This report contains the collective views of international groups of experts and does not necessarily represent the decisions or the stated policy of the United Nations Environment Programme, the International Labour Organization, or the World Health Organization.

Environmental Health Criteria 224

ARSENIC AND ARSENIC COMPOUNDS

Second edition

The first and second drafts of this monograph were prepared, under the coordination of Dr J. Ng, by the authors A. Gomez-Caminero, P. Howe, M. Hughes, E. Kenyon, D.R. Lewis, M. Moore, J. Ng, and by A. Aitio and G. Becking.

Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organization, and the World Health Organization, and produced within the framework of the Inter-Organization Programme for the Sound Management of Chemicals.

World Health Organization

Geneva, 2001

The International Programme on Chemical Safety (IPCS), established in 1980, is a joint venture of the United Nations Environment Programme (UNEP), the International Labour Organization (ILO), and the World Health Organization (WHO). The overall objectives of the IPCS are to establish the scientific basis for assessment of the risk to human health and the environment from exposure to chemicals, through international peer-review processes, as a prerequisite for the promotion of chemical safety, and to provide technical assistance in strengthening national capacities for the sound management of chemicals.

The Inter-Organization Programme for the Sound Management of Chemicals (IOMC) was established in 1995 by UNEP, ILO, the Food and Agriculture Organization of the United Nations, WHO, the United Nations Industrial Development Organization, the United Nations Institute for Training and Research, and the Organisation for Economic Co-operation and Development (Participating Organizations), following recommendations made by the 1992 UN Conference on Environment and Development to strengthen cooperation and increase coordination in the field of chemical safety. The purpose of the IOMC is to promote coordination of the policies and activities pursued by the Participating Organizations, jointly or separately, to achieve the sound management of chemicals in relation to human health and the environment.

WHO Library Cataloguing-in-Publication Data

Arsenic and arsenic compounds.

(Environmental health criteria ; 224)

1.Arsenic - toxicity

2.Arsenicals - toxicity

3.Environmental exposure

I. International Programme on Chemical Safety

II. WHO Task Group on Environmental Health Criteria for Arsenic and Arsenic Compounds

III.Series

ISBN 92 4 157224 8

(NLM Classification: QV 294)

ISSN 0250-863X

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CONTENTS

ENVIRONMENTAL HEALTH CRITERIA FOR ARSENIC AND ARSENIC COMPOUNDS

NOTE TO READERS OF THE CRITERIA

MONOGRAPHS

Every effort has been made to present information in the criteria monographs as accurately as possible without unduly delaying their publication. In the interest of all users of the Environmental Health Criteria monographs, readers are requested to communicate any errors that may have occurred to the Director of the International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland, in order that they may be included in corrigenda.

* * *

A detailed data profile and a legal file can be obtained from the International Register of Potentially Toxic Chemicals, Case postale 356, 1219 Châtelaine, Geneva, Switzerland (telephone no. + 41 22 - 9799111, fax no. + 41 22 - 7973460, E-mail irptc@unep.ch).

* * *

This publication was made possible by grant number 5 U01 ES02617-15 from the National Institute of Environmental Health Sciences, National Institutes of Health, USA, and by financial support from the European Commission.

The Commonwealth Department of Health and Aged Care, Australia, contributed financially to the preparation of this Environmental Health Criteria monograph. The Task Group meeting was arranged by the National Research Centre for Environmental Toxicology, Australia.

Environmental Health Criteria

Objectives

In 1973 the WHO Environmental Health Criteria Programme was initiated with the following objectives:

The first Environmental Health Criteria (EHC) monograph, on mercury, was published in 1976 and since that time an ever-increasing number of assessments of chemicals and of physical effects have been produced. In addition, many EHC monographs have been devoted to evaluating toxicological methodology, e.g. for genetic, neurotoxic, teratogenic and nephrotoxic effects. Other publications have been concerned with epidemiological guidelines, evaluation of short-term tests for carcinogens, biomarkers, effects on the elderly and so forth.

Since its inauguration the EHC Programme has widened its scope, and the importance of environmental effects, in addition to health effects, has been increasingly emphasized in the total evaluation of chemicals.

The original impetus for the Programme came from World Health Assembly resolutions and the recommendations of the 1972 UN Conference on the Human Environment. Subsequently the work became an integral part of the International Programme on Chemical Safety (IPCS), a cooperative programme of UNEP, ILO and WHO. In this manner, with the strong support of the new partners, the importance of occupational health and environmental effects was fully recognized. The EHC monographs have become widely established, used and recognized throughout the world.

The recommendations of the 1992 UN Conference on Environment and Development and the subsequent establishment of the Intergovernmental Forum on Chemical Safety with the priorities for action in the six programme areas of Chapter 19, Agenda 21, all lend further weight to the need for EHC assessments of the risks of chemicals.

Scope

The criteria monographs are intended to provide critical reviews on the effect on human health and the environment of chemicals and of combinations of chemicals and physical and biological agents. As such, they include and review studies that are of direct relevance for the evaluation. However, they do not describe every study carried out. Worldwide data are used and are quoted from original studies, not from abstracts or reviews. Both published and unpublished reports are considered and it is incumbent on the authors to assess all the articles cited in the references. Preference is always given to published data. Unpublished data are used only when relevant published data are absent or when they are pivotal to the risk assessment. A detailed policy statement is available that describes the procedures used for unpublished proprietary data so that this information can be used in the evaluation without compromising its confidential nature (WHO (1990) Revised Guidelines for the Preparation of Environmental Health Criteria Monographs. PCS/90.69, Geneva, World Health Organization).

In the evaluation of human health risks, sound human data, whenever available, are preferred to animal data. Animal and in vitro studies provide support and are used mainly to supply evidence missing from human studies. It is mandatory that research on human subjects is conducted in full accord with ethical principles, including the provisions of the Helsinki Declaration.

The EHC monographs are intended to assist national and international authorities in making risk assessments and subsequent risk management decisions. They represent a thorough evaluation of risks and are not, in any sense, recommendations for regulation or standard setting. These latter are the exclusive purview of national and regional governments.

Content

The layout of EHC monographs for chemicals is outlined below.

• Summary – a review of the salient facts and the risk evaluation of the chemical

• Identity – physical and chemical properties, analytical methods

• Sources of exposure

• Environmental transport, distribution and transformation

• Environmental levels and human exposure

• Kinetics and metabolism in laboratory animals and humans

• Effects on laboratory mammals and in vitro test systems

• Effects on humans

• Effects on other organisms in the laboratory and field

• Evaluation of human health risks and effects on the environment

• Conclusions and recommendations for protection of human health and the environment

• Further research

• Previous evaluations by international bodies, e.g. IARC, JECFA, JMPR

Selection of chemicals

Since the inception of the EHC Programme, the IPCS has organized meetings of scientists to establish lists of priority chemicals for subsequent evaluation. Such meetings have been held in Ispra, Italy, 1980; Oxford, United Kingdom, 1984; Berlin, Germany, 1987; and North Carolina, USA, 1995. The selection of chemicals has been based on the following criteria: the existence of scientific evidence that the substance presents a hazard to human health and/or the environment; the possible use, persistence, accumulation or degradation of the substance shows that there may be significant human or environmental exposure; the size and nature of populations at risk (both human and other species) and risks for environment; international concern, i.e. the substance is of major interest to several countries; adequate data on the hazards are available.

If an EHC monograph is proposed for a chemical not on the priority list, the IPCS Secretariat consults with the Cooperating Organizations and all the Participating Institutions before embarking on the preparation of the monograph.

Procedures

The order of procedures that result in the publication of an EHC monograph is shown in the flow chart on p. xvii. A designated staff member of IPCS, responsible for the scientific quality of the document, serves as Responsible Officer (RO). The IPCS Editor is responsible for layout and language. The first draft, prepared by consultants or, more usually, staff from an IPCS Participating Institution, is based initially on data provided from the International Register of Potentially Toxic Chemicals, and reference data bases such as Medline and Toxline.

The draft document, when received by the RO, may require an initial review by a small panel of experts to determine its scientific quality and objectivity. Once the RO finds the document acceptable as a first draft, it is distributed, in its unedited form, to well over 150 EHC contact points throughout the world who are asked to comment on its completeness and accuracy and, where necessary, provide additional material. The contact points, usually designated by governments, may be Participating Institutions, IPCS Focal Points, or individual scientists known for their particular expertise. Generally some four months are allowed before the comments are considered by the RO and author(s). A second draft incorporating comments received and approved by the Director, IPCS, is then distributed to Task Group members, who carry out the peer review, at least six weeks before their meeting.

The Task Group members serve as individual scientists, not as representatives of any organization, government or industry. Their function is to evaluate the accuracy, significance and relevance of the information in the document and to assess the health and environmental risks from exposure to the chemical. A summary and recommendations for further research and improved safety aspects are also required. The composition of the Task Group is dictated by the range of expertise required for the subject of the meeting and by the need for a balanced geographical distribution.

The three cooperating organizations of the IPCS recognize the important role played by nongovernmental organizations. Representatives from relevant national and international associations may be invited to join the Task Group as observers. Although observers may provide a valuable contribution to the process, they can only speak at the invitation of the Chairperson. Observers do not participate in the final evaluation of the chemical; this is the sole responsibility of the Task Group members. When the Task Group considers it to be appropriate, it may meet in camera.

All individuals who as authors, consultants or advisers participate in the preparation of the EHC monograph must, in addition to serving in their personal capacity as scientists, inform the RO if at any time a conflict of interest, whether actual or potential, could be perceived in their work. They are required to sign a conflict of interest statement. Such a procedure ensures the transparency and probity of the process.

When the Task Group has completed its review and the RO is satisfied as to the scientific correctness and completeness of the document, it then goes for language editing, reference checking and preparation of camera-ready copy. After approval by the Director, IPCS, the monograph is submitted to the WHO Office of Publications for printing. At this time a copy of the final draft is sent to the Chairperson and Rapporteur of the Task Group to check for any errors.

It is accepted that the following criteria should initiate the updating of an EHC monograph: new data are available that would substantially change the evaluation; there is public concern for health or environmental effects of the agent because of greater exposure; an appreciable time period has elapsed since the last evaluation.

All Participating Institutions are informed, through the EHC progress report, of the authors and institutions proposed for the drafting of the documents. A comprehensive file of all comments received on drafts of each EHC monograph is maintained and is available on request. The Chairpersons of Task Groups are briefed before each meeting on their role and responsibility in ensuring that these rules are followed.

WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR ARSENIC AND ARSENIC COMPOUNDS

Members

Dr C. Abernathy, Office of Water/Office of Science and Technology, Health and Ecological Criteria Division, US Environmental Protection Agency, Washington, D.C., USA (Chairperson)

Dr D. Chakraborti, School of Environmental Studies, Jadavpur University, Calcutta, India

Professor J.S. Edmonds, Department of Chemistry, De Montfort University, Leicester, United Kingdom

Dr H. Gibb, US Environmental Protection Agency, National Center for Environmental Assessment, Washington DC, USA

Dr P. Hoet, Industrial and Occupational Medicine Unit, Catholic University of Louvain, Brussels, Belgium

Dr C. Hopenhayn-Rich, Department of Preventive Medicine and Environmental Health, University of Kentucky, Lexington, KY, USA

Mr P.D. Howe, Centre for Ecology and Hydrology, Monks Wood Experimental Station, Abbots Ripton, Huntingdon, Cambridgeshire, United Kingdom

Dr L. Järup, Department of Epidemiology and Public Health, Imperial College School of Medicine, London, United Kingdom

Dr A.A. Meharg, Department of Plant and Soil Science, Aberdeen, United Kingdom

Professor M.R. Moore, Director, Queensland Health Scientific Services and National Research Centre for Environmental Toxicology, Queensland, Australia (Vice-Chairperson)

Dr J. C. Ng, National Research Centre for Environmental Toxicology, Brisbane, Australia

Dr A. Nishikawa, Division of Pathology, National Institute of Health Sciences, Tokyo, Japan

Dr L. Pyy, Director of the Deptartment, Oulu Regional Institute of Occupational Health, Oulu, Finland

Dr M. Sim, Unit of Occupational and Environmental Health, Department of Epidemiology and Preventive Medicine, Monash University, Victoria, Australia

Dr J. Stauber, CSIRO Energy Technology, Lucas Heights Science and Technology Centre, Bangor, NSW, Australia

Professor M. Vahter, Institute of Environmental Medicine, Karolinska Institute, Stockholm, Sweden

Observers/Representatives

Dr P. Imray, Scientific Adviser, Environmental Health Branch, Queensland Health, Brisbane, Australia

Dr L. Tomaska, Canberra, Australia (representing the Australia New Zealand Food Authority)

Mr D. Hughes, MIM Holdings Limited, Brisbane, Australia (representing the Mining Industry)

Secretariat

Dr A. Aitio, International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland

Dr G. Becking, Kingston, Ontario, Canada (Adviser to the Secretariat)

Dr K. Buckett, Director DHAC, Public Health Division, Canberra, Australia

Mr P. Callan, Assistant Director, National Health and Medical Research Council, Canberra, Australia

Dr M.F. Hughes, NHEERL/ET/PKB, US Environmental Protection Agency, Research Triangle Park, NC, USA

Dr E.M. Kenyon, NHEERL/ET/PKB, US Environmental Protection Agency, Research Triangle Park, NC, USA

Dr D.R. Lewis, Human Studies Division, NHEERL, US Environmental Protection Agency, Research Triangle Park, NC, USA

Dr M. Younes, International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland

WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR ARSENIC AND ARSENIC COMPOUNDS

The first and second drafts of this monograph were prepared, under the coordination of Dr J. Ng, by the authors A. Gomez-Caminero, P. Howe, M. Hughes, E. Kenyon, D.R. Lewis, M. Moore, J. Ng, and by A. Aitio and G. Becking. The group of authors met at National Health and Environmental Effects Research Laboratory, US. EPA, North Carolina, on 20–22 July 1998.

A WHO Task Group on Environmental Health Criteria for Arsenic and Arsenic Compounds met at the National Research Centre for Environmental Toxicology, Brisbane, Australia, on 15–19 November 1999. The group reviewed the draft and the peer review comments, revised the draft and made an evaluation of the risks for human health and environment from exposure to arsenic and arsenic compounds.

After the meeting, and based on the peer reviewer comments and Task Group advice, Drs Gibb, Hopenhayn-Rich, Järup, Sim, and Aitio revised and updated the section on Effects on Human Health. This section was then sent for review to a selected group of experts.

The document was revised on the basis of the peer review comments received, these revisions were verified, and the document was finalized by a Review Board, consisting of Drs D. Anderson, H. Gibb, L. Järup, M. Sim and A. Aitio, in TNO BIBRA, Carshalton, UK. The document was finally approved by the Task Group in a mail ballot.

The cut-off date for the literature searches for the document was the Task Group meeting, i.e. November 1999, with the exception of the section on effects on human health, for which the last literature searches were performed in November 2000.

Peer review comments at the first stage international review were received from:

Dr J. Ahlers, Umwelt Bundes Amt, Germany

Dr R. Benson, Region VIII, Environmental Protection Agency, USA

Professor GB Bliss, N.N. Petrov’s Research Institute of Oncology, Russian Federation

Dr M. Bolger, Food and Drug Administration, USA

Professor M. Cíkrt, Centre of Industrial Hygiene and Occupational Diseases, Czech Republique

Professor I. Dési, Albert Szent-Györgyi University, Hungary

Professor J Duffus, The Edinburgh Centre for Toxicology, UK

Dr P Edwards, Department of Health, UK

Dr H Falk, Agency for Toxic Substances and Disease Registry, USA

Dr H. Gibb, Environmental Protection Agency, USA

Dr N. Kurzeja European Environmental Bureau, Germany

Dr I. Mangelsdorf, Fraunhofer Institute, Germany

Dr TG Rossman, NYU School of Medicine

Professor H Taskinen, Finnish Institute of Occuational Health

Mr S Tsuda, Ministry of Halth and Welfare, Japan

Dr G. Ungváry, József Fodor National Center for Public Health, Hungary

Professor M. Vahter, Karolinska Institute, Sweden,

Bureau of Chemical Safety, Canada

Elf Atochem North America, USA

Environmental Protection Agency Office of Research and Development, USA

Eurometaux

Finnish Institute of Occupational Health, Finland

Comments on the revised section on effects on human health were received from members of the Task Group, and from:

Dr D Anderson, TNO BIBRA International, UK

Dr Michael Bates, Kenepuru Science Centre, New Zealand

Dr R. Calderon, National Health and Environmental Effects Research Laboratory, US EPA

Professor PE Enterline, University of Pittsburgh, USA

Dr A. Gomez-Caminero, National Health and Environmental Effects Research Laboratory, US. EPA

Dr J Lubin, National Cancer Institute, USA

Professor AH Smith, University of California, USA

Dr A. Aitio of the IPCS central unit was responsible for the scientific aspects of the monograph, and Kathleen Lyle for the technical editing.

The efforts of all, especially Queensland Health and the Natinal Research Centre for Environmental Toxicology, Australia, who helped in the preparation and finalization of the monograph are gratefully acknowledged.

ABBREVIATIONS

1. SUMMARY

1.1 Properties and analytical procedures

Arsenic is a metalloid widely distributed in the earth’s crust and present at an average concentration of 2 mg/kg. It occurs in trace quantities in all rock, soil, water and air. Arsenic can exist in four valency states: –3, 0, +3 and +5. Under reducing conditions, arsenite (As(III)) is the dominant form; arsenate (As(V)) is generally the stable form in oxygenated environments. Elemental arsenic is not soluble in water. Arsenic salts exhibit a wide range of solubilities depending on pH and the ionic environment.

There is a variety of instrumental techniques for the determination of arsenic. These include AAS, AFS, ICP-AES, ICP-MS and voltammetry. Some of these (e.g. ICP-MS) can serve as element-specific detectors when coupled to chromatographic separation techniques (e.g. HPLC and GC). These so-called "hyphenated" methods are used for determining individual arsenic species. Additional sensitivity for a limited range of arsenic compounds can often be achieved by the use of hydride generation techniques. A test kit based on the colour reaction of arsine with mercuric bromide is currently used for groundwater testing in Bangladesh and has a detection limit of 50–100 µg/litre under field conditions.

1.2 Sources and occurrence of arsenic in the environment

Arsenic is present in more than 200 mineral species, the most common of which is arsenopyrite.

It has been estimated that about one-third of the atmospheric flux of arsenic is of natural origin. Volcanic action is the most important natural source of arsenic, followed by low-temperature volatilization.

Inorganic arsenic of geological origin is found in groundwater used as drinking-water in several parts of the world, for example Bangladesh.

Organic arsenic compounds such as arsenobetaine, arsenocholine, tetramethylarsonium salts, arsenosugars and arsenic-containing lipids are mainly found in marine organisms although some of these compounds have also been found in terrestrial species.

Elemental arsenic is produced by reduction of arsenic trioxide (As₂O₃) with charcoal. As₂O₃ is produced as a by-product of metal smelting operations. It has been estimated that 70% of the world arsenic production is used in timber treatment as copper chrome arsenate (CCA), 22% in agricultural chemicals, and the remainder in glass, pharmaceuticals and non-ferrous alloys.

Mining, smelting of non-ferrous metals and burning of fossil fuels are the major industrial processes that contribute to anthropogenic arsenic contamination of air, water and soil. Historically, use of arsenic-containing pesticides has left large tracts of agricultural land contaminated. The use of arsenic in the preservation of timber has also led to contamination of the environment.

1.3 Environmental transport and distribution

Arsenic is emitted into the atmosphere by high-temperature processes such as coal-fired power generation plants, burning vegetation and volcanism. Natural low-temperature biomethylation and reduction to arsines also releases arsenic into the atmosphere. Arsenic is released into the atmosphere primarily as As₂O₃ and exists mainly adsorbed on particulate matter. These particles are dispersed by the wind and are returned to the earth by wet or dry deposition. Arsines released from microbial sources in soils or sediments undergo oxidation in the air, reconverting the arsenic to non-volatile forms, which settle back to the ground. Dissolved forms of arsenic in the water column include arsenate, arsenite, methylarsonic acid (MMA) and dimethylarsinic acid (DMA). In well-oxygenated water and sediments, nearly all arsenic is present in the thermodynamically more stable pentavalent state (arsenate). Some arsenite and arsenate species can interchange oxidation state depending on redox potential (Eh), pH and biological processes. Some arsenic species have an affinity for clay mineral surfaces and organic matter and this can affect their environmental behaviour. There is potential for arsenic release when there is fluctuation in Eh, pH, soluble arsenic concentration and sediment organic content. Weathered rock and soil may be transported by wind or water erosion. Many arsenic compounds tend to adsorb to soils, and leaching usually results in transportation over only short distances in soil.

Three major modes of arsenic biotransformation have been found to occur in the environment: redox transformation between arsenite and arsenate, the reduction and methylation of arsenic, and the biosynthesis of organoarsenic compounds. There is biogeochemical cycling of compounds formed from these processes.

1.4 Environmental levels and human exposure

Mean total arsenic concentrations in air from remote and rural areas range from 0.02 to 4 ng/m³. Mean total arsenic concentrations in urban areas range from 3 to about 200 ng/m³; much higher concentrations (> 1000 ng/m³) have been measured in the vicinity of industrial sources, although in some areas this is decreasing because of pollution abatement measures. Concentrations of arsenic in open ocean seawater are typically 1–2 µg/litre. Arsenic is widely distributed in surface freshwaters, and concentrations in rivers and lakes are generally below 10 µg/litre, although individual samples may range up to 5 mg/litre near anthropogenic sources. Arsenic levels in groundwater average about 1–2 µg/litre except in areas with volcanic rock and sulfide mineral deposits where arsenic levels can range up to 3 mg/litre. Mean sediment arsenic concentrations range from 5 to 3000 mg/kg, with the higher levels occurring in areas of contamination. Background concentrations in soil range from 1 to 40 mg/kg, with mean values often around 5 mg/kg. Naturally elevated levels of arsenic in soils may be associated with geological substrata such as sulfide ores. Anthropogenically contaminated soils can have concentrations of arsenic up to several grams per 100 ml.

Marine organisms normally contain arsenic residues ranging from < 1 to more than 100 mg/kg, predominantly as organic arsenic species such as arsenosugars (macroalgae) and arsenobetaine (invertebrates and fish). Bioaccumulation of organic arsenic compounds, after their biogenesis from inorganic forms, occurs in aquatic organisms. Bioconcentration factors (BCFs) in freshwater invertebrates and fish for arsenic compounds are lower than for marine organisms. Biomagnification in aquatic food chains has not been observed. Background arsenic concentrations in freshwater and terrestrial biota are usually less than 1 mg/kg (fresh weight). Terrestrial plants may accumulate arsenic by root uptake from the soil or by adsorption of airborne arsenic deposited on the leaves. Arsenic levels are higher in biota collected near anthropogenic sources or in areas with geothermal activity. Some species accumulate substantial levels, with mean concentrations of up to 3000 mg/kg at arsenical mine sites.

Non-occupational human exposure to arsenic in the environment is primarily through the ingestion of food and water. Of these, food is generally the principal contributor to the daily intake of total arsenic. In some areas arsenic in drinking-water is a significant source of exposure to inorganic arsenic. In these cases, arsenic in drinking-water often constitutes the principal contributor to the daily arsenic intake. Contaminated soils such as mine tailings are also a potential source of arsenic exposure. The daily intake of total arsenic from food and beverages is generally between 20 and 300 µg/day. Limited data indicate that approximately 25% of the arsenic present in food is inorganic, but this depends highly on the type of food ingested. Inorganic arsenic levels in fish and shellfish are low (< 1%). Foodstuffs such as meat, poultry, dairy products and cereals have higher levels of inorganic arsenic. Pulmonary exposure may contribute up to approximately 10 µg/day in a smoker and about 1 µg/day in a non-smoker, and more in polluted areas. The concentration of metabolites of inorganic arsenic in urine (inorganic arsenic, MMA and DMA) reflects the absorbed dose of inorganic arsenic on an individual level. Generally, it ranges from 5 to 20 µg As/litre, but may even exceed 1000 µg/litre.

In workplaces with up-to-date occupational hygiene practices, exposure generally does not exceed 10 µg/m³ (8-h time-weighted average [TWA]). However, in some places workroom atmospheric arsenic concentrations as high as several milligrams per cubic metre have been reported.

1.5 Kinetics and metabolism

Absorption of arsenic in inhaled airborne particles is highly dependent on the solubility and the size of particles. Both pentavalent and trivalent soluble arsenic compounds are rapidly and extensively absorbed from the gastrointestinal tract. In many species arsenic metabolism is characterized by two main types of reactions: (1) reduction reactions of pentavalent to trivalent arsenic, and (2) oxidative methylation reactions in which trivalent forms of arsenic are sequentially methylated to form mono-, di- and trimethylated products using S-adenosyl methionine (SAM) as the methyl donor and glutathione (GSH) as an essential co-factor. Methylation of inorganic arsenic facilitates the excretion of inorganic arsenic from the body, as the end-products MMA and DMA are readily excreted in urine. There are major qualitative and quantitative interspecies differences in methylation, to the extent that some species exhibit minimal or no arsenic methylation (e.g. marmoset monkey, guinea-pig, chimpanzee). However, in humans and most common laboratory animals, inorganic arsenic is extensively methylated and the metabolites are excreted primarily in the urine. Factors such as dose, age, gender and smoking contribute only minimally to the large inter-individual variation in arsenic methylation observed in humans. However, lower methylation efficiency in children has been observed in only one study out of three. Studies in humans suggest the existence of a wide difference in the activity of methyltransferases, and the existence of polymorphism has been hypothesized. Animal and human studies suggest that arsenic methylation may be inhibited at high acute exposures. The metabolism and disposition of inorganic arsenic may be influenced by its valence state, particularly at high dose levels. Studies in laboratory animals indicate that administration of trivalent inorganic arsenic such as As₂O₃ and arsenite initially results in higher levels in most tissues than does the administration of pentavalent arsenic. However, the trivalent form is more extensively methylated, leading to similar long-term excretion. Ingested organoarsenicals such as MMA, DMA and arsenobetaine are much less extensively metabolized and more rapidly eliminated in urine than inorganic arsenic in both laboratory animals and humans.

Levels of arsenic or its metabolites in blood, hair, nails and urine are used as biomarkers of arsenic exposure. Blood arsenic is a useful biomarker only in the case of acute arsenic poisoning or stable chronic high-level exposure. Arsenic is rapidly cleared from blood, and speciation of its chemical forms in blood is difficult. Arsenic in hair and nails can be indicators of past arsenic exposure, provided care is taken to prevent external arsenic contamination of the samples. Arsenic in hair may also be used to estimate relative length of time since an acute exposure. Speciated metabolites in urine expressed either as inorganic arsenic or as the sum of metabolites (inorganic arsenic + MMA + DMA) provide the best quantitative estimate of recently absorbed dose of arsenic. However, consumption of certain seafood, mainly seaweed and some bivalves, may confound estimation of inorganic arsenic exposure because of metabolism of arsenosugars to DMA in the body or the presence of DMA in the seafood. Such food should be avoided for 2–3 days before urine sampling for monitoring of exposure to inorganic arsenic.

1.6 Effects on laboratory animals and in vitro systems

Both inorganic and organic forms of arsenic may cause adverse effects in laboratory animals. The effects induced by arsenic range from acute lethality to chronic effects such as cancer. The degree of toxicity of arsenic is basically dependent on the form (e.g. inorganic or organic) and the oxidation state of the arsenical. It is generally considered that inorganic arsenicals are more toxic than organic arsenicals, and within these two classes, the trivalent forms are more toxic than the pentavalent forms, at least at high doses. Several different organ systems are affected by arsenic, including skin, respiratory, cardiovascular, immune, genitourinary, reproductive, gastrointestinal and nervous systems.

Several animal carcinogenicity studies on arsenic have been carried out, but limitations such as high dose levels, limited time of exposure and limited number of animals make these inconclusive. However, a recently reported animal model may be a useful tool for future carcinogenicity studies. In that study, female C57B1/6J mice exposed to arsenic in drinking-water containing 500 µg As(V)/litre over 2 years was associated with increased incidence in tumours involving mainly lung, liver, gastrointestinal tract and skin. Inorganic arsenic does not induce point mutations. However, arsenic can produce chromosomal aberrations in vitro, affect methylation and repair of DNA, induce cell proliferation, transform cells and promote tumours. One study has indicated that DMA may cause cancer of the urinary bladder in male rats at high doses.

1.7 Effects on human health

Soluble inorganic arsenic is acutely toxic, and ingestion of large doses leads to gastrointestinal symptoms, disturbances of cardiovascular and nervous system functions, and eventually death. In survivors, bone marrow depression, haemolysis, hepatomegaly, melanosis, polyneuropathy and encephalopathy may be observed.

Long-term exposure to arsenic in drinking-water is causally related to increased risks of cancer in the skin, lungs, bladder and kidney, as well as other skin changes such as hyperkeratosis and pigmentation changes. These effects have been demonstrated in many studies using different study designs. Exposure–response relationships and high risks have been observed for each of these end-points. The effects have been most thoroughly studied in Taiwan but there is considerable evidence from studies on populations in other countries as well. Increased risks of lung and bladder cancer and of arsenic-associated skin lesions have been reported to be associated with ingestion of drinking-water at concentrations £ 50 µg arsenic/litre.

Occupational exposure to arsenic, primarily by inhalation, is causally associated with lung cancer. Exposure–response relationships and high risks have been observed. Increased risks have been observed at cumulative exposure levels ³ 0.75 (mg/m³) × year (e.g. 15 years of exposure to a workroom air concentration of 50 µg/m³). Tobacco smoking has been investigated in two of the three main smelter cohorts and was not found to be the cause of the increased lung cancer risk attributed to arsenic; however, it was found to be interactive with arsenic in increasing the lung cancer risk.

Even with some negative findings, the overall weight of evidence indicates that arsenic can cause clastogenic damage in different cell types with different end-points in exposed individuals and in cancer patients. For point mutations, the results are largely negative.

Chronic arsenic exposure in Taiwan has been shown to cause blackfoot disease (BFD), a severe form of peripheral vascular disease (PVD) which leads to gangrenous changes. This disease has not been documented in other parts of the world, and the findings in Taiwan may depend upon other contributing factors. However, there is good evidence from studies in several countries that arsenic exposure causes other forms of PVD.

Conclusions on the causality of the relationship between arsenic exposure and other health effects are less clear-cut. The evidence is strongest for hypertension and cardiovascular disease, suggestive for diabetes and reproductive effects and weak for cerebrovascular disease, long-term neurological effects, and cancer at sites other than lung, bladder, kidney and skin.

1.8 Effects on other organisms in the environment

Aquatic and terrestrial biota show a wide range of sensitivities to different arsenic species. Their sensitivity is modified by biological and abiotic factors. In general, inorganic arsenicals are more toxic than organoarsenicals and arsenite is more toxic than arsenate. The mode of toxicity and mechanism of uptake of arsenate by organisms differ considerably. This may explain why there are interspecies differences in organism response to arsenate and arsenite. The primary mechanism of arsenite toxicity is considered to result from its binding to protein sulfhydryl groups. Arsenate is known to affect oxidative phosphorylation by competition with phosphate. In environments where phosphate concentrations are high, arsenate toxicity to biota is generally reduced. As arsenate is a phosphate analogue, organisms living in elevated arsenate environments must acquire the nutrient phosphorous yet avoid arsenic toxicity.

Arsenic compounds cause acute and chronic effects in individuals, populations and communities at concentrations ranging from a few micrograms to milligrams per litre, depending on species, time of exposure and end-points measured. These effects include lethality, inhibition of growth, photosynthesis and reproduction, and behavioural effects. Arsenic-contaminated environments are characterized by limited species abundance and diversity. If levels of arsenate are high enough, only species which exhibit resistance may be present.

2. PROPERTIES AND ANALYTICAL PROCEDURES

2.1 Identity

Elemental arsenic (As) is a member of Group 15 of the periodic table, with nitrogen, phosphorus, antimony and bismuth. It has an atomic number of 33 and an atomic mass of 74.91. The Chemical Abstract Service (CAS), National Institute for Occupational Safety and Health Registry of Toxic Effects of Chemicals (RTECS), Hazardous Substances Data Bank (HSDB), European Commission, and UN transport class numbers are 7440-38-2, HSB 509, CG 05235 000, 033-001-00-X and UN 1558, respectively.

This monograph deals with arsenic and inorganic and organic arsenic compounds, except arsine (AsH₃), for which a Concise International Chemical Assessment Document (CICAD) is being prepared.

2.2 Chemical and physical properties of arsenic compounds

Arsenic is a metalloid widely distributed in the earth’s crust. It can exist in four valency states; –3, 0, +3, and +5. In strongly reducing environments, elemental arsenic and arsine (–3) can exist. Under moderately reducing conditions, arsenite (+3) may be the dominant form, but arsenate (+5) is generally the stable oxidation state in oxygenated environments.

Arsenic and its compounds occur in crystalline, powder, amorphous or vitreous forms. They usually occur in trace quantities in all rock, soil, water and air. However, concentrations may be higher in certain areas as a result of weathering and anthropogenic activities including metal mining and smelting, fossil fuel combustion and pesticide use.

Arsenical salts exhibit a range of aqueous solubilities depending on the pH and the ionic environment.

There are many arsenic compounds of environmental importance. Representative marine arsenic-containing compounds, of which some are found in terrestrial systems, are shown in Table 1; their molecular structures are shown Fig. 1. Other arsenic compounds discussed in the text are listed in Table 2.

Table 1. Naturally occurring inorganic and organic As species
(see Fig. 1 for structures [1]–[22])

Speciation determines how arsenic compounds interact with their environment. For example, the behaviour of arsenate and arsenite in soil differs considerably. Movement in environmental matrices is a strong function of speciation and soil type. In a non-absorbing sandy loam, arsenite is 5–8 times more mobile than arsenate (Gulens et al., 1979). Soil pH also influences arsenic mobility. At a pH of 5.8 arsenate is slightly more mobile than arsenite, but when pH changes from acidic to neutral to basic, arsenite increasingly tends to become the more mobile species, though mobility of both arsenite and arsenate increases with increasing pH (Gulens et al., 1979). In strongly adsorbing soils, transport rate and speciation are influenced by organic carbon content and microbial population. Both arsenite and arsenate are transported at a slower rate in strongly adsorbing soils than in sandy soils.

Table 2. Other As compounds of environmental significance referred to in the text

Under oxidizing and aerated conditions, the predominant form of arsenic in water and soil is arsenate. Under reducing and waterlogged conditions (< 200 mV), arsenites should be the predominant arsenic compounds. The rate of conversion is dependent on the Eh and pH of the soil as well as on other physical, chemical and biological factors.

In brief, at moderate or high Eh, arsenic can be stabilized as a series of pentavalent (arsenate) oxyanions, H₃AsO₄, H₂AsO₄^–, HAsO₄^2– and AsO₄^3–. However, under most reducing (acid and mildly alkaline) conditions, arsenite predominates. A pH and Eh diagram is shown in Fig. 2.

2.3 Analytical procedures

Historically, colorimetric and gravimetric methods have been used for the determination of arsenic. However, these methods are either semi-quantitative or lack sensitivity. In recent years, atomic absorption spectrometry (AAS) has become the method of choice, as it offers the possibility of selectivity and sensitivity in the detection of a wide range of metals and non-metals including arsenic. Popular methods for generating atoms for AAS are flame and electrothermally heated graphite furnaces. However, a commonly used technique for the measurement of arsenic is the highly sensitive hydride generation atomic absorption spectrometric method (HGAAS). However, although it is suitable for total arsenic determination after appropriate digestion the technique is only routinely used to speciate a limited number of compounds – arsenite, arsenate, MMA, DMA, trimethylarsine oxide (TMAO).

Hydride generation followed by cryogenic trapping and AAS detection is a relatively inexpensive technique for the speciation of inorganic arsenic and its methylated metabolites (Ng et al., 1998a), although more expensive hyphenated techniques may also be used.

A number of other approaches have been reported for speciation of arsenic. Inductively coupled plasma-mass spectrometry (ICP-MS) offers very high sensitivity for the determination of arsenic, and coupled with HPLC enables equally sensitive estimation of a wide variety of arsenic species.

2.4 Sample preparation and treatment

2.4.1 Sampling and collection

Care must be taken to avoid contamination and prevent speciation changes during sample collection and storage. Plastic containers should be acid washed and traces of oxidizing and reducing agents avoided to preserve the oxidation state of arsenic compounds. Freezing samples to –80 °C has also been recommended (Crecelius, 1986). Concentrated hydrochloric acid (1 ml to 100 ml urine) has been added to urine to prevent bacterial growth (Concha et al., 1998a).

For particulates in air and aerosols sampling, various types of filter have been employed including polytetrafluoroethylene (Rabano et al., 1989), cellulose ester (Yager et al., 1997), glass microfibre (Beceiro-Gonzalez et al.,1997) and filter paper (Tripathi et al., 1997).

2.4.2 Oxidative digestion

Acid digestion (George & Roscoe, 1951) and dry ashing (George et al., 1973) are the two basic methods which have been widely employed for oxidative digestion of samples before analysis. In more recent years, microwave-assisted digestion has been used (Le et al., 1994b; Thomas et al., 1997). For analysis of biological soft tissues by ICP techniques, a simple partial digestion in a closed vessel at low temperature and pressure is often sufficient for the sample preparation and pretreatment step.

2.4.3 Extraction

For speciation of arsenic, solvent extraction is often required before analysis. For example, arsenite and arsenate in soil can be speciated after a hydrochloric acid and chloroform extraction procedure (Chappell et al., 1995; Ng et al., 1998b). Water has been used for the extraction of soluble arsenic compounds from soil with the aid of ultrasonic treatment (Hansen et al., 1992). Forms of arsenic compounds can also be separated by sequential extractions based on procedures described by Tessier et al. (1979). Aqueous methanol has been widely used for the extraction of organic arsenic species (Edmonds & Francesconi, 1981a; Shiomi et al., 1988a; Shibata et al., 1996; Kuehnelt et al., 1997). Yu & Wai (1991) and Laintz et al. (1992) described the use of sodium bis(trifluoroethyl) dithiocarbamate (NaFDDC) as a selective chelation reagent of arsenic followed by either a gas chromatograph (GC) detection or supercritical fluid chromatography (SFC) detection. The former gave a limit of detection of 10 µg As/litre in water and the latter gave similar sensitivity after 100–1000-fold preconcentration of the chelate complex in organic solvent.

2.4.4 Supercritical fluid extraction

There are very few publications on the use of supercritical fluid extraction (SFE) for the determination of arsenic. Wenclawiak & Krah (1995) reported a procedure for the measurement of arsenic species using SFE followed by GC or SFC detection. The authors described a rapid extraction of organic and inorganic arsenic species from spiked sand and soil samples by SFE with on-line derivatization using thioglycollic acid methylester (TGM) under supercritical conditions. The TGM derivatives are thermally stable, which makes them amenable to GC–SFC determination. The extracts were chromatographed without further clean-up steps. The limits of detection were 1 ng As/µl and 3 ng As/µl injection for DMA-TGM and MMA-TGM respectively.

2.5 Macro-measurement

Vogel (1954) described the historic Marsh test, a qualitative method based on the generation of arsine (AsH₃) by the addition of Zn granules to sulfuric acid. If the gas is mixed with hydrogen, and conducted through a heated glass tube, it decomposes into hydrogen and metallic arsenic which is deposited as a brownish-black "mirror" just beyond the heated part of the tube.

2.6 Colorimetric methods

George & Roscoe (1951) reported a spectroscopic emission measurement of the blue complex formed by the reaction of ammonium molybdate and hydrazine sulfate with arsenic in various biological materials. The sensitivity was about 0.01 µg.

George et al. (1973) carried out a collaborative study for a colorimetric measurement of arsenic in poultry and swine tissues using silver diethyldithiocarbamate (AgDDTC) as the complexing agent. The sensitivity was 0.1 mg/kg in tissues. Dhar et al. (1997) reported a detection limit of 0.04 mg/litre with 95% confidence limit using AgDDTC in chloroform with hexamethylenetetramine.

Gutzeit’s test (Vogel, 1954) is based on the generation of arsine from arsenic compounds by the addition of zinc granules to concentrated sulfuric acid. The arsine can be detected by means of a strip of filter paper moistened with silver nitrate or mercuric chloride. The arsine reacts with silver nitrate to give a grey spot, and with mercuric chloride to give a yellow to reddish-brown spot. The sensitivity is about 1 µg. A modification of this method, using mercuric bromide, is found in a test kit currently being used in Bangladesh for groundwater testing which has a limit of detection of 50–100 µg/litre under field conditions.

2.7 Methods for total inorganic arsenic

Methods for the analysis of inorganic arsenic based on its conversion to arsenic trichloride or arsenic tribromide by treatment with 6 mol/litre hydrochloric acid or hydrobromic acid have been described. The arsenic trihalide is separated from the remaining organic arsenic either by distillation (Maher, 1983) or by solvent extraction (Brooke & Evans, 1981). The methods have been applied routinely to the measurement of inorganic arsenic in a variety of foodstuffs, including those of marine origin where any inorganic arsenic is a small percentage of the total arsenic present (Flanjak, 1982; Shinagawa et al., 1983).

2.8 Atomic spectrometry

Common flame atomic absorption spectrometric methods are flame AAS (FAAS), electrothermal AAS (ETAAS) and hydride generation AAS (HGAAS). FAAS is relatively less sensitive for the determination of arsenic than ETAAS and HGAAS. Its detection limit is usually in the range of sub-milligram quantities per litre, and therefore it has limited application, especially for biological samples.

ETAAS, referred to also as graphite furnace-AAS (GFAAS), is generally one of the most sensitive atomic spectroscopic methods. Julshamn et al. (1996) reported factors that are known to interfere with the GFAAS determination of arsenic. The study was carried out by four participating laboratories using five marine standard reference materials. A mixture of palladium and magnesium salts has been recommended as a chemical modifier to avoid nickel contamination of the graphite furnace. The use of a pyrolytically coated graphite furnace tube with the L’vov platform improves sensitivity. Larsen (1991) achieved characteristic masses of about 16 pg of arsenic for arsenate, monomethylarsonate, DMA, arsenobetaine, arsenocholine and tetramethylarsonium ion calculated from aqueous standard solutions.

HGAAS is probably the most widely used method for the determination of arsenic in various matrices. Most of the reported errors in the determination of arsenic by HGAAS with NaBH₄ can be attributed to variation in the production of the hydride and its transport into the atomizer. The reaction and atomization of arsine have been reviewed and discussed by Welz et al. (1990). The addition of a solution of l-cysteine to a sample before hydride generation eliminates interference by a number of transition metals in the generation of arsine from arsenite and arsenate (Boampong et al., 1988), and improves responses of arsine generated from MMA and DMA in the presence of arsenite and arsenate (Le et al., 1994a).

Holak & Specchio (1991) described the determination of total arsenic, arsenite and arsenate in foods by HGAAS after a chloroform extraction procedure. The recovery was > 80%. Similar methods (Chappell et al., 1995; Ng et al., 1998a) have been developed for arsenic speciation in soils. Ybanez et al. (1992) described a HGAAS determination of arsenic in dry ashed mussel products and reported a detection limit of 0.017 µg As/g with a precision of 3%.

HGAAS has been used for arsenic speciation of inorganic arsenic and its urinary metabolites, MMA and DMA, since 1973, when Braman & Foreback (1973) introduced a cold-trapping step into a basic hydride generation system. Since then a number of improvements have been made to this method (Crecelius, 1978; Buchet & Lauwerys, 1981; Van Cleuvenbergen et al., 1988). Ng et al. (1998b) described an optimized procedure for the speciation of arsenic metabolites in the urine of occupationally exposed workers and experimental animals with detection limits of 1, 1.3 and 3 ng per reaction of inorganic arsenic, MMA and DMA (equivalent to 0.25 µg/litre, 0.325 µg/litre, and 0.75 µg/litre respectively), using 4 ml of urine per reaction.

HGAAS has also been widely employed for analysis of arsenic in water (Chen et al., 1994; Chatterjee et al., 1995; Mandal et al., 1996; Dhar et al., 1997; Biswas et al., 1998). Hasegawa et al. (1994) published the first report of trivalent methyl arsenicals, namely monomethylarsonous acid [MMA(III)] and dimethylarsinous acid [DMA(III)], being found and measured in natural waters. Arsenious acid, MMA(III) and DMA(III) were separated from the pentavalent species by solvent extraction using diethylammonium diethyldithiocarbamate (DDDC) and determined by HGAAS after cold trapping and chromatographic separations. The detection limits were 13–17 pmol/litre and 110–180 pmol/litre for the trivalent and pentavalent species respectively.

Atomic fluorescence spectrometry (AFS) has recently been used for the detection of arsenic hydrides in the ultraviolet spectral region because of the small background emission produced by the relatively cool hydrogen diffusion flame (Gomez-Ariza et al., 1998). The use of cold vapour or hydride generation, together with intense light sources, allows very low detection limits to be achieved. For example, arsenic species in seawater have been measured using hydride generation and cold trapping, coupled with AFS detection at 193.7 nm (Featherstone et al., 1998). They found detection limits of 2.3, 0.9, 2.4 and 3.7 ng/litre for arsenite, arsenate, MMA and DMA respectively (in a 5 ml sample), with a precision of 3.5%.

2.9 ICP methodologies

The main advantages of ICP-MS over ICP-AES are lower detection limits (sub-nanogram to sub-picogram) with wide linear range and isotope analysis capability of high precision. The detection limits of ICP-AES are typically in the range of sub-micrograms to sub-nanograms.

ICP-MS is more susceptible to isobaric interferences arising from the plasma. For example, hydrochloric acid and perchloric acid are not desirable for sample preparation, because the chloride ions generated in the plasma combine with the argon gas to form argon chloride (ArCl). This has the same mass as arsenic (75) which could lead to error if not corrected. Therefore, whenever possible, only nitric acid should be used in sample preparation. Careful sample preparation is as important as the final measurement, and special care should be taken to avoid contamination and losses by volatilization, adsorption and precipitation.

2.10 Voltammetry

Voltammetric stripping methods are mostly based on the chemical reduction of As(V) to As(III) before the deposition step, because it has been generally assumed that As(V) is electrochemically inactive. Mercury and gold (or gold-plated) electrodes are most commonly used for the determination of arsenic.

Sadana (1983) used differential pulse cathodic stripping voltammetry (DPCSV) coupled to a hanging mercury drop electrode (HMDE) to determine arsenic in drinking-water in the presence of Cu²⁺ and reported a detection limit of 1 ng/ml and a relative standard deviation of 6.4%. Zima & van den Berg (1994) reported a detection limit of 3 nmol/litre in seawater. DPCSV was employed by Higham & Tomkins (1993) to determine arsenic in canned tuna fish. They evaluated a number of digestion procedures and found the best procedure gave 93–96% recovery. No detection limit was reported.

A gold electrode affords better sensitivity than a mercury electrode. Hua et al. (1987) reported an automated determination of total arsenic in seawater by flow constant-current stripping analysis with a gold film fibre electrode, in which As(V) in the sample was reduced to As(III) with potassium iodide; the detection limit was 0.15 µg/litre. The reduction of As(V) to As(III) can also be achieved by reaction with sulfur dioxide or hydrazinium chloride for use with a gold electrode or HMDE respectively (Esteban et al., 1994).

Huiliang et al. (1988) have shown that As(V) can be reduced to elemental arsenic provided that extremely low reduction potentials are used. They used this method to measure As(V) and total arsenic in seawater and urine. The detection limit was 0.1 µg/litre using constant-current stripping voltammetry on a gold-coated platinum-fibre electrode. Greulach & Henze (1995) developed a cathodic stripping voltammetric method for the determination of As(V) in water and stream sediment on the basis that As(V) can be reduced in perchloric acid solution containing d-mannitol, combined with the accumulation of arsenic by co-precipitation with copper on an HMDE. The detection limit was 4.4 µg/litre.

Pretty et al. (1993) developed an on-line anodic stripping voltammetry (ASV) flow cell coupled to ICP-MS for the determination of arsenic in spiked urine. The detection limit was 130 pg/ml and the recovery was 94–113%.

2.11 Radiochemical methods

Orvini et al. (1974) reported a combustion technique for sample preparation and determination of arsenic, selenium, zinc, cadmium and mercury by neutron activation analysis (NAA) in environmental matrices including a range of standard reference materials. The recoveries were 98–100%. Sharif et al. (1993) described a NAA technique for the determination of arsenic in eight species of marine fishes caught in the bay of Bengal, Bangladesh.

Haddad & Zikovsky (1985) measured several elements including arsenic in air from workroom welding fumes by NAA and reported a detection limit of 0.17 ± 0.07 µg/m³. Landsberger & Wu (1995) reported the use of NAA to measure arsenic from environmental tobacco smoke in indoor air with a detection limit of 0.2 ng.

Chutke et al. (1994) described a radiochemical solvent extraction procedure for the determination of arsenite using an arsenic-76 tracer. The procedure is based on the complexation of arsenite with toluene-3,4-dithiol (TDT) at pH 2 and subsequent extraction in benzene. This isotopic dilution technique was employed to measure arsenic in a range of standard and certified reference materials. The detection limit was 250 ng with an accuracy of about 4% error and 170 ng with about 12% error.

2.12 X-ray spectroscopy

Particle-induced X-ray emission spectrometry (PIXES) is an analytical technique that entails the bombardment of a sample (target) with charged particles, resulting in the emission of characteristic X-rays of the elements present. PIXES is a multi-elemental technique with a detection limit of approximately 0.1 µg As/g. It has the advantage of using small samples (1 mg or less) and being a non-destructive technique. Applications of PIXES in the environmental field have mostly focused on atmospheric particulate material (aerosol samples) (Maenhaut, 1987).

Castilla et al. (1993) described the determination of arsenite and arsenate by X-ray fluorescence (XRF) spectroscopy in water with a detection limit of 3.1 ng/g. The recovery was 97 ± 2.1% and 103 ± 1.4% for arsenite and arsenate respectively. In this method, the water sample was acidified to pH 2 and arsenite co-precipitated with sodium dibenzyldithiocarbamate (DBDTC). Arsenate in the filtrate was then reduced to arsenite with potassium iodide before the co-precipitation step for the XRF measurement.

Although there are a variety of methods to determine the concentration and oxidation states of arsenic in coal and ash, there have been few attempts to determine the mineral forms of arsenic. Huffman et al. (1994) described the use of X-ray absorption fine structure (XAFS) spectroscopy and its capability of providing speciation information at realistic concentrations of 10–100 mg/kg. They identified arsenic present as arsenopyrite in one coal sample and as aluminosilicate slag and calcium orthoarsenate in combustion ashes.

2.13 Hyphenated techniques

Hyphenated techniques is a term referring to the coupling of more than two instrumental systems to form a single technique.

The combination of chromatographic separation with element-specific spectrometric detection has been proved to be particularly useful for the speciation of arsenic compounds at trace levels in environmental samples. Woller et al. (1995) used AFS detection in combination with ultrasonically nebulized liquid chromatography (LC) for on-line speciation of arsenic, but found that the technique had limited sensitivity owing to matrix interferences. More recently, Slejkovec et al. (1998) used LC and purge-and-trap GC interfaced with AFS to separate and quantify six arsenic species with detection limits of 0.5 ng/ml As (100 µl). Gomez-Ariza et al. (1998) coupled anion-exchange HPLC, hydride generation and AFS to achieve detection limits of 0.17, 0.45, 0.30 and 0.38 µg/litre for arsenite, DMA, MMA and arsenate respectively (using a 20 µl loop). Arsenobetaine was also determined by introducing an on-line photo-oxidation step after the chromatographic separation.

Ebdon et al. (1988) described a number of coupled chromatograph–atomic spectrometry methods for arsenic speciation including GC or HPLC with detection by atomic spectrometry, namely FAAS, flame atomic fluorescence spectrometry (FAFS) and ICP-AES. The FAAS system is capable of detection at less than 1 µg/kg (0.22–0.55 ng absolute for different species) when levels permit; HPLC–hydride generation–FAAS is probably the simplest routine method and HPLC–hydride generation–ICP-AES is preferred for multi-elemental analysis. HPLC–ICP-AES has been employed for the speciation of organic arsenic of aquatic origin (Francesconi et al., 1985; Gailer & Irgolic, 1996). Gjerde et al. (1993) described the coupling of microbore columns with direct-injection nebulization to ICP-AES and reported a detection limit of 10 µg/litre (100 pg). Microbore HPLC has the advantage of analysing small sample size using low flow rates (80–100 µl/min) of mobile phases.

Numerous methods (Shum et al., 1992; Larsen et al., 1993; Magnuson et al., 1996; Thomas et al., 1997; Le & Ma, 1997) have been developed for the speciation of arsenic using the separation power of chromatography coupled to the sensitivity of ICP-MS detection. Heitkemper et al. (1989) described an anion-exchange HPLC–ICP-MS method for the speciation of arsenite, arsenate, MMA and DMA in urine with absolute detection limits ranging from 36 to 96 pg (corresponding to 0.7–1.9 µg/litre in a 50 µl injection). Beauchemin et al. (1989) reported detection limits for arsenic species in DORM-1 (a dogfish muscle certified reference material) ranging between 50 and 300 pg using ion pairing and ion exchange HPLC-ICP-MS. Anion exchange is more tolerant because of the higher buffering capacity of the mobile phase. Cation pairing is more suitable for the determination of DMAA and arsenobetaine in biological samples containing high concentrations of salts. Pergantis et al. (1997) analysed and speciated animal feed additives using microbore HPLC–ICP-MS with detection limits ranging from 0.1 to 0.26 pg. Hakala and Pyy (1992) described an ion-pairing HPLC-HGAAS method for speciation of arsenite, arsenate, MMA and DMA in urine with detection limits of 1.0, 1.6, 1.2 and 4.7 µg/litre respectively.

Ding et al. (1995) described the coupling of micellar liquid chromatography (MLC) and ICP-MS for the speciation of arsenite, arsenate, MMA and DMA with detection limits of 90 pg for DMA and 300 pg for the other species. MLC is a type of chromatography that uses surfactants in aqueous solutions, well above their critical micelle concentration, as alternative mobile phases for reversed-phase liquid chromatography (RPLC). MLC extends the analyte candidates to almost all hydrophobic and many hydrophilic compounds providing they can partition to the micelles. Other advantages of MLC over RPLC include simultaneous separation of both ionic and non-ionic compounds, faster analysis times and improved detection sensitivity and selectivity.

Capillary electrophoresis (CE) is a versatile technique for the separation of a variety of analytes ranging from small inorganic ions to large biomolecules such as proteins and nucleic acids. CE-ICP-MS has been described for the speciation of arsenic by Liu et al. (1995) with detection limits of 100 pg arsenite/ml and 20 pg arsenate/ml and Olesik et al. (1995) with a detection limit of 8 µg/litre (1 pg injection).

Although techniques such as HPLC–ICP-MS and MLC–ICP-MS offer the advantages of high sensitivity and selectivity as well as low detection limits, species identification is based on the comparison of chromatographic retention times to those of available standards. When structure information is required, as well as quantification, electrospray HPLC–MS (Siu et al., 1991) and ionspray MS (Corr, 1997) should be considered. Corr & Larsen (1996) reported the use of LC–MS–MS for speciation of arsenic with a detection limit of 2 pg for the tetramethylarsonium cation.

3. SOURCES AND OCCURRENCE OF ARSENIC IN THE ENVIRONMENT

3.1 Natural sources

Arsenic is the main constituent of more than 200 mineral species, of which about 60% are arsenate, 20% sulfide and sulfosalts and the remaining 20% include arsenides, arsenites, oxides and elemental arsenic (Onishi, 1969). The most common of the arsenic minerals is arsenopyrite, FeAsS, and arsenic is found associated with many types of mineral deposits, especially those including sulfide mineralization (Boyle & Jonasson, 1973). The ability of arsenic to bind to sulfur ligands means that it tends to be found associated with sulfide-bearing mineral deposits, either as separate As minerals or as a trace of a minor constituent of the other sulfide minerals. This leads to elevated levels in soils in many mineralized areas where the concentrations of associated arsenic can range from a few milligrams to > 100 mg/kg.

Concentrations of various types of igneous rocks range from < 1 to 15 mg As/kg, with a mean value of 2 mg As/kg. Similar concentrations (< 1–20 mg As/kg) are found in sandstone and limestone. Significantly higher concentrations of up to 900 mg As/kg are found in argillaceous sedimentary rocks including shales, mudstone and slates. Up to 200 mg As/kg can be present in phosphate rocks (O’Neill, 1990).

Concentrations of arsenic in open ocean water are typically 1–2 µg/litre. The concentrations of arsenic in unpolluted surface water and groundwater are typically in the range of 1–10 µg/litre. Elevated concentrations in surface water and groundwater of up to 100–5000 µg/litre can be found in areas of sulfide mineralization (Welch et al., 1988; Fordyce et al., 1995). Elevated concentrations (> 1 mg As/litre) in groundwater of geochemical origins have also been found in Taiwan (Chen et al., 1994), West Bengal, India (Chatterjee et al., 1995; Das et al., 1995, 1996; Mandal et al., 1996) and more recently in most districts of Bangladesh (Dhar et al., 1997; Biswas et al., 1998). Elevated arsenic concentrations were also found in the drinking-water in Chile (Borgono et al., 1977); North Mexico (Cebrian et al., 1983); and several areas of Argentina (Astolfi et al., 1981; Nicolli et al., 1989; De Sastre et al., 1992). Arsenic-contaminated groundwater was also found in parts of PR China (Xinjiang and Inner Mongolia) and the USA (California, Utah, Nevada, Washington and Alaska) (Valentine, 1994). More recently, arsenic concentrations of < 0.98 mg/litre have been found in wells in south-western Finland (Kurttio et al., 1998). Levels as high as 35 mg As/litre and 25.7 mg As/litre have been reported in areas associated with hydrothermal activity (Kipling, 1977; Tanaka, 1990).

In nature, arsenic-bearing minerals undergo oxidation and release arsenic to water. This could be one explanation for the problems of arsenic in the groundwater of West Bengal and Bangladesh. In these areas the groundwater usage is very high. It has been estimated that there are about 4–10 million tube wells in Bangladesh alone. The excessive withdrawal and lowering of the water table for rice irrigation and other requirements lead to the exposure and subsequent oxidation of arsenic-containing pyrite in the sediment. As the water table recharges after rainfall, arsenic leaches out of the sediment into the aquifer.

However, recent studies seem to favour the reduction of Fe/As oxyhydroxides as the source for arsenic contamination in groundwater (Nickson et al., 1998; BGS, 2000; BGS & DPHE, 2001). Arsenic forms co-precipitates with ferric oxyhydroxide. Burial of the sediment, rich in ferric oxyhydroxide and organic matter, has led to the strongly reducing groundwater conditions. The process has been aided by the high water table and fine-grained surface layers which impede the penetration of air to the aquifer. Microbial oxidation of organic carbon has depleted the dissolved oxygen in the groundwater. The highly reducing nature of the groundwater explains the presence of arsenite (< 50%) in the water. The "pyrite oxidation" hypothesis is therefore unlikely to be a major process, and the "oxyhydroxide reduction" hypothesis (Nickson et al., 1998; Acharyya et al., 1999) is probably the main cause of arsenic contamination in groundwater. Although the oxyhydroxide reduction hypothesis requires further validation, there is no doubt that the source of arsenic in West Bengal and Bangladesh is geological, as none of the explanations for anthropogenic contamination can account for the regional extent of groundwater contamination. During the past 30 years the use of phosphate fertilizers has increased threefold in this region. The widespread withdrawal of groundwater may have mobilized phosphate derived from fertilizers and from the decay of natural organic materials in shallow aquifers. The increase in phosphate concentration could have promoted the growth of sediment biota and the desorption of arsenic from sediments, and the combined microbiological and chemical process might have increased the mobility of arsenic (Acharyya et al., 1999).

Marine organisms naturally accumulate considerable quantities of organic arsenic compounds. In marine animals the bulk of this arsenic is present as arsenobetaine, whereas marine algae contain most of the arsenic as dimethylarsinoylribosides. Humans are therefore exposed to these arsenic compounds through any diet that includes seafoods. This subject is fully discussed in Chapter 4.

Some arsenic compounds are relatively volatile and consequently contribute significant fluxes in the atmosphere. It has been estimated that the atmospheric flux of As is about 73 540 tonnes/year of which 60% is of natural origin and the rest is derived from anthropogenic sources (Chilvers & Peterson, 1987). Volcanic action is the next most important natural source of arsenic after low-temperature volatilization, and on a local scale it will be the dominant atmospheric source.

3.2 Sources of environmental pollution

3.2.1 Industry

It has long been recognized that the smelting of non-ferrous metals and the production of energy from fossil fuel are the two major industrial processes that lead to anthropogenic arsenic contamination of air, water and soil. Other sources of contamination are the manufacture and use of arsenical pesticides and wood preservatives.

Smelting activities generate the largest single anthropogenic input into the atmosphere (Chilvers & Peterson, 1987).

Tailings from metal-mining operations are a significant source of contamination, and can lead to contamination of the surrounding topsoils, and, because of leaching, sometimes the groundwater too. It has been estimated that several billion tons of tailings waste exist in the USA alone (Wewerka et al., 1978). As sulfur is often present in these tailings, exposure to the atmosphere in the presence of water leads to the production of an acid solution that can leach many elements including arsenic.

Elevated concentrations of arsenic in acid sulfate soils in Canada and New Zealand are associated with pyrite (Dudas, 1987). Concentrations of arsenic < 0.5% through lattice substitution of sulfur in this pyrite iron-rich bauxite have been recorded.

In the United Kingdom, the estimated arsenic releases (Hutton & Symon, 1986) were 650 tonnes/year from the non-ferrous metal industry, 9 tonnes/year emission into the atmosphere and 179 tonnes/year to landfill from iron and steel production, and 297 tonnes/year into the atmosphere and 838 tonnes/year to landfill from fossil fuel combustion. In 1996, the estimated total releases of arsenic to the air in the UK were 50 tonnes (DG Environment, 2000).

The working group of the European Union DGV (the directorate with responsibility for the environment) concluded that there were large reductions in the emissions of arsenic to air in several member countries of the European Union in the 1980s and early 1990s. In 1990, the total emissions of arsenic to the air in the member states were estimated to be 575 tonnes, of which 492 tonnes came from stationary combustion (mainly coal and oil combustion) and 77 tonnes from production processes, mainly from the iron and steel industry (35 tonnes) and the non-ferrous metal industry (31 tonnes) (DG Environment, 2000).

Arsenic is present in the rock phosphate used to manufacture fertilizers and detergents. In 1982, the United Kingdom imported 1324 × 10³ tonnes of rock phosphate with an estimated arsenic burden of 10.2 tonnes (Hutton & Symon, 1986).

3.2.2 Past agricultural use

In 1983, arsenical pesticides were one of the largest classes of biocontrol agent in the USA (Woolson, 1983). From the 1960s there was a shift, in herbicide use, from inorganic compounds (including lead and calcium arsenate and copper acetoarsenite) to inorganic and organic compounds (arsenic acid, sodium arsenate, MMA and DMA). Use of total arsenical pesticides, excluding wood preservatives, at the time of publication (1983) was estimated at 7–11 × 10³ tonnes As/year. Annual historical applications of lead arsenate to orchards in the USA ranged from 32 to 700 kg As/ha. Residues in orchard soils as high as 2500 mg/kg have been reported, but they are more commonly in the range of 100–200 mg/kg. In Australia between 1900 and 1950 As₂O₃ was widely used for controlling cattle ticks (Boophilus microplus), resulting in widespread arsenic contamination (Seddon, 1951).

3.2.3 Sewage sludge

The levels of arsenic in sewage sludge reflect the extent of industrialization of the area served by the local sewage system. Significant quantities may be added by arsenic-contaminated wastewater runoff derived from sources including atmospherically deposited arsenic, residues from pesticide usage, phosphate detergents and industrial effluent, particularly from the metal-processing industry. Levels of 0–188 mg As/kg dry weight have been reported in the United Kingdom (Woolson, 1983). Zhu & Tabatabai (1995) reported levels of 2.4–39.6 mg As/kg with a mean of 9.8 for sewage sludges from waste treatment plants in Iowa, USA.

O’Neil (1990) estimated that in the UK as a whole about 2.5 tonnes As/year is added to the agricultural land by use of sludge, compared to 6.1 tonnes As/year when phosphate fertilizer is used.

3.3 Uses

Arsenic is produced commercially by reduction of As₂O₃ with charcoal. As₂O₃ is produced as a by-product of metal-smelting operations. It is present in flue dust from the roasting of ores, especially those produced in copper smelting. In the 1960s, the pattern of use for As₂O₃ in the USA is believed to have been 77% as pesticides, 18% as glass, 4% as industrial chemicals and 1% as medicine. However, the pattern has changed over the years as the use of arsenic compounds for timber treatment has been increasingly popular since the late 1980s. Worldwide usage in the early 1980s was estimated to be 16 000 tonnes As/year as a herbicide, 12 000 tonnes As/year as a cotton desiccant/defoliant and 16 000 tonnes As/year in wood preservative (Chilvers & Peterson, 1987). By 1990, the estimated end-use of arsenic in the USA was 70% in wood preservatives, 22% in agricultural chemicals, 4% in glass, 2% in non-ferrous alloys and 2% in other uses including semiconductors (US DOI, 1991). Arsenic pentoxide and As₂O₃ are used as additives in alloys, particularly with lead and copper; arsenic and As₂O₃ are used in the manufacturing of low-melting glasses. High-purity arsenic metal and gallium arsenide are used in semiconductor products. Fowler’s solution (1% potassium arsenite solution) was used as a medication (Cuzick et al., 1992). As₂O₃ has been used for the treatment of acute promyelocytic leukaemia (Soignet et al., 1998).

Hutton & Symon (1986) reported that about 5000 tonnes/year As₂O₃ is imported to the United Kingdom for conversion to other arsenic compounds. These processes result in an estimated discharge of 87 tonnes As/year in manufacturing sludges on landfilled sites. Currently about 500 tonnes As/year is utilized in copper chrome arsenate (CCA) timber treatment, of which at most 5 tonnes/year is retained in sludges. Small amounts of arsenic are used in the production of glass, and most of the remainder is re-exported.

4. ENVIRONMENTAL TRANSPORT AND DISTRIBUTION

4.1 Transport and distribution between media

4.1.1 Air

Arsenic is primarily emitted into the atmosphere by high-temperature processes such as coal-fired power generation, smelting, burning vegetation and vulcanism. Natural low-temperature biomethylation and microbial reduction also release arsenic into the atmosphere; microorganisms can form volatile methylated derivatives of arsenic under both aerobic and anaerobic conditions, and can reduce arsenic compounds to release arsine gas (Cheng & Focht, 1979; Tamaki & Frankenberger, 1992) (see section 4.2.2). Arsenic is released into the atmosphere primarily as As₂O₃ or, less frequently, as one of several volatile organic compounds. Arsenic released to air exists mainly in the form of particulate matter (Coles et al., 1979). These particles are dispersed by the wind to a varying extent, depending on their size, and the particles are returned to the earth by wet or dry deposition. Arsines that are released from microbial sources in soils or sediments undergo oxidation in the air, reconverting the arsenic to less volatile forms that settle back to the ground (Wood, 1974; Parris & Brinckman, 1976).

Pacyna et al. (1989) studied atmospheric transport of arsenic from various sources in Europe to selected receptor sites in Norway. By modelling long-range transport they were able to calculate a dry deposition velocity for arsenic of 0.4 cm/second. Scudlark & Church (1988) measured arsenic in acid precipitation on the mid-Atlantic coast of the USA during 1985 and 1986. They calculated the total annual arsenic deposition rate to range from 38 to 266 µg/m², with dry deposition estimated to comprise 29–55% of the total. Davidson et al. (1985) calculated the annual dry deposition flux of arsenic to the Olympic National Park, Washington (USA) to range from 76.7 to 208 µg/m2. The average annual wet deposition of arsenic at Chesapeake bay (Maryland, USA) was found to be 49 µg/m2 (Scudlark et al., 1994).

Total atmospheric arsenic emissions from both natural and anthropogenic sources have been estimated to be 31× 10⁹ g/year, and total atmospheric arsenic removal was estimated to be 30–50 × 10⁹ g/year. The global tropospheric residence time of arsenic appears to be about 9 days (Walsh et al., 1979). Nakamura et al. (1990) estimated global atmospheric emissions into the atmosphere and deposition of arsenic. Total emissions were estimated at 36 × 10⁹ g/year, with the major source of atmospheric arsenic being anthropogenic emissions; the major natural source of arsenic was volcanic activity. Emissions from anthropogenic sources were estimated at 24 × 10⁹ g/year, representing 64% of total arsenic influxes. Depositions from the atmosphere to the land and the oceans were estimated at 24 × 10⁹ g/year and 9 × 10⁹ g/year respectively. Akeredolu et al. (1994) calculated the total annual transport of arsenic into the Arctic atmosphere at 285 t (285 × 10⁶ g) on the basis of a chemical transport modelling approach previously used for sulfur.

Arsenic in the atmosphere exists primarily adsorbed to particulate matter and mostly to particles < 2 µm in diameter (Coles et al., 1979). Waslenchuk (1978) found that atmospheric arsenate at the continental shelf of the south-eastern USA is associated exclusively with the particulate fraction. Rabano et al. (1989) collected size-fractionated aerosol samples at an urban site during 1987. A greater proportion (75%) of the arsenic was observed in the fine particles (< 2.5 µm). The As(III)/As(V) ratio for both fine and coarse (> 2.5 µm) particles was approximately 1. Similarly, Waldman et al. (1991) reported that 65% of the arsenic in aerosol samples collected at an urban site (China) was associated with fine particles (< 2.5 µm). Kelley et al. (1995) monitored arsenic in aerosol collected from the Kola Peninsula (Russia). They found 68% of arsenic associated with fine particles (< 1 µm), 26% with coarse particles (1–10 µm) and 7% with large particles (> 10 µm). The atmospheric residence time of particulate-bound arsenic depends on particle size and meteorological conditions, but a typical value is about 9 days (US EPA, 1982).

4.1.2 Freshwater and sediment

The dissolved forms of arsenic in the water column include arsenate, arsenite, monomethylarsonic acid (MMA) and dimethylarsinic acid (DMA) (Braman & Foreback, 1973). Some As(III) and As(V) species can interchange oxidation states depending on Eh, pH and biological processes (Ferguson & Gavis, 1972). Some arsenic species have an affinity for clay mineral surfaces and organic matter, and this can affect their environmental behaviour. Methylation and demethylation reactions are also important transformations controlling the mobilization and subsequent distribution of arsenicals (Mok & Wai, 1994). Transport and partitioning of arsenic in water depends on the chemical form of the arsenic and on interactions with other materials present. Arsenic may be adsorbed from water on to clays, iron oxides, aluminium hydroxides, manganese compounds and organic material (Callahan et al., 1979; Welch et al., 1988). The distribution and transport of arsenic in sediment is a complex process that depends on water quality, native biota and sediment type. There is a potential for arsenic release when there is fluctuation in Eh, pH, soluble arsenic concentration and sediment organic content (Abdelghani et al., 1981).

Ferguson & Gavis (1972) proposed an arsenic cycle for a stratified lake. In the aerobic epilimnetic water, reduced forms of arsenic tend to be oxidized to arsenate, which co-precipitates with ferric oxyhydroxide. Turbulent dispersion and convection transports some of the arsenate across the thermocline to the oxygen-depleted hypolimnion, where reduction to HAsO₂ and AsS₂^– takes place, depending on the sulfur concentration and the Eh. Co-precipitation, adsorption and epitaxial crystal growth cause arsenic to be removed to the sediments, where reduction of ferric iron, arsenate and arsenite result in either solubilization or stabilization as an insoluble sulfide or arsenic metal. Microbial reduction and methylation to arsine solubilize the arsenic (see section 4.2), and diffusion through the sediments or mixing by currents or burrowing organisms (see section 4.1.4) cause arsenic to re-enter the water column.

Aurilio et al. (1994) studied the speciation and fate of arsenic in three lakes of the Aberjona watershed (Massachusetts, USA). Speciation appeared to be controlled by reduction, methylation, and oxidation processes, and by adsorption to and desorption from particles. Biologically mediated reduction, at rates of 0.2–0.5% total arsenic/day, and methylation, at rates of 0.4–0.6% total arsenic/day, occurred in the mixed layers of these lakes. These processes are slow or even absent in the hypolimnion, however, allowing arsenate to accumulate in seasonally anoxic hypolymnetic waters. High micromolar concentrations of arsenic, predominantly arsenite, persisted in the saline, sulfidic monimolimnion of one lake.

Clement & Faust (1981) studied the release of arsenic from contaminated sediments. Anaerobic conditions led to aqueous levels of arsenic, principally as arsenite, about 10 times higher than concentrations reached with aerobic conditions. Under aerobic conditions arsenic in the overlying water comprised 70% arsenate and 20% organic arsenic. The authors found that adsorption–desorption equilibria and the amount of ‘available’ arsenic present in the sediment greatly influenced the soluble arsenic concentration found in the aqueous phase. In sediment under oxidized conditions arsenic solubility was low and 87% of the arsenic in solution was present as arsenate. On reduction, arsenite became the major arsenic species in solution and solubility increased (Masscheleyn et al., 1991b). Ahmann et al. (1997) identified rapid arsenic mobilization from aquatic sediments in upper Aberjona (Massachusetts, USA) sediment microcosms. The findings suggest that arsenic reduction by microorganisms may contribute to arsenic flux from anoxic sediments in this arsenic-contaminated watershed.

The predominant arsenic species in the water column of lakes is arsenate, as expected in oxidizing environments (Seyler & Martin, 1989). Arsenite is usually present and sometimes dominates in bottom water which contains high concentrations of Fe(II) and low oxygen. Peterson & Carpenter (1983) reported that the arsenate : arsenite concentration ratio was 15 : 1 in the oxic region of the water column and 1 : 12 in the anoxic zone. Seasonal trends reveal higher concentrations of arsenic in summer than in winter. The source of arsenic in the summer is most likely surface sediments that have become anoxic causing a release into the water column of arsenic adsorbed on iron and manganese oxides (Singh et al., 1988; Crecelius et al., 1994).

Pettine et al. (1992) found that arsenate was the predominant arsenic species in the river Po (Italy). The main factors affecting dissolved concentrations included flow and suspended matter concentration and biological activity. The ratio between oxidized and reduced species appears to be significantly influenced by iron and manganese oxides. Abdel-Moati (1990) monitored arsenic in the Nile delta lakes and found arsenate to be the dominant arsenic species (85–95%). Increased arsenite (14–33%) was found near local sewage discharge points. Dimethylarsenic was the dominant organic species, reaching 22% of the total dissolved arsenic.

A temporal study of arsenic speciation in Davis Creek Reservoir, a seasonally anoxic lake in northern California (USA), demonstrated that dimethylarsinic acid increased sufficiently to become the dominant form of dissolved arsenic within the surface photic zone during late summer and early autumn. Methylated forms decreased and arsenate increased when the lake ‘turned over’ in early December, suggesting a degradation of dimethylarsinic acid (Anderson & Bruland, 1991).

Aggett & O’Brien (1985) report that Lake Ohakuri (New Zealand) becomes stratified during the summer. During this period arsenic released from the sediment accumulates in the hypolimnion until turnover when it is mixed with epilimnetic water. It is estimated that this turnover effect causes a temporary increase in arsenic concentrations of 10–20%. Aggett & Roberts (1986) conclude that arsenate and phosphate are incorporated into Lake Ohakuri sediments by co-precipitation at the time of formation of the hydrous oxides rather than by adsorption on existing surfaces. Aggett & Kriegman (1988) show that in sediment cores from Lake Ohakuri over 90% of arsenic in interstitial waters was present as arsenite, an indication that reduction from arsenate, the predominant form adsorbed from the lake water, was taking place. When conditions at the sediment–water interface became anoxic, arsenite diffused across the interface into the hypolimnion.

Johnson & Thornton (1987) studied the seasonal variation of arsenic in the Carnon river, south-west England (UK). Approximately 85% of the arsenic was found to originate from mine waters. Arsenic is found to a large extent (~80%) in the particulate phase; the authors suggest that sorptive or co-precipitation processes are responsible for the regulation of dissolved concentrations of arsenic in these waters. These processes are largely independent of pH. Adsorption appears to be important in the removal of arsenic from solution, with 80% being removed on entering estuarine waters.

Both adsorption of arsenic on iron-rich oxides on the surface of the sediments and incorporation of arsenic into the sediments by co-precipitation with hydrous iron oxides are factors controlling mobilization of sediment arsenic. The major arsenic species leached was arsenate; release of arsenate was found to be pH dependent and related to the total iron and free iron oxides in the sediments (Mok & Wai, 1989). Arsenate and arsenite differ in adsorption characteristics, and this influences their mobilization and subsequent distribution during water–sediment interactions. The extent of adsorption and remobilization varies with the oxidation state of arsenic, the Eh and the pH of the water. The increase in mobility of arsenate under more reducing conditions is generally attributed to the reduction of Fe³⁺ to Fe(II), with subsequent release of arsenate, and reduction of arsenate to arsenite (Mok & Wai, 1994).

Brannon & Patrick (1987) found that arsenate added to sediment became associated with relatively immobile iron and aluminium compounds. Addition of arsenate to sediments before anaerobic incubation also resulted in accumulation of arsenite and organic arsenic in the interstitial water and exchangeable phases of anaerobic sediments. Seyler & Martin (1989) report that the presence of arsenic in the anoxic zone of a permanently stratified lake was due to adsorption on to iron and manganese.

Sorption of arsenate, MMA and DMA on anaerobic bottom sediments from the Menominee river, Wisconsin (USA) is described by Langmuir isotherms (Holm et al., 1979). Singh et al. (1988) found that the adsorption of arsenite from aqueous solution followed first-order adsorption expression obeying Langmuir’s model of adsorption. Similar findings were reported by Yadava et al. (1988) for adsorption of arsenite by china clay, with maximum adsorption at pH 8. Sediment adsorption of arsenate, monosodium methanearsonate (MSMA) and DMA was positively correlated with clay content (Wauchope & McDowell, 1984). Organic sediments adsorbed arsenic more strongly than sandy sediments (Faust et al., 1987b).

The extent of uptake and the rate of adsorption of arsenate decrease with an increase in temperature from 20 °C to 40 °C. The amount of arsenate adsorbed increases as the pH of the system increases and reaches its maximum at pH 4.2 for haematite and pH 6.2 for feldspar. The removal of arsenate from aqueous solution by adsorption on to geological materials such as haematite and feldspar follows first-order kinetics, and intraparticle diffusion seems to control the mass transfer (Prasad, 1994). The adsorption of arsenate on alumina, haematite, kaolin and quartz was influenced by the charge of the solid surface and the arsenic speciation in solution as determined by pH (Xu et al., 1988).

Adsorption of arsenate by fly ash was significantly greater at pH 4 than at pH 7 or 10 and was found to be almost irreversible. Adsorption fitted both the Freundlich and Langmuir adsorption models (Sen & De, 1987; Diamadopoulos et al., 1993). The partitioning of arsenic between acidic fly ash and leachate is controlled by sorption on iron oxyhydroxide. The leaching of arsenic is mainly controlled by sorption on hydroxylamine-extractable ("amorphous") iron oxyhydroxide; crystalline iron oxides appear to have little influence on the process (Van der Hoek & Comans, 1996). Thanabalasingam & Pickering (1986) found that arsenic sorption by humic acid varies with pH, adsorbate concentration and ash content of the substrate. At fixed pH, the amount of arsenic sorbed conformed to a Langmuir relationship, with calculated capacities in the region of maximum uptake (~pH 5.5) being of the order of 5250–6750 mg/kg for arsenite and 6750–8250 mg/kg for arsenate.

Gupta & Chen (1978) report that arsenic acid and arsenious acid species are effectively adsorbed in the pH range 4–7. Laboratory adsorption experiments indicated that arsenite is less effectively removed than arsenate. Adsorption of MMA and DMA on ferric oxyhydroxide and activated alumina decreased with increasing pH (4–11) (Cox & Ghosh, 1994).

Arsenic in porewater is controlled by the solubility of iron and manganese oxyhydroxides in the oxidized zone and metal sulfides in the reduced zone. Diagenetic sulfides are important sinks for arsenic in reduced, sulfidic sediments. During reduction, oxyhydroxides of iron and manganese dissolve, arsenic sulfides precipitate and arsenic is released to groundwater dominantly as arsenite (Moore et al., 1988). Therefore, mobilization of arsenic is more likely to occur in sediments low in iron and manganese oxyhydroxides, and calcium carbonate (Brannon & Patrick, 1987; Mok & Wai, 1990).

Bright et al. (1994) found that arsenite was the predominant arsenical in sediment porewater throughout a watershed receiving gold-mine effluent; dissolved arsenic in water column samples was mostly arsenate. Arsenic distribution in surficial sediments was controlled partially by the bulk movement of sediments, followed by burial with less-contaminated sediments in the upper reaches of the watershed. Particulate concentrations of arsenic contributed significantly (< 70%) to the total arsenic concentrations in the water column downstream of the gold-mine discharge. Azcue et al. (1994b) found that 66–83% of the arsenic in sediment porewater from a mine-polluted lake was arsenite. The concentration gradient of total dissolved arsenic indicated an upward diffusion of arsenic towards the water column, with the estimated annual fluxes being 0.8–3.8 µg/cm².

Mok & Wai (1990) reported that acid precipitation caused increased release of both arsenate and arsenite from contaminated sediments (pH 2.7). Arsenic release decreased with increasing pH; lowest levels of release were found at pH 8.3 for arsenite and pH 6.3 for arsenate. Release of arsenic increased at more alkaline pH values. Xu et al. (1991) concluded that environmental acidification would increase the leaching of arsenic from sediments to surface waters under reducing conditions as a result of the release of arsenite from iron oxyhydroxide phases, but could also reduce the mobility because of enhanced adsorption under oxidizing conditions. However, a large reduction in pH (to £ 4) would enhance the mobility of arsenic even under oxidizing conditions.

4.1.3 Estuarine and marine water and sediment

An arsenic cycle has also been outlined for the estuarine environment. Sanders (1980) found that the major inputs to the marine environment were river runoff and atmospheric deposition. Biological uptake caused changes in arsenic speciation resulting in measurable concentrations of reduced and methylated arsenic species. The overall cycle is similar to the phosphate cycle, but the regeneration time for arsenic is much slower. Arsenic flows into the estuary as arsenate and arsenite from river water and mine adits. There is oxidation of arsenite to arsenate, microbiological reduction of arsenate to arsenite and removal of arsenic by dilution with seawater and subsequent transport out of the estuary. Inorganic arsenic can be adsorbed on to charged particles of iron oxyhydroxides and manganese oxides and deposited as flocculated particles to sediment. There is subsequent release of dissolved arsenite and arsenate following the reduction and dissolution of the iron and manganese carrier phases in the anoxic sediments. Arsenate can be reduced, either microbially or chemically, to arsenite within the anoxic sediment, and arsenic (as arsenate or arsenite) can enter by sediment resuspension (Sanders, 1980; Knox et al., 1984). Studies on the pH dependence of arsenate and arsenite adsorption to soils and sediments and to minerals are not consistent. For example, greater adsorption of arsenate to fly ash occurred at pH4 than at pHs 7 and 10 (Sen & De, 1987; Diamadopoulos et al., 1993), whereas Mok & Wai (1990) reported that absorption increased as pH increased for sediments.

Arsenic entering unpolluted estuaries associated with particulates remains adsorbed, and accumulates in sediment. Remobilization has only a small effect (< 7%) on the dissolved arsenic concentration in the water column. Dissolved arsenic species form complexes with low-molecular-weight dissolved organic matter, and these tend to prevent adsorption and co-precipitation interactions between arsenic and flocculating iron oxyhydroxides and humics (Waslenchuk & Windom, 1978). Langston (1983) reports that more than 80% of arsenic entering Restronguet creek in southwest England (United Kingdom) was retained by the sediment, which consequently acts as a sink for riverine inputs and limits transport of dissolved species to coastal waters. Iron oxyhydroxide scavenging seems to be a predominant factor in the removal of arsenic from the Scheldt estuary (The Netherlands) (van der Sloot et al., 1985). Millward et al. (1997) estimated an annual arsenic budget for the Thames plume (United Kingdom) and found that cycling of arsenic by phytoplankton was the dominant process. Inorganic arsenic was removed from the water column by phytoplankton and recycled during phytoplankton degradation and consumption.

An arsenic budget for Puget sound (Washington, USA) revealed that sediments accumulate less than 30% of the arsenic entering the sound (Crecelius et al., 1975). Carpenter et al. (1978) found that sedimentation processes including adsorption–desorption reactions with natural Puget sound suspended matter remove less than 15% of the dissolved arsenic input, with iron oxyhydroxides dominating what removal does occur. Most of the arsenic entering the sound is removed by advection of surface waters out into the strait of Juan de Fuca . A similar budget for Lake Washington (USA) showed equal inputs of arsenic from the atmosphere and from rivers, and subsequent removal by outflowing water (45%) and by accumulation in the sediments (55%) (Crecelius, 1975).

Riedel (1993) studied the distribution of dissolved and solid arsenic species in contaminated estuarine sediment. Arsenite was the dominant dissolved and solid species in the deeper reduced sediment, and arsenate was dominant in the oxidized surface layer. Arsenite in the interstitial water diffused toward the surface layer, where it was mostly oxidized to arsenate.

Howard et al. (1988) found that the distribution of dissolved inorganic arsenic in an estuary appears to be determined by a combination of secondary inputs arising from old mine drainage and advective transport of arsenic-enriched sediment interstitial waters into the water column. Bioutilization of the element during the warmer months results in the release of dissolved monomethylarsenic and dimethylarsenic. Inorganic arsenite and methylated arsenic species can account for up to 41% and 70% of the dissolved arsenic respectively, but only when the water temperature exceeds 12 °C (Howard et al., 1984). Arsenate was the dominant form in a temperate estuary throughout the year except late winter when a dimethylarsenic species was dominant (Riedel, 1993).

Andreae (1978, 1979) monitored seawater samples from the northeast Pacific and southern Californian coast (USA). Methylation and reduction of arsenate to arsenite and methylarsenic acids occur in the photic zone. Arsenic is taken up by planktonic organisms in the surface waters and transported to deeper waters with biogenic debris. At intermediate levels regeneration of arsenate occurs. There was a good correlation between photosynthetic activity and concentration of methylated arsenicals. Andreae & Froelich (1984) and Sadiq (1990) found that arsenate is more abundant in oxic seawaters whereas arsenite is more abundant in anoxic seawaters.

Waslenchuk (1978) found that concentrations of arsenic species in continental shelf waters of the south-eastern USA are controlled mainly by simple mixing of shelf waters and Gulf Stream intrusions. Riverine and atmospheric arsenic inputs to the shelf waters were relatively insignificant, and uptake of arsenic by biota had only a minor effect on arsenic distribution.

Byrd (1988) studied the seasonal cycle of arsenic on the continental shelf of the South Atlantic. During periods of high winds in the winter and early spring, inorganic arsenic concentrations are reduced to as little as 20% of typical open-ocean concentrations by sorption on to suspended sediments or incorporation into phytoplankton. In the late summer and early autumn arsenic is remobilized and returned to the water column, elevating arsenic concentrations to 50% more than open-ocean concentrations. Belzile (1988) analysed vertical profiles of arsenic in cores from the Laurentian trough in the gulf of St Lawrence. The surface enrichment of solid arsenic and the increase of dissolved arsenic with depth suggested that the mobile portion of arsenic is associated with iron oxyhydroxides. It follows a redox pattern of dissolution in the suboxic zone, upwards diffusion, and precipitation near the sediment–water interface under non-steady-state conditions.

Nereis succinea, a burrowing polychaete, affected distribution and flux of arsenic from sediments by its production of irrigated burrows. These burrows increased both the effective surface area of the sediment and the diffusion of arsenic by a factor of five. Although physical suspension can produce large pulses of materials from contaminated sediments, it is the continuous biological activity that is likely to be more important in the mobilization of arsenic from sediments (Riedel et al., 1987).

Riedel et al. (1989) reported that three species of burrowing invertebrates (N. succinea, Macoma balthica and Micura leidyi) cause a measurable flux of arsenic out of contaminated sediments which was not measurable in the absence of fauna. Arsenic release from sediment was primarily arsenate and arsenite, with trace amounts of methylated arsenic compounds.

4.1.4 Soil

Arsenic from weathered rock and soil may be transported by wind or water erosion. However, because many arsenic compounds tend to adsorb to soils, leaching usually results in transportation over only short distances in soil (Moore et al., 1988; Welch et al., 1988). However, rainwater or snowmelt may leach soluble forms into surface water or groundwater, and soil microorganisms may reduce a small amount to volatile forms (arsines) (Woolson, 1977a; Richardson et al., 1978; Cheng & Focht, 1979; Turpeinen et al., 1999).

Under reducing conditions, arsenite dominates in soil (Deuel & Swoboda, 1972a; Haswell et al., 1985) but elemental arsenic and arsine can also be present (Walsh & Keeney, 1975). Arsenic would be present in well-drained soils as H₂AsO₄^– if the soil was acidic or as HAsO₄^2– if the soil was alkaline. Oxidation, reduction, adsorption, dissolution, precipitation and volatilization of arsenic reactions commonly occur in soil (Bhumbla & Keefer, 1994). In the porewater of aerobic soils arsenate is the dominant arsenic species, with small quantities of arsenite and MMA in mineralized areas.

The amount of arsenic sorbed from solution increases as the free iron oxide, magnesium oxide, aluminium oxide or clay content of the soil increases; removal of amorphous iron or aluminium components by treatment with oxalate eliminates or appreciably reduces the arsenic sorption capacity of the soil (Dickens & Hiltbold, 1967; Jacobs et al., 1970a; Galba, 1972; Wauchope, 1975; Livesey & Huang, 1981). Barry et al. (1995) examined the adsorption characteristics of a forest soil profile. The greatest sorption capacity for arsenic occurred at a depth of 30 cm in the profile, in the B2 horizon where there was a predominance of clay and oxyhydroxides of iron and aluminium. Adsorption of arsenic on soil colloids depends on the adsorption capacity and behaviour of these colloids (clay, oxides or hydroxides of aluminium, iron and manganese, calcium carbonates or organic matter). In general, iron oxides/hydroxides are the most commonly involved in adsorption of arsenic in both acidic and alkaline soils (Sadiq, 1997). Manning & Goldberg (1997) studied the adsorption of arsenic in three arid-zone soils. They found that the soil with the highest citrate–dithionite extractable iron and percentage of clay had the highest affinity for arsenite and arsenate and displayed adsorption behaviour similar to that of pure ferric oxide. Adsorption isotherms indicated that arsenate species adsorbed more strongly than arsenite.

The surfaces of aluminium oxides/hydroxides and clay may play a role in arsenic adsorption, but only in acidic soils. Carbonate minerals are expected to adsorb in calcareous soils (Sadiq, 1997), and Goldberg & Glaubig (1988) concluded that carbonates play a major role in arsenate adsorption at pH > 9. Phosphate substantially suppresses arsenate adsorption by soil, with the extent of the suppression varying from soil to soil (Livesey & Huang, 1981). Roy et al. (1986) found that the adsorption of arsenate was significantly reduced by competitive interactions with phosphate in three different soil types (clay, silt loam and ultisol). Darland & Inskeep (1997) found that phosphate effectively competed with arsenate for adsorption sites on sand in batch isotherms as well as in saturated transport studies. The phosphate competition was not, however, sufficient to desorb all of the applied arsenate either in simultaneously applied pulses, or in a column where arsenate was applied before a concentrated pulse of phosphate. Approximately 40% of the applied arsenate remained sorbed to the sand even after the total phosphate loading exceeded the column capacity by more than two orders of magnitude. The authors concluded that rates of arsenate desorption play an important role in transport of arsenate through porous media. Elkhatib et al. (1984) found that arsenite adsorption was not reversible, with only small amounts of sorbed arsenite released during subsequent desorption procedures. No significant correlation was found between arsenic adsorption and soil organic carbon or cation exchange capacity (CEC) (Hayakawa & Watanabe, 1982).

Jones et al. (1997) found that increased mobility of arsenic after liming appears to be consistent with the pH dependence of sorption reactions of arsenic on iron oxide minerals rather than dissolution–precipitation reactions of solid metal arsenates.

Sakata (1987) reports distribution coefficients (K_d) for arsenite for 15 subsurface soils from different sites in Japan with K_d values ranging from 75 to 1200. The distribution coefficient was significantly correlated with the extractable iron content of the soils.

Precipitation is another mechanism of arsenic removal from soil. Thermodynamic calculations showed that in acidic oxic and suboxic soils, iron arsenate may control arsenic solubility, whereas in anoxic soils, sulfides of arsenite may control the concentrations of the dissolved arsenic in soil solutions. In alkaline, acidic, oxic and suboxic soils, precipitation of both iron arsenate and calcium arsenate may limit arsenic concentrations in soil solutions (Sadiq et al., 1983; Sadiq, 1997). Carey et al. (1996) studied the sorption of arsenic in two free-draining sandy soils in New Zealand. They concluded that arsenate sorption occurred primarily through adsorption rather than a precipitation mechanism.

Many soil organisms are capable of converting arsenate and arsenite to several reduced forms, largely methylated arsines which are volatile (see section 4.2). Woolson (1977b) proposed that about 12% of the arsenic applied and present in a soil is lost through volatilization of alkylarsines each year. Woolson & Isensee (1981) report total losses of 14–15% per year from soil treated with sodium arsenite, DMA or MMA. Most of the loss was through volatilisation, although some apparent loss was caused by movement to or mixing with subsoil. Sandberg & Allen (1975) estimated an arsenic loss of 17–35% per year through volatilization. Sanford & Klein (1988) report that arsenic volatilization showed a direct relationship with nutrient levels and microbial growth in soil.

Leaching does not appear to be a significant route of arsenic loss from soil. Arsenic as MMA was applied to three soil types over a 6-year period. Percentage recovery of applied arsenic averaged 67%, 57% and 39% in a fine sandy loam, a silt loam and a sandy loam soil respectively. All of the arsenic recovered in the soils was detected in the ploughed layer (< 30 cm) with no evidence of leaching into deeper zones (Hiltbold et al., 1974). Elfving et al. (1994) monitored the movement of arsenic following the application of lead arsenate to fruit orchards for insect control. The rate of decrease in concentration of arsenic with depth was significantly greater in a sandy soil than in clay, suggesting that downward movement occurred less readily in the former. Peryea & Creger (1994) studied the vertical distribution of arsenic in six contaminated orchard soils. Most of the arsenic was restricted to the upper 40 cm, with maximum arsenic concentrations ranging from 57.8 to 363.8 mg/kg. Absolute soil enrichment with arsenic occurred to depths between 45 and > 120 cm, with arsenic concentrations of 5.3–47.3 mg/kg at 120 cm. The authors state that the deeper movement found in this study compared with many others is due to high loading rates of lead arsenate, coarse soil texture, low organic matter content and use of irrigation. The use of phosphate fertilizers significantly increases the amount of arsenic leached from soil contaminated with lead arsenate pesticide residues (Davenport & Peryea, 1991).

Masscheleyn et al. (1991a) found that at soil Eh levels of 200 and 500 mV arsenic solubility was low and the major part (65–98%) of the arsenic in solution was arsenate. Under moderately reduced soil conditions (at 0 and –100 mV) arsenic solubility was controlled by the dissolution of iron oxyhydroxides. Arsenic was co-precipitated as arsenate with iron oxyhydroxides and released on solubilization. On reduction to –200 mV the soluble arsenic content increased to 13 times what it was at 500 mV.

Richardson et al. (1978) monitored surface runoff of arsenic from a fine montmorillonitic clay after application of arsenic acid for desiccation of cotton (Gossypium hirsutum). They calculated that approximately 7% of the amount applied would be transported from the watershed by runoff and erosion, 38% in solution and 62% attached to sediment.

Tammes & de Lint (1969) calculated an average half-life of 6.5 ± 0.4 years for arsenic persistence on two Netherlands soils after application of arsenite.

4.2 Biotransformation

Most environmental transformations of arsenic appear to occur in the soil, in sediments, in plants and animals, and in zones of biological activity in the oceans. Biomethylation and bioreduction are probably the most important environmental transformations of the element, since they can produce organometallic species that are sufficiently stable to be mobile in air and water. However, the biomethylated forms of arsenic are subject to oxidation and bacterial demethylation back to inorganic forms (IPCS, 1981, section 4).

Three major modes of biotransformation of arsenic species have been found to occur in the environment: redox transformation between arsenite and arsenate, the reduction and methylation of arsenic, and the biosynthesis of organoarsenic compounds. There is biogeochemical cycling of compounds formed by these processes (Andreae, 1983).

Arsenic is released into the atmosphere primarily as As₂O₃ or, less frequently, in one of several volatile organic compounds, mainly arsines (US EPA, 1982). Trivalent arsenic and methyl arsines in the atmosphere undergo oxidation to the pentavalent state, and arsenic in the atmosphere is usually a mixture of the trivalent and pentavalent forms (Scudlark & Church, 1988). Photolysis is not considered an important breakdown process for arsenic compounds (Callahan et al., 1979).

Arsenic can undergo a complex series of transformations, including redox reactions, ligand exchange and biotransformation (Callahan et al., 1979; Welch et al., 1988). Factors affecting fate processes in water include the Eh, pH, metal sulfide and sulfide ion concentrations, iron concentrations, temperature, salinity, and distribution and composition of the biota (Callahan et al., 1979; Wakao et al., 1988).

4.2.1 Oxidation and reduction

Oscarson et al. (1980) observed oxidation of arsenite (10 mg/litre) to arsenate in sediments from lakes in Saskatchewan (Canada). The oxidation process was unaffected by flushing nitrogen or air through the system or by the addition of mercuric chloride. The authors therefore concluded that the oxidation was an abiotic process, with microorganisms playing a very minor role in the system. However, Scudlark & Johnson (1982) examined the oxidation of arsenite in seawater at low levels. They found that abiotic oxidation proceeded at a slow and constant rate with rapid oxidation occurring only in the presence of certain aquatic bacteria. The rate of abiotic oxidation, after spiking water with an initial arsenite concentration of 4 µg/litre (53 nmol/litre), was 0.2 µg/litre per day in distilled water and 0.3 µg/litre per day in artificial seawater. Baker et al. (1983a) found no methylated arsenic compounds in sterile lake sediments incubated in the presence of arsenate or arsenite.

Scudlark & Johnson (1982) studied the biological oxidation of arsenite in seawater in Narragansett bay (Rhode Island, USA). They found that oxidation was primarily due to microbial activity. Oxidation obeyed first-order kinetics with a rate constant of 0.06 h^–1 and half-lives ranging from 8.9 to 12.8 h for initial arsenite concentrations ranging from 7.5 µg/litre to 6.9 mg/litre (0.1–91.8 µmol/litre). Under aerobic conditions the mixed microbial cultures of lake sediments were able to reduce arsenate to arsenite and also to oxidize arsenite to arsenate. However, under anaerobic conditions only reduction was observed (Freeman et al., 1986).

In seawater containing free dissolved oxygen, arsenate is the thermodynamically stable form of the element. Arsenite is present in amounts exceeding those of arsenate only in reduced, oxygen-free porewaters of sediments and in anoxic basins such as the Baltic sea. However, significant amounts of arsenite (up to 10% of total arsenic) are found in the surface and deep waters of the oceans and, conversely, some arsenate is still present in anoxic water (Andreae, 1983). The presence of arsenite in seawater suggests that some reduction of arsenate occurs, and indeed Johnson (1972) demonstrated that bacterial arsenate reduction can take place under laboratory conditions. Matsuto et al. (1984) isolated a cyanobacterium (Phormidium sp.) from the coastal marine waters of Suruga bay (Japan) that was capable of readily reducing adsorbed arsenate to arsenite.

Freeman (1985) isolated an Anabaena oscillaroides–bacteria assemblage from the arsenic-rich Waikato river (New Zealand) capable of reducing arsenate to arsenite. In continuous culture the cyanophyte–bacteria assemblage could reduce arsenate to arsenite at a rate of 12 ng As/10⁶ cells per day. Wakao et al. (1988) detected microbial arsenite oxidation occurring in acid mine waters (pH 2.0–2.4) containing 2–13 mg As/litre. Ahmann et al. (1994) isolated a microorganism from arsenic-contaminated sediment in eastern Massachusetts (USA) which used the reduction of arsenate to arsenite to gain energy for growth. Similarly, Macy et al. (1996) found that an anaerobic bacterium Chrysiogenes arsenatis from gold-mine wastewater grew by reducing arsenate to arsenite using acetate as the electron donor and carbon source.

On the bassis of both aqueous and solid-phase observations, McGeehan (1996) found that arsenate was reduced to arsenite in flooded soil under batch conditions. Reduction of arsenate to arsenite has also been reported for both freshwater and marine macroalgae (Blasco, 1975; Johnson & Burke, 1978; Andreae & Klumpp, 1979; Wrench & Addison, 1981). Calculations based on the measured rates of reduction indicate that 15–20% of the total arsenic is reduced by phytoplankton during spring and autumn blooms on the continental shelf (Sanders & Windom, 1980).

4.2.2 Methylation

The biomethylation of arsenic was first recognized when arsines were produced from cultures of a fungus, Scopulariopsis brevicaulis (Challenger, 1945). Subsequently, the methylation of arsenic by methanogenic bacteria (McBride & Wolfe, 1971) and by reaction with methyl cobalamine (Schrauzer et al., 1972) or l-methionine-methyl-d3 (Cullen et al., 1977) has been demonstrated in laboratory work. Cox & Alexander (1973) showed that cultures of the fungus Candida humicola methylate arsenite, arsenate, methylarsonate and DMA to trimethylarsine. Further experiments have shown that growing cells of C. humicola can be induced to produce trimethylarsine from arsenate and DMA by preconditioning with DMA (Cullen et al., 1979b). Cullen et al. (1979a) incubated C. humicola in the presence of ⁷⁴As-arsenate, ¹⁴C-methylarsonate or ¹⁴C-DMA. They identified arsenite, methylarsonate, DMA and trimethylarsine oxide as intermediates in a biological synthesis of trimethylarsine. However, they tentatively conclude that methylarsonate does not occur as a free intermediate in the arsenate to trimethylarsine pathway.

McBride et al. (1978) reported that dimethylarsine was mainly produced by anaerobic organisms, whereas trimethylarsine resulted from aerobic methylation.

Methylated arsenic compounds were detected in aerobic sediments from various locations in Ontario (Canada) incubated with or without the addition of extraneous arsenic. Two pure bacterial cultures, Aeromonas sp. and Flavobacterium sp., isolated from lake water, were also found to methylate arsenic compounds in a synthetic medium (Wong et al., 1977).

Baker et al. (1983a) incubated lake sediment in the presence of arsenate or arsenite (7.5 mg As/litre). Methylation occurred over the pH range 3.5–7.5, with analysis revealing the presence of both methyl arsonic acid and dimethylarsinic acid. The amount of arsenic recovered in the methylated species ranged from 0 to 0.4% of the total inorganic arsenic added. Maeda et al. (1988) exposed the cyanobacterium Phormidium sp. (isolated from an arsenic-polluted environment) to arsenate (128 mg/kg) and found that 3.2% of the accumulated arsenic had been methylated.

Huysmans & Frankenberger (1991) isolated a Penicillium sp. from evaporation pond water capable of methylating and subsequently volatilizing organic arsenic. The conditions optimum for trimethylarsine production were a minimal medium containing 100 mg/litre methylarsonic acid, pH 5–6, a temperature of 20 °C and a phosphate concentration of 0.1–50 mmol/litre.

Reimer & Thompson (1988) found a strong positive correlation between the sum of the methylarsenic compounds and the total dissolved arsenic in marine interstitial waters influenced by mine tailings discharges indicating in situ microbial methylation. Laboratory studies have shown that microorganisms present in both natural marine sediments and sediments contaminated with mine tailings are capable of methylating arsenic under aerobic and anaerobic conditions (Reimer, 1989).

Biomethylation is primarily restricted to the high-salinity regions of estuaries with the presence of methylated arsenic at lower salinities predominantly as a result of the mixing of saline water (containing bioarsenicals) with river water (Howard & Apte, 1989).

Several authors have reported arsenic methylation in macroalgae, particularly in marine organisms (Edmonds & Francesconi, 1977; Andreae & Klumpp, 1979; Wrench & Addison, 1981; Maeda et al., 1987b; Cullen et al., 1994). In fact, most diatoms, dinoflagellates and macroalgae as well as freshwater higher plants, release protein-bound arsenic as a result of sequential methylation and adenosylation (Benson et al., 1988). Baker et al. (1983b) reported that freshwater green algae were capable of methylating sodium arsenite in lake water. Analysis revealed the presence of MMA, DMA and trimethylarsine oxide; however, volatile arsine and methylarsines were not detected. Similarly, Wrench & Addison (1981) identified MMA and DMA as polar arsenic metabolites synthesized by the marine phytoplankton Dunaliella tertiolecta. Maeda et al. (1987b) exposed five arsenic-resistant freshwater algae from an arsenic-polluted environment to arsenate. Small amounts of methylated arsenic compounds were detected and these were strongly bound with proteins or polysaccharides. Methylated arsenic compounds were found mainly in the lipid-soluble fractions and the major form was a dimethyl arsenic compound. No methylation occurred in algal cells (Chlorella vulgaris) exposed to arsenate under in vitro conditions; however, in vivo a small fraction of the arsenic accumulated was first transformed to methyl and dimethyl arsenic compounds during the early exponential phase and finally transformed to trimethylarsenic species (Maeda et al., 1992b). The marine algae Ecklonia radiata and Polyphysa peniculus methylated arsenate to produce a dimethylarsenic derivative. It was concluded that methionine or S-adenosylmethionine was the source of the methyl groups in this biological alkylation (Edmonds & Francesconi, 1988a; Cullen et al., 1994). S-adenosylmethionine is also likely to be the source of adenosyl and ribosyl groups in the arsenosugars.

The organic arsenical arsenobetaine was first identified in the late 1970s (Edmonds & Francesconi, 1981b) and has now been isolated in a variety of marine organisms (Edmonds & Francesconi, 1981b; Norin & Christakopoulos, 1982; Shiomi et al., 1984; Edmonds et al., 1992). Edmonds & Francesconi (1981a) identified arsenosugars isolated from brown kelp (Ecklonia radiata) as intermediates in the cycling of arsenic and stated that these compounds could be subsequently metabolized to arsenobetaine. Edmonds et al. (1982) have shown that the simpler arsenosugars in the brown alga are degraded under anaerobic conditions to dimethyloxarsylethanol. The transformation of dimethyloxarsylethanol to arsenobetaine would require both a reduction-methylation step and an oxidation step; these are probably bacterially mediated (Edmonds & Francesconi, 1987a, 1988b). Edmonds & Francesconi (1988b) concluded that arsenobetaine is probably formed by the conversion of arsenate to dimethyl(ribosyl)arsine oxides by algae, and that the microbially mediated transformation to arsenobetaine or its immediate precursors occurs in sediments. Phillips & Depledge (1985, 1986) proposed that phospholipids containing arsenoethanolamine or arsenocholine moieties may be formed as intermediates in the formation of arsenosugars and arsenobetaine. Edmonds et al. (1992) identified arsenocholine-containing lipids as natural products in the digestive gland of the rock lobster (Panulirus cygnus). Phillips & Depledge (1985) concluded that arsenic replaces nitrogen in phospholipid synthesis leading to a large number of arsenic-containing intermediates, which would be either water-soluble or lipid-soluble. Arsenic-containing compounds are catabolized as they pass through the food web, yielding arsenobetaine as a stable end-product.

Inorganic arsenic administered orally to brown trout (Salmo trutta) was detected in tissues as organoarsenical species, whereas arsenic administered by injection was taken up as inorganic arsenic and slowly converted to the organic form. It was concluded that biosynthesis of arsenic was occurring in the gastrointestinal tract (Penrose, 1975). Oladimeji et al. (1979) reported that arsenic given as an oral dose to rainbow trout (Oncorhynchus mykiss) was rapidly converted to organic forms. The ratio of total organic to inorganic increased with time in all tissues, with the organic arsenic fraction accounting for about 50% after 6 h and over 80% within 24 h. The major organic arsenical appeared to be an arsenobetaine-related compound. Similarly, Penrose et al. (1977) found that sea urchins (Strongylocentrotus droebachiensis) were also able to convert inorganic arsenic to an organic form, but to a more limited degree than trout. However, Wrench et al. (1981) concluded that organic arsenic synthesized in the brine shrimp (Artemia salina) is methylated by intestinal microflora and not by the filter feeder itself.

Maeda et al. (1990c) found that 85% of arsenic accumulated by the guppy (Poecilia sp.) was in the di- and tri-methylated forms. The percentage of organic species was much higher than that found in phytoplankton and zooplankton in the same model ecosystem. Similarly, Maeda et al. (1990a) found that biomethylation of arsenic increased successively with trophic level in another model ecosystem: goldfish (Carassius sp.) > zooplankton (Moina sp.) > alga (Chlorella sp.).

4.2.3 Degradation

4.2.3.1 Abiotic degradation

The rates of photochemical decomposition of arsenite, DMA, MMA and arsenobetaine have been studied in both distilled water and seawater. All species were found to degrade rapidly in aerated distilled water. In deaerated solutions the rate of oxidation of arsenite was almost two orders of magnitude slower. Half-lives for the degradation of DMA, MMA and arsenite were 9.2, 11.5 and 0.9 min respectively for aerated distilled water and 25, 19 and 8 min for deaerated distilled water. In seawater, the rates of photochemical decomposition were slower. For example, in seawater only 20% of DMA was converted to MMA after 300 min with no other products detected, whereas in distilled water DMA was completely degraded within 100 min (Brockbank et al., 1988). This study suggests that UV irradiation is of limited use for the pretreatment of saline samples to convert organoarsenic species to As(V) before analysis. The implications for photochemical decomposition of arsenic species in natural waters is not clear, because sunlight is deficient in the lower-wavelength bands generated by the mercury lamp used in this study. In addition, colloids and suspended particulates in the photic zone may play a significant role in arsenic decomposition in natural waters.

Von Endt et al. (1968) concluded that degradation of MSMA in soil was primarily due to soil microorganisms rather than abiotic factors. In 60-day tests in non-sterile soil 1.7–10% of the ¹⁴C-MSMA was degraded, whereas under steam-sterilized conditions only 0.7% was degraded.

4.2.3.2 Biodegradation

The predominant form of arsenic in water is usually arsenate (Callahan et al., 1979; Wakao et al., 1988), but aquatic microorganisms may reduce the arsenate to arsenite and a variety of methylated arsenicals.

Marine organisms tend to contain much higher levels of arsenic than terrestrial organisms; this is because of the high arsenate/phosphate ratio in oceans, which is a consequence of the very low phosphate concentration. Most of the arsenic accumulated in marine organisms is in a water-soluble form of arsenic, namely arsenobetaine. Hanaoka et al. (1987) incubated marine sediments in the presence of arsenobetaine and demonstrated microbial degradation, with arsenate, arsenite, MMA, DMA and arsenobetaine being identified. Further experiments revealed the formation of trimethylarsine oxide during aerobic incubation of bottom sediments with arsenobetaine as the carbon source (Kaise et al., 1987). Under aerobic conditions, arsenobetaine is converted to its metabolites to a much greater extent than other methylarsenicals. Under anaerobic conditions little or no degradation of arsenobetaine occurred, whereas trimethylarsine oxide and DMA were converted to less methylated compounds (Hanaoka et al., 1990). Degradation of arsenobetaine has also been demonstrated in the water column in the presence of suspended substances (Hanaoka et al., 1992).

Organoarsenical pesticides (e.g. MMA and DMA) applied to soil are metabolized by soil bacteria to alkylarsines, MMA, and arsenate (ATSDR, 1993). The half-time of DMA in soil is about 20 days (ATSDR, 1993).

Cheng & Focht (1979) added arsenate, arsenite, methylarsonate and DMA to three different soil types. Arsine was produced in all three soils from all substrates but methylarsine and dimethylarsine were only produced from methylarsonate and DMA respectively. Both Pseudomonas sp. and Alicaligenes sp. produced arsine as the sole product when incubated anaerobically in the presence of arsenate or arsenite. The authors concluded that reduction to arsine, not methylation to trimethylarsine, was the primary mechanism for gaseous loss of arsenicals from soil.

Degradation of MSMA by soil microorganisms was studied by Von Endt et al. (1968). In 60-day tests they found that 1.7–10% of the ¹⁴C-MSMA was degraded; four soil microorganisms isolated in pure cultures degraded 3–20% of ¹⁴C-MSMA to ¹⁴CO₂ when grown in liquid culture at 10 mg MSMA/litre. Woolson & Kearney (1973) showed that sodium DMA was degraded to arsenate in soil under aerobic conditions but not under anaerobic conditions. Degradation of MSMA has been shown to be associated with soil organic matter oxidation. In a loamy soil, degradation increased with increasing organic matter content (Dickens & Hiltbold, 1967). Akkari et al. (1986) studied the degradation of MSMA in soils at concentrations up to 5 mg As/kg. It was found that degradation followed first-order kinetics. The rate constant was temperature dependent only at soil water contents less than field capacity, and the temperature effect was less under flooded conditions. The differences in degradation rate under aerobic conditions and 20% water content were related to differences in the texture of the three soils. Half-lives for the clay and silty loam soils were 144 and 88 days respectively. Under anaerobic (flooded) soil conditions MSMA degradation occurs by reductive methylation to form arsenite and alkylarsine gases. The half-life values for the two soils indicate significantly faster degradation at 25 and 41 days respectively. The third soil, a sandy loam, produced the slowest degradation rate (t_½ = 178 days) probably because of its low organic matter content which may have supported fewer microorganisms.

The overall percentage of DMA (sodium salt) and MMA mineralized in a silty clay soil after 70 days ranged from 3% to 87% – values much higher than arsenic loss as volatile arsines (0.001–0.4%). Arsenate was the main metabolite from the degradation of both sodium DMA and MMA. The amount of sodium DMA mineralized was linearly related to the concentration of sodium DMA in the soil, indicating that the rate is first order. Mineralization of sodium DMA increased with increasing soil moisture and temperature. It was concluded that the loss of arsenic from some soils to the atmosphere may not be a major pathway and that inorganic arsenic may accumulate in soil from arsenical usage (Gao & Burau, 1997).

4.2.4 Bioaccumulation

Bioconcentration of arsenic under laboratory conditions occurs in aquatic organisms, primarily in algae and lower invertebrates. Bioconcentration factors (BCFs) measured in freshwater invertebrates for several arsenic compounds generally ranged up to 20; bioconcentration factors in fish were < 5; higher concentration factors have been observed in algae. Biomagnification in aquatic food chains does not appear to be significant (Callahan et al., 1979). Terrestrial plants may accumulate arsenic by root uptake from the soil or by adsorption of airborne arsenic deposited on the leaves, some species accumulating substantial levels.

4.2.4.1 Microorganisms

Maeda et al. (1987a) exposed cyanobacteria (Nostoc sp.) to arsenate concentrations of 1 and 10 mg As(V)/litre for 32 days with no effect on cell growth. Nostoc sp. accumulated 32 and 77 mg As/kg (dry cell) respectively at the two exposure concentrations.

Lindsay & Sanders (1990) report BCFs ranging from 1132 to 3688 for estuarine phytoplankton (Thalassiosira pseudomonas, Skeletonema costatum and Dunaliella tertiolecta) exposed to 25 µg As(V)/litre as arsenate for up to 48 h.

Phytoplankton take up arsenate readily and incorporate a small proportion into the cell. Most of the arsenate is reduced, methylated and released to the surrounding media. Phytoplankton batch cultures exposed to elevated levels of arsenate take up additional arsenic during the log phase of growth. Studies using ⁷⁴As indicate that the uptake rate varies from 0.15 ng As(V)/10⁶ cells per hour in unenriched cultures to 2.3 ng As(V)/10⁶ cells per hour in cultures containing 25 µg As(V)/litre. Cultured Skeletonema costatum increase their arsenic concentrations approximately 40% from 22 to 29 mg/kg (dry weight) in response to arsenate concentrations of 6–25 µg As(V)/litre (Sanders & Windom, 1980).

Phytoplankton readily incorporated dissolved arsenic, with average arsenic residues increasing from 5.7 to 17.7 mg/kg (dry weight) when cultured for 48–96 h at 25 µg As(V)/litre as arsenate (Sanders et al., 1989). Arsenate added to a freshwater model ecosystem was readily accumulated in plankton with arsenic residues of 37–47 mg/kg (dry weight) at 5 µg As(V)/litre and > 200 mg/kg at 50 µg As(V)/litre after 65-day exposures. Accumulation in other biota was much lower than for phytoplankton (Reuther, 1992).

Giddings & Eddlemon (1977) studied the uptake of radioactively labelled arsenic (added as sodium arsenate at 50 µg As(V)/litre) in model ecosystems (7 and 70 litres) for 5 weeks. Mean BCFs for algae ranged from 370 for sand microcosms to 4300 for lake mud microcosms. Algal arsenic concentrations were significantly greater in the 70-litre microcosms and in the sand microcosms than in the 7-litre and sediment microcosms.

Green algae (Chlorella vulgaris) exposed to arsenate concentrations of 7 to 9 mg As(V)/litre accumulated maximum residues of 3.75 g total As/kg (dry mass) within 10 days (Maeda et al., 1992c).

Maeda et al. (1985) found that arsenate uptake increased with an increase in the arsenic exposure concentration with C. vulgaris isolated from an arsenic-polluted environment. Maximum BCFs of 200–300 were observed during the log phase. At the highest exposure concentration (10 g As(V)/litre) algae were able to accumulate 50 g As/kg (dry weight). Approximately half of the arsenic taken up was estimated to be adherent to the extraneous coat of the cell with the remainder accumulated by the cell. Arsenate accumulation was affected by the growth phase; arsenic was most actively accumulated when the cell was exposed to arsenic during the early exponential phase (Maeda et al., 1992a).

Accumulation of arsenic (1 mg As(V)/litre as arsenate) by Dunaliella sp. was rapid, with equilibrium established within 8 h. Arsenic accumulation was studied at temperatures ranging from 10 °C to 33 °C, pH 4–10, light intensity from 0 to 10 000 lux and sodium chloride concentrations from 1 to 100 g/litre. Maximum arsenic residues under optimum conditions (22 °C; pH 8.2; 5000–10 000 lux and 20 g NaCl/litre) were 2000 mg As/kg. Increased phosphate significantly decreased the uptake of arsenic in the culture (Yamaoka et al., 1988). Yamaoka et al. (1992) found that D. salina accumulated more arsenic at nitrogen concentrations of 72 mg/litre than at 4.5 mg/litre.

4.2.4.2 Macroalgae

Fucus vesiculosus accumulated approximately 120 mg As/kg during an 85-day exposure to 7.5 µg As(V)/litre as arsenate. Filamentous algae and planaria accumulated less than 40 mg As/kg (dry weight), and cyanobacteria and various zooplankton accumulated less than 20 mg As/kg (Rosemarin et al., 1985).

Klumpp (1980) studied the effect of a variety of factors on the uptake of labelled arsenic by the seaweed Fucus spiralis. Neither pH (pH 7–9) nor salinity (9–36 g/litre) affected the uptake of arsenic; however, uptake at 30 °C was twice that at 16 °C. Arsenate uptake was reduced with increasing phosphate concentration (40–400 µmol/litre).

Lee et al. (1991) grew the aquatic plant Hydrilla verticillata in both mine-waste pool water and deionized distilled water contaminated with arsenate (0.4 and 0.8 mg As(V)/litre) for up to 16 days. Accumulation of arsenic reached steady state at 2–6 days in pool water at BCFs of 110–190. In deionized water maximum arsenic accumulation occurred after 8 days at a BCF of around 300. Phosphate (³ 12 mg/litre) inhibited the uptake of arsenic by H. verticillata.

4.2.4.3 Aquatic invertebrates

Sanders et al. (1989) studied the uptake of arsenic from water and from phytoplankton by the copepod Eurytemora affinis and the barnacle Balanus improvisus. In 24-h tests, E. affinis exhibited no uptake of dissolved arsenic; the arsenic content of copepods fed phytoplankton increased to 11.2 mg/kg (dry weight) compared with 8.9 mg/kg in controls. In 22-day tests, B. improvisus exposed to dissolved arsenate (55 µg As(V)/litre) in water did not accumulate arsenic, with levels remaining around 0.88 mg/kg; however, levels in shell material increased from 0.3 mg/kg to 2 mg/kg. Barnacles fed arsenic-contaminated phytoplankton (~18 mg/kg) exhibited an increase in total arsenic concentrations from 0.3 mg/kg to 1.7 mg/kg. In further experiments with oysters (Crassostrea virginica) no accumulation of arsenic from water was observed in 28-day tests, but tissue concentrations increased significantly from 5.3 mg/kg to 8.2 mg/kg in oysters fed arsenic-contaminated phytoplankton. Zaroogian & Hoffman (1982) reported maximum total arsenic residues in soft tissues of oysters (Crassostrea virginica) of 12.6, 12.7 and 14.1 mg/kg (dry weight) at arsenite exposure concentrations of 1.2 (control), 3 and 5 mg As(III)/litre during 112-day exposures. Generally, arsenic body burdens increased with increases in phytoplankton concentration and it appears that food contributes more to arsenic uptake than do seawater arsenic concentrations. No relationship between arsenic uptake and seawater arsenic concentrations was found.

Ünlü & Fowler (1979) exposed mussels (Mytilus galloprovincialis) to arsenate (⁷⁴As) concentrations ranging from 20 to 100 µg As(V)/litre at 12 °C and 21 °C. Mean concentration factors after 20 days were low, at respectively 8.8 and 12.1 for the two temperatures; however, mussels did accumulate significantly more arsenic at 21 °C than at 12 °C. Arsenic uptake was inversely related to salinity over the range 31–19 g/litre. Arsenic loss was essentially biphasic, with biological half-times of approximately 3 and 32 days for the fast and slow compartments respectively. The active secretion of arsenic in the byssal threads contributed to the total elimination of the element from the mussels. Similarly, Ünlü (1979) found a biphasal loss of arsenic from crabs (Carcinus maenas) during a 43-day depuration period. The elimination of ⁷⁴As by the crabs after ingestion of arsenic-contaminated mussels was dependent on the chemical form of the arsenic. After ingestion of mussel containing mostly lipid- and water-soluble arsenic species (undetermined), biological half-times were 3.4 and 19.6 days for the first and second phase of loss. After ingestion of mussel containing mostly arsenite and residual arsenic, half-times were 1.6 and 9.3 days respectively.

Naqvi et al. (1990) exposed red crayfish (Procambarus clarkii) to MSMA at concentrations of 0.5, 5 and 50 mg As/litre. Uptake of arsenic was dose-dependent but not time-dependent. Maximum whole-body residues were 1.36, 4.29 and 9 mg As/kg respectively for each of the exposure concentrations during the 8-week uptake period.

Gibbs et al. (1983) reported equilibrium BCFs, based on ⁷⁴As, for the cirratulid polychaete Tharyx marioni ranging from 4.5 at an exposure concentration of 10 mg As(V)/litre (as arsenate) to 111.6 at 0.01 mg/litre after 7 days. A lower BCF of only 15.9 at 0.01 mg/litre was reported for the polychaete Caulleriella caputesocis.

Shrimps exposed to water concentrations ranging from 0.1 to 1.5 mg As(V)/litre (as arsenate) or food (Chlorella sp.) containing, 1940 mg total As/kg contained arsenic residues ranging from 18.9 to 31.8 mg/kg (dry weight) (Maeda et al., 1992c).

Fowler & Ünlü (1978) reported BCFs of less than 10 for shrimps exposed to arsenate (⁷⁴As) concentrations of 20–100 µg As(V)/litre for 14 days. Arsenic loss was biphasic with half-lives of 3 and 26 days for the fast and slow compartments respectively. Moults shed during loss contained 2–5% of the shrimp’s ⁷⁴As body burden.

Lindsay & Sanders (1990) found no bioaccumulation of arsenate directly from the water (25 µg As(V)/litre) or from food for the grass shrimp (Palaemonetes pugio). Brine shimp (Artemia sp.) grown in elevated arsenic concentrations exhibited small, but significant, increases in arsenic content from an average of 16.8 mg/kg (dry weight) in controls to 17.8 mg/kg at 25 µg As(V)/litre; no accumulation was observed when brine shrimps were fed arsenic-contaminated food.

4.2.4.4 Fish

Barrows et al. (1980) exposed bluegill sunfish (Lepomis macrochirus) to 130 µg As(III)/litre of As₂O₃ for 28 days. The maximum BCF was found to be 4, with a half-life in tissues of 1 day. Nichols et al. (1984) found no accumulation of arsenic in a 6-month study on coho salmon (Oncorhynchus kisutch) exposed to As₂O₃ concentrations of < 300 µg As(III)/litre. Whole-body residues were below 0.4 mg As/kg (wet weight) and were not dose dependent.

Sorensen (1976) found that green sunfish (Lepomis cyanellus) exposed to higher arsenic concentrations of 100, 500 and 1000 mg As(V)/litre (as arsenate) accumulated whole-body arsenic concentrations of 33.4, 541.2 and 581.6 mg/kg (BCFs ranging from 0.3 to 1.1). Green sunfish exposed to 60 mg As(V)/litre for 6 days accumulated mean arsenic residues of 158.7, 47.7, 18.9 and 14.2 mg/kg in the gallbladder (plus bile), liver, spleen and kidney respectively (BCFs ranging from 0.2 to 2.6) (Sorensen et al., 1979).

Cockell & Hilton (1988) fed rainbow trout (O. mykiss) on diets containing As₂O₃ (180–1477 mg As/kg diet), disodium arsenate heptahydrate (DSA) (137–1053 mg As/kg diet), DMA (163–1497 mg As/kg diet) or arsanilic acid (193–1503 mg As/kg) for 8 weeks. For each of the arsenicals investigated, carcass arsenic concentration showed a dose–response relationship to dietary arsenic concentration and exposure rate. At lower levels of exposure (137 mg As/kg diet), dietary DSA yielded the highest mean carcass arsenic concentrations (6.9 mg As/kg), but at higher levels, dietary As₂O₃ (1477 mg As/kg diet) yielded the highest mean residues (21.6 mg As/kg). Inorganic arsenicals were accumulated from the diet to a greater degree than the organic forms. In a 16-week study, dietary DSA (8–174 mg As/kg diet) accumulated in the carcass (0.25–5.7 mg As/kg), liver (0.7–34.4 mg As/kg) and kidney (1.1–31.9 mg As/kg) in a dose-related manner (Cockell et al., 1991).

Oral administration of sodium arsenate to estuary catfish (Cnidoglanis macrocephalus) and school whiting (Sillago bassensis) resulted in an accumulation of trimethylarsine oxide in their tissues (Edmonds & Francesconi, 1987b). Yelloweye mullet (Aldrichetta forsteri) fed the organic arsenicals 2-dimethylarsinylethanol, 2-dimethylarsinylacetic acid or 2-dimethyllarsinothioylethanol showed no arsenic accumulation in their tissues; fish fed arsenate-contaminated food showed a small but significant increase in arsenic concentration (muscle tissue = 1 mg As/kg wet weight). However, administering arsenobetaine or arsenocholine in the diet led to muscle concentrations of around 24 mg As/kg (wet weight) (Francesconi et al., 1989).

Oladimeji et al. (1984) fed rainbow trout (O. mykiss) on a diet containing 10, 20 or 30 mg As(III)/kg (as sodium arsenite) (equivalent to 0.2, 0.4 and 0.6 mg/kg fish wet weight per day) for up to 8 weeks. Arsenic accumulation was dose related, with residues ranging from 1.28 to 1.52 mg/kg (dry weight) for muscle, 1.55 to 5.21 mg/kg for liver, 0.84 to 1.88 mg/kg for gills and 1.21 to 1.98 mg/kg for skin tissue.

4.2.4.5 Terrestrial plants

Arsenic species can enter into edible tissues of food crops through absorption (i.e. not just surface contamination) (Woolson, 1973; Helgesen & Larsen, 1998). Helgesen & Larsen (1998) demonstrated that bioavailability of arsenic pentoxide to carrots in soil from a wood preservative treatment plant (soil was contaminated with CCA) was 0.47 ± 0.06% of total soil arsenic burden. This study showed that arsenite, arsenate, MMA and DMA were present in carrot tissue, where only arsenite and arsenate were present in soil. In soils dosed with arsenate (0–500 µg/g) at the concentrations which inhibited growth of vegetable crops (green bean, lima bean, spinach, cabbage, tomato and radish), high levels of accumulation when found in the edible parts of radish (76 µg/g) spinach (10 µg/g) and green bean (4.2 µg/g). Arsenic accumulation in Lima bean, cabbage and tomato ranged from 0.7–1.5 µg/g. The studies of Woolson (1973) and Helgesen & Larsen (1998) highlight the potential of movement of arsenic species from soil into agronomic crops.

Uptake of arsenate (10 mg/litre [133 µmol/litre]) by moss (Hylocomium splendens) from nutrient solution displayed saturation kinetics at pH 5 that could be described in terms of Michaelis–Menten parameters with a mean K_m value of 31.4 mg/litre (418 µmol/litre). Phosphate was a competitive inhibitor of arsenate uptake with an inhibition constant (K_m phosphate) of 82 µmol/litre (Wells & Richardson, 1985).

Asher & Reay (1979) studied the uptake of arsenate (1 mg/litre [15 µmol/litre]) from nutrient solution by barley (Hordeum vulgare) seedlings. They found that uptake consisted of a rapid initial phase followed by a less rapid ‘steady-state’ phase, both of which were strongly inhibited by phosphate and positively correlated with temperature.

The marsh plant species Spartina alterniflora was grown in sediment treated with ~50 µg As(V)/litre (as arsenate) and accumulated significantly elevated total concentrations of arsenic after 9 days; new and old leaf blades contained mean arsenic concentrations of 6.3 and 5 mg/kg (dry weight) respectively, relative to 1 and 0.4 mg/kg in control plants (Sanders & Osman, 1985).

Meharg & Macnair (1991b) found that non-tolerant genotypes of Holcus lanatus accumulated significantly more arsenate than tolerant plants during a 6-h period of growth in 3.75 mg As(V)/litre (0.05 mol/m³) arsenate. They found that tolerant plants transported a much greater proportion of arsenic to their shoots than non-tolerant plants. Phosphate (0.05 or 0.5 mol/m³) decreased arsenate uptake in both tolerant and non-tolerant genotypes. Arsenate tolerance involves reduced accumulation of arsenate through suppression of the high-affinity phosphate–arsenate uptake system (Meharg et al., 1994).

Anastasia & Kender (1973) grew lowbush blueberry (Vaccinium angustifolium) plants in greenhouse soil at As₂O₃ concentrations ranging from 7.7 (controls) to 84.5 mg As(III)/kg for 17 weeks. Arsenic was accumulated in a dose-dependent manner with arsenic residues of 0.78–15 mg/kg for leaves, 0.27–13.3 mg/kg for stems and 2.4–164.2 mg/kg for roots.

Otte et al. (1990) grew Urtica dioica and Phragmites australis in soil containing up to 30 mg As/kg added as lead arsenate or sodium DMA. Concentrations of arsenic in shoots and roots of P. australis increased significantly only at the highest arsenic concentration in soil with mean values of up to 1 mg/kg (dry weight) in shoots and 44.3 mg/kg in roots, whereas the arsenic content of U. dioica increased by a factor of 4 at 5 mg As/kg with plants accumulating mean arsenic concentrations of up to 150 mg/kg in roots at the highest exposure.

Onken & Hossner (1995) grew rice (Oryza sativa) in two soil types treated with up to 45 mg As(III) or As(V)/kg (as arsenite or arsenate) for 60 days. The arsenic concentration of rice plants correlated with the mean soil solution arsenate concentration in the clay soil and to the mean soil solution arsenite for the silt loam. The rate of arsenic uptake by plants increased as the rate of plant growth increased.

4.2.4.6 Terrestrial invertebrates

Meharg et al. (1998) exposed earthworms (Lumbricus terrestris) to arsenate (40 mg/kg dry weight) for 23 days. There was a steady-state increase in residues for depurated and undepurated worms and by 12 days earthworm residues were equivalent to those of the soil. Arsenic residues were accumulated to three times soil levels by the end of the 23-day exposure in depurated worms; however, undepurated worms did not appear to bioconcentrate arsenic beyond the level of the surrounding soil.

4.2.4.7 Birds

Daghir & Hariri (1977) administered arsanilic acid (used as a feed medication for poultry) to White Leghorn laying hens at 50 and 100 mg/kg for 15 weeks. Maximum concentrations in eggs were reached after 4–5 weeks at 0.13 and 0.24 mg As/kg (dry weight) for the two dose levels respectively. Residual arsenic was negligible 2 weeks after the withdrawal of the drug from the feed.

Proudfoot et al. (1991) found a higher concentration of arsenic in liver and muscle of broilers that were fed arsanilic acid (99 mg/kg diet) compared with controls. Mean arsenic residues of up to 1.5 mg/kg and 0.4 mg/kg were measured for the two tissues respectively. Broilers fed a diet containing 100 or 500 mg/kg arsanilic acid accumulated up to 2.3 and 8 mg As/kg in liver tissue at the two exposure concentrations respectively. Lower levels were accumulated in muscle tissue, with arsenic concentrations of up to 0.15 and 0.67 mg/kg (VanderKop & MacNeil, 1989).

Holcman & Stibilj (1997) fed Rhode Island Red hens on diets containing 7.5, 15 or 30 mg As(III)/kg (as As₂O₃) for 19 days. Eggs were collected on days 8–19 day of the experiment, and arsenic residues were consistent throughout this period. Mean concentrations were respectively 0.2, 0.42 and 0.96 mg As/kg (dry weight) in egg yolk and in 0.06, 0.14 and 0.3 mg As/kg egg white for the three exposure concentrations.

Hoffman et al. (1992) fed mallard on a diet containing 200 mg As(V)/kg (as sodium arsenate) for 4 weeks. Arsenic accumulated in the liver at a concentration of 2.3 mg As/kg (wet weight). Birds maintained on a restricted protein and exposed to the same arsenic-contaminated diet accumulated 5.1 mg As/kg. Stanley et al. (1994) maintained mallards on diets containing 25, 100 or 400 mg As(V)/kg (as sodium arsenate) for 16–18 weeks. Arsenic was accumulated in a dose dependent manner; mean concentrations in adult livers were 0.49–6.6 mg As/kg (dry weight), in duckling livers from 0.65–33 mg/kg and in whole eggs from 0.46–3.6 mg/kg.

5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

5.1 Environmental levels

Arsenic is a natural component of the earth’s crust, and found in all environmental media. Concentrations in air in remote locations range from < 1 to 3 ng/m³, but concentrations in cities may range up to 100 ng/m³. Concentrations in water are usually < 10 µg/litre, although higher concentrations can occur near natural mineral deposits or anthropogenic sources. Natural levels in soils usually range from 1 to 40 mg/kg, but pesticide application or waste disposal can produce much higher values.

5.1.1 Air

Levels of arsenic in ambient air are summarized in Table 3. Examples are given of mean total arsenic concentrations in remote and rural areas ranging from 0.02 to 4 ng/m³. Levels of arsenic in outdoor air near to urban and industrial sources are summarized in Table 4. Examples are given of mean total arsenic concentrations in urban areas ranging from 3 to 200 ng/m³; much higher concentrations (> 1000 ng/m³) have been measured in the vicinity of industrial sources. Arsenic in ambient air is usually a mixture of arsenite and arsenate, with organic species being of negligible importance except in areas of substantial methylated arsenic pesticide application or biotic activity. Schroeder et al. (1987) reviewed worldwide arsenic concentrations associated with particulate matter. They identified arsenic levels ranging from 0.007 to 1.9 ng/m³ for remote areas, 1 to 28 ng/m³ for rural areas and 2 to 2320 ng/m³ in urban areas. The highest arsenic levels detected in the atmosphere were near non-ferrous-metal smelters.

Typical background levels for arsenic are now 0.2–1.5 ng/m³ for rural areas, 0.5–3 ng/m³ for urban areas and < 50 ng/m³ for industrial sites (DG Environment, 2000).

Table 3. Concentrations of As in ambient air^a

^a As = inorganic As; As = organic As; NS = not stated

^b Mean and ranges of total As unless stated otherwise

Table 4. Concentrations of As in outdoor air near urban and industrial sources

^a Mean and ranges of total As unless stated otherwise

NS = not stated

5.1.2 Precipitation

Arsenic has been detected in rainwater at mean concentrations of 0.2–0.5 µg/litre (Welch et al., 1988). Peirson et al. (1974) report mean arsenic concentrations in rainfall ranging from < 6 µg/litre for a rural site to 45 µg/litre at a North Sea gas platform. Arsenic concentrations in precipitation at the mid-Atlantic coast of the USA ranged from < 0.005 to 1.1 µg/litre with an average of 0.1 µg/litre (Scudlark & Church, 1988). Andreae (1980) collected rainwater samples from non-urban sites in California (USA) and state parks in Hawaii and found mean arsenic concentrations ranging from 0.013 to 0.032 µg/litre. Samples from a rural site in Washington state (USA) contained a mean concentration of 1.1 µg As/litre; the author states that the site is 154 km north of a large copper smelter. Vermette et al. (1995) monitored arsenic levels in wet deposition at three sites (Colorado, Illinois and Tennessee, USA) and found mean concentrations ranging from 0.09 to 0.16 µg/litre. Reimann et al. (1997) monitored rainwater samples during the summer of, 1994 in eight Arctic catchments. Median arsenic concentrations (0.45 µm) ranged from 0.07 µg/litre at the most remote site to 12.3 µg/litre near a smelter.

Barbaris & Betterton (1996) analysed fresh snowpack samples from high-elevation forests of north-central Arizona (USA) during late winter and early spring 1992–1994. Arsenic concentrations ranged from 0.02 to 0.4 µg/litre with a mean value of 0.14 µg/litre.

5.1.3 Surface water

Levels of arsenic in seawater are summarized in Table 5. Concentrations of arsenic in open ocean seawater are typically 1–2 µg/litre. The dissolved forms of arsenic in seawater include arsenate, arsenite, MMA and DMA, with adsorption on to particulate matter being the physical process most likely to limit dissolved arsenic concentrations (Maher & Butler, 1988). Levels of arsenic in estuarine water are summarized in Table 6. Tremblay & Gobeil (1990) noted that arsenic concentrations increased with increasing salinity (0–31 g/litre) from 0.5 to 1.4 µg/litre (6.6 to 18.9 nmol/litre) in the St Lawrence estuary (Canada) and from 0.1 to 1.4 µg/litre (1.1 to 18.7 nmol/litre) in its tributary Saguenay fjord. Penrose et al. (1975) monitored seawater in Moreton’s Harbour, Newfoundland near a long-term stibnite mine. Total inorganic arsenic concentrations were 5.3 µg/litre near the mine but declined to normal levels (1–2 µg/litre) within 200 m.

Howard et al. (1988) report that concentrations of methylated arsenic increased with salinity in the Tamar estuary (United Kingdom). Concentrations of monomethylarsenic ranged from 0.02 to 0.46 µg As/litre and dimethylarsenic from 0.02 to 1.27 µg As/litre; these two methylated forms of arsenic were typically 4% and 10% of the total soluble arsenic levels respectively.

Levels of arsenic in surface freshwaters are summarized in Table 7. Surveys of arsenic concentrations in rivers and lakes indicate that most values are below 10 µg/litre, although individual samples may range up to 1 mg/litre (Page, 1981; Smith et al., 1987; Welch et al., 1988). Mean total arsenic concentrations of 2000 µg/litre have been recorded near a pesticide plant, with MMA being the predominant arsenic species (Faust et al., 1983; 1987a). Crearley (1973) measured arsenic in two lakes near a manufacturing plant which had been producing arsenic-based cotton desiccants/defoliants for 30 years. Mean arsenic concentrations of 7900 and 3200 µg/litre were found. During the dry season total dissolved arsenic concentrations (< 0.45 µm) of up to 250 µg/litre were recorded near industrial discharges to the Xiangjiang river (China); however, maximum levels during the rainy season were generally less than 30 µg/litre (Chunguo & Zihui, 1988).

High levels of arsenic have been recorded in thermal waters. Tanaka (1990) found a mean concentration of 570 µg/litre in geothermal waters throughout Japan, with a maximum level of 25.7 mg/litre.

Table 5. Background concentrations of As in seawater

^a Mean and ranges of total As unless stated otherwise

NS = not stated

Table 6. Concentrations of As in estuarine waters

^a Mean and ranges of total As unless stated otherwise

NS = not stated

Table 7. Concentrations of As in surface freshwaters

^a Mean and ranges of total As unless stated otherwise

NS, not stated

5.1.4 Groundwater

Levels of arsenic in groundwater are summarized in Table 8. Arsenic levels in groundwater average about 1–2 µg/litre, except in areas with volcanic rock and sulfide mineral deposits where arsenic levels can range up to 3400 µg/litre (Page, 1981; Welch et al., 1988; Robertson, 1989). In some mining areas arsenic concentrations of up to 48 mg/litre have been reported (Welch et al., 1988). Korte & Fernando (1991) reported that arsenic levels in arsenic-contaminated water supply wells in southern Iowa and western Missouri (USA) ranged from 34 to 490 µg/litre. The authors state that the arsenic appears to be of natural origin. Similarly, Matisoff et al. (1982) found no evidence for an anthropogenic source contributing to elevated groundwater levels of arsenic (< 1 to 100 µg/litre) in north-eastern Ohio (USA). Arsenic levels in groundwater were found to exceed 10 µg/litre in 5.6–9.5% of samples collected in Germany during the period 1992–1994 (Umweltbundesamt, 1997). Varsanyi (1989) found arsenic concentrations in deep groundwater in Hungary to range from 1 to 174 µg/litre with an average value of 68 µg/litre. High arsenic levels originating from arsenic-rich bedrock were found in drilled wells in south-west Finland, with concentrations ranging from 17 to 980 µg/litre (Kurttio et al., 1998). Del Razo et al. (1990) monitored groundwater in the Lagunera region of northern Mexico. Total arsenic concentrations ranged from 8 to 624 µg/litre with over 50% of samples > 50 µg/litre. The predominant arsenic species in 93% of samples was arsenate, although in 36% of samples 20–50% arsenite was found. Chen et al. (1994) report that arsenic levels in the groundwater of south-west Taiwan contained mean dissolved arsenic levels of 671 µg/litre. Arsenic levels in the well-waters of Hsinchu (Taiwan) were less than 0.7 µg/litre.

Table 8. Concentrations of As in groundwater

^a Mean and ranges of total As unless stated otherwise

NS = not stated

Arsenic contamination of groundwater from arsenic-rich sediment has been reported in both India and Bangladesh. Chatterjee et al. (1995) analysed groundwater from six districts of West Bengal (India). Mean total arsenic levels ranged from 193 to 737 µg/litre with a maximum value of 3700 µg/litre. Mean arsenite levels in the groundwater were around 50% of the total arsenic. Mandal et al. (1996) reported that 44% of groundwater samples collected in West Bengal (India) up to January 1996 contained total arsenic levels > 50 µg/litre. Dhar et al. (1997) found that 38% of groundwater samples collected from 27 districts of Bangladesh contained total arsenic levels > 50 µg/litre.

During 1990 and 1991 Chatterjee et al. (1993) sampled groundwater in the vicinity of a chemical plant in Calcutta, India, which had produced the insecticide Paris green (acetocopper arsenite) for 20 years. Groundwater contained total arsenic levels ranging from < 0.05 to 58 mg/litre; the highest total arsenic level included 75% arsenite.

5.1.5 Sediment

Arsenic concentrations in sediments are summarized in Table 9. Sediments in aquatic systems often have higher arsenic concentrations than those of the water (Welch et al., 1988). Most sediment arsenic concentrations reported for rivers, lakes and streams in the USA range from 0.1 to 4000 mg/kg, with higher levels occurring in areas of contamination (Welch et al., 1988). Arsenic concentrations of < 10 000 mg/kg (dry weight) were found in surface sediments near a copper smelter (Crecelius et al., 1975). Sediment arsenic concentrations of < 3500 mg/kg were reported for lakes in the Northwest Territory (Canada) which had received past inputs from gold-mining activity (Wagemann et al., 1978). Mean total arsenic concentrations of 500 mg/kg (dry weight) were measured in sediment near a pesticide plant and at a lake 14–17 km downstream mean concentrations of almost 3000 mg/kg had accumulated (Faust et al., 1987a). Arsenate was the predominant arsenic species, with inorganic arsenic amounting to 70–90% of the total arsenic measured (Faust et al., 1983). Bright et al. (1996) found total arsenic concentrations ranging from 1043 to 3090 mg/kg in the top 10 cm of sediment from subarctic lakes contaminated by gold-mining activity. Total dissolved arsenic levels in porewater ranged from 800 to 5170 µg/litre (0.7% organic arsenic). Ebdon et al. (1987) reported that methylated arsenic species represented 1–4% of the total arsenic in sediment porewater from the Tamar estuary, south-west England (United Kingdom). Similar findings were reported by de Bettencourt (1988) for the Tagus Estuary (Portugal).

Table 9. Concentrations of As in sediment

^a Mean and ranges of total As unless stated otherwise

NS = not stated

Chunguo & Zihui (1988) studied arsenic accumulation in sediment of the Xiangjiang river (China), which receives inputs from a variety of industrial plants. Total arsenic concentrations upstream of industrial inputs were 13.2 mg/kg during the rainy season and 81.4 mg/kg during the dry season. Near to industrial discharges maximum total arsenic concentrations exceeded 1000 mg/kg during the dry season (approximately 70% as iron or aluminium arsenate) but rarely reached 100 mg/kg during the rainy season.

Farmer & Lovell (1986) monitored arsenic concentrations in surface sediments of Loch Lomond (Scotland, UK); no recent significant sources of environmental arsenic contamination were identified. They found natural enrichment of sediment to levels of up to 675 mg As/kg compared with typical background concentrations of 15–50 mg/kg.

5.1.6 Sewage sludge

Zhu & Tabatabai (1995) monitored total arsenic levels in sewage sludges from waste treatment plants in Iowa (USA). Concentrations ranged from 2.4 to 39.6 mg/kg with a mean of 9.8 mg/kg.

5.1.7 Soil

Levels of arsenic in soil are summarized in Table 10. Arsenic is found in the earth’s crust at an average level of 2 mg/kg. Most natural soils contain low levels of arsenic, but industrial wastes and pesticide applications may increase concentrations. Background concentrations in soil range from 1 to 40 mg/kg, with a mean value of 5 mg/kg (Bowen, 1979; Beyer & Cromartie, 1987). Naturally elevated levels of arsenic in soils may be associated with geological substrata such as sulfide ores. Anthropogenically contaminated soils can have concentrations of arsenic up to several percent (NAS, 1977; Porter & Peterson, 1977). Arsenic concentrations of up to 27 000 mg/kg were reported in soils contaminated with mine or smelter wastes (US EPA, 1982). Chatterjee & Mukherjee (1999) reported arsenic levels of 20 100–35 500 mg/kg in soil around the effluent dumping point of an arsenical pesticide manufacturing plant. Peat may contain considerable quantities of arsenic. Minkkinen & Yliruokanen (1978) found maximum arsenic concentrations in various Finnish peat bogs of between 16 and 340 mg/kg dry peat. However, Shotyk (1996) analysed peat cores from the Jura mountains (Switzerland) and found mean total arsenic concentrations of 3.6 mg/kg at a depth of < 30 cm and 0.16 mg/kg at between 69 and 84 cm. Higher levels of arsenic were found in the mineral sediments underlying the peat bogs, with mean concentrations of 6.4 mg/kg at 170 cm and 15.9 mg/kg at 650 cm. Soil on agricultural land treated with arsenical pesticides may retain substantial amounts of arsenic. Mean total arsenic concentrations of 50–60 mg/kg have been recorded for agricultural soils treated with arsenical pesticides (Takamatsu et al., 1982; Sanok et al., 1995). Walsh & Keeney (1975) reported that arsenic-treated soils contained up to 550 mg As/kg. Stilwell & Gorny (1997) found mean arsenic concentrations ranging from 9 to 139 mg/kg (dry weight) in soil (upper 5 cm) below decking treated with copper chrome arsenate (CCA).

Table 10. Concentrations of As in soil

^a Mean and ranges of total As unless stated otherwise

NS = not stated; TR = trace

Uptake and effects of arsenic on organisms are related to bioavailable arsenic rather than total arsenic. Xu & Thornton (1985) measured mean total arsenic levels of 300 mg/kg in garden soils (south-west England, UK) at sites of past mining activity; however, water-soluble and acid-fluoride extractable arsenic represented < 1% and < 2% of total arsenic respectively. Kavanagh et al. (1997) report total arsenic concentrations ranging from 174 to 477 mg/kg for agricultural soil and from 1200 to 22 290 mg/kg for mine waste in the Tamar valley, south-west England. The proportion of water-extractable arsenic in agricultural topsoils ranged from 0.05 to 0.3% and in mine wastes from 0.02 to 1.2%. Similarly, McLaren et al. (1998) found total arsenic concentrations of 37–3540 mg/kg (dry weight) in surface soil (0–10 cm) contaminated by cattle dip (sodium arsenite) compared with water-extractable arsenic concentrations in the same samples ranging from 0.2 to 22.4 mg/kg . The highest total arsenic level recorded was 14 800 mg/kg (water-extractable arsenic = 1.2 mg/kg) at a depth of 40–45 cm. Ng et al. (1998b) measured total arsenic concentrations of 32–1597 mg/kg in soil which had been contaminated 30 years previously with arsenical pesticides. Chemical speciation showed that arsenite ranged from 0.32–56% of total arsenic. In a rat model, the absolute bioavailability of these contaminated soils relative to arsenite and arsenate ranged from 1.02 to 9.87% and 0.26 to 2.98% respectively.

Doyle & Otte (1997) found that the presence of vegetation and burrowing organisms significantly increased the concentration and accumulation of arsenic in salt-marsh soils.

Chutke et al. (1995) analysed dust samples collected in Nagpur city (India) during 1992. Mean arsenic levels for residential/commercial areas, industrial areas and highways were respectively 10.2, 18 and 17 mg/kg (dry weight). Stone & Marsalek (1996) collected samples of sediment from road surfaces in an urban area of Ontario (Canada) during 1991 and found total arsenic concentrations ranging from 1 to 33 mg/kg with a mean of 3.4 mg/kg.

5.1.8 Biota

Background arsenic concentrations in living organisms are usually less than 1 mg/kg (fresh weight) in freshwater and terrestrial biota. The levels are higher in biota collected from mine waste sites, arsenic-treated areas, near smelters and mining areas, near areas with geothermal activity and near manufacturing sites of arsenical defoliants and pesticides (Eisler, 1988). Marine organisms, however, can normally contain arsenic residues ranging from 1–2 mg/kg to more than 100 mg/kg (Lunde, 1977; Maher & Butler, 1988; Phillips, 1990). Neff (1997) reviewed levels of total arsenic in marine organisms and calculated geometric means ranging from < 1 mg/kg for marine mammals to 50 mg/kg for snails. An overall geometric mean for a wide variety of marine biota was calculated to be 11 mg/kg (dry weight). There is a substantial number of publications on the levels of arsenic in biota, and the following examples have been chosen to provide an overview.

5.1.8.1 Freshwater

Freshwater plants in uncontaminated environments tend to contain arsenic concentrations < 10 mg/kg (Reay, 1972; Outridge & Noller, 1991). Reay (1972) reported a considerable accumulation of arsenic in freshwater plants in the Waikato river (New Zealand). The elevated arsenic concentrations in the water (30–70 µg/litre) arising from geothermal activity gave rise to concentrations of < 971 mg As/kg in aquatic plants. Arsenic concentrations of < 1200 mg/kg (dry weight) were reported by Mudroch & Capobianco (1979) for aquatic macrophytes growing in an area of the Lake Ontario drainage basin (Canada) contaminated with mine effluent. Wagemann et al. (1978) reported arsenic concentrations ranging from 150 to 3700 mg/kg for macrophytes in lakes (Northwest Territory, Canada) which had received past inputs from gold-mining activity. During 1983, Tanner & Clayton (1990) analysed macrophytes from Lake Rotoroa (New Zealand), which had been sprayed with sodium arsenite herbicide in 1959. Arsenic concentrations ranged from 540 to 780 mg/kg (dry weight) in surficial sediments and from 193 to 1200 mg/kg in macrophytes. Similar levels of arsenic accumulation to that seen in aquatic plants has been observed for zooplankton (700–2400 mg/kg) in lakes receiving mine drainage water (Wagemann et al., 1978).

Freshwater bivalves have been used to measure arsenic in several biomonitoring programmes. Leland & Scudder (1990) monitored freshwater bivalves (Corbicula fluminea) in the San Joaquin valley (California, USA), an area influenced by high levels of elements in irrigation wastewater. Mean concentrations of arsenic in bivalves ranged from 5.3 to 13.9 mg/kg (dry weight). A highly significant relationship was observed between arsenic residues and the HNO₃-extractable arsenic : iron ratio of suspended matter. Similarly, Johns & Luoma (1990) found mean arsenic levels ranging from 5.4 to 11.5 mg/kg (dry weight) for the same species for the Sacramento/San Joaquin river delta (California, USA). Arsenic levels in mussels from the St Lawrence river (Canada) ranged from 2.8 to 8.6 mg/kg (dry weight) (Metcalfe-Smith, 1994).

Freshwater fish have not been shown to accumulate arsenic to the same degree as lower aquatic organisms. Arsenic residues in freshwater fish have been monitored in the USA over a period of approximately 10 years. The geometric means (mg/kg wet weight), with the range in parentheses, of total arsenic concentrations were 0.27 (0.05–2.92) during 1976–1977, 0.14 (0.05–1.69) during 1980? 1981 and 0.14 (0.27–1.5) during 1984 (May & McKinney, 1981; Lowe et al., 1985; Schmitt & Brumbaugh, 1990). Mean total arsenic residues in freshwater fish near a copper smelter (Sweden) ranged from 0.05 to 0.24 mg/kg (wet weight) compared with 0.06 to 0.09 mg/kg for a control lake (Norin et al., 1985). Takatsu & Uchiumi (1998) analysed fish from a naturally acidified volcanic lake (Lake Usoriko, Japan) with low phosphate levels (< 0.02 mg/litre). Mean arsenic levels were 0.28 and 0.27 mg/kg (wet weight) for gills and bone respectively and 6.1 mg/kg for eye tissue. Arsenic residues have also been measured in fish from the San Joaquin valley area of California (USA) exposed to agricultural subsurface drainage water. Mean arsenic concentrations ranged from 0.18 to 0.44 mg/kg (dry weight) (maximum value 0.97 mg/kg) for bluegill sunfish (Lepomis macrochirus) and from 0.23 to 0.39 mg/kg (maximum value 1.5 mg/kg) for common carp (Cyprinus carpio) (Saiki & May, 1988). Mean arsenic concentrations for striped bass (Morone saxatilis) from the same area ranged from 0.23 to 0.65 mg/kg (dry weight) compared with mean values of 1.23–1.44 mg/kg for bass from San Francisco bay (California, USA) (Saiki & Palawski, 1990).

Clark et al. (1998) found a mean arsenic concentration of 6.87 mg/kg (wet weight) in tadpoles of the cricket frog (Acris crepitans) collected in 1994 downstream from Finfeather Lake (Texas, USA); the lake was contaminated during 53 years (1940–1993) of industrial production of arsenic-based cotton desiccants/defoliants.

5.1.8.2 Marine

Marine biota tend to accumulate much higher levels of arsenic than freshwater species (see section 4.2.3.2). Very little information is available on arsenic levels in natural phytoplankton populations. Benson & Summons (1981) reported 9 mg/kg total arsenic in a mixed marine phytoplankton population near Cape Ferguson (Queensland, Australia). Sanders (1979a) found that mean total arsenic concentrations in marine macroalgae ranged from 1.4 mg/kg (Rhodophyceae) to 10.3 mg/kg (Phaeophyceae). The absolute concentration of inorganic arsenic was not significantly different between groups, suggesting that the variation is due to metabolic differences between algal classes rather than to differences in the environmental concentration of arsenic. Mean total arsenic concentrations in macroalgae collected in the South Atlantic ranged from 5.3 to 70.2 mg/kg, with inorganic arsenic residues ranging from 0.2 to 2.0 mg/kg (Muse et al., 1989). Lai et al. (1998) report seasonal changes in arsenic speciation in the brown alga Fucus gardneri in Vancouver (Canada). During the summer algae contain 9 mg As/kg with most (79–98%) being extractable, whereas during the winter months residues range from 16 to 22 mg/kg with extraction efficiencies of 5.8–49%. Klumpp & Peterson (1979) found mean arsenic concentrations ranging from 83.7 to 141.4 mg/kg (dry weight) (maximum 189.3 mg/kg) for macroalgae in Restronguet creek, south-west England (UK) (an estuary influenced by past mining activity). Penrose et al. (1975) monitored marine biota near the site of a disused stibnite mine (pre–1916). Mean arsenic residues were 17.2 mg/kg for macroalgae and 3.8–11.5 mg/kg for invertebrates near the mine site compared with 9.8–12.1 mg/kg for macroalgae and 1.6–4.0 mg/kg for invertebrates at a control site.

Stronkhorst (1992) reported mean arsenic concentrations in mussels of 1 mg/kg (wet weight) for two Dutch estuaries. Similar levels (mean values ranging from 1.1 to 2.7 mg/kg) were reported in clams and oysters collected from U.S. coastal waters in use for shellfish production during 1985 and 1986 (Capar & Yess, 1996). Shellfish from the Arabian Gulf contained mean arsenic concentrations ranging from 3 to 15.8 mg/kg (wet weight) (Attar et al., 1992; Madany et al., 1996). Molluscs sampled in Restronguet creek contained arsenic concentrations ranging from 35 to 64 mg/kg (dry weight) (Klumpp & Peterson, 1979). Benson & Summons (1981) found that arsenic was accumulated to substantial levels in the kidney of molluscs from the Great Barrier Reef (Australia) with residues ranging from 481 to 1025 mg/kg (dry weight).

Langston (1980) found that the highest arsenic concentrations in estuarine benthic organisms (< 190 mg/kg [dry weight]) were found at sites where high arsenic : iron ratios exist in the sediment. Concentrations of arsenic in estuarine organisms correlated more significantly with the arsenic : iron ratio in sediments than arsenic levels alone.

Arsenic residues in marine fish appear to show substantial variation. Hellou et al. (1992) found mean arsenic residues of 3.2 mg/kg (dry weight) (1.6–4.2 mg/kg) in Atlantic tuna (Thunnus thynnus). Engman & Jorhem (1998) reported arsenic residues in marine fish muscle ranging from 0.59 to 17 mg/kg (fresh weight) with a mean value of 4.5 mg/kg. The mean value was 60 times greater than that found for freshwater fish in the same study. Several studies of marine fish from the Arabian Gulf have shown that in general mean arsenic concentrations range from < 1 to < 10 mg/kg in muscle (Tariq et al., 1991; Attar et al., 1992; Madany et al., 1996). However, higher concentrations have been reported, for example, Attar et al. (1992) found mean muscle concentrations of up to 32.3 mg As/kg (wet weight) in black-banded bream (Acanthopagrus bifasciatus). Bohn (1975) reported mean arsenic concentrations for marine fish from West Greenland ranging from 21.9 to 240 mg/kg (dry weight).

Maher (1983, 1988) analysed a variety of marine biota and found mean total arsenic concentrations (dry weight) ranging from 2.7 mg/kg (fish muscle) to 114 mg/kg (macroalgae: Cystophora moniliformis). Mean inorganic arsenic concentrations were low (0.02–3.6 mg/kg) in all marine organisms with organic arsenic representing 70–98% of the total arsenic. Arsenobetaine was the most abundant arsenic species found in marine invertebrates and fish muscle tissue (Edmonds & Francesconi, 1981c; Shiomi et al., 1984; Maher, 1985b; Matsuto et al., 1986).

Mean arsenic concentrations in the liver and muscle tissue of marine mammals were found to be generally less than 1 mg/kg (Julshamn et al., 1987; Muir et al., 1988; Skaare et al., 1990; Miles et al., 1992; Varanasi et al., 1994).

5.1.8.3 Terrestrial

The arsenic content of plants grown on soils that had never been treated with arsenic-containing pesticides varied from 0.02 to about 5 mg/kg (dry weight). Plants grown on arsenic-contaminated soils may, however, contain considerably higher levels, especially in the roots. Plants growing on arsenical mine wastes (south-west England, UK) contained mean arsenic levels ranging from 350 to 2040 mg/kg (dry weight); a maximum concentration of 6640 mg/kg was reported for Jasione montana (Porter & Peterson, 1975). Benson et al. (1981) reported mean arsenic concentrations of 1480 and 1070 mg/kg (dry weight) for the grasses Agrostis stolonifera and A. tenuis growing on arsenical mine waste. De Koe (1994) found arsenic concentrations of up to 1800 and 1900 mg/kg in senescent shoots and roots respectively of grass species growing on gold-mine spoils (north-east Portugal). Jonnalagadda & Nenzou (1997) report arsenic concentrations in couch grass (Cynodon dactylon) growing on or near gold/arsenic mine dumps (Zimbabwe) ranging from 200 to 1660 mg/kg (dry weight) in stems and from 1020 to 10 880 mg/kg in roots. Mean concentrations of arsenic in the leaves of plants growing near a copper mine (northern Peru) ranged from 111 to 1651 mg/kg (dry weight) (Bech et al., 1997). Temple et al. (1977) found mean arsenic levels of 5.8 mg/kg in grass samples and 7.4 mg/kg in tree and shrub foliage from within 700 m of a secondary lead smelter; samples collected at a control site contained < 1 mg/kg.

Grass growing on plots which had been previously treated (7–11 years before) with lead arsenate contained mean arsenic residues of 1.5 mg/kg, compared with 0.9 mg/kg in grass from untreated sites. After a further 2 years mean arsenic concentrations were 0.88 and 0.56 mg/kg for treated and untreated sites respectively (Chisholm & MacPhee, 1972). Merry et al. (1986) reports that pasture plants growing at sites formerly used as orchards (soil concentration 80 mg As/kg) contained less than 2.5 mg As/kg (dry weight).

Biomonitoring studies at six background sites in Norway found mean arsenic concentrations in moss (Hylocomium splendens) ranging from 0.1 to 2.2 mg/kg (Berg et al., 1995a); an overall mean of 0.36 (< 0.03–3.2) mg/kg was reported by Berg et al. (1995b). Similarly, Glooschenko & Arafat (1988) sampled sphagnum moss (Sphagnum fuscum) throughout northern Canada. A mean background concentration of 0.66 mg As/kg (dry weight) was found, with elevated levels (> 3 mg/kg; maximum 31 mg/kg) in the vicinity of mining and smelting areas. Lichen biomonitoring of arsenic in a geothermal area of central Italy revealed a mean concentration of 1.19 mg/kg (0.19–3.55 mg/kg) (dry weight) (Loppi & Bargagli, 1996).

Monitoring of conifer needles has been carried out at sites remote from pollution sources, with mean arsenic concentrations ranging from 5 to 58 µg/kg for Norway spruce (Picea abies) and from 2 to 8 µg/kg for balsam fir (Abies balsamea) (Lin et al., 1995; Wyttenbach et al., 1997). However, much higher concentrations (0.46–3.1 mg/kg) have been reported for leaves from loblolly pine trees (Pinus taeda) growing on land affected by coal-pile leachate (Carlson & Carlson, 1994). Dmuchowski & Bytnerowicz (1995) monitored Scots pine (Pinus sylvestris) needles at three sites in Poland during 1983–1985. Mean arsenic concentrations were 0.54 mg/kg (dry weight) in a primeval forest (eastern Poland), 0.88 mg/kg near the city of Warsaw and 1.5 mg/kg at a polluted site in Silesia. Mankovska (1986) analysed pine needles (Pinus silvestris) from the vicinity of a smelter and found arsenic concentrations ranging from ~15 to ~22 mg/kg within 1000 m of the smelter (soil concentrations ranged from 30 to > 120 mg/kg).

Byrne & Tusek-Znidaric (1983) found arsenic concentrations ranging from 34 to 182 mg/kg (dry weight) in caps and stalks of the common mushroom (Laccaria amethystina) from rural sites in Slovenia; soil arsenic concentrations ranged from 3.2 to 27 mg/kg.

Beyer & Cromartie (1987) analysed earthworms from a diverse variety of sites in Maryland, Pennsylvania and Virginia (USA). Arsenic concentrations ranged from trace levels to 0.8 mg/kg (dry weight) at uncontaminated sites, mining sites and industrial sites. However, a single earthworm sample at a mining site contained 10 mg/kg (soil concentration 20 mg/kg) although all other samples from mining sites contained only trace amounts of arsenic. Total arsenic concentrations ranging from 3.2 to 17.9 mg/kg (dry weight) were found in earthworms sampled from six sites in Austria. There was no correlation between the total arsenic concentrations in the earthworms and the soil. The major arsenic compounds detected in the earthworms were arsenous acid and arsenic acid; arsenobetaine, dimethylarsinic acid and two dimethylarsinoylribosides were also detected (Geiszinger et al., 1998).

Arsenic residues in birds tend to be low (< 1 mg/kg) with little accumulation even at sites with higher environmental concentrations (Martin & Nickerson, 1973; Blus et al., 1977; White et al., 1980; Ohlendorf et al., 1991; Pain et al., 1992; Vermeer & Thompson, 1992; Custer & Hohman, 1994; Guitart et al., 1994; Hothem & Welsh, 1994). Of 18 osprey (Pandion haliaetus) livers analysed by Wiemayer et al. (1980), 14 contained less than 1.5 mg/kg (wet weight); arsenic concentrations in the other four birds ranged from 2 to 16.7 mg/kg. The bird with the highest concentration was in a weak condition with very low fat reserves. Erry et al. (1999) analysed tissue samples from raptors in south-west England, an area with naturally and anthropogenically (through mining) elevated arsenic levels and compared the results with birds from another geographical area. Mean arsenic residues of 0.278, 0.346 and 0.187 mg/kg (dry weight) in the kidney, liver and muscle of kestrels (Falcio tinnunculus) were approximately three times higher in south-west England than in south-west Scotland. However, in another two raptors (sparrowhawk Accipiter nisus and barn owl Tyto alba) arsenic levels were not elevated in south-west England. The authors suggested that the difference could be attributed to differences in both diet and arsenic metabolism. Vermeer & Thompson (1992) analysed livers from birds collected in the vicinity of a copper mine; mean arsenic concentrations ranged from 0.08 to 3.23 mg/kg (wet weight). Goede (1985) found mean arsenic concentrations ranging from 0.5 to 3.2 mg/kg in the feather shafts of wading birds (Waddenzee, Netherlands); liver concentrations ranged from 4 to 14 mg As/kg (dry weight).

Elfving et al. (1979) analysed small mammals (voles and mice) from apple orchards which had received lead and calcium arsenate applications for many years. Arsenic concentrations in the soil ranged from 31 to 94 mg/kg (dry weight) and in the small mammals from 0.05 to 0.96 mg/kg (whole-body). Arsenic concentrations at a control site were 2.4 mg/kg in soil and < 0.03 mg/kg in small mammals. Ismael & Roberts (1992) monitored arsenic residues in vegetation and small mammals near an arsenic refinery. Mean arsenic levels in vegetation were 0.2 and 37.3 mg/kg for a control site and 250 m from the arsenic refinery respectively. Mean whole-body arsenic residues in four species of small mammal ranged from 0.4 to 3.2 mg/kg (fresh weight) at the control site and from 0.4 to 2.4 mg/kg near to the refinery. Significantly higher levels were found at the control site for three of the four species; the common shrew Sorex araneus, a carnivorous species, accumulated the highest levels of arsenic at both sites.

Arsenic was not detected (detection limit 5 µg/kg) in kidney tissue of mink (Mustela vison) collected in Georgia, North Carolina and South Carolina (USA) (Osowski et al., 1995). Langlois & Langis (1995) did not detect arsenic in muscle tissue (detection limit 50 µg/kg) of hares or martens in northern Quebec (Canada). Norstrom et al. (1986) found a mean arsenic concentration of 0.07 mg/kg (dry weight) in livers of polar bears (Ursus maritimus) in the Canadian Arctic. Norheim et al. (1992) reported mean arsenic concentrations of 0.06 and 0.04 mg/kg (wet weight) in the livers of adult and juvenile polar bears respectively at Svalbard (Norway).

5.2 General population exposure

Arsenic is widely distributed and human exposure is inevitable. Exposure of the general population to the various species of arsenic (inorganic and organic) will vary according to local geochemistry and the level of anthropogenic activity and can occur through the intranasal, oral and dermal routes.

5.2.1 Air

Arsenic in ambient air is associated with particulate matter and is predominantly a mixture of arsenite and arsenate. Organic species are of negligible significance except in areas where there has been substantial use of methylated arsenic pesticides or in areas with high biotic activity (ATSDR, 1993). As discussed in section 5.1.1 (see Tables 3 and 4), arsenic concentrations associated with particulate matter vary world wide as follows: 0.007–1.9 ng/m³ in remote areas; 1–28 ng/m³ in rural locations, and 2–2320 ng/m³ in urban environments (Schroeder et al., 1987). The highest concentrations are found near non-ferrous-metal smelters.

5.2.2 Food and beverages

Arsenic has been found in all foodstuffs analysed. Although most monitoring data is given as the concentration of total arsenic, arsenic in foods is a mixture of inorganic species and organoarsenicals including arsenobetaine. The actual total arsenic concentrations in foodstuffs from various countries will vary widely depending on the food type, growing conditions (type of soil, water, geochemical activity, use of arsenical pesticides) and processing techniques.

From monitoring studies in the USA (Gunderson, 1995, Yost et al., 1998; US NRC, 1999), in the United Kingdom (UK MAFF, 1997), Canada (Dabeka et al., 1993) and Australia (ANZFA, 1994), by far the highest concentrations of total arsenic is found in seafood. Meats and cereals have higher concentrations than vegetables, fruit and dairy products. On the basis of limited data, it has been estimated that the percentage of inorganic arsenic is about 75% in meats, 65% in poultry, 75% in dairy products, and 65% in cereals (US EPA, 1988; Yost et al., 1998). Tao & Bolger (1998) estimated an inorganic arsenic intake for US men and women aged 60–65 years of 13 and 10 µg respectively. Other age groups had lower estimated daily intakes of inorganic arsenic, varying from 1.3 µg for infants to 9.9 µg for men aged 25–30 years. Additional samples, and a wider range of foodstuffs, need to be analysed in various countries before a definite conclusion can be reached on the normal range of inorganic arsenic in foods. In fruits, and vegetables and seafood the organic species predominate, with inorganic arsenic contributing 10%, 5% and 0–10% respectively. On the basis of these preliminary data it has been estimated that approximately 25% of the daily intake of dietary arsenic is inorganic (US EPA, 1988, Yost et al., 1998). A report from the Netherlands (Vaessen & van Ooik, 1989) estimated that inorganic arsenic in seafood was 0.1 to 41% of the total. Edmonds & Francesconi (1993) reviewed all data on inorganic arsenic in seafoods (excluding algae) available at that time and concluded that inorganic arsenic represented less than 1% of total arsenic at low arsenic burdens and fell to about 0.5% of total arsenic at concentrations of about 20 mg/kg. Mohri et al. (1990) estimated that the customary Japanese diet contained 5.7% inorganic arsenic with an intake ranging from 27 to 376 µg total arsenic/day.

Concentrations of arsenic in various food groups found in Canada are given in Table 11. Analysis of various beverages from Denmark found 3–11 µg/litre (Pedersen et al., 1994). Very few data were found on the concentration of arsenic in human breast milk. One study of 10 lactating women by Concha et al. (1998c) found a range of 0.83–7.6 µg/kg fresh weight (median 2.3 µg/kg) in breast milk from women consuming > 200 µg arsenic per day from drinking-water. Thus breast-feeding provided 1–2 µg As/day, compared to 100–200 µg As/day from formula mixed with the arsenic-rich water.

Table 11. Total As concentrations in various food groups from Canada^a

Examples of mean total daily intakes of arsenic from food and beverages in different countries are given in Table 12. The variation in dietary intake of total arsenic in adults reflects in large part the variability in the consumption patterns of arsenic-rich food groups (fish/shellfish and meats) confirming the need to consider such regional variations in arsenic intake when assessing human health effects for arsenic.

Table 12. Estimated average dietary intake of As in various countries

The risk of arsenic exposure to populations living in or near arsenic-contaminated environments (i.e. mine-tailing sites, CCA and arsenical pesticide contamination soils), must be considered. In particular, contamination of home-grown vegetables and reared livestock, or wild collected foods must be considered. Helgesen & Larsen (1998) demonstrated that 0.47–0.6% of total soil arsenic (from a CCA plant) was bioavailable to carrot. Woolson (1973) dosed soils with 0–500 mg/kg arsenate and showed that at the arsenate dose that limited growth by 50%, arsenic in edible parts was up to 87 mg/kg (see section 4.2.4.5).

5.2.3 Drinking-water

Concentrations of arsenic in fresh surface water and groundwater, potential sources of drinking-water, are given in sections 5.1.3 and 5.1.4. Arsenate is the predominant species, but some groundwaters have been found to contain a high proportion of arsenite (section 5.1.4). Concentrations of methylated species in natural waters are usually less than 0.3 µg/litre (ATSDR, 1993). Unless stated otherwise in this section, monitoring data for drinking-water is reported as total arsenic.

A summary of the monitoring of drinking-water carried out in the USA by the US EPA during 1976–1993 has been published by Borum & Abernathy (1994). Concentrations of arsenic were in the range of < 2.5–28 µg/litre for surface waters and < 5–48 µg/litre for groundwater sources. Detection limits of 2 or 5 µg/litre preclude more accurate estimates of the lower limit of these ranges. On the basis of these data, it was estimated that approximately 2% of the population of the USA is exposed to > 10 µg/litre arsenic in drinking-water. Additional data sources (US EPA, 1993) provide support for this estimate, and have identified areas with higher concentrations of arsenic in drinking-water. In 1978 arsenic was detected in 67% of the 3834 drinking-water samples analysed (detection limit 0.1 µg/litre) with a mean concentration of 2.4 µg/litre (Borum & Abernathy, 1994). In areas with elevated geological concentrations of arsenic (e.g. California and Nevada) mean arsenic concentrations up to 80 µg/litre have been reported, with maximum reported levels of > 1400 µg/litre.

A limited summary of the monitoring data collected in 1985–1988 for total arsenic in drinking-water in six Canadian provinces has been published (NHW/DOE, 1993). Of the 717 samples of surface water, 3.6% were > 5 µg/litre and 5% of the 600 groundwater samples contained arsenic at concentrations > 5 µg/litre.

Although arsenic levels in natural waters are usually low (a few µg/litre), there are several areas in the world where humans consume drinking-water containing > 100 µg As/litre resulting from natural geochemical activity. In the West Bengal region of India it was estimated that over 1 000 000 people consume drinking-water containing > 50 µg/litre (up to 3.7 mg/litre) arising from normal geochemical processes (Das et al., 1995; Chowdhury et al., 1997). In the areas of West Bengal and Bangladesh, 38% of groundwaters sampled in 27 districts were > 50 µg/litre (Dhar et al., 1997). Natural geochemistry resulted in the pre–1970 exposure of about 100 000 people in the south-western coastal region of Taiwan to variable but high (10–1800 µg/litre, mean 500 µg/litre) concentrations of arsenic in drinking-water (Guo et al., 1994). A similar problem was reported in Chile where 100 000 people consumed drinking-water containing 800 µg As/litre between 1959 and 1970, when the concentration was lowered to about 50 µg/litre(Borgono et al., 1977). About 200 000 people in north central Mexico were reported to be exposed to >50 µg/litre arsenic in drinking-water (410 µg/litre in at least one village) (Cebrian et al., 1983).

In major Australian drinking-water systems levels of arsenic range up to 15 µg/litre, but typical concentrations are usually < 5 µg/litre (NHMRC, 1996).

5.2.4 Soil

Although ingestion of arsenic in soil and dust may not be a significant source of arsenic intake in adults, it may be significant for children, particularly in locations near industrial and hazardous waste sites. As described in section 5.1.7 (Table 10), background concentrations of total arsenic in soil are 1–40 mg/kg dry weight with a mean of 5 mg/kg (Beyers & Cromartie, 1987). The comparative bioavailability of arsenic in soil from a CCA-contaminated site and soil contaminated by arsenic solutions used in cattle tick control were reported by Ng & Moore (1996). In a rat model, soil from the cattle dip site had a bioavailability of 8.1 ± 4%; 14.4 ± 7% and 60 ± 3.4% when compared with orally administered sodium arsenite, calcium arsenite and sodium arsenate respectively. For CCA-contaminated soil the corresponding comparative bioavailabilities were 13.0 ± 4.5%; 32.2 ± 11.2% and 38.0 ± 13.2%. Also using a rat model, Ng et al. (1998b) have reported the absolute bioavailability of arsenic in soils containing 32–1597 µg As/kg (0.32–56% arsenite) from a combination of arsenical pesticides and natural geological formations in a residential area. The absolute bioavailability ranged from 1.02 to 9.87% relative to arsenite and from 0.26 to 2.98% relative to arsenate.

Freeman et al. (1993) determined both the absolute and comparative bioavailability of arsenic in soil from a smelter site using male rabbits and monkeys. When compared to the intravenous administration of sodium arsenate, the absolute bioavailability was reported as 25.9% in rabbits and 24.2% in monkeys. When compared to an oral dose (gavage) of sodium arsenate, the comparative bioavailabilities were 67.8% in rabbits and 43.6% in monkeys, in general agreement to findings of Ng et al. (1998b) in rats. Such data on availability of arsenic in soil needs to be considered in assessing human uptake of arsenic from soil (for more details on bioavailability see Table 15).

5.2.5 Miscellaneous exposures

Smokers are exposed to arsenic by the inhalation of mainstream cigarette smoke. It has been estimated that someone in the USA smoking 40 cigarettes per day would inhale about 10 µg of arsenic (ATSDR, 1993).

Proprietary herbal asthma medicines have been shown to contain up to 107 mg/g of inorganic arsenic (Chan, 1994).

5.3 Occupational exposures

There is the potential for significant occupational exposure to arsenic in several industries, in particular non-ferrous smelting, electronics, wood preservation, wood joinery shops, arsenic production, glass manufacturing, and the production and application of arsenical pesticides. Exposure is primarily through inhalation of arsenic-containing particulates, but ingestion and dermal exposure may be significant in particular situations. (e.g. preparation of CCA-treated timber). It is extremely rare for workers to be exposed to arsenic alone: the exposure is usually to arsenic in combination with other elements. Data on typical exposure levels of arsenic in the workplace are difficult to obtain and may vary considerably between different locations of the same industry because of the level of occupational hygiene in place and the chemical properties of the materials processed. Also, they are often out of date with regard to the current level of industrial hygiene. Currently, countries which have occupational regulations for arsenic have set the limit for inorganic arsenic between 0.01 and 0.1 mg/m³ (ILO, 1991; DFG, 1999; Ministerie van Soziale Zaken en Werkgelegenheid, 2000; OSHA, 2000). The following examples are given to illustrate levels found in specific industries in various locations worldwide and provide some information on present and past exposures of workers to arsenic. They should not be considered as representative of all similar industrial sites.

Some workplace exposures dating from before 1980 are summarized in IPCS (1981). For example, in a Swedish copper smelter during the mid-1950s levels of arsenic ranged between 0.06 and 2 mg/m³, but at the same facility in the 1970s levels of arsenic between 0.002 and 0.23 mg/m³ in the air breathed by the workers were reported. Several other studies described in IPCS (1981) reported levels of arsenic in non-ferrous metal production to be between 0.001 and 0.3 mg/m³ depending on the job location and the level of ventilation.

Welch et al. (1982), on the basis of industrial hygiene measurements made from 1943 to 1965, estimated average arsenic concentrations of workers in various departments of a copper smelter in the USA who were employed before 1956. Very high exposures (> 5 mg/m³) were estimated in the following departments: arsenic roaster (20 mg/m³), electrostatic precipitator (13 mg/m³), arsenic refinery (7.5 mg/m³), and main flue (7 mg/m³). High exposures (0.5–4.99 mg/m³) were estimated in these departments: masons’ shop (3 mg/m³), ore roaster (1 mg/m³), materials crushing (1 mg/m³) and reverberatory furnaces (0.6 mg/m³). Medium (0.1–0.49 mg/m³) or low exposures (< 0.1 mg/m³ ) were estimated in the 10 other departments of the smelter in which arsenic measurements were carried out. In another copper smelter in the USA, Pinto et al. (1976) reported an overall mean arsenic concentration of 0.05 mg/m³ (range 0.003–0.3 mg/m³) on the basis of data from 24 workers wearing personal air samplers on 5 consecutive days. For 1973, a more detailed exposure estimation within the 32 departments of the smelter was made on the basis of the individual 24-h urinary excretion of arsenic in 1000 workers (Pinto et al., 1978). The highest average urinary arsenic excretions (µg As/litre of urine) were calculated for the following departments: electrostatic precipitator (526 mg As/litre), arsenic plant (516 µg As/litre), roaster (414 µg As/litre) and boiler room (409 µg As/litre). Eleven other departments had urinary excretion levels between 289 and 201 µg As/litre of urine, and the remaining 17 areas had levels between 180 and 58 µg As/litre, the lowest level calculated for the refined casting department. On the basis of a study by Enterline & Marsh (1982) airborne arsenic levels (as µg/m³) will be about one-third of the urinary excretion concentrations (as µg As/litre of urine). These authors also summarized the airborne arsenic levels in the smelter between 1938 and 1957. Levels varied by department, but were all high. For example, between 1947 and 1953, in a total of 25 samples from the arsenic plant they found airborne arsenic concentrations ranging from 0.8 to 41.4 mg/m³.

Vahter et al. (1986) reported airborne arsenic levels (8 h TWA) of 1–194 µg/m³ in a copper smelter. Daily urinary excretion of total arsenic metabolites ranged from 16 to 328 µg As/g creatinine. Correlation between urinary excretion of arsenic species and 8-h TWAs of arsenic between 0.8–45 µg/m³ in 24 workers in a copper smelter and an As₂O₃ refinery were reported by Hakala & Pyy (1995). The best correlation was obtained between urinary excretion of the sum of arsenite and arsenate species in urine samples taken 8 h after exposure. An exposure to an 8-h TWA of 10 µg/m³ was calculated to lead to an inorganic arsenic concentration of 5 µg/litre in urine. Jakubowski et al. (1998) reported levels of arsenic in a copper smelter between 1 and 746 µg/m³ in the worker’s breathing zone (8-hTWA) resulting in daily urinary excretion of 2–850 µg As/g creatinine. On the basis of results from this study and three others, the authors calculated that daily exposure to arsenic concentrations of 10 or 50 µg/m³ corresponded to concentrations of total urinary metabolites of 30 µg/litre and 70 µg/litre (specific gravity 1.024) respectively. This compares to urinary excretions (total As metabolites) of 5–30 µg As/litre in people not excessively exposed via the workplace or from the consumption of seafood (Foa et al., 1984; Vahter et al., 1986). Simonato et al. (1994) reported urinary excretion of 183–205 µg As/g creatinine of arsenic metabolites in a cohort of gold-miners and refinery workers. Using airborne arsenic data for 1952–1991, Ferreccio et al. (1996) categorized workers’ exposure to arsenic in various units of a copper mine and smelter complex (in µg As/m³⁾ as follows: workshop and administration, 9.8; administrative area, 1.6; mine, 2.3; oxide production, 3.1; sulfur plant, 8.4; smelter, 201.7. In comparison, Offergelt et al. (1992) reported levels of arsenic (TWA) between 6 and 502 µg/m³ in a sulfuric acid plant. As part of an epidemiological investigation on lung cancer mortality of workers in non-ferrous mines, Liu & Chen (1996) measured airborne arsenic levels in 1978, 1981 and 1988. In chronological order, concentrations of arsenic reported (in mg/m³) were 0.23 (range 0.004–0.577 in 6 samples); 0.06 (range 0.003–0.166 in 14 samples) and 0.32 (range 0.028–1.442 in 8 samples).

Workers in certain glass-manufacturing industries may be exposed to airborne arsenic through the use of As₂O₃ (IARC, 1993). Workers in the heavy crystal industry in Germany were found to have urinary arsenic concentrations ranging from 3 to 114 µg/g creatinine, with 36% of the cases in 1976 and 18% of cases in 1981 being above the upper normal limit of 25 µg As/g creatinine (Schaller et al., 1982). A study in the USA of 35 crystal glassworkers within the mix-and-melt and batch-house areas indicated the potential for arsenic exposure (Chrostek et al., 1980). Personal air monitoring of 8 workers found airborne arsenic concentrations of 2–11 mg/m³. The mean urinary arsenic excretion in 18 workers involved in weighing and mixing chemicals in a specialist glass-manufacturing facility was 79.4 µg/g creatinine compared to 4.4 µg/g creatinine in controls (Farmer & Johnson, 1990). In a Belgian glass factory, Roels et al. (1982) measured urinary excretion of arsenic in 10 workers ranging between 10 and 941 µg/g creatinine compared to a range of 7.6 to 59 µg/litre in control workers. The authors concluded that the high urinary arsenic concentrations in the workers were more related to oral intake due to poor hygienic practices than to pulmonary uptake.

Airborne arsenic levels in a wood joinery shop handling treated wood were reported to be 0.043–0.36 mg/m³ (IPCS, 1981). In a more recent study of joinery shops (Nygren et al., 1992), airborne arsenic concentrations between 0.54 and 3.1 µg/m³ were reported. In two workshops machining wood impregnated with CCA, levels of arsenic in personal air samples were reported to be 30–67 µg/m³ in plant A (8 workers) and 10–62 µg/m³ in plant B (8 workers) (Subra et al., 1999).

Workers in coal-powered power plants may also be exposed to arsenic found in the coal, or more likely that found in the fly ash during cleaning. Yager et al. (1997) reported arsenic concentrations (8-h TWA) between 0.17 and 375.2 µg/m³ (mean 48.3) in the breathing zone of maintenance workers in a coal-fired power plant in Slovakia. The urinary excretion of total urinary arsenic metabolites ranged between 2.6 and 50.8 µg As/g creatinine (mean 16.9). The authors estimated a mean urinary excretion of 13.2 µg As/g creatinine, in workers exposed to fly ash, from an 8-h TWA exposure to 10 µg As/m³, suggesting that the bioavailability of arsenic in coal fly ash is approximately one-third that seen in smelters. Concentrations of arsenic in the breathing zone of underground gold-miners in Ontario (Canada) were reported to range between 2.4 and 5.6 µg/m³ (geometric mean) with urinary arsenic concentrations reported to range between 23.5 and 25.9 µmol As/mol creatinine (Kabir & Bilgi, 1993). The median total urinary arsenic concentration of the miners was significantly higher than that of a control group, but no correction was made for differences in dietary habits of the two groups. In a study relating arsenic exposures to lung cancer among tin-miners in Yunnan province (China), Taylor et al. (1989) reported mean concentrations of airborne arsenic to range from 0.42 mg/m³ in 1951 to 0.01 mg/m³ in 1980.

5.4 Total human intake of arsenic from all environmental pathways

For healthy humans who are not occupationally exposed the most significant pathway of exposure to arsenic is through the oral intake of food and beverages. In areas with elevated concentrations of arsenic in drinking-water, this source make a significant contribution to the total intake of inorganic arsenic. For example, a consumption of 1.4 litres of drinking-water containing > 50 µg As/litre could provide over 70 µg inorganic arsenic compared to an estimated daily intake, based on very preliminary data, of 12–14 µg inorganic arsenic from typical North American diets (Yost et al., 1998).

As shown in Table 12, the total estimated daily dietary intake of arsenic may vary widely, mainly because of wide variations in the consumption of fish and shellfish. Data in Table 12 are for total arsenic intake and do not reflect the possible variation in intake of the more toxic inorganic arsenic species (see sections 5.2.2 and 5.2.3). In areas where drinking-water contains > 50 µg As/litre, water may be the major source of inorganic arsenic. All other routes of intakes of arsenic (intranasal and dermal) are of minor importance in comparison to the oral route (ATSDR, 1993). For example, inhalation would add about 1 µg As/day from airborne particulates and approximately 6 µg As /day may be inhaled from 20 cigarettes.

The most appropriate approach to determining the internal dose of inorganic arsenic in individuals in specific populations is to measure the arsenic species in urine. Concentrations of total urinary arsenic and metabolites of inorganic arsenic (inorganic arsenic, MMA and DMA) provides estimates of the exposure (uptake) to total arsenic and inorganic arsenic, respectively (see section 6.3). Reported concentrations of metabolites of inorganic arsenic in urine with no known exposure to arsenic are generally < 10 µg/litre in European countries (Apel & Stoeppler, 1983; Valkonen et al., 1983; Foa et al., 1984; Vahter & Lind, 1986; Andren et al., 1988; Jensen et al., 1991; Buchet et al., 1996; Trepka et al., 1996; Kristiansen et al., 1997; Kavanagh et al., 1998). Similar or slightly higher concentrations are reported from studies in some parts of the USA (Smith et al., 1977, Morse et al., 1979; Binder et al., 1987; Kalman et al., 1990; Pollisar et al., 1990; Gottlieb et al., 1993; Bates et al., 1995; Lewis et al., 1999) and around 50 µg/litre in Japan (Yamamura et al., 1979; Yamauchi et al., 1992). In West Bengal and Bangladesh arsenic concentrations > 1 mg/litre have frequently been observed (Chatterjee et al., 1995; Das et al., 1995).

The concentration of arsenic metabolites in urine correlates well with the concentration of arsenic in the drinking-water. However, the relationship may vary considerably depending on the amount of water consumed and the amount of water used for preparation of drinks and food. For example, studies from California and Nevada (USA) showed that a water concentration of 400 µg/litre corresponded to about 230 µg/litre in urine (total arsenic) and 100 µg/litre in water corresponded to 75 µg/litre in urine (Valentine et al., 1979). Similarly, people in Alaska drinking water containing about 400 µg/litre had on average 180 µg/litre in urine, and those drinking water containing 50–100 µg/litre had on average 45 µg/litre in urine (Harrington et al., 1978). Thus, urinary arsenic concentration was about half of that in the water. However, people living in areas of northern Argentina with drinking-water containing 200 µg/litre had much higher arsenic concentrations in the urine (metabolites of inorganic arsenic) – on average 250–450 µg/litre (Vahter et al., 1995a; Concha et al., 1998a). The fluid intake of these people consisted mainly of drinking-water or drinks prepared at home from the drinking-water. Also, most of the food consumed was prepared at home using the local drinking-water. In areas in the north-east of Taiwan where drinking-water concentrations were 50–300 µg/litre, people had similar concentrations in the urine; about 140 µg/litre (Chiou et al., 1997a).

Soil (section 5.2.4) may be a significant source of arsenic intake, particularly for children. However, the bioavailability may vary considerably.

Some studies have been conducted for the purpose of evaluating whether there is an increased body burden of arsenic in children living near arsenic-contaminated sites relative either to children from areas of low arsenic exposure or to adults. For example, Binder et al. (1987) reported that total urinary arsenic excretion was significantly increased in children living in a Montana (USA) community with high levels of arsenic in soil (average ~400–700 mg/kg) compared to a community with low arsenic levels in soil (44 mg/kg). In first-morning urine samples taken in July in the high-arsenic community, mean total arsenic in urine averaged 54 µg/litre (53.8 µg/g creatinine) compared to 16.6 µg/litre (17.1 µg/g creatinine) in the low-arsenic community.

Trepka et al. (1996) studied differences in arsenic exposure among children of different age groups in various areas of Germany, as assessed by urinary arsenic excretion. They reported no marked age- or gender-related differences, although urinary arsenic excretion was significantly increased in children from the most polluted area (5.1 µg/litre vs. 4 µg/litre in the control area). However, the authors did not consider this increase to be toxicologically significant. In contrast, Diaz-Barriga et al. (1993) reported increases in urinary arsenic in children living closest to a copper smelter (median soil levels ~500 mg/kg arsenic; range 69–594 µg/g creatinine in urine) compared to children living 7–25 km away (median soil levels ~11–14 mg/kg arsenic). Urinary arsenic excretion (normalized to creatinine) was more than doubled, and arsenic levels in hair were more than 10-fold higher.

6. KINETICS AND METABOLISM IN LABORATORY ANIMALS AND HUMANS

Humans are exposed to many different forms of inorganic and organic arsenic species (arsenicals) in food, water and other media. Study of the kinetics and metabolism of arsenicals in animals and humans can thus be quite complex, as a result of differences in physicochemical properties and bioavailability of the various forms of arsenic. Arsenic metabolism is also characterized by relatively large qualitative and quantitative interspecies differences. Given the relatively large interspecies differences in arsenic metabolism, and that there is considerable information on human metabolism of arsenicals, discussion of animal studies focuses on areas where human data is inadequate or where animal data can serve to aid in the interpretation of toxic effects caused by arsenicals. This chapter covers both inorganic and organic arsenicals.

6.1 Inorganic arsenic

The metabolism and disposition of inorganic arsenic may depend on its valence state, particularly at high doses. The two most common valence states to which humans might be environmentally exposed are the trivalent and pentavalent forms. Since these two forms are readily interconverted, studies cited in this review were evaluated with particular attention to whether methods used were appropriate to ensure that inorganic arsenic was maintained in the intended valence state until the time of administration. Arsenite, but not arsenate, exists mainly in the non-ionized form at physiological pH and relatively low Eh.

6.1.1 Absorption

6.1.1.1 Respiratory deposition and absorption

Human inhalation exposure to inorganic arsenic can occur as a consequence of industrial activity (e.g. smelting of ores) and energy production (e.g. coal-fired power plants), and during cigarette smoking. Arsenic in air exists on particulate matter and thus respiratory absorption of arsenic is a two-stage process, involving deposition of the particles on to airway and lung surfaces, followed by absorption of arsenic from deposited particulates. The extent of deposition of inhaled arsenic will depend largely on the size of the inhaled particulates, and absorption of deposited arsenic is highly dependent on the solubility of the chemical form of arsenic.

a) Animal studies

Quantification of the relative amount of airborne arsenic that is deposited in various parts of the respiratory tract is not possible because there is a lack of such animal inhalation studies. However, intratracheal instillation studies provide information on the extent of absorption of various chemical forms of inorganic arsenic. In general, solubility appears to be the most important physicochemical property determining the extent of lung clearance, although wetting capacity and pulmonary toxicity may also have an important influence. It is important to note that clearance of particulates by the mucociliary escalator may also result in oral exposure.

Pershagen et al. (1982) found that lung concentrations of arsenic in hamsters given weekly intratracheal instillations of As₂O₃, arsenic trisulfide or calcium arsenate differed by a factor of approximately 10-fold after 4 weeks. The much more rapid clearance of As₂O₃ was attributed to its being much more soluble in vivo than the other two arsenicals. The authors speculated that the clearance of calcium arsenate was much slower than that of arsenic trisulfide because of its higher wetting capacity, which would result in more calcium arsenate being transported to the alveolar regions of the lung where clearance is slower. The authors also indicated that the pulmonary toxicity of calcium arsenate may have impaired normal clearance mechanisms which would have prolonged lung retention.

Marafante & Vahter (1987) reported that the extent of absorption of inorganic arsenicals from the lungs of hamsters after intratracheal instillation was directly correlated with their in vivo solubility as determined by the amount of radiolabelled arsenical retained at an intramuscular injection site. The lung retention of arsenic (2 mg As/kg) 3 days after an intratracheal instillation of sodium arsenite, sodium arsenate, arsenic trisulfide and lead arsenate was respectively 0.06%, 0.02%, 1.3% and 45.5% of the dose. Similar observations have been reported by Buchet et al. (1995), who found that 24 h after a single intratracheal instillation of soluble arsenic salts in hamsters (NaAsO₂ and Na₂HAsO₄) at 50 and 100 µg/kg, the amount of arsenic detected in the lung was not different from that found in the control animals. Buchet et al. (1995) also observed that both lung retention and urinary excretion indicate a prolonged contact of the lung tissue with particulate arsenic rather than soluble arsenic salts. Some 48 h after intratracheal administration of arsenic in the form of fly ash or copper smelter dust, lung retention amounted to 25–35% of the administered dose (50–100 µg As/kg). Rosner & Carter (1987) also reported, based on results of intratracheal instillation studies in hamsters, that the more soluble forms of arsenic, sodium arsenate and sodium arsenite (5 mg/kg dose), had a relative bioavailability 10-fold greater than gallium arsenide (GaAs). Webb et al. (1987) also reported that decreasing the particle size by a factor of 2 increased the in vivo dissolution rate and toxicity of GaAs in rats after intratracheal instillation.

b) Human studies

The available human data are insufficient to allow quantitative estimation of regional arsenic deposition in the respiratory tract. Occupational studies in which both the concentration of inorganic arsenic in the breathing zone and the urinary excretion of inorganic arsenic and its metabolites were determined provide information on arsenic absorption. These studies (e.g. Vahter et al., 1986; Yamauchi et al., 1989a; Offergelt et al., 1992; Hakala & Pyy, 1995; Yager et al., 1997) demonstrate that excretion of inorganic arsenic and sometimes total arsenicals and methylated metabolites are significantly increased in workers exposed to arsenic in their breathing zone. This indicates that arsenic is absorbed from the respiratory tract, but does not provide sufficient information for quantitative estimation of arsenic absorption after inhalation because the contribution of oral exposure after mucociliary clearance – and in some instances probably also from diet and drinking-water – cannot be assessed.

Comparison of studies relating occupational arsenic exposure in different industrial settings to urinary arsenic excretion suggests that there are differences in respiratory absorption depending on the form of arsenic. Using equations relating urinary arsenic excretion to air concentrations, Yager et al. (1997) noted that in several studies the predicted urinary arsenic output for workers exposed to 10 µg/m³ arsenic was more than one-third lower for boiler maintenance workers in a coal-fired power plant than it was for copper-smelter workers. This finding was attributed to the fact that the arsenic in coal fly ash in their study was predominantly in the form of calcium arsenate, whereas in the copper smelter work environment the arsenic was in the form of As₂O₃. Such an interpretation is consistent with the much greater retention of calcium arsenate than As₂O₃ in hamster lung that was reported by Pershagen et al. (1982).

6.1.1.2 Gastrointestinal absorption

Arsenic can be absorbed from the gastrointestinal tract after ingestion of arsenic-containing food, water, beverages or medicines, or as a result of inhalation and subsequent mucociliary clearance. The bioavailability of ingested inorganic arsenic will vary depending on the matrix in which it is ingested (e.g. food, water, beverages, soil), the solubility of the arsenical compound itself and the presence of other food constituents and nutrients in the gastrointestinal tract.

a) Animal studies

Soluble arsenates and arsenites are rapidly and extensively absorbed from the gastrointestinal tract of common laboratory animals after a single oral dose (Table 13). The mouse data of Vahter & Norin (1980) indicate that arsenite may be more extensively absorbed from the gastrointestinal tract than arsenate at lower doses (e.g. 0.4 mg As/kg), whereas the reverse appears to occur at higher doses (e.g. 4.0 mg As/kg). In these same studies (Vahter & Norin, 1980) about the same percentage faecal elimination was observed following the same dose given orally and subcutaneously, indicating nearly complete gastrointestinal absorption (Table 13).

Studies conducted by Odanaka et al. (1980) suggest that much less pentavalent arsenic is absorbed from the gastrointestinal tract after oral administration – 48.5% of dose (5 mg/kg) in urine, compared to the 89% of dose (4 mg/kg) in urine found by Vahter & Norin (1980) (Table 13). This difference may be attributable to the fact that the mice in the study of Vahter & Norin were not fed for at least 2 h before and 48 h after dosing, whereas the mice in the Odanaka et al. studies were not food restricted. Studies by Kenyon et al. (1997) suggest that feeding a diet lower in fibre or "bulk" to female B6C3F₁ mice increased absorption of sodium arsenate by ~10% compared to a standard rodent chow diet, after a single oral dose of 5 mg As/kg.

Table 13. Cumulative 48-h elimination (% of dose) of As in urine and faeces of laboratory animals after oral and parenteral administration of inorganic As

^a Data given as mean ± SEM

Yamauchi et al. (1986b) studied the absorption and metabolism of GaAs (a relatively insoluble arsenical compared to sodium arsenite and sodium arsenate) in hamsters after a single oral or intraperitoneal dose. Faecal elimination of total arsenic after 5 days averaged 87.5 ± 13.8%, 79.4 ± 10.6% and 77.9% after oral doses of 10, 100 and 1000 mg/kg GaAs, respectively. However, after a single intraperitoneal dose of 100 mg/kg GaAs only 0.38% of the total arsenic dose was eliminated in faeces after 5 days. During this same time period less than 1% of the dose was eliminated in urine irrespective of the route of administration, indicating that GaAs is minimally absorbed from the gastrointestinal tract. It is noteworthy that a consistently greater amount of DMA was excreted in urine over time after intraperitoneal administration than after oral administration. This indicates that arsenic liberated from GaAs undergoes methylation, and is consistent with results reported in hamsters after intratracheal instillation of GaAs (Rosner & Carter, 1987).

The bioavailability of arsenic from soils has been assessed using various animal models because this can be a significant issue in risk assessment for contaminated industrial sites where there is potential for arsenic exposure via soil ingestion. As summarized in Table 14, these studies indicate that oral bioavailability of arsenic in a soil or dust vehicle is often lower than that of the pure soluble salts typically used in toxicity studies. However, bioavailability is substantially dependent on the soil type. A study by Vahter (1988) showed that although some soil samples from former wood-treatment plants containing 1.1 mg As/g were highly toxic, other soil samples containing 9 mg As/g were without any effect when tested in mice. Similarly, the mean relative bioavailability of arsenic in mining wastes compared to that of sodium arsenate administered to young swine and analysed as arsenic in urine was found to range from 7–52% (US EPA, 1996). Davis et al. (1992) have pointed out that this is due mainly to mineralogical factors which control solubility in the gastrointestinal tract, such as solubility of the arsenic-bearing mineral itself and encapsulation within insoluble matrices (e.g. silica). The comparative bioavailabilities of arsenic in soil from a site contaminated with copper chrome acetate (CCA) and soil contaminated by arsenic solutions used in cattle tick control were reported by Ng & Moore (1996), using a rat model. Soil from the cattle dip site had a bioavailability of 8.1 ± 4%, 14.4 ± 7% and 60 ± 3.4% when compared with orally administered sodium arsenite, calcium arsenite and sodium arsenate respectively. For CCA-contaminated soil the corresponding comparative bioavailabilities were 13.0 ± 4.5%, 32.2 ± 11.2% and 38.0 ± 13.2%. Ng et al. (1998b), also using a rat model, have reported the absolute bioavailability of arsenic in soils containing 32–1597 µg As/kg (0.32–56% arsenite) from a combination of arsenical pesticides and natural geological formations in a residential area. The absolute bioavailability ranged from 1.02 to 9.87% relative to arsenite and from 0.26 to 2.98% relative to arsenate.

Table 14. Oral bioavailability of As from soil, based on studies in laboratory animals

Several older studies reviewed in the previous IPCS arsenic document (IPCS, 1981, sections 6.1.1 and 6.2) demonstrated that composition of the diet can alter gastrointestinal absorption of arsenic. Some more recent studies have examined the mechanism of arsenical uptake and interaction with nutrients at the intestinal level. Gonzalez et al. (1995), using isolated perfused rat small intestine, demonstrated that uptake of pentavalent arsenic is carried out by a saturable transport process and that addition of phosphate markedly decreased arsenic absorption, most likely because arsenate and phosphate can share the same transport mechanism. Hunder et al. (1993), using isolated rat jejunal segments, found that increasing concentrations of arsenite (2.5–250 µmol/litre) and arsenate (2.5–2500 µmol/litre) caused a dose-dependent decrease in the intestinal transfer of water, sodium, glucose and leucine, with arsenite being about 5-fold more potent than arsenate.

b) Human studies

In common with studies in experimental animals, controlled ingestion studies in humans indicate that trivalent and pentavalent arsenic are both well absorbed from the gastrointestinal tract (Table 15). For example, Pomroy et al. (1980) reported that healthy male human volunteers excreted 62.3 ± 4.0% of a 0.06-ng dose of ⁷⁴As-arsenic acid (As(V)) in urine over a period of 7 days, whereas only 6.1 ± 2.8% of the dose was excreted in the faeces. Few other controlled human ingestion studies have actually reported data on both urine and faecal elimination of arsenic. However, between 45% and 75% of the dose of various trivalent forms of arsenic is excreted in the urine within a few days (Table 15), which suggests that gastrointestinal absorption is both relatively rapid and extensive.

Table 15. Metabolism and urinary excretion of inorganic and organic arsenicals in humans after experimental administration

Subjects in all studies cited were adult males

^a IAs, sum of As(III) and As(V)

6.1.1.3 Dermal absorption

Wester et al. (1993) studied the percutaneous absorption of arsenic acid (H₃AsO₄) from water and soil both in vivo using rhesus monkeys and in vitro with human skin. In vivo, absorption of arsenic acid from water (loading 5 µl/cm² skin area) was 6.4 ± 3.9% at the low dose (0.024 ng/cm²) and 2.0 ± 1.2% at the high dose (2.1 µg/cm²). Absorption from soil (loading 0.04 g soil/cm² skin area) in vivo was 4.5 ± 3.2% at the low dose (0. 04 ng/cm²) and 3.2 ± 1.9% at the high dose (0.6 µg/cm²). Thus, in vivo in the rhesus monkey, percutaneous absorption of arsenic acid is low from either soil or water vehicles and does not differ appreciably at doses more than 10 000-fold apart. Wester et al. (19993) also reported that for human skin, at the low dose, 1.9% was absorbed from water and 0.8% from soil over a 24-h period.

Limited data suggest that the in vitro percutaneous absorption of inorganic arsenicals may differ substantially depending on chemical form or species used. Rahman et al. (1994) evaluated the percutaneous absorption of sodium arsenate in vitro using clipped full-thickness dorsal skin of B6C3F₁ mice. They found that a constant fraction of the applied dose was absorbed over a 24-h period irrespective of dose level (5–500 ng or 0.36–360 mg/kg for soil), but that the vehicle or vehicle volume had significant effects. Using 100 µl water as a vehicle resulted in ~60% of applied dose being absorbed, whereas using a volume of 250 µl water resulted in ~37% absorption, which was about the same percentage absorbed if the chemical was applied in the solid form. Absorption of sodium arsenate from soil was minimal (< 0.3%), which is similar to what was reported for arsenic acid in the studies of Wester et al. (1993).

6.1.1.4 Placental transfer

a) Animal studies

Both older and more recent studies have documented the ability of trivalent and pentavalent inorganic arsenic to cross the placenta in laboratory animals. Lindgren et al. (1984) reported that in pregnant mice given a single intravenous injection (4 mg As/kg) of sodium arsenate or sodium arsenite, both forms passed through the placenta easily and to approximately the same extent. These investigators also reported that the rate of placental transfer was lower in a marmoset monkey (non-methylating species) injected intravenously with arsenite than in mice, and suggested that this was a consequence of stronger binding in maternal tissues.

Hood et al. (1987) compared the fetal uptake of sodium arsenate after oral (40 mg/kg) or intraperitoneal (20 mg/kg) administration to pregnant CD-1 mice on day 18 of gestation. Arsenic levels peaked later and over 5-fold lower in fetuses of mice dosed orally, most likely reflecting both slower uptake from the gastrointestinal tract and greater opportunity for methylation in the liver before the arsenic reached the systemic circulation. The quantity of dimethylated metabolite present in the fetuses rose over time (to ~80% of total metabolites present for both routes of administration) and remained relatively constant from ~10 h after dosing until the study ended, 24 h after dosing.

Hood et al. (1988) also compared the fetal uptake of sodium arsenite after oral (25 mg/kg) or intraperitoneal (8 mg/kg) administration to mice that were 18 days pregnant. As was the case with arsenate, injected mice achieved both higher fetal and placental levels of arsenic more quickly than did mice dosed orally. Both valence forms followed similar time-course trends after oral administration. However, levels of arsenic in fetuses of dams injected with arsenite reached a plateau 12–24 h after dosing, whereas levels of arsenic in fetuses of dams injected with arsenate peaked at 2–4 h after dosing and then declined quickly. The proportion of arsenic present in fetuses as methylated metabolite increased over time to 88% and 79% after oral and intraperitoneal administration, respectively. A higher fraction of monomethylated arsenic was present in fetuses of dams dosed with arsenite than with arsenate. The authors concluded that much of the arsenic reaching the fetus has already been transformed to the less acutely toxic methylated metabolites.

b) Human studies

Case reports of arsenic poisoning in pregnant women resulting in death of the fetus accompanied by toxic levels of arsenic in fetal organs and tissues demonstrate that arsenite (As₂O₃) readily passes through the placenta (Lugo et al., 1969; Bollinger et al., 1992). In a more recent study, Concha et al. (1998b) reported that arsenic concentrations were similar in cord blood and maternal blood (~9 µg/litre) of maternal–infant pairs exposed to drinking-water containing high levels of arsenic (~200 µg/litre). Another study of an "unexposed" population in the southern USA found that concentrations of arsenic in cord blood and maternal blood (about 2 µg/litre) were also similar, and suggests that arsenic readily crosses the placenta (Kagey et al., 1977).

6.1.2 Distribution

6.1.2.1 Fate of inorganic arsenic in blood

a) Animal studies

Inorganic arsenic is rapidly cleared from the blood in most common laboratory animals, including mice, rabbits, and hamsters (Vahter & Norin, 1980; Marafante et al., 1982, 1985; Yamauchi & Yamamura, 1985). The notable exception to this is the rat in which the presence of arsenic is prolonged owing to accumulation in erythrocytes (Vahter, 1981; Marafante et al., 1982; Lerman & Clarkson, 1983). For example, Marafante et al. (1982) reported that levels of arsenic in erythrocytes were 2-fold, 102-fold and 268-fold higher at 1, 16 and 48 h after dosing in rats compared to rabbits that received the same intraperitoneal dose of arsenite; in the plasma of these same animals the rat : rabbit ratio of arsenic never exceeded 1. Lerman & Clarkson (1983) further noted that higher levels of arsenic were achieved much more rapidly in the blood of rats dosed intravenously with arsenite than with arsenate, and that 95% or more of the arsenic in erythrocytes was in the form of DMA by 4 h after dosing. It appears that rat haemoglobin specifically binds DMA, and this greatly increases the biological half-life of inorganic arsenic and DMA in rats (Vahter, 1981; Vahter et al., 1984).

Although clearance of both arsenate and arsenite from blood in other mammalian species is rapid, differences dependent on both valence state and dose have been observed. Vahter & Norin (1980) reported that at a high oral dose of arsenic (4 mg As/kg), arsenite-dosed mice had a higher erythrocyte to plasma ratio (~2–3), whereas in arsenate-dosed mice the ratio was much closer to 1. No such difference was observed at a lower oral dose (0.4 mg As/kg) of arsenate or arsenite. Delnomdedieu et al. (1994b) investigated the in vitro uptake of arsenite and arsenate in intact rabbit erythrocytes. They reported that ~76% of arsenite, compared to ~25% of arsenate, was taken up within 0.5 h, and that arsenite subsequently bound with intracellular glutathione (GSH), whereas arsenate entered the phosphate pathway, depleting ATP and increasing inorganic phosphate levels.

Yamauchi & Yamamura (1985) characterized the forms of arsenic present and their distribution over time in whole blood, plasma and erythrocytes in male Syrian golden hamsters given a single oral dose of 4.5 mg/kg As₂O₃ (As(III)). Arsenic levels in whole blood had dropped to control levels by 72 h after dosing, indicating rapid clearance from the blood. In whole blood, inorganic arsenic, MMA and DMA concentrations peaked 1, 12 and 24 h after dosing, respectively; the levels of MMA and DMA achieved relative were 1/3 and 1/10 as high as those of inorganic arsenic. Inorganic arsenic and MMA were found mainly in erythrocytes and DMA occurred chiefly in plasma.

Marafante et al. (1985) also characterized the forms of arsenic present and their distribution over time in blood of rabbits dosed with arsenate. Male New Zealand white rabbits were injected intravenously with 0.4 mg As/kg and blood samples were taken at intervals from 15 min to 6 h after dosing. Within 15 min after dosing, 10% of the arsenic in the plasma was in the form of arsenite, 30% was present as arsenate and 60% was bound to plasma proteins. Arsenate was rapidly cleared from plasma (first-order t_1/2 ~1 h). Both arsenite and plasma protein-bound arsenic exhibited biphasic kinetics, with half-times of 10 min and 2 h for arsenite and 15 min and 2.5 h for protein-bound arsenic. In studies of giant Flemish rabbits using carrier-free ⁷⁴As (arsenate), DeKimpe et al. (1996) reported biphasic blood clearance rates of about 2 h and 58 h in plasma and 8.6 h and 170 h in erythrocytes. Taken together, the findings of Marafante et al. (1985) and DeKimpe et al. (1996) suggest a triphasic, rather than biphasic, clearance in plasma. These findings indicate that binding of arsenic to plasma protein is not strong. The kinetics and concentration of arsenite in erythrocytes were quite similar to that seen in plasma, but arsenate concentrations in erythrocytes were only about 1/10 as high. DMA levels peaked at about 4 h after dosing in both plasma and erythrocytes, but the concentration in erythrocytes was only about 20% of that found in plasma.

b) Human studies

Inorganic arsenic is reported to be rapidly cleared from blood. Results from some older studies reviewed in the previous arsenic IPCS document (IPCS, 1981, section 6.1.2), suggest that the kinetics of arsenic clearance in plasma and erythrocytes are similar, although levels in erythrocytes tended to be approximately 3-fold higher a few hours after exposure (similar to findings in laboratory animals). Mealey et al. (1959) measured the plasma and erythrocyte levels of radioactive arsenic after intravenous injection of labelled arsenite. The rate of decline of arsenic in the erythrocytes was comparable with that in plasma, but the erythrocytes contained about 3 times more arsenic than the plasma 10 h after the injection. The plasma curve showed a three-compartment model (IPCS, 1981, section 6.1.3). The first half-life seemed to be very short, and the bulk of the arsenic was removed from the plasma at a high rate. Some 24 h after dosing, less than 0.1% of the dose remained. The second phase of the curve showed a half-time of about 30 h. The third phase of the curve, beginning about 1 week after the injection, showed a very low rate of disappearance with a half-time of over 200 h.

Zhang et al. (1996a,b, 1997, 1998a,b) have reported on the distribution of arsenical species in serum and arsenic–protein binding in serum of patients with renal disease. The predominant species of arsenic present in serum were DMA (~15–30%) and arsenobetaine (~54–76%), with the remainder being protein-bound and inorganic arsenic; MMA was undetectable (Zhang et al., 1996a, 1997). Zhang et al. (1998a,b) further reported that only inorganic arsenic was bound to serum proteins, and that transferrin is the main carrier protein. It should be noted that since individuals with renal disease tend to accumulate arsenic in serum, these results may not be typical of the general population.

6.1.2.2 Tissue distribution

a) Animal studies

Studies in rabbits, rats, mice, hamsters and monkeys demonstrate that arsenic, administered orally or parenterally, in either the trivalent or pentavalent form, is rapidly distributed throughout the body (Lindgren et al., 1982; Marafante et al., 1982; Vahter et al., 1982; Vahter & Marafante, 1985; Yamauchi & Yamamura, 1985). Many of these studies have used radiolabelled arsenic, and it is noteworthy that arsenic-derived radioactivity is generally present in all tissues examined (Lindgren et al., 1982; Marafante et al., 1982; Vahter et al., 1982; Vahter & Marafante, 1985).

Comparative studies of arsenate and arsenite distribution at comparable dose levels provide insights on the influence of valence state on arsenic distribution. Lindgren et al. (1982) studied the distribution of arsenic in male C57BL mice intravenously administered 0.4 mg As/kg as either sodium arsenate or sodium arsenite at 0.5, 6, 24 and 72 h after dosing. Highest concentrations of arsenic-derived radioactivity were present in liver, kidney and gallbladder at 0.5 h after administration of either arsenate or arsenite, reflecting their rapid elimination (see section 6.1.4). Arsenate administration resulted in much lower arsenic concentrations in liver and gallbladder, but higher concentrations in kidney compared to administration of arsenite within 0.5 h after dosing, indicating valence-dependent differences in route of elimination. In general, concentrations of arsenic in organs tended to be higher after administration arsenite than of arsenate, with the notable exception of the skeleton at all time points (Table 16). This latter finding was ascribed to arsenate being a structural analogue of phosphate and substituting for it in the apatite crystal of bone. The greater retention of arsenite in tissues is a consequence of its reactivity and binding with tissue constituents, most notably sulfhydryl groups (Vahter & Marafante, 1983).

Table 16. Comparison of tissue distribution over time in mice given a single intravenous injection of ⁷⁴As-As (0.4 mg As/kg) as either sodium arsenate or sodium arsenite^a

^a Table assembled from Lindgren et al. (1982)

Both species-specific and valence-state-dependent differences have been demonstrated in the biliary excretion of arsenic. Studies reviewed in the previous arsenic document (IPCS, 1981, 6.1.2) indicate that excretion of trivalent arsenic into the bile is much more extensive in rats than in rabbits or dogs (Klaassen, 1974). The studies of Lindgren et al. (1982) suggest that arsenite is excreted to a greater extent than arsenate in the bile of mice (Table 16); these authors also attribute the higher concentrations of arsenite-derived radioactivity in the duodenum of mice to greater biliary excretion of arsenite. Excretion of arsenite into the bile of rats is also more rapid and efficient than that of arsenate – 19% vs. 6% of the dose in 2 h (Gyurasics et al., 1991). Mechanistic studies indicate that transport of either arsenate or arsenite into the bile of rats is dependent on GSH, since agents that decrease hepatobiliary transport of GSH (e.g. diethyl maleate) also decrease hepatobiliary transport of arsenic (Alexander & Aaseth, 1985; Gyurasics et al., 1991). It has been demonstrated in recent studies that arsenite as well as other trivalent arsenicals directly form complexes with GSH (Scott et al., 1993; Delnomdedieu et al., 1994a).

Numerous studies reveal that skin, hair, and tissues high in squamous epithelium (e.g. mucosa of the oral cavity, oesophagus, stomach and small intestine) have a strong tendency to accumulate and maintain higher levels of arsenic (e.g. Lindgren et al., 1982; Yamauchi & Yamamura, 1985). This is apparently a function of the binding of arsenic to keratin in these tissues (Lindgren et al., 1982). Autoradiographic studies have also revealed a tendency for arsenic to accumulate in the epididymis, thyroid and lens of the eye of mice (Lindgren et al., 1982).

Arsenic can cross the blood–brain barrier; it is found in brain tissue after oral or parenteral administration of trivalent or pentavalent inorganic arsenic in all species studied (e.g. see Table 16). However, the levels are uniformly low both across time and relative to other tissues, which indicates that arsenic (when administered in the form of sodium salts) does not readily cross the blood–brain barrier or accumulate in brain tissue after acute dosing (Lindgren et al., 1982; Marafante et al., 1982; Vahter et al., 1982; Vahter & Marafante, 1985; Yamauchi & Yamamura, 1985; Itoh et al., 1990).

Relatively few studies have examined the distribution of arsenic metabolites in tissues, owing to limitations in availability of appropriate analytical techniques. Given the vigorous treatment necessary to extract arsenicals from tissues, and the ease with which arsenate and arsenite are interconverted, any reports that distinguish between arsenate and arsenite in tissues should be interpreted with caution. DeKimpe et al. (1996) reported the tissue distribution of arsenic metabolites at 4, 20 and 120 h after intraperitoneal injection of a trace amount of ⁷⁴As-arsenate in male Flemish giant rabbits. The analytical methodology used was ion-exchange chromatography separation of ultrafiltrates with radiometric detection. The predominant metabolite present in tissues was DMA, followed by the inorganic arsenic species, with MMA being generally detected in all tissues, although making up a smaller percentage of the total metabolites in most cases. The percentage of total metabolite present as DMA increased steadily in bone marrow, heart, liver, muscle, pancreas, small intestine and spleen, but levelled off or declined in kidney and lung. Marafante et al. (1982) also found that inorganic arsenic was the predominant form of arsenic in rat and rabbit liver and kidney ultrafiltrable fraction (cut-off 25 000 Da) 1 h after intraperitoneal injection of 50 µg As/kg sodium arsenite; the analytical methodology used was ion-exchange chromatographic separation with radiometric detection. However, the fraction of DMA was higher at 16 h after injection in kidney of both rats and rabbits, but only in the liver of rats. The fraction present as MMA was uniformly low, generally less that one-tenth that of inorganic arsenic.

Yamauchi & Yamamura (1985) studied the tissue distribution over time of arsenic metabolites in male Syrian golden hamsters given a single oral dose of 4.5 mg/kg As₂O₃. Hydride generation atomic absorption spectrophotometry (HGAAS) with a cold trap was used to speciate arsenicals in whole tissues after alkaline digestion. The predominant form of arsenic present in all tissues at 1, 6, 12, 24, 72 and 120 h after dosing was inorganic arsenic. Interestingly, concentrations of MMA in tissue were uniformly 2–4-fold greater than DMA at all time points, although much more DMA (22% of the dose) was excreted in urine over 5 days than MMA (2.5% of the dose). Highest concentrations of MMA were achieved in lungs and spleen at 12 h and kidney at 24 h, and highest concentrations of DMA occurred in liver, lung and kidney at 24 h.

The subcellular distribution of total arsenic administered as either sodium arsenate or sodium arsenite has been studied in mice, rats, rabbits and marmoset monkeys (Marafante et al., 1982; Vahter et al., 1982; Vahter & Marafante, 1983; Vahter & Marafante, 1985). In general the subcellular localization and retention of arsenic accounts for its much slower elimination in rats than in other species. In rats arsenic is strongly associated with high-molecular-weight cellular components in liver and kidney, whereas in rabbits it is associated with low-molecular-weight, more readily diffusible cellular components (Marafante et al., 1982). This research group also reported that in the marmoset monkey arsenic, administered as arsenite or arsenate, shows a unique strong tendency to bind with the rough endoplasmic reticulum in the liver that they had not observed in other laboratory animals (Vahter et al., 1982; Vahter & Marafante, 1985).

b) Human studies

As in experimental animals, postmortem analysis of human tissues reveals that arsenic is widely distributed in the body after either long-term relatively low-level exposure or poisoning (Dang et al., 1983; Gerhardsson et al., 1988; Raie, 1996). Dang et al. (1983) used neutron activation analysis (NAA) to measure total arsenic in various tissues of individuals (age and sex not specified) dying in accidents in the Bombay area (India) (Table 17). Notable results from this study are that arsenic concentrations are quite low in both blood and brain relative to other tissues and that arsenic concentration in any given tissue was quite variable.

Table 17. As levels in human tissues from accident victims in Bombay area of India^a