Reviews of Environmental Contamination and Toxicology, Vol. 177

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Reviews of Environmental Contamination and Toxicology VOLUME 177

Reviews of Environmental Contamination and Toxicology Continuation of Residue Reviews

Editor

George W. Ware Editorial Board Lilia A. Albert, Xalapa, Veracruz, Mexico F. Bro-Rasmussen, Lyngby, Denmark ⴢ D.G. Crosby, Davis, California, USA Pim de Voogt, Amsterdam, The Netherlands ⴢ H. Frehse, Leverkusen-Bayerwerk, Germany O. Hutzinger, Bayreuth, Germany ⴢ Foster L. Mayer, Gulf Breeze, Florida, USA D.P. Morgan, Cedar Rapids, Iowa, USA ⴢ Douglas L. Park, Washington DC, USA Raymond S.H. Yang, Fort Collins, Colorado, USA Founding Editor Francis A. Gunther

VOLUME 177

Coordinating Board of Editors DR. GEORGE W. WARE, Editor Reviews of Environmental Contamination and Toxicology 5794 E. Camino del Celador Tucson, Arizona 85750, USA (520) 299-3735 (phone and FAX) DR. HERBERT N. NIGG, Editor Bulletin of Environmental Contamination and Toxicology University of Florida 700 Experimental Station Road Lake Alfred, Florida 33850, USA (941) 956-1151; FAX (941) 956-4631 DR. DANIEL R. DOERGE, Editor Archives of Environmental Contamination and Toxicology 6022 Southwind Drive N. Little Rock, Arkansas, 72118, USA (501) 791-3555; FAX (501) 791-2499

Springer-Verlag New York: 175 Fifth Avenue, New York, NY 10010, USA Heidelberg: Postfach 10 52 80, 69042 Heidelberg, Germany Library of Congress Catalog Card Number 62-18595. Printed in the United States of America. ISSN 0179-5953 Printed on acid-free paper.  2003 Springer-Verlag New York, Inc. All rights reserved. This work may not be translated or copied in whole or in part without the written permission of the publisher (Springer-Verlag New York, Inc., 175 Fifth Avenue, New York, NY 10010, USA), except for brief excerpts in connection with reviews or scholarly analysis. Use in connection with any form of information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed is forbidden. The use in this publication of trade names, trademarks, service marks, and similar terms, even if they are not identified as such, is not to be taken as an expression of opinion as to whether or not they are subject to proprietary rights. Printed in the United States of America. ISBN 0-387-00214-6

SPIN 10903537

www.springer-ny.com Springer-Verlag New York Berlin Heidelberg A member of BertelsmannSpringer Science+Business Media GmbH

Foreword

International concern in scientific, industrial, and governmental communities over traces of xenobiotics in foods and in both abiotic and biotic environments has justified the present triumvirate of specialized publications in this field: comprehensive reviews, rapidly published research papers and progress reports, and archival documentations. These three international publications are integrated and scheduled to provide the coherency essential for nonduplicative and current progress in a field as dynamic and complex as environmental contamination and toxicology. This series is reserved exclusively for the diversified literature on “toxic” chemicals in our food, our feeds, our homes, recreational and working surroundings, our domestic animals, our wildlife and ourselves. Tremendous efforts worldwide have been mobilized to evaluate the nature, presence, magnitude, fate, and toxicology of the chemicals loosed upon the earth. Among the sequelae of this broad new emphasis is an undeniable need for an articulated set of authoritative publications, where one can find the latest important world literature produced by these emerging areas of science together with documentation of pertinent ancillary legislation. Research directors and legislative or administrative advisers do not have the time to scan the escalating number of technical publications that may contain articles important to current responsibility. Rather, these individuals need the background provided by detailed reviews and the assurance that the latest information is made available to them, all with minimal literature searching. Similarly, the scientist assigned or attracted to a new problem is required to glean all literature pertinent to the task, to publish new developments or important new experimental details quickly, to inform others of findings that might alter their own efforts, and eventually to publish all his/her supporting data and conclusions for archival purposes. In the fields of environmental contamination and toxicology, the sum of these concerns and responsibilities is decisively addressed by the uniform, encompassing, and timely publication format of the Springer-Verlag (Heidelberg and New York) triumvirate: Reviews of Environmental Contamination and Toxicology [Vol. 1 through 97 (1962–1986) as Residue Reviews] for detailed review articles concerned with any aspects of chemical contaminants, including pesticides, in the total environment with toxicological considerations and consequences. Bulletin of Environmental Contamination and Toxicology (Vol. 1 in 1966) for rapid publication of short reports of significant advances and discoveries in the fields of air, soil, water, and food contamination and pollution as well as v

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Foreword

methodology and other disciplines concerned with the introduction, presence, and effects of toxicants in the total environment. Archives of Environmental Contamination and Toxicology (Vol.1 in 1973) for important complete articles emphasizing and describing original experimental or theoretical research work pertaining to the scientific aspects of chemical contaminants in the environment. Manuscripts for Reviews and the Archives are in identical formats and are peer reviewed by scientists in the field for adequacy and value; manuscripts for the Bulletin are also reviewed, but are published by photo-offset from cameraready copy to provide the latest results with minimum delay. The individual editors of these three publications comprise the joint Coordinating Board of Editors with referral within the Board of manuscripts submitted to one publication but deemed by major emphasis or length more suitable for one of the others. Coordinating Board of Editors

Preface

Thanks to our news media, today’s lay person may be familiar with such environmental topics as ozone depletion, global warming, greenhouse effect, nuclear and toxic waste disposal, massive marine oil spills, acid rain resulting from atmospheric SO2 and NOx, contamination of the marine commons, deforestation, radioactive leaks from nuclear power generators, free chlorine and CFC (chlorofluorocarbon) effects on the ozone layer, mad cow disease, pesticide residues in foods, green chemistry or green technology, volatile organic compounds (VOCs), hormone- or endocrine-disrupting chemicals, declining sperm counts, and immune system suppression by pesticides, just to cite a few. Some of the more current, and perhaps less familiar, additions include xenobiotic transport, solute transport, Tiers 1 and 2, USEPA to cabinet status, and zerodischarge. These are only the most prevalent topics of national interest. In more localized settings, residents are faced with leaking underground fuel tanks, movement of nitrates and industrial solvents into groundwater, air pollution and “stay-indoors” alerts in our major cities, radon seepage into homes, poor indoor air quality, chemical spills from overturned railroad tank cars, suspected health effects from living near high-voltage transmission lines, and food contamination by “flesh-eating” bacteria and other fungal or bacterial toxins. It should then come as no surprise that the ‘90s generation is the first of mankind to have become afflicted with chemophobia, the pervasive and acute fear of chemicals. There is abundant evidence, however, that virtually all organic chemicals are degraded or dissipated in our not-so-fragile environment, despite efforts by environmental ethicists and the media to persuade us otherwise. However, for most scientists involved in environmental contaminant reduction, there is indeed room for improvement in all spheres. Environmentalism is the newest global political force, resulting in the emergence of multi-national consortia to control pollution and the evolution of the environmental ethic. Will the new politics of the 21st century be a consortium of technologists and environmentalists or a progressive confrontation? These matters are of genuine concern to governmental agencies and legislative bodies around the world, for many serious chemical incidents have resulted from accidents and improper use. For those who make the decisions about how our planet is managed, there is an ongoing need for continual surveillance and intelligent controls to avoid endangering the environment, the public health, and wildlife. Ensuring safety-

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Preface

in-use of the many chemicals involved in our highly industrialized culture is a dynamic challenge, for the old, established materials are continually being displaced by newly developed molecules more acceptable to federal and state regulatory agencies, public health officials, and environmentalists. Adequate safety-in-use evaluations of all chemicals persistent in our air, foodstuffs, and drinking water are not simple matters, and they incorporate the judgments of many individuals highly trained in a variety of complex biological, chemical, food technological, medical, pharmacological, and toxicological disciplines. Reviews of Environmental Contamination and Toxicology continues to serve as an integrating factor both in focusing attention on those matters requiring further study and in collating for variously trained readers current knowledge in specific important areas involved with chemical contaminants in the total environment. Previous volumes of Reviews illustrate these objectives. Because manuscripts are published in the order in which they are received in final form, it may seem that some important aspects of analytical chemistry, bioaccumulation, biochemistry, human and animal medicine, legislation, pharmacology, physiology, regulation, and toxicology have been neglected at times. However, these apparent omissions are recognized, and pertinent manuscripts are in preparation. The field is so very large and the interests in it are so varied that the Editor and the Editorial Board earnestly solicit authors and suggestions of underrepresented topics to make this international book series yet more useful and worthwhile. Reviews of Environmental Contamination and Toxicology attempts to provide concise, critical reviews of timely advances, philosophy, and significant areas of accomplished or needed endeavor in the total field of xenobiotics in any segment of the environment, as well as toxicological implications. These reviews can be either general or specific, but properly they may lie in the domains of analytical chemistry and its methodology, biochemistry, human and animal medicine, legislation, pharmacology, physiology, regulation, and toxicology. Certain affairs in food technology concerned specifically with pesticide and other food-additive problems are also appropriate subjects. Justification for the preparation of any review for this book series is that it deals with some aspect of the many real problems arising from the presence of any foreign chemical in our surroundings. Thus, manuscripts may encompass case studies from any country. Added plant or animal pest-control chemicals or their metabolites that may persist into food and animal feeds are within this scope. Food additives (substances deliberately added to foods for flavor, odor, appearance, and preservation, as well as those inadvertently added during manufacture, packing, distribution, and storage) are also considered suitable review material. Additionally, chemical contamination in any manner of air, water, soil, or plant or animal life is within these objectives and their purview.

Preface

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Normally, manuscripts are contributed by invitation, but suggested topics are welcome. Preliminary communication with the Editor is recommended before volunteered review manuscripts are submitted. Tucson, Arizona

G.W.W.

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Table of Contents

Foreword ....................................................................................................... Preface ..........................................................................................................

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Role of Phosphorus in (Im)mobilization and Bioavailability of Heavy Metals in the Soil–Plant System ................................................................. NANTHI S. BOLAN, DOMY C. ADRIANO, AND RAVI NAIDU

1

Environmental Fate of Methyl Bromide as a Soil Fumigant ...................... SCOTT R. YATES, JAY GAN, AND SHARON K. PAPIERNIK

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Disposal and Degradation of Pesticide Waste ............................................. 123 ALLAN S. FELSOT, KENNETH D. RACKE, AND DENIS J. HAMILTON Index ............................................................................................................. 201

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 Springer-Verlag 2003

Rev Environ Contam Toxicol 177:1–44

Role of Phosphorus in (Im)mobilization and Bioavailability of Heavy Metals in the Soil–Plant System Nanthi S. Bolan, Domy C. Adriano, and Ravi Naidu Contents I. Introduction .......................................................................................................... II. Sources of Heavy Metals in Soil Environment .................................................. III. Reactions of Metals in Soils ............................................................................... A. Adsorption ...................................................................................................... B. Complexation .................................................................................................. C. Precipitation .................................................................................................... D. Solid-Phase Speciation ................................................................................... IV. Reactions of Phosphate Compounds in Soils ..................................................... A. Water-Soluble Compounds ............................................................................ B. Water-Insoluble Compounds .......................................................................... V. Mechanisms for (Im)mobilization of Heavy Metals by Phosphate Compounds ................................................................................... A. Phosphate Compounds as a Metal Source ..................................................... B. Physiological Phosphorus-Metal Interactions in Plants ................................. C. Adsorption/Desorption of Metals ................................................................... D. Precipitation of Metals ................................................................................... E. Rhizosphere Modifications ............................................................................. Summary .................................................................................................................... Acknowledgments ...................................................................................................... References ..................................................................................................................

1 4 5 6 7 8 8 10 10 10 12 12 20 21 24 29 34 35 35

I. Introduction The term heavy metal in general includes elements (both metals and metalloids) with an atomic density greater than 6 g cm−3 [with the exception of arsenic (As), boron (B), and selenium (Se)]. This group includes both biologically essential

Communicated by George W. Ware. N.S. Bolan ( ) Soil & Earth Science Group, Massey University, Palmerston North, New Zealand D.C. Adriano University of Georgia, Savannah River Ecology Laboratory, Drawer E, Aiken, SC 29802, USA R. Naidu CSIRO Land and Water, Adelaide, South Australia 5064

1

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N.S. Bolan, D.C. Adriano, and R. Naidu

[e.g., cobalt (Co), copper (Cu), chromium (Cr), manganese (Mn), and zinc (Zn)] and nonessential [e.g., cadmium (Cd), lead (Pb), and mercury (Hg)] elements. The essential elements (for plant, animal, or human nutrition) are required in low concentrations and hence are known as trace elements or micronutrients. The nonessential metals are phytotoxic and/or zootoxic and are widely known as toxic elements (Adriano 2001). Both groups are toxic to plants, animals, and humans at exorbitant concentrations (Adriano 2001; Alloway 1990). Consequently, metals are extensively researched in the life, agricultural, and environmental sciences. Among the metals, Cd, Cr, Cu, Hg, Pb, and Zn have been intensively studied. Other metals, such as silver (Ag) and tin (Sn), are important especially in the mining environment and around industries that process ores containing these metals (Table 1). Health authorities in many parts of the world are becoming increasingly concerned about the effects of heavy metals on environmental and human health and the potential implications to international trade. For example, Cd accumulating in the offal (mainly kidney and liver) of grazing animals not only makes it unsuitable for human consumption but also imperils the access of offal products to overseas markets (Bramley 1990; Roberts et al. 1994). Similarly, bioaccumulation of Cd in potato, wheat, and rice crops has serious implications for local and international commodity marketing. For these reasons, there is global urgency to ensure that the heavy metal content of foodstuffs produced complies with regulatory standards and compares well with those from other countries. Effective action in the long term will depend on gaining an understanding of the causes of heavy metal accumulation and a proper appreciation of these issues for public health. Because of the ever-increasing production of livestock and poultry products for human consumption, more and more organic wastes from these industries have to be handled so as to abide by environmental regulations, including safe disposal onto land. Large quantities of organic amendments, such as poultry manure compost and biosolids, are used as a source of nutrients and also as a soil conditioner to improve the physical properties and fertility of soils. With increasing demand for safe disposal of wastes generated from agricultural and industrial activities, soil is not only considered as a source of nutrients for plant growth but is also used as a sink for the removal of contaminants from these waste materials (Power and Dick 2000). As land treatment becomes one of the important waste management practices, soil is increasingly being seen as a major source of metals reaching the food chain, mainly through plant uptake and animal transfer. Such waste disposal has led to significant buildup of a wide range of metals, such as Cu, Cr, Pb, Cd, Hg, and Zn, and metalloids, such as As, Cr, and Se. Entry of soilborne metals into the food chain depends on the amount and source of metal input, the properties of the soil (especially soil pH and organic matter), the rate of uptake by plants, and the extent of absorption by grazing animals. Nriagu (1988) stated that “this very profound experiment, in which one billion (109) human guinea pigs are being exposed to undue insults of toxic metals, has yet to receive scientific attention that it clearly deserves.”

Table 1. Sources of heavy metals in soils and their expected ionic species in soil solution.

Metal

Density (g cm−3) 5.73

Cadmium (Cd)

8.64

Chromium (Cr)

7.81

Copper (Cu)

8.96

Lead (Pb) Manganese (Mn)

11.35 7.21

Mercury (Hg)

13.55

Molybdenum (Mo) Nickel (Ni)

10.2 8.90

Zinc (Zn)

7.13

As(III): As(OH)3, AsO3− 3 ; As(V): H2As−4, HAsO2− 4 Cd2+, CdOH+, CdCl−, CdHCO+3, CdSO04 Cr(III): Cr3+, CrO−2, CrOH2+, 2− Cr(OH)−4; Cr(VI): Cr2O2− 7 , CrO4 2+ 2+ Cu (II), Cu (III)

Pb2+, PbOH+, PbCl−, PbHCO+3, PbSO04 Mn2+, MnOH+, MnCl−, MnCO03, MnHCO+3, MnSO04 Hg2+, HgOH+, HgCl02, CH3Hg+, Hg(OH)02 − 0 MoO2− 4 , HMoO4, H2MoO4 2+ + 0 Ni , NiSO4, NiHCO3 NiCO03 Zn2+, ZnSO04, ZnCl+, ZnHCO+3, ZnCO03

Contaminant sources Timber treatment, paints, pesticides, geothermal Electroplating, batteries, fertilizers Timber treatment, leather tanning, pesticides, dyes Fungicides, electrical, paints, pigments, timber treatment, fertilizers, mine tailings Batteries, metal products, preservatives, petrol additives Fertilizer

Toxicitya Toxic to plants, humans, and animals Toxic to plants, humans, and animals Cr (VI) toxic to plants, humans, and animalsb Toxic to plants, humans and animals Toxic to plants, humans, and animals Toxic to plants

Instruments, fumigants, geothermal

Toxic to humans and animals

Fertilizer Alloys, batteries, mine tailings

Toxic to animals Toxic to plants, humans, and animals Toxic to plants

Galvanizing, dyes, paints, timber treatment, fertilizers, mine tailings

Phosphorus (Im)mobilization of Heavy Metals

Arsenic (As)

Ionic species in soil solution

aMost

likely to be observed at elevated concentrations in soils and water. is very mobile and highly toxic; however, Cr(III) is essential in animal and human nutrition and is generally immobile in the environment. Source: Adriano (2001).

bCr(VI)

3

4


With greater public awareness of the implications of contaminated soils on human and animal health, there has been increasing interest within the scientific community in the examination of the transformation and the fate of metals in soils and the development of technologies to remediate contaminated sites. Unlike organic contaminants, most metals do not undergo microbial or chemical degradation, and the total concentration of these metals in soils persists for a long time after their introduction (Adriano 2001). The mobilization of metals in soils for plant uptake and leaching to groundwater can, however, be minimized by reducing the bioavailability of metals through chemical and biological immobilization. Recently, there has been increasing interest in the immobilization of metals using a range of inorganic compounds, such as lime, phosphate (P) compounds (e.g., apatite rocks), and alkaline waste materials, and organic compounds, such as exceptional quality biosolids (Basta et al. 2001; Knox et al. 2000). Although traditionally investigative efforts have examined the potential value of hydroxyapatite in the immobilization of metals, there is growing interest in assessing the effect of other P compounds on metal transformation in soils. Because most of these P compounds are applied to soils as a nutrient source, some of them are also referred to as P fertilizers in this review. In many countries, increasing amounts of phosphate rocks (PRs) are added directly to soils as a source of phosphorus (P) (Bolan et al. 1990; Rajan et al. 1996). Regular application of P fertilizers has been identified as the main source of heavy metal contamination of soils in these countries (McLaughlin et al. 1996; Roberts et al. 1994). Some of these P fertilizers, which act as a source of heavy metal contamination of agricultural soils, have also been found to act as a sink for the immobilization of these metals (McGowen et al. 2001). Phosphate amendment has often been proposed as a practical remediation option for sites with Pb-contaminated soils (Hettiarachchi et al. 2000). This method is considered to be more economical and less disruptive than the conventional remediation option of soil removal. Application of P compounds to soils, however, can have no effect, induce mobilization, or enhance immobilization of metals, the effect being dependent on the nature of P compound, soil type, and metal species (Knox et al. 2000). Although a number of studies have examined the potential value of P compounds in the immobilization of metals in contaminated soils, there has been no comprehensive review on the mechanisms involved in the P-induced (im)mobilization of these metals. Following a brief overview of the reactions of metals and common P compounds that are used as fertilizer in soils, the review focuses on the mechanisms for the (im)mobilization of metals by P compounds. The practical implications of P compounds on the transformation of metals are discussed in relation to sequestration and phytoavailability of metals in soils.

II. Sources of Heavy Metals in Soil Environment Heavy metals reach the soil environment through both pedogenic (or geogenic) and anthropogenic processes. Most metals occur naturally in soil parent materials, chiefly in forms that are not available for plant uptake. Because of their low


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bioavailability, the metals present in the parent materials are often not available for plant uptake and cause minimum impact to soil organisms. Often the concentrations of metals released into the soil system by the natural pedogenic (or weathering) processes are largely related to the origin and nature of the parent material. Apart from Se (Dhillon and Dhillon 1990) and As (Chakraborty and Saha 1987; Naidu and Skinner 1999), other elements (e.g., Cr, Ni, Pb) derived via geogenic processes have limited impact on soil. Unlike pedogenic inputs, metals added through anthropogenic activities typically have high bioavailability (Naidu et al. 1996a). Anthropogenic activities, primarily associated with industrial processes, manufacturing, and the disposal of domestic and industrial waste materials, are the major source of metal enrichment in soils (Adriano 2001) (see Table 1). Atmospheric pollution from Pb-based petrol is a major issue in many developing countries where there is no constraint on the usage of leaded gasoline. Phosphate fertilizers are considered to be the major source of heavy metal input, especially Cd, in pasture soils in Australia and New Zealand (refer to Section V).

III. Reactions of Metals in Soils

cip

sso

tio

ion

Soluble Complexes

ion

Biomass

ion

zat

li era

n

Mi

Layer Silicate Clays

ch Ex

Soil Solution

pt sor

Ab

e

ang

Ion

n

lut

Soluble “Free” Ions

Groundwater

Di

ita

Leaching to

Pre

Plant

Precipitates

Uptake

Metal ions can be retained in the soil largely by (ad)sorption, precipitation, and complexation reactions (Fig. 1). Sorption is defined as the accumulation of matter at the interface between the aqueous solution phase and a solid adsorbent (Sposito 1984). This process can include ion exchange, formation of surface complexes, precipitation, and diffusion into the solid. In many situations, adsorption is believed to be the precursor for subsequent precipitation, and it is difficult to define the boundary separating adsorption and precipitation pro-

Ad

sor

De

sor

pti

on

pti

on

Humus, Oxides, and Allophane

Fig. 1. Reactions of metals in soils. [Source: Adriano (2002)]

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cesses (Corey 1981). The lower the metal solution concentration and the more sites available for adsorption, the more likely that adsorption/desorption processes will determine the soil solution concentration (Tiller 1988). Following input into the soil environment, metals interact with the soil mineral and organic phases. However, the fate of metals in the soil environment is dependent on both soil properties and environmental factors. A. Adsorption Charged solute species (ions) are attracted to the charged soil surface by electrostatic attraction and through the formation of specific bonds (Barrow 1985). Retention of charged solutes by charged surfaces is broadly grouped into specific and nonspecific retention (Bolan et al. 1999a; Sposito 1984). In general terms, nonspecific adsorption is a process in which the charge on the ions balances the charge on the soil particles through electrostatic attraction; in contrast, specific adsorption involves chemical bond formation between the ions and the sorption sites on the soil surface (Sposito 1984). If the nonspecific adsorption process solely controls metal adsorption, then the adsorption capacity of the soil is dictated by its cation-exchange capacity (CEC). However, in many soils the amount of metal sorbed exceeds the CEC of the soils (Bolan et al. 1999a). This observation infers that, in addition to nonspecific adsorption, other processes, such as specific adsorption, precipitation, and complex formation, also contribute to metal retention in soils. Both soil properties and soil solution composition determine the dynamic equilibrium between metals in solution and the soil solid phase. The concentration of metals in soil solution is influenced by the pH (Adriano 2001) and the nature of both organic and inorganic ligands (Bolan et al. 1999b; Harter and Naidu 1995; Naidu et al. 1994; Shuman 1986). The effect of pH values above 6 in lowering free metal ion activities in soils has been attributed to the increase in pH-dependent surface charge on oxides of Fe, Al, and Mn, chelation by organic matter, or precipitation of metal hydroxides (Adriano 2001). The effect of pH on the activity of metals in solution in naturally acidic soils is found to decrease with increasing pH. The gradual decrease in heavy metal activity with increasing pH, especially in variable charge soils, is attributed to increasing CEC (Shuman 1986). In general, both the CEC and the total amount of metal removed from soil solution increase with increasing soil pH (Adriano 2001). Three reasons have been given for the effect of inorganic and organic anions on the adsorption of metals (Naidu et al. 1994). First, anions form complexes with metals, thereby reducing their adsorption onto soil particles. Second, the specific adsorption of ligand anions is likely to increase the negative charge on soil particles, thereby increasing the adsorption of heavy metal cations. And third, specifically sorbed anions, such as phosphate (H2PO−4), strongly compete with heavy metal anions, such as arsenate and selenate, resulting in their desorption. Phosphate-induced metal adsorption/desorption reactions in relation to (im) mobilization of heavy metals are discussed in Section V.


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B. Complexation Metals form both inorganic and organic complexes with a range of solutes in soils. A number of studies have examined the effect of inorganic anionic complex formation on the adsorption of Cd2+ by soils (Boekhold et al. 1993; Bolan et al. 1999b; Naidu et al. 1994). Most of these studies have indicated that chloride has often been found to form a complex with Cd2+ as CdCl+, thereby decreasing the adsorption of Cd2+ onto soil particles (Naidu et al. 1994, 1996b). However, when the activity of Cd2+ was corrected for complex formation, the adsorption curves for Cd2+ in nitrate (NO−3) and chloride (Cl−) media coincided. O’Connor et al. (1984) showed that while the presence of Cl− ions decreased 2+ adsorption of Cd2+, sulfate (SO2− 4 ) ions increased Cd adsorption relative to com− parable concentrations of chlorate (ClO4) in three calcareous soils. Cadmiumchloro complexation was identified as the active process reducing Cd2+ retention. 2+ The increased retention in the presence of SO2− 4 was attributed to the low Ca 2+ ion activity available for competition with Cd as a result of the formation of the soluble CaSO04 complex. In contrast to inorganic ligand ions, Haas and Horowitz (1986) found that Cd2+ adsorption by kaolinite, a variable charge mineral, was enhanced by the presence of organic matter, which was attributed to the formation of an adsorbed organic layer on the clay surface. As might be expected, the organic component of soil constituents has a high affinity for metal cations because of the presence of ligands or groups that can form chelates with metals (Harter and Naidu 1995). With increasing pH, the carboxyl, phenolic, alcoholic, and carbonyl functional groups in soil organic matter dissociate, thereby increasing the affinity of ligand ions for metal cations. The general order of affinity for metal cations complexed by organic matter is as follows (Adriano 2001): Cu2+ > Cd2+ > Fe2+ > Pb2+ > Ni2+ > Co2+ > Mn2+ > Zn2+ The extent of metal–organic complex formation however, varies with a number of factors including temperature, steric factors (e.g., geometry), and concentration. All these interactions are controlled by solution pH and ionic strength, nature of the metal species, dominant cation, and inorganic and organic ligands present in the soil solution. Metal interactions involving organic ligands were reported for variable charge oxide surfaces where certain organic ligands were found to enhance the adsorption of Cu2+ and Ag+. An alternate mechanism, which appears to be important in temperate soils, involves metal–ligand complexation in solution and subsequent reduction in cation charge, which probably reduces adsorption (Harter and Naidu 1995). The formation of aqueous complexes of Cd with low molecular weight organic acids (LMWOA) from root exudates is expected to dominate the solution chemistry of Cd in rhizosphere. Based on differential pulse anode stripping voltametric and cation-exchange resin extraction data, the dissolved Cd in soil solutions was found to be almost completely complexed with organic matter (Sauve

8


et al. 2000). Krishnamurti et al. (1997a) observed significant solubilization of Cd from neutral to slightly acidic soils with 0.1 to 1 mM concentrations of acetic, succinic, oxalic, and citric acids, suggesting that Cd release is related to the stability constant of the Cd-LMWOA complex. C. Precipitation Precipitation appears to be the predominant process of metal immobilization in 2− high-pH soils in the presence of anions, such as SO2− 4 , carbonate (CO3 ), hydrox− − ide (OH ), and H2PO4, especially when the concentration of heavy metal ion is high (Adriano 2001). Metalloids such as Cr and As that form anionic species at field soil pH have been reported to form precipitates with cations, such as Ca2+ (Avudainayagam et al. 2001). Coprecipitation of metals, especially in the presence of iron (Fe) and aluminum (Al) oxyhydroxides, has also been reported and often such interactions lead to significant changes in the surface chemical properties of the substrate. Precipitation as metal phosphates is considered to be one of the primary mechanisms for the P-induced immobilization of metals, especially in substrates containing high concentration of metals; this is discussed in more detail in Section V. Liming is often found to increase the retention of metals. For example, Bolan and Thiyagarajan (2001) have observed an increase in the retention of Cr(III) with an increase in pH due to liming. The pH of the lime-treated soil ranged from 7.18 to 8.04, which coincides with the effective precipitation range for Cr(III) as Cr(OH)3 (Rai et al. 1987). Thus, the increased retention of Cr(III) in the presence of the lime is likely due to the formation of Cr(OH)3. The increase in pH from liming is also likely to increase the negative charge of these variable charge soils, which may have enhanced Cr(III) adsorption. D. Solid-Phase Speciation Fractionation studies are often used to examine the effect of amendments, such as lime, P compounds, and biosolid, on the immobilization of metals. Irrespective of the nature of interaction between the metals and soil colloidal particles, following adsorption metal ions redistribute among organic and mineral soil constituents. Fractionation studies suggest that the majority of the metals are associated with organic matter, Fe and Al oxides, and silicate clay minerals in soils. Factors affecting the distribution of metal among different forms include pH, ionic strength of the soil solution, the solid and solution components and their relative concentration and affinities for the metal, and time (Shuman 1991). A large number of sequential extraction schemes have been used for soils, generally attempting to identify metals held in any of the fractions that include soluble, adsorbed/exchangeable, carbonate-bound, organic-bound, amorphous ferromanganese hydrous oxide-bound, crystalline ferromanganese hydrous oxide-bound, and residual or lattice mineral-bound. The most commonly used schemes are modifications of Tessier et al. (1979).


9

Metal fractionations using the sequential extraction techniques have primarily been used to identify the fate of metals applied in sewage sludges and in soils contaminated with smelter and mine drainage wastes (Dudka and Chlopecka 1990; Sposito et al. 1982). These studies suggest that treating the soils with sludges or wastes shifts the solid phases of the metals away from immobile fractions to forms that are potentially more mobile, labile, and bioavailable. For example, Dudka and Chlopecka (1990) found that with sewage sludge application the residual forms of Cd2+, Cu2+, and Zn2+ in soil decreased from 34%–43% to 6%–34%, with a corresponding increase in the readily bioavailable forms. The treatment of metal-contaminated soils with P compounds tends to cause the opposite effect in relation to solid-phase metal fractions (Basta et al. 2001; Seaman et al. 2001). Elaborate sequential extraction schemes have frequently been used to identify the distribution of different species of the metal among the various fractions (Krishnamurti 2000; Ross 1994). However, very few attempts have been made to identify the particular species of the metal that contributes to bioavailability. Using an innovative extraction scheme (Krishnamurti and Naidu 2000; Krishnamurti et al. 1995), which estimates the species associated with metal–organic complexes, the importance of this form in the bioavailability of Cd in native soils has been established. Likewise, the importance of metal–fulvic complexes in the phytoavailability of Cu and Zn has also been documented (Krishnamurti and Naidu 2000). Based on the differential Fourier transform infrared (differential FTIR) spectra of the metal–organic complexes extracted by 0.1 M sodium pyrophosphate extractant, Krishnamurti et al. (1997b) showed that Cd in soils was bonded to the carboxyl and the phenolic groups. A logical approach to minimize plant uptake and subsequent contamination of the food chain is to render the trace metals in the soil immobile. The phytoavailability of the different forms of the solid-phase species generally decreases in this order: soluble > exchangeable/adsorbed > organic-bound > carbonate-bound > ferromanganese hydrous oxide-bound > residual or refractory (i.e., fixed in mineral lattice). Immobilization of metals such as Pb, Zn, and Cd could be achieved by additives, such as zeolites (Chlopecka and Adriano 1997; Gworek 1992; Seaman et al. 2001), apatite (Basta et al. 2001; Ma et al. 1993), Mn oxides (Fu et al. 1991; Hettiarachchi et al. 2000), and clay-hydroxy Al polymers (Mench et al. 1994), which may not produce any detrimental byproduct or alter the physicochemical environment of the soils to affect plant growth. Physiologically based in vitro chemical fractionation schemes are increasingly being used to examine the biovailability of metals (Basta and Gradwohl 2000; Ruby et al. 1996). These schemes include the physiologically based extraction test (PBET), potentially bioavailable sequential extraction (PBASE), and the gastrointestinal (GI) test. These improved tests are capable of predicting the bioavailability of metals for both plant uptake and certain soil organisms.

10


IV. Reactions of Phosphate Compounds in Soils The P-induced (im)mobilization of metals depends on the form of P compounds used and their reactions in soils. Phosphate compounds that are used as fertilizer are broadly grouped into water-soluble (fast-release) and water-insoluble (slowrelease) fertilizers (Bolan et al. 1993). The important water-soluble P fertilizers include: single superphosphate (SSP), triple superphosphate (TSP), monoammonium phosphate (MAP), and diammonium phosphate (DAP). The important water-insoluble P fertilizers include phosphate rocks (PRs) and basic slag. Partially acidulated phosphate rocks (PAPR) and superphosphate and reactive rock mixtures (e.g., Longlife super in New Zealand) contain both water-soluble and water-insoluble P components. Monocalcium phosphate (MCP) and ammonium phosphate (AMP) are the principal P components present in superphosphates (SSP and TSP) and ammonium phosphates (MAP and DAP), respectively. It is important to understand the reactions of these P compounds in soils to predict the effect of these on the (im)mobilization of metals. A. Water-Soluble Compounds When superphosphate fertilizers are added to soils, the dissolution of MCP results in the formation of slowly soluble dicalcium phosphate (DCP) with a release of phosphoric acid close to the fertilizer granules (Eq. 1). Phosphoric acid subsequently dissociates into H2PO−4 and hydrogen ions (protons–H+). The protons reduce the pH around the fertilizer granules to a very low level (pH < 2). When ammonium phosphate fertilizers are added to soil, they dissociate into ammonium (NH+4) and H2PO−4 ions. The subsequent oxidation of NH+4 to NO−3 results in the release of protons (Eq. 2). Ca (H2PO4)2 + H2O → CaHPO4. H2O + H3PO4 + 4

− 3

+

NH + 2O2 → NO + 2H + H2O

(1) (2)

The acidic solution around the fertilizer granules dissolves the Fe and Al compounds in the soil, resulting in the adsorption and precipitation of P. The pH around the ammonium phosphate fertilizer granules, however, is unlikely to be as low as that around superphosphate fertilizers, causing less adsorption of H2PO−4 ions. The concentration of plant-available P in soil solution decreases with time of contact of fertilizer granules in soils, the decrease depending on the amount of Fe and Al compounds in the soil. The retention of P by soils decreases the amount of P available for both plant uptake and leaching to groundwater. The acidity generated can also have important implications to the mobilization of metals (Section V). B. Water-Insoluble Compounds The solubility of common P compounds that are added to soils and formed as reaction products is given in Table 2. When insoluble P fertilizers such as PRs are added to soil, they must be dissolved in soils for the P to become plant

Table 2. Equilibrium dissolution reaction and the solubility of common crystalline phosphate compounds in soil.

Chemical formula

Equilibrium dissolution reaction

Calcium dihydrogen phosphate

Ca(H2PO4)2

Ca(H2PO4)2(s) ↔ Ca2+ + 2H2PO4−

CaHPO4

CaHPO4(s) ↔

Tricalcium phosphate

Ca3(PO4)2

Ca3(PO4)2(s) ↔

Hydroxy apatite

Ca10(PO4)6(OH)2

Carbonate apatite

Log Ksp

Solubility (g/100 g)

−1.14

18

−6.6

0.14

−24.0

0.02

Ca10(PO4)6(OH)2(s) ↔ 10Ca2+ + 6PO43− + 2OH−

−55.9

Insoluble

Ca10(PO4)6CO3

Ca10(PO4)6CO3(s) ↔ 10Ca2+ + 6PO43− + CO32−

−108.3

Insoluble

Fluroapatite

Ca10(PO4)6F2

Ca10(PO4)6F2(s) ↔

−110.2

Insoluble

Variscite

AlPO4ⴢ2H2O

AlPO4(s) ↔

−21.0

Insoluble

Strengite

FePO4ⴢ2H2O

FePO4ⴢ2H2O(s) ↔ Fe3+ + PO43− + 2H2O

−26.0

Insoluble

Vivianite

Fe3(PO4)2ⴢ8H2O

Fe3(PO4)2ⴢ8H2O (s) ↔ Fe2+ + 2PO43− + 8H2O

Calcium monohydrogen phosphate

Ca2+

−

+ HPO4

Ca2+

Al3+

+ 2PO4

3−

10Ca2+

+ PO4

3−

+ 6PO4

3−

+

2F−

+ 2H2O

−3.11

Insoluble


Phosphate compound

Sources: Whitelaw (2000); Stumm and Morgan (1995); Aylward and Findlay (1994); Snoeyink and Jenkins (1980); Lindsay (1971).

11

12


available. Dissolution of PRs is a prerequisite not only for the plant availability of P (Rajan et al. 1996) but also for the immobilization of metals through precipitation as metal phosphates (Laperche and Traina 1998). In soils, PRs dissolve by using the acid produced in the soils (Eq. 3); this is a major reason why PRs are very effective as a nutrient source mainly in acid soils (pH < 6.5) (Bolan et al. 1990) and as a metal-immobilizing agent in acid mine drainage (Evangelou and Zhang 1995). The rate of dissolution also depends on the chemical nature and the particle size of the PR. Dissolution rate increases with decreasing particle size. North Carolina phosphate rock (NCPR from the United States), Sechura PR (from Peru), Gafsa PR (from North Africa), and Chatham rise phosphorite (from New Zealand) are considered to be highly reactive. Once the PR is dissolved, the P released undergoes similar adsorption and precipitation reactions as in the case of soluble P fertilizers. Ca10(PO4)6F2 + 12H+ → 10Ca2+ + 6H2PO−4 + 2F−

(3)

Fertilizer brands such as Microbial Phosphate (New Zealand) and Coastal super (in Australia) contain a small amount of water-soluble P, PR, and elemental sulfur (S0). Microbial oxidation of S0 in these fertilizers releases sulfuric acid, which results in the solubilization of PRs (known as the biosuper effect) (Rajan et al. 1996) and the mobilization of metals (Loser et al. 2001; Schippers and Sand 1998).

V. Mechanisms for (Im)mobilization of Heavy Metals by Phosphate Compounds As indicated earlier, P compounds act both as a source and a sink for heavy metals in soils. As a source they enhance the mobilization of metals and as a sink they induce their immobilization. Phosphate compounds affect the (im)mobilization of metals in the soil–plant system through various processes (Fig. 2), which include P–micronutrient imbalance in plant nutrition, direct metal adsorption by P compounds, phosphate anion-induced metal adsorption and desorption, direct precipitation of metals with solution P as metal phosphates, precipitation through the liming action of PRs, and rhizosphere modification through acidification and mycorrhizal association. Both mobilization and immobilization of metals in soils treated with a range of P compounds have been reported, and the probable mechanisms for the P-induced (im)mobilization of metals reported in the literature are given in Tables 3 and 4. A. Phosphate Compounds as a Metal Source Phosphate compounds contain a range of metals (McLaughlin et al. 1996; Mortvedt 1996; Syers et al. 1986) (Table 5). According to Nriagu (1984), “virtually every known element has been found, at least in trace amounts, in a phosphate mineral.” Phosphate minerals are particularly favored as the host of uranium, thorium, and many other rare elements. Addition of P compounds to soils not


Fig.2. Plausible mechanisms by which phosphate compounds enhance the (im)mobilization of metals in soils.

13

14


Table 3. Selected references on the mobilization of heavy metals by water-soluble and water-insoluble phosphate compounds.

Phosphate compound

Method of Heavy metal investigationa

Water-soluble phosphate compounds NPK fertilizers

Cd

PB

NPK fertilizers

Cd

PB

SSP

Cd

PB

SSP

Cd

PB

TSP

Cd

PB

TSP

Se

PB

TSP

Cu

PB

Ca(H2PO4)2 Ca(H2PO4)2 KH2PO4

As Mo, B, Se Mo

Proposed primary mechanism

Cd source dissolution Cd source dissolution Cd source dissolution Cd source dissolution Cd source dissolution Competitive adsorption Acidification

TL Desorption TL Desorption AD, TL, PB Desorption

KH2PO4

Mo

AD

Desorption

KH2PO4

Cr

AD

KH2PO4

Se

PB

NaH2PO4

Se

PB

Competitive adsorption Competitive root absorption Desorption

As

CF, TL

NaH2PO4

As

CF

NaH2PO4

As

PB

NaH2PO4

As

PB

(NH4)2HPO4

Cd

CF, PB

Competitive adsorption Competitive adsorption Competitive adsorption Competitive adsorption Cd source dissolution

Reference

Singh and Myhr (1998) He and Singh (1994) Williams and David (1976) Gray et al. (1999) Sparrow et al. (1993) Carter et al. (1972) Timmer and Leydon (1980) Qafoku et al. (1999) Qafoku et al. (2001) Nuenhauserer et al. (2001) Xie et al. (1993); Xie and Mackenzie (1991) Aide and Cummings (1997) Broyer et al. (1972)

Hopper and Parker (1999) Creger and Peryea (1994) Reynolds et al. (1999) Woolson et al. (1973) Livesey and Huang (1981) Loganathan et al. (1996)


15

Table 3. (Continued).

Phosphate compound



NH4H2PO4, Ca(H2PO4)2

As

AD

Competitive adsorption

NH4H2PO4, Ca(H2PO4)2 KH2PO4

As

TL

Mo

AD

Competitive adsorption Competitive adsorption

As

CF

PR

Cd

PB

PR

Cd

PB

PR

Cd

PB

PR

Se

PB

Water-insoluble phosphate compounds Hydroxyapatite

Competitive adsorption Cd source dissolution Cd source dissolution Cd source dissolution Competitive adsorption

Reference Peryea (1991); Peryea and Kammereck (1997) Davenport and Peryea (1991) Barrow (1973)

Biosson et al. (1999) Gray et al. (1999) Loganathan et al. (1996) He and Singh (1994) Carter et al. (1972)

SSP: single superphosphate; TSP: triple superphosphate; PR: phosphate rock. aMethod of investigation is as follows: Chemical: adsorption/desorption (AD); chemical fractionation (CF); solubility diagram (SD); transport/leaching (TL). Biological: phytoavailability bioassay (PB).

only helps to overcome the deficiency of some of the essential trace elements, such as Mo, but also introduces toxic metals, such as Cd (McLaughlin et al. 1996). In this regard, Cd contamination of agricultural soils is of particular concern because this metal reaches the food chain through regular use of Cdcontaining fertilizer materials, such as SSP and TSP. This pathway is one of the main reasons this element has been studied extensively in relation to soil and plant factors affecting its bioavailability Accumulation of Cd in soils through regular fertilizer use has been observed in many countries (see Table 3). For example, in New Zealand and Australia, most of the Cd accumulation in pasture soils has been derived from the use of P fertilizers containing a high Cd concentration. The Cd in most P fertilizers originates mainly from the PRs used for manufacturing P fertilizers. It is important to stress that PRs deposits vary in their Cd content from trace to the parts per million (ppm) range, depending on the source of PR. Thus, manufactured P fertilizers also vary accordingly in their Cd content. The Cd in superphosphates is water soluble, and high-analysis P fertilizers, such as TSP, PAPR, and ammonium phosphates, generally contain lower Cd relative to P.

16


Table 4. Selected references on the immobilization of heavy metals by water-soluble and water-insoluble phosphate compounds.

P compound



Water-soluble phosphate compounds Ca(H2PO4)2 Pb CF, PB, XRD Precipitation as metal phosphates Ca(H2PO4)2 Cd, Zn PB Precipitation as metal phosphates K2HPO4, Pb, Zn, Cd CF, PB Phosphate(NH4)2HPO4 induced adsorption K2HPO4 Pb, Zn, Cd XRD, MB Precipitation as metal phosphates KH2PO4, Cd AD, SD PhosphateCa(H2PO4)2 induced adsorption KH2PO4 Cd AD, SD Phosphateinduced adsorption KH2PO4 Cd, Pb CF Phosphateinduced adsorption K2HPO4 Ni XRD, CF Precipitation as metal phosphates (NH4)2HPO4 Cd, Pb, Zn PBET, TL Precipitation as metal phosphates NO4H2PO4, Pb TL Precipitation as Ca(H2PO4)2 metal phosphates Na2HPO4 Pb XRD, XRF, Precipitation as EMPA metal phosphates Na2HPO4 Pb, Zn SEM, EDX, Precipitation as XRD metal phosphates NaH2PO4 Se PB Desorption Soil P

Pb, Cd

SD

Reference

Hettiarachchi et al. (2000) MacLean (1976)

Pierzynski and Schwab (1993) Pearson et al. (2000)

Bolan et al. (1999b)

Naidu et al. (1994)

Tu et al. (2000)

Pratt et al. (1964)

McGowen et al. (2001)

Davenport and Peryea (1991) Ruby et al. (1994)

Cotter-Howells and Capron (1996)

Hopper and Parker (1999) Precipitation as Santillan-Medrano and metal Jurinak (1975) phosphates


17


P compound


Water-insoluble phosphate compounds Hydroxyapatite Cd Hydroxyapatite

Zn

Hydroxyapatite

Cd

Hydroxyapatite

Pb

Hydroxyapatite

Pb

Hydroxyapatite

Cd, Zn

Hydroxyapatite

Pb

Hydroxyapatite

Zn, Pb, Cu, Cd

Hydroxyapatite

Pb

Hydroxyapatite

Pb

Hydroxyapatite

Cd, Pb, Zn

PR (NCPR)

Pb, Zn, Cd

PR

Pb

PR

Pb, Zn, Cd

CF, PB


Cation exchange Adsorption

Reference

Jeanjean et al. (1995)

Chlopecka and Adriano (1996) AD Adsorption, ion Mandjiny et al. (1998) exchange, surface complexation, precipitation CF Precipitation as Berti and Cunningham metal (1997) phosphates SEM, EDX, Precipitation as Laperche and Traina IR, EXAPS metal (1998) phosphates Surface Xu et al. (1994) complexation and coprecipitation AD Precipitation as Xu et al. (1994) metal phosphates CF Precipitation/ Boisson et al. (1999) adsorption/ cation exchange AD Precipitation as Zhang et al. (1997) metal phosphates AD, CF Precipitation as Ma et al. (1993) metal phosphates AD, TCLP Adsorption/ Seaman et al. (2001) precipitation CF, GI, Precipitation as Basta et al. (2001) PBET, PB metal phosphates CF Precipitation as Ma and Rao (1997); Ma metal et al. (1997) phosphates XRD, SEM Adsorption/ Chen et al. (1997b) precipitation

18


Table 4. (Continued). Proposed primary mechanism


P compound PR

Zn

AD

PR waste clay

Cd

CF, PB

Soil P P rich biosolid

Mn Cd

SD AD, SD

Adsorption/ precipitation Adsorption/ precipitation Precipitation Adsorption/ precipitation

Reference Prasad et al. (2001) Gonzalez et al. (1992) Schwab (1989) Soon (1981)

aMethod of investigation is as follows: Chemical: adsorption/partitioning (AD); chemical fractionation (CF); solubility diagram (SD); transport/leaching (TL); toxicity characteristics leaching procedure (TCLP). Minerological: X-ray diffraction (XRD); X-ray fluorescence (XRF); scanning electron microscopic (SEM); electron microprobe analysis (EMPA); energy-dispersive X-ray analysis (EDX); infrared (IR). Biological: phytoavailability bioassay (PB); gastrointestinal test (GI); physiologically based extraction test (PBET); microfauna/macrofauna bioassay (MB). PR: phosphate rock.

Table 5. Metal concentration in phosphate compounds from various sources. Phosphatea compound GPR NFPR JPR NCPR SPR MPR NIPR APR MIPR CRP IRP SSPb TSPb DAPb

Concentration (mg kg−1) As

Cd

Co

Cu

Zn

Mn

4 7 12 23 5 3 3 7 2

38 3 4 48 11 8 100 12 10 2

3 5 1 indicate greater effectiveness of mycorrhizal compared to nonmycorrhizal plants. Source: Bolan (1991).


33

crease in P uptake from PRs is identical to that from soluble P fertilizers. Similarly, the increase in the availability of P for mycorrhizal plants from P that had been allowed to react with soil is similar to that from freshly applied P. These studies were not able to provide conclusive evidence for the dissolution of water-insoluble fertilizers by mycorrhizae (Ness and Vlek 2000). Somewhat contradicting, Bolan (1991) noticed that the increase in P uptake with mycorrhizal infection varied with the P solubility, with greatest benefit arising from the least soluble P source. The logical explanation for the increased uptake by mycorrhizal plants in this case is greater surface volume in the bulk soil explored by the fungi. A further possibility is that mycorrhizal hyphae may be able to chemically modify the availability of less-soluble P sources by producing organic compounds with chelating properties, such as citrate. In addition, production of phosphatases by ectomycorrhizal fungi is important in the solubilization of organic phytates, which constitute a large fraction of total P in humic soils (Bartlett and Lewis 1973; Mitchell and Read 1981; Williamson and Alexander 1975). Similarly, ectomycorrhizae have been shown to produce large amounts of calcium oxalate (Lapeyrie 1988; Malajczuk and Cromack 1982), which may be involved in the chelation of Fe and Al, thereby releasing P for plant uptake (Graustein et al. 1977; Treeby et al. 1989). Some experimental evidence for direct chemical modification of P availability by endomycorrhizal plants is available (Abbott and Robson 1982; Gianinazzi-Pearson and Gianinazzi 1978). Parfitt (1979) suggested that the increased uptake of P from goethite–phosphate complexes by mycorrhizal plants might result from increased production of citrate and other organic compounds. Similarly, Jayachandran et al. (1989) have observed that in the presence of synthetic chelates (EDDHA) mycorrhizae caused greater uptake of P than in the absence of these chelates, whereas nonmycorrhizal plants were unable to take advantage of the P released by chelation. In this case, siderophore production by mycorrhizal fungi or other soil microbes could significantly increase P availability in lowpH soils, and this is a feasible mechanism by which mycorrhizal plants could acquire P sources unavailable to nonmycorrhizal plants. Differences in the absorption of anions and cations by mycorrhizal and nonmycorrhizal plants may lead to differences in rhizosphere pH (Buwalda et al. 1983), which may then affect the availability of adsorbed P and metals to plants (Hedley et al. 1982). Because mycorrhizal plants utilize NH+4−N more efficiently than nonmycorrhizal plants (Smith et al. 1985), this could create pH differences in the rhizosphere. It is logical that H+ extrusion, which is an inevitable consequence of NH+4−N assimilation in cells (Bolan et al. 1991; Raven and Smith 1976), would occur from the hyphae as well as from the roots (Raven et al. 1978). This extrusion could reduce the pH around the infected root, thereby enhancing the availability of P and the associated metals from sparingly soluble P sources such as PRs. Colonization of plant roots with mycorrhizal fungi could have synergistic or antagonistic effects on the rhizosphere microflora (Ames et al. 1984; Meyer and Linderman 1986). Consequently the changes in rhizosphere microflora indirectly

34


affect the availability of both organic and inorganic P sources and metals to plants (Azcon et al. 1976). The effect of P on mycorrhizal infection and the subsequent uptake of trace elements by plants has been explored (Gildon 1983; Gildon and Tinker 1983; Lambert et al. 1979; Pacovsky 1986; Timmer and Leydon 1980). Findings indicate that high levels of bioavailable P in soils inhibited mycorrhizal infection, concomitantly reducing the uptake of metals. For example, Lambert et al. (1979) observed that high levels of P in soils reduced mycorrhizal infection, lowering Zn uptake. Similarly, Pacovsky (1986) observed that mycorrhizal plants increased the uptake of Zn and Cu but this decreased with increasing level of soil P. In examining the relationship of mycorrhizal infection to P-induced Cu deficiency in sour orange seedlings, Timmer and Leydon (1980) concluded that, in the case of nonmycorrhizal plants, P induces Cu deficiency by stimulating growth of plants until Cu becomes a limiting nutrient, whereas in the case of mycorrhizal plants Cu deficiency occurs because of direct P-induced inhibition of mycorrhizal development. The antagonistic effect of P on mycorrhizal infection is likely to have a strong influence on phytoremediation of metals in contaminated soils. There has been some indication that mycorrhizal infection, particularly with the heavy metal tolerant isolate, could protect plants against heavy metal phytotoxicity (Gildon and Tinker 1981; Griffioen 1994; Griffioen et al. 1994). These infections might be important in revegetation of polluted sites, and any attempt to immobilize metals using high levels of soluble P compounds is likely to affect the mycorrhizae, leading to poor establishment of plant colonies.

Summary A large number of studies have provided conclusive evidence for the potential value of both water-soluble (e.g., DAP) and water-insoluble (e.g., apatite, also known as PRs) P compounds to immobilize metals in soils, thereby reducing their bioavailability for plant uptake. It is, however, important to recognize that, depending on the nature of P compounds and the heavy metal species, application of these materials can cause either mobilization or immobilization of the metals. Furthermore, some of these materials contain high levels of metals and can act as an agent of metal introduction to soils. Accordingly, these materials should be scrutinized before their large-scale use as immobilizing agent in contaminated sites. Although mobilization by certain P compounds enhances the bioavailability of metals, immobilization inhibits their plant uptake and reduces their transport in soils and subsequent groundwater contamination. Whenever phytoremediation of contaminated sites is practicable, appropriate P compounds can be used to enhance the bioavailability of metals for plant uptake. Removal of metals through phytoremediation techniques and the subsequent recovery of the metals or their safe disposal are attracting research and commercial interests. Phosphate


35

compounds can be used to enhance the solubilization of metals, leading to their increased uptake by plants. However, when it is not possible to remove the metals from the contaminated sites by phytoremediation, other viable options such as in situ immobilization should be considered as an integral part of risk management. One way to facilitate such immobilization is by altering the physicochemical properties of the metal–soil complex by introducing a multipurpose anion, such as phosphate, that enhances metal adsorption via anion-induced negative charge (i.e., CEC) and metal precipitation. It is important to recognize that large-scale use of P compounds can lead to surface and groundwater contamination of this element. It is therefore, important that future research should aim to focus on the role of P compounds on in situ remediation and natural attenuation in metal-contaminated sites, with minimum impact of P on quality of water sources.

Acknowledgments The U.S. Department of Energy contract number DE-FC-09–96SR18546 with the University of Georgia’s Savannah River Ecology Laboratory supported writing and editing time for Drs. Bolan and Adriano.

References Abbott KL, Robson AD (1982) The role of vesicular-arbuscular mycorrhizal fungi in agriculture and the selection of fungi for inoculation. Aust J Agric Res 33:389–408. Adriano DC (2001) Trace Elements in Terrestrial Environments; Biogeochemistry, Bioavailability and Risks of Metals, 2nd Ed. Springer, New York. Agbenin JO (1998) Phosphate-induced zinc retention in a tropical semi-arid soil. Eur J Soil Sci 49:693–700. Aide MT, Cummings MF (1997) The influence of pH and phosphorus on the adsorption of chromium (VI) on boehmite. Soil Sci 162:599–603. Alloway BJ (1990) Cadmium. In: Alloway BJ (ed) Heavy Metals in Soils. Wiley, New York, pp 100–124. Ames RN, Reid CPP, Ingham ER (1984) Rhizosphere bacterial population responses to root colonization by a vesicular-arbuscular mycorrhizal fungus. New Phytol 96: 555–563. Amijee F, Tinker PB, Stribley DP (1989) The development of endomycorrhizal root systems. 7. A detailed study of effects of soil-phosphorus on colonization. New Phytol 111:435–446. Ando J (1987) Thermal phosphate. In: Nielsson FT (ed) Manual of Fertilizer Processing. Dekker, New York, pp 93–124. Andrew CS, Johnson AD (1976) Effect of calcium, pH and nitrogen on the growth and chemical composition of some tropical and temperate pasture legumes. II. Chemical composition (calcium, nitrogen, potassium, magnesium, sodium and phosphorus). Aust J Agric Res 27:625–636. Avudainayagam S, Naidu R, Kookana RS, Alston AM, McClure S, Smith LH (2001) Effects of electrolyte composition on chromium desorption in soils contaminated by tannery waste. Aust J Soil Res 39:1077–1089

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Ayers AS, Hagihara HH (1953) Effect of anion on the sorption of potassium by some humic and hydrol humic Latosols. Soil Sci 75:1–17. Aylward G, Findlay T (1994) SI Chemical Data, 3rd Ed. Wiley, Brisbane, Australia. Azcon R, Barea JM, Hayman DS (1976) Utilization of rock phosphate in alkaline soils by plants inoculated with mycorrhizal fungi and phosphate solubilizing bacteria. Soil Biol Biochem 8:135–138. Barrow NJ (1973) On the displacement of adsorbed anions from soil. 1. Displacement of molybdenum by phosphate and by hydroxide. Soil Sci 116:423–431. Barrow NJ (1985) Reactions of anions and cations with variable-charge soils. Adv Agron 38:183–230. Barrow NJ (1987) The effects of phosphate on zinc sorption by a soil. J Soil Sci 38: 453–459. Bartlett EM, Lewis DH (1973) Surface phosphatase activity of mycorrhizal roots of beech. Soil Biol Biochem 5:249–257. Bartlett RJ (1991) Chromium cycling in soils and water: links, gaps, and methods. Environ Health Perspect 92:17–24. Bartlett RJ, James BR (1979) Behavior of chromium in soils. III. Oxidation. J Environ Qual 8:31–35. Bartlett RJ, Kimble JM (1976) Behavior of chromium in soils. II. Hexavalent forms. J Environ Qual 5:383–386. Basta NT, Gradwohl R (2000) Estimation of Cd, Pb, and Zn bioavailability in smeltercontaminated soils by a sequential extraction procedure. J Soil Contam 9:149–164. Basta NT, Tabatabai MA (1992) Effect of cropping systems on adsorption of metals by soils. II. Effect of pH. Soil Sci 153:195–204. Basta NT, Gradwohl R, Snethen KL, Schroder JL (2001) Chemical immobilisation of lead, zinc and cadmium in smelter-contaminated sols using biosolids and rock phosphate. J Environ Qual 30:1222–1230. Berti WR, Cunningham SD (1997) In-place inactivation of Pb in Pb-contaminated soils. Environ Sci Technol 31:1359–1364. Boekhold AE, Temminghoff EJM, van der Zee SEATM (1993) Influence of electrolyte composition and pH on cadmium adsorption by an acid sandy soil. J Soil Sci 44: 85–96. Boisson J, Mench M, Vangronsveld J, Ruttens A, Kopponen P, DeKoe T (1999) Immobilization of trace metals and arsenic by different soil additives: Evaluation by means of chemical extractions. Commun Soil Sci Plant Anal 30:365–387. Bolan NS (1991) A critical review on the effect of mycorrhizal fungi on the uptake of phosphorus by plants. Plant Soil 134:189–207. Bolan NS, Thiyagarajan S (2001) Retention and plant availability of chromium in soils as affected by lime and organic amendments. Aust J Soil Res 39:1091–1103. Bolan NS, Barrow NJ, Posner AM (1985) Describing the effect of time on the sorption of phosphate by iron and aluminum hydroxides. J Soil Sci 36:187–196. Bolan NS, Syers JK, Tillman RW (1986) Ionic strength effects on surface charge and adsorption of phosphate and sulfate by soils. J Soil Sci 37:379–388. Bolan NS, Hedley MJ, White RE (1991) Nitrogen fixation and soil acidification with emphasis on legume based pastures. Plant Soil 134:53–63. Bolan NS, White RE, Hedley MJ (1990) A review of the use of phosphate rock as fertilizer for direct application in Australia and New Zealand. Aust J Exp Agric 30: 297–313.


37

Bolan NS, Hedley MJ, Loganathan P (1993) Preparation, forms and properties of slowrelease phosphate fertilizers. Fert Res 35:13–24. Bolan NS, Naidu R, Syers JK, Tillman RW (1999a) Surface charge and solute interactions in soils. Adv Agron 67:88–141. Bolan NS, Naidu R, Syers JK, Tillman RW (1999b) Effect of anion sorption on cadmium sorption by soils. Aust J Soil Res 37:445–460. Bolland MDA, Posner AM, Quirk JP (1977) Zinc adsorption by goethite in the absence and presence of phosphate. Aust J Soil Res 15:279–286. Bramley RGV (1990) Cadmium in New Zealand agriculture. N Z J Agric Res 33: 505–519 Broyer TC, Johnson CM, Huston RP (1972) Selenium and nutrition of Astragalus. II. Ionic sorption interactions among selenium, phosphate, and the macro- and micronutrient cations. Plant Soil 36:651–669. Brudevold F, Steadman LT, Spinelli MA, Amdur BH, Gron P (1963) A study of zinc in human teeth. Arch Oral Biol 8:135–144. Buwalda JG, Stribley DP, Tinker PB (1983) Increased uptake of anions by plants with vesicular-arbuscular mycorrhizas. Plant Soil 71:463–467. Carter DL, Robbis CW, Brown MJ (1972) Effect of phosphorus fertilization on the selenium concentration in alfalfa (Medicago sativa). Soil Sci Soc Am Proc 36:624–628. Chakraborty AK, Saha KC (1987) Arsenical dermatosis from tube well water in West Bengal. Ind J Med Res 85:326–334. Chen XB, Wright JV, Conca JL (1997a) Effects of pH on heavy metal sorption on mineral apatite. Environ Sci Technol 31:624–631. Chen XB, Wright JV, Conca JL, Peurrung LM (1997b) Evaluation of heavy metal remediation using mineral apatite. Water Air Soil Pollut 98:57–78. Chlopecka A, Adriano DC (1996) Mimicked in situ stabilization of metals in a cropped soil: bioavailability and chemical form of zinc. Environ Sci Technol 30:3294–3303. Chlopecka A, Adriano DC (1997) Influence of zeolite, apatite and Fe-oxide on Cd and Pb uptake by crops. Sci Total Environ 207:195–206. Corey RB (1981) Adsorption vs precipitation. In: Anderson MA, Robin AJ (eds) Adsorption of Inorganics at Solid–Liquid Interfaces. Ann Arbor Science, Ann Arbor, MI, pp 51–90. Cotter-Howells J, Capron S (1996) Remediation of contaminated land by formation of heavy metal phosphates. Appl Geochem 11:335–342. Creger TL, Peryea FJ (1994) Phosphate fertilizer enhances arsenic uptake by apricot liners grown in lead-arsenate-enriched soil. Hortic Sci 29:88–92. Curtin D, Syers JK (1990) Mechanism of sulfate adsorption by two tropical soils. J Soil Sci 41:295–304. Davenport JR, Peryea FJ (1991) Phosphate fertilizers influence leaching of lead and arsenic in a soil contaminated with lead arsenate. Water Air Soil Pollut 57(8):101– 110. Dhillon KS, Dhillon DK (1990) Selenium toxicity in soil-plant-animal system: a case study. Trans 14th Int Congr Soil Sci Commun IV VI:300–305. Dudka S, Chlopecka A (1990) Effect of solid-phase speciation on metal mobility and phytoavailability in sludge-amended soil. Water Air Soil Pollut 51:153–160. Evangelou VP, Zhang YL (1995) A review: pyrite oxidation mechanisms and acid mine drainage prevention. CRC Crit Rev Environ Sci Technol 252:141–199. Feng D, Aldrich C, Tan H (2000) Treatment of acid mine water by use of heavy metal precipitation and ion exchange. Miner Eng 13:623–642.

38


Ford RG, Sparks DL (2000) The nature of Zn precipitates formed in the presence of pyrophyllite. Environ Sci Technol 34:2479–2483. Frazer L (2001) Probing the depths of a solution for acid mine drainage. Environ Health Perspect 109:486–489. Fu G, Allen HE, Cowan CE (1991) Adsorption of cadmium and copper by manganese oxide. Soil Sci 152:72–81. Gianinazzi-Pearson V, Gianinazzi S (1978) Enzymatic studies on the metabolism of vesicular-arbuscular mycorrhiza. II. Soluble alkaline phosphatase specific to mycorrhizal infection in onion roots. Physiol Plant Pathol 12:45–53. Gildon A (1983) The relationship of vesicular arbuscular mycorrhizal infection and copper nutrition. J Sci Food Agric 34:56–57. Gildon A, Tinker PB (1981) A heavy metal-tolerant strain of a mycorrhizal fungus. Trans Br Mycol Soc 77:648–649. Gildon A, Tinker PB (1983) Interactions of vesicular arbuscular mycorrhizal infections and heavy-metals in plants. 2. The effects of infection on uptake of copper. New Phytol 95:263–268. Gonzalez RX, Sartain JB, Miller WL (1992) Cadmium availability and extractability from sewage-sludge as affected by waste phosphatic clay. J Environ Qual 21:272– 275. Graustein WC, Cromack K, Sollins P (1977) Calcium oxalate occurrence in soils and effect on nutrient and geochemical cycles. Science 198:1252–1254. Gray CW, McLaren RG, Roberts AHC, Condron LM (1999) Effect of soil pH on cadmium phytoavailability in some New Zealand soils. N Z J Crop Hortic 27:169–179. Griffioen WAJ (1994) Characterization of a heavy metal-tolerant endomycorrhizal fungus from the surroundings of a zinc refinery. Mycorrhiza 4:197–200. Griffioen WAJ, Ietswaart JH, Ernst WHO (1994) Mycorrhizal infection of an Agrostis capillaris population on a copper-contaminated soil. Plant Soil 158:83–89. Gworek B (1992) Lead inactivation in soils by zeolites. Plant Soil 143:71–74. Haas CI, Horowitz ND (1986) Adsorption of cadmium to kaolinite in the presence of organic material. Water Air Soil Pollut 27:131–140. Harter RDR, Naidu R (1995) Role of metal-organic complexation in metal sorption by soils. Adv Agron 55:219–264. He QB, Singh BR (1994) Plant availability of cadmium in soils. 2. Factors related to the extractability and plant uptake of cadmium in cultivated soils. Acta Agric Scand 43: 142–150. Hedley MJ, White RE, Nye PH (1982) Plant-induced changes in the rhizosphere of rape (Brassica napus var. Emerald) seedlings. II. Origin of the pH changes. New Phytol 91:31–44. Helyar KR, Munns DN, Burau RG (1976) Adsorption of phosphate by gibbsite. II. Formation of a surface complex involving divalent cations. J Soil Sci 27:315–323. Hettiarachchi GM, Pierzynski GM, Ransom MD (2000) In situ stabilization of soil lead using phosphorus and manganese oxide. Environ Sci Technol 34:4614–4619. Hingston FJ (1981) A review of anion adsorption. In: Anderson MA, Robin AJ (eds) Adsorption of Inorganics at Solid-Liquid Interfaces, Ann Arbor Science, Ann Arbor, MI, pp 51–90. Hodgson JF, Tiller KG, Fellows M (1964) The role of hydrolysis in the reaction of heavy metals in soil forming materials. Soil Sci Soc Am Proc 28:42–46. Hopper JL, Parker DR (1999) Plant availability of selenite and selenate as influenced by the competing ions phosphate and sulfate. Plant Soil 210:199–207.


39

James BR (1996) The challenge of remediating chromium contaminated soil. Environ Sci Technol 30:248–251. James BR, Petura JC, Vitale RJ, Mussoline GR (1995) Hexavalent chromium extraction from soils: a comparison of five methods. Environ Sci Technol 29:2377–2380. Jayachandran K, Schwab AP, Hetrick BAD (1989) VA mycorrhizal mediation of phosphorus availability: effect of synthetic iron chelate on phosphorus solubilization. Soil Sci Soc Am J 35:1701–1706. Jeanjean J, Fedoroff M, Faverjon F (1995) Influence of pH on the sorption of cadmium ions on calcium hydroxyapatite. J Mater Sci 30:6156–6160. Khan AG, Kuek C, Chaudhry TM, Koo CS, Hayes WJ (2000) Role of plants, mycorrhizae and phytochelators in heavy metal contaminated land remediation. Chemosphere 41:197–207. Khattak R, Page AL, Parker DR, Baker D (1991) Accumulation and interactions of arsenic, selenium, molybdenum and phosphorus in alfalfa. J Environ Qual 20:165–168. Knox AS, Seaman JC, Mench MJ, Vangronsveld J (2000) Remediation of metal and radionuclides-contaminated soils by in situ stabilization techniques. In: Iskandar IK (ed) Environmental Restoration of Metals-Contaminated Soils. Lewis, New York, pp 21–60. Krishnamurti GSR (2000) Speciation of heavy metals: an approach for remediation of contaminated soils. In: Wise DL (ed) Remediation Engineering of Contaminated Soils. Dekker, New York, pp 693–716. Krishnamurti GSR, Naidu R (2000) Speciation and phytoavailability of cadmium in selected surface soils of South Australia. Aust J Soil Res 38:991–1004. Krishnamurti GSR, Huang PM, Van Rees KCJ, Kozak LM, Rostad HPW (1995) Speciation of particulate-bound cadmium of soils and its bioavailability. Analyst 120:659– 665. Krishnamurti GSR, Huang PM, Van Rees KCJ (1996) Studies on soil rhizosphere: speciation and availability of cadmium. Chem Speciat Bioavailab 8:23–28. Krishnamurti GSR, Cieslinski G, Huang PM, Van Rees KCJ (1997a) Kinetics of cadmium release from soils as influenced by organic acids. Implication in cadmium availability. J Environ Qual 26:271–277. Krishnamurti GSR, Huang PM, Van Rees, KCJ, Kozak LM, Rostad HPW (1997b) Differential FTIR study of pyrophosphate-extractable material of soils: implication in Cd-bonding sites and availability. In: Prost R (ed) Contaminated Soils. Proceedings, 3rd International Conference on Biogeochemistry of Trace Elements. Institut National de la Recherche Agronomique, Paris, pp 1–10. Krishnamurti GSR, Huang PM, Kozak LM (1999) Sorption and desorption kinetics of cadmium in soils: influence of phosphate. Soil Sci 164:888–898. Kuo S (1986) Concurrent adsorption of phosphate and zinc, cadmium, or calcium by a hydrous ferric oxide. Soil Sci Soc Am J 50:1412–1419. Lambert DH, Baker DE, Cole H (1979) The role of mycorrhizae in the interactions of phosphorus with zinc, copper and other elements. Soil Sci Soc Am J 43:976–980. Laperche V, Traina SJ (1998) Immobilization of Pb by hydroxyapatite. In: Everett JA (ed) Adsorption of Metals by Geomedia: Variables, Mechanisms, and Model Applications. Academic Press, Orlando, pp 225–276. Laperche V, Logan TJ, Gaddam P, Traina SJ (1997) Effect of apatite amendments on plant uptake of lead from contaminated soil. Environ Sci Technol 7:91–102. Lapeyrie F (1988) Oxalate synthesis from soil bicarbonate by the mycorrhizal fungus Paxillus involutus. Plant Soil 110:3–8.

40


Lazic S, Vukovic Z (1991) Ion exchange of strontium on synthetic hydroxyapatite. J Radioanal Nucl Chem 19:161–168. Lindsay WL (1971) Chemical Equilibria in Soils. Wiley, New York. Livesey NT, Huang PM (1981) Adsorption of arsenate by soils and its relations to selected chemical properties and anions. Soil Sci 131:88–94. Loganathan P, Hedley MJ, Gregg PEH, Currie LD (1996) Effect of phosphate fertiliser type on the accumulation and plant availability of cadmium in grassland soils. Nutr Cycl Agroecosyst 46:169–178. Loneragan JF, Grove TS, Robson AD, Snowball K (1979) Phosphorus toxicity as a factor in zinc-phosphorus interactions in plants. Soil Sci Soc Am J 43:966–972 Loser C, Seidel H, Hoffmann P, Zehnsdorf A (2001) Remediation of heavy metalcontaminated sediments by solid-bed bioleaching. Environ Geol 40:643–650. Ma LQ, Rao GN (1997) Effects of phosphate rock on sequential chemical extraction of lead in contaminated soils. J Environ Qual 26:788–794. Ma LQ, Traina SJ, Logan TJ, Ryan JA (1993) In situ Pb immobilization by apatite. Environ Sci Technol 27:1803–1810. Ma LQ, Choate AL, Rao GN (1997) Effects of incubation and phosphate rock on lead extractability and speciation in contaminated soils. J Environ Qual 26:801–807. Maclean AJ (1976) Cadmium in different plant species and its availability in soils as influenced by organic-matter and additions of lime, P, Cd and Zn. Can J Soil Sci 56: 129–138. Mahler RJ, Bingham FT, Page AL (1978) Cadmium-enriched sewage sludge application to acid and calcareous soils—effect on yield and cadmium uptake by lettuce and chard. J Environ Qual 7:274–281. Malajczuk N, Cromack K Jr (1982) Accumulation of calcium oxalate in the mantle of ectomycorrhizal roots of Pinus radiata and Eucalyptus marginata. New Phytol 92: 527–531. Malone CP, Koeppe DE, Miller RJ (1974) Localisation of lead accumulated by corn plants. Plant Physiol 53:388–394. Mandjiny S, Matis KA, Fedoroff M (1998) Calcium hydroxyapatites: evaluation of sorption properties for cadmium ions in aqueous solution. J Mater Sci 33:5433–5439. Marschner H (1995) Mineral Nutrition of Higher Plants, 2nd Ed. Academic Press, London. McGowen SL, Basta NT, Brown GO (2001) Use of diammonium phosphate to reduce heavy metal solubility and transport in smelter-contaminated soil. J Environ Qual 30: 493–500. McLaughlin MJ, Tiller KG, Naidu R, Stevens DP (1996) Review: the behaviour and environmental impact of contaminants in fertilizers. Aust J Soil Res 34:1–54. Mench MJ, Didier VL, Loeffler M, Gomez A, Masson P (1994) A mimicked in situ remediation study of metal contaminated soils with emphasis on cadmium and lead. J Environ Qual 23:58–63. Meyer JR, Linderman RG (1986) Selective influence on population of rhizosphere or rhizoplane bacteria and actinomycete by mycorrhizas formed by Gloumus fasciculatus. Soil Biol Biochem 18:191–196. Middelburg JJ, Comans MJ (1991) Sorption of cadmium on hydroxyapatite. Chem Geol 90:45–53. Misra DN, Bowen RL, Wallace BM (1975) Adhesive bonding of various materials to hard tooth tissues. 8. Nickel and copper ions on hydroxyapatite—role of ion-exchange and surface nucleation. J Colloid Interface Sci 51:36–43


41

Mitchell DT, Read D (1981) Utilization of inorganic and organic phosphates by the mycorrhizal endophytes of Vaccinium macrocarpon and Rhododendron ponticum. Trans Br Mycol Soc 76:255–260. Mortvedt JJ (1996) Heavy metal contaminants in inorganic and organic fertilizers. Fert Res 43:55–61. Naidu R, Skinner HCW (1999) Arsenic contamination of rural ground-water supplies in Bangladesh and India: implications for soil quality, animal and human health. In: Barber C, Humphries B, Dixon J (eds) Proceedings, International Conference on Diffuse Pollution, 16–20 May 1999, Perth, pp 407–417. Naidu R, Syers JK, Tillman RW, Kirkman JH (1990) Effect of liming and added phosphate on charge characteristics of acid soils. J Soil Sci 41:157–164. Naidu R, Bolan NS, Kookana RS, Tiller KG (1994) Ionic-strength and pH effects on the adsorption of cadmium and the surface charge of soils. Eur J Soil Sci 45:419–429. Naidu R, Kookana RS, Oliver DP, Rogers S, McLaughlin MJ (1996a) Contaminants and the Soil Environment in the Australasia-Pacific region, Kluwer, London. Naidu R, Kookana RS, Sumner ME, Harter RD, Tiller KG (1996b) Cadmium adsorption and transport in variable charge soils: a review. J Environ Qual 26:602–617. Ness RLL, Vlek PLG (2000) Mechanism of calcium and phosphate release from hydroxyapatite by mycorrhizal hyphae. Soil Sci Soc Am J 64:949–955. Neunhauserer C, Berreck M, Insam H (2001) Remediation of soils contaminated with molybdenum using soil amendments and phytoremediation. Water Air Soil Pollut 128:85–96. Nriagu JO (1984) Formation and stability of base metal phosphates in soils and sediments. In: Nriagu JO, Moore PB (eds) Phosphate Minerals. Springer-Verlag, New York, pp 318–329. Nriagu JO (1988) A silent epidemic of environmental metal poisoning. Environ Pollut 50:139–161. O’Connor GA, O’Connor C, Cline GR (1984) Sorption of cadmium by calcareous soils: Influence of solution composition. Soil Sci Soc Am J 48:1244–1247. Pacovsky RS (1986) Micronutrient uptake and distribution in mycorrhizal or phosphorusfertilized soybeans. Plant Soil 95:379–388. Parfitt RL (1979) The availability of P from phosphate-goethite bridging complexes: desorption and uptake of ryegrass. Plant Soil 53:55–65. Pearson MS, Maenpaa K, Pierzynski GM (2000) Effects of soil amendments on the bioavailability of lead, zinc, and cadmium to earthworms. J Environ Qual 29:1611– 1617. Peryea FJ (1991) Phosphate-induced release of arsenic from soils contaminated with lead arsenate. Soil Sci Soc Am J 55:1301–1306. Peryea FJ, Kammereck R (1997) Phosphate-enhanced movement of arsenic out of lead arsenate-contaminated topsoil and through uncontaminated subsoil. Water Air Soil Pollut 93:243–254. Pierzynski GM, Schwab AP (1993) Bioavailability of zinc, cadmium and lead in a metal contaminated alluvial soil. J Environ Qual 22:247–254. Power JF, Dick WA (2000) Land Application of Agricultural, Industrial, and Municipal By-Products. Soil Science Society of America, Inc., Madison, WI. Prasad M, Saxena S, Amritphale SS, Chandra N (2001) Detoxification of aqueous zinc using fluorapatite-bearing lean grade rock phosphate. Environ Technol 22:367–371. Pratt PF, Blair FL, McLean GW (1964) Reactions of phosphate with soluble and exchangeable nickel. Soil Sci Soc Am Proc 28:363–365.

42


Qafoku NP, Kukier U, Sumner ME, Miller WP, Radcliffe DE (1999) Arsenate displacement from fly ash in amended soils. Water Air Soil Pollut 114:185–198. Qafoku NP, Dudka S, Sumner ME, Miller WP (2001) Arsenic, boron, selenium, and molybdenum displacement and transport in a fly ash amended soil leached with calcium phosphate solution. Commun Soil Sci Plant Anal 32:1499–1512. Rai D, Earty LE, Zachara JM (1987) Chromium(III) hydrolysis constant and solubility of chromium(III) hydroxide. Inorg Chem 26:345–349. Rajan SSS (1978) Sulfate adsorbed on hydrous alumna, ligands displaced, and changes in surface charge. Soil Sci Soc Am J 42:39–44. Rajan SSS, Perrott KW, Saunders WMH (1974) Identification of phosphate-reactive sites of hydrous alumna from proton consumption during phosphate adsorption at constant pH values. J Soil Sci 25:438–447. Rajan SSS, Watkinson JH, Sinclair AG (1996) Phosphate rocks for direct application to soils. Adv Agron 57:77–159. Raven JA, Smith FA (1976) Nitrogen assimilation and transport in vascular land plants in relation to intercellular pH regulation. New Phytol 76:415–431. Raven JA, Smith SE, Smith FA (1978) Ammonium assimilation and the role of mycorrhizas in climax communities in Scotland. Trans Bot Soc Edinb 43:27–35. Reynolds JG, Naylor DV, Fendorf SE (1999) Arsenic sorption in phosphate-amended soils during flooding and subsequent aeration. Soil Sci Soc Am J 63:1149–1156. Roberts A, Longhurst RD, Brown MW (1994) Cadmium status of soils, plant and grazing animals in New Zealand. N Z J Agric Res 33:119–129. Ross SM (1994) Retention, transformation and mobility of toxic metals in soils. In: Ross SM (ed) Toxic Metals in Soil–Plant Systems. Wiley, New York, pp 63–152. Rothbaum HP, Goguel RL, Johnson AE, Mattingly GEG (1986) Cadmium accumulation in soils from long continued application of superphosphate. J Soil Sci 37:99–107. Ruby MV, Davis A, Schoof R, Eberle S, Sellstone CM (1994) In situ formation of lead phosphates in soils as a method to immobilise lead. Environ Sci Technol 28:646–654. Ruby MV, Davis A, Schoof R, Eberle S, Sellstone CM (1996) Estimation of bioavailability using a physiologically based extraction test. Environ Sci Technol 30:420–430. Rumberg CB, Engel RE, Meggitt WF (1960) Effect of phosphorus concentration on the absorption of arsenate by oats from nutrient solution. Agron J 52:452–453. Ryden JC, Syers JK (1976) Calcium retention in response to phosphate adsorption by soils. Soil Sci Soc Am J 40:845–846. Santillan-Medrano J, Jurinak JJ (1975) The chemistry of lead and cadmium in soil: solid phase formation. Soil Sci Soc Am Proc 39:851–856. Sauve S, Norvell WA, McBride M, Hendershot W (2000) Speciation and complexation of cadmium in extracted soil solutions. Environ Sci Technol 34:291–296. Sawhney BL (1974) Charge characteristics of soils as affected by phosphate sorption. Soil Sci Soc Am J 38:159–160. Schalscha EB, Pratt PF, Soto D (1974) Effect of phosphate adsorption on the cationexchange capacity of volcanic ash soils. Soil Sci Soc Am J 38:539–540. Schippers A, Sand W (1998) Bacterial leaching of metal sulfides proceeds by two indirect mechanisms via thiosulfate or via polysulfides and sulfur. Appl Environ Microbiol 65:319–321. Schwab AP (1989) Manganese-phosphate solubility relationships in an acid soil. Soil Sci Soc Am J 53:1654–1660. Seaman JC, Arey JS, Bertsch PM (2001) Immobilization of nickel and other metals in contaminated sediments by hydroxyapatite addition. J Environ Qual 30:460–469.


43

Shuman LM (1986) Effect of ionic strength and anions on zinc adsorption by two soils. Soil Sci Soc Am J 50:1438–1442. Shuman LM (1991) Chemical forms of micronutrients in soils. In: Mortvedt JJ, Cox FR, Shuman LM, Welch RM (eds) Micronutrients in Agriculture. Soil Science Society of America, Madison, WI, pp 113–144. Singh BR, Myhr K (1998) Cadmium uptake by barley as affected by Cd sources and pH levels. Geoderma 84:185–194. Smith SE, St John BJ, Smith FA, Nicholas DJD (1985) Activity of glutamine synthetase and glutamine dehydrogenase in Trifolium subterraneum L. and Allium cepa L.: effect of mycorrhizal infection and phosphate nutrition. New Phytol 99:211–217. Smith E, Naidu R, Alston AM (1998) Arsenic in the soil environment: a review. Adv Agron 66:149–195. Smith E, Naidu R, Alston AM (1999) Chemistry of arsenic in soils. I. Sorption of arsenate and arsenite by four Australian soils. J Environ Qual 28:1719–1726. Smith E, Naidu R, Alston A (2002) Chemistry of arsenic in soils. II. Effect of pH and ionic strength. J Environ Qual 31:557–563. Soon YK (1981) Solubility and sorption of cadmium in soils amended with sewage sludge. J Soil Sci 32:85–95. Soon YK, Miller MH (1977) Changes in the rhizosphere due to NH+4 and NO−3 fertilization and phosphorus uptake by corn seedlings (Zea mays L.). Soil Sci Soc Am Proc 41:77–80. Snoeyink VL, Jenkins D (1980) Water Chemistry, Wiley, New York. Sparrow LA, Salardini AA, Bishop AC (1993) Field studies of cadmium in potatoes (Solanum tuberosum L). 1. Effects of lime and phosphorus on cv. russet burbankaust. J Agric Res 44:845–853. Sposito G (1984) The Surface Chemistry of Soils. Oxford University Press, New York. Sposito G, Lund LJ, Chang AC (1982) Trace metal chemistry in arid zone field soils amended with sewage sludge. I. Fractionation of Ni, Cu, Zn, Cd and Pb in solid phases. Soil Sci Soc Am J 46:260–264. Street JJ, Sabey BR, Lindsay WL (1978) Influence of pH, phosphorus, cadmium, sewage sludge, and incubation time on the solubility and plant uptake of cadmium. J Environ Qual 7:286–290. Stumm W, Morgan JJ (1995) Aquatic Chemistry: Chemical Equilibria and Rates in Natural Waters, 3rd Ed. Wiley, New York. Suzuki T, Hatsushika T, Hayakawa Y (1981) Synthetic hydroxyapatites employed as inorganic cation-exchangers. J Chem Soc Faraday Trans 77:1059–1062. Syers JJ, MacKay AD, Brown MW, Currie LD (1986) Chemical and physical characteristics of phosphate rock materials of ranging reactivity. J Sci Food Agric 37:1057– 1064. Taylor G (1975) Leaching rates of heavy metal ions in forest soils. Water Air Soil Pollut 9:137–148. Tessier A, Campbell PGC, Bissom M (1979) Sequential extraction procedure for the speciation of particulate trace metals. Anal Chem 51:844–850. Tiller KG (1988) Heavy metals in soils and their environmental significance. Adv Soil Sci 9:113–142. Timmer LW, Leydon RF (1980) The relationship of mycorrhizal infection to phosphorusinduced copper deficiency in sour orange seedlings. New Phytol 85:15–23. Treeby M, Marschner H, Ro¯mheld V (1989) Mobilization of iron and other micronutrient

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cations from a calcareous soil by plant-borne, microbial and synthetic metal chelators. Plant Soil 114:217–226. Tu C, Zheng CR, Chen HM (2000) Effect of applying chemical fertilizers on forms of lead and cadmium in red soil. Chemosphere 41:133–138. Wann SS, Uehara G (1978a) Surface charge manipulation in constant surface potential soil colloids. I. Relation to sorbed phosphorus. Soil Sci Soc Am J 42:565–570. Wann SS, Uehara G (1978b) Surface charge manipulation in constant surface potential soil colloids. II. Effect on solute transport. Soil Sci Soc Am J 42:886–888. Whitelaw MA (2002) Growth promotion of plants innoculated with phosphate-solubilizing fungi. Adv Agron 69:99–151. Williams CH, David DJ (1976) The accumulation in soil of cadmium residues from phosphate fertilizers and their effect on the cadmium content of plants. Soil Sci 121: 86–93. Williamson B, Alexander I (1975) Acid phosphatases localized in the sheath of beech mycorrhizas. Soil Biol Biochem 7:194–198. Woolson EA, Axley JH, Kearney P (1973) The chemistry and phytotoxicity of arsenic in soil: II. Effect of time and phosphorus. Soil Sci Soc Am Proc 37:254–258. Xie RJ, Mackenzie AF (1991) Molybdate sorption-desorption in soils treated with phosphate. Geoderma 48:321–333. Xie RJ, Mackenzie AF, Lou ZJ (1993) Causal-modeling pH and phosphate effects on molybdate sorption in 3 temperate soils. Soil Sci 155:385–397. Xu Y, Schwartz FW, Traina SJ (1994) Sorption of Zn2+ and Cd2+ on hydroxyapatite surfaces. Environ Sci Technol 28:1472–1480. Zhang P, Ryan JA, Yang J (1997) In vitro soil Pb solubility in the presence of hydroxyapatite. Environ Sci Technol 32:2763–2768. Manuscript received February 16; Accepted March 7, 2002.

 Springer-Verlag 2003

Rev Environ Contam Toxicol 177:45–122

Environmental Fate of Methyl Bromide as a Soil Fumigant Scott R. Yates, Jay Gan, and Sharon K. Papiernik Contents I. Introduction ....................................................................................................... 46 A. Environmental Concerns .............................................................................. 46 B. Economic Concerns ..................................................................................... 47 II. Chemical and Physical Properties of Methyl Bromide ................................... 49 III. Methyl Bromide Use as a Soil Fumigant ........................................................ 50 A. History and Scope of Use ........................................................................... 50 B. Application Methods .................................................................................... 51 IV. Ozone Depletion and Methyl Bromide ............................................................ 52 A. Reactions with Ozone .................................................................................. 52 V. Sampling and Analysis of Methyl Bromide in the Air ................................... 54 A. Container Methods ....................................................................................... 56 B. Adsorbent Methods ...................................................................................... 56 C. Other Methods .............................................................................................. 58 VI. Processes Affecting Environmental Fate of Methyl Bromide ........................ 58 A. Transformation ............................................................................................. 58 B. Phase Partitioning ........................................................................................ 68 VII. Simulating the Environmental Fate of Methyl Bromide ................................. 70 A. Transport Model .......................................................................................... 70 B. Mobility Indices ........................................................................................... 73 C. Simulating Transport in Relatively Dry Soils ............................................. 74 VIII. Methyl Bromide Diffusion in Soils ................................................................. 77 A. Diffusion Coefficient ................................................................................... 77 B. Methyl Bromide Diffusion in Soils ............................................................. 77 IX. Assessing Methyl Bromide Volatilization from Soil ...................................... 81 A. Measurement Methods ................................................................................. 81 B. Field Experiments to Determine Methyl Bromide Volatilization .............. 89 X. Potential Methods for Minimizing Volatilization ............................................ 94 A. Containment ................................................................................................. 94 B. Soil Conditions ............................................................................................. 102 C. Effect of Degradation Rate on Emissions ................................................... 104

Communicated by George W. Ware. S.R. Yates ( ), S.K. Papiernik USDA-ARS, George E. Brown Jr. Salinity Laboratory, 450 West Big Springs Road, Riverside, CA 92507-4617, USA J. Gan Department of Environmental Science, University of California, 417 Geology Riverside, CA, USA 92521

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S.R. Yates, J. Gan, and S.K. Papiernik

XI. Considerations for Developing Alternatives to Methyl Bromide ................... 106 Summary .................................................................................................................... 108 References .................................................................................................................. 109

I. Introduction Methyl bromide (bromomethane, MeBr) has been used widely since the 1940s as an effective preplant soil fumigant for controlling nematodes, plant pathogens, weeds and insects (UNEP 1995; Noling and Becker 1994). Fumigant use is vital for the economic viability of many crops, including strawberries, tomatoes, peppers, eggplants, tobacco, ornamentals, nursery stock, vines, and turf (Anderson and Lee-Bapty 1992; NAPIAP 1993; Ferguson and Padula 1994). The total global sales of MeBr were 7.16 × 107 kg in 1992, about 75% of which was used as a preplant soil fumigant (UNEP 1995). Methyl bromide is also widely used as a structural and commodity fumigant, as well as for quarantine or regulatory purposes (Anderson and Lee-Bapty 1992; NAPIAP 1993; Ferguson and Padula 1994; UNEP 1995). Its success as a fumigant is largely due to its wide spectrum of activity against pests at many stages of life, its ability to penetrate the fumigated zones, and the ease of application. Because of its high volatility, it leaves very low residue levels in the soil that may be phytotoxic or accumulated in plants, a problem commonly associated with the use of many other modern pesticides. A. Environmental Concerns In 1991, MeBr was identified as a potential ozone-depleting compound (Chakrabarti and Bell 1993). In 1992, on the Fourth Conference of the Parties to the Montreal Protocol, MeBr was officially added to the list of ozone-depleting chemicals, with its production suggested to be frozen at the 1991 level, effective from 1995. The inclusion of MeBr in the ozone-depleting chemicals list naturally brought this fumigant within the scope of the U.S. Environmental Protection Agency (EPA) Clean Air Act, which has an amendment that mandates discontinuation of any chemical with an ozone depletion potential (ODP) greater than 0.2 at the beginning of 2001. The ODP index for MeBr was determined to be 0.60–70 in 1992 (UNEP 1995); the ODP estimate was reduced to 0.4 in 1998. In March 1993, EPA announced that MeBr was scheduled for a phaseout in the United States by the year 2001 (USEPA 1993). This date was later changed to 2005 (USEPA 2000). During the past decade, there has been an increased research effort devoted to understanding the effects of halogenated gases emitted into the atmosphere on the depletion of the stratospheric ozone layer. According to the Ozone Assessment Synthesis Panel of the United Nations Environmental Programme (UNEP), the hole in the Antarctic ozone layer is due primarily to increases in chlorine- and bromine-containing chemicals in the atmosphere. Even though most of the ozone loss is due to chlorinated compounds (90%–95%; Watson et al. 1992), attention has been focused more recently on MeBr because strato-

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spheric bromine is believed to be 40 times more efficient than chlorine in breaking down ozone on a per atom basis (Wofsy et al. 1975). Although the largest effects from ozone-depleting gases have been observed in the southern hemisphere, there are indications that atmospheric ozone is also decreasing in the northern hemisphere. There is a great deal of uncertainty in estimates of the global MeBr budget. In the early 1990s, the ocean was viewed as a net source of MeBr. More recent global balances account for larger sinks than sources (Yvon-Lewis and Butler 1997), with the ocean acting as a net sink of MeBr, the magnitude of which is being refined (Lobert et al. 1997; King et al. 2000). Soil fumigation is thought to contribute 32 Gg yr−1 (1 Gg is equivalent to 1000 metric tons) of MeBr to the atmosphere, or ⬃20% of the total MeBr source (Yvon-Lewis and Butler 1997). The oceans represent the largest known source of atmospheric MeBr, followed by fumigation (Butler 2000). Other natural sources of atmospheric MeBr include biomass burning and production by plants, salt marshes, and fungi (Butler 2000). In recent global budgets, only 60% of the MeBr sinks were accounted for by the quantified source terms, and the “missing source” outweighed all other sources in the budget (Butler 2000). Agricultural use of MeBr, including soil fumigation, may be responsible for 3%–10% of stratospheric ozone depletion (USDA 2001). The relative significance of each global source of MeBr, including that from agricultural uses, needs to be better quantified to assist in developing rational national and worldwide policy. B. Economic Concerns An economic assessment conducted by the U.S. Department of Agriculture (USDA) indicated that the phase-out of MeBr as a fumigant will have a substantial impact on many commodities because current alternative control practices are either less effective or more expensive than MeBr (NAPIAP 1993; Ferguson and Padula 1994). A recent estimate of the annual economic loss to U.S. producers and consumers resulting from a ban of agricultural uses of MeBr is $500 million (Carpenter et al. 2000). In addition, currently a single chemical alternative that can completely replace MeBr does not exist (Anderson and Lee-Bapty 1992; Duafala 1996). Under these circumstances, MeBr has become the topic of many heated discussions, and the “methyl bromide issue” has received widespread attention (Anonymous 1994; Noling and Becker 1994; Taylor 1994; Sauvegrain 1995; Butler 1995, 1996; Duafala 1996; Thoms 1996). The many unanswered questions have also stimulated extensive research on the environmental fate of MeBr, particularly on estimating the relative contribution from each source, obtaining accurate estimates for volatilization losses from fumigated fields, understanding the processes and factors that affect the volatilization, and identifying and developing emission reduction techniques. The USDA National Agricultural Pesticide Impact Assessment Program conducted one of the first assessments of the economic impact of eliminating MeBr (NAPIAP 1993). This assessment determined that there would be a substantial

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adverse impact on the agricultural community and that this would be most strongly noticed in two states, California and Florida, the primary users of MeBr. It was estimated that a MeBr phase-out for preplant soil fumigation would cause $1.5 billion dollars in annual lost production in the U.S. This estimate ignored postharvest, nonquarantine uses, and quarantine treatments of imports and other future economic aspects such as lost jobs, markets, etc. The report predicted that the major crop losses would occur with tomatoes ($350M), ornamentals ($170M), tobacco ($130M), peppers ($130M), strawberries ($110M), and forest seedlings ($35M). More recently, the National Center for Food and Agricultural Policy (NCFAP) conducted an assessment of the economic implications of the methyl bromide ban (Carpenter et al. 2000). The NCFAP estimates a smaller economic loss of $479M to producers and consumers with the ban of preplant uses of methyl bromide. These losses are due to decreases in yield with use of alternative pest control strategies, increased production costs, and changes in the marketing window in response to supply and demand. The NCFAP estimates that losses of $235M may occur in annual crops (such as tomatoes, strawberries, and peppers), $143M in perennials (orchards and grapes), and $101M in nurseries and ornamentals. As research on methyl bromide alternatives continues to progress and regulatory issues surrounding soil fumigation take effect, the economic impact of the MeBr ban will become more clearly defined. The purpose of this review is to summarize studies on the transformation and transport processes of MeBr in soil, the interactions of these processes, and their effect on volatilization of MeBr into the atmosphere. Special emphasis is given to recent field, laboratory, and modeling studies that have been conducted for determining MeBr volatilization losses under various conditions and for identifying approaches to minimize these losses. Such a review has not been written, although a number of reviews have appeared on other aspects of MeBr use, such as the toxicological effects of MeBr (Alexeeff and Kilgore 1983; Yang et al. 1995), efficiency and phytotoxicity (Maw and Kempton 1973), and application and movement of fumigants in general (Goring 1962; Hemwall 1962; Hoffmann and Malkomes 1978; Siebering and Leistra 1978; Munnecke and Van Gundy 1979; Lembright 1990). A purpose of this review is to summarize relevant information concerning MeBr to provide a reference source to decision makers as well as to scientists in related industrial and academic areas. It also contains detailed information on measurements of volatilization on various scales (e.g., field, miniplots, and laboratory columns) that may be applicable for studies of other volatile pesticides or organic compounds in general. Reported studies on the use of MeBr in fumigating stored products, commodities, and structures are not included. Most of the MeBr studies completed in the early years are not cited because these studies dealt with issues of efficacy, rather than environmental fate. As the use of MeBr in soil fumigation has spanned more than five decades, there is a wealth of research on this topic, and many interesting research investigations have been conducted. Failure to include an

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article in this review should not be construed as a negative implication and is the result of our attempt to prepare a relatively concise review article. It is our belief that such a review is not only timely for MeBr itself but also offers lessons that will be valuable in finding and developing successful alternative fumigants to replace MeBr. Many of the factors affecting the phase-out of MeBr also arise for other chemical fumigants, most notably their potential to volatilize and contaminate the atmosphere. Fortunately, the emission control strategies discussed herein will also assist in reducing emissions of other fumigant compounds (e.g., 1,3-dichloropropene, chloropicrin, methyl isothiocyanate) and emerging potential chemical alternatives (e.g., methyl iodide, propargyl bromide).

II. Chemical and Physical Properties of Methyl Bromide Some of the basic physical and chemical properties of MeBr are listed in Table 1. Because of its high vapor pressure, MeBr can readily penetrate many matrices and is extremely difficult to contain even in the laboratory. As MeBr is colorless and odorless at room temperature, even at potentially toxic concentrations, severe exposure can occur unknowingly (Yang et al. 1995). In commercial formulations of MeBr, various percentages (0.5%–33%) of chloropicrin are added as a warning agent to protect workers and residents during and immediately after MeBr applications and to assist in protecting plants from disease. However, it Table 1. Selected physical properties of methyl bromide. Property Synonyms Trade name

Molecular weight, g mole−1 Vapor density, g L−1 Dipole moment, debye Liquid density, g cm−3 20 °C 25 °C Solubility, mg L−1 20 °C 20 °C 25 °C aReid

Value Bromomethane Brom-O-Gas, Meth-O-Gas, Terr-O-Gas 94.94a 3.974e 1.8a 1.676c 1.737a 16,000d 17,500e 13,400b

Property CAS Registry Number Freezing point, °C

Boiling point (at 1.0 atm), °C Koc, cm3/(g %Foc) Log(Kow) Vapor pressure, mmHg 20 °C Henry’s law constant 20 °C 21 °C

Value 74-83-9 −93.7c 3.56e 22b 1.19f 1395 ± 19a,b,d,e 0.23g 0.30 ± 0.02h

et al. (1987): properties handbook; bWauchope et al. (1992); cCRC (1996); dGoring (1962); Index (1996); fSchwartzenbach et al. (1993); gEstimated from mean vapor pressure and solubility data; hGan and Yates (1996). eMerck

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should be noted that because the vapor pressure of MeBr is many times that of chloropicrin, the safety of using low ratios of chloropicrin in the mixture as a warning agent is questionable (van Assche 1971). Methyl bromide is considered to be acutely toxic, with an 8-hr time-weight-averaged limit for human exposure in air of only 5 ppm (ACGIH 1988). Acute toxicity to workers following exposure to its vapor has been a major concern in the many years of MeBr use and is one of the reasons for some early modifications of its application method (e.g., use of surface tarp, mixing with chloropicrin, and use of buffer zones). Fatalities and injuries resulting from exposure to MeBr have been reported, but most incidents are related to structural fumigations rather than soil fumigations (Yang et al. 1995).

III. Methyl Bromide Use as a Soil Fumigant A. History and Scope of Use The insecticidal activity of MeBr was first reported by Le Goupil (1932) and has been subsequently shown on a wide variety of pests including insects, nematodes, rodents, bacteria, viruses, fungi, mites, and weeds (Alexeeff and Kilgore 1983). The early application of MeBr as a fumigant was mainly on stored products (Thompson 1966), but it has been largely replaced by phosphine due to issues of safety and ease of use (Taylor 1994). The concept of fumigating soil with MeBr was introduced later, around the beginning of the 1960s (Hague and Sood 1963), and it has since experienced a steady increase in production and sales. For instance, in 1960, the total MeBr production in the U.S. was 5.7 × 106 kg (Alexeff and Kilgore 1983), which increased to 2.93 × 107 kg by 1992 (UNEP 1995). The total worldwide sales of MeBr increased steadily from 4.56 × 107 kg in 1984 to 7.16 × 107 in 1992 (UNEP 1995). Of the 7.16 × 107 kg of MeBr sold in 1992, 75% was used as a soil fumigant and 15% as a commodity and structure fumigant (UNEP 1995). The U.S., Europe, and Asia are the largest users, consuming 41%, 26%, and 24%, respectively, of the total MeBr in 1992 (UNEP 1995). On a global basis (excluding the U.S.), the predominate use of MeBr as a soil fumigant is in the production of tomatoes (22%), strawberries (14%), and curcurbits (11%) (UNEP 1995). Total methyl bromide preplant use in the U.S. was estimated at ⬃1.6 × 107 kg for the period 1992–1996, most of which was used in the production of tomatoes (30%), strawberries (19%), peppers (14%), grapes (7%), and nursery stock (6%) (Carpenter et al. 2000). Methyl bromide is currently scheduled to be incrementally phased out. In the U.S. and other developed countries that are parties to the Montreal Protocol, MeBr production and importation will be reduced from 1991 baseline levels by 25% in 1999, 50% in 2001, 70% in 2003, and 100% in 2005. Preshipment and quarantine uses are currently exempt from MeBr phase-out. The MeBr phaseout schedule for nonindustrialized countries indicates a 100% reduction by 2015. The MeBr phase-out has produced a wealth of research for chemical and nonchemical alternatives to MeBr fumigation. Research activities are currently

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underway investigating alternative chemicals such as 1,3-dichloropropene (Shaw and Larson 1999; Riegel et al. 2000; Nelson et al. 2001), methyl isothiocyanate (Borek et al. 1997; Shaw and Larson 1999), chloropicrin (Massicotte et al. 1998; Hutchinson et al. 2000), methyl iodide (Eayre et al. 2000; Hutchinson et al. 2000; Zhang et al. 1998), and propargyl bromide (Yates and Gan 1998; Papiernik et al. 2000; Ma et al. 2001), to mention a few. Several nonchemical methods have also been proposed, such as solarization/soil heating (Katan 1992; Porter et al. 1991; Yu¨cel 1995; Stapleton 2000). Brassica roots have been shown to produce isothiocyanates and other allelochemicals, and incorporation of Brassica cover crops may have some efficacy against nematodes and pathogenic fungi (Angus et al. 1994; Gardiner et al. 1999; Walker and Morey 1999). B. Application Methods As a preplant soil fumigant, MeBr is applied by distinctively different methods within the U.S. and outside the U.S. (UNEP 1995). In the U.S., mechanized injection into subsoils is the predominant application method, whereas in many other countries, tarped-surface applications are used. In 1992, about 54% of the overall soil fumigation was conducted with the tarped-subsoil injection method and 40% with the tarped-surface application method (UNEP 1995). The application rate for outdoor mechanical injection is about 24–48 g m−2, whereas a higher rate, normally 45–135 g m−2, is used for the tarped-surface application. A variety of mixtures of MeBr and chloropicrin, with chloropicrin content varying from 0.5% to 37%, are normally used in outdoor fumigation. The formulation used for tarped-surface application in glasshouses is usually a mixture with 2% chloropicrin (UNEP 1995). In outdoor mechanized application, MeBr from pressurized cylinders is injected via tractor-driven shanks or chisels at depths of 20–90 cm into the soil. For shallow injections, the chisels are spaced about 25 cm apart, and the fumigated area is covered with plastic films (almost exclusively low-density or highdensity polyethylene sheets) immediately after injection. The practice of tarping is necessary for controlling weeds near the surface (Lembright 1990), although it also has implications for reducing MeBr emissions, thereby providing better protection for field workers and nearby residents. After liquid MeBr is introduced into the soil, it absorbs heat from the environment around the injection point and rapidly vaporizes. During the initial moments, the distribution pattern of MeBr in the field can be perceived as narrow horizontal cylinders with high concentration of MeBr gas. Driven by the large concentration gradient, MeBr vapor diffuses outward from the lines of injection, and the cylinders of fumigated soil overlap and form uniform concentration profiles in the horizontal plane (Lembright 1990). Plastic covers are removed 2–7 days after fumigation, so planting can begin shortly after the treatment. Untarped deep injection (60–80 cm) is mainly used for eradicating pests for deep-rooted plants such as in orchards and vineyards (UNEP 1995). The spacing between injection lines for deep applications is generally around 1.5–2.0 m.

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In many other countries, manual application of MeBr underneath a presealed plastic tarp is the principal method and essentially the exclusive method for fumigating soil in greenhouses (Maw and Kempton 1973; UNEP 1995). This method is the so-called hot-gas application, in which liquid MeBr is vaporized in a heat exchanger and hot MeBr vapor is delivered through tubing lines with side openings into the space between the plastic tarp and the soil surface (Maw and Kempton 1973; UNEP 1995). Another technique is using liquid MeBr directly instead of the hot gas. Small cans containing MeBr are placed in the surface soil pretarped with plastics, and a special opener is used to release MeBr vapors under the tarp. Following tarped-surface application, MeBr penetrates into deeper soil layers by gas diffusion. The high permeability of MeBr through the traditional polyethylene film was found to be a problem that caused inadequate pest control as well as increased air pollution in the late 1970s. Since then, new materials with reduced permeability have been tested and used for soil fumigation (Kolbezen and Abu-El-Haj 1977; Munnecke et al. 1977; Van Wambeke 1983; de Heer et al. 1983; Daponte 1995; Gamliel et al. 1997). As a result, relatively impermeable plastics have been used on a small scale in some glasshouse fumigations (de Heer et al. 1983; Gamliel et al. 1997).

IV. Ozone Depletion and Methyl Bromide There are both technical and legislative limitations to MeBr use as a soil fumigant. Methyl bromide is toxic to humans, and in most countries its application is restricted to trained applicators, and buffer zones also are required between fumigation sites and nearby residential areas. Fatalities and injuries resulting from MeBr use have been recorded, particularly during fumigation of structures (Yang et al. 1995). Bromide ion residues produced from MeBr in soil have been shown to cause damage to certain sensitive plants such as carnations and wheat (Brown and Jenkinson 1971; Kempton and Maw 1974). In addition, plant accumulation of Br− at excessive levels is considered hazardous for human consumption, and there are maximum limits for Br− concentrations in various fruits and vegetables (Kempton and Maw 1972, 1973; Masui et al. 1978, 1979; Helweg and Rasmussen 1982). These negative effects have influenced some of the application techniques or regulatory restrictions on MeBr use. For instance, leaching fumigated soil to reduce Br− accumulation in plants has been used in glasshouse fumigation in some European countries (Vanachter et al. 1981a,b; Lear et al. 1983; de Heer et al. 1986; Van Wambeke et al. 1988). However, the primary reason that future use of MeBr in agriculture will be eliminated is its involvement in stratospheric ozone depletion. A. Reactions with Ozone It has been known for some time that bromine can catalytically destroy stratospheric ozone (Wofsy et al. 1975; Yung et al. 1980; McElroy et al. 1986; Salawitch et al. 1988; Anderson et al. 1989; Prather and Watson 1990). Reactions

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involving bromine are believed to be responsible for 20%–25% of the Antarctic ‘ozone hole’ that develops each austral spring (Anderson et al. 1989), which implies that a bromine atom is approximately 40 times more efficient than a chlorine atom in destroying ozone (Wofsy et al. 1975; Salawitch et al. 1988; Solomon et al. 1992). Methyl bromide is unique because it is a significant source of bromine to the stratosphere (Wofsy et al. 1975; Yung et al. 1980; Penkett et al. 1985; Cicerone et al. 1988; Schauffler et al. 1993). However, the case for restricting its use is not clear-cut. Unlike the chlorofluorocarbons (CFCs), atmospheric MeBr is not entirely contributed by human activities. Atmospheric MeBr has abundant natural and anthropogenic sources. Also, its sinks result not only from reactions in the atmosphere but also from interaction with the oceans and land. Thus, estimating the contribution of MeBr fumigation (currently ⬃80% of the entire anthropogenic source) to the depletion of stratospheric ozone is much more complex than it is for other regulated halogenated compounds. To justify the pending suspension of MeBr use in agriculture, it should be established that the known sources of atmospheric MeBr surpass the sinks, and the surplus is contributed by anthropogenic emissions. However, current estimates of global MeBr are out of balance, with sinks exceeding sources by a wide margin (Yvon-Lewis and Butler 1997). The total atmospheric burden of MeBr is believed to be around 145 Gg yr−1 (100–194 Gg yr−1), and the concentration about 10 parts per trillion by volume (pptv), increasing at 0.1–0.3 pptv yr−1 (Khalil et al. 1993; Singh and Kanakidou 1993). The sinks currently thought to remove MeBr from the atmosphere include reactions with OH radicals in the atmosphere (accounting for ⬃86 Gg y−1 MeBr), removal by oceans (⬃77 Gg yr−1), degradation in soil (42 Gg yr−1), and uptake and degradation by plants. The relative strength of each of these sinks is not well quantified. The estimated lifetime of atmospheric MeBr is 0.7 yr (range, 0.4–0.9 yr), with a calculated ozone depletion potential (ODP) of 0.4 (range, 0.2–0.5), according to the World Meteorological Organization 1998 Scientific Assessment of Ozone Depletion (WMO 1999). The known sources of atmospheric MeBr include oceanic emissions, biomass burning, automobile emissions from leaded gasoline, and fumigation. Together, these emissions combine to produce 122 Gg yr−1 of MeBr (range, 43–244 Gg yr−1) (WMO 1999). The 1998 Scientific Assessment of Ozone Depletion estimated oceanic MeBr emissions to be 60 Gg yr−1, with the ocean acting as a net MeBr sink of −21 Gg yr−1 (WMO 1999). Recent research has indicated that the magnitude of the oceanic sink may be −11 to −20 Gg yr−1 (King et al. 2000). Biomass burning (Mano¨ and Andreae 1994) is another significant natural source of atmospheric MeBr, and its contribution is poorly quantified. Global emission of MeBr from biomass burning is estimated to be 20 Gg yr−1 (range, 10–40 Gg yr−1) (WMO 1999). It has also been demonstrated that automobile exhaust from the combustion of leaded gasoline, which contains bromine compounds, can include measurable amounts of MeBr (Harsch and Rasmussen 1977). Emissions from this source could range from 0 to 5 Gg yr−1 (WMO 1999). Additional

54


potential MeBr sources that have recently been identified include production by plants (Gan et al. 1998a), salt marshes (Rhew et al. 2000), and fungi (Butler 2000). Salt marshes may be a globally important source of MeBr (contributing 7–29 Gg yr−1) (Rhew et al. 2000) and production of MeBr has been observed for a variety of plants (Gan et al. 1998a); therefore, plant sources may account for a large proportion of the “missing source” in current MeBr budgets. Some anthropogenic emissions, such as fumigation of structures, perishables, and durables, are relatively well quantified, because nearly 100% of the applied MeBr is vented into the air during these fumigation processes. The use of MeBr for these fumigations accounts for about 15% of the total production. Trapping and/or decomposing MeBr during structural fumigation can drastically decrease atmospheric emissions of MeBr during these operations. Approximately 85% of the industrially produced MeBr is used as a soil fumigant, equivalent to ⬃65 Gg yr−1 in 1996. The actual discharge of MeBr from fumigated fields into the air is largely determined by the proportion of the applied MeBr that is emitted from the treated soil, which can be reduced through management practices (Wang et al. 1997a; Yates et al. 1998; Gan et al. 1998d). New measurements of the sources and sinks of MeBr are still being actively obtained, as evidenced by many recent reported studies. It is a fact that the relative contribution of MeBr used in fumigation practices is far from well quantified. Despite this, being a significant controllable source, the agricultural use of MeBr becomes a natural target for elimination. It is also assumed that, due to a short atmospheric lifetime of less than 1 yr, the effect of cessation of anthropogenic MeBr emissions on the restoration of stratospheric ozone will be nearly immediate. In comparison, all the chlorofluorocarbons (CFCs) have extremely long lifetimes, and with a complete elimination of emissions of these compounds, it may take many years, or even centuries, to reduce the atmospheric burden of the CFCs to an insignificant level (Butler 1995).

V. Sampling and Analysis of Methyl Bromide in the Air In the course of monitoring MeBr in workplace, field, and ambient atmospheres, sampling and analytical methods of different sensitivities and complexities have been developed. Depending on the sampling device that is used for collecting air samples, MeBr can be either in a contained atmosphere (such as canisters) or adsorbed on a solid adsorbent (such as activated carbon or a porous polymer) before analysis. For the past two decades, quantitation of MeBr has used gas chromatography (GC) exclusively, and electron-capture detectors (ECD) are usually selected over the other types of detectors because of their high sensitivity to halogenated compounds (Scudamore 1988), although very high sensitivity is also found with photoionization detectors (PID) (Dumas and Bond 1985). The main reported sampling and analytical methods of analyzing atmospheric MeBr are summarized in Table 2.

Table 2. Sampling and analytical methods for atmospheric methyl bromide. Sampling method

Tenax-GC Charcoal tube Charcoal tube Charcoal tube Charcoal tube Cold traps Silica capillary Teflon tubing

GC detector

Throughput

Sensitivity

Reference

Cryogenic Cryogenic Cryogenic Direct injection Direct injection

FID ECD FID FID PID

Low Low Low Intermediate Intermediate

Intermediate High Intermediate Low Very high

Jayanty (1989) Gholson et al. (1990) Yagi et al. (1993, 1995) Kolbezen et al. (1974) Dumas and Bond (1985)

Thermodesorptioncryogenic Thermodesorption Solvent extraction Solvent extraction Headspace-GC Headspace-GC

MS

Low

High

Krost et al. (1982)

High Intermediate Intermediate Intermediate Intermediate

Dumas (1982) Eller (1984) Lefevre et al. (1989) Woodrow et al. (1988) Gan et al. (1995b)

Very high Very high

Kallio and Shibamoto (1988) Kerwin et al. (1996)

FID Low FID High ECD (FID) Intermediate ECD Intermediate ECD Very high

Thermo-evaporation ECD Thermo-evaporation ECD

Low Low

Methyl Bromide

Containers Canister Canister Canister Syringe Syringe Adsorbents Polymeric beads

Sample preparation

FID: flame ionization detector; ECD: electron-capture detector; PID: photoionization detector; MS: mass spectrometry.

55

56


A. Container Methods Steel canisters were used for sampling volatile toxic chemicals in air, such as MeBr, by Jayanty (1989) and Gholson et al. (1990), and good stability and sensitivity were achieved for all the selected analytes. Cryogenic preconcentration was required before the delivery of samples into the GC column. Yagi et al. (1993, 1995) used 500-mL canisters for sampling MeBr to obtain flux measurements under field conditions. Sampling with canisters is labor intensive because the container must be evacuated before sampling, and the contents must be cryogenically concentrated before injection, which limits the number of samples that can be collected and analyzed. Sampling with canisters is therefore not suitable for extensive sampling as needed in volatilization flux measurement under field conditions, although the sensitivity could be very high if a proper detector is used. Using canisters is also not compatible with active (flowthrough) chambers that are used for continuous sampling of the atmosphere. Another container method involves collecting an air sample using a gastight syringe and injecting the contents directly into a GC. In a study of the transport of MeBr in soil after fumigation, Kolbezen et al. (1974) used glass syringes to take and temporarily store soil air samples. The plungers were coated with a film of Triton X-100 to eliminate rapid leakage, and the needle was embedded in a MeBr-impervious sponge. Loss of MeBr was determined to be insignificant within 6 hr, but 5%–7% was lost after 22 hr. The analysis was made by direct injection of the air sample in the syringe into a GC. This method has also been employed in small-scale laboratory experiments (Gan et al. 1998a). B. Adsorbent Methods The most commonly used method for sampling atmospheric MeBr is pumping a relatively large volume of air through one or a series of adsorbent tubes. Methyl bromide in the air stream is trapped in the sample tube containing the solid adsorbent due to its high affinity to the adsorbent. Two types of adsorbent material have been recorded for use with MeBr: activated carbon (charcoal) (Eller 1984; Woodrow et al. 1988; Lefevre et al. 1989; Gan et al. 1995a,b; Majewski et al. 1995; Yates et al. 1996a–c, 1997; Wang et al. 1997a) and porous polymeric adsorbent such as Tenax GC (Brown and Purnell 1979; Dumas 1982; Dumas and Bond 1985; Krost et al. 1982). Activated carbon or charcoal tubes are low in cost (about $1 each), can accommodate large sample volumes, and need minimum preparation before sampling. A typical charcoal tube consists of two adsorption beds: a primary bed (A) and a backup bed (B) in a sealed glass tube. The charcoal can be derived from either coconut or petroleum. Polyurethane spacers are used to separate the two adsorption beds, and a plug of glass wool is usually placed in front of the primary bed to hold the charcoal in the sample tube. Before use, a tube is broken at both ends and then connected to a vacuum source to draw the air to be sampled into the tube. Depending on the sampled volume, air flow rate, and

Methyl Bromide

57

MeBr concentration, multiple tubes connected in series may be required to eliminate loss through breakthrough (Gan et al. 1995a). The number of tubes should be increased when a high flow rate or a long sampling interval is used. Gan et al. (1995a) found, for a single 600-mg coconut charcoal tube at a flow rate of 100 ml min−1, that a sampling interval of ≤2 hr resulted in no breakthrough loss for spiked MeBr masses up to 3.9 mg. Methyl bromide adsorbed in charcoal tubes may be analyzed by two different methods: solvent extraction followed by quantitation in the solvent phase and the so-called headspace-GC method. In solvent extraction, charcoal is transferred into a vial, a known amount of extracting solvent such as carbon disulfide (CS2) is added into the vial, and the vial is sealed (Eller 1984; Lefevre et al. 1989). After the solvent–charcoal mixture is mechanically shaken, an aliquot of the solvent is injected into a GC. This method has the drawbacks of manual sample preparation and presence of other compounds in the final sample solution that may elute with or interfere with MeBr during chromatography (Gan et al. 1995b). This method allows for multiple injections of each sample so that multiple analytes may be measured using different methods or detectors. An alternative method is the headspace-GC method. In headspace-GC analysis, the charcoal is equilibrated with an organic solvent in a closed headspace vial at an elevated temperature for a given period of time, and an aliquot of the headspace containing the analyte is then introduced into the GC column for detection. Benzyl alcohol is often used as the solvent due to its high boiling point (210 °C) (Woodrow et al. 1988; Gan et al. 1995b). When the vial size, solvent volume, and equilibrating temperature and time are fixed, automated headspace injectors give high reproducibility and sample throughput. Gan et al. (1995b) found that the equilibration temperature and time in the headspace autosampler, the size of headspace vials, and the amount of solvent all had an effect on the signal output for a given sample. The sensitivity of analysis can thus be maximized by choosing an optimal combination of these parameters. For instance, to analyze a sample tube containing 600 mg coconut charcoal, the best conditions were determined to be 9-mL headspace vials, 1.0 mL benzyl alcohol, 110 °C equilibration temperature, and 15 min equilibration time (Gan et al. 1995b). Using this method, analysis of a MeBr-containing sample tube takes only 3–4 min, and as many as 300 samples can be analyzed with 24 hr. This method is appropriate when a large number of samples is required, such as when analyzing samples from large-scale field studies measuring MeBr volatilization (Yates et al. 1996b,c, 1997; Wang et al. 1997a). This method has the disadvantage of being destructive, where each charcoal sample can be analyzed with only a single injection. Many kinds of porous polymer adsorbent materials have been used for collecting volatile compounds in the air, including the Chromosorb series, the Porapak series, Ambersorb XE-340, and others. The most popular adsorbent, however, is Tenax-GC, which is a polymer of 2,6-diphenyl-p-phenylene oxide. Brown and Purnell (1979) estimated the safe sampling volume for MeBr to be

58


0.14 L for sample tubes packed with 0.13 g Tenax-GC. When coupled with a cryofocusing technique, the whole sample can be introduced into the GC column following thermal desorption, which greatly enhances the sensitivity. Detection limits of 500 pg L−1 (Krost et al. 1982) and 35 ng (Dumas and Bond 1985) were reported when this method was used. Compared with charcoal tubes, polymer samplers need to be conditioned before sampling, the safe sampling volume is smaller, the cost is higher, and each analysis takes a longer time. C. Other Methods Other than the container method and the adsorbent method, cryogenic concentration in a cold trap has also been used for collecting MeBr (Kallio and Shibamoto 1988; Kerwin et al. 1996). The cold traps include mixtures of dry ice–acetone, liquid nitrogen, and dry ice–2-propanol. The weakness of this technique is the long time and many steps involved in handling one sample, but it is useful when sample throughput is not a factor and very low detection limits are sought. When extremely high sensitivity is pursued, such as in the case of monitoring MeBr in ambient air, a technique called O2-doping could be useful (Grimsrud and Miller 1978; Kerwin et al. 1996). Grimsrud and Miller (1978) first reported that addition of a fraction of O2 in the carrier gas drastically increased the sensitivity of ECD detection of halogenated methanes including MeBr. When 3%–5% of O2 was added to the carrier gas, signal response for MeBr was enhanced about two orders of magnitude. Using cryogenic concentration and O2-doping, Kerwin et al. (1996) reported a detection limit as low as 0.23 pmol or 22 pg.

VI. Processes Affecting Environmental Fate of Methyl Bromide The efficacy of MeBr fumigation and the extent of volatilization from the soil surface depend on the concentration of MeBr in the soil and the time for which those concentrations are maintained. After MeBr is applied to soil, it is subject to numerous interactions with the soil that may affect its concentration. These processes include transformation and phase partitioning. Transformation also results in the production of Br−, which is considered harmful under certain circumstances. Factors relating to soil and meteorological conditions and the method of application affect MeBr volatilization losses by acting on the transformation and phase-partitioning processes. An analysis of these processes and factors is critical for understanding the fate of MeBr as a soil fumigant, as well as for developing strategies to minimize its volatilization losses. A. Transformation Transformation or degradation of MeBr is an irreversible process that depletes MeBr from the soil–water–air system before it reaches the soil surface and volatilizes into the air. Extremely rapid transformation may deplete MeBr concentrations so quickly that efficacy is compromised. The actual transformation

Methyl Bromide

59

of MeBr in an agricultural soil is the sum of its hydrolysis in water, reactions with soil constituents, and decomposition by soil microorganisms. 1. Water Degradation of MeBr in water is important because it contributes to MeBr degradation in moist soil as well as to its fate in the overall environment. Based on its chemical structure, MeBr is an electrophile, and −Br is reactive as a leaving group and may participate in various nucleophilic substitution reactions (SN1 and SN2 types) in the environment. Water is a weak nucleophile, and therefore hydrolysis of MeBr in water is anticipated: CH3Br + H2O → CH3OH + Br− + H + −

CH3Br + OH → CH3OH + Br

−

Reaction I Reaction II

(1) (2)

The reaction rate constants for reactions I and II are approximately 5 × 10−9 and 10−4 M−1s−1, respectively (Schwarzenbach et al. 1993). In pure water where the OH− concentration is extremely low, reaction I dominates, and the calculated pseudo-first-order half-dissipation time (t1/2) of MeBr should be around 30 d. Mabey and Mill (1978) and Papiernik et al. (2000) report a t1/2 of 20 d, Arvieu (1983) reported a t1/2 of 46 d for MeBr in water at 20 °C, and Gentile et al. (1989) reported t1/2 of 36–50 d in well waters at 18 °C. The relatively slow hydrolysis of MeBr in water was also noted by Herzel and Schmidt (1984). In an attempt to correlate MeBr hydrolysis and pH, Gentile et al. (1992) measured MeBr degradation in buffer solutions with pH 3.0–8.0 and found MeBr hydrolysis rates generally increased with increasing pH. However, in their experiments, they used buffer solutions composed of phosphate and citrate, and apparently nucleophiles other than OH− caused the enhanced hydrolysis in solutions with elevated pH. From the rate constant of reaction II, the MeBr hydrolysis rate should not increase significantly when pH is changed from 7 to 10. In waters that are rich in nucleophiles, such as the supernatant of a salt marsh containing sulfide, MeBr degradation may be accelerated (Oremland et al. 1994a). The reaction produces methanethiol: CH3Br + HS − → CH3SH + Br−

(3)

and further reaction with MeBr produces dimethylsulfide CH3SH + CH3Br → (CH3)2S + Br− + H +

(4)

MeBr was observed to degrade rapidly in anaerobic salt marsh slurries containing sulfide, with a reported transformation half-life ⲏ1 d. Production of methanethiol in slurries doped with sulfide exhibited very rapid reaction, with a MeBr half-life ⬃1 hr (Oremland et al. 1994a). Accelerated transformation by MeBr in aqueous solution containing other nucleophiles (for example, aniline) has also been reported. Reaction with aniline in aqueous solution with a molar ratio of aniline : MeBr of 10 : 1 formed N-methylaniline and N,N-dimethylaniline with a MeBr transformation half-life of 2.9 d (Gan and Yates 1996). The mechanism of photoinduced hydrolysis of MeBr in water was first re-

60


ported by Castro and Belser (1981). When a pen-ray UV lamp emitting UV at 254 nm was used to irradiate MeBr–water solution in a 4-L closed flask, MeBr was gradually converted to methanol and Br−. The following mechanism was proposed by these authors: CH3Br + hν → (CH3Br)* + H2O → CH3OH + H + + Br−

(5)

Photohydrolysis caused faster dissipation of MeBr in sunlight under sealed conditions (Castro and Belser 1981) and under UV irradiation (Gentile et al. 1989). The significance of reactions with nucleophiles and UV in MeBr transformation in soil water, however, has not been investigated. Hydrolysis of MeBr in aqueous solutions may bear limited significance in determining its fate as a water contaminant. The loss of MeBr in stirred and ventilated waters, or water that had a high surface-to-volume ratio, was found to be very rapid due to volatilization (Gentile et al. 1992). In a study following MeBr kinetics in surface water, Wegman et al. (1981) found that the average half-life for MeBr in surface water at a water temperature of 11 °C was only 6.6 hr. Over its many years of use, contamination of water sources with the parent MeBr has never been a topic of concern. 2. Soil Hydrolysis in water is not the only pathway, and in many cases not even an important pathway, that causes MeBr degradation in soil; this is evidenced by shorter t1/2 values obtained in soil degradation studies, particularly with organic matter-rich soils, as opposed to the t1/2 in water. Two other pathways, i.e., reaction with soil organic matter and microbial degradation, have been identified as contributing to MeBr degradation in soil. Soil organic matter (OM) contains nucleophilic sites such as −NH2, −NH, −OH, and −SH functional groups. Methyl bromide may react with these groups via nucleophilic substitution, as in the hydrolysis reactions: CH3Br + OM − NH2 → OM − NH − CH3 + Br− + H + −

CH3Br + OM − SH → OM − S − CH3 + Br + H

+

(6) (7)

As a result of these reactions, soil organic matter is methylated, and inorganic bromide ion is released. The reaction of MeBr with soil organic matter is supported by the general observation of the close dependence of MeBr degradation on soil organic matter content: in organic matter-rich soils, degradation is consistently more rapid than in organic matter-poor soils (Brown and Jenkinson 1971; Brown and Rolston 1980; Arvieu 1983; Arvieu and Cuany 1985; Gan et al. 1994). For a soil containing 2.81% organic matter, 63 ppm of Br− was generated after exposure to 500 ppm MeBr in closed flasks for 24 hr whereas only 25 ppm of Br− was produced in a soil with 0.93% organic matter (Brown and Jenkinson 1971). Arvieu (1983) studied the rate of MeBr degradation rate in eight soils and found that t1/2 decreased with increasing organic matter content. In a soil with 0.23% organic matter, the t1/2 was 49 d, but in a soil with 5.11% organic matter, the t1/2 was shortened to only 3.6 d. After measuring MeBr degra-

Methyl Bromide

61

dation rates in selected soils, Gan et al. (1994) found that the MeBr degradation rate (based on Br− production) and soil organic matter content were highly correlated, and the measured correlation coefficient (r2) ranged from 0.95 to 1.00. Papiernik et al. (2000) also reported an increase in MeBr degradation rate with increasing soil organic matter content. Spiking soil samples with 14C-labeled MeBr resulted in the formation of nonextractable (bound) residues of MeBr, which increased as extractable MeBr decreased. These bound residues represented transformed (not sorbed) MeBr, as evidenced by the release of equimolar amounts of Br− for each mole of MeBr lost (Papiernik et al. 2000). The reliance of MeBr degradation on soil organic matter content also has implications for MeBr degradation in subsurface layers. Because soil organic matter content normally decreases with increasing depth, MeBr degradation may be much slower in the deep layers, and the overall persistence could be much longer than suggested by the degradation data generated from surface soils. This idea was verified in laboratory incubation studies using soils collected from 0 to 300 cm (Gan et al. 1994). In a Greenfield sandy loam, the t1/2 for MeBr degradation was about 8 d for the 0–30 cm layer, but increased gradually with depth to a t1/2 of 21 d for the 270–300 cm layer. This decrease closely corresponded to the decrease in soil organic matter content (Gan et al. 1994). Biodegradation of MeBr has been documented for isolated bacteria, including the nitrifying bacteria Nitrosomonas europaea, Nitrosolobus multiformis, and Nitrosococcus oceanus (Rasche et al. 1990; Hyman and Wood 1984) and for Methylomonas methanica and Methylococcus capsulatus (Colby et al. 1975, 1977; Meyers 1980; Oremland et al. 1994b). The nitrifier-catalyzed degradation suggests the involvement of ammonia monooxygenase, whereas the consumption of MeBr by the methane-oxidizing bacteria indicates that methane monooxygenases are responsible. Increasing the activity of nitrifying bacteria may increase the rate of biodegradation of MeBr in soil (Ou et al. 1997). Other bacteria capable of degrading MeBr have been isolated from soil. Miller et al. (1997) isolated a facultative methylotroph that could use MeBr as a source of carbon and energy. Oremland et al. (1994b) demonstrated that, at high concentration, biodegradation of MeBr in methanotrophic soils was inhibited due to the toxicity of MeBr itself but became significant at concentrations lower than 1000 ppm. Shorter et al. (1995) suggested that microbial degradation of MeBr at low concentrations (ppb) in surface soils may be important in removing MeBr from the atmosphere, thus reducing its lifetime in the atmosphere and lowering its ozone depletion potential. They observed that MeBr removal from the headspace in closed systems containing soil was more rapid in live soils than in autoclaved soils, and the degradation rate decreased with the depth from which the soil was sampled, which corresponded to the methane oxidation activity of the soil. Hines et al. (1998) reported that at low atmospheric mixing ratios (5 pptv to 1 ppmv) rapid degradation of MeBr was effected by aerobic soil bacteria, resulting in half-lives on the order of minutes for a variety of soils. As the initial concentrations around the injection point are normally on the order of 104 ppmv (Kolbezen et al. 1974) and the MeBr-degrading bacteria are low in population in

62


normal agricultural soils, bacteria-mediated degradation may be insignificant under typical circumstances. However, studies have indicated that MeBr oxidation can occur in field-fumigated soil. High rates of 14C-MeBr oxidation to 14CO2 were observed in the first few days following soil fumigation where the MeBr concentration was >9.5 µg/g soil (Miller et al. 1997). This oxidation was inhibited by the addition of chloropicrin at concentrations >1.6 µg/g soil. 3. Enhancing Transformation Rates Fumigant degradation in soil varies with soil type and organic matter content (Leistra et al. 1991; Shorter et al. 1995; Gan et al. 1994). However, with reported half-lives of the order of days to months, indigenous soil degradation alone may not be sufficient to significantly affect the emission rate unless amendments are added to soil to enhance degradation. Various organic amendments have been used to increase the degradation rate of MeBr and other fumigants in soil (Gan et al. 1998b,c; Dungan et al. 2002). Gan et al. (1998b) used a biosolid–manure mix and a composted manure to increase the rate of MeBr degradation in soil. They found that the composted manure was more effective in increasing MeBr degradation compared to the biosolid–manure mix and that the degradation rate increased linearly with increasing concentrations of composted manure in the soil. MeBr degradation was nearly 2 times faster than in unamended soil when composted manure was added to soil at a ratio of 1 : 40 dry weight and was 10 times faster at a ratio of 1 : 8 (dry weight). Because similar degradation rates were observed in both sterilized and nonsterilized samples, the increase was attributed to enhanced abiotic reactions. The reactivity of MeBr toward nucleophiles was further demonstrated by its enhanced transformation by thiosulfate salts in solutions and soils. In the reaction, bromide is replaced by thiosulfate (S2O32−), forming a methyl–thiosulfate complex and bromide ion (Gan et al. 1998d). In ammonium thiosulfate (ATS)amended soil, the rate of MeBr degradation was dependent on the level of ATS. At room temperature, MeBr half-life was reduced from 5 d to less than 5 hr when ATS (2 µmole/g) was added to the soil containing MeBr (0.5 µmole/g) (Gan et al. 1998d). It was further shown that after the reaction, the acute toxicity of the aqueous solution substantially decreased (Wang et al. 2000). Thiosulfateinduced transformation occurred at similar rates in different soils, and increased with increasing temperature (Gan et al. 1998d). This reaction has also been used in the development of a simple, safe, and potentially cost-effective method to remediate activated carbon used as a chemical trap in commodity fumigation (Gan and Yates 1998; Yates and Gan 1999). The method will destroy methyl bromide that has been adsorbed to activated carbon. The method only requires that fumigation gases be pumped through a charcoal bed containing a thiosulfate solution, which could be easily accomplished at the fumigation site. 4. Br− Residues in Soil and Plants following Methyl Bromide Fumigation Degradation of MeBr in the soil-water environment produces Br− as the degradation

Methyl Bromide

63

end product, regardless of the transformation pathway. Numerous studies (Table 3) have reported the appearance of elevated concentrations of Br− in soils and plants grown in the soil after MeBr fumigation (Hoffmann and Malkomes 1974). Some of these studies are summarized here to provide a brief review of the problem. Bromide residues produced by MeBr fumigation have importance because excessive uptake of plant materials containing Br− is considered harmful to humans. In addition, some plants, mainly carnations, are sensitive to high Br− levels in the soil. FAO/WHO recommended a maximum acceptable daily intake (ADI) of 1 mg bromine/kg body weight. Different countries and organizations have different tolerance limits in different foodstuffs, but the general limit is 50 ppm for cereals, 20–30 ppm for fresh fruits, 35–300 ppm for dried fruits, and 5–50 ppm for leafy vegetables. The concern for Br− toxicity from edible plant products grown in fumigated fields was the main reason for the suspension of MeBr in Germany (Anonymous 1980), the adoption of more restrictive regulations such as lower doses and leaching with water after fumigation in the Netherlands in the early 1980s, and the ultimate ban in the Netherlands in the early 1990s (UNEP 1995). The production of Br− from MeBr in fumigated soil is a result of the transformation of MeBr in soil. All known and proposed mechanisms of MeBr transformation in soil and water release Br−. All MeBr that is not volatilized from the soil surface will be converted to Br− in the soil. Thus, concentrations of Br− resulting from MeBr fumigation are dependent on the factors that impact MeBr degradation and volatilization and on the rate and extent of Br− removal following its formation. These factors include methods of application, application rates, surface cover removal time, soil organic matter content and texture, permeability of surface tarp, irrigation or precipitation after fumigation, and timing of successive fumigations. Masui et al. (1978, 1979) fumigated soil beds with incremental dosages of MeBr under vinyl film-tarped conditions, and found that soil Br− levels generally increased with increasing dosages in the same soil but that 10 times more Br− was formed in a soil that had a higher clay content. Nazer et al. (1982) found that soil Br− concentration increased proportionally with application rates from 45 to 135 g m−1. Hass and Klein (1976) reported that more Br− was formed in the soil when it was fumigated with liquid MeBr than with gaseous MeBr, and organic amendment greatly stimulated the conversion of MeBr to Br−. Fransi et al. (1987) noted, after fumigation at 90 g m−2 in the field, that Br− residues ranging from 5 to 10 ppm were distributed to a depth of 50–60 cm, where a compacted layer existed. In packed soil columns treated with MeBr under polyethylene film-tarped conditions, Vanachter et al. (1981a) found that nearly three times more Br− was produced in a loamy soil with 7.22% organic matter than in a sandy soil with 2.15% organic matter. In outdoor broadcast fumigation (Yates et al. 1996a; Yagi et al. 1995), because typically a much lower rate is used, Br− production in the soil profile is significantly lower than that found in fumigated glasshouse soils. After a tarped, shallow (25–30 cm) injection at 24 g m−2, noticeable increases of Br− were

64

Table 3. Br− residues in soil and plant tissues after soil fumigation (in mg kg−1). Fumigation typea

Dosage (g m−2)

Soil typeb

Soil Br−c (mg kg−1)

Plant tissued

96

Unknown

⬃12

Wheat leaves

Tarped-PE

24 49 98 24

Unknown

Lettuce

24

Loam/peat (7 : 3)

Tarped-PE (gas)

24 73 24 73 24 73 73

Loam Loam Loam/peat (4 : 1) Loam/peat (4 : 1) Loam/peat (1 : 4) Loam/peat (1 : 4) Sandy loam

⬃2 ⬃3 ⬃5 ⬃17 ⬃14 ⬃12 ⬃18 ⬃18 ⬃16 ⬃10 ⬃23 ⬃13 ⬃32 ⬃20 ⬃52 ⬃33

Tarped-PE

73

Sandy loam

⬃35

Tarped-PE (2 d) (+ discing) (+flooding) Tarped-PE (5 d) (+ discing) (+flooding) Tarped-PE

Loam/peat (7 : 3)

Carnation

Carnation

Tomato Tomato Tomato Tomato

leaves fruit leaves fruit

Reference

3500–6100 (1st yr) 2400–4200 (2nd yr) 460–2500 (3rd yr) ⬃430 ⬃770 ⬃870 1100 1300 800 1800 1800 1600

Brown and Jenkinson (1971)

9300 720

Kempton and Maw (1972)




Tarped-PE

Plant Br− (mg kg−1)


Fumigation

typea

Tarped-vinyl

Tarped-vinyl

Tarped-PE Tarped-PE (gas)

PE: polyethylene.

Loam

Clay

Clay (Iwata)

Clay (Takamatsu)

Unknown

Soil Br−c (mg kg−1) ⬃9 ⬃7 ⬃9 ⬃4 ⬃6 ⬃12 ⬃6 ⬃12 ⬃4 ⬃12 ⬃4 ⬃12 ⬃19 ⬃73 ⬃165 14–27

90

14–49

135

14–62

75 50 75 50 80

Clay

Sandy

⬃17 ⬃12 ⬃12 ⬃5 ⬃32

Plant tissued Muskmelon fruit Muskmelon fruit Muskmelon fruit Cucumber fruit Cucumber fruit Cucumber fruit Tomato fruit Pepper fruit Eggplant fruit Strawberry fruit

Tomato fruit (wet) Cucumber fruit (wet) Tomato fruit (wet) Cucumber fruit (wet) Tomato fruit (wet) Cucumber fruit (wet) Cucumber fruit (wet) Cucumber fruit (wet) Cucumber fruit (wet) Cucumber fruit (wet) Tomato fruit

Plant Br− (mg kg−1) 164 303 987 97 98 189 118–136 162–196 98–312 214–383 197–229 360–368 164–195 167–255 228–328 6–11 8–11 8–13 9–14 8–13 13–17 ⬃25 ⬃15 ⬃19 ⬃17 55 (wet)

Reference Masui et al. (1978)

Masui et al. (1979)

Nazer et al. (1982)

Malathrakis and Sarris (1983)

Fallico and Ferrante (1991)

65

Tarped-PE

200 400 600 200 400 600 100 200 100 200 100 200 100 200 300 45

Soil

typeb

Methyl Bromide

Tarped-PE

Dosage (g m−2)

66


found from the surface to about 100 cm below surface (Yates et al. 1996a). The highest Br− concentrations were found near the soil surface, and about 39% of the injected MeBr was transformed to Br−. Drastic variations were noticed in the distribution of Br− in soil, and numerous soil cores had to be sampled to obtain a meaningful average (Yates et al. 1996a). These variations could be caused by the nonuniform distribution of MeBr in soil or the effect of soil heterogeneity in degrading MeBr. The accumulation of Br− in plant tissues is usually proportional to the Br− concentration in the soil. Therefore, factors affecting the production and retention of Br− in the soil around plant root systems also affect Br− uptake by the plant (Lavergne and Arvieu 1983). In addition, the accumulation and distribution of Br− in a plant is also affected by plant species, age, and position of the organs. Brown et al. (1979) established a linear correlation (r2 = 0.74) between Br− concentrations in soil and barley plants from numerous plots. They also found that the Br− level in plants varied from one location to another and from one species to another. Soil texture, temperature, leaching, and nonuniform distribution of MeBr were among the other factors that might have caused the observed variations. Kempton and Maw (1972) reported that the accumulation of Br− by lettuce increased with application rates from 49 to 98 g m−1, and Br− residues as high as 10,100 ppm were recovered in the dried tissue. Staerk and Suess (1974) found that in Germany, the natural bromine content in fresh vegetables varied from 0.5 to 3.0 ppm. After the treatment of soil with MeBr, the bromine content of the vegetables increased by factors of 200 or more. They found that the bromine uptake was influenced by several parameters such as plant species, soil type, MeBr dose, vegetation time, and rotation. In England and Wales, lettuce grown on unfumigated soil contained less than 10 ppm Br−, while most lettuce grown on MeBr-fumigated soil contained higher levels of Br−, with a proportion in excess of 1,000 ppm (Roughan and Roughan 1984a,b). Basile and Lamberti (1981) reported that tomato, string beans, radish, egg plant, zucchini, and pepper grown in MeBr-treated soil (60 g m−2) all had markedly higher Br− concentrations than those grown in untreated soil. The accumulation of Br− also appeared to be related to the interval between soil fumigation and planting as well as to the frequency of fumigation (Roughan and Roughan 1984a,b). White and Hunt (1983) found that fumigating seedbeds with MeBr gave high Br− concentrations in the seedlings of cabbage, cauliflower, and Brussels sprouts, but 6 wk after transplanting, Br− level in the plants decreased to nearbackground level. Kempton and Maw (1973) found that uptake of Br− by tomato plants was directly influenced by fumigation rates and the interval between fumigation and planting. The concentration in the leaves decreased from the base to the tip of the plants and increased with the age of the tissue, but the accumulation in fruits was less than that in the leaves (Kempton and Maw 1973). Very high Br− concentrations derived from MeBr fumigation also exhibited phytotoxicity to certain plants. Brown and Jenkinson (1971) noticed that wheat plants were discolored (“scorched”) as a result of MeBr fumigation, particularly near the injection sites. They associated this phenomenon with Br− uptake by

Methyl Bromide

67

the plants, and found as high as 6100 ppm Br− in the scorched plants. Kempton and Maw (1974) found that carnations were extremely sensitive to inorganic Br−, and many carnation plants were injured or even killed when planted in MeBr-fumigated soil. Flooding the soil or incorporating peat into soil alleviated the injury due to their effects on lowering soil Br− level. Possible Br− residues in edible plant parts is an important objection against the use of MeBr as a soil fumigant in some European countries, where MeBr is often repeatedly used in glasshouses. Various methods have been studied to reduce Br− uptake by plants. Among the tested methods, the most effective approaches are leaching the fumigated soil with abundant amounts of water and reducing the application dosage by using less-permeable tarps or mixtures of MeBr with a secondary fumigant. In packed soil columns, the amount of Br− leached out of soil was found to be directly dependent on the amount of water added, and the amount of water needed was dependent on the soil type (Vanachter et al. 1981a,b). After leaching with 40 cm water, 47% of the produced Br− was removed from a sandy soil column and 92% was eliminated from a loamy soil column. MaCartney and Price (1988) reported that after flooding the treated soil with 15 cm of water, the Br− content in the surface soil layers decreased from 34 to 4 ppm. Roorda et al. (1984) also found that leaching effectively reduced soil Br− derived from fumigation, and in some cases, soil Br− level became lower than that before the fumigation. In addition to leaching, application methods that allow a lower input of MeBr into the soil were found to result in lower Br− levels in the fumigated soil. Coosemans and Van Assche (1976) compared soil Br− production following applications of pure MeBr and a mixture of 70% MeBr and 30% chloropicrin and found that soil Br− concentrations were considerably lower with the application of the mixture. There are some negative effects associated with leaching. Trace amounts of MeBr were carried in greenhouse leachate water into surface water when leaching was started directly after the films were removed (1 wk after the fumigation). As a result, some small fish were killed near the drainage discharge point (Wegman et al. 1981). The concentration of MeBr in drainage water was found to be related to the covering time and the time that irrigation started (de Heer et al. 1986). Leaching fumigated greenhouse soil also resulted in high Br− concentrations in surface and drainage waters near the fumigation site, but it was found that the Br− was rapidly diluted after leaching was stopped (Wegman et al. 1981; Vanachter et al. 1981b; de Heer et al. 1986). Calculation of Br− mass balance over a year in a polder district of the Netherlands revealed that fumigation with MeBr contributed 68% of the total Br− found in the precipitation, surface water, and groundwater, which corresponds to conversion of 14% of the applied MeBr (Wegman et al. 1983). 5. Production of MeBr by Plants Production of MeBr by plants has been demonstrated in laboratory studies (Saini et al. 1995; Gan et al. 1998a). Floating plant leaf disks on solutions containing high concentrations of halide ions resulted in evolution of MeBr and other methyl halides (Saini et al. 1995). The

68


capability of plant leaves to methylate halide ions and release methyl halides appears to be common among higher plants. A high proportion of species tested demonstrated the capability to methylate halide ions, with the family Brassicaceae being particularly efficient (Saini et al. 1995). Plants may contain a variety of methyl transferase enzymes that effect this reaction. Intact plants grown in soil containing Br− were also observed to produce MeBr. In a study by Gan et al. (1998a), several species from the Brassicaceae family were grown in soil containing 0.4–100 mg/kg Br−. Production of MeBr per gram of plant mass was correlated to the soil Br− concentration. No MeBr production was observed in Br−-treated soil without plants. No MeBr was observed in microcosms from which the aboveground plant mass was removed, indicating that MeBr production or release could not be caused by plant roots alone. Because low concentrations of Br− are ubiquitous in soils, terrestrial plants may be an important component of the MeBr cycle. Flux of MeBr and other methyl halides from rice paddies was measured by Redeker et al. (2000). Results indicated production of MeBr by rice paddies, although the study did not include a plant-free control, so it was not verified that plants were responsible for MeBr production. The flux of MeBr depended on the rice growth stage, and maximum MeBr flux occurred during heading and flowering. Plots in which the rice straw was incorporated into the soil had increased soil organic matter content and increased MeBr production. At one site, MeBr production was proportional to soil Br− concentrations. Muramatsu and Yoshida (1995) also observed production of methyl halide (methyl iodide) by rice plants, with seasonal variation in MeI production. Unplanted soil controls also produced measurable MeI, but both flooded and unflooded soils produced less MeI than the treatments containing rice plants. Gan et al. (1998a) estimated a global production level of MeBr by rapeseed and cabbage plants to be 6.6 ± 1.6 Gg/yr and 0.4 ± 0.2 Gg/yr, respectively. Rice paddies may contribute approximately 1.3 Gg/yr (Redeker et al. 2000). These values are of the same order of magnitude as some anthropogenic sources of MeBr (Yates et al. 1998). It seems likely that the contribution of plants is important in the MeBr cycle, although it is often not accounted for in current MeBr budgets. B. Phase Partitioning 1. Henry’s Law Constant It has been long established that the movement of a volatile chemical in soil is controlled by its distribution behavior over the soil–water–air phases. The reported Kh for MeBr at 20 °C varies from 0.24 to 0.30 (Siebering and Leistra 1978; Gan and Yates 1996), and changes with temperature. With Kh of this magnitude, it can be expected that the movement of MeBr in unsaturated soil is mainly driven by its diffusion via the vapor phase (soil air). The temperature dependence of the Henry’s law constant for MeBr is shown in Fig. 1, including Arrhenius equations and fitted parameters. 2. Adsorption The other distribution factor, the adsorption coefficient, Kd (mL g−1), is important as a retaining force in slowing down MeBr transport through

Methyl Bromide

69

Fig. 1. Vapor pressure, solubility, and Henry’s constant as a function of temperature.

70


the soil. There are a few published measured or estimated Kd and Koc values for MeBr. The reported Koc ranges from 9 to 22 (Briggs 1981; Karickhoff 1981; Rao et al. 1985), which corresponds to a Kd of 0.09–0.22 in a soil with 1% organic carbon. Arvieu (1983) measured MeBr adsorption and desorption and found different characteristics for soil with different organic matter contents. In organic matter-poor soils, the adsorption of MeBr is very weak unless the soil is very dry. In organic matter-rich soils, the adsorption is considerably greater. The same author also noted that the adsorbed MeBr became resistant to desorption. Gan and Yates (1996) observed that degradation of MeBr during the equilibration in adsorption studies might have contributed to the observed increased adsorption in soils with high organic matter content. A noticeable fraction of the spiked MeBr was degraded to Br− during a 16-hr shaking period in organic matter-rich soils. This phenomenon may be also responsible for the irreversibility found in MeBr desorption isotherms (Arvieu 1983). After correcting for the degraded fraction, MeBr adsorption became negligible in all the tested soils (Gan and Yates 1996). Therefore, MeBr can be considered to be a nonadsorbing chemical in soil with normal water content.

VII. Simulating the Environmental Fate of Methyl Bromide Models of various complexity are available to simulate the transport of water, heat, and MeBr in variably saturated soils. A comprehensive simulation requires the use of conservation (i.e., mass or energy balance) equations as well as other equations that describe the mass or energy flux, reactions, and interactions that are important to describe the problem. Under natural conditions existing in agricultural soils, many factors affect the transport process to varying degrees, making it virtually impossible to use models for reliable prediction in all cases. The fate and transport of MeBr is strongly affected by temperature. To conduct an accurate simulation under temperature-dependent conditions, a governing equation is necessary that describes the transport of energy. In general, this equation will be coupled to both the water flow and chemical transport equations and considerably increases the model and solution complexity. A. Transport Model A common approach for simulating the fate and transport of MeBr for saturated and unsaturated water flow conditions, with consideration of variable soil temperature, includes descriptions for at least three governing processes: water flow, heat transport, and fate and movement of MeBr. Programs exist that numerically solve the nonlinear partial differential equations for one- and two-dimensional systems, nonequilibrium coupled transport of water, heat, and solute (in both liquid and gaseous phases) in a variably saturated porous medium. Degradation is usually described using a first-order decay reaction, and often the degradation rate in each phase (liquid, vapor, and solid) can be specified. The governing transport equations can be written as follows (S˘imu˚nek and van Genuchten 1994).

Methyl Bromide

Water Transport:

71

册

冋

∂h ∂θ ∂ = Kij(h) + Kiz(h) − S ∂t ∂xi ∂xj

(8)

where θ is the volumetric water content [L3 L−3], h is the pressure head [L], Kij are components of the unsaturated hydraulic conductivity tensor [L t−1], and S is a sink term [t−1]; t is time, x is distance [L], and indices i and j represent the horizontal and vertical directions. Heat Transport: Ch(θ)

册

冋

∂T ∂ ∂T ∂T = λij(θ) − Cwqi ∂t ∂xi ∂xj ∂xi

(9)

where Ch and Cw are the volumetric heat capacity for the porous media [J m−3 K−1] and liquid, respectively, and λij is the apparent thermal conductivity [Wm−1 K−1]. Solute Transport:

冋

册冋

册

∂θCL ∂ρbCS ∂ηCg ∂ ∂CL ∂Cg ∂ + + = θDijw + ηDgij ∂t ∂t ∂t ∂xi ∂xj ∂xi ∂xj +

(10)

∂qiCL − (µwθCL + µgηCg + µSρbCS) − SCr ∂xi

where CL [M L−3], CS[M M−1], and Cg[M L−3] are solute concentrations for the liquid, solid, and gaseous phases, respectively; q is the volumetric flux density; µw , µs , and µg are first-order rate constants [t−1] for solutes in the liquid, solid, and gas phases, respectively; θ is the volumetric water content, ρ is the soil bulk density, η is the soil air content, S is the sink term in the water flow equation [t−1], Cr is the concentration of the sink term, Dijw is the dispersion coefficient tensor for the liquid phase [L2 t−1], and Dgij is the diffusion coefficient tensor for the gas phase. Numerous computer programs have been developed to evaluate the effects of interacting processes and factors on pesticide movement through the root zone and to the groundwater. The approach used in developing the programs varies with the intended use of the model. Some of these include GLEAMS (Leonard et al. 1987), LEACHM (Wagenet and Hutson 1987), PRZM (Carsel et al. 1985, 1998), PESTAN (Enfield et al. 1982), and SESOIL (Bonazountas and Wagner 1984). Some of these models are not capable of predicting pesticide movement when water is applied in a controlled manner by furrow or subsurface drip irrigation systems. This limitation has led to the development of processbased models that can be used to predict the transport in irrigated agriculture: CHAIN-2D (Sˇimu˚nek and van Genuchten 1994), HYDRUS-2D (Sˇimu˚nek et al. 1996), and PESTLA (van den Berg and Boesten 1997). 1. Volatilization Boundary Condition For methyl bromide, critical processes affecting the fate and transport in soils are vapor diffusion and volatilization.

72


Volatilization is an especially important route of dissipation because of MeBr’s large vapor pressure and Henry’s law constant, as demonstrated in recent field experiments (Yagi et al. 1995; Majewski et al. 1995; Yates et al. 1996b). Excessive volatilization is associated with many problems, such as a reduction in the amount of material available to control pests and increased potential for contamination of the atmosphere. Emission losses to the atmosphere pose an increased risk to persons living near treated fields. When simulating MeBr emissions to the atmosphere, the approach used to describe the soil surface– atmospheric boundary condition strongly affects the simulated emission response. The most common volatilization boundary condition used in current models is based on stagnant boundary layer theory (Jury et al. 1983). This approach assumes that a thin stagnant air layer occurs at the soil–atmosphere interface and that chemical movement across the layer is the result of vapor diffusion. The controlling parameter is the mass transfer coefficient, which is expressed as the ratio of the binary diffusion coefficient (i.e., air and MeBr) to the boundary layer thickness (Jury et al. 1983). A limitation of this approach is the estimation of the thickness of the stagnant boundary layer. Further, for some atmospheric conditions (e.g., changes in barometric pressure), it is likely that chemical transport occurs by both advection and diffusion. For these situations, assuming a stagnant boundary layer is inappropriate and more complex boundary conditions are required (Massmann and Farrier 1992; Chen et al. 1995). An advantage of the stagnant boundary layer approach is that information about atmospheric conditions is unnecessary. However, adopting this boundary condition produces MeBr emission histories that are very regular and often do not resemble the erratic behavior commonly observed in the field (Majewski et al. 1995; Yates et al. 1996b, 1997). A more accurate description of the volatilization process requires the coupling of soil-based processes with those operating in the atmosphere, which led Baker et al. (1996) to develop an alternate formulation for the mass transfer coefficient that includes atmospheric resistance terms. This boundary condition depends on several aerodynamic parameters, such as the roughness Reynolds number, the Schmidt number, the friction velocity, the wind speed, and an atmospheric stability term. The atmospheric resistance to diffusion near the soil surface and aerodynamic resistance from the diffusive layer to the measurement height affects the predicted emission rate. Further research is needed to evaluate the effectiveness of this approach in simulating the volatilization boundary condition, especially for agricultural fumigation. Several studies have been conducted to determine whether conventional modeling approaches can accurately predict the rate of MeBr volatilization from bare soils (Wang et al. 1997b; Yates et al. 2002). Wang et al. (1997b) used CHAIN-2D to simulate methyl bromide emissions from a 3.5-ha field and compared the simulation results to emissions measured in a field experiment (Yates et al. 1996c). They found that the model simulated the total emission within a few percent of the measured value but the pattern of instantaneous emission rate

Methyl Bromide

73

was much more regular than the measured values and, at times, a value of the predicted volatilization rate could be very different from the measured value. Yates et al. (2002) conducted a similar study using the same experimental data and the volatilization boundary condition of Baker et al. (1996). They found that the predicted emissions had a more realistic temporal pattern compared with a simulation based on the stagnant boundary layer. The total emissions were also within a few percent of the measured value. When discrepancies occur, it cannot be determined whether the model or the measured values were in error because the measured volatilization rate is also subject to uncertainty (Majewski 1997). Further research is needed to improve the accuracy of volatilization measurements and simulation models. Research is also needed to develop and test methods for coupling atmospheric and soil processes in models so that more accurate predictions of the volatilization rate can be obtained. B. Mobility Indices Along with highly sophisticated numerical approaches, there is a need for simple methods for ranking pesticides by their potential to contaminate the environment. Jury et al. (1983, 1984a,b) described a screening model for assessing the relative volatility, mobility, and persistence of pesticides and other trace organic chemicals in the soil. Other mobility indices reported in the literature include the retardation and attenuation factors (Rao et al. 1985) and the convective and diffusive mobility times (Jury et al. 1984b). The retardation factor (RF) is an index of the relative time needed for a pesticide to move past some specified depth compared to a nonadsorbing tracer. The attenuation factor (AF) is the fraction of pesticide mass that is likely to move past some specified depth and includes the effects of adsorption and degradation. The convective and diffusive mobility times (tc and tD) give a measure of the time needed for a pesticide to travel a specified distance by convection or diffusion, respectively. This information is important in determining the suitability of candidate pesticides and in developing management practices. Jury et al. (1984b) classified MeBr as one of the most mobile compounds considered, with a class 5 convective and class 3 diffusive mobility. In terms of persistence, MeBr was found to be class 5, very short lived. Mobility indices were used by Yates and Gan (1998) to compare the transport potential of MeBr to several alternative fumigants, herbicides, and insecticides (Table 4). The results demonstrate that MeBr is highly mobile, an ideal characteristic for a fumigant, but it is somewhat more persistent in soil than the other soil fumigants. In general, all the fumigant materials, with the possible exception of methyl isothiocyanate (MITC) have high diffusive mobility and are quickly attenuated in soil. The mobility indices can provide insight into observed fumigant behavior. For example, some fumigants (e.g., MITC) have relatively low diffusive mobility compared to MeBr and often suffer from difficulties achieving a uniform soil concentration throughout the treatment zone. Olson and No-

74


Table 4. Mobility indices for MeBr and selected alternative fumigants and herbicides.

Pesticide

RF

AF R = 25 cm

tc R = 10 cm

tD R = 10 cm

Class

Methyl bromide 3-Bromopropyne (Z)-1,3-D (E)-1,3-D Methyl isothiocyanate 2,4-D (acid) Atrazine Carbaryl Lindane

2.37 1.25 2.81 2.79 1.34 2.1 6.52 17.6 61.7

0.59 0.18 0.042 0.019 0.37 0.33 0.56 0.0001 0.45

7.12 3.75 8.44 8.38 4.01 6.31 19.62 52.69 185.19

4.03 11.54 16.19 27.6 55.18 —a —a —a —a

3 3 3 2 2 1 1 1 1

RF: retardation factor; AF: attenuation factor. aGreater than 105.

ling (1994) found that MITC may not completely control root gall in tomatoes and strawberries because of poor chemical distribution, causing others to develop better methods for distributing MITC in soil (Juzwik et al. 1997; McHenry 1994). MeBr has relatively high diffusive mobility and low retention on soils and therefore provides a very uniform concentration distribution and pest control. C. Simulating Transport in Relatively Dry Soils For certain situations, it is possible to simplify the mathematical model given by Eqs. 8 to 10. Within the context of predicting MeBr transport in the subsurface under typical application conditions, it is often reasonable to neglect the flow of water. For shank injection, this can be a reasonable assumption because the fumigation occurs in relatively dry soils. For dry-soil conditions, the magnitude of the effective diffusion coefficient is determined primarily by the vaporphase component (Fig. 2); this is a result of the magnitude of the gas-phase diffusion coefficient, which is approximately four orders of magnitude greater than the liquid-phase diffusion coefficient (Jury et al. 1983). Further, given the short time scale of a soil fumigation, water movement (redistribution) is also very small. The MeBr transport process can be approximated mathematically with a single transport equation that is not coupled to the Richard’s water flow equation. Such an assumption would not be appropriate if the fumigant were applied in moist soils or delivered in irrigation waters. 1. Using Models to Extrapolate Laboratory Data to Field Situations Comparing experimental measurements from laboratory and field studies can be problematic. In the laboratory, soil columns are often of limited length and have impermeable surfaces at the lower end of the column. In fields, there are generally no impermeable boundaries and MeBr gases can diffuse unimpeded. There-

Methyl Bromide

75

Fig. 2. Contribution of vapor and liquid components to the effective soil diffusion coefficient as a function of water content. For water contents below 0.375 (dotted line), the vapor-phase component dominates. Soil porosity = 0.4. Devap and Deliquid use the % of De scale.

fore, deep diffusion in the field promotes additional soil residence time (and concomitant degradation) and reduces emissions. This effect was illustrated by Gan et al. (1997a), who found that total MeBr emission values from an untarped 60-cm laboratory column could overpredict field values by 11%–58%, respectively, for a 20-cm and 60-cm injection depth. For a 60-cm injection, the measured value of 60% considerably overestimated a field measurement of 21% (Yates et al. 1997). This finding led to the development of a method to extrapolate laboratory measurements conducted in soil columns to values more representative of field conditions. The corrected emissions were obtained by using two analytical solutions to a partial differential equation describing vapor diffusion in a one-dimensional homogeneous soil system. For both, it was assumed that the chemical and material properties are spatially and temporally constant. These assumptions allow simulation using a simple differential equation: η

∂Cg ∂Cl ∂Cs ∂2Cg ∂2Cl +θ + ρb = Dg + D − ηµgCg − θµlCl − ρbµsCs (11) l ∂t ∂t ∂t ∂x2 ∂x2

where Cg, Cl, and Cs are the vapor-, liquid-, and solid-phase concentrations [mg/ m3], Dl and Dg are the liquid and gaseous diffusion coefficients [m2/d], and θ, η, ρb, and µi are the water content, air content, bulk density, and degradation coefficient for the ith phase. The three terms on the left-hand side of Eq. 11 describe the time rate of change in the gas-, liquid-, and sorbed-phase concentration. On the right-hand side of Eq. 11, the first two terms describe gas- and liquid-phase diffusion and the remaining terms describe first-order degradation in each phase.

76


For both analytical solutions, linear relationships are used to characterize the partitioning between the solid and liquid phases and for liquid–vapor partitioning: C s = K d Cl C g = K h Cl

(12)

with initial and boundary conditions ∂C(0, t) Surface Boundary = h [C(0, t) − Cair] ∂x ∂C(x, t) (13) *x→L = 0 Lower Boundary ∂x C(x, 0) = Co [u(x − xo − δ) − u(x − xo + δ)] Initial Condition Ds

where Ds is the effective soil diffusion coefficient [m2/d] and h is a mass transfer coefficient [m/d] that characterizes the resistive nature of the interface between soil and the atmosphere. The initial condition assumes that the chemical is injected into the column at a specified depth, xo, with a pulse width of 2δ. For a laboratory column with an impermeable barrier at the bottom, the solution to Eq. 11, given the initial and boundary conditions in Eq. 13, is this: C(x, t) = e−µt

∞

∑a

2

n

e−knDst Cos[kn (F − x)]

(14)

i=1

where an =

4Co{Sin[kn (F − zo + δ)] − Sin[kn (F − zo − δ)]} 2knF + Sin(2knF)

(15)

where δ is some small interval and kn are the roots of the following equation: kn → h Cos[F kn] − Dskn Sin[F kn] = 0

(16)

There are an infinite number of values for kn that satisfy Eq. 16. A second solution for transport in a homogeneous field soil that does not have an impermeable vapor barrier at depth (i.e., when L → ∞ in Eq. 13) is this:

冋册冋册冋√ 册冋√ 册冋 √ 册冋 √ 册

z + z2 z − z2 2 C(x, t) µ t e = Erfc + Erfc Co √4Dst √4Dst − Erfc

z + z1

4Dst

− Erfc

+ 2 eh[ht+(z+z1)]/Ds Erfc − 2 eh[ht+(z+z2)]/Ds Erfc

where z1 = F − zo − δ and z2 = F − zo + δ.

z − z1

4Dst

2ht + z + z1 4Dst

2ht + z + z2 4Dst

(17)

Methyl Bromide

77

Equations 14 and 17 can be used to extrapolate laboratory measurements to analogous field values by finding the transport parameters that produce the best fit of Eq. 14 to the laboratory measurements, then using these parameters in Eq. 17 to obtain corresponding corrected (or “field”) values. Values obtained through this sort of procedure should be more representative of field values compared to the original laboratory measurements. An application using this method is discussed in Section X.A.1.

VIII. Methyl Bromide Diffusion in Soils A. Diffusion Coefficient The value of the diffusion coefficient of a chemical in the vapor phase is generally 104 times larger than that in the liquid phase (Jury et al. 1983). The diffusion coefficient can be estimated using a variety of methods (Reid et al. 1987), including the Fuller correlation Dab =

0.00143T 1.75

冋

1/3 1/3 P M1/2 ab (Σν)a + (Σν)b

册

2

(18)

where Dab is the binary diffusion coefficient (cm2/s), T is the absolute temperature (K), Mab = 2/(M a−1 + Mb−1), Ma and Mb are the molecular weights of air and MeBr, respectively, P is the pressure (bars), and Σν is obtained using the atomic diffusion volumes (Reid et al. 1987). Using Eq. 18 yields an estimated diffusion coefficient for MeBr of 0.114 cm2 s−1 at 20 °C and 1 atmosphere ambient pressure. The temperature dependence of the diffusion coefficient, shown in Fig. 3, appears to be nearly linear over the temperature range 0°–60 °C. The temperature dependence of the binary diffusion coefficient can be described using Eq. 18 or using activation energy and the Arrhenius equation as shown in Fig. 3. Using a screening model, Jury et al. (1991) found that the movement of a chemical is dominated by vapor-phase diffusion if the air-to-water partition coefficient, or the Henry’s law coefficient (Kh) is Ⰷ10−4. As the Kh for MeBr is approximately 0.25, transport in the vapor phase is important in describing the fate and transport in soil. B. Methyl Bromide Diffusion in Soils Because of its large air–water distribution coefficient (e.g., Kh) and insignificant adsorption to soil, a large fraction of MeBr partitions into the soil air in unsaturated soils. For instance, in a soil with 20% volumetric water content and 20% air space, as much as 24% of the total MeBr may be present in the soil air with negligible adsorption onto the soil. This partitioning forms the basis of rapid vertical movement of MeBr immediately following a typical subsoil injection, and the potential for significant volatilization losses into the air if movement

78


Fig. 3. The temperature dependence of the binary diffusion coefficient for MeBr in air. The solid line is from Eq. 18, and an equivalent Arrhenius equation is shown.

across the soil surface is not suppressed. Figure 4 contains a series of concentration curves measured at different time intervals following an injection at 30 cm into a homogeneously packed Greenfield sandy loam column (Gan et al. 1997a). It is clear that after the injection MeBr quickly expands and moves vertically in both directions. Unless the soil surface is sealed with a relatively impermeable

Fig. 4. Methyl bromide distribution in the soil gas phase after injection at 30 cm. Soil surface tarped with 1 mil HDPE.

Methyl Bromide

79

tarp or saturated with water, a large fraction of the applied MeBr would be expected to escape into the air. Decreases in the total area under a curve in Figure 4 are due to both soil degradation and volatilization. Although the transport of MeBr in soil is largely governed by its intrinsic chemical and physical properties, some soil and meteorological conditions can influence this process by affecting MeBr partitioning over the phases. Of these factors, temperature, soil water content, and content and continuity of soil air spaces are important. With increases in soil temperature, more MeBr is expected to partition into the soil air, the diffusion coefficient is increased, and therefore the movement should be accelerated (Goring 1962). However, little experimental effort has been devoted to measuring MeBr transport in soil, or volatilization from soil, as a function of temperature. Soil water content was shown to be important in controlling MeBr penetration in soil (Kolbezen et al. 1974; Abdalla et al. 1974). Methyl bromide diffuses more rapidly, deeply, and widely in drier soils than in wet soils. Altering soil water content has also been shown to be effective in reducing MeBr volatilization losses from soil. Lower volatilization loss occurred from Greenfield sandy loam with a water content of 18% than from the same soil with 6% water content (Gan et al. 1996). This effect was even more drastic when water was applied to soil surface under tarped conditions (Jin and Jury 1995). Only about 4% of the MeBr was lost after three surface irrigations of 16 mm water in a tarped packed soil column. Soil water content affects MeBr transport by altering the soil void ratio and the continuity of air-filled pores. Theoretically, when a soil is saturated with water, gas-phase diffusion is eliminated and the movement of a volatile chemical will be greatly reduced. Other factors that may affect soil air spaces and distribution include disturbance and compaction of the soil. During mechanized application, void channels are often created by shanks or chisels along the injection lines. It is a common practice during untarped fumigation of MeBr to use a roller behind the shanks to compact the soil surface and to close some of the openings caused by injection. Compaction has been found to result in reduced MeBr diffusion in soil (Knavel et al. 1965). In packed soil columns, MeBr diffused more slowly and less MeBr was lost through volatilization from a soil column packed at 1.7 g cm−3 than from the same soil at 1.40 g cm−3 (Gan et al. 1996). The diffusion of fumigants in soil has been studied using mathematical and numerical simulation techniques since the 1950s. Hemwall (1959) was one of the first to utilize computers to numerically solve a two-dimensional diffusion equation similar to Eq. 11. Two principal goals of this work were the determination of the diffusion pattern and the calculation of the biological control function, also called the concentration–time index: i

B = ∫ C(t) dt

(19)

0

where the value of B gives an indication of the level of biological control. The simulation was used to obtain information on fumigant application methods and the soil MeBr distribution to ensure that efficacious dosages and biological con-

80


trol were achieved. From the simulation, it was found that during the first day the diffusion process was basically radial with rapidly increasing losses to the atmosphere. At later times, the soil concentration was generally uniform with a constantly decreasing flux rate. Since then, numerous studies have been conducted that adopted similar assumptions. Siebering and Leistra (1978) developed a diffusion model to study how moisture, bulk density (e.g., tillage), dosage, and cover time affect control, assessed from concentration–time curves. From the results of their model, they discuss how various soil, chemical, and application factors affect the movement and dissipation of methyl bromide in soils. In a more recent study, Reible (1994) used a one-dimensional diffusion model together with more recent estimates of the model parameters to determine the fate and transport of methyl bromide in soils compared to that of a methyl bromide alternative, methyl isothiocyanate (MITC). Reible suggested that because MeBr is more dense than air, the downward movement of MeBr would occur at a rate of approximately 42 cm/d (assuming 30% porosity and 10−8 cm−2 permeability), but density-driven mass flow was not considered in the model. They also found that a shallow water table would act as a barrier to downward movement and thus enhance volatilization. The simulation included the effects of a film that was assumed to have a constant permeability of 8.31 L/hr/m2. The effect of temperature on the permeability of polyethylene film was not considered, although has been shown to be important (Kolbezen and Abu-El-Haj 1977; Yates et al. 1996c). Additionally, the effect of temperature on the volatilization rate was not considered. Therefore, the estimated flux curves do not show the diurnal fluctuations that are commonly observed in field experiments. Also, the parameters that were used are appropriate for a mean temperature of approximately 20°–25 °C, and the effects of atmospheric resistance above the tarp were not included in the model. The results indicate that approximately 45% of the applied MeBr would be lost to the atmosphere within approximately 14 d when injected at 25 cm and the soil covered with plastic that is removed after 2 d. Several field experiments have shown that the time response to the MeBr flux is highly irregular (Yagi et al. 1993, 1995; Majewski et al. 1995; Yates et al. 1996b) and is caused, in part, by temperature effects on behavior of the plastic (Kolbezen and Abu-El-Haj 1977) and the diffusion coefficient (Reid et al. 1987; Wang et al. 1997b), along with diurnal changes in pressure (Massman and Farrier 1992; Chen et al. 1995) and atmospheric processes (Baker et al. 1996). Sorption of many pesticides does not follow the linear equilibrium assumption, especially when the concentration is constantly changing due to movement in the gas phase. The assumption of a linear equilibrium sorption response is often adopted to simplify the mathematics. Brown and Rolston (1980) showed that the transport of volatile compounds was better described by a first-order kinetic model than by a linear equilibrium assumption. Their experiments were conducted in a laboratory with a steady flow of gas. They found that the retardation factor obtained using the linear equilibrium approach was dependent on the

Methyl Bromide

81

flow rate through the column. The parameters for the first-order kinetic model, however, were not sensitive to flow rate, which is evidence that this model may be more appropriate. To obtain a model suitable to describe field studies, transport in two or three dimensions must be considered. Rolston and Glauz (1982) developed a twodimensional model of MeBr diffusion that included the effects of reversible first-order sorption. They assumed that the input of MeBr into soil was a saturated cylinder of radius r0 and that gas entered the soil after diffusing out of the cylinder. They investigated the use of plastic covers and compared soil gas concentrations measured midway between the injection points to simulated values. The authors found that the simulated values without plastic film matched the field data with film better than the simulation with a film present. This result indicates that either the model did not include important processes, such as heat transport, or one or more of the input parameters were not correctly specified. This represents a challenging test of the model because the data were reported in the literature nearly 10 years earlier; this also points to the difficulty in predicting pesticide transport in field studies. Particularly interesting are the predicted Br− concentrations that show a peak value at the injection depth. Comparing this with field data collected by Yates et al. (1996a) shows that the maximum Br− concentration occurred at the soil surface, which may be caused by higher soil organic matter content in the near-surface soil enhancing the effective degradation coefficient in this region; soil evaporation may also transport Br− to the soil surface.

IX. Assessing Methyl Bromide Volatilization from Soil A. Measurement Methods There are numerous methods for estimating the gas flux from soils to the atmosphere. Three methods used recently to estimate MeBr emissions to the atmosphere include estimating total loss from Br− appearance (i.e., from MeBr degradation, micrometeorological, and enclosure-based methods). 1. Estimating Total MeBr Loss from Br− Appearance The Br−-appearance method assumes that the difference between the MeBr mass applied and mass degraded (i.e., Br− produced) was released into the atmosphere. Therefore, measuring Br− in the soil provides a method for estimating the total atmospheric emission. An advantage of this method is the ease of analyzing the Br− content of soils. A disadvantage is the large number of soil samples necessary to obtain an accurate field-scale estimate of degradation at all depths (Jury 1985; Yates et al. 1996a). Also, no information about the dynamics of MeBr emissions can be obtained using this method. Figure 5 shows the Br− [mg/kg] concentration measured on a dry soil weight basis taken before and after MeBr application during tarped-soil, shallow injection (Fig. 5A) and bare-soil deep injection (Fig. 5B). An estimate of the total

82


Fig. 5. Bromide ion distribution in soil after methyl bromide application at 25 cm and covering the soil surface with high-density polyethylene (HDPE) (A) and after injection into bare soil at 68 cm (B).

MeBr lost to the atmosphere can be obtained from the difference between the initial and final curves and converting from Br− mass to MeBr mass. From Fig. 5A it was estimated that 325 (±164) kg or 39% (±19%) of the applied MeBr was degraded to Br−. Because the MeBr mass (parent material) remaining in the field at the time of Br− sampling was estimated to be less than 0.05% at the time of sampling, the total loss from volatilization is approximately 518 (±164) kg, or 61% (±19%) of the applied MeBr mass. The spatial and measurement variability was found to introduce considerable uncertainty into the Br− mass calculation, as shown by the standard error of ±164 kg, and also produces uncertainty in the estimate of MeBr volatilization. Figure 5B shows the measured soil Br− content before and 3 mon after a deep-injection (68 cm), bare-soil fumigation experiment. From this information it was determined that 78% of the amount applied was detected as Br− (22% volatilized). A second sampling 6 mon after application was conducted to verify the accuracy of the amount degraded and was found to be within 3% of the first sampling, indicating that an accurate field-scale Br− concentration measurement was obtained. The nearly equivalent values for the total mass degraded from each sampling (75%–78%) indicate that sufficient samples were collected to obtain an accurate field-scale average. Field-scale variability has been shown to have considerable effect on the spatial average solute concentration sampled in the field (Jury 1985). This effect has also been shown for MeBr (Yates et al. 1996a) and can be illustrated as follows. If the estimate of the MeBr mass degraded shown in Figure 5A was obtained using 45 background concentrations and 1100 samples taken at the completion of the experiment, the total volatilization loss was estimated to be 298 kg or 35.3%. If, instead, 500 background samples were used, the total volatilization loss was estimated to be 435 kg or 48.4%. When 1100 background samples were used, the MeBr mass loss was estimated to be 518 kg or 61%, which more closely matched the total loss estimates determined by using other

Methyl Bromide

83

methods. This result demonstrates that large errors are possible when solute concentrations are determined in too few samples in a field soil. 2. Chamber Methods The two principal approaches for using chambers to estimate the volatilization rate from soil into the atmosphere are passive (closed system, assumed perfect mixing) and active (or flowing system) chamber methods. The chamber method allows measurement of the volatilization rate over small surface areas compared to other methods, which is both an advantage and a disadvantage. Chambers may offer the only means of measuring volatilization for situations where the source has a small areal extent. For large fields, however, the volatilization rate measured using chambers may be highly variable because of a combination of field-scale spatial variability of properties that affect the mass transfer and the chamber’s small measurement area (Yagi et al. 1993; Yates et al. 1996c; Williams et al. 1999). Each measurement of chemical concentration in the chamber indicates the instantaneous flux. Multiple chambers are required to provide information on the spatial variability of MeBr flux. Flux estimates are sensitive to the placement of the chambers relative to the position of MeBr injection, i.e., distance to the source, and the presence of a chamber can affect the area sampled, especially the local temperature and relative humidity (Clendening 1988). These characteristics of chambers can have a tremendous effect on experimental uncertainty. To characterize the dynamics of MeBr flux and to accurately determine cumulative emissions, numerous samples are required from each chamber over the time period for which volatilization is occurring. Emissions of MeBr from treated fields follow a diurnal variation (Yates et al. 1997), and appropriate sampling intervals must be chosen to adequately represent volatilization throughout the experiment. There have been numerous reviews of the chamber methods, including those by Rolston (1986), Wesely et al. (1989), and Livingston and Hutchinson (1995), as well as articles that describe the use of these methods (Matthias et al. 1980; Hutchinson and Mosier 1981; Reicosky 1990; Harrison et al. 1995) for measuring the surface fluxes of trace gases. The reviews cite several advantages of chambers compared to meteorological methods. For example, chambers are easy to construct, relatively inexpensive to operate and conceptually simple. Chambers can be used to measure gas losses from very small surfaces, which make them suitable for use in laboratory studies to obtain surface emission data to aid in understanding the physical, chemical, and biological aspects of transport and fate of volatile contaminants in soils columns. Passive Chambers Passive or closed chamber systems consist of a container with an open bottom that is placed over the soil surface for a short time period. The concentration of the target gas in the chamber increases with time as the chemical moves from the soil matrix into the chamber. The concentration of MeBr in the chamber is determined at the completion of some time interval (∆t) elapsed since the placement of the chamber on the soil surface. From this

84


information and use of various equations, the flux into the chamber can be estimated. A linear model is most commonly used: flux =

V ∆C A ∆t

(20)

where V is the chamber volume and A is the area of soil sampled. This approach assumes complete mixing inside the chamber, constant emissions during the placement period, and no loss mechanism (Matthias et al. 1980; Rolston 1986; Flessa et al. 1995). It has been generally recognized that the simple model in Eq. 20 underestimates the volatilization rate due to a reduction in the gradients as the chemical concentration builds up inside the chamber (Jury et al. 1982; Rolston 1986; Livingston and Hutchinson 1995). To overcome this, more comprehensive nonlinear models have been developed that do not require the constant flux assumption (Hutchinson and Mosier 1981; Livingston and Hutchinson 1995; de Mello and Hines 1994; Valente et al. 1995). For example, Hutchinson and Mosier (1981) use the following: flux =

冋

册

V (C1 − Co)2 C1 − C o ln A ∆t (2 C1 − C2 − Co) C2 − C 1

(21)

where Co is the initial concentration in the chamber. To use Eq. 21, the concentrations C1 and C2 must be measured at two times, ∆t and 2 ∆t. Active Chambers Another simple method for measuring the rate of emission of volatile compounds is the flowthrough flux chamber method (Hollingsworth 1980; Clendening 1988). The flowthrough flux chamber is a closed system device that allows the pesticide emission from a small surface area to be collected. A continuous and uniform flow rate is maintained through the chamber. The flow rate is chosen to avoid MeBr accumulation inside the chamber and to minimize negative pressure inside the chamber, which could cause advective mass flow. Chemical concentrations are determined in the air entering (Cin) and exiting (Cout) the chamber. Once this information is known, the flux is determined using flux =

(flow rate)(Cout − Cin) (sampled area)

(22)

Adsorbent tubes are often used to accumulate MeBr mass from the chamber effluent over some time interval, typically 2–4 hr. Temperature Effects on Chambers One disadvantage of chamber methods is that the pesticide flux measurement is affected by the presence of the chamber covering the sampling area. For example, the temperature and relative humidity have been shown to be affected by the presence of a chamber (Clendening 1988). During a series of experiments, the air temperature inside the chamber was found to be as much as 5 °C higher than the outside air temperature. The

Methyl Bromide

85

relative humidity inside the chamber was found to be from 20% to 50% higher than external values. Using chambers of the same design, Yates et al. (1996c) found that the internal chamber temperature could be as high as 30 °C above the ambient air temperature under certain conditions. A few factors that can alleviate temperature increases inside chambers include using high flow rates and opaque materials whenever possible. Because the permeability of polyethylene plastic to MeBr has been shown to be strongly affected by temperature (Kolbezen and Abu El-Haj 1977; Yates et al. 1997; Papiernik et al. 2002), increased temperature inside a flux chamber can produce biased MeBr volatilization rates. Yates et al. (1996c) demonstrated that a correction should be used to provide a more accurate estimate of the MeBr flux density because of this effect whenever significant heating occurs. In their experiment, uncorrected flux chamber data provided an estimate of ⬃96% of the applied MeBr mass being lost to the atmosphere, nearly 50% greater than the total loss estimates from micrometeorological and the appearance of Br− methods. After correcting for temperature increases inside the active chambers, estimated total MeBr emission was approximately 59%, about 5% less than estimated from micrometeorological and appearance of Br− methods. This finding indicates the importance of designing chambers that minimize internal heating. 3. Micrometeorological Methods The meteorological techniques require that emissions occur over a large surface area, so that relatively stable gas concentration profiles can be established and detected above the soil surface. Reviews and comparisons of various meteorological techniques are available (Fowler and Duyzer 1989; Wesely et al. 1989; Majewski et al. 1990; Denmead and Raupach 1993). Methods for measuring the volatilization rate using micrometeorological information are fairly complex, require numerous measurements of MeBr concentration and other atmospheric parameters, and may require assumptions concerning the behavior of the atmosphere. Advantages are that the methods are well tested, they provide a field-scale average total emission rate, and they provide information on the dynamics of the volatilization process. Aerodynamic Method The aerodynamic method is based on atmospheric gradients of wind speed, temperature, and concentration and provides a measurement of the pesticide flux from the soil surface (Parmele et al. 1972; Brutsaert 1982; Majewski et al. 1989). The method requires a spatially uniform source and a relatively large upwind fetch so that the atmospheric gradients are fully developed. The fetch requirements are generally assumed to be from 50 to 100 times the height of the instruments, which for large agricultural fields is typically greater than 0.5 m in height. Chemical concentrations, wind speed, and temperature are determined at multiple heights at the sampling location to define the gradient. To characterize the temporal variability in flux, samples are collected over relatively short time intervals, typically 2–4 hr. Adsorbent tubes are often used to accumulate the volatilized mass throughout the sampling interval.

86


The aerodynamic method was originally developed for use under neutral atmospheric conditions. Using empirical relations, however, the method can be extended to stable and unstable atmospheric conditions. Numerous stability corrections have been proposed (Fleagle and Businger 1980; Brutsaert 1982; Rosenberg et al. 1983). The aerodynamic equation, suitable for general atmospheric stability conditions, is: fz(0, t) = k2

[c¯ 1(t) − c¯ 2(t)][u¯ 2(t) − u¯ 1(t)] ϕm(t) ϕc(t) ln(z2/z1)2

(23)

where k ⬇ 0.4 is von Ka´rma´n’s constant, fz(0, t) is the interval-averaged vertical flux at the soil surface, u¯(t) is the interval-averaged wind speed [m/s], z is height above the soil surface [m] (note: z2 > z1), and c¯1(t) is the interval-averaged concentration [µg/m3] at height z1 above the soil surface. The gradient-based stability corrections for a particular time interval, t, can be written as (Rosenberg et al. 1983): 1

ϕc = ϕm = (1 − 16Ri)− 3

Ri < 0

unstable

1 3

Ri > 0

stable

ϕc = ϕm = (1 + 16Ri)

+

(24)

where it is assumed that the stability functions for momentum and the concentration are the same. Majewski et al. (1995) used a slightly different equation to account for atmospheric stability, and several other stability corrections have been proposed (Fleagle and Businger 1980; Brutsaert 1982; Rosenberg et al. 1983). The Richardson number, Ri, is defined as Ri =

冋册

g ∂T ∂u −2 T ∂z ∂z

(25)

where g is the gravitational acceleration (i.e., 9.8 m/s2), and T(t) is the absolute temperature [K]. The gradient Richardson’s number is one means for characterizing the importance of buoyancy and mechanical mixing on the turbulence. Theoretical Profile Shape Method The theoretical profile shape method (Wilson et al. 1982) can be used to determine the volatilization rate from field experiments conducted on a circular plot. This method has advantages over the aerodynamic method in that (1) the large fetch requirement is not necessary, (2) measurements of the air concentration and wind speed are needed at only one height, and (3) the sensor is placed at a height that is relatively insensitive to the atmospheric stability so temperature and wind gradients and stability corrections are unnecessary. This approach is based on the trajectory simulation model described by Wilson et al. (1981a–c). Wilson et al. (1983) and Majewski et al. (1990) have used this method, among others, to determine the rate of pesticide and ammonia volatilization from field experiments. Yates et al. (1996b, 1997) adapted the method so that MeBr volatilization from rectangular fields could be estimated.

Methyl Bromide

87

The flux density is estimated from [u¯ (t) c¯ (t)] *Zinst (26) Ω where interval-average values of the wind speed, u¯(t), and air concentration, c¯(t), are obtained at the instrument height, Zinst. Flux can be obtained by determining the ratio of the horizontal to vertical flux, Ω, using the trajectory simulation model discussed below. This ratio depends on surface roughness and upwind fetch distance (i.e., the radius of the circular plot) but does not depend on wind speed. fz(0, t) =

Trajectory Simulation The theoretical profile shape method is based on the trajectory simulation model. This model is used to simulate pesticide transport in the atmosphere using a particle-tracking algorithm and assuming that the atmosphere experiences conditions of inhomogeneous turbulence. Wilson et al. (1981a–c) give a complete description of the method. The simulation of a particle of air mass proceeds in a hypothetical atmosphere in which the Eulerian velocity, σw(z), time [τ(z)] and length [Λ(z) = σw(z)τ(z)] scales are assumed to vary only in the vertical direction. It is further assumed that the horizontal wind speed varies only in the z direction, the timeaveraged value of the vertical wind speed is zero, the pesticide source is spatially uniform, there are no sources of pesticide outside the treated area, no degradation of pesticide occurs once it is in the atmosphere, and the surface roughness of the field, zo, is assumed to be constant. A source area extending from 0 ≤ x ≤ Xmax is discretized into M sections of equal length. A large number of particles is emitted from each section and tracked until the particles reach the collector (located at Xmax) which coincides with the position of the sampling mast in the field. During the simulation, the instantaneous vertical position of each particle is obtained and used to determine the current horizontal position increment for the current time step. Once a particle has reached the collector, the count of the appropriate height level, Z, is incremented. As the number of particles released increases, the statistical character of the vertical distribution becomes fixed and the ratio of horizontal to vertical flux, Ω, as well as the instrument height, Zinst, becomes known. To obtain Ω, the simulation is conducted for strongly stable, strongly unstable and neutral atmospheric conditions. Plotting the results against height produces a curve similar to that shown in Fig. 6. The height where the three curves more or less converge is the height where the sensor is positioned. At this location, the effects of atmospheric stability are minimized. One difficulty using this method is the determination of the instrument height before initiating the experiment; this involves estimating surface roughness, which may not be known until after the experiment begins (i.e., when the plastic is placed on the field, at the time of application). Because the upwind source distance varies with the wind direction for rectangular fields, the trajectory simulation must be conducted for several upwind source distances ranging from the smallest to the largest distance between the

88


Fig. 6. Using the trajectory simulation model (Wilson et al. 1982), a height in the atmosphere can be found that is relatively insensitive to atmospheric stability. For this example, the instrument height is approximately 288 cm and Ω is approximately 15.

sampling mast and edge of the field. Then, a relationship can be developed between wind direction, instrument height, and Ω, which may be used to estimate the flux using the average wind direction for the sampling interval. Integrated Horizontal Flux Method The integrated horizontal flux method (Denmead et al. 1977; Wilson et al. 1982; Majewski et al. 1990) can be used to estimate the surface flux when the concentration, c(z), and horizontal wind speeds, u(z), in the atmosphere are known as a function of height. Measurements of chemical concentration and wind speed gradients are similar to those used for the aerodynamic method. Assuming a spatially uniform source, the flux is estimated from a statement of mass balance, that is: fz(0, t) =

1 L

∞

∫

u¯ (z) c¯ (z) dz

(27)

0

From this equation, the mass emitted from the surface upwind from a sampling point is equal to the mass that passes through a vertical plane of sufficient height to capture all the mass (i.e., of infinite extent) located at the sampling point. To use this method, the concentration profile at several heights must be determined, and the distance of the source area upwind from the sampling mast must be known. An advantage of this method over the aerodynamic method is that cor-

Methyl Bromide

89

rections for atmospheric stability are not needed because this approach is based on principles of mass balance. B. Field Experiments to Determine Methyl Bromide Volatilization Since the mid-1990s, several experiments have been conducted to obtain information on MeBr emissions from typical agricultural operations. The results from these studies are summarized in Table 5. Various methods were used to estimate the emission rate, including an increase in soil Br− concentration as a result of MeBr degradation (Yates et al. 1996a), the atmospheric flux method (Majewski et al. 1995; Yates et al. 1996b), and the enclosed flux chamber method (Yagi et al. 1993, 1995; Yates et al. 1996c). Every method has advantages and disadvantages that often make the interpretation of the experimental results somewhat difficult. However, for determining the total emission, all the methods should provide reasonably accurate results. 1. Yagi et al. (1993) Yagi et al. (1993) conducted an experiment in Irvine, California, to measure the MeBr emission from a fumigated southern California field using four passive flux chambers. MeBr was applied at a depth of approximately 25 cm and the soil surface was covered with low-density polyethylene plastic film. The authors originally estimated that 87% of the total MeBr applied to the field escaped into the atmosphere. This estimate was revised to 74% ± 5% (Williams et al. 1999) by eliminating the data from a chamber that covered tarp material with a hole. The estimates of MeBr emissions measured during this study are the highest reported for MeBr injection at shallow depth and the soil surface covered with plastic. The high emission rates are probably due to a combination of factors such as use of low-density polyethylene plastic, which is permeable to MeBr vapors (Kolbezen and Abu-El-Haj 1977), the presence of a high bulk density and moist soil layer at 60 cm depth. This value is also higher than expected given other estimates based on mathematical models (Albritton and Watson 1992; Singh and Kanakidou 1993), but was similar in magnitude to the losses observed in glasshouse studies (de Heer et al. 1983). To verify these results, the authors returned to the field to collect Br− information to provide mass balance information (Yagi et al. 1995). 2. Yagi et al. (1995) The investigators conducted a second experiment in a nearby field using the same procedures as their first experiment (Yagi et al. 1993). For this experiment, high-density polyethylene (HDPE) plastic was used to cover the field and five flux chambers were used to measure emissions. They found that only 34% of the applied MeBr escaped to the atmosphere. This value is more than 50% lower than the result of their first experiment, which included a low-density polyethylene tarp. Variability in the emission measurement is expected for several reasons: (1) only 10–15 samples of the volatilization rate were obtained during each 7-d experiment, generally at the high point during the day; (2) only a few soil samples were taken to measure Br− concentrations,

90

Table 5. Total amount of methyl bromide volatilized during the experiment and mass balance.

Flux chamber Flux chamber Aerodynamic, profile Aerodynamic, profile Br− appearance Aerodynamic, discrete Aerodynamic, profile Theoretical profile shape Integrated horizontal flux Flux chamber Br− Appearance Aerodynamic, discrete Aerodynamic, profile Theoretical profile shape Integrated horizontal flux Flux chamber Flux chamber (location 1) Flux chamber (location 1) Flux chamber (location 2)

Surface cover

25–30 35 25–30 25–30 25–30

PE PE Bare PE HDPE

25–30

HDPE

69

25–30

Bare

LDPE

Number of sampling intervals 14 14 31 33* 2 105 105 105 105 107 2 164 164 164 164 173 12 17 14

MeBr Percent Number of Field mass MeBr measurements Total number of size applied volatilized per interval measurements (ha) (kg/ha) (%) 4 chambers 5 chambers 6 heights 6 heights 1100 samples 2 heights 6 heights 6 heights 6 heights 3 chambers 1050 samples 2 heights 9 heights 9 heights 9 heights 4 chambers 3 chambersa 8 chambers 5 chambersb

56 70 186 198 2200 210 630 630 630 321 2098 328 1476 1476 1476 692 36 136 70

17 1 6 4 3.5

256 243 199 262 240

3.5

240

3.5

322

17 17 1

256 208 243

Mass balance (%)

87 — 34 104 32 — 89 — 61 — 62 101† 67 106† 60 99† 70 108† 59 97† 21 ± 2.6 — 4.5 ± 1.3 83† 3.1 ± 0.1 82 2 81 1.9 ± 0.2 81 4.9 ± 3.1 84 74 ± 5a 97 ± 5a 63 ± 12 94 ± 12 36 ± 6b 106 ± 11b

Reference Yagi et al. (1993) Yagi et al. (1995) Majewski et al. (1995) Yates et al. (1996a)

Yates et al. (1997)

Williams et al. (1999)


Method used

Injection depth (cm)


Method used

Injection depth (cm)

Surface cover

35

HDPE

25

Flux chamber (10-d cover period)

25

Flux chamber (15-d cover period)

25

Br− Appearance

25

Br− Appearance

60

HDPE Hytibar Hytibar HDPE Hytibar Hytibar HDPE Hytibar Hytibar Bare HDPE HDPE + water Bare HDPE Hytibar

16 17 13 114 162 114 114 162 114 114 162 114 2 2 2 2 2 2

MeBr Percent Number of Field mass MeBr measurements Total number of size applied volatilized per interval measurements (ha) (kg/ha) (%) 7 chambers 8 chambers 7 chambers 2 2 2 2 2 2 2 2 2 75 75 75 75 75 75

112 136 91 228 324 228 228 324 228 228 324 228 150 150 150 150 150 150

1 1 — 0.002c

0.002c

0.002c

0.002c

0.002c

233 257 310 280 210 140 280 210 140 280 210 140 345 241 363 427 368 508

24 ± 5 45 ± 8 50 ± 9 68 36 39 56 2 3 67 3 1 87 59

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