Factors Influencing Sludge Utilization Practices in Europe

FACTORS INFLUENCING SLUDGE UTILISATION PRACTICES IN EUROPE

Proceedings of a Round-Table seminar organised by the Commission of the European Communities, Directorate-General Science, Research and Development, Environment Research Programme, held in Liebefeld, Switzerland, 8–10 May 1985

FACTORS INFLUENCING SLUDGE UTILISATION PRACTICES IN EUROPE Edited by R.D.DAVIS Water Research Centre, Marlow, UK H.HAENI Forschungsanstalt für Agrikulturchemie und Umwelthygiene, Liebefeld, Switzerland and P.L’HERMITE Commission of the European Communities, Brussels, Belgium

ELSEVIER APPLIED SCIENCE PUBLISHERS LONDON and NEW YORK

ELSEVIER APPLIED SCIENCE PUBLISHERS LTD Crown House, Linton Road, Barking, Essex IG11 8JU, England This edition published in the Taylor & Francis e-Library, 2005. “To purchase your own copy of this or any of Taylor & Francis or Routledge’s collection of thousands of eBooks please go to www.eBookstore.tandf.co.uk.” Sole Distributor in the USA and Canada ELSEVIER SCIENCE PUBLISHING CO., INC. 52 Vanderbilt Avenue, New York, NY 10017, USA WITH 24 TABLES AND 30 ILLUSTRATIONS © ECSC, EEC, EAEC, BRUSSELS AND LUXEMBOURG, 1986 British Library Cataloguing in Publication Data Factors influencing sludge utilisation practices in Europe. 1. Sewage sludge as fertilizer— Environmental aspects 2. Slurry— Environmental aspects I. Davis, R.D. II. Haeni, H. III. L’Hermite, P. 363.7′384 TD196.S4 ISBN 0-203-21495-1 Master e-book ISBN

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FOREWORD

This publication constitutes the proceedings of a seminar held at the Federal Research Station for Agricultural Chemistry and Hygiene of the Environment, Liebefeld (Switzerland), 8–10 May 1985, under the auspices of the Commission of the European Communities, as part of the Concerted Action COST 681 ‘Treatment and Use of Sewage Sludge and Organic Waste’. The seminar was convened by Working Party 5 (Environmental Effects) of the Concerted Action to consider factors influencing sludge utilisation practices in Europe. Increased treatment of sewage in Europe to improve the quality of effluents for discharge to rivers inevitably generates, as a by-product, sewage sludge which has to be disposed of safely and economically. Utilisation on agricultural land is an important disposal route for sewage sludge with the advantages of recycling organic matter and supplying nutrients for crop growth. But there are attendant problems associated in particular with possible effects of contaminants in the sludge on crop growth and composition, soil fertility and contamination of groundwater and eutrophication. Management of sludge utilisation on land therefore requires a sound understanding of the associated risks and benefits leading to a scientific basis for disposal policy. The seminar provided a forum for the exchange of recent findings on most aspects of sludge utilisation on agricultural land, including contributions from researchers in Europe, Scandinavia and Canada. Ten papers are included with discussions and conclusions.

CONTENTS

Foreword

v

Geological and pedological factors affecting the heavy-metal load capacity of soils H.KUNTZE (Lower Saxony Geological Survey, Institute of Soil Technology, Bremen, Federal Republic of Germany)

2

Session I

Heavy metal availability in long term experiments C.JUSTE and P.SOLDÂ (Station d’Agronomie, I.N.R.A., Centre de Recherches de Bordeaux, Pont-de-la-Maye, France)

12

Basis for metal limits relevant to sludge utilization E.VIGERUST (Department of Soil Fertility and Management) and A.R.SELMEROLSEN (Chemical Analytical Laboratory, Agricultural University of Norway)

25

Soil ingestion by grazing animals: a factor in sludge-treated grassland G.A.FLEMING (An Foras Taluntais, Johnstown Castle Research Centre, Wexford, Republic of Ireland)

40

Groundwater quality in relation to land application of sewage sludge M.AHTIAINEN (National Board of Waters, Water District Office of North Karelia, Finland)

46

Predictability of estimated mobilizable N pool in sludge and soils A.RUDAZ and S.K.GUPTA (Swiss Federal Research Station for Agricultural Chemistry and Hygiene of Environment, Liebefeld-Berne, Switzerland)

56

Summary of investigations in Italy into effects of sewage sludge on soil microorganisms S.COPPOLA (Istituto di Microbiologia agraria e Stazione di Microbiologia industriale, Università di Napoli, Italy)

62

Effects of long-term sludge additions on microbial biomass and microbial processes in soil S.P.McGRATH and P.C.BROOKES (Soils and Plant Nutrition Department, Rothamsted Experimental Station, Harpenden, UK)

69

Session II

vii

Political and administrative considerations in the formulation of guidance for sludge utilization H.M.J.SCHELTINGA (Ministerie van Volkshuisve sting, Ruimtelijke Ordening en Milieubeheer, Arnhem, The Netherlands) and T.CANDINAS (Forschungsanstalt für Agrikulturchemie und Umwelthygiene, Liebefeld-Bern, Switzerland)

78

Future developments in sludge disposal strategies M.D.WEBBER S.BERGLUND (Swedish Environment Protection Board, Solna, Sweden)

91

Conclusions

102

List of Participants

104

Index of Authors

110

SESSION I

Chairman Rapporteurs

: :

H.HAENI S.GUPTA G.PORTEOUS S.MCGRATH

Geological and pedological factors affecting the heavy-metal load capacity of soils Discussion

Heavy metal availability in long term experiments Discussion

Basis for metal limits relevant to sludge utilization Discussion

Soil ingestion by grazing animals; a factor in sludge-treated grassland Discussion

Groundwater quality in relation to land application of sewage sludge Discussion

GEOLOGICAL AND PEDOLOGICAL FACTORS AFFECTING THE HEAVY-METAL LOAD CAPACITY OF SOILS Herbert KUNTZE Lower Saxony Geological Survey Institute of Soil Technology Bremen

Summary The present German Sewage Sludge Regulations are inadequate to sufficiently differentiate the sources and availabilities of heavy metals in different soils for different plants. Because they are primary constituents of rocks, heavy metals are often present in natural soils in higher concentrations than allowed by the Sewage Sludge Regulations without being plant available. These heavy metals are mobilized, i.e. converted to soluble forms, by the processing of mineral resources or the weathering of rocks and associated soil formation. Important parameters influencing the mobility of heavy metals in soils are:clay content, type of clay minerals, presence of organic material and degree of humification, pH, metal oxides, and salt content. Heavy metals are adsorbed more readily by soils than desorbed. A project has been begun in the Fed. Rep. of Germany to quantify the sources of heavy metals in the soil. Several hundred standard profiles in different subsoil types are being surveyed in terms of soil chemistry and physics using the vegetation typical for each site. The results will be used as the basis for a later survey to map the heavy metal load and capacity with a better geoscientific differentiation in order to revise the German Sludge Regulations. This is also a contribution to the Soil Protection Programme. 1. CRITIQUE OF THE GERMAN SEWAGE SLUDGE REGULATIONS The German Sewage Sludge Regulation of 1983 issued following the Federal Waste Disposal Law of 1972 gives values for the maximum concentration of heavy metals, that may be tolerated in soils. Before sewage sludge is added to soils, the total concentration of Cd, Ni, Hg, Cu, Pb, and Zn are to be determined in both soil and sludge after digestion in hot aqua regia. It is assumed, that this procedure will show the potential maximum, long-term hazard with respect to heavy metals. But the different sources and thus the solubility of heavy metals are not considered, nor are the properties of the soils (pH, clay, iron, and organic matter content) and plant species that influence these. This is inadequate from a pedological and ecological/ chemical viewpoint.

3

FIGURE 1—Heavy metals—Definitions

2. SOURCES AND MOBILITIES OF HEAVY METALS Heavy metals are not artificial substances in our environment, in contrast to many of the organic pollutants. They belong to the natural system of elements (fig. 1). Elements with a density higher than 4, 5 g/cm3 are called heavy metals. Therefore, even iron (Fe) is a heavy metal but a non toxic one. Some heavy metals are essential elements or micro-nutrients (e.g. copper) as long as their concentration remains rather small. Others (e.g. Cd, Pb, Hg) are not essential. They are originally toxic for plants, animals, and especially for men. Using modern analytical methods (e.g. atomic adsorption spectroscopy and X-ray spectrophotometry), heavy metals can be detected in extremely small concentrations (ppm, ppb). In these low concentrations they are called trace elements or micro-elements. Many rocks and minerals contain very small amounts of heavy metals. From there they will come into the soils and the food chain by two ways (see figure 2): a) Weathering and soil forming processes. b) Industrial processing and emissions. Soils become the central part of heavy metal accumulation, mobilization or immobilization within the ecosystems. Heavy metals are subject to changes in their type of binding, solubility and plant availability within ecosystemar cycles. 2.1 GEOLOGICAL FACTORS Heavy metals are present in rocks in the crystal lattice of minerals, e.g. silicates or sulfides. In this state they are nearly insoluble. Several rock types from Northwest Germany are listed in Table I, which can contain certain heavy metals at concentrations in the order of magnitude of the values given in the German Sewage Sludge Regulations. But, the significance of these background values of naturally occurring heavy metals is not yet known. The industrial processing of raw materials that contain heavy metals always involves the danger, that these heavy metals will be released into the environment. Industrial dusts from high temperature processings mostly contain oxides of heavy metals. These are ± soluble. Heavy metal emissions lead to a diffuse distribution; deposition of wastes containing heavy metals or the long term use of sewage sludge in agriculture create a high concentration in one place. The heavy metals in these new accumulations are no

4

FIGURE 2 The Cycle of Heavy Metals

longer bound in their natural insoluble form, but are present as unfixed ion, exchangeable adsorbed by clay minerals and humic substances, as free oxide or sulfate, and thus more easily enter the food chain (s. fig. 2). The various weathering processes also free the heavy metals from their insoluble form in the original rocks. Even rocks, that have a very low heavy metal content, can lead to soils with a high concentration of heavy metals. This enrichment can be explained by the fact, that several meters of limestone must be weathered to produce several decimeters of soil. The primary clay minerals in the soil or the secondary ones formed by weathering, are then available as sorbent for heavy metal ions. Table I Rock Units in Lower Saxony with Elevated Heavy Metal Contents According to Archive Documents of the NLfB sampled till May 1982 Series

Rock Type

L. Cretaceous Lias Keuper

Alb phosphorite clay –

Zn

Pb

Cu

Cd

+ +

+

+

+

Ni

Cr

+ +

+

Co

5

Series

Rock Type

Zn

Pb

Cu

Cd

Ni

Cr

U. Muschelkalk trochite limestone + + + M. Muschelkalk dolomite + + Buntsandstein Hardegsener clay + + + + Zechstein copper schist + + + + + + Rotliegendes basalt + + U. Carboniferous gabbro and shale + + + L. Carboniferous shale + + + + U. Devonian banded shale + + M. Devonian calciola shale + + + L. Devonian shale + (+means concentration greater than the maximum tolerable value according to the German Sewage Sludge Regulations)

Co

+

+

2.2 PEDOLOGICAL FACTORS A natural acidification of the soil takes place in humid climates. This leads to migration (leaching) and at least to decomposition (podsolization) of clay minerals, which in turn results in a mobilization and accumulation of heavy metals in specific soil horizons. Therefore, it is not sufficient to analyze only the top soil for heavy metals as set forth in the German Sewage Sludge Regulation. Also it must be taken into consideration, that heavy metals in an acid soil are more mobile than those in rocks.

FIGURE 3

Heavy metals on ion exchangers are replaced especially by hydrogen ions. Therefore, the pH of the soil solution is important; and this is affected by the input of acids through precipitation, fertilization, plant roots and other biological activities. For example, we have determined a 14 % decrease in heavy metal uptake in test plants by decreasing the acid content of the soil by one pH unit (12).

6

FIGURE 4—Solubility of Cadmium depending on pH and soil constituents acc. Federal Environmental Office Berlin 1980

Anthropogenic heavy metals introduced in the soil via sewage sludge, for example, are highly mobile at first, but are immobilized within a relatively short period of time (Fig. 3). At first the heavy metals remain organically bound in the soil solution. Depending on their binding force and concentration and the cation exchange capacity of the soil, they are adsorbed soon by clay minerals, organic substances, and metal oxides in the soil. Their adsorption and desorption depends on the specific ion exchanger. Mumbrum & Jackson (11) concluded on the basis of infrared studies of copper and zinc smectite, vermiculite, and kaolinite, that heavy metal cations are bound on the OH-groups of layer silicates. Thus, the kind of clay mineral present has more influence than the clay content on how much heavy metal ions will be adsorbed. Figure 4 shows an example. The solubility of Cd increases with decreasing pH more in the presence of Kaolinite than Illite or organic matter. The greatest solubility will be reached in equilibrium with hematite below its ZPC at pHPb, Cr. In a pot experiment increasing temperatures resulted in higher content of Mn, Fe, Zn and Cd, while the content of Ni, Cu and Mo were little influenced. 1. INTRODUCTION The guidlines for sludge utilization with regard to metal limits are very different in different countries. Extra applications of metals to soil through sludge is more doubtful if the strain from other sources e.g. air pollution is high. There can be large variations between districts as regards this strain. Although there are differences in the farming managements, the types of crops grown and the heavy metal strain in different countries, there is probably little basis for such great differences in the guidelines for sludge utilization. Recent, relative extensive research on the heavy uptake from sludge—amended soil has just been carried out in Norway (VIGERUST and SELMER-OLSEN, 1985). Investigations on the effect of temperature on heavy metal uptake have also been reported (SIRIRATPIRIYA, VIGERUST and SELMER-OLSEN, 1985). Even if the problem in Norway may be different from the problems in other countries, the above mentioned investigations may serve as a basis for a more general discussion of the problem. The magnitude of the experimental data may, however, restrict reference in short and simple term.

*Department of Soil Fertility and Management **Chemical Analytical Laboratory Agricultural University of Norway

26

2. EXPERIMENTAL RESULTS a. Sludge for agricultural purpose For agricultural purpose 16 field experiments with increasing applications of sludge were carried out. Most of the trials lasted for 2 growing seasons and results from 28 harvests are refered here. Lime treated sludge is excluded. The main crop was cereal but grass and fodder rape have also been used in some of the experiments. Analyses for heavy metals were carried out on yield from plots without sludge but with N-fertilizer and plots with sludge but without N. These were compared. Mean heavy metal content in sludge used in the trials was, mg/kg DM; Cd 3

Pb 108

Zn 693

Ni 36

Cu 356

Mn 271

Cr 71

Hg 3

The mean values obtained here are in close agreement with the mean and median values from the analyses of sludge from different sewage purification plants in Norway (HALLBERG and VIGERUST, 1981). The content of heavy metal in sludge used in the trials can be characterized as low but representative. For reasons of simplification mean values for each metal, based on all the harvests and all crops, were calculated and expressed as relative figures (concentration for treatment without sludge=100). The relative figures for each sludge amount was calculated based on just the same experiments with control treatments. Table 1. Heavy metal content in plants from field experiment, expressed as relative numbers, without sludge=100. Numbers of experiment harvests in brackets. Sludge, tons DM per hectare Cd Pb Zn Ni Cu Mn Fe

0

15

20

45

60

120

100 100 100 100 100 100 100

65(3) 97(2) 109(4) 89(4) 92(4) 100(3) 101(2)

91(6) 93(3) 131(11) 111(11) 119(11) 102(5) 130(3)

93(14) 99(2) 141(18) 99(16) 130(18) 112(12) 105(8)

108(7) 104(1) 166(8) 145(8) 128(8) 120(4) 110(1)

112(10) 208(12) 154(11) 134(12) 124(7) 96(2)

Sludge application has, only to a little degree changed the Cd-content of the plants while the content of Pb was practically unchanged. The Zn amounts in the plants have been greatly increased by increased sludge applications but the levels were scarcely phytotoxic. Sludge application rates of 60 and 120 t/ha DM lead to a marked increase of the Ni content in plants. This was observed in 2 experiments in the first year. This can partly be attributed to the decline in pH upon sludge application, but this pH effect lasted for only one year. Mn and Cu levels in the plants were continuously raised by increasing sludge doses. Crop response to sludge application was somewhat greater for meadow grasses and rape than for cereal grains.

27

b. Trials with different crops grown on 0, 10 and 40 cm decomposed sludge Green areas commonly require higher applications of sludge compost than agricultural land. Our trials indicate that a 4–10 cm layer is adequate. As our experiments on the use of sludge on agricultural land indicated that the metal content in yields, except for Zn, is changed within a certain margin of error, it was considered necessary to lay out our present trials using high rates viz: 0, 10 and 40 cm layers of decomposed sludge. The 10 cm layer was incorporated into the top 20 cm of the soil. The soil was a sandy loam with 8% organic matter. Altogether 28 crops were grown for each treatment. Some of the crops were grown for 3 years, some for 2 and some for only 1 year. The heavy metal content in the sludge used was, mg/kg DM: Cd 5.5

Pb 271

Zn 516

Cu 492

Cr 148

Mn 268

Ni 42

pH in the top soil was ordinarily 6.2. During the first summer a drop in the pH in pure sludge to 5.0 was observed in the upper 15 cm. Even though a rise in the pH was later recorded in sludge and sludge-treated soils, this never reached the pH level of the top soil without sludge. Cd, Fig. 1. The Cd-content in different plants and different parts of the plants differed greatly. The content in the plants as little influenced by the Cd-amounts applied with the sludge. Crops such as lettuce, maize and red beets had, however, high Cd contents as a results of sludge treatment. Cereal grains and potato tubes, important food crops in Norway, had low Cd-contents and, compared to the other food stuffs, were little affected by the treatment. Pb Sludge application increased the Pb content in lettuce while the other crops were little or not affected. Zn, Fig. 2. The Zn-content in plants increased with increased amounts of applied sludge. Crops such as fodder beet, lettuce, red beet and red clover had especially high Zn-content in the leaves. Some of these crops did not grow well,probably because of toxic Zn-effect. High rates of sludge or waste compost may result in prolonged crop failure and the Zn-content may be a decisive factor in this connection. Ni, Fig. 3. Increasing sludge applications have enhanced the nickel content in the plants. The increase in the plants was, however, considered moderate in relation to the high amounts of applied Ni,. although the element is regarded as relatively easily plant available. Cu, Fig. 4. Variations in sludge applications did not lead to great differences in Cu-content in plants. Plant response to Cu was not as regular as was found for Cd and Zn. Mn, Fig. 5. Incorporating a 10 cm sludge layer into the soil resulted in only a slight change in the Mn-contents in the plants. Grown in pure sludge on the other hand resulted in high concentrations of the metal in the plants or parts of the plants. This may be attributed to sludge decomposition, pH-changes and red/ox conditions leading to a greater availability of Mn. Fe There was a clear tendency that high sludge applications resulted in decreased Fe-content in the plants. This is probably connected to the sludge decomposition and the effect can be altered with time. DOWDY and LARSON (1975) pointed out that sludge applied to acid soil decreased the Fe-content in the plants.

28

They supposed that the reason might be that Fe forms complexes with other compounds in the soil and is tendered less available. Other elements The treatment had little or no effect on the content of Cr, Hg or Co in plants. Grown in pure sludge enhanced the B-content in the vegetative parts of the plants. For barley, wheat and partly oat, this lead to symptoms of B-toxicity which became less after one or two years. Recapitulatior The mean content of each metal in all the crops including parts of the plant has been calculated and presented as a guide (table 2). FIGURE 1—mg Cd pr. kg dry matter Table 2. Concentration of heavy metals, mean values for different crops and part of plants. Treatment with sludge calculated as relative numbers, concentration without sludge=100. Applied sludge

Cd Pb Zn Ni Cu Cr Mn

0 mg/kg

10cm relative

40cm numbers

0.5 3.4 46 2,4 10.6 1.0 53

127 115 269 130 114 110 109

168 127 420 218 157 110 485

Increasing amounts of sludge affected the content of the different heavy metals in this order: Zn>Mn*>Ni>Cd>Cu>Pb, Cr Some crops as beans, red beet, maize and partly lettuce did not grow quite well in pure sludge. This might be caused by a toxic effect of one or more of the heavy metals. Several crops showed higher concentrations of Zn and partly Cu than what DAVIS and CARLTONSMITH (1980) and SAUERBECK (1982) give as indication of critical limits. The content of Ni, Cd or Pb, however, did not exceed their indicated limits. c. Effect of liming Application of 8000 kg CaO/ha to pure sludge increased pH only 1 unit, As an average for different crops this lime application only reduced the metal content in plants by 5–20% as compared to unlimed sludge. This effect of liming on metal uptake must be characterized as very little compared to the common lime effect on soils. It could probably be related to the increased decomposition as a result of liming and the metal-fixing power of intermediate decay products. According to CHRISTENSEN and TJELL (1983) increased CO2 production might also lead to metal precipitation. d. Effect of organic matter In a frame experiment the following yields of ryegrass, g/m2, was obtained: a) b) c) d)

Control (including N-fertilizer) 60 tons sludge/ha (No N-fertilizer) Heavy metals as b), given as chlorides (N-fertilizer) Sludge 20cm layer on top (No N)

384 389 322 491

29

FIGURE 2—mg Zn pr. kg dry matter

e)

Heavy metals as d), given as chloride (N-fertilizer)

0

30

FIGURE 3—mg Ni pr. kg dry matter

31

FIGURE 4—mg Cu pr. kg dry matter

32

FIGURE 5—mg Mn pr. kg dry matter

These results show that the plants can tolerate much higher amounts of metals when they are part of the sludge compared to application as inorganic salts. e. Effect of temperature

33

Fig. 6. The effect of declining temperatures (24–9°C) upon heavy metal content in lettuce expressed relatively; concentration at 24°C=100. (From: SIRIRATPIRIYA, VIGERUST and SELMER-OLSEN, 1985).

Temperatures: 24, 21, 18, 15, 12 and 9°C

Pot trials conducted in phytotrone showed that different temperatures (9, 12, 15, 18, 21 and 24°C) have an effect on the content of some heavy metals in lettuce (fig. 6) Decreasing temperatures very clearly lowered the content of Zn, Mn, and Fe in the plants. This was also true for Cd but not so regular for both harvests. Ni, Cu and Mo were not significantly affected by temperature. 3. RESULTS OF SOME PUBLISHED INVESTIGATIONS Published results from trials with sludge are collected and the mean values are calculated for different crops and presented in Tab. 3. Although reservations must be made for such a presentation, it is desirable to give a picture of the effects. It is obvious that high variations in sludge application and composition of sludge exist in the material. In some of these trials sludge with high or even very high content of heavy metals was used, and only in a few trials the sludge was at the same low level as sludge used in our investigations. For each crop an average metal content in plants is calculated for treatments without sludge (mg/kg DM). For treatment with sludge these mean values are given as relative numbers, concentration without sludge= 100. Tab. 3 represent data from 26 publications, references number: 1, 2, 4, 6, 7, 8, 12, 13, 16, 17, 18, 20, 2 1, 22, 23, 25, 26, 27, 28, 29, 31, 32, 34, 35, 36, 38. The calculations include more than 117 experiment harvests, The calculated average of sludge application rates was high, and is given in a separate column. Tab. 3 indicates that application of sludge has affected the content in plants in this order: Zn>Cd>Ni>Cu>Pb, Hg, Cr Compared to this data our results showed less content of most of the metals from sludge treated soils, especially Cd and Cu. * The Mn-content was little changed at 10 cm, while pure sludge gave high contents in plants.

34

Tab. 3. Average content of metals in crops for treatment without sludge (O) (mg/kg DM) and relative numbers for treatment with sludge (S) , without sludge=100. Numbers of experiment referred in brackets. Cd O

S

BAR LEY

O

Pb

Hg

Ni

Zn

Cu

Cr

S

O

S

O

S

O

S

O

S

O

S

tons DM, ha

grain 0.10 (10)

130

0.90 (6)

100

50

0.3 (5)

133

58 (13)

109

4.9 (13)

173

0.33 (4)

106

straw

0.17 (9)

158

300 (9)

113

0. 024 (8) 95

1.3 (7)

92

36 (9)

208

4.9 (11)

116

0.90 (4)

89

57

OAT

grain 0.11 (6)

163

100

1.2 (2)

446

35 (6)

140

3.6 (6)

131

0.42 (2)

95

straw

0.12 (6)

158

1.70 (6)

135

0. 010 (2) 93

1.7 (6)

100

54 (6)

117

6.1 (6)

125

0.38 (6)

126

50

Whea t

grain 0.07 (22)

214

67

5.4 (10)

259

36 (17)

228

4.3 (17)

221

0.30 (6)

93

straw

0.19 (11)

1.10 (11)

91

0. 030 (2) 84

0.6 (7)

150

23 (12)

700

3.3 (9)

239

429 (6)

80

139

0.9 (6) 106 2.4 (8)

189

25 (6) 58 (9)

240

4.6 (6) 6.5 (9)

172

1.10 (4) 0.85 (2)

127

60

88

40

132

2.8 (6)

204

73 (6)

223

7.4 (6)

142

0.75 (2)

61

73

150

1.2 (7)

225

50 (7)

610

5.4 (7)

361

0.33 (7)

148

103

9.0 (4) 121

108

46 (7) 135

135

6.5 (6) 109

122

91

140

1.0 (2) 8.8 (2) 100

170

21 (12) 82 (4) 125

129

200

RYE

grain 0.15 186 (6) GRASS 0.21 105 (8)

0. 041 (7) 0.42 90 (4) 0. 042 (6) 0.44 61 (13) 0. 148 (3) 0.93 108 (2) 2.40 100 (8)

CLOVER

0.13 (6)

223

131 (6)

100

RAPE

0.32 (7)

213

0.78 (7)

85

0.77 (7) 105

187

1.83 (3) 5.15 107 (2)

86

CAR ROT tops

POT ATO RED BEET tops BEA NS

root 1.53 (4)

0. 020 (1) tuber 0.23 113 1.50 133 (4) (2) root 0.35 287 15.0 89 (4) (2) 1.20 292 15.0 67 (4) (2) seed 0.18 194 1.57 51 (10) (3)

0. 035 (4) 0. 041 (2) 0. 014 (1)

160

4.3 (4)

8.2 (2)

4.6 (3)

129

63 (4)

202 450 (2) 161

121

8.5 (4)

131

1.10 (2) 1.60 119 (2)

5.1 143 (12) 122 13.9 89 (2) 27.6 97 (2) 36 150 4.4 189 (10) (8)

1.20 (1)

57

70

114

55

100

90 29

29 54

35

Cd O

S

veg. 0.50 part (2) TOM fruit ATO veg. 0.80 part (3) SAL AD Mean rel. numbers

O

Pb S

O

5.10 92 (2) 0.46 235 5.60 (8) (6) 75 5.60 98 (3) 0.81 285 4.6 (19) (11) 184

Hg S

O

160

Ni

Zn

S

O

S

O

S

O

5.0 (2)

52

34 (8)

259

6.4 (6) 146

173

132

117 98

4.42 (2)

Cu

100 104

4.2 (6)

30 (3) 119 164

24 (8) 140 43 (11)

6.7 (3) 300 217

Cr S

tons DM, ha 56

8.3 (7) 128

133

10.0 (13)

120

180 90

156

79 102

78

4. DISCUSSION Different factors are important when considering the actual limits and regulations in sludge utilization: – Utilization of sewage sludge is especially precarious where the heavy metal strain as a whole is high. – Results obtained in sludge experiments in Norway clearly show that only a very little part of applied metals through sludge is taken up by the plants per year. This indicates that a possible problem might be long lasting. According to ELINDER et al (1978) the Cd-concentration in kidney cortex of Swedes in 1974 were not at a level where injury could be expected. This can, however, be changed in the future if the Cd-strain is increased. In addition it is shown that Cd in kidney accumulates with age (l.c.). This also underlines that the strain over several years are important. – The most effective way to prevent problem caused by sludge is to ensure that only sludge with low concentration of critical metals are allowed to be spread. – Limitation of the yearly amount of sludge which can be spread will ensure good distribution of metals applied. This will, however, not reduce the total amounts of metals applied to agriculture land and will scarsely change the total metal-content in the agricultural products as a whole. By spreading sludge to area where e.g. cereals are the main crop, an eventual enhanced content may very effectively be “diluted”, because the grain from different areas are mixed. Also products which are marketed (sold) are over years “mixed” with others and thus an effective dilution takes place. – In general the heavy metal uptake is higher in vegetables than in e.g. grain of cereals. The total amount of metals brought into circulation by the products are relatively high for vegetables. In Norway we recommend that sludge should not be used on area where vegetables normally are grown. – Utilization of sludge on typical green areas should not involve any risk for contamination of food. In the free marketing of sludge—or waste—compost there is a certain risk for use of high doses in private gardens including soil where vegetables are grown for their own use. – Composting is expensive and the product ought to be sold if this treatment shall be an actual alternative. The marketing is, however, difficult if certain regulations are required.

36

– High doses of sludge, which seem to be of special interest on green areas, involve a certain risk for phytotoxic effects. Our investigations indicate that, at a normal ratio in concentration between different metals in sludge, this risk follows this order: Zn>Cu>Ni Mn-toxicity may occur if the compost is unripe or if anaerobic conditions exist especially in the first one or two years. – Liming is an effective way to reduce the plant availability of most of the heavy metals, according to our results in this order: Mn>Zn>Ni, Cd>Cu>Cr, Pb, Fe Liming reduces the metal uptake by plants in the short term. The metals, however, are preserved in the top soil and could be potentially available. The problem can therefore be more long lasting. In areas or countries where the heavy metal problem is considered only as a long term risk, little is gained by requiring a certain limestatus in soil when sludge is to be spread. Where metals, for different reasons, might have toxic effects, liming can be an effective counteractor regardless of the fact that the toxicity is to plant, animal or man. – Cd is considered to be the most precarious as far as health risks are concerned. The plant uptake of the metal, however, is reduced by increasing the amount of available Zn in soil, (CHANEY 1974). It is desirable that the Cd-content in the sludge does not exceed 0.5% of the Zn-content (l.c) According ELINDER and KESSLER (1983) the Cd absorption from the gastrointestinal tract is relatively low if there is a simultaneous low intake of, among other metals, Zn. Our results do show that sludge application gives especially high Zn uptake in plants. Zn applied through sewage sludge, therefore, may reduce possible harmful effects of Cd both by reduced uptake by plants and lower absorption from food. Everything which reduces the metal uptake tends to make the problem more long lasting. Liming, high Zn-content, high organic matter content in soil and sludge postpone the uptake in plants. These factors reduce the uptake of Cd and increase the possibilities of dilution in our food. – In our pot experiment increasing temperatures increased the content of Zn, Mn, Fe and Cd in lettuce. If this is true, also under field conditions, one could expect the heavy metal problem being less in colder than in warmer climate. – Considerations of the heavy metal problem from sewage sludge are based on a series of experiments. Many of these experiments are, however, carried out with higher doses of sludge or with higher heavy metal content in the sludge than what is recommended or allowed to be used in practice. Some data may also be based on pot experiments, which are not directly comparable to field experiments (De VRIES et al, 1978; KUNTZE et al, 1983). On one hand it is possible that the experimental data partly overestimate the uptake from sludge. On the other hand short term experiments do not give realistic answers on the future situation when sludge is spread for decades. – We have calculated the amount of sludge produced in different geographical parts of our country in relation to the total area of agricultural land. The amount of sewage sludge produced in Oslo and Akershus per unit cultivated area is about 10 times higher than the equivalent amount in any other county (fylke) in Norway. The total amount of metals applied to land by sludge on agricultural land close to towns are much higher than in typical rural districts. In general the pollution from other sources are also higher in these areas. We also must take into account that it is precisely these areas that are especially important for production of food for the population center of the district. In the nordic countries almost only dewatered sludge is utilized, facilitating transportation over fairly long distances. Liquid sludge which has a high N-value, is often applied in lower doses than dewatered sludge. In the long run, however, liquid sludge can give higher strain than dewatered sludge on the agricultural land close to the

37

bigger towns because of limited transport possibilities. We believe that this type of distribution in the long run is much more precarious than if certain areas receive high doses for 10 or 20 years. – Compared to most of the EEC-countries the heavy metal problems as a whole seem to be moderate in Norway. The metal content in sludge is normally low or very low. The farming management is dominated by one-sided cereal production in areas where use of sludge has particular interest. We also have a relatively cold climate. The farmers are interested in high sludge applications per unit of area. It is, however, strongly recommended to utilize the sludge primarily on soil in poor physical conditions. The sludge should not be used on soils where vegetables normally are grown. REFERENCES 1 2 3 4

5

6 7 8 9 10

11 12 13 14 15 16 17

() ANDERSSON, A. and K.O.NILSSON, 1976. Influence on the levels of heavy metals in soil and plant from sewage sludge used as fertilizer. Swedish J. agric. res., 6:151–159. () BÆRUG, R. and J.H.Martinsen, 1977. The influence of sewage sludge on the content of heavy metals in potatoes and on tuber yield. Plant and Soil, Vol 47, No. 2:407–417. () CHANEY, R.L. 1974. Recommendations for management of potentially toxic elements in agricultural and municipal wastes. USDA National Program Staff Soil, Water and Air Sciences Beltsville, U.D. () CHANEY, R.L., S.B.HORNICK and P.W.SIMON, 1977. Heavy metal relationships during land utilization of sewage sludge in the northeast. Land as a Waste Management Alternative . U.S. Dep. of Agric. Beltsville, Maryland: 283–310. () CHRISTENSEN, T.H. and J.C.TJELL, 1983. Interpretation of experimental results on cadmium crop uptake from sewage sludge amended soil. CEC-Proceedings of the Third International Symposium, Brighton. Sept. 83: 358–369. () CUNNINGHAM, R.D., D.R.KEENEY and J.H.RYAN, 1975. Yield and metal composition of corn and rye forage grown on sewage sludge amended soil. J. Environ. Qual.Vol.4, No. 4:448–454. () DAMGAARD-LARSEN, S., K.E.LARSEN og P.SØNDERGAARD KLAUSEN, 1979. Årlig tilførsel af slam fra rensningsanlæg til landbrugsjord. Tidsskr. f. Planteavl. 83:349–386. () DAMGAARD-LARSEN, S., P.SØNDERGAARD KLAUSEN og K.E.LARSEN, 1979. Engangstilførsel af slam fra rensningsanlaeg til landbrugsjord. Tidsskr. f. Planteavl. 83:387–403. () DAVIS, R.D. and C.CARLTON-SMITH, 1980. Crops as indicators of the significance of contamination of soil by heavy metals. Water Research Centre T.R. 140, 44p. Medmenham Laboratory. () DE VRIES, M.P.C, and K.G.TILLER, 1978. Sewage sludge as a soil amendment, with special reference to Cd, Cu, Mn, Ni, Pb and Zn-Comparison of results from experiments conducted inside and outside a glasshouse. Environ. Pollut, 16:231–240. () DOWDY, R.H. and W.E. LARSON, 1975. Metal uptake by barley seedlings grown on soils amended with sewage sludge. J. Environ. Qual. Vol.4, No.2:229–233. () DOWDY, R.H. and W.E.LARSON, 1975. The availability of sludge borne metals to various vegetable crops. J. Environ. Qual., Vol.4, No. 2:278–282. () DUDAS, M.J. and S.PAWLUK, 1975. Trace elements in sewage sludges and metal uptake by plants grown on sludge-amended soils. Can. J. Soil Sci. 55, 239–243. () ELINDER, C.-G., L.FRIBERG og M.PISCATOR, 1978. Hälsoeffekter av kadmium. Läkartidningen. Vol.75, Nr.47:4365–4367. () ELINDER, C.G. and E.KESSLER, 1983. Toxicity of metals. CEC-Proceedings of a Seminar Uppsala 1983. () FURR, A.K., W.C.KELBY, C.A.BACHE, H. GUTENMANN and D.J.LISK, 1976. Multielement absorption by crops grown in pots on municipal sludge-amended soil. J. AGRIC. Food Chem. 24,889–892. () GAYNOR, J.D. and R.L.HALSTEAD, 1976. Chemical and plant extractability of metals and plant growth on soils amended with sludge. Can. J. Soil Sci. 56:1–8.

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18 19 20 21 22

23 24

25 26

27 28 29 30

31 32 33 34 35 36 37 38

() GIORDANO, P.M., J.J.MORTVEDT and D.A.MAYS, 1975. Effect of municipal wastes on crop yields and uptake of heavy metals. J. Environ. Qual. Vol. 4:394–399. () HALLBERG, P.A. og E.Vigerust, 1981. Slamdisponering 3. Tungmetaller i kloakkslam. Prosjekt 2.2.15. Utvalg for fast avfall, NTNF, 82s. () HAMAR, T.O., 1974. Kloakkslam til plantedyrking. Hovedoppgave Inst. f. jordkultur, NLH. 127s. () JOHN, M.K. and C.J.VAN LAERHOVEN, 1976. Effects of Sewage Sludge Composition. Application Rate, and Lime Regime on Plant Availability og Heavy Metals. J. Environ. Qual. Vol.5, No. 2:246–251. () KELLING, K.A., D.R.KEENEY, L.M.WALSH and J.A.RYAN, 1977. A field study of the agricultural use of sewage sludge: III Effect on uptake and extractability of sludge-borne metals. J. Environ. Qual. Vol.6, No. 4: 352–358. () KOSKELA, J., 1978. Disposal of municipal sludges containing heavy metals in agriculture. Seminar on Heavy Metals, Technological Methods for the Limitation of Discharges. Copenhagen 4.–7. June. 3.5:1–12. () KUNTZE, H., E.PLUQUET, J.STARK and S.COPPOLA, 1983. Current techniques for the evaluation of metal problems due to sludge . Prosessing and use of sewage sludge. Proc. of the Third International Symp. Birghton. Sept. 1983 D.Reidel. () LATTERELL, J.J., R.H.DOWDY and W.E.LARSON, 1978. Correlations of extractable metals and metal uptake of snop bean grown on soil amended with sewage sludge. J. Environ. Qual. Vol. 7, No.3:435–440 () LINNMAN, L., A.ANDERSSON, K.O.NILSSON, B.LIND, T.KJELLSTRØM and L.FRIBERG, 1973. Cadmium uptake by wheat from sewage sludge used as a plant nutrient source. Archiveo. Environ. Health, Vol. 27:45–47. () MARTINSEN, J.H., 1976. Kloakkslam som gjødsel og jordforbedringsmiddel. Inst. f. jordkultur, NLH. Stensiltrykk. 221s. () MCINTYRE, D.R., W.J.SILVER and K.S.GRIGGS, 1977. Trace elements uptake by field-grown plants fertilized with waste water sewage sludge. Compost Sci. 18, 5:22–29. () SABEY, B.R. and W.E.HART, 1975. Land application of sewage sludge: I. Effect on growth and chemical composition of plants. J. Environ. Qual . Vol. 4, No. 2:252–256. () SAUERBECK, D., 1982. Welche Schwermetallgehalte in Pflanzen dürfen nicht überschritten werden, um Wachstumsbeeinträchtigungen zu vermeiden. Kongressband 1982. Vorträge 94. VDLUFA-Kongress, Münster. Sept. 1982, s.108–129. () SCHAUER, P.S., W.R.WRIGHT and J.PELCHAT, 1980. Sludge-borne heavy metal availability and uptake by vegetable crops under field conditions. J. Environ. Qual. Vol. 9, No. 1:69–73. () SHEAFFER, C.C., A.M.DECKER, R.L.CHANEY and L.W.DOUGLASS, 1979. Soil temperature and sewage sludge effects on metals in crop tissue and soils. J. Environ. Qual. Vol. 8, No. 4:455–459. () SIRIRATPIRIYA, O., E.VIGERUST and A.R.SELMER-OLSEN (in press) Effect of temperature and heavy metal application on metal content in lettuce. Sci. reports of the agric. university of Norway. () SORTEBERG, A., 1972. Kloakkslam og tungmetaller. Norsk Landbruk nr. 22. 7s. () TORP, M., 1980. Kloakkslam til jordbruksformål. Hovedoppgave Inst. f. jordkultur, NLH. 125s . () VALDMAA, K., 1972. Jordbrukets krav på slamkvalität. Åttonde nordiska symposiet om vattenforskning. Nordforsk. Publikation 1972–3: 223–236. () VIGERUST, E. og A.R.SELMER-OLSEN, 1985. Tungmetallopptak ved bruk av kloakkslam. Rapport fra Inst. f. jordkultur, NLH. Serie B 2/85, 58s () ZWARICH, M.A. and J.G.MILLS, 1979. Effects of sewage sludge application on the heavy metal content of wheat and forage crops. Can. J. Soil Sci. 59:231–239.

DISCUSSION KUNTZE (FRG):

You have suggested that the organic matter in sludge causes a more rapid decrease in availability of metals than occurs in soil alone, but here you are comparing two different systems. Soil is better aerated and therefore has a

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higher redox potential than sludge. You also observed that with increasing temperature the crop uptake of cadmium, zinc, manganese and iron increased. Microbial activity and lime in sewage sludge both have a warming effect and contribute towards the higher metal uptake from sludge than from soil. VIGERUST (Norway): We did not grow plants in sludge alone. The temperatures in the soil with and without sludge were similar throughout the seasons during our experiment. SCHELTINGA (Netherlands): You seemed to suggest that heavy additions of metals to the soil was acceptable because of mixing of agricultureal products such as grain and wide distribution in marketing. Is this policy in Norway? VIGERUST: I did not suggest we should allow high levels of metals in food; low levels are essential. We grow only 20% of food and this is mixed with imported food. In general metal levels are low in Norwegian sludges. GOMEZ (France): In your studies on the effect of temperature on metal availability, was there a difference between soil and atmospheric temperatures? VIGERUST: The temperature was uniform for roots and tops in each plant growing room, but different growing rooms were maintained at different temperatures. GOMEZ: You compared the effects of metal salts and metal-containing sludge. The sludge contained large amounts of iron, did you add iron in the metal salt treatments? VIGERUST: No we did not add iron. GOMEZ: Can you explain the initial drop in pH after sludge addition? VIGERUST: The change in pH was not understood. In some cases it increased, and in others it decreased. FROSSARD (Switzerland): Was the humidity in the laboratories controlled? VIGEREST: We tried to control humidity. Water consumption was greater at higher temperatures. FROSSARD: The transport of elements such as copper and nickel by mass transfer in the transpiration stream is affected by temperature.

SOIL INGESTION BY GRAZING ANIMALS; A FACTOR IN SLUDGE-TREATED GRASSLAND G.A.Fleming, An Foras Taluntais, Johnstown Castle Research Centre, Wexford, Republic of Ireland INTRODUCTION While the application of metal-rich sewage sludge to land has received considerable attention in recent years most studies have been concerned with its incorporation into tillage soils. On grassland different conditions obtain necessating different approaches and strategies for application. While its effect on grass production may be considerable (O’Riordan 1983) the surface accumulation of metals in grass swards can be undesirable under grazing systems. Some indication of the extent of this surface accumulation is shown in Fig. 1.

Fig. 1. Surface accumulation of zinc in a grassland soil after application of a high Zn sludge. (O’Riordan 1983).

SOIL INGESTION BY ANIMALS It is well known that involuntary ingestion of soil together with herbage takes place during grazing. The beneficial effects of such ingestion were recognised in the eighteenth century (Fraser 1794) when cobalt deficiency in sheep in England was successfully treated by dosing with a suspension of soil in water.

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Likewise the classical cure for piglet anaemia involved the provision of upturned sods from which the young pigs received sufficient amounts of iron to offset the condition. On the other hand increased tooth wear of sheep resulting from ingested soil has been reported (Healy and Ludwig 1965, Nolan and Black 1970). The amount of soil ingested by grazing animals varies to a considerable extent and close grazing animals such as sheep may under extreme conditions ingest over 400 gm of soil per day (Field and Purves, 1964). Normally amounts are much smaller and typical data are shown in Table 2. Table 1. Intake of herbage and soil by animals. Animal

Herbage

Soil

(gm/day D.M.) Sheep Cow D.M.=Dry matter.

600 13,000

Soil intake % of D.M.

100 1,000

14 8

FACTORS AFFECTING SOIL INGESTION Soil type.

Soils with good structure are associated with low soil contamination of pasture and vice versa. When structure is poor and drainage bad, treading by animals results in much poaching. Consumption of grass is then attended by relatively large soil intakes by animals. In modern farming practice, however, soils with good drainage and structure are more likely to suffer from over stocking and thus soil ingestion becomes of greater consequence. Stocking rate and Season

It is to be expected that soil intake by animals will increase when stocking rate is increased. Likewise seasonal effects can be anticipated; when for instance rainfall becomes more abundant, soil contamination of pasture is enhanced and thus soil intake rises. Some quantification of these factors is shown in Fig. 2. In this experiment (McGrath et al. 1982), sheep at two stocking rates were used and soil intake measured. In both the high (HSR) and the low (LSR) stocking rates, animal numbers were reduced as the season advanced and grass on offer became more scarce but the mean stocking rates were—HSR 30 and LSR 20 sheep per hectare. The data clearly indicate greater soil ingestion by the animals on the high stocking rate while for both rates ingestion increased as the season advanced and the quantity of grass declined. This is especially evident in the case of the highly stocked animals. Stock management.

In modern management systems animals may be either set stocked or rotationally grazed. In the former instance a group of animals are confined to a grazing area until all grass has been consumed while in the latter case animals are moved at frequent intervals—often daily—and they are returned to the original paddock or grazing area in three to four weeks. The less frequent the changes the closer is the approximation to set stocking. The increase in soil intake by sheep on a seven day grazing cycle as opposed to a twenty-eight day one is shown by the data of Table 2.

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Fig. 2. Effects of stocking rate and season on amounts of soil ingested by sheep. (McGrath et al. 1982). Table 2. Soil ingestion and grazing pattern1 Stocking rate ewes/ha

Grazing cycle (days)

Soil in faeces (% D.M.)

37 37 1 Healy (1973).

7 28

37 10

Other factors influencing the degree to which soil is ingested include the soil content of conserved material (silage) and the duration of supplementary winter feeding (meals and concentrates). Earthworn activity

Increased dry matter production is associated with greater earthworm activity and in wet periods especially, more worm casts appear on the soil surface. Movement of animals spreads these casts with resultant contamination of pasture and cast height and shape facilitates their consumption. Worm casts also tend to proliferate in open type swards rather than on those having a surface mat. Newly reseeded pastures of high producing ryegrasses are characterically of the open sward type. MEASUREMENT OF SOIL INGESTION The amounts of soil ingested by grazing animals can be measured by estimation of the soil content of faeces. This may be done by collection of fresh faecal material in the field or more precisely by collection of faeces from animals fitted with a suitable harness. In this case, daily faecal output can be readily measured and determination of the ash content of the faeces provides an estimate of soil content. A refinement of this method involves treatment of the ash with dilute mineral acid thus correcting for ash contribution from undigested herbage. The acid insoluble residue (AIR) is then taken as a measure of soil ingested and is related to it by the formula.

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AIR in faeces is normally calculated by the method of Healy and Ludwig (1965). About 4 gm of faeces are ashed overnight and a portion of the ash (usually 0.5 gm) digested with 6 N HCl (20 ml) on the water bath for 1 hour. The residue is then washed free of acid on filter paper and reignited in the furnace. AIR is calculated as a percentage of digested ash in oven-dry faeces. Soil in faeces may also be measured by relating the titanium (Ti) content of faeces to that of the soil. The method is based on the principal that the Ti content of faeces can be taken as being essentially derived from soil only. This assumption is based on the known wide disparity between the Ti level in soils (0.2–1%) compared with that in uncomtaminated herbage (0.5–2 ppm). Soil ingestion is then calculated from the following formula. (Thornton and Abrahams 1983).

Titanium may be determined by atomic absorption or by X-ray fluorescence (XRF) as described by Healy (1968). SOME NUTRITIONAL IMPLICATIONS FOR ANIMALS It is now recognised that ingested soil is an important pathway for metal absorption by animals especially in areas which are naturally contaminated (geochemical pollution) or where enhanced metal levels have resulted from industrial activities. Ingested soil as it passes through the animal’s alimentary tract it is subjected to a range of conditions inherent in the digestive process. In particular the differences in pH between the abonasum (c. pH 3.0) and the rumen (c. pH 7.0) are important and must exercise significant effects on metal absorption. Other processes such as complexation may also be involved. Alimentary tract conditions therefore may increase or decrease heavy metal availability to the animal. Lead poisoning in cattle in Ireland has resulted from consumption of soil-contaminated beet tops from an old mining area and from dust blown on to soil and herbage adjacent to mine tailings ponds. Studies in England in areas where metalliferous mining and smelting were carried out in the past, have revealed that under the prevailing conditions up to 80% of lead and up to 90% of arsenic assimilated by cattle could be attributed to ingested soil (Thornton 1983). Similarly selenium poisoning of animals which was reported in Ireland many years ago (Fleming 1962) was subsequently observed to be more prevalent in dry summers (Twomey et al. 1977). Under such conditions when forage becomes scarce animals ingest greater quantities of soil than normal, and this increased soil intake would seem to offer at least a partial explanation for the observed increase in selenosis in dry summers.

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Even where soils do not contain anomalous levels of elements it is thought that soil ingestion exercises some effects. Thus Statham and Bray (1975) considered that iodine intake by sheep from ingestion of a clay soil was greater than that from a sandy one, as sheep grazing on the latter had more goitrous lambs than those on the former. One of the most common trace element disorders associated with livestock production is that of copper deficiency. This is frequently associated with high levels of molybdenum in herbage and is exacerbated by high intakes of sulphur. Iron has now been implicated in the complex mechanisms that bring about hypocupraemia in animals (Suttle et al. 1982) and this being so, soil intake is brought into sharp focus, as it can be the major source of ingested iron. Recently it has been suggested that a possible mechanism for copper depletion may first involve precipitation of FeS in the rumen followed by release of sulphide in the more acid abomosum milieu with resultant trapping of potentially absorbable copper as insoluble CuS (Suttle et al. 1984). Other elements such as zinc and cadmium may also be involved and can also act as antagonists towards copper absorption. The occurrence of the swayback condition in lambs results from a copper shortage in the parent ewe. The condition has been related to the length of time of snow cover on winter pastures insofar as less swayback is experienced when snow lies on the ground for relatively long periods. Under these conditions ewes obviously ingest less soil and also receive more supplementary feed thereby increasing their chances of receiving more copper (Thornton 1983). The above examples serve to illustrate some of the effects of soil ingestion on animals. Disposal of sewage sludge on grassland must take these factors into account and—apart from the more obvious dangers associated with organic pathogens—the possible repercussions from enhanced heavy metal intake from soil ingestion must be considered. REFERENCES FIELD, A.C. and PURVES, D. (1964). The intake of soil by grazing sheep. Proc. Nut. Soc. 23:24. FLEMING, G.A. (1962). Seleniun in Irish soils and plants. Soil Sci. 94: 28–35 28–35. FRASER, R. (1794). General view of the county of Devon. Quoted by Hopkirk, C.S.M. and Patterson, J.B.E. “The story of cobalt deficiency in Animal Health”. Mond-Nickel, Co. London. HEALY, W.B. (1968). Ingestion of soil by dairy cows. N.Z. Jl. agric. Res. 11:487–499. HEALY, W.B. (1973). Nutritional Aspects of Soil Ingestion by Grazing Animals. In “Chemistry and Biochemistry of Herbage” Vol. 1. ed. G.W. Butler and R.W.Bailey. Acad. Press London, 567–588. HEALY, W.B. and LUDWIG, T.G. (1965). Wear of sheep’s teeth 1. The role of ingested soil. N. Z. Jl. agric. Res. 8: 737–752. MC GRATH, D., POOLE, D.B.R., FLEMING, G.A. and SINNOTT, J. (1982). Soil ingestion by grazing sheep. Ir. J. agric. Res. 21:135–145. NOLAN, T. AND BLACK, W.J.M. (1970). Effect of stocking rate on tooth wear in ewes. Ir. J. agric. Res. 9:187–196. O’RIORDAN, E.G. (1983). The chemical composition of Irish sewage sludges and their effects on soil properties and grass growth. Unpub. Ph. D. thesis, University College, Dublin pp 591. STRATHAM, M. and BRAY, A.C. (1975). Congenital goitre in sheep in Southern Tasmania. Aust. J. agric. Res. 26: 751–768. SUTTLE, N.F. ABRAHAMS, P. and THORNTON, I. (1982). The importance of soil type and dietary sulphur in the impairment of copper absorption in sheep which ingest soil. Proc. Nut. Soc. 41:83A. SUTTLE, N.F. ABRAHAMS, P. and THORNTON, I. (1984). The role of a soil x dietary sulphur interaction in the impairment of copper absorption by ingested soil in sheep. J. agric. Sci. Camb. 103:81–86.

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THORNTON, I. (1983). Soil-plant-animal interactions in relation to the incidence of trace element disorders in grazing livestock. In “Trace Elements in Animal Production and Veterinary Practice”. Occ. Publ. No. 7 Br. Soc. Anim. Prod. Ed. N.F.Suttle, R.G.Gunn, W.M.Allen, K.A.Linklater and G.Weiner. BSAP Edinburgh 39–49. THORNTON, I. and ABRAHAMS, P. (1983). Soil ingestion—a major pathway of heavy metals into livestock grazing contaminated land. Sci. Total Envir. 28:287–294. TWOMEY, T., CRINION, R.A.P. and GLAZIER, D.B. (1977). Seleniun toxicity in cattle in Co. Meath. Ir. vet. J. 31: 41–46.

DISCUSSION McGRATH (UK):

You related the percentage of acid insoluble residue to percent soil ingestion. Is there a formula for the relationship? FLEMING (Eire): Yes, this is published. If an X-ray fluorescence spectrometer is available the titanium marker method is better. WILLIAMS (UK): In reference to copper deficiency you mentioned FeS formation. Does this take place in the animal? FLEMING: Yes, a black deposit is formed in the gut. JUSTE (France): Do the stomach contents of the animal extract metals from the ingested soil and does the liquor give a good indication of the heavy metal intake? FLEMING: Yes, metals particularly selenium and cobalt but not copper are extracted from ingested soil. However, the liquors of the duodenum and ilium have different pH values and therefore solubilise different amounts of metals. COPPOLA (Italy): It has been shown that germ-free animals have different metal uptake. Should the effects of micro-organisms in the gut be included in the factors influencing metal status? FLEMING: Yes, this is an important factor and one which is not familiar to most of us. WILLIAMS: Have you established whether there should be limits for any elements such as fluorine or lead to safeguard against contamination of herbage? FLEMING: No, we have not investigated limits, but this could be a subject for research. FROSSARD (Switzerland): Is the splashing of soil onto plants by rain important? FLEMING: Yes, herbage splashed by soil has been found to have a much higher cobalt analysis than uncontaminated herbage. In dry conditions, blown soil dust can also cause herbage contamination.

GROUNDWATER QUALITY IN RELATION TO LAND APPLICATION OF SEWAGE SLUDGE M.Ahtiainen National Board of Waters, Water District Office of North Karelia, Finland

Summary At present the most hazardous pollution phenomenon in groundwaters is the nitrateproblem. The extent of N-removal will determine the feasibility of most sludge application projects. From a practical standpoint, the annual loading rate can be estimated using a N-mass balance. The annual loading rate can be adjusted in accordance with proper monitoring technique. The upper limit for nitrogen fertilizers has been emphasized to be 100 kg/N/ha, when the groundwater pollution is hindered. Several research studies confirmed that little, if any, migration of heavy metals occur through the soil, if moderate amounts of sewage sludge have been applied. The heavy metal load into groundwaters will become more serious, if the acidification process of the soils is allowed to continue with acid rains. Consequently a more effective pretreatment of industrial effluents must be reconsidered in the near future. The fecal bacteria are known to accumulate better in the top soil layer than viruses. More information is required about the behaviour of virus percolation and survival in groundwaters. Our knowledge of toxic organic chemicals is scarce concerning the filtration processes into groundwaters. At present the runoff phenomena of these compounds appear in surface waters in a more prominent and harmful way than in groundwaters. When focusing on the groundwater contamination we should keep in mind the numerous other sources, which in the long run can seriously affect the groundwater quality. In a wider context, land application of sewage sludge plays a minor role, but locally land application should be carried out with extreme care. 1. Introduction In recent years concern has been expressed for the pollution question of European groundwaters. When discussing the application of sewage sludge it should also be borne in mind that more effective agriculture, industrialization as well as urbanization are also to be blamed for. Most of nothern Europe has been glaciated. The area is therefore covered with quarternary stratifications. A high proportion of the groundwater consumed thus derives from these stratifications, which are of great importance. A simple geohydrological model of a landscape in central Sweden can serve as a basis for estimates of where the risk of the impact of land application is great and where it is absent or small (Figure 1).

47

Naturally only the groundwater in the permeable layers is of interest in terms of water supply. The magnitude of the impact of land application on groundwater quality is dependent on the local conditions such as climate, type of soil, type and amount of sludge, and chemical conditions (7). In central Europe the aquifers are in sedimented rock layers, which are significantly larger and deeper than the Scandinavian ones. The border of ekers and rivers are of special concern. Naturally different kinds of crackers and ruggedness in soil can assist the sludge infiltration into the deeper soil layers as well. 2. Nitrate-nitrogen The nitrate-nitrogen concentrations have elevated in every European country. At the same time plant and animal production and the use of fertilizers in agriculture have increased. The nitrate content of groundwater is usually observed to be higher in areas where water can easily percolate through the soil than in clay soil areas (20). In Denmark the nitrate content of groundwater has increased threefold in the past 20–30 years and is continuously growing (14). In Holland in cultivated areas the nitrate concentrations of shallow aquifers is hudredfold compared with the natural state of groundwaters in sandy soils (15). Because of slow infiltration rate the effects of fertilizers cannot be observed immediately. In Sweden the present NO3-concentrations are said to reflect the fertilization levels of the 1960s. Therefore, it is predicted that NO3-N concentrations of groundwaters will be elevated in the near future. (1). If the amount of fertilizers exceeds the need for plants, the amount of percolating nitrogen increases. If the fertilization limit is over 170 kg N/ha it is the main NO3-source of groundwaters (17). According to Andersson (1) the critical level is 100 kg/ha N. For sound fertilization and water management sludge application rates that would not elevate groundwater or soilwater NO3-N levels above the 10 mg/1 potable water standards were recommended as 9.5 dry Mg of undigested sludge/ha of red pine plantation, 16 dry Mg of digested sludge/ha of red pine and white pine plantation and 19 dry Mg digested sludge/ha of aspen sprout stand (Figure 2). These rates differ because of the faster N-mineralization rate in undigested sludge and the greater nutrient uptake rate in the young aspen stand (2). According to Higgins aerobically digested liquid sewage sludge was applied to Sassafras sandy loam soils for 3 consecutive years. Sludge was applied at rates of 22.4 and 44.8 Mg of dry solids/ha. Annual application rates of 22.4 Mg of dry solids/ha resulted in the contamination of the groundwater beneath the test plots and immediately offsite. Annual application rates of 44.8 Mg of dry solids/ha grossly contaminated the groundwater. Nitrate-N concentrations in groundwater sampled from within the plots receiving sludge increased to≤50 mg/1 in proportion to the amount of sludge applied. Once sludge application ceased, normal levels quickly resumed; therefore it appears that the effects of sludge application≤44.8 Mg of dry solids/ha on groundwater NO3-N are temporary and easily recoverable in this area. The results of this study show that leaching, particularly NO3-N, will occur in late fall and winter. Cover crops will mitigate the leaching effects on groundwater. Application of the 22.4 Mg of dry solids of sludge is clearly at the upper limit to ensure protection of the groundwater quality on this site according to the federal drinking water standards. Application rates, at or slightly below the 22.4 Mg of dry solids/ha application rate are sufficient for providing plant nutrient for corn and rye. (10)

48

Fig. 1 A geohydrological model of a landscape in central Sweden (7)

In conclusion I would like to emphasize the fact that the nitrogen deposition of the air has increased everywhere. What is more, other activities such as the use of artificial fertilizers, liquid manures, furfarming, and clear cutting of forests have elevated NO3-N concentrations. 3. Heavy metals Contrary to the nitrogen and contagion problems, which can be managed adequately with existing knowledge and technology, the metals involve rather complicated considerations before a satisfactory assessment of the problem can be reached (5). The mobility of heavy metals in the soil, their retention and release is governed by precipitation, complexation, adsorption and oxido-reduction. The total maximum acceptable amounts of sludges are calculated in function of the heavy metal contents of the sludge and the receiving soil. From this it appears that the most limiting elements are Cu, Zn, Cd and Cr. The yearly dose will, however, in the first order be restricted by the nitrogen content, while heavy metals accumulate from year to year restricting their further input in the distant future (4). To estimate the effect of pollution of the soil with heavy metals and other elements from sewage sludge on surface and groundwater quality, data on the physical and chemical parameters involved in the

49

Fig. 2. Recommended sludge application rates as related to soil water and groundwater NO3-N concentrations. (10)

adsorption processes in soils are essential. (5). Adsorption is affected by speciation of the elements in the soil solution and by pH, Eh, ionic strength and composition of the soil solution as well as by the clay and organic matter content of the soil. Cadmium is considered the most harmful among heavy metals, mercury and lead are other clearly toxic elements. According to Huylebroeck raised concentrations of heavy metals in groundwater have been raported to a minor extent. According to several Danish and Dutch researchers (6, 8, 9) even high doses of sludge (up to 30 tn/ha DM) did not increase leaching of heavy metals (Cd, Pb, Ni) into subdrains and groundwater. Mostly Cd has been found to accumulate in the top soil layers. Only lead and zinc had been noticed to leach below 0, 5m and only the content of zinc had increased in groundwater. In the field experiment of Liperi and Maaninka elevated concentrations of heavy metals could not be detected either (13). In an early mentioned study (10) sewage sludge application—44.8 Mg of dry solids/ha did not contribute heavy metals to groundwater samples taken within the test-plots nor 30.5 m off-site. According to this

50

writer application of 22.4 Mg of dry solids of ha sludge is clearly at the upper limit to ensure protection of groundwater quality on this site according to the federal drinking standards. Hence the results indicate that no detectable contaminant of the groundwater occur as a result of sludge application even at the highest rate. These results confirm previous research findings that little, if any, migration of heavy metals occurs through the soil. There is little risk from a single sludge application on arable land, but it is necessary to examine the situation created by an increasing sludge production and its future disposal. Therefore, the only acceptable attitude is to keep the total amounts of heavy metals in the soil below sufficiently safeguard limits concerning the increasing sludge production, and the limited absorption capacity of soils for heavy metals. The treatment of effluents will have to be conducted in the future in such a way that heavy metals should be eliminated where necessary (4). Moreover, there may be a natural heavy metal explosion to be expected, if the acid rains do lower the soil pH below a certain level. 4. Patogens It would be fair to assume that the microbial load of the sewage will vary from community to community and that within each community it will vary according to the enteric infections prevalent at the time of testing. It is important that the patogen load in the sludge be known before operations commence. It is equally important that the characteristics of the soil with regard to pH, clay content, proximity to the water table, infiltration capacity and drainage pattern be known (19). In addition, it is essential that accurate records be kept of sludge application to land so that retrospective investigation can establish the operating conditions in case of problems. Salmonella are probably the most important bacteria present in sludge. Patogens might cause human illness by contaminating surface waters by runoff and percolation. It appears that patogen movement in soil is related directly to hydraulic infiltration rate and inversely to media particle diameter, in general patogens do not travel to any great extent in soils and remain principally in the upper layers. It has been presented that most of the total and fecal coli-forms removed to the top soil, but fecal streptococcs were found as deep as 14 m and at the distance of 183 m horizontally from the infiltration area of waste waters. (Table 1 after Lahti). The possibility of viral contamination of surface and groundwaters as a result of the land application of domestic wastes poses one of the major concerns. (Table 2 after 12). Vaughn studied lateral movement of groundwaters entailed viruses from a subsurface waste-water disposal system in Long Island New York. Virus movement could be traced at least up to 67.05 m the septic source. There was also no correlation between the presence of viruses in groundwater and the counts of fecal coliform bacteria. It has been stated however that the land disposal of sewage sludge is much safer than landfilling, incineration and discharge to water bodies.

51

5. Organic chemicals The impact on groundwater of diffuse land spreading of municipal sludge is practically unknown, because fewer than 1% of the present municipal sludge land spreading facilities are monitored for effect on water quality. Groundwater pollution by organic chemicals can be more evident from leaching of toxic chemicals from industrial wastewater landfills, municipal solid waste disposal sites or agricultural use of pesticides. Pesticides are organic products that have very diverse natures and physical and chemical properties. Studies on the development of pesticides in the ground are many, but they are always difficult because of the complexity of the phenomena involved: solubility, complexing with the humic acids and hydroxides present in soil, possibilities of leaching. Nevertheless, most authors agree that pesticides are generally halted in the ground and stay there long enough to be partially destroyed On the other hand, they are more easily transported to surface water by soil erosion during run-off and they may then be found in wells bordering rivers because of the induced recharge. Even if the doses which reach drinking water are small, prudence in accepting then is appropriate—The Americans have adopted strict standards for pesticides—because they are susceptible of concentrating in the food chain. (3). Since there is no information available on groundwater and surface water pollution from organic chemicals in waste water and sludge applied to land, effects of toxic organic chemicals leaching from industrial chemicals landfills are briefly mentioned because of their potential for groundwater pollution. Examples where groundwater pollution occured from leachates from a toxic waste landfill, where residential drinking water supplies were contaminated by chlorinated pesticide waste chemicals and petrochemicals in USA are Toone Tennessee, Dover Township, New Jersey. (US EPA). References 1 2 3 4 5 6

7 8

9

10

() ANDERSSON, A. (1982). Växtnärings förluster från åker och skog. Vatten 38 205–225. () BROCKWAY, D.G. and URIE, D.H. (1983). Determing Sludge Fertilization Rates from Nitrate-N in Leachate and Groundwater. J. Environ. Qual. Vol 12, no. 4, 1983 487–492. () CLOUET D’ORVAL, M. Protection of aquifers against pollution 1 General Report 1–12. () COTTONIE, A. (1981). Sludge treatment and disposal in relation to heavy metals. International Conference Heavy Metals in the Environment. Amsterdam—September 1981 167–175 () GERRITSE, R.G., VRIESEMA, R., DALENBERG, H., DE ROOS, H.P. Trace element mobility in soils; effect of sewage sludge. International Conference Heavy Metals in the Environment, Amsterdam Sept. 1981. () GRANT, R.O. and OLESEN, S.E. (1983). Sludge utilization in spruce plantations on sandy soils. Utilization of Sewage Sludge on Land: Rates of Application and Long-Term Effects of Metals. Proceedings of a Workshop held a Seminar, June 7–9, 1983. () GUSTAFSON, A. (1982). Leaching of nitrate from arable land into groundwater in Sweden. Ecohydrologi 12. Sveriges Lantbruksuniversitet. Avd. for Vattenvård 37–45. () DE HAAN, S. (1981). Cadmium in soils, crops and drainage water, as affected by cadmium applied in soluble salts, sewage sludges, town refuse compost and dredged soils from Rotterdam harbours. Treatment and use of sewage sludge, Commission of the European Communities. () HANSEN, J.A. and TJELL, J.C. (1982). Sludge application to land—overwiew of the cadmium problem. Environmental Effects of Organic and Inorganic Contaminants in Sewage Sludge. Proceedings of a Workshop held at Stevenage, May 25–26, 1982. () HIGGINS, A.J. (1984). Land Application of Sewage Sludge with Redard to Cropping Systems and Pollution Potential. J. Environ. Qual., Vol 13, no 3, 1984.

52

13 14 15

17

19 20

(11) HUYLEBROECK, J. (1981). Review of research projects on groundwater pollution from agricultural use of sewage sludge. Treatment and use of sewage sludge. Commission of the European Communities. Cost project 68 Bis. Final report. III Technical annexes 491–521. (12) LAHTI, K. (1981). Suolistoperäisten bakteerien ja virusten aiheuttama pohjavesien pilaantuminen. Vesihallitus, tiedotus 208, 43. Helsinki. () MELKAS, M., MELANEN, M., JAAKKOLA, A., AHTIAINEN, M. and MATINVESI, J. Leaching resulting from land-application of sewage sludge and slurry (unpublished) () OVERGAARD, K. (1984). Nitrate pollution of groundwater. Nordforsk. Tjugonde nordiska symposiet om Vatten-forskning “Jordbrukets förorening av vattenmiljö”, Hägersten, 8.–10. 1984 6. () RIJTEMA, P.E. (1982). Effects of regional water management on N-pollution in areas with intensive agriculture. Nonpoint Nitrate Pollution on Municipal Water Supply Sources; Issues of analysis and control. 15–42 Luxenburg. (16) SCHAUB, S.A. and SORBER, C.A. (1977). Virus and bacteria removal from wastewater by rapid indiltration through soil. Appl. Environ. Microbiol. 33 609–619. () SPALDING R.P., EXNER M.E., LINDAU, C.W. and EADEN, D.W. (1982). Investigation of sources of groundwater nitrate contamination. Washington, USA. Journal of Hydrology 58 307–324. (18) VAINIO, E. (1984). Typpiyhdisteet maatalousalueiden kaivovesissä. Vesihallituksen monistesarja 240 1984. Helsinki 62. () WALLIS, P.M. and LEHMANN, D.L. Biological Risks of Sludge disposal to land in cold climates. Kananaskis Centre for Environmental Research, The University of Calgary 1–374. () VAUGHN, J.M., LANDRY, E.P., BARANOSHY, C.A. BECKWITH, M.C. DAHL , DALIHAS, N.C. (1978). Survey of human virus occurence in wastewater recharged groundwater in Long island, Appl. Environ. Microbiol. 36 47–51 (21) YRJÄNÄ, E-R. (1982). Esiselvitys korkeiden nitraattipitoisuuksien esiintymisestä pohjavesissä, Vesihallituksen monistesarja 156 6 1983, Helsinki 49.

TABLE 1 SURVIVAL TIMES OF FECAL BACTERIA IN GROUNDWATER (12) ORGANISM

RED. % TIME, H

TEMP. °C

REF.

COLIFORMS ENTEROCOCCIS STREPTOCOCCUS EQUINUS STREPTOCOCCUS BOVIS SHIGELLA DYSENTERIAE SHIGELLA SONNEI SHIGELLA FLEXNERI SALMONELLA ENTERTIDIS SER. PARATYPHI A S. ENTERTIDIS PARATYPHI D S. ENTERTIDIS SER. TYPHIMURIUM S. TYPHI VIBRIO CHOLERAE S. ENTERTIDIS SER. PARATYPHI B E. COLI

50 50 50 50 50 50 50 50

17,0 22, 0 10, 0 4, 3 22, 4 24, 5 26, 8 16, 0

9, 5–12, 5 WELLWATER ” ” ” ”

MECFETERS YM 1974 ” ” ” ”

” ”

” ”

50

19, 2

”

”

50 50 50

16, 0 6, 0 7, 2

” ” ”

” ” ”

50 99, 99

2, 4 ” 3–3, 5 mon. GROUNDWATER, FIELD

” KUDRYAVTSEVA 1972

53

ORGANISM

RED. % TIME, H

E. COLI E. COLI S. FAECALIS

99, 99 99 90

TEMP. °C

REF.

4–4, 5 mon. GROUNDWATER, LAB. 1 mon. 2–15 1 mon. 2–15

HAGERDORN YM. 1978

TABLE 2 SURVIVAL TIMES OF VIRUSES IN NATURAL AND GROUNDWATERS (12) VIRUS

TEMPERATURE ° REDUCTION % TIME, D SAMPLE C

REF.

POLIOVIRUS 1

23–27

90

1,0

” ” ” POLIOVIRUS 3 ” COXSACKIEVIR US A 13 ” COXSACKIEVIR US B-l ” ” POLIOVIRUS 1

12–20 7–17 4–8 23–27 12–20 23–27

90 90 90 90 90 90

1,3 1,5 1,9 0,8 1,0 0,3

RIVERWATER, IN SITU ” ” ” ” ” ”

O’BRIEN & NEWMAN 1977 ” ” ” ” ” ”

12–20 12–20

90 90

0,5 1,2

” ”

” ”

7–17 4–8 22

90 90 90

1,8 2,4 3,4

”

22

90

5,4

” ” YEAGER & O’BRIEN 1979 ”

”

22

90

12,0

COXSACKIEVIR US A 2

8–10

99

14,0

” ”

20–23 20–23

99 99

5,0 47,0

”

8–10

99

160,0

” ” ” POLIOVIRUS 1

20–23 8–10 20–23 –

99 99 99 99

102,0 >272,0 100,0 4–14,0

” ” RIVERWATER, LAB. GROUNDWATER, LAB. SEWAGEWATER, LAB. SLIGHTLY POLL. RIVERWATER LAB. ” POLLUTED RIVERWATER AUTOCLAVETED RIVERWATER ” DEST. WATER ” LAKEWATER

”

21–26

99,9

20,0

”

4–16

99,9

40,0

SEAWATER, IN SITU ”

” CLARKE YM. 1956

” ” ” ” ” ” HERRMANN YM. 1974 LO YM. 1976 ”

54

VIRUS

TEMPERATURE ° REDUCTION % TIME, D SAMPLE C

REF.

”

2–19

”

99,9

32,0

”

DISCUSSION WEBBER (Canada):

Groundwater pollution by nitrate from fertiliser, sewage sludge and animal slurry is important. Is there any evidence of groundwater pollution from heavy metals? AHTIAINEN (Finland): Large scale water suppliers do not experience problems from metals, but there could be local problems from industrial wastes particularly where groundwater has shallow permeable ground cover. TJELL (Denmark): Why is Norway leaving COST 681? Is it because you have low permissible limits for heavy metals in soil? VIGERUST: The reasons are economic and political. TJELL: What are you views on the present Norwegian limits? VIGERUST: They are strict because farmers like to use sludge liberally and safe metal levels must be maintained.

SESSION II

Chairman Rapporteurs

: :

R.D.DAVIS G.C.PORTEOUS S.MCGRATH

Predictability of estimated mobilizable N pool in sludge and soils Summary of investigations in Italy into effects of sewage sludge on soil microorganisms Effects of long-term sludge additions on microbial biomass and microbial processes in soil Discussion

Political and administrative considerations in the formulation of guidance for sludge utilization Discussion

Future developments in sludge disposal strategies Discussion

PREDICTABILITY OF ESTIMATED MOBILIZABLE N POOL IN SLUDGE AND SOILS A.Rudaz and S.K.Gupta Swiss Federal Research Station for Agricultural Chemistry and Hygiene of Environment CH-3097 Liebefeld-Berne

Summary Mineralized nitrogen from sewage sludge and soil can cause an enrichment of surface soil and groundwater. In order to prevent this Kind of pollution and to estimate the nitrogen supplying capacity of sludge and soils, potential mineralizable nitrogen of sludge should be determined. This paper discusses a few models to determine the potential mineralizable nitrogen content of sludges and soils. Almost all of the models fail to predict the mineralizable nitrogen content of different soils and sludges. It is suggested that to improve the predicting ability of a model one needs to incorporate in it some correction factor specific to the composition of organic matter. 1. INTRODUCTION The application of sewage sludge in soils can cause N enrichment of surface soil and groundwater. To prevent this kind of pollution the sewage sludge has to be spread on the soil at the correct time and in reasonable quantities. The form of nitrogen present in sludges must be studied in order to estimate the quantity of plant available nitrogen. The mineral nitrogen for example: NH4 -N; NO2 -N; NO3 -N is easily absorbed by the crops. The organic form of nitrogen must be mineralized through microorganisms before they are available to the plants. For the estimation of nitrogen value of sludge the amount of mineralizable nitrogen must be determined. In this paper a few models assessing the potentially mineralizable nitrogen will be discussed. 2. NITROGEN CYCLE IN ECOSYSTEM Nitrogen is an essential element for the growth of plants. Application of nitrogen to soils should be done at optimum levels because if the nitrogen content in soils exceeds a particular limit it can be harmful to plant growth. Estimating the optimum fertilizing values of nitrogen with regard to given crops based on two fundamental principles: the response curve and the N balance (1). The two methods pose a few problems when estimating available nitrogen in mineral fertilizers. In the case of organic fertilizers, the two methods frequently fail to predict the potentially available nitrogen content. The organic nitrogen present in these wastes must first be mineralized through microorganisms before it can be utilized by the plants, then

57

Fig 1: NITROGEN CYCLE in Ecosystem

nitrified and finally denitrified (Fig. 1). A part of mineralized nitrogen immobilizes if this C: N ratio is greater than 10:1. 3. FORMS OF SLUDGE NITROGEN The total nitrogen content of sludge ranges between 1–7 % of the total dry matter. Total nitrogen is subdivided in two fractions: the organic and the ammoniacal (Table 1). Organic nitrogen is found in large quantities (30–95%) of the total nitrogen (2). The most important are proteins, hexoamines and other unidentified components. The ammoniacal nitrogen amounts to 5–70% of the total nitrogen (2). The quantity is influenced by the type of sludge and its water content. For example it has been observed that approximately 60% of inorganic nitrogen are lost due to centrifugation, whereas 80% are lost during drying in open fields (3). 4. SLUDGE NITROGEN VALUE Ammoniacal fertilizer is only 90 % as efficient as mineral fertilizer. This is due to loss by volatilization. The fertilizer value of organic nitrogen is more difficult to estimate. This mainly because of the complexe nature of organic nitrogen which has to undergo microbial decomposition. The residual nitrogen effect of sludge is principally caused by the slowly decomposable organic molecules present in sludge. Keeny and coworkers (1975) in (4) estimated that 15 to 20% of the sludge N are mineralized the first year and after initial application about 6,4; 2% of the remaining N are released over the 3 subsequent years. Nevertheless,

58

many scientists attempted to estimate the nitrogen value of sludge. Furrer and Bolliger (5) suggested the following formula which is based on pot experiments. The results indicate that available nitrogen content of sludge was as follows:

The nitrogen value of Swiss sludges is approximately 40–50%. In order to predict the potential mineralizable nitrogen which is either taken up by the plants or lost in drainage water, one must know the biochemical behaviour of sludge organic nitrogen. 5. ESTIMATION OF POTENTIALLY MINERALIZABLE NITROGEN In order to predict the amount of nitrogen available to crops and the nitrogen susceptible to leaching it is necessary to know the N-mineralizable capacity of soils and the sludge-soil system. Many methods have been developed in several laboratories and some of them are discussed below: Incubation Methods These short time methods follow the production of mineral nitrogen versus time. This incubation is carried out either aerobically or anaerobically. In order to accelerate the rate of N decomposition, this experiment Table 1: Forms of sludge nitrogen Type of sludge

NTotal (% OM)

NH4-N (% NT)

Norganic (% NT)

Raw Aerobically digested Anaerobically digested Deshydrated

2–4 7 3, 5–5, 1 1–2

5–10 5–10 5–70 95

is conducted at a temperature above 20°C. The results obtained in laboratories were measured at high temperatures and could even be extrapolated for field soils. Although in general there is a good correlation between potentially mineralizable nitrogen estimated by anaerobic incubations and the plant production, we must not forget that estimates may sometimes be too high because of a possible inhibition of immobilisation.

59

Chemical methods Incubation methods being time consuming, scientists have attempted to discover chemical extractants to estimate the quantity of mineralizable nitrogen. These methods are sometimes contradictory to those based on incubations. The soil properties were not taken into account in estimation of mineralizable nitrogen by both methods. In addition, the incubation methods are carried out for short periods of time, that is 1–2 weeks, and are not sufficient to estimate the total pool of mineralized nitrogen. Some models exist, which can predict the potentially mineralizable nitrogen, but only for soils. A few of them are discussed below: Model 1 (Stanford and Smith 1972) This model presents the cumulative nitrogen mineralization over 30 weeks in aerobic conditions at 35°C with periodic leaching of mineralized nitrogen (Fig. 2). In Figure 2, the relationship between mineralized nitrogen and time is presented. N−MOB is function of the square root of time (6). This first model has the disadvantage that at time zero (0) the calculated N−MOB is sometimes negative, which is impossible to find in a real situation. The model cannot be applied to predict the mineralizable potential nitrogen (N−MOB) for different soils. In the second model, the authors tried to quantify the mineralizable potential nitrogen. They suggested that N0 values obtained after 2 weeks of incubation correspond approximately to N−MOB. N0 values were then used to calculate the constant rate (k) for different soils log (N0−NT)= log N0−K/2, 303 and were found more or less constant for different soils (0.059±0.006 week−1). With this k value, they calculated then the potential mineralizable nitrogen (N-MOB) from 0–30 weeks . Model 2 (Tabatabai and Al-Khafaji 1980) This model describes the nitrogen mineralization over 26 weeks incubation at 20 to 35°C under aerobic conditions.The authors found a linear relationship between N−MOB and the incubation time (N−MOB=a+b . t] (7) in opposition to Stanford and Smith. They also observed that N mineralized at time could be either negative or positive, depending on soil types. This is contrary to the real situation. Model 3 (Stadelmann and al. 1982) This model uses the results of experiments carried out under anaerobic conditions at 35°C (Fig. 2). They differentiate organic mineralizable nitrogen in two fractions NL (easily degradable) and NS (difficultly degradable) (8). This model presents advantages over others because the investigation lasts only forteen days and no more than 2 and 3 measurements are required to estimate the NL and the rate of decomposition (v). This model, however, cannot explain some of the results observed. For example a soil Table 2: Effect of soil properties on N-MOB parameters calculated from model 3 Soils

Soil characteristics

C (%)

N (%)

total

organic

1

5,3

0,555

Ca CO3 (%)

7 days

14 days

35 days

0,547

0

6,7+1,9

10,6+0,9

15,8+1,1

12,2

407

140

60

Fig 2: Different models to calculate mobilisable N in soils

Soils

Soil characteristics

C (%)

N (%)

total

organic

2 3

1,8 10,5

0,22 1,03

Ca CO3 (%)

7 days

14 days

35 days

0,215 0,905

0 32

3,5+0,2 1,1+0,1

3,3+0,7 2,6+0,1

6,9+0,2 4,8+0,2

2,6 9,0

53 107

162 798

rich in organic nitrogen (N0) (Soil No. 3) compared to a soil poor in organic nitrogen (Soil No. 1): the calculated N-MOB is higher in a soil poor in organic N (Table 2). It is clear that in estimations of N-MOB in soils or sludges not only the organic matter quantity but its chemical composition should be taken into consideration. At our research station more experiments will be conducted in future to get a deeper insight into the problem of N-MOB. LITERATURE 1 2 3 4

5 6 7

. CATROUX, G. (1982) Nitrogen and phosphorus value of sewage sludges, published by COMMISSION OF THE EUROPEAN COMMUNITIES, SL/82/82 . . HALL, J.E. Predicting the nitrogen values of sewage sludges. In Proceedings of the Third International Symposium on the Processing and Use of Sewage Sludge. Brighton, 1983, CEC. . FURRER, O.J., CANDINAS, T.Klärschlamm-Rohstoff oder Schadstoff? In Sonderdruck aus: Gewässerschutz. Wasser. Abwasser, Aachen 1984. . KEENEY, D. R., REE, K.W. and WALSH, L.M. (1975) Guidelines for the Application of Wastewater Sludge to Agriculture Land in Wisconsin. Technical Bulletin 88 Wisconsin Department of Natural Ressources, Madinson, Wisconsin, USA. . FURRER, O.J., BOLLIGER, R. Die Wirksamkeit des Stickstoffes im Klärschlamm. Schweiz. landw. Forsch. 17 (3/4), 137–147, 1978. . STANFORD, G. and SMITH, S.J. (1972) Nitrogen mineralization potentials of soils. Soil Sci. Soc. Amer. Proc. 36, 465–472. . TABATABAI, M.A. and AL-KHAFAJI, A.A. (1980) Comparison of nitrogen and sulfur mineralization in soils. Soil Sci. Soc. Am. J. 44. 1000–1006.

61

8

. STADELMANN, F.X., FURRER, O.J., GUPTA, S.K. and LISCHER, P. (1983) Einfluss von Bodeneigenschaften, Bodennutzung und Bodentemperatur auf die N-Mobilisierung von Kulturböden. Zeitschrift für Pflanzenern. und Bodenkunde. 146 (2), 228–242.

SUMMARY OF INVESTIGATIONS IN ITALY INTO EFFECTS OF SEWAGE SLUDGE ON SOIL MICROORGANISMS S.COPPOLA Istituto di Microbiologia agraria e Stazione di Microbiologia industriale Università di Napoli—I 80055 Portici, Italia

1. INTRODUCTION In Italy as well as in other countries, the agricultural utilization of sewage sludge has been considered as a practice able to return nutrients to the agroecosystem. Moreover, because italian pedoclimatic conditions, also the organic matter of sewage sludge has been emphasized as capable to improve soil fertility (2). Normally sewage sludge showes a considerable microbial content. Both nutrients and organic matter can then support the growth of soil microflora after sludge application. At last some organic or inorganic sludge contami nants can affect soil microorganisms.The application of sewage sludge must be therefore considered producing an important influence upon soil bacteria. Nature, intensity and duration of such an influence have been among the topics of various researches. Investigations carried out at the University of Pisa and at the C.N.R. Institute for Plant Ecophysiology, near Rome, have mainly concerned effects of sludge and compost applications on biological activities in the rhizo sphere and on mycorrhizal fungi. At the University of Naples microbiologic al characteristics of six different soils have been investigated after treatment with composted sewage sludge at three application rates. At the same Laboratory, biological properties of a soil treated with the same sludge stabilized by different methods have been studied and some effects of Cadmium-bearing sewage sludge on microorganisms have been estimated in two types of soil occurring in the Southern Italy. This report summarizes the most interesting experiments, putting into evidence some practical conclusions as well as needs for further research arising from the available results. 2. EXPERIMENTS AND RESULTS A sandy-loam soil (pH=6.5) has been treated by Tomati et al. (18–19) with three doses of a domestic sludge in order to study the effects on soil biological activities in the presence of maize as test crop. In pot trials the most beneficial application rate resulted 8 g of sludge (d.w.) per Kg of soil, corresponding to common quantities utilized in agriculture.Higher (16 g Kg−1) or lower (4 g Kg−1) doses equally enhanced total microflora, oxygen uptake and protease activity in the plant rhizosphere. The highest microbial counts were recorded 16÷21 days after sludge treatment for bacte ria, later for fungi and actinomycetes. Biochemical activities reached their maximum after 27 days. The same Authors (20) have carried out four years experiments in the field (500 m2 plots) on a sandy soil pH=5.8. Maize was grown every year after treatment with 2 tons per hectar of organic matter in the form of

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aerobic and anaerobic sewage sludges. Control plots were treated with inorganic fertilizers. Microbiological investigations have been carried out before each treatment and again at the plant emergence time. No significant difference was determined for fungi and actinomycetes before the treatment. Bacteria and oxygen consumption significantly increased after the third year, more in aerobic sludge treated soil. At the emergence time, since the first year, all microbial counts in the plant rhizosphere resulted always enhanced, higher in the soil treated with aerobic sludge. Oxygen consumption by the soil sampled from the rhizosphere of plants grown on sludge treated plots progressively rised after three and four years. Crop yield and quali ty weakly increased. So Authors concluded that the fertility of the poor soil used for their experiments (0.9% of organic matter) could be consider ed significantly improved after four years, especially employing aerobical ly digested sludge. The same soil and test crop have been utilized by Pera et al. (16) to study the influence on physiological groups of soil microorganisms and on mycorrhizal symbiosis following application of liquid aerobic sludge (about 20 tons of organic matter per hectar), anaerobic (about 24 tons of organic matter ha−1), aerobic sludge composted in mixture (40:60) with organic fraction of urban solid wastes (about 23 tons), anaerobic sludge composted in mixture (20:80) with organic fraction of urban wastes (about 20 tons) and farmyard manure (about 24 tons of o.m.ha−1). The experimental design (500 m2 plots) included a control plot and a plot treated with mineral fer tilizers. Microbiological analyses have been carried on monthly after plant emergence and during 6 months, considering the rhizosphere soil and myco rrhiza. Variations in number of bacteria, actinomycetes and fungi resulted to be more related both to plant growth phase and to plant nutrients availabi lity than tothe different organic or inorganic additions. The lowest counts resulted in the rhizosphere of plants grown on the control soil. As far as physiological groups were concerned, proteolytic and ammonia producing bacteria did not show any significant variation among the treatments; auto trophic ammonia oxidizers were generally depressed by the application of organic matter, whilst nitrite oxidizers widely enhanced; aerobic and free-living nitrogen fixing bacteria increased during late stages of plant grow th and in the experiments where organic fertilizers were used. Authors concluded that the rhizosphere-effect is more important for microbial grow th in the soil-root interface than any other external effect, including tratments with organic or inorganic fertilizers. Vesicular-arbuscular mycorrhizal infections in control plants reached 30%, while in plants treated with inorganic fertilizers was 15% only. The infection was absent in all the other treatments. Authors have cited studies of Gerdemann (13) and of Menge et al. (15) to explain the results as a consequence of high levels of available phosphate in soil. The influence of applications of sewage sludge composted in mixture with solid urban wastes upon rhizosphere microorganisms of Sorghum plants has been studied through pot trials in two soils: a clay soil with 0.63% of organic matter and pH=8.70 and an alluvial sandy soil with 1% of organic matter and pH=7.90, by Pera et al. (17). Three different application rates of compost have brought the organic content of the clay soil to 1.25, 2.41 and 4.79 per cent; to 2.11, 4.09, 8.23 per cent in the sandy soil. Microbiological analyses were carried out 5, 15, 45 and 75 days after the plant emergence. The results have demonstra ted beneficial effects of compost applications in both soils. Heterotrophic aerobic bacteria and fungi were intensely enhanced by the treatments. Acti nomycetes were significantly stimulated in the clay soil. Cellulolytics resulted scarcely influenced in clay soil, enhanced by increasing doses of compost in the sandy soil. Nitrite-oxidizing bacteria showed a better res ponse to compost in the sandy soil. All the microbial groups evaluated reached their maximum 45 days after the plant emergence; cellulolytic after 75 days. Microbiological evaluations resulted quite in agreement with the plant response, as far as root and epigeous development at different concentrations of organic matter was concerned.

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Incubation studies have been carried out by Coppola and co-workers (3–5–6) with six different types of soils occurring in the Southern Italy treated with three doses (exceeding included) of sewage sludge composted in mixture with wood chips, in order to ascertain the effect of heavy applica tion rates on microbiological processes as a function of pedological chara cteristics. A fine texured, volcanic, sandy soil, with 1.67% of organic matter, an alluvial sandy-loam soil with 2.96% of organic matter, a sample of “terra rossa”, with high percentage of kaolinite-type clay and 3.35% of organic matter, a clay-loam soil with 2.06% of organic matter and two sandy-clay-loam soils with 4.5% of organic matter, the first with 28% of CaCO3 , have been assayed. Composted sludge was mixed with each soil to realize applications of 12.5, 25.0 and 50.0 tons organic matter per hectar. The water content of the mixtures was carefully brought to 50% of the satu ration value and incubation was carried at room temperature. Mineralization was monitored during 2÷3 months. Then microbiological analyses were carried out to evaluate the ability of control and sludged soils to mineralize organic carbon and nitrogen, to oxidize , to fix molecular ni trogen. Soil respiration ran proportionally to the organic matter added after compost application, but in the two soils with the highest percentage of humus the heaviest sludge applications depressed biological activity during the first month of incubation. After 75 days, sludged and control soils were treated with 5% of wheat straw to verify carbon mineralization, with 5% of Casaminoacids to study ammonification and with 15 mg per cent of ammonia nitrogen to evaluate nitrification. Nitrogenase activity was gaschromatographically assayed by the acetylene-ethylene reduction test. Organic carbon mineralization resulted enhanced in all the treated soil samples, excepting in “terra rossa”, the richest in clay. The lowest applic ation rates resulted generally more beneficial for all the soils. Ammonification appeared widely depressed, mainly after high application rates and in soils with lower clay content. The most probable number of ammonifier microorganisms did not significantly differ among the various experimental conditions. Since the cumulative ammonia released from the treated soils was often drammatically reduced in comparison with the contr ol (more than 50%), nitrogen immobilization due to an excess of organic carbon cannot justify the extent of the negative influence of the composted sludge. The decrease of microbial amino-hydrolase activity in the sludged soils could be therefore attributed to organic and/or inorganic inhibitors. Nitrification occurred without troubles in all the soils in the presence of every dose of compost. Any way, in the two richest soils (sandy-clay-loam, with about 4.5% of organic matter) the heaviest applications weakly reduced the constant rate of the process. Dinitrogen fixation resulted remarkably stimulated in two poor sandy soils as well as in the calcareous one, slight ly enhanced or variably affected in the other samples. These experiments have altogether shown that single compost applica tions in agreement with a correct practice of fertilization do not appear to endanger at an important extent soil microbiological properties. Beneficial effects are evident in poor soils. Negative influences, especial ly upon mineralization processes, can rise from very heavy applications. These effects are more evident on microbial activities than on the number of responsible microorganisms, therefore due to chemical and/or phisical actions on microbial enzymes. Considering that environmental effects of sludge application to agri cultural lands can also depend on the type of sludge, field and laboratory experiments have been carried out to study the effects of the same sludge stabilized by different methods: liquid aerobically digested, composted in mixture with wood chips and aerobically stabilized in the solid phase in mixture with synthetic inert bulking agents. Within a first experiment (8) a volcanic sandy-loam soil of Naples area (pH=7.1, organic carbon about 3%) has been treated with the three types of sludge on the basis of their inorganic nitrogen content to supply 120 Kg of per hectar supporting the growth of potato. Control plots and equivalent mineral fertilization were included in the experimental design. Microbiological investigations on the control and treated soil samples followed the harvest. Statistical evaluation of the results showed that soil respiration, soil

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microbial biomass measured accor ding the technique of Jenkinson and Powlson (14) modified by Chaussod and Nicolardot (1), ammonification, nitrification and nitrogenase activity were significantly increased only in the samples treated with sludge composted in mixture with wood chips. This treatment had supplied about 8 tons of organic matter per hectar, a very higher quantity in comparison to the other applications (1.77 tons ha−1 in the form of sludge stabilized with inert agents and 0.25 tons ha−1 as liquid sludge). Within a second experiments (10) the three types of sludge have been applicated on the basis of their organic matter content: 9 tons ha−1 of organic matter and, for composted sludges, a second rate of 36 tons ha −1. Mineralization of organic carbon resulted significantly faster in the soil treated with liquid sludge, where an amount of organic carbon higher than the applicated one was mineralized during the experiment (positive “priming effect”). The process appeared more intense also in soils treated with the heaviest doses of composted sludges. Mineral nitrogen reached the highest values in the soil treated with sludge stabilized with inerts after two months, in the soil treated with sludge composted with wood chips after four months. Microbial biomass resulted enhanced in the soil treated with liquid sludge in incubation experiments, in the soil treated with sludge composted in mixture with wood chips in the field. Hydrolytic activity on FDA (fluorescein diacetate) was affected by application rates of materials; in the field, it resulted significantly increased by the treatments with sludge composted with wood chips and, at a lower extent, by the sludge stabilized with inert bulking agents. The various treatments did not produce considerable effects on counts of ammonia producing microorganisms, NH4+– oxidizers and N02− -oxidizers, confirming the evaluation of soil microbial activities be more suitable than counting viable cells in order to assess biological conditions of soils as affected by sludge applications. Heavy metal toxicity towards soil microorganisms is receiving more and more attention (11–12). As for plants, only a fraction of total metal con centration in soil remains available affecting microbial populations too, but this fraction cannot be easily quantified (4). A CdS04 -spiced sewage sludge has been utilized to realize total metal concentrations in soil of 2, 4, 8 and 16 p.p.m. Within a first trial (7) two different soils were treated: a volcanic neutral sandy soil, with a very low clay content and high percentage of vitreous material and a sample of “terra rossa”, rich in iron oxides, with about 70% of clay. Microbiologi cal analyses followed crops growth. The quantification of microbial biomass by ATP measurements has not put into evidence significative differences among the treatments. An adverse influence by 16 p.p.m. of Cadmium only resulted in the volcanic soil. Mineralization of organic carbon proceeded at a similar extent in both the soils, unaffected by metal concentration. Functional groups of soil microorganisms are quite resulted more evidently affected in the volcanic soil than in “terra rossa”. Ammonifiers appeared the most sensitive ones to the action of the metal and their activity resulted really depressed in proportion to Cadmium concentration, especially in the volcanic soil. In the presence of low counts of nitrogen-fixing microorganisms recorded in both soils, nitrogenase activity resulted consi derably influenced by Cadmium, but, surprisingly, more in “terra rossa”, where 4 p.p.m. of total metal reduced at about 50% the efficiency of the process. Cadmium sensitivity of Azotobacter strains isolated from this soil resulted indeed very higher in comparison to those isolated from the volca nic one. These results show the importance of specific and diversified bio-tests evaluating the environmental impact of heavy metals as a correct ecotoxicological procedure. A last experiment (9) carried on a neutral sandyloam soil with 3% of organic carbon, treated with Cd-spiced sewage sludge to get total metal concentrations of 2, 4 and 8 p.p.m. has shown that soil respiration and FDA-hydrolytic activity were unable to evidentia te significant differences among the samples, whilst microbial biomass measured according to the above mentioned modified method of Jenkinson and Powlson allowed to define the adverse influence of Cadmium. Appropriate analytical techniques must be therefore developped for routine controls of soil microbiological properties of sludged and polluted soils.

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3. PRACTICAL CONCLUSIONS Investigations have shown that sludge application at rational rates does not disturb soil microbiological processes. In the soil-plant system the rhizosphere-effect seems to be more important for microbial growth than the effect arising from the addition of sludge. Mycorrhizal infections can disappear, probably as a consequence of the increased phosphorus availability. Repeated correct applications progressively improve biological soil properties. The sandy poorest soils show the most beneficial effects. In the field, microbial biomass is increased by applications of sludge stabilized in the solid phase by composting. Heavy application rates can depress mineralization processes, mainly in soils with a low clay content. Such negative influence is more evident on microbial activities than on the number of responsible microorganisms. Accumulating heavy metals in soil through sludge applications, important microbial processes as ammonification and dinitrogen fixation can result negatively affected at low metal concentration too. The extent of such an influence depends both on soil properties (metal bonding capaci ties) and on genetic characteristics of microbial populations involved. Therefore, only specific bio-assays can quantify the phenomenon. 4.NEEDS FOR FURTHER RESEARCH The quantification of beneficial effects of sludge application on soil biological properties in terms interesting to farmers should be achieved through appropriate experiments. Indeed microbiological investigations appear rarely as unique topics of research. Long-term experiments should be promoted in order to verify the possi bility to improve biological fertility of derelict lands through repeated sludgeing. Limits for heavy metals depressing microbial activities must be defi ned in the field, particularly ascertaining if low metal concentrations can adversely affect microorganisms without negative effects on plant growth. Screening for microbiological methods should suggest more suitable analytical techniques to control the treated soils. ACKNOWLEDGMENTS This paper reports results obtained within researches supported by the Consiglio Nazionale delle Ricerche and by the Ministero per la Pubblica Istruzione, Rome. REFERENCES 1

2

) CHAUSSOD, R. et B.NICOLARDOT (1982)—Mesure de la biomasse microbienne dans les sols cultivés. I. Approche cinétique et estimation simplifiée du carbone facilment minéralisable. Rev. Ecol. Biol. Sol, 19, 501–512. ) CONSIGLIO NAZIONALE DELLE RICERCHE (1982)—Guida alla utilizzazione in agricoltura dei fanghi derivanti dai trattamenti biologici delle acque di scarico urbane. Collana del Progetto Finalizzato “Promozione della qualità dell’ambiente”, AQ/2/13, Roma.

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3

4

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8 9

10

11 12 13 14 15

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) COPPOLA, S. (1981)—Effect of composted sewage sludge on mineralization of organic carbon, ammonification and nitrification in soil. In CEC: “Characterization, treatment and use of sewage sludge”, Proc. of the 2nd Eur. Symp., Vienna, October 21–23, 1980 (L’Ermite, P. and H.Ott Eds.), D.Reidel Publ. Co., Dordrecht, 482–492. ) COPPOLA, S. (1983)—Effect of heavy metals on soil microorganisms. In CEC: “Environmental effects of organic and inorganic contaminants in sewage sludge”, Proc. of a Workshop held in Stevenage, May 25–26 (Davis, R.D., Hucker, G. and P.L’Hermite Eds.), D.Reidel Publ. Co., Dordrecht, 233–243. ) COPPOLA, S. (1983)—Soil microbial activities as affected by applica. tion of composted sewage sludge. In CEC: “The influence of sewage slud ge application on physical and biological properties of soils”, Proc. of a Seminar held in Munich, June 23–24, 1981 (Catroux, G., L’Hermite, P. and E.Suss Eds.), D.Reidel Publ. Co., Dordrecht, 170–195. ) COPPOLA, S. (1983)—Organic soil conditioners from sludges. In: “Biological reclamation and land utilization of urban wastes”, Proc. of the Int. Symp. held in Naples, October 11–14 (Zucconi, F., De Bertoldi, M. and S.Coppola Eds.), La Buona Stampa, Ercolano, 317–342. ) COPPOLA, S., DUMONTET, S., PONTONIO, M., BASILE, G. and P.MARINO (1985) -Effect of Cadmiumbearing sewage sludge on crop plants and microorga nisms in two different soils. Submitted for pubblication to Agriculture, Ecosystems & Environment. ) DUMONTET, S. and E.PARENTE (1984)—Effets d’une boue de station d’épu ration sur la microflore tellurique et sur ses activités. Ann. Fac. Sci. Agr. Univ. Napoli, Portici, IV, 18, 9–36. ) DUMONTET, S., PARENTE, E. e F.VILLANI (1985)—Effetto del Cadmio sulla biomassa microbica di un suolo vulcanico. Conv. Naz. su “Inquinamento idrico e conservazione dell’ecosistema”, Vico Equense, 22–23 febbraio. ) DUMONTET, S., PARENTE, E. and S.COPPOLA (1984)—Mineralisation of organic matter in soil treated with sewage sludge stabilized by differ ent methods. CEC Seminar on “Long term effects of sewage sludge and farm slurries application”, Pisa, September 25–27. ) DUXBURY, T. (1981)—Toxicity of heavy metals to soil bacteria. FEMS Microbiol. Letters, 11, 217–220. ) DUXBURY, T. and B.BICKNELL (1983)—Metal-tolerant bacterial populati ons from natural and metalpolluted soils. Soil Biol. Biochem., 15, 243–250. ) GERDERMANN, J.W. (1975)—Vesicular-arbuscular mycorrhizae. In: “The development and function of roots” (Torrey-Clarkson Eds.), 229. ) JENKINSON, D.S. and D.S.POWLSON (1976)—The effect of biocidal treat ment on metabolism in soil. V. A method for measuring soil biomass. Soil Biol. Biochem., 8, 209–213. ) MENGE, J.A., STEIRLE, D., BAGYARAI, D.J., JOHNSON, E.L.U. and R.T. LEONARD (1978)— Phosphorus concentrations in plants responsible for inhibition of mycorrhiza infection. New Phytolog., 80, 575–578. ) PERA, A., GIOVANNETTI, M., VALLINI, G. and M. DE BERTOLDI (1983)—Land application of sludge: effects on soil microflora. In CEC: “The influ ence of sewage sludge application on physical and biological properties of soils”, Proc. of a Seminar held in Munich, June 23–24, 1981 (Catroux G., L’Hermite, P. and E.Suss Eds.), D.Reidel Publ. Co., Dordrecht, 208–228. ) PERA, A., VALLINI, G., SIRENO, I., BIANCHIN, M.L. and M.DEBERTOLDI (1983)—Effect of organic matter on rhizosphere microorganisms and root development of Sorghum plants in two different soils. Plant and Soil, 74, 3–18. ) TOMATI, U., GRAPPELLI, A. and E.GALLI (1981)—Biological activities in a soil-plant system after treatment with different amounts of digest ed sludge. Pot experiments. In CEC: “Characterization, treatment and use of sewage sludge”, Proc.of the 2nd Eur. Symp., Vienna, October 21–23, 1980 (L’Hermite, P. and H. Ott Eds.), D.Reidel Publ. Co., Dordrecht, 553–561. ) TOMATI, U., GRAPPELLI, A. and E.GALLI (1983)—Sludge effect on soil and rhizosphere biological activities. In CEC: “The influence of sewage sludge application on physical and biological properties of soils”,

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Proc. of a Seminar held in Munich, June 23–24, 1981 (Catroux, G., L’Hermite, P. and E.Suss Eds.), D.Reidel Publ. Co., Dordrecht, 229–241. ) TOMATI, U., GRAPPELLI, A. and E.GALLI (1984)—Soil microorganisms and log-term fertility. CEC Seminar on “Long term effects of sewage sludge and farm slurries application”, Pisa, September 25–27.

EFFECTS OF LONG-TERM SLUDGE ADDITIONS ON MICROBIAL BIOMASS AND MICROBIAL PROCESSES IN SOIL S.P.MCGRATH & P.C.BROOKES Soils and Plant Nutrition Department, Rothamsted Experimental Station, Harpenden, Herts., AL5 2JQ., U.K.

Summary Metal-contaminated sewage sludge was applied to a field experiment from 1942 to 1961 and now, more than 20 years after the last application, metal contents of the soils are still at about current UK maximum permitted limits. In 1983, there was only about half as much total microbial biomass in sludge-treated soils compared to those receiving farmyard manure. Nitrification of added NH4-N, soil dehydrogenase activity and nitrogen fixation (measured by acetylene reduction) was also less in contaminated soils. In contrast, mineralisation of native soil organic N and soil phosphatase activity were unaffected. Counts of bacteria, fungi, actinomycetes and protozoa were not significantly different in treated soils with high or low metal contents. 1. INTRODUCTION The implications of additions of metal-contaminated sewage sludge to soil on soil microbial activity are not as well known as the effects on higher plants. This is because there are no straightforward techniques to measure microbial activity. The methods of measuring either total amounts of microbial biomass or the contributions of different taxa are often time-consuming and subject to large estimation errors. Hence, the relative sensitivities of various microbial processes to metals in soils under realistic conditions have not been studied. However, newer, more precise techniques to measure the entire biomass have been developed recently (1, 2) and if used with suitable controls they can give useful information on the effects of various soil treatments (3). The problems of interpreting the results of such measurements centre on the fact that the size of the biomass reflects management and soil factors such as the substrate inputs as well as the physicochemical environment (3, 4). Therefore field experiments with treated plots in a reasonable state of equilibrium, are ideal. This avoids short-term effects such as the presence of excess soluble salts applied with fresh sludge, which could obscure metal effects or the lack of equilibrium between recently applied inorganic metal salts and other soil constituents. The field experiment from which samples were taken began in 1 942 at Woburn, England, on a sandy loam soil, with 9% clay and CEC of 12 meq 100g−1 to test the effect of various bulky organic manures each of which was applied at two rates, single and double. For the laboratory measurements soils from four replicates of the following treatments were sampled: 1) anaerobically-digested, lagoon-dried sewage sludge from West London, applied at 8.2 and 16.4 t ha−1 year−1 organic matter (OM) on a dry matter basis from

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1942–61, 2) sewage sludge-straw compost, 5.9 and 11.8 t ha−1 year−1 OM from 1942–61, 3) farmyard manure (FYM), 5.2 and 10.4 t ha−1 year−1 OM from 1942–67, and 4) inorganic fertilizers (NPK) from 1942– 67. After the organic manure treatments ceased all plots were given uniform annual additions of inorganic (NPK) fertilizers. Mainly vegetable crops were grown until 1972, the whole experiment was then put down to grass. In the autumn of 1982 the grass was ploughed in. The plots which received organic manures were of similar total carbon, mineral N, available P & K and pH status (Table 1). FYM treatments are used here to assess the effect of metals added in the metalcontaminated sludge treatments because the FYM added OM but only small amounts of metals (5). The range of extractable metal concentrations in the soils are summarised in Table 2. Soils from sludge and sludge compost-treated soils are referred to as ‘high metal’ and FYM and inorganic fertilizer treatments as ‘low metal’. Table 1. Total carbon, pH, available phosphorus and potassium and mineral nitrogen in soils from the Market Garden Experiment, Woburn in 1983 (means and standard errors). %C

pH* mg kg−1

Treatment Inorganic fertilizer FYM Sludge * in 1:25 (w/v) soil: H2O.

1.16±0.133 1.61±0.119 1.75±0.059

6.78 ±0.025 6.66±0.050 6.53±0.031

115±5 131 ±5 130±3

229±11 226 ±11 164±5

3.0±0.25 3.7±0.23 4.9±0.45

2. EXPERIMENTAL In 1983, 22 years after the last application of sewage sludge, thirty soil cores were taken to a depth of 23 cm from each plot with a 5 cm diameter stainless steel semicylinder auger and bulked. The soil was sieved moist to >Incineration>Sea>Unspecified, including storage. This order is unlikely to change rapidly. A variety of novel sludge disposal strategies is considered including, e.g., earthworm conversion, use as an animal feed supplement, thermal conversion to liquid and solid fuels. Most are unlikely to account for large amounts of sludge in the future, however, thermal conversion to liquid and solid fuels appears promising and might largely replace incineration. 1. INTRODUCTION Sewage sludge is a nutrient rich largely organic by-product of municipal wastewater treatment. It must be removed from the treatment facility and disposed of in an environmentally acceptable and cost-effective manner. It represents a major waste management problem for cities of all sizes and the magnitude of the problem is increasing due to expansion of wastewater treatment facilities and improved treatment methods. Sludge treatment and disposal is frequently the most costly phase of municipal wastewater treatment and can account for as much as 60% of the operating budget. Consequently, there is a considerable economic incentive to improve sludge treatment and disposal methods which results in continually changing practice. The purpose of this report is to update Working Party 5 of the CEC COST 68 USE OF SLUDGE program concerning current and probable future sludge production and disposal practices in Europe and North America.

1Environment Canada, Environmental Protection Service, Wastewater Technology Centre, Burlington, Ontario, L7R 4A6, Canada 2RIVM-LAE Antonie van Leeuwenhoeklaan 9, Postbus 1, 3720 BA Bilthoven, The Netherlands 3Swedish Environment Protection Board, Box 1302, 17125 Solna, Sweden

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2. SLUDGE PRODUCTION AND CURRENT DISPOSAL STRATEGIES Several recent inventories of sludge production and disposal in the European countries have been published (1, 2, 3, 4, 5). They vary as regards the number of countries considered and the nature and amount of data presented, however, comparable data generally agree favorably. The format of the inventory by Matthews (3) was best suited to the objectives of this report and it was adopted for Table 1. TABLE 1. INVENTORY OF MUNICIPAL SLUDGE PRODUCTION AND DISPOSAL IN EUROPE AND NORTH AMERICA Country

Population Sludge Disposed 106

Disposal Methods %

103 dry tonnes/ Agricultural year Land Incl. Hortic ulture & Gardens

CEC COUNTRIES Ireland 3.4 21 5 FRG 61.7 1690 32 France 53.6 510 24 Belgium 9.9 57 — UK 55.9 1210 45 Italy 56.7 800 30 Netherlands 13.9 258 58 Luxembourg 0.4 11 90 Greece 9.5 3 — Denmark 5.1 156 32 OTHER COUNTRIES Norway 4.1 70 40 Finland 4.8 130 41 Sweden 8.3 250 60 Switzerland 6.4 170 71 Spain 37.0 45 60 Austria 7.5 150 47 Canada 24 287 42 USA 220 7000 42 Sources: Primarily (2, 8) but also (1, 3, 4, 5, 6, 7).

Other Land Mostly Landfill

Incineration Sea

Unspecified

57 56 45 Most 15 50 27 10 100 27

— 10 31 — 4 20 1 — — 41

38 2 — — 31 — 9 — — —

— — — — 6, most to agr. — 5 — — —

40 37 30 — 20 — 18 15

— — — 29 — 20 40 27

5–10 — — — 20 — — 4

10–15 22 10 — — 33 — 12

In general, estimated sludge production (Table 1) increases with increasing population. However, some countries with similar populations (e.g., Norway and Finland, and Belgium, Greece and Sweden) produce widely different amounts of sludge. These differences probably result from different proportions of the populations being served by sewage treatment and different degrees of treatment. Estimates of future sludge production are not presented because experience has proven them to be unreliable. They do not take into account changing economic and political constraints and frequently are based upon different assumptions.

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However, it is reasonable to assume that sludge production will increase in the future with improved treatment technology and increasing populations served. Estimated land disposal (i.e., agricultural and other land) accounts for a much larger proportion (frequently>70%) of the sludge production in Europe and North America than other disposal methods (Table 1). Values differ among countries but, on average, equal proportions of sludge production are utilized for agricultural and horticultural purposes as are disposed on other land, mainly in landfills. Sludge disposal by incineration, dumping at sea and unspecified methods generally account for small proportions of sludge production. 3. FUTURE SLUDGE DISPOSAL STRATEGIES The choice of a sludge disposal strategy is based on a variety of considerations including costs, regulations and public opinion. Costs may be highly variable depending upon economic conditions and a number of site specific factors including transport distance and fuel, land and labour requirements. Increased regulation generally increases cost due to increased monitoring and surveillance requirements. Strong public opposition to a strategy may force its abandonment. The following comments concerning future sludge disposal strategies assume no radical change from present conditions for any of the above factors. Agricultural Land Including Horticulture And Gardens Sludge disposal on agricultural land including horticulture and gardens is practiced widely in Europe and North America and, although it is an environmentally sensitive outlet and is in direct competition withanimal manures, it will likely continue to be practiced widely in the future. Recent initiatives such as the development of national sludge use guidelines and the CEC COST 68 USE OF SLUDGE program have focused interest on this disposal method and are providing the information framework required for its management. It is frequently the most economic, simple and convenient method of sludge disposal and is particularly well suited to small and medium sized urban communities with easy access to land. It is a way of recycling a valuable resource and, with some exceptions, represents a free supply of nitrogen and phosphorus to farmers thus reducing their expenditures on artificial fertilizers. Moreover, spreading sludge thinly over land significantly reduces air, surface water and groundwater pollution problems which may result from concentrating large amounts of it in a small area. Current information indicates that heavy metal, particularly cadmium, buildup in soil represents the most significant human health risk associated with sludge use on food producing land. There is considerable uncertainty concerning safe concentrations of heavy metals in soil and this is reflected by large differences among national sludge use guidelines (9). However, this is an active area of research and it is probable that the uncertainty will be resolved in the future. In the meantime, there is concern, particularly on the part of countries with the more lenient guidelines, that decisions which may curtail their land utilization operations should be based upon accepted scientific evidence and not on economic or political pressures. Composting: Composting is a sludge treatment process involving aerobic thermophilic decomposition of organic matter and it is practiced widely as a component of disposal by land utilization. Its recent use has increased dramatically, particularly in the USA (10), and its future use is likely to continue to increase with increasing sludge production and increasing constraints on disposal techniques. The process improves the acceptability of sludge to farmers and householders; the product has a favourable odour and appearance and a high humus content similar to muck soil. Its primary value is as a soil replacement (11) or conditioner and

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not as a fertilizer material because it contains only small concentrations of the major plant nutrients; nitrogen, phosphorus and potassium. Composting may be done in either confined (mechanical) or unconfined (windrow or static aerated pile) systems. Confined systems are more widely used in the European countries, particularly the FRG, than in North America. In many European countries municipal sludge is mixed and composted with a variety of other solid wastes including municipal refuse, leaves and bark and much of the product is used in public gardens and recreation areas. A variation on sludge composting is employed in the Netherlands to produce “black earth”. Sand or peat is mixed with sludge and the mixture is composted in windrows. The heavy metal content of black earth is regulated and it is recommended for use as a soil replacement, however, research into its hygienic quality is ongoing. There is a large demand for the product. Sale or give-away: The sale or give-away of bagged or bulk, processed sludges (composted, heat dried, etc.) has been practiced in many countries. Frequently, the processed sludge is given a proprietary name such as “Milorganite”, “Philorganic”, “Compro”, etc. and in some cases it is fortified with inorganic fertilizer to enhance its nutrient, particularly nitrogen, content. Processing improves the acceptability of sludge as a soil replacement or amendment. However, it is generally expensive and economic analysis should be conducted prior to its implementation to determine the likelihood of cost-recovery. The sale or give-away of composted sludge is likely to increase in the future with increasing composting, but of heat dried sludge is likely to decrease as heat drying is phased out due to high energy costs. Other Land, Mostly Landfills Landfills: In many European countries disposal by landfilling (tipping) accounts for as much or more municipal sludge as agricultural utilization (Table 1). Landfilling is practiced widely at the present time and is likely to be practiced widely in the future. Recently, its use in the FRG has increased due to increasing regulation of agricultural utilization and there are indications that this may also occur in other countries. Landfilling involves application of sludge to land and subsequent covering with a layer of soil. The sludge should be stabilized and dewatered to a solids content of ≥15% prior to landfilling (>35% in the Netherlands). Co-disposal of sludge with municipal solid waste (refuse) is the most common landfilling method. The sludge and refuse are thoroughly mixed prior to spreading, compaction and coverage with soil. Liquid sludge may be co-disposed with refuse provided that the absorptive capacity of the refuse for water is not exceeded. Landfilling is simple and convenient, however, it offers no opportunity for resource recovery. Moreover, there are increasing concerns about groundwater pollution resulting from landfills receiving sludge. In the Netherlands all new landfills must be lined (generally with plastic) and leachate quality must be monitored. A recent expansion of the Burlington, Ontario landfill was approved subject to no sludge disposal. Dedicated Disposal Sites: Sludge disposal on other land presents similar advantages and disadvantages as does landfilling. It is practiced to a limited extent in the UK on former sewage farms and in the USA. Its use may increase in the USA where it has been suggested that sludge application on limited land areas may be preferable to wide spread application. In the UK, agricultural produce grown on sewage farms is monitored to insure that its consumption will not result in human and animal health problems. In the USA, it is recommended that food chain crops not be grown on dedicated disposal sites but if they are grown, monitoring similar to that in the UK is required. Surveillance of surface and groundwater qualities is required. Non-Arable Land, Land Reclamation and Waste Cover: Sludge disposal on non-arable land and for land reclamation and waste cover (derelict land) is practiced sparingly at the present time but is likely to

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increase in the future. For example, it is estimated that 33 000 ha of land in the UK require reclamation, affording a potential outlet for 3.3 million dry tonnes of sludge applied at 100 dry tonnes per ha (12). A guide for revegetation of mined land in the eastern USA using municipal sewage sludge has been prepared (13). Derelict land refers in the main to. areas without an established soil cover such as colliery waste, shale tips, mine waste, china clay waste tips, sand dunes and the sides of embankments and cuttings, or pits filled with overburden from open-cast coal mining or with urban refuse and then covered with a layer of soil. In most cases the surface material is severely deficient in organic matter, leading to a lack of water holding capacity and poor soil structure, and it is almost always deficient in nitrogen and phosphorus. The composition of sewage sludge is almost exactly complimentary to the needs of derelict land. It contains large quantities of organic matter, nitrogen, phosphorus (especially in limed sludges) and trace elements. A major advantage of sludge over most inorganic amendments for land reclamation is that the nutrients (particularly nitrogen) are released slowly such that one application of sludge may enhance growth for several years. In experiments conducted in Norway (11), the UK (12, 14) and the USA (15), reclamation of disturbed lands with sludge has proven much more successful than with inorganic soil amendments. The reclaimed soil may or may not be suitable for the production of food chain crops depending upon the degree of contamination of the original soil material and the amounts of heavy metals added in sludge. Forest Land: Sludge disposal on forest land is practiced sparingly at the present time but is of considerable interest and is likely to increase in the future. It is perceived by the public to be preferable to application on agricultural land because very little forest vegetation is included in the human and animal food chain. Experiments conducted in Denmark (16), France (17), the Netherlands (18), and the USA (15) indicate that it generally results in greatly increased tree growth. Draft guidelines for sludge utilization on forest lands in the USA have recently been prepared (19). However, forest application of sludge is not a panacea. Forests are frequently located large distances from the sludge supply and application is difficult or impossible due to inaccessability. In Washington State, USA an expensive all terrain vehicle with a tank and high pressure gun has been developed to spray sludge into established tree stands from trails cut at regular intervals. Application to forest land is much simplified following tree harvest but results in greatly enhanced vegetative growth which competes with seedling trees and attracts grazing animals which may damage the trees. In addition, sludge application rates must be regulated to avoid nitrate-nitrogen contamination of groundwater (18, 19) and long-term application may be detrimental to conifers due to increased soil pH, increased mineralization of the humus ground cover and increased danger of root rot (18). Thus, there are many unsolved problems associated with forest application of sludge and a final assessment of its suitability for wide scale practice is not possible at the present time. Incineration The proportion of sludge disposed of by incineration in the European countries is unlikely to increase, and may decrease, in the future whereas in North America it is likely to remain constant or increase. Objections to incineration include high cost and the risk of atmospheric pollution. Several incinerators have been shutdown in the UK due to the high cost of energy and an application for a new incinerator was refused on environmental planning grounds (20). Sludge incineration is prohibited in Sweden due to the possible release into the atmosphere of mercury and dioxins. In North America incineration continues to be used for sludge disposal from large urban centres where transport costs to agricultural land are high and road haulage would create a considerable traffic nuisance. However, high energy costs have resulted in a change in approach to incineration which has included improvements in sludge dewatering and the installation of systems to recover some of the heat generated by the combustion process.

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Sea The future of sludge disposal at sea is unclear at the present time. It accounts for appreciable proportions of sludge production in only a few countries (Table 1), however, particularly in the UK, it is considered to be an important and environmentally acceptable management alternative (21). Sludge disposal at sea comes under the purview of the Oslo Convention (1974) relating to the prevention of marine pollution by dumping from ships and aircraft, and the London Convention (1975) relating to the global prevention of marine pollution by dumping wastes and other matter. Both of these conventions provide for: (a) prohibition of the disposal of specified dangerous substances (List 1) e.g., mercury, cadmium and organochlorine compounds; (b) control by issue of special permits of the disposal of other less toxic substances (List 2) e.g., lead, zinc, copper and arsenic; and (c) control by general permits of List 3 materials which include sewage sludge. Provided List 3 materials contain only traces of Lists 1 and 2 substances, they qualify for disposal under a List 3 general permit. The definition of “trace” in this context is not yet fully agreed but a reasonable yardstick is that the Lists 1 and 2 substances should not be present in sludge at such levels that disposal of the sludge leads to unacceptable concentrations of these substances in marine organisms. Several years of research at longstanding UK sea disposal sites have identified no serious adverse effects on the marine environment (21). Similar findings have been reported for most USA sea disposal sites (22), however, heavy degradation has been reported at the relatively confined site 12-miles off the New York Bight Apex (23). It is generally acknowledged that sludge has contributed to the degradation, and further sludge disposal at this site was banned by the U.S. EPA as of 1 April, 1985. Unspecified Records are incomplete and some sludge disposal is unspecified. Long-term storage probably accounts for much of the sludge in this category, however, increasing production and high land costs are expected to render storage uneconomic in the future. Novel Uses of Sludge In addition to the foregoing conventional sludge disposal techniques, there is interest in a large number of novel sludge uses, some of which are offshoots or components of the conventional techniques. With few exceptions, they have not been adopted on a large scale but are under research and development or are practiced at individual locations. A resume of several such uses follows. Earthworm Conversion: Earthworm conversion (vermicomposting) of sludge is under study in the USA (10, 24) and Canada. Vermicomposting, like conventional composting, improves the acceptability of sludge to farmers and householders. It is a process in which earthworms convert large volumes of organic wastes to earthworm manure, called castings. Dewatered aerobically digested sludges appear to be the most satisfactory substrates, however, raw sludges have also been used. Anaerobically digested sludges generally exhibit toxicity and require pretreatment prior to vermicomposting. The conversion of sludge to castings increases the surface area enormously, and this accelerates drying and microbial activity. The castings are a finished compost that will not burn plants. In addition, they crumble easily and have a favourable odour and appearance. They are the main product of the process and excess worms, if any, are considered a byproduct. Worm additions may be required to maintain the process because the worms reproduce via egg capsules which may be lost when the castings are removed. The City of Lufkin, Texas has practised full-scale vermicomposting since 1979, however, due to the experimental nature of the project, the state Health Department does not allow sale of either the castings or

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worms (10). Based on Lufkin data it is estimated that the cost of sludge processing is from $175—$200 US per dry tonne. In the UK a ground bait producer has successfully used both raw and activated sludges as substrates for earthworm production (24). Microbial quality and concentration in the sludge determined growth rate of the worms. An adequate level of cellulosic material was required for efficient nitrogen utilization. Feed supplement: A considerable number of studies have been conducted with sludge as a feed supplement for cattle, swine and poultry (24, 25). The sludge has usually constituted ≤20% dry wt of the diet and has not reduced its palatability. However, reduced palatability for cows of a feed supplement containing 62% sludge solids was reported. Cadmium, zinc, copper and lead concentrations in the kidney, liver or bone tissues of experimental subjects have generally exhibited significant increases due to sludge consumption, however, the subjects have remained healthy and there have been no visual symptoms of pathology. Inspection of carcasses following the experiments has generally confirmed the absence of pathology. Short-term studies of sludge supplementation to fish diets were considered encouraging but indicated a need for long-term studies (24). Although sludges contain protein, lipid and vitamin B12 which are of interest as feed supplements, they contain low levels of digestible nutrients and metabolizable energy and there is increasing evidence that their main effect is to act as a diet diluent. This being so they may serve a useful purpose in maintenance diets, but are unsuitable as a constituent of high energy diets. Lipid extraction: Studies conducted in the UK indicated that primary sludge contains 30–35% extractable lipid on a dry wt basis (24). Two thirds of the isolate was found to be saponifiable and the remaining material consisted of long chain hydrocarbons, aliphatic esters and long chain primary alcohols. The saponifiable fraction resembled tallow and was a suitable raw material for detergent manufacture. Economic analysis indicated that lipid extraction could become financially attractive and might comprise an important element of future sludge disposal strategy. Protein extraction: A further process was developed to extract protein from sludge following the lipid extraction described above (24). Yields of 15% protein isolate on a dry wt basis were obtained. The isolate contained approximately 50% true protein with an amino acid profile and content similar to soybean meal. Toxic metal carry-over into the isolate was sufficient to severely limit its inclusion in animal feedstuffs. Methods to prevent the carry-over have been developed but have not been tested on a commercial scale. Vitamin B12 extraction: Study in the UK indicated that vitamin B12 could be extracted from sewage sludge, however, the process would be uneconomic unless it offered some advantage in addition to production of the vitamin (24). Metal recovery: Efforts have been undertaken to evaluate the potential for metal recovery from municipal sewage sludges. Both acid and heat treatment processes have been considered, however, these processes require considerable chemical and/or energy inputs and in general have not proven cost-effective. A promising discovery was that sludge incinerator ash from Palo Alto, California contained levels of platinum and silver (26) and from Toronto, Ontario contained levels of gold and silver (27) higher than many commercial ores. Precious metals are being recovered from these incinerator ashes. Chemical fixation and polyethylene and asphalt encapsulation: The purpose of these processes is to treat waste so that it can be safely landfilled. They are expensive and large scale use has been limited to hazardous industrial wastes, however, they may have application for municipal sewage sludges. There has been limited small scale testing with sewage sludges (28). Chemical fixation involves mixing various chemicals such as cement, limestone, fly ash, or lime with sludge. The product of this process has improved physical properties and is less likely to leach than untreated sludge when landfilled.

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Polyethylene encapsulation involves fusing (at 180°C) sludge within a 6 mm thick seamless layer of polyethylene. Leaching tests indicate virtually no release of chemical constituents from sludge treated in this manner. Asphalt encapsulation involves mixing waste with asphalt at 150°C so that each particle is completely coated. The coated particles are then placed in containers where they cool and form a solid non-porous mass which is highly resistant to leaching, mechanical damage and bacterial attack. About 1 kg of asphalt is required for each kg of waste dry solids. Overseas shipment: The overseas shipment of municipal sewage sludge from the USA has been examined extensively (26). Shipment in ore boats, oil tankers or sludge ships to Caribbean, Central American, African and Middle and Far Eastern countries for use as soil conditioner has been proposed. To date, no such projects for large scale shipment of USA sludge have been implemented. It is illegal to export wastes from the Netherlands. Brick manufacture: Laboratory and commercial scale tests using municipal sewage sludge as a clay replacement in brick manufacture have been conducted in the Netherlands and the USA (29). Results in the two countries were very similar and indicated that sludge-amended brick had the look, feel and smell of standard brick. Bricks with 30–40% sludge (the Netherlands) and ≤30% sludge (the USA) by volume in the wet mix met the appropriate technical standards. Suggested advantages of sludge-amended manufacture include economic use of clay, water and energy, and a light-weight product with improved water absorption and transfer properties. Further research is necessary on air quality effects, fugitive metal and organic emissions and the possibility of occupational hazard to plant personnel. Thermal conversion to liquid and solid fuels: Numerous sludge processing options including anaerobic digestion, starved air incineration gasification and liquefaction have the potential to convert sludge organic matter into energy (30). Anaerobic digestion is widely practiced and about 25% of the organics are converted resulting in an energy recovery of about 5 MJ/kg of dry sludge fed to the digester. The most advanced new energy recovery technology is the Hyperion Energy Recovery System currently being installed in Los Angeles, California. This system comprises digestion, dewatering, Carver-Greenfield dehydration and starved air fluid bed incineration of the sludge derived fuel. A total of 25 MW of electricity will be produced from 265 dry tonnes of sludge per day. This process is designed assuming a net energy production of 8.15 MJ/kg of dry sludge. By contrast, sludge liquefaction technology is still in its infancy, and data reported in the literature are based on laboratory scale systems. Researchers from Batelle Northwest Laboratories have reported on a sophisticated process consisting of sludge alkaline pretreatment and subsequent autoclaving at 320°C for one hour at approximately 10 000 kPa under an argon atmosphere. This process produces oil, asphalt and char, with oil yields ranging up to 15% of sludge dry wt. Total thermal efficiency ranges up to 70% and the net energy production is about 5.9 MJ/kg of dry sludge. This is based on the assumption that the oil represents the net energy. In another approach, both raw and digested dry sludge were processed with a carrier oil in an autoclave at temperatures ranging from 396–420°C under hydrogen at 10000–13000 kPa. Oils and asphaltenes were produced, with oil yields up to 30% of sludge dry wt. The most promising technology has been studied in the FRG where researchers from Tubingen University have demonstrated a very simple process to convert sewage sludge to oil and char (31). The process comprises heating dried sludge to 300–350°C in an oxygen free environment for about 30 minutes. Significant advantages of this process are that it operates at only slightly above atmospheric pressure and that no additives are required. It is postulated that vapour phase catalysed reactions convert the organics in sludge to straight chain hydrocarbons much like those present in crude oil. The FRG researchers have

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demonstrated oil yields ranging from 18–27% and char yields from 50–60% of sludge dry wt. The heating value of the oil was approximately 39 MJ/kg and of the char approximately 15 MJ/kg. Energy balance calculations indicate that the process is a net producer of energy provided the sludge is mechanically dewatered to about 20% solids. It was estimated that a net energy production of 7–10 MJ/kg sludge dry wt could be demonstrated at full scale. Experiments using both batch and continuous reactor systems, haveconfirmed that Canadian sludges are amenable to conversion to liquid and solid fuels (30). Oil yields greater than 25% appear to be readily achievable and thermal efficiencies greater than 95% easily attained. Net energies of 7–10 MJ/kg of dry sludge should be achievable for the thermal conversion process provided mechanical dewatering of the sludge is practiced. It is estimated that by 1985, 350 000 tonnes of sewage sludge will be incinerated annually in Canada. Thermal conversion of this sludge could produce 700 000 barrels of oil per year, with a market value of at least $21 million. ACKNOWLEDGEMENTS The authors express appreciation to Dr. K.Aichberger, Austria, Dr. E.Vigerust, Norway and Dr. P.J.Matthews, UK, for supplying up-to-date sludge production and disposal data for their countries. REFERENCES 1

2 3

4

5 6 7 8 9

10 11 12

() Klein, L. La politique de 1’environnement des communautes Europeenes et la valorisation des dechets en agriculture. In Characterization, Treatment and Use of Sewage Sludge, eds. P.L’Hermite and H.Ott, EUR 7076, D.Reidel Publishing Co., pp 3–10, 1981. () Vincent, A.J. and R.F.Critchley. A Review of Sewage Sludge Treatment and Disposal Practices in Europe. Water Research Centre Report 442-M/1, Medmenham, 1983. () Matthews, P.J.Sewage sludge utilization in European countries which are members of EWPCA and sewage sludge utilization in the EEC—An overview. In Sewage Sludge In Agriculture, House of Lords Select Committee on the European Communities, Session 1983–84 1st Report, HMSO London, pp 122–141, 1983. () Dam Kofoed, A. Optimum use of sludge in agriculture. In Utilization of Sewage Sludge on Land: Rates of Application and Long-Term Effects of Metals, eds. S.Berglund, R.D.Davis and P.L’Hermite, EUR 8822, D.Reidel Publishing Co. pp 2–20, 1984. () Concerted Action Treatment and Use of Sewage Sludge, COST 68 ter. Final Report of the European Community—Cost Concertation Committee, I. General Report, 1983. () Webber, M.D. Waste metals—The Canadian approach for limiting metals on land from municipal sludges. Abstracts, NE Branch, ASA; E Section CSA, pp 5–9, 1984. () Booz-Allen and Hamilton, Inc. Description and Comparison of Municipal Sewage Sludge Generation and Disposal Data Bases . Office of Solid Waste, U.S. Environmental Protection Agency, Washington, D.C., 1982. () Eurowater. ed., P.J.Matthews. Thunderbird Enterprises Ltd., 1984. () Webber, M.D., A.Kloke and J.Chr.Tjell. A review of current sludge use guidelines for the control of heavy metal contamination in soils. In Processing and Use of Sewage Sludge, eds. P.L’Hermite and H.Ott, EUR 9129, D.Reidel Publishing Co., pp 371–386, 1984. () Managing Sludge by Composting, eds., the staff of BioCycle, Journal of Waste Recycling. The JG Press, Inc., 1984. () Vigerust, E. Use of sewage sludge on green area. Ibid ref. 4, pp 36–46, 1984. () Hall, J.E. and E.Vigerust. The use of sewage sludge in restoring disturbed and derelict land to agriculture. Ibid ref. 4, pp 91–102, 1984.

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13

14

15 16 17 18 19

20 21 22 23 24 25 26

27 28 29 30 31

() Sopper, W.E. and E.M.Seaker. A Guide for Revegetation of Mined Land in Eastern United States using Municipal Sludge. School for Forest Resources and Institute for Research on Land and Water Resources, The Pennsylvania State University, University Park, Pennsylvania, 1983. () Coker, E.G., R.D.Davis, J.E.Hall and C.H.Carlton-Smith. Field Experiments on the Use of Consolidated Sewage Sludge for Land Reclamation: Effects on Crop Yield and Composition and Soil Conditions, 1976–1981. Water Research Centre Technical Report, TR 183, Stevenage, 1982. () Land Reclamation and Biomass Production with Municipal Wastewater and Sludge, eds., W.E.Sopper, E.M.Seaker and R.K.Bastian, The Pennsylvania State University Press, 1982. () Grant, R.O. and S.E.Olesen. Sludge utilization in spruce plantations on sandy soils. Ibid ref. 4, pp 79–90, 1984. () Thomann, Th. Experimental study on the use of urban sewage sludge on Mediterranean forests. Ibid ref. 4, pp 61–78, 1984. () van den Burg, J. Forestry and waste application in the Netherlands. A paper presented to the Eighth World Forestry Congress, Jakarta, 1978. () Urie, D.H., J.H.Cooley and A.R.Harris. Guidelines for Land Treatment of Wastewater and Sewage Sludge on Forests and Wildlands (a draft). North Central Forest Experiment Station, U.S. Department of Agriculture— Forest Service, East Lansing, Michigan, 1983. () Bruce, A.M., H.W.Campbell and P.Balmer. Developments and trends in sludge processing techniques. Ibid ref. 9, pp 19–38, 1984. () Fish, H.Sea disposal of sludge—the U.K. experience. Wat. Sci. Tech. 15; 77–87, 1983. () Ocean Disposal Systems for sewage sludge and Effluent. A National Research Council Report, National Academy Press, Washington, D.C., 1984. () EPA denies 12-mile site for ocean dumping of sludge. Sludge Newsletter 10; 58, 1985. () Progress in Research. Department of the Environment/National Water Council Standing Committee on the Disposal of Sewage Sludge, London, 1984. () Sludge—Health Risks of Land Application, eds. G. Bitton, B.L. Damron, G.T.Edds and J.M.Davidson, Ann Arbor Science Publ. Inc., 1980. () Bastian, R.K. EPA comprehensive review of municipal sludge management alternatives. In Proceedings of a National Conference on Municipal and Industrial Sludge Utilization and Disposal, Atlantic City. The Hazardous Materials Control Research Institute, Silver Springs Md., pp 3–13, 1983. () Preliminary Evaluation of Precious Metals Extraction from Incinerated Sewage Sludge. A report prepared by R.L.Pyne Metallurgical Consultants Inc., Toronto, 1984. () Process Design Manual for Sludge Treatment and Disposal. U.S. Enviromental Protection Agency, EPA 625/1– 79–011, pp 13–1 to 13–9, 1979. () Alleman, J.E. Beneficial use of sludge in building components. Final Report to the National Science Foundation, Program for Civil and Environmental Engineering, 1983. () Bridle, T.R., H.W.Campbell, A.Sachdev and I.Marvan. Thermal conversion of sewage sludge to liquid and solid fuels. In Proceedings of Canadian Society for Chemical Engineering Conference, Toronto, 1983. () Bayer, E. and M.Kutubbudin. Low temperature conversion of sludge and waste to oil. In Proceedings of the International Recycling Congress, Berlin, 1982.

DISCUSSION McGRATH (UK):

What are the problems in dealing with the waste remaining from the thermal processes producing liquid and solid fuel from sewage sludge? WEBBER (Canada): The wastes are often more difficult to deal with than the original materials. The oil extraction process leaves a material which is similar to incinerator ash and may be disposed in landfill. The wastes from acid extraction of metals do present a disposal

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problem. I have no information at this time on the disposal of the residue from lipid extraction processes.

CONCLUSIONS

1. Improved knowledge of natural geochemical background levels of metals in soils is needed in order to give better assessment of soil pollution from anthropogenic sources. The latter is likely to be confined to the surface or cultivated layer of agricultural soils. Therefore, it can be separated from the pedological and geological load of metals naturally associated with the soil by examining the depthwise distribution of metals in soil profiles. Often the natural background metal load of the soil is less available for crop uptake and hence less hazardous than metals introduced to the soil from anthropogenic sources such as sewage sludge. Thus the background level of metal in soils may in some circumstances be subtracted from the guideline soil metal limits to be applied where soils receive sewage sludge. (Kuntze’s paper). 2. Cadmium uptake from sludge-treated soils was reduced due to an antagonistic effect of nickel in an experiment in which high levels of these metals (600 kg Cd/ha; 1000 kg Ni/ha) were applied. Soil temperature appeared to influence crop uptake of these metals. Application of metal-rich sludge reduced iron and manganese availability partly due to changes in soil pH value. (Juste’s paper) 3. In pot experiments with sewage sludge enhanced crop uptake of manganese, iron, zinc and cadmium was observed following increased soil temperature caused by microbial degradation of sludge organic matter. In these trials the heavy metal content of plants varied widely according to the species group. (Vigerust’s paper). 4. Inadvertent soil ingestion from pasture land by grazing animals can significantly affect their trace element status especially as regards copper and selenium for example. Ingestion of sludge following surface dressing of grassland could be an important pathway for the transfer of lipophilic organic contaminants into milk. (Fleming’s paper). 5. Rates of application of sewage sludge and animal slurry in Swiss agriculture are limited by several constraints including climatic conditions, topography and accessibility of suitable land, time for application (long winters), phosphate levels in the soil and concerns about runoff and eutrophication of surface water. (Fürrer’s paper). 6. Excessive application of sewage sludge or animal slurry to shallow surface soil can cause nitrate pollution of groundwater. There tends to be significant leaching of nitrate following applications of sludge to forest soils. As regards organic contaminants, surface runoff is likely to be more of a problem than leaching. (Ahtiainen’s paper). 7. Assessment of the value of sludge as a source of available nitrogen for crops is difficult to estimate because of the problem of predicting the quantity of nitrogen likely to be mineralised from sludge organic matter. Chemical methods were inadequate for this purpose. In Switzerland the content of plant-available nitrogen in sludge was taken to be 90% of the ammoniacal N plus 25% of the N in the

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sludge organic matter. So in general it was found that the nitrogen content of sludge was approximately 40–50% as available for crop uptake as the nitrogen in inorganic fertiliser. (Rudaz’s paper). 8. Moderate applications of sludge to agricultural soils have little effect on soil microbiological processes and can progressively improve the fertility of poor, sandy soils. Microbial biomass in soil can be increased by applications of composted sludge. Mycorrhizal activity may be depressed by sludge applications, probably as a result of increased phosphate availability. Excessive applications of sludge to soils of low cation exchange capacity can adversely affect mineralisation processes and ammonification and nitrogen fixation capability. Deleterious effects were observed in some circumstances when the soil cadmium content was increased by just 0.4 ppm. (Coppola’s paper). 9. Long-term effects of sludge on soil microbial activity can be assessed by applying sensitive techniques to soils which had received heavy applications of sludge historically but which had received no further sludge in the recent past before testing. In such soils it seemed that soil microbial biomass, as measured by chloroform fumigation, and nitrogen fixation capability as measured by acetylene reduction, tended to be reduced compared with soils which had received farmyard manure or inorganic fertiliser instead of sludge. These effects could be due to the residual effects of metals in the sludge. (McGrath’s paper). 10. Whilst the desirability of recycling sludge to agricultural land was usually accepted, it was essential to observe good management practices, particularly as regards metal limits, and to avoid excessive applications of plant nutrients especially to land already receiving animal slurries. This latter problem was acute in the Netherlands, Belgium, Switzerland and parts of FRG where there was already insufficient land to meet requirements for the disposal of animal slurry. In these circumstances metal limits for sludges and soils receiving them should be kept as low as possible to avoid potential toxicity problems in the long-term. Marketing policy should be developed to improve public acceptability of sludge spreading on land. (Scheltinga and Candinas’s paper). 11. Although utilisation of sludge on agricultural land was likely to continue as a major disposal route, it was important to be aware of the environmental and economic aspects of alternative outlets. The main alternative outlets were disposal to landfill (likely to remain stable), use for land reclamation (could increase), incineration (problem of atmospheric pollution), composting and dumping at sea (important for coastal towns in the UK and USA). Technology was available for a variety of novel uses for sludge viz. earthworm conversion (castings for compost), animal feed supplement, extraction of lipids, protein extraction, vitamin B12 extraction, metal recovery, fixation and encapsulation for safer landfilling, overseas shipment, eg for reclamation of infertile soils, brick manufacture, and thermal conversion to liquid and solid fuels. More knowledge was needed about all these outlets to assist in planning the best practicable environmental option for sludge disposal according to particular circumstances. (Webber, DuVoort Van Engers and Berglund’s paper).

LIST OF PARTICIPANTS

AHTIANEN, M. Pohjois—Karjala Water District Office P.O. Box 69 SF—80101—JOENSVU 10 FINLAND BÖHRINGER, J. Motor Columbus Parkstr. 27 D—5400 BADEN GERMANY BONJOUR, R.A. Forschungsanstalt für Agriculturchemie und Umwelthygiene Schwarzenburgstr. 155 CH—3097—LIEBEFELD (BE) SWITZERLAND BOVAY, E. Forschungsanstalt für Agriculturchemie und Umwelthygiene Schwarzenburgstr. 155 CH—3097 LIEBEFELD (BE) SWITZERLAND CANDINAS, A. Forschungsanstalt für Agriculturchemie und Umwelthygiene Schwarzenburgstr. 155 CH—3097 LIEBEFELD (BE) SWITZERLAND COPPOLA, S. Universita degli studi di Napoli Istituto di Microbiologia Agraria e Stazione di Microbiologia Industriale I—80055 PORTICI

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ITALY DAVIS, R.D. Water Research Centre Medmenham Laboratory Henley Road Medmenham P.O. Box 16 UK—SL7 2HD—MARLOW, BUCKS UNITED KINGDOM DESAULES, A. Forschungsanstalt für Agriculturchemie und Umwelthygiene Schwarzenburgstr. 155 CH—3097—LIEBEFELD (BE) SWITZERLAND DETTWILER, J. Federal Office for Environmental Protection Hattwylstr. 4 CH—3003 BERNE SWITZERLAND DUBOIS, J.P. Ecole Polytechnique Fédérale de Lausanne I6R—Pedologie CH—1015 LAUSANNE SWITZERLAND DUVOORT-VAN-ENGERS, L.E. RIVM—LAE Antonie van Leeuwenhoeklaan 9 Postbus 1 NL—3720 BA Bilthoven THE NETHERLANDS FLEMING, G.A. The Agricultural Institute Johnstown Castle Research Centre IRL—WEXFORD IRELAND FROSSARD, R. Forschungsanstalt für Agriculturchemie und Umwelthygiene Schwarzenburgstr. 155 CH—3097—LIEBEFELD (BE) SWITZERLAND FURRER, O.J. Forschungsanstalt für Agriculturchemie und Umwelthygiene

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Schwarzenburgstr. 155 CH—3097—LIEBEFELD (BE) SWITZERLAND GASSES, U. ILW, LFO E24 ETH—Zürich CH—8092 ZURICH SWITZERLAND GOMEZ,A. I.N.R.A. Station d’Agronomie Centre de Recherches de Bordeaux Domaine de la Grande Ferrade F—33140—PONT DE LA MAYE FRANCE GUNS, M. Inst. voor Scheikundig Onderzoek 5 Museumlaan B—1980—TERVUREN BELGIUM GUPTA, S. Forschungsanstalt für Agriculturchemie und Umwelthygiene Schwarzenburgstr. 155 CH—3097—LIEBEFELD (BE) SWITZERLAND HAENI, H. Forschungsanstalt für Agriculturchemie und Umwelthygiene Schwarzenburgstr. 155 CH—3097—LIEBEFELD (BE) SWITZERLAND JUCHTER, S. ETH—Zürich Bodenkunde, ILW CH—8092 ZURICH SWITZERLAND JUSTE, C. I.N.R.A. Station d’Agronomie Centre de Recherches de Bordeaux Domaine de la Grande Ferrade F—33140—PONT DE LA MAYE FRANCE KUNTZE, R. Niedersächsisches Landesamt

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für Bodenforschung Bodentechnologisches Institut Friedrich-Missler-Str. 46–50 D—2800—BREMEN GERMANY L’ HERMITE, P. Commission of the European Communities Environment Protection Research Programme Rue de la Loi 200 B-1049 Brussels LESCHBER, R. Institut fuer Wasser-, Boden und LuftHygiene Corrensplatz 1 1000—BERLIN 33 GERMANY McGRATH, S. Department of Soils and Plant Nutrition Rothamsted Experimental Station AL5—HARPENDEN, HERTS UNITED KINGDOM MATHYS, E. Institut für Pflanzenbau ETH Versuchsstation CH—8307—ESCHIKON-LINDAU SWITZERLAND G.C.PORTEOUS Department of the Environment Room B4/54 Romney House 43 Marsham Street UK—LONDON SW1P 3PY UNITED KINGDOM RICHNER, G. Institut für Pflanzenbau ETH Versuchsstation CH—8307—ESCHIKON-LINDAU SWITZERLAND RUDAZ, A. Forschungsanstalt für Agriculturchemie und Umwelthygiene Schwarzenburgstr. 155 3097—LIEBEFELD (BE) SWITZERLAND SCHELTINGA, H.M.J.

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Staatstoezicht op de Volksgezondheid Pels Rijckenstraat 1 P.O. Box 9013 6800 DR—ARNHEM THE NETHERLANDS SCHMITT, H.W. ETH—Zürich Bodenkunde, ILW CH—8092 ZURICH SWITZERLAND SIEGENTHALLER, A. Forschungsanstalt für Agriculturchemie und Umwelthygiene Schwarzenburgstr. 155 CH—3097—LIEBEFELD (BE) SWITZERLAND STADELMANN F.X. Forschungsanstalt für Agriculturchemie und Umwelthygiene Schwarzenburgstr. 155 CH—3097 LIEBEFELD (BE) SWITZERLAND TIMMERMAN, F. Institut für Planzenernährung & Bodenkunde (FAL) Bundesallee 50 D—3300—BRAUNSCHWEIG GERMANY TJELL, J.Ch. Technical University of Denmark Dept. of Environmental Engineering Building 115 DK—2800—LYNGBY DENMARK VIGERUST, E. Agricultural University Boks 28 N—1432—AS-NLH NORWAY WEBBER, M.D. Waste Water Technology Center Environmental Protection Service P.O. Box 5050 867, Lakeshore Road

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L7R 4A6—BURLINGTON ONTARIO CANADA WILLIAMS, J.H. Ministry of Agriculture Fisheries and Food Woodthorne Wolverhampton UK—West Midlands WV6 8TQ UNITED KINGDOM

INDEX OF AUTHORS

AHTIAINEN, M., 51 BERGLUND, S., 103 BROOKES, P.C., 80 CANDINAS, T., 90 COPPOLA, S., 72 DUVOORT-VAN ENGERS, L., 103 FLEMING, G.A., 43 GUPTA, S.K., 64 JUSTE, C., 13 KUNTZE, H., 2 MCGRATH, S.P., 80 RUDAZ, A., 64 SCHELTINGA, H.M.J., 90 SELMER-OLSEN, A.R., 26 SOLDÂ, P., 13 VIGERUST, E., 26 WEBBER, M.D., 103

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