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% po