling rats. Fundam. Appl. Toxicol. 7,299-308. Garrettson, L. (1997). Commentary-DEET: Caution for children still needed. J. Toxicol. Clin. Toxicol. 35, 443-445. Goldenthal, E. 1. (1971). A compilation of LD50 values in newborn and adult animals. Toxicol. Appl. Pharmacol. 18, 185-207. Goldman, L. R. (1995). Case studies of environmental risks to children. Future. Child 5,27-33. Hall, L. L., Fisher, H. L., Sumler, M. R., Hughes, M. F., and Shah, P. V. (1992). Age-related percutaneous penetration of 2-sec-butyl-4,6-dinitrophenol (dinoseb) in rats. Fundam. Appl. Toxicol. 19,258-267. Harp, P., Tanaka, D. J., and Pope, C. N. (1997). Potentiation of organophosphorus-induced delayed neurotoxicity following phenyl saligenin phosphate exposures in 2-, 5- and 8-week-old chickens. Fundam. Appl. Toxicol. 37, 64-70. Jinna, R. R., Uzodinma, J. E., and Desaiah, D. (1989). Age-related changes in rat brain ATPases during treatment with chlordecone. J. Toxicol. Environ. Health 27,199-208. Johnson, D. E., Seidler, F. J., and Slotkin, T. A. (1998). Early biochemical detection of delayed neurotoxicity resulting from developmental exposure to chloropyrifos. Brain Res. Bull. 45,143-147. Johnson, M. K. (1976). Mechanism of protection against the delayed neurotoxic effects of organophosphorus esters. Fed. Proc. 35, 73-74. Johnson, M. K. (1980). The mechanism of delayed neuropathy caused by some organophosphorus esters: Using the understanding to improve safety. J. Environ. Sci. Health [Bl 15,823-841. Johnson, M. K., and Barnes, J. M. (1970). Age and the sensitivity of chicks to the delayed neurotoxic effects on some organophosphorus compounds. Biochem. Pharmacol. 19,3045-3047. Karanth, S., and Pope, C. (2000). Carboxylesterase and A-esterase activities during maturation and aging: Relationship to the toxicity of chlorpyrifos and parathion in rats. Toxicological Sci. 58,282-289. Kemp, J. R., and Wallace, K. B. (1990). Molecular determinants of the speciesselective inhibition of brain acetylcholinesterase. Toxicol. Appl. Pharmacol. 104, 246-258. Kiran, R., and Varma, M. N. (1988). Biochemical studies on endosulfan toxicity in different age groups of rats. ToxicoZ. Lett. 44, 247-252. Knisely, J. S., and Hamm, R. J. (1989). Physostigmine-induced analgesia in young, middle-aged, and senescent rats. Exp. Aging Res. 15,3-11. Lankas, G. R., Minsker, D. H., and Robertson, R. T. (1989). Effects of ivermectin on reproduction and neonatal toxicity in rats. Food Chem. Toxicol. 27,523-529.

884

CHAPTER 41


Lee, C., Stonestreet, B. S., Oh, W., Outerbridge, E. W., and Cashore, W. J. (1995). Postnatal maturation of the blood-brain barrier for unbound bilirubin in newborn piglets. Brain Res. 689, 233-238. Li, W. E, Costa, L. G., and Furlong, C. E. (1993). Serum paraoxonase status; a major factor in determining resistance to organophosphates. 1. Toxicol. Environ. Health 40, 337-346. Li, W. E, Furlong, C. E., and Costa, L. G. (1995). Paraoxonase protects against chlorpyrifos toxicity in mice. Toxicol. Lett. 76,219-226. Lifshitz, M., Shahak, E., Bolotin, A., and Sofer, S. (1997). Carbamate poisoning in early childhood and in adults. 1. Toxicol. Clin. Toxicol. 35,25-27. Little, D. N. (1995). Children and environmental toxins. Prim. Care 22, 69-79. Long, G. G., Scheidt, A. B., Everson, R. J., and Oehme, E W. (1986). Age related susceptibility of newborn pigs to the cutaneous application of chlorpyrifos. Vet. Hum. Toxicol. 28, 297-299. Lotti, M., Caroldi, E., Capodicasa, E., and Moretto, A. (1991). Promotion of organophosphate-induced delayed polyneuropathy by phenylmethylsulfonyl fluoride. Toxicol. Appl. Pharmacol. 108,234-241. Lovell, R. A. (1990). Ivermectin and piperazine toxicoses in dogs and cats. Vet. Clin. North Am. Small. Anim. Pract. 20, 453-468. Lu, C., and Fenske, R. A. (1999). Dermal transfer of chlorpyrifos residues from residential surfaces: Comparison of hand press, hand drag, wipe, and polyurethane foam roller measurements after broadcast and aerosol pesticide applications. Environ. Health Perspect. 107,463-467. Lu, E c., Jessup, D. c., and Lavallee, A. (1965). Toxicity of pesticides in young versus adult rats. Food Cosmet. Toxicol. 3, 591-596. Mactutus, C. E, Unger, K. L., and Tilson, H. A. (1982). Neonatal chlordecone exposure impairs early learning and memory in the rat on a multiple measure passive avoidance task. NeuroToxicology 3, 27-44. Maxwell, D. M. (1992). The specificity of carboxylesterase protection against the toxicity of organophosphorus compounds. Toxicol. Appl. Pharmacol. 114,306-312. Mehendale, H. M. (1980). Aldrin epoxidase activity in developing rabbit lung. Pediatr. Res. 14,282-285. Mendoza, C. E. (1976). Toxicity and effects of malathion on esterases of suckling albino rats. Toxico!. App!. Pharmacol. 35, 229-238. Meneguz, A., Bisso, G. M., and Michalek, H. (1992). Age-related changes in acetylcholinesterase and its molecular forms in various brain areas of rats. Neuroehem. Res. 17, 785-790. Meyer, E. M., and Crews, E T. (1984). Effects of aging on rat cortical presynaptic cholinergic processes. Neurobiol. Aging 5,315-317. Michalek, H., Fortuna, S., Volpe, M. T., and Pintor, A. (1990). Age-related differences in the recovery rate of brain cholinesterases, choline acetyltransferase and muscarinic acetylcholine receptor sites after a subacute intoxication of rats with the anticholinesterase agent, isofluorophate. Acta Neurobiol. Exp. (Warn) 50, 237-249. Mileson, B. E., Chambers, J. E., Chen, W. L., Dettbarn, w., Ehrich, M., Eldefrawi, A. T., Gaylor, D. W., Hamernik, K., Hodgson, E., Karczmar, A. G., Padilla, S., Pope, C. N., Richardson, R. J., Saunders, D. R., Sheets, L. P., Sultatos, L. G., and Wallace, K. B. (1998). Common mechanism of toxicity: A case study of organophosphorus pesticides. Toxico!. Sci. 41, 8-20. Moretto, A., Capodicasa, E., Peraica, M., and Lotti, M. (1991). Age sensitivity to organophosphate-induced delayed polyneuropathy. Biochemical and toxicological studies in developing chicks. Biochem. Pharmacol. 41, 14971504. Mortensen, S. R., Chanda, S. M., Hooper, M. J., and Padilla, S. (1996). Maturational differences in chlorpyrifos-oxonase activity may contribute to agerelated sensitivity to chlorpyrifos. 1. Biochem. Toxicol. 11,279-287. Mortensen, S. R., Hooper, M. J., and Padilla, S. (1998). Rat brain acetylcholinesterase activity: Developmental profile and maturational sensitivity to carbamate and organophosphorus inhibitors. Toxicology 125,13-19. Moser, V. C. (1999). Comparison of aldicarb and methamidophos neurotoxicity at different ages in the rat: Behavioral and biochemical parameters. Toxieo!. Appl. Pharmacol. 157,94-106. Moser, V. c., Chanda, S. M., Mortensen, S. R., and Padilla, S. (1998). Ageand gender-related differences in sensitivity to chlorpyrifos in the rat reflect developmental profiles of esterase activities. Toxicological Sci. 46, 211222.

Moser, V. C., and Padilla, S. (1998). Age- and gender-related differences in the time course of behavioral and biochemical effects produced by oral chlorpyrifos in rats. Toxicol. Appl. Pharmacol. 149, 107-119. Mussalo-Rauhamaa, H., Pyysalo, H., and Moilanen, R. (1984). Influence of diet and other factors on the levels of organochlorine compounds in human adipose tissue in Finland. 1. Toxicol. Environ. Health 13,689-704. National Academy of Sciences (NAS) (1993). "Pesticides in the Diets ofInfants and Children." National Academy Press, Washington, DC. Natural Resources Defense Council (NRDC) (1989). "Intolerable Risks: Pesticides in Our Children's Food." Natural Resources Defense Council, New York, NY. Nostrandt, A. c., Padilla, S., and Moser, V. C. (1997). The relationship of oral chlorpyrifos effects on behavior, cholinesterase inhibition, and muscarinic receptor density in rat. Pharmacol. Biochem. Behav. 58, 15-23. Osimitz, T. G., and Murphy, J. V. (1997). Neurological effects associated with use of the insect repellent N,N -diethyl-m-toluamide (DEET). 1. Toxicol. Clin. Toxicol. 35,435--441. Overstreet, D. H. (2000). Organophosphate pesticides, cholinergic function and cognitive performance in advanced age. NeuroToxicology 21, 75-81. Padilla, S., Buzzard, J., and Moser, V. C. (2000). Comparison of the role of esterases in the differential age-related sensitivity to chlorpyrifos and methamidophos. NeuroToxicology 21, 49-56. Pedata, E, Antonelli, T., Lambertini, L., Beani, L., and Pepeu, G. (1983). Effect of adenosine, adenosine triphosphate, adenosine deaminase, dipyridamole and aminophylline on acetylcholine release from electrically-stimulated brain slices. Neuropharmacology 22, 609-614. Peraica, M., Capodicasa, E., Moretto, A., and Lotti, M. (1993). Organophosphate polyneuropathy in chicks. Biochem. Pharmacol. 45, 131-135. Pond, A. L., Chambers, H. w., and Chambers, J. E. (1995). Organophosphate detoxication potential of various rat tissues via A-esterase and aliesterase activities. Toxicol. Lett. 78, 245-252. Pope, C. N. (1999). Organophosphorus pesticides: do they all have the same mechanism of toxicity? 1. Toxico!. Environ. Health B. Crit. Rev. 2, 161181. Pope, C. N., Chakraborti, T. K., Chapman, M. L., Farrar, J. D., and Arthun, D. (1991). Comparison of in vivo cholinesterase inhibition in neonatal and adult rats by three organophosphorothioate insecticides. Toxicology 68, 5161. Pope, C. N., Chapman, M. L., Tanaka, D. J., and Padilla, S. (1992). Phenylmethylsulfonyl fluoride alters sensitivity to organophosphorus-induced delayed neurotoxicity in developing animals. NeuroToxicology. 13,355-364. Pope, C. N., and Liu, J. (1997). Age-related differences in sensitivity to organophosphorus pesticides. Environ. Toxicol. Pharmacol. 4, 309-314. Pope, C. N., and Padilla, S. (1990). Potentiation of organophosphorus-induced delayed neurotoxicity by phenylmethylsulfonyl fluoride. 1. Toxico!. Environ. Health 31, 261-273. Pope, C. N., Tanaka, D., Jr., and Padilla, S. (1993). The role of neurotoxic esterase (NTE) in the prevention and potentiation of organophosphorusinduced delayed neurotoxicity (OPIDN). Chem. BioI. Interact. 87, 395406. Poul, J. M. (1988). Effects of perinatal ivermectin exposure on behavioral development of rats. Neurotoxico!' Teratol. 10,267-272. Rivera, S., Rosa, R., Martinez, E., Sunol, C., Serrano, M. T., Vendrell, M., Rodriguez-Farre, E., and Sanfeliu, C. (1998). Behavioral and monoaminergic changes after lindane exposure in developing rats. Neurotoxicol. Teratol. 20, 155-160. Rivera, S., Sanfeliu, C., and Rodriguez-Farre, E. (1990). Behavioral changes induced in developing rats by an early postnatal exposure to lindane. Neurotoxicol. Teratol. 12,591-595. Robinson, C. P., and Smith, P. W. (1977). Lack of involvement of monoamine oxidase inhibition in the lethality of acute poisoning by chlordimeform. 1. Toxico!. Environ. Health 3, 565-568. Romijn, H. J., Hofman, M. A., and Gramsbergen, A. (1991). At what age is the developing cerebral cortex of the rat comparable to that of the full-term newborn human baby? Early Hum. Dev. 26,61-67. Saegusa, J. (1987). Age-related susceptibility to dibromochloropropane. Toxico!. Lett. 36, 45-50.

References

Samanta, L., and Chainy, G. B. (1997). Age-related differences of hexachlorocyclohexane effect on hepatic oxidative stress parameters of chicks. Indian J. Exp. Bio!. 35, 457-461. Schildkraut, J. M., Demark-Wahnefried, W., DeVoto, E., Hughes, c., Laseter, J. L., and Newman, B. (1999). Environmental contaminants and body fat distribution. Cancer Epidemiol. Biomarkers. Prey. 8, 179-183. Serrano, M. T., Vendrell, M., Rivera, S., Serratosa, J., and Rodriguez-Farre, E. (1990). Effect oflindane on the myelination process in the rat. Neurotoxicol. Teratol. 12,577-583. Shah, P. v., Fisher, H. L., Sumler, M. R., Monroe, R. J., Chernoff, N., and Hall, L. L. (1987). Comparison of the penetration of 14 pesticides through the skin of young and adult rats. J. Toxicol. Environ. Health 21, 353-366. Sheets, L. P., Doherty, J. D., Law, M. w., Reiter, L. w., and Crofton, K. M. (1994). Age-dependent differences in the susceptibility of rats to deltamethrin. Toxicol. Appl. Pharmacol. 126, 186--190. Skopets, B., Wilson, R. P., Griffith, J. w., and Lang, C. M. (1996). Ivermectin toxicity in young mice. Lab. Anim. Sci. 46, 111-112. S1otkin, T. A. (1999). Developmental cholinotoxicants: Nicotine and chlorpyrifos. Environ. Health Perspect. 107 (Suppll), 71-80. Small, D. H., Reed, G., Whitefield, B., and Nurcombe, V. (1995). Cholinergic regulation of neurite outgrowth from isolated chick synpathetic neurons in culture. J. Neurosci. 15, 144-151. Smialowicz, R. J., Riddle, M. M., Rogers, R. R., Luebke, R. w., and Copeland, C. B. (1989). Immunotoxicity of tributyltin oxide in rats exposed as adults or preweanlings. Toxicology 57, 97-111. Song, X., Seidler, E J., Saleh, J. L., Zhang, J., Padilla, S., and Slotkin, T. A. (1997). Cellular mechanisms for developmental toxicity of chlorpyrifos: Targeting the adenylyl cyclase signaling cascade. Toxicol. Appl. Pharmacol. 145,158-174. Song, X., Violin, J. D., Seidler, E J., and Slotkin, T. A. (1998). Modeling the developmental neurotoxicity of chlorpyrifos in vitro: Macromolecule synthesis in PCl2 cells. Toxicol. Appl. Pharmaco!' 151, 182-191. Soni, M. G., and Mehendale, H. M. (1998). Role of tissue repair in toxicologic interactions among hepatotoxic organics. Environ. Health Perspect. 106 (SuppI6), 1307-1317. Takahashi, R. N., Poli, A., Morato, G. S., Lima, T. C., and Zanin, M. (1991). Effects of age on behavioral and physiological responses to carbaryl in rats. Neurotoxicol. Terato!' 13,21-26. Talts, U., Fredriksson, A., and Eriksson, P. (1998a). Changes in behavior and muscarinic receptor density after neonatal and adult exposure to bioallethrin. Neurobiol. Aging 19, 545-552.

885

Talts, U., Talts, J. E, and Eriksson, P. (1998b). Differential expression of muscarinic sUbtype mRNAs after exposure to neurotoxic pesticides. Neurobiol. Aging 19,553-559. Tenenbeim, M. (1987). Severe toxic reactions and death following the ingestion of diethyltoluamide-containing insect repellants. J. Amer. Med. Assoc. 258, 1509-1511. Tilson, H. A. (1998). Developmental neurotoxicology of endocrine disruptors and pesticides: Identification of information gaps and research needs. Environ. Health Perspect. 106 (SuppI3), 807-811. Tilson, H. A., and Mactutus, C. E (1982). Chlordecone neurotoxicity: A brief overview. NeuroToxicology 3,1-8. Tilson, H. A., Squibb, R. E., and Burne, T. A. (1982). Neurobehavioral effects following a single dose of chlordecone (Kepone) administered neonatally to rats. NeuroToxicology 3, 45-57. Veronesi, B., Jones, K., and Pope, C. N. (1990). The neurotoxicity of subchronic acetylcholinesterase (AChE) inhibition in rat hippocampus. Toxicol. Appl. Pharmacol. 104, 440-456. Vickroy, T. W., and Cadman, E. D. (1989). Dissociation between muscarinic receptor-mediated inhibition of adenylate cyclase and autoreceptor inhibition of [3H] acetylcholine release in rat hippocampus. J. Pharmacol. Exp. Ther. 251, 1039-1044. Watkinson, W. P. (1985). Effects of chlordimeform on cardiovascular functional parameters: Part 1. Lethality and arrhythmogenicity in the geriatric rat. J. Toxicol. Environ. Health 15, 729-744. Watkinson, W. P. (1986). Effects of chlordimeform on cardiovascular functional parameters: Part 3. Comparison of different routes of administration in the postweanling rat. 1. Toxicol. Environ. Health 19,207-217. Weiler, M. H. (1989). Muscarinic modulation of endogenous acetylcholine release in rat neostriatal slices. J. Pharmacol. Exp. Ther. 250, 617--623. Wennberg, R. P. (1993). Animal models of bilirubin encephalopathy. Adv. Vet. Sci. Comp. Med. 37, 87-113. Whitney, K. D., Seidler, E J., and Slotkin, T. A. (1995). Developmental neurotoxicity of chlorpyrifos: cellular mechanisms. Toxicol. Appl. Pharmacol. 134, 53--62. Wilkinson, C. E, and Ginevan, M. E. (1989). "A Critical Review of the Natural Resources Defense Council's Report 'Intolerable Risks: Pesticides in Our Children's Food'." RiskFocus, Versar, Inc., Springfield, VA. Wynne, H., Mutch, E., James, O. E, Rawlins, M. D., and Woodhouse, K. W. (1987). The effect of age on mono-oxygenase enzyme kinetics in rat liver microsomes. Age. Ageing 16, 153-158.

CHAPTER

42 Emerging Issues: Children's Exposure to Pesticides in Residential Settings John L. Adgate and Ken Sexton University of Minnesota School of Public Health

42.1 INTRODUCTION During the 1990s, there has been increasing concern about the potential effects of pesticides on children's health, much of it driven by mounting evidence from both animal toxicological studies and epidemiological investigations that children may suffer adverse health effects (e.g., neurobehavioral) from exposure to organophosphate (OP) pesticides. It is now widely recognized that health risk assessments should take special account of children because they may be both more exposed and more biologically susceptible than adults [Guzelian et aI., 1992; National Research Council (NRC), 1993; U.S. Environmental Protection Agency (U.S. EPA), 1995a]. Among the reasons children may be at potentially greater risk are their lower body weights, developing organs, higher metabolic rates, and unique behavior patterns. For example, the differences in body weight between children and adults are illustrated in Table 42.1, which summarizes average body weights of residents of the United States (U.S. EPA, 1999a). The importance of understanding children's exposure to pesticides was highlighted by the National Research Council in its 1993 report, Pesticides in the Diet of Infants and Children. The NRC recommended" ... that certain changes be made in current regulatory practice. Most importantly, estimates of expected total exposure to pesticide residues should reflect the unique characteristics of the diets of infants and children and should account also for all nondietary intake of pesticides" (NRC, 1993, p. 7). Three years later, Congress passed the Food Quality Protection Act (FQPA) of 1996 (P.L. 104-170, P.L. = Public Law), which requires that children's exposure to pesticides be evaluated for all potential pathways, both dietary (i.e., consumption of food and beverages) and nondietary (i.e., intake of pesticides in air, water, and soil or dust). The FQPA codified the need for more and better exposure data to help in the process of risk-based decision making, and mandated an examination of children's aggregate exposure, which means analysts must conduct assessments of nondietary pesticide exposure by mUltiple routes: that is, exposures by inhalation of airborne chemicals; Handbook of Pesticide Toxicology Volume 1. Principles

dermal absorption of chemicals in contact with the skin; and ingestion of chemicals in nonfoods such as soil and house dust. This chapter surveys existing and emerging methods for assessing children's nondietary exposures to pesticides in residential settings, and it comments on the implications of these methods for health risk assessment. We begin by describing basic principles of exposure assessment; then we examine the sources and pathways for children's exposure to pesticides. Methods for measuring pesticides are discussed next, followed by a discussion of biomarker measurements to estimate exposure and dose. The chapter concludes by looking ahead at the research needed to reduce uncertainties in children's risk assessments.

42.2 PRINCIPLES OF EXPOSURE ASSESSMENT Although the terms "exposure" and "dose" are well-established concepts familiar to all environmental health scientists, their meaning can vary depending on the nature of the discussion. For the purposes of exposure assessment, however, it is important that these and related terms be defined precisely in the context of an environmental health paradigm. 42.2.1 EXPOSURE AND DOSE Exposure is a deceptively simple concept, defined as contact at a body boundary between a person and an environmental stressor (biological, chemical, or physical) over time (Sexton et aI., 1995b; U.S. EPA, 1992). This simple definition masks the fact that a rigorous exposure analysis can be a complex endeavor, requiring collection and analysis of multiple variables, such as concentration and duration of exposure, as well as descriptive factors that influence contact rates and, therefore, determine the magnitude of exposure. A minimal description of exposure for a particular route (i.e., inhalation, ingestion, dermal absorption)

887


888

CHAPTER 42 Emerging Issues: Children's Exposure Table 42.1 Body Weight Values for Specified Age Groups in the United Statesa Body weight Age (years)

(kg)

Comments

Infants (0.5-1.5)

10

Mean of values for males and females in the 6-11

Toddlers (3)

15

Mean of values for male and female 3-year-olds

Children (6)

23

Mean of values for male and female 6-year-olds

Youth (10--12)

41

Mean of values for males and females age 10, 11,

Adult reproductive females

60

Mean for females age 13-54 years

Adults

72

Mean for males and females 18 years and older

month and I year age groups

and 12 years

aSource: Data from U.S. EPA (l999a).

must include at least two related attributes: concentration of the stressor in the carrier medium (exposure concentration); and the time of contact (duration). If the exposure concentration is integrated over the duration of contact (Eq. 1), the area under the resulting curve is the magnitude of the exposure in units of concentration times time (e.g., mg/l-day, mg/kg-day, Il-g/m3 -hr). This is the method of choice to describe and estimate short-term exposures, where integration times are on the order of minutes, hours, or days (NRC, 1991; Sexton et a!., 1995a; U.S. EPA, 1992): 12

E=

1

C(t)dt

(1)

11

where E is magnitude of exposure, t2 - t1 is duration of exposure, and C(t) is exposure concentration as a function of time. Over periods of months, years, or decades, exposures to most stressors, including pesticides, occur intermittently rather than continuously. Yet long-term health effects, such as cancer, are customarily evaluated based on average exposure or dose over the period of interest (typically years), rather than as a series of intermittent exposures. Consequently, long-term exposures or doses are usually estimated by summing across discrete exposure episodes and then calculating an average for the period of interest. Although the integration approach can also be used to estimate long-term exposures or doses, its application to time periods longer that about a week is usually difficult and inconvenient (Sexton et a!., 1995a). Duration of exposure is a key element of pesticide exposure assessment because it is directly related to adverse outcomes (U.S. EPA, 1997). Accordingly, pesticide exposures are typically divided into four general categories by duration: Acute exposures are less than a day in duration; short-term exposures last between 2 and 7 days; intermediate-term exposures last from one week to several months; and chronic exposures persist over a substantial portion of an individual's lifetime. Examples of pesticide uses that could result in acute or short-term exposures include treatments to turf in parks or applications indoors at schools; intermediate-term or chronic exposures might occur

because of agricultural pesticide residues in the food supply or continual use of pesticides inside the residence. Dose is a more complicated concept intimately related to exposure. Once a stressor, such as a chemical pesticide, enters the body it is described as a dose. Several different types of dose are relevant to exposure estimation. Potential (or administered) dose is the amount of a chemical that is actually ingested, inhaled, or in contact with the skin. Applied dose is the amount of a chemical directly in contact with the body's absorption barriers, such as the skin, respiratory tract, or gastrointestinal tract. Internal (or absorbed) dose is the amount of the chemical absorbed and, therefore, available to undergo metabolism, transport, storage, or elimination. Delivered dose, sometimes referred to as body burden, is the portion of the absorbed dose that reaches a tissue of interest, such as blood, urine, or hair. Biologically effective (or target) dose is that portion of the delivered dose that reaches a site of toxic action, such as the liver or brain (Sexton et a!., 1995a; U.S. EPA, 1992). All these dose types can be represented by a general equation (Eq. 2) that incorporates the intake and uptake factors that modify these distinct dose types (Lioy, 1990): D

=

10

= 10

1

D(t)dt 1

J(x)g(ab)p(as, rd, me, el)C(t) dt

(2)

where D is the integrated dose at a target tissue; D(t) is the time-varying function for dose; J (x) is the contact rate; g( ab) is the target organ or system specific bioavailability that determines absorption (ab); and p (as, rd, me, el) describes the nature of a contaminant's assimilation (as), cell repair or damage (rd), elimination (el), and metabolism (me).

42.2.2 THE DOMAIN OF EXPOSURE ASSESSMENT The chain of events depicted in Fig. 42.1 is an "environmental health paradigm"-a simplified representation of key steps be-

42.2 Principles of Exposure Assessment

Intake and Uptake Processes

Environmental Health Paradigm

889

Important Determinants (Mechanisms)

- eompensatlon -damage • repair

Figure 42.1 The domain of exposure assessment in relation to the environmental health paradigm. (Reproduced with permission Sexton et aI., 1995a.)

tween release of hazardous agents into the environment and potential morbidity or mortality in humans (Sexton et al., 1995b). This sequential series of steps serves as a useful construct to aid in understanding and evaluating environmental health risks. The figure also illustrates the domain of exposure assessment, which extends from identifying emission sources to characterizing the biologically effective dose, and includes developing an understanding of important events (e.g., contact between people and pollution), mechanisms (e.g., pharmacokinetics), and processes (e.g., intake, uptake). An important aspect of exposure assessment typically involves characterizing the critical "exposure pathway," which refers to the specific course of movement for a particular chemical from its source through various environmental media (i.e., air, water, soil) to ultimate contact with people. Common residential nondietary pathways for exposure to pesticides occur indoors (e.g., a consumer product applied to

a carpet results in child being exposed through hand-to-mouth activity-dose occurs via ingestion) and outdoors (e.g., application of consumer product to lawn or garden results in child being exposed through dermal contact-dose occurs by dermal absorption). It is important to remember that children can also be exposed to pesticides in many nonresidential settings, such as day care centers, schools, and playgrounds. As shown schematically in Fig. 42.1, exposure assessment is necessarily broad-based and complicated, involving acquisition and interpretation of information about key steps in the environmental health paradigm and related data on exposure factors, intake, uptake, and pharmacokinetics. Basically, assessing exposure to chemical pesticides involves the qualitative description and quantitative estimation of the chemical's contact with and entry into the body. Although no two exposure assessments are exactly alike, most address three areas that are critical for re-

890

CHAPTER 42

Emerging Issues: Children's Exposure

alistic risk assessments and contribute substantially to informed risk management decisions: (1) the number of people exposed at specific concentrations for the time period of interest; (2) the resulting dose, especially to the target tissue; and (3) the relative contributions of important sources and pathways to exposure or dose. In addition to these areas, a comprehensive exposure assessment should include an analysis of variability (e.g., within and between individuals) and uncertainty (e.g., statistical error in measurements and model parameters) (Sexton et aI., 1995a; U.S. EPA, 1992).

drawbacks of this approach are that it tends to be intrusive, resource intensive, and generally does not provide information on pathways and routes of exposure. This approach is also constrained by the lack of specific physiologically-based pharmacokinetic models for the pesticides of interest. Without a good understanding of the relevant pharmacokinetics, reconstruction of internal dose and calculation of previous exposures are highly uncertain (Sexton et aI., 1995a; U.S. EPA, 1992). These problems are even more acute for studies involving children.

42.2.3 EXPOSURE ASSESSMENT APPROACHES

42.2.3.3 Scenario-Based Assessment

In practice, exposure assessment for environmental chemicals involves use of both qualitative and quantitative data to describe contact with and entry into the human body. The quantitative estimation of chemical pesticide exposure can be approached in three ways: personal measurements, reconstructive analysis, or scenario-based assessment (Sexton et aI., 1995a; U.S. EPA, 1992). 42.2.3.1 Personal (Point-of-Contact) Measurement

Personal (or point-of-contact) measurements document exposures as they occur by measuring the pesticide concentration at the point of contact between the person and the environmental (or carrier) medium. Examples include use of pumps and filters to measure airborne concentrations near the breathing zone, duplicate diet food samples to measure dietary levels, or skin patch samples to measure dermal concentrations. The major strength of this approach is that it measures exposure directly during the monitoring period, which typically is on the order of minutes, hours, or, at the most, days. The problems with personal measurements are that they are costly and timeconsuming, can be burdensome for the study participants, and suitable monitoring devices are not available for all pesticides and pathways of interest (Sexton etal., 1995a; U.S. EPA, 1992). Because these problems are exacerbated in the case of children, personal monitoring has rarely been attempted in this subpopulation (Weaver et aI., 1998). 42.2.3.2 Reconstructive Analysis

Reconstructive exposure analysis uses measurements of dose (i.e., body burden, elimination levels), in conjunction with information or assumptions about rates of intake and uptake, to derive (or reconstruct) estimates of past pesticide exposure. Use of this approach requires valid measures of exposure biomarkers in accessible human tissues so that internal dose can be realistically reconstructed, and adequate information to estimate rates of intake, uptake, and metabolism accurately. As discussed in more detail in the section on biomonitoring, urine levels are the biomarker of choice for assessing children's pesticide exposures, largely because of the relative ease of collection. The strength of this approach is its capacity to demonstrate unequivocally that exposure and uptake have occurred. The primary

In principle, personal (point-of-contact) measurements and reconstructive analysis are complementary methods, which, because they are based on direct measurements in the exposed population, are the preferred approaches for pesticide exposure assessment. In practice, however, they are rarely used because necessary data are not available for the situations and populations of interest. Consequently, the most common approach is scenario-based exposure assessment, which requires the analyst to use available facts (e.g., environmental measurements, databases), in combination with inferences and professional judgment, to construct a plausible set of assumptions (i.e., a scenario) that describes quantitatively how contact occurs between people and chemical pesticides (Sexton et aI., 1995a). A typical scenario-based approach estimates pesticide exposure by merging two separate but essential components of exposure: (1) concentration of the chemical in the environmental (carrier) medium, estimated by using monitoring data or making assumptions about source-path way-exposure interactions; and (2) people's contact time with the carrier medium, estimated by using existing demographic, geographic, and timeactivity data or by making reasonable assumptions about activity patterns, lifestyle characteristics, residential proximity to sources, and other factors. The related doses are estimated by using knowledge and assumptions about relevant pharmacokinetic processes. Variations of the scenario-based approach include both (a) "microenvironmental" methods, which combine measurements in important microenvironments (e.g., inside the residence, outdoors in the community) with data on timeactivity patterns and (b) pathway--exposure factor ("PEF") methods, which combine measurements in important environmental media (e.g., air, water, food, soil, dust) with "off-theshelf" exposure factors (U.S. EPA, 1999a) (e.g., volume of air breathed or water consumed per day, body weight, skin surface area). Examples of different types of data used in constructing exposure scenarios are listed in Table 42.2. The primary advantage of scenario-based approaches is that they enable assessors to estimate pesticide exposure and dose in cases where data are limited or lacking (a common occurrence). On the other hand, the uncertainty introduced by the need to make assumptions and inferences in the face of inadequate or inappropriate information is also their major disadvantage. Scenario-based assessments typically do not include a complete description of the exposure and dose distribution for

42.2 Principles of Exposure Assessment

891

Table 42.2 Examples of Types of Measurements Useful in Construction of Scenarios for Pesticide Exposure Assessmentsa Additional information needed to Type of measurement

Element estimated

Examples

characterize exposure

Fixed location

Environmental medium

Air and water quality

Population location and activities

monitoring

samples used to establish

networks

relative to monitoring location; fate of pesticides over distance between

long-term media levels

monitor and variation in concentration

and trends

at point of exposure Short-term media monitoring

Environmental or ambient medium; samples used to

Special studies of spray drift, indoor air

of pesticides between measurement

establish a snapshot of

point and point of exposure; time

quality of medium over

variation of pesticide concentration at point of exposure

relatively short time Food samples

Population location and activities; fate

Concentrations of

U.S. Food and Drug

Dietary habits of various age, sex, or

contaminants in food

Administration Total

cultural groups; relationship between

supply

Diet Study, market

food items sampled and groups

basket studies, shelf

(geographic, ethnic, demographic)

studies, cooked-food

studied; relationships between

sampling

concentrations in uncooked versus prepared foods

Drinking water samples

Concentrations of

Ground Water Supply

Fate and distribution of pesticides

pesticides in drinking

Survey, Community

from point of sample to point of

water supply

Water Supply Survey, tap

consumption; population served

water samples

by specific facilities and consumption rates; for exposure due to other uses (e.g., cooking, showering); need to know activity patterns and volatilization and degradation rates

Consumer product samples

Concentration levels of the products

Shelf surveys, e.g.,

Establish use patterns and market

solvent concentration in

share of particular products,

household cleaners

individual exposure at various usage levels, extent of passive exposure

Breathing zone measurements

Exposure to airborne

Indoor air studies

Location, activities, and time spent relative to monitoring locations;

chemicals

protective measures or avoidance Microenvironmental measurements

Ambient medium in a defined area, e.g., kitchen

Special studies of indoor air, house dust, and contaminated surfaces

Surface soil samples

Degree of contamination of soil available for contact

Soil samples at contaminated sites

Activities of study populations relative to monitoring locations and time exposed Fate of pesticide on or in soil; bioavailability, contact, and ingestion rates as a function of activity patterns and age

Fish tissue samples

Extent of contamination of edible fish tissue

National Shellfish Survey

Relationship of samples to food supply of individuals or population of interest; consumption habits; preparation habits

aBased on U.S. EPA (1992).

892

CHAPTER 42


the population of interest, instead providing only "point estimates" of specific locations on the population distribution of exposures. Historically, the points of interest on the exposure distribution have been the median or average exposures, the "high-end" exposure (the 90th percentile and above is defined as the high end of the exposure distribution), and the most exposed individual in the population. Although the scenario-based approach is used in the vast majority of pesticide exposure assessments, it is most useful when the analyst has some insight into the completeness, soundness, and validity associated with underlying assumptions and inferences, and understands their overall effect on the uncertainty of estimated values for exposure and dose. Yet despite their obvious limitations, scenariobased approaches remain the only viable method for estimating pesticide exposure and dose in the absence of direct measurements. Scenario-based approaches are specified in the Environmental Protection Agency's standard operating procedures (SOPs) for pesticide exposure assessment in residential settings, which provide standard default methods for assessment of both handler (i.e., the individual applying the pesticide) and postapplication exposures when chemical or site-specific data are limited (U.S. EPA, 1997). These SOPs are for "high-end" exposures so that the residential lawn scenario, for example, is assumed to represent the upper end of the distribution of exposures that could occur from lawns, parks, playgrounds, recreational areas, athletic fields, and other turf areas. These scenarios typically rely on one or more upper-percentile assumptions, such as the 90th percentile value for exposure duration and the 90th percentile value for skin surface area. They are intended to represent "Tier 1" assessments that can be used to indicate whether a more detailed assessment is warranted, possibly including collection of chemical-specific or site-specific data. A Tier 1 assessment uses a "conservative" screening scenario to make "worst-case" or "bounding" estimates (e.g., maximum pesticide application to 100% of turf) (U.S. EPA, 1995b, 1997). If this analysis suggests a possible problem, estimates can then be refined in subsequent tiers using progressively more realistic assumptions and values (e.g., average pesticide residues on percentage of turf actually treated). Typically, Tier 1, scenario-based exposure assessments combine information and assumptions about environmental medium concentrations, contact rate, and exposure duration using the general equation E

= ex CR x ED

(3)

where E = exposure to the chemical pesticide (mass per unit time, e.g., J.lg/day); C = concentration in environmental media (mass of substance per unit volume, e.g., J.lg/m3); CR = contact rate (volume per unit time, e.g., m3/day); and ED = exposure duration (fraction of a day exposed, e.g., minutes exposed per total minutes in a day). Better and more complete data are often available for inhalation exposures and these assessments frequently involve relatively straightforward assumptions about inhalation rates and

absorption of the target chemicals. Adequate and appropriate data are available less often for both ingestion and dermal exposures, which are also intrinsically more complicated because of difficulties quantifying relevant behaviors, activities, and physical processes that determine actual exposures. Scenario-based approaches also typically go on to calculate dose using the general equation D

=

E x AF/BW

(4)

where D = dose (in micro grams per kilogram of body weight per day, J.lg/kg-bw-day); E = exposure to a chemical pesticide (mass per unit time, e.g., J.lg/day); AF = absorption factor (unitless); BW = body weight (kg). Because population-specific and situation-specific data are rarely available to estimate parameters in the simplified exposure and dose equations, values are typically obtained by using EPA-approved "exposure factors," which are derived from an amalgamation of sources (U.S. EPA, 1999a).

42.3 SOURCES AND PATHWAYS FOR CHILDREN'S RESIDENTIAL EXPOSURE TO PESTICIDES The most repeated yet underappreciated precept in children's environmental health is the factual statement that "children are not little adults," which means that it is not appropriate to estimate exposure and dose in children by merely scaling down data from adults. Postnatally, there are at least three important differences between children and adults that are crucial for assessing exposure: factors relating to (a) sources, (b) intake, and (c) uptake (Plunkett et aI., 1992). Children tend to eat less varied and different diets than adults, they may consume food that has come in contact with the floor and other contaminated surfaces, and they are more likely to spend time on or near treated surfaces. Children have higher intake rates of air, water, soil, and food per unit of body weight and surface area than do adults. Moreover, differences in uptake and pharmacokinetics can result in children receiving a higher proportionate dose for a given exposure than adults. Although it is common to speak of children as a homogeneous group, there are important differences in exposure-related attributes by age. For example, dietary and nondietary exposures are likely to be substantially different for a 2-year-old toddler, a 6-year-old elementary school student, and a 14-yearold adolescent, even if they live in the same household. In fact, experts have identified at least seven distinct and sometimes overlapping developmental periods in the life of a child that are likely to have important effects on exposure-related attributes. These stages and their duration, important developmental milestones, and related behaviors, activities, and potential exposures are summarized in Table 42.3 (Carmichael, 1946; Freeman, 1999; Kelly, 1983; Larsen and Pascal, 1998; NRC, 1993). Although this chapter focuses primarily on exposures outside the womb, exposures prior to birth can be an important

42.3 Sources and Pathways for Children's Residential Exposure to Pesticides

893

Table 42.3 Developmental Stages in Children and Associated Milestones and Exposure-Related Behaviorsa Important developmental milestones and

Examples of potential exposures and related

stage

Time period

average age of occurrence

behaviors and activities

Embryonic

8 days to 8

Human organogenesis approximately days

Mother's acute or short-term exposures to current

Developmental

weeks of

use pesticides or release of persistent pesticide stores

20-60 of gestation

pregnancy Fetal

8 weeks of pregnancy to

Control of autonomic nervous system at 24 weeks

current use pesticides or release of persistent

birthb

Perinatal

29 weeks of pregnancy to 7

pesticides Preterm viability outside the womb 24-38

End transplacental exposure; begin breast-feeding transfer of persistent compounds and recent maternal

weeks

days after birth Neonatal

Birth-28 days

exposure to compounds with short half-lives Beginning of ocular control and head

Continued breast-feeding

motion Infancy

Birth-I 2 months

Transplacental transfer from mother's exposures to

Begin potential inhalation and dermal exposures

Rolling over at 2-3 months

Breast-feeding usually completed by end of 1st year

Sitting starts at 3 months

Transition to solid food begins at 6-9 months: low

Maturation of xenobiotic metabolism and

diet diversity leads to higher relative consumption of

elimination processes at 3 months

some foods compared to adults and older children

Standing supported at 6 months

Mouthing nonfood objects begins as early as 4

Crawling usually mastered by 9 months

months and continues to at least age 2 Dermal and inhalation exposures from sitting on

Walking begins at 10-17 months Weight triples and height increases to

~20

cm

treated surfaces

during first year Childhood

I year-12 years

Receptive to adult language and self-feeding at

~1

year

Bladder and bowel control at

Hand to mouth behavior common from ages 1-3 year olds

~2

years

Mastery of common motor skills by 5 years

Periodic consumption of nonfood objects (pica) exhibited in half of 1- to 3-year-olds Increasing time spent outside of the home

Adolescence

> 12-18 years

Maturation of organ systems to adult size and weight

Potential occupational exposures in young farm workers

aCompiled from Carmichael (1946), Freeman (1999), Guzelian et at. (1992), Kelly (1983), Larsen and Pascal (1998) and NRC (1993). bNormal term birth is at 40 ± 2 weeks of pregnancy.

health issue for which only limited exposure data exist. Typically, embryonic and fetal in utero exposures are estimated based on the mother's body burden for a particular chemical. Many children are exposed in utero to low levels of persistent pesticides (i.e., chlorinated organic compounds such as DDT and its metabolites). Although many of these pesticides have been banned in the United States for many years, they are still present in the food web due to both past and current use worldwide and their tendency to bioaccumulate as a result of their persistence in the environment. Recent studies using highly sensitive measurement techniques indicate that many of these compounds are present in adipose tissue (Kutz et aI., 1992) of U.S. residents and in the breast milk of women from both developed and less developed nations (Banerjee et aI., 1997; BasriUstunbas et aI., 1994; Dogheim et aI., 1996; Furst et aI., 1994; lohansen et aI., 1994; Klopov et aI., 1998; Kroger, 1972; Quinsey et aI., 1995; Saleh et aI., 1996; Schade and Heinzow, 1998). Residue levels measured in breast milk have been steadily declining in the United States and no studies have demonstrated adverse effects in breast-fed children. Current scientific consen-

sus supports the conclusion that the benefits of breast-feeding outweigh the risks from exposure to persistent organic pesticides in human milk (NRC, 1993). Beginning with the neonatal stage and progressing through infancy (birth to 12 months), children are influenced by a special set of pesticide sources and pathways as they experience rapid increases in mobility and important changes in behaviors and activities. For example, activities such as crawling, rolling, climbing, and finally walking during the first year or so of life have profound effects on young children's contact with surfaces, which increases the potential for dermal absorption of pesticide residues. Infants often put their hand and other objects in their mouth, resulting in non dietary ingestion from contaminated surfaces. Teething occurs largely during the first year of life and coincides with very high levels of hand-to mouth and object-to-mouth activities (Freeman, 1999). From an exposure perspective, children between ages 1 and 12 can be subdivided into four age groups [young toddlers (12-24 months), older toddlers (24-36 months), preschoolers (3-5 years), and

894

CHAPTER 42


school-aged children (>5 years)], based on behavior and activity patterns likely to influence pesticide exposures. 42.3.1 SOURCES OF NONDIETARY RESIDENTIAL PESTICIDE EXPOSURES FOR CHILDREN

For children less than 12 years old, their residence is likely to be the primary setting in which they come into contact with pesticides. Children spend an average of 21 hours a day indoors (D.S. EPA, 1999a), most of it at home, in school or in daycare. A wide variety of consumer pesticide products are used in the home, including preparations for flies, ants, roaches, and fleas, as well as products for care of pets and houseplants (Nigg et aI., 1990). Pesticide application inside residences falls generally into four categories: (1) broadcast (i.e., pressurized fan spray with dilute pesticide formulation on open surfaces), (2) crack and crevice (i.e., application method similar to broadcast but applied to these enclosed locations), (3) total aerosol releases (e.g., foggers), and (4) structural applications (e.g., treatments of walls and foundations for termites) (Fenske et aI., 1991). In addition, many cleaners and disinfectants, room deodorizers, and laundry aids also contain pesticides, though users are seldom aware of this. Typical active ingredients found in indoor application products include chlorpyrifos, propoxur, diazinon, malathion, dichlorvos, and pyrethroids (Ness, 1994). Other pesticide products are used outdoors around the residence. For example, it is common for building occupants to use insecticides on lawns and gardens, fungicides for structural protection and lawn treatments, and herbicides for weed control. Prevalent active ingredients in these products are the insecticides carbaryl, chlorpyrifos, diazinon, and malathion, fungicides captan and chlorothalonil, and herbicides such as 2,4-D and glyphosate (Ness, 1994). Probability-based sampling studies of the prevalence of pesticide storage for the D.S. population indicate that >90% of households have at least one pesticidecontaining product. In 1988, a national in-home survey found an average of 3.8 pesticide products (95th percent confidence interval 3.3-4.3) per home (Whitmore et aI., 1992). A recent probability-based survey conducted in more than 300 Minnesota households with children, representing more than 49,000 households in the census tracks sampled, reported a mean of 6.0 pesticide products stored, and a mean of 3.1 products used in the previous year (Adgate et al., 2000). Exposure and dose may be influenced by the formulation of the pesticide itself. In quantifying exposure to these compounds it is important, therefore, that the physical, chemical, and biological properties of the specific compound be examined. For example, pesticides are made and distributed in numerous formulations, some of which are designed to improve their efficacy or enhance storage or safety (Fenske, 1997; Ware, 1994). Common formulations are aerosols, dusts, or granular materials, which have varying persistence in the environment. Insecticides or fungicides applied indoors or tracked in from the outdoors on shoes or by pets can become embedded in carpets

and other surfaces indoors. Although data are limited, it is possible that pesticides in these indoor locations are protected from natural processes that might otherwise degrade them (e.g., sunlight, extremes of temperature, rainfall, and microbial action) and, therefore, may represent a long-term potential source of exposure. 42.3.2 NONDIETARY RESIDENTIAL EXPOSURE PATHWAYS FOR PESTICIDES

The sUbpopulations of children thought to be at greatest risk from residential exposures are children whose families use pesticides inside or outside the home, and those whose parents or siblings work with pesticides. Children of applicators and agricultural workers are at risk for higher exposures because of their contact with work-contaminated clothing and the increased likelihood of pesticides being tracked indoors from work applications (Fenske, 1997). Several studies have shown that airborne concentrations of pesticides outside residences are typically an order of magnitude lower than indoor levels (Nigg et aI., 1990). Investigations of children's nondietary pesticide exposures in the residential setting have focused principally on two pathways: potential indoor exposures occurring upon reentry after a pesticide has been applied ("reentry exposures"), and potential outdoor exposures as a result of garden and turf treatments. An intrinsic problem in these kinds of exposure assessments is determining the amount of pesticide product that is "dislodgeable," which is defined as the amount that can be removed from lawn and garden foliage or indoor surfaces and is therefore available for subsequent ingestion or skin absorption. Examples of possible residential exposure scenarios for children are shown in Table 42.4. Although the table is not exhaustive, it does include the residential exposure scenarios deemed most likely to cause elevated pesticide exposures in children.

42.4 PESTICIDE MEASUREMENT

METHODS RELEVANT TO SCENARIO-BASED ASSESSMENTS The most commonly used environmental sampling methods relevant to assessing exposures in children are air monitoring, measurements of house dust, and measures of dislodgeab1e residues on both indoor and outdoor surfaces. Personal (e.g., breathing zone) air samples are rarely collected for young children (1-12 years old) because of their inability to comply with extensive study protocols and the relatively large size of the monitoring equipment. Instead, air samples have been collected in important microenvironments (e.g., main activity room of the residence) using area monitors. It has been demonstrated, for example, that airborne chlorpyrifos levels after broadcast application are higher near the floor, which is in much closer proximity to a toddler's breathing zone than to an adult's (Fenske et aI., 1990).

42.4 Pesticide Measurement Methods Relevant to Scenario-Based Assessments

895

Table 42.4 Examples of Nondietary Residential Exposure Scenarios and Important Routes of Exposure for Children Exposure routes Exposure scenarios

Dermal

Investigation

Inhalation

References

x

x

x

Byme et al. (1998), Fenske et al. (1990);

x

x

Post application contact indoors Residues from crack and crevice, broadcast, or

Gurunathan et al. (1998); Ross et al. (1991)

aerosol treatments on surfaces House dust contaminated from indoor or

Bradman et al. (1997); Loewenherz et at. (1997);

structural treatments, track-in, or spray drift

Nishioka et at. (1996); Simcox et at. (1995);

from outdoor sources

Whitmore et at. (1994)

Transfer from pet treatments or flea collars

x

Insect repellents applied to clothes or skin

x

Mouthing or consumption of impregnated

Ames et at. (1989); Chambers (1996)

x

Lipscomb et at. (1992); Osimitz and Murphy (1997)

x

V.S. EPA (1997)

x

Nishioka et at. (1996); V.S. EPA (1997)

materials, such as antimicrobial shower curtains Postapplication contact outdoors Pesticide product or residues on treated turf

x

or soils Fogging or spraying for mosquitoes and other

x

Moore et at. (1993)

x

Bradman et at. (1997); Bradman et al. (1994);

pests near homes Spray drift from large-scale treatments near homes or schools

Because the behaviors and activities of young children are more likely to bring them into contact with contaminated surfaces, it is important to understand the levels of pesticides in these locations. Available methods for sampling indoor surfaces include (1) deposition pad samples, typically used on any indoor surface, (2) wipe sampling techniques, used on relatively smooth surfaces such as floors, counter tops, and window sills, and (3) vacuum techniques, which have been used to collect house dust samples from both hard floor surfaces and carpets. Deposition sampling is performed postapplication using aluminum foil (Fenske et aI., 1991), gauze (Ross et aI., 1991), or cotton cloth pads (Byrne et aI., 1998; Krieger et aI., 1997) as the collection medium. Wipe sampling techniques have been used to collect chlorpyrifos samples from deposition samplers using water as a solvent (Fenske et aI., 1991) and from surfaces using octadecylbonded silica disks and methanol and hexane as solvents (Gurunathan et aI., 1998). Choice of method used to remove pesticide residues from surfaces can have a significant effect on estimated exposures. For example, investigators (Byrne et aI., 1998) using deposition pads wiped on surfaces and toys subsequently extracted with isooctane report either nondetectable or lower chlorpyrifos exposures compared to investigators (Gurunathan et aI., 1998) using hexane to remove pesticides directly from surfaces and toys. Use of organic solvents directly on a surface apparently results in more complete removal of chlorpyrifos residues, but may overestimate doses obtained in dermal contact or hand-to-mouth activities, where the solvent would be saliva, sweat, or the sebum layer on the skin.

Loewenherz et at. (1997); Marty et at. (1994)

A number of vacuum sampling systems have been developed to collect house dust samples from carpets, rugs, and bare floors, mostly for sampling lead (Rinehart and Rogers, 1995). A specialized high-volume vacuum sampler (HVS3) was developed specifically to obtain samples of semivolatile pesticides in house dust [American Society of Testing and Materials (ASTM), 1993; Roberts et al. (1991)]. The HVS3 has been used in several field studies to collect carpet and smooth-surface samples (Bradman et aI., 1997; Lewis et al., 1994; Nishioka et aI., 1996; Simcox et aI., 1995). In a recent field test comparing this method with samples from 15 home vacuum cleaner bags, the HVS3 obtained higher upper bound concentrations for up to 26 pesticides, but prevalence of measurable pesticides and median dust concentrations were similar for both methods (Colt et aI., 1998). Three techniques have been developed to measure dislodgeable residues on indoor and outdoor surfaces and to characterize transfer from one location to another. The California Department of Food and Agriculture (CD FA) developed a 12-kg roller made of polyvinyl chloride (PVC) pipe covered with polyurethane foam (PUF), which is used to roll over cotton sheets placed on a treated surface (Ross et al., 1991). Results of this method were compared to residue levels on the clothes of adult subjects wearing dosimeter clothing while performing a set of standardized aerobic dance routines (Jazzercise) on treated surfaces (Ross et aI., 1990). The upper bound estimate of total dislodgeable residues, 1-3% of residues present on surfaces, was comparable between the CDFA roller and lazzercise methods. A second PUF roller method has been developed to simulate the force a crawling 9-kg child applies to a

896

CHAPTER 42


surface (Lewis et aI., 1994). Using this method, investigators estimated 2,4-D and dicamba track-in rates onto carpets after outdoor turf applications (Nishioka et aI., 1996). Transfer was estimated to be 3% of dislodgeable residues, which were J.1-0.2% of overall turf application levels. A subsequent study by these same investigators in 13 homes after homeowner and commercial applications indicated that children and pet activities were the most significant factors determining residue levels indoors (Nishioka et aI., 1999). A "drag sled" method has also been developed that uses a 100-cm2 patch of denim affixed to the bottom of a sledlike device, whose weight approximates the force exerted by a 1O-kg child on a surface (Byrne et aI., 1998; Vaccaro et aI., 1996). In theory, these three methods should give similar results, but no studies have directly compared these methods. Dermal absorption is a major occupational route of pesticide exposure in adults and a variety of direct and indirect methods have been developed using various dosimeters and indirect estimation methods to measure pesticide loading on the skin. Surrogate skin (Fenske, 1997) and fluorescent tracer methods (Fenske, 1988, 1990; Fenske et aI., 1986) have been used to monitor occupational exposures in adults, but their use in additional studies on children has been limited because of concerns about intentional exposure to subjects and the introduction of fluorescent tracer into the pesticide formulation prior to application is required. Hand wipe methods have been developed that use either isopropanol and gauze wipes (Geno et aI., 1996; Lewis et aI., 1994; Lu and Fenske, 1999) or a 10% isopropanoldistilled water mixture wrapped around a subject's hand using a plastic bag (Edwards and Lioy, 1999; Fenske and Lu, 1994) to remove pesticides. Although some data suggest that hand wipes may remove deeply embedded compounds that may not be removable by typical soap-and-water washing, data from controlled mass-balance experiments also suggests that dermal wash methods may significantly underestimate exposure because they typically remove approximately 20-40% of the available compound, with the remaining amounts likely absorbed through the skin (Fenske and Lu, 1994). Only one study has compared pesticide residue levels measured by wipe, roller, and hand wipe or press methods. Indoor chlorpyrifos levels after broadcast application indicate that wipe and PUF roller measurements estimate a dermal loading that is 23-36 times greater than estimates based on hand press or drag samples (Lu and Fenske, 1999). Overall, the relation between residue levels measured by these methods and the actual exposures experienced by children remain unclear, although evidence suggests the methods may provide a reasonable upper bound estimate of surface transferable residues. Information about activity patterns is an important component of scenario-based assessments that is combined with data about environmental concentrations to estimate exposure. Several small field studies have collected time-activity data for children (Lewis et aI., 1994), and most existing time-activity databases are summarized in the U.S. EPA Exposure Factors Handbook (U.S. EPA, 1999a). This important reference describes how older children (e.g., 9-11 year olds) spend their

time (Schwab et aI., 1991), but has little information about the frequency of certain important exposure-related activities for younger age groups, such as toy mouthing, consumption of nonfood items, or dermal contact rates. The National Human Activity Pattern Survey (NHAPS) is the largest probabilitybased survey ever conducted in the contiguous United States (n = 9386), but relatively few children (approximately 500) less than 4 years old were surveyed (Nelson et aI., 1994). Although NHAPS does not provide information on variability in activities by age or season of the year, it contains data important for assessing upper bound exposures associated with time spent in specific locations. These data can be useful for estimating inhalation exposures, but provide little insight for developing estimates of dermal exposure or nondietary ingestion by handto-mouth behaviors. Recently, videotaping has been used to provide a more detailed examination of children's activity patterns, including quantifying specific behaviors, such as hand-to-mouth rates, using video translation software (Zartarian et aI., 1995, 1997a, b) or trained video observers (Reed et aI., 1999). Both methods derive similar results for hand-to-mouth rates in preschool children over approximately a 1O-hr period. Combining this information with environmental media concentrations, skin surface concentrations, and data on contact and removal rates will improve exposure estimates for the hand-to-mouth ingestion pathway. Videotaping methods may eventually assist in developing more rigorous estimates for both nondietary ingestion of pesticides and dermal contact rates.

42.5 USING BIOMARKER MEASUREMENTS TO ESTIMATE PESTICIDE EXPOSURE AND DOSE Many of the dose-related steps in the environmental health paradigm (Fig. 42.1) occur at inaccessible sites in the body (e.g., liver, brain). Biological markers (biomarkers) are indicators of these significant but inaccessible events that can be measured in accessible human tissues (e.g., blood) and excreta (e.g., urine), using either invasive or noninvasive (sample collection does not require penetration of the body) methods. For example, invasive methods may involve collection of blood, lung tissue, bone marrow, amniotic fluid, liver tissue, or adipose tissue, whereas noninvasive techniques might involve collection of expired air, saliva, semen, urine, hair, feces, breast milk, skin, or fingernails. A particular biomarker may be an indicator of exposure (e.g., pesticide metabolites in urine), effect (e.g., DNA hyperploidy caused by benzidine exposure), or susceptibility (e.g., immunoglobulin levels in blood) (Sexton et aI., 1995a). For pesticide exposures, biomarker measurements can serve as exposure and dose indicators and, when linked with physiologically based pharmacokinetic models, can assist in estimating past exposure and related doses (Cashman et aI., 1996; Dong et aI., 1994, 1996). Biomarker measurements can also provide a "reality check" to verify the realism and reasonableness of scenario-based exposure assessments. The value of a

42.5 Using Biomarker Measurements to Estimate PesticideExposure and Dose

particular biomarker for exposure assessment depends generally on the half-life of the marker in the body, the specificity of the marker for the pesticide(s) of interest, the relative ease of sample collection, and the difficulty or cost of chemical analysis.

42.5.1 STUDIES MEASURING BIOMARKERS OF PESTICIDE EXPOSURE IN HUMAN POPULATIONS A 1998 review by Fenske (Fenske, 1998) of the literature on biological monitoring for pesticides found relatively few references directly relevant for assessing nonoccupational exposures. Fenske divided the relevant publications into five groups: (1) 9 review articles, reports, or compendia published in the past 10 years; (2) 8 small-scale or controlled urine monitoring studies; (3) 12 general or subpopulation urine monitoring studies; (4) 3 blood or serum monitoring studies; and (5) 7 studies measuring pesticides in breast milk and 1 in adipose tissue (Fenske, 1998). Urinary metabolites are the most commonly used biomarkers of organophosphate exposure in both adults and children. Although these samples are relatively easy to collect and analyze, their utility is limited by the lack of validated pharmacokinetic models for most compounds of interest (Woollen, 1993), and the fact that urinary metabolites are elimination products rather than direct markers of exposure or internal dose (Fenske, 1997). Urine samples are typically collected either as "spot" samples, which can be first morning voids or convenience samples taken as they become available during the day, or as aggregate 24-hr or week-long samples. Although repeat spot samples are desirable because they provide data on rate of excretion and furnish a measure of within-person variability, insuring subject compliance can be complicated and timeconsuming. A variety of methods exist for quantitative analysis of urine samples (Hill et aI., 1995b; Nigg et aI., 1990), which are normally adjusted by dividing measured biomarker concentrations by creatinine mass to normalize for differences in metabolic rate and urine dilution. Because children typically have higher creatinine excretion rates than adults due to higher metabolic and tissue turnover rates, comparing adjusted data between adults and children is not straightforward. Relatively few studies involving measurement of urinary pesticide biomarkers have been conducted in human popu1ations, and most have not collected data on children. Fenske (1998) identified 16 published studies that use urine monitoring to estimate nonoccupational pesticide exposures in the United States, Israel, and Canada (Table 42.5). Most studies have been small-scale, convenience samples of at-risk populations, such as occupants of treated residences (Esteban et aI., 1996; Richter et aI., 1992b), people likely to be exposed by aerial sprayings (Dong et aI., 1994), and families of farm workers (Shealy et aI., 1997). Larger-scale, urinary biomarker studies in adults have been conducted in conjunction with the National Health and Nutrition Examination Survey (NHANES) 11 (1976-1980)

897

(Kutz et aI., 1992) and III (1988-1991) (Hill et aI., 1995a, c), which uses a probability-based sampling scheme for the U.S. population. Results from 1000 adults in NHANES III have been used to establish "reference ranges" for 12 pesticide biomarkers, where a reference range is defined as the distribution of biomarker concentrations in a population with no known exposure or only minimal exposure to the toxic ant of interest (Hill et aI., 1995c). A reference range provides an indication of median and high-end biomarker levels in the adult D.S. population, and serves as a basis for comparison with data from children that might be collected in the future. The reference ranges show that p-dichlorobenzene is ubiquitous and most adults 29-59 years of age have been exposed to naphthalene, chlorpyrifos, and pentachlorophenol. Over the 1O-year period between samples, time trend data indicate exposures to chlorpyrifos are increasing whereas exposures to pentachlorophenol are apparently decreasing (Hill et aI., 1995c). Although a few biomarker studies have been conducted in children, virtually all utilize convenience samples of potentially at-risk groups, such as children of pesticide workers (Bradman et aI., 1997; Loewenherz et aI., 1997). One of the largest is an Arkansas study using single first-morning-void urine samples to examine 197 children thought to be exposed to chlorinated phenols and phenoxy herbicides (Hill et aI., 1989). As in adults, p-dichlorobenzene and pentachlorphenol were detected in nearly 100% of the urine samples, with median concentrations of 9 and 14 ppb, respectively. Measurable levels of some herbicide metabolites were found in more than 50% of the samples, with others present in only 10-20% of the samples. In another study of at-risk children, investigators examined dimethylphosphate metabolite levels in 160 spot urine samples from children of 48 families of pesticide applicators and 14 reference families in central Washington's applegrowing region (Loewenherz et aI., 1997). Findings demonstrated that children of pesticide applicators had significantly higher exposures than children from the reference families, even though they all lived in the same community. Results also suggest that decreasing age and closer residential proximity to orchards were associated with higher exposures. Although the study does not use a probability-based sampling frame, it does involve collection of repeat biomarker measurements over a 3- to 7-day period in each child, thereby allowing for examination of both within and between individual variability in exposures. Because neither of the two studies just described reported measurements of pesticide concentrations in all relevant environmental media, it was not possible to determine the primary pathways and routes of exposure. Persistent chlorinated pesticides or their metabolites can be measured readily in blood and serum, and these biomarkers have been used to provide an estimate oflong-term body burden (Pirkle et aI., 1995). However, most active ingredients in current use, such as organophosphate compounds and pyrethrins and pyrethoids, have relatively short half-lives in the body and are not easily measurable in blood or serum. Measurable levels of the pesticides ethion and carbaryl and the herbicide atrazine

CHAPTER 42 Emerging Issues: Children's Exposure

898

Table 42.5 Published Studies Using Urine Monitoring to Estimate Nonoccupational Exposures in the United States and Canadaa Study

Scenario

Population

Pesticides

Harris and Solomon (1992)

Controlled exposure to treated turf

Adults

2,4-D

Harris et al. (1992)

Residential exposure to treated turf

Adult applicators and bystanders

2,4-D

Richter et al. (1992b)

Indoor residential exposure from

Family: 2 adults, 2 children

Diazinon

Dong et al. (1994, 1996)

Modeling of population exposure from

Adults

Malathion

Adults

Disodium octaborate tetrahydrate

commercial application aerial application Krieger et al. (1996)

Controlled exposure to indoor carpet treatment

Shealy et al. (1997)

Farm pesticide application

Farm families

Carbaryl, dicamba, 2,4-D, others

Hill et al. (1989)

Community exposure to hazardous

Children

Phenoxy herbicides

waste site Kutz et al. (1992)

NHANES II subsample

U.S. adults

Pentachlorophenol, chlorpyrifos,

Richter et al. (1992a)

Residential treatments and agricultural

Adult kibbutz workers and

OP pesticides

Thompson and Treble (1994);

Population survey

parathion, others drift

residents Various ages

Pentachlorophenol

Treble and Thompson (1996) Hill et al. (1995a)

NHANES III subsample

1000 adults

p-Dichlorobenzene

Hill et al. (1995c)

NHANES III subsample

1000 adults

Chlorpyrifos, pentachlorophenol,

Esteban et al. (1996)

Indoor residential pest control

Residents

Methyl parathion

Davies and Peterson

Community exposure from agricultural

General population of Dade

OP pesticides, chlorpyrifos

others treatments (1997) Loewenherz et al. (1997)

and residential uses of pesticides Community exposure from agricultural pesticide use

County, FL Children 1--6 years old

Azinphos-methyl, phosmet, parathion, chlorpyrifos

aSource: Data from Fenske (1998). Reproduced with permission.

have been observed in saliva, but few human studies have been conducted and none in children (Fenske, 1997; Nigg and Wade, 1992). Beyond the NHANES reference range data, few populationbased studies of pesticide exposures, using either environmental or biomarker measurements, have been conducted and, until the late 1990s, none had been done in a sample of children. In the 1980s, the Non-Occupational Exposure Assessment Study (NOPES) was one of the first population-based studies of pesticide exposures in the U.S. (Whitmore et aI., 1994). It focused on estimating adults' exposure (using outdoor, indoor, and personal air samples, as well as some surface residue samples) to 32 pesticides in two U.S. cities: Jacksonville, Florida, selected to representative high pesticide usage; and Springfield, Massachusetts, selected to represent lowto-moderate pesticide usage. Results demonstrated that a number of widely used home and garden pesticides were detectable in indoor residential air in the two cities. Subsequent indirect analysis suggested that exposures via the dietary pathway were likely to be greater than exposures occurring by either inhalation or dermal absorption for most of the 32 pesticides examined, with cyclodiene termiticides being the notable exception.

42.5.2 THE NEXT GENERATION OF CHILDREN'S PESTICIDE EXPOSURE STUDIES

Fenske (1998) has identified 24 ongoing studies in the United States and Canada that involve biological measurements of nonoccupational pesticide exposures. The 23 that use urinary biomarkers are summarized in Table 42.6. In the near future, this next generation of biomarker studies will begin providing more and better data on children's pesticide exposures. Many of the studies focus particularly on children, including children offarm families, children living along the U.S.-Mexico border, and children living in urban and nonurban settings. Several of these studies, especially the Phase 1 field studies that are part of the National Human Exposure Assessment Survey (NHEXAS), focus on measuring multipathway exposures to mUltiple pesticides using a combination of environmental, personal, biological and time-activity measurements. The goal is to measure or estimate the contributions to exposure from important pathways and to collect and analyze data so that it is possible to assess both aggregate exposures (total exposure from ingestion, inhalation, and dermal contact for a single pesticide) and cu-

42.5 Using Biomarker Measurements to Estimate PesticideExposure and Dose

8~

Table 42.6 Current Pesticide Biological Monitoring Studies in the United States and Canadaa Study

Location

Study design

Target pesticides

Univ.ofMN

Minnesota

State population-based probability

Atrazine, malathion, chlorpyrifos, carbaryl

design Univ.ofMN

Minneapolis-St.

Urban-rural comparison of

Paul and rural

Atrazine, malathion, chlorpyrifos, carbaryl, 2,4-[

3- to 12-year-old children

counties Univ.ofMN

Minneapolis

Comparison of K-5 school

Malathion, chlorpyrifos, carbaryl

children from low-income neighborhoods Univ. of AZ

Arizona

State population-based probability


design Univ. of AZ

U.S.-Mexico border

Geographic-based and pop.-based


probability design Univ. of AZ

Yuma County, AZ

Subpopulation of farmworker

Chlorpyrifos, diazinon

children EmoryUniv.

Maryland

Population-based probability


design Univ.ofWA

Chelan-Douglas and

Clinic-based urban-rural

King Counties, WA Univ.ofWA

Chelan-Douglas

Subpop. defined by proximity to

Counties, WA Univ.ofWA

Yakima County, Monterey County,

Probability-based within defined

OP pesticides (dialkylphosphates)

communities Clinic-based enrollment during

CA OR Health Science


treated farmland

WA Univ. CA Berkeley


comparison


prenatal care

Oregon

Subpop. of farmworker children


Controlled study

Adult volunteers

Chlorpyrifos

Riverside, CA

Convenience sample of pesticide-

Chlorpyrifos

Univ. Univ.ofCA Riverside Univ.ofCA

using population

Riverside Univ. of Guelph

Ontario, Canada

Convenience sample of home

Chlorpyrifos

users of pesticides National Cancer

Iowa, North

ATSDR

Subpop of farm families from

Carolina

Institute

Texas border area

Extensive pesticide screen

Agricultural Health Study Component of epidemiological study of neural tube defects

ATSDR

U.S.-Mexico border

Comparison of high and low

EPAlORD

Ohio

Convenience sample of pesticide-

Methyl parathion, atrazine, 2,4-D, chlorpyrifos, carbaryl, malathion OP pesticides (dialkylphosphates)

potential exposure (children) 2,4-D

using population EPAlORD

California and North

Clinic-based high-risk pediatric

Carolina Health Canada

Ontario, Canada


population Probability-based sample of farm family population

a Source: Data from Fenske (1998). Reproduced with permission.

Phenoxy herbicides

900

CHAPTER 42


mulative risks (sum of health risks from exposure to multiple pesticides). The completed Minnesota Children's Pesticide Exposure Study (MNCPES) is an example of the new generation of exposure monitoring studies (Adgate et aI., 2000; Quackenboss et aI., 2000). The primary objective was to characterize children's exposure to selected pesticides through a combination of complementary monitoring approaches, including concurrent biomarker, personal, environmental, and activity-pattern measurements. The design strategy focused on testing the hypothesis that ingestion of organophosphate pesticides from the dietary pathway is the primary route of exposure for children. Emphasis was placed on measuring exposures to four pesticide compounds (i.e., chlorpyrifos, diazinon, malathion, atrazine) that were selected based on their frequent use, presence in multiple environmental media, expected population exposures, and related human toxicity. The study was conducted during the summer of 1997 and involved a stratified random sample of households with children ages 3-12 years located in either (a) the cities of Minneapolis and St. Paul (urban households) or (b) Rice and Goodhue Counties (nonurban households) just south of the metropolitan area. The results from a residential survey documenting storage and use of products containing target pesticides were used to screen homes and preferentially select households where children were likely to have higher exposures. The study suc-

cessfully obtained pesticide exposure data on lO2 children, including measurements of personal exposures (air, hand rinse, duplicate diet), environmental concentrations (residential indoor and outdoor air; drinking water; residential surfaces; soil), activity patterns (questionnaire, diary, videotaping), and internal dose (metabolites in urine and blood). MNCPES demonstrates that it is feasible and practical to use a population-based sample to obtain environmental, personal, and biological measurements, as well as time-activity data, from children and their households as part of an intensive, six-day monitoring protocol. The extensive exposure monitoring (multiple pathways, routes, and pesticides) conducted for each of the lO2 children in the MNCPES is unprecedented, and the resulting database will provide the most comprehensive and in-depth characterization of children's exposures to organophosphate pesticides ever undertaken.

42.6 LOOKING AHEAD As we enter the twenty-first century there are legitimate reasons to be concerned about possible adverse health effects in children from pesticide exposures. The overarching need right now is for sound science to help answer critical questions. Which pesticides are of primary concern based on health-based criteria? Who is exposed to levels above health-related benchmarks? Personal (point of Contact) Measurements

Environmental Concentration Measurements

I I

Outdoor

Air

I I I

t Soil (Entranceway)

Inhal~tion

Breathing Zone Air

I

I I I I I I I I

8 - - - -1---------------, I I

: ~

I

I I

r----------------------Indoor Air

Dust on Non-floor Surfaces

----r~~~~±~~~~~~:

l_ _ _ _ _ _ _ _ _ _ __ _ _ _.._ --

---------- ---- --- - ----- ---- --- --

----

Ingestion

Food, Beverages

I I I I I I I I Dermal

...

i

1

!

1

r----------'

Dermal Contact

Dust on Floors, Carpet, etc. I

Biomarker Measurements

,

Ab~orption

I

I I I I

Figure 42.2 Schematic representation of (a) important residential pesticide exposure pathways, (b) potential cross-media transfers, and (c) the relationship between environmental, personal, and biomarker measurements. Source U.S. EPA (l999b).

Body Burden (e.g., urine, blood, hair)

References

How, why, and when are they exposed to elevated levels? Which actions will be most effective and efficient in preventing or reducing exposures? Unfortunately, the data necessary to answer these and related questions with an appropriate degree of certainty are scarce, uneven, and fragmented. The 1993 NRC report Pesticides in the Diets of Infants and Children and the Food Quality Protection Act of 1996 has focused attention on the critical need to better understand children's exposure to pesticides from both dietary and nondietary pathways. Nevertheless, because of the existing paucity of adequate and appropriate data virtually all assessments of children's exposure to pesticides must resort to scenario-based approaches. The major scientific uncertainties inherent in residential pesticide exposure assessments for children include uncertainties about transfer of pesticide residues from surfaces to skin, rates of dermal absorption, ingestion of pesticide residues adhering to hands, and quantitative relationships between exposure, dose, and biomarkers. To address these and other important data needs, a variety of research approaches are required. Biomarkers must be linked to improved pharmacokinetic models that can be used to calculate internal dose and exposure, and which serve to validate exposure estimates from scenario-based approaches. Predictive multimedia, multipathway, and multichemical models need to be developed and validated to assist in characterizing aggregate exposure and cumulative risk. Firstgeneration aggregate exposure methods have been developed for some chemicals (International Life Sciences Institute, 1998; Shurdut et aI., 1998), but much more work is needed to apply them to children. These models must also address key exposure issues, such as the potential for cross-media transfer to nontarget surfaces or sinks, the timing of multiple exposures relative to one another, and the role of developmental or behavioral characteristics in determining high-end and maximally exposed individuals. The schematic in Fig. 42.2 provides a simplified representation of nondietary pesticide exposure pathways and potential cross-media transfers in the residential setting. It summarizes our current understanding of how children can be exposed to pesticides in and around their dwellings. The individual boxes indicate locations where exposure-related measurements can be made. Solid lines between or within media, such as those between outdoor air and indoor air, represent pathways driven largely by physical or chemical processes; dotted lines represent pathways and exposures that are largely influenced by human activities. Pesticides in outdoor and indoor air, tap water, dust, and soil can come into contact with children during their normal daily activities, and cause exposures through inhalation, ingestion, or dermal contact. Once a pesticide enters the body by either intake or uptake, that portion which reaches an accessible tissue or excreta of interest (i.e., delivered dose or body burden) can often be measured using sophisticated analytical techniques. To adequately characterize these various pathways and determine their relative contributions to actual exposures and doses under real-life conditions and situations, it is necessary to undertake a systematic research program involving well-

901

designed and complementary studies. As shown in Fig. 42.2, this will necessarily entail three interrelated types of measurements: (1) measurements of pesticide concentrations in environmental media; (2) measurements of personal pesticide exposures; and (3) measurements of exposure biomarkers in urine or other appropriate biological material. Although each category of information is important in its own right, the value of these data increase dramatically when they are interconnected. If, for example, the three types of measurements are conducted concurrently or in a complementary sampling frame, then the data can be used to elucidate the entire sequence of events depicted in Fig. 42.2. By collecting and analyzing matched data on environmental concentrations, personal exposures, and body burden, it becomes possible to understand the residential conditions, including critical pathways, which cause elevated pesticide exposures in children. We can then make informed choices about whether associated health risks are acceptable and, if not, which mitigation strategies are most cost-effective. Instead of asking "how much do these studies cost?" we should be asking the more important question, "can we afford not to obtain the necessary information?"

ACKNOWLEDGEMENT During the writing of this Chapter Drs. Adgate and Sexton were supported in part by U.S. EPA STAR Grant R825283 to the University of Minnesota.

REFERENCES Adgate, l. L., Kukowski, A., Stroebel, c., Shubat, P. l., Morrell, S., Quackenboss, l. l., Whitmore, R. w., and Sexton, K. (2000). Household pesticide storage and use patterns in Minnesota. J. Expo. Anal. Environ. Epidemiol. 10(2), 159-167. Ames, R. G., Brown, S. K., Rosenberg, l., lackson, R. l., Stratton, l. W., and Quenon, S. G. (1989). Health symptoms and occupational exposure to flea control products among California pet handlers. Am. Ind. Hyg. Assoe. J. 50, 466-472. American Society of Testing and Materials (ASTM) (1993). "Standard Practice for Collection of Dust from Carpeted Floors for Chemical Analysis." Standard Practice D 5438-93. Am. Soc. Testing and Materials, Philadelphia. Banerjee, B., Zaidi, S., Pasha, S., Rawat, D., Koner, B., and Hussain, Q. (1997). Levels of HCH residues in human milk samples from Delhi, India. Bull. Environ. Contam. Toxieol. 59(3), 403-406. Basri-Ustunbas, R, Ozturk, M., Hasanoglu, E., and Dogan, M. (1994). Organochlorine pesticide residues in human milk in Kayseri. Hum. Exp. Toxieol. 13,299-302. Bradman, M. A., Harnly, M. E., Draper, w., Seidel, S., Teran, S., Wakeham, D., and Neutra, R. (1997). Pesticide exposures to children from California's central valley: Results of a pilot study. J. Expo. Anal. Environ. Epidemiol. 7(2), 217-234. Bradman, M. A., Harnly, M. E, Goldman, L. R., Marty, M. A., Dawson, S. v., and Dibartolomeis, M. l. (1994). Malathion and maloxon environmental levels used for exposure assessment and risk characterization of aerial applications to residential areas of southern California, 1989-1990. J. Expo. Anal. Environ. Epidemiol. 4(1), 49-63. Byrne, S. L., Shurdut, B. A., and Saunders, D. G. (1998). Potential chlorpyrifos exposure to residents following standard crack and crevice treatment. Environ. Health Perspeet. 106(11),725-731.

902

CHAPTER 42


Carmichael, L., ed. (1946). "Manual of Child Psychology." Wiley, New York. Cashman, J. R., Perotti, B. Y., Berkman, C. E., and Lin, J. (1996). Pharmacokinetics and molecular detoxication. Environ. Health Perspect. 104(Suppl. 1), 23-40. Chambers, J. (1996). "Potential Exposure of Children to Organophosphate Pesticides from Pets." EPA STAR Grant Abstract, Mississippi State University. Available at http://es.epa.gov/ncerqa/grants/96/expochil.html. Colt, J. S., Zahm, S. H., Camann, D. E., and Hartge, P. (1998). Comparison of pesticides and other compounds in carpet dust samples collected from used vacuum cleaner bags and from a high-volume surface sampler. Environ. Health Perspect. 106(11),721-724. Davies, J. E., and Peterson, J. C. (1997). Surveillance of occupational, accidental, and incidental exposure to organophosphate pesticides using urine alkyl phosphate and phenolic metabolite measurements. Ann. N. Y. Acad. Sci. 837, 257-268. Dogheim, S. M., Mohamed el, Z., Gad Alia, S. A., el-Saied, S., Emel, S. Y., Mohsen, A. M., and Fahmy, S. M. (1996). Monitoring of pesticide residues in human milk, soil, water, and food samples collected from Kafr EI-Zayat Governorate. J. AOAC Int. 79(1),111-116. Dong, M., Draper, W., Papenek, P., Ross, J., Woloshin, K., and Stephens, R. (1994). Estimating malathion does in California's medfly eradication campaign using a physiologically based pharmacokinetic model. In "Environmental Epidemiology;' Advances in Chemistry Series, Vol. 241, pp. 189208. Am. Chem. Soc., Washington, DC. Dong, M., Ross, J., Thongsinthusak, T., and Krieger, R. (1996). Use of spot urine sample results in physiologically-based pharmacokinetic modeling of absorbed malathion doses in humans. In "Biomarkers for Agrochemicals and Toxic Substances," ACS Symposium Series, Vol. 643, pp. 229-241. Am. Chem. Soc., Washington, DC. Edwards, R., and Lioy, P. (1999). The EL sampler: A press sampler for the quantitative estimation of dermal exposure to pesticides in housedust. J. Expo. Anal. Environ. Epidemiol. 9(5), 521-529. Esteban, E., Rubin, c., Hill, R., Olson, D., and Pearce, K. (1996). Association between indoor residential contamination with methyl parathion and urinary para-nitrophenol. J. Expo. Anal. Environ. Epidemiol. 6(3), 375-387. Fenske, R. A. (1988). Visual scoring system for fluorescent tracer evaluation of dermal exposure to pesticides. Bull. Environ. Contam. Toxicol. 41(5), 727736. Fenske, R. A. (1990). Nonuniform dermal deposition patterns during occupational exposure to pesticides. Arch. Environ. Contam. Toxicol. 19(3), 332337. Fenske, R. A. (1997). Pesticide exposure assessment of workers and their families. Occupational Med.: State of the Art Rev. 12(2),221-237. Fenske, R. A. (1998). "Status Report on Biological Monitoring Research Relevant to Aggregate Exposure Assessment under the Food Quality Protection Act." Subcommittee on Aggregate Exposure Assessment, International Life Sciences Institute (ILSI), Seattle, WA. Fenske, R. A., and Lu, C. (1994). Determination of handwash removal efficiency: Incomplete removal of the pesticide chlorpyrifos from skin by standard handwash techniques. Am. Ind. Hyg. Assoc. J. 55(5), 425-432. Fenske, R. A., Black, K. G., Elkner, K. P., Lee, C. L., Methner, M. M., and Soto, R. (1990). Potential exposure and health risks of infants following indoor residential pesticide applications. Am. J. Public. Health 80(6), 689693. Fenske, R. A., Curry, P. B., Wandelmaier, F., and Ritter, L. (1991). Development of dermal and respiratory sampling procedures for human exposure to pesticides in indoor environments. J. Expo. Anal. Environ. Epidemiol. 1(1), 11-30. Fenske, R. A., Wong, S. M., Leffingwell, J. T., and Spear, R. C. (1986). A video imaging technique for assessing dermal exposure. 11. Fluorescent tracer testing. Am. Ind. Hyg. Assoc. J. 47(12), 771-775. Freeman, N. C. G. (1999). Variation in hand-to-mouth behavior with age, personal communication. Furst, P., Furst, C., and Wilmers, K. (1994). Human milk as a bioindicator for body burden of PCDDs, PCDFs, organochlorine pesticides, and PCBs. Environ. Health Perspect. 102(Suppl. 1), 187-193.

Geno, P. W., Camann, D. E., Harding, H. F., Villalobos, K., and Lewis, R. G. (1996). Handwipe sampling and analysis procedure for the measurement of dermal contact with pesticides. Arch. Environ. Contam. Toxieol. 30, 132138. Gurunathan, S., Robson, M., Freeman, N., Buckley, B., Roy, A., Meyer, R., Bukowski, J., and Lioy, P.1. (1998). Accumulation of chlorpyrifos on residential surfaces and toys accessible to children. Environ. Health Perspect. 106(1),9-16. Guzelian, P., Henry, C., and Olin, S., eds. (1992). "Similarities and Differences between Children and Adults: Implications for Risk Assessment." International Life Sciences Institute Press, Washington, DC. Harris, S. A., and Solomon, K. R. (1992). Human exposure to 2,4-D following controlled activities on recently sprayed turf. J. Environ. Sci. Health [Bl 27(1),9-22. Harris, S. A., Solomon, K. R., and Stephenson, G. R. (1992). Exposure of homeowners and bystanders to 2,4-dichlorophenoxyacetic acid (2,4-D). J. Environ. Sci. Health [Bl 27(1),23-38. Hill, R. H., Jr., Ashley, D. L., Head, S. L., Needham, L. L., and Pirkle, J. L. (1995a). p-Dichlorobenzene exposure among 1,000 adults in the United States. Arch. Environ. Health 50(4), 277-280. Hill, R. H., Jr., Shealy, D. B., Head, S. L., Williams, C. c., Bailey, S. L., Gregg, M., Baker, S. E., and Needham, L. L. (1995b). Determination of pesticide metabolites in human urine using an isotope dilution technique and tandem mass spectrometry. J. Anal. Toxieol. 19(5),323-329. Hill, R. H., Jr., To, T., Holler, J. S., Fast, D. M., Smith, S. J., Needham, L. L., and Binder, S. (1989). Residues of chlorinated phenols and phenoxy acid herbicides in the urine of Arkansas children. Arch. Environ. Contam. Toxicol. 18(4),469-474. Hill, R. H. Jr., Head, S. L., Baker, S., Gregg, M., Shealy, D. B., Bailey, S. L., Williams, C. C., Sampson, E. J., and Needham, L. L. (1995c). Pesticide residues in urine of adults living in the United States: Reference range concentrations. Environ. Res. 71, 99-106. International Life Sciences Institute (1998). "Aggregate Exposure Assessment: An ILSI Risk Sciences Institute Workshop Report." International Life Sciences Institute, Washington, DC. Johansen, H., Becher, G., Polder, A., and Skaare, J. (1994). Congener-specific determination of polychlorinated biphenyls and organochlorine pesticides in human milk from Norwegian mothers living in Oslo. J. Toxicol. Environ. Health 42, 157-171. Kelly, P., ed. (1983). "First Year Baby Care: An Illustrated Step-by-Step Guide for New Parents." Meadowbrook Press, Deephaven, MN. Klopov, v., Odland, J. 0., and Burkow, I. C. (1998). Persistent organic pollutants in maternal blood plasma and breast milk from Russian arctic populations.Int. J. Circumpolar Health 57(4),239-248. Krieger, R. I., Dinoff, T. M., and Peterson, J. (1996). Human disodium octaborate tetrahydrate exposure following carpet flea treatment is not associated with significant dermal absorption. J. Expo. Anal. Environ. Epidemiol. 6(3), 279-288. Krieger, R. I., Rosenheck, L. A., and Schuester, L. L. (1997). Adult and infant abamectin exposures following Avert 310 and pressurized gel. Bull. Environ. Contam. Toxicol. 58(5),681-687. Kroger, M. (1972). Insecticide residues in human milk. J. Pediatries 80(3), 401405. Kutz, F. w., Cook, B. T., Carter-Pokras, O. D., Brody, D., and Murphy, R. S. (1992). Selected pesticide residues and metabolites in urine from a survey of the U.S. general population. J. Toxicol. Environ. Health 37, 277-291. Larsen, J. c., and Pascal, G. (1998). Workshop on the applicability of the ADIto infants and children: Consensus summary. Food Addit. Contam. 15(Suppl.), 1-9. Lewis, R. G., Fortmann, R. c., and Camann, D. E. (1994). Evaluation of methods for monitoring the potential exposure of small children to pesticides in the residential environment. Arch. Environ. Contam. Toxicol. 26, 37-46. Lioy, P. J. (1990). Assessing total human exposure to contaminants. Environ. Sci. Technol. 24(7), 938-945. Lipscomb, J. w., Kramer, J. E., and Leikin, J. B. (1992). Seizure following brief exposure to the insect repellent N,N-diethyl-m-toluamide. Ann. Emerg. Med. 21(3), 315-317.

References

Loewenherz, C., Fenske, R. A., Simcox, N. J., Bellamy, G., and Kalman, D. (1997). Biological monitoring of organophosphorus pesticide exposure among children of agricultural workers in central Washington state. Environ. Health Perspeet. 105, 1344-1353. Lu, e., and Fenske, R. (1999). Dermal transfer of chlorpyrifos residues from residential surfaces: Comparison of hand press, hand drag, wipe, and polyurethane foam roller measurements after broadcast and aerosol pesticide applications. Environ. Health Perspeet. 107(6),463-467. Marty, M. A., Dawson, S. v., Bradman, M. A., Hamly, M. E., and Dibartolomeis, M. J. (1994). Assessment of exposure to malathion and maloxon due to aerial application over urban areas of southern California. J. Expo. Anal. Environ. Epidemiol. 4(1), 65-81. Moore, J. e., Dukes, J. C., Clark, J. R., Malone, J., Hallmon, e. F., and Hester, P. G. (1993). Downwind drift and deposition of malathion on human targets from ground ultra-Iow volume mosquito sprays. J. Am. Mosq. Control Assoe. 9(2),138-142. Nelson, W. e., Ott, w. R., and Robinson, J. P. (1994). ''The National Human Activity Pattern Survey (NHAPS): Use of Nationwide Activity Data for Human Exposure Assessment." EPA 600/A-94/147, Atmospheric Research and Exposure Assessment Lab, U.S. Environmental Protection Agency, Research Triangle Park, Ne. Ness, S. A. (1994). "Surface and Dermal Monitoring for Toxic Exposures." Van Nostrand Reinhold, New York. Nigg, H. N., and Wade, S. E. (1992). Saliva as a monitoring medium for chemicals. Rev. Environ. Contam. Toxieol. 129,95-119. Nigg, H. N., Beier, R. C., Carter, 0., Chaisson, e., Franklin, e., Lavy, T., Lewis, R. G., Lombardo, P., McCarthy, J. F., et al. (1990). Exposure to pesticides. In "The Effect of Pesticides on Human Health," pp. 35-130. Princeton Sci. Publ., Princeton, NJ. Nishioka, M. G., Burkholder, H. M., Brinkman, M. e., Gordon, S. M., and Lewis, R. G. (1996). Measuring transport of lawn-applied herbicide acids from turf to home: Correlation of dislodgeable 2,4-D turf residues with carpet dust and carpet surface residues. Environ. Sci. Teehnol. 30,3313-3320. Nishioka, M. G., Burkholder, H. M., Brinkman, M. C., and Lewis, R. G. (1999). Distribution of 2,4-dichlorophenoxyacetic acid in floor dust throughout homes following homeowner and commercial lawn applications: Quantitative effects of children, pets, and shoes. Environ. Sei. Teehnol. 33, 13591365. National Research Council (NRC) (1991). "Human Exposure Assessment for Airborne Pollutants. Advances and Opportunities." National Academy Press, Washington, De. National Research Council (NRC) (1993). "Pesticides in the Diets of Infants and Children." National Academy Press, Washington, De. Osimitz, T. G., and Murphy, J. V. (1997). Neurological effects associated with use of the insect repellent N,N-diethyl-m-toluamide (DEET). J. Toxieol. Clin. Toxieol. 35(5), 435-441. PirkIe, J. L., Needham, L. L., and Sexton, K (1995). Improving exposure assessment by monitoring human tissues for toxic chemicals. J. Expo. Anal. Environ. Epidemiol. 5(3),405-424. Plunkett, L., Turnbull, D., and Rodricks, J. (1992). Differences between adults and children affecting exposure assessment. In "Similarities and Differences between Children and Adults: Implications for Risk Assessment," (P. GuzeIian, e. Henry, and S. OIin, eds.). International Life Sciences Institute Press, Washington, DC. Quackenboss, J., PeIIizzari, E., Shubat, P., Whitmore, R., Adgate, J., Thomas, K, Freeman, N., Stroebel, C., Lioy, P., et al. (2000). Design strategy for a multipathway pesticide exposure study in children. J. Expo. Anal. Environ. Epidemiol. 10(2), 145-158. Quinsey, P. M., Donohue, D. e., and Ahokas, J. T. (1995). Persistence of organochlorines in breast milk of women in Victoria, Australia. Food. Chem. Toxieol. 33(1), 49-56. Reed, K. J., Jimenez, M., Freeman, N. C., and Lioy, P. J. (1999). Quantification of children's hand and mouthing activities through a videotaping methodology. J. Expo. Anal. Environ. Epidemiol. 9(5), 513-520. Richter, E., Chuweres, P., Levy, Y., Gordon, M., Grauer, F., Marzouk, J., Levy, S., Barron, S., and Gruener, N. (1992a). Health effects from expo-

903

sure to organophosphate pesticides in workers and residents in Israel. Israel J. Med. Sei. 28, 584-598. Richter, E. D., Kowalski, M., Leventhal, A., Grauer, F., Marzouk, J., Brenner, S., Shkolnik, I., Lerman, S., Zahavi, H., et al. (1992b). Illness and excretion of organophosphate metabolites four months after household pest extermination. Arch. Environ. Health 47(2), 135-138. Rinehart, R., and Rogers, J. (1995). "Sampling House Dust for Lead. Basic Concepts and Literature Review." Westat Inc., RockviIIe, MD (and EPA 747-R-95-007, U.S. EPA, Washington, DC). Roberts, J. W., Budd, W. T., Ruby, M. G., Bond, A. E., Lewis, R. G., Wiener, R. W., and Camann, D. E. (1991). Development and field testing of a high volume sampler for pesticides and toxics in dust. J. Expo. Anal. Environ. Epidemiol. 1(2), 143-155. Ross, J., Fong, H. R., Thongsinthusak, T., Margetich, S., and Krieger, R. (1990). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: An interim report. Chemosphere 20(3-4),349-360. Ross, J., Fong, H. R., Thongsinthusak, T., Margetich, S., and Krieger, R. (1991). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: Using the CDFA roller method. Interim report n. Chemosphere 22(9-10), 975-984. Saleh, M., Afify, A., Ragab, A., EI-Baroty, G., Kamel, A., and EI-Sebae, A. (1996). Breast milk as a biomarker for monitoring human exposure to environmental pollutants. In "Biomarkers for Agrochemicals and Toxic Substances," ACS Symposium Series, Vol. 643, pp. 114-125. Am. Chem. Soc., Washington, DC. Schade, G., and Heinzow, B. (1998). Organochlorine pesticides and polychlorinated biphenyls in human milk of mothers living in northern Germany: Current extent of contamination, time trend from 1986 to 1997 and factors that influence the levels of contamination. Sci. Total Environ. 215(1-2), 3139. Schwab, M., Terblanche, A. P., and Spengler, J. D. (1991). Self-reported exertion levels on time/activity diaries: Application to exposure assessment. J. Expo. Anal. Environ. Epidemiol. 1(3),339-356. Sexton, K, Callahan, M. A., and Bryan, E. F. (1995a). Estimating exposure and dose to characterize health risks: The role of human tissue monitoring in exposure assessment. Environ. Health Perspeet. 103(Suppl. 3), 13-29. Sexton, K, Callahan, M. A., Bryan, E. F., Saint, e. G., and Wood, W. P. (I 995b). Informed decisions about protecting and promoting public health: Rationale for a National Human Exposure Assessment Survey. 1. Expo. Anal. Environ. Epidemiol. 5(3), 233-256. Shealy, D. B., Barr, J. R., Ashley, D. L., Patterson, D. G., Jr., Camann, D. E., and Bond, A. E. (1997). Correlation of environmental carbaryl measurements with serum and urinary I-naphthol measurements in a farmer applicator and his family. Environ. Health Perspeet. 105(5), 510-513. Shurdut, B. A., Barraj, L., and Francis, M. (1998). Aggregate exposures under the Food Quality Protection Act: An approach using chlorpyrifos. Regul. Toxieo/. Pharmacal. 28(2),165-177. Simcox, N. J., Fenske, R. A., Wolz, S. A., Lee, I.-e., and Kalman, D. A. (1995). Pesticides in household dust and soil: Exposure pathways for children of agricultural families. Enviran. Health Perspeet. 103(12), 1126-1134. Thompson, T. S., and Treble, R. G. (1994). Preliminary results of a survey of pentachlorophenol levels in human urine. Bull. Environ. Contam. Taxiea/. 53(2), 274-279. Treble, R. G., and Thompson, T. S. (1996). Normal values for pentachlorophenol in urine samples collected from a general population. J. Anal. Taxieol. 20(5),313-317. U.S. Environmental Protection Agency (U.S. EPA) (1992). Guidelines for exposure assessment. Federal Register 57(104),22888-22938. U.S. Environmental Protection Agency (U.S. EPA) (1995a). "Guidance for Risk Characterization." Science Policy Council, U.S. Environmental Protection Agency, Washington, De. U.S. Environmental Protection Agency (U.S. EPA) (l995b). "Series 875: Occupational and Residential Exposure Test Guidelines, Group B-Post Application Exposure Monitoring Test Guidelines (Working Draft, Version 5.1)." Office of Prevention, Pesticides, and Toxic Substances, U.S. EPA, Washington, De.

904

CHAPTER 42


V.S. Environmental Protection Agency (V.S. EPA) (1997). "Standard Operating Procedures (SOPs) for Residential Exposure Assessments." Office of Pesticide Programs, Health Effects Division, V.S. EPA, Washignton, DC., and Versar, Inc. V.S. Environmental Protection Agency (V.S. EPA) (1999a). "Exposure Factors Handbook." EPA/600/C-99/00I, Office of Research and Development, V.S. Environmental Protection Agency, Washington, DC. V.S. Environmental Protection Agency (V.S. EPA) (1999b). "FIFRA Scientific Advisory Panel Meeting, February 23-24, 1999. Session Ill-A Set of Scientific Issues Being Considered by the Environmental Protection Agency Regarding: Consultation on Development of Draft Aggregate Exposure Assessment Guidance Document for combining Exposure from Multiple Sources and Routes." SAP Report 99-02C, Scientific Advisory Panel, Washington, DC. Vaccaro, J., Nolan, R., Murphy, P., and Berbrich, D. (1996). "The Vse of a Vnique Study Design to Estimate Exposure to Adults and Children to Surface and Airborne Chemicals," pp. 166-183. ASTM Spec. Tech. Pub!. 1287, Am. Soc. Testing and Materials (ASTM), Philadelphia. Ware, G. W. (1994). 'The Pesticide Book." Thomson, Fresno, CA. Weaver, V. M., Buckley, T. J., and Groopman, J. D. (1998). Approaches to environmental exposure assessment in children. Environ. Health Perspect. l06(Supp!. 3), 827-832.

Whitmore, R. W., Immerman, F. W., Camann, D. E., Bond, A. E., Lewis, R. G., and Schaum, J. L. (1994). Non-occupational exposures to pesticides for residents of two V.S. cities. Arch. Environ. Contam. Toxicol. 26(1),47-59. Whitmore, R. w., Kelly, J. E., and Reading, P. L. (1992). "Executive Summary, results, and Recommendations," National Home and Garden Pesticide Vse Survey. Final Report, Vo!. 1. NTIS PB92-1747471NZ, prepared by Research Triangle Institute for Office of Pesticides and Toxic Substances, V.S. EPA, Washington, DC. Woollen, B. H. (1993). Biological monitoring for pesticide absorption. Ann. Occup. Hyg. 37(5), 525-540. Zartarian, V. G., Ferguson, A. C., and Ledde, J. O. (1997a). Quantified dermal activity data from a four-child pilot field study. J. Expo. Anal. Environ. Epidemiol. 7(4), 543-552. [Erratum 1. Expo. Anal. Environ. Epidemiol. (1998) 8(1), 109.] Zartarian, V. G., Ferguson, A. c., Ong, C. G., and Ledde, J. O. (1997b). Quantifying videotaped activity patterns: Video translation software and training methodologies. J. Expo. Anal. Environ. Epidemiol. 7(4), 535-542. Zartarian, V. G., Streicker, J., Rivera, A., Cornejo, C. S., Molina, S., Valadez, O. E, and Leckie, 1. O. (1995). A pilot study to collect micro-activity data of two- to four-year-old farm labor children in Salinas Valley, California. J. Expo. Anal. Environ. Epidemiol. 5(1), 21-34.

CHAPTER

43 Pesticide Percutaneous Absorption and Decontamination Ronald C. Wester and Howard 1. Maibach University of California, San Francisco

Human skin is a primary body organ that contacts the environment and is a route by which pesticides enter the body. Pesticides, in actuality, are designed poisons intended to protect agricultural production and home, work, and play environments. Contact with humans is not a question of if, but of when and how much, a question regulators and risk assessors have been addressing. Human skin is designed to retain body fluids and act as a barrier to the environment. This barrier can be crossed by a process called percutaneous absorption. Human absorption differs in regions of the body, and clothing can be a partial barrier. However, pesticides have managed to cross those barriers, leading to illness and death. The major processes and outcomes are discussed here.

Figure 43.1 shows human systemic parathion absorption from dermal exposure. Parathion is predicted to be lethal not only for total systemic absorption but also for exposure to limited regions. The LD 50 used for parathion is 14 mg/kg. Given a body weight of 70 kg, systemic absorption of 980 mg might result in 50% mortality. Thus, parathion lethal toxicity levels can be reached at 8-hr and longer exposures. This was unfortunately validated in the agricultural fields of California and elsewhere.

43.2 PERCUTANEOUS ABSORPTION METHODOLOGY 43.2.1 ABSOLUTE TOPICAL BIOAVAILABILITY

43.1 INTRODUCTION Percutaneous absorption is a primary focal point for dermatotoxicology and dermatopharmacology. Local and systemic toxicity depend on a chemical penetrating the skin. The skin is both a barrier to absorption and a primary route to the systemic circulation. The skin's barrier properties are impressive. Fluids and precious chemicals are reasonably retained within the body; at the same time hundreds of foreign chemicals are restricted from entering the systemic circulation. Even with these impressive barrier properties, the skin is a primary body organ that contacts the environment and is a route by which many chemicals enter the body. Some chemicals applied to the skin have proved to be toxic. These include pesticides which in actuality are designed poisons. Table 43.1 summarizes the 30-year lesson with parathion. Absorption of parathion was established for human skin contact, but other species similarly absorb the compound. Mathematical models based on quantitative structure-activity relationships now can predict a human skin permeability coefficient, but the accuracy of the predicted coefficient is not fully validated to in vivo man. Skin absorption amounts combined with toxicity data can predict potential human health hazard. Handbook of Pesticide Toxicology


The only way to determine the absolute bioavailability of a topically applied compound is to measure the compound by specific assay in blood or urine after topical and intravenous administration. This is extremely difficult to do in plasma because concentrations after topical administration are often low. However, as advances in analytical methodology bring forth more sensitive assays, estimates of absolute topical bioavailability will become more available (Wester and Maibach, 1999).

43.2.2 RADIOACTIVITY IN EXCRETA Percutaneous absorption in vivo is usually determined by the indirect method of measuring radioactivity in excreta after topical application of the labeled compound. In human studies, plasma levels of compound are extremely low after topical application, often below assay detection level, so it is necessary to use trace methodology. The compound, usually labeled with 14C or tritium, is applied and the total amount of radioactivity excreted in urine or urine plus feces determined. The amount of radioactivity retained in the body or excreted by some route not assayed (C02, sweat) is corrected for by determining the amount of radioactivity excreted after parenteral administration. This final

905


906

CHAPTER 43

Percutaneous Absorption and Decontamination

Table 43.1 Summary of Parathion Percutaneous Absorption Parathion

o ,O-Diethyl O-(4-nitrophenyl) phosphorothioate Other names: ethylparathione, parathion-ethyl

CAS: 56-38-2; mol. wt. 291.26 Molecular formula: ClOH14N05PS Nonsystemic contact and stomach-acting insecticide and acaricide with some fumigant action Nonphytotoxic except to some ornamentals and under certain weather conditions; absorption takes place readily through any portal; fatal human poisoning has followed skin exposure Skin absorption (species) Human (forearm): solvent, acetone;

Mouse (dermal): no solvent; 1.4%, 1 hr (excretion analysis)

10%,5 days (excretion analysis)a

Mouse (dermal): solvent, acetone; 32%, days (patch)d Human: solvent, acetone; forearm 8.6%,

Frog (dermal): solvent, acetone; 33%, 1 hr (patch)e

palm 11.8%, foot 13.5%, abdomen 18.5%, hand, dorsum 21.0%, forehead 36.3%, axilla 64.0%,

Quail (dermal): solvent, acetone; 40%, 1 hr (patch)'

jaw 33.9%, fossal cubitalis 28.4%, scalp 32.1 %, ear canaI46.6%, scrotum 101.6%b Rat (dermal): solvent, acetone; 59%, 1 hr

Roach (dermal): solvent, acetone; 14%, 1 hr (patch),

(patchY

Hornworm (dermal): solvent, acetone; 8%, 1 hr (patch), Skin absorption (mathematical model) kp (crnlhr)

LogP(Kow)

1.59 x 10- 2

3.83

= -2.74 + [0.71 x log P(Kow )]- [0.0061 x MW], where kp = permeability coefficient, = partition coefficient in octanol compared to water, MW = molecular weight!

based on the formula logkp P(Kow )

Toxicity Rat: Oral, male LD 50: 13-15 mg/kg

Skin, male LD 50: 21 mg/kg

Oral, female LD 50: 3-3.6 mg/kg

Skin, female LD 50: 6.8 mg/kg

aFeldmann and Maibach (1974). bMaibach et al. (1974). cKnaak et al. (1984). dMarty (1976) eShah et al. (1983). !Guy and Potts (1992).

amount of radioactivity is then expressed as the percentage of applied dose that was absorbed (Feldmann and Maibach, 1974). The equation used to determine percutaneous absorption is absorption (%) total radioactivity after topical administration --------------------------------- x 100 total radioactivity after parenteral administration

taining the chemical that has been absorbed through skin. The skin flap can be used to study percutaneous absorption in vivo or in vitro. The absorption of chemicals through skin and metabolism within the skin can be determined by assay of the perfusate (Wester and Maibach, 1997). 43.2.4 STRIPPING METHOD

43.2.3 SKIN FLAPS The methodology is to isolate surgically a portion of skin so that a singular blood supply is created to collect blood con-

The stripping method determines the concentration of chemical in the stratum corneum during an application period and predicts the percutaneous absorption of that chemical. The chemi-

43.3 Regional Variation in Percutaneous Absorption Distribution of systemic absorption Compound name: Parathion Dose: 4 \lg/cm' on whole body area (1.8 m') continuous/infinite dose application Exposure: 24 hours

Total systemic absorption: 4.918 grams· Head, Neck and Arms = 1.116 grams· Estimated systemic LD50 of Parathion is 980 mg (human, 70 kg) • indicates 50% lethality dose

Figure 43.1 Simulated parathion human skin exposure to regions of the body. As early as 8 hr following exposure lethality is possible. At 24 hr, lethality is possible if only certain body regions are exposed, for example, head and neck of a field worker.

cal is applied to skin of animals or humans, and at various skin application times the stratum corneum is removed by successive tape application and removal. The tape strippings are assayed for chemical content. Pharmacokinetic Cmax , Tmax, and area-under-curve (AUC) parameters can be calculated for stratum corneum bioavailability.

43.2.5 BIOLOGICAL RESPONSE

Another in vivo method of estimating absorption is to use a biological or pharmacological response. Here, a biological assay is substituted for a chemical assay and absorption is estimated. An obvious disadvantage to the use of a biological response is that it is only good for compounds that will elicit an easily measurable response. An example of a biological response would be the vasoconstrictor assay in which the blanching effect of one compound is compared to that of a known compound. This method is perhaps more qualitative than quantitative. The best known use of this method is in the comparison of various hydrocortisone products for skin dermatitis (Wester and Maibach, 1997).

907

43.2.6 IN VITRO METHODOLOGY In vitro percutaneous absorption is best done with human skin. The skin should be used as soon as possible and stored in the refrigerator no longer than 7 days. In vitro penetration into skin gives results suitable for distinguishing drug formulations, especially in cases where the drug will not partition into reservoir fluid. Material balance in an in vitro study design adds to the overall data presentation. In vivo verification of skin absorption, preferably in humans, adds relevance to the in vitro data. The human skin sample can be kept viable if stored properly in the refrigerator (freezing kills skin viability) and used appropriately (Wester et aI., 1984). Table 43.2 gives the in vitro human skin and in vivo percutaneous absorption of several chemicals from a variety of vehicles. The in vitro absorption is divided into skin content and receptor fluid (either buffered saline or human plasma) accumulation. Generally, receptor fluid accumulation does not agree with in vivo percutaneous absorption. The reason for this is lack of solubility in the receptor fluid. In some cases, skin content (see DDT) was reflective of in vivo absorption because the chemical was able to penetrate skin (and, lacking solubility, failed to partition into receptor fluid). Chemicals with high log P (octanol:water partition coefficient) will not partition into receptor fluid (Wester and Maibach, 1997, 1999).

43.3 REGIONAL VARIATION IN HUMAN AND ANIMAL PESTICIDE PERCUTANEOUS ABSORPTION Feldmann and Maibach (1967) first explored the potential for regional variation in percutaneous absorption. The first absorption studies were done with ventral forearm, because this site is convenient to use. However, skin exposure to chemicals exists over the entire body. They first showed regional variation with the absorption of parathion (Fig. 43.2). The scrotum was the highest-absorbing skin site (scrotal cancer in chimney sweeps was the key to identifying this fact). Skin absorption was lowest for the sole and highest around the head and face. Table 43.3 gives the effect of anatomical region on the percutaneous absorption of pesticides in humans (Maibach et aI., 1971). There are two major points in this study. First, regional variation was confirmed with the different chemicals; note that parathion and malathion are chemically related to some chemical warfare agents. Second, those skin areas that would be exposed to the pesticides, the head and face, were of the higher absorbing sites. Body areas most exposed to environmental contaminants are among the areas with the higher skin absorption. Table 43.4 gives site variability for parathion skin absorption with time. Soap-and-water wash in the first few minutes after exposure is not a perfect decontaminant. Site variation is

908

CHAPTER 43

Percutaneous Absorption and Decontamination

Ta ble 43.2 In vitro versus in vivo Percutaneous Absorptiona Percentage dose In vitro Compound

Logpa

DOT

6.9

Vehicle

Skin

Acetone Soil

Benzo [a ]pyrene

5.97

Acetone

Chlordane

5.58

Acetone

Pentachlorophenol

5. 12

Acetone

Receptor fl uid

In vivo

18.1 ± 13.4

0.08 ± 0.02

18.9 ± 9.4

1.0 ± 0.7 23.7 ± 9.7

0.04 ± 0.01 0.09 ± 0.06

1.4 ± 0.9 10.8 ± 8.2 0.3 ± 0.3

0.0 1 ± 0.06 0.Q7 ±0.06 0.04 ± 0.05

3.3 ± 51.0 ± 13.2 ± 6.0 ± 4.2 ±

0.6 ± 0.09 0.01 ±O.OO

29.2 ± 5.8 24.4 ± 6.4

Soil Soil

3.7 ± 1.7 0.11 ± 0.04

Soil PCBs (1242)

21.4 ± 8.5 18.0 ± 8.3 20.8 ± 8.3

Acetone

High

TeB

PCBs (1254)

M ineral oil

6.4 ± 6.3

Soil

1.6 ± 1.1

0.3 ± 0.6 0.04 ± 0.05

10.0 ± 16.5

0. 1 ± 0.07

2.8 ± 2.8

0.04 ± 0.05

1.6 ± 0.2 1.0 ± 1.0 0.3 ±0.2 6.7 ± 4.8 0.09 ± 0.03 28.5 ± 6.3 7.9 ± 2.2

0.02 ± 0.01 0.9 ± l.l 0.4 ± 0.5 0.4 ± 0.2 0.03 ± 0.02 0.Q7 ± 0.01

14.1 ± 1.0 14.6 ±3.6

Acetone

High

28.0 ± 8.3 20.4 ± 8.5

T CB Mineral oil Soil 2,4 Dichlorophenoxy-

2.81

13.8 ± 8.6 ± 15.9 ± 2.0 ± 3.2 ±

Acetone

acetic acid (2,4-0)

Soil

Arsenic

Water

Cadmium

Water

Soil Soil Water

Mercury

Soil

0.5 22.0 3.4 2.8 1.8

2.7 2.1 4.7 1.2 1.9

0.06 ±0.01

aNote that a log P of 6 means that 106 (1,000,000) molecules will partition into octanol for each (1) molecule which will partition into water.

200

0

w CD a:

•m

0

(J)

et

w (J) 0

FOFEARM

100

w

r

.....

1

.....

J

.....

J'"

,

.1 .... Qr LRapId PertusedJ....

.....

,

Kidney

1..... Qk

1.... Qf

Fat

r

.....

1 Diaphragm

(DFP hydrolysis by A-EST) (4) (5)

- KCDE • CCE ... C

(OI

... en .s::.

8

P

7

en

6

G)

5

c

B

4

VP

C NP

40

Duration

60

80 100

(USecl

Figure 49.16B Log-log plot for inflection region (40-500 ).is) of strengthduration curves from biventer cervices nerve muscle preparation days 15-16 after treatment of hens with PSP (2.5 mg/kg im). The dosing regimen is given in the legend to Fig. 49.16A. Control (C) e_, PSP (P) = A-A, nifedipine plus PSP (NP) = v-v, verapamil plus PSP (VP) = 0-0. Reproduced with permission from El-Fawal, H. A., Jortner, B. S., and Ehrich, M. (1990). Modification of phenyl saligenin phosphate-induced delayed effects by calcium channel blockers: in vivo and in vitro electrophysiological assessment. NeuroToxicology 11, 573-592. Copyright © 1990 by Intox Press, Little Rock, AR.

Recently, it was discovered that administration of certain NTE inhibitors after neuropathy-inducing OP compounds will initiate or exacerbate clinical manifestations of OPIDN (Lotti et aI., 1991; Pope and Padilla, 1990). This phenomenon, called

1004

CHAPTER 49

Organophosphorus-Induced Delayed Neuropathy

Figure 49.17 Cross sections of distal levels of the biventer cervicis nerve from chickens dosed with 2.5 mg/kg im of phenyl saligenin phosphate 15 days earlier. The nerve in A is from a hen that only received the toxicant and shows extensive loss of myelinated fibers. The nerve in B was from a hen that had received the toxicant, plus the calcium channel blocker verapamil at 7 mg/kg/day for 4 days, beginning one day prior to the phenyl saligenin phosphate administration. Examination of this nerve (B) shows that the verapamil dosing was protective to the myelinated nerve fibers, many of which have a normal morphological appearance (arrow). Toluidine bluesafranin stain; bar = 100 }.tm.

promotion or potentiation, has subsequently been reproduced in other laboratories, with clinical manifestations and nervous system lesions included among the endpoints (Johnson and Read, 1993; Massicotte et aI., 1999; Pope et aI., 1992; Randall et aI., 1997; Richardson, 1995). With administration of NTE inhibitors after dosing with a neuropathy-inducing OP compound, OPIDN in hens has been exaggerated beyond what would be expected if the neuropathy-inducing OP compound were given alone. The exacerbation of OPIDN appears to be due to a quantitative rather than qualitative difference, as observed in hens given several different OP compounds (e.g., DFP, DBDCVP, PSP) and several different NTE inhibitors, with phenyl methanesulfonyl fluoride (PMSF) being used most often (Massicotte et aI., 1999; Moretto et aI., 1992a; Osman et aI., 1996; Peraica et aI., 1995; Randall et al., 1997). To date, all promotors of OPIDN are NTE inhibitors, yet most are those that do not lose a side group after attachment to the enzyme (in other words, the promotor-enzyme complex does not have to "age"). There have been reports recently of promotion occurring with a dose of an NTE inhibitor below that necessary for inhibition of the enzyme (Moretto et al., 1994; Osman et aI., 1996). During promotion, OPIDN can appear at subclinical doses of a neuropathy-inducing OP compound or in test subjects normally not susceptible to this condition (e.g., chicks, rats) (Harp et aI., 1997; Lotti et aI., 1993, 1995; Moretto et aI., 1992a,

1992b; Pope et aI., 1992, 1993, 1995). A recent report indicated that the only enzyme consistently inhibited by promoters was NTE, suggesting that some fraction or isoform of NTE may be the molecular target for promotion of OPIDN (Milatovic et aI., 1997). However, because NTE can be maximally inhibited without subsequent OPIDN by the OP compound first given, with administration of a second NTE inhibitor being followed by OPIDN, others have suggested that NTE is unlikely to be the target of OPIDN for promotion (Gardiman et aI. , 1999; Johnson, 1993; Lotti, 1992, 1995, 1997; Lotti and Moretto, 1999; Lotti et aI., 1993; Moretto et aI., 1994; Osman et al., 1996; Pope et aI., 1993). Factors other than NTE inhibition may be involved in OPIDN and promotion of OPIDN, because a recent study indicated that a soluble factor released in the spinal cord after exposure to a neuropathic OP compound had dramatic effects on cell growth (Pope et aI., 1995).

49.9 TESTING FOR ORGANOPHOSPHORUS-INDUCED DELAYED NEUROPATHY Registration of OP compounds for pesticide use under FIFRA recommends that they be tested in hens 8-14 months old without designation of breed or strain. Since 1991, this testing has

49.10 Summary

AChE and NTE activities

gc 0

U

'0 ::.!! 0

100 80

i

~. ,

~

~\

- - mouse AChE -.- mouse NTE

'1

\

\

60

b··o

40

.

20

\

\~.

11, ,

~

b

0 I

10"

i

10"

I

I

10-7

I

10-5

10""

I

10'"

-iJ-.

human AChE human NTE

- 0-

human AChE

.. 0··

\

\

.~

.... I

10.3

paraoxon (M)

(A)

AChE and NTE Activities 100

gc 0

....u0

'*

80 60 40 20

~ k

",.". >.... ........ ", ..... . .. " ~:.:.~\

~

~

"-

"- ....

"-

.

.:!

"-

"- ....

- - mouse AChE -iJ-. human NTE ..... mouse NTE

'0.. ,

'--;;

0

PSP (M)

(B) Figure 49.18 Concentration response curves for inhibition of acetylcholinesterase (AChE) and neuropathy target esterase (NTE) in neuroblastoma cells of human and murine origin by organophosphorus compounds. Cells were incubated with OP compounds for 1 h before assay. Paraoxon causes acute cholinergic crisis (AChE inhibition) rather than organophosphorus-induced delayed neuropathy (OPIDN); PSP causes OPIDN. Point-to-point composite curves are provided to aid visualization (Prism; GraphPad, San Diego). Each curve represents at least three different assays that included at least three concentrations of OP compounds that provided values between 10 and 90% of values in vehicle-treated cells. Reproduced with permission from Ehrich, M., Correll, L., and Veronesi, B. (1997). Acetylcholinesterase and neuropathy target esterase inhibitions in neuroblastoma cells to distinguish organophosphorus compounds causing acute and delayed neurotoxicity. Fundam. App!. Taxiea!. 38,55-63. Copyright © 1997 by Academic Press, San Diego.

included NTE and acetylcholinesterase determinations, clinical observations, and neuropathology following single- and multiple-dosing procedures (EPA, 1991). In the initial testing procedure, brain and spinal cord samples are collected from a subset of the dosed hens within 48 h of administration of a single dose of the test OP insecticide and NTE and acetylcholinesterase activities determined. The remaining hens are observed over the next 3 weeks, with in situ perfusionfixation and removal of brain, spinal cord, and peripheral nerves for histopathological examination at that time. Multipledose testing (28 days) may also be necessary. With these tests, the relative sensitivity of NTE to inhibition compared to acetylcholinesterase inhibition identifies those OP compounds

1005

capable of causing OPIDN even before clinical signs and morphological changes appear. Suggestions have been made that NTE measurements in cultured cells could be used to predict potential for OPIDN without the need to run this test in animals (Barber et aI., 1999a, 1999b; Ehrich and Veronesi, 1995, 1999; Ehrich et aI., 1994, 1997; Funk et aI., 1994b, 1994c; Knoth-Anderson and Abou-Donia, 1993; Knoth-Anderson et aI., 1992; Nostrandt and Enrich, 1992, 1993; Sogorb et aI., 1996, 1997; Veronesi, 1992; Veronesi and Ehrich, 1993; Veronesi et aI., 1997). Investigations indicated that NTE activity could be found in both primary cultures of avian and bovine origin and continuous cell lines of human and rodent origin (e.g., SH-SY5Y, PC-12, NB41A3). A recent, thorough concentration-response study with 11 active esterase inhibitors, including 7 that cause OPIDN and 4 that do not, indicated that either a human cell line or a murine cell line was capable of identifying the neuropathy-inducing OP compounds based on the relative sensitivity of NTE to inhibition compared to acetylcholinesterase (Ehrich et aI., 1997) (Fig. 49.18). Concentrations of OP compounds needed to inhibit NTE and acetylcholinesterase were far below those cytotoxic to the cultures. A similar result was noted in another study in which new, very sensitive NTE inhibitors were examined in cell lines of rodent origin (Li and Casida, 1997). Although it appeared that cell cultures did not have sufficient oxidative capability to convert protoxicant phosphorothioates to active enzyme inhibitors (Ehrich, 1995; Ehrich and Veronesi, 1995; Ehrich et aI., 1994, 1997), recent studies have indicated that this can be overcome by preincubation of OP protoxicants with a bromine solution or a microsomal preparation (Barber et aI., 1999a, 1999b). The results of recent studies enhance the possibility that OP compounds may one day be screened for potential to induce OPIDN by using an in vitro system.

49.10 SUMMARY Organophosphorus-induced delayed neuropathy (OPIDN) is a generally progressive, irreversible disorder that causes clinical manifestations appearing days to weeks after humans and certain species of animals are exposed to OP compounds that can essentially irreversibly inhibit most of the available neuropathy target esterase (NTE, neurotoxic esterase). The severity of OPIDN, as indicated by clinical and anatomical manifestations, depends on species and age of test animals and extent of NTE inhibition. Chickens have proven the most sensitive test species. OPIDN is manifest clinically by ataxia and weakness progressing to paralysis, associated with bilateral degeneration of distal and terminal regions of long myelinated nerve fibers. The neuropathy can be prevented by pretreatment with NTE inhibitors; yet these same compounds promote OPIDN when given after a neuropathy-inducing OP compound. Although the precise mechanism of OPIDN has not been determined, there appears to be a role for calcium, as calcium blockers ameliorated the effect, and changes on CaM kinase II activity and cytoskeletal protein phosphorylation appear after

1006

CHAPTER49


administration of neuropathy-inducing OP compounds. Recent studies indicate that NTE purification is imminent, and that neuropathy-inducing OP compounds and their effects on NTE can be studied in cultured cells.

REFERENCES Abou-Donia, M. B. (1981). Organophosphorus ester-induced delayed neurotoxicity. Annu. Rev. Pharmacol. Toxicol. 21,511-548. Abou-Donia, M. B. (1993). The cytoskeleton as a target for organophosphorus ester-induced delayed neurotoxicity (OPIDN). Chem.-Biol. Interact. 87 383-393. Abou-Donia, M. B. (1995). Organophosphorus pesticides. In "Handbook of Toxicology" (L. W. Chang and R. S. Dyer, eds.), pp. 419-473. Dekker, New York. Abou-Donia, M. B., and Graham, D. G. (1978). Delayed neurotoxicity from long-term low-level topical administration ofleptophos to the comb of hens. Toxicol. Appl. Pharmacol. 46, 199-213. Abou-Donia, M. B., and Lapadula, D. M. (1990). Mechanisms of organophosphorus ester-induced delayed neurotoxicity: Type I and type 11. Annu. Rev. Pharmacol. Toxicol. 30,405-440. Abou-Donia, M. B., and Preissig, S. H. (1976a). Delayed neurotoxicity of leptophos: Toxic effects on the nervous system of hens. Toxicol. Appl. Pharmacol. 35, 269-282. Abou-Donia, M. B., and Preissig, S. H. (1976b). Delayed neurotoxicity from continuous low-dose oral administration ofleptophos to hens. Toxicol. Appl. Pharmacol. 38, 595--608. Abou-Donia, M. B., Graham, D. G., and Komeil, A. A. (1979). Delayed neurotoxicity of O-ethyl 0-2,4-dichlorophenyl phenylphosphonothioate: Effects of a single oral dose on hens. Toxicol. Appl. Pharmacol. 49,293-303. Abou-Donia, M. B., Trofatter, L. P., Graham, D. G., and Lapadula, D. M. (1986). Electromyographic, neuropathologic, and functional correlates in the cat as the result of tri-o-cresyl phosphate delayed neurotoxicity. Toxicol. Appl. Pharmacol. 83, 126-141. Abou-Donia, M. B., Viana, M. E., Gupta, R. P., and Anderson, J. K. (1993). Enhanced calmodulin binding concurrent with increased kinase-dependent phosphorylation of cytoskeletal proteins following a single subcutaneous injection of diisopropyl phosphorofluoridate in hens. Neurochem. Int. 22, 165-173. Abou-Donia, M. B., Wilmarth, K. R., Abdel-Rahman, A. A., Jensen, K. F., Oehme, F. W., and Kurt, T. L. (1996). Increased neurotoxicity following concurrent exposure to pyridostigmine bromide, DEET, and chlorpyrifos. Fundam. Appl. Toxicol. 34,201-222. Anderson, R. J., Robertson, D. G., Henderson, J. D., and Wilson, B. W. (1988). DFP-induced elevation of strength-duration threshold in hen peripheral nerve. NeuroToxicology 9, 47-52. Baker, T., and Stanec, A. (1985). Methylprednisolone treatment of an organophosphorus-induced delayed neuropathy. Toxicol. Appl. Pharmacol. 79, 348-352. Baker, T., Drakontides, A. B., and Riker, W. F. (1982). Prevention of the organophosphorus neuropathy by glucocorticoids. Exp. Neurol. 78, 397-408. Baker, T., Lowndes, H. E., Johnson, M. K., and Sandborg, I. C. (1980). The effects of phenylmethanesulfonyl fluoride on delayed organophosphorus neuropathy. Arch. Toxicol. 46,305-311. Barber, D., Correll, L, and Ehrich, M. (1999a). Comparison of two in vitro activation systems for protoxicant organophosphorous esterase inhibitors. Toxicol. Sci. 47, 16-22. Barber, D., Correll, L, and Ehrich, M. (1999b). Comparative effectiveness of organophosphorous protoxicant activating systems in neuroblastoma cells and brain homogenates. 1. Toxicol. Environ. Health 56, 101-112. Barril, J. B., Vilanova, E., and Pellin, M. (1988). Sciatic nerve neuropathy target esterase: Methods of assay, proximo-distal distribution and regeneration. Toxicology 49, 107-114. Beck, B. E., Wood, C. D., and Whanham, G. R. (1977), Triaryl phosphate poisoning in cattle. Vet. Pathol. 14, 128-137.

Bertoncin, D., Russolo, A., Caroldi, S., and Lotti, M. (1985). Neuropathy target esterase in human lymphocytes. Arch. Environ. Health 40, 139-144. Bischoff, A. (1967). The ultrastructure of tri-ortho-cresyl phosphate poisoning. I. Studies on myelin and axonal alterations in the sciatic nerve. Acta Neuropathol. 9, 158-174. Bischoff, A. (1970). Ultrastructure of tri-ortho-cresyl phosphate poisoning in the chicken. 11. Studies on spinal cord alterations. Acta Neuropathol. 15, 142-155. Borhan, B., Ko, Y., Mackay, C., Wilson, B. w., Kurth, M. J., and Hammock, B. D. (1995). Development of surrogate substrates for neuropathy target esterase. Biochim. Biophys. Acta 1250, 171-182. Bouldin, T. W., and Cavanagh, J. B. (1979a). Organophosphorus neuropathy. I. A teased-fiber study of the spatio-temporal spread of axonal degeneration. Am. J. Pathol. 94, 241-252. Bouldin, T. W., and Cavanagh, J. B. (1979b). Organophosphorus neuropathy. 11. A fine-structural study of the early stages of axonal degeneration. Am. J. Patho!. 94,253-270. Brailowsky, S. (1988). Therapeutic approaches in subjects with brain lesions. In "Pharmacological Approaches to the Treatment of Brain and Spinal Cord Injury" (D. G. Stein and B. A. Sabel, eds.), pp. 1-21. Plenum, New York. Bursian, S. J., Lehning. E. J., Correll, L., and Ehrich, M. (1989). Effect of betanaphthoflavone on o-tolyl saligenin phosphate-induced delayed neuropathy in two lines of chickens. 1. Toxicol. Environ. Health 28,461-471. Bush, D. M., Lehning, E. J., and Bursian, S. J. (1995). The effects of diisopropylphosphorofluoridate (DFP) on the ganglioside profile in the chicken (Gallus domesticus) hindbrain. NeuroToxicology 16,55--61. Capildeo, R. (1989). "Steroids in Diseases of the Central Nervous System." Wiley, New York. Carboni, D., Ehrich, M., Dyer, K., and Jortner, B. S. (1992). Comparative evolution of mipafox-induced delayed neuropathy in rats and hens. NeuroToxicology 13,723-733. Caroldi, S., and Lotti, M. (1982). Neurotoxic esterase in peripheral nerve: Assay, inhibition, and rate of resynthesis. Toxicol. App!. Pharmacol. 62, 498-501. Carrera, v., Barril, J., Mauricio, M., Pellin, M., and Vilanova, E. (1992). Local application of neuropathic organophosphorus compounds to hen sciatic nerve: Inhibition of neuropathy target esterase and peripheral neurological impairments. Toxicol. Appl. Pharmacol. 117,218-225. Carrera, v., Diaz-Alejo, N., Sogorb, M. A, Vicedo, J. L., and Vilanova, E. (1994). In vivo inhibition by mipafox of soluble and particulate forms of organophosphorus neuropathy target esterase (NTE) in hen sciatic nerve. Toxico!. Lett. 71,47-51. Carrington, C. D. (1989). Prophylaxis and the mechanism for the initiation of organophosphorus compound-induced delayed neurotoxicity. Arch. Toxicol. 63, 165-172. Carrington, C. D., and Abou-Donia, M. B. (1985a). Axoplasmic transport and turnaround of neurotoxic esterase in hen sciatic nerve. J. Neurochem. 44, 616--621. Carrington, C. D., and Abou-Donia, M. B. (1985b). Characterization of [3Hldiisopropyl phosphorofluoridate-binding proteins in hen brain: Rates of phosphorylation and sensitivity to neurotoxic and non-neurotoxic organophosphorus compounds. Biochem. 1. 228, 537-544. Carrington, C. D., and Abou-Donia, M. B. (1986). Kinetics of substrate hydrolysis and inhibition by mipafox of paraoxon-preinhibited hen brain esterase activity. Biochem. J. 236,503-507. Carrington, C. D., Brown, H. R., and Abou-Donia, M. B. (1988). Histopathological assessment of triphenyl phosphite neurotoxicity in the hen. NeuroToxicology 9, 223-233. Carrington, C. D., Lapadula, D. M., and Abou-Donia, M. B. (1989). Acceleration of anterograde axonal transport in cat sciatic nerve by diisopropyl phosphorofluoridate. Brain Res. 476, 179-182. Cavalleri, A., and Cosi, V. (1980). Polyneuropathy in shoe factory workers. In "Advances in Neurotoxicology" (L. Manzo, ed.), pp. 193-200. Pergamon, New York. Cavanagh, J. B. (1954). The toxic effects of tri-ortho-cresyl phosphate on the nervous system: An experimental study in hens. 1. Neurol. Neurosurg. Psych. 17, 163-172.

References

Cavanagh, J. B. (1964). Peripheral nerve changes in ortho-cresyl phosphate poisoning in the cat. J. Pathol. Bacteriol. 87,365-383. Cavanagh, J. B., and Patangia, G. N. (1965). Changes in the central nervous system in the cat as the result of tri-o-cresyl phosphate poisoning. Brain 88, 165-180. Chemnitius, J. M., Haselmeyer, K. H., and Zech, R. (1983). Neurotoxic esterase, identification of two isoenzymes in hen brain. Arch. Toxicol. 53, 235-244. Chemnitius, J. M., Holling, M., Meyer, J. H., Schmidt, P. E, Schomburg, E. D., Steffens, H., and Zech, R. (1988), Influence of the organophosphorus compound DFP on inhibitory motor systems and esterase activity in the spinal cord of cats. Neurosci. Res. 6, 257-263. Chemiack, M. G. (1986). Organophosphorus esters and polyneuropathy. Ann. Int. Med. 104, 264-266. Chemiack, M. G. (1988). Toxicological screening for organophosphorusinduced delayed neurotoxicity: Complications in toxicity testing. NeuroToxicology 9, 249-272. Classen, w., Gretener, P., Rauch, M., Weber, E., and Krinke, G. J. (1996). Susceptibility of various areas of the nervous system of hens to TOCP-induced delayed neuropathy. NeuroToxicology 17,597--604. Clothier, B., and Johnson, M. K. (1979). Rapid aging of neurotoxic esterase after inhibition by di-isopropyl phosphorofluoridate. Biochem. 1. 177,549558. Correll, L., and Ehrich, M. (1991). A microassay method for neurotoxic esterase determinations. Fundam. Appl. Toxicol. 16, 110-116. Cranmer, J. M., and Hixson, E. J. (1984). "Delayed Neurotoxicity." Intox Press, Little Rock, AR. Davis, C. S., and Richardson, R. J. (1980). Organophosphorus compounds. In "Experimental and Clinical Neurotoxicology" (P. S. Spencer and H. H. Schaumburg, eds.), pp. 527-544. Williams and Wilkins, Baltimore. Davis, C. S., and Richardson, R. J. (1987). Neurotoxic esterase: Characterization of the solubilized enzyme and the conditions for its solubilization from chicken brain microsomal membranes with ionic, zwitterionic, or nonionic detergents. Biochem. Phamacol. 36, 1393-1399. Davis, C. S., Barth, M. L., Dudek, B. R., and Richardson, R. J. (1980). Inhibitor characteristics of native, solubilized, and lipid-reconstituted neurotoxic esterase. Dev. Toxicol. Environ. Sci. 8, 63--66. Drakontides, A B., and Baker, T. (1983). An electrophysiologic and ultrastructural study of the phenylmethanesulfonyl fluoride protection against a delayed organophosphorus neuropathy. Taxicol. Appl. Pharmacol. 70,411422. Drakontides, A B., Baker, w., and Riker, W. E (1982). A morphological study of the effect of glucocorticoid treatment on delayed organophosphorus neuropathy. NeuroToxicology 3, 165-178. Dretchen, K. L., Bowles, A M., and Raines, A. (1986). Protection by phenytoin and calcium channel blocking agents against the toxicity of diisopropylfluorophosphate. Toxicol. Appl. Pharmacol. 83, 584-589. "Drug Facts and Comparisons" (2001). Facts and Comparisons, St. Louis. Dudek, B. R., and Richardson, R. J. (1982). Evidence for the existence of neurotoxic esterase in neural and lymphatic tissue of the adult hen. Biochem. Pharmacal. 31, 1117-1121. Dunnington, E. A., Siegel, P. B., and Ehrich, M. (1989). Differences in response of chickens from two genetic lines to diisopropyl phosphorofluoridate. NeuroToxicology 10,71-78. Durham, H. D., and Ecobichon, D. J. (1984). The function of motor nerves innervating slow tonic skeletal muscle in hens with delayed neuropathy induced by tri-o-tolyl phosphate. Can. J. Physiol. Pharmacol. 62, 1268-1273. Dyer, K., Ehrich, M., and Jortner, B. S. (1996). Neuropathology of organophosphorus ester induced delayed neuropathy (OPIDN) in hens, lesions of spinal cord gray matter revisited. Fundam. Appl. Toxicol. 308, 300. Dyer, K. R., EI-Fawal, H. A, and Ehrich, M. E (1991). Comparison of organophosphate-induced delayed neuropathy between branches of the tibial nerve and the biventer cervicis nerve in chickens. NeuroToxicology 12, 687--695. Dyer, K. R., Jortner, B. S., Shell, L. G., and Ehrich, M. (1992). Comparative dose-response studies of organophosphorus ester-induced delayed

1007

neuropathy in rats and hens administered mipafox. NeuroToxicology 13, 745-755. Ecobichon, D. J. (1994). Organophosphorus insecticides. In "Pesticides and Neurological Diseases" (D. J. Ecobichon and R. M. Joy, eds.), pp. 171-249. CRC Press, Boca Raton, FL. Ehrich, M. (1995). Using neuroblastoma cell lines to address differential specificity to organophosphates. Clin. Exp. Pharmacol. Physiol. 22,291-292. Ehrich, M. (1996). Neurotoxic esterase inhibition: Predictor of potential for organophosphorus-induced delayed neuropathy. In "Biomarkers for Agrochemicals and Toxic Substances" (1. N. Blancato, R. N. Brown, C. C. Dary, and M. A. Saleh, eds.), pp. 79-93. Am. Chem. Soc., Washington, DC. Ehrich, M., and Gross, W. B. (1982). Effect of corticosterone on toxicity of malathion and TOTP in hens. Neurobehav. Toxicol. Teratol. 4, 789-792. Ehrich, M., and Gross, W. B. (1983). Modification of tri-ortho-tolyl phosphate toxicity in chickens by stress. Toxicol. Appl. Pharmacol. 70, 249-254. Ehrich, M., and Gross, W. B. (1986). Effect of supplemental corticosterone and social stress on organophosphorus-induced delayed neuropathy in chickens. Toxical. Lett. 31, 9-13. Ehrich, M., and Veronesi, B. (1995). Esterase comparison in neuroblastoma cells of human and rodent origin. Clin. Exp. Pharmacol. Physiol. 22, 385386. Ehrich, M., and Veronesi, B. (1999). In vitro neurotoxicology. In "Neurotoxicology" (H. A. Tilson and G. J. Harry, eds.), 2nd ed., pp. 37-51. Taylor & Francis, Philadelphia. Ehrich, M., Brites, R. w., Briles, W. E., Dunnington, E. A, Martin, A, Siege!, P. B., and Gross, W. B. (1986a). Neurotoxicity of tri-ortho-tolyl phosphate in chickens of different genotypes in the presence and absence of deoxycorticosterone. Poult. Sci. 65, 375-379. Ehrich, M., Correll, L., and Veronesi, B. (1994). Neuropathy target esterase inhibition by organophosphorus esters in human neuroblastoma cells. NeuroToxicology 15, 309-313. Ehrich, M., Correll, L., and Veronesi, B. (1997). Acetylcholinesterase and neuropathy target esterase inhibitions in neuroblastoma cells to distinguish organophosphorus compounds causing acute and delayed neurotoxicity. Fundam. Appl. Toxicol. 38, 55--63. Ehrich, M., Jortner, B. S., and Gross, W. B. (1985). Absence of a protective effect of corticosterone on O-O-diisopropyl phosphorofluoridate (DFP) induced delayed neurotoxicity in chickens. NeuroToxicology 6, 87-92. Ehrich, M., Jortner, B. S., and Gross, W. B. (1986b). Dose-related beneficial and adverse effects of dietary corticosterone on organophosphorus-induced delayed neuropathy in chickens. Toxicol. Appl. Pharmacal. 83, 250-260. Ehrich, M., Jortner, B. S., and Gross, W. B. (1988). Types of adrenocorticoids and their effect on organophosphorus-induced delayed neuropathy in chickens. Taxicol. Appl. Pharmacol. 92,214-223. Ehrich, M., Jortner, B. S., and Padilla, S. (1993a). Relationship of neuropathy target esterase inhibition to neuropathology and ataxia in hens given organophosphorus esters. Chem.-Biol. Interact. 87,431-437. Ehrich, M., Jortner, B. S., and Padilla, S. (1995). Comparison of the relative inhibition of acetylcholinesterase and neuropathy target esterase in rats and hens given cholinesterase inhibitors. Fundam. Appl. Toxicol. 24,94-101. Ehrich, M., Shell, L., Rozum, M., and Jortner, B. S. (1993b). Short-term clinical and neuropathologic effects of cholinesterase inhibitors in rats. J. Am. Coli. Toxical. 12, 55--68. Eisenbrandt, D. L., Mattsson, J. L., Albee, R. R., Spencer, P. J., and Johnson, K. A. (1990). Spontaneous lesions in subchronic neurotoxicity testing of rats. Toxicol. Pathol. 18, 154-164. EI-Fawal, H. A., and Ehrich, M. E (1993). Calpain activity in organophosphorus-induced delayed neuropathy (OPIDN): Effects of a phenylalkylamine calcium channel blocker. Annu. N. Y. Acad. Sci. 679, 325-329. EI-Fawal, H. A., Correll, L., Gay, L., and Ehrich, M. (1990a). Protease activity in brain, nerve, and muscle of hens given neuropathy-inducing organophosphates and a calcium channel blocker. Toxicol. Appl. Pharmacal. 103, 133-142. EI-Fawal, H. A., Jortner, B. S., and Ehrich, M. (1989). Effect of verapamil on organophosphorus-induced delayed neuropathy in hens. Toxicol. Appl. Pharmacol. 97,500-511.

1008

CHAPTER 49


EI-Fawal, H. A, Jortner, B. S., and Ehrich, M. (1990b). Modification of phenyl saligenin phosphate-induced delayed effects by calcium channel blockers: in vivo and in vitro electrophysiological assessment. NeuroToxicology 11, 573-592. EI-Fawal, H. A., Jortner, B. S., and Ehrich, M. (1990c). Use of the biventer cervicis nerve-muscle preparation to detect early changes following exposure to organophosphates inducing delayed neuropathy. Fundam. Appl. Toxicol. 15, 108-120. EI-Fawal, H. A, Jortner, B. S., Eyre, P., and Ehrich, M. (1988). The biventer cervicis nerve-muscle preparation of adult hens: Effects of phenyl saligenin phosphate administration. NeuroToxicology 9, 625-636. EI-Sebae, A. H., Soliman, S. A., Abo EI-Amayen, M., and Ahmed, N. S. (1977). Neurotoxicity of organophosphorus insecticides leptophos and EPN. 1. Environ. Sci. Healt, B 12, 269-288. Environmental Protection Agency (EPA) (1991). "Pesticide Assessment Guidelines, Subdivision E. Hazard Evaluation: Human and Domestic Animals. Addendum 10: Neurotoxicity, Series 81,82, and 83." National Technical Information Service, Springfield, VA. Escudero, M. A, Cespedes, M. v., and Vilanova, E. (1997). Chromatographic discrimination of soluble neuropathy target esterase isoenzymes and related phenyl valerate esterases from chicken brain, spinal cord, and sciatic nerve. J. Neurochem. 68,2170-2176. Funk, K. A, Henderson, J. D., Liu, C. H., Higgins, R. J., and Wilson, B. W. (1994a). Neuropathology of organophosphate-induced delayed neuropathy (OPIDN) in young chicks. Arch. Toxicol. 68, 308-316. Funk, K. A, Liu, C. H., Higgins, R J., and Wilson, B. W. (1994b). Avian embryonic brain reaggregate culture system. Il. NTE activity discriminates between effects of a single neuropathic or nonneuropathic organophosphorus compound exposure. Toxicol. Appl. Pharmacol. 124,159-163. Funk, K. A, Liu, C. H., Wilson, B. w., and Higgins, R J. (1994c). Avian embryonic brain reaggregate culture system. I. Characterization for organophosphorus compound toxicity studies. Toxicol. Appl. Pharmacol. 124, 149-158. Gallo, M. A., and Lawryk, N. J. (1991). Organic phosphorus pesticides. In "Handbook of Pesticide Toxicology" (W. J. Hayes and E. R J. Laws, eds.), Vol. 2, pp. 917-1123. Academic Press, San Diego. Gardiman, G., Moretto, A, and Lotti, M. (1999). Influence of dithiocarbamates on the development of organophosphate induced delayed polyneuropathy (OPIDP). Toxicol. Sci. 488, 100. Glazer, E. J., Baker, T., and Riker, W. F., Jr. (1978). The neuropathology ofDFP at cat soleus neuromuscular junction. 1. Neurocytol. 7,741-758. Glynn, P., Hotton, J. L., Nolan, C. c., Read, D. J., Brown, L., Hubbard, A., and Cavanagh, J. B. (1998). Neuropathy target esterase: Immunolocalization to neuronal cell bodies and axons. Neuroscience 83, 295-302. Glynn, P., Read, D. J., Quo, R, Wylie, S., and Johnson, M. K. (1994). Synthesis and characterization of a biotinylated organophosphorus ester for detection and affinity purification of a brain serine esterase: Neuropathy target esterase. Biochem. J. 301, 551-556. Glynn, P., Read, D. J., Lush, M. J., Li, Y., and Atkins, J. (1999). Molecular cloning of neuropathy target esterase (NTE). Chem.-Biol. Interac. 119-120, 513-518. Glynn, P., Ruffer-Tumer, M., Read, D., Wylie, S., and Johnson, M. K. (1993). Molecular characterisation of neuropathy target esterase: Proteolysis of the [3HJDFP-Iabelled polypeptide. Chem.-Biol. Interact. 87,361-367. Goldstein, D. A., McGuigan, M. A., and Ripley, B. D. (1988). Acute tricresylphosphate intoxication in childhood. Hum. Toxieol. 7, 179-182. Gupta, RP., and Abou-Donia, M. B. (1993). Comparison ofCa2+/calmodulindependent protein kinase Il purified from control and diisopropyl phosphorofluoridate (DFP)-treated hens. Neuroehem. Res. 18,259-269. Gupta, RP., and Abou-Donia, M. B. (1994). In vivo and in vitro effects of diisopropyl phosphorofluoridate (DFP) on the rate of hen brain tubulin polymerization. Neuroehem. Res. 19,435-444. Gupta, R. P., and Abou-Donia, M. B. (1995a). Diisopropyl phosphorofluoridate (DFP) treatment alters calcium-activated proteinase activity and cytoskeletal proteins of the hen sciatic nerve. Brain Res. 677, 162-166. Gupta, RP., and Abou-Donia, M. B. (1995b). Neurofilament phosphorylation and [125IJcalmodulin binding by Ca2+ jcalmodulin-dependent protein ki-

nase in the brain subcellular fractions of diisopropyl phosphorofluoridate (DFP)-treated hen. Neuroehem. Res. 20, 1095-1105. Gupta, RP., Abdel-Rahman, A., Wilmarth, K. w., and Abou-Donia, M. B. (1997). Alteration in neurofilament axonal transport in the sciatic nerve of the diisopropyl phosphorofluoridate (DFP)-treated hen. Bioehem. Pharmacol. 53, 1799-1806. Gupta, RP., Lapadula, D. M., and Abou-Donia, M. B. (1992). Ca2+ j calmodulin-dependent protein kinase Il from hen brain: Purification and characterization. Bioehem. Pharmaeol. 43, 1975-1988. Haley, R W., and Kurt, T. L. (1997). Self-reported exposure to neurotoxic chemical combinations in the Gulf War. 1. Am. Med. Assoe. 277, 231-237. Haley, R w., Horn, J., Roland, P. S., Bryan, W. w., Van Ness, P. C., Bonte, F. J., Devous, M. D., Mathews, D., Fieckenstein, J. L., Wians, F. H., Wolfe, G. I., and Kurt, T. L. (1997). Evaluation of ncurologic function in Gulf War veterans.1. Am. Med. Assoe. 277, 223-230. Harp, P., Tanaka, D., Jr., and Pope, C. N. (1997). Potentiation of organophosphorus-induced delayed neurotoxicity following phenyl saligenin phosphate exposures in 2-, 5-, and 8-week-old chickens. Fundam. Appl. Toxieol. 37,64-70. Hierons, R., and Johnson, M. K. (1978). Clinical and toxicological investigations of a case of delayed neuropathy in man after acute poisoning by an organophosphorus pesticide. Arch. Toxieol. 40, 279-284. Hollingshaus, J. G. (1983). Chemistry and metabolism of delayed neurotoxic organophosphorus esters. In "14th Conference on Environmental Toxicology," pp. 76-105. National Technical Information Service, Springfield, VA. Iliis, L., Pantangia, G. N., and Cavanagh, J. B. (1966). Boutons terminaux and tri-ortho-cresyl phosphate neurotoxicity. Exp. Neural. 14, 160-174. Inui, K., Mitsumori, K., Harada, T., and Maita, K. (1993). Quantitative analysis of a neuronal damage induced by tri-ortho-cresyl phosphate in Wistar rats. Fundam. Appl. Toxieol. 20, 111-119. Ishakawa, Y., Chow, E., McNamee, M. G., McChesney, M., and Wilson, B. W. (1983). Separation of paraoxon and mipafox sensitive esterases by sucrose density gradient sedimentation. Toxieol. Lett. 17,315-320. !toh, H., Kishida, H., Tadokoro, M., and Oikawa, K. (1984). Studies on the delayed neurotoxicity of organophosphorus compounds. Il. 1. Toxieol. Sci. 9,37-50. !toh, H., Kishida, H., Takeuchi, E., Tadokoro, M., Uchikoshi, T., and Oikawa, K. (1985). Studies on the delayed neurotoxicity of organophosphorus compounds. Ill. 1. Toxieo!. Sei. 10, 67-82. Jensen, K. F., Lapadula, D. M., Anderson, J. K., Haykal-Coates, N., and AbouDonia, M. B. (1992). Anomalous phosphorylated neurofilament aggregations in central and peripheral axons of hens treated with tri-ortho-cresyl phosphate (TOCP). J. Neurosei. Res. 33, 455-460. Johnson, M. K. (1975a). The delayed neuropathy caused by some organophosphorus esters: Mechanism and challenge. Crit. Rev. Toxieol. 3, 289-316. Johnson, M. K. (1975b). Structure-activity relationships for substrates and inhibitors of hen brain neurotoxic esterase. Bioehem. Pharmaeol. 24, 797805. Johnson, M. K. (1980). Irreversible phosphorylation of brain neurotoxic esterase: The primary event leading to the delayed neuropathy caused by some organophosphorus esters. Mon. Neural Sci. 7, 99-105. Johnson, M. K. (1982). The target for initiation of delayed neurotoxicity by organophosphorus esters: Biochemical studies and toxicological application. In "Reviews in Biochemical Toxicology" (E. Hodgson, J. R Bond, and R. M. Philpot, eds.), Vol. 4, pp. 141-212. Elsevier Biomedical, New York. Johnson, M. K. (1992). Molecular events in delayed neuropathy: Experimental aspects of neuropathy target esterase. In "Clinical and Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 90-113. Butterworth-Heinemann, Oxford. Johnson, M. K. (1993). Symposium introduction: Retrospect and prospects for neuropathy target esterase (NTE) and the delayed polyneuropathy (OPIDP) induced by some organophosphorus esters. Chem.-Biol. Interact. 87,339346. Johnson, M. K., and Glynn, P. (1995). Neuropathy target esterase (NTE) and organophosphorus-induced delayed polyneuropathy (OPIDP): Recent advances. Toxieol. Lett. 82-83, 459-463.

References

Johnson, M. K., and Read, D. J. (1987). The influence of chirality on the delayed neuropathic potential of some organophosphorus esters: Neuropathic and prophylactic effects of stereoisomeric esters of ethyl phenylphosphonic acid (EPN oxon and EPN) correlate with quantities of aged and unaged neuropathy target esterase in viva. Taxical. Appl. Pharmacal. 90, 103-115. Johnson, M. K, and Read, D. J. (1993). Prophylaxis against and promotion of organophosphate-induced delayed neuropathy by phenyl di-npentylphosphinate. Chem.-Bial. Interact. 87,449-455. Johnson, M. K., and Safi, J. M. (1993). The R-(+)isomer of O-n-hexyl S-methyl phosphorothioamidate causes delayed neuropathy in hens after generation of a form of inhibited neuropathy target esterase (NTE) which can be reactivated ex viva. Chem.-BiaI. Interact. 87,443-448. Johnson, M. K, Vilanova, E., and Read, D. J. (1991). Anomalous biochemical responses in tests of the delayed neuropathic potential of methamidophos (O,S-dimethyl phosphorothioamidate), its resolved isomers and of some higher O-alkyl homologues. Arch. Taxical. 65,618-624. Jortner, B. S. (1984). Pathology of organophosphorus-induced delayed neurotoxicity. In "14th Conference on Environmental Toxicology," pp. 106-117. National Technical Information Service, Springfield, VA. Jortner, B. S. (1988). Organophosphorus ester-induced delayed neuropathy, rat. In "Monographs on Pathology of Laboratory Animals-Nervous System" (T. C. Jones, U. Mohr, and R. D. Hunt, eds.), pp. 41-47. Springer-VerJag, New York. Jortner, B. S., and Ehrich, M. (1987). Neuropathological effects of phenyl saligenin phosphate in chickens. NeuraTaxicalagy 8, 303-314. Jortner, B. S., Perkins, S. K., and Ehrich, M. (1999). Immunohistochemical study of phosphorylated neurofiloments during the evolution of organophosphorus ester-induced delayed neuropathy (OPIDN). NeuraTaxicalagy 20, 971-976. Jortner, B. S., Pope, A M., and Heavner, J. E. (1983). Haloxon-induced delayed neurotoxicity: Effect of plasma A (aryl) esterase activity on severity oflesions in sheep. NeuroTaxicalagy 4, 241-246. Jortner, B. S., Shell, L., El-Fawal, H., and Ehrich, M. (1989). Myelinated nerve fiber regeneration following organophosphorus ester-induced delayed neuropathy. NeuraTaxicalagy 10, 717-726. Kidd, J. G., and Langworthy, O. R. (1933). Jake paralysis: Paralysis following the ingestion of Jamaica ginger extract adulterated with tri-artha-cresyl phosphate. Bull. lahns Hapkins Hasp. 52, 39-65. Knoth-Anderson, J., and Abou-Donia, M. B. (1993). Differential effects of triphenylphosphite and diisopropyl phosphorofluoridate on catecholamine secretion from bovine adrenomedullary chromaffin cells. 1. Taxicol. Enviran. Health 38, 103-114. Knoth-Anderson, J., Veronesi, B., Jones, K., Lapadula, D. M., and AbouDonia, M. B. (1992). Triphenyl phosphite-induced ultrastructural changes in bovine adrenomedullary chromaffin cells. Taxical. Appl. Pharmacal. 112, 110-119. Krinke, G., Ullmann, L., Sachsee, K., and Hess, R. (1979). Differential susceptibility of peripheral nerves of the hen to tri-artha-cresyl phosphate and to trauma. Agents Actians 9, 227-231. Lapadula, D. M., Kinnes, C. G., Somjen, G. G., and Abou-Donia, M. B. (1982). Monosynaptic reflex depression in cats with organophosphorus neuropathy: Effects of tri-O-cresyl phosphate. NeuraTaxicalagy 3, 51-61. Lapadula, D. M., Patton, S. E., Campbell, G. A., and Abou-Donia, M. B. (1985). Characterization of delayed neurotoxicity in the mouse following chronic oral administration of tri-a-cresyl phosphate. Taxical. Appl. Pharmacal. 79, 83-90. Lapadula, E. S., Lapadula, D. M., and Abou-Donia, M. B. (1991). Persistent alterations of calmodulin kinase II activity in chickens after an oral dose of tri-a-cresyl phosphate. Biachem. Pharmacal. 42, 171-180. Lapadula, E. S., Lapadula, D. M., and Abou-Donia, M. B. (1992). Biochemical changes in sciatic nerve of hens treated with tri-a-cresyl phosphate: Increased phosphorylation of cytoskeletal proteins. Neurochem. Int. 20, 247-255. Larsen, C., Jortner, B. S., and Ehrich, M. (1986). Effect of neurotoxic organophosphorus compounds in turkeys. 1. Taxical. Enviran. Health 17, 365-374.

1009

Lehning, E. J., Tanaka, D., Jr., and Bursian, S. J. (1996). Triphenyl phosphite and diisopropylphosphorofluoridate produce separate and distinct axonal degeneration patterns in the central nervous system of the rat. Fundam. Appl. Taxical. 29, 110-118. LeVay, S., Meier, c., and Glees, P. (1971). Effects of tri-artha-cresyl-phosphate on spinal ganglia and nerves of chickens. Acta Neurapathal. 17, 103-113. Li, and Casida, J. E. (1997). Actions of two highly potent organophosphorus neuropathy target esterase inhibitors in mammalian cell lines. Taxical. Let!. 92, 123-130. Lidsky, T. 1., Manetto, C., and Ehrich, M. (1990). Nerve conduction studies in chickens given phenyl saligenin phosphate and corticosterone. 1. Taxical. Enviran. Health 29, 65-75. Lotti, M. (1992). The pathogenesis of organophosphate polyneuropathy. Cri!. Rev. Taxical. 21, 465-488. Lotti, M. (1995). A key step forward in understanding the pathogenesis of organophosphate polyneuropathy. Hum. Exp. Taxical. 14,69-70. Lotti, M. (1997). The concept and target of promotion ofaxonopathies. Arch. Taxical. Suppl. 19,331-336. Lotti, M., and Moretto, A. (1993). The search for the physiological functions of NTE: Is NTE a receptor? Chem.-Bial. Interact. 87,407-416. Lotti, M., and Moretto, A. (1999). Promotion of organophosphate induced polyneuropathy by certain esterase inhibitors. Chem.-Bial. Interact. 119120,513-518. Lotti, M., Caroldi, S., Capodicasa, E., and Moretto, A. (1991). Promotion of organophosphate-induced delayed polyneuropathy by phenylmethanesulfonyl fluoride. Taxical. Appl. Pharmacal. 108,234-241. Lotti, M., Caroldi, S., Moretto, A., Johnson, M. K, Fish, C. J., Gopinath, c., and Roberts, N. L. (1987). Central-peripheral delayed neuropathy caused by diisopropyl phosphorofluoridate (DFP): Segregation of peripheral nerve and spinal cord effects using biochemical, clinical, and morphological criteria. Taxical. Appl. Pharmacal. 88,87-96. Lotti, M., Moretto, A., Bertolazzi, M., Peraica, M., and Fioroni, F. (1995). Organophosphate polyneuropathy and neuropathy target esterase: Studies with methamidophos and its resolved optical isomers. Arch. Taxical. 69, 330-336. Lotti, M., Moretto, A, Capodicasa, E., Bertolazzi, M., Peraica, M., and Scapellato, M. L. (1993). Interactions between neuropathy target esterase and its inhibitors and the development of polyneuropathy. Taxical. Appl. Pharmacal. 122, 165-171. Lotti, M., Moretto, A., Zoppellari, R., Dainese, R., Rizzuto, N., and Barusco, G. (1986). Inhibition oflymphocytic neuropathy target esterase predicts the development of organophosphate-induced delayed polyneuropathy. Arch. Taxical. 59, 176-179. Lush, M. J., Li, Y., Read, D. J., Willis, A. c., and Glynn, P. (1998). Neuropathy target esterase and a homologous Drasaphila neurodegeneration-associated mutant protein contain a novel domain conserved from bacteria to man. Biachem. 1. 332,1-4. Mackay, C. E., Hammock, B. D., and Wilson, B. W. (1996). Identification and isolation of a 155-kDa protein with neuropathy target esterase activity. Fundam. Appl. Taxical. 30,23-30. Maroni, M., and Bleecker, M. L. (1986). Neuropathy target esterase in human lymphocytes and platelets. 1. Appl. Taxical. 6, 1-7. Massicotte, C., Dyer Inzana, K, Ehrich, M., and Jortner, B. S. (1999). Neuropathologic effects of phenylmethylsulfonyl fluoride (PMSF)-induced promotion and protection in organophosphorus ester-induced delayed neuropathy (OPIDN) in hens. NeuraTaxicolagy, 20, 749-760. McCain, W. c., Flaherty, D. M., Correll, L., Jortner, B., and Ehrich, M. (1996). Catecholamine concentrations and contractile responses of isolated vessels from hens treated with cyclic phenyl saligenin phosphate or paraoxon in the presence or absence of verapamil. 1. Taxicol. Environ. Health 48, 397-411. McCain, W. c., Wilcke, J., Lee, J. c., and Ehrich, M. (1995). Effect of cyclic phenyl saligenin phosphate and paraoxon treatment on vascular response to adrenergic and cholinergic agents in hens. 1. Taxical. Environ. Health 44, 167-187. Metcalf, R. L. (1984). Historical perspective of organophosphorus esterinduced delayed neurotoxicity. In "Delayed Neurotoxicity" (J. M. Cranmer and E. J. Hixson, eds.), pp. 7-22. Intox Press, Little Rock, AR.

w.,

1010

CHAPTER49


MiIatovic, D" and Johnson, M, K. (1993), Reactivation of phosphorodiamidated acetylcholinesterase and neuropathy target esterase by treatment of inhibited enzyme with potassium fluoride, Chem.-Biol. Interact. 87,425430. MiIatovic, D., Moretto, A., Osman, K. A., and Lotti, M. (1997). Phenyl valerate esterases other than neuropathy target esterase and the promotion of organophosphate polyneuropathy. Chem. Res. Toxicol. 10,1045-1048. Morazina, R, and Rosenberg, P. (1970). Lipid changes in tri-o-cresyl phosphate-induced neuropathy. Toxicol. Appl. Pharmacol. 16,461-471. Moretto, A., Bertolazzi, M., Capodicasa, E., Peraica, M., Richardson, R J., Scapellato, M. L., and Lotti, M. (1992a). Phenylmethanesulfonyl fluoride elicits and intensifies the clinical expression of neuropathic insults. Arch. Toxicol. 66,67-72. Moretto, A., Bertolazzi, M., and Lotti, M. (1994). The phosphorothioic acid O-(2-chloro-2,3,3-trifluorocyclobutyl) O-ethyl S-propyl ester exacerbates organophosphate polyneuropathy without inhibition of neuropathy target esterase. Toxicol. Appl. Pharmacol. 129, 133-137. Moretto, A., Capodicasa, E., and Lotti, M. (1992b). Clinical expression of organophosphate-induced delayed polyneuropathy in rats. Toxicol. Lett. 63, 97-102; Erratum (1992). Toxicol. Lett. 63,355. Moretto, A., Capodicasa, E., Peraica, M., and Lotti, M. (1991). Age sensitivity to organophosphate-induced delayed polyneuropathy: Biochemical and toxicological studies in developing chicks. Biochem. Pharmacol. 41, 14971504. Moretto, A., Lotti, M., Sabri, M. I., and Spencer, P. S. (1987). Progressive deficit of retrograde axonal transport is associated with the pathogenesis of di-nbutyl dichlorvos axonopathy. 1. Neurochem. 49, 1515-1522. Moretto, A., Lotti, M., and Spencer, P. S. (1989). In vivo and in vitro regional differential sensitivity of neuropathy target esterase to di-n-butyl2,2-dichlorovinyl phosphate. Arch. Toxicol. 63, 469-473. National Institutes of Health (NIH) (1994). "The Persian Gulf Experience and Health." National Institutes of Health, Bethesda, MD. Nostrandt, A. C., and Ehrich, M. (1992). Development of a model cell culture system in which to study early effects of neuropathy-inducing organophosphorus esters. Toxicol. Lett. 60, 107-114. Nostrandt, A. C., and Ehrich, M. (1993). Modification ofmipafox-induced inhibition of neuropathy target esterase in neuroblastoma cells of human origin. Toxicol. Appl. Pharmacol. 121, 36-42. Osman, K. A., Moretto, A., and Lotti, M. (1996). Sulfonyl fluorides and the promotion of diisopropyl fluorophosphate neuropathy. Fundam. Appl. Toxicol. 33,294-297. OsterIoh, J., Lotti, M., and Pond, S. M. (1983). Toxicologic studies in a fatal overdose of 2,4-D, MCPP, and chlorpyrifos. 1. Anal. Toxicol. 7, 125-129. Padilla, S., and Veronesi, B. (1985). The relationship between neurological damage and neurotoxic esterase inhibition in rats acutely exposed to triortho-tolyl phosphate. Toxicol. Appl. Pharmacol. 78,78-87. Padilla, S., and Veronesi, B. (1988). Biochemical and morphological validation of a rodent model of organophosphorus-induced delayed neuropathy. Toxicol. Ind. Health 4,331-337. Patton, S. E., Lapadula, D. M., and Abou-Donia, M. B. (1986). Relationship of tri-o-cresy1 phosphate-induced delayed neurotoxicity to enhancement of in vitro phosphorylation of hen brain and spinal cord proteins. 1. Pharmacol. Exp. Ther. 239, 597-605. Patton, S. E., Lapadula, D. M., O'Callaghan, J. P., Miller, D. B., and AbouDonia, M. B. (1985). Changes in in vitro brain and spinal cord protein phosphorylation after a single oral administration of tri-o-cresyl phosphate to hens. 1. Neurochem. 45, 1567-1577. Patton, S. E., O'Callaghan, J. P., Miller, D. B., and Abou-Donia, M. B. (1983). Effect of oral administration of tri-o-cresyl phosphate on in vitro phosphorylation of membrane and cytosolic proteins from chicken brain. 1. Neurochem. 41, 897-901. Peraica, M., Capodicasa, E., Moretto, A., and Lotti, M. (1993). Organophosphate polyneuropathy in chicks. Biochem. Pharmacol. 45, 131-135. Peraica, M., Moretto, A., and Lotti, M. (1995). Selective promotion by phenylmethanesulfonyl fluoride of peripheral and spinal cord neuropathies initiated by diisopropyl phosphorofluoridate in the hen. Toxicol. Lett. SO, 115-121.

Perdrizet, J. A., Cummings, J. E., and Lahunta, A. (1985). Presumptive organophosphate-induced delayed neurotoxicity in a paralyzed bull. Cornell Veterinarian 75, 401-410. Perkins, S. K., Ehrich, M., and Jortner, B. S. (1995). Morphological and immunocytochemical study of phosphorylated neurofilaments during the evolution of organophosphorus ester-induced delayed neuropathy (OPIDN). Toxicologist 15, 209. Pope, c., diLorenzo, K., and Ehrich, M. (1995). Possible involvement of a neurotrophic factor during the early stages of organophosphate-induced delayed neurotoxicity. Toxicol. Lett. 75, 111-117. Pope, C. N., and Padilla, S. (1989a). Modulation of neurotoxic esterase activity in vitro by phospholipids. Toxicol. Appl. Pharmacol. 97, 272-278. Pope, C. N., and Padilla, S. S. (1989b). Chromatographic characterization of neurotoxic esterase. Biochem. Pharmacol. 38, 181-188. Pope, C. N., and Padilla, S. (1990). Potentiation of organophosphorus-induced delayed neurotoxicity by phenylmethylsulfonyl fluoride. 1. Toxicot. Environ. Health 31, 261-273. Pope, C. N., Chapman, M. L., Tanaka, D., and Padilla, S. (1992). Phenylmethylsulfonyl fluoride alters sensitivity to organophosphorus-induced delayed neurotoxicity in developing animals. NeuroToxicology 13,355-364. Pope, C. N., Tanaka, D., and Padilla, S. (1993). The role of neurotoxic esterase (NTE) in the prevention and potentiation of organophosphorus-induced delayed neurotoxicity (OPIDN). Chem.-Biol. Interact. 87, 395-406. President's Advisory Council on Gulf War Veterans (1996). "Final Report." U.S. Gov. Printing Office, Washington, DC. Prineas, J. (1969). The pathogenesis of dying-back polyneuropathies. 1. An ultrastructural study of experimental tri-ortho-cresyl phosphate intoxication in the cat. 1. Neuropathol. Exp. Neurol. 28, 571-597. Randall, J. c., Yano, B. L., and Richardson, R J. (1997). Potentiation of organophosphorus compound-induced delayed neurotoxicity (OPIDN) in the central and peripheral nervous system of the adult hen: Distribution of axonal lesions. 1. Toxicot. Environ. Health 51, 571-590. Richardson, R J. (1992). Interactions of organophosphorus compounds with neurotoxic esterase. In "Organophosphates: Chemistry, Fate, and Effects" (J. E. Chambers and P. E. Levi, eds.), pp. 299-323. Academic Press, San Diego. Richardson, R. J. (1995). Assessment of the neurotoxic potential of chlorpyrifos relative to other organophosphorus compounds: A critical review of the literature. 1. Toxicol. Environ. Health 44, 135-165. Richardson, R. J., Moore, T. B., Kayyali, U. S., and Randall, J. C. (1993). Chlorpyrifos: Assessment of potential for delayed neurotoxicity by repeated dosing in adult hens with monitoring of brain acetylcholinesterase, brain and lymphocyte neurotoxic esterase, and plasma butyrylcholinesterase activities. Fundam. Appl. Toxicol. 21, 89-96. Roberts, D. V. (1977). A longitudinal electromyographical study of six men occupationally exposed to organophosphorus compounds. Int. Arch. Occup. Environ. Health 38, 221-229. Robertson, D. G., Mattson, A. M., Bestervelt, L. L., Richardson, R J., and Anderson, R J. (1988). Time course of electrophysiologic effects induced by di-n-butyl-2,2-dichlorovinyl phosphate (DBCVP) in the adult hen. 1. Toxicol. Environ. Health 23, 283-294. Robertson, D. G., Schwab, B. W, Sills, R. D., Richardson, R J., and Anderson, R. J. (1987). Electrophysiologic changes following treatment with organophosphorus-induced delayed neuropathy-producing agents in the adult hen. Toxicol. Appl. Pharmacal. 87,420-429. Ruffer-Tumer, M. E., Read, D. J., and Johnson, M. K. (1992). Purification of neuropathy target esterase from avian brain after prelabelling with [3H]diisopropyl phosphorofluoridate. 1. Neurochem. 58, 135-141. Sanders, D. E., Lahunta, D. E., Cummings, J. F., and Sanders, J. A. (1985). Progressive paresis in sheep due to delayed neurotoxicity of triaryl phosphates. Cornell Veterinarian 75, 493-504. Schaeppi, U., Krinke, G., and Kobel, W (1984). Prolonged exposure to triethylphosphate induces sensory motor neuropathy in the dog. Neurobehav. Toxicol. Teratol. 6, 39-50. Schlaepfer, W. W. (1971). Experimental alterations of neurofilaments and neurotubules by calcium and other ions. Exp. Cell Res. 67, 73-80.

References

Schlaepfer, W. W. (1987). Neurofilaments: Structure, metabolism and implications in disease. 1. Neuropathol. Exp. Neurol. 46, 117-129. Schlaepfer, W. W., and Hasler, M. B. (1979). Characterization of the calciuminduced disruption of neurofilaments in rat peripheral nerve. Brain Res. 168, 299-309. Schlaepfer, W. W., and Zimmerman, U. P. (1984) Calcium activated protease and the regulation of the axonal cytoskeleton. In "Axonal Transport in Neuronal Growth and Regeneration" (J. S. Elam and P. Cancalon, eds.), pp. 261-273. Plenum, New York. Schlimmer, B. P., and Parker, K. 1. (1996). Adrenocorticotropic hormone; adrenocortical steroids and their synthetic analogs; inhibitors of the synthesis and actions of adrenocortical hormones. In "Goodman & Gilman's The Pharmacological Basis of Therapeutics" (1. G. Hardman, L. E. Limbird, P. B. Molinoff, R. W. Ruddon, and A. G. Gilman, eds.), 9th cd., pp. 1459-1485. McGraw-Hill, New York. Schwab, B. W., and Richardson, R. J. (1986). Lymphocyte and brain neurotoxic esterase: Dose and time dependence of inhibition in the hen examined with three organophosphorus esters. Toxicol. Appl. Pharmacol. 83, 1-9. Schwab, B. w., Davis, C. S., Miller, P. H., and Richardson, R. J. (1985). Solubilization of hen brain neurotoxic esterase in dimethylsulfoxide. Biochem. Biophys. Res. Commun. 132,81-87. Seifert, J., and Wilson, B. W. (1994). Solubilization of neuropathy target esterase and other phenyl valerate carboxylesterases from chicken embryonic brain by phospholipase A2. Comp. Biochem. Physiol. 108,337-341. Shell, L., Jortner, B. S., and Ehrich, M. (1988). Assessment of organophosphorus-induced delayed neuropathy in chickens using needle electromyography.1. Toxicol. Environ. Health 25, 21-33. Sogorb, M. A., Bas, S., Gutierrez, L. M., Vilanova, E., and Viniegra, S. (1997). Bovine chromaffin cells as an in vitro model for the study of non-cholinergic toxic effects of organophosphorus compounds. Arch. Toxicol. Suppl. 19, 347-355. Sogorb, M. A., Vilanova, E., Quintanar, J. L., and Viniegra, S. (1996). Bovine chromaffin cells in culture show carboxylesterase activities sensitive to organophosphorus compounds. Int. 1. Biochem. Cell BioI. 28,983-989. Somkuti, S. G., Tilson, H. A., Brown, H. R., Campbell, G. A., Lapadula, D. M., and Abou-Donia, M. B. (1988). Lack of delayed neurotoxic effect after tri-o-cresyl phosphate treatment in male Fischer 344 rats: Biochemical, neurobehavioral, and neuropathological studies. Fundam. Appl. Toxicol. 10, 199-205. Sprague, G. L., Sandvik, L. L., Bickford, A. A., and Castles, T. R. (1980). Evaluation of a sensitive grading system for assessing acute and subchronic delayed neurotoxicity in hens. Life Sci. 27, 2523-2528. Stumpf, A. M., Tanaka, D. J., Aulerich, R. J., and Bursian, S. J. (1989). Delayed neurotoxic effects of tri-o-tolyl phosphate in the European ferret. 1. Toxicol. Environ. Health 26, 61-73. Suwita, E., Lapadula, D. M., and Abou-Donia, M. B. (1986a). Calcium and calmodulin stimulated in vitro phosphorylation of rooster brain tubulin and MAP-2 following a single oral dose of tri-o-cresyl phosphate. Brain Res. 374, 199-203. Suwita, E., Lapadula, D. M., and Abou-Donia, M. B. (1986b). Calcium and calmodulin-enhanced in vitro phosphorylation of hen brain cold-stable microtubules and spinal cord neurofilament triplet proteins after a single oral dose oftri-o-cresyl phosphate. Proc. Natl. Acad. Sci. U.SA 83,6174-6178. Tanaka, D., Jr., Bursian, S. J., and Aulerich, R. J. (1994). Age-related effects of triphenyl phosphite-induced delayed neuropathy on central visual pathways in the European ferret (Mustela putorius Juro). Fundam. Appl. Toxicol. 22, 577-587. Tanaka, D., Bursian, S. J., and Lehning, E. J. (1992). Neuropathological effects of triphenyl phosphite on the central nervous system of the hen. Fundam. Appl. Toxicol. 18, 72-78. Tanaka, D., Bursian, S. J., and Lehning, E. (1990a). Selective and terminal degeneration in the chicken brains tern and cerebellum following exposure to bi(1-methylethyl)phosphorofluoridate (DFP). Brain Res. 519, 200-208. Tanaka, D., Bursian, S. J., Lehning, E. J., and Aulerich, R. J. (1990b). Exposure to triphenyl phosphite results in widespread degeneration in the mammalian central nervous system. Brain Res. 531, 294-298.

1011

Tanaka, D. J., and Bursian, S. J. (1989). Degeneration patterns in the chicken central nervous system induced by ingestion of the organophosphorus delayed neurotoxin tri-ortho-tolyl phosphate: A silver impregnation study. Brain Res. 484, 240-256. Tanaka, D. J., Bursian, S. J., Lehning, E. J., and Auterich, R. J. (1991). Delayed neurotoxic effects of bis(1-methylethyl) phosphorofluoridate (DFP) in the European ferret: A possible mammalian model for organophosphorusinduced delayed neurotoxicity. NeuroToxicology 12, 209-224. Thomas, T. C., Ishakawa, Y., McNamee, M. G., and Wilson, B. W. (1989). Correlation of neuropathy target esterase activity with specific tritiated di-isopropyl phosphorofiuoridate-labelled proteins. Biochem. 1. 257, 109116. Thomas, T. c., Szekacs, A., Hammock, B. D., Wilson, B. W., and McNamee, M. G. (1993). Affinity chromatography of neuropathy target esterase. Chem.-Biol. Interact. 87,347-360. Thomas, T. c., Szekacs, A., Rojas, S., Hammock, B. D., Wilson, B. W., and McNamee, M. G. (1990). Characterization of neuropathy target esterase using trifluoromethyl ketones. Biochem. Pharmacol. 40, 2587-2596. Tormo, N., Gimeno, 1. R., Sogorb, M. A., Diaz-Alejo, N., and Vilanova, E. (1993). Soluble and particulate organophosphorus neuropathy target esterase in brain and sciatic nerve of the hen, cat, rat, and chick. 1. Neurochem.

61,2164-2168. "USP DI" (2001). United States Pharmacopia, Micromedix, Englewood, CO. Varghese, R. G., Bursian, S. J., Tobias, c., and Tanaka, D., Jr. (1995). Organophosphorus-induced delayed neurotoxicity: A comparative study of the effects of tri-ortho-tolyl phosphate and triphenyl phosphite on the central nervous system of the Japanese quail. NeuroToxicology 16,45-54. Vasilescu, C., Alexianu, M., and Dan, A. (1984). Delayed neuropathy after organophosphorus insecticide (Diperterex) poisoning: A clinical, electrophysiological and nerve biopsy study. 1. Neurol. Neurosurg. Neuropsych.

47,543-548. Veronesi, B. (1984). Effect of metabolic inhibition with piperonyl butoxide on rodent sensitivity to tri-ortho-cresyl phosphate. Exp. Neurol. 85, 65 1-660. Veronesi, B. (1992). The use of cell culture for evaluating neurotoxicity. In "Neurotoxicology" (H. A Tilson and C. L. Mitchell, eds.), pp. 21-49. Raven Press, New York. Veronesi, B., and Dvergsten, C. (1987). Triphenyl phosphite neuropathy differs from organophosphorus-induced delayed neuropathy in rats. Neuropathol. Appl. Neurobiol. 13, 193-208. Veronesi, B., and Ehrich, M. (1993). Using neuroblastoma cell lines to examine organophosphate neurotoxicity. In Vitro Toxicol., 6, 57-65. Veronesi, B., Enrich, M., Blusztajn, J. K., Oortgiesen, M., and Durham, H. (1997). Cell culture models of interspecies selectivity to organophosphorous insecticides. NeuroToxicology 18,283-297. Veronesi, B., PadiIIa, S., Blackmon, K., and Pope, C. (1991). Murine susceptibility to organophosphorus-induced delayed neuropathy (OPIDN). Toxicol. Appl. Pharmacol. 107,311-324. Veronesi, B., Padilla, S., and Lyerly, D. (1986). The correlation between neurotoxic esterase inhibition and mipafox-induced damage in rats. NeuroToxicology 7, 207-216. Vilanova, E., Barril, J., and Carrera, V. (1993). Biochemical properties and possible toxicological significance of various forms of NTE. Chem.-Biol. Interact. 87, 369-381. Vilanova, E., Barril, J., Carrera, v., and PeIIin, M. C. (1990). Soluble and particulate forms of the organophosphorus neuropathy target esterase in hen sciatic nerve. 1. Neurochem. 55, 1258-1265. WiIliams, C. H., Johnson, H. J., and CasterIine, J. L. (1966). Cholesterol content of spinal cord and sciatic nerve of hens after organophosphate and carbamate administration. 1. Neurochem. 13,471-474. WiIliams, D. G., and Johnson, M. K. (1981). Gel-electrophoretic identification of hen brain neurotoxic esterase, labelled with tritiated di-isopropyl phosphorofluoridate. Biochem. 1. 199, 323-333. Wu, S. Y., and Casida, J. E. (1994). Neuropathy target esterase inhibitors: Enantiomeric separation and stereospecificity of 2-substituted-4H-l,3,2benzodioxaphosphorin 2 oxides. Chem. Res. Toxicol. 7, 77-8\.

1012

CHAPTER 49


Wu, S. Y., and Casida, J. E. (1995). Ethyl octylphosphonofluoridate and analogs: Optimized inhibitors of neuropathy target esterase. Chem. Res. Toxicol. 8, 1070-1075. Yoshida, M., Tomizawa, M., Wu, S. Y., Quinstad, G. B., and Casida, J. E. (1995). Neuropathy target esterase of hen brain: Active site reactions with 2-[octyl-3Hjoctyl-4H-,3,2-benzodioxaphosphorin 2 oxide and 2-octyl-4H-, 3,2-[aryl-3Hjbenzodioxaphosphorin 2 oxide. J. Neurochem. 64, 16801687.

Yoshida, M., Wu, S. Y., and Casida, J. E. (1994). Reactivity and stereospecificity of neuropathy target esterase and a-chymotrypsin with 2-substituted4H-l,3,2-benzodioxaphosphorin 2 oxides. ToxieoZ. Lett. 74, 164-176. Zech, R., and Chemnitius, J. M. (1987). Neurotoxicant sensitive esterase: Enzymology and pathophysiology of organophosphorus ester-induced delayed neuropathy. Prog. Neurobiol. 29, 191-218.

CHAPTER

50 Understanding the Toxic Actions of Organophosphates Kai Savolainen Finnish Institute of Occupational Health

50.1 INTRODUCTION Organophosphates (OP) are widely used as insecticides and thus exposure to these compounds still represents a genuine health risk. The overall mechanisms of action of organophosphate-induced neurotoxic effects are well known, but the underlying molecular mechanisms of toxic actions are surprisingly poorly known. However, the introduction of a number of OP pesticides and highly toxic OP nerve agents has emphasized the importance of understanding the mechanisms of toxicity of these OP compounds in detail. It is fundamental to appreciate that their toxicity stems largely from excess acetylcholine (ACh) due to the inhibition of acetylcholinesterase (AChE) and subsequent accumulation of ACh in the target tissues, especially those cells in the vicinity of cholinergic receptors that are responsible for mediating the effects of ACh. The dramatic effects seen in OP intoxication include brain activation, epileptiformic convulsions, muscular tremors, which lead ultimately to flaccid paralysis, increased sweating and salivation, profound bronchial secretion, bronchoconstriction, increased activity of the intestine and diarrhea, miosis, hypertension, lowered body temperature, and hyperglycemia. When the effects of OP compounds are compared, marked differences are evident between them. This is most likely due to the marked differences in their ability to bind with their prime target, AChE, and the differences in the rapidity of ACh accumulation in and close to the targets of ACh. The consequences of excess ACh are primarily mediated via cholinergic muscarinic and nicotinic receptor activation. Muscarinic receptors are found in the central nervous system (CNS), blood vessel walls, and endocrine and exocrine glands. Nicotinic receptors are located in autonomic nervous ganglia, in the CNS, in the adrenals, and in the neuromuscular junction, the area specialized for transmission of neuronal impulses to striated muscles. Muscarinic receptors are G-protein-coupled, slow reacting transmembrane proteins. After activation, their effects are mediated into the cells via formation of calcium-mobilizing phosphoinositide-derived second Handbook of Pesticide Toxicology Volume 2. Agents

messengers or inhibition of adenylate cyclase, leading to increased formation of cyclic adenosine monophosphate (cAMP). Nicotinic receptors are ion channels. Their activation leads to increased influx of sodium into the cell. There are several subtypes of both receptors, and the mode of action of different muscarinic receptors and different nicotinic receptors may markedly differ from each other. In the CNS, there are more muscarinic than nicotine receptors, and muscarinic receptor activation in the brain, as in peripheral tissue, has profound effects on neuronal signalling, and can alter the numbers of many different receptors, as well as modify gene expression and the expression of proteins encoded by these genes. Cholinergic muscarinic activation also dramatically facilitates brain metabolism and induces major electrophysiological effects, often associated with overt convulsions. The CNS effects of OPs can be modified by drugs, typically cholinergic antagonists such as atropine, but also with y-aminobutyric acidergic (GABAergic) agonists such as benzodiazepines and antagonists of glutamatergic receptors. In fact, anticholinergics, like atropine, and diazepam, which belong to the benzodiazepine group of drugs, are the most effective antidotes against OP poisoning. In addition, interaction of cholinergic stimulation with lithium markedly amplifies cholinergic-induced neuronal signalling and convulsions, most likely due to an interaction at the G-protein level. Nicotinic receptors have their most dramatic effects in autonomic ganglia and the neuromuscular endplates. These alterations are characterized by a decrease in membrane potential, membrane resistance, and a decrease in afterhyperpolarization. Many of these effects can, surprisingly, be inhibited by atropine, most likely due to an interaction of muscarinic and nicotinic receptors in autonomic ganglia. The most effective blockers of nicotinic receptors are d-tubocurarine and its more modem analogs. At the neuromuscular endplate, OPs induce (1) repetitive activity in response to single nerve stimulus and (2) decremental responses to repetitive nerve stimulation. OPs typically also induce accelerated spontaneous release of ACh, leading to increases in the miniature endplate potentials

1013


1014

CHAPTER 50 Toxic Actions of Organophosphates

(MEPP) frequency, but at high OP concentrations, the end result is depolarization of the neuromuscular endplate and endplate regions. The effects of OPs on neuromuscular endplates can be prevented with AChE reactivators such as pyradine-2-aldoxime methiodide (2-PAM) that can restore, in part, neuromuscular transmission. The cardiovascular and respiratory systems are particularly sensitive to the effects of OPs because both are under strict cholinergic control. A more detailed understanding of the effects of these toxic agents seems to be warranted because of their dramatic effects on the CNS. In particular, muscarinic receptor-mediated effects have been overlooked in the past. Recent observations also suggest that cholinergic stimulation of cholinergic muscarinic receptors might, in fact, be a trigger that activates many other neurotransmitter systems. For example, it is now becoming clear that cholinergic brain stimulation is the trigger that sets off the propagation of convulsive waves.

Table 50.1 Mllior Actions of Organophosphate AntichoIinesterases at Various Sites in the Body Receptor

Target organ

Symptoms and signs

Central

Central nervous system

Giddiness, anxiety, restlessness, headache, tremor, confusion, failure to concentrate, convulsions, respiratory depression

Muscarinic

Glands Nasal mucosa

Bronchorrhea

Sweat

Sweating

Lachrymal

Lachrymation

Salivary

Salivation

Smooth muscle Iris

50.2 HISTORY AND BACKGROUND Organophosphorous compounds were first synthetized in 1854, but their remarkable toxicity was not recognized until the 1930s (see Minton and Murray, 1988). The first synthesized OP pesticide, tetraethyl pyrophosphate (TEPP), was developed in Germany before World War 11 to replace the highly toxic and lipid-soluble botanical insecticide, nicotine. At the same time, the highly toxic nerve gas agents, tabun and sarin, also were developed, but they were not used during the war (Minton and Murray,1988). Anticholinesterase OP compounds have been widely used as insecticides because it was quickly appreciated that insects are highly susceptible to these agents, even though their toxicity to mammals also proved to be high (Minton and Murray, 1988). OPs represent a large group, many members of which cause toxicity by inhibiting the key enzyme, acetylcholinesterase. This enzyme promotes the hydrolysis of acetylcholine, the principal physiological cholinergic agonist of nicotinic and muscarinic receptors in the body. Thus, the toxicity of OP compounds in many respects can be considered to be ACh toxicity. Due to the ubiquitous distribution of both nicotinic and muscarinic cholinergic receptors throughout the body, exposure to OP compounds has widespread toxic consequences in several target organs. The actions of OPs include neuronal excitation in the brain and subsequent epileptic convulsions, muscular tremors, increased sweating and salivation, increased bronchial secretion, bronchoconstriction, increased activity of the intestine and diarrhea, miosis of the eyes, blurred vision, tachycardia and hypertension, hyperglycemia, and lowered body temperature. For example, when adult baboons were exposed to soman, a highly toxic anticholinesterase OP nerve agent (for soman, see Churchill et aI., 1985), via an intravenous infusion, the animals had a very rapid onset of OP-type cholinergic signs of intoxication, including overt muscular fasciculations resembling epileptic convulsions, stridorous breathing, copious secretions, and atrioventricular arrhythmias (Anzueto et al., 1986).

Rhinorrhea

Bronchial mucosa

Nicotinic

Miosis

Ciliary muscle

Failure of accommodation

Gut

Abdominal cramps, diarrhea

Bladder

Frequency, involuntary micturition

Heart

Bradycardia

Autonomic ganglia

Sympathetic effects: pallor,

Skeletal muscle

Weakness, fasciculation

tachycardia, hypertension

Source: Modified from Fuortes et al. (1993), and Marrs (1996).

OP-induced effects are mediated through the activation of nicotinic or muscarinic receptors. The distribution of these receptors varies in different parts of the body. Table 50.1 summarizes the most important toxic actions of OP compounds and classifies them according to the receptor type that is behind each toxic action. A number of recent reviews have summarized the key toxic actions of different OP compounds (Choi et al., 1995; De Bleecker et al., 1992; Ecobichon, 1996; Fuortes etal., 1993; Gunderson et al., 1992; Gutmann and Besser, 1990; Marrs, 1996; McDonough and Shih, 1997; Minton and Murray, 1988). Intoxication with OPs can take place a number of ways. Previously, when less was known about their toxicity, careless use of OPs often lead to accidental acute poisonings, and even some deaths of workers using these agents. It was estimated in 1990 that the number of annual pesticide poisonings was about 3 million cases in the world and that the incidence was thus about 57 cases per potentially exposed 100,000 individuals. There were about 20,000 cases of pesticide-induced fatalities in the world, most due to excessive exposure to OP compounds. In industrialized countries, occupational exposure to OP compounds is, for the most part, well controlled and the number of poisonings is relatively small. However, in developing countries, in which the use of OP compounds is particularly widespread because of the hot climatic conditions, the number of deaths may be high. For example, pesticide poisonings are relatively common in countries such as Sri Lanka, Venezuela, Indonesia, South Africa, and Brazil (see Choi et aI., 1995).

50.2 History and Background

Whereas OP compounds are usually highly lipid-soluble, they readily penetrate the skin, and exposure in the occupational environment takes place mainly through the skin. This is illustrated in Fig. 50.1 by the excellent correlation between alterations of the levels of mevinphos, a highly toxic OP insecticide, on the foliage in greenhouses, and the dermal contamination rate, as well as the decline in the acetylcholinesterase activity in red blood cells in the exposed workers as a function of time (Kangas et al., 1993). In some cases, exposure via the lungs also may play a role, but it is usually of minor importance (Savolainen and Kangas, 1995). In some cases, inhalation can be relatively important as the route of the compound into the body. These situations are, however, unlikely to be associated with an exposure to OP that would have remarkable health consequences. A rare example could be the use of OP nerve agents when OP concentrations in the air are likely to be high. However, in occupational environments and in situations in which OP-induced effects are likely to take place, dermal exposure predominates (Kangas et al., 1993; Savolainen and Kangas, 1995; Storm et al., 2000). Exposure to OPs via food, in household use, and thorough oral routes is nonsignificant when suicidal cases are excluded. Lethalities associated with OPs in the past contained a large number of suicides; for example, in the 1950s, parathioncontaining formulations became popular for this purpose in several countries (Choi et al., 1995; Hayes et aI., 1978; 10vanovic et al., 1990), including Finland, until their availability was strictly restricted (AI ha, 1967; Minton and Murray, 1988; Marrs, 1996). However, the most toxic OP compounds are the anticholinesterase OP nerve agents, which were created by defence industries of many countries. These include sarin (isopropyl methylphosphonoftuoridate), soman (pinacolyl methylphosphonofluoridate), tabun (ethyl N -dimethylphosphoramidocyanidate), and YX (O-ethyl-S-[2(diisopropylamine)-ethyl)methyl-phosphonothionate) (see Abdallah et al., 1992). Even though these compounds have been used rarely (e.g., the Iraqis against the Iranians and the Kurds, as well as in the well known attack by a terrorist group in a Tokyo underground station), their very existence carries a potential continuous threat (Gunderson et aI., ] 992; Ecobichon, 1996). The extreme toxicity of many OP compounds highlights the need for a more complete understanding of their mechanisms of toxic actions. Even though the overall toxicity and the general mechanisms of toxic actions of OP compounds have been rather well clarified over the years and seem to be quite similar within the group, a more thorough understanding of the cascades of cellular and subcellular toxic events of these compounds in the nervous systems is needed for effective prevention and treatment of OP-induced poisonings. For this purpose, studies on mechanisms of OP nerve agents are important, because state-of-the-art studies on mechanisms of occupationally used OPs are, for the most part, lacking.

1015

Mevlnphos (og/cm1) 80r---------------------------~

20

oL---------------------------~ 10 20 30 40

o

Time (br) aIler appllcatlon

(a) Dermal exposure rate ijlg/hr) 20 ,------nr-------------------~

16

12

8

•

ID

20

.to

30

40

50

Time (br) after appUcatlon

(b) Mevlnpbos (ngfanl) 1~,-------------------------~~ 120

90

8

12

16

20

Dermal exPOSRre rate ()lIIbr)

(c)

Figure 50.1 Dennal absorption is important in the absorption of organophosphates. Absorption of mevinphos into the body after exposure in greenhouses is used as an example in an occupational setting. (a) The decrease in the amount of mevinphos in foliar samples from flowers grown in greenhouses as a function of time. The equation of the curve is y = -0.026 + 1.789; r = 0.96. (b) The dennal exposure rate of workers exposed to mevinphos after the application of the compound to fl owers grown in greenhouses. The equation of the curve is y = 3352.7- 2.4; r = -0.67. (c) The correlation (y = 7.2 + 3.5; r 0.97) between the decrease of the amount of mevinphos on the foliage and the dennal exposure rate to mevinphos via the hands. These data provide evidence that in this situation the skin is the most important exposure route to mevinphos. Reprinted with pennission from J. Kangas et al., Am. J. lnd. Hyg. 54(4), 150--157 (1993).

=

1016

CHAPTER 50

Toxic Actions of Organophosphates

50.3 CHEMISTRY OF

ORGANOPHOSPHORUS COMPOUNDS Anticholinesterase OPs are derivatives of phosphonic or phosphoric acid or their sulfur-containing analogs, notably phosphorothioic, phosphonothionic, phosphorodithioic, or phosphonodithioic acids (see Fig. 50.2). Phosphonic acid or its derivatives does not generally inhibit AChE activity. OPs with AChE activity usually have two alkyl groups and a third group, the leaving group, that is often an aryl group or a heterocyclic group. However, in most of the OP warfare nerve agents, the leaving group contains fluorine (see Fig. 50.3; Holmstedt, 1963; Marrs, 1996; Minton and Murray, 1988; World Health Organization, 1986). The leaving group is more susceptible to hydrolysis than the alkyl groups. Typically, OPs with the P=S

Type ofOP

configuration have little or no inhibitory action on AChE unless they have been activated through enzymatic or nonenzymatic oxidative desulfuration to the corresponding oxon that contains the P=O configuration. Such compounds are termed indirect inhibitors of AChE, and include many important insecticides such as malathion and parathion (Aldridge, 1996; Ecobichon, 1996; Hirvonen et al., 1993; Savolainen et al., 1991). Holmstedt (1963) classified the OPs into four categories based on the structure of the leaving group. In category I, the leaving group contains a quaternary nitrogen. An example of thc compounds in this group is the drug ecothiopate. Category 11 includes the warfare nerve agents soman and sarin, as well as diisopropyl phosphofluoridate (DFP) (see Churchill et aI., 1987; Savolainen et al., 1988a, b), where the leaving group is fluorine. In category Ill, the leaving group is cyanide, cyanate, thio-

Example

Structure

Phosphates

Monocrotophos Chlorfenvinphos Chlorpyrifos-methyl Dichlorvos Tri-o-cresyl phosphate

Phosphonates

Trichlorfon

Phosphinates

Glufosinate"

Phosphorothioates

Pirimiphos-methyl Bromophos Diazinon Triazophos

(8=)

EPN Leptophos

Phosphonothioates (8=1

Demeton-S-methyl Ecothiopate

Phosphorothioates (S·substituted)

Phosphonothioates (S-substituted)

R

vx

(continuedl Figure 50.2 Organic derivatives of phosphoric acid.

50.3 Chemistry of Organophosphorus Compounds

cyanate, or a halogen other than fluorine. For example, the nerve agent tabun belongs to category III (see Holmstedt, 1963). The fourth group, category IV, is the most heterogeneous in terms of the structure of the leaving group, and contains a large number of pesticides. Derivatives of pyrophosphoric acid include compounds such as TEPP, sulfotep, monothiotep, schradan, and tetraisopropyl pyrophosphoramide (iso-OMPA) that is extensively used in laboratories as a specific inhibitor of butyrylcholinesterase (see Koelle, 1963; Savolainen et al., 1984). These compounds do not, in fact, conveniently fit into the classification created by Holmstedt (1963). What they have in common is an ability to inhibit AChE activity (Marrs, 1996). Some of the other OPs that express some AChE activity do not have a clearly defined leaving group. For example, S,S,S-tributyl phosphorotrithioite (DEF) and phosphorotrithioite (merphos), where all three substituents are S-butyl moieties, and perhaps ethephon, an accelerator of fruit ripening and a mono(chloroalkyl)

derivative of phosphonic acid, all belong to this poorly defined group of OPs (Marrs, 1996). In terms of OP chemistry and nonenzymatic degradation, it is important to keep in mind that most OPs are poorly watersoluble, have a high oil-water partition coefficient, and a low vapor pressure. Most of the OPs, with the exception of dichlorvos, are not particularly volatile, and all are degraded by hydrolysis, especially in alkalic conditions, yielding water-soluble products that are generally considered to be nontoxic. Knowledge of these general chemical properties of OPs has practical implications in decontamination of skin exposed to OP compounds because scrubbing the skin with (an alkali) soap causes rapid hydrolysis of the compound. Extensive reviews on the treatment and management of acute OP compound poisoning are available (see De Bleecker et al., 1992; Ecobichon, 1996; Marrs, 1996; McDonough and Shih, 1997; Minton and Murray, 1988; World Health Organization, 1986).

o

Rs-~-SR

Phosphorodithioates

Malathion Azinphos-ethyl and methyl Dimethoate Disulfoton Phosmet Phosalone

dR

o

Phosphorotrithioates

RS-~-SR s~ o

Phosphorarnidates

Phosphoramidothioates

RO-~-~ I

DEF menos

R Fenamiphos

,

OR

R

S

R

dR

or ' R

RO-~-~

0

R

dR

Methamidophos R Propetamphos

Rs-~-~

Phosphorofiuoridates

DFP

Phosphonofiuoridates

Soman Sarin

GF Pyrophosphates

(Continued).

0

"

I'

RO- P -O-P-O-R

"Not an anticholinesterase. Figure 50.2

o

6R

1017

6R

TEPP Sulfotep Schradan (OMPAl

1018

CHAPTER 50 Toxic Actions of Organophosphates

Examples of the Four Main Categories of Organic Phosphorus Compounds Group

X constituents

Example

substituted quaternary nitrogen

40%)

Nausea, vomiting, diarrhea, salivation/lachrymation, miosis bronchoconstriction, increased bronchial secretions, bradycardia Same as above plus pupils unreactive to light, urinary/fecal incontinence Same as above

Usually none

Headache, dizziness

Muscle fasciculation (fine muscles)

Same as above plus dysarthria, ataxia

Same as above plus muscle fasciculation (diaphragm and respiratory muscles)

Same as above plus coma, convulsions

Moderate (20% < RBCAChE < 40%) Severe (RBC AChE < 20%)

aModified from Lotti (1991).

by low levels of plasma BuChE (see Section 51.4.3.2). Several methods are available to measure these blood enzymes (among the many are Doctor et aI., 1987; Ellman et al., 1961; GarciaLopez and Monteoliva, 1988; Lewis et aI., 1981; London et aI., 1995; St. Omer and Rottinghaus, 1992; Wilson et aI., 1996). However, because hospital laboratories rarely measure RBC AChE activity, in most circumstances one should rely on measurements of plasma BuChE, for which kits are easily available. RBC AChE inhibition confirms the diagnosis of acute OP poisoning. Whole blood AChE also may be measured, considering that only about 10% of the activity is due to the plasma enzyme (Worek et aI., 1999b). Usually there is a good correlation between the severity of signs and symptoms of poisoning and the degree of inhibition of RBC AChE (Table 51.5). Nevertheless, because acute poisoning usually requires prompt treatment, treatment should not be delayed while laboratory confirmation is sought. Therefore, measurement of RBC AChE has limited value in an emergency because diagnosis is exclusively clinical and severe poisoning is inevitably associated with high RBC AChE inhibition. It is more difficult to interpret the relatively low levels of RBC and plasma cholinesterases such as those observed in cases of poisoning that present with equivocal symptoms. This may be the case in mild poisoning or in the initial phase of poisoning caused by OPs that are slowly disposed. Reasons for these difficulties are manyfold and include the following: • Large inter- and intraindividual variability of both RBC AChE and plasma BuChE, making the distinction between physiologically low and inhibited activities impossible. Methods to address these difficulties have recently been developed but they have not been applied yet in clinical settings (Polhuijs et aI., 1997; see Section 51.4.3.2). • Different sensitivity of AChE and BuChE to the same inhibitor. In many cases, the identity of the chemical involved and knowledge of its biochemical characteristics are unknown, thus hampering the interpretation of a significant inhibition of plasma BuChE associated with little

or no inhibition of RBC AChE. This may be the case in both a mild poisoning or the early phase of a more severe poisoning when poisonings are caused by OPs that preferentially inhibit BuChE (see Section 51.4.3.2). • The ratio between the inhibition of RBC AChE and that in the synapses may vary according to the compound. Inhibition of the RBC enzyme may be detected without clinical signs of toxicity in cases of exposures to OPs that do not easily cross the blood-brain barrier (see Section 51.4.3.2). • If reactivators (oximes) are administered, the pharmacological effect depends on the ratio between inhibited and aged enzyme. This ratio may be different in blood and in the nervous tissue, and the pharmacological reactivation of inhibited blood enzymes will be more effective than that of enzymes in the nervous system. Under these circumstances relatively small inhibition in blood enzymes would not correlate with symptomatology (see Section 51.4.3.2). • The method of cholinesterase measurement involves dilutions, which in the case of inhibitors such as carbamates and dimethyl phosphates, may favor the spontanous reactivation of inhibited enzymes (see Section 51.1.2.2). These assay related problems would understimate the actual in vivo inhibition. In conclusion, measurements of blood cholinesterases may confirm the diagnosis, but are not essential: clinical observation remains the cornerstone for diagnosis. For the same reasons, repeated measurements of blood cholinesterase during poisoning have no prognostic value (Nouira et aI., 1994) and they cannot be used to assess the efficacy of treatment. When AChE is irreversibly inhibited by OPs, the reappearance of RBC activity depends on new erythrocytes entering the blood stream. Whereas the average lifespan of RBCs is 120 days, in most cases, observed reappearance of RBC AChE occurs at a rate of about 1% per day. The corresponding rate for plasma BuChE, which derives from liver synthesis, is about 5% per day (see Section 51.4.3.2).

51.1 The Cholinergic Syndrome

1051

100

10 0;

El

SCo

~

0

I>Il

::!.

0.1

O.oI 0

2

4

6

8

10

DAYS

Figure 51.4

Time course of blood concentration of several OPs in acutely poisoned patients. Data for

(+) methamidophos, (_) fenitrothion, (J.) methylparathion, and (e) parathion from (Lotti, 2000); (100 0.12

AChE 150 NTEIso

16

0.06

400

0.2

0.2 12 0.003

0.05 >1 39

aData from Lotti and Johnson (1978), Bertolazzi et al. (1991), and Capodicasa et al. (1991). bDirect acting OPs or metabolites (chlorpyrifos oxon from chlorpyrifos and phenylsaligenin phosphate from TOCP).

but just another toxic effect of OPs, in the case of insecticides it will develop at a much higher dose than that which causes cholinergic overstimulation. This is indirectly shown in Table 51.8, where sensitivity to various inhibitors of AChE (target of cholinergic toxicity) and of NTE (target of OPIDP) derived from human tissues are compared. OP insecticides that caused OPIDP in humans (Table 51.9) produced OPIDP only after cholinergic toxicity. Mipafox was never developed as an insecticide and caused OPIDP after mild cholinergic toxicity. Phenyl saligenin phosphate is the active metabolite of TOCP (not an insecticide), which caused several cases of OPIDP without cholinergic toxicity. Similar differences were seen with hen enzymes, which in turn correlate with the capability of a given OP to cause OPIDP, relative to that of causing death (Lotti and Johnson, 1978). Therefore, compounds with AChE 150INTE 150 ratios> 1 may cause OPIDP without cholinergic toxicity, whereas those with a ratio < 1 will cause OPIDP only after cholinergic toxicity and appropriate antidotal treatment. OPIDP displays a characteristic age-related sensitivity in both experimental animals and humans. Children are resistant to OPIDP (Goldstein et aI., 1988) and when they are affected, they recover much quicker than adults, usually within a few months (Senanayake, 1981). In addition to the case reports listed in Table 51.9, other reports can be found where development of OPIDP was associated with single or short-term exposures to certain OPs, although clinical and toxicological evidence was not convincing. EPN and leptophos cause OPIDP in hens and are NTE inhibitors, but only at doses that cause severe cholinergic toxicity (John son, 1975; Ohkawa et aI., 1980). However, a report of OPIDP that involved several workers who had long-tenn exposure to EPN indicated little or no evidence of cholinergic overstimulation and most clinical details were missing. Moreover, during the release of EPN, these workers were also exposed to other chemicals derived from an explosion and fire in a manufacturing facility (Xintaras and Burg, 1980). An outbreak of neurological disorders occurred in a plant that manufactured leptophos (Xintaras et aI., 1978), which is

51.3 Delayed Polyneuropathy

1065

Table 51.9 OP Insecticides that Cause Delayed Polyneuropathy in Humans a Compound

Number of cases

Circumstances

References

Chlorpyrifos

2

Suicide

Lotti et aI., 1986; Martinez-Chuecos et al., 1992

Dichlorvos

3

Suicide

Vasilescu and Florescu, 1980; Wadia et aI., 1985

Isofenphos b

3

Suicide

Catz et aI., 1988; Tracey and Gallagher, 1990; Moretto and Lotti, 1998

Methamidophos

Several

Suicide/occupational

Senanayake and Johnson, 1982; Moretto and Lotti, 1998; McConnell et aI., 1999

Mipafox

2

Occupational

Bidstrup et aI., 1953

Trichlorfon

Several

Suicide

Hierons and Johnson, 1978; Johnson, 1981; Shiraishi et aI., 1983; Niedziella et aI., 1985;

Trichlomate

2

Suicide

Jedrzejowska et aI., 1980; Willems, 1981

Csik et aI., 1986

a Modified

from Lotti (2000). All cases displayed preceding cholinergic toxicity. bOne of these cases was a combined exposure with phoxim.

known to cause OPIDP in experimental animals (Hollingshaus et aI., 1981). Three subjects had signs compatible with OPIDP at medical examination and six had symptoms in a retrospective study. Several subjects, however, had neurological signs unrelated to OPIDP and all were exposed to a variety of neurotoxic chemicals, including n-hexane. Another group of case reports suggested OPIDP development after exposures to omethoate, parathion, mecarbam, fenthion, and mevinphos. However, these pesticides are not NTE inhibitors and negative results have always been reported in the hen test (FAOIWHO, 1987, 1996, 1997; Johnson, 1975; Lotti, 1992b). A polyneuropathy compatible with OPIDP developed after a suicide attempt with omethoate (Curtes et aI., 1981), but in a subsequent report of a man who died shortly after an acute omethoate poisoning, no NTE inhibition was detected in post mortem nerve tissues (Lotti et aI., 1981). Parathion was associated with OPIDP after massive poisoning together with methanol (de Jager et aI., 1981). Toxicological evidence of parathion in body fluids was missing and the clinical description does not support evidence ofRBC AChE inhibition or methanol poisoning (Lotti and Becker, 1982b). Moreover, several cases of severe parathion poisoning resulted in no OPIDP (Namba et aI., 1971). Mecarbam was reported to cause neuropathy, but nerve biopsy revealed segmental demyelination without axonal degeneration (Stamboulis et aI., 1991), a morphological lesion not expected in OPIDP (see Section 51.3.3.3). A case of delayed neuropathy apparently developed 1 month after acute poisoning by fenthion, which was followed by intermediate syndrome (Karademir et aI., 1990). It is not clear whether diagnosis was made on clinical grounds or exclusively on electromyography (EMG). Because the results ofEMG studies were not reported, interpretation of this case is difficult. One further case of delayed polyneuropathy by fenthion was reported (Martinez-Chuecos et aI., 1992), but no clinical details were given. Moreover, another patient poisoned with fenthion, and belonging to the same series, did not develop neuropathy. A female patient who attempted suicide with methylparathion, fell into a deep coma that lasted 4 weeks (Nisse et aI.,

1998). Electromyography was normal 3 weeks after poisoning, whereas 1 week later it showed signs of mild distal sensory motor polyneuropathy. Diffuse myogenic alterations were also detected, but electrophysiological data were not reported. Because the neuropathy disappeared within 4 weeks, this was unlikely a case of OPIDP and was probably a consequence of prolonged coma. A case of severe poisoning by mevinphos was reported to have been complicated by polyneuropathy (Hsiao et aI., 1996). No clinical or electrophysiological data were given. It is said that nerve conduction studies confirmed the neuropathy. In such a case, OPIDP would be unlikely, because conduction is usually, at the most, slightly affected in axonopathy. No followup was reported. An isolated case of Guillan-Barre-like syndrome was described in a patient after exposure to merphos, a defoliant with little anticholinesterase activity (Fisher, 1977). Exposure was dermal and likely very low. Four days later the patient started complaining of upper and lower limb weakness; 14 days after exposure, he developed facial diplegia. EMG and clinical features were suggestive of Guillan-Barre syndrome. This is the only case in the literature of Guillan-Barre-like signs after acute exposure to OP, but the clinical and electrophysiological characteristics of OPIDP are quite different. In conclusion, combined clinical and experimental evidence allows firm conclusions on OPIDP only for a few chemicals (Table 51.9). Nevertheless, it should be pointed out that neuropathic impurities may be present in commercial formulations, which perhaps accounts for some of these cases (Johnson, 1984). 51.3.2 PATHOGENESIS

The initial event in OPIDP is the inhibition of NTE, followed by aging of the phosphoryl NTE complex. These molecular changes occur within a few hours of exposure, but almost nothing is known about what happens between these events and the clinical, morphological, and electrophysiological onset of

1066

CHAPTER 51

Clinical Toxicology of Anticholinesterase Agents

OPIDP 2-3 weeks later (Johnson, 1990; Lotti, 1992b). Limited evidence suggests that development of OPIDP in humans also involves inhibition of NTE. In two fatal cases of OP poisoning, NTE activity was measured post mortem in the peripheral nerve. As expected from animal studies, NTE was found to be inhibited in a case of chlorpyrifos poisoning (Osterloh et aI., 1983), whereas it was not in a case of omethoate poisoning (Lotti et aI., 1981). Lymphocytic NTE (see Section 51.3.3.2) was found to be inhibited soon after exposure in one case of poisoning with chlorpyrifos (Lotti et aI., 1986) and in two cases with methamidophos that all developed OPIDP weeks later (Moretto and Lotti, 1998; McConnell et aI., 1999). Lymphocytic NTE inhibition was also found in a patient poisoned with isofenphos who died on day 32; OPIDP might have developed after this time. In severe poisoning by compounds not known to cause OPIDP, lymphocytic NTE inhibition was not detected (Moretto and Lotti, 1998). After occupational exposures to DEF, substantial NTE inhibition in lymphocytes was measured, but found not to be associated with the development of OPIDP or with electrophysiological changes. In this case, lymphocytic NTE did not represent a good mirror of peripheral nerve NTE, probably because of the particular pharmacokinetics of this compound (Lotti et aI., 1983). 51.3.3 CLINICAL MANIFESTATIONS 51.3.3.1 General Features Symptoms of OPIDP begin 2-3 weeks after single doses when, as in the case of insecticides, cholinergic symptoms have subsided (Lotti et aI., 1984). The lag time between single or short-term exposure and the clinical onset of OPIDP depends on both the chemical involved and the dose (Bidstrup et aI., 1953; Lotti et aI., 1986; Senanayake and Johnson, 1982). OPs with slow pharmacokinetics may cause OPIDP after a prolonged period following exposure (up to 4 weeks), whereas higher doses of OPs that are powerful in causing OPIDP may shorten this period to about 10 days. Clinical features of OPIDP are usually fully expressed within a few days of the onset of symptoms and signs, and no progression has been observed in the absence of further exposure. After repeated exposures, such as those to nonanticholinesterase OPs, the onset of symptoms and their full development is more variable and less definible (Vasilescu, 1982). The usual initial complaint is cramping muscle pain in the legs (Susser and Stein, 1957), followed by distal numbness and paresthesia (Senanayake and Jeyaratnam, 1981; Vasilescu et aI., 1984). Progressive leg weakness occurs, together with depression of tendon reflexes. Symptoms and signs may also appear in the arms and forearms following those in the legs, but always after severe exposures (Bidstrup et aI., 1953; Moretto and Lotti, 1998; Senanayake and Jeyaratnam, 1981; Vasilescu, 1982; Vasilescu et aI., 1984). Physical examination reveals distal symmetrical predominantly motor polyneuropathy, with wasting and flaccid weakness of distal limbs muscles, especially in the legs. Signs

include a characteristic high-stepping gait associated with bilateral foot drop (Senanayake and Jeyaratnam, 1981). Severe OPIDP may result in quadriplegia with foot and wrist drop as well as mild pyramidal signs. In time, there is complete functional recovery if spinal cord axons have been spared by smaller doses (Senanayake, 1981); otherwise, pyramidal and other signs of central neurological involvement may become more evident. The degree of pyramidal involvement determines the prognosis for functional recovery, and spastic ataxia may represent a permanent outcome of severe OPIDP (Morgan and Penovich, 1978; Tosi et aI., 1994; Vasilescu, 1982). Objective evidence of sensory loss is usually slight or absent. In one group of patients poisoned with methamidophos, some sensory symptoms, but no objective signs, were recorded (Senanayake and Johnson, 1982). In two patients who developed OPIDP after exposure to chlorpyrifos and isofenphos, slight sensory alterations were detected during both physical and electrophysiological examination (Moretto and Lotti, 1998). However, in a series of patients, cases were reported where purely sensory peripheral neuropathy was displayed after repeated low exposures to chlorpyrifos that caused some symptomatology, and no signs or mild signs of cholinergic overstimulation (Kaplan et aI., 1993). This contrasts with the known toxicological characteristics of chlorpyrifos, which is a better inhibitor of AChE than NTE (Capodicasa et aI., 1991; Richardson, 1995), and the clinical features observed in two cases of OPIDP induced by chlorpyrifos, where OPIDP was always preceded by severe cholinergic overstimulation (Lotti et aI., 1986; Martinez-Chuecos et aI., 1992). Whereas the exposure assessment in the Kaplan series was limited and based almost exclusively on medical history, interpretation of these discrepancies is difficult. 51.3.3.2 Laboratory Findings They are no specific changes in common laboratory tests, including chemical and morphological analysis of spinal fluid. Increased serum levels of immunoglobulin G autoantibodies to glial fibrillary acidic protein and to neurofilament 200 have been detected in a case of methamidophos poisoning after the development of OPIDP (McConnell et aI., 1999), probably reflecting peripheral nerve damage. Lymphocytic NTE NTE activity was found in lymphatic tissues in humans (Moretto and Lotti, 1988) and its characteristics in circulating lymphocytes led to the conclusion that the level of this blood enzyme is similar to that in the nervous system (Bertoncin et aI., 1985). On this basis, suggestions were made to measure and use lymphocytic NTE like RBC AChE activity is used in the clinical setting and in the biomonitoring of occupational exposures (Lotti, 1987). NTE activity has also been detected in humans platelets (Maroni and Bleecker, 1986). Only on one occasion has the ratio between NTE inhibition in lymphocytes and peripheral nerves been measured, and it was found to be about 1 (Osterloh et aI., 1983), although it is anticipated that it will not be always so, given the different pharmacokinetics of OPs. Inhibition of lymphocytic NTE soon after poisoning

51.4 Long-Term Exposures

was predictive of OPIDP development when measured several days before the onset of OPIDP in two cases (McConnell et aI., 1999; Moretto and Lotti, 1998). Given the relatively high turnover of blood lymphocytes and the usually rapid disappearance of OPs from the blood, measurement of lymphocytic NTE should be made in the early days after poisoning because the activity may be back to almost normal at the onset of OPIDP. However, because no treatment for OPIDP is available, the detection of lymphocytic NTE soon after poisoning has limited clinical value. Similarly, measurements of lymphocytic NTE to monitor occupational exposures to OP insecticides have no practical value because such exposures preferentially inhibit blood cholinesterases and, therefore, lymphocytic NTE rarely would be affected (see Section 51.3.1). Electrophysiology Electrophysiological changes are usually detected concurrently with the onset of clinical symptoms and signs of OPIDP. When performed during the symptom-free period between the disappearance of cholinergic toxicity and the clinical onset of OPIDP, the electrophysiological examination is normal (Lotti et aI., 1986). The electrophysiological picture accords well with the histopathological findings of distal axonopathy (Jedrzejowska et aI., 1980; Lotti et aI., 1986; Moretto and Lotti, 1998; Vasilescu and Florescu, 1980; Vasilescu et aI., 1984; see Section 51.3.3.3). The evaluation reveals partial denervation of affected muscles, with increased insertional activity, abnormal spontaneous activity (fibrillation potentials and positive sharp waves), and a reduced interference pattern; large polyphasic motor unit potentials also may be found after a few weeks. The compound muscle action potentials to supramaximal stimulation of motor nerves are reduced in amplitude, and terminal motor latencies are delayed; maximal conduction velocity is usually normal or slightly reduced. Minimal electrophysiological abnormalities of sensory function are occasionally detected. About 1 year after poisoning, normalization of electrophysiological parameters parallels that of clinical signs unless the pyramidal tract is involved. In such a case, findings may resemble those of amyotrophic lateral sclerosis (Vasilescu, 1982). 51.3.3.3 Pathology The histopathology of OPIDP has rarely been described in humans (Aring, 1942; Jedrzejowska et aI., 1980; Lotti et aI., 1986; Vasilescu et aI., 1984), although there are no major differences from what has been extensively observed in experimental animals (Abou-Donia and Lapadula, 1990; Cavanagh, 1973; Tanaka and Bursian, 1989). The central peripheral distribution of lesions is similar to that of toxic neuropathies of other origins. The vulnerability of nerve fibers is directly related to axonal length and diameter; large-diameter and long fibers are more susceptible than small and short ones. Spinal cord changes involve mainly the anterior horn cells and the pyramidal tracts. Lesions in the tract of Goll were less constant and no lesions were seen in the tract of Burdach (Aring, 1942). Sural nerve biopsies indicated axonal-type lesions with an even degree of involvement of myelinated fibers of differ-

1067

ent sizes and a lesser degree of involvement of unmyelinated fibers. Dark and swollen axoplasm due to the accumulation ofaxoplasmic organelles can be observed association with aspects of axonal degeneration. On teased fiber preparations, some ovoids arranged in linear rows were identified. Electron microscopy found myelin debris in the Schwannian profiles. Depending on the time of biopsy after poisoning, various stages of regeneration and remyelination can be observed. Segmental demyelination is not observed (Jedrzejowska et aI., 1980; Lotti et aI., 1986). These changes indicate a process that is a primary distal axonopathy with moderate, secondary, and distalloss of myelin. 51.3.3.4 Differential Diagnosis The unequivocal suggestion for diagnosis of OPIDP caused by insecticides is the presence of an episode of acute cholinergic toxicity in the recent past medical history. More difficult is differential diagnosis when substantial exposures to nonanticholinesterase OPs that cause OPIDP is overlooked. Symmetricalleg involvement with additional involvement of upper limbs only in severe cases, lack of involvement of cranial nerves and the autonomic system, and electromyographic changes consistent with distal axonal neuropathy are all indicative of OPIDP. Medical history aimed at identification of possible sources of OP exposure remains, in these cases, the only way to etiologically attribute neuropathy. 51.3.3.5 Treatment There is no specific treatment for OPIDP. Intensive programs of physical therapy are indicated to ameliorate muscle trophism during the recovery from peripheral nerve lesions. In later stages, if spasticity develops, GABA antagonists may be used.

51.4 LONG-TERM EXPOSURES Long-term exposures to OPs may cause cholinergic syndrome if both the size of repeated doses and the intervals between them overcome AChE resynthesis in the nervous tissues. In such a case, a buildup of AChE inhibition may occur and when threshold is reached, symptoms that are indistinguishable from those observed after single or short-term exposures are produced. Therefore, only signs and symptoms unrelated to overt cholinergic toxicity will be considered in this section. Many reports on the effects of long-term exposures to OPs lack follow up studies, particularly after cessation of exposures. Therefore, most of the described effects may not be chronic (i.e., longstanding or irreversible) and probably reflect the effects of current exposures. Therefore, it is advisable not to talk about chronic effects, but rather of effects of low-level exposures, either during exposure or shortly afterward, and keep them distinct from effects detectable several months or years after cessation of exposure. Moreover, the major problem of these studies is often the insufficient assessment of exposures, that obviously hampers the interpretation of results.

1068

CHAPTER 51


51.4.1 NEUROLOGICAL, PSYCHIATRIC, AND BEHAVIORAL EFFECTS The large amount of literature that describes neurological, psychiatric, and behavioral effects has been reviewed in several articles (Brown and Brix, 1998; ECETOC, 1998; Eyer, 1995; Steenland, 1996; Ray, 1998a, b). In one review, these effects were all ranked under the heading of chronic OP-induced neuropsychiatric disorders (COPIND, phenomenon 2; Jamal, 1997). As previously discussed (see Section 51.1.3.9), the various effects will be discussed separately for better comprehension and because there is no evidence that they collectively represent a single nosological entity. 51.4.1.1 Neurological Effects on Central Nervous System Clinical reports A case of parkinsonism was described in a subject with reported past and prolonged exposure to OPs. Apparently he also had several episodes of acute poisoning with parathion and malathion said to have required treatment with oral doses of atropine to control symptoms and signs (Davis et aI., 1978). The past history was not fully reported and it is doubtful that oral atropine would have been effective because it is known to be poorly and unreliably absorbed. Consequently, this case remains an isolated and anecdotal report. A visual syndrome, known as Saku disease, which is characterized by reduced visual field, myopia, astigmatism, lesions of the optic nerve, and abnormal retinal functions, was associated with the extensive use of OPs during the 1960s in one area of Japan (Saku). These effects, exclusively reported by Japanese investigators, have been summarized by PleStina and Piukovic-Plestina (1978) and Dementi (1994). However, the symptoms were not consistant among various OPs to which patients were allegedly exposed and often, but not always, associated with AChE inhibition. This inconsistency raises the question whether the effects are compound-specific and related to RBC AChE inhibition. These results have been criticized and the etiologicallink between OPs and Saku disease remains, for the time being, speCUlative (Erikson-Lamy and Grant, 1992). Veterans who took part in the Persian Gulf War reported higher rates of many symptoms, including neurological ones, and had a decreased perception of well-being (Ismail et aI., 1999; NIH, 1994). Although a single consistent pattern of symptoms and signs is far from being defined (Gray et aI., 1996), this mysterious ailment is now known as the Gulf War syndrome or illness, and several hypotheses have been made concerning causes (Lotti, 1999). One theory states that wartime exposure to a combination of OPs and other cholinesterase inhibiting chemicals synergistically produced the syndrome and the neurological signs in particular, (Haley and Kurt, 1997). Among these chemicals, pyridostigmine bromide was the only defined risk factor (Shen, 1998) because it was given to soldiers who served in the Gulf, apparently for several weeks, as prophylaxis for possible nerve gas attacks. The dosing regime was a 30 mg tablet every 8 hours and it aimed to cause reversible

inhibition of AChE at nerve endings, thereby preventing irreversible inhibition of the enzyme by OP weapons. None of the soldiers ever experienced acute cholinergic symptoms and signs compatible with OP or pyridostigmine poisoning. In a study based on a questionnaire submitted by 249 Gulf War veterans from a single battalion of 606 soldiers, factor analysis of symptoms yielded several syndrome factors (possibly variants of a single syndrome) that suggested various neurological dysfunctions (Haley et aI., 1997a). Subjects with the highest factor scores on syndrome 1 (impaired cognition), syndrome 2 (confusion-ataxia), and syndrome 3 (arthro-myo-neuropathy), for a total of 23 cases, were evaluated for neurological functions and compared with 20 controls from the same battalion, 10 of whom had been deployed in the war region but had no complaints and 10 of whom had not been deployed (Haley et aI., 1997b). Brain dysfunction was evidenced by changes in auditory evoked potentials, interocular asymmetry of nystagmic velocity, asymmetry of saccadic velocity, and somatosensory evoked potentials. However, no clinical differences between cases and controls were detected on neurological examination. Exposures to anticholinesterase chemicals of either cases or controls were not reported. Therefore, whether the Gulf War syndrome exists, whether it affects the nervous system, and whether the clinical findings are due to anticholinesterases cannot be ascertained from these studies. Occupational Exposure Studies Minimal EEG disturbances were reported in a study on 50 workers engaged in the manufacture of a range of unspecified OPs (Metcalf and Holmes, 1969). These changes were not seen in 22 controls and mirrored, to a lesser degree, the more severe disturbances usually seen after acute exposures. Work history and exposure data were insufficient, although it was claimed that the workers were also exposed to chlorinated hydrocarbons. Certain neuropsychological changes were also reported, but it is not clear whether they were associated with such persistent EEG changes. In another study, 32 workers exposed to both OPs and organochlorine (dieldrin) pesticides were subdivided into two equal groups, low and high exposure, based on occupational history. Plasma cholinesterase levels were the same in both groups. EEG and neuropsychological changes were found in the high exposure group (Korsak and Sato, 1977). Quantitative exposure data were not given and EEG changes were different from those reported in the above-mentioned study. A selective effect on the left frontal hemisphere, as derived from EEG and neuropsychological results, was detected. This seems inconsistent with a toxic effect. In a seven country biomonitoring and cross-sectional epidemiological survey of low-dose occupational exposure to OPs, changes in EEG were reported only from some countries. Data from one country showed postseason slow wave activity, whereas data from another country reported different changes (Richter, 1993). These results are difficult to assess given the lack of information on exposure and other confounding factors at the time of testing.


51.4.1.2 Neurological Effects on Peripheral Nervous System Clinical Reports In a study on volunteers, mevinphos (25 !1-gkg-1) was administered daily for 28 days to eight subjects, whereas placebo was given to eight controls (Verberk and Salle, 1977). RBC AChE was depressed by 19%, but no correlation was found with the detected changes. At the end of exposure, a 7% decrease in slow fiber motor nerve conduction velocity and a 38% increase in Achilles tendon reflex force were found (as percentages of preexposure values). No effects on neuromuscular transmission were detected. The authors concluded that the significance of such effects with regard to health is unclear. Sensory neuropathies on a series of patients with low-level exposures to chlorpyrifos have been consistently associated with mild or no cholinergic symptoms (Kaplan et aI., 1993) although there is little evidence, if any, for a causal relationship between sensory neuropathy and low-level exposures to chlorpyrifos (see Section 51.3.3.1). In a pilot study, 14 Gulf War veterans were examined for peripheral nerve dysfunction and compared with a control group. Although differences were detected in some parameters (cold threshold, sural nerve latencies, and median nerve sensory action potential), the authors' conclusion was that there may be dysfunction in veterans, but more studies are required (Jamal et aI., 1996). Moreover, the hypothesis that anticholinesterases, other than pyridostigmine, represented a risk factor for veterans of the Gulf War remains to be demonstrated. A clinical study was performed on 72 selected subjects with long-term exposure to sheep dip OPs identified in pilot field studies (Pilkington et aI., 1999). According to defined criteria and neuropathy scores, 23 workers were ranked as having probable/definite neuropathy, 34 workers had possible neuropathy, and 15 workers had no neuropathy. Clinical evaluation, quantitative sensation testing, nerve conduction studies, and electromyography were performed. Results showed neurological signs in 10% of subjects and some small fiber abnormalities in 65% of electrophysiological tests. None of subjects with symptoms was in the no neuropathy subgroup. This neuropathy was thought to be different from classical OPIDP (see Section 51.2), because sensory fiber almost exclusively and small fiber more than large fiber populations were affected. Clinical and electrophysiological assessments were apparently performed at variable times after cessation of peak exposures (up to 1.5 years). Unlike other toxic neuropathies, this apparently new entity also affected upper limbs at early stages, but unlike OPIDP caused by commercial OP pesticides, there was no preceding cholinergic toxicity. Results from this study are difficult to evaluate because they are not presented analytically. Electrophysiological changes were overwhelmingly more frequent than neurological ones, but it is impossible to derive what electrophysiological abnormalities were found in patients with clinically detectable neuropathy. Moreover, the relevance of electrophysiological alterations that are not associated with signs and symptoms is unclear. According to the criteria of

1069

sample selection, a causal relationship with exposure cannot be inferred; there is also a lack of relationship with the estimated cumulative dose. Finally, we should ask why, after such a long period since cessation of exposure, a very mild neuropathy did not recover. It seems that perhaps the study better represents a validation of a screening system to detect minor electrophysiological signs to be used in the field than a demonstration of a causal relationship between low-level exposure to OPs during sheep dipping and the development of a new form of toxic peripheral sensory neuropathy. Occupational Exposure Studies Neuromuscular function was assessed with surface electrodes on the upper limbs of workers exposed to OPs and organochlorine pesticides (Jager et aI., 1970). A higher incidence of electromyographic changes (repetitive activity and reduced amplitude) was detected in workers exposed to both chemical classes (n = 36) as compared to those exposed to organochlorine only (n = 24) and controls (n = 28). The biological significance of these small changes is unclear, in part because of the use of surface electrodes. There was insufficient information, only statements, concerning exposures. Changes were thought to be related to synaptic dysfunction because they were similar to changes found in myastenic patients who were overtreated with anticholinesterase drugs. However, when observed in such patients, these changes are associated with substantial inhibition of AChE, whereas changes in workers were not associated with whole blood AChE inhibition. Fifty-three workers exposed to both OPs and organochlorine were examined shortly after the start and toward the end of a spraying season (Drenth et aI., 1972). Surface electrode electromyography records of 12 subjects changed from normal to abnormal, whereas those of 13 subjects changed from abnormal to normal. No differences in blood AChE were detected. No evidence of exposure was given. Therefore, the conclusion of the authors that electromyography abnormalities represent only an indication of the need for more protection of workers and no evidence of immediate health problems is not substantiated. Minimal electromyographic changes were detected by surface electrodes in 102 workers exposed to OPs when compared to an unmatched control group of 75 subjects (Roberts, 1976). Fifty-six workers were examined before and after a holiday period. Subjects who displayed these changes somewhat improved after the holidays, whereas some unspecified variability was observed in exposed subjects with normal electromyography. No exposure data were available. In a longitudinal study on six workers exposed to OPs over a 7-9 month period, surface electrode electromyography indicated that voltages varied in a manner that reflected a vague assessment of the pattern of exposure (Roberts, 1977). It is difficult to evaluate these results given the lack of exposure data and the methods used. Neuromuscular function was assessed with surface electrodes in a group of 11 spraymen exposed to OPs (including

1070

CHAPTER 51


bromophos, diazinon, chlorpyrifos, and malathion) on a recurrent basis (Stalberg et aI., 1978). Plasma cholinesterases were significantly reduced after work, whereas RBC AChE was not. A slight reduction in sensory conduction velocity and increased fiber density was detected in some workers, but was unrelated to lowered plasma cholinesterase activity. Although exposure was not assessed in this study, plasma enzyme inhibition indirectly suggests a lack of correlation between degree of exposure and detected electrophysiological changes. Another study was conducted on four groups of subjects in which approximate exposure was assessed: 42 highly exposed to OP pesticides, 14 seasonal workers exposed to OPs and reexamined after exposure, 129 agricultural workers with low exposure to OPs, and 26 agricultural workers not exposed (Jusic et aI., 1980). The authors concluded that synapse testing with needle electromyography and clinical examination did not detect latent OP intoxication. A study was conducted on workers exposed to the defoliant DEF, where needle electromyography and biochemical studies (lymphocytic NTE and blood cholinesterases) were performed before and after the spraying season (Lotti et aI., 1983). Air and dermal exposure were measured on a typical working day. No electrophysiological changes were detected, although lymphocytic NTE was about 60% inhibited. NTE inhibition after exposure without correlation with electrophysiological changes was explained by the pharmacokinetics of DEF, which requires metabolic activation to be an esterase inhibitor and occurs mainly in the liver; the active metabolite formed is extremely reactive and unless large amounts are formed, it will not reach the nervous system, but will reach the blood, where it inhibits lymphocytic NTE. Twenty-four workers exposed to fenthion were examined with surface electromyography before and after exposure and compared with 19 unexposed controls (Misra et aI., 1988). Serum AChE was also measured. Electrophysiological findings after exposure were no different from controls. However, mean values of some electrophysiological parameters were altered in the exposed group when results obtained during exposure were compared with follow-up data collected 3 weeks after withdrawal from exposure. Also, mean cholinesterase values increased after the end of exposure, but remained within the normal range. Although each exposed individual was his or her own control, results are difficult to interpret because the intraindividual variability of measured parameters of controls was not reported. Two-hundred twenty-nine workers at a pesticide plant were examined clinically (for neurological impairment) and biochemically (lymphocytic NTE and plasma cholinesterase), and tested for tactile sensitivity and motor performance (Otto et aI., 1990). These workers were engaged in the production of a variety of OPs including diazinon, dimethoate, malathion, phentoate, EPN, leptophos, methamidophos, and trichorfon. Results were compared with those obtained from 180 workers from a fertilizer plant and 167 workers from a textile plant. Mean serum cholinesterases and lymphocytic NTE were lower in pesticide workers, although they were within normal ranges.

The proportion of workers with abnormal neurological findings (involuntary tremors and vibration sense) varied between plants. Tactile thresholds in the finger of the nondominant hand were higher in workers in the pesticide plant and the authors stated that this symptom was the most sensitive index of pesticide neurotoxicity. Toes were not tested. No changes were detected in the neurobehavioral tests. Assessment of exposure was missing (although some OPs that potentially cause OPIDP were manufactured), the incidence of various diseases, including neurological ones, was particularly high (in both cases and controls), and the fact that upper limb neuropathy is not expected in mild OPIDP creates problems in interpretating results. An epidemiological study on 90 pesticide applicators (Stokes et aI., 1995) led to the conclusion that prolonged OP exposure is associated with loss of peripheral nerve function. Exposure was assessed by means of urinary excretion of dimethylthiophosphate, one metabolite of azinphos-methyl. However, workers had been exposed to several other OPs and pesticides. The authors' conclusion was based on a significant increase in mean vibration threshold sensitivity for applicators' hands as compared to a matched control group. Feet were not affected. Long-term exposure was determined by questionnaire, but it is unclear whether poisonings had occurred in the past. Subjective symptoms were collected off and on season: headache, weight loss, and nightmares were reported more frequently among pesticide workers, but only headache was statistically increased during the season. Because toxic polyneuropathy does not exclusively affect upper limbs and because workers were exposed to many chemicals, there is no evidence that long-term lowlevel exposures to OPs cause loss of peripheral nerve functions. A cross-sectional study compared 168 spray operators with long-term exposures to OPs with 84 controls (London et aI., 1997). No evidence was found between exposure and loss of vibration sense. However, small differences were found on neurobehavioral test batteries based on information-processing parameters. The authors concluded that there was a small overall evidence of chronic effects of OP exposures, but indicated that exposure misclassification may have contributed to these findings. In another study, the same authors (London et aI., 1998) investigated neurological symptoms, vibration sense, and tremors in much the same population during the peak spraying season. Eighty-three nonspraying workers were used as the control group. Exposure, as in the previous study, was derived from a job-exposure matrix for pesticides in agriculture. Applicators significantly reported more dizziness, sleepiness, and headache, and had a higher overall neurological symptom score. Vibration sense and tremor outcome were not associated with past longterm OP exposure. A correlation was found between symptoms and either current exposure or episodes of past OP poisoning (see also Section 51.1.3.9). The effects of low-level exposure to foliar OP residues (primarily to azinphos methyl and possibly to phosmet and methylparathion) during one season were assessed in a cross-sectional study on 67 workers and 68 matched controls (Engel et at., 1998). Sensory and motor nerve conduction velocities, neuro-


muscular junction testing, and RBC AChE were measured. No differences were found between exposed and controls. 51.4.1.3 Psychiatric Effects Clinical Reports Schizophrenic and depressive reactions, with severe impairment of memory and difficulty in concentration were reported in 16 workers after variable exposures to OPs (Gershon and Shaw, 1961). This report was criticized because of serious flaws, including the lack of evidence for exposure, the detailed clinical description of only a few cases, and the inconsistency with larger studies (Barnes, 1961; Bidstrup, 1961). An anecdotal report suggested a causal relationship between psychiatric disturbances and exposure to a variety of pesticides including OPs in two pilots (Dille and Smith, 1964). Another anecdotal report linked the onset of psychosis in a farmer with previous spraying of demeton-S-methyl, but no casual relationship was established (Bradwell, 1994). Geographical Studies A geographical study was carried out to determine whether areas of high OP usage in Australia had a higher proportion of admissions for psychiatric disorders than low-usage areas (Stoller et aI., 1965). No evidence was found that schizophrenia, depressive states, psychoneuroses, or personality disorders were more common in high-usage areas than elsewhere. Increased risk of suicide was associated to pesticide exposure (mainly OPs) in an agricultural area (Parr6n et aI., 1996a). The rate of suicides was compared with rates where exposure to pesticides was low. Most suicide cases involved pesticides, but other factors that influenced suicide attitudes were not analyzed. Occupational Exposure Studies Workers who had unspecified exposures to OPs were compared with a control group on personality tests, structured interview, and cholinesterase levels (Levin et aI., 1976). Commercial sprayers, but not farmers, showed higher levels of anxiety and lower plasma cholinesterase than controls. The authors concluded that these findings were tentative until confirmed by additional studies. The effect of exposure stress in the absence of exposure was reported during a manufacturing accident with malathion (Markowitz et aI., 1986). The reactions of allegedly exposed workers were compared with a matched group. The exposed group showed high demoralization scores, particularly among those who admitted to little knowledge about toxic chemicals. Twenty-five greenhouse workers were compared to controls and showed a higher incidence of symptoms of depression and tremors (Parr6n et aI., 1996b). Exposure was not measured and blood cholinesterases were normal. Two groups, one of pesticide formulators (208 individuals) and another of applicators (172 individuals), were compared with matched controls (72 and 151 individuals, respectively; Amr et aI., 1997). Exposures to a variety of pesticides, including OPs, carbamates, organochlorine, and pyrethroids, were not

1071

quantified. Both exposed groups had a higher incidence of total psychiatric disorders, whereas formulators had a higher incidence of depressive neurosis that was related to the duration of employment. It is difficult to assess the role, if any, of the OPs, given the variety of pesticides to which these workers were exposed. A case control study investigated the link between exposure to pesticides and suicide in Canadian farmers (Pickett et al., 1998). Results excluded exposure to pesticides as an important risk factor for suicide among farmers. However, the chemicals used were not identified and were only divided between herbicides and insecticides. The latter certainly included OPs. Therefore, it cannot be ascertained from this study whether exposures to OPs were involved. 51.4.1.4 Neurohehavioral Effects The neurobehavioral effects of long-term exposures to low levels of OPs have been extensively reviewed over the last few years (D'Mello, 1993; ECETOC, 1998; Eyer, 1995; Jamal, 1995; Mearns et aI., 1994; Ray, 1998a, b; Steenland, 1996). Although much information has been published, results are contradictory and whether such exposures are linked with an increased risk of behavioral effects in humans is controversial. A study compared two groups of 53 and 68 asymptomatic workers with varying degrees of unquantified and unspecified exposure to OPs to controls (25 and 22 subjects, respectively) on a complex reaction time test. Results showed there was no indication that exposure at levels insufficient to produce clinical illness had any important effect on mental alertness (Durham et aI., 1965). Another study selected 23 workers who regularly used OPs and had used them within 2 weeks of the testing date (Rodnitzky et aI., 1975). Recent exposure was confirmed by lower plasma cholinesterase, but RBC AChE was normal. Types of pesticides were not reported. The results of tests for memory, signal processing, vigilance, language, and proprioceptive performance were no different from those of a matched control group. Neurobehavioral tests were performed before and after work shifts on 99 pest control workers with low-level, short-term exposure to diazinon (Maizlish et aI., 1987). Exposure was measured by means of the urinary metabolite diethylthiophosphate before and after the end of shifts. No changes in neurobehavioral functions were detected on a battery of seven tests. Similarly, no changes were seen when workers were subdivided according to the degree of exposure. Neuropsychological performance was assessed by test battery in a group of 49 pesticide applicators prior to and I month after the end of a 6-month pesticide spraying season. Results were compared with 40 controls (Daniell et a/., 1992). The nature and extent of pesticide exposure were assessed and reported in another paper (Karr et al., 1992). The comparison of seasonal RBC AChE changes according to exposure levels showed lower cholinesterase among higher exposure groups compared with lesser exposure group. No evidence of sig-

1072

CHAPTER 51


nificant decrements in neuropsychological performance was reported. The neurobehavioral status was assessed in workers and kibbutz residents differently exposed to OPs and other pesticides (Richter et aI., 1992). Subjects were examined during the spraying season and afterward. Most neurobehavioral scores were poorer during the season. Exposure data were not reported and the authors drew attention to other risk factors such as work load and heat stress. In a cross-sectional study, neuropsychological performance in 146 sheep farmers was compared to 143 quarry workers (Stephens et aI., 1995a, b). The selection and the testing procedures for workers who belonged to the experimental group were different from those of controls. Long-term exposure data were assessed by means of a retrospective exposure questionnaire that used the number of sheep, dips, and years of employment as a surrogate. Farmers performed significantly worse than controls in tests to assess sustained attention and speed of information processing (simple reaction time, symbol digit substitution, and syntactic reasoning). A dose-response relationship was found only for one test (syntactic reasoning). Moreover, in another article, no association was found between the experience of acute symptoms and performance on neuropsychological tests (assessed on a subset of workers), and it was concluded that neuropsychological data reflect chronic effects that occur independently of acute effects (Stephens et al., 1996). Given the large differences between OP and quarry workers, it is doubtful that the small changes detected in the former should be attributed to low-level long-term exposures to OPs. Fifty-seven licensed applicators were compared on several neuropsychological tests to a control group of 34 farmers who had no exposure to pesticides (Fiedler et aI., 1997). Exposure to OPs was assessed with a questionnnaire on work history, but details were not given. None of the applicators had episodes of acute poisoning, and RBC AChE values were normal. Except for slower reaction time, no other difference in neuropsychological performances was detected between exposed and nonexposed subjects. Subclinical morbidity patterns, including symptoms, aiming at digit symbol tests, and measurement of RBC AChE, were investigated in 226 established farm workers and were compared with an equal number of controls and with 92 new farm workers (Gomes et al., 1998). Results indicated a higher incidence of symptoms (irritated conjunctiva, watery eyes, blurred vision, dizziness, headache, and muscular pain and weakness), reduced performance on the aiming and digit symbol tests, and reduced AChE activity in the group of established farmworkers. Although RBC AChE inhibition implies OP exposure, no actual exposure data were reported. Moreover, reduction of AChE was still within the coefficient of variation of the test. Nevertheless, because the above-reported symptoms are consistent with cholinergic overstimulation, it is likely that differences between groups reflect the effects of current exposures to OPs.

51.4.2 OTHER EFFECTS

Several toxic effects have been associated with long-term exposures to OPs. However, most of them are either isolated reports or are based on circumstantial evidence of exposure and probably are simply coincidental. A case report, for instance, suggested congestive cardiomyopathy caused by long-term exposure to OPs without signs of severe acute poisoning. Evidence of exposure was limited and the patient had a previous myocardial infarction, therefore, cardiomyopathy was likely a consequence of myocardial infarction rather than OP exposure (Fazekas and Kiss, 1980). Scveral cases of influenza-like symptoms associated with OP use in farmers during the sheep dipping season apparently were unrelated to RBC AChE inhibition (Murray et aI., 1992). Contact dermatitis and asthma have been linked to exposures to OP pesticides (Bryant, 1985; Deschamps et al., 1994; Xue, 1992). Whether the effects are due to sensitization or irritation, or if they are linked to other ingredients always present in commercial formulations of pesticides is unclear. Immunotoxicity of OPs has been suggested, although human data are mostly based on in vitro studies (Newcombe and Esa, 1992; Rodgers et aI., 1992; Sharma and Tomar, 1992). Hypotheses propose that exposures to OPs are linked to cancer development (Newcombe, 1992). One study was performed on workers engaged in the production of several OPs (trichlorfon, chlorfenvinphos, malathion, dichlorvos, fenitrothion, and formothion; Hermanowicz and Kossman, 1984). Exposure involved other chemicals and its assessment was rather approximate. RBC AChE and plasma ChE were reported on a group basis and correlated with the assessment of exposure. A marked impairment in neutrophil chemotaxis was found in workers who were likely to be exposed to OPs. The frequency of upper respiratory tract infections was higher in the exposed group compared with controls. No other types of infections showed increased frequency. The authors themselves concluded that a distinction cannot be made between OPs and other chemicals as possible factors for the described effects. Certainly more information is needed to ascertain whether immunotoxicity is an effect of OPs and what pathophysiological significance should be attributed to it. There is no evidence so far that any form of immunomediated clinical effect is linked to OP exposures. 51.4.3 BIOMONITORING OCCUPATIONAL EXPOSURES

Long-term occupational exposures to OPs are commonly monitored by measuring either urinary excretion of alkylphosphates or blood cholinesterases. The goal is to prevent adverse effects from OPs. 51.4.3.1 Assessment of Urinary Alkylpbospbates

Although OPs may be excreted unchanged, they are usually hydrolyzed, and the acidic and alcoholic moieties can be found in the urine of exposed subjects. Measurement of metabolites


is common practice in workers exposed to OPs, and several alkylphosphates have been identified, including dimethylphosphates, dimethy lthiophosphates, dimethy ldithiophosphates, and dimethylphosphorothioates derived from dimethylated OPs and the corresponding metabolites derived from diethylated OPs (Coye et aI., 1986). Several gas-chromatographic methods to measure urinary dialkylphosphates have been developed (Aprea et al., 1996; Nutley and Cocker, 1993). Measurement of excretion of the alcoholic moiety in exposed workers has been used less frequently. Examples include the measurement of 3,5,6-trichloropyridinol after exposure to chlorpyrifos and chlorpyrifos-methyl (Nolan et al., 1984), p-nitrophenol after exposure to parathion and parathion-methyl (Morgan et al., 1977), and mono- and dicarboxylic acid after exposure to malathion (Bradway and Shafik, 1977). Despite the numerous field studies where exposures have been assessed by means of urinary metabolites, not many data are suitable for a toxicological interpretation. The reasons are many. The first problem is related to usual agricultural practices, which lead to concurrent exposures to several OPs. As stated earlier, different OPs, each with its own toxicity, may be metabolized, yielding the same product. It is, therefore, difficult to assess the toxicological risk associated with such exposures. For instance, certain exposures to either parathionmethyl or chlorpyrifos-methyl caused comparable excretion of dimethylphosphates and dimethylthiophosphates. However, the risk derived from each OP is quite different, because they display a 3 orders of magnitude difference in acute toxicity (Moretto et aI., 1995). Depending on the OP, route of exposure, metabolism, and distribution, peak metabolite excretion might be reached at different times after the end of exposure. For instance, certain compounds such as chlorpyrifos show peak urinary excretion of ethylphosphates several hours after the end of exposure (Fenske and Elkner, 1990; Moretto et al., 1995). On the contrary, peak excretion of ethylphosphates derived from exposures to parathion occurs within a much shorter time (Morgan et aI., 1977). Moreover, alkylphosphates may have a different time course of excretion. For instance, diethylphosphate peaks earlier than diethylphosphorothioate after exposure to diazinon (Sewell et al., 1999). Therefore, the timing of urine sampling is crucial to assess the significance of a given concentration; for many OPs, the relevant timing information is missing. Finally, very little is known about the correlation between urinary metabolite excretion and the inhibition of AChE and/or plasma BuChE. In many cases, enzyme inhibition was not found (Griffin et aI., 1999; Kraus et aI., 1977, 1981; Krieger and Thongsinthusak, 1993; Popendorf et aI., 1979); in a few cases, minimal inhibition was found (Jauhiainen et aI., 1992; Spear et aI., 1977). Only at a time when occupational exposures were probably much more severe than nowadays, was a good correlation found between p-nitrophenol and RBC AChE inhibition in parathion exposed workers (Arterberry et aI., 1961). In conclusion, for the time being, it is difficult to give these data a toxicological significance beyond that of being a qualitative exposure index.

1073

51.4.3.2 Monitoring Blood Cholinesterase Blood cholinesterase activities have been used extensively to monitor the effects of occupational exposures to OPs. Guidelines have been developed on methods, interpretation of results, and actions to be taken (EPA, 1992; Plestina, 1984; WHO, 1986). However, these suggestion should be taken as general indications, particularly when interpreting single data, because the following issues must be considered (Lotti, 1995). Relationship between Inhibition of Blood Enzymes and Cholinergic Toxicity The postulate for using blood cholinesterases to biomonitor occupational exposures to OPs is that inhibition of these enzymes reflects either or both the degree of exposure or the corresponding enzyme inhibition in the nervous tissues. Because no physiological functions have been attributed to BuChE (confirmed by the fact that homozygote carriers of defective BuChE are healthy subjects; see succeeding text), the inhibition of this enzyme in any tissue most likely has no significance in terms of health. However, its inhibition in plasma means that exposure has occurred. This statement may not be true in the case of diseases (unlikely in occupational exposures) that cause depression of plasma BuChE, such as parenchimalliver diseases, acute infections, some anemias, and malnutrition. In this respect, an interesting observation is that patients with liver diseases not only have low plasma cholinesterases, but also may show a further reduction as a result of a level of exposure to OPs that causes no change of enzyme activity in normal persons or in persons affected by other diseases (Hayes, 1982). A possible explanation is the reduced ability of patients affected by liver diseases to detoxify certain OPs. Obviously, if BuChE inhibition is associated with inhibition of RBC AChE, then different conclusions should be drawn. How accurately inhibition of RBC AChE reflects that in the synapses is unknown and extrapolation is difficult, given the different access various OPs have to the blood and the nervous system. Animal data suggest that sometimes inhibition is similar, more often, inhibition of blood enzyme is higher (Su et aI., 1971) due to the particular protection of the nervous system that is offered by the blood-brain barrier. After recovery from exposure, extrapolation is even more difficult, given the different rates of recovery of AChE in the RBCs and nervous tissues, respectively. Nevertheless, based on clinical and toxicological data, a rough estimate of the levels of RBC AChE inhibition that require action are reported in the Table 51.10. It is also clear that because the access of xenobiotics to blood is always easier than to brain and because no evidence exists that OPs accumulate in the nervous system, the inhibition of RBC AChE usually overestimates the level in the brain. Spontaneous Reactivation and Reappearance of Blood Enzymes When measuring blood cholinesterase for biomonitoring purposes, the rates of reappearance of activities after inhibition should be taken into account. Such rates depend on spontaneous reactivation (in the case of carbamates and

1074

CHAPTER 51


Table 51.10 Relationship between RBC AChE Inhibition and Preventive Actions When Monitoring Occupational Exposures to opsa

Table 51.11 Sensitivity of Plasma BuChE and RBC AChE in Humans to Various Insecticides a

RBCAChE

RBC AChE most inhibited

(% inhibition from

Dimefox (Edson, 1964)

Chlorfenvinphos (FAOIWHO, 1995)

Mevinphos (Rider et aI., 1972, 1975)

Chlopyrifos (Eliason et aI., 1969)

Improve hygenic conditions

Methyl-parathion (Rider et aI., 1970)

Demeton (Moeller and Rider, 1965)

As above plus removal of

Parathion (Hayes, 1982)

preexposure values)

Significance

Preventive action

20-29

Evidence of exposure

30-50

Hazard

subject from exposure >50

Poisoning

Plasma cholinesterase most inhibited

Diazinon (FAOIWHO, 1967) Dichlorvos (Rasmussen et aI., 1963)

Admit subject to the hospital

Fenitrothion (Vandekar, 1965) Malathion (Elliot and Barnes, 1963)

aData from WHO (1986) and Lotti (1995).

Monocrotophos (FAOIWHO, 1996) Trichlorfon (Abdel-Al et aI., 1970)

dimethylphosphates) and resynthesis of new enzymes. As previously discussed (Section 51.1.2), the rate constants of spontaneous reactivation and of aging vary according to the phosphorylating agent. Because rate constants also depend on the enzyme that is phosphorylated, they will be different when measured in RBC AChE or plasma BuChE (Aldridge and Reiner, 1972). However, the few studies in humans do not necessarily confirm these theoretical considerations. Data from school children treated orally with trichlorfon against schistosomiasis showed that plasma BuChE as well as RBC AChE recovered much slower than was predicted from in vitro spontanous reactivation studies (Reiner and Ple§tina, 1979). The synthesis of AChE occurs in the bone marrow, and its presence in the blood depends on the normal turnover of RBCs (i.e., 120 days). The synthesis of BuChE occurs in the liver and its turnover in plasma corresponds to about 20 days. Resynthesis of both enzymes after irreversible inhibition by OPs in the nervous systems of animals seems to occur at similar rates, corresponding to a half-life of 5-7 days. However, resynthesis is reflected in the blood quite differently because the localizations of AChE and BuChE differ. Thus, the reappearance of RBC AChE has been shown to occur at a rate of about 1% per day, whereas the rate of plasma BuChE is about 5% per day (Hayes, 1982). Sensitivity of Blood Enzymes to Inhibitors As stated previously, RBC AChE and plasma BuChE are different enzymes, and, therefore, they display different substrate specificity. Whereas interactions of OPs with esterases are analogous to interactions of substrates with esterases, it is clear that blood cholinesterases are differently inhibited by a given ~P. As shown in Table 51.11, plasma BuChE is generally more sensitive to inhibition than RBC AChE by most OPs used as insecticides. In cases of mild exposures, plasma BuChE may be the only inhibited enzyme. This observation should be interpreted as a sign of exposure, but not of poisoning. In cases of severe exposures, profound inhibition of plasma BuChE is always associated with similar inhibition of RBC AChE as it occurs in poisoned patients. The situation may be more complex when substantial repeated exposures are involved. In such a case, even if the plasma enzyme is more sensitive, a buildup of RBC

a Circumstances

of exposure vary.

AChE inhibition can occur, thus equalizing the inhibition of both enzymes at a certain time because of the different rates of reappearance of the two enzymes. Therefore, equal inhibition of the enzymes may represent the consequence of either a single substantial exposure or less severe but repeated exposures. Finally, RBC AChE may be more inhibited, even if the plasma enzyme is more sensitive when subjects are recovering from substantial exposures, given the prolonged life of the RBCs that carry inhibited AChE. Inter- and Intraindividual Variability of Blood Enzymes Intraindividual variability of both plasma BuChE and RBC AChE is high. Samples taken at intervals ranging from a few days to several years indicate that the coefficients of variation of both enzymes in unexposed subjects vary from 7 to 11 %. In some cases, an intraindividual variability of plasma enzyme up to 100% was detected in the course of several months (Hayes, 1982). Interindividual variability of these enzymes is even greater. The coefficients of variation of RBC AChE in unexposed subjects vary from 10 to 40%, whereas the corresponding values for plasma enzyme vary from 12 to 46%. Minor differences exist according to gender, age, and race (Hayes, 1982). A few people who have normal levels of RBC AChE are genetically deprived of plasma BuChE (0stergaard et al., 1992). It was observed that some of these people who were treated with succinylcholine during surgery exhibited an abnormal prolonged period of muscular paralysis after usual dosages of the drug. These patients were found to have a plasma BuChE much lower than normal (Bourne et al., 1952). Moreover, the enzyme is also qualitatively different from the norm, as for instance, in sensitivity to inhibitors. Because of this, a test was developed based upon lesser inhibition of the enzyme by dibucaine (Kalow and Staron, 1957). The dibucaine number (degree of plasma cholinesterase inhibition by dibucaine) discriminates three phenotypes: normal, intermediate, and atypical. The approximate frequency of these phenotypes has been estimated at 96, 3.9, and 0.05%, respectively (Harris and Whittaker, 1962). Other

References

tests are available to discriminate abnormal BuChE, based upon fluoride number (Harris and Whittaker, 1961), chloride number (Whittaker, 1968), scoline number (King and Griffin, 1973), and urea number (Hanel and Viby-Mogensen, 1977; see Section 1.3.6.4).

51.4.3.3 Detection of Hypersusceptible SUbjects Whereas OPs are inhibitors of plasma BuChE and are largely hydrolyzed by A-esterases, such as paraoxonase (PONI) and other carboxylesterases (aliesterases), inherited or acquired deficits of scavenger (plasma BuChE) or detoxifying (esterases) abilities have been suggested as potential factors for increased susceptibility to OPs (Loewenstein-Lichtenstein et aI., 1996; Saxena et aI., 1997). Although there is some evidence in experimental animals for hypersusceptibility based upon reduced ability to detoxify OPs, only one example is known in humans. An unexpected outbreak of malathion poisoning arose in workers occupationally exposed to commercial brands of malathion that contained high amounts of impurities. Among these, isomalathion was the most relevant because it inhibited the carboxylesterases that inactivativate malathion by hydrolyzing its carboxyl-ester linkages (Baker et aI., 1978). Recurrent suggestions have been made that genetically determined low levels of plasma BuChE increased the susceptibility to OP toxicity because of a reduced scavenger capability. However, a relationship between abnormal plasma BuChE and hypersusceptibility to OP poisoning has never been reported. Hypersusceptibility to succinylcholine would never have been discovered if some unusual people had not undergone succinylcholine treatment, an event that is probably no more common than heavy exposure to OPs. Moreover, when the biochemical characteristics of the normal and of the genetically determined defective enzymes were compared stoichiometrically with the plasma concentrations of inhibitors, no scavenger functions to plasma BuChE could be detected (Lotti and Moretto, 1995). Based on the polymorphism of PON 1 in human populations and the known role of this enzyme in the detoxification of some OPs (Davies et al., 1996; Mueller et aI., 1983), it has been inferred that the expression of this enzyme is involved in determining hypersusceptibility to OPs (Mackness et aI., 1998). However, so far, proof for this has been obtained only in animals (Costa et aI., 1999).

REFERENCES Abdel-AI, A. M. A., EI-Hawary, M. F. S., Kamel, H., Abdel-Khalek, M. K., and EI-Diwany, K. M. (1970). Blood cholinesterases, hepatic, renal and haemopoietic functions in children receiving repeated doses of "Dipterex." l. Egypt. Med. Assoe. 53,265-271. Abend, Y., Goland, S., Evron, E., Sthoeger, Z. M., and Geltner, D. (1994). Acute renal failure complicating organophosphate intoxication. Renal. Fail. 16, 415-417. Abou-Donia, M. B., and Lapadula, D. M. (1990). Mechanisms of organophosphorus ester-induced delayed neurotoxicity: Type I and type 11. Ann. Rev. Pharmaco!' Toxicol. 30, 405-440.

1075

Adams, R. G., Verma, P., lackson, A. l., and Miller, R. L. (1982). Plasma pharmacokinetics of intravenously administered atropine in normal human subjects. l. Clin. Pharmacol. 22,477-481. Ahlgren, l. D., Manz, H. l., and Harvey, 1. C. (1979). Myopathy of chronic organophosphate poisoning: A clinical entity? S. Afr. Med. l. 72,555-563. Aiuto, L. A., Pavlakis, S. G., and Boxer, R. A. (1993). Life-threatening organophosphate-induced delayed polyneuropathy in a child after accidental chlorpyrifos ingestion. l. Pediatr. 122, 658-660. Aldridge, W. N., and Reiner, E. (1972). "Enzyme Inhibitors as Substrates." North-Holland, Amsterdam. Alkondon, M., and Albuquerque, E. X. (1989). The nonoximc bispyridinium compound SAD- I 28 alters the kinetic properties of the nicotinic acetylcholine receptor ion channel: A possible mechanism for antidotal effects. l. Pharmacal. Exp. Ther. 250, 842-852. American EIectroencephalographic Society (1987). Statement on the clinical use of quantitative EEG. l. Clin. Neurophysiol. 4, 75. Ames, R. G., Steenland, K., Jenkins, B., ChrisIip, D., and Russo, J. (1995). Chronic neurologic sequelae to cholinesterase inhibition among agricultural pesticide applicators. Arch. Environ. Health 50, 440-443. Amr, M. M., Halim, Z. S., and Moussa, S. S. (1997). Psychiatric disorders among Egyptian pesticide applicators and formulators. Environ. Res. 73, 193-199. Aprea, c., Sciarra, G., and Lunghini, L. (1996). Analytical method for the determination of urinary alkylphosphates in subjects occupationally exposed to organophosphorus insecticides and in the general population. l. Anal. Taxica!. 20,559-563. Aring, C. D. (1942). The systemic nervous affinity of triorthocresyl phosphate (Jamaica Ginger Palsy). Brain 65, 34-47. Arterberry, J. D., Durham, W. F., Elliot, J. w., and Wolfe, H. R. (1961). Exposure to parathion: Measurement by blood cholinesterase level and urinary p-nitrophenol excretion. Arch. Environ. Health 3, 112-121. Arterberry, l. D., Bonifaci, R. w., Nash, E. w., and Quinby, G. E. (1962). Potentiation of phosphorus insecticides by phenothiazine derivatives. Possible hazard, with report of a fatal case. lAMA 182, 848-850. Baker, D. 1., and Sedgwick, E. M. (1996). Single fibre electromyographic changes in man after organophosphate exposure. Human Exp. Toxiea!. 15, 369-375. Baker, E. L., Zack, M., Miles, l. W., Alderman, L., Warren, M., Dobbin, R. D., Miller, S., and Teeters, W. R. (1978). Epidemic malathion poisoning in Pakistan malaria workers. Lancet i, 31-34. Balali-Mood, M., and Shariat, M. (1998). Treatment of organophosphate poisoning. Experience of nerve agents and acute pesticide poisoning on the effects of oximes. l. Physio!. Paris 92, 375-378. Bardin, P. G., and van Eeden, S. F. (1990). Organophosphate poisoning: Grading the severity and comparing treatment between atropine and glycopyrrolate. Crit. Care Med. 18, 956-960. Bardin, P. G., van Eeden, S. F., and Joubert, J. R. (1987). Intensive care management of acute organophosphate poisoning. A 7-year experience in the western Cape. S. Ajr. Med. J 72,593-597. Bardin, P. G., van Eden, S. F., Moolman, l. A., Foden, A. P., and Joubcrt, J. R. (1994). Organophosphate and carbamate poisoning. Arch. Intern. Med. 154, 1433-1441. Bames, J. M. (1961). Psychiatric sequelae of chronic exposure to organophosphorus insecticides. Lancet ii, 102-103 Barr, A. M. (1966). Further experience in the treatment of severe organic phosphate poisoning. Med. l. Aust. 1, 490-492. Benson, B. l., Tolo, D., and McIntire, M. (1992). Is the intermediate syndrome in organophosphate poisoning the result of insufficient oxime therapy? Clin. Toxieal. 30, 347-349. Bentur, Y., Nutenko, 1., Tsipiniuk A., Raikhlin-Eisenkraft, B., and Taitelman, U. (1993). Pharmacokinetics of obidoxime in organophosphate poisoning associated with renal failure. Clin. Toxicol. 31, 315-322. Bertolazzi, M., Caroldi, S., Moretto, A., and Lotti, M. (1991). Interaction of methamidophos with hen and human acetylcholinesterase and neuropathy target esterase. Arch. Toxico!. 65, 580-585. Bertoncin, D., Russolo, A., Caroldi, S., and Lotti, M. (1985). Neuropathy target esterase in human Iymphocytes. Arch. Environ. Health 40, \39-144.

10'16

CHAPTER 51


Besser, R., Gutmann, L., and Weilemann, L. S. (1989a). Inactivation of end-plate acetylcholinesterase during the course of organophosphate intoxications. Arch. Toxicol. 63,412-415. Besser, R., Gutmann, L., Dillmann, U., Weilemann, L. S., and Hopf, H. C. (1989b). End-plate dysfunction in acute organophosphate intoxication. Neurology 39,561-567. Besser, R., Vogt, T., and Gutmann, L. (1990). Pancuronium improves the neuromuscular transmission defect of human organophosphate intoxication. Neurology 40, 1275-1277. Besser, R., Vogt, T., Gutmann, L., and Wessler, 1. (1991). High pancuronium sensitivity of axonal nicotinic-acetylcholine receptors in humans during organophosphate intoxication. Muscle Nerve 14, 1197-1201. Besser, R., Weilemann, L. S., and Gutmann, L. (1995). Efficacy of obidoxime in human organophosphorus poisoning: Determination by neuromuscular transmission studies. Muscle Nerve 18, 15-22. Betrosian, A., Balla, M., Kafiri, G., Kofinas, G., Makri, R., and Kakouri, A (1995). Multiple systems organ failure from organophosphate poisoning. CZin. Toxicol. 33, 257-260. Bidstrup, P. L. (1961). Psychiatric sequelae of chronic exposure to organophosphorus insecticides. Lancet ii, 103. Bidstrup, P. L., Bonnell, J. A, and Beckett, A. G. (1953). Paralysis following poisoning by a new organic phosphorus insecticide (mipafox). Report on two cases. Brit. Med. J. 1, 1068-1072. Bismuth, C., Inns, R. H., and Marrs, T. C. (1992). Efficacy, toxicity and clinical use of oximes in anticholinesterase poisoning. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 555-577. Butterworth-Heinemann, Oxford. Bjornsd6ttir, U. S., and Smith, D. (1999). South African religious leader with hyperventilation, hypophosphataemia, and respiratory arrest. Lancet 354, 2130. Boskovic, B., Kovacevic, v., and Jovanovic, D. (1984). PAM-2 Cl, HI-6, and HGG-12 in soman and tabun poisoning. Fundam. Appl. Toxicol. 4, SI06S115. Bourne, J. E., Collier, H. O. J., and Somers, G. F. (1952). Succinylcholine (succinoylcholine). Muscle relaxant of short action. Lancet i, 1225-1229. Bowers, M. B., Jr., Goodman, E., and Sim, V. M. (1964). Some behavioral changes in man following anticholinesterase administration. J. Nerv. Mental. Dis. 138,383-389. Bradway, D. E., and Shafik, T. M. (1977). Malathion exposure studies. Determination of mono- and dicarboxylic acids and alkyl phosphates in urine. Agr. Food Chem. 25, 1342-1344. Bradwell, R. H. (1994). Psychiatric sequelae of organophosphorus poisoning: A case study and review of the literature. Behav. Neurol. 7, 117-122. Brill, D. M., Maisel, A. S., and Prabhu, R. (1984). Polymorphic ventricular tachycardia and other complex arrhythmias in organophosphate insecticide poisoning. J. Electracardial. 17,97-102. Brown, J. H., and Taylor, P. (1996). Muscarinic receptor agonists and antagonists. In "Goodman and Gilman's the Pharmacological Basis of Therapeutics" (J. G. Hardman and L. E. Limbird, eds.), 9th ed., pp. 141-160. McGraw-Hill, New York. Brown, M. A., and Brix, K. A. (1998). Review of health consequences from high-, intermediate- and low-level exposure to organophosphorus nerve agent. J. Appl. Taxicol. 18, 393-408. Bryant, D. H. (1985). Asthma due to insecticide sensitivity. Aust. N. Z. J. Med. 15,66-68. Buckley, N. A, Dawson, A. H., and Whyte, 1. M. (1994). Organophosphate poisoning: Peripheral vascular resistance-a measure of adequate atropinization. CZin. Toxical. 32, 61-ti8. Burchfiel, J. L., Duffy, F. H., and Sim, V. M. (1976). Persistent effects of sarin and dieldrin upon the primate electroencephalogram. Taxical. Appl. Pharmacal. 35, 365-379. Calesnick, B., Christensen, J. A., and Richter, M. (1967). Human toxicity of various oximes. 2-Pyridine aldoxime methyl chloride, its methane sulfonate salt, and 1,1'-trimethylenebis-(4-formylpyridinum chloride). Arch. Environ. Health 15, 599-608.

Capodicasa, E., Scapellato, M. L., Moretto, A., Caroldi. S., and Lotti, M. (1991). Chlorpyrifos-induced delayed polyneuropathy. Arch. Toxicol. 65, 150--155. Castillo, E., Rubin, R. T., and Holsboer-Trachsler, E. (1989). Clinical differentiation between lethal catatonia and neuroleptic malignant syndrome. Am. J. Psychiatry 146, 324-328. Catz, A., Chen, B., Jutrin, 1., and Mendelson, L. (1988). Late onset of isofenphos neurotoxicity. J. Neural. Neurosurg. Psychiatry 51, 1338-1340. Cavanagh, J. B. (1973). Peripheral neuropathy caused by toxic agents. Crit. Rev. Taxical. 2,365-417. Chambers, H. W. (1992). Organophosphorus compounds: An overview. In "Organophosphates. Chemistry, Fate, and Effects" (J. E. Chambers and P. E. Levi, eds.), pp. 3-17. Academic Press, San Diego. Choi, P. T.-L., Quinonez, L. G., Cook, D. J., Baxter, F., and Whitehead, L. (1998). The use of glycopyrrolate in a case of intermediate syndrome following acute organophosphate poisoning. Can. 1. Anaesth. 45, 337-340. Chuang, E R., Jang, S. w., Lin, J. L., Chern, M. S., Chen, J. B., and Hsu, K. T. (1996). QTc prolongation indicates a poor prognosis in patients with organophosphate poisoning. Am. J. Emerg. Med. 14,451-453. Compton, J. A E (1987). "Military Chemical and Biological Agents." Telford Press, Caldwell, NJ. Costa, L. G., Li, W. E, Richter, R. J., Shih, D. M., Lusis, A, and Furlong, C. E. (1999). The role of paraoxonase (PONI) in the detoxication of organophosphates and its human polymorphism. Chem.-Bial. Interact. 119120, 429-438. Coudray-Lucas, c., Prioux-Guyonneau, M., Sentenac, H., Cohen, Y., and Wepierre, J. (1983). Brain catecholamine metabolism changes and hypothermia in intoxication by anticholinesterase agents. Acta Pharmacal. Taxical. 52,224-229. Coye, M. J., Lowe, J. A, and Maddy, K. J. (1986). Biological monitoring of agricultural workers exposed to pesticides: n. Monitoring of intact pesticides and their metabolites. J. Occup. Med. 28, 628-636. Csik, v., Motika, D., and Marosi, G. Y. (1986). Delayed neuropathy after trichlorfon intoxication. J. Neurol. Neurosurg. Psychiatry 49, 222. Curtes, J. P., Develay, P., and Hubert, J. P. (1981). Late peripheral neuropathy due to an acute voluntary intoxication by organophosphorus compounds. CZin. Toxicol. 18, 1453-1462. Dagli, A. J., and Shaikh, W. A. (1983). Pancreatic involvement in malathionanticholinesterase insecticide intoxication: A study of 75 cases. Br. J. CZin. Pract. 37, 270--272. Daniell, w., Barnhart, S., Demers, P., Costa, L. G., Eaton, D. L., Miller, M., and Rosenstock, L. (1992). Neuropsychological performance among agricultural pesticide applicators. Enviran. Res. 59, 217-228. Davies, H. G., Richter, R. J., Keifer, M., Broomfield, C. A., Sowalla, J., and Furlong, C. E. (1996). The effect of human serum paraoxonase polymorphism is reversed with diazoxon, soman and sarin. Nat. Genet. 14,334-336. Davies, J. E., Barquet, A., Freed, V. H., Haque, R., Morgade, C., Sonnebom, R. E., and Vaclavek, C. (1975). Human pesticide poisonings by a fat-soluble organophosphate insecticide. Arch. Environ. Health 30, 608613. Davies, P., and Landy, M. (1998). Suxamethonium and mivacurium sensitivity from pregnancy-induced plasma cholinesterase deficiency. Anaesthesia 53, 1109-1116. Davis, K. L., Yesavage, J. A, and Berger, P. A. (1978). Possible organophosphate-induced parkinsonism. J. Nerv. Ment. Dis. 166, 222-225. Davis, K. L., Thai, L. J., Gamzu, E. R., Davis, C. S., Woolson, R. E, Gracon, S. 1., Drachman, D. A, Schneider, L. S., Whitehouse, P. J., Hoover, T. M., Morris, J. c., Kawas, C. H., Knopman, D. S., Earl, N. L., Kumar, v., Doody, R. S., and the Tacrine Collaborative Study Group. (1992). A double blind, placebo-controlled multicenter study of tacrine for Alzheimer's disease. N. Engl. J. Med. 327, 1253-1259. Dawson, A., Buckley, N., and Whyte, 1. (1997). What target pralidoxime concentration? CZin. Toxicol. 35, 227-228. Dawson, R. M. (1994). Review of oximes available for treatment of nerve agent poisoning. J. Appl. Taxical. 14,317-331. De Bleecker, J. L. (1992). Transient opsoclonus in organophosphate poisoning. Acta Neurol. Scand. 86,529-531.

References

De Bleecker, J. L. (1993). Intermediate syndrome: Prolonged cholinesterase inhibition. Clin. Toxicol. 31, 197-199. De Bleecker, J. L., De Reuck, J. L., and WiIlems, J. L. (1992a). Neurological aspects of organophosphate poisoning. Clin. Neurol. Neurosurg. 94, 93103. De Bleecker, J., Vogelaers, D., Ceuterick, C., Van Den Neucker, K., WiIlems, J. L., and De Reuck, J. L. (1992b). Intermediate syndrome due to prolonged parathion poisoning. Acta Neurol. Scand. 86,421-424. De Bleecker, J., Van Den Neucker, K, and Colardyn, E (1993). Intermediate syndrome in organophosphorus poisoning: A prospective study. Crit. Care Med. 21,1706-1711. de Jager, A. E. J., van Weerden, T. W., Houthoff, H. J., and de Monchy, J. G. R. (1981). Polyneuropathy after massive exposure to parathion. Neurology 31, 603-605. de Jong, L. P. A., and Ceulen, D. I. (1978). Anticholinesterase activity and rate of decomposition of some phosphylated oximes. Biochem. Pharmacol. 27, 857-863. de Kort, W. L. A. M., Kiestra, S. H., and Sangster, B. (1988). The use of atropine and oximes in organophosphate intoxications: A modified approach. Clin. Toxicol. 26, 199-208. Dementi, B. (1994). Ocular effects of organophosphates: A historical perspective of Saku disease. 1. Appl. Toxicol. 14,119-129. de Reuck, J., and WiIlems, J. (1975). Acute parathion poisoning: Myopathic changes in the diaphragm. 1. Neurol. 208,309-314. de Reuck, J., Colardyn, E, and WiIlems, J. (1979). Fatal encephalopathy in acute poisoning with organophosphorus insecticides. A cIinico-pathologic study of two cases. Clin. Neurol. Neurosurg. 81, 247-254. Deschamps, D., Questel, E, Baud, E J., Gervais, P., and Dally, S. (1994). Persistent asthma after acute inhalation of organophosphate insecticide. Lancet 344, 1712. de Silva, H. J., Wijewickrema, R., and Senanayake, N. (1992). Does pralidoxime affect outcome of management in acute organophosphorus poisoning? Lancet 339, 1136-1138. de Silva, H. J., Sanmuganathan, P. S., and Senanayake, N. (1994). Isolated bilateral recurrent laryngeal nerve paralysis: A delayed complication of organophosphorus poisoning. Hum. Exp. Toxico!. 13,171-173. Dettbarn, W. D. (1992). Anticholinesterase induced myonecrosis. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. BaIlantyne and T. C. Marrs, eds.), pp. 167-179. ButterworthHeinemann, Oxford. De Wilde, v., Vogelaers, D., Colardyn, E, Vanderstraeten, G., Van Den Neurcker, K., De Bleecker, J., De Reuck, J., and Van den Heede, M. (1991). Postsynaptic neuromuscular dysfunction in organophosphate induced intermediate syndrome. Klin. Wochenschr. 69,177-183. DiIle, J. R., and Smith, P. W. (1964). Central nervous system effects of chronic exposure to organophosphates insecticides. Aerospace Med. 35, 475-478. D'MeIlo, G. D. (1993). Behavioural toxicity of anticholinesterases in humans and animals-A review. Hum. Exp. Toxico!. 12,3-7. Doctor, B. P., Toker, L., Roth, E., and Silman, I., (1987). Microtiter assay for acetylcholinesterase. Ana!' Biochem. 166, 399-403. Drenth, H. J., Ensberg I. E G., Roberts, D. v., and Wilson, A. (1972). Neuromuscular function in agricultural workers using pesticides. Arch. Environ. Health 25, 395-398. Dressel, T. D., Goodale, R. L., Jr., Ameson, M. A., and Bomer, J. W (1979). Pancreatitis as a complication in anticholinesterase insecticide intoxication. Ann. Surg. 189, 199-204. Duffy, E H., and Burchfiel, J. L. (1980). Long term effects of the organophosphate sarin on EEGs in monkeys and humans. Neurotoxicology 1,667-689. Duffy, EH., Burchfiel, J. L., Bartels, P. H., Gaon, M., and Sim, V. M. (1979). Long-term effects of an organophosphate upon the human electroencephalogram. Toxicol. Appl. Pharmacol. 47, 161-176. Duffy, EH., Bartels, P. H., and Neff, R. (1986). A response to Oken and Chiappa. Ann. Neuro!. 19, 494-496. Durham, W. E, and Hayes, W. J., Jr. (1962). Organic phosphorus poisoning and its therapy. Arch. Environ. Health 5, 21-47.

1077

Durham, W E, Wolfe, H. R., and Quinby, G. E. (1965). Organophosphorus insecticides and mental alertness. Studies in exposed workers and in poisoning cases. Arch. Environ. Health 10, 55-66. ECETOC (1998). "Organophosphorus Pesticides and Long-Term Effects on the Nervous System." Tech. Rep. 75, European Center for Ecotoxicology and Toxicology of Chemicals, Brussels. Ecobichon, D. J., Ozere, R. L., Reid, E., and Crocker, J. E S. (1977). Acute fenitrothion poisoning. Can. Med. Assoc. 1. 116, 377-379. Edmundson, R. S. (1988). "Dictionary of Organophosphorus Compounds." Chapman and Hall, London. Edson, E. E (1964). No-effect levels of three organophosphates in the rat, pig, and man. Food Cosmet. Toxico!. 2, 311-316. Eliason, D. A., Cranmer, M. E, von Windeguth, D. L., Kilpatrick, J. W, Suggs, J. E., and Schoof, H. E (1969). Dursban premises application and their effect on the cholinesterase levels in spraymen. Mosq. News 29, 591595. EIIiot, R., and Barnes, J. M. (1963). Organophosphorus insecticides for the control of mosquitos in Nigeria. Bull. WH.O. 28, 35-54. EIlman, G. L., Courtney, K D., Andres, v., Jr., and Featherstone, R. M. (1961). A new and rapid colorimetric determination of acetylcholinesterase activity. Biochem. Pharmacol. 7,88-95. Engel, A. G., Lambert E. H., and Santa, T. (1973). Study of long-term anticholinesterase therapy. Effects on neuromuscular transmission and on motor end-plate fine structure. Neurology 23, 1273-1281. Engel, L. S., Keifer, M. c., Checkoway, H., Robinson, L. R., and Vaughan, T. L. (1998). Neurophysiological function in farm workers exposed to organophosphate pesticides. Arch. Environ. Health 53, 7-14. EPA (1992). In "Proceedings of the EPA Workshop on Cholinesterase Methodologies," Office of Pesticide Programs, United States Environmental Protection Agency, Washington, DC. Erdmann, W. D., Zech, R., Franke, P., and Bosse, I. (1966). Zur Frage der therapeutischen Wirksarnkeit von Esterase-Reaktivatoren bei der Vergiftung mit Dimethoat. Arzneimittelforschung 16, 492-494 (in German). Erikson-Lamy, K, and Grant, W M. (1992). Ophthalmic toxicology of anticholinesterases. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. BaIlantyne and T. C. Marrs, eds.), pp. 180-194. Butterworth-Heinemann, Oxford. Eyer, P. (1995). Neuropsychopathological changes by organophosphorus compounds-A review. Hum. Exp. Toxico!. 14, 857-864. FAOIWHO (1967). Evaluations of some pesticide residues in food. 1966. In "Joint Meeting of the Food and Agriculture Organization Working Party and the World Health Organization Expert Committee on Pesticide Residues," World Health Organization, Geneva, WHOIF00dAdd./667.32. FAOIWHO (1987). Pesticide residues in food, 1986. In "Evaluations of the Joint Meeting of the Food and Agriculture Organization Panel of Experts on Pesticide Residues in Food and Environment and the Food and Agriculture Organization Expert Group on Pesticide Residues. Part II-Toxicology," Food and Agriculture Organization, Rome, Plant Production and Protection Paper 7812, pp. 90-91. FAOIWHO (1995). Pesticide residues in food, 1994. Report. In "Joint Meeting of the Food and Agriculture Organization Panel of Experts on Pesticide Residues in Food and Environment and the Food and Agriculture Organization Expert Group on Pesticide Residues," Food and Agriculture Organization, Rome, FAO Plant Production and Protection Paper 127. FAOIWHO (1996). "Pesticide residues in food, 1995. In "Evaluations of the Joint Meeting of the Food and Agriculture Organization Panel of Experts on Pesticide Residues in Food and Environment and the Food and Agriculture Organization Expert Group on Pesticide Residues. Part IIToxicological and Environmental," World Health Organization, Geneva, WHOIPCS/96.48, pp. 141-142. FAOIWHO (1997). Pesticide residues in food, 1996. In "Evaluations of the Joint Meeting of the Food and Agriculture Organization Panel of Experts on Pesticide Residues in Food and Environment and the Food and Agriculture Organization Expert Group on Pesticide Residues. Part II-Toxicological and Environmental," World Health Organization, Geneva, WHOIPCS/97. I, pp. 187-188.

1078

CHAPTER 51


Fazekas, T., and Kiss, Z. (1980). Organophosphate cardiomyopathy. Congestive cardiomyopathy caused by long-term organic phosphoric acid ester exposure. Z. Kardiol. 69,584-586 (in German). Fenske, R. A., and Elkner, K. P. (1990). Multi-route exposure assessment and biological monitoring of urban pesticide applicators during structural control treatments with chlorpyrifos. Toxicol. Ind. Health 6, 349-371. Fernandez, G., Dfaz G6mez, M. I., and Castro J. A. (1975). Cholinesterase inhibition by phenothiazine and nonphenothiazine antihistaminics: Analysis of its postulated role in synergizing organophosphate toxicity. Toxicol. Appl. Pharmacol. 31, 179-190. Fiedler, N., Kipen, H., Kelly-McNeil, K., and Fenske, R. (1997). Long-term use of organophosphates and neuropsychological performance. Am. l. Ind. Med. 32,487-496. Fisher, J. R. (1977). Guillain-Barre syndrome following organophosphate poisoning. lAMA 238, 1950-1951. Futagami, K., Otsubo, K., Nakao, Y., Aoyama, T., Iimori, E., Urakami, S., Ide, M., and Oishi, R. (1995). Acute organophosphate poisoning after disulfoton ingestion. Clin. Toxicol. 33, 151-155. Gadoth, N., and Fisher, A. (1978). Late onset of neuromuscular block in organophosphorus poisoning. Ann. Intern. Med. 88, 654-655. Garcia-Lopez, J. A., and Monteoliva, M. (1988). Physiologial changes in human erythrocyte cholinesterase as measured with the "pH-stat." Clin. Chem. 34,2133-2135. Gershon, S., and Shaw, E H. (1961). Psychiatric sequelae of chronic exposure to organophosphorus insecticides. Lancet i, 1371-1374. Gesztes, T. (1966). Prolonged apnoea after suxamethonium injection associated with eye drops containing an anticholinesterase agent. Br. l. Anaesth. 38, 408-409. Glow, P. H., and Rose, S. (1965). Effects of reduced acetylcholinesterase levels on extinction of a conditioned response. Nature 206, 475-477. Goldstein, D. A., McGuigan, M. A., and Ripley, B. D. (1988). Acute tricresylphosphate intoxication in childhood. Human Toxicol. 7, 179-182. Golsousidis, H., and Kokkas, V. (1985). Use of 19 590 mg of atropine during 24 days of treatment, after a case of unusually severe parathion poisoning. Human Toxicol. 4, 339-340. Gomes, J., Lloyd, 0., Revitt, M. D., and Basha, M. (1998). Morbidity among farm workers in a desert country in relation to long-term exposure to pesticides. Scand. l. Environ. Health 24, 213-219. Good, J. L., Khurana, R. K., Mayer, R. E, Cintra, W. M., and Albuquerque, E. X. (1993). Pathophysiological studies of neuromuscular function in subacute organophosphate poisoning induced by phosmet. l. Neurol. Neurosurg. Psychiatry 56, 290-294. Goswamy, R., Chaudhuri, A., and Mahashur, A. A. (1994). Study of respiratory failure in organophosphate and carbamate poisoning. Heart Lung 23, 466472. Granacher, R. P., and Baldessarini, R. J. (1975). Physostigmine. Its use in acute anticholinergic syndrome with antidepressant and antiparkinson drugs. Arch. Gen. Psychiatry 32,375-380. Gray, G. c., Coate, B. D., Anderson, C. M., Kang, H. K., Berg, S. W., Wignail, E S., Knoke, J. D., and Barrett-Connor, E. (1996). The postwar hospitalization experience ofU.S. veterans of the Persian Gulf War. N. Engl. l. Med. 335, 1505-1513. Green, M. D., Jones, D. E., and Hilmas, D. E. (1985). Sarin intoxication elevates plasma pralidoxime. Toxicol. Let!. 28, 17-21. Greenaway, c., and Orr, P. (1996). A foodborne outbreak causing a cholinergic syndrome. l. Emerg. Med. 14, 339-344. Greenberger, N. J., Toskes, P. P., and Isselbacher, K. J. (1998). Acute and chronic pancreatitis. In "Harrison's Principles of Internal Medicine" (A. S. Fauci, E. Braunwald, K. J. Isselbacher, J. D. Wilson, J. B. Martin, D. L. Kasper, S. L. Hauser, and D. L. Longo, eds.), 14th ed., pp. 17411752. McGraw-Hill, New York. Griffin, P., Mason, H., Heywood, K., and Cocker, J. (1999). Oral and dermal absorption of chlorpyrifos: A human volunteer study. Occup. Environ. Med. 56, 10-13. Grob, D. (1956). The manifestations and treatment of poisoning due to nerve gas and other organic phosphate anticholinesterase compounds. Arch. Intern. Med. 98, 221-239.

Grob, D., and Harvey, J. C. (1958). Effects in man of the anticholinesterase compound sarin (isopropyl methyl phosphonofluoridate). l. Clin. Invest. 37, 350-368. Grob, D., Harvey, A. M., Langworthy, O. R., and Lilienthal, J. L., Jr. (1947). The administration of di-isopropyl fluorophosphate (DFP) to man. Ill. Effect on the central nervous system with special reference to electrical activity of the brain. Bull. lohns Hopkins Hosp. 81, 257-266. Guillermo, E P., Pretel, C. M. M., Royo, E T., Macias, M. J. P., Ossorio, R. A., Gomez, J. A. A., and Vidal, J. C. (1988). Prolonged suxamethoniuminduced neuromuscular blockade associated with organophosphate poisoning. Br. l. Anaesth. 61, 233-236. Gunderson, C. H., Lehmann, C. R., Sidell, ER., and Jabbari, B. (1992). Nerve agents: A review. Neurology 42, 946-950. GUven, M., Dnliihizarci, K., Gokta~, Z., and Kurto3lu, S. (1997). Intravenous organophosphate injection: An unusual way of intoxication. Human Exp. Toxicol. 16, 279-280. Haley, R. W. and Kurt, T. L. (1997). Self-reported exposure to neurotoxic chemical combinations in the Gulf War. A cross sectional epidemiologic study. lAMA 277,231-237. Haley, R. W., Kurt, T. L., and Horn, J. (1997a). Is there a Gulf War Syndrome? Searching for syndromes by factor analysis of symptoms. lAMA 277, 215222. Haley, R. W. Horn, J., Roland, P. S., Bryan, W. w., VanNess, P. c., Bonte, E J., Devous, M. D., Mathews, D., Fleckenstein, J. L., Wians, EH., Wolfe, G. 1., and Kurt, T. L. (1997b). Evaluation of neurologic function in Gulf War Veterans. A blinded case-control study. lAMA 277, 223-230. Hanel, H. K., and Viby-Mogensen, J. (1977). The inhibition of serum cholinesterase by urea. Mechanism of action and application in the typing of abnormal genes. Br. l. Anaesth. 49, 1251-1257. Hantson, P., Hainaut, P., Vander Stappen, M., and Mahieu, P. (1996). Regulation of body temperature after acute organophosphate poisoning. Can. l. Anaesth. 43,755. Harris, H., and Whittaker, M. (1961). Differential inhibition of human serum cholinesterase with fluoride. Recognition of two new phenotypes. Nature 191, 469-498. Harris, H., and Whittaker, M. (1962). The serum cholinesterase variants. Study of twenty-two families selected via the "intermediate" phenotype. Ann. Human Genet. 26, 59-72. Harrison, L. I., Smallridge, R. C., Lasseter, K. c., Goldlust, M. B., Shamblen, E. C., Gam, V. w., Chang, S. E, and Kvam, D. C. (1986). Comparative absorption of inhaled and intramuscularly administered atropine. Am. Rev. Respir. Dis. 134, 254-257. Hart, P. S., McCarthy, G. J., Brown, R., Lau, M., and Fisher, D. M. (1995). The effect of plasma cholinesterase activity on mivacurium infusion rates. Anesth. Analg. 80,760-763. Hata, S., Bernstein, E., and Davis, L. E. (1986). Atypical ocular bobbing in acute organophosphate poisoning. Arch. Neurol. 43, 185-186. Hatta, K., Miura, Y., Asukai, N., and Hamabe, Y. (1996). Amnesia from sarin poisoning. Lancet 347, 1343. Haubenstock, A., Hruby, K., and Jager, U. (1983). More on the triad ofpancreatitis, hyperamylasemia, and hyperglycemia. lAMA 249, 1563. Hayes, M. M. M., Van der Westhuizen, N. G., and Gelfand, M. (1978). Organophosphate poisoning in Rhodesia. A study of the clinical features and management of 105 patients. S. Afr. Med. 1. 54,230-234. Hayes, W. J., Jr. (1982). "Pesticides Studies in Man." William and Wilkins, Baltimore. Hayes, W. J., Jr., and Laws, E. R. (1991). "Handbook of Pesticide Toxicology." Academic Press, San Diego. He, E, Xu, H., Qin, E, Xu, L., Huang, J., and He, X. (1998). Intermediate myasthenia syndrome following acute organophosphates poisoning-An analysis of21 cases. Human Exp. Toxicol. 17,40-45. Heath, A. J. w., and Meredith, T. (1992). Atropine in the management of anticholinesterase poisoning. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 543-554. Butterworth-Heinemann, Oxford. Hermanowicz, A., and Kossman, S. (1984). Neutrophil function and infectious disease in workers occupationally exposed to phosphoorganic pesticides:

References

Role of mononuclear-derived chemotactic factor for neutrophils. Clin. Immunol. Immunopathol. 33, 13-22. Hierons, R., and Johnson, M. K (1978). Clinical and toxicological investigations of a case of delayed neuropathy in man after acute poisoning by an organophosphorus pesticide. Arch. Toxicol. 40, 279-284. Hinderling, P. H., Gundert-Remy U., and Schmidlin, O. (1985). Integrated pharmacokinetics and pharmacodynamics of atropine in healthy humans I: Pharmacokinetics. 1. Phann. Sci. 74,703-710. Holland, P., and Parkes, D. C. (1976). Plasma concentrations of the oxime pralidoxime mesylate (P2S) after repeated oral and intramuscular administration. Br. 1. Ind. Med. 33, 43-46. Hollingshaus, J. G., Nishioka, T., March, R. B., and Fukuto, T. R. (1981). Effect of impurities on the delayed neurotoxicity of 0 -(4-bromo-2,5dichlorophenyl) O-ethyl phenylphosphonothioate administered orally to hens. 1. Agric. Food Chem. 29, 593-600. Hollis, G. J. (1999). Organophosphate poisoning versus brainstem stroke. Med. 1. Aust. 170,596-597. Holmstedt, B. (1959). Pharmacology of organophosphorus cholinesterase inhibitors. Phannacal. Rev. 11, 567-688. Holmstedt, B. (1963). Structure-activity relationships of the organophosphorus anticholinesterase agents. In "Handbuch der Experimentellen Pharmakologie. Cholinesterases and Anticholinesterase Agents" (0. Eichler, A. Farah, and G. B. Koelle, eds.), pp. 428-485. Springer-Verlag, Berlin. Hoskins, B., Fernando, J. C. R., Dulaney, M. D., Lim, D. K, Liu, D. D., Watanabe, H. K, and Ho, I. K. (1986). Relationship between the neurotoxicities of soman, sarin and tabun, and acety !cholinesterase inhibition. Toxicol. Left. 30, 121-129. Hsiao, c.-T., Yang, C.-C., Deng, J. E, Bullard, M. J., and Liaw, S.-J. (1996). Acute pancreatitis following organophosphate intoxication. Clin. Toxicol. 34,343-347. Hui, K S. (1983). Metabolic disturbances in organophosphate insecticide poisoning. Arch. Pathol. Lab. Med. 107, 154. Indudharan, R., Win, M. N., and Noor, A. R. (1998). Laryngeal paralysis in organophosphorus poisoning. 1. Laryngol. Otol. 112, 81-82. Inoue, N., Fujishiro, K., Mori, K, and Matsuoka, M. (1988). Triorthocresyl phosphate poisoning. A review of human cases. 1. UOEH 10, 433-442. Ismail, K, Everitt, B., Blatchley, N., Hull, L., Unwin, c., David, A., and Wessley, S. (1999). Is there a Gulf War Syndrome? Lancet 353, 179-182. Jager, KW., Roberts, D. V., and Wilson, A. (1970). Neuromuscular function in pesticide workers. Br. 1. Industr. Med. 27,273-278. Jamal, G. A. (1995). Long term neurotoxic effects of organophosphate compounds. Adverse Drug React. Toxicol. Rev. 14, 85-99. Jamal, G. A. (1997). Neurological syndromes of organophosphorus compounds. Adverse Drug React. Toxicol. Rev. 16, 133-170. Jamal, G. A., Hansen, S., Apartopoulos, E, and Peden, A. (1996). The "Gulf War Syndrome." Is there evidence of dysfunction in the nervous system? 1. Neurol. Neurosurg. Psychiatry 60, 449-451. Jastrzebski, J., Zlotorowicz, M., and Szczepanski, M. (1994). Activation of blood coagulation induced by organophosphate pesticide. Mater. Med. Po!. 26,33-34. Jauhiainen, A., Kangas, J., Laitinen, S., and Savolainen, K (1992). Biological monitoring of workers exposed to mevinphos in greenhouses. Bull. Environ. Contam. Toxicol. 49,37-43. Jedrzejowska, H., Rowinska-Marcinska, K, and Hoppe, B. (1980). Neuropathy due to phytosol (agritox). Report of a case. Acta Neuropathol. (Berlin) 49, 163-168. Johnson, M. K (1975). Organophosphorus esters causing delayed neurotoxic effects. Mechanism of action and structure/activity studies. Arch. Toxicol. 34, 259-288. Johnson, M. K (1981). Delayed neurotoxicity-Do trichlorphon and/or dichlorvos cause delayed neuropathy in man or in test animals? Acta Pharmacol. Toxicol. 49(Suppl V), 87-98. Johnson, M. K (1984). Check your paraoxon and parathion for neurotoxic impurities. In "Delayed Neurotoxicity Workshop. Proceedings of the Delayed Neurotoxicity Workshop" (1. M. Crammer and J. E. Hixson, eds.). IUTOX Press, Little Rock, AK

1079

Johnson, M. K. (1990). Organophosphates and delayed neuropathy-Is NTE alive and well? Toxicol. Appl. Pharmacol. 102,385-399. Johnson, M. K, and Vale, J. A. (1992). Clinical management of acute organophosphate poisoning: An overview. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 528-535. Butterworth-Heinemann, Oxford. Johnson, M. K., Vale, J. A., Marrs, T. c., and Meredith, T. J. (1992). PraIidoxime for organophosphorus poisoning. Lancet 340, 64. Joubert, J., and Joubert, P. H. (1988). Chorea and psychiatric changes in organophosphate poisoning. A report of 2 further cases. S. Afr. Med. 1. 74, 32-34. Joubert, J., Joubert, P. H., van der Spuy, M., and van Graan, E. (1984). Acute organophosphate poisoning presenting with choero-athetosis. Clin. Toxicol. 22, 187-19l. Jovanovic, D. (1989). Pharmokinetics of pralidoxime chloride. A comparative study in healthy volunteers and in organophosphorus poisoning. Arch. Toxicol. 63,416-418. Jusic, A., Jurenic, D., and Milic, S. (1980). Electromyographical neuromuscular synapse testing and neurological findings in workers exposed to organophosphorus pesticides. Arch. Environ. Health 35,168-175. Kalow, W, and Staron, N. (1957). On the distribution and inheritance of atypical forms of human serum cholinesterase as indicated by dibucaine numbers. Can. 1. Biochem. Physiol. 35, 1305-1320. Kanto, J., and Klotz, U. (1988). Pharmacokinetic implications for the clinical use of atropine, scopolamine and glycopyrrolate. Acta Anaesthesio!' Scand. 32,69-78. Kaplan, 1. G., Kcssler, l., Rosenberg, N., Pack, D., and Schaumburg H. H. (1993). Sensory neuropathy associated with Dursban (chlorpyrifos) exposure. Neurology 43, 2193-2196. Karademir, M., Ertiik, F., and Ko~ak, R. (1990). Two cases of organophosphate poisoning with development of intermediate syndrome. Human Exp. Toxicol. 9, 187-189. Karalliede, L. (1999). Organophosphorus poisoning and anaesthesia. Anaesthesia 54,1073-1088. Karalliedde, L., and Henry, J. A. (1993). Effects of organophosphates on skeletal muscle. Human Exp. Toxicol. 12, 289-296. Karalliedde, L., and Senanayake, N. (1999). Organophosphorus insecticide poisoning. JIFCC 11, 4-9. Karalliedde, L., Senanayake, N., and Ariaratnam, A. (1988). Acute organophosphorus insecticide poisoning during pregnancy. Human Toxicol. 7,363-364. Karr, C., Demers, P., Costa, L. G., Daniell, WE., Barnhart, S., Miller, M., Gallagher, G., Horstrnan, S. W, Eaton, D., and Rosenstock, L. (1992). Organophosphate pesticide exposure in a group of Washington State orchard applicators. Environ. Res. 59,229-237. Kassa, J., and Cabal, J. (1999). A comparison of the efficacy of acetylcholinesterase reactivators against cyclohexyl methylphosphonofluoridate (GF agent) by in vitro and in vivo methods. Pharmacol. Toxicol. 84,41-45. Kaulla, K, and Holmes, J. H. (1961). Changes following anticholinesterase exposures. Arch. Environ. Health. 2, 168-177. Kecik, Y., Yorukoglu, D., Saygin, B., and Sekerci, S. (1993). A case of acute poisoning due to organophosphate insecticide. Anaesthesia 48, 141-143. Keeler, J. R., Hurst, C. G., and Dunn, M. A. (1991). Pyridostigmine used as a nerve agent pretreatment under wartime conditions. lAMA 266, 693-695. KentaIa, E., Kaila, T., lisalo, E., and Kanto, J. (1990). Intramuscular atropine in healthy volunteers: A pharmacokinetic and pharmacodynamic study. Int. J. Clin. Pharmacal. Ther. Toxicol. 28, 399-404. Khurana, D., and Prabhakar, S. (2000). Organophosphorus intoxication. Arch. Neurol. 57, 600-602. King, J., and Griffin, D. (1973). Differentiation of serum cholinesterase variants by succinylcholine inhibition. Br. 1. Anaesth. 49,450-454. Kiss, Z., and Fazekas, T. (1979). Arrhythmias in organophosphate poisonings. Acta Cardiol. 34, 323-330. Kiss, Z., and Fazekas, T. (1982). Organophosphate poisoning and complete heart block. J. R. Soc. Med. 73, 138-139. Kopman, A. F., Strachovsky, G., and Lichtenstein, L. (1978). Prolonged response to succinylcholine following physostigmine. Anesthesiology 49, 142-143.

1080

CHAPTER 51


Korsak, R J., and Sato, M. M. (1977). Effects of chronic organophosphate pesticide exposure on the central nervous system. Clin. Toxicol. 11, 83-95. Kraus, J. F., Richards, D. M., Borhani, N. O. Mull, R, Kilgore, W. w., and Winterlin, W. (1977). Physiological response to organophosphate residues in field workers. Arch. Environ. Contam. Toxicol. 5,471-485. Kraus, J. F., Mull, R, Kurts, P., Winterlin, w., Franti, C. E., Borhani, N., and Kilgore, W. (1981). Epidemiologic study of physiological effects in usual and volunteer citrus workers from organophosphate pesticide residues at reentry. J. Toxicol. Environ. Health 8, 169-184. Krieger, R. I., and Thongsinthusak, T. (1993). Metabolism and excretion of dimethoate following ingestion of overtolerance peas and a bolus dose. Food Chem. Toxicol. 31, 177-182. Kusic, R., Boskovic, B., Vojvodic, v., and Jovanovic, D. (1985). HI-6 in man: Blood levels, urinary excretion, and tolerance after intramuscular administration of the oxime to healthy volunteers. Fund. Appl. Toxicol. 5, S89-S97. Lamminpaa, A, and Riihimaki, V. (1992). Pesticide-related incidents treated in Finnish hospitals-A review of cases registered over a 5-year period. Human Exp. Toxicol. 11,473-479. Lankisch, P. G., Miiller C.-H., Niederstadt, H., and Brand, A. (1990). Painless acute pancreatitis subsequent to anticholinesterase insecticide (parathion) intoxication. Am. J. Gastroenterol. 85, 872-875. Le Blanc, F. N., Benson, B. E., and Gilg, A. D. (1986). A severe organophosphate poisoning requiring the use of an atropine drip. J. Toxicol. Clin. Toxicol. 24,69-76. Lee, w.-c., Yang, c.-c., Deng, J.-F., Wu, M.-L., Ger, J., Lin, H.-C., Chang, F.-Y., and Lee, S.-D. (1998). The clinical significance ofhyperamylasemia in organophosphate poisoning. Clin. Toxicol. 36, 673-681. Lefkowitz, R J., Hoffman, B. B., and Taylor, P. (1996). Neurotransmission. The autonomic and somatic motor nervous systems. In "Goodman and Gilman's the Pharmacological Basis of Therapeutics" (J. G. Hardman and L. E. Limbird, eds.), 9th ed., pp. 105-139. McGraw-Hill, New York. Levin, H. S., Rodnitzky, R L., and Mick, D. L. (1976). Anxiety associated with exposure to organophosphate compounds. Arch. Gen. Psychiatry 33, 225228. Lewis, P. J., Lowing, R. K., and Gompertz, D. (1981). Automated discrete kinetic method for erythrocyte acetylcholinesterase and plasma cholinesterase. Clin. Chem. 27,929-929. Lipp, J. A. (1973). Effect of benzodiazepine derivatives on soman-induced seizure activity and convulsions in the monkey. Arch. Int. Pharmacodyn. 202, 244-251. Loewenstein-Lichtenstein, Y., Glick, D., Gluzman, N., Sternfeld, M., Zakut, H., and Soreq, H. (1996). Overlapping drug interaction sites of human butyrylcholinesterase dissected by site-directed mutagenesis. Mol. Pharmacol. 50, 1423-1431. London, L., Thompson, M. L., Sacks, S., Fuller, B., Bachmann, O. M., and Myers, J. E. (1995). Repeatability and validity of a field kit for estimation of cholinesterase in whole blood. Occup. Environ Med. 52, 57-64. London, L., Meyers, J. E., Nell, v., Taylor, T., and Thompson, M. L. (1997). An investigation into neurologic and neurobehavioral effects of long-term agrichemical use among deciduous fruit farm workers in the Western Cape, South Africa. Environ. Res. 73, 132-145. London, L., Nell, v., Thompson, M. L., and Meyers, J. E. (1998). Effects of long-term organophosphate exposures on neurological symptoms, vibration sense and tremor among South African farm workers. Scan. J. Work Environ. Health 24, 18-29. Lotti, M. (1987). Organophosphate-induced delayed polyneuropathy in humans: Perspectives for biomonitoring. Trends Pharmacol. Sci. 8, 175-176. Lotti, M. (1991). Treatment of acute organophosphate poisoning. Med. J. Aust. 154,51-55. Lotti, M. (1992a). Central neurotoxicity and behavioural effects of anticholinesterases. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 75-83. Butterworth-Heinemann, Oxford. Lotti, M. (1992b). The pathogenesis of organophosphate delayed polyneuropathy. Crit. Rev. Toxicol. 21, 465-487. Lotti, M. (1995). Cholinesterase inhibition: Complexities in interpretation. Clin. Chem. 41, 1814--1818.

Lotti, M. (1999). Causes of the Gulf War Syndrome: testing hypotheses. Muscle Nerve 22,663-665. Lotti, M. (2000). Organophosphorus compounds. In "Experimental and Clinical Neurotoxicology" (P. S. Spencer, H. S. Schaumburg, and A. C. Ludolph, eds.), 2nd ed., pp. 898-925. Oxford University Press, New York. Lotti, M., and Becker, C. E. (1982a). Treatment of acute organophosphate poisoning: Evidence of a direct effect on central nervous system by 2-PAM (pyridine-2-aldoxime methyl chloride). J. Toxicol.-Clin. Toxicol. 19, 121127. Lotti, M., and Becker, C. E. (1982b). Polyneuropathy and exposure to parathion. Neurology 32, 217. Lotti, M., and Johnson, M. K. (1978). Neurotoxicity of organophosphorus pesticides: Predictions can be based on in vitro studies with hen and human enzymes. Arch. Toxicol. 41,215-221. Lotti, M., and Moretto, A (1995). Cholinergic Symptom and Gulf War Syndrome. Nature Med. 1, 1225-1226. Lotti, M., Ferrara, S. D., Caroldi, S., and Singaglia, F. (1981). Enzyme studies with human and hen autopsy tissue suggest omethoate does not cause delayed neuropathy in man. Arch. Toxico!. 48, 265-270. Lotti, M., Becker, C. E., Aminoff, M. J., Woodrow, J. E., Seiber, J. N., Talcott, R. E., and Richardson, R J. (1983). Occupational exposure to the cotton defoliants DEF and merphos. A rational approach to monitoring organophosphorus-induced delayed neurotoxicity. J. Occup. Med. 25, 517522. Lotti, M., Becker, C. E., and Aminoff, M. J. (1984). Organophosphate polyneuropathy: Pathogenesis and prevention. Neurology 34, 658-662. Lotti, M., Moretto, A, Zoppellari, R, Dainese, R, Rizzuto, N., and Barusco, G. (1986). Inhibition oflymphocytic neuropathy target esterase predicts the development of organophosphate-induced delayed polyneuropathy. Arch. Toxico!. 59, 176-179. Ludomirsky, A., Klein, H. 0., Sarelli, P., Becker, B., Hoffman, S., Taitelman, U., Barzilai, J., Lang, R., David, D., Disegni, E., and Kaplinsky, E. (1982). Q-T prolongation and polymorphous ("Torsade de Pointes") ventricular arrhythmias associated with organophosphorus insecticide poisoning. Am. J. Cardia!. 49, 1654--1658. Luo, c., Saxena, A., Smith, M., Garcia, G., Radi<e, Z., Taylor, P., and Doctor, B. P. (1999). Phosphoryl oxime inhibition of acetylcholinesterase during oxime reactivation is prevented by edrophonium. Biochemistry 38, 99379947. Luzhnikov, E. A., Yaroslavsky, A. A., Molodenkov, M. N., Shurkalin, B. K., Evseev, N. G., and Barsukov, U. F. (1977). Plasma perfusion through charcoal in methylparathion poisoning. Lancet i, 38-39. Lyon, J., Taylor, H., and Ackerman, B. (1987). A case report of intravenous malathion injection with determination of serum half-life. Clin. Toxieo!. 25, 243-249. Mackness, B., Durrington, P. N., and Mackness, M. I. (1998). Human serum paraoxonase. Gen. Pharmaeol. 31, 329-336. Mahieu, P., Hassoun, A., Van Binst, R, Lauwerys, R., and Deheneffe, Y. (1982). Severe and prolonged poisoning by fenthion. Significance of the determination of the anticholinesterase capacity of plasma. J. Toxieo!. Clin. Toxieo!. 19, 425-432. Maizlish, N., Schenker, M., Weisskopf, c., Seiber, J., and Samuels, S. (1987). A behavioral evaluation of pest control workers with short-term, low-level exposure to the organophosphate diazinon. Am. J. Ind. Med. 12, 153-172. Maltby, N., Broe, G. A, Creasey, H., Jorm, A. F., Christensen, H., and Brooks, W. S. (1994). Efficacy of tacrine and lecithin in mild to moderate Alzheimer's disease: Double blind trial. Br. Med. J. 308, 879-883. Maresch, W. (1957). Die Vergiftung durch Phosphorsaureester (E 605, Parathion, Thiophos). Arehiv fur Toxikologie 16,285-319 (in German). Markowitz, J. S., Gutterman, E. M., and Link, B. G. (1986). Self-reported physical and psychological effects following a malathion pesticide incident. J. Oeeup. Med. 28, 377-383. Maroni, M., and Bleecker, M. L. (1986). Neuropathy target esterase in human lymphocytes and platelets. J. App!. Toxieol. 6, 1-7. Marrs, T. C. (1991). Toxicology of oximes used in treatment of organophosphate poisoning. Adverse Drug React. Toxiea!. Rev. 10,61-72. Marrs, T. C. (1993). Organophosphate poisoning. Pharmae. Ther. 58, 51-66.

References

Marsh, W. H., Vukov, G. A., and Conradi, E. C. (1988). Acute pancreatitis after cutaneous exposure to an organophosphate insecticide. Am. J. Gastroentero!' 83, 1158-1160. Martinez-Chuecos, J., Jurado, M. C., Gimenez, M. P., Martinez, D., and Menendez, M. (1992). Experience with hemoperfusion for organophosphate poisoning. Crit. Care Med. 20, 1538-1543. Maselli, R. A., and Leung, C. (1993). Analysis of anticholinesterase-induced neuromuscular transmission failure. Muscle Nerve 16, 548-553. Maselli, R. A., and Soliven, B. C. (1991). Analysis of the organophosphateinduced electromyographic response to repetitive nerve stimulation: Paradoxical response to edrophonium and d-tubocurarine. Muscle Nerve 14, 1182-1188. Maselli, R. A., Jacobsen, J. H., and Spire, J.-P. (1986). Edrophonium: An aid in the diagnosis of acute organophosphate poisoning. Ann. Neuro!' 19,508510. Massumi, R. A., Mason, D. T., Amsterdam, E. A., DeMaria A., Miller, R. R., Scheinman, M. M., and Zelis, R. (1972). Ventricular fibrillation and tachycardia after intravenous atropine for treatment of bradycardias. N. Eng!. J. Med. 287, 336-338. Matsumiya, N., Tanaka, M., Iwai, M., Kondo, T., Takahashi, S., and Sato, S. (1996). Elevated amylase is related to the development of respiratory failure in organophosphate poisoning. Human Exp. Toxieo!. 15, 250-253. McConnell, R., Keifer, M., and Rosenstock, L. (1994). Elevated quantitative vibrotactile threshold among workers previously poisoned with methamidophos and other organophosphate pesticides. Am. J. Ind. Med. 25, 325334. McConnell, R., Delgado-T€llez, E., Cuadra, R., T6rres, E., Keifer, M., Almendarez, J., Miranda, J., EI-Fawal, H. A. N., Wolff, M., Simpson, D., and Lundberg, I. (1999). Organophosphate neuropathy due to methamidophos: Biochemical and neurophysiological markers. Arch. Toxieol. 73, 296-300. McLeod, C. G., Jr. (1985). Pathology of nerve agents: Perspectives on medical management. Fund. App!. Toxieol. 5, SIO-SI6. Mearns, J., Dunn, J., and Lees-Haley, P. R. (1994). Psychological effects of organophosphate pesticides: A review and call for research by psychologists. J. Clin. Psycho!. 50, 286-294. Medicis, J. J., Stork, C. M., Howland, M. A., Hoffman, R. S., and Goldfrank, L. R. (1996). Pharmacokinetics following a loading plus a continuous infusion of pralidoxime compared with the traditional short infusion regimen in human volunteers. Clin. Toxieo!. 34, 289-295. Merrill, D. G., and Mihm, F. G. (1982). Prolonged toxicity of organophosphate poisoning. Crit. Care Med. 10,550-551. Metcalf, D. R., and Holmes, J. H. (1969). EEG, psychological, and neurological alterations in humans with organophosphorus exposure. Ann. N. Y. Aead. Sci. 160, 357-365. Michaleck, H., and Stavinoha, W. B. (1978). Effect of chlorpromazine pre-treatment on the inhibition of total cholinesterases and butyrylcholinesterase in brain of rats poisoned by physostigme or dichlorvos. Toxicology 9, 205-218. Michotte, A., Van Dijck, I., Maes, v., and D'Haenen, H. (1989). Ataxia as the only delayed neurotoxic manifestation of organophosphate insecticde poisoning. Eur. Neuro!. 29,23-26. Millard, C. B., and Broomfield, C. A. (1995). Anticholinesterases: Medical applications of neurochemical principles. J. Neuroehem. 64, 1909-1918. Minton, N. A., and Murray, V. S. G. (1988). A review of organophosphate poisoning. Med. Toxieo!. 3,350-375. Misra, U. K., Nag, D., Khan, W. A., and Ray, P. K. (1988). A study of nerve conduction velocity, late responses and neuromuscular synapse functions in organophosphate workers in India. Arch. Toxieo!. 61,496-500. Mizutani, T., Naito, H., and Oohashi, N. (1991). Rectal ulcer with massive haemorrhage due to activated charcoal treatment in oral organophosphate poisoning. Human Exp. Toxieo!. 10, 385-386. Moeller, H. C., and Rider, J. A. (1965). Furtber studies on the anticholinesterase effect of systox and methyl parathion in humans. Fed. Proe., Fed. Am. Soc. Exp. Bio!. 24, 641. Molphy, R., and Rathus, E. M. (1964). Organic phosphorus poisonings and therapy. Med. l. Aust. 2, 337-340.

1081

Moore, W. K. S. (1956). Two cases of poisoning with di-isopropylfiuorophosphonate (D.F.P.). Br. l. Industr. Med. 13,214-216. Moore, P. G., and James, o. F. (1981). Acute pancreatitis induced by acute organophosphate poisoning. Postgrad. Med. l. 57,660-662. Moretto, A., and Lotti, M. (1988). Organ distribution of neuropathy target esterase in man. Bioehem. Pharmaeo!' 37,3041-3043. Moretto, A., and Lotti, M. (1998). Poisoning by organophosphorus insecticides and sensory neuropathy. l. Neuro!. Neurosurg. Psychiatry 64, 463-468. Moretto, A., Capodicasa, E., Bertolazzi, M., De Paris, P., Saia, B. 0., and Lotti, M. (1995). Biological monitoring of occupational exposures to organophosphorus insecticides. In "Agriculture Health and Safety: Workplace, Environment, Sustainability" (H. H. McDuffie, J. A. Dosman, K. M. Semchuk, S. A. Olenchock, and A. Senthilselvan, eds.), pp. 217-221. CRC Press, Boca Raton, FL. Morgan, J. P. (1982). The Jamaica ginger paralysis. lAMA 248, 1864-1867. Morgan, J. P., and Penovich, P. (1978). Jamaica ginger paralysis. Forty-sevenyear follow-up. Arch. Neuro!' 35, 530-532. Morgan, D. P., Hetzler, H. L., Slach, E. F., and Lin, L.I. (1977). Urinary excretion of paranitrophenol and alkyl phosphates following ingestion of methyl or ethyl parathion by human subjects. Arch. Environ. Contam. Toxieol. 6, 159-173. Mueller, R. F., Homung, S., Furlong, C. E., Anderson, J., Giblett, E. R., and Motulsky, A. G. (1983). Plasma paraoxonase polmorphism: A new enzyme assay, population, family, biochemical, and linkage studies. Am. l. Hum. Genet. 35,393-408. Murray, V. S. G., Wiseman, H. M., Dawling, S., Morgan, I., and House, I. M. (1992). Health effects of organophosphate sheep dips. Br. Med. l. 305, 1090. Nagler, J., Braeckman, R. A., Willems, J. L., Verpooten, G. A., and De Broe, M. E. (1981). Combined hemoperfusion-hemodialysis in organophosphate poisoning. J. App!. Toxico!. 1, 199-201. Namba, T., Nolte, C. T., Jackrel, J., and Grob, D. (1971). Poisoning due to organophosphate insecticides. Acute and chronic manifestations. Am. l. Med. 50,475-492. Neuvonen, P. J., and Olkkola, K. T. (1988). Oral activated charcoal in the treatment of intoxications. Role of single and repeated doses. Med. Toxieo!. 3, 33-58. Newcombe, D. S. (1992). Immune surveillance, organophosphorus exposure, and Iymphomagenesis. Lancet 339,539-541. Newcombe, D. S., and Esa, A. H. (1992). Immunotoxicity of organophosphorus compounds. In "Clinical Immunotoxicology" (D. S. Newcombe, N. R. Rose, and J. C. Bloom, eds.), pp. 349-364. Raven Press, New York. Niedziella, S. w., Gopel, w., and Banzhaf, E. (1985). Akute Alkylphosphatintoxikation (Trichlorphon) mit intervaIHirem Polyneuropathie-Syndrom. Z. Ges. Inn. Med. lahrg. 40, 237-239 (in German). NIH Technology Assessment Workshop Panel (1994). The Persian Gulf experience and health. JAMA 272, 391-395. Nilsson, E. (1982). Physostigmine treatment in various drug-induced intoxications.Ann. Clin. Res. 14, 165-172. Nisse, P., Forceville, X., Cezard, C., Ameri, A., and Mathieu-Nolf, M. (1998). Intermediate syndrome with delayed distal polyneuropathy from ethyl parathion poisoning. Vet. Human Toxieo!. 40, 349-352. Nolan, R. J., Rich, D. L., Freshour, N. L., and Saunders, J. H. (1984). Chlorpyrifos: Pharmacokinetics in human volunteers. Toxieol. App!. Pharmaeo!' 73,8-15. Nordgren, I., Holmstedt, B., Bengtsson, E., and Finkel, Y. (1980). Plasma levels of metrifonate and dichlorvos during treatment of schistosomiasis with bilarcil® . Am. 1. Trop. Med. Hyg. 29,426-430. Norton, S. (1986). Toxic responses of the central nervous system. In "Casarett and Doull's Toxicology" (c. D. Klaassen, M. O. Amdur, and J. Doull, eds.), 3rd ed., pp. 359-386. Macmillan, New York. Nouira, S., Abroug, F., Elatrous, S., Boujdaria, R., and Bouchoucha, S. (1994). Prognostic value of serum cholinesterase in organophosphate poisoning. Chest 106,1811-1814. Nutley, B., and Cocker, J. (1993). Biological monitoring of workers occupationally exposed to organophosphorus pesticides. Pest. Sei. 38, 315-322.

1082

CHAPTER 51


Ochi, G., Watanabe, K, Tokuoka, H., Hatakenaka, S., and Arai, T. (1995). Neuroleptic malignant-like syndrome: A complication of acute organophosphate poisoning. Can. J. Anaesth. 42, 1027-1030. Ohbu, S., Yamashina, A., Takasu, N., Yamaguchi, T., Murai, T., Nakano, K, Matsui, Y., Mikami, R., Sakurai, K, and Hinohara, S. (1997). Sarin poisoning on Tokyo subway. S. Afr. Med. 1. 90, 587-593. Ohkawa, H., Oshita, H., and Miyamoto, J. (1980). Comparison of inhibitory activity of various organophosphorus compounds against acetylcholinesterase and neurotoxic esterase of hens with respect to delayed neurotoxicity. Biochem. Phannacol. 29,2721-2727. Oken, B. S., and Chiappa, K. H. (1986). Statistical issues conceruing computerized analysis of brainwave topography. Ann. Neurol. 19,493-497. Okumura, T., Takasu, N., Ishimatsu, S., Miyanoki, S., Mitsuhashi, A, Kumada, K, Tanaka, K, and Hinohara, S. (1996). Report on 640 victims of the Tokyo subway sarin attack. Ann. Emerg. Med. 28, 129-135. 0stergaard, D., Jensen, E S., Jenson, E., Skovgaard, L. T., and Viby-Mogensen, J. (1992). Influence of plasma cholinesterase activity on recovery from mivacurium-induced neuromuscular blockade in phenotypically normal patients. Acta Anaesthesio!. Scand. 36, 702-706. Osterloh, J., Lotti, M., and Pond, S. M. (1983). Toxicologic studies in a fatal overdose of2,4-D, MCPP, and chlorpyrifos. J. Analyt. Toxico!. 7, 125-129. Otto, D. A., Soliman, S., Svendsgaard, D., Soffar, A, and Ahmed, N. (1990). Neurobehavioral assessment of workers exposed to organophosphorus pesticides. In "Advances in Neurobehavioural Toxicology: Applications in Environmental and Occupational Health" (B. L. Johnson, ed.), pp. 305-322 Lewis Publishers, Chelsea, M1. Panieri, E., Krige, l. E., Bomman, P. c., and Linton, D. M. (1997). Severe necrotizing pancreatitis caused by organophosphate poisoning. J. Clin. Gastroenterol. 25, 463-465. Parron, T., Hemandez, A. E, and Villanueva, E. (1996a). Increased risk of suicide with exposure to pesticides in an intensive agricultural area. A 12-year retrospective study. Foren. Sci. Int. 79, 53-63. Parron, T., Hemandez, A. E, Pia, A., and Villanueva, E. (1996b). Clinical and biochemical changes in greenhouse sprayers chronically exposed to pesticides. Human Exp. Toxicol. 15,957-963. Perron, R., and lohnson, B. B. (1969). Insecticide poisoning. N. Eng!. J. Med. 281,274-275. Pickett, w., King, W. D., Lees, R. E. M., Bienefeld, M., Morrison, H. 1., and Brison, R. l. (1998). Suicide mortality and pesticide use among Canadian farmers. Am. J. Ind. Med. 34,364-372. Pilkington, A., lamal, G. A., Gilham, R., Hansen, S., Buchanan, D., Kidd, M., Azis, M. A, lulu, P. 0., Al-Rawas, S., Ballantyne, l. P., Hurley, l. E, and Soutar, C. A. (1999). "Epidemiological Study of the Relationshps between Exposure to Organophosphate Pesticides and Indices of Chronic Peripheral Neuropathy, and Neurophysiological Abnormalities in Sheep Farmers and Dippers. Phase 3. Clinical Neurological, Neurophysiological and Neuropsychological Study." Technical Memorandum Series TMl99/02c, Institute of Occupational Medicine, Edinburgh. Pimentel, l. M., and Carrington da Costa, R. B. (1992). Effects of organophosphates on the heart. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 145-148. Butterworth-Heinemann, Oxford. Plestina, R. (1984). "Prevention, Diagnosis and Treatment of Insecticide Poisoning." Report WHONBC/84.889, World Health Organization, Geneva. Plestina, R., and Piukovic-Plestina, M. (1978). Effect of anticholinesterase pesticides on the eye and on vision. Crit. Rev. Toxicol. 6, 1-23. Polhuijs, M., Langenberg, J. P., and Benschop, H. P. (1997). New method for retrospective detection of exposure to organophosphorus anticholinesterases: Application to alleged sarin victims of lapanese terrorists. Toxico!. App!. Phannaco!' 146, 156-161. Popendorf, W. J., Spear, R. C., Leffingwell, l. T., Yager, l., and Kahn, E. (1979). Harvester exposure to Zolone® (phosalone) residues in peach orchards. 1. Occup.Me~ 21, 189-194. Pullicino, P., and Aquilina, J. (1989). Opsoclonus in organophosphate poisoning. Arch. Neurol. 46, 704-705.

Quinby, G. E., Loomis, T. A., and Brown, H. W. (1963). Oral occupational parathion poisoning treated with 2-PAM iodide (2-pyridine aldoxime methiodide). N. Eng!. J. Med. 268, 639-643. Rasmussen, W. A, lensen, l. A, Stein, W. l., and Hayes, W. l., lr. (1963). Toxicological studies of DDVP for disinsection of aircraft. Aerosp. Med. 34, 594-600. Ray, D. E. (1998a). "Organophosphorus Esters: An Evaluation of Chronic Neurotoxic Effects." MRC Institute for Environment and Health, Leicester, UK Ray, D. E. (1998b). Chronic effects of low level exposure to anticholinesterases-A mechanistic view. Toxicol. Lett. 102-103,527-533. Reidy, T. l., Bowler, R. M., Rauch, S. S., and Pedroza, G. 1. (1992). Pesticide exposure and neuropsychological impairment in migrant farm workers. Arch. Clin. Neuropsycho!. 7, 85-95. Reiner, E., and Plestina, R. (1979). Regeneration of cholinesterase activities in humans and rats after inhibition by O,O-dimethyl-2,2-dichlorovinyl phosphate. Toxicol. Appl. Pharmacol. 49,451-454. Relman, A. S. (1991). Tacrine as a treatment for Alzheimer's dementia. N. Engl. J. Med. 324, 349. Rengstorff, R. H. (1985). Accidental exposure to sarin: vision effects. Arch. Toxico!. 56,201-203. Rengstorff, R. H. (1994). Vision and ocular changes following accidental exposure to organophosphates. J. App!. Toxicol. 14, 115-118. Richardson, R. l. (1995). Assessment of the neurotoxic potential of chlorpyrifos relative to other organophosphorus compounds: A critical review of the literature. 1. Toxico!. Environ. Health 44, 135-165. Richter, E. (1993). "Organophosphorus Pesticides: A Multinational Epidemiologic Study." World Health Organization, Copenhagen, Denmark. Richter, E. D., Chuwers, P., Levy, Y., Gordon M., Grauer, E, Marzouk, l., Levy, S., Barron, S., and Gruener, N. (1992). Health effects from exposure to organophosphate pesticides in workers and residents in Israel. Isr. J. Med. Sci. 28,584-597. Rider, l. A., Swader, l. 1., and Puletti, E. J. (1970). Methyl parathion and guthion anticholinesterase effects in human subjects. Fed. Prod. Fed. Am. Soc. Exp. Bio!. 29, 349. Rider, l. A., Swader, l. 1., and Puletti, E. l. (1972). Anticholinesterase toxicity studies with guthion, phosdrin, di-syston, and trithion in human subjects. Fed. Prod. Fed. Am. Soc. Exp. Bio!. 31, 520. Rider, l. A., Puletti, E. l., and Swader, l. 1. (1975). The minimal oral toxicity level for mevinphos in man. Toxico!. App!. Phannaco!. 32,92-100. Roberts, D. V. (1976). E. M. G. voltage and motor nerve conduction velocity in organophosphorus pesticide factory workers. Int. Arch. Occup. Environ. Health 36, 267-274. Roberts, D. V. (1977). A longitudinal electromyographic study of six men occupationally exposed to organophosphorus compounds. Int. Arch. Occup. Environ. Health 38, 221-229. Rodgers, K. E., Devens, B. H., and Imamura, T. (1992). Immunotoxic effects of anticholinesterases. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 211-222. Butterworth-Heinemann, Oxford. Rodnitzky, R. L., Levin, H. S., and Mick, D. L. (1975). Occupational exposure to organophosphate pesticides. A neurobehavioral study. Arch. Em'iron. Health 30,98-103. Rosenstock, L., Keifer, M., Daniell, W. E., McConnell, R., CIaypoole, K, and the Pesticide Health Effects Study Group (1991). Chronic central nervous system effects of acute organophosphate pesticide intoxication. Lancet 338, 223-227. Rosenthal, N. E., and Cameron, C. L. (1991). Exaggerated sensitivity to an organophosphate pesticide. Am. J. Psychiatry 148, 270. Saadeh, A M., AI-Ali, M. K, Farsakh, N. A., and Ghani, M. A. (1996). Clinical and sociodemographic features of acute carbamate and organophosphate poisoning: A study of 70 adult patients in North Jordan. Clin. Toxicol. 34, 45-51. Saadeh, A. M., Farsakh, N. A, and AI-Ali, M. K (1997). Cardiac manifestations of acute carbamate and organophosphate poisoning. Heart 77, 461-464.

References

Sahin, M., Bemay, I., Cantiirk, F., and Demircali, A. E. (1994). Reflex sympathetic dystrophy syndrome secondary to organophosphate intoxication induced neuropathy. Ann. Nuc!. Med. 8, 299-300. Sakamoto, T., Sawada, Y., Nishide, K., Sadamitsu, D., Yoshioka, T., Sugimoto, T., Nishii, S., and Kishi, H. (1984). Delayed neurotoxicity produced by an organophosphorus compound (Sumithion). A case report. Arch. Toxicol. 56, 136-138. Savage, E. P., Keefe, T. J., Mounce, L. M., Heaton, R. K., Lewis, J. A., and Burcar, P. J. (1988). Chronic neurological sequelae of acute organophosphate pesticide poisoning. Arch. Environ. Health 43, 38-45. Saxena, A., Maxwell, D. M. Quinn, D. M., Radic, Z., Taylor, P., and Doctor, B. P. (1997). Mutant acetylcholinesterase as potential detoxification agents for organophosphate poisoning. Biochem. Pharmacol. 54,269-274. Schexnayder, S., James, L. P., Keams, G. L., and Farrar, H. C. (1998). The pharmacokinetics of continuous infusion pralidoxime in children with organophosphate poisoning. Clin. Toxicol. 36,549-555. Schuman S. H., and Wagner, S. L. (1991). Pesticide intoxication and chronic CNS effects. Lancet 338, 948. Scott, R. J. (1986). Repeated asystole following PAM in organophosphate selfpoisoning. Anaesth. Int. Care 14, 458-460. Sedgwick, E. M., and Senanayake, N. (1997). Pathophysiology of the intermediate syndrome of organophosphorus poisoning. J. Neurol. Neurosurg. Psychiatry 62, 201-202. Selden, B. S., and Curry, S. C. (1987). Prolonged succinylcholine-induced paralysis in organophosphate insecticide poisoning. Ann. Emerg. Med. 16, 215-217. Sellstrom, A. (1992). Anticonvulsants in anticholinesterase poisoning. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 578-586. ButterworthHeinemann, Oxford. Senanayake, N. (1981). Tri-cresyl phosphate neuropathy in Sri Lanka: A clinical and neurophysiological study with a three year follow up. J. Neurol. Neurosurg. Psychiatry 44,775-780. Senanayake, N., and Jeyaratnam, J. (1981). Toxic polyneuropathy due to gingili oil contaminated with tri-cresyl phosphate affecting adolescent girls in Sri Lanka. Lancet i, 88-89. Senanayake, N., and Johnson, M. K. (1982). Acute polyneuropathy after poisoning by a new organophosphate insecticide. N. Engl. J. Med. 306, 155157. Senanayake, N., and Karalliedde, L. (1987). Neurotoxic effects of organophosphorus insecticides. An intermediate syndrome. N. Engl. J. Med. 316, 761-763. Senanayake, N., and Sanmuganathan, P. S. (1995). Extrapyramidal manifestations complicating organophosphorus insecticide poisoning. Human Exp. Toxicol. 14, 600-604. Sewell, c., Pilkington, A., Buchanan, D., Tannahill, S. N., Kidd, M., Cherrie, B., and Robertson, A., (1999). "Epidemiological Study of the Relationships between Exposure to Organophosphate Pesticides and Indices of Chronic Peripheral Neuropathy, and Neuropsychological Abnormalities in Sheep Farmers and Dippers. Phase 1. Development and Validation of an Organophosphate Uptake Model for Sheep Dippers." Technical Memorandum Series, Institute of Occupational Medicine, Edinburgh. Sharma, R. P., and Tomar, R. S. (1992). Immunotoxicity of anticholinesterase agents. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 203-210. Butterworth-Heinemann, Oxford. Shen, Z.-X. (1998). Pyridostigmine bromide and Gulf War syndrome. Med. Hypothesis 51, 235-237. Shiraishi, S., Inoue, N., Murai, Y., Onishi. A., and Noda, S. (1983). Dipterex (trichlorfon) poisoning. Clinical and pathological studies in human and monkeys. 1. UOEH 5(Suppl.), 125-132. Sidell, F. R., and Groff, W. A. (1971). Intramuscular and intravenous administration of small doses of 2-pyridinium aldoxime methochloride to man. J. Pharm. Sci. 60, 1224-1228. Sidell, F. R., Groff, W. A., and Kaminskis, A. (1972). Toxogonin and pralidoxime: Kinetic comparison after intravenous administration to man. J. Pharmacol. Sci. 61, 1765-1769.

1083

Singh, G., Avasthi, G., Khurana, D., Whig, J., and Mahajan, R. (1998a). Neurophysiological monitoring of pharmacological manipulation in acute organophosphate (OP) poisoning. The effects of pralidoxime, magnesium sulphate and pancuronium. Electroenceph. Clin. Neurophysiol. 107, 140148. Singh, G., Mahajan, R., and Whig, J. (1998b). The importance of electrodiagnostic studies in acute organophosphate poisoning. J. Neurol. Sci. 157, 191-200. Singh, G., Sidhu, U. P. S., Mahajan, R., Avasthi, G., and Whig, J. (2000). Phrenic nerve conduction studies in acute organophosphate poisoning. Muscle Nerve 23, 627-632. Singh, S., Batra, Y. K., Singh, S. M., Wig, N., and Sharma, B. K. (1995). Is atropine alone sufficient in acute severe organophosphorus poisoning?: Experience of a North West Indian Hospital. Int. J. Clin. Pharmacol. Ther. 33, 628-630. Spear, R. C., Popendorf, W. J., Leffingwell, J. T., Milby, T. H., Davies, J. E., and Spencer, W. F. (1977). Fieldworker's response to weathered residues of parathion. J. Occup. Med. 19,406-410. StaIberg, E., Hilton-Brown, P., Kolmodin-Hedman, B., Holmstedt, B., and Augustinsson, K. B. (1978). Effect of occupational exposure to organophosphorus insecticides on neuromuscular function. Scand. J. Work. Environ. Health 4, 255-261. Stamboulis, E., Psimaras, A., Vassilopoulos, D., Davaki, P., Manta, P., and Kapaki, E. (1991). Neuropathy following acute intoxication with mecarbam (OP ester). Acta Neurol. Scand. 83, 198-200. Stavinoha, W. B., Modak, A. T., and Weintraub, S. T. (1976). Rate of accumulation of acetylcholine in discrete regions of the rat brain after dichlorvos treatment. J. Neurochem. 27, 1375-1378. Steenland, K. (1996). Chronic neurological effects of organophosphate pesticide. Br. Med. J. 312, 1312-1313. Steenland, K., Jenkins, B., Ames, R. G., O'Malley, M., Chrislip, D., and Russo, J. (1994). Chronic neurological sequelae to organophosphate pesticide poisoning. Am. J. Public Health 84, 731-736. Stephens, R., Spurgeon, A., Beach, J., Calvert, I., Berry, H., Levy, L., and Harrington, J. M. (1995a). "An Investigation into the Possible Chronic Neuropsychological and Neurological Effects of Occupational Exposure to Organophosphates in Sheep Farmers." Contract Research Report 74/1995, Health & Safety Executive, Sheffield, UK. Stephens, R., Spurgeon, A., Calvert, I. A., Beach, J., Levy, L. S., Berry, H., and Harrington, J. M. (1995b). Neuropsychological effects of long term exposure to organophosphates in sheep dip. Lancet 345, 1135-1139. Stephens, R., Spurgeon, A., and Berry, H. (1996). Organophosphates: The relationship between chronic and acute exposure effects. Neurotoxicol. Teratol. 18, 449-453. Sterri, S. H., Rognerud, B., Fiskum, S. E., and Lyngaas, S. (1979). Effects of toxogonin and P2S on the toxicity of carbamates and organophosphorus compounds. Acta Pharmacol. Toxicol. 45, 9-15. Stokes, L., Stark, A., Marshall, E., and Narang, A. (1995). Neurotoxicity among pesticide applicators exposed to organophosphates. Occup. Environ. Med. 52, 648-653. Stoller, A., Krupinski, J., Christophers, A. J., and Blanks, G. K. (1965). Organophosphorus insecticides and major mental illness. An epidemiological investigation. Lancet i, 1387-1388. St. Omer, V. E. V., and Rottinghaus, G. E. (1992). Biochemical determination of cholinesterase activity in biological fluids and tissues. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 15-27. Butterworth-Heinemann, Oxford. SU, M.-Q., Kinoshita, F. K., Frawley, J. P., and DuBois, K. P. (1971). Comparative inhibition of aliesterases and cholinesterase in rats fed eighteen organophosphorus insecticides. Toxicol. Appl. Pharmacol. 20,241-249. Summers, W. K., Majovski, L. v., Marsh, G. M., Tachiki, K., and Kling, A. (1986). Oral tetrahydroaminoacridine in long-term treatment of senile dementia, Alzheimer type. N. Engl. J. Med. 315, 1241-1245. Sundwall, A. (1960). Plasma concentration curves of P2S after intramuscular, intravenous and oral administration in man. Biochem. Pharmacol. 8,413417.

1084

CHAPTER 51


Sundwall, A. (1961). Minimum concentrations of N-methylpyridinium-2aldoxime methane sulphonate (P2S) which reverse neuromuscular block. Biochem. Phannacal. 8,413-417. Susser, M., and Stein, Z. (1957). An outbreak of tri-ortho-cresyl phosphate (T. O. C. P.) poisoning in Durban. Br. J. Indust. Med. 14, 111-120. Sussman, J. L., Harel, M., Frolow, E, Oefner, c., Goldman, A., Toker, L., and Silman, I. (1991). Atomic structure of acetylcholinesterase from Torpedo californica: A prototypic acetylcholine-binding protein. Science 253, 872879. Suzuki, T., Morita, H., and Ono, K. (1995). Sarin poisoning in Tokyo subway. Lancet 345, 980. Tabershaw, I. R., and Cooper, W. C. (1966). Sequelae of acute organic phosphate poisoning. J. Occup. Med. 8, 5-20. Tafuri, J., and Roberts, .T. (1987). Organophosphate poisoning. Ann. Fmerg. Med. 16, 193-202. Tanaka, D., Jr., and Bursian, S. J. (1989). Degeneration patterns in the chicken central nervous system induced by ingestion of the organophosphorus delayed neurotoxin tri-ortho-tolyl phosphate. A silver impregnation study. Brain Res. 484, 240-256. Taylor, P. (1996a). Anticholinesterase agents. In "Goodman and Gilman's the Pharmacological Basis of Therapeutics" (J. G. Hardman and L. E. Limbird, eds.), 9th ed., pp. 161-176. McGraw-Hill, New York. Taylor, P. (1996b). Agents acting at the neuromuscular junction and autonomic ganglia. In "Goodman and Gilman's the Pharmacological Basis of Therapeutics" (1. G. Hardman and L. E. Limbird, eds.), 9th ed., pp. 177-197. McGraw-Hill, New York. Thompson, J. W, and Stocks, R. M. (1997). Brief bilateral vocal cord paralysis after insecticide poisoning. A new variant of toxcity syndrome. Arch. Otalaryngal. Head Neck Surg. 123, 93-96. Tomlin, C. D. S. (1997). "The Pesticide Manual," llth ed. British Crop Protection Council, Surrey, UK. Tosi, L., Righetti, C., Adami, L., and Zanette, G. (1994). October 1942: A strange epidemic paralysis in Saval, Verona, Italy. Revision and diagnosis 50 years later of tri-ortho-cresyl phosphate poisoning. J. Neural. Neurosurg. Psychiatry 57, 810-813. Tracey, J. A., and Gallagher, H. (1990). Use of glycopyrrolate and atropine in acute organophosphorus poisoning. Human Exp. Taxicol. 9,99-100. Tsao, T. C-Y., Juang, y-c., Lan, R.-S., Shieh, W-B., and Lee, C.-H. (1990). Respiratory failure of acute organophosphate carbamate poisoning. Chest 98,631-636. Tsatsakis, A. M., Aguridakis, P., Michalodimitrakis, M. N., Tsakalov, A. K., Alegakis, A K., Koumantakis, E., and Troulakis, G. (1996). Experiences with acute organophosphate poisonings in Crete. Vet. Human Taxcial. 38, 101-107. Tush, G. M., and Amstead, M.1. (1997). Pralidoxime continuous infusion in the treatment of organophosphate poisoning. Ann. Pharmacather. 31,441-444. Vale, J. A, and Scott, G. W. (1974). Organophosphorus poisoning. Guy's Hasp. Rep. 123, 13-25. Valero, A, and Golan, D. (1967). Accidental organic phosphorus poisoning: The use of propranolol to counteract vagolytic cardiac effects of atropine. Isr. J. Med. Sci. 3, 582-584. Vandekar, M. (1965). "Observations of the Toxicity of Two Organophosphorus and One Carbamate Insecticide in a Village Trial Performed by WHO Insecticide Testing Unit in Lagos During 1964", WHO Work. Doc. 65fToxl2.64, U.S. Govt. Printing Office, Washington, DC. Van Meter, W G., Karczmar, A. G., and Fiscus, R. R. (1978). CNS effects of anticholinesterases in the presence of inhibited cholinesterases. Arch. Int. Pharmacadyn.231,249-260. Vasilescu, C. (1982). Neuropathy after organophosphorus compounds poisoning. J. Neural. Neurosurg. Psychiatry 45, 942. Vasilescu, C., and Florescu, A (1980). Clinical and electrophysiological study of neuropathy after organophosphorus compounds poisoning. Arch. Taxicol. 43,305-315. Vasilescu, c., Alexianu, M., and Dan, A. (1984). Delayed neuropathy after organophosphorus insecticide (dipterex) poisoning: A clinical, electrophysiological and nerve biopsy study. J. Neural. Neurasurg. Psychiatry 47, 543-548.

Verberk, M. M., and Salle, H. J. A (1977). Effects of nervous function in volunteers ingesting mevinphos for one month. Taxicol. Appl. Pharmacal. 42, 351-358. Verpooten, G. A., and De Broe, M. E. (1984). Combined hemoperfusionhemodialysis in severe poisoning: Kinetics of drug extraction. Resuscitation 11,275-289. Wadia, R. S., Sadagopan, C., Amin, R. B., and Sardesai, H. V. (1974). Neurological manifestations of organophosphorus insecticide poisoning. J. Neural. Neurasurg. Psychiatry 37, 841-847. Wadia, R. S., Shinde, S. N., and Vaidya, S. (1985). Delayed neurotoxicity after an episode of poisoning with dichlorvos. Neural. India 33, 247-253. Wadia, R. S., Chitra, S., Amin, R. B., Kiwalkar, R. S., and Sardesai, H. v. (1987). Electrophysiological studies in acute organophosphate poisoning. J. Neural. Neurosurg. Psychiatry 50, 1442-1448. Wang, A-G., Liu, R.-S., Liu, J.-H., Teng, M. M.-H., and Yen, M. Y. (1999). Positron emission tomography scan in cortical visual loss in patients with organophosphate intoxication. Ophthalmology 106, 1287-1291. Wecker, L., Mrak, R. E., and Dettbam, W. D. (1985). Evidence of necrosis in human intercostal muscle following inhalation of an organophosphate insecticide. J. Enviran. Pathol. Taxicol. Oncal. 6, 171-175. Weeks, D. B., and Ford, D. (1989). Prolonged suxamethonium-induced neuromuscular block associated with organophophate poisoning. Br. J. Anaesth. 62,327. Weir, S., Minton, N., and Murray, V. (1992). Organophosphate poisoning in the U.K.: The National Poisons Information Service experience during 1984-1987. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 463-470. Butterworth-Heinemann, Oxford. Weizman, Z., and Sofer, S. (1992). Acute pancreatitis in children with anticholinesterase insecticide intoxication. Pediatrics 90, 204-206. Whittaker, M. (1968). The pseudocholinesterase variants. Differentiation by means of sodium chloride. Acta Genet. 18, 566-562. WHO (1986). "Organophosphorus Insecticides: A General Introduction." Environmental Health Criteria 63, World Health Organization, Geneva. WHO (1998). "The WHO Recommended Classification of Pesticides by Hazard and Guidelines to Classification 1998-1999." WHO/PCS/98.21, World Health Organization, Geneva. Whorton, M. D., and Obrinsky, D. L. (1983). Persistence of symptoms after mild to moderate acute organophosphate poisoning among 19 farm field workers. J. Taxicol. Environ. Health 11,347-354. Willems, J. L. (1981). Poisoning by organophosphate insecticide: Analysis of 53 human cases with regard to management and drug treatment. Acta Med. Milit. (Belg) 134, 7-14. Willems, J. L., and Belpaire, EM. (1992). Anticholinesterase poisoning: An overview of pharmacotherapy. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 536-544. Butterworth-Heinemann, Oxford. Willems, J. L., Langenberg, J. P., Verstraete, A. G., De Loose, M., Vanhaesebroeck, B., Goethals, G., Belpaire, EM., Buylaert, W. A., Vogelaers, D., and Colardyn, E (1992). Plasma concentrations of pralidoxime methylsulphate in organophosphorus poisoned patients. Arch. Taxical. 66, 260-266. Willems, J. L., De Bisschop, H. c., Verstraete, A. G., Declerck, c., Christiaens, Y., Vanscheeuwyck, P., Buylaert, W. A., Vogelaers, D., and Colardyn, E (1993). Cholinesterase reactivation in organophosphorus poisoned patients depends on the plasma concentrations of the oxime pralidoxime methylsulphate and of the organophosphate. Arch. Toxical. 67, 79-84. Wilson, B. W, Padilla, S., Henderson, J. D., Brimijoin, S., Dass, P. D., Elliot, G., Jaeger, B., Lanz, D., Pearson, R., and Spies, R. (1996). Factors in standardizing automated cholinesterase assays. J. Taxicol. Environ. Health 48, 187-195. Worek, E, Kirchner, T., Backer, M., and Szinicz, L. (1996). Reactivation by various oximes of human erythrocyte acetylcholinesterase inhibited by different organophosphorus compounds. Arch. Taxicol. 70,497-503. Worek, E, Eyer, P., and Szinicz, L. (1998a). Inhibition, reactivation and aging kinetics of cyclohexylmethylphosphonofluoridate-inhibited human cholinesterases. Arch. Taxicol. 72,580-587.

References

Worek, E, Widmann, R, Knopff, 0., and Szinicz, L. (1998b). Reactivating potency of obidoxime, pralidoxime, HI 6 and HL5 7 in human erythrocyte acetylcholinesterase inhibited by highly toxic organophosphorus compounds. Arch. Toxicol. 72,237-243. Worek, E, Diepold, c., and Eyer, P. (1999a). Dimethylphosphoryl-inhibited cholinesterases: Inhibition, reactivation, and aging kinetics. Arch. Toxicol. 73,7-14. Worek, E, Mast, U., Kiderlen, D., Diepold, c., and Eyer, P. (l999b). Improved determination of acetylcholinesterase activity in human whole blood. Clin. Chim. Acta 288, 73-90. Xintaras, C., and Burg, J. R (1980). Screening and prevention of human neurotoxic outbreaks: Issues and problems. In "Experimental and Clinical Neurotoxicology" (P. S. Spencer and H. H. Schaumburg, eds.), pp. 663674. Williams & Wilkins, Baltimore. Xintaras, c., Burg, J. R, Tanaka, S., Lee, S. T., Johnson, B. L., Cottrill, C. A., and Bender, J. (1978). "NIOSH Health Survey of Velsicol Pesticide Workers, Occupational Exposure to Leptophos and Other Chemicals. " DHEW (NIOSH) Publication 78-136, U.S. Govt. Printing Office, Washington, DC. Xue, S. Z. (1992). Acute anticholinesterase poisoning in China. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B.

1085

Ballantyne and T. C. Marrs, eds.), pp. 502-510. Butterworth-Heinemann, Oxford. Yilmazalar, A., and Ozyurt, G. (1997). Brain involvement in organophosphate poisoning. Environ. Res. 74, 104-109. Yokoyama, K., Ogura, Y., Kishimoto, M., Hinoshita, F., Hara, S., Yamada, A., Mimura, N., Seki, A., and Sakai, O. (1995). Blood purification for severe sarin poisoning after the Tokyo subway attack. lAMA 274, 379. Yoshida, M., Shimada, E., Yamanaka, S., Aoyama, H., Yamamura, Y., and Owada, S. (1987). A case of acute poisoning with fenitrothion (sumithion). Human Toxicol. 6,403-406. Zadik, Z., Blachar, Y., Barak, Y., and Levin, S. (1983). Organophosphate poisoning presenting as diabetic ketoacidosis. l. Toxicol.-Clin. Toxicol. 20, 381-385. Zoppellari, R, Borron, S. W., Chieregato, A., Targa, L., Scaroni, I., and Zatelli, R (1997). Isofenphos poisoning: Prolonged intoxication after intramuscular injection. Clin. Toxicol. 35,401-404. Zwiener, R. J., and Ginsburg, C. M. (1988). Organophosphate and carbamate poisoning in infants and children. Pediatrics 81,121-126.

CHAPTER

52 Carbamate Insecticides Donald J. Ecobichon Queen's University

52.1 INTRODUCTION Early testing of the natural carbamate, physostigmine (eserine) from the calabar bean (Physostigma venenosum) and the synthetic derivative, neostigmine, revealed that these highly polar compounds possessed no insecticidal activity. Aliphatic esters of carbamic acid were synthesized in the early 1930s and, while showing herbicidal and fungicidal activities, were not insecticidal. These agents will be discussed in other chapters (Chapter 66 for herbicides and Chapter 77 for fungicides). Interest in the carbamates was not renewed until the mid-1950s when there was a search for insecticides having anticholinesterase activity, more selectivity, and less mammalian toxicity than some of the organophosphorus esters then in use. This led to the synthesis of several potent aryl esters of methyl carbamic acid, these agents becoming the insecticides of the 1960s and 1970s. While a large number of carbamates have been synthesized, relatively few were developed further, the pesticide market being limited to less than 20 agents. The early history of carbamate insecticide development has been discussed by Kuhr and Dorough (1976) and Cremlyn (1978).

52.2 NOMENCLATURE As is shown in Fig. 52.1, the structure of all carbamate insecticides is based on carbamic acid (the monoamide of carbon dioxide), a highly unstable compound decomposing into carbon dioxide and ammonia. Carbamic acid may be stabilized by forming salts such as ammonium carbamate or by synthesizing alkyl or aryl esters. Replacement of one hydrogen associated with the nitrogen by a methyl group results in the formation of N-monomethy1carbamic acid which, when combined with an aryl ester substituent, results in significant alterations in various physicochemical properties and introduces insecticidal activity (e.g., bendiocarb, carbaryl, propoxur). An additional group of carbamate insecticides are derivatives of aliphatic oximes rather than esters, resembling aldehydes or ketones, known collectively as methy1carbamoyloximes, and possessing a high degree of toxicity (e.g., aldicarb and methomyl). Handbook of Pesticide Toxicology Volume 2. Agents

The majorIty of the carbamate insecticides in use are N-monomethyl carbamates, frequently referred to as N-methylcarbamates or just methy1carbamates. In this chapter, the insecticides will be referred to by the name commonly used in the literature, thereby simplifying discussion. Table 52.1 lists the methy1carbamates currently used in pest control with their chemical names and chemical structures (Baron, 1991).

52.3 CHEMISTRY The nature of the substituent groups alters both the physicochemical properties of the insecticide and the biological activity. Most of these insecticides dissolve readily in organic solvents but are only slightly soluble in water, thereby conferring varying degrees of lipid solubility. The exceptions are the methy1carbamoyloximes, the "oxime" carbamates aldicarb and methomyl, which are highly water soluble. A wide range of melting points (50 to 150°C) is found for these agents, determined largely by the size of the substituent group. Vapor pressures range from less than 5 x 10- 6 to 5 X 10-2 mmHg (Melnikov, 1971). While high melting points and low vapor pressures enhance the environmental stability of the compound, decomposition can be markedly enhanced by increased temperatures, a 10°C increase raising the hydrolysis rate two- to three-fold (Aly and EI-Dib, 1971; Fukuto et al., 1967). The environmental stability of carbamates is severely affected by photodegradation at short ultraviolet wavelengths (254 nm) and by oxidation upon exposure to air. These aspects of decomposition are discussed succinctly by Kuhr and Dorough (1976). Alkyl esters tend to be relatively unstable in the environment, in contrast to aryl esters. Stability can be enhanced by attaching additional substituents either to the aryl structure or to the carbamoylated nitrogen. While carbamates decompose slowly in water at an acidic pH, alkalinity enhances degradation since the substituent groups tend to draw electrons from around the ester linkage, thereby weakening it and accelerating the hydrolysis by hydroxide ions. Considering some of the structurally different agents shown in Table 52.1, incubation in an alkaline solution (0.01 M sodium barbital buffer, pH 9.3)

1087

Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved

1088

CHAPTER 52

Carbamate Insecticides I

I

!

0

I

I

I

I

I

,

H-;O-C-N-;-Hl I

Aryl alcohol Aliphatic oxime (R-N-O-) alcohol (R-C-O-)

H2 \ Methyl Dimethyl

Figure 52.1 The basic structure of carbamic acid, the monoamide of carbon dioxide, shoving the positions of substituant methyl, aryl, or aliphatic groups to produce methy1carbamate insecticides.

resulted in the determination of biological half-lives for methiocarb, carbaryl, mexacarbate, and propoxur of 0.4,0.5,2.3, and 3.1 hours, respectively (Abdel-Wahab et aI., 1966). Mono- or dimethylation of the carbamoyl nitrogen results in stabilization of the ester bond. N-monomethylcarbamates degrade slowly in the environment; for example, carbaryl at pH 7.0 has a halflife of 10 days. Dimethylcarbamates are exceedingly stable, the half-life of dimetilan (l-dimethyl-carbamoyl-5-methylpyrazol3-yl dimethylcarbamate) being approximately 100 days at pH ranging from 6 to 10 (Kuhr and Dorough, 1976). In addition to the direct-acting, anticholinesterase methylcarbamates, certain derivative agents such as benfuracarb, carbosulfan, mecarbam, and thiodicarb, known as procarbamates, have insecticidal activity but low mammalian toxicity until they are biotransformed to release biologically active agents or to yield nontoxic, readily excreted products. Fukuto (1983) showed that substitution of the remaining hydrogen on the carbamyl nitrogen of methylcarbamates reduced the mammalian toxicity due to slower conversion of the derivative to the original toxic insecticide. For example, carbosulfan and thiocarb are sulfide derivatives of carbofuran and methomyl, respectively. Most of the carbamate ester insecticides have low vapor pressures, which results in poor volatility at usual temperatures (Fig. 52.2). However, as is shown for aldicarb, oxamyl, and pirimicarb, increasing the temperature can markedly alter the vapor pressure, a factor that must be considered when using these agents in tropical countries. If other carbamate esters behave in the same manner, many would become highly volatile in climates having high temperatures. The propoxur toxicity incident in southern Nigeria, discussed by Vandekar (1965), is an example in point. Propoxur has a vapor pressure of 6.5 x 10-6 mmHg at 20°C; spraying huts and roofs at ambient temperatures of 70°C (140°F) caused acute toxicity among spraymen who wore some protective equipment. The effects were considered to be due to revolatilized propoxur from surfaces rather than from the suspended spray aerosol. Most of the carbamates in commercial use have relatively low water solubility (Table 52.2), a high level of solubility in polar solvents (ethanol, isopropanol, methanol, acetone), and limited-to-moderate solubility in nonpolar solvents (benzene, toluene, xylenes). This lipophilicity enhances the insec-

ticidal potency, the agents readily penetrating insect cuticles and tissues, but it also presents problems of oral and/or dermal absorption in other animal species, and enhanced storage in tissues. There are, however, exceptions, the high water solubility of some carbamates playing important roles in absorption, distribution (both in vivo and environmental), storage, and elimination, as well as governing the regulation of use. Example: note the degree of water solubility of the "oxime" carbamates, aldicarb, methomyl, oxamyl (Table 52.2). These agents are restricted for use on crops with a low water content. There has been illegal use of aldicarb resulting in consumer poisonings from melons and hydroponically grown cucumbers and widespread contamination of groundwater and community drinking water (Fiore et aI., 1986; Goes et aI., 1980; Goldman et aI., 1990a, 1990b; Zaki et aI., 1982).

52.4 TOXICOKINETICS 52.4.1 ABSORPTION

The most likely route of exposure to carbamates is via the skin in an occupational setting. The lipophilicity of this class of agents and the fact that most formulations contain organic solvents and emulsifiers insure a rapid dermal penetration and absorption into the systemic circulation. Temperature and humidity play important roles; high temperature and relative humidity enhance absorption, environmental conditions being reflected in less clothing being worn, greater areas of skin being exposed, and greater subdermal vasodilatation and perspiration, all resulting in a more complete absorption. Carbamates are readily absorbed in the gastrointestinal tract, the efficiency of absorption being somewhat guided by the vehicle(s) in which they are administered or are formulated. Exposure to low levels of carbamate residues in fresh fruits and vegetables may occur where regulatory tolerances have been established for food crop use. Residues in edible foods may be less efficiently absorbed, being trapped or bound in the food bolus. Under certain circumstances, inhalation may be an important route of exposure. The vapor pressures of some carbamates (Fig. 52.2) make them vulnerable to rapid revolatilization when applied under climatic conditions of high temperature in excess of 60-70°C. The previously mentioned accidental poisonings by propoxur in southern Nigeria are a case in point (Vandekar, 1965). The spraymen, applying a 5.0% suspension of propoxur on hut walls and roofs, had to terminate the operation within 2 to 3 hours when severe symptoms of toxicity were observed due to revolatilized agent rather than to the spray aerosol. Similar conditions would be encountered in greenhouses and mushroom barns, areas of high temperature and high humidity. These conditions alter the behavior of aerosols, the change being primarily to keep them suspended in air for long periods of time.

52.4 Toxicokinetics

1089

Table 52.1 Structure, Common Names, and Chemical Names of Carbamate Insecticides Agent

CAS number

Aldicarb

CAS 116-06-3

IUPAC chemical name

Structure Me

I

2-methyl-2-(methylthio)pro-

I

pionaldehyde O-methyl-

MeS.C.CH: N.O.CO.NHMe

TEMIKTM Bendiocarb

Me

CAS 22781-23-3

OC ~

FICAM™ ROTATE™

carbamoyloxime

I

°\;Me

/\e

2,2-dimethyl-l,3-benzodioxol4-yl methylcarbamate

°

O.C.NHMe

8

Carbaryl

CAS 63-25-2

Carbofuran

CAS 1563-66-2

I-naphthyl methylcarbamate

w,~

2,3-dihydro-2,2-dimethyl-7 benzofuranyl methylcarbamate

MeNHCO.O

Carbosulfan

CAS 55285-14-8

2,3-dihydro-2,3-dimethyl7 -benzofuranyl( (dibutyl-

ADVANTAGE™

amino )thio) methylcarbamate

MARSHAL™

Formetanate

CAS 23422-53-9 HCl

HCl CARZOL™

aminophenyl methylcarbamate hydrochloride

I

O.CO.NHMe

DICARZOL™ Methiocarb

0-'' '0',

3-dimethylaminomethylene-

CAS 2032-65-7

4-methylthio-3,5-xylyl methylcarbamate

Methomyl

CAS 16752-77-5

~

MeS"

C-N-OCNHMe

M./

LANNATETM Mexacarbate

CAS 315-18-4

ZECTRAN

Me

o.~~ }'I"~

S-methyl N-(methylcarbamoyloxy) thioacetimidate 3,5-dimethyl-4-(dimethylamino )phenyl methylcarbamate

Me

Oxamyl

CAS 23135-22-0

o

0

11 11 Me,N.C.?=N.O.C.NHMe

SM.

VYDATE™ Pirimicarb

CAS 23103-98-2

N,N-dimethyl-2-methylcarbamoyloxyimino-2-(methylthio )acetamide 2-dimethylamino-5,6-dimethy1pyrimidin-4-yl dimethylcarbamate

(continues)

1090

CHAPTER 52

Carbamate Insecticides

Table 52.1 (continued) Agent

CAS number

Propoxur

CAS 114-26-1

Structure

IUPAC chemical name 2-isopropoxyphenyl methylcarbamate

Thiodicarb

CAS 59669-26-0

Dimethyl N,N-(thiobis(methyl-

rH, ~ CH,-S-C=N-O-C-7-CH,

imino)carbony loxy)-bis-

S

(ethanimidothioate)

I

CH,-S-r=N-O-ji-N-CH, CH,

52.4.2 BIOTRANSFORMATION A number of excellent reviews consider the biotransformation of carbamate insecticides, induding those of Knaak (1971), Ryan (1971), Fukuto (1972), Kuhr and Dorough (1976), Wilkinson (1976), Kulkarni and Hodgson (1980), and the IPCS

W21

0

(1986). The initial response of any species exposed to a carbamate ester is to convert the chemical into more polar forms for ready excretion via the urine. To achieve this, the organism calls upon Phase I and Phase 11 detoxification mechanisms in tissues to create water-soluble, easily excreted, and less toxic by-products. While these insecticidal esters are suscep-

CARBAMATE ESTERS VAPOR PRESSURE (mmHg)

1/100

•

VP (mmHg)

w3

1/1,000

1/10,000

DlllETILAN

·~A e

0

ALllOXYCARB

PIRIMICARB

/

/0~OMETHOMYV oo )y 1

CARBARYL

THIODICARB

METHACARB

1/100,000

BUFENCARB

0/ ... 0

/ PROPOXUR

o o

w6

CARBOFllRAN

BENDIOCARB

1/1.000,000 o

10

20

i

CARBOSULF AN

30

40

50

60

70

TEMPERATURE (CC) Figure 52.2 The vapor pressures of carbamate insecticides determined at 20°C, 25°C, or 30°, with examples of altered vapor pressures at elevated temperatures as might be encountered in tropical countries.

52.5 Mechanism Table 52.2 Relative Water Solubility of Carbamate Ester Insecticidesa,b Agent

Solubility (g/L)

Aldicarb

6.0

Bendiocarb

0.04

Carbaryl

0.7

Carbofuran

0.7

Carbosulfan Fonnetanate HCl Methiocarb Methomyl Mexacarbate Oxamyl Pirimicarb

0.0003 >500 0.01 58 0.1 280 2.75

Propoxur

2.0

Thiodicarb

0.035

a Data from Baron (1991) and the Merck Index, 12th edition (1996).

bMeasured at 20--25°C.

tible to a variety of enzyme-catalyzed detoxification reactions, the principal biotransformation pathways involve oxidation and hydrolysis, with conjugation of some of the cleaved products (Ecobichon, 1994a). The nature and position of the substituent groups on the ether oxygen or the nitrogen exert an important role over the rate and pathway of biotransformation. Being esters, carbamate insecticides are susceptible to hydrolysis by nonspecific carboxylesterases ubiquitously distributed throughout the tissues of species from insects to humans. The products formed are identical to many of those produced by chemical (alkali, water) hydrolysis in that an aryl alcohol plus methyl- or dimethyl-carbamic acid will be formed. The unstable methylated carbamic acids will rapidly decompose into carbon dioxide and mono- or dimethylamine. Rates of hydrolysis in vivo are governed by the molecular structure of the agent, the specificity or selectivity of the carboxylesterases for particular agents, and interspecies differences. Carbamate esters are actually poor substrates for many tissue esterases. The hydrolysis of the various carbamate esters is highly individualistic, only a certain percent hydrolysis occurring with different agents (Schlagbauer and Schlagbauer, 1972). A generalization that carbamates can be hydrolyzed by tissue enzymes requires rigorous testing with several carbamates as substrates. The ubiquitous distribution of the reactive hemoprotein, cytochrome P-450, and the various isoenzymatic forms in tissues of all life forms, point to a commitment to the oxidative detoxification of a broad spectrum of both endogenous and exogenous chemicals as a protective measure. These hemoprotein isoenzymes, in conjunction with molecular oxygen, flavoproteins, cytochrome-b5 and reduced nicotinamide adenine dinucleotide phosphate (NADPH), can initiate a variety of enzymatic oxidative/reductive reactions depending upon the nature of the substituent groups on the carbamate ester. Oxidative reactions can be simplified into two groups: (1) oxidation of appropriate side chains, for example, hydroxylation of N-methyl groups

1091

and/or hydroxylation of methyl substituents on aryl moieties to form hydroxymethyl groups, N-demethylation of secondary amines attached to the aryl moiety; and (2) ring hydroxylation through the formation of an epoxide intermediate. In addition, thiocarbamates may undergo S-oxidation by these same oxidative mechanisms; for example, aldicarb can be converted into a sulfoxide and/or a sulfone, depending upon the species being studied. In conjugative or Phase 11 detoxification reactions, a functional group on the molecule, introduced as a consequence of hydrolytic or oxidative biotransformation, is enzymatically reacted with an endogenous substance in the tissues of the life form to produce water-soluble, biologically inactive, and readily excreted products. Depending upon the species of plant or animal being studied, a variety of products may be formed but, in general, the products may be classified as sulfates, glucuronides, glucosides, amino acid conjugates, acetylated amines, or glutathione conjugates, the last being excreted as mercapturic acid derivatives. In mammalian species, the cleaved aryl substituent(s) are conjugated to produce sulfates, glucuronides, and mercapturates. Biotransformation/degradation in aquatic systems, plants, and by microorganisms has been reviewed (IPCS, 1986). Hydrolysis of the carbamate ester bond is the major degradation pathway in soils. In plants, oxidative processes result in ring hydroxylation followed by conjugation with either amino acids (cysteine), phosphates, or sugars to form glycosides. Hydrolysis can occur in some plant species. 52.4.3 ELIMINATION

There is little evidence of extensive carbamate bioaccumulation since biotransformation is relatively rapid. There is at least one report of persistent toxicity in a human intoxication, the signs and symptoms disappearing slowly when the afflicted individual was removed from the source (Branch and Jacqz, 1986a). However, this effect might have been related to slow recovery from agent-induced neuropathy, or altered metabolism, rather than the clearance of any body burden. Excretion of the water-soluble by-products of detoxification occurs relatively rapidly via the urine and/or feces in most vertebrate species. Glucuronide and sulfate derivatives of the aryl substituents are the major products found in the urine. Small amounts of the parent carbamate may be excreted in the urine. Mercapturates are usually found in mammalian feces if they are not broken down in the intestinal tract, reabsorbed systemically, and recycled to form other products to be excreted in the urine.

52.5 MECHANISM Like the organophosphorus ester insecticides, the carbamates elicit toxicity by inhibiting nervous tissue acetylcholinesterase (AChE). However, it is a transient, reversible inhibition, since there is a relatively rapid reactivation of the enzyme in the

1092

CHAPTER 52


~Ka k

k

k

EH ~EHAB ~EA ~EH

+

AB

-1

+

+

BH

AOH

Figure 52.3 A schematic diagram shoving the mechanism of interaction between a methylcarbamate insecticide (AB) and acetylcholinesterase (EH), depicting the unstable intermediate complex (EHAB), the carbambylated enzyme (EA), the leaving group (BH), and the spontaneously decarbamoylated enzyme (EH) and the released methyicarbamic acid (AOH).

presence of "tissue" water. The biological effects of the accumulating acetylcholine (ACh) tend to be of short duration, in terms of hours rather than in days to weeks as is seen with organophosphorus esters. As is shown in Fig. 52.3, a reversible carbamate-AChE complex (EHAB) is formed, followed by the hydrolysis of the ester bond and the loss of the aryl or alkyl substituent (BH), the result being a carbamylated enzyme (EA) which is unstable and hydrolyzes in the presence of water to release free and active enzyme (EH) (Ecobichon, 1996). The differences between organophosphorus and carbamate ester inhibitors lie in the rate constants for the various steps in the reaction(s) (Table 52.3). Both classes of insecticides have high affinity constants (Ka = k-ll kl) for the active center of the enzyme, the interaction with the enzyme (EHAB) being almost instantaneous (Hastings et aI., 1970; Reiner, 1971). The rate of carbamy1ation of the enzyme depends largely on molecular complementarity and reactivity, the latter depending on the nature of the leaving group, for example, phenolic and oxime substituents being somewhat better than benzyl alcohols. While carbamylation appears to be reversible from the point of view of the enzyme, it is not reversible from the point of view of the carbamate which is cleaved and loses anticholinesterase potency in the process (Baron, 1991). Thus, the carbamylation constant, K2, will vary considerably between carbamate esters. Acetylcholinesterase inhibition varies in degree with the rate of the EHAB-to-EA complex formation and the relative Ka of each compound. The decarbamylation constant, K3, would be the same for all N-methylcarbamates, the moiety (A) adhering to the enzyme being identical in all cases, with aqueous hydrolysis at the same rate resulting in the formation of free, uninhibited enzyme (EH). By contrast, the phosphorylation of AChE is regulated by (1) the electron-withdrawing power of the "leaving" substituent, which is highly variable between chemicals; and (2) the nature of the alkyl (methyl, ethyl, isopropyl, methylarnido, ethylamido, etc.) substituents on the ester. The rate of reactivation of AChE is governed by the rate constants, K2 and K3, frequently quite different from those for carbamate esters (Table 52.2). The phosphorylated enzyme can be quite stable, aqueous hydrolysis being very slow in many cases. The degree of inhibition of nervous tissue AChE and/or plasma pseudocholinesterase (PChE) by carbamates is variable, being dependent upon the specificity of the agent for the active site of the enzyme, the rate constants for complex formation,

Table 52.3 Kinetic Rates of Inhibition of Cholinesteerases By Carbamate and Organophosphorus Esters Kinetic

Reaction rates

Parameter

constantsa

Organophosphorus

Carbamate

Complex Formation

LJ/kJ

Rapid (high affinity)

Rapid (high affinity)

Inhibition Rate

k2

Rapid to moderately rapid

Variable

Reactivation Rate

k3

Slow to extremely slow

Relatively rapid

aSee Fig. 52.3.

spontaneous reversal of the complex, the carbamylation of the enzymes, and the decarbamylation stage. Carbamate variability is reflected in the relative rate(s) of recovery of AChE and PChE and the level of exposure. In mild-to-moderate cases of intoxication, carbamates may have little effect on PChE while severely inhibiting the AChE (both erythrocytic and nervous tissue). In severe intoxications, both PChE and AChE will be markedly inhibited. As an example, in a case of a suicidal attempt with a propoxur formulation, the blood sample taken within an hour of visiting the emergency room revealed no activity of either erythrocytic AChE or PChE, but the sample taken 6 hr later showed 60% inhibition of erythrocytic AChE and no residual inhibition of PChE (Ecobichon, unpublished). The transient nature of carbamate-induced inhibition of AChE poses several problems in the attempt to measure the level of inhibition. Care must be taken to keep blood and tissue samples cold or frozen during transportation to the laboratory prior to analysis. For example, blood samples should be kept on ice, centrifuged under refrigerated conditions to recover both the plasma and the erythrocyte fractions, and frozen at -20°C immediately until assayed. Spontaneous reversal of the inhibition is rapid and can be accelerated by (1) the time interval between sampling and analysis; (2) the dilution of the sample; (3) the addition of substrate, usually acetylcholine at high concentration, which competes successfully for the enzymatic active site in either of the EHAB and EA complexes; and (4) the duration of the assay time. Laboratory assays of cholinesterase inhibition must be very rapid (less than 3 minutes), and must employ minimal dilution and minimal amounts of substrate. Modifications can be made to the colorimetric assay of Ellman et al. (1961) to meet the restrictive criteria mentioned above. It is fallacious to measure cholinesterase activities in biological fluids and tissues collected in subchronic and chronic exposure studies 24 hours after the last exposure. Such assays should be done immediately following the last treatment since, as was seen in the case of acute aminocarb (4-dimethylaminom-tolyl methylcarbamate) toxicity in rats, recovery of the vital cholinesterases was complete by 6 hours post-treatment (Vassilieff and Ecobichon, 1983). Little or no inhibition would be observed if the activities were measured 24 hours following the last exposure.

52.6 Toxicology

1093

Table 52.4 Signs and Symptoms of Anticholinesterase Insecticide Poisoning Nervous tissue and receptors affected

Site affected

Manifestations

Parasympathetic autonomic

Exocrine glands

Increased salivation, lacrimation, perspiration

Eyes

Miosis (pinpoint and nonreactive), ptosis, blurring of vision, conjunctival

Gastrointestinal tract

Nausea, vomiting, abdominal tightness, swelling and cramps, diarrhea,

Respiratory tract

Excessive bronchial secretions, rhinorrhea, wheezing, edema, tightness in chest,

Cardiovascular system

Bradycardia, decrease in blood pressure

(muscarinic receptors)

injection, "bloody tears"

postganglionic nerve fibers

tenesmus, fecal incontinence bronchospasms, bronchoconstriction, cough, bradypnea, dyspnea

Parasympathetic and sympathetic

Bladder

Urinary frequency and incontinence

Cardiovascular system

Tachycardia, pallor, increase in blood pressure

Skeletal muscles

Muscle fasciculalions (eyelids, fine facial muscles), cramps, diminished

autonomic fibers (nicotinic receptors) Somatic motor nerve fibers

tendon reflexes, generalized muscle weakness in peripheral and respiratory

(nicotinic receptors)

muscles, paralysis, flaccid or rigid tone Restlessness, generalized motor activity, reaction to acoustic stimuli, tremulousness, emotionallability, ataxia Brain (acetylcholine receptors)

Central nervous system

Drowsiness, lethargy, fatigue, mental confusion, inability to concentrate, headache, pressure in head, generalized weakness Coma with absence of reflexes, tremors, Cheyne-Stokes respiration, dyspnea, convulsions, depression of respiratory centers, cyanosis

Source: From Ecobichon and Joy (1982).

The toxicity of carbamates in mammals can be predicted in vitro by the degree to which they inhibit AChE activity, and in vivo by the severity of the clinical manifestations (Feldman, 1999).

52.6 TOXICOLOGY 52.6.1 MODE OF ACTION The insecticidal carbamates, like organophosphorus esters, exert their effects by inhibiting nervous tissue AChE found in the synaptic spaces and on the postsynaptic membranes of all neurons, using acetylcholine as a chemical neurotransmitter. The role of this enzyme is to terminate, by hydrolysis, the biological actions of the neurotransmitter, thereby restoring the acetylcholine receptors to a state where they can receive the next chemical stimulus. With the loss of this regulating mechanism, the accumulating, nondetoxified acetylcholine (ACh) continues to stimulate specific receptor types, eliciting a spectrum of characteristic clinical signs and symptoms of intoxication (Cranmer, 1986; Ecobichon, 1994b, 1996). Due to the transient nature of carbamate-inhibited nervous tissue AChE, acute intoxication by carbamates is generally resolved within a few hours. Depending upon the level of exposure, the clinical signs and symptoms may appear quite rapidly, be of mild-to-severe intensity, but last for a relatively short duration, disappearing within six hours.

Acetylcholine is an important neurotransmitter at parasympathomimetic, postganglionic nerve endings that are not under voluntary control (autonomic pathways) and which include the exocrine glands, the eyes, the gastrointestinal tract, the respiratory tract secretions, the cardiovascular system, and the bladder (Ecobichon, 1994b). Such neuronal junctions are stimulated specifically by the chemical muscarine and are blocked by atropine, an agent used in treating intoxications to alleviate what are called muscarinic effects, which frequently appear early in any carbamate intoxication. Acetylcholine is also a neurotransmitter at the interneuron ganglia of both the parasympathomimetic and the sympathomimetic divisions of the autonomic nervous system, the major effects seen being a stimulation of the ganglia of sympathetic, adrenergic neurons and the adrenal medulla (releasing epinephrine), with observed clinical signs in the cardiovascular system (tachycardia, vasoconstriction) resulting in increased heart rate and blood pressure and pallor. These neuronal junctions are also stimulated by nicotine, giving rise to the term nicotinic receptors. Acetylcholine stimulates skeletal neuromuscular junctions under voluntary control (the somatic nervous system), these neuromuscular receptors characteristically being stimulated by nicotine and blocked by the agents d-tubocurarine and succinyldicholine. These receptors are known as nicotinic receptors. Overstimulation of such receptors by acetylcholine causes generalized increased motor activity with muscle fascicula-

1094

CHAPTER 52


tions. An excess of neurotransmitter may lead to receptor blockade' the evident clinical signs being skeletal muscle paralysis and/or generalized muscle weakness, as well as respiratory distress due to paralysis of the diaphragmatic and intercostal muscles (Ecobichon, 1996). Acetylcholine has important roles in the central nervous system, cholinergic brain receptors being both muscarinic and nicotinic in nature. The respiratory center is cholinergic in nature, controls the respiratory rate (overstimulation causes blockade' respiration is impaired or stops), and responds to atropine treatment. Convulsions are elicited through centrally located neurons, and a host of other effects (disorientation, anxiety, memory loss, drowsiness, lethargy, fatigue, general malaise) appear to have central origins. The acronym "MUDDLES" (i.e., miosis, urination, diarrhea, diaphoresis, lacrimation, excitation of the CNS, salivation) is an accurate description of the principal effects of AChE inhibition (O'Malley, 1997). A detailed listing of clinical signs and symptoms observed in animals and humans is presented in Table 52.4 (Ecobichon, 1996). The appearance of none, some, or all of the symptomatology is largely dependent upon the compound and the level of exposure (Cranmer, 1986; Vandekar, 1965; Vassilieff and Ecobichon, 1983). The rate(s) of recovery will be dependent upon the rate(s) of biotransformation and excretion of the particular chemical, most intoxications being brief but with some signs, particularly in humans, persisting for weeks after exposure. The persistent peripheral- and central-mediated symptoms will be addressed in a later section. The acute toxicity of different carbamate insecticides correlates well with their anticholinesterase activity, particularly with the inhibition of erythrocytic AChE (Vandekar et aI., 1971; Vassilieff and Ecobichon, 1983). Intoxications showing obvious cholinergic signs of toxicity may be accompanied by little or no inhibition of cholinesterase activity, this phenomenon being due to a number of assay problems: (1) the selection of the proper enzyme for assay, the plasma PChE being less sensitive to carbamates than the erythrocytic AChE; (2) the selection of an inappropriate substrate for the enzyme being assayed; (3) the ease with which the carbamoylated cholinesterase spontaneously reactivates following dilution, lysis in the case of erythrocytes, or addition of substrate, all factors related to the assay method being used; and (4) the interval between exposure and blood sampling, during which time the carbamate may be degraded or the inhibition may be reversed in vivo (Berry, 1971; Ecobichon and Comeau, 1973; Iverson, 1975; Reiner, 1971; Wilhelm and Reiner, 1973). Particular attention should be paid to the analytical method, which should incorporate minimum exposure-to-collection intervals, minimum dilution of sample, minimum assay time, minimum substrate concentration, and the appropriate pH.

52.6.2 ACUTE TOXICITY-ANIMAL One index of acute toxicity is reflected by an LD50 value determined in suitable animal species, the agent being administered

Table 52.5 Acute Oral Toxicity of Carbamate Insecticides (Technical)Q Chemical Aldicarb

Species

Sex

rat

both

0.46-1.23

mouse

both

0.38-1.50

rabbit

1.3

guinea pig Bendiocarb

rat

mouse

Carbaryl

1.0 both

34--156

M

138

both

350--657

both

28-45

M

175

both

173-380

rabbit

both

35-40

guinea pig

F

35

dog

both

rat

both

mouse

both

108-650

rabbit

?

710

guinea pig cat

233-850

250--795

?

swine

125-250 1500--2000 >1000

monkey rat

ca. 300

280

dog

Carbofuran

LDso (mg/kd

M

5.3-13.2 2.0

mouse

19

dog rat

both

90--250

mouse

both

33-124

rabbit

both

37-53

rat

both

15-26

mouse

both

13-25

dog

both

19

rat

both

13-135

guinea pig

both

14--100

dog

both

10--25

Methomyl

rat

both

12-48

Mexacarbate

rat

both

8.5-12.0

Oxamyl

rat

both

2.5-16.0

mouse

both

2.3-3.3

guinea pig

M

7.1

rat

F

mouse

F

107

dog

both

100--200

rat

both

80--191

mouse

both

Carbosulfan

Formetanate HCl

Methiocarb

Pirimicarb

Propoxur

37-109 40

guinea pig Thiodicarb

68-221

rat

both

mouse

both

39-136 226

guinea pig

M

160

rabbit

both

556

monkey

both

467.2

aData modified from Baron (1991). bValues determined using different vehicles.

52.6 Toxicology Table 52.6 Acute Dennal Toxicity of Carbamate Insecticides (Technical)a Sex

LDso (mg/kg)b

rat

both

rabbit

M

3.2->10 5.0--20 566 >5000 >1000 >2000 >2000 >10200 >300-->5000 >2000 >1000-->2400 556->1500 >2000 >1200 740 >500 >500 1000-->2400 >500 2540 >6310

Chemical

Species

Aldicarb Bendiocarb

rat

both

Carbaryl

rat

both

Carbofuran

rat

both

rabbit

both

Carbosulfan

rabbit

both

Fonnetanate HCl

rabbit

both

Methiocarb

rat

both

rabbit

both

Methomyl

rat

M

rabbit

both

Mexacarbate

rabbit

both

Oxamyl

rat

M

rabbit

M

rat

F

Pirimicarb

rabbit Propoxur Thiodicarb

rat

both

rabbit

M

rat

M

rabbit

both

QData modified from Baron (1991). bYalues detennined using different vehicles.

via the route(s) by which humans are most likely to acquire the chemical (Ecobichon, 1996). To this end, for comparative purposes, Tables 52.5, 52.6, and 52.7 list the oral, dermal, and inhalation LD50s of the carbamate ester insecticides of commercial interest, these tables being reproduced from the 1991 edition of Hayes' and Laws' Handbook of Pesticide Toxicology (Baron, 1991). The specific references for any particular LD50 may be found in that text. Table 52.7 Acute Inhalation Toxicity of Carbamate Insecticides (Technical)a Chemical

Species

Carbaryl

rat

Carbosulfan

rat

Sex

both

Fonnetanate HC]

rat

both

Methiocarb

rat

both

Methomyl

rat

M

Oxamyl

rat

both M

Pirimicarb

rat

?

Propoxur

rat

M

Thiodicarb

rat

both

QData modified from Baron (1991). bYalues detennined over different time intervals (1--6 hr).

LDso (mg/L)b

0.005-0.023 0.61-1.53 0.29-2.8 >0.322 0.45 0.12-0.17 0.064 ca. 0.3 >1.44 0.116-0.22 >0.20

1095

The signs and symptoms of carbamate-induced, acute toxicity observed in various animal species should be comparable for the different insecticides, given that adequate, toxic doses have been administered. Considering carbaryl as a prototype carbamate ester of moderate toxicity, the following signs will be seen in mammals in approximate order of appearance, beginning some 15 to 30 minutes after oral administration: salivation, lacrimation, increased respiration with rales due to bronchial secretions, urination, defecation, and muscle fasciculations and tremors progressing to mild-to-moderate convulsions within 90 minutes of treatment. More severe intoxications may be characterized by pupillary constriction, profuse salivation, chromodacryorrhea, respiratory difficulty, loss of bladder and bowel control, muscular spasms and weakness, prostration, and incoordination. While most of the symptoms will disappear within 6 hours of exposure, a few, such as diarrhea, chromodacryorrhea, and muscle weakness, may persist beyond 24 hours posttreatment. Death is due to respiratory collapse if intoxication is severe. A number of studies have examined the behavioral effects of anticholinesterase-type insecticides immediately following treatment. Carbaryl produces CNS depressant effects, making it obvious that ACh plays a significant role in memory, cognitive, and motor functions; many of the adverse effects are ameliorated by such cholinolytic agents as atropine or scopolamine (Kurtz, 1977; Takahashi et aI., 1991). The acute administration of carbaryl (1.0, 3.0, 5.0, and 10 mg/kg) to rats resulted in a dose-related decrease in variable interval response rates in a learned procedure of pushing a lever to receive a food pellet (Anger and Wilson, 1980). The rate decreases were 55 to 77,81 to 94, and 88 to 100 percent at 3.0, 5.0, and 10 mg/kg, depending upon the route (ip or im) of administration. In other acute experiments, both propoxur and carbaryl caused post-treatment reductions in motor activity (open field and figure eight mazes) in a dose-dependent manner (Ruppert et aI., 1983). However, maze activity recovered within 30 and 60 minutes, while the brain AChE activities remained depressed for 120 to 240 minutes for propoxur and carbaryl, respectively. These results suggest several possibilities, including no association between behavior and AChE or some threshold effect of ACh counteracted by the spontaneous recovery of sufficient AChE activity. A more recent intoxication in both sheep and humans involved aldicarb contamination of a buckwheat field into which the sheep had been moved (Grendon et aI., 1994). Of the 318 sheep, 288 died rapidly from acute poisoning, exhibiting respiratory distress, hypersalivation, miosis, diarrhea, and seizures. Reduced erythrocytic AChE activity was measured in five animals tested, and levels of aldicarb ranging from 0.19 to 344 ppm were detected in the rumen contents of 13 of the exposed animals. The remaining live sheep, given atropine, showed some clinical improvement but continued to have poor appetites, showed body weight loss, and, within 3 weeks, either had died or were euthanized. The shepherd was affected with difficulty in breathing and a burning sensation in his throat. Those arriving to assist the owner experienced classical acute signs and

1096

CHAPTER 52


symptoms. Chronic symptoms, evident in some of the humans, will be considered in a later section. From acute animal studies, reports in the literature suggest that carbamates possess another mechanism of action in addition to that of inhibition of nervous tissue AChE. In some experiments, animals died within a few minutes of receiving the agent, seemingly from a marked anesthetic-like or "narcotic" effect accompanied by severe respiratory difficulty (dyspnea) and eventual respiratory failure (Vandekar et al., 1971). These effects have been observed with intravenous and intraperitoneal administration but not with oral administration. The "narcotic" effect was produced only by carbamates of low toxicity (high LDso values) (Ecobichon, 1994a). Hypotheses have suggested that such agents cause a complete blockade of nerve conduction by direct action at the level of sodium ion transport across axonal membranes and/or at motor end plate, postsynaptic, ACh receptors, both effects indicating a possible interaction of the agent with membranes to cause perturbation. A similar effect was noted following the intravenous injection of some organophosphorus ester insecticides of low toxicity (Heath, 1961). 52.6.3 ACUTE TOXICITY-HUMAN

Despite statements to the effect that "most" carbamate ester insecticides are relatively safe and produce only transient, short-term toxicity in animals following acute administration, carbamate toxicity does occur in humans, particularly in cases of ingestion by accident or with suicidal intent (Ecobichon, 1994a; Hayes, 1982). Invariably, acute toxicity in humans is associated with one of the more acutely toxic carbamates such as aldicarb, methomyl, or propoxur. Because of long and extensive use, several reported carbaryl intoxications have been summarized in the literature (Cranmer, 1986; Dickoff et aI., 1987; Farago, 1969; Hayes, 1982). Fatalities have occurred, one particular case being well documented (Farago, 1969). In this situation, a 39-year-old male purposefully drank approximately 500 mL of Sevin-80™ (80% concentration of carbaryl), with death occurring some 6 hours after ingestion, even with prompt hospitalization, gastric lavage, and antidote administration. Quantitative analysis of tissues and fluids revealed carbaryl concentrations (ppm): stomach lavage fluid (2,446), stomach contents (148), intestinal contents (176), blood (14), liver (29), kidney (25), and urine (31). No measurements of cholinesterases were reported (Farago, 1969). Overexposure to mexacarbate as a consequence of leakage from a high-pressure pump line in a cockpit resulted in acute intoxication of a copilot (Richardson and Batteese, 1973). Approximately 1lO minutes post-exposure, the copilot experienced the characteristic cholinergic symptoms. On landing, the affected individual was unable to stand, shook uncontrollably, and developed paresthesia and paralysis of the hands and arms and slurred speech while in transit to the hospital. After atropine treatment was initiated at the hospital, the symptoms disappeared rapidly and the patient was discharged three hours

after admission, the only residual effects being headache and weakness for the remainder of the day. Carbofuran-induced occupational intoxication has occurred, two plant employees being affected while preparing a lO% granular formulation (Tobin, 1970). Taken to their respective physicians within 3 hours after the onset of symptoms, one patient received atropine, while the second was not treated since the symptoms appeared to be regressing. The atropine-treated patient recovered fully within 30 minutes, while the untreated patient recovered over the course of 2 to 3 hours. Hayes (1982) reported several interesting poisoning and voluntary consumption cases never published in the literature, giving insight to the amount(s) required to elicit toxicity. In one case, a physician, testing the efficacy of carbaryl as an anthelmintic, ingested 250 mg of carbaryl (2.8 mg/kg) and experienced sudden, violent epigastric pain and profuse sweating within 20 to 30 minutes, followed by a gradually developing lassitude and vomiting (twice). By 3 hours after self-administration, and after having taken 3.0 mg of atropine, improvement of the symptoms occurred and, by 4 hours, the individual had recovered completely. In a second case described by Hayes (1982), arising from a personal communication from a professional colleague who was also testing the anthelmintic efficacy of carbaryl, an oral dose of 420 mg of carbaryl (5.45 mg/kg) was ingested on an empty stomach. The signs and symptoms appeared in the order shown in Table 52.8. The severity of the symptoms reached a maximum by 120 minutes after ingestion of the carbaryl. By the third hour after ingestion, the symptoms were dissipating, and the patient felt "practically normal" by the fourth hour. In a limited volunteer study reported by Wills et al. (1968), volunteers receiving acute oral doses of carbaryl (0.5, 1.0, and 2.0 mg/kg) showed Table 52.8 Time Sequence of Appearance of Symptoms in a Carbaryl Intoxicationa Time interval Symptoms observed

(min)

Blurred vision persisting for 10 to IS min Nausea Lightheaded Nausea, lightheaded, continuing sweating Hyperperistalsis Persistent nausea without vomiting or diarrhea

120

Weakness Pulse rate-normal 64/min Respiratory rate-I8/min No pinpoint pupil No lacrimation, salivation or rales Answered questions readily and correctly Improvement of symptoms

180

Some increase in strength Practically normal, walking about

240 a Data from Hayes, Jr. (1982). b Atropine

administered, 20 mg at 10 min 2.8 mg at 17 min.

52.6 Toxicology no subjective or objective effects. In a further study in which carbaryl doses of 0.06,0.12, and 0.13 mg/kg/day were administered for six consecutive days to volunteers, few conclusive abnormalities attributable to carbaryl were observed. The suicidal poisoning described by Dickoff et al. (1987) details the case of a 23-year-old male who swallowed 100 mL of Ortho-Liquid Sevin™ containing 27% carbaryl, showed the classical signs and symptoms reported earlier but, following recovery from the acute cholinergic toxicity, developed an acute weakness in the arms and legs associated with a peripheral, axonal neuropathy. The apparent delayed neurotoxicity will be addressed in a later section. One unpublished incident reported by Hayes (1982) involved the aerial application of carbofuran in place of carbaryl on corn, with the rapid onset of mild-to-moderate symptoms and the rapid recovery from the intoxication. Within 12 hours of application, 142 teenaged boys and girls entered the sprayed field to remove tassels from the plants. Within 6 hours, 74 complained of dizziness, nausea and/or blurred vision; some 45 received medical aid, with 29 being hospitalized and all but one individual being released within a few hours. Propoxur has frequently been associated with occupational intoxications, volunteer trials, accidents and suicidal attempts (Hayes, 1982; Vandekar et aI., 1968, 1971). Voluntary ingestion of propoxur at concentrations of 1.5 mg/kg resulted in a depression of erythrocytic AChE to 27% of normal at 15 minutes, with a return to normal activity by 120 minutes postingestion. Symptoms such as discomfort, head pressure, blurred vision, pallor, nausea, sweating, increased pulse rate (from 76 to 140/minute), increased blood pressure (from 130/90 to 175/95 mmHg) were observed by 60 minutes after ingestion and, by 30 minutes, pronounced nausea, repeated vomiting, and profuse sweating were observed, these symptoms persisting through 45 minutes after ingestion. By 60 minutes, the individual was feeling better, the signs and symptoms disappearing and, by 120 minutes, was feeling well enough to eat (Vandekar et aI., 1971). Ingestion ofpropoxur at a level of 0.36 mg/kg caused a rapid decrease in erythrocytic AChE to 57% of normal activity within 10 minutes and recovery within 180 minutes, and produced initial abdominal discomfort, blurred vision, moderate facial redness, and sweating lasting about 5 minutes, with recovery by 3 hours after ingestion (Vandekar et aI., 1971). A more recent, suicidal attempt, using a tickand-flea preparation of propoxur, has been described, giving a detailed list of classic signs and symptoms of severe toxicity over an 8-hour period, including: unconsciousness, labored breathing, bilateral pinpoint pupils, salivation, reduced respiratory movements, regular heart rhythm but with frequent premature ventricular contractions, incontinence with watery stool, cyanosis in the extremities, sweating, no response to painful stimuli, no gag or deep-tendon reflexes, downward deflection of toes during the plantar reflex, myotonic jerks of all extremities, and grand-mal seizure (Remaley et aI., 1988). The patient spontaneously awakened approximately 8 hours after hospital admission and was discharged 4 days after the episode.

1097

The most toxic of the carbamate esters is the systemic insecticide aldicarb, registered for use on citrus fruits, cotton, potatoes, peanuts and soybeans. It is not registered for use on any fruits or vegetables having a high water content. Surprisingly, this highly water-soluble chemical has been the source of periodic outbreaks of poisoning, usually associated with the inappropriate, even illegal, use in hydroponically grown cucumbers (CDe, 1979; Goes et aI., 1980), various melons (CDC, 1986; Goldman et al., 1990a, b), or the contamination of drinking water in New York and Wisconsin (Fiore et aI., 1986; Sterman and Varma, 1983; Zaki et aI., 1982). Levels of aldicarb in cucumbers ranged from 6.6 to 10.7 ppm (Goes et aI., 1980). In the melons, the active ingredient was not the parent insecticide but the equally water-soluble, biologically active metabolite, aldicarb sulfoxide (Goldman et al., 1990a). Estimates of the amounts of aldicarb sulfoxide ingested and responsible for intoxications ranged from 0.0011 to 0.06 mg/kg body weight (Goldman et aI., 1990a). In drinking water derived from groundwater sources, levels of aldicarb ranged between 8 and 75 ).Lg/L in Suffolk County, NY, while in Wisconsin, the levels ranged between 1.0 and 61 ).Lg/L (Fiore et aI., 1986; Zaki et aI., 1982). In all of the cucumber and melon intoxications, classic cholinergic symptoms (diarrhea, nausea, vomiting, sweating, blurred vision, abdominal pain, dyspnea, muscle fasciculations, headache, and, in some cases, loss of function of arms and legs) were observed, persisting some 4 to 12 hours followed by complete recovery (Risher et al., 1987). The exposures to aldicarb in drinking water were less conclusively related to specific signs and symptoms of intoxication (Sterman and Varma, 1983; Zaki et aI., 1982). In an experimental study with human subjects, aldicarb was administered in single oral doses of 0.025,0.05, 0.1 mg/kg body weight, with consequent manifestations of a variety of cholinergic symptoms at the highest level, all of which disappeared by 6 hours after administration (Union Carbide report quoted by Risher et aI., 1987). While abnormal reductions in erythrocytic AChE activity (25% of pre-exposure activity) were measured at the highest dose, the inhibition was rapidly reversible and preceded the disappearance of the symptoms. Methomyl appears periodically as an insecticidal toxic ant in humans, either from occupational exposure or through accidental or suicidal ingestion (Simpson and Bermingham, 1977). One recorded acute poisoning in Jamaica involved unleavened bread prepared with methomyl mistakenly used in place of common table salt, a level of some 1000 ppm being detected (Liddle et aI., 1979). Consumption of the bread was rapidly fatal to three individuals; another was asymptomatic, while the fifth showed generalized muscle fasciculations and respiratory distress. It was estimated that, in those who died, the amounts ingested were equivalent to 12 to 15 mg/kg body weight. In Japan, a 31-year-old woman committed suicide by incorporating methomyl in food, this being eaten by her three children as well (Araki et aI., 1982). A 9-year-old child survived. Autopsies revealed congestion of the stomach lining and lungs, edema, and hemorrhaging due to acute circulatory failure. The amounts ingested were estimated at 55 mg/kg for the mother and 13 mg/kg for a 6-year-old child.

1098

CHAPTER 52


Concerning the more toxic carbamate ester insecticides, aldicarb and methomyl, it has been demonstrated that central effects of these toxicants are more severe in children than in adults; symptoms in adults were miosis, muscle fasciculations, slowing of the heart, and broncorrhea, whereas in children, stupor/coma, hypotonia, and diarrhea were significant effects. Feldman (1999) suggested that the observed differences might be reflected in differences in the permeability of the bloodbrain barrier of children and adults. 52.6.4 CHRONIC TOXICITY-ANIMAL

Subchronic and chronic studies have been conducted in various animal species, including mice, rats, dogs, swine, and monkeys, although many have never been published since they were confidential documents submitted to regulatory agencies in support of product registrations. Summaries of the general chronic effects of carbamate insecticides on various target sites have been published (Cranmer, 1986; Ecobichon, 1994a; IPCS, 1986). Much of the toxicity is associated with the nervous systems, neuromuscular dysfunction, and neurobehavioral changes. The chronic toxicity of carbaryl has been the most extensively reviewed (Cranmer, 1986). Baron (1991) has published a relatively complete inventory of the effects resulting from the subchronic and chronic treatment of mice, rats, and dogs with several carbamate insecticides; much of this data was gleaned from unpublished reports from chemical manufacturers submitted at the time of product registration. Consistently, depression of erythrocytic and plasma cholinesterases were seen associated with cholinergic effects, particularly at higher dosages. Only with some agents were changes seen in food consumption, growth and development, organ weights, hematological and clinical chemistry measurements, urinalysis, and gross and histopathological parameters. The data from chronic dietary exposure studies are summarized in Table 52.9. 52.6.4.1 Neurophysiological Effects

There is evidence from animal studies that long-term exposure to carbamates such as sodium diethylthiocarbamate (a copperchelating agent) and tetraethylthiuram (a rubber vulcanizer and a therapeutic agent in chronic alcoholism) can elicit neurological effects, possibly due to metabolism to carbon disulfide, a neuropathic agent (Barry, 1953; Gardner-Thorpe and Benjamin, 1971; Moddel et aI., 1978; Waibel et aI., 1957; Watson et al., 1980). Detailed descriptions of anatomic lesions including degeneration and vacuolization in the peripheral and central nervous systems of rabbits and chickens have been published, suggesting that the lesion pattern was similar to the "dying back" process described for some organophosphorus esters (Cavanagh, 1969, 1973; Edington and Howell, 1966, 1969; Howell and Edington, 1968). The elicited neurotoxicity has been attributed to the biotransformation of the dithiocarbamates to yield carbon disulfide, a known neuropathic agent (Brugnone et aI., 1993; Johnson et aI., 1998).

Miller et al. (1969) demonstrated that a single dose of carbaryl (20 mg/kg) given to miniature swine caused a 44% and 75% inhibition of cerebral cortex and brain stem AChE, respectively, and caused a hindlimb paralysis even though no obvious effects were seen upon histopathological examination. Severe carbaryl-induced neuromuscular effects were observed in another study in swine (Smalley et aI., 1969). High dosages (150 mg/kg/day over 73 to 83 days or 150 mg/kg daily for 28 days followed by 300 mg/kg/day over the next 18 or 57 days) caused severe neuromuscular effects. Reluctance to stand was observed first, followed by a peculiar stance, the rear legs being carried well forward under the body, the animals appearing to be "walking" on their dew claws. There was greatly exaggerated flexion of the rear legs, the animals having difficulty in backing up or sitting down. Forcing the animals to move caused marked incoordination, ataxia, muscle tremors, and clonic contractions. Muscular lesions consisted of a myodegeneration of traumatic or ischemic nature, an acute hyaline and vacuolar degeneration, and an acute degenerative process associated with dystrophic calcification (Smalley et aI., 1969). Carbamate esters have caused severe neuropathy in adult chickens following repeated oral administration, although this neuropathy is different from organophosphorus-induced delayed neuropathy (OPIDN) (Fisher and Metcalf, 1983; Hollingshaus and Fukuto, 1982). In young (3-week-old) chicks, both carbaryl and aldicarb affected locomotor activity for 6 weeks after cessation of the subacute (7 days) exposure (FarageElawar, 1989a, 1989b). The treated chicks walked with an abnormal gait, taking shorter steps but with a wider stance suggestive of ataxia. Some paralysis was seen for up to 40 days post-treatment. Once again, this neuropathy was unlike that seen in OPIDN. While the short-term exposure of rats and dogs to repeated oral doses of carbaryl, carbofuran, and propoxur resulted in the inhibition of plasma, erythrocytic, and brain cholinesterases accompanied by typical acute signs of toxicity, there was little evidence of persistent effects on the central and/or peripheral nervous systems (Cranmer, 1986; IPCS, 1986; Krechniak and Foss, 1982). While rats dying in an acute carbaryl study showed congestion of the brain and meninges, such morphological changes have not been reported in animals treated with less than near-lethal concentrations (Boyd and Boulanger, 1968). Cranmer (1986) cited studies in which morphological changes in the brain ganglia in rabbits and increased brain protein in rats were reported, but suggested that such effects occurred only at doses that reduced cholinesterase activities. Rats and dogs receiving oral aldicarb for up to two years showed no adverse effects (Risher et aI., 1987). Two studies have suggested that long-term exposure to carbamate esters can cause neurotoxicity. In a two-year rat study of carbaryl at levels causing no inhibition of blood cholinesterases or clinical signs, the animals showed electroencephalographic (EEG) changes and had a decreased maze performance (Desi et aI., 1974). In monkeys, the EEG patterns were not adversely affected at carbaryl levels of 1.0 mg/kg/day (Santolucito and Morrison, 1971). Tolerance to carbamates has been reported

1099

52.6 Toxicology

Table 52.9 Chronic Toxicity of Carbamate Insecticidesa Max. dosage

Food consump.

+

Chemical

Species

(mglkg)

Aldicarb

Rat

0.5

Rat

0.3

Dog

0.25

Bendiocarb Carbaryl Carbofuran

Carbosulfan

Rat

200ppm

Dog

500ppm

Rat

400ppm

Dog

1250 ppm

Mouse

500ppm

Rat

100ppm

Dog

500ppm

Rat

500 ppm 2500ppm

Methiocarb

Dog

IOOOppm

Mouse

2500ppm

Rat Dog

Methomyl

Mexacarbate

Oxamyl

Mouse

Wt.

Death

Clin.

tology

Urina-

chem.

Pathology

ysis

+ +

+

+ + + + + + + + + +

+

+

+ +

+ +

+ +

+

+ +

+ + + +

+

+

+

+

+ +

+

+

60ppm

+

800ppm 20--26

Dog

1000ppm 300ppm

Rat

250ppm

Dog

325 ppm

Rat

150ppm

Dog Pirimicarb

+ + + + +

600 ppm

Rat Mouse

+

Hema-

Organ Growth

+ + + + + +

+ + + +

+ +

250--750

Dog

4.0ppm

+

+ + +

150ppm

Rat

+

+

+

ppm Propoxur

Rat

2000 ppm

Rat

750 ppm

Dog

2000 ppm

+ +

+ +

+

"+," while no effects are represented by "-." Maximum dosage is in milligram per kilogram of body weight (mglkg) unless presented as parts per million (ppm) in food.

a Data derived from Baron (1991). Positive effects on a parameter are indicated by

(Costa et aI., 1982). For example, male rats receiving carbaryl (200 mg/kg/day) orally for 3 days/week for 90 days showed no overt toxicity (Dikshith et aI., 1976). In the above-mentioned carbaryl swine study, Smalley et al. (1969) described a clinical syndrome of chronic intoxication characterized by progressive myasthenia, incoordination, ataxia, intention tremor, and clonic muscular contractions terminating in paraplegia and prostration. Moderate-to-severe edema was found in myelinated tracts of the cerebellum, brain stem, and upper spinal cord, as well as fragmentation of myelin sheaths, moderate swelling and rupture ofaxons, necrosis of cellular components in spinocerebellar tracts, and vascular congestion, hypertrophic endothelium, and vascular degeneration and hemorrhage in the gray columns. The authors attributed the pathological effects to vascular changes induced by carbaryl (Smalley et aI., 1969).

52.6.4.2 Neurobehavioral Effects Behavioral changes have been associated with carbamate exposure in many animal studies, although such adverse effects were usually detected immediately following acute administration of sufficient chemical to inhibit cholinesterases, suggesting that the effects were the result of cholinergic-mediated stress. Most studies have involved carbaryl (Cranmer, 1986). Single intraperitoneal doses of carbaryl (8.0 mg/kg) reduced shock avoidance by 50% in treated rats for 30 minutes, while a dose of7.3 mg/kg caused a 50% reduction in a positive reward response test (food presentation) (Goldberg et aI., 1965). In a food reward test in cats, inhaled carbaryl (40 mg/m 3 ) caused a deficit only immediately after exposure (Yakim, 1967). Spontaneous locomotor activity was reduced in rats in a 60-minute period after acute oral carbaryl (0.56 and 2.24 mg/kg), whereas the

1100

CHAPTER 52


daily administration of a higher dose for 14 consecutive days had no effect on "wheel-turning" (Singh, 1973). In contrast, acute intoxication of rats with carbaryl (10 mg/kg) resulted in increased locomotor activity in a familiar environment but reduction in exploratory activity in a new environment (Singh, 1973). A decrease in rats' working memory was carbaryl dosedependent shortly after treatment (Heise and Hudson, 1985). In a feeding study of carbaryl and arprocarb (2-isopropoxyphenylN-methylcarbamate) in rats, the amounts given in the diet to achieve body dosages of 10 and 20 mg/kg over 50 days resulted in the animals showing increased difficulty in performing tasks, forgetting already learned skills (De si et aI., 1974). Over 4-month periods, rats inhaling carbaryl (12 to 23 mg/m 3 for 4 hours daily) showed decreased performance in a maze task for food reward, but if the treatment was extended every two weeks for four months, performance was normal (Viter, 1978). A possible association could be suggested between behavioral alterations and cholinesterase inhibition/acute toxicity, the adverse effects being ACh-related. Learning in monkeys appeared to be affected by small, chronic doses of carbaryl. This led to the use of the technique called "chain acquisition" task in which monkeys learned one set of equivalent response sequences each day; they were trained to make four out of the 12 possible response sequences in a certain order, with the correct responses being changed

every day (Anger and Setzer, 1979). Carbaryl was administered orally (up to 50 mg/kg) or intramuscularly (1.0,3.0,5.0, and 10 mg/kg) to trained macaque monkeys. Oral carbaryl caused no consistent effects on performance, whereas the injected carbaryl elicited significant decreases in total session time and increased errors in performance at and above dosages of 3.0 mg/kg.

52.6.4.3 Mutagenicity

As a class, methylcarbamates are not mutagenic, negative results being obtained in an overwhelming majority of in vitro gene mutation assays using microbial systems, cultured mammalian cells, and such in vivo systems as Drosophila and dominant lethal mutation tests (Baron, 1991; IPCS, 1986). Table 52.10 lists the methylcarbamate insecticides that showed mutagenicity in in vitro and in vivo test systems. While weak mutagenicity has been identified as a property of several carbamates, high, almost toxic dosages were used and, frequently, the results either could not be replicated or were derived from nonstandardized protocols that could not be compared with results from acceptable techniques.

Table 52.10 Mutagenic Potential of Carbamate Insecticides Chemical

Test systems

Aldicarb

Carbaryl

Effectsa

References Rashid and Mumma (1986)

S. typhimurium

DNA damage

Mouse bone marrow cells

CA

Debuyst and Van Larebeke (1983)

Cultured human lymphocytes

SCE

Gonzales and Matos (1984)

Saccharomyces cerevisiae

M

Guerzoni et al. (1976)

S. typhimurium

M

Egert and Greim (1976)

Cultured rodent cells

M

Ahmed et al. (1977)


CA

Ishidate and Odashima (1977)


spindle poison

Onfelt (1983)

Human fibroblasts

DNA damage

Ahmed et al. (1977)

D. melanogaster

CA

Hoque (1972)

Rats

mitotic abnormalities

Baron (1991)

S. typhimurium

M

Moriya et al. (1983)


M

Wojciechowski et al. (1982)


CA


Formetanate HCl

Cultured human lympho-

M

Baron (1991)

Methomyl


M


Human lymphocytes

SCE


Cultured mammalian cells

CA

Baron (1991)


M


Mouse dominant lethal

CA

Baron (1991)

Carbofuran

cytes

Mexacarbate Propoxur Pirimicarb Thiodicarb a Abbreviations:

Cultured mammalian cells

CA

Pilinskaya (1981, 1982)

Human lymphocytes

CA

Pilinskaya (1981)


DNA damage

Baron (1991)

M, mutagenic; CA, chromosomal aberrations; SCE, sister-chromatid exchanges.

52.6 Toxicology 52.6.4.4 Reproductive Effects

There is little evidence that methylcarbamate insecticides cause reproductive anomalies in mammals. The common, positive findings have been embryotoxicity and/or fetotoxicity associated with the administration of high dosages and concomitant maternal and possibly nutritional toxicity (IPCS, 1986). There is little evidence of these carbamates causing teratogenicity other than through nutrition-related problems (IPCS, 1986). 52.6.4.5 Carcinogenicity

There has been no evidence of the potential of methylcarbamate insecticides to cause carcinogenicity (IPCS, 1986). However, in studies where positive effects have been found, caution must be observed in interpreting the data because of design inadequacies (limited dosage range, duration of study, numbers of animals, etc.) and the extremely high doses administered. Most of these studies are summarized in the IPCS document (IPCS, 1986); more recent reports have not been seen. 52.6.5 CHRONIC TOXICITY-HUMAN

As has been observed with organophosphorus ester intoxications, persistent effects may result following either acute, single, high-level or repeated, even long-duration, low-level exposure to carbamate esters (Ecobichon, 1994a, b). Involvement of both the peripheral and central nervous systems has given rise to distinctive neurophysiological and/or behavioral anomalies. In 1982, a bizarre case of a 55-year-old male was described following a thorough soaking of skin and clothing by a waterwettable preparation of carbaryl (Ecobichon, 1982). Within 3 to 4 weeks of initial antibiotic treatment for "bacterial meningitis," the patient was reporting headaches, blurred vision, photophobia, peripheral numbness, tingling sensations in the hands and legs, muscle weakness, vertigo, incoordination, lethargy, tiredness, forgetfulness, and loss of recent memory. More alarmingly, behavioral changes persisting for several months manifested themselves as frustrated rage, inability to control temper, severe headaches, and short periods of blackouts. Even at 18 months post-exposure, the patient was unable to drive a car, still experienced photophobia, had persistent short-term memory loss and mild peripheral paresthesia, and suffered from lassitude, lethargy, and muscle weakness. The behavioral aspects persisted, being partially controlled by anticonvulsant and antipsychotic drugs. At the time of this case, no other reports had been published concerning slow developing and/or persistent symptoms arising from exposure to carbamate ester insecticides. Subsequently, a number of such cases have appeared in the literature, all reporting persistent adverse effects (Ecobichon, 1994a; Feldman, 1999). 52.6.5.1 Neurophysiological Effects

Indicative of delayed neurotoxicity following acute exposure to high levels of carbamate insecticides, Dickoff et al. (1987) reported a case involving the ingestion of a liquid preparation of

1101

carbaryl (500 mg/kg), showing a weakness in the arms and legs as well as sensory loss following the acute cholinergic crisis, with electrophysiological alterations consistent with a peripheral axonal neuropathy. Recovery began one week after exposure and progressed for 9 months. However, while recovery appeared to be complete, bilateral severe ankle and toe weakness persisted, accompanied by reduced propioception in the toes and tactile preception below mid-calf. In a second, suicidal case, the ingestion of a metolcarb (m-tolyl-methylcarbamate) formulation (estimated dosage of 1.0-1.2 mg/kg) resulted in neurological damage (Umehara et aI., 1991). Fibrillation potentials and sharp positive waves were observed at rest, with reduced recruitment patterns during muscle contractions. A sural nerve biopsy, performed 38 days after the poisoning, revealed reduced densities oflarge and small myelinated fibers, degenerated axons, and denervated Schwann cell clusters. At 3 months post-exposure, motor symptoms had improved, with a reduction in numbness, although deep tendon reflexes in the extremities were absent. By 6 months, upper motor neuronal signs were no longer observed. In a more recent case of acute exposure to a1dicarb, men handling poisoned sheep showed persistent symptoms similar to those mentioned above, five of the six individuals still experiencing persistent neurotoxicity some three years after exposure (Grendon et aI., 1994). Feldman (1999) described two cases of chronic symptoms following an aerial overspray of carbaryl in a forest spraying program. Both individuals had symptoms persisting for more than six years after the exposure, one (a 35-year-old male) showing hoarseness, bronchorrhea, dizziness, and a peripheral neuropathy. The second individual (a 53-year-old male) suffered recurrent bouts of abdominal cramps and diarrhea, anxiety attacks with associated flashback memories, fatigue, numbness in the feet which gradually affected the hands, a clumsy gait, reduced sensation to mid-calf developing within three months of exposure, and some effects (slowed nerve conduction, reduced muscle action potential amplitude and motor activity in the lower extremities) for some five years postexposure. Electromyographical and nerve conduction studies indicated a peripheral neuropathy with chronic denervation and reinnervation in distal muscles. In contrast to the acute-exposure situation, reports linking incidents of-low-to-moderate, subchronic or chronic exposure to chronic toxicity are almost nonexistent. In one study in which volunteers took daily oral doses of carbaryl (0.06 or 0.13 mg/kg) for six weeks, with concomitant monitoring of plasma and whole blood cholinesterase activities, other blood and urinary biochemical parameters, and electroencephalograms, no deleterious changes were attributable to the agent (Wills et aI., 1968). However, the level of exposure could be considered to be low. One case report of a low-level exposure to carbaryl stands out as being contrary to everything known and expected of carbamate-related toxicity. The propositus patient, a 75-yearold male, and his family (wife and son) were exposed over a period of 8 to 10 months to carbaryl (a 10% dust formulation) applied some six times to the basement area of a home to control fleas (Branch and Jacqz, 1986a). The insecticide

1102

CHAPTER 52


was dispersed throughout the home by the central air conditioning located in the basement. While the entire family experienced a range of acute influenza-like symptoms (headache, malaise, epigastric discomfort, abdominal colic, diarrhea, nausea, muscle spasms, cough) within 3 days of the initial application, the wife and son appeared to recover. However, the father's signs and symptoms became progressively more severe (headache with intense pressure, tinnitus, vertigo, mild ataxia, rhinorrhea, excessive lacrimation, weakness in major skeletal muscle groups, fasciculations, somnolence, agitation, mental confusion). Both the plasma and erythrocytic cholinesterase activities were below normal values and consistent with an anticholinesterase intoxication. During intervals of living away from the home, the patient's symptoms dissipated, but they returned when he reentered the house for any extended period. Leaving the house permanently, the patient's well-being improved, but it was two months before the plasma PChE was within normal range. Symptoms that failed to abate included alterations in sleep patterns (episodic awakening with headache and tinnitus) and mental confusion that persisted for a further two years. A computerized tomographic scan revealed cerebral atrophy which had not been observed on a pre-exposure scan. A persistent, neurological deficit became more severe, being defined as a stocking-and-glove peripheral neuropathy. There has been considerable criticism of this last case report, including that of drawing a valid inference of carbaryl intoxication based on one case in the absence of any occupational exposure literature of similar agent-related toxicity. No indoor measurements of aerial or surface concentrations of carbaryl were ever made, thereby preventing an estimation of exposure. The advanced age of the patient could have played a role in the rate of carbaryl detoxification. The concomitant treatment of the patient with cimetidine, an H2-receptor antagonist used in treating gastric acidity and peptic ulcers, might have had an effect. In animal studies, cimetidine has been shown to inhibit carbaryl biotransformation, thereby increasing the systemic bioavailability of the agent (Branch and Jacqz, 1986b). However, taken in context with other case reports, the claim of a carbaryl-induced intoxication appears valid. Regular occupational exposure to methomyl, particularly in the packaging area of the manufacturing plant, resulted in a high incidence of anticholinesterase symptoms (constricted pupils, nausea, vomiting, blurred vision, increased salivation, muscle weakness, fatigue) and a high number of hospitalizations (Morse et aI., 1979). A decreased vibratory sensation was noted in 19.8% of the workforce, with hospitalized workers having significantly more vibratory sensory loss than nonhospitalized workers. However, no depression of either plasma or erythrocytic cholinesterase activities was noted. It is of interest to note that some acute carbamate intoxications have resulted in the later development of persistent respiratory problems that were exacerbated by exposure to other chemicals such as solvents, household pesticides, hairsprays, and perfumes (Feldman, 1999; Grendon et aI., 1994). The descriptions are reminiscent of multiple chemical sensitivity, an olfactory sensitizing and triggering phenomenon described by

Ashford and Miller (1998). One of the earliest reports of odor aversion as a consequence of pesticide exposure was that of Tabershaw and Cooper (1966), who reported that several of their cases could no longer tolerate smelling or contact with pesticides, reacting with nausea and vomiting after even a whiff of the agents and being forced to give up work involving contact with agrochemicals. The genesis of this idiosyncratic reaction has not yet been ascertained but appears to have both neurological and psychogenic components. 52.6.5.2 Neurobehavioral Effects Sufficient numbers of cases of both acute and chronic intoxications by organophosphorus ester, anticholinesterase insecticides have revealed a recognizable pattern of delayed and/or persistent neurobehavioral anomalies that can be detected and assessed by appropriate neuropsychological evaluation (Ecobichon, 1994b, 1999; Feldman, 1999). Not surprisingly, close examination of acute and chronic carbamate insecticide intoxications reveals a similar pattern of persistent behavioral effects even though neuropsychological testing of affected individuals has not been reported. The long-term follow-up of acutely or chronically carbamate-exposed individuals has been poor but, in several of the incidents discussed in the previous section, observed symptoms have included vertigo, incoordination, disturbed sleep patterns, anxiety attacks, mood changes, chronic lethargy and fatigue, agitation, mental confusion, and difficulty in performing simple tasks or making decisions (Branch and Jacqz, 1986a; Ecobichon, 1982; Grendon et aI., 1994). Feldman (1999) describes the detailed neuropsychological assessment of two individuals some five or six years following an overspraying by carbaryl applied aerially, using the Wechsler Adult Intelligence Scale (WAIS) test battery. As has been observed with organophosphorus ester insecticide intoxications, many of the component tests were within normal values, but impairment, deficiency, or slowing of performance were detected in: a confrontational naming task (Boston Naming Test), memory tests requiring a delayed recall of information; psychomotor (Digit Symbol Test) evaluation, and tasks of visual spatial organization, cognitive tracking, and reasoning. The Profile of Mood States (POMS) and the Minnesota Multiphasic Personality Inventory (MMPI) revealed persistent fatigue, depression and/or anxiety states. Discrepancies were revealed between verbal and performance intelligence quotients. It is obvious that greater use could be made of refined neurological and behavioral test batteries to evaluate short- and long-term neurological effects of carbamate exposure. 52.6.5.3 Reproductive Effects Few specifics are known about the potential of short- and longterm influence of carbamate insecticide exposure on any aspects of reproduction in either male or female agricultural workers. One paper has reported an increase in sperm abnormalities (abnormally shaped heads) in production workers who had been exposed to carbaryl at the time of sampling (Wyrobek et aI., 1981). Neither the sperm count nor the presence of double

1103

52.8 Exposure Limits

fluorescent bodies was changed. Formerly exposed workers, removed from carbaryl-related occupational activities for an average of 6.3 years, showed only a marginally significant elevation in sperm abnormalities, results suggesting that the carbarylinduced morphological effects may not be reversible or may be only slowly so. A dose-dependent change in sperm morphology could not be established. More research needs to be conducted in this area of toxicology.

52.7 TREATMENT The symptoms (Table 52.8) associated with carbamate insecticide intoxication are associated with the accumulating, unmetabolized neurotransmitter ACh at the nerve endings of the parasympathetic and sympathetic autonomic ganglia, the postganglionic parasympathetic nerve endings, and the neuromuscular junctions of the somatic, motor neurons, as a consequence of the inhibition of the nervous tissue AChE. The reversible nature of the AChE inhibition would suggest that the symptoms would be transient although, depending upon the level of exposure, possibly moderate to severe in nature. Atropine is the antidote of choice, antagonizing the action of ACh by blocking the receptor site for the transmitter at parasympathetic nerve fibers innervating exocrine glands, gastrointestinal tract, respiratory tract, eyes, heart, and bladder, as well as exerting a direct, central effect on the respiratory center (Ecobichon et aI., 1977; Feldman, 1999; Namba et al., 1971). Alleviation of these muscarinic signs will be best achieved by administering frequent small doses (0.5 to 1.0 mg) subcutaneously until there is dilatation (mydriasis) of the pupils and the face becomes flushed and/or sweating disappears. The patient should be carefully titered using these signs as physiological endpoints since excess atropine can cause severe toxicity. This is of particular importance in carbamate ester poisoning where the enzyme-insecticide complex is unstable; the enzyme may be decarbamoylated by the excess ACh and the carbamate ester may be metabolized in a short period of time. Atropine is ineffective in counteracting the nicotinic, neuromuscular effects of the accumulated ACh. Acetylcholinesterase reactivators such as the pyridinium oximes, 2-PAM, P2S, and toxogonin, have been used in carbamate-induced intoxications but with mixed results, their use remaining controversial. Signal animal (rats, dogs) experiments involving carbaryl intoxications revealed that the protective effect of atropine was markedly reduced by the concomitant administration of 2-PAM (Carpenter et aI., 1961; Natoff and Reiff, 1973; Sanderson, 1961). This observation was confirmed in one human carbaryl-related poisoning where it was noted that the patient's condition deteriorated rapidly following the administration of2-PAM (Farago,1969). This led to a generalized condemnation of oxime use as an antidote in carbamate insecticide intoxication (Harris et aI., 1989; Lifshitz et aI., 1994; Natoff and Reiff, 1973; Sterri et al., 1979). However, beneficial effects of oxime treatment were seen in treating aldicarb,

Table 52.11 Exposure Limits For Carbamate Insecticidesa Chemical

OS HA (PEL)

ACGIH

NIOSH REL

IDLH

(TLY)

Carbaryl

5.0

5.0

100

5.0

Carbofuran

0.1

ND

0.1

Methomyl

O.lb 2.5 b

2.5

ND

2.5

Propoxur

0.5 b

0.5

ND

0.5

aYalues represent concentrations (mg/m3) in air. REL values represent a timeweighted average for a lO-hr exposure, while TLY values represent a timeweighted average for an 8-hr exposure. bYa1ues "vacated" by OS HA in 1993.

mecarbam (S-(N-ethoxycarbonyl-N-methyl-carbamoylmethyl) O,O-diethyl phosphorodithioate) and methomyl intoxications (Natoff and Reiff, 1973; Sterri et aI., 1979). In cases involving carbofuran, methiocarb, mexacarbate, thiodicarb, and trimethacarb, 2-PAM was ineffective but did not exacerbate the clinical symptoms or interfere with the antidotal effectiveness of atropine (Baron, 1991; FAOIWHO, 1982). Similar results were seen for P2S with pirimicarb and for toxogonin with methomyl (FAOIWHO, 1982; Sanderson, 1961). Overall, the studies to date suggest the efficacious use of oximes with aliphatic oxime carbamate (aldicarb, methomyl, and possibly mecarbam) intoxications but not their use in carbaryl- or other carbamate-related intoxications (Feldman, 1999). The myorelaxant agent, diazepam, should be considered in treatment regimens of all but the mildest cases of carbamate intoxications to relieve anxiety, to counteract some central nervous system-related symptoms not affected by atropine.

52.8 EXPOSURE LIMITS Given the importance of methyl carbamate insecticides in the agricultural industry, it is surprising that so few exposure limit values have been established for these agents by the Occupational Safety and Health Administration (OSHA), the National Institute for Occupational Safety and Health (NIOSH), and the American Conference of Governmental Industrial Hygienists (ACGIH). As is shown in Table 52.11, the promulgated values for a few carbamates include permissible exposure levels (PELs), recommended exposure limits (RELs), immediately dangerous to life and health (IDLH) levels, and threshold limit values (TLVs) or time-weighted average (TWA) exposure levels for 8 to 10 hours. In addition, the U.S. Environmental Protection Agency (EPA) has introduced maximum drinking water contamination levels (MCLs) only for aldicarb (0.007 mg/L) and carbofuran (0.04 mg/L), chemicals that have been associated with problems of surface and groundwater contamination.

1104

CHAPTER 52


REFERENCES Abdel-Wahab, A. M., Kuhr, R. J., and Casida, J. E. (1966). Fate of 14C_ carbonyl-Iabeled aryl methylcarbamate insecticide chemicals in and on bean plants. 1. Agric. Food Chem. 14, 290-298. Ahmed, E E., Lewis, N. J., and Hart, R W (1977). Pesticide-induced ouabain resistant mutants in Chinese hamster V79 cells. Chem. Biol. Interact. 19, 369-374. Aly, O. M., and E1-Dib, M. A. (1971). Studies on the persistence of some carbamate insecticides in the aquatic environment. 1. Hydrolysis of SevinTM, Baygon™, PyroIan™ and DimetiIan™ in waters. Water Res. 5,1191-1200. Anger, W. K., and Setzer, J. V. (1979). Effect of oral and intramuscular carbaryl administration on repeated acquisition in monkeys. 1. Toxicol. Environ. Health 5, 793-808. Anger, W. K, and Wilson, S. M. (1980). Effect of carbaryl on variable interval response rates in rats. Neurobehav. Toxicol. 2, 21-24. Araki, M., Yonemitsu, K, Kambe, T., Idaka, D., Tsunenari, S., Kanda, M., and Kambara, T. (1982). Forensic toxicological investigations on fatal cases of carbamate pesticide methomyl (Lannate®) poisoning. Ipn. 1. Legal Med. 36,584-588. Ashford, N., and Miller, C. (1998). "Chemical Exposures. Low Levels and High Stakes," 2nd ed. Van Nostrand Reinhold, New York. Baron, R. L. (1991). Carbamate insecticides. In "Handbook of Pesticide Toxicology" (W J. Hayes, Jr. and E. R Laws, Jr., eds.), Vol. 3, Ch. 17, pp. 11251189. Academic Press, New York. Barry, W K (1953). Peripheral neuritis following tetraethylthiuram-disulfide treatment. Brit. Med. 1. 2, 937. Berry, W. K (1971). Acceleration by free carbamate of the spontaneous reactivation of carbamylated acetylcholinesterase. Biochem. Pharmacol. 20, 3236-3238. Boyd, E. M., and Boulanger, M. A. (1968). Insecticide toxicology. Augmented susceptibility to carbaryl toxicity in albino rats fed purified casein diets. 1. Agric. Food Chem. 16, 834-838. Branch, R. A., and Jacqz, E. (l986a). Subacute neurotoxicity following longterm exposure to carbaryl. Am. J. Med. 80, 741-745. Branch, R A., and Jacqz, E. (1986b). Is carbaryl as safe as its reputation? Does it have a potential for causing chronic neurotoxicity in humans? Am. J. Med. 80, 659-664. Brugnone, E, Maranelli, G., Guglielmi, G., Ayyad, K., Soleo, L., and Elia, G. (1993). Blood concentrations of carbon disulfide in dithiocarbamate exposure and in the general population. Int. Arch. Occup. Environ. Health 64, 503-507. Carpenter, C. P., Weil, C. S., Palm, P. E., Woodside, M. W., Nair, rn, J. H., and Smyth, Jr., H. E (1961). Mammalian toxicity of I-naphthyl-Nmethylcarbamate (Sevin insecticide). J. Agric. Food Chem. 9, 30-38. Cavanagh, J. B. (1969). Toxic substances and the nervous system. Brit. Med. Bull. 25, 268-273. Cavanagh, J. B. (1973). Peripheral neuropathy caused by chemical agents. CRC Crit. Rev. Toxicol. 2, 365-417. Centers for Disease Control (CDC) (1979). Suspected carbamate intoxications-Nebraska. Morbidity and Mortality Weekly Report 28, 133-134. Centers for Disease Control (CDC) (1986). Aldicarb food poisoning from contaminated melons-California. Morbidity and Mortality Weekly Report 35, 254-255. Costa, L. G., Schwab, B. W, and Murphy, S. D. (1982). Tolerance to anticholinesterase compounds in mammals. Toxicology 25, 79-97. Cranmer, M. E (1986). Carbaryl: A toxicological review and risk analysis. Neurotoxicology 7, 247-328. Cremlyn, R (1978). "Pesticides. Preparation and Mode of Action." John WiIey and Sons, New York. Debuyst, B., and Van Larebeke, N. (1983). Induction of sister-chromatid exchanges in human Iymphocytes by aldicarb, thiofonax and methomyl. Mutat. Res. 113, 242-243. Desi, 1., Gonczi, L., Simon, G., Farkas, 1., and Kneffel, Z. (1974). Neurotoxicologic studies of two carbamate pesticides in subacute animal experiments. Toxicol. Appl. Pharmacol. 27, 465-476.

Dickoff, D. J., Gerber, 0., and Turovsky, Z. (1987). Delayed neurotoxicity after ingestion of carbamate pesticide. Neurology 37, 1229-1231. Dikshith, T. S. S., Gupta, P. K, Gaur, J. S., Datta, K K, and Mathur, A. K (1976). Ninety day toxicity of carbaryl in male rats. Environ. Res. 12, 161170. Ecobichon, D. J. (1982). Carbamic Acid Ester Pesticides. In "Pesticides and Neurological Diseases" (D. J. Ecobichon and R M. Joy, eds.), Ch. 6, pp. 220-221. CRC Press, Boca Raton, FL. Ecobichon, D. J. (l994a). Carbamic acid ester insecticides. In "Pesticides and Neurological Diseases" (D. J. Ecobichon and R. M. Joy, eds.), 2nd ed., Ch. 5, pp. 258-262. CRC Press, Boca Raton, FL. Ecobichon, D. J. (l994b). Organophosphorus ester insecticides. In "Pesticides and Neurological Diseases" (D. J. Ecobichon and R M. Joy, eds.), 2nd ed., Ch. 4, pp. 211-220. CRC Press, Boca Raton, FL. Ecobichon, D. J. (1996). Toxic effects of pesticides. In "Casarett and Doull's Toxicology. The Basic Science of Poisons" (c. D. Klaassen, ed.), 5th ed., Ch. 22, pp. 655-662. McGraw-HiII, New York. Ecobichon, D. J. (1999). Biological monitoring: Neurophysiological and behavavioral assessments. In "Occupational Hazards of Pesticide Exposures: Sampling, Monitoring, Measuring" (D. J. Ecobichon, ed.), Ch. 8, pp. 209230. Taylor and Francis, Philadelphia. Ecobichon, D. J., and Comeau, A. M. (1973). Pseudocholinesterases of mammalian plasma: Physicochemical properties and organophosphate inhibition in eleven species. Toxicol. Appl. Pharmacol. 24, 92-100. Ecobichon, D. J., and Joy, R. M. (1982). Carbamic acid ester pesticides. In "Pesticides and Neurological Diseases." CRC Press, Boca Raton, FL. Ecobichon, D. J., Ozere, R L., Reid, E., and Crocker, J. E S. (1977). Acute fenitrothion poisoning. Can. Med. Assoc. J. 116, 377-379. Edington, N., and Howell, J. M. (1966). Changes in the nervous system of rabbits following the administration of sodium-diethylthiocarbamate. Nature 210, 1060-1062. Edington, N., and Howell, J. M. (1969). The neurotoxicity of sodium diethylthiocarbamate in the hen. Acta Neuropathol. 12,339-347. Egert, G., and Greim, H. (1976). Formation of mutagenic nitroso-compounds from ephedrine and the pesticides carbaryl, dodin and prometryn in the presence of nitrite at pH 1. Naunyn-Schmiedberg 's Arch. Pharmacol. 293, Supp. R66. Ellman, G. L., Courtney, K. D., Andres, Jr., v., and Featherstone, R. M. (1961). A new and rapid colorimetric determination of acetylcholinesterase activity. Biochem. Pharmacol. 7, 88-95. Farage-Elawar, M. (1989a). Enzyme and behavioral changes in young chicks as a result of carbaryl treatment. J Toxicol. Environ. Health 26, 119-131. Farage-Elawar, M. (I 989b). Toxicity of aldicarb in young chicks. Neurotox. Teratol. 10, 549-554. Farago, A. (1969). Suicidal, fatal Sevin® (1-naphthyl-N-methyl carbamate) poisoning. Arch. Toxicol. 24, 309-315. Feldman, R G. (1999). Carbamates. In "Occupational and Environmental Neurotoxicology" (R. G. Feldman, ed.), Ch. 23, pp. 442-465. Lippincott-Raven Publishers, Philadelphia. Fiore, M. c., Anderson, H. A., Hong, Z., Golubjatnikov, R, Seiser, J. E., Nordstrom, D., Hanrahan, L., and Belluck, D. (1986). Chronic exposure to aldicarb-contaminated groundwater and human immune function. Environ. Res. 41, 633-645. Fisher, S. W, and Metcalf, R L. (1983). Production of delayed ataxia by carbamic acid esters. Pestic. Biochem. Physiol. 19, 243-253. Food and Agricultural Organization/World Health Organization (FAOIWHO) (1982). "Pesticide Residues in Food: 1981 Evaluations: The Monographs." FAO Plant Product Protection Paper No. 42. Food Agric. Organ. V.N., Rome. Fukuto, T. R (1972). Metabolism of carbamate insecticides. Drug Metab. Rev. 1, 117-150. Fukuto, T. R (1983). Structure-activity relationships in derivative of anticholinesterase insecticides. In "Pesticide Chemistry: Human Welfare and the Environment" (J. Miyamoto and P. C. Keamey, eds.), Vol. I, pp. 203213. Pergamon Press, New York.

References

Fukuto, T. R., Fahmy, M. A. H., and Metcalf, R. L. (1967). Alkaline hydrolysis, anticholinesterase and insecticidal properties of some nitro-substituted phenyl carbamates. 1. Agric. Faad Chem. 15,273-277. Gardner-Thorpe, and Benjamin, S. (1971). Peripheral neuropathy after disulfiram adminstration. 1. Neural. Neurosurg. Psychiat. 34, 253-259. Goes, A. E., Savage, E. P., Gibbons, G., Arronson, M., Ford, S. A., and Wheeler, H. W. (1980). Suspected foodborne carbamate pesticide intoxications with the ingestion of hydroponic cucumbers. Am. 1. Epidem. 111, 254-259. Goldberg, M. E., Johnson, H. E., and Knaak, J. B. (1965). Inhibition of discrete avoidance behavior by three anticholinesterase agents. Psychapharmacalagia 7, 72-76. Goldman, L. R., Beller, M., and Jackson, R. J. (1990a). Aldicarb food poisonings in California. 1985-1988: Toxicity estimates for humans. Arch. Enviran. Health 45,141-147. Goldman, L. R., Smith, D. E, Neutra, R. R., Saunders, L. D., Pond, E. M., Stratton, J., Walker, K, Jackson, R. J., and Kizer, K. W. (l990b). Pesticide food poisoning from contaminated watermelons in California. Arch. Enviran. Health 45, 229-236. Gonzales, C. M., and Matos, E. (1984). Induction of sister-chromatid exchanges in cultured human lymphocytes by aldicarb, a carbamate pesticide. Mutat. Res. 138, 175-179. Grendon, J., Frost, E, and Baum, L. (1994). Chronic health effects among sheep and humans surviving an aldicarb poisoning incident. Vet. Human Taxical. 36,218-223. Guerzoni, M. E., DelCupolo, L., and Ponti, 1. (1976). Mutagenic activity of pesticides. Riv. Sci. Tecnal. Alimenti. Nutr. 6, 161-165. Harris, L. w., Talbot, B. G., Lennox, W. J., and Anderson, D. R. (1989). The relationship between oxime-induced reactivation of carbamylated acetylcholinesterase and antidotal efficacy against carbamate intoxication. Taxicol. Appl. Pharmacal. 98, 128-133. Hastings, E L., Main, A. R., and Iverson, E (1970). Carbamylation and affinity constants of some carbamate inhibitors of acetylcholinesterase and their relation to analogous substrate constants. 1. Agric. Faad Chem. 18, 497-502. Hayes, Jr., W. J. (1982). Carbamate pesticides. In "Pesticides Studied in Man" (w. J. Hayes, Jr., ed.), Ch. 8, pp. 436-462. Williams and Wilkins, Baltimore. Heath, D. E (1961). Abnormal effects. In "Organophosphorus Poisons. Anticholinesterases and Related Compounds," Ch. XVI, pp. 338-339. Pergamon Press, New York. Heise, G. A., and Hudson, J. D. (1985). Effects of pesticides and drugs on working memory in rats: Continuous delayed response. Pharmacal. Biachem.

c.,

Behav.23,591-598. HoIIingshaus, J. G., and Fukuto, T. R. (1982). The effect of exposure to pesticides on delayed neurotoxicity. In "Effects of Chronic Exposures to Pesticides on Animal Systems" (J. Chambers and J. D. Yarborough, eds.), pp. 85-120. Raven Press, New York. Hoque, M. Z. (1972). Carbaryl, a new chemical mutagen. Curr. Sci. 41, 855856. Howell, J. M., and Edington, N. (1968). The neurotoxicity of sodium diethyldithiocarbamate in the hen. 1. Neuropathal. Exp. Neural. 27,464-472. International Program on Chemical Safety (IPCS) (1986). "Carbamate Pesticides: A General Introduction." Environmental Health Criteria 64. World Health Organization, Geneva. Ishidate, Jr., M., and Odashima, S. (1977). Chromosome tests with 134 compounds on Chinese hamster cells in vitro. A screening for chemical carcinogens. Mutat. Res. 48, 337-354. Iverson, E (1975). Affinity and carbamylation rate constants of propoxur in reaction with erythrocyte and serum cholinesterase. Biachem. Pharmacal. 24, 1537-1538. Johnson, D. J., Graham, D. G., Amamath, v., Amamath, K, and Valentine, W. M. (1998). Release of carbon disulfide is a contributing mechanism in the axonopathy produced by N,N-diethyldithiocarbamate. Taxical. Appl. Pharmacal. 148, 288-296. Knaak, J. B. (1971). Biological and nonbiological modifications of carbamates. Bull. W.H.O. 44,121-131. Krechniak, J., and Foss, W. (1982). Cholinesterase activity in rats treated with propoxur. Bull. Enviran. Cantam. Taxical. 29, 599-604.

1105

Kuhr, R. J., and Dorough, H. W. (1976). "Carbamate Insecticides: Chemistry, Biochemistry and Toxicology." CRC Press, Boca Raton, PL. Kulkami, A. P., and Hodgson, E. (1980). Metabolism of insecticides by mixed function oxidase systems. Pharmacal. Ther. 8, 379-475. Kurtz, P. J. (1977). Behavioral and biochemical effects of the carbamate insecticide, Mobam. Pharmacal. Biachem. Behav. 6, 303-310. Liddle, J. A., Kimbrough, R. D., Needham, L. L., Cline, R. E., Smrek, A. L., Yert, L. w., and Bayse, D. D. (1979). A fatal episode of accidental methomyl poisoning. Clin. Taxical. 15, 159-167. Lifshitz, M., Rotenberg, M., Sofer, S., Tamiri, T., Shahak, E., and Almog, S. (1994). Carbamate poisoning and oxime treatment in early children: A clinical and laboratory study. Pediatrics 93, 652-655. Lifshitz, M., Shahak, E., Bolotin, A., and Sofer, S. (1997). Carbamate poisoning in early childhood and in adults. Clin. Taxicol. 35, 25-27. Lima, J. S., Alberto, C., and Reis, G. (1995). Poisoning due to illegal use of carbamates as a rodenticide in Rio de Janeiro. Clin. Taxicol. 33, 687-690. Melnikov, N. N. (1971). Chemistry of pesticides. Residue Rev. 36,1-480. Merck Index, 12th ed. (1996). Merck Research Laboratories, Division of Merck and Co., Inc., Whitehouse Station, NJ. Miller, E., Reinwall, J., Brouwer, J., Ear, E L., and Loon, E. J. (1969). Effects of acute administration of carbaryl on cholinesterase levels in the CNS of swine. Taxical. Appl. Pharmacal. 14, 622-623. Moddel, G., Bilbao, J. M., Payne, D., and Ashby, P. (1978). Disulfiram neuropathy. Arch. Neurol. 35, 658-660. Moriya, M., Ohta, T., Watanabe, K., Miyazawa, T., Kato, K., and Shirasu, Y. (1983). Further mutagenicity studies on pesticides in bacterial reversion assay systems. Mutat. Res. 116, 185-216. Morse, D. L., Baker, E. L., Kimbrough, R. D., and Wisseman, C. L. (1979). Propanil-chloracne and methomyl toxicity in workers of a pesticide manufacturing plant. Clin. Taxical. 15, 13-21. Namba, T., Nolte, C. T., Jackrel, J., and Grob, D. (1971). Poisoning due to organophosphate insecticides. Am. 1. Med. 50, 475-492. Natoff, I. L., and Reiff, B. (1973). Effect of oximes on the acute toxicity of anticholinesterase carbamates. Taxical. Appl. Pharmacal. 25, 569-573. O'Malley, M. (1997). Clinical evaluation of pesticide exposure and poisonings. Lancet 349, 1161-1166. Onfelt, A. (1983). Spindle disturbances in mammalian cells. I. Changes in the quantity of free sulfhydryl groups in relation to survival and C-mitosis in V79 Chinese hamster cells after treatment with coIcemid, diamide, carbaryl and methyl mercury. Chem. Bial. Interact. 46, 201-217. Pilinskaya, M. A. (1981). Study of the cytogenetic effect of a number of pesticides in human peripheral blood Iymphocyte culture at various initial levels of chromosomal aberrations. Cytal. Genet. 15,74-76. Pilinskaya, M. A. (1982). The cytogenetic effect of pesticide pirimor in a human peripheral blood lymphocyte culture in viva and in vitro. Cytal. Genet. 16, 45-49. Rashid, K A., and Mumma, R. O. (1986). Screening pesticides for their ability to damage bacterial DNA. 1. Enviran. Sci. Health Part B 21, 319-334. Reiner, E. (1971). Spontaneous reactivation of the phosphorylated and carbamylated cholinesterases. Bull. W.H.O. 44, 109-112. Remaley, A. T., Hicks, D. G., Kane, M. D., and Shaw, L. M. (1988). Laboratory assessment of poisoning with a carbamate insecticide. Clin. Chem. 34, 1933-1936. Richardson, E. M., and Batteese, R. 1. (1973). An incident ofZectran poisoning. 1. Maine Med. Assac. 64, 158-159. Risher, J. E, Mink, E L., and Stara, J. E (1987). The toxicologic effects of the carbamate insecticide aldicarb in mammals: A review. Environ. Health Res. 72,267-281. Ruppert, P. H., Cook, L. L., Dean, K. E, and Reiter, L. W. (1983). Acute behavioral toxicity of carbaryl and propoxur in adult rats. Pharmacal. Biachem. Behav. 18, 569-584. Ryan, A. J. (1971). The metabolism of pesticidal carbamates. CRC Crit. Rev. Taxicol. 1,33-54. Sanderson, D. M. (1961). Treatment of poisoning by anticholinesterase insecticides in the rat. 1. Pharm. Pharmacal. 13, 435-442.

1106

CHAPTER 52


Santolucito, J. A., and Morrison, G. (1971). EEG of Rhesus monkeys following prolonged low-level feeding of pesticides. Toxicol. Appl. Pharmacol. 19, 147-154. Schlagbauer, B. G. L., and Schlagbauer, A. W. J. (1972). The metabolism of carbamate pesticides-A literature analysis. Part I and Part n. Residue Rev. 42,1-84;42,85-90. Simpson, G. R., and Bermingham, S. (1977). Poisoning by carbamate pesticides. Med. 1. Austr. 2, 148-149. Singh, J. M. (1973). Decreased performance behavior with carbaryl-An indication of clinical toxicity. Clin. Toxicol. 6, 97-108. Smalley, H. E., O'Hara, P. J., Bridges, C. H., and Radeleff, RD. (1969). The effect of chronic carbaryl administration on the neuromuscular system of swine. Toxicol. Appl. Pharmacol. 14,409-419. Sterman, A. B., and Varma, A. (1983). Evaluating human neurotoxicity of the pesticide aldicarb: When man becomes the experimental animal. Neurobehav. Toxicol. Terato!' 5,493-495. Sterri, S. H., Rognerud, B., Fiskum, S. E., and Lyngaas, S. (1979). Effect of toxogonin and P2S in the toxicity of carbamates and organophosphorus compounds. Acta Pharmacol. Toxicol. 45, 9-15. Tabershaw, I. R., and Cooper, W. C. (1966). Sequelae of acute organic phosphate poisoning. 1. Occup. Med. 8, 5-20. Takahashi, RN., Poli, A., Morato, G. S., Lima, T. C. M., and Zanin, M. (1991). Effect of age on behavioral and physiological responses to carbaryl in rats. Neurotox. Terato!' 13,21-26. Tobin, J. S. (1970). Carbofuran a new carbamate insecticide. 1. Occup. Med. 12, 16-19. Umehara, E, Izumo, S., Arimura, K, and Osame, M. (1991). Polyneuropathy induced by m-tolyl methyl carbamate intoxication. 1. Neurol. 238, 47-48. Vandekar, M. (1965). Observations on the toxicity of carbaryl, folithion and 3-isopropoxyphenyl-N-methyIcarbamate in a village-scale trial in Southern Nigeria. Bull. WH.O. 33, 107-115. Vandekar, M., Heyadat, S., Plestina, R, and Ahmady, G. (1968). A study of the safety of O-isopropoxyphenyl-methyIcarbamate in an operational field trial in Iran. Bull. WH.O. 38, 609-623.

Vandekar, M., Plestina, R, and Wilhelm, K (1971). Toxicity of carbamates for mammals. Bull. WH.O. 44, 241-249. Vassilieff, I., and Ecobichon, D. J. (1983). Acute toxicity of aminocarb in male rats and inhibition of their esterases. Bull. Environ. Contam. Toxicol. 31, 326-330. Viter, V. E (1978). Continuous and intermittent effect of carbaryl on certain behavior reactions of experimental animals. Gig. Sanit. 43, 33-34. Waibel, P. E., Johnson, E. L., Pomeroy, B. S., and Howard, L. B. (1957). Toxicity of tetraethylthiuram disulfide in chicks, poults and goslings. Poultry Sci. 36,697-701. Watson, C. P., Ashby, P., and Bilbao, J. M. (1980). Disulfiram neuropathy. Can. Med. Assoc. 1.123,123-126. Wilhelm, K, and Reiner, E. (1973). Effect of sample storage on human blood cholinesterase activity after inhibition by carbamates. Bull. WH.O. 48, 235238. Wilkinson, C. E, ed. (1976). "Insecticide Biochemistry and Physiology." Plenum Press, New York. Wills, J. H., Jameson, E., and Coulston, E (1968). Effects of oral doses of carbaryl on man. Clin. Toxicol. 1,265-271. Wojciechowski, J. P., Kaur, P., and Sabharwal, P. S. (1982). Induction of ouabain resistance in V79 cells by four carbamate pesticides. Environ. Res. 29,48-53. Wyrobek, A. J., Watchmaker, G., Gordon, L., Wong, K, Moore, n, D., and Whorton, D. (1981). Sperm shape abnormalities in carbaryl-exposed employees. Environ. Health Perspec. 40, 255-265. Yakim, V. S. (1967). The maximum permissible concentration of Sevin in the air of the work zone. Hyg. Sanit. 32, 32-37. Zaki, M. H., Moran, D., and Harris, D. (1982). Pesticides in ground-water: The aldicarb story in Suffolk County, New York. Am. 1. Public Health 72,13911395.

CHAPTER

53 Aldicarb: Current Science-Based Approaches in Risk Assessment Abraham J. Tobia, Pierre-Gerard Pontal, Peter McCahon and Neil G. Carmichael Aventis CropScience

J oseph P. Rieth lSC, Inc.

Rick Williams RTI, Inc.

53.1 INTRODUCTION Toxicology-based human health risk assessment is evolving continuously. As new data are developed allowing greater understanding of chemical effects, dose relationships, and modes of action, improvements in the reliability of the subsequent risk assessments follow. This is particularly true in the case of pesticides, because this class of chemicals has undergone a virtually continuous process of registration and reregistration since the advent of the Environmental Protection Agency in the United States (U.S. EPA) as well as increased regulatory activity in Europe and Japan since the early 1970s. As a result, each currently registered pesticide has a very robust toxicological database from which one can assess potential health risks, and the wealth of information continues to grow as new study types are developed and conducted. With the advent of the Food Quality Protection Act of 1996 (FQPA) in the United States further emphasis has been placed on the quality of the risk assessment. Whereas previous risk assessment practices did indeed ensure the protection of infants and children, with FQPA emphasis is now placed on increasing the certainty of the assessment. A variety of new science-based techniques have been and are being developed to increase the certainty of the risk assessment process. This chapter describes new approaches, which reduce the uncertainty in the risk assessment of aldicarb. It begins with a description of aldicarb chemistry, uses, and biological mode of action. This is followed by a brief overview of risk assessment practices in the crop protection industry. The next Handbook of Pesticide Toxicology Volume 2. Agents

section describes the available toxicology and exposure studies used to evaluate the potential risk of the product. Finally, two elements of aldicarb risk assessment are presented. The first demonstrates the use of the pharmacokinetics of reversibility of cholinesterase inhibition following aldicarb exposure to adjust the exposure component of the risk assessment. The second describes special dermal toxicity studies used to evaluate the potential toxicity to occupational users of the aldicarb-formulated product Temik

53.2 ALDICARB: DESCRIPTION, USE AND BIOLOGICAL MODE OF ACTION Technical aldicarb belongs to the N -methyl carbamate chemical family. The pure (technical) material is a white crystalline solid with a water solubility of approximately 6000 ppm at 25°C and is stable at room temperature. The empirical formula of aldicarb is C7H14N202S, with a molecular weight of 190.3. The structural formula is shown in Fig. 53.1, and the (IUPAC) and Chemical Abstract Service (CAS) names are as follows: IUPAC-2-methyl-2-(methylthio)propionaldehyde O-(methylcarbamoyl) oxime CAS-2-methyl-2-(methylthio)propanalO-[(methylamino) carbonyl] CAS number-116-06-30 Aldicarb is a carbamate insecticide used in agriculture for the control of insects, mites, and nematodes. The product is mar-

1107

Copyright © 200 1 by Academic Press. All rights of reproduction in any fonn reserved.

1108

CHAPTER 53

Figure 53.1

Aldicarb Risk Assessment

Structural formula of aldicarb.

keted under the trade name Temik as a 15% (active ingredient) granule in the United States, and as a 5, 10, or 15% granule worldwide. Aldicarb has a number of unique characteristics which make it an invaluable tool for crop protection. First, aldicarb controls pests from three divergent animal groups: insects, mites, and nematodes. This range of activity for a single product is rare for this class of crop protection chemicals, and because of it, a single application of aldicarb can replace two or more applications of alternative pesticides. Second, aldicarb has systemic activity whereby the product can be applied beneath the soil, and uptake through the roots allows distribution throughout the plant, with subsequent control of chewing and sucking pests. Aldicarb is formulated solely as a dust-free granule and is not produced as a liquid formulation. This type of formulation significantly reduces potential dermal and inhalation exposures, which makes the product much safer from an occupational perspective. Granular Temik is also much safer for the environment than liquid-formulated insecticides. It affords longer control, reducing the number of applications. Also, it is applied beneath the surface of the soil to a depth of up to several inches. Both of these factors significantly reduce the negative impact on beneficial insects, fish, birds, and other wildlife because the product is not available for exposure. Aldicarb was discovered by Union Carbide Corporation in 1962. The first U.S. registration of Temik was received in 1970 for use on cotton. The four major crops on which Temik is currently used are citrus, cotton, peanuts, and potatoes, and it is also registered for use on nine other crops in the United States. Temik is typically applied via tractor-mounted equipment used to place the granules at a depth of 2-6 inches beneath the surface of the soil. Developments in positive displacement metering devices allow the application of precise amounts of the material into the application area. It is often applied in conjunction with other cultural practices such as planting, cultivating, or fertilizing. Most often, application is once per use season. Technical aldicarb is produced as an integrated 35% solution and then formulated into a granular product. Two types of granules are produced, one made of corncob grit and the other with gypsum clay. Both granules have binding agents and the production method produces a virtually dust-free product. Carbamate insecticides are reversible cholinesterase inhibitors for which recovery is primarily a function of the rate at which the active chemical is hydrolytic ally decarbamylated by the cholinesterase enzyme (Alvarez, 1992; O'Brien, 1967; Rotenberg and Almog, 1995). This process, commonly called spontaneous reactivation, is often measured in minutes. This

is in contrast to the organophosphates, which are generally irreversible inhibitors of cholinesterase; the inhibition is due to significantly stronger binding between the chemical and enzyme by the process of phosphorylation. Recovery following exposure to organophoshates is primarily through prolonged reactivation of the inactivated enzyme and synthesis of new enzyme, and is typically measured in days or weeks. It is primarily because of the difference in recovery times that carbamates are considered to pose significantly less risk to exposed humans, relative to organophosphate repeated exposure.

53.3 CURRENT PRACTICES IN PESTICIDE RISK ASSESSMENT The general process of human health risk assessment based on toxicological data is similar for all chemicals, that is, for pharmaceuticals, crop protection chemicals and industrial and naturally occurring chemicals. At a minimum, two types of information are required for a reliable risk assessment, knowledge of the "hazard" the chemical may cause, and data on possible levels of exposure to humans (Cohrssen and Covello, 1989). Hazard refers to the potential toxic effect or effects the material may cause; this information is typically gleaned either from descriptive toxicity tests in animals or from controlled studies in humans such as clinical trials or epidemiological studies (or a combination of both). Exposure data may be known exactly, as in the case of the prescribed dose of a pharmaceutical, or can be measured on those individuals coming into contact with the chemical, using a variety of techniques (U.S. EPA, 1987). In essence, human health risk assessment involves the determination of a level of exposure to a chemical which is expected to be safe for humans. This process is based on the well-established principles that for every toxic effect there must be a dose sufficiently large to cause that effect. Comparison of hazard and exposure data will not provide a safe exposure level, but can indicate if an exposure level is safe. Hazard information for registered pesticides is based on a very extensive toxicological database. Table 53.1 provides a listing of the types of toxicological studies conducted on aldicarb, which have been cited, reviewed and accepted by multiple international regulatory authorities [Baron, 1994; California EPA, 1998; Food and Agriculture Organization-World Health Organization (FAOIWHO), 1992; International Agency for Research on Cancer (IARC), 1991; International Programme on Chemical Safety (IPCS), 1991]. Studies are performed to examine the full spectrum of possible toxicological effects. These include acute, short-term, and long-term studies; tests for carcinogenicity and mutagenicity; reproductive and developmental effects; and neurotoxicity tests. The studies are conducted by various routes of exposure, for example, oral, dermal, and inhalation, and the studies are typically designed to approximate the route and duration of a potential human exposure.

53.3 Current Practices in Risk Assessment Table 53.1 Toxicological Studies Conducted on Aldicarb Acute Oral LDSO Dermal LDsO Inhalation LCSO Potential for eye irritation

1109

Table 53.2 Human Health Risk Assessments for Aldicarb Acute dietary exposure Chronic dietary exposure Short-term dermal occupational exposure Intermediate-term dermal occupational exposure Long-term dermal occupational exposure

Potential for skin irritation Potential for sensitization Subchronic 7-day dietary study in rats

Short-term inhalation occupational exposure Intermediate-term inhalation occupational exposure Long-term inhalation occupational exposure

90-day dietary study in rats 7-day dietary study in mice 14-day dietary study in dogs 90-day dietary study in dogs Chronic 2-year dietary study in rats (two studies) 2-year dietary study in rats with a mixture of aldicarb, aldicarb sulfoxide, and aldicarb sulfone Neurotoxicity Acute neurotoxicity study 90-day neurotoxicity study Developmental neurotoxicity study Human volunteer studies Controlled clinical studies (two studies) Worker exposure monitoring Studies on product formulations 21-day dermal study Oral LDso Dermal LDso Inhalation LCSO Potential for eye irritation Potential for skin irritation

The risk assessment is performed by first determining the human exposure scenario of interest and then selecting the toxicological study of most relevance to the human situation as well as the relevant end point in this study (usually the most sensitive). Thus, human health risk assessment involves a number of different risk assessments, which reflect each particular human exposure scenario. A listing of the types of risk assessments required for aldicarb can be found in Table 53.2. The potential routes of exposure for aldicarb include oral exposure, from low-level residues in food, and dermal and inhalation exposure of workers handling the product. Once the end point and study of interest for the appropriate risk assessment have been determined, the next step in the process is the determination of the no-observed-adverse-effect level (NOAEL). The NOAEL for the study is the dose level at which no hazard has been detected. Because these studies are generally conducted in animals, the NOAEL for a given study represents a safe dose for the species tested. The prediction of a safe dose for humans is derived by dividing the NOAEL from the animal study by an appropriate safety factor (also called

uncertainty factor); thus the safe human dose is lower than the NOAEL from the animal study. The safe dose for human intake is described in a number of ways, for example, a dose "without appreciable health risk" (WHO, 1987) and reference dose (RID; V.S. EPA, 1993), which was previously known as allowable or acceptable daily intake (ADI). The safety factor historically considered appropriate to provide a safe dose for human exposure for pesticides is 100 when it is derived from an appropriate animal study; typically the RID is set as the NOAEL from an appropriate animal study divided by 100. For excellent reviews of the origin and justification for the use of a lOO-fold safety factor in human health risk assessment see Swartout et al. (1998), Dourson et al. (1992), and Renwick and Lazarus (1998). The lOO-fold safety factor takes into consideration two sources of uncertainty in the risk assessment: (i) the toxicology study was conducted in animals but the objective is the protection of humans (interspecies extrapolation); (ii) there is variability in the human population and sensitive individuals need to be protected (intraspecies extrapolation). A lO-fold margin of safety is generally considered to provide adequate protection for each of these sources of uncertainty, hence the margin of safety of 100, which accounts for both sources simultaneously in the risk assessment. For the protection of infants and children, an additional safety factor with a default value of lOx has been mandated by FQPA (1996). When implemented, this will create a safety margin of 1000. Once the RID has been established, it can be compared to the expected exposure. If anticipated human exposures are less than the RID, then such exposures are considered safe. A related approach is to divide the NOAEL by the exposure value and calculate the margin of safety value directly. With this method, also called the margins of exposure (MOE) approach, values greater than 100 (or other numbers if considered appropriate) are considered adequate to protect human health. In cases where the initial risk assessment indicates that the exposure is higher than the RID, additional measures must be taken to ensure human safety. The risk assessment can be improved by reducing the uncertainty contained within it. This can be accomplished by developing new information, which allows a better understanding of the biological processes resulting in the hazard, and also by improving the knowledge of the exposures actually taking place. Direct measurement of the true

1110

CHAPTER 53


exposure levels through the conduct of a worker exposure study may demonstrate that the real exposure values are less than the estimated values. In addition exposure levels can be closely estimated from knowledge of related products for which exposure studies have been conducted. There are also ways in which improved toxicology information can be used to increase the accuracy of risk assessments by reducing the uncertainty. For example, if the mechanism of toxicity in the animal model can be demonstrated with certainty, and if it can be shown that this mechanism does not exist in humans, then a higher NOAEL based on another toxicity end point may be appropriate. This would have the effect of increasing the safety relative to the exposure value. Another way to reduce the uncertainty is to study the effects of the chemical in humans if human exposure will occur as a result of the product use. If the information used in the risk assessment has been determined in a well-conducted study in humans and demonstrates that there is concordance between the animal and human databases, then the safety factor used for the interspecies extrapolation is not necessary, and a safety factor of 10 (for the intraspecies extrapolation only) is considered adequate. Therefore, it is normal and appropriate practice to utilize the human study to augment the derivation a NOAEL and the RID. In the past, risk assessments were conducted primarily for long-term or lifetime exposure to pesticides. However, more recently, regulatory agencies (such as V.S. EPA) have also begun to conduct risk assessments for acute (1 day), short-term (1-7 days), and intermediate (7 days to 3 months) time periods for oral (dietary), dermal, and inhalation routes of exposure. The process is generally the same as described previously, except that the RID approach is used to assess chronic risk whereas the MOE approach is used for the short-term and intermediate assessments. There are other methodologies used to calculate the short-term assessments, but this chapter will focus mainly on the methodology utilized by the V.S. EPA. It should be noted that there are other risk assessment methods which do not rely on the NOAEL, such as the benchmark dose (BMD) approach, which describe the dose-response data mathematically; then a point-of-departure (POD) approach is used in which a predefined effect level, such as dose predicted to give a 10% effect (EDlO), is used rather than the NOAEL. The safety factor is then applied to the dose-response curve with the POD as the starting point for the hazard component of the risk assessment. These types of models have advantages in certain situations, for example, where no NOAEL has been established in a study. These models have not been widely used in a regulatory context because there is currently no scientific consensus to drive policy decisions.

as a developmental neurotoxIcIty study. Aldicarb has high acute toxicity. Toxicity is that commonly associated with acetylcholinesterase inhibition (ChEI) caused by a carbamate pesticide, that is, cholinergic symptoms. These symptoms are dose-dependent, are rapidly reversible, and do not occur at expected human exposure levels. Aldicarb is neither genotoxic nor carcinogenic. It does not cause developmental or reproductive effects in the absence of maternal toxicity. The degradation pathway for aldicarb involves a combination of oxidation to aldicarb sulfoxide and then aldicarb sulfone and hydrolysis of parent, sulfoxide, and sulfone, to low toxicity compounds.

53.4.2 ALDICARB ACUTE TOXICITY Tables 53.3 and 53.4 provide acute toxicity data for aldicarb technical and for the formulated product Temik 15G. Aldicarb technical is highly toxic by the oral, dermal, and inhalation routes. Aldicarb is not a sensitizer. Aldicarb sulfoxide has similar potency with regard to acetylcholinesterase inhibition as aldicarb itself. Aldicarb sulfone is approximately 25 times less toxic than either aldicarb or aldicarb sulfoxide.

53.4.3 ALDICARB SUB CHRONIC TOXICITY In assessing the subchronic toxicity of aldicarb, the most sensitive indicator of exposure is cholinesterase inhibition. A number of subchronic and subacute (e.g., 14-day) oral studies have been conducted on aldicarb, aldicarb sulfoxide, and aldicarb sulfone. Results from study to study are consistent; for the sake of simplicity, only the longer term oral studies and a 21-day dermal study are discussed. Table 53.5 provides a summary of these studies. In an oral study, rats were fed aldicarb in their diet for 93 days at dose levels of 0, 0.02, 0.1, and 0.5 mg/kg/day. The noTable 53.3 Acute Toxicity Data for Aldicarb Technicala Toxicity Species

Findings

Acute oral

Rat

LDSO = 1.2 mg/kg

Rabbit

LDSO = 544 mg/kg

Rat

LCSO = 0.0039 mg/l

Rabbit

Moderately irritating

III

Rabbit

Slightly irritating

IV

Guinea

Not a sensitizer

Not

toxicity Acute dermal

53.4.1 SUMMARY OF ALDICARB TOXICITY Aldicarb has a very robust toxicity database including developmental, reproductive, and neurotoxicity studies, as well

II

toxicity Acute inhalation toxicity Primary eye

53.4 ALDICARB TOXICOLOGY PROFILE

category

Study

irritation Primary dermal irritation Dermal sensitization

pig

applicable

a Aldicarb technical is approximately 35% aldicarb in dichloromethane.

53.4 Toxicology Profile

1111

Table 53.6 Acute Neurotoxicity Study in Rats: Cholinesterase Inhibition in FemalesPercentage Relative to Control, 0.75 h

Table 53.4 Acute Toxicity Data For Temik IS G Toxicity Study

Species

Findingsa

Acute oral

Rat

LDSO = 2.14 (m)/2.46 (t)

Rabbit

LDso > 2000 mglkg

Rat

Not applicable-

Rabbit

Moderately irritating

III

Rabbit

Not irritating

IV

Guinea

No data-technical

toxicity Acute dermal

mg/kg III

toxicity Acute inhalation

sensitization

0.05

46.6

8.6

5.1

0.1

73.3

30.6

15.6

0.5

94.1

54.2

50.4

(21-day dermal study)

irritation Dermal

Brain

Plasma

53.4.4 ALDICARB NEUROTOXICITY STUDIES

irritation Primary dermal

RBC

Dose (mg/kg)

granular product

toxicity Primary eye

% Inhibition

category

pig

is not a sensitizer

Not applicable

am = male; f = female.

observed-adverse-effect level was 0.5 mg/kg/day based on the lack of effects on red blood cell ChE. In addition, body weight and food consumption were decreased at the highest dose level. An oral dog study was conducted to investigate the ChEI dose-response curve of aldicarb. During the 5-week study, the dogs were fed diets mixed with aldicarb technical at levels of 0, 0.35, 0.7, and 2 ppm (0.013, 0.023, and 0.069 mg/kg/day in males, and 0.012, 0.025, and 0.067 mg/kg/day in females). There was neither mortality nor any changes in body weight, food consumption, or clinical observation data indicative of a compound effect. A NOAEL based on erythrocyte cholinesterase was established at 0.07 mg/kg/day. There have been a number of subchronic and subacute studies using aldicarb sulfoxide and aldicarb sulfone. As stated in the acute toxicity section, the sulfoxide metabolite is comparable in toxicity to aldicarb; the sulfone metabolite is less toxic. In each case, ChEI is the most sensitive indicator of exposure.

There is a complete neurotoxicity database on aldicarb consisting of acute, subchronic, and developmental neurotoxicity studies. In addition, there is a "time to peak behavioral effects" study of a single oral administration of aldicarb technical. Also, there are acute neurotoxicity studies on both aldicarb sulfoxide and aldicarb sulfone. As discussed earlier, effects on ChEI are always the most sensitive indicators of both exposure and toxicity in the case of aldicarb. The aldicarb dose-effect relationship for ChEI is consistent across studies. A dose of 0.05 mg/kg/day gives the first indications of erythrocyte cholinesterase inhibition with no concomitant brain cholinesterase inhibition or behavioral changes. At 0.2 mg/kg/day, marked erythrocyte ChEI is observed accompanied by measurable inhibition in the brain and moderate clinical signs. Higher dose levels result in marked erythrocyte and brain ChEI and clinical signs, the magnitude of which increases with dose. A summary of the ChEI effects from three aldicarb neurotoxicity studies is provided in Tables 53.6-53.8. In these tables, data on ChEI are shown for females only, to simplify the presentation, but the effects for males show the same magnitude of effects.

Table 53.5 Subchronic Toxicity Data for Aldicarb Dose Level

NOAEU

LOEU

Study

Species

(mg/kg/day)

(mg/kg/day)

(mg/kg/day)

Toxicological endpoints

Oral toxicity,

Rat

0,0.02,0.1,0.5

0.1

0.5

Plasma ChEI at the highest dose tested. In addition, food consumption and body weight were decreased at this

93-day

dose level. RBC ChEI was not affected at all dose levels. Oral toxicity,

Dog

5-week

0,0.35,0.7, and 2 ppm; 0, 0.013, 0.023,

0.023 (m), 0.025 (t)

0.069 (m), 0.067 (t)

and 0.069 mg/kglday (m);

Plasma ChE! was over 20% compared to controls at the highest dose tested. RBC ChE! was not affected at all dose levels.

0,0.012,0.025, and 0.067 mg/kg/day (t) Dermal toxicity,

Rat

0, 100, 250, 500

100

250

Plasma ChE! at 250 mg/kglday dose level. RBC NOAEL

21-day,

was 250 mg/kglday; brain ChE! NOAEL was at least

Temik 15G

500 mg/kglday.

am

= male; f = female.

1112

CHAPTER 53


Table 53.7 Time to Peak Behavioral Effects Study in Rats, Females-Percentage Relative to Predose Value, I h DOSE (mg/kg)

% Inhibition plasma

% Inhibition RBC

0.1

79.7

39.5

0.4

93.1

62.2

0.6

93.9

78.4

In an oral feeding developmental neurotoxicity study in rats, the dose levels were 0, 0.05, 0.1, and 0.3 mg/kg/day. This study provides strong evidence that aldicarb does not cause permanent effects on the nervous system, and that the young are not more sensitive to the effects of aldicarb than mature animals. The maternal NOAEL was 0.05 mg/kg/day based on miosis at 0.1 mg/kg/day. The developmental NOAEL was 0.05 mg/kg/day based on postweaning body weight decrement, reduced hindlimb grip strength, and foot splay in F] females on postpartum day 35. Cholinesterase (ChE) activity was measured in the maternal animals on gestation day 7, and lactation days 7 and 11. Cholinesterase inhibition was not detected at 0.05 mg/kg/day, probably because the blood was collected two hours postdose and the enzyme had spontaneously reactivated by this time. In a subchronic toxicity study, this same dose level resulted in 24% erythrocyte ChE!. Thus, the systemic toxicity NOAEL was 0.05 mg/kg/day and the maternal NOAEL for red blood cell (RBC) ChEI was similar. These results demonstrate the lack of increased sensitivity to developing animals relative to adults because there were no developmental effects even in the presence of slight maternal ChEL In an acute gavage study, rats were treated with aldicarb sulfoxide at doses of 0, 0.05, 0.1, and 0.5 mg/kg/day. Cholinesterase activity was not measured in this study. There were no deaths in the study and no significant effects on body weights or body weight gain. Food intake values were reduced for males in the 0.5 mg/kg/day dose. Significant functional observation battery (FOB) effects were seen at the 0.5-mg/kg/day dose level at the time of peak effect. Significant decreases in motor activity were also seen at this same dose. There were no neuropathology effects. The NOAEL for FOB and motor activity was 0.1 mg/kg/day; the NOAEL for histopathology was 0.5 mg/kg/day. Table 53.8 13-Week Neurotoxicity Study in Rats: Cholinesterase Inhibition Females at 4 Weeks Relative to Control, 0.75 h

In an acute neurotoxicity gavage study in rats, the dose levels for aldicarb sulfone were 0, 1, 10, and 20 mg/kg/day. Cholinesterase activity was not measured in this study. No deaths occurred in the study. At the 20-mg/kg/day dose level, males and females showed significant decreases in food consumption, and males exhibited a significant reduction in body weight gain. At the time of peak effect, significant FOB effects were seen at the 10- and 20-mg/kg/day dose levels. Neuropathological evaluations revealed no effects at any dose. The NOAEL for FOB and motor activity was 1 mg/kg/day; the NOAEL for histopathology was 20 mg/kg/day.

53.4.5 ALDICARB DEVELOPMENTAL AND REPRODUCTIVE TOXICITY STUDIES

There is a complete developmental and reproductive toxicity database including a developmental neurotoxicity study (discussed in the previous section). Aldicarb does not cause developmental or reproductive effects in studies in the absence of maternal (or parental) toxicity. The following section discusses the study results, and these are summarized in Table 53.9. In an oral gavage developmental study, rats were given doses of 0, 0.125, 0.25, and 0.5 mg/kg/day. Maternal toxicity was indicated by maternal death and clinical signs were observed at the upper dose levels. Gestational parameters were not affected. No increased incidence of malformations was observed in the absence of clear maternal toxicity. The NOAEL for fetal toxicity was 0.25 mg/kg/day; fetal effects at the highest dose included dilated ventricles. In an oral gavage rabbit developmental study with doses of 0, 0.1,0.25, and 0.5 mg/kg/day, there were no fetal effects. Maternal toxicity was clearly established in the upper two dose levels with an increase in severity being observed at the highest dose level tested. The maternal NOAEL was 0.1 mg/kg/day based on decreased body weight and clinical signs. In a two-generation reproductive toxicity study, rats were fed a diet with 0,2,5, 10, or 20 ppm (ca. 0, 0.1, 0.25, 0.5, or 1.0 mg/kg/day). Cholinesterase inhibition and body weight changes in parents were observed at the upper dose levels. The maternal NOAEL was 0.25 mg/kg/day. The reproductive NOAEL was 0.5 mg/kg/day based on decreased pup weight and reduced viability. There were no reproductive effects in the absence of parental toxicity.

53.4.6 ALDICARB MUTAGENICITY STUDIES % Inhibition Dose (mg/kg/day)

Plasma

0.05

64.9

24.0

2.1

0.2

92.7

71.1

33.1

0.3

94.9

70.3

56.7

RBC

Brain

Studies covering gene mutations, chromosomal aberrations, unscheduled DNA synthesis, and dominant lethal effects were all negative for aldicarb. There is no concern for mutagenicity for aldicarb. A limited battery of genotoxicity studies on aldicarb sulfoxide and sulfone are also negative.

53.4 Toxicology Profile

1113

Table 53.9 Developmental and Reproductive Toxicity Data for Aldicarb NOAEL

LOEL

Study

Species

Dose level

(mg/kg/day)

(mg/kg/day)


Developmental

Rat

0,0.125,0.25,

Maternal toxicity, 0.125;

Maternal,0.25;

Maternal toxicity was indicated by death, reduced

toxicity

fetal toxicity, 0.25

0.5 mg/kg/day

body weight gain and food consumption, and

fetal,0.5

clinical signs of cholinesterase inhibition at 0.5 mg/kg/day, and reduced food consumption and body weights at 0.25 mg/kg/day. Fetal toxicity was indicated by increased dilated ventricles and reduced ossification of the 6th sternebrae. Developmental

Rabbit

toxicity

0,0.1,0.25,0.5

Maternal toxicity, 0.1;

Maternal, 0.25;

mg/kg/day

fetal toxicity: 0.5

fetal, >0.5

Signs of maternal toxicity included pale kidneys, hydroceles on oviducts, and decreased body weight. There was no fetal toxicity.

Two-generation

Rat

0,2,5,10,20

Parental,0.25; reproductive, 0.5

Parental: 0.5;

Maternal toxicity included plasma and RBC Cheri

reproductive

ppm; ca.

toxicity

0,0.1,0.25,0.5,

The reproductive LOEL is based on decreased pup

1.0 mg/kg/day

weights and reduced viability.

53.4.7 ALDICARB CHRONIC TOXICITY AND ONCOGENICITY STUDIES

Aldicarb has been shown to have no oncogenic potential when administered to rats and mice in lifetime experiments. Cholinesterase inhibition is the most sensitive indicator of ex-

reproductive, 1.0

and body weight changes at the upper dose levels.

po sure in chronic studies in rats and dogs. A discussion of chronic toxicity and oncogenicity data follows and Table 53.10 summarizes the study results. In a 2-year study, rats were fed aldicarb at levels of 0, I, 10, and 30 ppm in the diet (equivalent to ca. 0,0.05,0.5, and 1.5 mg/kg/day). There were no compound-related effects on sur-

Table 53.10 Chronic Toxicity and Oncogenicity Data for Aldicarb NOAEU

LOELa

Study

Species

Dose levels

(mg/kg/day)

(mg/kg/day)


Chronic toxicity or

Rat

0, I, 10,30 ppm;

0.05 (m), 0.59 (f)

0.5 (m), 1.5 (f)

Highest dose tested equivalent to greater than an LDSO

oncogenicity

ca. 0, 0.05, 0.5, 1.5

dosage when administered by gavage. Only clinical effect

mg/kg/day

was limited use of the tail at the highest dose tested. Body weights and body weight gains reduced at this dose level. Atrophy of the iris at the high dose.

Oncogenicity

Rat

0,2,6ppm;

Not evaluated

Not evaluated

None.

Not evaluated

Not evaluated

There were slight increases in mortality and slight

ca. 0,0.1,0.3 Chronic toxicity or

Rat

oncogenicity

0, 0.3 mg/kg/day for aldicarb; other

depressions in growth at certain stages for some of

doses for other

the test materials.

materials (e.g., a1dicarb sulfoxide) Oncogenicity

Mouse

0,0.1,0.3,0.7

Not evaluated

Not evaluated

Mortality and an increase in hematomas and lymphoid

Not evaluated

Not evaluated

None.

0.027

0.055

Plasma ChEI occurred at 0.055 mg/kg/day. Brain ChEI

mg/kg/day Oncogenicity

Mouse

0, 2, 6 ppm; ca. 0,

neoplasia were observed at the highest dose tested.

0.29, and 0.86 mg/kg/day Chronic toxicity

Dog

0, 1,2,5, 10 ppm; ca. 0, 0.027, 0.055, 0.13 mg/kg/day

am = male; f = female.

occurred at 0.13 mg/kg/day.

1114

CHAPTER 53


vival. It should be noted that the high dose of 1.5 mg/kg/day was greater than the LDso and was tolerated every day over the course of the study. This was possible because the aldicarb was administered via the diet, and the total dose was ingested in fractionated amounts throughout the day, allowing for ChEI reversibility between consumption periods. The principal clinical effect observed was limited use of the tail in high-dose males and females. Body weights and body weight gains were reduced in high-dose males and females. Also, atrophy of the iris occurred in this dose group. There was no evidence of direct organ toxicity, and no evidence of oncogenic effects. The NOAELs were 0.05 mg/kg/day in males and 0.59 mg/kg/day in females based on erythrocyte ChE!. It is noteworthy that the rats could tolerate such a dosing regimen over their entire lifespan, demonstrating that recovery is complete, accumulation of aldicarb does not occur, and there are no persistent effects following such exposure. In a National Cancer Institute (NCI) study, rats were fed aldicarb in the diet at concentrations of 0, 2, and 6 ppm (equivalent to ca. 0, 0.1, and 0.3 mg/kg/day). There was no mortality attributed to aldicarb and no effect on body weight was noted. It was concluded that aldicarb was not oncogenic; the NOAEL was the highest dose tested. In a third rat study, rats were fed aldicarb at dose levels of 0 and 0.3 mg/kg/day. In addition, other groups were fed aldicarb sulfoxide at dose levels of 0, 0.3, and 0.6 mg/kg/day, aldicarb sulfone at dose levels of 0,0.6, and 0.24 mg/kg/day, or a mixture of aldicarb sulfoxide and aldicarb sulfone at doses of 0, 0.5, and 1.2 mg/kg/day. Neither aldicarb nor either of its major metabolites was found to be oncogenic. There were slight increases in mortality and slight depressions in growth at certain stages for some of the test materials. Cholinesterase activity was measured at 6, 12, and 24 months during the study. Plasma, erythrocyte, and brain ChE activity were examined 24 hours after animals were removed from test diets. No ChEI was noted other than a possible slight inhibition with respect to plasma ChE. There have been three mouse oncogenicity studies conducted on aldicarb. The first is a National Cancer Institute study in which mice were fed 0, 2, or 6 ppm of aldicarb in the diet (equivalent to ca. 0, 0.29, and 0.86 mg/kg/day). It was concluded that aldicarb was not oncogenic. No effects on mortality or body weights were noted. In a second study, mice were fed aldicarb at doses of 0, 0.1, 0.2,0.4, and 0.7 mg/kg/day. Mortality was evident in males at the two highest dose levels, and in females at the three highest dose levels during the first few months of the study. Following this period, aldicarb was mixed in the diet in a different manner which eliminated the acute toxicity. Based on the mortality observed in the study, these data are not considered appropriate for the evaluation of an oncogenic response. In a third study, mice were fed aldicarb at dose levels of 0, 0.1,0.3, and 0.7 mg/kg/day. There was no effect on mortality or growth. Inclusion of aldicarb in the diet did not result in an increased incidence of oncogenic response. In a one-year study in dogs, groups of beagles were fed dietary concentrations of 0, 1,2, 5, and 10 ppm daily for 52 weeks

(equivalent to ca. 0, 0.027, 0.055, 0.13, and 0.24 mg/kg/day). The study was designed to produce maximum ChE! by limiting feeding time to 2 hours per day to mimic a bolus administration of aldicarb. Cholinesterase activity was measured from blood samples approximately 2 hours after the feeding period. There were no observable effects other than ChE!. The NOAEL for erythrocyte ChEI was 0.027 mg/kg/day. The lowest observed effect levels (LOELs) for erythrocyte and brain ChEI were 0.055 and 0.24 mg/kg/day, respectively. In another one-year dog feeding study, aldicarb sulfone was administered at dietary concentrations of 0, 5, 25, and 100 ppm (ca. 0, 0.125, 0.625, and 2.5 mg/kg/day). Cholinesterase determinations were taken approximately 2 hours after feeding to measure maximum ChE!. No mortality or treatment related clinical signs were seen. At the high dose, slight changes in spleen and thyroid-parathyroid weights and slight effects in the mandibular lymph nodes and adrenal cortex were observed. The NOAEL based on erythrocyte ChEI was 0.625 mg/kg/day. 53.4.8 HUMAN VOLUNTEER STUDIES

In a series of studies reported in 1973, groups of four adult male volunteers were administered aldicarb orally in aqueous solution at dose levels of 0.025,0.05, and 0.1 milligram per kilogram of body weight (mg/kg bw). Clinical signs were recorded and whole blood cholinesterase activity was measured up to 6 hours after administration of the sample. Total urine voided was collected and aldicarb-excretion patterns for the initial 8 hours after dosing were evaluated. In addition, spot samples were taken at 12 and 24 hours. In other studies, two additional subjects were administered aldicarb in water solution at dose levels of 0.05 and 0.26 mg/kg bw. Dose levels of 0.1 and 0.26 mg/kg bw are considered to be high doses. Acute signs, typical of anticholinesterase agents, were observed at the high-dose levels (0.1 and 0.26 mg/kg bw) within 1 hour of aldicarb administration. Cholinesterase depression at these very high dose levels was observed in all volunteers within 1-2 hours after treatment. Within the first 6 hours of treatment, cholinesterase depression and clinical signs of poisoning were within normal levels. There were no signs of treatment observed at the 0.05-mg/kg bw dose level. Urine analysis showed that approximately 10% of the administered dose was excreted as carbamates within the first 8hour interval. Cholinesterase analyses confirmed the same rapid inhibition and recovery pattern with man as had been observed in experimental animals. In 1992 Aventis Crop Science conducted a human volunteer study, which was conducted and performed under globally accepted ethical guidelines established for such work. This was a double blind, placebo controlled study, in which aldicarb was administered as a single oral dose to healthy male and female subjects. The doses administered were: placebo (22 subjects16 males and 6 females); 0.01 mg/kg bw (8 males); 0.025 mg/kg bw (8 males and 4 females); 0.05 mg/kg bw (8 males and 4 females); and 0.075 mg/kg bw (4 males). Volunteers were screened before entry for general medical history by examination and laboratory tests including hematology, clinical

53.5 Use of Pharmacokinetics in Aldicarb Risk Assessments chemistry, and urinalysis. Clinical measurements were made at intervals before and after dosing. These included vital signs (systolic and diastolic blood pressure, pulse rate), pulmonary function tests, pupil size, electrocardiographs (ECGs), salivation, and clinical signs of nausea, vomiting, diarrhea, sweating, abdominal cramps, involuntary movement, or slurred speech. Samples were taken for urinalysis, clinical chemistry (including red blood cell and plasma cholinesterase activity), and hematology before and after dosing. There were no clinically significant changes in vital signs, pupil size, pulmonary function, ECGs, salivation, clinical signs, clinical chemistry (apart from cholinesterase), hematology, or urinalysis in the study. Cholinesterase activity was the only parameter affected during the study. Red blood cell and plasma cholinesterase was maximally depressed at 1 hour after dosing and had recovered by 8 hours in all subjects. The fall in cholinesterase activity and recovery was dose-related. No biologically significant depression of erythrocyte cholinesterase activity (> 20%) was seen in subjects treated with 0.01 or 0.025 mg/kg bw or in plasma cholinesterase at 0.01 mg/kg bw. Depression in cholinesterase activity > 20% was seen in erythrocytes at 0.05, and 0.075 mg/kg bw and in plasma at 0.025, 0.05, and 0.075 mg/kg bw. A single volunteer (0.075-mg/kg bw group, actual dose 0.06 mg/kg bw) reported some sweating, which is not considered to be related to aldicarb. Nevertheless, the NOAEL for clinical signs was reported as 0.05 mg/kg bw and the NOAEL based on erythrocyte cholinesterase inhibition was 0.025 mg/kg bw.

53.5 USE OF PHARMACOKINETICS IN ALDICARB RISK ASSESSMENTS Although it has been known since the initial discovery of the carbamates that this class of compounds present reduced risk compared to organophosphates due to the rapid reversibility of cholinesterase inhibition, a quantitative analysis of ChE reversibility is desirable because it allows the conduct of more precise risk assessments taking into account the exposure patterns (route of exposure, duration, and repetition). The kinetics of reversibility of cholinesterase inhibition for aldicarb has been studied through the development of a mathematical model, which describes the reversal of inhibition using data generated during the 1992 study in human volunteers. As previously described, healthy male and female volunteers were administered aldicarb in a study to evaluate its effects on plasma (PChE) and red blood cell (RChE) cholinesterase activity. Cholinesterase activity was measured at three pretreatment time points ( -16 and - 3 hr, and immediately predose) and 1,2, 4, 6, 8, and 21 hr post dose. Measurement of PChE and RChE activity at these time points allowed the determination of the maximum inhibitory effect for each dose level as well as the time required for spontaneous reactivation of the enzyme activity through hydrolysis of the carbamate. In addition, samples were taken for urinalysis, clinical chemistry, and hematology evaluations.

1115

The data presented in this chapter will focus on the male RChE data because these are more robust (a larger number of male than female subjects were observed), and because RChE data are generally considered to have more biological relevance than PChE in regard to potential peripheral and central nervous system effects (Carlock et aI., 1999; V.S. EPA, 2000). A graphical representation of the male RChE data in the volunteer study is shown in Fig. 53.2. The data show a very stable baseline activity level and rapid decrease in RChE activity following exposure to the bolus dose. The RChE inhibition is then subsequently reversed during a time period ranging from minutes to approximately 2 hours depending on the magnitude of the initial inhibition event (i.e., the smaller the inhibition, the faster the recovery time). The mathematical model was developed to quantitatively describe the data from the study. With such a model, one can predict the degree of cholinesterase inhibition which would be expected to occur following exposure to a dose that had not been tested experimentally. Additionally, one can calculate the time needed to completely reverse the effects of cholinesterase inhibition following exposure to a given dose. The approach presented here is based on the knowledge that the clinical determination of the reversal of cholinesterase inhibition following exposure to a carbamate is in effect the measurement of the removal of the carbamate from the body. This is so because the reactivation of the cholinesterase enzyme occurs via hydrolysis of the carbamate; hydrolysis results in the production of inactive metabolites and the full restoration of the functional capability of the enzyme. If the enzyme is not inhibited, then there is no longer free aldicarb in the blood compartment available for binding to the enzyme. Thus a mathematical description of the reactivation of cholinesterase activity following cholinesterase inhibition due to carbamate exposure can be regarded as the inverse of the model for removal of the carbamate. The latter model can be used in risk assessments for human health issues surrounding carbamate exposure because, when the initial exposure level is known, the amount remaining after a given point in time can be calculated. In general terms, the model is f(tUPI,

2, 1>3)

= 1>d 1 -

1>2 exp[ -1>3(t -

1)]}

where t is time in hours after exposure (t ~ 1), 1>1 is the horizontal asymptote that is approached over time (t -+ 00), 1>1 (l - 1>2) is the minimum value at 1 hour, and 1>3 is a scale parameter related to the rate of recovery. The value lOO1>2 is the percentage reduction from the asymptotic maximum value to the minimum value at time 1 hour. In biological terms, 1>1 is the baseline cholinesterase activity level, 1>2 is a parameter describing the degree of inhibition as a function of dose, and 1>3 describes the rate of recovery of enzymatic activity (Williams et aI., 2000). The model belongs to the general class of nonlinear mixed effects growth curve models containing both fixed and random effects. The fixed effects permit estimation of the average curves describing the cholinesterase activity levels following exposure to a given dose; the random effects account for individual variability observed in a repeated measures study of

1116

CHAPTER 53


Activity (mU/mL)

it.:·.S

~:4.?

13. i

--------------.--~.(!! -~~

~

:\

~---------

*""--------

-t:iBJ

0

'

-

~:=--=-! --III

t-,- ,-------------.

::::\W......:

+------------+'~

~(/·+-.i

••

W' .::/

0 :~ Cl

~o ?....

'"' lR~V)(i) =M

Cl

\

r

1

:~~"'O

\ 'mf°ll~Cl t Cl j/ )1

C

man (v) R(V) (i)

R(v}

rearranqement[Cl~

DIELDRIN

micerv)

H(v)

(ri)

",,0

".

0

'.A ~

s

f1'r;I'-:/",

r-

~ (Z) 0

.'-Q.'

Figure 55.3 Biotransformation routes of dieldrin, photodieldrin, and endrin. Wavy line indicates only partial ring structure shown. H, housefly; I, some insects; mo, mosquito; m, microorganisms; P, pig; Ra, rabbit; R, rat; S, sheep; v, in vivo; i, in vitro; S ---+, sesoxane inhibits. See text for references.

roach were confirmed (Schroeder et aI., 1977; Shankland and Schroeder, 1973), but, based on the less intense neuroactive effect of the diols and their very slow action in vivo, the diols were concluded to be detoxification products in the cockroach. Both diols are produced as metabolites of dieldrin by rats and mice and appear to be detoxification products in these mammals. These pharmacokinetic studies raise questions about possible internal barriers to the penetration of such molecules and their metabolites to critical sites in the nervous system. Similar problems are apparent throughout the series and are difficult to resolve experimentally. Moreover, a particular metabolite might be a bioactivation product in one species but a detoxification product in another. The tendency for molecular rearrangements in the environment (e.g., from exposure to sunlight) and in vivo has complicated investigations on residues and metabolites. Photoconversion products are frequently more toxic than the parent insecticides and may themselves be further metabolized; for example, photodieldrin (PD; 2, Fig. 55.3) is oxidatively dechlorinated to the pentachloroketone (3, Fig. 55.3; "Klein's ketone";

55.2 Discovery of Polychlorocycloalkane Metabolism as a Factorin Toxicity Klein et al., 1970) in rats and insects (Baldwin and Robinson, 1969; Baldwin et al., 1972; Khan et al., 1970; Matthews and Matsumura, 1969); PD has a much shorter half-life (23 days) than dieldrin (10-13 days) in rat adipose tissue but is two- to four-fold more toxic to rodents and insects (Table 55.2). Dieldrin-treated rats excrete 9-hydroxy-dieldrin (9HD; 4, Fig. 55.3) in the feces and the pentachloroketone in the urine (Richardson et al., 1968), and these are considered to arise by alternative modes of attack from beneath the ring system (Fig. 55.3). The same pentachloroketone (3) was produced, along with varying amounts of 9-HD and cis- and trans-DDA, in American cockroaches, German cockroaches (Blattella germanica), and houseflies (Nelson and Matsumura, 1973). The pentachloroketone (3) was reported to be more toxic than photodieldrin to mosquitoes and houseflies (Khan et al., 1970) but less toxic and slower acting than PD to the German cockroach (Kadous and Matsumura, 1982; Reddy and Khan, 1977) indicating that PD itself is the active toxic ant in this insect. PD acted four-fold more rapidly (LDso, 0.01 ).Lg/insect) than dieldrin (LDso, 0.05 ).Lg/insect) and two-fold more rapidly than the pentachloroketone (LDso, 0.13 ).Lg/insect) observations that suggest it has pharmacokinetic properties more favorab1e for toxicity than the other compounds. 9-HD (4) appeared to be more toxic (LDso, 0.02 ).Lg/insect) than dieldrin to the German cockroach and may contribute to dieldrin's toxicity in this insect; the cis- and trans-DDAs appeared to be relatively nontoxic when injected. From other experiments on the American cockroach, it seems clear that these metabolites can enter the nerve cord from the insect body and are also produced in small amounts by metabolism in the nerve itself. Isodrin was found to be epoxidized to endrin (Fig. 55.3) in houseflies (Brooks, 1960) and subsequently by liver microsomes from rats and rabbits, as a result of mixed-function oxidase (MFO) action (Nakatsugawa et al., 1965; Wong and Terriere, 1965). Endrin incubated with pig or rat liver microsomes in the presence of reduced nicotinamide adenine dinucleotide phosphate (NADPH) gave a monohydroxy-derivative, formation of which was inhibited by sesoxane, indicating MFO involvement (Brooks, 1969). It soon became clear from mammalian studies that the inversion of the unchlorinated norbornene nucleus in isodrin and endrin (as compared with aldrin and dieldrin) exposes this ring to enzymatic hydroxylation in vivo and greatly increases the rate of elimination of these compounds from mammalian tissues, in contrast to their behavior in insect tissues. Endrin is generally more toxic to vertebrates and less toxic to some insects than dieldrin; whereas the latter undergoes 9-hydroxylation syn to the epoxide ring and 9-HD (4, Fig. 55.3) is eliminated by conjugation in mammals, endrin is both syn- (slowly) and anti- (rapidly) hydroxy1ated; the antiderivative (5, Fig. 55.3) is rapidly conjugated and excreted but the syn-isomer (6, Fig. 55.3) is further oxidized to 9-keto-endrin (9-KEN, also called 12-keto-endrin; 7, Fig. 55.3), a remarkable example of the profound influence of stereochemistry on metabolic pathways. Bridge-end (tertiary) hydroxylation also occurs and endrin trans-diol is a minor metabolite. 9-KEN is some five-fold more

1135

toxic than endrin to rats and appears to be the ultimate toxic metabolite of endrin (Bedford et al., 1975a; Hutson et al., 1975). Species differences are evident, since Kadous and Matsumura (1982) reported the order of endrin metabolite toxicity to male German cockroaches as 5-0H > anti-9-0H > 9-keto-, whereas the order on topical application was 9-keto '" syn-9OH > endrin » anti-9-0H to houseflies and 9-syn-OH > 9-keto-> endrin » anti-9-0H to blowflies (Brooks and Mace, 1987). Also in this report, syn-9-hydroxydieldrin (9-HD; 4) was essentially nontoxic to houseflies and blowflies, whereas the order 9-oxadie1drin (9-0D; 13, Fig. 55.1) '" dieldrin > 9-ketodieldrin (9-KD; 8, Fig. 55.3) and 9-oxadie1drin (9-0D) ~ 9-KD > dieldrin, respectively, was found for houseflies and blowflies. Toxic 9-KD is apparently not formed from 9-HD in vivo, possibly because, in contrast to the situation with endrin, steric hindrance prevents enzymic attack on the hydroxyl group. Each set of toxicities lies within a narrow range and the toxicities of 9-KD and 9-0D might be expected to be similar, because 9-0D is an isostere of 9-KD, in which -C=O has been replaced by the more compact 5,8-bridged oxirane structure. These results show that several of the oxidative metabolites of these insecticides retain insect toxicity and may contribute to the toxic effect of the parent insecticides. 55.2.3 HEPTACHLOR, CHLORDENE, DIHYDROHEPTACHLOR, CHLORDANE,ANDISOBENZAN Further chlorination of the feebly toxic chlordene, the DielsAlder adduct of "hex" and cyclopentadiene, gave heptachlor, the dihydroheptachlor isomers (Table 55.3 and Fig. 55.4), and the chlordane isomers (Fig. 55.5). The nontoxic adduct of "hex" and cis-2-butene-1 A-diol, namely 5,6-bis(hydroxymethyl)hexachloronorbornene-2-ene, is the precursor to which isobenzan, endosulfan (Fig. 55.6), bromocyclen (Bromodan®), and chlorbicyclen (Alodan®) (Fig. 55.7; 20 and 21, respectively) are related. The last two compounds were once used to control animal ectoparasites because of their low mammalian toxicity; endosulfan is still used extensively today, whereas isobenzan was discontinued in 1965. A preparation of heptachlor is mentioned in the original Hyman patent (Hyman, 1949) on chlordane. Numerous investigations from 1951 demonstrated the formation of heptachlor exo-epoxide, m.p. 160°C (HE160; Fig. 55.4); the less insecticidal endo-epoxide, m.p. 90°C (HE190) is not formed in vivo but can be obtained indirectly by chemical synthesis. The biotransformations of chlordene and heptachlor involve allylic hydroxylation for chlordene, hydrolysis of allylic chlorine for heptachlor, epoxidation (Miles et al., 1969), and epoxide ring hydration (Brooks, 1966; Fig. 55.4). Microorganisms can degrade heptachlor by removing the allylic chlorine, either reductively or by hydrolysis, so that the degradation routes for chlordene can then be followed (Miles et al., 1969); in some soils, the production of 1-hydroxychlordene (1, Fig. 55.4) is comparable to HE160 production. The hydroxylated metabolites appear to

1136

CHAPTER 55

Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners

Table 55.2 Toxicity Data for Some PolychlorocycIohexane Insecticides and Their Transformation Products

Compound

Rodent acute oral LDSOa

Topical 24-h LDsOb

(mg/kg)

housefly

Lindane

90-190

Aldrin

38-60

6,7-Dihydroaldrin

1.5 1.5 40 (3.0)

5,8-0xadihydroaldrin Dieldrin

(~g1g)

30 (1.5) 47

1.0

77 (m) HCE

>400

90 (2.0) 200-400 (m)

HEOM Photodieldrin

>500 10

0.12

7 (m) Didechlorodieldrin (DD)

0.2

0.9 1.4

9-Hydroxydieldrin (9-HD)

>400 (m)

trans-Dihydroaldrin-diol (t-DDA)

1,250 (m)

Isodrin

750 >750

12-17

6,7-Dihydroisodrin

3.0 39 (4.0)

Photoisodrin

>2,000

15 (3.0)

Endrin

5.6

2.0

29 (m) 9-Keto-endrin

1.0

0.95

anti-9-Hydroxy-endrin (AHEN)

2.5-5.5

syn-9- H ydroxy-endrin

1.2

Heptachlor epoxide (HEI60)

60

>100 1.2 1.0

Heptachlor epoxide (HE90)

6.0 2,400-4,600

1-H ydroxychlordene

Inactive 50 (20)

Chlordene Chlordene exo-epoxide

35 (4.0)

1-Hydroxy-chlordene exo-epoxide

Inactive

trans-Chlordane

11

1,100 500-600

eis-Chlordane

4.0 5.0

alpha-Endosulfan

76

beta-Endosulfan

240

9.0

Endosulfan sulfate

76

9.5

Endosulfan diol

>15,000

>500

Endosulfan ether

>15.000

>500

alpha-Hydroxy-endosulfan ether

1,750

>500

Endosulfan lactone

306 (m)

>500

Isobenzan

3-10

1.0

6.0 (m) Bromocyclen (Bromodan®)

13,000

11.5

Chlorbicyclen (Alodan®)

15,000

15.5

Mirex Chlordecone Toxaphene (technical) Toxaphene (component B) a For

600->3,000 125 90-270 75 (m; ip)

rat unless marked (m) for mouse. bparenthetic values in housefly column are toxicities measured with sesoxane (5 ~g), preapplied before the insecticide to inhibit microsomal oxidases. Data from Brooks and Harrison (1964a), Buchel et al. (l966a, 1966b), Jager (1970), Khan et al. (1970), Korte (1967), Maier-Bode (1968), Miles et al. (1969), and Bedford et al. (1975a), Smith (1991).

55.2 Discovery of Polychlorocycloalkane Metabolism as a Factorin Toxicity

be detoxification products in mammals. This is difficult to prove in insects, however. Chlordene (2, Fig. 55.4) and its exo-epoxide (3, Fig. 55.4) have a weak housefly toxicity, which is synergized lO-fold by sesoxane, suggesting that the biotransformations observed in microsomal preparations are detoxications (Brooks, 1966; Brooks and Harrison, 1964a, 1967a, b). Is heptachlor much more toxic than chlordene because the allylic chlorine inhibits hydroxylation in this position and also ensures that heptachlor is converted into the metabolically stable epoxide ?-a question reminiscent of the aldrin/dieldrin situation. The view that heptachlor is intrinsically toxic is supported by the toxicity (Table 55.3) of the alpha- and beta-dihydroheptachlor isomers (Fig. 55.4), formed by the addition of hydrogen chloride to the double bond of chlordene. Their housefly toxicity also is synergized by sesoxane, which suggests that metabolic hydroxylation, which for them replaces the epoxidation of heptachlor, results in detoxication. Beta-dihydroheptachlor (beta-DH; Fig. 55.4; 2, Table 55.3) is particularly interesting because of its low mammalian toxicity (Buchel et aI., 1966a, 1966b). In the presence of NADPH, pig liver microsomes converted alpha, beta-, and gamma-DH into a variety of hydroxylation products, which are illustrated for beta-DH in Fig. 55.4. These were chlorohydrins, obtained by simple hydroxylation of the cyclopentane rings; alcohols, formed by elimination of the single chlorine atom on the cyclopentane ring; dihydroxy-compounds; and a ketone (e.g., 2-keto-dihydrochlordene from beta-DH, which may afford the corresponding alcohol via a ketoreductase reaction). The 2-0H-dihydrochlordene excreted by rats fed beta-DH (Korte, 1967) may arise in this way. Sim-

1137

ilar metabolites were produced in housefly microsomes, although no dihydroxy-compounds were detected. Sesoxane inhibited the hydroxylations, which doubtless explains the synergism against houseflies observed in vivo (Table 55.3). The metabolism of the two chlordane isomers, alpha- (= trans-l ,2-dichlorodihydrochlordene) and beta- (= cis-l,2-dichlorodihydrochlordene), is complex (Fig. 55.5). Either isomer might give heptachlor by dehydrochlorination and hence HE160 and all the metabolites arising therefrom. In fact, the metabolites in rats include l-exo,2-dichlorochlordene (1, Fig. 55.5), oxychlordane (2), l-exo-hydroxy-2-chlorochlordene (3), l-exo-hydroxy-2-chloro-2,3-epoxychlordene (4), l-exohydroxy, 2-endo-chlorodihydrochlordene (chlordene chlorohydrin), 1,2-trans-dihydroxydihydrochlordene, and the metabolites of heptachlor (Brimfield and Street, 1979; Brimfield et aI., 1978; Tashiro and Matsumura, 1977). A similar series of compounds was excreted in the form of unidentified conjugates in the urine of rabbits treated with these chlordane isomers (Balba and Saha, 1978). These biotransformations demonstrate the remarkable versatility of the drug-metabolizing enzymes. In particular, the formation in rats of oxychlordane (Fig. 55.5), analogous to heptachlor epoxide and said to be more toxic than trans-chlordane (Street and Blau, 1972), is a bioactivation due to the unexpected formation of a stable epoxide in vivo, presumably following an enzymatic desaturation that introduced a 2,3-double bond. There was no evidence for epoxide formation from either cis- or trans-chlordane in houseflies, however. Notably, transchlordane was three-fold less toxic than eis-chlordane to this insect, and neither isomer was synergized by sesoxane (Brooks

Table 55.3 Toxicities of Dihydroheptachlor and Chlordane Isomers to Housefly and Mouse exoB, exo-

endo-

Compound

A

(1) a-

fJ-

(2)

Housefly LDso

Housefly LDSO (flg/f1y with

(flg/fly)a

sesoxane)a

Mouse acute oral LDso (mg/kg)

B

C

D

Cl

H

H

H

0.26

0.015

1,285

H

Cl

H

H

0.16

0.015

>9,000

(3) y-

H

H

Cl

H

1.8

0.07

>6,000

(4)b

Cl

H

Cl

H

0.22

0.22

1,100

(5)b

Cl

Cl

H

H

0.08

0.08

(6)

H

Cl

Cl

H

0.04

>600

(7)b

Cl

H

H

Cl

0.04

31

Alodan

0.31

0.05

Dieldrin

0.02

0.02

aTopicaI application; sesoxane applied (5 flgl20 mg fly) before insecticide. b 4, trans-chlordane; 5, eis-chlordane; 7, 8-chlordane. cLDSO for rat. Data compiled from Brooks and Harrison (l964a, 1967b) and Buchel et al. (l966a, b).

500-600

15,000e 75-100

1138

CHAPTER 55

Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocyc1oalkanes and Recent Congeners

tY

P::(H

!~:;;

;::!iJ) (J

/

~7

~O

rrY/°PI(i)

'l'"

~~~)

B

3..

~

~ HE 160

~ H(,t~~Yh ~ Cl

Cl +

td

RCv)

p:j~

PICI)

CHLORDANE

3

lC~LPHA_(TRANS_) BETA_(CIs_)~16 ~: 1

CHLORDANE

PI(I) H(I)

p:jH

rrY

CI

MCv)(I); MCV)

rr-,

1-'.--1 ALPHA- ~ GAMMA- el

BETA-

I

+

DIHYDROHEPTACHLOR ISOMERS

~

~

~CHLOROHYDRINS,

~V) Cl

PCI) 2,3-EPOXIDE 4--~ 'OXYCHLORDANE' R(v) Cl (2)

Y'V

HEPTACHLO~j{V)(I)J

SYNTHESIS CHCl GAS)

RA(V~V)

~~(V)

~CH~~~DENE ' ~ \

R(v)

'"""011;"'1"

Cl

E,\:

-

OH

~:~)

W 6

(1)

R(v)

RA'vy /

,RC?):

rr;,

1

1

M : I (I) I

I

~

RCI)

RA'v)

OH

Cl

OH

;J;;jq) :

/H(vHI)

r

HEPTACHLOR, 3-oH-CHLORDANE

~OH

;:t;J0H

~H(V){I}~~ (2) H(v){l)

~

HE 190

ETC.

AS FOR BETA-ISOMER

OH Figure 55.4 Biotransformations of chlordene, heptachlor, and the dihydroheptachlor isomers. B, bacteria; H, housefly; I, some insects; M, mammals generally; Pi, pig; R, rat; m, microorganisms; Ra, rabbit; E, abiotic conversion; v, in vivo; i, in vitro. All structures contain the fully chlorinated norbomene moiety.

and Harrison, 1964a), indicating that intrinsic toxicities were being measured. Alpha- and beta-DH were as toxic as heptachlor when synergized; synergized gamma-DH was four-fold less toxic than synergized alpha- and beta-DH and as toxic as eis-chlordane (Table 55.3). This suggests that the 2-endochlorine atoms in trans-chlordane and gamma-DH contribute less to toxicity than the exo-chlorines present in alpha- and beta-DH and eis-chlordane. Moreover, an additional chlorine introduced into the 2-exo-position of gamma-DH (Table 55.3) increases its toxicity more than 40-fold, so that the resulting gem-dichloro-compound is as toxic as beta-DH having the single exo-chlorine in this position. Does this extra exo-chlorine simply reduce the possibilities for metabolic detoxification that are more likely for gamma-DH (exo-side of the ring exposed to enzymatic attack) and transchlordane, or do exo-chlorines increase the affinity of these molecules for a critical binding site in the nervous system? That synergized gamma-DH is as toxic as eis-chlordane (unaffected by sesoxane) may suggest that metabolism is the only factor involved and that the exo- or endo-disposition of the chlorines is immaterial. There is also the interesting question of the

(.!.>

Figure 55.5 Biotransformations of the eis- and trans-chlordane isomers. Abbreviations as in Fig. 55.3. All structures contain the fully chlorinated norbomene moiety.

role of symmetry; the most insecticidal compounds in this series are beta-DH, the 2,2-gem-dichloro-analog (gamma-DH), and l-exo, 3-exo-dichlorodihydrochlordene (delta-chlordane; Table 55.4), all having a plane of symmetry, in contrast to the other molecules discussed, for which the enantiomers may differ in toxicity (see later discussion on the heptachlor epoxide enantiomers in Section 55.3.1). Production of isobenzan (Telodrin® ceased in 1965 (Jager, 1970), but this molecule (Fig. 55.6) remains of theoretical interest as a cyclic ether analog of gamma-chlordane, which, like the latter, has high insect and mammalian toxicity. Enzymatic attack on the chlorinated cyclic ether structure of isobenzan analogous to the biotransformations noted for the chlordane isomers results in hydrophilic metabolites such as derivatives of the gamma-hydroxy-acid (1, Fig. 55.6), which afforded the lactone (2) and alcohol (3) on hydrolysis. Alternatively, these might arise directly by oxidative or hydrolytic elimination of chlorine atoms from the cyclopentane ring. Of particular interest because they illustrate the variety of structures having toxicity in this series is the mixture of two interconvertible isomeric ketones (14 and 15, Fig. 55.7), with high insect and mammalian toxicity (housefly LD50, 0.5 j..lg/g; rat LDso, 7 mg/kg) , which can be obtained chemically from 5,6-bis(hydroxymethyl)-HCNB ("endosulfan-diol"). Transannular dehydrochlorination affords an even more toxic cage ketone (16, Fig. 55.7; housefly LDso, 0.25 j..lg/g; rat acute oral LDso, 1.0 mg/kg). These analogs of isobenzan are more compact versions of the various cage molecules formed from dieldrin and provide further evidence that the dichloroethylene moiety of cyclodienes can be replaced by other polar moieties without loss of toxicity and with increased toxicity in some cases.

55.2 Discovery of PolychlorocycIoalkane Metabolism as a Factorin Toxicity

1139

Table 55.4 Insect Toxicity of Aldrin and Dieldrin Relatives, Including Some Molecules with Fewer Chlorine Atoms

x =

y

=

carbon,

except for compounds (3 ) and (4 ) Chlorination in aldrin analog Chemical

2

3

4

lO-syn

lO-anti

HFa

GRb 2.0

(1)

---------

6-Cl

- - - - - - (aldrin)

0.55

(2)

---------

6-Cl

---

---------

6,7 -epoxide: dieldrin

1.0

1.0

(3)

---------

6-Cl

---

---------

6,7-N=N-

3.6

5.3

(4)

---------

---------

6,7-N=N(--+ 0)-

4.45

2.1

lOa

2.1

9.3

6-Cl

---

(5)

H

H

4

lOs

(6)

H

H

4

lOs

lOa (6,7-epoxide: DD)

3.8

8.0

(7)

H

H

4

H

H

0.03

0.8

(8)

2

3

4

H

lOa

0.08

0.65

(9)

2

3

4

lOs

H

0.02

Inactive

a Housefty toxicity compared with dieldrin (1.0) by direct spray.

bGerman cockroach toxicity compared with dieldrin (1.0) by exposure to dry films on paper. Data compiled from Soloway (1965).

55.2.4 ENDOSULFAN (THIODAN) Technical endosulfan is a 7:3 mixture of the alpha- (m.p. lO9°C) and beta- (m.p. 213°C) isomers, the former (Fig. 55.6) having an "extended," dieldrin-like structure (see also Section 55.4.3) and the latter having a cagelike structure resembling endrin stereochemically. The alpha-isomer is more toxic than the beta-isomer to mammals and houseflies; both are oxidized in vivo to endosulfan sulfate (4, Fig. 55.6), which resembles beta-endosulfan stereochemically and has similar toxicity to alpha-endosulfan, so that this conversion is analogous to the aldrin-to-dieldrin one. The cyclic sulfite (and sulfate) ester structures completely alter the behavior of the endosulfans, which disappear quite rapidly from living tissue, partly by hydrolysis to the parent nontoxic endosulfan-diol and metabolites similar to those formed from isobenzan. The sulfate is formed faster from the alpha- than from the beta-isomer in houseflies and is as toxic as beta-endosulfan to these insects (Barnes and Ware, 1965); cyclodiene-resistant flies eliminated these isomers more rapidly than normal (S-) flies, but the tissues contained only the toxic sulfate, which also appears in the body fat of mammals but disappears rapidly when exposure ceases. Endosulfan-treated locusts excreted the sulfate, endosulfan ether, alpha-hydroxy-endosulfan ether (3, Fig. 55.6), and the corresponding lactone (2). Endosulfan-treated mice stored the sulfate transiently in their fat and excreted endosulfan, the sulfate, and the parent diol in feces (Maier-Bode, 1968). It is evident that endosulfan is a relatively nonpersistent compound in mammals (Dorough et aI., 1978) and has generally favorable environmental properties, apart from high fish toxicity, which requires caution in aquatic situations. With the exception of the

toxic sulfate, metabolites of endosulfan isomers are undoubtedly detoxication products.

55.2.5 TOXAPHENE, MIREX, CHLORDECONE (KEPONE) Toxaphene (camphechlor) is a complex mixture of some 177 compounds obtained by chlorinating camphene to a 67-69% chlorine content (Pollock and Kilgore, 1980; Saleh et aI., 1979). The identified compounds are actually chlorinated bornanes arising from the Wagner-Meerwein rearrangement of the camphene skeleton, among which the octachloronorbornanes; 2,2,5-endo, 6-exo-8,8,9,1O-octachloro-norbornane (17, Fig. 55.7) and 2,2,5-endo,6-exo-8,9,9, lO-octa-chloronorbornane (18, Fig. 55.7), are highly potent, with mouse ip LDso values of 2-3 mglkg (Turner et aI., 1977). The less toxic 2,2,5endo,6-exo,8,9,lO-heptachloronorbornane (19; compound B; LDso, 75 mglkg) was potentiated eight-fold by PBO administered prior to the insecticide, suggesting the possibility of oxidative detoxication mechanisms for this compound. Experiments with rat liver preparations confirmed metabolism by MFO, and the formation of glutathione and glucuronide conjugates (Chandurkar and Matsumura, 1979) could be demonstrated (see Smith, 1991). The positioning of the added chlorine substituents in compound B seems to be critical; at the 3-exo-position and in the 10-chloromethyl moiety, an additional chlorine greatly reduces mouse toxicity, as does the combination of 3-exochlorination and 5,6-dehydrochlorination, to give a vinylic chlorine atom. Notably also, the simpler (less bulky) molecules

1140

CHAPTER 55 Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners

~o, ~ ~~S02

•

(4)ENDOSULFAN SULFATE

1

HYDROLYSIS

/() ENDOSULFAN 'ETHER'

ENVIRONMENT BIOTIC, ABIOTIC

~

i

Cl

,

mfo ?

CI~I 0 Cl ,0'7),' _LOC_U_ST_)~ I

0"

HYDROLYSIS

(l)

~o

0

(2) ENDOSULFAN 'LACTONE'

Cl ALPHA-ENDOSULFAN

Y1'CH20S03H p......COOH

\ '4

'-\ OXIOOREDUCTASE?

/~.

~CI ISOBENZAN

Figure 55.6 Major transfonnations of alpha-endosulfan and isobenzan. All structures contain the fully chlorinated norbomene moiety. Note that - -0- - indicates the skewed ("trans") position of the second oxygen in the "twist-chair" (asymmetric) configuration of alpha-endosulfan (Schmidt et aI., 1997).

hexachloronorbornene-2,5-diene and heptachloronorborn-2ene used to prepare cyclodiene insecticides lack toxicity, which only appears when halomethyl groups are introduced into the nucleus as in bromocyclen (Bromodan®; 20, Fig. 55.7) and chlorobicyclen (Alodan®, 21). Both are synergized lO-fold by sesoxane in houseflies (Brooks and Harrison, 1964a; Table 55.2) and are quite good insecticides with very favorable mammalian toxicity (rat acute oral LDsos, 13000-15000 mg/kg); that is, they appear to be considerably more selective (insect versus mammal) than the most toxic components of toxaphene. Mirex (22, Fig. 55.7) is the fully chlorinated cage molecule, formed by the self-condensation of two molecules of "hex," and might be expected to be rather resistant to enzymatic attack. Animal tissue levels plateau only slowly on exposure and decrease very slowly when exposure ceases. One chlorine atom is reductively replaced in the environment to give photomirex (8-monohydro-mirex), which appears to behave like mirex in the rat (Chu et aI., 1979; Hallett et aI., 1978). Reductive dechlorination can occur in vivo; 2,8-dihydromirex and 5,10dihydromirex have been identified as rat metabolites. Whereas 2,8-dihydromirex does not appear to be further metabolized, 5,IO-dihydromirex appears to be converted into more polar metabolites, which appear in rat urine (Yarbrough et aI., 1983). Mirex has low mammalian toxicity (rat oral LDso ranging from 600 to > 3000 mg/kg) and its signs of poisoning differ from those produced by the less chlorinated cyclodienes. The metabolism of mirex in houseflies is equally slow, and Shankland (1982) compared its slow insecticidal action with the delayed onset of dieldrin poisoning discussed earlier. The onset of poisoning following topical application of

lethal doses of mirex to the American cockroach occurred only after 3 days. Moreover, when isolated sixth abdominal ganglia were irrigated with suspensions of 5 x 10-4 M mirex for 4 h, there was no change in the patterns of spontaneous activity or elicited postganglionic responses. Ganglia excised from symptomatic cockroaches showed, however, spontaneous after-discharge behavior characteristic of poisoning following dieldrin treatment. Hemicholinium-3, which depletes Ach stars, eliminated the neuroactivity in giant fibers, but the ganglia remained responsive to nicotine, as is found in dieldrin poisoning. Because mirex appears to be highly resistant to biotransformation, Shankland concluded that the delayed action was unlikely to involve a requirement for bioactivation and must arise from the intrinsic properties of this highly chlorinated molecule, such as slow penetration through diffusion barriers in the insect central nervous system. Chlordecone (Kepone, 23, Fig. 55.7) differs from mirex in having a carbonyl group, which is probably responsible for its moderately rapid clearance from animal tissues. In humans and pigs, this is via the alcohol (chlordecol), and a cytosolic ketoreductase, which can effect this reduction, has been found in gerbil and human liver (Molowa et aI., 1986). Bloomquist and Shankland (1983) found that chlordecone produced the same signs of poisoning as mirex in the American cockroach and concluded that chlordecone has the same mode of action as dieldrin, although, like mirex, it acts more slowly. From experiments on the displacement of [3H] picrotoxinin (PTX) binding by mirex and chlordecone from American cockroach head membranes, Tanaka et al. (1984) concluded that chlordecone interacts with the PTX-binding site, as expected, whereas mirex was much less potent in this respect; moreover, dieldrin-resistant

55.3 Structure-Toxicity Relationship and Mode of Action

Cl

II

~

Cl

CIO

CI~CI Cl

~~

11

o U4,15)

1141

10

C

O~

(6)

Cl

Cl ..... ,~~__ Cl CH 2CI

(2() Bromodan®

(8) Toxaphene

~ (23)

Mirex

'0

Chlordecone (Kepone)

Cl

Figure 55.7

Chemical structures of compounds mentioned in the text.

German cockroaches were resistant to chlordecone but not to mirex. Chlordecone is also known to have inhibitory effects on neurotransmitter uptake in mammals and such an action may also contribute to its insect toxicity.

55.3 STRUCTURE-TOXICITY RELATIONSHIP AND MODE OF ACTION 55.3.1 FULLY CHLORINATED CYCLODIENES: SUBSTITUTED HEXACHLORONORBORNENES (HCNB) Soloway (1965) published a comprehensive review on the structure-activity relationships of cyclodiene insecticides at a time when information on their metabolism was just beginning to appear, so his review makes only passing reference to the possible influence of metabolism but provides a

great deal of information about toxicity trends in numerous series of cyclodiene analogs. Initially, he emphasized the similarity between heptachlor epoxide HE160/cis-chlordane and HE90ltrans-chlordane (Figs. 55.4 and 55.5), each pair having two similarly oriented electronegative atoms (i.e., oxygen and chlorine), with toxicity greater in the first (exo,exo) orientation than in the second (exo,endo) orientation of these substituents. Delta-chlordane (Table 55.3) with l-exo,3-exo chlorine substituents is a highly insecticidal symmetrical variant of the orientation found in the HE 160lcis-chlordane pair. Interestingly, delta-chlordane is an analog of alodan in which the two side-chain chlorines have become fixed in the exopositions by the extra carbon atom of the cyclopentane ring and their insect toxicities are of the same order when alodan is synergized by sesoxane (Table 55.3). The second (3-exo) chlorine in delta-chlordane has a severe effect on mammalian toxicity, because this molecule has a much higher rodent toxicity than either alodan, alpha- (1, Table 55.4) or beta-dihydroheptachlor(2) (DH), or the chlordane isomers (4, 5). As noted already, the order of housefly toxicity of the DH-isomers is beta-DH > alpha-

1142

CHAPTER 55


DH > gamma-DH; beta-DH has an exo-chlorine and is also symmetrical, alpha-DH has an exo-chlorine but is asymmetrical, whereas gamma-DH has an endo-chlorine, which, being "hidden" beneath the ring system, may be less accessible to a critical binding site and also leaves the exo-face of the ring more exposed to metabolic attack from the exo-side (compare the metabolism of endrin in Section 55.2.2). Soloway presented insect toxicity data for many derivatives of HCNB of the chlordane, isobenzan, endosulfan, aldrin, and isodrin series, together with lindane, which already appeared to have the same mode of action (Busvine, 1964). He concluded that high insecticidal activity required the presence of two electronegative centers within a narrow range of distance and direction with respect to one another and placed on or across the plane of symmetry defined by the CClz-bridge. Many cyclodienes fulfill these requirements but some, such as dihydroaldrin, dihydroisodrin, and photoisodrin (9, 10, and 11, respectively, Fig. 55.1) (only one electronegative center) and bromodan (20, Fig. 55.7), alpha-DH, exo-chlordene epoxide (Fig. 55.4), and HCE (8, Fig. 55.1) (asymmetrical), do not, yet are indicated to have high intrinsic toxicities when their metabolism is suppressed in vivo. Evidently, the involvement of a second electronegative center such as an epoxide ring in binding to the site of action may increase the affinity of the molecule for this site, by hydrogen bonding, for example. Thus, lack of a second electronegative center may explain the earlier noted slow action of the dihydro-compounds, which, in the absence of an inhibitor of metabolic oxidations, may afford them increased opportunity for both detoxication and binding to inert storage sites. Notably, the cage molecule photoisodrin is more polar than the related dihydroisodrin and dihydroaldrin and acts rapidly, especially when synergized; it may have the more favorable pharmacokinetic properties of the more rapidly acting epoxides, although apparently lacking their second electronegative center (Brooks, 1973; Brooks and Harrison, 1963). There is limited information about the relative toxicities of enantiomeric forms of chiral cyclodienes, which are obviously of interest in this context. The epoxide hydrolases of pig liver microsomes selectively hydrate the same enantiomers of chlordene epoxide, HCE and HE90 (Brooks et aI., 1968). The isolated residual epoxides appeared to have the same order of toxicity to houseflies as their respective racemates, which are not detectably hydrated by this insect (Brooks et aI., 1970). Miyazaki et al. (1978, 1979, 1980) synthesized the pure enantiomers of chlordene, chlordene exo-epoxide, HEl60, 2-chloroheptachlor (Fig. 55.8), and 3-chloroheptachlor and found that their toxicities to the German cockroach (topical LD50, !-1g/g) were in the order (+ )-chlordene (148) > racemic chlordene (>300) > (-)-chlordene (inactive); (-)chlordene epoxide (74) > racemic chlordene epoxide (158) > (+ )-epoxide (inactive); racemic heptachlor (2.64) > (+)heptachlor (3.38) > (- )-heptachlor (5.32); (+ )-HE160 (1.29) > racemate (1.82) > (- )-HE160 (2.98); (+ )-2-chloroheptachlor (20) > racemate (50) > (- )-2-chloroheptachlor (100); 3-chloroheptachlor (enantiomers and racemate inactive).

Miyazaki et al. concluded that (- )-chlordene is intrinsically nontoxic to the German cockroach, observing that the corresponding (+ )-epoxide (nontoxic) formed in vivo is metabolized to the expected oxidative and hydrolytic products (Brooks and Harrison, 1965) at about the same rate as the toxic ( - )-epoxide from observably toxic (+ )-chlordene. However, they also considered (+ )-chlordene to be intrinsically inactive, therefore requiring bioactivation by conversion into the toxic (- )-epoxide in vivo. Unfortunately, these experiments did not include a synergist to suppress oxidative metabolism. Experiments with houseflies showed that both chlordene and dihydrochlordene had low but measurable toxicities to that insect, which were synergized by sesoxane (Brooks, 1966; Brooks and Harrison, 1964a); in fact, synergized chlordene was only fivefold less toxic than synergized chlordene exo-epoxide. The role of epoxidation in the toxicities of the heptachlor enantiomers (or those of 2-chloroheptachlor) has not been reported. The (+)- to (- )-heptachlor toxicity ratio for German cockroach was 1.56; for (+)- to (-)-HE160 it was 2.3 and for (+)- to (-)-2chloroheptachlor it was 5.0, with the more toxic (+ )-antipodes

_-..~o ~ (+~

~ .~. C1 6

(-)-,

Chlordene

Chlordene

~o '

C1 6

~

~ .-

(+)-

Heptachlor

~-epoxide

Cl

,

(-)-

C16

2

Cl

~

Heptachlor exo-epoxide (HE 160)

2-Chloroheptachlor

Figure 55.8 Absolute stereochemical configuration of the enantiomers of chlordene, chlordene epoxide, heptachlor, heptachlor epoxide (HEI60), and 2-chloroheptachlor as established by Miyazaki et al. (I978, 1979, 1980).

55.3 Structure-Toxicity Relationship and Mode of Action and toxic (- )-chlordene epoxide all having the same absolute stereochemistry (Fig. 55.8). Apart from (+ )-chlordene epoxide and the antipodes of 3-chloroheptachlor, the other antipodes are clearly all active but the ratio of 2.3 for the HEl60 antipodes is likely to be the safest measure of comparative intrinsic toxicities in this series because the known stability of this epoxide should avoid or minimize the complication of metabolism in vivo. Thus, although one absolute configuration of HEl60 is favored, both are toxic, which might be expected if the critical binding site is in a symmetrical (or nearly symmetrical) cylinder of about the same diameter as the molecules discussed, so that either antipode can interact reasonably well with such a site in the bore of the structure, now known to be the chloride ionophore of the GABAA -receptor (Section 55.4). Notably, alpha-DH must exist in enantiomeric forms, which, if superimposed, give a symmetrical "composite" molecule that resembles both delta-chlordane and isobenzan (its oxygen isostere). Likewise, superimposition of the enantiomers of both HEl60 and HE90 gives "composites" that are similar to both delta-chlordane and isobenzan. Such symmetrical molecules might be expected to interact particularly well with a closefitting cylindrical binding site. 55.3.2 COMPOUNDS WITH FEWER, OR NO CHLORINE ATOMS 55.3.2.1 Reductive Dechlorination of Cyclodienes Early information (Soloway, 1965) indicated that the unchlorinated methano-bridge of aldrin could be replaced by 9-syn-CI-CH- and that of isodrin by -CH2CH2- or spirocyclopropane, but the overall molecular length could not exceed that delineated by dieldrin or alpha-endosulfan. There were, however, interesting indications that some of the chlorine atoms in the hexachloronorbomene moiety could be replaced by hydrogen (Table 55.4). Species differences were evident; an aldrin analog (7, Table 55.4) having only the two one- and four-bridge chlorines was reported to be nearly as toxic as dieldrin to the German cockroach, although nontoxic to other insects tested. In aldrin, the methano-bridge chlorine atom anti to the chlorinated double bond was found to be more important for toxicity than the syn-chlorine (compare 8 and 9, Table 55.4), and replacement of the two ethylenic chlorines in dieldrin by hydrogen to give didechloro-dieldrin (DD; 6, Table 55.4) increased housefly toxicity four-fold and toxicity to the German cockroach eightfold; the latter insect appears to be particularly sensitive to these compounds. The high toxicity of diazaaldrin and its N-oxide (3 and 4, Table 55.4) should also be noted. Of interest was the possibility that if the increase in toxicity effected by replacement of the ethylenic chlorines in dieldrin proved to be a general phenomenon for cyclodiene insecticides, it might be possible to combine the change to a tetrachloronorbomene moiety with a more labile epoxide ring or other labile system (e.g., cyclic sulfite as in endosulfan) to produce useful insecticides having both oxidative and hydrolytic

1143

detoxication routes that would be more selective and environmentally acceptable. Selective dechlorination of several series of cyclodienes was undertaken to test this possibility (Brooks, 1975,1977,1980,1985; Brooks and Mace, 1987; Brooks et aI., 1981). It transpired that the effect of dechlorination was not uniform but depended on the structure of the molecule as a whole. The most consistent observation for all series was the greater importance of the anti- versus the syn-chlorine atom in the pentachloronorbomene moiety, as noted for aldrin by Soloway (1965). These two changes for dieldrin were combined to give the 1,4, anti-l O-trichloro-analog of dieldrin (DSD; 24, Fig. 55.7), which approached dieldrin in toxicity to houseflies and blowflies and was 7-20-fold more toxic than its 10syn-chloro-isomer with the chlorine atom adjacent to the double bond. Thus, the pentagonal arrangement of chlorine atoms evident in lindane and in the cyclodienes derived from HCNB (Busvine, 1964) is not sacrosanct for cyclodienes. 55.3.2.2 Structural Convergence of Cyclodienes and Their Dechlorinated Analog with Other Cage Convulsants Acting at the Chloride Ionophore Further information arose from comparisons with the naturally occurring convulsant picrotoxinin (PTX; Fig. 55.9), which Hathway et al. (1965) found to have effects similar to those of dieldrin and isobenzan on ammonia metabolism in rat brain. PTX antagonizes the action of GABA by blocking the chloride ion channel associated with its receptor (Kadous et aI., 1983; Takeuchi and Takeuchi, 1966, 1972). Evidence was then reported that cyclodienes and lindane compete with PTX at a commmon binding site in cockroach brain (Matsumura and Ghiasuddin, 1983; Tanaka et aI., 1984), and this was the site of convulsant action of these compounds, a proposal supported by the similarity in the neurophysiological effects of of the cyclodienes and PTX and the cross-resistance to PTX shown by cyclodiene-resistant cockroaches. The structural similarities among PTX, HE160, and lindane led Ozoe and Matsumura (1986) to elaborate the two electronegative center hypothesis of Soloway (1965) in a series of PTX analogs that emphasized the importance of the bulky trans-substituent on the lactone ring as a third requirement for interaction with the PTX binding site. This is the point of convergence (Fig. 55.9) with the cyclodienes (B) and lindane (C), for which the lO-anti-chlorine and the central axial chlorine (or the central equatorial chlorine), respectively, appear to provide the appropriate bulky substituent (Brooks and Mace, 1987) and the highly toxic cage compounds such as t-butylbicyclophosphorothionate (TBPS) and t-butylbicycloorthobenzoate(TBOB), in which the t-butylmoiety is the necessary bulky substituent (Casida et aI., 1985; Palmer and Casida, 1985). The highly insecticidal orthobenzoate EBOB (Fig. 55.10) proved to be the best ligand for the insect GABA-receptor chloride ionophore, and eH]-EBOB has been used subsequently in binding displacement studies with numerous putative channel blockers at the PTX binding site. Attempts to simplify the orthobenzoate ring system of these potent

1144

CHAPTER 55

Interactions with the garnma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners A (trans-J CB

a Cl

A

3

1

Cl S

cl

e

c Ya 8 ' 'Cl

~lc~_",...~.!

~

B

a

Cl

e

Cl

A

PTX

B Cyclodienes

antl-lO-CI

C12

C

Lindane

trans-lsopropenyl

Cl3

D

E

Figure 55.9 Structures of (A) picrotoxinin (PTX) with its bulky beta-isopropenyl group trans- (or anti-) to the lactone ring; (B) heptachlor epoxide; (C) lindane, showing the aaaeee configuration of chlorine atoms essential for toxicity. A, anti-; S, syn-; a, axial; e, equatorial substituents. D shows the highly chlorinated face of a cycIodiene insecticide, as in heptachlor epoxide (B above), and E, the lactone ring system as in PTX, both viewed from the right ("end-on" position). In the corresponding view of lindane (C viewed from below), note that the chlorine equivalent to the syn-IO-chlorine of fully chlorinated cycIodienes is replaced by a hydrogen (double arrow). However, the electrostatically more favored superimposition of lindane on PTX (Calder et aI., 1993) is C viewed from the right, in which the three-axial chlorines of lindane are together equivalent to the trans-bulky isopropenyl substituent and lactone ring of PTX.

cage convulsants and the search for structural changes to confer selective toxicity in favor of mammals versus insects produced numerous insecticidal dithianes, oxathianes, and their sulfoxides and sulfones (Palmer and Casida, 1995; Wacher et aI., 1992). Figure 55.10 further shows the convergence of structural changes between partly (Brooks and Mace, 1987) and totally (Ozoe et aI., 1990) dechlorinated alpha-endosulfan (1, Fig. 55.10) (which surprisingly retains measurable housefly toxicity (LD50, ca. 43 J.l.g/g; 20-fold less toxic than alphaendosulfan, when co-applied with sesoxane), the toxic t-butyl trioxabicyclooctanes (TBOs) such as 2 (Fig. 55.10) (Palmer and Casida, 1985), and the dithianes described by Elliott et al. (1990). Although the 1,3-dioxan (not shown but analogous to structure 7) corresponding to 2 (Fig. 55.10) is weakly insecticidal (Palmer and Casida, 1995), the hybrid molecule (3) of deschloro-alpha-endosulfan (1) and that 1,3-dioxan was more toxic than fully dechlorinated alpha-endosulfan (1), indicating that extra rigidity conferred on the 1,3-dioxan structure can improve insecticidal activity. In the hybrid molecule (3), the norbomene moiety appears to provide the bulky substituent (compare t-butyl in 2), while possibly increasing the conformational rigidity of the dioxan moiety. Furthermore, Ozoe et al. (1993) showed that the extended structure 4

(Fig. 55.10) is highly insecticidal and that compound 5 has intermediate toxicity, which is entirely lost by any chlorination in the norbomene nucleus. Thus, the dioxan (5) loses its housefly toxicity (LD50, 5.5 J.l.g/g) completely when oneor two-bridge chlorines are introduced and there is evidently a crossover point between the two structural types, because the fully chlorinated but not extended dioxan (6) has the same toxicity, which is much reduced in various partially dechlorinated analogs (see Section 55.4.3 for further discussion). Bridge bis-chlorination of fully dechlorinated alpha-endosulfan (Fig. 55.10) greatly reduces its housefly toxicity (Ozoe et aI., 1993), but the 10-anti-chlorine analog retains measurable toxicity and several dechlorinated analogs having this Cl atom in combination with one ethylenic chlorine and the two bridgeend chlorines are highly toxic, showing that the additional chlorines are required for binding this shortened molecule in the critical site (Brooks and Mace, 1987), implying the requirement for a minimum of four chlorines for high toxicity. Nevertheless, at least one of the bridge-end chlorines in cyclodienes can be replaced, because a dieldrin analog (25, Fig. 55.11), which has one bridge-end carbon atom replaced by nitrogen, has appreciable insect toxicity (Gladstone and Wong, 1977).

55.3 Structure-Toxicity Relationship and Mode of Action

1145

anti-

H~CL 4H 0,

H~O, syn-

'V

O'S

~

0-

~

anti-lO, pentachloroisomer (1.0)

arious dechlor endosulfans

~

~9

0-

CL Cl ~O"

P=S TBPS

0 .... ~

0

4H

~

6H~-

, rE0'

S 1\

0

(265)

----.'

(>500)

6H

~~R R=

(1) deschloro-endosulfan (43)

(3) 4-Br-Ph (15)

R= phenyl (Ph) (TBOB) (2) R= 4-Br-Ph (0.8) 4-CHaC-Ph various TBOs 4-CN-Ph

(4) 4-CII;C-Ph (0.33) (5) 4-CN-Ph (5.5)

~ o

6CI

~

(6)

(5.5)

>

U

f:I

~~t'"-·-~ (8)

f,

H axial

~)2

Ph-4-C",CH

(0.24)

Figure 55.10 Convergence between exploration of reductively dechlorinated alpha-endosulfan analog (norbornene type) (Brooks and Mace, 1987; Ozoe et aI., 1990, 1993), the trioxabicyclooctane-derived cage convulsants, and the more recent dioxans and dithianes (Pulman et al., 1996). The bracketed number following a chemical number or structure is the housefly topical LDSO (J.1g/g; measured in the presence of sesoxane or piperonyl butoxide).

Many trioxabicyclooctanes are highly toxic to both mammals and insects but remarkable selectivity can be conferred on some structures by appropriate derivatization; thus, the trimethylsilyl-derivative (26, Fig. 55.11) of the 4-n-butyl analog of EBOB (Fig. 55.10) is highly toxic to houseflies (LDso, 0.43 Ilg/g) but poorly toxic to mice (LDso, > 400 mg/kg). This derivative appears to be oxidatively reconverted into its toxic ethynyl precursor in houseflies, whereas in the mouse this oxidative bioactivation is much less important (Palmer et aI., 1991b). Because all the compounds mentioned and some other classes of cage convulsants are believed to act at the PTX binding site, a considerable array of compact molecules is now available with which to delineate this site. A 4-ethynylphenylsubstituent in the 2-position of 5-t-butyl-l,3-dioxane (7, Fig. 55.10) or 1,3-dithiane (8, Fig. 55.10) was found to be more effective than a 4-bromophenyl-substituent; the trans-(linear) ethynylphenyl dithiane (8) was somewhat more toxic to houseflies than the analogous trans-dioxane (7) and cis- (angular) isomers were generally equally toxic to or less toxic than trans-isomers (Palmer and Casida, 1995; Pulman et aI., 1996). With the possibility of oxidation at sulfur in vivo, which may enforce additional conformational rigidity and also increase binding propensity, the situation becomes more complex (see Section 55.4.3).

55.3.3 LINKS BETWEEN POLYCHLOROCYCLOALKANEAND RECENT HETEROCYCLICS APPARENTLY ACTING AT THE CHLORIDE IONOPHORE

Recently, arylpyrazoles, such as fipronil (27, Fig. 55.11), and various 5-alkyl-2-arylpyrimidines (28) and 1,3-thiazines (29) (Pulman et aI., 1996), in which the planar heterocyclic ring replaces the spacers formed by the TBO and 1,3-dioxane and dithiane structures, have been added to the list of chloride ionophore blockers. Insecticidal activity was also found in triazoles (30, 31) (Boddy et aI., 1996; Von Keyserlingk and Willis, 1992) and pyrimidinones (32) (Whittle et aI., 1995) and a spirosultam (33) (Bloomquist et aI., 1993), demonstrating the diversity of structures that probably act at this site. Cole et al. (1994) examined the inhibition of eH]-alphaendosulfan binding in housefly head membranes by lindane and several cyclodienes and concluded that these insecticides are the only GABA-receptor ionophore blockers that consistently inhibit the binding in these membranes not only of the earlier used ligands such as eSS]-TBPS and eH]-EBOB but also of eH]-alpha-endosulfan. However, a representative dithiane, EBOB, fipronil, and other pyrazoles were less effective in inhibiting [3H]-alpha-endosulfan binding than the chlorinated insecticides, from which it appeared that the latter compete

1146

Cl

CHAPTER 55


~ )~ ~

r'

~"'PH-4-C=CSi (CH3)3

0

(26)

c.l

t

'-.FN ~N ~NQ- PH-4-C=CH ~ (28)

PH-4-C=CH (29)

Cl Figure 55.11 35).

Chemical structures of compounds mentioned in the text (25-

directly for the endosulfan site, whereas the others bind with different inhibition kinetics or at a site more closely coupled to the EBOB than to the endosulfan binding domain. Notably, the channel activator avermectin Ba did not inhibit endosulfan binding. An even more suitable ligand for the chlorinated insecticides is [3H]-BIDN (34, Fig. 55.12) (Holyoke et aI., 1994), a simple norbomene derivative, which has high insect and mammalian toxicity (KOlbl et aI., 1981; Middleton and Bingham, 1982). Several putative affinity probes for the binding site have also been described (Casida and Pulman, 1994). When aryl pyrazoles synthesized as herbicides were found to be insecticidal, their convulsive activity was not immediately recognised to result from GABA-antagonism (Klis et aI., 1991). Co le et al. (1993) reported subsequently that several compounds, including fipronil (27, Fig. 55.11) (Colliot et aI., 1992; Hatton et aI., 1988), which has become a commercially successful insecticide, blocked the GABA-gated chloride ionophore with higher potency for a site in housefly than in mouse brain, offering the possibility of selective toxicity. Fipronil has relatively low acute mammalian toxicity (Section 55.4.3). It inhibits eH]-EBOB binding to housefly head membranes and dieldrin-resistant flies show some resistance to it (Cole et al.,

1993; Colliot et aI., 1992), providing a clue to its mode of action. The cyclodiene insecticides and lindane were found to be potent displacers of eSS]-TBPS binding to GABA-receptors in rat brain and inhibitors of GABA-dependent 36 Cl ion flux into rat brain microsacs, from which it was suggested that these PCCAs act as noncompetitive blockers of GABAAreceptors (Abalis et aI., 1985; Gant et aI., 1987; Lawrence and Casida, 1984). Potency in these assays correlates with toxicity (Casida et aI., 1988) but TBPS is not a potent insecticide and [3S S]_TBPS is unsuitable as a radioligand for insect studies; it appears that the structural features required for binding at the housefly GABA-receptor are different from those for the mammalian one and eH]-EBOB, a highly potent insecticide, was ultimately designed as a superior ligand for insect binding studies (Deng et aI., 1991) and generally provides a good correlation between its displacement by PCCAs and their housefly toxicities. By use of this ligand, it was concluded that PCCA, PTX, dithiane-related compounds, and phenylpyrazoles all have the same mode of insecticidal action, a view supported by the up to 27-fold cross-resistance to EBOB shown by dieldrin-resistant houseflies (Cole et aI., 1993). Moreover, the naturally occurring insecticide avermectin Bla and derived moxidectin (Fisher, 1997), which behave as GABA-agonists, stimulating rather than inhibiting chloride ion influx, are potent noncompetitive inhibitors of EBOB binding. This implies that averrnectin action involves the chloride ionophore but that it is bound at a site different from that involving EBOB and PCCA; nor, in contrast to EBOB, is there cross-resistance to dieldrin, so that the channel modification that confers dieldrin resistance does not apparently involve the avermectin binding site (Deng et aI., 1991). Based on ligand binding studies, Deng et al. (1993) proposed four partly associated sites in the housefly GABA chloride ionophore that are relevant to insecticidal action: site A, interacting with EBOB and its isosteres; B with TBPS and isosteres; C with phenylpyrazo1es; and D with averrnectins. Action at sites A and C gives similar signs of poisoning and crossresistance to dieldrin; PCCA and some TBPS isosteres may act at both A and B. The avermectin site D is coupled in some way with A and C but not to the TBPS site B, which is also distinct from the phenylpyrazole site C. Thus, the reduced affinity for [3H]-EBOB binding observed in dieldrinresistant houseflies is due to its reduced affinity for the PCCA binding site, and the cross-resistance noted for TBOs, lindane, toxaphene, cyclodienes, dithianes, arylsilatranes (35, Fig. 55.11), and PTX suggests that the structural modifications in the EBOB binding site are involved in resistance to all these insecticides (Hawkinson and Casida, 1993) but fortunately do not confer resistance to averrnectins, which have very high toxicity against agricultural and household insect pests, phytophagous mites, and plant and animal nematodes.

55.4 Molecular Mechanism of Action

1147

Chloride ion channel

OUTSIDE

:,--,

.

,"" • • 1

"I

,- -~

,I :

I

LIPID BILAYER

,.. - - ' l

, I

I

-, I

CYTOPLASM

-- -

-

(INSIDE)

binding site for noncompetitive blockers (cyclodienes, lindane, PTX, arylpyrazoles, etc.)

Figure 55.12 Schematic representation of the GABAA -receptor of mammalian brain, showing five transmembrane glycoprotein subunits, each with their four trans-membrane helices (M \-M4), of which the M2 segments (shown as cylinders, and black circles in the plan view) are believed to form the pore of the integral chloride ion channel (MacDonald and Olsen, 1994). A modified subunit carrying cyclodiene resistance in Drosophila (Rdl) shows homology with the mammalian brain beta-subunit (ffrench-Constant et ai., 1991).

55.4 MOLECULAR MECHANISM OF ACTION 55.4.1 TOPOGRAPHY OF THE GAMMA-AMINOBUTYRIC ACID A-RECEPTOR GABA is the principal neurotransmitter of the mammalian and insect central nervous system (CNS) and the insect neuromuscular junction. In mammals, baclofen-sensitive GABABreceptors are coupled to calcium and potassium channels and the action of GABA is mediated by G-proteins. In contrast, GABAA -receptors, of interest here, are members of the super family of ligand-gated ion channels that contain a chloride ionophore (Schofield et aI., 1987). Simplistically, an inhibitory GABA-ergic nerve terminal abutting on the presynaptic terminal of another nerve that releases a neurotransmitter (e.g., acetylcholine, ACh) releases GABA when stimulated. GABA then diffuses to the presynaptic terminal of the other nerve, where it binds to a GABAA -receptor, causing entry of chloride ions and resulting in hyperpolarization of the terminal and inhibition of release of the other neurotransmitter. Thus, postsynaptic stimulation of the other nerve by its transmitter (e.g., ACh) is reduced. This inhibitory mechanism explains the apparent cholinergic effects of dieldrin and lindane on Ameri-

can cockroach ganglia (Shankland and Schroeder, 1973; Uchida et aI., 1978), because disinhibition (blockade of a presynaptic chloride ionophore) of the presynaptic terminal of a cholinergic nerve should result in uninhibited ACh release and consequent hyperstimulation of the postsynaptic terminal, as is observed. The same basic mechanism for disinhibition may, of course, affect nerve terminals involving neurotransmitters other than ACh (Joy, 1982) with a variety of possible effects, depending on the species and on differing nerve architecture. The GABAA -receptor of human brain consists of four or five 50-60 kDa glycoprotein subunits, each of which contains four (M 1-M4) hydrophobic domains (alpha-helices) that traverse the membrane (Fig. 55.12) (MacDonald and Olsen, 1994; Schofield et aI., 1987) and contribute to and stabilize the walls of the chloride ionophore. The five M2-domains are believed to be arranged so as to form the 5.6-A-diameter lumen of the channel, with the side chains of their threonines and serines forming hydrophilic rings that contribute to the induction of ion flow.

55.4.2 MOLECULAR BIOLOGY OF CYCLODIENE RESISTANCE Recently, a cyclodiene resistance-conferring gene, Rdl, from the fruit fly, Drosophila melanogaster, has been cloned and shows homology with the mammalian brain beta-subunit

1148

CHAPTER 55

Interactions with the gamma-Aminobutyric Acid A-Receptor: PolychlorocycIoalkanes and Recent Congeners

(ffrench-Constant et al., 1991). Dieldrin resistance was subsequently found to be associated with the point mutation alanine 302 to serine (Ala302-Ser) within the M2 membranespanning domain, near the site conferring charge selectivity in the closely related nicotinic ACh-receptor (ffrench-Constant et al., 1993a, b). Homooligomeric, wild-type Rdl-receptors expressed in Xenopus oocytes showed the expected electrophysio10gica1 properties of GABA-receptors; channels containing the Ala302 _ Ser mutation, when similarly expressed, were identical with the wild-type ones but had consistently lower sensitivity to dieldrin and PTX, sensitivity to these being reduced about lOO-fold. Notably, Lee et al. (1995) reported no detectable specific [3H]-EBOB binding in resistant D. melanogaster strains carrying the Ala302 -rep1acement, despite high specific binding to membranes from susceptible flies, indicating the involvement of Ala 302 in EBOB binding to GABA-receptors containing Rdl subunits. Similar results were found for the dieldrin resistance mutation Ala302 -glycine found in Drosophila simulans (ffrench-Constant et al., 1993a, b). Furthermore, EBOB blocks chloride ion currents generated by Rdl homomultimers expressed in insect cells and the Ala302 -Ser replacement reduces sensitivity to this block lO-fold (Lee et al., 1995). Examination of the Rdl gene from three different insect orders has revealed that in all cases Ala302 is replaced by either a serine or a glycine (less effectively), indicating that this mechanism is universal. The change confers reduced sensitivity to PTX, lindane, and TBPS, lower channel conductances, extended channel open times and shorter closed times, and a markedly reduced rate of GABA-induced receptor desensitization. From a simple model to represent binding and allosteric changes, it has been suggested that the preceding mutations are the only ones that can directly weaken cyclodiene binding to the desensitized (antagonist favored) conformation of the receptor and simultaneously destabilize the antagonist-favored conformation through an allosteric mechanism, resulting in a powerful dual-resistance mechanism (ffrench-Constant et al., 1995; Zhang et al., 1994). In this model, the antagonist associates with the open channel but binds much more tightly when the channel next changes into the desensitized (closed) state, so that this configuration is stabilized. If homomultimers are present in vivo, which may not, however, be the case (Zhang et al., 1995), the preceding mutation could lead to a resistant ion channel with a ring of up to five serines replacing the five alanines in the wild-type Rdl ion channel, which would greatly increase the polarity of this region (5-CH20H replacing 5-CH3) and considerably alter its affinity for the various toxicants under discussion; even one or two added hydroxyl groups introduced here in a heteromultimer might have a significant effect and also reduce the energy barriers for ion permeation by participating in hydrogen bonding with water (Leonard et al., 1988). Analogies with the nACh receptor are evident and the closed configuration of the mutated chloride channel may remain somewhat ion permeable (Zhang et al., 1994; Revah et al., 1991).

55.4.3 MOLECULAR TOXICOLOGY OF NONCOMPETITIVE CHLORIDE IONOPHORE BLOCKERS

The localization of Ala302 to the PTXlcyclodiene binding site in Rdl prompts some further consideration of the information on the structure-toxicity relationships outlined in earlier sections. The cylinder formed by the amino-acid sequences Leu 303 , Ala302 , Va1 301 in five adjacent M2 domains (helices) provides a quasi-centrosymmetricallipophilic pocket into which cyclodiene insecticides and other noncompetitive chloride ionophore blockers (NCBs) might fit. The molecular dimensions of cyclodienes, taking into account the known range of allowable molecular substituents in these rather compact molecules, are sufficient to block a 5.6-A pore. If the pore is more or less centrosymmetrical, a cyclodiene molecule could fill the pocket and interact with all of its walls simultaneously because similar binding sites are presented around the lumen even if the arrangement is not a homomultimer. This would explain the observed toxicity of both enantiomeric forms of asymmetric molecules such as heptachlor epoxide; the forms may differ somewhat in toxicity, however, if the channel is not completely symmetrical, as is found (Miyazaki et al., 1980). Symmetrical molecules should be particularly effective, because they may be able to offer a symmetrically distributed electronegative center (or centers) to similar binding sites on opposite sides of the channel, as in the case of delta-chlordane and isobenzan, each having two symmetrically substituted chlorines on their fivemembered rings. These molecules may be viewed as symmetrical composites of the enantiomeric forms of alpha-DH and heptachlor epoxide (HE160), respectively, as suggested in Section 55.3.1 In a hypothetical model (Fig. 55.13) in which the HCNB moiety of cyclodienes is presumed bound at the synaptic end of the lipophilic pocket so that its gem-dichloro-bridge is presented to the channel wall, then the second electronegative center in, for example, dieldrin, is directed toward the cytoplasmic end of the pocket with its epoxide ring and unchlorinated methano-bridge fixed in an inward direction toward the channel lumen. This "cytoplasmic" end of the molecule is then close to the critical ring of Ala302 -methyl groups around the channellumen, which when replaced by -CH20H groups inhibits the binding of NCBs. Dieldrin and alpha-endosulfan (extended molecules) appear to provide the limiting acceptable molecular "lengths," as noted earlier, whereas isobenzan, endrin, and betaendosulfan are more compact. On this model, it might be argued also that the anti-1 O-chlorine atom (Fig. 55.9 and Table 55.4) of the dichloromethano-bridge of cyclodienes is better accommodated in the lipophilic pocket than the syn-10-chlorine, which may interfere sterically with the large side-chain alkyl groups of the ring of Leu 303 s that lie at the synaptic end of the pocket, making this syn-chlorine universally unfavorable for toxicity. The same argument might explain the four-fold increase in dieldrin toxicity effected by removal of the ethylenic chlorines (in DD, Table 55.4), because these chlorines might also interfere sterically in this region. This increase in toxicity is not uni-


,

CYTOPLASI\

-=-.

I I:

- 5-'

-

\ §..=

-'

•

\ ... 5

",

,,= H 1,\-

.........:;.--H,c ,-

SYNAPSF

Figure 55.13 Dieldrin (A) is oriented in a hypothetical binding site in or near the chloride ion channel lumen (in the region of Leu 303 ?) with chlorine X (I O-anti-chlorine) located in a subsite P that accommodates a bulky substituent. Its epoxide ring then penetrates a three-dimensional region S near Ala302 , which may interact, especially in the closed channel configuration, with the electronegative moieties of various cyclodienes when similarly oriented. If, however, Y (the IO-syn-chlorine) is presented to P, then the epoxide ring cannot so readily interact with zone S (dieldrin orientation B). C indicates the approximate position of the epoxide ring of endrin and also of the sulfur of beta-endosulfan, when either X or Y in these molecules is bound to subsite P; D is the approximate position of the endrin methano-bridge when its IO-antichlorine is located in P, corresponding to dieldrin orientation A. In this model, the dotted cylinder is the region containing a ring of leucine side chains. Note that a substituted benzene ring and some other extensions are permissible in the arrowed directions when the bridging system is unchlorinated (see Fig. 55. IO and related discussion, Section 55.4.3).

versal for cyclodienes, however, and the syn-chlorine atom is still present in DD, so that its adverse effect on toxicity is more than offset by removal of the ethylenic chlorine atoms. Further reductive replacement of the syn-10-chlorine atom in DD to give the trichloro-derivative DSD (24, Fig. 55.7) reduces toxicity to the level of dieldrin again (Brooks and Mace, 1987). The syn-lO-chloro- isomer of DSD is significantly less toxic than DSD or dieldrin, again indicating the greater importance of the 10-anti-chlorine for toxicity. The difference in toxicities between syn- and anti-l O-monodechloro-isomers is less marked for endrin and beta-endosulfan (the endrin-like isomer) but remains evident for alpha-endosulfan. This observation was discussed (Brooks, 1992) in connection with the insect cross-resistance spectrum for lindane/cyclodienes first noted by Busvine (1964), in which lindane, isobenzan, endrin, and the endosulfan isomers retain measurable toxicity to dieldrin-resistant insects (Brooks and Harrison, 1964a; Busvine, 1964). The first three molecules and beta-endosulfan are rather compact compared with dieldrin and alpha-endosulfan; the latter has been considered to be extended

1149

and dieldrin-like (Fig. 55.6), but recent structural studies indicate a more complex situation (see later). If the 10-anti-chlorine of dieldrin (X in Fig. 55.13) corresponds to the bulky anti-substituent found in PTX and must be presented to an appropriate lipophilic pocket in a binding subsite (P, Fig. 55.13) so as to place the epoxide ring in a correct position (in the region of S) with respect to the remainder of the binding site, then the lO-syn-chlorine (Y in Fig. 55.13), if similarly presented, cannot place the epoxide ring in the same position. If this latter position is modified in resistance to prevent interaction with the epoxide ring, dieldrin can no longer bind; however, either bridge-chlorine of endrin or beta-endosulfan can be offered to the bulky substituent binding subsite (P) such that the epoxide ring or sulfite moiety will still be placed in approximately their original positions (near to C), still able to interact with the critical site (S), on account of the more compact "cage" shape of these molecules. Consequently, these molecules may still be able to interact to some extent with the binding site that has been modified to exclude binding with dieldrin. On this model, alpha-endosulfan is anomalous in retaining effectiveness, because the same arguments apply as to dieldrin yet this molecule was actually somewhat more toxic than beta-endosulfan to dieldrin-resistant houseflies (Brooks and Harrison, 1964a). Two further observations may be significant, however. First, endosulfan sulfate (4, Fig. 55.6) formed in vivo from both endosulfan isomers may be the critical toxicant; it has the same structural (cage) configuration as beta-endosulfan (Forman et aI., 1965) and is formed faster from alpha- than from beta-endosulfan in some living organisms. Second, alpha-endosulfan has recently been reported (Schmidt et aI., 1997) to exist in the asymmetrical "twist-chair" conformation, in which the c-o bonds are "trans," not parallel as usually depicted (Fig. 55.6). Molecular models suggest that this twisted configuration may be more flexible, allowing the S=O moiety to occupy several spatial positions between the extremes represented by beta-endosulfan and the extended alpha-structure shown in Fig. 55.6. Consequently, the alphaisomer might be expected to be intrinsically at least as effective as the beta-isomer in terms of their interactions with the resistance-modified binding site, regardless of possible oxidation to the sulfate in vivo. In the case of endrin, the 9-keto(l2-keto-) metabolite is presumed to be the ultimate toxicant in mammals (Hutson et aI., 1975) and may contribute to endrin toxicity in insects (Kadous and Matsumura, 1982); notably, this oxidation places a second, additional, electronegative center at D, near to the upper sub site S (Fig. 55.13), which may improve binding potency toward the resistance-modified binding site. Lindane resembles a very compact cyclodiene and might bind without conflict with a subsite modified for dieldrin resistance or in more than one orientation, and similar arguments apply to isobenzan. Interestingly, isobenzan may be regarded as a "composite" of the HEl60 enantiomers, in which an "in plane" oxygen replaces the epoxide rings. Dieldrin resistance normally confers total resistance to HE160 (Brooks and Harrison, 1964a; Busvine, 1964) so the more compact placement of oxygen in isobenzan, combined with a possible increase in binding affinity

1150

CHAPTER 55


associated with the symmetrical chlorine substituents, appears to overcome both dieldrin and HE160 resistance to some extent. Notably, diaza-aldrin (3, Table 55.4), in which the unchlorinated double bond is replaced by -N=N- and which is probably converted into its N-oxide (4, Table 55.4) in vivo, is much more toxic than dieldrin to some insects (Busvine, 1964; Soloway, 1965) but dieldrin-resistant insects are immune to it. Using the TB PS binding assay in rat brain membranes, it has recently been confirmed that this molecule inhibits the binding competitively and therefore interacts directly with the PTX binding site (Ozoe et aI., 1995). Other modifications of the dichloroethylene moiety of dieldrin are acceptable to the binding site, as in photodieldrin (converted in vivo into Klein's ketone, however; Fig. 55.3) and the ketone analogs (14, 15, 16, Fig. 55.7) derived from isobenzan. From the analogy with PTX, the dichloroethylene moiety, unchlorinated double bond, and ketone derivatives in these various analogs may correspond to the lactone system of PTX. In lindane, the best superimposition with PTXlcyclodienes is apparently that in which two of the axial chlorines substitute for the lactone ring of PTX and the third axial chlorine provides the equivalent of the bulky trans-isopropenyl group of PTX or anti-l0-chlorine of cyclodienes. This last configuration of lindane is favored electrostatically (Calder et aI., 1993), although in it the bridge-end chlorines, which complete the pentagonal arrangement of chlorine atoms seen in cyclodienes (Brooks and Mace, 1987), are replaced by hydrogens. Replacement of the bridge-end chlorines by hydrogen reduces the toxicity of some cyclodienes (Soloway, 1965) and, notably, whether bulky alkyl groups in the alpha- and beta-positions of gammabutyrolactones (cf. PTX) stabilize the open or closed states of the ionophore and hence induce anticonvulsant or convulsant activity depends markedly on the stereochemistry of these substituents relative to the carbonyl group (Holland et aI., 1995; Klunk et aI., 1983; Peterson et aI., 1994). However, the replacement of one bridge-end carbon atom of dieldrin by nitrogen (available for hydrogen bonding) gives azadieldrin (25, Fig. 55.11) without great loss in toxicity compared with dieldrin (Gladstone and Wong, 1977), another indication that some modification in this region of cyclodienes is possible. The preceding discussion indicates that numerous modifications of the HCNB moiety retain binding capacity to the critical site. In this context, it may be noted that BIDN (34, Fig. 55.11), in which the simple norbomene nucleus carries two strongly electron-withdrawing substituents (gem-diCN and gem-di-CF3), is highly toxic to both mammals and insects (K61bl et aI., 1981). The allowable length compatible with toxicity of the fully chlorinated cyclodienes appears to be restricted, however (see earlier). TBPS has approximately the same molecular length as cyclodienes such as dieldrin, but there is the question of how the binding of cyclodienes to the critical site relates to that of the extended unchlorinated molecule EBOB or similar molecules in which an aromatic substituent replaces the bulky 4-alkyl group (Palmer et aI., 1991a). Superimposition of cyclodienes, aryl-TBOs, aryldithianes, arylsilatranes (35, Fig. 55.11), and PTX by CoMFA (comparative

molecular field analysis) (Calder et aI., 1993) supports the view that all act at the same or overlapping sites. Interesting additional information is available from work on the molecular hybrids of dechlorinated alpha-endosulfan (Fig. 55.10) (Ozoe et aI., 1993) and insecticidal dioxans (Palmer and Casida, 1995) mentioned previously (Section 55.3.2.2). In this series (Ozoe et aI., 1993), the fully chlorinated molecule (6, Fig. 55.1 0) cannot be extended but has the same housefly toxicity as the unchlorinated extended molecule (5, Fig. 55.10) related to the compounds reported by Casida. The latter molecule cannot be chlorinated; a single bridge chlorine abolishes its housefly toxicity. Assuming that the hybrid molecule (5) is a rigid form of the corresponding dioxan in which the unchlorinated norbomene moiety serves as the bulky substituent (e.g., t -butyl) (Ozoe et aI., 1990) and occupies a spatial region equivalent to that occupied by the HCNB moiety of cyclodienes, it is evident that the combination of substituted phenyl, dioxan, and norbomene moieties can bind to the receptor and afford potentially excellent insect toxicity, as found in 5-t-butyl-2-(4-ethynylphenyl)-1,3-dioxane (7, Fig. 55.10). However, chlorination may disrupt this binding in the corresponding deschoro-alpha-endosulfan analog with a 4-cyanophenyl substituent by forcing the "extended" molecule into a position it cannot occupy in the ion channel for steric reasons. Thus, the compact unchlorinated norbomene moiety may not by itself have sufficient binding potency in the lipophilic pocket bounded by the Leu 301 , Ala302 , and Va1 303 rings, but an added aromatic ring with an appropriate 4-substituent (particularly ethynyl), which, according to this model, binds additionally in a region of the channel pore beyond Ala302 and near to Va1 301 and Arg 300, may greatly reinforce the interaction. Conversely, the bulky, fully chlorinated cyclodiene interacts well with the closed configuration of the spherical pocket, but extension is impossible in this case because the added benzene ring in the extended HCNB moiety would be forced into steric hindrance with the valine isopropyl groups and the narrow cytoplasmic end of the channel; this contains the large guanidiny1 side chains of Arg 300 and might admit only "sticklike" structures such as the ethynyl moiety. Support is given to this hypothesis by the application (Akamatsu et aI., 1997) of CoMFA analysis to compare alphaendosulfan analogs (their series 2) with the hybrid extended norbomene derivatives (series 1) (Ozoe et aI., 1993; Fig. 55.10) analogous to the 1,3-dioxanes reported by Cas id a (Fig. 55.10). When the simple cyclodienes were closely superimposed on the extended molecules with a 2-(4-cyanophenyl) substituent, the housefly toxicity of some of the extended molecules was not well predicted until, in the superposition, the extended molecules were rotated 15° clockwise about their common bond on the norbomene ring junction (C4a-C8a in Table 55.4). This rotation enabled the separate correlation equations for the two series derived by CoMFA on the basis of close superimposition to be combined satisfactorily into a single equation representing the housefly toxicities of both series. The two series compared could not, however, be brought together in the same way when the measure of biological activity was the dis-


placement of eSS]-TBPS binding from rat brain membranes. In compound 5 (Fig. 55.10), the previous rotation turns the aromatic ring toward the center of the ion channel as modeled in Fig. 55.13, away from a region sterically forbidden according to CoMFA, and incidentally may "rock" the unchlorinated norbornene moiety into closer contact with the lipophilic pocket than it can achieve when chlorinated. Conversely, chlorination of the norbornene moiety of these extended molecules would have the reverse effect, forcing the aromatic ring toward the channel wall, into sterically forbidden space. The extending group in TBOs need not necessarily be aromatic, because short alkyl chains with a terminal ethynyl group give insecticidal activity, especially when synergized (Smith et aI., 1993) as do 4-ethyny1cyclohexyl groups in TBOs and dithianes (Weston et aI., 1995). A further interesting feature of the dithianes is their conversion into sulfoxides and sulfones, which doubtless occurs in vivo; the equatorial (trans- with respect to t-butyl; linear) 5-t-butyl-2-(4-Br-phenyl)-1,3-dithiane (8, Fig. 55.10; two-substituent is equatorial 4-Br-Ph) is twofold more toxic than the eis- (2-axial; angular) isomer, and the corresponding 4-ethynylphenyl-isomers have equal toxicity when synergized; for each isomer, conversion into the isomeric monosulfoxides and monosulfone progressively increases toxicity (Palmer and Casida, 1992). The not dissimilar toxicities of the trans- (linear) and eis- (angular) dithianes is intriguing. If, in the pore model (Fig. 55.13), the bulky t-butyl substituent is placed in the lipophilic pocket as for the other bulky substituents discussed previously, then the sulfoxides and sulfones from the isomeric dithianes occupy rather different positions in this lipophilic site; only in the linear isomer do these moieties occupy positions near the two ester oxygens of alpha-endosulfan and the epoxide ring of dieldrin, toward the cytoplasmic end of the channel and near to Ala302 . For the angular isomer, the SO (or S02) moieties are placed at the synaptic end of the lipophilic pocket and lie toward the channel center, in this case, the aromatic ring overlies the trioxabicyclooctane ring of TBOs and its 4-substituent reaches only the aromatic ring of the linear isomer, so that the position of this part of the angular molecule is foreshortened in the cytoplasmic direction relative to the linear one. All of the molecules discussed can be superimposed, and with the cyclodiene molecules oriented in this way, the sulfite moiety of beta-endosulfan (and the S02 of endosulfan sulfate) lies in the region occupied by the aromatic ring of the angular dithiane isomer. It should be noted, however, that the possibility of eis- to trans-rearrangement exists for the dithianes in vivo (Pulman et aI., 1996), in which case, the eis-isomers would merely be precursors of the linear (trans-) molecules, which are more readily accommodated in the model. Among the heterocyclic compounds of recent interest that are believed to act by blocking the GABA-gated chroride ionophore, the experimental spirosultam LY 219048 (33, Fig. 55.11) contains an obvious lipophilic bulky substituent in the form of the cyclohexane ring, in analogy with the compounds discussed previously. Probably because this ring may be susceptible to metabolic attack, the compound is not very toxic

1151

to insects but its toxicity to mice is similar to that of endrin (Bloomquist et aI., 1993). Other insecticidal compounds such as the phenylpyrazoles (e.g., fipronil; 27, Fig. 55.11) have little obvious structural similarity to the compounds discussed in previous sections yet inhibit eH]-EBOB binding to housefly head membranes and are believed to act at the PCCAlPTX binding site (Cole et aI., 1993). Some other compounds containing the common 2,6-dichloro-4-trifluoromethylphenyl moiety combined with a pyrimidinone (e.g., 32, Fig. 55.11) and other small heterocyclic rings were described by Whittle et al. (1995) but these lacked the broad-spectrum insecticidal activity shown by fipronil, which was introduced in 1993 (Colliot et aI., 1992) and now has a wide range of applications for crop protection by foliar, soil or seed treatment. Its mammalian toxicity is generally moderate (rat acute oral LDso, 97 mg/kg; mouse acute oral LDso, 95 mg/kg) and it is readily converted into degradation products, including the corresponding sulfone, in the environment (Tomlin, 1997). The substitution pattern in the phenyl and heterocyclic rings of these compounds places these ring planes at right angles (Whittle et aI., 1995) so that the molecules are "space filling" in these two planes and are relatively rigid, with the pyrazole ring skewed with respect to the benzene ring because of the pyramidal linking nitrogen. The substituents (-NH2, -CN, -SOCF3) are attached to double bonds and are fixed in the plane of the pyrazole ring, although they can rotate about their attaching bonds, which may be particularly important for the critical -SOCF3 group. This group confers insect (housefly LDso, 0.3 J.Lg/g) and vertebrate (mouse ip LDso, 30 mg/kg) toxicity, even in the analogous molecule lacking both the - NH2 and -eN substituents, which, however, clearly optimize the binding properties of fipronil. In these molecules, the sulfur atom requires bioactivation through sulfoxide/sulfone formation to confer toxicity. How do the phenylpyrazoles bind to the chloride ionophore in relation to the other molecules discussed previously? They inhibit EBOB binding in housefly head membranes noncompetitively, which may involve irreversible or slowly reversible inhibition or action at an allosteric but coupled site (Cole et a!., 1993). Whittle et al. (1995) explored a series of compounds, including pyrimidinone 32 (Fig. 55.11) on the basis of their commonality of dipole direction with phenylpyrazoles; the positive end of the dipole lies toward the benzene ring in active compounds. If this ring, with its strongly electron-withdrawing -CF3 moiety, is placed in the channel pore binding site so that the -CF3 is in a similar position to that occupied by the 4-ethynyl group of EBOB, according to the model discussed earlier, that is, in the region of Arg 300 , then the benzene ring with its chlorine atoms and the pyrazole ring with its substituents are well placed to bind with channel components in this quasi-centrosymmetrical ionophore and, notably, the -SOCF3 moiety is then placed near to Ala302 in a region occupied by the endosulfan ester oxygens, the oxygens of TBOs, the oxygens or sulfurs of dioxans and oxidized trans- (linear) dithianes, and the second electronegative center (e.g., the epox-

1152

CHAPTER 55

Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocyc1oalkanes and Recent Congeners

ide ring of dieldrin) of cyclodienes, when these molecules are located in the lipophilic pocket bounded by Leu 303 . These molecules might penetrate the chloride ionophore from either the extracellular (synaptic) mouth or the cytoplasmic mouth or by first penetrating the lipid bilayer and then entering the channel laterally. Current understanding of the nACh-receptor (nAChR) ionophore, which has analogies with the GABAA ionophore, may be relevant. According to Unwin (1995), the M2 cylinders (alpha-helices) (Fig. 55.12) are kinked inward in the region of a ring of leucines when the ionophore is closed; in the open configuration, the cylinders are twisted laterally, moving the leucine side chains away from the pore. By analogy, a cyclodiene (for example) may enter a similar open conformation in the chloride ionophore and then become tightly bound when the pore reverts to the closed conformation, as suggested by ffrench-Constant et al. (1995). Nakanishi et al. (1997) provide an interesting discussion on the question of mode of channel entry, based on research on philanthotoxins (PhTXs) binding to the nAChR. In particular, they noted that an n-butyl side chain introduced into the hydrophilic polyamine chain of PhTX-433 increases potency eight-fold. The n-butyl moiety is so placed as to reach the area of Leu 251 when the molecule is inserted linearly into the AChR ionophore from the cytoplasmic end, and the authors speculate that the increased potency might result from additional hydrophobic binding between the n-butyl side chain and the alkyl groups of Leu 251 ; potency decreases below that of the parent PhTX-433 when this side chain is placed in other positions along the polyamine chain. Hydrophobic binding must also have a significant role in the interaction of the cage convulsants with the chloride ionophore. Here we recall that the hybrid (unchlorinated) molecule 4 (Fig. 55.10) combines the bulky norbornene (hydrophobic binding) and dioxan moieties with the additional binding capacity evidently conferred by the benzene ring with its electronegative 4-ethynyl substituent; it is more toxic (when oxidative metabolism is inhibited) than many fully chlorinated cyclodienes. No doubt work with the affinity probes currently under development (Casida and Pulman, 1994) will resolve some of these questions regarding the exact location of binding sub sites in the chloride ionophore.

55.5 CONCLUSIONS Some 45 years after serious toxicological research on the PCCA insecticides began, there is now a broad understanding of their mode of action and, due to rapid progress in the application of molecular biology, the tantalizing mechanism of insect resistance to them has been illuminated. Furthermore, new structural classes of chemicals have now emerged that appear to act in the same way. Although many chemicals found to interact with the picrotoxinin binding site in the GABAA -receptor chloride ionophore are highly toxic to both mammals and insects, binding studies with radioligands have indicated differences between their GABA-receptors that offer the prospect of selective insect toxicity involving this target.

The commercially successful insecticide fipronil appears to fulfill these expectations and other new chemicals are likely to follow. There is also the possibility of "building in" selectivity by using the "propesticide" approach, which exploits differences between insects and nontarget organisms in their biotransformation routes for chemically derivatized toxic ants and has been applied successfully to alleviate the mammalian toxicity of other classes of insecticides. Meanwhile, many questions remain to be answered that are of fundamental importance in understanding the molecular action of neurotransmitters and insecticides on ion channels. The intense interest in this subject, stimulated by rapid advances in molecular biology, will ensure a prominent use for the PCCA insecticides and the newer chemicals with related actions, as tools in these explorations.

REFERENCES Abalis, I. M., Eldefrawi, M. E., and Eldefrawi, A. T. (1985). High affinity stereospecific binding of cyclodiene insecticides and gammahexachlorocyclohexane to gamma-aminobutyric acid receptors. Pestic. Biochem. Physiol. 24,95-102. Akamatsu, M., Ozoe, Y., Fujita, T, Mochida, K, Nakamura, T, and Matsumura, F. (1997). Sites of action of noncompetitive GABA antagonists in houseflies and rats: Three dimensional QSAR analysis. Pestic. Sci. 49, 319-332. Akkermans, L. M. A., van den Bercken, 1., van der Zalm, 1. M., and van Stranten, H. W. M. (1974). Effects of dieldril and some of its metabolite on synaptic transmission in the frog motor end plate. Pestic. Biochem. Physiol. 4,313-324. Akkermans, L. M. A., van den Bercken, 1., and van der Zalm, 1. M. (l975a). Effects of aldrin-transdiol on neuromuscular facilitation and depression. Eur. J. Pharmacol. 31, 166-175. Akkermans, L. M. A., van den Bercken, 1., and Versluijs-Helder, M. (l975b). Excitatory and depressant effects of dieldrin and aldrin-transdiol on the spinal cord of the toad (Xenopus laevis). Eur. J. Pharmacol. 34,133-142. Balba, H. M., and Saha, J. G. (1978). Studies on the distribution, excretion and metabolism of alpha- and gamma-isomers of 4 C] chlordane in rabbits. J. Environ. Sci. Health, B13, 211-233. Baldwin, M. K (1971). "The Metabolism of the Chlorinated Insecticides Aldrin, Dieldrin, Endrin and Isodrin." Ph.D. Thesis, University of Surrey, Guildford, UK Baldwin, M. K, and Robinson, 1. (1969). Metabolism in the rat of the photoisomerisation product of dieldrin. Nature (London) 224, 283-284. Baldwin, M. K., Robinson, 1., and Parke, D. V. (1972). A comparison of the metabolism of HE OD (Dieldrin) in the CFI mouse with that in the CFE rat. Food Cosmet. Toxicol. 10,333-351. Barnes, w., and Ware, G. W. (1965). The absorption and metabolism of C 14 _ endosulfan in the housefly. 1. Eeon. Entomo!" 58, 286-291. Bedford, C. T. (1974). Von BaeyerlIUPAC names and abbreviated chemical names of metabolites and artifacts of aldrin (HHDN), dieldrin (HEOD) and endrin. Pestie. Sei. 5,473-489. Bedford, C. T, Harrod, R. K, Hoadley, E. c., and Hutson, D. H. (l975a). The metabolic fate of endrin in the rabbit. Xenobiotiea, 5,485-500. Bedford, C. T, Hutson, D. A., and Natoff, I. L. (l975b). The acute toxicity of endrin and its metabolites to rats. Toxieol. Appl. Pharmacol. 33, 115-121. Bender, H. (1935). Benzene hexachloride and other chlorinated carbocyclics. U.S. Patent 2010841. Bloomquist, 1. R., and Shankland, D. L. (1983). The mode of action and neurotoxicity of mirex, chlordecone and four hydrogenated mirex analogues. Pestic. Bioehem. Physiol. 19, 235-242.

e

References

Bloomquist, J. R., Jackson, J. L., Karr, L. L., Ferguson, H. J., and Gajewski, R. P. (1993). Spirosultam LY 219048: A new chemical class of insecticide acting upon the GABA receptor/chloride ionophore complex. Pes tic. Sci. 39,195-202. Boddy, K. L., Briggs, G. G., Harrison, R. P., Jones, T. H., O'Mahony, M. J., Marlow, I. D., Roberts, B. G., Bardsley, R., and Reid, J. (1996). The synthesis and insecticidal activity of2-aryl-1,2,3-triazoles. Pestic. Sci. 48,189196. Brimfield, A. A., and Street, J. C. (1979). Mammalian biotransformation of chlordane: in vivo and primary hepatic comparisons. Ann. N.y. Acad. Sci. 320, 247-256. Brimfield, A. A., Street, J. c., Futsell, J., and Chatfield, D. A. (1978). Identification of products arising from the metabolism of cis- and trans-chlordane in rat liver microsomes in vitro: Outline of possible metabolic pathway. Pestic. Biochem. Physiol. 9,84-95. Brooks, G. T. (1960). Mechanisms of resistance of the adult housefly (M. domestica) to cyclodiene insecticides. Nature (London) 186, 96-98. Brooks, G. T. (1966). Progress in metabolic studies of the cyclodiene insecticides and its relevance to structure-activity correlations. World Rev. of Pest ControlS, 62-84. Brooks, G. T. (1969). The metabolism of diene-organochlorine (cyclodiene) insecticides, Residue Rev. 27, 81-138. Brooks, G. T. (1973). The design of insecticidal chlorohydrocarbon derivatives. In "Drug Design" (E. J. Ariens, ed.), Vol. 4, pp. 379-444. Academic Press, New York. Brooks, G. T. (1974). "Chlorinated Insecticides," Vol. I, pp. 87-98. CRC Press, Cleveland. Brooks, G. T. (1975). The insect toxicities of biodegradable derivatives of chlorinated norbomenes. In "Proceedings of the Eight British Insecticide and Fungicide Conference," Vol. 2, pp. 381-387. British Crop Protection Council, Famham, Surrey, UK. Brooks, G. T. (1977). Action and inaction of certain non-anticholinesterase insecticides. In "Proceedings of the 1977 British Crop Protection Conference-Pests and Diseases," Vol. 3, pp. 731-740. British Crop Protection Council, Farnham, Surrey, UK. Brooks, G. T. (1980). The preparation of some reductively dechlorinated analogues of dieldrin, endosulfan and isobenzan. 1. Pestic. Sci. (Tokyo) 5, 565574. Brooks, G. T. (1985). The preparation of reductively dechlorinated analogues of endrin and some other cyclodiene insecticides. 1. Pestic. Sci. (Tokyo) 10, 241-245. Brooks, G. T. (1992). Progress in structure-activity studies on cage convulsants and related GABA receptor chloride ionophore antagonists. In "Insecticides: Mechanism of Action and Resistance" (D. OUo and B. Weber, eds.), pp. 237-242. Intercept, Andover, UK. Brooks, G. T., and Harrison, A. (1963). Relations between structure, metabolism and toxicity of the "cyclodiene" insecticides. Nature (London) 198,1169-1171. Brooks, G. T., and Harrison, A. (1964a). The effect of pyrethrin synergists, especially sesamex, on the insecticidal potency of hexachlorocyclopentadiene derivatives ("cyclodiene" insecticides) in the adult housefly, M. domestica L. Biochem. Pharmacal. 13, 827-840. Brooks, G. T., and Harrison, A. (1964b). The metabolism of some cyclodiene insecticides in relation to dieldrin resistance in the adult housefly, M. domestica L. 1. Insect Physiol. 10,633-641. Brooks, G. T., and Harrison, A. (1965). Structure-activity relationships among insecticidal compounds derived from chlordene. Nature (London) 205, 1031-1032. Brooks, G. T., and Harrison, A. (1967a). The metabolism of dihydrochlordene and related compounds by housefly (M. domestica L.) and pig liver microsomes. Life Sci. 6, 681-689. Brooks, G. T., and Harrison, A. (1967b). The toxicity of alphadihydroheptachlor and related compounds to the housefly (M. domestica L.) and their metabolism by housefly and pig liver microsomes. Life Sci. 6, 1439-1448.

1153

Brooks, G. T., and Harrison, A. (1969a). The oxidative metabolism of aldrin and dihydroaldrin by houseflies, housefly microsomes and pig liver microsomes and the effect inhibitors. Biochem. Pharmacal. 18, 557-568. Brooks, G. T., and Harrison, A. (1969b). Hydration of HEOD (dieldrin) and the heptachlor epoxides by microsomes from the livers of pigs and rabbits. Bull. Environ. Contam. Toxicol. 4,352-361. Brooks, G. T., and Mace, D. W. (1987). Toxicity and mode of action of reductively dechlorinated cyclodiene insecticide analogues on houseflies (M. domestica L.) and other diptera. Pestic. Sci. 21, 129-142. Brooks, G. T., Barlow, F., Hadaway, A. B., and Harris, A. G. (1981). The toxicities of some analogues of dieldrin, endosulfan and isobenzan to bloodsucking diptera, especially tsetse flies. Pestic. Sci. 12, 475-484. Brooks, G. T., Harrison, A., and Lewis, S. E. (1970). Cyclodiene epoxide ring hydration by microsomes from mammalian liver and houseflies. Biochem. Pharmacal. 19, 255-273. Brooks, G. T., Lewis, S. E., and Harrison, A. (1968). Selective metabolism of cyclodiene enantiomers by pig liver microsomal enzymes. Nature (London) 220, 1034-1035. Buchel, K. H., Ginsberg, A. E., and Fischer, R. (1966a). Synthese and struktur von heptachlor-methano-tetrahydrindanen. Chem. Ber. 99,405-415. Buchel, K. H., Ginsberg, A. E., and Fischer, R. (1966b). Synthese und struktur von isomeren des chlordans. Chem. Ber. 99,421-430. Burt, P. E. (1973). "Mode of Action of Insecticides." Rothamsted Experimental Station Annual Report, p. 171. Busvine, J. R. (1964). The insecticidal potency of gamma-BHC and the chlorinated cyclodiene compounds and the significance of resistance to them. Bull. Entomol. Res. 55, 271-288. Calder, J. A., Wyatt, J. A., Frenkel, D. A., and Casida, J. E. (1993). CoMFA validation of the superposition of six classes of compounds which block GABA receptors non-competitively. 1. Comput.-Aided Mol. Des. 7, 45-60. Casida,1. E., and Pulman, D. A. (1994). Recent advances in heterocyclic insecticides acting as GABA antagonists. In "Advances in the Chemistry of Insect Control m" (G. G. Briggs, ed.), pp. 36-51. Royal Society of Chemistry, Cambridge. Casida, J. E., Palmer, C. J., and Cole, L. (1985). Bicycloorthocarboxylate convulsants: Potent GABAA receptor antagonists. Mol. Pharmacal. 28, 246253. Casida, J. E., Nicholson, R. A., and Palmer, C. J. (1988). Trioxabicyclooctanes as probes for the convulsant site of the GABA-gated chloride channel in mammals and arthropods. In "Neurotox 88, Molecular basis of drug and pesticide action" (G. G. Lunt, ed.), Elsevier Science Publishers, Amsterdam, 125-144. Chandurkar, P. S., and Matsumura, F. (1979). Metabolism of toxaphene components in rats. Arch. Environ. Contam. Toxicol. 8, 1-24. Chipman, J. K., and Walker, C. H. (1979). The metabolism of dieldrin and two of its analogues: the relationship between rates of microsomal metabolism and rates of excretion of metabolites in the male rat. Biochem. Pharmacal. 28, 1337-1345. Chu, I., Villeneuve, D. c., Secours, v., Becking, G. C., Viau, A., and Benoit, F. (1979). The absorption, distribution and excretion of photomirex in the rat. Drug. Metab. Dispos. 7, 24-27. Cole, L. M., Nicholson, R. A., and Casida, J. E. (1993). Action ofphenylpyrazole insecticides at the GABA-gated chloride channel. Pestic. Biochem. Physiol. 46,47-54. Cole, L. M., Saleh, M. A. and Casida, J. E. (1994). House fly head GABAgated chloride channel: [3Hl-endosulfan binding in relation to polychlorocycloalkane insecticide action. Pestic. Sci. 42, 59-63. Colliot, F., Kukorowski, K. A., Hawkins, D. w., and Roberts, D. A. (1992). Fipronil: a new soil and foliar broad spectrum insecticide. British Crop Protection Conference-Pests and Diseases, British Crop Protection Council, Famham, UK I, 29-34. Craven, A. C. C., Brooks, G. T., and Walker, C. H. (1976). The inhibition of HEOM epoxide hydrase in mammalian liver microsomes and insect pupal homogenates. Pestic. Biochem. Physiol. 6, 132-141. Deng, Y., Palmer, C. J., and Casida, J. E. (1991). Housefly brain gammaaminobutyric acid-gated chloride channel: Target for multiple classes of insecticides. Pestic. Biochem. Physiol. 41, 60-65.

1154

CHAPTER 55


Deng, Y., Palmer, C. J., and Casida, J. E. (1993). Housefly head GABA-gated chloride channel: four putative binding sites differentiated by [3HJ EBOB and [35 SJ TBPS. Pestic. Biochem. Physiol. 47,98-112. Dorough, H. w., Huhtanen, K., Marshall, T. C. M., and Bryant, H. E. (1978). Fate of endosulfan in rats and toxicological considerations of apolar metabolites. Pestic. Biochem. Physiol. 8, 241-252. EIIiott et al. (1990). EIIiott, M., Pulman, D. A., Larkin, J. P., and Casida, J. E. (1992). Insecticidal 1,3-dithianes. 1. Agric. Food Chem. 40, 147-151. El Zorgani, G. A., Walker, C. H., and Hassall, K. A. (1970). Species differences in the in vivo metabolism of HEOM, a chlorinated cyclodiene epoxide. Life Sci. 9,415-420. Estabrook, R. W., Cooper, D. Y., and Rosenthal, o. (1963). The light-reversible carbon monoxide inhihition of the steroid C-21 hydroxylation system of the adrenal cortex. Biochem. Z. 338, 741-755. ffrench-Constant, R. H., Steichen, J. c., and Ode, P. J. (1993b). Cyclodiene insecticide resistance in Drosophila melanogaster (Meigen) is associated with a temperature-sensitive phenotype. Pestic. Biochem. Physiol. 46, 7377. ffrench-Constant, R. H., Steichen, J. C., Rocheleau, T. A., Aronstein, K., and Roush, R. T. (l993b). A single amino-acid substitution in a gamma-butyric acid subtype A receptor locus is associated with cyclodiene insecticide resistance in Drosophila populations. Proc. Natl. Acad. Sci. U.S.A.90, 19571961. ffrench-Constant, R. H., Rocheleau, T. A., Steichen, J. c., and Chalmers, A. E. (1993a). A point mutation in a Drosophila GABA receptor confers insecticide resistance. Nature (London) 363, 449-451. ffrench-Constant, R. H., Zhang, H.-G., and Jackson, M. B. (1995). Biophysical analysis of a single amino-acid replacement in the resistance to dieldrin gamma-aminobutyric acid receptor: Novel dual mechanism for cyclodiene insecticide resistance. In "Molecular Action of Insecticides on Ion Channels," ACS Symposium Series 591 (1. Marshall Clark, ed.), pp. 192-204. Am. Chem. Soc., Washington, DC. Fisher, M. H. (1997). Structure-activity relationships of the avermectins and milbemycins. In "Phytochemicals for Pest Control," ACS Symposium Series 658 (P. A. Hedin, R. M. HoIIingwortb, E. P. Masler, J. Miyamoto, and D. G. Thompson, eds.), Am. Chem. Soc., Washington, DC. pp. 220-238. Forman, S. E., Durbetaki, A. J., Cohen, M. v., and OIofson, R. A. (1965). Conformational Equilibria in Cyclic Sulfites and Sulfates: the Configurations and Conformations of the Two Isomeric Thiodans. 1. Org. Chem. 30, 169175. Gant, D., Eldefrawi, E, and Eldefrawi, A. T. (1987). Cyclodienes inhibit GABAA -receptor-regulated chloride transport. Toxicol. Appl. Pharmacol. 88,313-321. Giannotti, 0., Metcalf, R. L., and March, R. B. (1957). The mode of action of aldrin and dieldrin in Periplaneta americana (L.). Ann. Entomol. Soc. Am. 49, 588-592. Gladstone, C. M., and Wong, J. L. (1977). Azadiene chemistry 4. Insecticidal activities and chemical reactivities of azadieldrin and azaaldrin. Comparison with aldrin and dieldrin. 1. Agric. Food Chem. 25,489-493. Hallett, D. J., Khera, K. S., Stoltz, D. R., Chu, I., and ViIIeneuve, D. C. (1978). Photomirex: Synthesis and assessment of acute toxicity, tissue distribution and mutagenicity. 1. Agric. Food Chem. 26,288-291. Hathway, D. E., and Mallinson, A. (1964). Effect of telodrin on the liberation and utilisation of ammonia in rat brain. Biochem. 1. 90, 5 1-60. Hathway, D. E., MaIIinson, A., and Akintonwa, D. A. A. (1965). Effects of dieldrin, picrotoxin and telodrin on the metabolism of ammonia in brain. Biochem. 1. 94, 676-686. Hatton, L. R., Hawkins, D. w., Pearson, C. J., and Roberts, D. A. (1988). Derivatives of N-phenylpyrazoles. European Patent 295,117. Hawkinson, J. E., and Casida, J. E. (1993). Insecticide binding sites on gammaaminobutyric acid receptors of insects and mammals. In "Pest Control with Environmental Safety," ACS Symposium Series 524 (S. O. Duke, J. J. Menn, and J. R. Plimmer, eds.), pp. 126-143. Am. Chem. Soc., Washington, DC. Holland, K. D., Mathews, G. c., Bolos-Sy, A. M., Tucker, J. B., Reddy, P. A., Covey, D. E, Ferrendeli, J. A., and Rothman, S. M. (1995). Dual modula-

tion of the gamma-aminobutyric acid type A receptor/ionophore by alkylsubstituted gamma-butyrolactones. Mol. Pharmacol. 47, 1217-1223. Holyoke, C. W., Rauh, J. J., Kleier, D. A., Schnee, M. A., Cordova, D., Benner, E. A., Watson, M. K., Bai, D., Howard, M. H., and Sattelle, D. B. (1994). "Advances in the Chemistry of Insecticides." Royal Society of Chemistry, Cambridge. Hutson, D. H., Baldwin, M. K., and Hoadley, E. C. (1975). Detoxication and bioactivation of endrin in the rat. Xenobiotica 5, 697-714. Hyman, J. (1949). Improvements in, or related to method of forming halogenated organic compounds and the products resulting there from. British Patent 618432. Jager, K. W. (1970). "Aldrin, Dieldrin, Endrin and Telodrin: An Epidemiological and Toxicological Study of Long-Term Occupational Exposure." Elsevier, Amsterdam. Joy, R. M. (1982). Mode of action of lindane, dieldrin and related insecticides in the central nervous system. Neurobehav. Toxicol. Teratol. 4, 813-823. Kadous, A. A., and Matsumura, E (1982). Toxicity of metabolites of dieldrin, photodieldrin and endrin in the cockroach. Arch. Environ. Contam. Toxicol. 11, 635-643. Kadous, A. A., Ghiasuddin, S. M., Matumura, E, Scott, J. G., and Tanaka, K. (1983). Differences in the picrotoxin receptor between the cyclodiene resistant and susceptible strains of the German Cockroach. Pestic. Biochem. Physiol. 19, 157-166. Kearns, C. w., Ingle, L., and Metcalf, R. L. (1945). New chlorinated hydrocarbon insecticides. 1. Econ. Entomol. 38,661-668. Khan, M. A. Q., Sutherland, D. J., Rosen, J. D., and Carey, W. E (1970). Effect of sesamex on the toxicity and metabolism of cyclodienes and their photoisomers in the housefly. 1. Econ. Entomol. 63,470-475. K1ein, A. K., Dailey, R. E., Walton, M. S., Beck, v., and Link, J. D. (1970). Metabolites isolated from urine of rats fed 14C-photodieldrin. 1. Agric. Food Chem. 18,705-708. K1ingenberg, M. (1958). Pigments of rat liver microsomes. Arch. Biochem. Biophys. 75, 376-386. Klis, S. E L., Nijman, N. J., Vijverberg, H. P. M., and van den Bercken, J. (199 I). Pheny Ipyrazoles, a new class of pesticide: Effects on neuromuscular transmission and acetyl choline responses. Pestic. Sci. 33,213-222. K1unk et al. (1982). Klunk, W. E., Kalman, B. L., Ferrendelli, J. A., and Covey, D. E (1983). Computer assisted modelling of the picrotoxinin and gamma-butyrolactone receptor site. 23, 511-518. Kolbl, H., Gompper, Behrenz, W., Hammann, I., Homeyer, B., and Hermann, G. (1981). "Agents for Pest Control." DE 3141 119 AI, Bayer AG. Korte, E (1967). Metabolism of 14C-Iabelled insecticides in microorganisms, insects and mammals. Botyu kagaku 32, 46-59. Korte, E, and Arent, H. (1965). Metabolism of insecticides. IX. Isolation and identification of dieldrin metabolites from the urine of rabbits after oral administration of dieldrin- 14 C. Life Sci. 4,2017-2026. Kunze, E M., and Laug, E. P. (1953). Toxicants in tissues of rats on diets containing dieldrin, aldrin, endrin and isodrin. Fed. Proc. 12,339. Lauger, D., Martin, H., and Muller, P. (1944). Uber konstitution und toxische wirkung von naturlichen und neuen synthetischen insektentotenden stoffen. Helv. Chim. Acta 27, 892. Lawrence, J., and Casida, J. E. (1984). Interactions of lindane, toxaphene and cyclodienes with brain specific t-butylbicyclophosphorothionate receptor. Life Sci. 35, 171-178. Lee, H. J., Zhang, H.-G., Jackson, M. B., and ffrench-Constant, R. H. (1995). Binding and physiology of 4'-ethynyl-4-n-propylbicycloorthobenzoate (EBOB) in cyclodiene-resistant Drosophila. Pestic. Biochem. Physiol. 51, 30-37. Leonard, R. J., Labarca, C. G., Charnet, P., Davidson, N., and Lester, H. A. (1988). Evidence that the M2 membrane spanning region lines the ion channel pore of the nicotinic receptor. Science 242, 1578-1581. Lewis, S. E. (1967). Effect of carbon monoxide on metabolism of insecticides in vivo. Nature (London) 215, 1408-1409. Ludwig, G., Weis, J., and Korte, E (1964). Excretion and distribution of aldrin14C and its metabolites after oral administration for a long period of time. Life Sci. 3, 123-130.

References

MacDonald, R. L., and Olsen, R. W. (1994). GABAa-receptor channels. Ann. Rev. Neurosci. 17, 569-602. Maier-Bode, H. (1968). Properties, effect, residues and analytics of the insecticide endosulfan. Residue Rev. 22, 1-44. Matsumura, E, and Ghiasuddin, S. M. (1983). Evidence for similarities between cyclodiene type insecticides and picrotoxinin in their action mechanisms. J. Environ. Sci. Health, B 18, 1-14. Matthews, H. B., and Matsumura, E (1969). Metabolic fate of dieldrin in the rat. J. Agric. Food Chem. 17,845-852. Middleton, W J., and Bingham, E. M. (1982). Fluorine containing 1,1dicyanoethylenes: Their preparation, Diels-Alder reaction and derived norbomenes and norbomanes. J. Fluorine Chem. 20, 397-418. Miles, J. R. W, Tu, C. M., and Harris, C. R. (1969). Metabolism of heptachlor and its degradation products by soil microrganisms. J. Econ. Entomo!' 62, 1334-1338. Miyazaki, A., Hotta, T., Marumo, S., and Sakai, M. (1978). Synthesis and absolute stereochemistry and biological activity of optically active cyclodiene insecticides. J. Agric. Food Chem. 26,975-977. Miyazaki, A., Sakai, M., and Marumo, S. (1979). Comparative metabolism of enantiomers of chlordene and chlordene epoxide in German cockroaches, in relation to their remarkably different insecticidal activity. J. Agric. Food Chem. 27, 1403-1405. Miyazaki, A., Sakai, M., and Marumo, S. (1980). Synthesis and biological activity of optically active heptachlor, 2-chloroheptachlor and 3-chloroheptachlor. J. Agric. Food Chem. 28, 1310-1311. Molowa, D. T., Wrighton, S. A, Blank, R. v., and Guzelian, P. S. (1986). Characterization of a unique aldo-keto-reductase responsible for the reduction of chlordecone in the liver of gerbil and man. 1. Toxico!. Environ. Health 17, 375-384. Nakanishi, K, Huang, D., Monde, K, Tokiwa, Y., Fang, K, Liu, Y., Jiang, H., Huang, X., Matile, S., Usherwood, P. N. R., and Berova, N. (1997). Philanthotoxins and the nicotinic acetylcholine receptor. In "Phytochemicals for Pest Control," ACS Symposium Series 658 (P. A Hedin, R. M. Hollingworth, E. P. Masler, J. Miyamoto, and D. G. Thompson, eds.), pp. 339-353. Am. Chem. Soc., Washington, DC. Nakatsugawa, T., Ishida, M., and Dahm, P. A (1965). Microsomal epoxidation of cyclodiene insecticides. Biochem. Pharmacol. 14, 1853-1865. Nelson, J. 0., and Matsumura, E (1973). Dieldrin (HEOD) metabolism in cockroaches and houseflies. Arch. Environ. Contam. Toxico!. 1, 224-244. Owe, Y., and Matsumura, E (1986). Structural requirements for bridged bicyclic compounds acting on picrotoxinin receptor. J. Agric. Food Chem. 34, 126-134. Owe, Y., Matsumoto, K, Mochida, K, Nakamura, T., and Matsumura, E (1995). Nitrogen analogues of aldrin as non-competitive antagonists of GABAA -receptor. J. Pestic. Sci. (Tokyo) 20, 317-319. Owe, Y., Sawada, Y., Mochida, K, Nakamura, T., and Matsumura, E (1990). Structure-activity relationships in a new series of insecticidally active dioxatricycloalkenes derived by structural comparison of the GABA antagonists bicycloorthocarboxylates and endosulfan. J. Agric. Food Chem. 38, 12641268. Owe, Y., Takayama, T., Sawada, Y., Mochida, K, Nakamura, T., and Matsumura, E (1993). Synthesis and structure-activity relationships of a series of insecticidal dioxatricyclododecenes acting as the noncompetitive antagonists of GABAA receptors. 41, 2135-2141. Palmer, C. J., and Casida, J. E. (1985). 1,4-Disubstituted 2,6,7-trioxabicyclo [2.2.2] octanes: a new class of insecticides. J. Agric. Food Chem. 33, 967980. Palmer, C. J., and Casida, J. E. (1992). Insecticidal 1,3-dithianes and 1,3dithiane 1,I-dioxides. J. Agric. Food Chem. 40, 492-496. Palmer, C. J., and Casida, J. E. (1995). Insecticidal 1,3-oxathianes and their oxides. J. Agric. Food Chem. 43,498-502. Palmer, C. J., Cole, L. M., Larkin, J. P., Smith, I. H., and Casida, J. E. (1991a). (4-ethynylphenyl)-4-substituted -2,6,7-trioxabicyclo[2.2.2] octanes: Effect of 4-substituent on toxicity to houseflies and mice and potency at the GABA-gated chloride channel. J. Agric. Food Chem. 39, 1329-1334. Palmer, C. 1., Cole, L. M., Smith, I. H., Moss, M. D. v., and Casida, J. E. (1991 b). Silylated 1-(4-ethynylphenyl)-2,6, 7 -trioxabicyclo[2.2.1] octanes:

1155

Structural features and mechanisms of proinsecticidal action and selective toxicity. 1. Agric. Food Chem. 39, 1335-1341. Peterson, E. M., Kun Xu, Holland, K D., McKeon, A c., Rothman, S. M., Ferrendelli, J. A., and Covey, D. E (1994). Alpha-spirocyclopentyl and alpha-spirocyclopropyl-gamma-butyrolactones: Conformationally constrained derivatives of anticonvulsant and convulsant alpha, alphadisubstituted gamma-butyrolactones. J. Med. Chem. 37, 275-286. Pollock, G. A., and Kilgore, W. W. (1980). Toxicities and descriptions of some toxaphene fractions: Isolation and identification of a highly toxic component. 1. Toxico!. Environ. Health 6, 115-125. Pulman, D. A, Smith, I. H., Larkin, J. P., and Casida, J. E. (1996). Heterocyclic insecticides acting at the GABA-gated chloride channel: 5-Alkyl-2arylpyrimidines and 1,3-thiazines. Pestic. Sci. 46,237-245. Ray, J. W (1967). The epoxidation of aldrin by housefly microsomes and its inhibition by carbon dioxide. Biochem. Pharmaco!' 16,99-107. Reddy, G., and Khan, M. A Q. (1977). Metabolism of [I4C]_photodieldrin in house flies. J. Agric. Food Chem. 25, 25-28. Revah, E, Bertrand, D., Galzi, J. L., Devillers-Theiry, A., Mulle, c., Hussy, N., Bertrand, S., Ballivet, M., and Changeux, J. P. (1991). Mutations in the channel domain alter desensitisation of a neuronal nicotine receptor. Nature (London) 353,846-849. Richardson, A., Baldwin, M. K., and Robinson, J. (1968). Metabolites of dieldrin in the urine and faeces of rats. Chem. Ind. (London) 588-590. Ryan, K J., and Engel, L. L. (1957). Hydroxylation of steroids at carbon-21. J. Bio!. Chem. 225, 103-114. Saleh, M. A, Skinner, R. E, and Casida, J. E. (1979). Comparative metabolism of 2,2,5-endo,6-exo,8,9,IO-heptachloronorbomene and toxaphene in six mammalian species and chickens. J. Agric. Food Chem. 27,731-737. Schmidt, W E, Hapeman, C. J., Fettinger, J. c., Rice, C. P., and Bilboulian, S. (1997). Structure and asymmetry in the isomeric conversion of beta- to alpha-endosulfan. J. Agric. Food Chem. 45, 1023-1026. Schofield, P. R.., Darlison, M. G., Fujita, N., Burt, D. R., Stephenson, E A, Rodriguez, H., Rhee, L. M., Ramachandran, J., Reale, v., Glencorse, T. A, Seeburg, P. H., and Barnard, E. A (1987). Sequence and functional expression of the GABAA -receptor shows a ligand-gated receptor super-family. Nature (London) 328, 221-227. Schroeder, M. E., Shankland, D. E., and Hollingworth, R. M. (1977). The effects of dieldrin and isomeric diols on synaptic transmission in the American cockroach and their relevance to the dieldrin poisoning syndrome. Pestic. Biochem. Physio!. 7,403-415. Shankland, D. L. (1982) Neurotoxic action of chlorinated hydrocarbon insecticides. Neurobehav. Toxico!. Terato!' 4, 805-811. Shankland, D. L., and Schroeder, M. E. (1973). Pharmacological evidence for a discrete neurotoxic action of dieldrin (HEOD) in the American cockroach, Perplaneta americana (L.). Pestic. Biochem. Physio!. 3, 77-86. Slade, R. E. (1945). The gamma-isomer of hexachlorocyclohexane (gammexane). Chem.Ind. (London) 40, 314-319. Slade, M., Brooks, G. T., Hetnarski, H., and Wilkinson, C. E (1975). Inhibition of the enzymatic hydration of the epoxide HEOM in insects. Pestic. Biochem. Physio!. 5, 35-46. Smith, A G. (1991). Chlorinated hydrocarbon insecticides. In "Handbook of Pesticide Toxicology" (W J. Hayes, Jr., E. R. Laws, Jr., eds.), Vol. 2, pp. 731-915. Academic Press, New York. Smith, I. H., Budd, T. c., Sills, J. H., and Casida, J. E. (1993). Insecticidal 1-(alkynyl alkyl)-3-cyano-2,6,7-trioxabicyclo[2.2.1] octanes. J. Agric. Food Chem. 41, 1114-1117. Soloway, S. B. (1965). Correlation between biological activity and molecular structure of the cyclodiene insecticides. Adv. Pest Control Res. 6, 85-126. Stemburg, J., Keams, C. W, and Moorefield, H. (1954). DDTdehydrochlorinase, an enzyme found in DDT-resistant flies. J. Agric. Food Chem. 2, 1125. Street, J. C., and Blau, S. E. (1972). Oxychlordane. Accumulation in rat adipose tissue on feeding chlordane isomers or technical chlordane. J. Agric. Food Chem. 20,395-397. Sun, Y. P., and Johnson, E. R. (1960). Synergistic and antagonistic actions of insecticide-synergist combinations and their mode of action. J Agric. Food Chem. 8, 261-266.

1156

CHAPTER 55


Takeuchi, A" and Takeuchi, N. (1966). On the permeability of the presynaptic terminal of the crayfish neuromuscular junction during synaptic inhibition and the action of GABA. J. Physiol. 183,433-449. Takeuchi, A., and Takeuchi, N. (1972). Actions of transmitter substances on the neuromuscular functions of vertebrates and invertebrates. Adv. Biophys. 3, 45-95. Tanaka, K., Scott, J. G., and Matsumura, F. (1984). Picrotoxinin receptor in the central nervous system of the American cockroach: Its role in the action of cyclodiene-type molecules. Pestic. Biochem. Physiol. 22, 117-127. Tashiro, S., and Matsumura, F. (1977). Metabolic routes of cis- and transchlordane in rats. J. Agric. Food Chem. 25, 872-880. Tomlin, C. D. S., ed. (1997). "Pesticidehonval," 4th ed., PPC, 545-547. British Crop Protection Council, Farnham, Surrey, UK. Turner, W. v., Engel, J. L., and Casida, J. E. (1977). Toxaphene components and related compounds: preparation and toxicity of some hepta-, octaand nonachloronorbornanes, hexa- and heptachlorobornenes, and a hexachlorobornadiene. J. Agric. Food Chem. 25, 1394-1401. Uchida, M., Fujita, T., Kurihara, N., and Nakajima, M. (1978). Toxicities of gamma-BHC and related compounds. In "Pesticide and Venom Neurotoxicity" (D. L. Shankland, R. M. Hollingworth, and T. Smyth, Jr., eds.), pp. 133151. Plenum, New York. Ullman and Wong et al. (1972). Unwin, N. (1995). Acetylcholine receptor channel imaged in the open state Nature (London) 373, 37-43. Von Keyserlingk, H. c., and Willis, R. J. (1992). The Gaba-activated chloride channel in insects as target for insecticide action-A physiological study. In "Insecticides: Mechansim of Action and Resistance" (D. Otto and B. Weber, eds.), pp. 205-236. Intercept, Andover, UK.

Wacher, V. J., Toia, R. F., and Casida, J. E. (1992). 2-Aryl-S-tert-butyl-1,3dithianes and their S-oxidation products: Structure-activity relationships of potent insecticides acting at the GABA-gated chloride channel. J. Agric. Food Chem. 40, 497-505. Wang, C. M., Narahashi, T., and Yamada, M. (1971). The neurotoxic action of dieldrin and its derivatives in the cockroach. Pestic. Biochem. Physiol. 1, 84-91. Weil, E. D., Colson, J. G., Hoch, P. E., and Gruber, R. H. (1969). Toxic chlorinated methanoisobenzofuran derivatives. J. Heteroeyc. Chem. 6,643-649. Weston, J. B., Larkin, J. P., Pulman, D. A., Holden, 1., and Casida, J. E. (1995). Insecticidal isomers of 4-tert-butyl-1-(4-ethynylcyclohexyl)-2,6,7trioxabicyclo[2.2.1J octane and S-tert-butyl-2-(4-ethynylcyclohexyl)-1,3dithiane. Pestic. Sci. 44, 69-74. Whittle, A. J., Fitzjohn, S., Mullier, G., Pearson, D. P. J., Perrior, T. R., Taylor, R., and Salmon, R. (1995). The use of computer-generated electrostatic surface maps for the design of new GABA-ergic insecticides. Pestic. Sci.

44,29-31. Wong and Terriere (1965). Yarbrough, J. D., Grimley, J. M., Karl, P. 1., Chambers, J. E., Case, R. S., and Alley, E. G. (1983). Tissue disposition, metabolism and excretion of cisand trans-S,lO-dihydrogen mirex. Drug Metab. Dispos. 11,611-614. Zhang, H.-G., ffrench-Constant, R. H., and Jackson, M. B. (1994). A unique amino acid of the Drosophila GABA-receptor with influence on drug sensitivity by two mechanisms. J. Physiol. 479,65-75. Zhang, H.-G., Lee, H.-J., Rocheleau, T., ffrench-Constant, R. H., and Jackson, M. B. (1995). Subunit composition determines picrotoxinin and bicuculline sensitivity of Drosophila GABA-receptors. Mol. Phannacol. 48, 835-840.

CHAPTER

56 The Avermectins: Insecticidal and Antiparasitic Agents Jim Stevens and Charles B. Breckenridge Syngente Crop Protection

56.2 CHEMISTRY AND FORMULATIONS

56.1 INTRODUCTION The avermectins are macrocyclic lactones isolated from the fermentation broth of the soil actinomycete, Streptomyces avermitilis. Included in this avermectin group are abamectin and emamectin benzoate, which are used as insecticides, and ivermectin, which is sold for parasite control in human and veterinary medicine. Because the avermectins act as GABAA receptor agonists in vertebrates, their general safety for use as pest control agents in mammals depends on an intact bloodbrain barrier in juvenile and adult animals and an intact bloodplacental barrier in utero. Inherent in the integrity of these barriers is the substance P-glycoprotein. Intact P-glycoprotein barriers are present in human adults (male and female, pregnant and nonpregnant), newboms, and children. This fact is supported by significant clinical evidence from in excess of 50 million people; a genetically varied population throughout the world, including thousands of pregnant women, have been administered therapeutic doses (0.15-0.20 mg/kg) of ivermectin. However, there is also experimental evidence that certain laboratory animals, such as genetically polymorphic CF-l mice and rat pups early postnatally, do not possess intact P-glycoprotein blood-brain barriers. Unfortunately, in the process of establishing the hazard profile for the mectins, both of these models were used. This has resulted in very low NOELs (no observed effect levels) based on neurotoxicity in the CF-l polymorphic mouse. The World Health Organization's JMPR and the U.S. Environmental Protection Agency have concluded that the CF-l polymorphic mouse is not an appropriate model for human risk assessment for the avermectins; the JMPR has recognized the flawed nature of the rat pup model for testing of these avermectins. However, a consistent understanding of the inappropriateness of applying results from standard animal models used for toxicity testing to human risk assessment is yet to be fully appreciated. Handbook of Pesticide Toxicology Volume 2. Agents

Abamectin belongs to a general class of closely related macrocyclic lac tones either produced directly by the actinomycete Streptomyces avermitilis or generated through semisynthetic modifications (Fisher and Mrozik, 1989). The structure for the natural avermectins is given in Fig. 56.1. The basic structural motif of the avermectins is evident in the natural product avermectin B 1a, which is the principal constituent of the insecticide abamectin. As used in pesticides, abamectin consists of 80% or more of avermectin Bla and 20% or less of avermectin BIb and is called avermectin BI (Fisher and Mrozik, 1989). Their structures are shown in Fig. 56.2. Chemical modification of avermectin Bla has yielded a number of semisynthetic materials. Emamectin (4/1-epimethylamino-4/1 -deoxyavermectin B la) benzoate is shown in Fig. 56.3. Emamectin and ivermectin differ from avermectin B I by having only a single bond at the C22C23 position (instead of a double bond). The major manufacturers, trade names, and formulations for abamectin, emamectin benzoate, and ivermectin are given in Table 56.1.

56.3 USES Abamectin and emamectin benzoate (Novartis) are used as insecticides and ivermectin (Merck) is sold for parasite control in human and veterinary medicine. Abamectin and emamectin migrate into treated leaves, exhibit oral activity against insect pests, and display rapid breakdown in sunlight; all of these features favor their use in integrated pest management (Bloomquist, 1999). Abamectin is used primarily to control mites, and emamectin benzoate has been designed primarily for control of lepidopterian species in vegetable, cotton, and tobacco. Ivermectin has found great favor in both the pharmaceutical and the veterinary product marketplaces. In human medicine,

1157


1158

CHAPTER 56 The Avermectins: Insecticidal and Antiparasitic Agents ivermectin has been used as an anthelmintic in the treatment of infection of intestinal threadworm, river blindness (onchocerciasis), and lymphatic filariasis. Its uses in veterinary medicine have been as anthelmintic and antiparasitic agents, including treatment of heartworm, hookworm, threadworm, and whipworm (FAOIWHO, 1992; 1993; Greene et aI., 1989).

Component A: Component B: Component 1: Component 2: Component a: Component b:

Figure 56.1

RS;: CH3 RS;: H X '"' -CH=CHX = -CH2CHOH R26 ;: C2HS R26 = CH3

56.3.1 MODE OF ACTION OF THE AVERMECTINS

Structure of the natural avermectins.

H HC:0 0 0:o 3

HO ....

3

H

3

H COO,

H

3

HaC ....

x .. -CH=CH-

: Abamectin

X .. -C~-C~- : lvermectin Figure 56.2

Structures of ivermectin and abamectin.

-.X

H3

HO .... H

cJ....o}oUHa HaC H

3

0

0,

H H3C'"

Figure 56.3

The avermectins were first demonstrated to possess anthelmintic activity in 1979 (Burg et aI., 1979). The avermectins act as chloride channel-blocking insecticides, causing hyperexcitability and convulsions. Arena et al. (1995) demonstrated that in insects stimulation of glutamate (inhibitory) chloride channels is the most sensitive target site for the avermectins. The glutamate-gated chloride channels of insect and nematode skeletal muscle are especially important as they mediate avermectin-induced muscle paralysis in these organisms. These effects are mediated via a specific, high-affinity (10- 10 M) binding site (Turner and Schaeffer, 1989). In vertebrates, the effects occur via poisoning of the central nervous system (CNS) through reactions at the receptor for the inhibitory neurotransmitter y-aminobutyric acid (GABA); see Fig. 56.4. The avermectins open the GABAA receptor chloride channel by binding to the GABA recognition site (receptor protein) and act as partial agonists (Abalis et aI., 1986). Chloride ions then flow into the postsynaptic neuron. This chloride permeability increase can significantly hyperpolarize (make more negative) the membrane potential, which has a dampening effect on nerve impulse firing. There is also a reversible dosedependent increase in chloride ion permeability in response to very low doses of avermectins. In GABA-insensitive neurons with no inhibitory innervation, the avermectins induce an irreversible increase in chloride ion conductance through interacting with voltage-dependent chloride channels. Avermectin intoxication in mammals begins with hyperexcitability, tremors, and incoordination and later develops into ataxia and coma-like sedation. This is similar to the mode of action of ethanol and barbiturates (Eldefawi and

Structure of emamectin benzoate.

Table 56.1 Chemical Structures, Trade Names, Major Manufactures, and Formulations for the Avermectins

Chemical Abamectin

Trade

Major

names

nanufactures

Agri-mek

Novartis

Emamectin Proclaim Ivomec Mectizan Stromectal

0

0~

0.15Ib/gal 0.15Ib/gal

Novartis

benzoate Denim Ivermectin

Vesicles

Formulation

Zephyr

Presynaptic Nerve Terminals

5% granule 0.16Ib/gal

Merck & Co.

Tablets Injectables

Convulsant • • block Avermeetlns • Ictlvate

G

GABA..>l

Postsynaptic Cell

GABA Receptor I Cl Channel Complex

"1/

Block of Aetion Potential

Inhibitory Postsynaptic Potential

Figure 56.4 Depiction of the mechanism of action for the avermectins in the brain (after Bloomquist, 1999).

56.3 Uses

Eldefawi, 1987) and benzodiazepine sedatives (Williams and Yarbrough, 1979). However, the avermectins are less specific in their action and can affect a variety of other ligand- and voltage-gated chloride channels. The general safety of the avermectins depends on the presence of an intact P-glycoprotein blood-brain barrier.

56.3.2 IMPORTANCE OF THE P-GLYCOPROTEIN BLOOD-BRAIN BARRIER

In mammals, the circulatory system is the most important transport system in the body. The microvasculature, which includes an estimated 40 billion capillaries in the human, serves as a site of substance exchange and is vital for the transportation of chemicals throughout the body (Aigner et aI., 1997). These capillaries consist of a single layer of endothelial cells (continuous, discontinuous, and fenestrated) surrounded by a basement membrane. The capillaries with continuous endothelial cells are only present in a few organs, including the intestine, bile duct, liver, pancreas, kidney, adrenal, testes, placenta, and brain (Thiebaut et aI., 1987). In addition to this structural feature, two other pathways exist at some sites in the brain and the placenta for the further elimination of xenobiotics and intracellular metabolites. These pathways are (1) biotransformation and (2) direct transport by transmembrane pumps such as P-glycoprotein (Fisher and Sikic, 1995; Gottesman and Pastan, 1993). P-glycoproteins are large-membrane proteins (150-180 kDa) consisting of two identical subunits each with a single adenosine 5' -triphosphate (ATP)-binding site and several transmembrane domains (Juliano and Ling, 1976). They are highly expressed in endothelial cells at areas that have a barrier function (e.g., the blood-brain barrier and the blood-placental barrier). P-glycoprotein barriers have also been identified in the adrenal gland, colon, testes, and the gravid uterus (Tiirikainen and Krusius, 1991). In addition, its localization along the apical surfaces of the intestines, proximal tubule of the kidney, and the bile ducts of the liver imply that P-glycoprotein is probably involved in secretory functions in humans. P-glycoprotein is a member of a highly conserved multigene family with isoforms identified in a wide variety of mammalian species, including humans, rats, mice, hamsters, pigs, guinea pigs, rhesus monkeys, orangutans, cows, and chickens (Saunders, 1977). Further, P-glycoprotein has been identified in fruit flies (Wu et aI., 1991) and in tobacco hornworms and budworms (Lanning et aI., 1996). In addition, P-glycoprotein has been found in mussels and sponges (Kurelec and Pivcevic, 1991), in yeast (McGrath and Varshavsky, 1989), in parasites (Descoteaux et aI., 1992), and in nematodes (Lincke et aI., 1992). The fact that P-glycoprotein is conserved across phylogenic lines suggests that it is an ancient protein associated with fundamental cellular functions.

1159

56.3.2.1 Implications of an Incomplete P-glycoprotein Blood-Brain Barrier in Hazard Testing

Unlike other transporters, P-glycoprotein transports a variety of chemically unrelated compounds. These compounds are commonly large lipophilic molecules that contain at least one aromatic ring and a positively charged nitrogen atom. Initially, P-glycoprotein interactions were thought to be limited to only the natural products such as anthracyclines, vinca alkaloids, actinomycin D, epipodophyllotoxins, taxol, and taxotere. However, it is now known that P-glycoprotein also transports steroid hormones, peptide antibiotics, immunosuppressive agents, and calcium channel blockers (Ueda et aI., 1997). Recently, pesticides have also been demonstrated to interact with P-glycoprotein. Such agents include abamectin (Didier and Loor, 1996), ivermectin (Schinkel et aI., 1995), 2-acetylaminofluorene and pentachlorophenol (Toomey and Epel, 1995), thiodicarb, and chlorpyrifos (Lanning et aI., 1996). P-glycoprotein is encoded as three isoforms. Mouse P-glycoproteins are known as mdrla (also called mdr3 or Pgyl), mdrlb (also called mdrl or Pgy2), and mdr2 (Pgy3) (Borst et aI., 1993). The isoform mdrla is primarily found in the intestinal brush border epithelium, the microvessel endothelial cells in the brain and testis, and the microvillus border of the trophoblast and Hofbauer cells of the placenta (MacFarland et aI., 1994; Nakamura et aI., 1997). The isoform mdrlb has been reported in the adrenal gland, kidney, and gravid uterus. Both mdrla and mdrlb play roles in the multidrug resistance phenotype and are able to pump xenobiotics out of the cells. The isoforms mdr 1a and mdr 1b of the mouse are comparable to the human MDRl gene in this regard (Mauad et aI., 1994). The isoform mdr2 has been shown necessary for bile production and is probably the phosphatidyl choline transporter. Mouse mdr2 is comparable to the human MDR3 (also called MDR2) gene and does not confer multidrug resistance to cells. P-glycoprotein transport processes have been conserved across species, probably because such a transporter system is essential for adaptation and survival (Saunders, 1977). It is therefore probably not surprising that a population genotypically recessive for P-glycoprotein has not been identified. However, it has been possible to develop a "knockout" mouse for the mdrla gene (Schinkel et aI., 1994; 1995). In addition, polymorphism for mdrla P-glycoprotein gene expression has been reported for the CF-l mouse (Umbenhauer et aI., 1997) and the collie dog (Lankas et aI., 1997). 56.3.2.2 Impact of an Incomplete P-glycoprotein Blood-Brain Barrier in Animal Model

Animals with a recessive mdrla (- / -) genotype do not have an intact blood-brain or blood-placental barrier because they are deficient in P-glycoprotein expression. Studies using knockout mice homozygous for disruption of mdr1a (Schinkel et al., 1994; 1995) have clearly demonstrated that the presence/absence of P-glycoprotein is a major determinant of drug entry in the brain. Studies with CF-1 mice polymorphic for

1160

CHAPTER 56

The Avermectins: Insecticidal and Antiparasitic Agents

P-glycoprotein (Umbenhauer et aI., 1997) have shown this same response. Mice with disrupted P-glycoprotein and CF-1 mice without P-glycoprotein were shown to be significantly more susceptible to the effects of neurotoxicants. Brain levels of ivermectin in the knockout mice that do not express P-glycoprotein [mdrla (- / -) genotype] were elevated approximately 90-fold over the wild type [mdr 1a (+/ +) genotype] and sensitivity to ivermectin was increased (Schinkel et aI., 1994; 1995). Further, CF-1 mice show a unique developmental response to avermectins due to the polymorphic nature of the P-glycoprotein gene. These mdr 1a ( - / -) animals could be present in a population of CF-1 mice in a range from 0 to 100%, depending on the genotype of the parental animals (Umbenhauer et aI., 1997). Therefore, experiments carried out with P-glycoprotein substrates in the heterogeneous population of the CF-1 mouse must be interpreted with caution and may be unsuitable for risk assessment. Besides the CF-1 mouse model, there are other unique features noted in the standard hazard testing models. The SpragueDawley rat pup does not establish a complete P-glycoprotein blood-brain barrier until appropriately 3 weeks postpartum, making it highly vulnerable to neurotoxic effects (Lankas et aI., 1989).

56.4 HAZARD IDENTIFICATION AND DOSE RESPONSE As previously indicated, in vertebrates, the avermectins increase membrane permeability to chloride ions and act as GABAA agonists; this is similar to the mode of action of the benzodiazepine sedatives (Turner and Schaeffer, 1989). Their toxicity follows this mode of action in overdose scenarios or in animal models with compromised mdrla P-glycoprotein barriers. The acute toxicology profiles for ivermectin and abamectin (EPA, 1999a; Lankas and Gordon, 1989) and emamectin benzoate (EPA, 1999b) are shown in Table 56.2. Table 56.2 Acute Oral Toxicity Studies on Ivennectin, Abamectin (Lankas and Gordon, 1989; EPA, 1999a), and Emamectin Benzoate (EPA, 1999b) LD50 (mg/kg) Species SD rat SD rat, neonatoa

Ivennectin 50 2

11

CF-l mouseb

25

Beagle dogs

80 > 24

Emamectin

Monkey Ivennectin

14-24

Humans

Abamectin

Therapeutic dose

Ivennectin 0.2 mg/kg

Peak plasma levels

20 ng/ml

Minimum effect level

2 mg/kg

2 mg/kg

Peak plasma levels

IlOng/ml

76 ng/ml

Signs of toxicity

Emesis

Emesis

8 mg/kg

8 mg/kg

Toxic effect level

6.6-8.6 mg/kga

Peak plasma levels

270 ng/ml

150 ng/ml

Unknown

Signs of toxicity

Emesis

Emesis

Emesis, mydriasis,

24 mg/kg

24 mg/kg

Peak plasma levels

680 ng/ml

390 ng/ml

Signs of toxicity

Emesis,

Emesis,

mydriasis,

mydriasis,

sedation

sedation

sedation Overdose level

aOverdose in humans.

These three avermectin-derived materials responded quite similarly in the different laboratory models. A comparison of the response of ivermectin and abamectin in the monkey as well as the response noted in the human with ivermectin is presented in Table 56.3. Signs of overdosing noted at 24 mg/kg of ivermectin or abamectin in the monkey were the same as observed in the human overdose at approximately 9 mg/kg. These signs were essentially identical to those observed in 10 adults who accidentally ingested tablets or solutions intended for veterinarian use (Greene, 1991). Subchronic dietary exposure to the three avermectins in the rat, mouse, and dog yielded similar results, as shown in Table 56.4. Although there are slight differences between the NOELs for abamectin and emamectin in the CD-1 mouse (probably the result of dose selection), the responses noted in the rat and dog were similar for the three avermectins. The aver-

Table 56.4 90-Day Dietary Toxicity Studies with Ivennectin (FAOIWHO, 1994), Abamectin (EPA, 1999a), and Emamectin Benzoate (EPA, 1999b)

76-88 Exposure (mg/kg/day)

1.5 220

CD-l mouse

Rhesus monkeys

Abamectin

Table 56.3 Acute Toxicity and Plasma Concentrations of Ivennectin and Abamectin (Lankas and Gordon, 1989)

107-120 22-31

Ivennectin

Abamectin

Emamectin

mg/kg

mg/kg

mg/kg

Study > 24

ap-glycoprotein-deficient blood-brain barrier is seen in neonatal rats (Lankas et aI., 1989). bCF-l mice tested were polymorphic for P-glycoprotein (Umbenhauer et aI., 1997).

SDrat

0.8

0.4

1.4

0.4

2.5

0.5

CD mouse

Nsa

NS

8

4

5.4

0.5

Beagle dog

2.0

0.5

> 1.0

0.5

0.5

0.25

aNo study available.

56.4 Hazard identification and dose response Table 56.5 Chronic Dietary Toxicity Studies with Ivermectin (FAOIWHO, 1994), Abamectin (EPA, 1999a), and Emamectin Benzoate (EPA, 1999b)

Exposure (mg/kg/day) Ivermectin

Abamectin

Emamectin

mglkg

mg/kg

mg/kg

Study SDrat

0.8

0.4

2.0

1.5

2.5

0.25

8

4

5.0 (M)

2.5

(105 weaks)a CD-1 mouse (18 months)

7.5 (F) 1.0

Beagle dog

0.5

0.5

0.25

0.5

0.25

(12 months) aRat studies with ivermectin and abamectin were only 53 weeks in duration.

mectins are equally well tolerated following chronic dietary administration, as shown in Table 56.5. The NOELs found in the chronic dog and mouse oncogenicity studies were comparable for ivermectin, abamectin, and emamectin. The NOEL in the chronic rat study was higher for abamectin than for emamectin or ivermectin. This apparent difference between abamectin and emamectin was most likely due to differences in dose selection between the studies. The avermectins are not genotoxic as has been demonstrated in a variety of standard tests for mutagenicity, clastogenicity, and unscheduled deoxyribonucleic acid (DNA) synthesis, as presented in Table 56.6. The maternotoxicity and developmentallfetotoxicity NOELs and LOELs for these three avermectins are shown in Table 56.7. In the CF-l mouse, SD rat, and rabbit developmental toxicity studies with ivermectin, cleft palate and clubbed feet (rabbit only) were observed at maternally toxic doses (Lankas and

1161

Gordon, 1989). Similar findings were noted in the CF-l mouse and rabbit studies with abamectin. Neither of these effects was noted with emamectin (EPA, 1999b). Sedation was observed in overdosed rabbit dams. Severe neurotoxicity (tremors, convulsion, and coma) was observed in some of the polymorphic CF-l mice with a compromised blood-brain barrier and blood-placental barrier (Umbenhauer et aI., 1997). These effects were also observed with ivermectin administered to the rnrdla knockout mouse (Schinkel et aI., 1994; 1995). Furthermore, the incidence of cleft palate correlated with the maternal mortality in a CF-l mouse study (Lankas et aI., 1997). The incidence of cleft palate was also linked to the polymorphism of mdrla in the CF-l mouse (Umbenhauer et aI., 1997). Developmental toxicity studies have also been conducted with the 8,9-Z-isomer of abamectin in the CF-l mouse to further evaluate the phenomenon of the linkage of developmental toxicity to the blood-placental barrier (Table 56.8). The NOEL for maternal and developmental toxicity was 0.1 mg/kg/dayand 0.05 mg/kg/day, respectively. In young adult CF-l mice, which were genotyped for their P-glycoprotein expression, the brain concentrations of the isomer 8 h after treatment were 60 times higher in (-/-) males and females than in the (+/ +) male and female CF-l mice. Brain concentrations of the delta-8,9-isomer of avermectin B 1a in ( -/-) CF-l fetuses were higher than in ( -/ +) fetuses, which, in turn, were higher than in (+ / +) fetuses. To study the development of P-glycoprotein in the placenta in the CF-l mouse, normal homozygous (+ / +) female Table 56.7 Developmental Toxicity Studies with Ivermectin (FAOIWHO, Abamectin (EPA, 1999a), and Emamectin Benzoate (EPA, 1999b)

1994),

Dose (mg/kg/day) Ivermectin Study

LOEL

Abamectin

NOEL

LOEL

NOEL

Emamectin LOEL

NOEL

Matemotoxicity Table 56.6 Genotoxicity Studies with Ivermectin (FAOIWHO, 1994), Abamectin (EPA, 1999a), and Emamectin Benzoate (EPA, 1999b)

CF-l mousea.b

0.2

0.1

6.0

3.0

0.075 2.0

0.05

Rabbit

1.0

6.0

3.0

SDrat

10.0

5.0

2.0

1.6

4.0

2.0

Mectin Fetotoxicity Tests Mutation

Ivermectin Abamectin Emamectin

Ames (+/- activation)

Negative

Mouse lymphoma

Negative

v- 79 Chinese hamster lung

Negative

Negative

CF-1 mousea Rabbit

3.0

SDrat Negative

Negative

Mouse bone marrow in vivo

Negative

Negative

Chinese hamster ovary

0.4

1.5

2.0

0.2 1.0

6.0C

1O.0c

2.0

1.6

4.0

Developmental

(+ / - activation) Clastogenicity

0.8 C

CF-1 mousea

0.4

0.2

0.4

0.2

Rabbit

3.0

1.5

1.0

6.0C

Negative

SDrat

10.0

5.0

2.0 2.0d

1.6

4.0

Negative

aCF-1 animals tested were polymorphic for P-glycoprotein (Umbenhauer et aI.,

in vitro

Other

Alkaline elution/rat hepatocyte Unscheduled DNA synthesis Negative in human fibroblasts

Negative

1997). bNot evaluated with emamectin benzoate.

cNo adverse effects at the highest dose tested. dNo cleft palates seen at the highest dose tested.

1162

CHAPTER 56 The Avennectins: Insecticidal and Antiparasitic Agents

Table 56.8 Genotyping Study of CF-1 Mice Treated with

~8,9-isomer

(Wise et aI., 1997)

Control +/-F; +/-M Number of fetuses

108

5 mg/kg

+/- F; +/- M

+/-F;+/+M

+/- F; +/+M

+/- F;+/+M

105

141

125

127

o

0

18

80

9

12

12

12

o

0

6

11

31

examined Number of fetuses with cleft palate Number of litters

8

examined Number of litters with cleft palate Pups with - / - genotype

19

50

NFb

NF

Pups with + / - genotype

15

NF

NF

41

29

Pups with + / + genotype

32

NF

39

31

NF

- / - with cleft palate

0%

0%

NF

NF

97%

+ / - with cleft palate

0%

NF

NF

39%

45%

+ / + with cleft palate

0%

NF

0%

0%

NF

aGenotype for P-glycoprotein: +, functional gene; -, defective gene. bNF, genotype not found.

and homozygous (+ / +) males were mated and the concentration of P-glycoprotein was measured in the placenta (Lankas et aI., 1989). The human population is known to be homozygous positive for this gene. The human fetus is therefore protected in utero due to a good placental-blood barrier with proper expression of P-glycoprotein (MacFarland et aI., 1994; Nakamura et aI., 1997). Furthermore, this protein has been identified in the capillaries of the brain of the human fetus as early as the third trimester (28 weeks) (Van Kalken et aI., 1992). The level of expression at 28 weeks is already the same as that of an

Table 56.9 Multigeneration Reproductive Toxicity Studies in SD Rats with the Avermectins (EPA, 1999a; 1999b; Lankas and Gordon, 1989) Ivermectin

Mg/kg/day Effect

LOEL

NOEL

0.4

0.2

Abamectin LOEL

NOEL

0.4

0.12

Emamectin LOEL

NOEL

3.6

0.6

Neonatal

Neonatal

Clinical

mortality

mortality

signs

adult.

Ivermectin has been extensively used worldwide at the high doses administered to humans (in monitored clinical trials as well as in more general therapeutic applications) (FAOIWHO, 1991). If a subpopulation of humans without P-glycoprotein existed, it would have been readily identified. Because humans and other primates have not been shown to have subpopulations deficient in P-glycoprotein, the CF-l mouse developmental toxicity data are not particularly relevant for use in human risk assessments to the avermectins (FAOIWHO, 1997). The critical levels and effects for multigeneration reproduction studies with the avermectins are shown in Table 56.9. The LOEL (lowest observed effect level) and NOEL values for ivermectin and abamectin were quite similar; the LOEL and NOEL for emamectin benzoate were somewhat higher. Early postpartum rat pup mortality has been observed with all three avermectins (EPA, 1999a; 1999b; Lankas et aI., 1989). Lankas et al. (1989) observed a significant increase in mortality between days 7 and 14 postpartum in treated dams nursing treated and control pups. In contrast, the mortality, growth, and development of treated and control pups nursing from con-

trol dams were similar. Because toxicity was only observed in control and treated pups cross-fostered to treated dams, it was concluded that neonatal toxicity of ivermectin in rats was a function of postnatal lactation exposure only and not due to in utero exposure. Further, these investigators administered purified, tritium-Iabeled avermectin B1a (ivermectin B1a) and sampled plasma and milk from dams treated orally with 2.5 mg/kg/day of radiolabeled ivermectin Bra for 61 days. The pattern for pup mortality, milk concentration, pup liver, plasma, and brain concentration of ivermectin, and percentage of adult levels for P-glycoprotein in the pup brain barrier are presented in Table 56.10. In contrast to rodents, the blood-brain barrier is formed prenatally in many species, including humans (Betz and Goldstein, 1981; Bohr and Mollgard, 1974; Jette et aI., 1995; Lankas et aI., 1989; Saunders, 1977; Van Kalken et aI., 1992). Furthermore, the blood-placental barrier is also intact in human infants at birth (MacFarland et aI., 1994; Nakamura et aI., 1997). Rats also differ from humans by having an increased utilization of fats at the time of parturition (Amano, 1967; Chiu et aI.,

56.5 Humans: Experience with Ivermectin

1163

Table 56.10 Pup Mortality, Milk and Pup Tissue Toxicokinetics, and Brain P-Glycoprotein Level Following a Daily Dose of 2.5 mg/kg/day of Tritiated Ivermectin for 61 Days (Lankas et aI., 1989) Day dose

Pup

(mg/kg)

mortality

Day I

-0.22/day"

Maternal milk

Pup plasma

Pup liver

Pup

level

level

level

brain level

p-gp level (%)b

(I-lg/g)

(I-lg/g)

(I-lg/g)

(I-lg/g)

0.094

1.640

0.100

2.324

0.276

3.918

0.251

1.482

0.804

6.106

0.318

1.052

0.893

6.648

0.264

Day 2

6.5

Day 4 Day 5

Rat pup brain

19.3/day

Day 6

5.7

Day 7 Day 8

16.9/day

Day 10

4.4

Day 11

6.9

Day 14 Day 15

19 0.86/day

Day 17

37

Day 20

89

aCorrected for background by subtracting the pup mortality observed in the control group. bCukierski (1995); P-glycoprotein expressed as percent of adult level.

1986; Scow et aI., 1964). This results in a greater release of lipophilic compounds such as abamectin and ivermectin from body fat into milk. In addition, rat milk has a much higher fat content than human milk and this leads to an increased transfer of lipophilic xenobiotic compounds to the neonatal animal compared to what would be anticipated in nursing humans. Ogbuokiri et al. (1993) reported that after a single oral dose up to 12 mg ivermectin administered to lactating women who were not breastfeeding the peak concentration in plasma was seen at 4 h posttreatment. The peak concentration in milk occurred at the same time point, but was 2-3 times lower than what was seen in plasma. These findings contrast with those found in rats where the concentration of ivermectin observed in rat milk was about threefold higher than that in plasma. Based on the uniqueness of the time profile of P-glycoprotein development and milk concentration of lipophilic toxicants in rats, it can be safely concluded that these pup deaths cannot be extrapolated to humans.

56.5 HUMANS: EXPERIENCE WITH IVERMECTIN Ivermectin has been used clinically for over a decade for the control of Onchocerca volvulus and other parasites in veterinary and human medicine (FAOIWHO, 1993). Onchoceriasis is endemic in large areas of Africa and Latin America. It is estimated that nearly 20 million people are infected and another 85 million are at risk. Large and diverse populations have been treated with ivermectin in Africa (Bumham, 1993; Chijioke and Okonkwo, 1992; Chippaux et al., 1993; De Sole et aI., 1989a; 1989b; Doumbo et al., 1992; Gardon et aI., 1997;

Ogunba and Gemade, 1992; Pacque et al., 1990; 1991; Whitworth et aI., 1991), Polynesia (Cartel et aI., 1992), and Latin America (Collins et aI., 1992). Typical human doses range from 0.1 to 0.2 mg/kg (FAOIWHO, 1993). The major effect noted following the administration of ivermectin is a severe inflammatory response, called the Mazzotti reaction. The Mazzotti reaction is secondary to the efficacy of ivermectin in killing the microfiliae, which dislodge from their site of infestation and are subsequently transported in the blood and body fluids (Ackerman et aI., 1990). This acute exacerbated immune response can be characterized by pruritis, erythema, edema, vesicle formation, papule formation, and scaling. Adenitis, fever, and hypotension may occur, and severe inflammatory changes may be noted in both the anterior and the posterior segments of the eye. The World Health Organization reviewed reports on the response to treatment for over 26,000 patients administered ivermectin for parasite control (FAOIWHO, 1993). Single oral doses up to 0.2 mg/kg (bw) produced no major effects except for those resulting from the eradication of the parasite infestation (the Mazzotti reaction). The effects observed in over 200,000 patients treated with ivermectin are summarized in Table 56.1l. Although the primary effect noted following the administration of ivermectin was the Mazzotti reaction, there were two cases of serious neurological response in two patients out a population of 17,877 treated with ivermectin (Gardon et aI., 1997). Headache was a common side effect noted, but no association between headache and treatment was observed in a doubleblind study on 7148 people conducted by Bumham (Bumham, 1993). During the first year of treatment, pain, edema, itch-

1164

CHAPTER 56


Table 56.11 Observations from Patients Treated with Ivennectin Population treated 14,911

Dose Incidence of 52 (0.35%)

(~g/kg)

Main effects observed

Reference

130--200

37 (0.25%) cases of severe

De Sole et al. (1989a)

symptomatic postural hypotension 13 (0.09%) cases of severe fever 2 (0.01 %) cases of dyspnea 118,925

835 (0.7%)

150

230 (0.19%) cases of headache

Ogunba and Gemade (1992)

210 (0.17%) cases of general pains 150 (0.12%) cases ofpruritis 120 (0.10%) cases of edema 80 (0.06%) cases offever 20 (0.02 %) cases of dizziness 15 (0.01 %) cases of vomiting 10 (0.01 %) cases of diarrhea 7,566

992 (13.1%)

150--200

Primarily Mazzotti reaction

Ogunba and Gemade (1992)

460 cases of headache 50,929

93 (1.83%)

150--200

49 cases of severe symptomatic

Chijioke and Okonkwo (1992)

postural hypotension 34 cases of severe fever 3 cases of severe dyspnea 3 cases of severe pain 7,699

100 (1.3%)

150

Primarily Mazzotti reaction

ing, and rash were found statistically associated with treatment. These reactions diminished in the second year and disappeared by the third year. Hence, in this large human study, patients treated with ivermectin did not exhibit any of the expected neurological side effects that would have occurred if the bloodbrain barrier had been compromised. The populations treated have included not only adults, but also children of all ages and, inadvertently, pregnant women. Epidemiological follow-up of more than 1000 pregnant women treated with ivermectin (primarily in the first trimester) did not yield any indication of an increase in the incidence of miscarriage, stillbirths, or congenital malformations (Burnham, 1993; Chippaux et aI., 1993; Pacque et aI., 1990). Based on this extensive human database, there should be little concern that neurotoxicity or birth defects might occur in humans exposed to the avermectins at doses less than 0.2 mg/kg.

56.6 RISK CHARACTERIZATION Previous joint meetings in 1992 (FAOIWHO, 1992) and 1994 (FAOIWHO, 1994) had established the acceptable daily intake (ADI) for abamectin at 0.0002 mg/kg bw using the NOEL (pup toxicity) of 0.12 mg/kg/day derived from the multigeneration reproduction study conducted in Sprague-Dawley rats to which a 500-fold uncertainty factor was applied (Table 56.12). This 500 x factor was based on the standard 10 x for interspecies differences and lOx for interindividual differences plus an extra 5 x due to concern over the teratogenicity of the abamectin's

De Sole et al. (1989b)

delta-8,9-isomer in the CF-l strain of mouse. In 1997, the World Health Organization's JMPR reexamined the basis for setting the ADI for abamectin and declared that the CF-l mouse was not suitable for human risk assessment because of its heterozygous (+/ - ) or homozygous ( - / - ) genetics for P-glycoprotein(FAOIWHO, 1997; Lankas eta!., 1997; Umbenhauer et aI., 1997). This same rationale, that is, the absence of P-glycoprotein blood-brain barrier in rat pups, postnatally (Betz and Goldstein, 1981; Lankas et al., 1989; Terao et al., 1996), led the JMPR to readjusted the ADI for abamectin to 0.002 mg/day (bw) by reducing the uncertainty factor from 500to 50-fold for the ADI from the rat multigeneration study. The EPA also declared that the CF-l mouse was unsuitable for human risk assessment but failed to consider the compromised unprotected postpartum period unique to rat pups and a lOO-fold uncertainty factor was applied to the NOEL for the multigeneration reproduction study conducted with abamectin (EPA, 1999a). Temporary tolerances were also set for emamectin benzoate in 1999 (EPA, 1999b); unfortunately, the EPA appears to have not applied the same criteria with regard to the inappropriateness of the CF-l mouse, polymorphic to P-glycoprotein, for human risk assessment. Not only was the NOEL for a IS-day neurotoxicity study in the CF-l mouse used, but 300-fold uncertainty factor was also applied to the NOEL derived from this study in order to calculate the reference dose (RID). Under appropriate testing conditions, these avermectins are not developmental toxins or reproductive toxins; neither are the

1165

References

Table 56.12 Chronic ADIIRtD Values Established for the Avermectins ADIor Study/

Uncertainty

RID

factor

mglkg/day

Avermectin

Organization

incidence used

Ivermectin

JECFA

Developmental

(FAOIWHO, 1992)

Abamectin

JMPR (FAOIWHO,1994)

10000intraspecies

toxicity (CF-I mouse) Multigeneration

10000intraspecies

reproduction (SD rat)

0.001

10---interspecies 0.0002

10---interspecies 5-teratogenic concerns

JMPR (FAOIWHO,1997) EPA,1999a

Multigeneration

10000intraspecies

reproduction (SD rat) Multigeneration

10---intraspecies

reproduction (SD rat) Emamectin

EPA,1999b

15-day neurotoxicity study (CF-1 mouse)

benzoate

0.002

5-interspecies 0.0012

10---interspecies 10---intraspecies

0.00025

10---interspecies 3-short duration of study used

genotoxic or carcinogenic. Further, the hazard profiles for the three averrnectins evaluated are qualitatively and often quantitatively similar. Because of these facts, it would appear to be appropriate to use critical values from the avermectin used extensively clinically for the risk characterization of this class of chemical. Therefore, for acute risk characterization, a clinical dose of 0.2 mg/kg could be used as the NOEL (FAOIWHO, 1993). Conservatively, a 10 x interindividual uncertainty factor could be applied, as well as the 3 x -uncertainty factor proposed by the EPA, that is, 30 x, for an acute RID of 0.0067 mg/kg. Likewise, using the common mechanism approach for the averrnectins, the chronic RID based on the NOEL for the I-year dog study (the most sensitive species in the chronic studies) would be suitable. The NOEL for the dog with ivermectin and abamectin (Lankas and Gordon, 1989) and emamectin benzoate (EPA, 1999b), that is, 0.5, 0.25, and 0.25 mg/kg/day, respectively, should be used instead of relying on the data from studies of shorter duration. The chronic RID of emamectin, based on either study, would be 0.25 mg/kg/day divided by lOOx and 3 x, or 0.00083 mg/kg/day, which is essentially the same as the JMPR AD! for abamectin (FAOIWHO, 1997).

56.7 CONCLUSIONS The use of inappropriate animal models for characterizing human risk unfortunately is not well recognized. The hazard assessment of the avermectins provides a better understanding of just two instances in which our surrogate models for humans fail. First, CF-l mice are more sensitive to the avermectins due to its heterozygous expression of P-glycoprotein. Second, the increased postnatal pup mortality in rats is due to a lack of expression of P-glycoprotein at birth and a high milk concentration of lipophilic toxicants unique to this species. The clinical use of ivermectin as well as special studies provides reassUf-

ance that humans are homozygous positive (unimodal) for the P-glycoprotein gene. Further, humans express P-glycoprotein fully at birth. Therefore, as long as animal models continue to be used to characterize potential risks to the human population, it will be critical to appreciate the appropriateness or inappropriateness of the genetics that drive the biological responses in animal surrogates.

REFERENCES Ackerman, S. J., Kephart, G. M., Francis, H., Awadzi, K., Gleich, G. J., and Ottesen, E. A. (1990). Eosinophil degranulation: An immunologic determinant in the pathogenesis of the Mazzotti reaction in human onchoceriasis. J.Immunol. 144,3961-3969. Abalis, I. M., Eldefawi, A. T., and Eldefawi, M. E. (1986). Actions of avermectin B la on the gamma-aminobutyric acid A receptor and chloride channels in rat brain. J. Biochem. Toxicol. 1, 69-82. Aigner, A., Wolf, S., and Gassen, H. G. (1997). Transport and detoxication: Principles, approaches, and perspectives for research on the blood-brain barrier. Angew. Chem., Int. Ed. Engl. 36,24-41. Amano Y. (1967). Changes of the levels of blood glucose during pregnancy in the rat. Jpn. J. Pharmacol. 17, 105-114. Arena, J. P., Lui, K. K., Paress, P. S., Frazier, E. G., Cully, D. E, Mrozik, H., and Schaeffer, J. M. (1995). The mechanism of action of avermectins in Caenorhabditis elegans: Correlation between activation of glutamatesensitive chloride current, membrane binding, and biological activity. J. Parasitol. 82,286-291. Betz, L., and Goldstein, G. W. (1981). Developmental changes in metabolism and transport properties of capillaries isolated from rat brain. J. Physiol. 312,365-376. Bloomquist, J. R. (1999). "Insecticides: Chemistries and Characteristics." Virginia Polytechnic Institute and State University, Blacksburg, VA. Available at http://ipmworld.umn.edulchapterslbloomq.htm. Bohr, v., and Mollgard, K. (1974). Tight junctions in human fetal choroid plexus visualized by freeze etching. Brain Res. 81, 314-318. Borst, P., Schinkel, A. H., Smit, J. J. M., Wagennar, E., van Deemter, L., Smith, A. J., Eijdems, E. W. H. M., Baas, E, and Zaman, G. J. R. (1993). Classical and novel forms of multidrug resistance and the physiological functions ofP-glycoproteins in mammals. Pharmacol. Ther. 60, 289-299.

1166

CHAPTER 56


Burg, R. w., Miller, B. M., Baker, E. E., Bumbaum, J., Cunie, S. A., Hartman, R., Kong, Y.-L., Monaghan, R. L., Olson, G., Putter, I., Tunac, J. B., Wallick, H., Stapley, E. 0., Oiwa, R., and Omura, S. (1979). Avermectins, new family of potent anthelmintic agents; producing organism and fermentation. Antimicrab. Agents Chemother. 15,361-367. Bumham, G. M. (1993). Adverse reactions to invermectin treatment for onchocerciasis: Results of a placebo-controlled, double-blind trial in Malawi. Trans. R. Soc. Trap. Med. Hyg. 87, 313-317. Cartel, J. L., Nguyen, N. L., Moulia-Pelat, J. P., Plichart, R., Martin, P. M. v., and Spiegel, A (1992). Mass chemoprophylaxis of lymphatic filariasis with a single dose of ivermectin in a Polynesian community with a high Wuchereria bancrofti infection rate. Trans. R. Soc. Trop. Med. Hyg. 86, 537-540. Chijioke, C. P., and Okonkwo, P. o. (1992). Adverse events following mass ivermectin therapy for onchocerciasis. Trans. R. Soc. Trap. Med. Hyg. 86, 284-286. Chippaux, J. P., Gardon-Wendel, N., Gardon, J., and Emould, J. C. (1993). Absence of any adverse effects of inadvertent ivermectin treatment during pregnancy. Trans. R. Soc. Trap. Med. Hyg. 87, 118. Chiu, S. H., Sestokas, E., Taub, R., Buhs, R. P., Gleen, M., Sestokas, R., Vandenheuval, W. J., Arison, B. H., and Jacob, T. A. (1986). Metabolic disposition of ivermectin in tissues of cattle, sheep, and rats. Drug Metab. Dispos. 14, 590-600. Collins, R. c., Gonzales-Peralta, c., Castro, J., Zea-Flores, G., Cupp, M. S., Richards, F 0., Jr., and Cupp, E. W. (1992). Ivermectin: Reduction in prevalence and infection intensity of Onchocerca volvulus following biannual treatments in five Guatemalan communities. Am. J. Trop. Med. Hyg. 47, 156-169. Cukierski, M. A (1995). "Exploratory study of P-glycoprotein development in Rat Fetuses and Pups:' Unpublished report, Merck Project Number TT #94-739-0, Merck & Co., West Point, PA. Descoteaux, S., Ayala, P., Orozco, E., and Samuelson, J. (1992). Primary sequences of two P-glycoprotein genes of Entamoeba histolytica. Mol. Biochem. Parasitol. 54, 201-212. De Sole, G., Awadzi, K., Remme, J., Dadzie, K. Y., Giese, J., Karam, M. FM., and Opuku, N. O. (1989a). A community trial of ivermectin in the onchocerciasis focus of Asubende, Ghana. H. Adverse reactions. Trap. Med. Parasitol. 40, 375-382. De Sole, G., Remme, J., Awadzi, K., Accorsi, S., Alley, E. S., Ba, 0., Dadzie, K. Y., Giese, J., Karam, M., and Keita, F M. (1989b). Adverse reactions after large-scale treatment of onchocerciasis with ivermectin: Combined results from eight community trials. Bull. World Health Org. 67,707719. Didier, A., and Loor, F (1996). The abamectin derivative ivermectin is a potent P-glycoprotein inhibitor. Anti-Cancer Drugs 7,745-751. Doumbo, 0., Soula, G., Kodio, B., and Perrenoud, M. (1992). Invermectine et Grossesses en Traitement de Masse au Mali. Bull. Soc. Pathol. Exp. 88, 247-251. Eldefawi, A. T., and Eldefawi, M. E. (1987). Receptors for g-aminobutyric acid and voltage-dependent chloride channels as targets for drugs and toxicants. FASEB J. 1, 262-271. Endicott, J. A., and Ling, V. (1989). The biochemistry of P-glycoprotein mediated multidrug resistance. Annu. Rev. Biochem. 58, 137-171. FAOIWHO (1991). "Pharmaceuticals: Ivermectin." Available at http://www. inchem.org/documents/pims/pharmlivermect.htm. FAOIWHO (1992). "Toxicological Evaluation of Certain Veterinary Drug Residues in Food." Report of the 36th meeting of the Joint FAOIWHO Expert Committee on Food Additives (JECFA), WHO Food Additive Series 27, pp. 10-18. FAOIWHO (1993). "Toxicological Evaluation of Certain Veterinary Drug Residues in Food." Report of the 40th Meeting of the Joint FAOIWHO Expert Committee on Food Additives (JECFA), WHO Food Additive Series 31, pp. 23-36. FAOIWHO (1994). Pesticide Residues in Food-I 994. "Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Expert Group on Pesticide Residues:' FAO Plant Production and Protection Paper 127, pp. 15-17.

FAOIWHO (1997). Pesticide Residues in Food-1997. "Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Expert Group on Pesticide Residues." September 22October I, 1997, pp. 22-34. Fisher, G. A., and Sikic, B. I. (1995). Clinical studies with modulators of multidrug resistance. Drug Resist. Clin. Oncol. Hematol. 9, 363-382. Fisher, M. H., and Mrozik, H. (1989). Chemistry. In "Ivermectin and Abamectin" (w. c. Campbell, Ed.), pp. 1-23. Springer-VerJag. New York. Gardon, J., Gardon-Wendel, N., Ngangue, D., Kamgno, J., Chippaux, J. P., and Boussinesq, M. (1997). Serious reactions after mass treatment of onchocerciasis with ivermectin in an area endemic for Loa loa infection. Lancet 350, 18-22. Gottesman, M. M., and Pastan, I. (1993). Biochemistry of multidrug resistance mediated by the multidrug transporter. Annu. Rev. Biochem. 62, 385-427. Greene, B. M. (1991). Expert report on the safety of ivermectin. In "Ivermectin-Report to JECFA," Vol. I. Unpublished Report, MSD Research Laboratories, Lauterbach, Germany. Greene, B. M., Brown, K. R., and Taylor, H. R. (1989). Use of ivermectin in humans. In "Ivermectin and Abamectin" (w. C. Campbell, Ed.), pp. 311323. Springer-VerJag, New York. Jette, L., Murphy, G. F, Lec1erc, J. M., and Beliveau, R. (1995). Interaction of drugs with P-glycoprotein in brain capillaries. Biochem. Pharmacol. 50, 1701-1709. Juliano, R. L., and Ling, V. (1976). A surface glycoprotein modulating drug permeability in Chinese hamster ovary cell mutants. Biochim. Biophys. Acta 455, 152-162. Kurelec, B., and Pivcevic, B. (1991). Evidence for a multi-xenobiotic resistance mechanism in the mussel Mytilus gallopravincialis. Aquat. Toxicol. 19,291-302. Lankas, G. R., and Gordon, L. R. (1989). Toxicology. In "Ivermectin and Abamectin" (w. C. Campbell, Ed.), pp. 89-112. Springer-VerJag, New York. Lankas, G. R., Cartwright, M. E., and Umbenhauer, D. (1997). P-glycoprotein deficiency in a subpopulation of CF-I mice enhances avermectin-induced neurotoxicity. Toxicol. Appl. Pharmacol. 143, 357-365. Lankas, G. R., Minsker, D. H., and Robertson, R. T. (1989). Effects of ivermcctin on reproduction and neonatal toxicity in rats. Food Chem. Toxicol. 27,523-529. Lanning, C. L., Fine, R. L., Corcoran, J. J., Ayad, H. A, Rose, R. L., and AbouDonia, M. B. (1996). Tobacco budworm P-glycoprotein: Biochemical characterization and its involvement in pesticide resistance. Biochim. Biophys. Acta 1291, 155-162. Lincke, C. R. I., van Groenigen, M., and Borst, P. (1992). The P-glycoprotein gene family of Caenorhabditis elegans: Cloning and characterization of genomic and complementary DNA sequences. J. Mol. BioI. 228,701-711. MacFarland, A, Abramovich, D. R., Ewen, S. W. B., and Pearson, C. K. (1994). Stage-specific distribution of P-glycoprotein in first-trimester and full-term human placenta Histochem. 1. 26,417-423. Mauad, T. H, van Nieuwkerk, C. M. J., Dingemans, K. P., Smit, J. J. M., van den Bergh Weerman, M. A., Verkruisen, R. P., Groen, A. K., Oude Elferink, R. P. J., van der Valk, M. A, Borst, P., and Offerhaus, G. J. A. (1994). Mice with homozygous disruption of the mdr2 P-glycoprotein gene: A novel animal model for studies of nonsuppurative inflammatory cholangitis and hepatocarcino-genesis. Am. J. Pathol. 145, 1237-1245. McGrath, J.P., and Varshavsky, A. (1989). The yeast STE6 gene encodes a homologue of the mammalian multidrug resistance P-glycoprotein. Nature 340,400-404. Nakamura, Y., Ikeda, S.-I., Furukawa, T., Sumizawa, T., Tani, A., Akiyama, S.-I., Nagata, Y. (1997). Function of P-glycoprotein expressed in placenta and mole. Biochem. Biophys. Res. Comun. 235, 849-853. Ogbuokiri, J. E., Ozumba, B. C., and Okonkwo, P. O. (1993). Ivermectin levels in human breast milk. Eur. J. Clin. Pharmacol. 45,389-390. Ogunba, R. O. and Gemade, F I. I. (1992). Preliminary observations on the distribution of ivermectin in Nigeria for control of river blindness. Ann. Trop. Med. Parasitol. 86,649-655.

References

Pacque, M" Munoz, B., Greene, B. M., and Taylor, H. R. (1991). Communitybased treatment of onchocerciasis with ivermectin: Safety, efficacy, and acceptability of yearly treatment. J. Infect. Dis. 163,381-385. Pacque, M., Munoz, B., Poetschke, G., Foose, 1., Greene, B. M., and Taylor, H. R. (1990). Pregnancy outcome after inadvertent ivermectin treatment during community-based distribution. Lancet 338, 486-489. Saunders, N. R. (1977). Ontogeny of the blood-brain barrier. Exp. Eye Res. Suppl. 523-550. Schinkel, A. H., Smit, 1. 1. M., van Tellingen, 0., Beijnen, 1. H., Wagennar, E., van Deemter, L., Mol, C. A. A. M., van der Valk, M. A., Robanus-Maandag, E. c., te Riele, H. P. 1., Berns, A. 1. M., and Borst, P. (1994). Disruption of the mouse mdrla P-glycoprotein gene leads to a deficiency in the bloodbrain barrier and to increased sensitivity of drugs. Cell 77, 491-502. Schinkel, A. H., Wagennar, E., van Deemter, L., Mol, C. A. A. M., and Borst, P. (1995). Absence of the mdr1a P-glycoprotein in mice affects tissue distribution and pharmacokinetics of dexamethasone, digoxin and cyc1osporin A. J. Clin. Invest. 96, 1698-1705. Scow, R. 0., Chernick, S. S., and Brinley, M. S. (1964). Hyperliperdemia and ketosis in the pregnant rat. Am. 1. Physiol. 206, 796-804. Terao, T., Hisanaga, E., Sai, Y., Tamai, 1., and Tsuji, A. (1996). Active secretion of drugs from the small intestinal epithelium in rats by P-glycoprotein functioning as an absorption barrier. J. Pharm. Pharmaco!' 48, 1083-1089. Thiebaut, F., Tsuruo, T., Hamada, H., Gottesman, M. M., Pastan, 1., and Willingham, M. C. (1987). Cellular localization of the multidrug-resistance gene product P-glycoprotein in normal human tissues. Proc. Natl. Acad. Sci. U.s.A. 84, 7735-7738. Tiirikainen, M., and Krusius, T. (1991). Multidrug resistance. Ann. Med. 23, 509-520. Toomey, B. H., and Epel, D. (1995). A multi-xenobiotic transporter in Urechis caupo embryos: Protection from pesticides? Marine Environ. Res. 39, 299300.

1167

Turner, M. 1., and Schaeffer, 1. M. (1989). Mode of action of ivermectin. In "Ivermectin and Abamectin" (w. c. Campbell. Ed.), pp. 73-87. SpringerVerlag, New York. Ueda, K., Taguchi, Y., and Morishoma, M. (1997). How does P-glycoprotein recognize its substrate? Cancer BioI. 8, 151-159. Umbenhauer, D. R., Lankas, G. R., Pippert, T. R., Wise, D., Cartwright, M. E., Hall, S. 1., and Beare, C. M. (1997). Identification of a P-glycoproteindeficient subpopulation in the CF-l mouse strain using a restriction fragment length polymorphism. Toxico!. Appl. Pharmacol. 146, 88-94. U.S. Environmental Protection Agency (EPA) (1999a). Avermectin; pesticide tolerance for emergency exemptions: Final rule. Fed. Reg. 64, 1684316850. U.S. Environmental Protection Agency (EPA) (1999b). Emamectin benzoate; pesticide tolerance: Final rule. Fed. Reg. 64,27192-27200. Van Kalken, C. K., Giaccone, G., van der Valk, P., Kuiper, C. M., Hadisaputro, M. M. N., Bosma, S. A. A., Scheper, R. 1., Meijer, C. 1. L. M., and Pineda, H. M. (1992). Multidrug resistance gene (P-glycoprotein) expression in the human fetus. Am. J. Patho!. 141,963-1072. Whitworth, 1. A. G., Morgan, D., Maude, G. H., Downham, M .D., and Taylor, D. W. (1991). A community trial ofivermectin for onchocerciasis in Sierra Leone: Adverse reactions after the first five treatments rounds. Trans. R. Soc. Trop. Med. Hyg. 85, 501-505. Williams, M., and Yarbrough, G. G. (1979). Enhancement of the in vitro binding and some of the pharmacological properties of diazepam by a novel antihelmintic agent, avermectin BIb. Eur. J. Pharmacol. 56,1273-1276. Wise, L. D., Lankas, G. R., Umbenhauer, D. R, Pippert, T. R, and Cartwright, M. E. (1997). CF-1 mouse sensitivity to abamectin-induced cleft palate correlates with fetal/placental P-glycoprotein genotype. Teratology 55, 41. Wu, C. T., Budding, M., Griffin, M. S., and Croop, 1. M. (1991). Isolation and characterization of Drosophila multidrug resistance gene homologs. Mol. Cell. Bio!. 11,3940--3948.

CHAPTER

57 Inhibitors and Uncouplers of Mitochondrial Oxidative Phosphorylation Robert M. Hollingworth Michigan State University

57.1 INTRODUCTION TO OXIDATIVE PHOSPHORYLATION: FUNCTIONS AND DYSFUNCTIONS 57.1.1 GENERAL CONCEPTS Oxidative phosphorylation (oxphos) is the primary process by which the energy derived from the catabolism of carbohydrates, fats, and proteins is used to synthesize ATP in virtually every cell of eukaryotic organisms. Since ATP is the universal source of chemical energy in the cell, any events that significantly disrupt its production or enhance its degradation will have widespread, multiple, and possibly severe physiological consequences. It is therefore not surprising that the complex, specialized machinery of oxphos is the target for a large variety of natural and synthetic toxicants, among which are a number of pesticides. Compounds that have a primary action on oxphos have a long history of use to control pests. Today, compounds that disrupt oxphos in the target species are widely used as fungicides and have important uses as insecticides and acaricides. They are of much lower significance as herbicides. Naturally occurring oxphos poisons such as rotenone have been used by native peoples for centuries. Oxphos was also the target of some of the earliest synthetic organic pesticides such as 2,4-dinitro-o-cresol which was in use to control tussock moth caterpillars in the 1890s. In last two decades, a large number of newer pesticides have been discovered that affect oxphos and many are now on market with others in the later stages of development. One problem in covering these new materials is that there is little detailed, publicly available toxicological information available, partly because of their newness and partly because of their safer nature. Many of these newer compounds have very favorable toxicological characteristics and may never generate the intense toxicological scrutiny given to some of their predecessors. On the other hand, many older compounds Handbook of Pesticide Toxicology Volume 2. Agents

in this class with a more extensive toxicological literature have been, or are in the process of being, replaced because they present risks that, by current standards, are unacceptable (e.g., arsenicals, dinitrophenols, and some uses of organotins). In other cases (e.g., rotenone) they have become limited in use because of their low intrinsic activity compared to their modem counterparts. Many of these older compounds were covered in depth in the first edition of this handbook and, because of their declining importance, are given less attention here. Finally, several older compounds affecting mitochondria that continue to be widely used are covered in other chapters of this edition [e.g., pentachlorophenol (Chapter 65), the bipiridyl herbicides such as paraquat (Chapter 70), and the fumigant phosphine and the metal phosphides that generate it (Chapter 86). Pesticides (or their active metabolites) can be divided into several broad classes according to the potency and specificity with which they affect oxphos and related mitochondrial functions:

1. Compounds which are potent disrupters of oxphos in vitro and the consequences of oxphos disruption can plausibly explain many or all of their toxic effects. This chapter focuses primarily on these compounds. 2. Compounds which are of low potency in affecting mitochondria but which may achieve sufficient concentration in vivo at high doses to impact oxphos. There is a continuum between compounds in classes 1 and 2. 3. Compounds for which other specific targets are primary in causing toxicity (e.g., neurotoxic insecticides), but for which oxphos is a possible secondary target which could be involved in some types of adverse responses. 4. Compounds which are generally reactive and have multiple sites of action, one of which is likely to be oxphos. Many older fungicides, and most fumigants, are multisite inhibitors, with a relatively nonspecific mode of action. They

1169

Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved

1170

CHAPTER 57

Pesticides Affecting Oxidative Phosphorylation

probably impact oxphos and respiration as part of their primary actions but also have many other deleterious effects on cell functions. Examples include methyl isothiocyanate and its generators (dazomet and metam-sodium), thiophosgene generators (captan, captafol, folpet), electrophilic alkylating agents (chloropicrin, ethylene dibromide, and methyl bromide), the carbon disulfide generator, sodium tetrathiocarbonate, sulfhydryl reactive compounds (acrolein, chlorothalonil, dithianon, quintozene, and some dithiocarbamate fungicides such as maneb, mancozeb, zineb, ziram), fungicidal cationic surfactants which act as general membrane disruptants (dodine, guazatine, and iminoctadine), and metal salts and derivatives (e.g., copper-, arsenic-, and mercury-containing pesticides). 5. Compounds which have no direct effect on oxphos but which can affect oxphos indirectly through the consequences of their primary action. These compounds include lipid, nucleic acidic, or protein biosynthesis inhibitors, and extramitochondrial generators of reactive oxygen species which impact oxphos functions as well as other cellular events. Generally much less is known about the specific role of mitochondria in the toxic action of the groups 2 through 5 than those in group 1. In particular, data obtained by incubating mitochondria with relatively high concentrations of pesticides in vitro may demonstrate effects on oxphos, but establishing that these have significance in vivo is much more difficult and often is not attempted. Nevertheless such actions may be toxicologically significant and this is considered briefly in Section 57.2. The criterion for inclusion in this chapter is that a compound must be known to have its most important primary toxic effect on oxphos in vertebrates or, if this unclear, the compound has a primary effect on mitochondrial oxphos in target species with a probability of similar action in vertebrates. Most of the sites at which these newer pesticides act are highly conserved across all eukaryotes, and frequently it has been shown that there is little difference in the sensitivity of the target system in a fungus or insect compared to that in a vertebrate. It is not unreasonable therefore to assume initially that these and vital oxphos sites in nontarget species are likely to be important in generating at least some of their toxic effects in non-Karget species also. The data regarding the general properties, uses, and toxicological profiles of individual compounds are drawn from a variety of sources. Many are cited individually as each compound is described. However, two sources contributed to many of the compound descriptions, The Pesticide Manual (Tomlin, 2000) and The Farm Chemicals Handbook (Meister, 2000), and these are acknowledged here as a general resource for these data.

lular energetics and, particularly, the production of ATP by oxphos. This process is now quite well understood. The mitochondrion is bounded by two membranes. The outer one is relatively permeable and allows free exchange of small molecular weight solutes with the external cellular environment. The inner membrane is not readily permeable to ions, is highly specialized, and is unusual in that it contains large amounts of the phosopholipid, cardiolipin, but little cholesterol. The control of its permeability is the key to energy conservation during oxphos. The individual catalytic components responsible for oxphos are located in and span across the inner membrane. Also spanning the inner membrane are a number of proteins which act as translocators for ions and for the precursors and products of oxphos such as tricarboxylic acid cycle substrates, and inorganic phosphate, ADP and ATP. Within the inner membrane is a matrix which contains a number of enzymes responsible for feeding the products of intermediary metabolism of carbohydrates, fats, and proteins to oxphos such as the tricarboxylic acid cycle, fatty acid oxidation, and the enzymes of the urea cycle which is responsible for elimination of wastes from the cell. The mitochondrial DNA which codes for a number of subunits of components of the respiratory chain is also located in this matrix. In addition to oxphos the mitochondrion has a number of other significant cellular functions including an important role in Ca2+ regulation within the cell, thermoregulation in some situations, and as a regulator of apoptosis and cell death. Despite these common roles, mitochondria vary widely in number, form, and activity between different tissues. These differences in functional significance and sensitivity to disruption play an important role in determining which tissues are likely to be injured by exposure to mitochondrial poisons. Information on the structure, functions, and bioenergetics of mitochondria are presented in greater detail in several books (Ernster, 1992; Nicholls and Ferguson, 1992; Scheffier, 1999; Tyler, 1992). 57.1.3 OXIDATIVE PHOSPHORYLATION

Oxidative phosphorylation (Fig. 57.1) consists of two closely coupled processes; electron transport and the phosphorylation

MATRIX Succinate

57.1.2 THE MITOCHONDRION AND THE MACHINERY OF OXIDATIVE PHOSPHORYLATION

Mitochondria are organelles that occur in virtually every eukaryotic cell. Much of the earliest work in understanding mitochondrial functions focused on their central role in cel-

CYTOPLASM

Figure 57.1

Schematic overview of mitochondrial oxidative phosphorylation.

57.1 Introduction to Oxidative Phosphorylation: Functions and Dysfunctions of ADP. In electron transport, the oxidation of intermediates from carbohydrates, fats, and proteins creates NADH and reduced ftavoproteins. These reduced carriers are reoxidized by the transfer of electrons down a chain of redox carriers culminating in the reduction of molecular oxygen and its incorporation into water. As the electrons pass from a higher energy state to a lower one, the free energy change at three points is sufficient to drive the pumping of protons outward across the inner mitochondrial membrane which is proton-impermeable. This creates an energy gradient across the membrane, termed the proton electrochemical gradient (~.uH+) or, when expressed in electrical potential units, the proton-motive force U..,.p). This consists of both an electrical component, the mitochondrial membrane potential (~I¥m) due to the charge separation, and a chemical component (~pH) due to the unequal distribution of protons across the membrane. In mitochondria most of ~p (typically about 200-220 mV) can be attributed to ~ I¥m (typicallyabout 150-180mV) (Nicholls and Budd, 2000). The value of ~pH is usually in the range of 0.5-1.0 pH units. The passage of these protons back across the inner membrane discharges the electrochemical gradient and occurs primarily by passage through channels in the membrane-spanning mitochondrial ATP synthase. The passage of protons through this enzyme drives the phosphorylation of ADP to ATP. The proton gradient is also used to power other transport processes across the membrane such as ionic regulation and amino acid uptake. The profound importance of oxphos can be judged, not only from the severe toxicological effects of compounds that disrupt it, but also from the fact that a normally active human synthesizes (and utilizes) approximately 40 kg of ATP daily. Under conditions of high muscular activity, this rate may increase by 10 to 20 fold. In addition to muscular tissues, the nervous system is also a heavy user of ATP. The human brain, which contributes 2% to the body weight, produces 20% of the ATP, mainly to power ion pumps. Since there are no appreciable other energy storage forms and ATP has a half life in cells of only a few seconds, continual resynthesis is crucial. An interruption of ATP biosynthesis of only a few minutes leads to permanent brain damage. Oxphos is the major, but not the sole, source of ATP synthesis in the cell. Glycolysis provides less than 10% of the ATP under conditions of full oxidation of glucose. However, since glycolysis produces ATP much more rapidly than oxphos, it is the major short term source of energy for skeletal muscle contraction in vertebrates. This cannot continue long since it incurs an "oxygen debt" due to the large amounts of lactic acid produced by glycolysis under anaerobic conditions, which must be reoxidized later by lactate dehydrogenase in well oxygenated tissues. When mitochondria are compromised and cannot transfer electrons to oxygen, this reoxidation cannot occur, glycolysis will slow, and lactate will accumulate in the tissues. Under extreme circumstances this can cause lactic acidosis, a potentially life-threatening condition. Finally it is worth noting that mitochondria tend to congregate in the cell around areas with a high rate of ATP utilization

1171

which is consumed very rapidly after synthesis. As shown by Aw and Jones (1985) there is a gradient of ATP away from the mitochondrion. In a condition in which ATP synthesis is partially compromised it is reasonable to assume that those functions furthest from the mitochondria will suffer the greatest deficit in ATP availability. Knowledge of the structure and mechanism of the components of oxphos has shown remarkable advances in the last decade and has recently been reviewed by Saraste (1999) and Scheffter (1999). 57.1.3.1 Electron Transport

The mitochondrial electron transport chain is illustrated in Fig. 57.2. The primary source of electrons is NADH which donates them to the chain through mitochondrial complex I (NADH: ubiquinone oxidoreductase). In sequence these electrons pass to complex III (ubiquinol : cytochrome c oxidoreductase) and then to complex IV (cytochrome oxidase) in which molecular oxygen is reduced and protonated to form water. In this process ubiquinone (Q) and cytochrome c act as mobile electron carriers within the inner mitochondrial membrane or the intermembrane space. A second, and significant, route by which electrons feed into the chain is by enzymes such as succinate dehydrogenase which reduce FAD. This passes electrons into the chain through complex 11 (succinate: ubiquinone oxidoreductase). Glycerol-3-phospate dehydrogenase (involved in shuttling extramitochondrial NADH across the impermeable inner membrane) and ,B-hydroxybutyrate dehydrogenase (not shown in Fig. 57.2) are similar FAD-linked enzymes located in the inner membrane with significance in some mitochondria. During the passage of electrons down this chain, protons are pumped outward across the inner membrane by complexes I, Ill, and IV. Thus substrates which yield NADH ultimately activate all three proton pumps whereas those that yield FADH only activate pumps in complexes III and IV, yielding proportionally less energy and ATP.

NADH FMN

~

(Fe-S)n Matrix

~

COUOOH2 ..(;:e-s~ PSST-"""

(COOH. CoO

N-2 Cytoplasm

I Rotenone, Pyridaben etc.

Figure 57.2 Major elements of mitochondrial complex I (NADH : ubiquinone oxidoreductase) and the site of action of inhibitors.

1172

CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation

Complex I (NADH: Ubiquinone Oxidoreductase) This is the largest and least well understood of the four complexes forming the respiratory chain. In vertebrates it consist of at least 42 polypeptide subunits which are thought to be arranged in an L-shaped configuration (Fig. 57.2). The shorter arm of the "L" extends into the matrix and is the locus of the NADH binding site. The longer arm is integrated into the inner membrane and contains the ubiquinone binding site or sites. Current models propose the presence of two Q binding sites acting in a Q cycle analogous to that established for complex III and invoke a ubi semiquinone radical ion as an intermediate resulting from a one electron transfer to ubiquinone (e.g., see Degli Esposti and Ghelli, 1994). In between these binding sites is a series of electron acceptors and donors including a flavoprotein and several (perhaps seven) iron-sulfur centers whose exact relationship remains to be determined. Four protons are translocated across the membrane for the reduction of each ubiquinone molecule to ubiquinol. The iron-sulfur complex designated N2 is believed to be the final carrier that passes electrons to ubiquinone. This is the region at which most of the known potent inhibitors act, including rotenone and a series of newer acaricide-insecticides such as pyri daben. Complex 11 (Succinate: Ubiquinone Oxidoreductase) In contrast to complex I, complex 11 is relatively simple consisting of only four subunits (Fig. 57.3). Although no high resolution structure for this complex is yet available, the structure of the closely related fumurate reductase of E. coli has recently been resolved at 3.3 angstroms (Iverson et aI., 1999). Despite its apparent simplicity, important features of the structure and mechanism of internal electron transfer of complex 11 remain to be established. Two of the four subunits extend into the matrix. One of these carries the catalytic site for the conversion of succinate to fumarate. The two electrons released in this oxidation are collected by an FAD cofactor and passed on through three iron-sulfur clusters in the second unit. Together, these two units Succinate

Fumarate

~ ~

FAD I

Carboxamides

Matrix

FADH2

FP IP (Fe-S)

Cytoplasm

Figure 57.3 Major elements of mitochondrial complex II (succinate: ubiquinone oxidoreductase) and the site of action of inhibitors.

represent the enzyme succinate dehydrogenase. The other two subunits act as membrane anchors for the succinate dehydrogenase. They share a b-type cytochrome of unknown function. Two quinone binding sites are present, one near the inner face and the other near the outer face of the membrane in an arrangement reminiscent to that of complex Ill, although in the case of complex 11 no protons are pumped outward as occurs in the Q cycle of complex Ill. Complex 11 is the locus of action for a group of fungicides collectively termed carboxamides.

Complex III (Ubiquinol: Cytochrome c Oxidoreductase) Complex III (which is often termed cytochrome c reductase or the bCl complex) consists of 11 peptide subunits in vertebrates. The understanding of its structure has been remarkably advanced by high resolution x-ray crystallography (Iwata et aI., 1998; Xia et aI., 1997), including definition of the binding sites of the (E)-,B-methoxyacrylate inhibitor, myxothiazol, and antimycin A (lwata et aI., 1998). However, some details of electron flow through the complex and the proton pumping mechanism are still unknown. Much of the complex protrudes into the matrix. The electron transfer chain from ubiguinol involves, in sequence, an iron-sulfur complex (named after Rieske), cytochrome b, and cytochrome Cl, arranged as shown diagrammatically in Fig. 57.4. The reaction is completed by the transfer of an electron from cytochrome Cl to the mobile carrier, cytochrome c. During these electron transfers, four protons are transferred from the matrix to the cytoplasmic side of the inner membrane. Two reaction centers for ubiquinone are present. One designated Qi (or QN) is located toward the matrix (inner, negative) face of the inner membrane, while the other (Qo or Qp) is located near the cytoplasmic (outer, positive) face. Together they are postulated to operate in the "Q cycle" in which the electron pair received from ubiquinol is split, one being passed on to the Rieske iron-sulfur center and thence to cytochromes Cl. The other electron is passed back to ubiquinone via cytochrome b yielding ubiquinol. A point of toxicological significance is that during the operations of the Q cycle, ubi semiquinone radical ions (Q-) are created as partial reaction products. By interactions with molecular oxygen, these can act as a source for reactive oxygen species (ROS) such as superoxide, hydroxyl radicals, hydrogen peroxide, and peroxynitrile radicals which may cause lipid peroxidation and other damaging cellular oxidations. Most inhibitors interact with either the Qo site (e.g., myxothiazol) or the Qi site (e.g., antimycin A). This complex is inhibited by a few insecticides, but, more significantly, it is the site of action of some very important new fungicides strobilurins that bind at the myxothiazol site. Complex IV Cytochromes Oxidase The final complex in the respiratory chain is cytochrome C oxidase, otherwise know as cytochrome oxidase. During the action of complex IV, cytochrome C is reoxidized and the electrons from the respiratory chain are used to reduce molecular oxygen resulting in the formation of water. During this process two protons are pumped outward from the matrix for each atom of oxygen reduced. The complex structure of the 13 subunit bovine cytochrome

57.1 Introduction to Oxidative Phosphorylation: Functions and Dysfunctions Matrix

2W

2 QH """

Q,

Center [

Qi

(

@ Q,

(

Center

""

1173

Phosphine is discussed elsewhere in this work (Chapter 86) and complex IV is not considered further in this chapter.

l [

57.1.4 THE SYNTHESIS OF ATP

----",Q

~Q

~ 2Q-' .JI 2QH2

57.1.4.1 Complex V (ATP Synthase)

Cytoplasm

Figure 57.4 Major elements of mitochondrial complex III (ubiquinol: cytochrome c oxidoreductase) and the site of action of inhibitors.

oxidase has been determined with high resolution (Tsukihara et al., 1996). It contains two hemes (cytochrome a and cytochrome a3), two copper centers, and two centers in which magnesium and zinc respectively are coordinated and which are probably involved in stabilizing the complex rather than in its redox reactions. The terminal cytochrome a3 operates in conjunction with one of the copper atoms to cleave molecular oxygen. A point of toxicological significance (Scheffler, 1999) is that during the reduction of oxygen by complex IV, reactive partially oxidized species such as the superoxide radical are formed. Leakage of such ROS from complex IV could contribute to oxidative stress in the cell, although leakage of electrons from higher potential sites earlier in the respiratory chain is perhaps a more general source of ROS. Much more detail might be given on the intricate operations of this oxidase, but since it is not known to be an important site of action for current pesticides, the reader is referred to Scheffler (1999) for additional discussion. Two exceptions to this generalization regarding lack of pesticidal significance are hydrogen cyanide, which is still used as a fumigant, and a second fumigant, phosphine (generated from such precursors as aluminum phosphide), which is also known to inhibit cytochrome oxidase.

ADP. -+-*x''J I Pi

ATP

Cytoplasm

Figure 57.5 Major elements of mitochondrial complex V (ATP synthase) and the site of action of inhibitors.

The linked processes of discharge of the electrochemical gradient and the synthesis of ATP is carried out by complex V (ATP synthase). This remarkable structure (considerably simplified in Fig. 57.5) has been shown to be a molecular-sized motor consisting of three major components (Abrahams et al., 1994; Boyer, 1997; Stock et al., 1999; see Scheffter, 1999 and Saraste, 1999 for overviews). A rotor system consisting of9-12 "e" subunits is embedded in the inner membrane. Although the mechanism of operation of the rotor system is still debated, it is likely that protons pass through the rotor and at the same time cause it to spin at rates up to 100 or more revolutions per second. This membrane-located portion of the complex is the Fo component of the ATP synthase. Attached to the rotor is an axle (y subunit) which is eccentric or bent and which contacts the third component, a head consisting of six subunits (alternating ex and fJ) which is attached to the inner side of the membrane (not shown in diagram). As the rotor spins and turns the axle, the axle alternately squeezes and relaxes the fJ units causing a configurational change that drives the combination of ADP and inorganic phosphate to synthesize ATP. As the fJ unit relaxes, its binding site opens to accept ADP and Pi. As the unit is squeezed by the rotating axle, ATP is expelled. Thus three ATP molecules are synthesized for each full rotation of the axle. The head portion of the complex is the FI component of the synthase. At least three, but perhaps four, protons are needed to drive the synthesis of one ATP molecule, paralleling the stoichiometry of three to four" e" subunits for each ATP catalytic site. Since the ATP synthase machinery (as with the electron transport chain) is reversible, it can hydrolyze ATP back to ADP under appropriate circumstances, at the same time pumping protons back out of the matrix. When operating in this mode it is often referred to as the mitochondrial Mg-dependent ATPase or FIFa-ATPase. The isolated FI portion of the complex can also act as an ATPase but without the capability to translocate protons. Several groups of chemicals inhibit this machinery, primarily by interference with the rotary mechanism (Fa component). These include antibiotics such as oligomycin, carbodiimides or carbodiimide generators such as the pesticide diafenthiuron, and the organotin biocides. The existence of this proton circuit comprising the pumping of protons out of the matrix by the respiratory chain and their reentry through the ATP synthase means that in healthy mitochondria with an intact inner membrane there is a close coupling between the consumption of ATP and the rate of respiration. As ATP is consumed, ADP builds up, the rate of ATP synthase rises, protons reenter the matrix, the value of the electrochemical gradient tends to decline, and electron flow through

1174

CHAPTER 57


the respiratory chain increases to maintain it. Substrate oxidation and ADP phosphorylation are therefore coupled together and respond rapidly and efficiently to the varying energy demands of the cell.

57.1.5 SIGNIFICANCE OF QUINONE BINDING SITES FOR PESTICIDE ACTION It is notable that most of the pesticides that inhibit mitochondrial electron transport do so by binding in or around the ubiquinone binding sites of complexes I, 11, or Ill. Complex TV, which contains no quinone binding site, is inhibited only by compounds which complex the metal redox centers such as cyanide or azide. The nature of these quinone binding sites and their roles as targets for pesticides and other inhibitors have been reviewed by Rich (1996), Rich and Fisher (1999), and Berry et al. (1999). The abounding variety of natural and synthetic inhibitors that act on these site in complexes I and III indicates that structural requirements for inhibitors are not highly restrictive. As pointed out by Degli Esposti (1998) and Miyoshi (1998), the main requirements for many inhibitors of complex I are a cyclic head attached to a lipophilic chain that approximates the structure of ubiquinone.

Cytoplasm

Matrix

+ +

A·- HA

Electron Transport Chain

OH"

OH"

57.1.6 MECHANISMS BY WHICH CHEMICALS DISRUPT OXPHOS Pesticides may act in several different ways to disrupt oxphos and hence cause damage to cells and tissues. First, they may inhibit the operations of the electron transport chain and the pumping of protons across the inner membrane so that the electrochemical potential gradient is not maintained. Second, they may prevent this gradient from being coupled to the synthesis of ATP. This uncoupling action generally occurs through by a compound's ability to increase the permeability of the inner membrane to protons, or other ions, and to discharge the energy stored in the gradient wastefully. This mechanism is shown in its simplest form in Fig. 57.6, where a lipophilic weak acid shuttles protons across the membrane from the outer side to the inner side, driven only by the simple physicochemical processes of association at the lower pH on the outer side of the membrane, dissociation at the more alkaline inner face, and diffusion along its chemical and electrical potential gradients. The net result is for such a compound to act as a shuttle which continually and rapidly transports protons back across the membrane and thereby discharges the electrochemical gradient formed by the activity of the electron transport chain. The third mechanism for interference with oxphos is to block the machinery of the ATP synthase and bring it to a halt. A fourth mechanism by which pesticides may interfere with mitochondrial functions and lead to tissue injury is by diverting electrons from the electron transport chain which wastes energy unproductively and can create ROS such as superoxide anion radicals and hydroxyl radicals leading to oxidative damage to the mitochondrion and other cellular constituents in the process

Figure 57.6 Basic mechanism of action of protonophoric uncouplers (HA) and suggested mechanism of uncoupling by triorganotin pesticides (R3SnOH).

often termed "oxidative stress" (e.g., see Kowaltkowski and Vercesi, 1999). Oxphos can also be impacted by compounds that inhibit the transport mechanisms that convey ATP precursors (ADP and inorganic phosphate) into the matrix and ATP itself out into the cytoplasm. Finally, arsenical compounds that yield arsenate ions in vivo have a special mechanism termed arsenolysis that prevents ATP biosynthesis. The first three of these mechanisms by which pesticides disrupt oxidative phosphorylation have somewhat different effects on cellular energetics and integrity depending on the degree to which the ATPase activity of complex V is stimulated, and whether the mitochondrial membrane potential (~Wm) can be preserved. These ideas are well reviewed by Nicholls and Budd (2000) and Wallace and Starkov (2000).

57.1.6.1 Uncoupling of Ox phos Uncouplers, generally acting as protonophores, discharge the electrochemical gradient which prevents further synthesis of ATP and the operation of other processes linked to this energy gradient such as ion pumping and the production of ROS. In addition the loss of the gradient also causes complex V (ATP synthase) to run in reverse as an ATPase which rapidly eliminates residual ATP from the cell. Even though glycolysis may continue to provide some ATP in the uncoupled cell, this too is

57.1 Introduction to Oxidative Phosphorylation: Functions and Dysfunctions

subject to the rapid ATPase activity. In almost all cells this represents an energy catastrophe and the consequences are rapid and drastic. 57.1.6.2 Inhibition of Electron Transport Whether acting at complexes I, 11, Ill, or IV, electron transport inhibitors have essentially the same effect in that electron flow through the chain is prevented. Inhibition of the respiratory chain reduces or eliminates its ability to pump protons and to maintain Llp,H+, ATP synthesis, and related reactions. Although both complexes I and 11 can independently reduce ubiquinone, the inhibition of either complex alone eventually blocks the respiratory chain through feedback effects on the citric acid cycle which provides NADH for complex I and succinate for complex 11. As Llp,H+ falls, the action of complex V tends to reverse which destroys extramitochondrial ATP, but in so doing pumps protons back out from the matrix side and so tends to maintain the Llp,H+ value close to that of uninhibited mitochondria. However, in contrast to uncouplers, the permeability of the inner membrane to protons is not increased, so that the protons now being pumped from the matrix cannot readily return and ATP is eliminated from the cell more slowly than is the case with uncouplers. Glycolysis may then, at least in the short term, be able to maintain a level of ATP in the cell that prevents catastrophic consequences depending on the relative kinetics of ATP synthesis and destruction. In the longer term, ATP synthesis may be unsustainable and again drastic effects on cellular integrity result. Since there is reserve capacity in some elements of the respiratory chain, a fairly high percentage inhibition of activity may be necessary before electron transport is limited [e.g., in brain mitochondria, complex I activity can be inhibited by 70 to 75% before it becomes the limiting factor in electron transport through the chain of respiratory carriers (Davey and Clark, 1996), and in human osteosarcoma-derived cells, it can be inhibited by 35-40% before respiration declines (Barrientos and Moraes, 1999)].

1175

57.1.6.4 Redox Cycling and Generation ofROS An additional group of compounds that can severely impact oxphos and mitochondrial integrity are those which have redox potentials within the span of that of the electron transport chain. In the simplest mode this allows them to accept electrons from carriers in the chain (e.g., complex I or complex Ill), thus diverting electron flow from the useful conservation of energy. Such alternative acceptors are likely to be particularly dangerous if they can be reoxidized efficiently, for example, by reaction with oxygen or tissue thiols, or by returning the electrons to the electron transport chain at a point of lower redox potential (e.g., complex IV). In either case this establishes a futile redox cycle in which the compound continually diverts electrons from the chain and bypasses some or all of the energy conservation sites (Fig. 57.7; compound). In the case ofthose compound that are reoxidized by oxygen or tissue thiols, the additional production of reactive free radicals, particularly the superoxide anion radical and other ROS, presents a second threat to the integrity and efficiency of oxphos and general cellular integrity (compounds B and C). Depending on the kinetics of their reduction and reoxidation, they may greatly increase oxygen consumption and discharge the electrochemical gradient and generate heat. This resembles the effects of uncouplers, since the respiratory chain continues to accept electrons from NADH, but this is not coupled to the generation of ATP. Few pesticides are known to be capable of this type of mitochondrial toxicity, but paraquat and related herbicides (see Chapter 70) are highly efficient electron acceptors which generate large amounts of ROS as a primary mechanism of action. Naphthoquinones and nitrosamines which can be formed as pesticide metabolites are also known to be capable of redox cycling. The acaricide acequinocyl (Section 57.5.2.2) is a naphthoquinone derivative, but it is has not been reported to be involved in redox cycling reactions with mitochondria. Several other pesticides (e.g., anthraquinone, dithianon, and quinoclamine) are naphthoquinones and might be involved in such cycling. If it were to occur, this would severely impact oxidative phosphorylation and mitochondrial functions in a

57.1.6.3 Inhibition of ATP Synthase In addition to preventing the synthesis of ATP from the electrochemical gradient, inhibitors of ATP synthase also block the ATPase activity of this complex. This allows glycolysis to continue to provide ATP to the cell. Since electron transport is not eliminated and LlP,H+ is not discharged (it may in fact be somewhat increased), the mitochondrial membrane potential is maintained so that other processes driven by this potential such as ion transport across the inner membrane and the generation of ROS continue unabated. The length of time for which the cell can maintain its integrity under these circumstances will depend on the balance between its capability to produce ATP by glycolysis and its rate of ATP consumption. Different cell types vary considerably in this regard (Nicholls and Budd, 2000), but even those with a vigorous glycolytic capability are likely to fail eventually in the face of continued ATP synthase inhibition.

-

A(ox)

A(red)

,J®

:::x:\~

CoQ

~~(C(OX)XOi C(red)

j/ Cyl

A(red)

~

A(ox)

C

O

/

2

_

'\0. e

O2

H2 0

Figure 57.7 Diagrammatic representation ofredox cycling and the generation of reactive oxygen species by the electron transport chain.

1176

CHAPTER 57


manner similar to that shown with menadione (Henry et aI., 1995) and adriamycin. This type of action is reviewed in greater detail by Wallace and Starkov (2000). 57.1.6.5 Inhibition of ADP/ATP Exchange

In support of oxphos, the adenine nucleotide transporter (ANT), that conveys ADP into the matrix and ATP outward from it across the impermeable inner membrane, is critical to maintain the synthesis and supply of ATP to the cell. Inhibitors of this transport system are known and have toxic effects. Bongkrekic acid and atractylosides block the ANT, acting specifically at the ATP and ADP binding sites, respectively (Fiore et aI., 1998). Recently, a new fungicide, MON 65500 (N -allyl-4,5dimethyl-2-( trimethylsil yl)thiophene-3-carboxamide) with the proposed common name of silthiofam (Beale et aI., 1998), has been shown to block the ATP transporter in fungal mitochondria as its primary fungicidal action (Joseph-Horne et aI., 2000). It is unclear whether it has the same capability in vertebrate mitochondria, but since this compound shows very low toxicity to many other fungi and to vertebrates (the acute oral LDso for the rat is >5000 mg/kg), it maybe relatively selective for the ANT of the target species. 57.1.6.6 Arsenolysis

Arsenates (lead or calcium) and arsenites were once widely used as pesticides. Their acute toxicity, concerns regarding their carcinogenicity and environmental accumulation, and lack of high efficacy have led to their general demise in this use. Their mechanisms of toxicity are multiple, but one effect, specific to the arsenate ion, involves effects on oxphos. In this case arsenate is able to replace phosphate during conversion of ADP to ATP. The resulting ADP-arsenate ester is unstable and rapidly hydrolyzes to ADP and arsenate (Moore et aI., 1983). Thus, although the mitochondrial electrochemical gradient is discharged and ADP is consumed, no energy conservation occurs. Although loosely referred to as uncoupling, since the consequences parallel those of uncoupling in increasing respiration and the wastage of energy as heat, this is not really accurate. Respiration and phosphorylation remain coupled and effectively transfer energy from substrates to an esterified ADP product. The biochemical perturbation arises from the instability of this product. The stimulation of mitochondrial respiration by arsenate is also distinguishable from that caused by uncouplers since it is blocked by the ATP synthase inhibitor oligomycin (Welle and Slater, 1967). 57.1.7 TOXICOLOGIC CONSEQUENCES OF

DISRUPTING OXIDATIVE PHOSPHORYLATION 57.1.7.1 General Effects

The mitochondrion is an extremely intricate machine. The disruption of oxphos leads to a web of complex interrelated consequences. A few examples of these interlocking processes,

which often involve positive and negative feedback circuits, include the fact that the membrane potential is responsible for driving ATP synthesis but also is directly involved in regulating ion accumulation such a Ca2+, in driving the reduction of NADPH by NADH, and in generating ROS through components of the electron transport chain. These ROS, unless effectively neutralized, are capable of destroying critical components of the electron transport chain (e.g., see Zhang et aI., 1990). The presence of increased levels of ROS also leads to damage to the mitochondrial inner membrane which, in turn, impacts both calcium regulation and the ability of the mitochondrion to maintain the electrochcmical gradient across it. At another level, the NADPH produced by mitachondrial is essential for the reduction of glutathione as an important antagonist of ROS-induced cellular injury. Calcium dysregulation also leads to the increased generation of ROS and finally may trigger events which lead to apoptosis. The many interactions between these events can be hard to fully comprehend but they are reviewed by Jabs (1999) and Nicholls and Budd (2000). Beyond the direct impact on the mitochondrion itself, the failure of ATP biosynthesis leads to many problems for the cell. Fatty acid, nucleic acid, protein, and sterol biosynthesis all require large amounts ATP and are likely to fail rapidly in an energy-compromised cell. Thus, a severe reduction in the availability of energy to the cell can lead to a variety of effects these range from a blockage in cell division, to impaired maintenance of the cellular cytoskeleton and intracellular transport due to effects on ATP-dependent tubulin polymerization, to dysfunction in ionic regulation through effects on ion pumps that are ATPdependent, and beyond. However, a reduction in ATP-dependent cellular functions is not the only deleterious effect of a compromised oxphos. As noted, the electron transport chain has the capability of producing large amounts of ROS, particularly free radicals [see Nicholls and Budd (2000) for an excellent review]. Even when operating in a healthy cell, up to 2% of the oxygen consumed by the respiratory chain is converted to superoxide radicals and thence to other ROS through the "leakage" from the electron transport chain (Boveris, 1984; Kowaltkowski and Vercesi, 1999). The superoxide is produced when electrons are transferred from intermediate sites such as ftavoproteins, Fe-S complexes, or ubiquisemiquinone radicals directly to oxygen rather than passed down the usual carriers to cytochrome oxidase. This electron leakage may arise in several locations including a site upstream of the rotenone inhibition site in complex I (Cadenas et aI., 1977; Hensley et aI., 1998; Herrero and Barja, 1997). A variety of studies have shown that the partial inhibition of complex I increases the level of ROS, free radical production, and subsequent lipid peroxidation in submitochondrial particles (Hasegawa et aI., 1990; Takeshige and Minakami, 1979; Turrens and Boveris, 1980), isolated mitochondria (Hensley et aI., 1998; Pitkanen and Robinson, 1996), and cells in culture (Barrientos and Moraes, 1999). The critical role of ROS in the lethal action of rotenone, a strong complex I inhibitor, in cultured cells was demonstrated by Seaton et al. (1997) who showed that a variety of antioxidants and free

57.1 Introduction to Oxidative Phosphorylation: Functions and Dysfunctions radical scavengers antagonized the ability of rotenone to induce apoptosis in rat PC12 cells. This enhanced oxidative stress induced by partial inhibition of complex I, rather than ATP depletion, has been cited as the critical event in rotenone's ability to cause apoptosis (Seaton et aI., 1998), and in support of the hypothesis that a deficiency in complex I activity (either genetic or chemically induced) is a cause of the oxidative cell death in the dopaminergic tracts of the substantia nigra that underlies Parkinson's disease. Barrientos and Moraes (1999) provide a particularly illuminating and complete analysis and comparison of the two potential causes of cell death caused by partial inhibition of complex I. Using rotenone and other tools, they assessed the relationship between complex I inhibition, respiration, ROS production, lipid peroxidation, the membrane potential, and the occurrence of apoptosis. Their conclusion is that the key event in causing apoptosis is an increased production of ROS rather than a decrease in ATP levels in the cells. Complex III also generates large amounts of ROS under appropriate circumstances. The site of electron transfer here is probably at the Qo site (Fig. 57.4) where the highly reactive ubisemiquinone radical can transfer an electron to molecular oxygen, creating a superoxide anion. Different types of complex III inhibitors have different effects on this ROS generation. Antimycin-A type inhibitors increase superoxide production by blocking the normal pathway for electron transfer from Qo to Qj. On the other hand, myxathiazol (and by extension other structurally related commercial fungicides based on strobilurin) decrease ROS production by preventing the formation of the ubi semiquinone radical. The complex and often contradictory results regarding the generation of ROS by mitochondria and the effects of inhibitors, including pesticides such as rotenone and the carboxamide fungicide, carboxin, on this process are reviewed by McLennan and Degli Esposti (2000). Under different experimental conditions such inhibitors may either increase or decrease the production of ROS. On the other hand, mitochondrial uncouplers, which reduce the mitochondrial membrane potential that drives ROS production, generally decrease the rate of ROS production by mitochondria. An effective complex of antioxidative defenses is needed to prevent significant injury to the cell. When these defenses, such as superoxide dismutase, peroxidases, and tissue thiols, such as GSH, are inactivated or depleted, the general phenomenon of oxidative stress can occur which is characterized by lipid oxidation, membrane destruction, and DNA and protein oxidations (Kowaltkowski and Vercesi, 1999).

57.1.7.2 Apoptosis and Necrosis A wave of recent interest in the machinery of a poptosis (programmed cell death) has done much to shed light on the complex web of events which follows interference with oxidative phoshorylation, and at the same time has highlighted the central role of the mitochondrion in controlling apoptosis as well as in cellular energetics. Apoptosis may be regarded as the orderly destruction and resorption of cells (whether during organ development or due to cellular injury) which occur constantly in

1177

the body and is characterized by blebbing of the plasma membrane, shrinkage of the cell, condensation of the chromatin and digestion of the DNA in a "ladderlike" fashion, cell shrinkage, controlled digestion of the plasma membrane to form small apoptotic bodies, and resorption of these by neigh boring cells without an inflammatory response. This stands in contrast to cellular necrosis (accidental cell death) in which more drastic damage to cellular functions and bioenergetics leads to a catastrophic collapse of all cellular functions, swelling of mitochondria and the cell in general, and rupture of the cell membrane with release of its contents and the concomitant possibility of an inflammatory response. In fact these two processes probably lie at the extreme ends of a continuum of modes of cell death. While necrosis is always a pathological event, apoptosis has important physiological functions such as the shaping of tissues and organs and the removal of injured cells. Inhibition of apoptosis can lead to anatomical malformations, autoimmune disease, and cancer. On the other hand excessive apoptosis removes healthy, vital cells and may be involved in degenerative diseases such as Parkinsonism (Kroemer et aI., 1998) and immunotoxicity. One critical determinant of whether apoptosis or necrosis occurs in an injured cell is the status of cellular energetics as defined by levels of ATP. Key events in apoptosis depend on the availability of a minimal level of ATP. If the ATP level drops too far, apoptosis cannot occur and necrosis is the likely result. Thus compounds that cause apoptosis are often found to cause necrosis when present at higher concentrations. Clearly, compounds which affect ATP levels may lead to either type of cell death and resulting organ injury. A detailed discussion of the biochemistry of apoptosis and the possible role of mitochondria and oxphos lies beyond the scope of this chapter and can be obtained from recent reviews (Brown et aI., 1999; Hengartner, 2000; Jabs, 1999; Kroemer et aI., 1998; Robertson and Orrenius, 2000). However, an outline of what is known is provided here to help to illuminate the critical role of mitochondria and oxidative phosphorylation in this process. Apoptosis may be initiated by humoral factors acting on cellular receptors or by actions within the mitochondrion. In the latter case, the key event is thought to be the opening of a very large pore, the permeability transition pore, spanning both the outer and inner mitochondrial membranes, in a process termed the mitochondrial permeability transition. This immediately causes drastic changes in mitochondrial function including discharge of the potential across the inner membrane, which drives ATP biosynthesis and ionic regulation, and the loss of the soluble cytochrome c and other apoptogenic factors from the mitochondrion. The appearance of cytochrome c in the cytoplasm activates a series of cysteine proteases (caspases) which in turn activate endonucleases that begin to digest the nuclear material. The mitochondrial permeability transition can be trigged by internal calcium dysregulation. This may result from a series of causes (often interlinked) such as loss of the ATP necessary for intracellular calcium regulation, the generation of reactive oxygen species in amounts that damage mitochondrial membranes, and changes in the status of antioxidative mech-

1178

CHAPTER 57


anisms (e.g., cellular sulfuydryl groups) which protect against the mitochondrion's constant production of free radicals. The mitochondrial permeability transition is thought to be created by the interaction of at least three proteins which have other functions in mitochondrial activity (Crompton et aI., 1999): the ANT, the voltage-dependent anion channel (also termed porin) which is a large voltage-dependent pore in the outer mitochondrial membrane that upon opening allows solutes up to 5000 Daltons to enter the mitochondrion, and cyclophilin D, a soluble protein in the matrix which may be involved in protein folding through catalysis of cis-trans isomerization of prolyl peptide bonds. The manner in which these come together to form the transition pore and the possible presence of other components remain to be established. Inhibitors of the normal functions of these components (e.g., bongkrekic acid for ANT or cyclosporin A for cyclophilin D) antagonize apoptosis. The pesticides described in this chapter may be involved in the apoptotic process in several ways [e.g., rotenone may exert an anticancer action at high doses by encouraging apoptosis of tumor-initiated cells through its effects on the bioenergetics of the cell (Section 57.3.2.1)]. Organotins cause the apoptosis of thymus cells and consequent loss of cellular immunity, probably by directly triggering the mitochondrial permeability transition (Section 57.6.2.1). More speculatively, diafenthiuron reacts with and changes the characteristics of porin in some species, although it not clear whether this action has any implications for apoptosis (Section 57.6.2.2). Obviously, there is much more to learn about the relationship between oxphos disruption by pesticides and both the mechanisms and toxicological significance of cellular apoptosis and necrosis.

57.2 OXIDATIVE PHOSPHORYLATION AS A TARGET FOR PESTICIDE ACTION AND ITS RELEVANCE FOR TOXICITY 57.2.1 GENERAL CONSIDERATIONS

Somewhere in the range of 40 to 50 pesticide active ingredients, acting either as respiratory inhibitors or uncouplers, have effects on mitochondrial oxphos in vitro at nanomolar to micromolar concentrations. Taken in conjunction with signs of poisoning and other evidence, it seems reasonable to suppose that their acute toxicity and many, though perhaps not all, of their chronic toxic effects arise by these disruption of mitochondrial energetics. These compounds are the main focus of this chapter and are discussed in detail in Sections 57.3 through 57.7. However, they represent only a part of the total spectrum of the interactions of pesticides with oxphos. It is probable that other pesticides interact with oxphos in vertebrates with some specificity but with relatively low potency, and that others interact with multiple cellular targets, among which is oxphos, but they are not specific to this site in their biochemical mechanisms of cellular disruption. Examples of the latter

group are sulfuydryl reactive compounds and general membrane disruptants. Although mitochondria may be involved in their toxicity, other sites may be equally or more important and it is often difficult to determine which are the primary and critical biochemical injuries and which are secondary or noncritical. In the former group (specific for oxphos, but weakly active) are a number of compounds, particularly fungicides and herbicides, which tend to have their pesticidal actions on physiological systems that are not present in animals including photosynthesis, essential amino acid biosynthesis, and cell wall biosynthesis. These often have very low acute toxicities to vertebrates. At oral doses of 1000-5000 mg/kg, and assuming reasonably efficient uptake, internal concentrations in the cells of nontarget species might reach 1 mM or higher. Many pesticides can be shown to perturb mitochondrial functions at such concentrations in vitro. However, only rarely has research been conducted to investigate the occurrence and significance of such low potency mitochondrial actions in the toxicology of relatively safe compounds in vivo. Fairly typical results that show the frequency with which pesticides in general exert such low potency effects on mitochondria are provided in a pair of papers by Yamano and Morita (1993, 1995). They examined the effects of 48 common pesticides of varied structures on respiration in isolated rat liver mitochondria. These compounds were not chosen because of any known ability to interact with mitochondria. Of the 48 compounds, 42% (20) were found to impact oxphos at concentrations of 1 mM or lower. Half of these were uncouplers and half were electron transport inhibitors. However, it is significant to note that only 1 (the obsolete herbicide trichlamide, acting as an uncoupler) was active at a concentration as low as 1 ).lM, and only 1 more (the insecticide amitraz acting as an inhibitor) was active at 10 ).lM. There was no obvious correlation between the ability of these compounds to disrupt oxphos and their toxicities to isolated rat hepatocytes at 1 mM. Thus, although the ability to alter oxphos in vitro is a rather common property of pesticides, this often occurs at relatively high concentrations that mayor may not have toxicological significance in vivo. A similar conclusion can be drawn from a study of the in vitro effect of 47 structurally and functionally varied pesticidal chemicals and metabolites on rat liver mitochondrial energetics by Abo-Khatwa and Hollingworth (1974). Eight of the compounds chosen were already known to be active as uncouplers or inhibitors of oxidative phosphorylation and gave appropriate results. Of the other 39 compounds, only 10 (26%) had appreciable effects on oxidation or phosphorylation at concentrations below 100 ).lM. Of the others, 41 % had threshold activities at 0.1 to 1.0 mm and 33% were inactive even at 1 mM. Thus again, many varied pesticides had the capability to interact with mitochondria, but few showed high potency. Read et al. (1998) have also published data on the action of a 164 chemicals on oxphos functions in beef heart submitochondrial particles (SMPs fragments of the inner mitochondrial membrane). Of these compounds, approximately 50 have pesticidal uses, including all classes of pesticidal activity and a wide range of structural types. Twelve of these 50 are compounds

57.2 Oxidative Phosphorylation as a Target for Pesticide Action and Its Relevance for Toxicity

widely recognized to have potent effects on mitochondrial respiration that can explain most, if not all, of their toxic effects (e.g., dinitrophenols and other phenolic uncouplers, and organotins). All of these known oxphos-perturbing compounds had EC50 values of 3 ppm (roughly 10 I-lM) or less. The most active compounds were rotenone and fentin hydroxide with EC50 values of 10 and 33 nM, respectively. Among the 37 compounds which are not generally recognized to act through oxphos effects, 46% had EC50 values of 100 I-lM or higher, and 32% had EC50 values between 10 and 100 I-lM. The remaining 8 compounds with EC50 values below 10 I-lM were all highly lipophilic neurotoxicants. In general, the potency of these compounds against SMPs was higher than in the studies of Yamano and Morita (1993, 1995) with intact mitochondria, perhaps because penetration barriers are removed during the disruption of mitochondria to form the SMPs. In the study by Read et al. (1998), the correlation of potency in causing oxphos effects with whole organism toxicity (using log transformations) is surprisingly good (e.g., an r2 value of 0.763 was obtained in correlating the EC50 on SMPs with toxicity to fish for 104 varied chemicals, and a value of 0.859 was obtained for a subset of 19 neurotoxic insecticides). However, it is unlikely that these compounds have their toxic action primarily by effects on oxphos rather than the nervous system, More likely, their high lipophilicity bestows the ability to concentrate in lipid bilayer membranes which is a prerequisite for both neurological and mitochondrial actions. 57.2.2 SPECIFIC HERBICIDES AND OXPHOS IN VERTEBRATES

There are a number of examples of compounds that act as herbicides by affecting specific plant functions and that are of unknown mode of action in vertebrates, but which have been shown to interfere with oxphos in vertebrate mitochondria in vitro. These compounds include the chlorophenoxy herbicides (2,4-D, 2,4,5-T, and MCPA) that act in plants as hormone (auxin) mimics and that are weak uncouplers in rat liver mitochondria in vitro (Abo-Khatwa and Hollingworth, 1974; Zychlinski and Zolnierowicz, 1990). At 1-10 mM, 2,4-D decreased ATP, glutathione, and NADH level in rat hepatocytes (Palmeira et aI., 1994a), inhibited complexes 11 and III strongly, and uncoupled mitochondria at 0.5 mM in isolated rat liver mitochondria. However, 2,4-D is 1OOO-fold less potent as an uncoupler than the dinitrophenolic compound dinoseb (Palmeira et aI., 1994b). Nevertheless, signs of poisoning in humans include several such as hyperventilation, tachycardia, and pyrexia and sweating that are typically seen with uncouplers (Flanagan et aI., 1990). The phenylcarbamate herbicide, terbutol, and its N -demethyl metabolite were toxic to rat hepatocytes at 1 mM in vitro. Cytotoxicity was accompanied by loss of ATP and free sulfhydryl groups, including glutathione. These herbicides also impaired the respiration of rat liver mitochondria in vitro (Suzuki et aI., 1997). However, the relationship of these observation to the in vivo toxicity of terbutol is unknown.

1179

The pyridazinone herbicide chloridazon is a photosynthetic electron transport inihibitor in plants (Tomlin, 2000). It has a low acute toxicity to vertebrates (acute oral LD50 of at least 800 mg/kg in rats) and the symptoms of poisoning bear a close resemblance to those of a mitochondrial uncoupler, including apathy, dyspnea, hyperventilation, death in clonic convulsions, and very rapid rigor mortis. However, its effects on rat liver mitochondria were complex and hard to interpret involving an inhibition of succinate oxidation, an increase in respiration with glutamate as substrate, and an increase in ATPase activity, but with an increase in the efficiency of respiratory control with both substrates (Guzy et aI., 2000; Mlynarcikova et aI., 1999). Mitochondrial disruption may therefore play a central role in its acute toxicity to vertebrates, but the mechanism is unclear. In each of these cases, even though the herbicides are only weakly active on oxphos, this may be the major biochemical target in vertebrates, but clear proof is lacking. 57.2.3 SECONDARY EFFECTS OF COMPOUNDS ACTING ON OTHER TARGETS IN VERTEBRATES: NEUROTOXICANTS

There are many reports in the literature of pesticides that are known to be potent neurotoxic ants in vertebrates also affecting mitochondrial functions in vitro. This usually occurs at relatively high concentrations compared to those needed to affect ion channels enzyme and receptors in the vertebrate nervous system. These reports include several organophosphate anticholinesterases (Carlson and Ehrich, 1999; Holmuhamedov et aI., 1996; Moreno and Madeira, 1990; Sitkiewicz et aI., 1980), organochlorines such as DDT (Moreno and Madeira, 1991), and pyrethroids (Gassner et aI., 1997; Read et aI., 1998) which affect sodium channels, the DDT metabolite DDE which is not neurotoxic (Ferreira et aI., 1997), and the organochlorines chlordane, dieldrin, endosulfan, and heptachlor, which act on GABA-gated chloride channels (Kannan et aI., 2000; Meguro et aI., 1990; Mishra and Shukla, 1995). Such effects could well explain the cytotoxicity of these compounds in vitro (e.g., the apoptotic effects found by Kannan et al. with endosulfan in a human leukemic cell line at 10-200 I-lM), but their acute toxicity in vivo is much more likely to result from specific neurological effects at lower concentrations. For example, the formamidine insecticide/acaricides (chlordimeform and amitraz) have been shown to affect oxphos by acting as uncouplers at 10-100 I-lM concentrations in vitro (Abo-Khatwa and Hollingworth, 1973, 1974; Yamano and Morita, 1993). However, it is probable that their acute toxic effects in vivo in vertebrates are related to their effects as aadrenergic agonists (e.g., see Costa et aI., 1989; Hsu et aI., 1988) and local anesthetics (e.g., see Pfister et aI., 1978) rather than to their mitochondrial actions. Another example is provided by the organochlorine insecticide chlordane. This inhibits respiration (states 3 and 4) in isolated rat liver mitochondria at 50-100 I-lM (Ogata et aI., 1989)

1180

CHAPTER 57


but quite high doses in vivo (100 mg/kg daily for four days) caused no major changes in functions of liver mitochondria in rats (Ogata and lzushi, 1991). Similarly, the anticholinesterase insecticide parathion at 0.1-1.0 mM caused several effects on oxphos in rat liver mitochondria including inhibition of complex II and ATP synthase and partial inhibition of the Pi transporter (Moreno and Madeira, 1990), and parathion and methyl parathion alter the fluidity of the mitochondrial membrane in vitro at 0.05 mM (Antunes-Madeiraet aI., 1994; Lopes et aI., 1997). However, methyl parathion has been reported to have no effect on liver mitochondria in vivo in rats (Mihara et aI., 1981). Nor did chronic exposure to several structurally related organophospates in rats cause diminished mitochondrial activity and ATP production in the brain (Fukushima et aI., 1997). While it is not possible to rigorously exclude the possibility that some toxic effects of organophosphorus insecticides result from their effects on oxphos rather than from their very potent ability to inhibit acetylcholinesterase, the claim that oxphos effects may play an important role in their toxicity based on such in vitro assays (e.g., see Sitkiewicz et aI., 1980) is quite speculative. On the other hand, endrin, an organochlorine insecticide, acts as a blocker of GABA-gated chloride channels which clearly accounts for many of its acute toxic effects. Endrin induced the formation of reactive oxygen species in isolated rat peritoneal macrophages and in hepatic mitochondria and microsomes at submicromolar concentrations, and it decreased microsomal membrane fluidity in vitro (Bagchi and Stohs, 1993). Interestingly, the same effects were observed in vivo in rats when endrin was given orally at a sublethal dose (4.5 mg/kg) (Bagchi et aI., 1993). Thus it would be a mistake to routinely dismiss the mitochondrial effects of potent neurotoxic ants observed in vitro as likely to be relatively unimportant in vivo. While it certainly does not explain all the results, one underlying reason for the reports of large number of pesticides that affect mitochondrial oxphos at high concentration in vitro may be that many of these pesticides are significantly lipophilic, and, at relatively high concentrations in vitro, can affect mitochondria through the simple disordering of the membrane integrity that is critical for mitochondrial functions and survival. Such effects on membrane structure, fluidity, and integrity have been shown to lead to a changed membrane environment for the components of oxphos and consequent alterations in their activity, membrane leakage, and uncoupling, and to the increased generation of ROS and oxidative stress (e.g., see Antunes-Madeira and Madeira, 1979; Stolze and Nohl, 1994). Examples of studies showing various types of changes in membrane ordering and fluidity by pesticides, generally at quite high concentrations (10 J.LM to 1 mM) compared to their most potent known effect on target receptors or enzymes, include chlorinated hydrocarbons, e.g., Moreno and Madeira (1991) with DDT, Ferreira et al. (1997) with its nontoxic metabolite, DDE, Antunes-Madeira and Madeira (1979) with aldrin, and Bagchi et al. (1993) and Bagchi and Stohs (1993), with en-

drin. Similar studies with organophosphates include malathion (Antunes-Madeira and Madeira, 1979), parathion (AntunesMadeira et aI., 1994), and methyl parathion (Lopes et aI., 1997). Lipophilic organotins such as tributyl- and triphenyltin also are well known to have general membrane disruptive effect (e.g., causing disruption of erythrocyte membranes at concentrations as low as 5 J.LM) (Gray et aI., 1987). A final factor that makes interpretation of some in vitro results with mitochondria difficult is that many discrepancies can be found between different reports of such weakly active compounds on mitochondria. In their survey of a range of pyrethroid insecticides Yamano and Morita (1993, 1995) found no consistent pattern of effects on rat liver mitochondrial respiration and none had an effect at a concentration less than 100 J.LM. On the other hand, Gassner et al. (1997), also using rat liver mitochondria, reported that the pyrethroids permethrin and cyhalothrin inhibit mitochondrial complex I with ICso values near 10 J.LM. Similarly, in their studies with bovine heart submitochondrial particles, Read et al. (1998) found that permethrin was one of the more potent inhibitors, active at less than 1 J.LM. By contrast, permethrin is reported to have no effect on rat liver mitochondria at 1 mM by Yamano and Morita (1993). The numerous contradictions in the published reports of the effects ofDDT on oxphos are reviewed by Moreno and Madeira (1991). It has been described by different authors as either a stimulator or an inhibitor of oxphos and related mitochondrial ATPase activity. Organophosphate insecticides were generally found to have no consistent effect on mitochondrial respiration at 1 mM (Yamano and Morita, 1993, 1995) but other authors have reported varied effects at much lower concentrations. The phenoxy acetic acid herbicide MCPA is reported to act like oligomycin in increasing the mitochondrial membrane potential through inhibition of mitochondrial ATP synthase in rat hepatocytes exposed at 2.5 mM (Camatini et aI., 1996) whereas Zychlinski and Zolnierowicz (1990) concluded that MCPA and related herbicides are weak uncouplers for rat hepatic mitochondria in vitro, an action that would tend to discharge the membrane potential. Variations between the biological preparations and techniques in these studies, and the fact that some compounds have biphasic or multiple effect on oxphos as their concentration is increased, may underlie some of these discrepancies, but others are not easily explained, and considerable caution is appropriate in interpreting the results from any single study, particularly in terms of possible toxic effects in vivo. In closing this section, it is worth noting the report that a formulation of Tordon herbicide (a mixture of picloram and 2,4-D) inhibited mitochondrial complex I in vitro (Pereira et aI., 1994). Subsequent investigation revealed that all the mitochondrial effects were due to a surfactant in the formulation and not to the active pesticidal ingredients either singly or in combination (Oakes and Pollak, 1999). This nicely illustrates the dangers of using commercial formulations in such studies in vitro and the ability of detergents to damage mitochondrial membranes.

57.3 Inhibitors of Complex I

CH,~OCH'O OH

57.3 INHIBITORS OF COMPLEX I 57.3.1 INTRODUCTION Until recently, the only pesticide thought to act primarily through the inhibition of complex I was rotenone. Although this is a familiar compound to toxicologists and biochemists, it currently has only minor uses as an insecticide because of its relatively low level of activity and short duration of action. Its extremely high toxicity to fish underlies its continuing use as a piscicide, but this, too, is a negative factor for its widespread use against insects. In the last decadc, a number of new pesticides that act on complex I with powerful acaricidal activity and some insecticidal actions have been developed and several are now used worldwide. These compounds, fenazaquin, fenpyroximate pyridaben, pyrimidifen, and tebufenpyrad, have the broad structural commonality of being lipophilic nitrogen heterocycles. The properties, mechanism of action, and toxicology of these compounds have been reviewed by Hollingworth and Ahammadsahib (1995). Because these compounds have the capability to inhibit complex I in vertebrates as well as in invertebrates with high potency, they generally possess a higher degree of acute toxicity to mammals than most modem pesticides. They also resemble rotenone in having very high toxicities to aquatic species in most cases. The structure-activity relations of several groups of complex I inhibitors, including rotenone and the lipophilic heterocyclic pesticides, has been reviewed by Miyoshi (1998) with the broad conclusion that necessary structural features for the agrochemical inhibitors are a heterocyclic ring with two nitrogens and a hydrophobic tail structure. Akagi et al. (1996) mode led the three-dimensional conformations of tebufenpyrad, fenpyroximate, and pyridaben (see Fig, 57.8 for structures) and concluded that a common structure featuring a lone electron pair in the heterocyclic ring and a hydrophobic extension with a terminal tert-Bu-substituted phenyl group existed among these compounds. The active conformations were nonplanar with the heterocyclic group and the hydrophobic tail held at about a 90° angle. This is also believed to be the active configuration of rotenone (Miyoshi, 1998). The emergence of these inhibitors has provided new tools to investigate the nature of the rotenone binding domain in complex I and its relationship to the binding sites for other complex I inhibitors and to the coenzyme Q reduction site(s). These results tend to be to confusing and there remains substantial disagreement regarding the number and relationship of binding sites for inhibitors in this region. The reader is referred to reviews by Degli Esposti (1998), Liimmen (1998,1999), Ohnishi et al. (1999), and Okun et al. (1999) for additional details. However, the conclusion seems to be generally accepted that these lipophilic nitrogen heterocycles bind at, or very close to, a high affinity rotenone binding site, which, in turn is located close to the site where coenzyme Q is reduced (see Fig. 57.2). Two (or even three) Q binding sites may be present and it has been proposed that different types of inhibitors bind preferentially to

1181

B

0

C

Rotenon:

H,C

\Q{

0

~~~H

'y Fenazaqu in

Ir\\ ---",0

\'N-{)~O-C(CH,),

CH, Fenpyroxirrale

~'

H~

CI~ CH,CH,

o

CH'CH'

0

Py r i daben

Py r i mi d i fen

C'HS~H'~(CH,), -

Cl

r;' o H

r; 0

Tebufenpyrad

Figure 57.8

9

Ir\\ ~

C'HSP;_H, Cl

CH,

N H

Tol fenpyrad

Pesticides that act as inhibitors of complex L

these different Q binding sites or can otherwise be divided into classes based on the kinetics of their interactions with complex I (Degli Esposti, 1998; Degli Esposti and Ghelli, 1994; Friedrich et al., 1994; Liimmen, 1999) but the specific details vary among these models and in some cases are contradictory. As few as one binding site near the Q reduction site (Liimmen, 1999) to as many as three (Degli Esposti, 1998) have been proposed. The case for two sites corresponding to the putative two quinone binding sites is reviewed by Degli Esposti and Ghelli (1999). These results are analyzed by Liimmen (1999) and the reasons for the apparent disparity in results are discussed. However, it is clear that complex I is indeed complex and the specific binding loci of these inhibitors remain to be established definitively. The culmination of these studies is the identification, using a photoaffinity label derived from the acaricide pyridaben, of the 23 kDa PSST subunit of complex I as the high affinity binding site for this compound and probably also for rotenone and several other complex I inhibitors (Schuler et al., 1999). This subunit is believed to link electron transfer from the terminal iron-sulfur cluster, N2, to the ubiquinone reduction site (Fig. 57.2).

57.3.2 PROPERTIES OF SPECIFIC COMPOUNDS 57.3.2.1 Rotenone (2R,6aS, 12aS)-1 ,2,6,6a, 12, 12a-hexahydro-2-isoprenyl-8,9-dimethoxychromeno[3,4-b ]furo[2,3-h ]chromen-6-one (Fig. 57.8) has CAS Reg. No. 83-79-4.

1182

CHAPTER 57


Rotenone occurs in a large number of leguminous plants, but for commercial use, it has primarily been derived from the roots of Derris species (D. elliptica, D. longicarpa, and D. mallaccensis) from Southeast Asia and Lonchocarpus species (L. urucu, L. nicou, and L. utilis) from South America. The commercial material derived from Lonchocarpus is termed cube, barbasco, nekoe, or timbo, while that from Derris is called derris root or tuba root. The root and its extracts contain a variety of compounds related to rotenone which are collectively termed rotenoids. Cube root from Peru, which is currently the only commercial source of rotenone in the United States, may be used as a ground powder (5-20% total rotenoids) or further extracted with organic solvents. Additional purification and the addition of stabilizers to prevent oxidation and microbial contamination may follow. High quality commercial cube resin typically contains 80-90% total rotenoids. In addition to rotenone itself, these rotenoids include deguelin, rotenolone, sumatrol, tephrosin, and toxicarol. They have variable but lesser insecticidal activity than rotenone. In cube, rotenone itself constitutes approximately 40% of these rotenoids with deguelin being the second most common constituent at about 20%. These two components are the most active inhibitors of mitochondrial complex I in cube resin (Fang and Casida, 1998) and are responsible for virtually all of its acute toxicity to insects, fish, mammals and cells in culture (Fang and Casida, 1997; Fang et aI., 1997). A large number of minor components of cube resin have also been identified and their biological potencies assessed (Fang and Casida, 1997, 1999a, 1999b). The composition, origins, and toxicology of rotenoids have been extensively reviewed (Metcalf, 1955; Negherbon, 1959; Fukami and Nakajima, 1971; Ha1ey, 1978; Gosse1in et aI., 1984; Ray, 1991). The reader is referred to these sources for coverage of much of the earlier work on rotenone and its toxicology. The discussion below focuses primarily on more recent results. General Properties Pure natural rotenone consists of colorless dimorphic crystals, m.p. 163°C or 181°C depending on form, v.p. 5000 ppm and the dietary NOEL in a one-generation reproduction study was 287 ppm. Fenazaquin does have a very high toxicity to fish and other aquatic species (Table 57.1). The no-effect concentration in a 63-day early life stage study in rainbow trout was extremely low, (0.96 ppb). Aquatic invertebrates are equally sensitive, with a no-effect concentration of 1.4 ppm for Daphnia growth and reproduction in a 3-week study. Shell growth is inhibited in the Eastern oyster at 5.4 ppb, and the acute LCso for brown shrimp is 15 ppb (Anonymous, 1993a). A mean BCF of 500-520 was found in trout but on depuration, 80% of the radioactivity was eliminated in 24 hr. Based on this result, Perkins et al. (1992) concluded that fenazaquin is unlikely to represent a bioaccumulation threat in fish. An aquatic microcosm study was used to simulate drift and runoff after application at 5 times the highest recommended use rate. No adverse effects were seen on aquatic life. Fenazaquin is rapidly and strongly adsorbed onto soil particles with Kd values ranging from 54 to 687 and Koc values from 18,700 to 41,200. The half-life for photolysis on the soil surface is 15 days under the summer sun at 40° latitude and 25°C, and the aerobic half-life in soils varies from 33 to 114 days with degradation mainly depending on microbial activity. It therefore has a low potential for leaching and accumulation in soils (Anonymous, 1993a). The half-life for photolysis in water is about 15 days under the same conditions as the soil photolysis. Fenazaquin is extremely stable to hydrolysis at neutral and alkaline pH (half-life is over a year) but it is hydrolyzed much more rapidly under acidic conditions. It is resistant to degradation by microbial action. In an aqueous microcosm study fenazaquin had a half-life in the water of 1 to 2 days as it partitioned onto solids where its half-life was about 5 months (Perkins et aI., 1992). Fenpyroximate: General Properties and Uses Tert-butyl (E)-a-(1,3-dimethyl-5-phenoxypyrazol-4-ylmethyleneaminooxy)-p-toluate (Fig. 57.8), CAS Reg. No. 134098-61-6, was discovered by Nihon Nohyaku Co and is described by Konno et al. (1990). The development of fenpyroximate has been reviewed by Hamaguchi et al. (1995). It exists as white crystals, m.p. 101-102°C, v.p. 7.4 x 10-6 Pa (25°C), W.s. 14.6 ppb (20°C), log P5.0I. Fenpyroximate can exist as E (trans) and Z (cis) geometrical isomers. The commercial insecticide is the E-form. In solution, fenpyroximate is readily degraded by photolysis to the Z-isomer with a half-life of 1.5 hr under conditions replicating those of sunlight. The Z-isomer then degrades with a half-life of 10.5 hr (Swanson et aI., 1995).

1189

Fenpyroximate (NNI-850, HOE 555-02A) is a widely used acaricide. Trade names include Acaben, Akari, Danitron, Dintron, Dynamite, Meteor, Naja, Ortus, Pamanrin, and Sequel.

Toxicology Profile The toxicological properties of fenpyroximate have been reviewed (FAOIWHO, 1996; U.S. EPA, 1999a) based almost entirely on unpublished studies submitted for registration. These are the source of much of the information below. An earlier summary of toxicity studies on fenpyroximate has also been published (Anonymous, 1992a). Acute Toxicity Fenpyroximate is moderately toxic after oral dosing in rodents but is somewhat more toxic by inhalation (Table 57.1). Signs of acute toxicity include hypoactivity and hypopnea. At necropsy, irritation of the gastrointestinal tract after oral dosing, and of the respiratory system after inhalation, was observed. IrritationlSensitization Fenpyroximate is not a skin irritant. It is a mild to moderate eye irritant and a moderate dermal sensitizer. Occular and dermal irritation have been noted among workers manufacturing fenpyroximate. Subchronic Toxicity In mice fed fenpyroximate at levels as high as 2000 ppm (175 mg/kg/day) for four weeks the only effects observed were several changes in hematological parameters and reduced food consumption and body weight gain. At 100 ppm (7.4 mg/kg/day) results in rats were similar with minimal liver hypertrophy and decreased white blood cell counts and plasma protein levels. At 500 ppm females had lowered acetyl- and butyrylcholinesterase levels in the blood. A subchronic oral feeding study with dogs caused some mortality at 50 mg/kg/day after appetite and body weight loss. Organ weight changes and signs of histopathological changes in the liver and kidney were also seen at this dose. Slight bradycardia and increased diarrhea were recorded at lower doses down to 10 mg/kg/day, the LOEL. Chronic Toxicity Non-neoplastic effects seen in chronic toxicity studies in rats at the highest dose tested (150 ppm; 6.9 mg/kg/day) included depressed growth, gastric ulceration, and pancreatic lobular degeneration in males and interstitial proliferation of the ovary and distention of the uterus in females. Pituitary neoplasia which compressed the brain was observed in males. There was a high incidence of pituitary adenoma in all treatment groups which nevertheless fell within the historical control range. In a lifetime feeding study in mice with doses as high as 72 mg/kg/day, no adverse effects were observed except reduced weight gain which was first noted at 10 mg/kg/day. In dogs, the results were quite similar to those in the subchronic study. Carcinogenicity Studies were negative in lifetime dietary studies in mice and rats and it was concluded that there was no evidence of compound-related carcinogenicity.

1190

CHAPTER 57


MutagenicitylGenotoxicity Results were uniformly negative in a battery of in vitro and in vivo tests including point mutations in S. typhimurium and E. coli, mutation in Chinese hamster lung cells, chromosomal aberrations in human lymphocytes in vitro, micronucleus formation in vivo in CD-l mice, DNA repair in Bacillus subtilis, and unscheduled DNA synthesis in rat hepatocytes. Reproductive Toxicity No negative effects were observed in reproductive performance in rats in a two-generation study. Reduced weight gain in both the parents and offspring occurred at the highest dose tested (100 ppm; 8.6 mg/kg/day). Developmental ToxicitylTeratogenicity No evidence of teratogenocity or fetotoxicity was observed in rats and rabbits at doses that were not also maternally toxic. An increase in the number of thoracic ribs was observed at the highest dose, 25 mg/kg/day, in rats. An increase in incidence of retinal folds in rabbits at 5 mg/kg/day fell within the historical control range. Neurotoxicity No evidence of delayed neurotoxicity was observed in hens at two oral doses of 5000 mg/kg given 21 days apart. Biochemical Mechanism of Action A variety of studies show that fenpyroximate is a powerful inhibitor of mitochondrial respiration, acting specifically at complex I. This was first described by Motoba et al. (1992) who obtained an ICso for complex I from the mite (Tetranychus) in vitro of 80 nM compared to 400 nM for that from rat liver. Fenpyroximate in vivo depleted ATP in mites and caused malformation of the mitochondria, particularly in peripheral nerves. Friedrich et al. (1994) confirmed that fenpyroximate is a high potency inhibitor of vertebrate complex I and assigned it to the same inhibitor class as piericidin A. Okun et al. (1999), using a bovine mitochondrial preparation, characterized the relative potencies of several complex I-inhibiting acaricides as pyrimidifen > fenazaquin > fenpyroximate > rotenone. The Iso values in this study ranged from 2 to 20 nM. Degli Esposti (1998) obtained a similar result with an ICso of 4.6 nM for beef heart submitochondrial particles, making fenpyroximate slightly more active than rotenone. Jewess (1994) reported that fenpyroximate strongly inhibits NADH oxidation in blowfly flight muscle mitochondria and displaces 3H-dihydrorotenone from its specific binding site on complex I in blowfly muscle submitochondrial particles. The high specificity of fenpyroximate as a miticide is not based primarily on differences in target site sensitivity since it inhibits the mitochondrial complex I from rat liver and from spider mites with less than a 10-fold difference in potency (Motoba et aI., 1992). The main mechanism of selectivity has been shown to depend on differential rates of metabolic detoxification, particularly through removal of the t-Bu group yielding the free carboxylic acid analog. This metabolite is inactive as a complex I inhibitor. This apparent hydrolysis is largely catalyzed by cytochrome P-450 through hydroxylation of the t-Bu

group followed by intramolecular ester cleavage. Oxidative ester cleavage was rapid in the several mammals, fish, and insects tested but it did not occur in mites (Motoba et aI., 2000). Absorption, Metabolism, and Elimination Fenpyroximate is well absorbed, extensively metabolized, and rapidly excreted in rats after an oral dose, primarily (70-90%) in the feces (FAOIWHO, 1996; Nishizawa et aI., 1993). The elimination half-life was 6-9 hr at 2 mg/kg and 35-49 hr at 400 mg/kg. Considerable biliary excretion occurred (approximately 50% in 48 hr). Multiple sites of metabolism were identified including oxidation of the tert-butyl group and pyrazole methyl group, ester and oxime ether hydrolysis, aryl hydroxylation, N -demethylation, and EIZ isomerization. The uptake of fenpyroximate after dermal application in rats was very low. Environmental Fate and Toxicity Fenpyroximate has a very low toxicity to birds, and it is less toxic to fish and Daphnia than most of the other members of this group of pesticides, but it is still potent enough to be classified as highly toxic by the D.S. Environmental Protection Agency (Table 57.1). It is moderately persistent in soils with a half-life of 26-50 days where it is degraded primarily by microbial action (Izawa et aI., 1993). Pyridaben: General Properties and Uses 2- Tert-butyl-5-( 4tert-butylbenzylthio )-4-chloropyridazin-3(2H)-one (Fig. 57.8), CAS Reg. No. 96489-71-3, was discovered by Nissan Chemical Industries and is described by Hirata et al. (1988). The development of pyridaben is discussed by Hirata et al. (1995). It exists as white crystals, m.p. Ill-112°C, v.p. 2.5 x 10-4 Pa (20°C), W.s. 12 ppb (24°C), log P6.37. It is relatively stable to heat and hydrolysis but sensitive to photolytic decomposition with a half-life of about 30 min at pH 7. Pyridaben (NC-129, BAS 3001) is widely used as an acaricide with a long residual action and as an insecticide mainly against sucking insects. Trade names include Nexter, Oracle, Poseidon, Pyramite, Sanmite, and Starling. Toxicology Profile The general sources of data are D.S. EPA (1997a, 1998a, 2000a). Another useful source is the detailed summary of the regulatory toxicological studies of pyridaben provided by Igarashi and Sakamoto (1994).

Acute Toxicity Pyridaben shows moderate to low acute toxicity to mammals (Table 57.1). The intraperitoneal LDso was 68 mg/kg in male rats (Igarashi and Sakamoto, 1994). The dermal toxicity is low but toxicity by the inhalation route is quite high. With sublethal doses in mice and rats, clinical signs included decreased food consumption, diarrhea, hypothermia, bradycardia, bradypnea, decreased spontaneous motor activity, abnormal gait, prostration, eye closing, amd piloerection. At near lethal or lethal doses (300 mg/kg or more) depression of the central nervous and cardiovascular systems were stronger, but no change occurred in motor functions, including coordination, muscle strength, and neuromuscular transmission, or in

57.3 Inhibitors of Complex I sensory functions. Early gastric lavage was effective in presenting poisoning in rats and loperamide was efficacious in reducing the diarrhea that occurred at low doses. IrritationlSensitization Pyridaben caused slight and readily reversible eye irritation in rabbits, but it was not a skin irritant or sensitizer in guinea pigs in either the maximization test or the modified Buelher test. However, moderate to severe skin reactions were seen in a dermal exposure study when pyridaben was applied to pregnant Himalayan rabbits at 70 mg/kg/day over 2 weeks. SubchroniclChronic Toxicity In subchronic and chronic feeding studies with mice, rats, and dogs, toxicological observations were unremarkable with endpoints being decreased food intake and weight gain, and sporadic changes in clinical chemistry or organ weights. The most sensitive endpoint was in the dog with an LOEL in a one-year dietary study of 0.5 mg/kg/day based on increased clinical signs (emesis, salivation) and decreased body weight. Carcinogenicity Pyridaben was not oncogenic in typical lifetime feeding studies in the rat and mouse. It is classified by the U.S. Environmental Protection Agency as a Group E compound (no evidence for carcinogenicity to humans). MutagenicitylGenotoxicity Pyridaben was negative in a battery of microbial and mammalian tests (Ames Salmonella assay, Chinese hamster V79 cell point mutation assay, Chinese hamster lung cell cytogenetic damage in vitro, micronucleus assay in mice in vivo, the rec-assay in Bacillus subtilis, and DNA damage and repair test in E. coli). Reproductive Toxicity No adverse reproductive effects were observed in a multigenerational study in rats at dietary doses up to 80ppm. Developmental ToxicitylTeratogenicity Delayed ossification was seen in rats and rabbits after oral administration to pregnant animals during organogenesis, but this was believed to be a secondary result of maternal stress. Fetal and placental weights were decreased in rats at 30 mg/kg/day and abortions occurred in rabbits at the highest dose (15 mg/kg/day). In both species these events occurred only at doses that were clearly maternally toxic. No evidence for teratogenicity was seen. Neurotoxicity Pyridaben caused only a low degree of acute neurotoxicity in a standard battery of neurobehavioral tests when given at a single oral dose of 200 mg/kg in males. Effects included piloerection, hypoactiviy, tremors, and lowered body temperature, but these were sporadic and transient. In a longer term (90 day) study in rats, no neurotoxicity or neuropathology was seen at oral doses up to 27 mg/kg/day, but plasma cholinesterase activity was reduced in females.

1191

Biochemical Mechanism of Action Pyridaben inhibited respiration in insects in vivo and blocked respiration at complex I in several mitochondrial preparations in vitro (Hollingworth et aI., 1994). The inhibitory potency was extremely high with IC50 values of 0.8 nM for isolated beef heart complex I and 4.0 nM for rat liver mitochondria. The very high activity of pyridaben as an inhibitor of complex I has been confirmed by Degli Esposti (1998). As already described in the Introduction (Section 57.3.1), using photoaffinity label methodology, the binding site for pyridaben in bovine heart submitochondrial particles has been shown to be the PSST subunit of complex I (Schuler et aI., 1999). Like fenazaquin (Section 57.3.2.2) and rotenone (Section 57.2.2), pyridaben also blocks complex I and the induction of ornithine decarboxylase in human breast cancer cells in vitro, an action that correlates with antiproliferative and anticancer activity (Rowlands and Casida, 1998). Oxidation of the sulfur group in pyridaben yields the corresponding sulfoxide and sulfone. Compared to pyridaben, these compounds have a reduced ability to inhibit complex I and reduced toxicity to vertebrates. However, they show an increased toxicity to mammalian cells in vitro. This may indicate a second mechanism of toxicity for pyridaben in which sulfur oxidation activates the molecule to become reactive with nucleophiles such as tissue thiols (Schuler and Casida, 1998). The occurrence and possible toxicological significance of such an action in vivo are unknown. Absorption, Metabolism, and Elimination An oral dose of pyridaben in rats was absorbed fairly well (38-46% of the dose), rapidly and completely metabolized, and eliminated mainly in the feces (80-97%) within 96 hr. Nearly 20% of the excreted residue in the feces was the parent compound. A large number of metabolites (20-30) were detected in the urine and feces with none predominant. They arose primarily from oxidation of the two tert-Bu groups and glutathione conjugation with the pyridazinone ring, probably following oxidation of the sulfur atom (Hirata et aI., 1995). Considerable biliary excretion occurred (22-30% in 24 hr) and there was evidence of enterohepatic circulation of these metabolites. Environmental Fate and Toxicity Pyridaben has a low acute toxicity to birds, but it is extremely toxic to aquatic species (Fig. 57.8). Its persistence in soil is relatively brief due to rapid microbial degradation (e.g., the half-life under aerobic conditions is reported to be less than 3 weeks). In natural water in the dark, the half-life is about 10 days, due mainly to microbial action since pyridaben is stable to hydrolysis over the pH range 5-9. The half-life including aqueous photolysis is about 30 min at pH 7 (Tomlin, 2000). Pyrimidifen: General Properties and Uses 5-Chloro- N{2-[4-(2-ethoxyethy1)-2,3-dimethylphenoxy]ethyl }-6-ethylpyrimidin-4-amine (Fig. 57.8), CAS Reg. No. 105779-78-0, was discovered by Ube Industries with joint development by the Sankyo Company.

1192

CHAPTER 57


It exists as a white crystal, m.p. 69-71 QC, v.p. 1.6 x 10-7 Pa (25°C), w.s. 2.17 ppm (25°C), log P4.59.1t is stable to hydrol-

ysis over a broad pH range. Pyrimidifen (E-787, SU-8801, SU-9118) is used as an acaricide and insecticide under the trade name Miteclean. Toxicology Profile (1999b).

The general data source is Anonymous

Acute Toxicity Pyrimidifen has a higher acute oral toxicity to mammals and birds than most other members of this group (Table 57.1). Signs of poisoning in rats include apathy, decreased respiration, abnormal gait, and decreased urinary output. IrritationlSensitization Pyrimidifen causes slight eye irritation but no dermal irritation or sensitization. Subchronic Toxicity In a 90-day dietary study in rats, toxicological effects were rather nonspecific including decreased weight gain, some changes in blood chemistry, and changes in organ weights. An increase in liver weight was seen in females at 0.69 mg/kg/day, the lowest dose tested. Similar results were seen in mice and the LOEL was established at 17.7 mg/kg/day. In dogs the LOEL was 0.5 mg/kg/day based on observations of diarrhea. Increased salivation and reduced body weight gain were seen at the highest dose, 4.5 mg/kg/day. Chronic Toxicity In a 2-year study in rats, dietary levels at 100 ppm (3.9 mg/kg/day) increased the incidence of benign adrenal pheochromocytoma. Increased weight and discoloration of the kidney and lipofuscin deposition in the tubules were also seen. The signs of toxicity in parallel long-term feeding studies in dogs and mice were unremarkable. Carcinogenicity Long-term feeding studies in rats, mice, and dogs proved negative for carcinogenicity. MutagenicitylGenotoxicity Pyrimidifen gave negative results in the Ames assay and a bacterial DNA repair test (Rec-assay). It did not cause chromosome aberrations in the Chinese hamster lung fibroblast assay. Reproductive Toxicity None was seen in a two-generation study in rats at doses of 7.6-9.5 mg/kg/day. Developmental ToxicitylTeratogenicity Some minor skeletal abnormalities were found when pyrimidifen was fed to pregnant rats and rabbits at 20-25 mg/kg/day, but these doses also caused maternal toxicity and it was concluded that pyrimidifen is not teratogenic.

Biochemical Mechanism of Action Pyrimidifen is an extremely potent inhibitor of complex I in bovine submitochondrial particles (150 of 2 nM) and competes with high affinity for a binding domain that is in common with fenazaquin, fenpyroximate, and rotenone (Okun et aI., 1999).

Tebufenpyrad: General Properties and Uses N-(4-tertbutylbenzyl )-4-chloro-3-ethyl-1-methylpyrazole-5-carboxamide (Fig. 57.8), CAS Reg. No. 119168-77-3, was discovered by Mitsubishi Kasei Corporation and was developed in partnership with the American Cyanamid Co. It is described by Kyomura et al. (1990) and Inoue and Fukuchi (1994). The structural conformation of tebufenpyrad has been determined by x-ray crystallography (Osano et aI., 1991). The synthesis and structure-acaricidal activity relations of tebufenpyrad and related compounds are described by Okada et al. (1991). A closely related compound, tolfenpyrad (OMI-88; CAS Reg. No. 129558-76-5; Fig. 57.8), with stronger insecticidal activity than tebufenpyrad and reasonable mammalian safety is currently under development in Japan (Okada et aI., 1999). General Properties Tekufenpyrael exists as white crystals, m.p. 61-62°C, v.p. 1 x 10-5 Pa (25°C), w.s. 2.8 ppm (25°C), log P4.61-5.04. It is stable to aqueous hydrolysis with a halflife over 28 days at pH 5-9. Uses Tebufenpyrad(MK-239, SAN-831, AC 801,757) is used primarily as an acaricide but it also has activity against sucking insects. Trade names include Comanch6, Masa'i, Oscar, and Pyranica. Toxicology Profile The general sources of the data are Anonymous (l993b) and Mitsubishi Chemical Industries (1995). Acute Toxicity Tebufenpyrad shows moderate acute toxicity to mammals (Table 57.1). The toxicity to rabbits is higher with an acute oral LDso between 40 and 100 mg/kg. By the dermal and inhalation routes its toxicity is low. Clinical signs of poisoning include slowed respiration, decreased locomotor activity, and prostration. Increased salivation is also seen after respiratory exposure. In an attempt to test therapeutic strategies, rabbits dosed orally at 100 mg/kg (LDlOO) were given dimorpholamine intravenously at 3 or 6 mg/kg intravenously when respiration had decreased to 50% of normal. This treatment improved respiratory function significantly but it is not stated whether mortality was decreased. IrritationlSensitization Tebufenpyrad caused slight, reversible eye irritation in rabbits, but it was not a skin irritant or sensitizer. Subchronic Toxicity A uniform picture emerges for 90-day dietary or gavage studies in rats, mice, and dogs. The only effects consistently observed were decreased weight gain and increased liver weight at the highest doses [about 28 mg/kg/day in the rat, 160 (M) to 220 (F) mg/kg/day in mice, and 10 mg/kg/day in dogs]. Some liver hypertrophy was seen in rats, and vomiting and diarrhea occurred in dogs. Chronic Toxicity Longer term exposures in these animals caused essentially the same responses as in the 90-day studies, but at somewhat lower doses. The NOELs in both rats

57.4 Inhibitors of Complex II (24-month dietary exposure) and dogs (12-month oral dosing) were found at 1 mg/kg/day. Chronic gastritis was observed in the dogs at the highest dose (20 mg/kg/day). The NOEL in an 18-month dietary study in mice was 4 mg/kg/day. Carcinogenicity No evidence of carcinogenicity was seen in the lifetime dietary studies in mice and rats. MutagenicitylGenotoxicity Tebufenpyrad is not mutagenic or clastogenic in a typical battery of bacterial and mammalian tests including the Ames Salmonella, Chinese hamster ovary cell, chromosomal aberrations in lymphocytes, mouse micronucleus, and unscheduled DNA synthesis assays. Reproductive Toxicity No serious adverse reproductive effects were observed in a two-generation dietary study in rats. A decrease in pup weight gain was recorded at the highest dose, 16 mg/kg/day, which was the LOEL for the study. Developmental ToxicitylTeratogenicity No developmental or teratogenic effects were observed when tebufenpyrad was given orally during organogenesis to rats at 150 mg/kg/day or to rabbits at 40 mg/kg/day, the highest doses tested. The NOEL for maternal effects was about 15 mg/kg/day in each study.

Biochemical Mechanism of Action Like the other compounds in this group, tebufenpyrad is a powerful and specific inhibitor of complex I with an Iso value of 6 nM for bovine heart submitochondrial particles (Degli Esposti, 1998) and 2 nM for mitochondria from housefly flight muscle (Liimmen, 1998). Absorption, Metabolism, and Elimination Tebufenpyrad is rapidly metabolized and cleared after an oral dose in rats. The metabolism of tebufenpyrad in rats in vitro and in vivo was investigated by Ogawa and Ihashi (1993). Hydroxylations of the ethyl and t-Bu groups were the predominant reactions both in vitro, using an S-9liver fraction, and in vivo. Subsequent oxidation of these initial alcohols to carboxylic acids and conjugation of the alcohols with sulfate occurred in vivo. Little cleavage of the amide bond was observed. Environmental Fate and Toxicity Tebufenpyrad is relatively safe to birds (Table 57.1). In 8-day feeding studies in mallard ducks and quail, the LCso was > 5000 ppm in both species. Like the other members of this group it is highly toxic to fish and to Daphnia (Table 57.1). Other aquatic invertebrates are also highly sensitive to tebufenpyrad (e.g., the LCso for mysid shrimps is 22 ppb and the ECso for the inhibition of shell growth in the Eastern oyster is 62 ppb). The 22-day no-effect concentration for reproduction in Daphnia is very low at 2.4 ppb. The uptake, metabolism, and excretion of tebufenpyrad by carp have been studied by Saito et al. (1994). The BCF at steady state, which was reached within about 4 days of exposure, was 864. However, less than 4% of the radioactivity in the body was unchanged tebufenpyrad, so the BCF for the parent is only 29. The major metabolites were formed by sequential oxidation of the

1193

tert-butyl group to the alcohol and then to the carboxylic acid and their subsequent conjugation with sulfate and glucuronic acid residues. The half-life for elimination during the depuration phase was about 12 hrwith 98% of the radioactivity cleared within 7 days. Thus, although tebufenpyrad is lipophilic, is readily taken up from water, and might be expected to show a high level of bioaccumulation, it is also rapidly metabolized and cleared which greatly decreases the degree of accumulation. The acylation of the amide nitrogen between the two rings in analogs of tebufenpyrad produces compounds that are improved as acaricides and also show much lower toxicity to fish, perhaps due to differential rates of metabolic deacylation to release the active compound in these organisms (Obata et aI., 1999). Tebufenpyrad binds firmly to soil organic matter with Koc values from 1380 to 4930. Together with its low water solubility, this indicates a very low potential for leaching. It has a moderate persistence in aerobic soils with a half-life of about 1 to 2 months.

57.4 INHIBITORS OF COMPLEX 11 57.4.1 INTRODUCTION Relatively few pesticides are thought to have their primary toxic action through effects on complex 11. Important pesticides that do act in this way are fungicides in the carboxamide (carboxanilide) group. These have been reviewed by Kulka and von Schmeling (1995). In some cases the information regarding these compounds that is available in the open literature or toxicological databases is quite minimal. Based on this rather limited information, they seem to have virtually no notable adverse effects and they have attracted minimal toxicological interest beyond the studies required by regulatory authorities to obtain approval for use. 57.4.2 PROPERTIES OF SPECIFIC COMPOUNDS 57.4.2.1 Carboxamides The forerunner of the group is the fungicide carboxin which was discovered in the mid-1960s (Kulka and von Schmeling, 1995). A relatively large number of carboxamides have subsequently been used as agricultural fungicides, but several of these are not now produced commercially [e.g., benodanil, mebenil, methfuroxam, metsulfovax, and pyracarbolid; Tomlin (2000)]. On the other hand, new members are still being added to the group (e.g., thifluzamide and furametpyr were both first registered for use as pesticides in the 1990s). The carboxamides are systemic fungicides with both protectant and curative actions which are used in seed treatments and in foliar and soil treatments, primarily to control Basidiomycetes. They are often used in mixtures with other fungicides or insecticides. They generally have low or very low toxicities to

1194

CHAPTER 57


terrestrial vertebrates and aquatic species, earthworms, and insects. Their chronic toxicity is equally unremarkable. The lack of recorded incidents of human poisoning appears to substantiate their safety under practical conditions of use. However, the toxicological information available in the open literature is quite limited for most of these compounds. They are rapidly metabolized and eliminated by mammals. It is interesting to note that fungicides within the carboxin group have proved effective as potential therapeutic agents for the human immunodeficiency virus (HIV-l) by inhibiting the viral reverse transcriptase (Bader et aI., 1991). Considerable work has followed from this discovery but this lies outside the scope of the chapter. Succinate-ubiquinone oxidoreductase was shown to be the target for carboxin and oxycarboxin in fungi 30 years ago (Mathre, 1971; White, 1971). The extensive work on this topic is reviewed in detail by White and Georgopoulos (1992) and Schewe and Lyr (1995). Despite the statement by Schewe and Lyr (1995) that mammalian mitochondria "show a sensitivity to carboxins comparable to those of sensitive fungi," a view that is also supported by the work of Mowery et al. (1976, 1977), other results indicate that complex 11 from vertebrate mitochondria is often significantly less sensitive to inhibition by carboxamides than that of target fungi (Mathre, 1971; Motoba et aI., 1988; Shimizu et aI., 1992). It appears that results vary with the type of preparation, the electron acceptor used in the assay, and the structure of the carboxamide tested. The complex 11 from nontarget fungal and plant mitochondria is generally found to be markedly less sensitive than that of target fungi. This allows for the use of these compounds in plant protection without phytotoxicity (Schewe and Lyr, 1995; Shimizu et aI., 1992). The precise binding site for the carboxamides is not known. Solubilized succinic dehydrogenase from bovine heart or fungal mitochondria, consisting of the FP and IP subunits (Fig. 57.3), is not inhibited by carboxin and other carboxanilides (Mowery et aI., 1976; Schewe and Lyr, 1995; Shimizu et aI., 1992), so the binding site probably lies within the membrane-associated anchor portion of the complex. Inhibition arises from a disruption of the transfer of electrons from the iron-sulfur center to ubiquinone, a situation analogous to the acaricide-insecticides that inhibit complex I (Section 57.3.1). This is in accord with photoaffinity labeling studies that indicate that carboxamides bind to a site associated with the ubiquinone-binding proteins of complex 11 (CII-3 and CII-4, Fig. 57.3) and not to the flavoprotein-containing or iron-sulfur proteins (White and Georgopoulos, 1992). A recent study by Matsson et al. (1998) using carboxamide-resistant mutants of Paracoccus denitrificans concluded that the key mutation is in one of these two membrane-located anchor polypeptides at a location adjacent to the Fe-S cluster. It was suggested that this may form part of the carboxamide binding site. On other hand, in the fungus Ustilago maydis, a study of resistance to carboxin showed that it is associated with a point mutation in the gene encoding the third iron-sulfur cluster in the IP subunit (Keon et al., 1994). These authors concluded that carboxamides are probably interposed between this high poten-

fY

Handbook of Pesticide Toxicology, Two-Volume Set: Principles and Agents: Pt. 1 & 2

Hayes' Handbook of Pesticide Toxicology, Two-Volume Set, Third Edition volume 1 & 2

Pesticide Toxicology and International Regulation (Current Toxicology)

Handbook of Toxicology of Chemical Warfare Agents

Elementary Set Theory: Pt. 1

The Blue Annals (Pt. 1 & 2)

Principles and Methods of Toxicology

Principles of Biochemical Toxicology

Principles of Toxicology Testing

Principles of Environmental Toxicology

Principles of toxicology

The Principles of Toxicology

Principles of Toxicology Testing

Stock & Station Agents' Handbook

Poisoning and Toxicology Handbook

Handbook of toxicology

Principles of Bone Biology 2 Vol. Set

Clinical Toxicology: Principles and Mechanisms

Clinical Toxicology: Principles and Mechanisms

Chemical Warfare Agents: Toxicology and Treatment

Encyclopedia of African American Education (1 & 2 Volume Set)

Principles and Practice of Mixtures Toxicology

Principles and Practice of Skin Toxicology

Principles and Practice of Skin Toxicology

Handbook of Spectroscopy (2 Vol. Set)

Accountants' Handbook, 2 Volume Set (Accountant's Handbook)

Encyclopedia of Toxicology, Four-Volume Set, Second Edition: 1 - 3

Handbook of Spectroscopy (2 Vol. Set)

Handbook of Analytical Techniques, 2 Volume Set

Esercitazioni di matematica Vol. 2, Pt. 1

Handbook of Semiconductor Technology (2 Volume Set)

Handbook of Pesticide Toxicology, Two-Volume Set: Principles and Agents: Pt. 1 & 2

Hayes' Handbook of Pesticide Toxicology, Two-Volume Set, Third Edition volume 1 & 2

Pesticide Toxicology and International Regulation (Current Toxicology)

Handbook of Toxicology of Chemical Warfare Agents

Elementary Set Theory: Pt. 1

The Blue Annals (Pt. 1 & 2)

Principles and Methods of Toxicology

Principles of Biochemical Toxicology

Principles of Toxicology Testing

Principles of Environmental Toxicology

Principles of toxicology

The Principles of Toxicology

Principles of Toxicology Testing

Stock & Station Agents' Handbook

Poisoning and Toxicology Handbook

Handbook of toxicology

Principles of Bone Biology 2 Vol. Set

Clinical Toxicology: Principles and Mechanisms

Clinical Toxicology: Principles and Mechanisms

Chemical Warfare Agents: Toxicology and Treatment

Encyclopedia of African American Education (1 & 2 Volume Set)

Principles and Practice of Mixtures Toxicology

Principles and Practice of Skin Toxicology

Principles and Practice of Skin Toxicology

Handbook of Spectroscopy (2 Vol. Set)

Accountants' Handbook, 2 Volume Set (Accountant's Handbook)

Encyclopedia of Toxicology, Four-Volume Set, Second Edition: 1 - 3

Handbook of Spectroscopy (2 Vol. Set)

Handbook of Analytical Techniques, 2 Volume Set

Esercitazioni di matematica Vol. 2, Pt. 1

Handbook of Semiconductor Technology (2 Volume Set)

Recommend Documents