西安建筑科技大学：《水资源利用与保护》参考资料_Wastewater as a resource_Review of Arsenic in natural water - Smedly

Applied Geochemistry PERGAMON Applied Geochemistry 17(2002)517-568 www.elsevier.com/locate/apgeochem Review a review of the source behaviour and distribution of arsenic in natural waters P L. Smedley*, D.G. Kinniburgh on OX10 8BB UK Received 1 March 2001; accepted 26 October 2001 Editorial handling by R. Fuge Abstract The range of As concentrations found in natural waters is large, ranging from less than 0.5 ug I-I to more than 5000 I-I. Typical concentrations in freshwater are less than 10 Hg l-l and frequently less than 1 ug l-l. Rarely, much higher concentrations are found, particularly in groundwater. In such areas, more than 10% of wells may be ' affected (defined as those exceeding 50 ug 1-)and in the worst cases, this figure may exceed 90%. Well-known high-As groundwater areas have been found in Argentina, Chile, Mexico, China and Hungary, and more recently in West Bengal(India), Bangladesh and Vietnam. The scale of the problem in terms of population exposed to high As con- entrations is greatest in the Bengal Basin with more than 40 million people drinking water containing'excessive'As These large-scale 'natural As groundwater problem areas tend to be found in two types of environment: firstly, inland or closed basins in arid or semi-arid areas, and secondly, strongly reducing aquifers often derived from alluvium. Both environments tend to contain geologically young sediments and to be in flat, low-lying areas where groundwater flow is sluggish. Historically, these are poorly flushed aquifers and any As released from the sediments following burial has been able to accumulate in the groundwater. Arsenic-rich groundwaters are also found in geothermal areas and, on a more localised scale. in areas of mining activity and where oxidation of sulphide minerals has occurred. The As content of the aquifer materials in major problem aquifers does not appear to be exceptionally high, being normally in the range 1-20 mg kg. There appear to be two distincttriggers' that can lead to the release of s on a large scale. The first is the development of high pH(>8.5) conditions in semi-arid or arid environments usually as a result of the combined effects of mineral weathering and high evaporation rates. This pH change leads either to the desorption of adsorbed As(especially As(V) species) and a range of other anion-forming elements (v, B, F, Mo, Se and u) from mineral oxides, especially Fe oxides, or it prevents them from being adsorbed. The second trigger is the development of strongly reducing conditions at near-neutral pH values, leading to the desorption of As from mineral oxides and to the reductive dissolution of Fe and Mn oxides, also leading to As release. Iron (II) and As(lll are relatively abundant in these groundwaters and SO4 concentrations are small(typically 1 mg I-I or less). Large concentrations of phosphate, bicarbonate, silicate and possibly organic matter can enhance the desorption of As because of competition for adsorption sites. A characteristic feature of high groundwater As areas is the large degree of spatial variability in As concentrations in the groundwaters. This means that it may be difficult, or impossible, to predict reliably the likely concentration of As in a particular well from the results of neighbouring wells and means that there is little alternative but to analyse each well. Arsenic-affected aquifers are restricted to certain environments and appear to be the exception rather than the rule. In most aquifers, the majority of wells are likely to be unaffected, even when, for example, they contain high concentrations of dissolved Fe C 2002 Published by elsevier Science Ltd. All rights reserved nding author. Fax: +44-1491-692345 ddress. pls(@ bgs ac uk(P L. Smedley) 0883-2927/02/. see front matter C 2002 Published by Elsevier Science Ltd. All rights reserved. PII:S0883-2927(02)00018-5

Review A review of the source, behaviour and distribution of arsenic in natural waters P.L. Smedley*, D.G. Kinniburgh British Geological Survey, Wallingford, Oxon OX10 8BB, UK Received 1March 2001; accepted 26 October 2001 Editorial handling by R. Fuge Abstract The range of As concentrations found in natural waters is large, ranging from less than 0.5 mg l
1 to more than 5000 mg l
1 . Typical concentrations in freshwater are less than 10 mg l
1 and frequently less than 1 mg l
1 . Rarely, much higher concentrations are found, particularly in groundwater. In such areas, more than 10% of wells may be ‘affected’ (defined as those exceeding 50 mg l
1 ) and in the worst cases, this figure may exceed 90%. Well-known high-As groundwater areas have been found in Argentina, Chile, Mexico, China and Hungary, and more recently in West Bengal (India), Bangladesh and Vietnam. The scale of the problem in terms of population exposed to high As con￾centrations is greatest in the Bengal Basin with more than 40 million people drinking water containing ‘excessive’ As. These large-scale ‘natural’ As groundwater problem areas tend to be found in two types of environment: firstly, inland or closed basins in arid or semi-arid areas, and secondly, strongly reducing aquifers often derived from alluvium. Both environments tend to contain geologically young sediments and to be in flat, low-lying areas where groundwater flow is sluggish. Historically, these are poorly flushed aquifers and any As released from the sediments following burial has been able to accumulate in the groundwater. Arsenic-rich groundwaters are also found in geothermal areas and, on a more localised scale, in areas of mining activity and where oxidation of sulphide minerals has occurred. The As content of the aquifer materials in major problem aquifers does not appear to be exceptionally high, being normally in the range 1–20 mg kg
1 . There appear to be two distinct ‘triggers’ that can lead to the release of As on a large scale. The first is the development of high pH (>8.5) conditions in semi-arid or arid environments usually as a result of the combined effects of mineral weathering and high evaporation rates. This pH change leads either to the desorption of adsorbed As (especially As(V) species) and a range of other anion-forming elements (V, B, F, Mo, Se and U) from mineral oxides, especially Fe oxides, or it prevents them from being adsorbed. The second trigger is the development of strongly reducing conditions at near-neutral pH values, leading to the desorption of As from mineral oxides and to the reductive dissolution of Fe and Mn oxides, also leading to As release. Iron (II) and As(III) are relatively abundant in these groundwaters and SO4 concentrations are small (typically 1mg l
1 or less). Large concentrations of phosphate, bicarbonate, silicate and possibly organic matter can enhance the desorption of As because of competition for adsorption sites. A characteristic feature of high groundwater As areas is the large degree of spatial variability in As concentrations in the groundwaters. This means that it may be difficult, or impossible, to predict reliably the likely concentration of As in a particular well from the results of neighbouring wells and means that there is little alternative but to analyse each well. Arsenic-affected aquifers are restricted to certain environments and appear to be the exception rather than the rule. In most aquifers, the majority of wells are likely to be unaffected, even when, for example, they contain high concentrations of dissolved Fe. # 2002 Published by Elsevier Science Ltd. All rights reserved. 0883-2927/02/$ - see front matter # 2002 Published by Elsevier Science Ltd. All rights reserved. PII: S0883-2927(02)00018-5 Applied Geochemistry 17 (2002) 517–568 www.elsevier.com/locate/apgeochem * Corresponding author. Fax: +44-1491-692345. E-mail address: pls@bgs.ac.uk (P.L. Smedley)

518 P L. Smedley, D.G. Kinniburgh/ Applied Geochemistry 17(2002 )517-568 Contents 1. Introduction 2. Arsenic in natural waters 520 2.2. Abundance and distribution 2.2.1. Atmospheric precipitation 2.2.2. River water 2.2.3. Lake water 2. 24. Seawater and estuaries 2.2.5. Groundwater 2.2.6. Mine drainage 2.2.7. Sediment porewaters. 2.2.8. Oilfield and other brines 2.3. Distribution of arsenic species in water bodies 2.4. Impact of redox kinetics on arsenic speciation 3.1 Minerals 3.1.1. Major arsenic minerals 3. 2. Rock-forming mineral 3.2. Rocks sediments and soils 3.2.1. Igneous rocks. 3.2.2. Metamorphic rocks 530 3.2.3. Sedimentary rocks 530 3.2.4. Unconsolidated sediments 3.2.5. Soils 3.2.6. Contaminated surficial deposits 3.3. The atmosphere 4. Mineral-water interactions 41. Controls on arsenic mobilisation 4.2 Arsenic associations in sediments 4.3. Reduced sediments and the role of iron oxides 4.4. Arsenic release from soils and sediments following reduction.... 4.5. Speciation of elements in sediments and the role of selective extraction techniques 4.6. Transport of arsenic 539 5. Groundwater environments with high arsenic concentrations 5.1. World distribution of groundwater arsenic problems 5.2. Reducing environments 5.2.1. Bangladesh and India (West Bengal 5.2.2. Taiwa 5.2.3. Northern china 233177 5.2.4.Ⅴ letan 5.2.5. Hungary and Romania.…… 5.3. Arid oxidising environments. 5.3. 2. Chile 5.3.3. 549 5.4. Mixed oxidising and reducing environments 5.4. 1. South-western USA 5.5. Geothermal sources 5.6. Sulphide mineralisation and mining-related arsenic problems 5.6.1. Thailand

Contents 1. Introduction ........................................................................................................................................................... 519 2. Arsenic in natural waters ....................................................................................................................................... 520 2.1. Aqueous speciation........................................................................................................................................ 520 2.2. Abundance and distribution.......................................................................................................................... 520 2.2.1. Atmospheric precipitation ................................................................................................................. 521 2.2.2. River water........................................................................................................................................ 523 2.2.3. Lake water......................................................................................................................................... 524 2.2.4. Seawater and estuaries....................................................................................................................... 525 2.2.5. Groundwater ..................................................................................................................................... 525 2.2.6. Mine drainage.................................................................................................................................... 525 2.2.7. Sediment porewaters.......................................................................................................................... 525 2.2.8. Oilfield and other brines.................................................................................................................... 526 2.3. Distribution of arsenic species in water bodies.............................................................................................. 526 2.4. Impact of redox kinetics on arsenic speciation.............................................................................................. 527 3. Sources of arsenic................................................................................................................................................... 528 3.1. Minerals......................................................................................................................................................... 528 3.1.1. Major arsenic minerals ...................................................................................................................... 528 3.1.2. Rock-forming minerals...................................................................................................................... 529 3.2. Rocks, sediments and soils ............................................................................................................................ 530 3.2.1. Igneous rocks..................................................................................................................................... 530 3.2.2. Metamorphic rocks ........................................................................................................................... 530 3.2.3. Sedimentary rocks ............................................................................................................................. 530 3.2.4. Unconsolidated sediments ................................................................................................................. 532 3.2.5. Soils ................................................................................................................................................... 533 3.2.6. Contaminated surficial deposits......................................................................................................... 533 3.3. The atmosphere ............................................................................................................................................. 533 4. Mineral-water interactions ..................................................................................................................................... 533 4.1. Controls on arsenic mobilisation................................................................................................................... 533 4.2. Arsenic associations in sediments.................................................................................................................. 534 4.3. Reduced sediments and the role of iron oxides ............................................................................................. 534 4.4. Arsenic release from soils and sediments following reduction....................................................................... 537 4.5. Speciation of elements in sediments and the role of selective extraction techniques ..................................... 538 4.6. Transport of arsenic ...................................................................................................................................... 539 5. Groundwater environments with high arsenic concentrations ............................................................................... 542 5.1. World distribution of groundwater arsenic problems ................................................................................... 542 5.2. Reducing environments ................................................................................................................................. 543 5.2.1. Bangladesh and India (West Bengal)................................................................................................. 543 5.2.2. Taiwan............................................................................................................................................... 547 5.2.3. Northern china .................................................................................................................................. 547 5.2.4. Vietnam ............................................................................................................................................. 547 5.2.5. Hungary and Romania...................................................................................................................... 548 5.3. Arid oxidising environments.......................................................................................................................... 548 5.3.1. Mexico............................................................................................................................................... 548 5.3.2. Chile .................................................................................................................................................. 548 5.3.3. Argentina........................................................................................................................................... 549 5.4. Mixed oxidising and reducing environments................................................................................................. 549 5.4.1. South-western USA........................................................................................................................... 549 5.5. Geothermal sources ....................................................................................................................................... 550 5.6. Sulphide mineralisation and mining-related arsenic problems ...................................................................... 551 5.6.1. Thailand ............................................................................................................................................ 551 518 P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568

P L. Smedley, D.G. Kinniburgh/ Applied Geochemistry 17(2002 )517-568 5. 6.2. Ghana 5.6.3. United States 5.6. 4. Other areas 6. Common features of groundwater arsenic problem areas 6. 1. A hydrogeochemical perspective 6.2. The source of arsenic 6.3. Arsenic mobilisation--the necessary geochemical trigger .3.1. Desorption at high ph under oxidising conditions 6.3.2. Arsenic desorption and dissolution due to a change to reducing conditions 6.3.3. Reduction in surface area of oxide minerals 6.3.4. Reduction in binding strength between arsenic and mineral surfaces 6.3.5. Mineral dissolution 6.4. Transport--historical groundwater flows 6.5. Future developments in arsenic research 6.6. Identification of at riskaquifers 7. Concluding remarks 559 Acknowledgements… References 1. ntroduction a variety of sources depending on local availability: sur- face water(rivers, lakes, reservoirs and ponds), ground- The recent finding that groundwaters from large areas water(aquifers) and rain water. These sources are very of West Bengal, Bangladesh and elsewhere are heavily variable in terms of As risk. Alongside obvious point enriched with As has prompted a reassessment of the sources of As contamination, high concentrations are factors controlling the distribution of As in the natural mainly found in groundwaters. These are where the environment and the ways in which As may be mobilised. greatest number of, as yet unidentified, sources are likely Arsenic is a ubiquitous element found in the atmosphere to be found this review therefore focuses on the factors ils and rocks, natural waters and organisms. It is mobi- controlling As concentrations in groundwaters. Hoy ed through a combination of natural processes such ever. the authors also review the occurrence of s weathering reactions, biological activity and volcanic broad range of natural waters since these may indirectly f anthropogenic be involved in the formation of As-rich groundwaters and activities. Most environmental As problems are the result can also provide a useful background against which to of mobilisation under natural conditions. However, view groundwater As concentrations. Furthermore, many man has had an important additional impact through of the processes involved in the uptake and release of A mining activity, combustion of fossil fuels, the use of are common to a wide range of natural environments rsenical pesticides, herbicides and crop desiccants and Following the accumulation of evidence for the the use of As as an additive to livestock feed particularly chronic toxicological effects of As in drinking water for poultry. Although the use of arsenical products such recommended and regulatory limits of many authorities as pesticides and herbicides has decreased significantly in are being reduced. The Who guideline value for As in the last few decades, their use for wood preservation drinking water was provisionally reduced in 1993 from still common. The impact on the environment of the use 50 to 10 ug l-l. The new recommended value was based of arsenical compounds, at least locally, will remain for on the increasing awareness of the toxicity of As, parti some cularly its carcinogenicity, and on the ability to measure Of the various sources of as in the envir nt, it quantitatively(WHO, 1993). If the standard basis for drinking water probably poses the greatest threat to risk assessment applied to industrial chemicals were human health. Airborne As, particularly through occu- applied to As, the maximum permissible concentration pational exposure, has also given rise to known health would be lower still. The ec maximum admissible con problems in some areas. Drinking water is derived from centration (MAC)for As in drinking water has been

1. Introduction The recent finding that groundwaters from large areas of West Bengal, Bangladesh and elsewhere are heavily enriched with As has prompted a reassessment of the factors controlling the distribution of As in the natural environment and the ways in which As may be mobilised. Arsenic is a ubiquitous element found in the atmosphere, soils and rocks, natural waters and organisms. It is mobi￾lised through a combination of natural processes such as weathering reactions, biological activity and volcanic emissions as well as through a range of anthropogenic activities. Most environmental As problems are the result of mobilisation under natural conditions. However, man has had an important additional impact through mining activity, combustion of fossil fuels, the use of arsenical pesticides, herbicides and crop desiccants and the use of As as an additive to livestock feed, particularly for poultry. Although the use of arsenical products such as pesticides and herbicides has decreased significantly in the last few decades, their use for wood preservation is still common. The impact on the environment of the use of arsenical compounds, at least locally, will remain for some years. Of the various sources of As in the environment, drinking water probably poses the greatest threat to human health. Airborne As, particularly through occu￾pational exposure, has also given rise to known health problems in some areas. Drinking water is derived from a variety of sources depending on local availability: sur￾face water (rivers, lakes, reservoirs and ponds), ground￾water (aquifers) and rain water. These sources are very variable in terms of As risk. Alongside obvious point sources of As contamination, high concentrations are mainly found in groundwaters. These are where the greatest number of, as yet unidentified, sources are likely to be found. This review therefore focuses on the factors controlling As concentrations in groundwaters. How￾ever, the authors also review the occurrence of As in a broad range of natural waters since these may indirectly be involved in the formation of As-rich groundwaters and can also provide a useful background against which to view groundwater As concentrations. Furthermore, many of the processes involved in the uptake and release of As are common to a wide range of natural environments. Following the accumulation of evidence for the chronic toxicological effects of As in drinking water, recommended and regulatory limits of many authorities are being reduced. The WHO guideline value for As in drinking water was provisionally reduced in 1993 from 50 to 10 mg l
1 . The new recommended value was based on the increasing awareness of the toxicity of As, parti￾cularly its carcinogenicity, and on the ability to measure it quantitatively (WHO, 1993). If the standard basis for risk assessment applied to industrial chemicals were applied to As, the maximum permissible concentration would be lower still. The EC maximum admissible con￾centration (MAC) for As in drinking water has been 5.6.2. Ghana................................................................................................................................................ 551 5.6.3. United States ..................................................................................................................................... 551 5.6.4. Other areas ........................................................................................................................................ 551 6. Common features of groundwater arsenic problem areas...................................................................................... 552 6.1. A hydrogeochemical perspective ................................................................................................................... 552 6.2. The source of arsenic..................................................................................................................................... 552 6.3. Arsenic mobilisation—the necessary geochemical trigger ............................................................................. 552 6.3.1. Desorption at high pH under oxidising conditions ........................................................................... 553 6.3.2. Arsenic desorption and dissolution due to a change to reducing conditions .................................... 554 6.3.3. Reduction in surface area of oxide minerals ..................................................................................... 555 6.3.4. Reduction in binding strength between arsenic and mineral surfaces ............................................... 555 6.3.5. Mineral dissolution............................................................................................................................ 556 6.4. Transport—historical groundwater flows...................................................................................................... 556 6.5. Future developments in arsenic research....................................................................................................... 558 6.6. Identification of ‘at risk’ aquifers .................................................................................................................. 558 7. Concluding remarks ............................................................................................................................................... 559 Acknowledgements...................................................................................................................................................... 560 References ................................................................................................................................................................... 560 P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568 519

P L. Smedley, D G. Kinniburgh/Applied Geochemistry 17(2002 )517-568 reduced to 10 ug I. The Japanese limit for drinking clay or organic matter. In contrast, most oxyanions water is also 10 ug I-I while the interim maximum including arsenate tend to become less strongly sorbed acceptable concentration for Canadian drinking water is as the ph increases (dzombak and Morel, 1990). Under 25 ug l-l. The US-EPA limit was also reduced from 50 some conditions at least, these anions can persist in to 10 ug I-I in January 2001 following prolonged debate solution at relatively high concentrations (tens of ug 1-) over the most appropriate limit. However, this rule is even at near-neutral ph values. Therefore the oxyanion now(September 2001) being reconsidered given the high forming elements such as Cr, As, U and Se are some of cost implications to the US water industry, estimated at the most common trace contaminants in groundwaters S200 million per year. Whilst many national authorities Relative to the other oxyanion-forming elements, As is are seeking to reduce their limits in line with the WHo among the most problematic in the environment because guideline value, many countries and indeed all affected of its relative mobility over a wide range of redox condi- developing countries, still operate at present to the 50 ug tions. Selenium is mobile as the selenate(seo4 )oxyanion I-I standard, in part because of lack of adequate testing under oxidising conditions but is immobilized under facilities for lower concentrations reducing conditions either due to the stronger adsorp- Until recently, As was often not on the list of con- tion of its reduced form, selenite (Seo3 ), or due to its ituents in drinking water routinely analysed by reduction to the metal. Chromium can similarly be national laboratories, water utilities and non-govern mobilized as stable Cr(vi)oxyanion species under oxi- menta organizations(NGOs)and so the body of info dising conditions, but forms cationic Cr(lll) species in mation about the distribution of As in drinking water is reducing environments and hence behaves like other trac not as well known as for many other drinking-water cations (i.e. is relatively immobile at near-neutral pH constituents. In recent years, it has become apparent values). Other oxyanions such as molybdate, vanadate, water sources, and often unexpectedly so. Indeed, Asa g- de yI and rhenate also appear to be less mobile under that both the WHo guideline value and current nationalura standards are quite frequently exceeded in drink cing conditions. In S-rich, reducing environment many of the trace metals also form insoluble sulphides F are now recognised as the most serious inorganic con- Arsenic is distinctive in being relatively mobile under taminants in drinking water on a worldwide basis. In reduced conditions. It can be found at concentrations in areas of high As concentrations, drinking water provides the mg I-I range when all other oxyanion-forming a potentially major source of As in the diet and so its elements are present in the ug l-range arly detection is of considerable importance. Redox potential (Eh) and ph are the most important factors controlling As speciation. Under oxidising con- 2. Arsenic in natural waters H 6.9). whilst at higher ph, Hasha becomes domi- nant(H3AsO? and AsOa- may be present in extremely 2.1. Aqueous speciation acidic and alkaline conditions respectively). Under reducing conditions at pH less than about pH 9.2, the Arsenic is perhaps unique among the heavy metal- uncharged arsenite species H3 AsOg will predominate loids and oxyanion-forming elements(e. g. As, Se, Sb, Mo,(Fig. 1; Brookins, 1988; Yan et al., 2000). The distribu- V, Cr, U, Re) in its sensitivity to mobilisation at the ph tions of the species as a function of ph are given in values typically found in groundwaters(pH 6.5-8.5)and Fig. 2. In practice, most studies in the literature report under both oxidising and reducing conditions. Arsenic can speciation data without consideration of the degree of ccur in the environment in several oxidation states(3, protonation. In the presence of extremely high con centrations of reduced s, dissolved As-sulphide specie inorganic form as oxyanions of trivalent arsenite can be significant. Reducing, acidic conditions favour LAs(IDi or pentavalent arsenate [As(v]. Organic As precipitation of orpiment (As S3), realgar(AsS)or other ulphide minerals containing coprecipitated As(Cullen surface waters, but are rarely quantitatively important. and Reimer, 1989). Therefore high-As waters are not Organic forms may however occur where waters are expected where there is a high concentration of free significantly impacted by industrial pollution. sulphide(moore et al., 1988) Most toxic trace metals occur in solution as cations (e.g. Pbt, Cu+, Ni2+, Cd+, Co2+, Zn+)which gen- 2. 2. Abundance and distribution erally become increasingly insoluble as the pH increases. At the near-neutral ph typical of most groundwaters, the Concentrations of As in fresh water vary by more solubility of most trace-metal cations is severely limited than four orders of magnitude (table 1)depending on the by precipitation as, or coprecipitation with, an oxide, source of As, the amount available and the local geo- hydroxide, carbonate or phosphate mineral, or more chemical environment. Under natural conditions, the likely by their strong adsorption to hydrous metal oxides, greatest range and the highest concentrations of As are

reduced to 10 mg l
1 . The Japanese limit for drinking water is also 10 mg l
1 while the interim maximum acceptable concentration for Canadian drinking water is 25 mg l
1 . The US-EPA limit was also reduced from 50 to 10 mg l
1 in January 2001following prolonged debate over the most appropriate limit. However, this rule is now (September 2001) being reconsidered given the high cost implications to the US water industry, estimated at $200 million per year. Whilst many national authorities are seeking to reduce their limits in line with the WHO guideline value, many countries and indeed all affected developing countries, still operate at present to the 50 mg l
1 standard, in part because of lack of adequate testing facilities for lower concentrations. Until recently, As was often not on the list of con￾stituents in drinking water routinely analysed by national laboratories, water utilities and non-govern￾mental organizations (NGOs) and so the body of infor￾mation about the distribution of As in drinking water is not as well known as for many other drinking-water constituents. In recent years, it has become apparent that both the WHO guideline value and current national standards are quite frequently exceeded in drinking￾water sources, and often unexpectedly so. Indeed, As and F are now recognised as the most serious inorganic con￾taminants in drinking water on a worldwide basis. In areas of high As concentrations, drinking water provides a potentially major source of As in the diet and so its early detection is of considerable importance. 2. Arsenic in natural waters 2.1. Aqueous speciation Arsenic is perhaps unique among the heavy metal￾loids and oxyanion-forming elements (e.g. As, Se, Sb, Mo, V, Cr, U, Re) in its sensitivity to mobilisation at the pH values typically found in groundwaters (pH 6.5–8.5) and under both oxidising and reducing conditions. Arsenic can occur in the environment in several oxidation states (
3, 0, +3 and +5) but in natural waters is mostly found in inorganic form as oxyanions of trivalent arsenite [As(III)] or pentavalent arsenate [As(V)]. Organic As forms may be produced by biological activity, mostly in surface waters, but are rarely quantitatively important. Organic forms may however occur where waters are significantly impacted by industrial pollution. Most toxic trace metals occur in solution as cations (e.g. Pb2+, Cu2+, Ni2+, Cd2+, Co2+, Zn2+) which gen￾erally become increasingly insoluble as the pH increases. At the near-neutral pH typical of most groundwaters, the solubility of most trace-metal cations is severely limited by precipitation as, or coprecipitation with, an oxide, hydroxide, carbonate or phosphate mineral, or more likely by their strong adsorption to hydrous metal oxides, clay or organic matter. In contrast, most oxyanions including arsenate tend to become less strongly sorbed as the pH increases (Dzombak and Morel, 1990). Under some conditions at least, these anions can persist in solution at relatively high concentrations (tens of mg l
1 ) even at near-neutral pH values. Therefore the oxyanion￾forming elements such as Cr, As, U and Se are some of the most common trace contaminants in groundwaters. Relative to the other oxyanion-forming elements, As is among the most problematic in the environment because of its relative mobility over a wide range of redox condi￾tions. Selenium is mobile as the selenate (SeO4 2
) oxyanion under oxidising conditions but is immobilized under reducing conditions either due to the stronger adsorp￾tion of its reduced form, selenite (SeO3 2
), or due to its reduction to the metal. Chromium can similarly be mobilized as stable Cr(VI) oxyanion species under oxi￾dising conditions, but forms cationic Cr(III) species in reducing environments and hence behaves like other trace cations (i.e. is relatively immobile at near-neutral pH values). Other oxyanions such as molybdate, vanadate, uranyl and rhenate also appear to be less mobile under reducing conditions. In S-rich, reducing environments, many of the trace metals also form insoluble sulphides. Arsenic is distinctive in being relatively mobile under reduced conditions. It can be found at concentrations in the mg l
1 range when all other oxyanion-forming elements are present in the mg l
1 range. Redox potential (Eh) and pH are the most important factors controlling As speciation. Under oxidising con￾ditions, H2AsO4
is dominant at low pH (less than about pH 6.9), whilst at higher pH, HAsO4 2
becomes domi￾nant (H3AsO4 0 and AsO4 3
may be present in extremely acidic and alkaline conditions respectively). Under reducing conditions at pH less than about pH 9.2, the uncharged arsenite species H3AsO3 0 will predominate (Fig. 1; Brookins, 1988; Yan et al., 2000). The distribu￾tions of the species as a function of pH are given in Fig. 2. In practice, most studies in the literature report speciation data without consideration of the degree of protonation. In the presence of extremely high con￾centrations of reduced S, dissolved As-sulphide species can be significant. Reducing, acidic conditions favour precipitation of orpiment (As2S3), realgar (AsS) or other sulphide minerals containing coprecipitated As (Cullen and Reimer, 1989). Therefore high-As waters are not expected where there is a high concentration of free sulphide (Moore et al., 1988). 2.2. Abundance and distribution Concentrations of As in fresh water vary by more than four orders of magnitude (Table 1) depending on the source of As, the amount available and the local geo￾chemical environment. Under natural conditions, the greatest range and the highest concentrations of As are 520 P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568

P L. Smedley, D G. Kinniburgh/Applied Geochemistry 17(2002 )517-568 1200 20 (a) Arseni HA pH 02468101214 As-Or-H,O at 25C and I bar total pressure. found in groundwaters as a result of the strong influence HAsO 2- of water-rock interactions and the greater tendency in aquifers for the physical and geochemical conditions be favourable for As mobilization and accumulation (b)Arsenate The range of concentrations for many water bodies is arge and hence ' typical values are difficult to derive 3- Many studies of As reported in the literature have also preferentially targeted known problem areas and hence aSo reported ranges are often extreme and unrepresentative of natural waters as a whole. Nonetheless, the following compilation of data for ranges of As concentrations found in various parts of the hydrosphere and lithosphere gives a 34567891011 broad indication of the expected concentration ranges and their variation in the environment Fig. 2.(a) Arsenite and(b)arsenate speciation as a function of H (ionic strength of about 0.01 M). Redox conditions have 2.2.1. Atmospheric precipitation been chosen such that the indicated oxidation state dominates Arsenic enters the atmosphere through inputs from the speciation in both cases wind erosion, volcanic emissions, low-temperature volatilisation from soils, marine aerosols and pollutie d is returned to the earths surface by wet and dry smelter operations, coal burning and volcanic emissions deposition. The most important anthropogenic inputs are generally higher. Andreae(1980) found rainfall are from smelter operations and fossil-fuel combustion. potentially affected by smelting and coal burning to The As appears to consist of mainly As(In2O3 dust have As concentrations of around 0.5 ug I(Table 1), particles(Cullen and Reimer, 1989). Nriagu and Pacyna although higher concentrations(average 16 ug l-)have (1988) estimated that anthropogenic sources of atmo- been found in rainfall collected in Seattle some 35 km spheric arsenic(around 18, 800 tonnes a-)amounted to downwind of a Cu smelter(Crecelius, 1975). values around 70% of the global atmospheric As flux. While it given for Arizona snowpacks(Table 1; Barbaris and is accepted that these anthropogenic sources have an Betterton, 1996) are also probably slightly above base important impact on airborne As compositions, their line concentrations because of potential inputs of air infuence on the overall As cycle is not well established. borne As from smelters, power plants and soil dust. In Baseline concentrations of As in rainfall and snow in general however, sources of airborne As in most indus al areas are invariably low at typically less than 0.03 trialized nations are limited as a result of air-pollution ug I-I Table 1). Concentrations in areas affected by control measures. Unless significantly contaminated

found in groundwaters as a result of the strong influence of water-rock interactions and the greater tendency in aquifers for the physical and geochemical conditions to be favourable for As mobilization and accumulation. The range of concentrations for many water bodies is large and hence ‘typical’ values are difficult to derive. Many studies of As reported in the literature have also preferentially targeted known problem areas and hence reported ranges are often extreme and unrepresentative of natural waters as a whole. Nonetheless, the following compilation of data for ranges of As concentrations found in various parts of the hydrosphere and lithosphere gives a broad indication of the expected concentration ranges and their variation in the environment. 2.2.1. Atmospheric precipitation Arsenic enters the atmosphere through inputs from wind erosion, volcanic emissions, low-temperature volatilisation from soils, marine aerosols and pollution and is returned to the earth’s surface by wet and dry deposition. The most important anthropogenic inputs are from smelter operations and fossil-fuel combustion. The As appears to consist of mainly As(III)2O3 dust particles (Cullen and Reimer, 1989). Nriagu and Pacyna (1988) estimated that anthropogenic sources of atmo￾spheric arsenic (around 18,800 tonnes a
1 ) amounted to around 70% of the global atmospheric As flux. While it is accepted that these anthropogenic sources have an important impact on airborne As compositions, their influence on the overall As cycle is not well established. Baseline concentrations of As in rainfall and snow in rural areas are invariably low at typically less than 0.03 mg l
1 (Table 1). Concentrations in areas affected by smelter operations, coal burning and volcanic emissions are generally higher. Andreae (1980) found rainfall potentially affected by smelting and coal burning to have As concentrations of around 0.5 mg l
1 (Table 1), although higher concentrations (average 16 mg l
1 ) have been found in rainfall collected in Seattle some 35 km downwind of a Cu smelter (Crecelius, 1975). Values given for Arizona snowpacks (Table 1; Barbaris and Betterton, 1996) are also probably slightly above base￾line concentrations because of potential inputs of air￾borne As from smelters, power plants and soil dust. In general however, sources of airborne As in most indus￾trialized nations are limited as a result of air-pollution control measures. Unless significantly contaminated Fig. 1. Eh-pH diagram for aqueous As species in the system As–O2–H2O at 25 C and 1bar total pressure. Fig. 2. (a) Arsenite and (b) arsenate speciation as a function of pH (ionic strength of about 0.01M). Redox conditions have been chosen such that the indicated oxidation state dominates the speciation in both cases. P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568 521

P L. Smedley, D.G. Kinniburgh/ Applied Geochemistry 17(2002 )517-568 Table I Typical As concentrations in natural waters Water body and location As concentration average Reference Terrestrial (w USA) 0.013-0.032 Coastal (Mid-Atlantic, USA) 0.1(<0.005-1.1) Scudlark and Church(1988) 0.14(0.02-0.42) restrial rain Andreae et al. (1983): Froelich et al. (1985) Seyler and Martin(1991) 25(<0.02-1.1) Lenvik et al. ( 1978) outh-east USA 5-0.45 Waslenchuk(1979 (1988) Po river tal Pettine et al. (1992) Polluted European river River Danube. Bavaria 0.75-3.8(upto30) Andreae and Andreae(1989) High-As groundwater influenced Northern chile 90-21800 Northern chile 400-450 Sancha etal.(1992) Cordoba, Argentina Lerda and Prosperi (1996) Geothermal influenced Sierra Nevada USA 0.20-264 enson and Spencer(1983) Waikato. New Zealand 32(28-36) McLaren and Kim(1995) Madison and missouri Rivers. USA 44(1967 Robinson et al. (1995) 10-370 Nimick et al. (1998 Mining influenced Ron phibun. thailand 218(48-583) Williams et al. (1996) Ashanti. Ghana 284(<2-7900) Smedley et al. (1996) British Columbia. Canada 17.5(<0.2-556 Azcue et al. (1994) Baselin British Columbia 0.28(<0.2-0.42) Azcue et al.(1994. 1995 Ontario 0.73-9.2 high Fe) apan 0.38-19 Sweden 0.06-1.2 Reuther(1992) Geothermal infuenced Western usa 0.38-1000 Benson and Spencer(1983 orthwest Territories. Canada 270(64-530) Bright (1996) Ontario Canada Baseline Oslofjord, Norway 0.7-2.0 Abdullah et al. (1995) Saanich Inlet. British Columbia 12-2.5 Peterson and Carpenter (1983)

Table 1 Typical As concentrations in natural waters Water body and location As concentration average or range (mg l
1 ) Reference Rain water Baseline Maritime 0.02 Andreae (1980) Terrestrial (w USA) 0.013–0.032 Andreae (1980) Coastal (Mid-Atlantic, USA) 0.1(<0.005–1.1) Scudlark and Church (1988) Snow (Arizona) 0.14 (0.02–0.42) Barbaris and Betterton (1996) Non-baseline: Terrestrial rain 0.46 Andreae (1980) Seattle rain, impacted by copper smelter 16 Crecelius (1975) River water Baseline Various 0.83 (0.13–2.1) Andreae et al. (1983); Froelich et al. (1985); Seyler and Martin (1991) Norway 0.25 (<0.02–1.1) Lenvik et al. (1978) South-east USA 0.15–0.45 Waslenchuk (1979) USA 2.1Sonderegger and Ohguchi (1988) Dordogne, France 0.7 Seyler and Martin (1990) Po River, Italy 1.3 Pettine et al. (1992) Polluted European rivers 4.5–45 Seyler and Martin (1990) River Danube, Bavaria 3 (1–8) Quentin and Winkler (1974) Schelde catchment, Belgium 0.75–3.8 (up to 30) Andreae and Andreae (1989) High-As groundwater influenced: Northern Chile Northern Chile 190–21800 400–450 Ca´ceres et al. (1992) Sancha (1999) Co´rdoba, Argentina 7–114 Lerda and Prosperi (1996) Geothermal influenced Sierra Nevada, USA Waikato, New Zealand Madison and Missouri Rivers, USA 0.20–264 32 (28–36) 44 (19–67) 10–370 Benson and Spencer (1983) McLaren and Kim (1995) Robinson et al. (1995) Nimick et al. (1998) Mining influenced Ron Phibun, Thailand Ashanti, Ghana 218 (4.8–583) 284 (<2–7900) Williams et al. (1996) Smedley et al. (1996) British Columbia, Canada 17.5 (<0.2–556) Azcue et al. (1994) Lake water Baseline British Columbia 0.28 (<0.2–0.42) Azcue et al. (1994, 1995) Ontario 0.7 Azcue and Nriagu (1995) France 0.73–9.2 (high Fe) Seyler and Martin (1989) Japan 0.38–1.9 Baur and Onishi (1969) Sweden 0.06–1.2 Reuther (1992) Geothermal influenced Western USA 0.38–1000 Benson and Spencer (1983) Mining influenced Northwest Territories, Canada Ontario, Canada 270 (64–530) 35–100 Bright et al. (1996) Azcue and Nriagu (1995) Estuarine water Baseline Oslofjord, Norway 0.7–2.0 Abdullah et al. (1995) Saanich Inlet, British Columbia 1.2–2.5 Peterson and Carpenter (1983) (continued on next page) 522 P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568

P.L. Smedley, D.G. Kinniburgh/ Applied Geochemistry 17(2002)517-568 Table I Water body and location n average Rhone Estuary, France 2.2(1.1-3.8) Seyler and Martin(1990) Krka Estuary. Yugoslavia 0.13-1.8 Seyler and Martin(1991) to 16 er and Martin(1990) Howard et al.(1988) stuary, Belgium 1.8-4.9 Andreae and Andreae(1989) Deep Pacific and Atlantic 10-1.8 allen and Reimer (1989) Coastal Malaysia 1.0(0.7-1.8) Y usof et al. (1994) Coastal Spain 1.5(0.5-3.7) Navarro et al. (1993) 1.3(1.1-1.6) Maher (1985) Baseline uK Edmunds et al. (1989) Assinc proentne gg: icengalthe 10-5000 Das et al. (1995): BGS and DPHE(2001) Nicolli et al.(1989): Smedley et al. (2001a): Del Razo et al. (1990); Luo et al. (1997) su et al. (1997); Va Williams et al. (1996) Geothermal water aur and Onishi(1969): White et al, ( 1963). Ellis and Mahon(1977) Arsenical herbicide plant, Texas er(1997ab) Various USA <1-34,000 Plumlee et al. (1999) Iron mountain upto850.000 Nordstrom and Alpers(1999) Ural mountains Gelova(1977) Baseline, Swedish Estuary Widerlund and Ingri(1995) Baseline, clays, Saskatchewan, 3.2-99 Yan et al. (2000) Canada Baseline. Amazon shelf sediments up to 300 Sullivan and Aller (1996 Mining.contam'd. British Columbia 50-360 Tailings impoundment, Ontario. McCreadie et al. (2000) Canada Oilfield and related brine Ellis Pool. Alberta. 230 White et al. (1963) Searles lake brine. california upto243.000 White et al.( 1963) with industrial sources of As, atmospheric precipitation also found low average concentrations of about 0.25 Hg contributes little As to surface and groundwater bodies I-I in rivers draining basement rocks in Norway,the lowest being in catchments on Precambrian rocks 2. 2.2. River water Waslenchuk(1979) found concentrations in river waters Baseline concentrations of As in river waters are also from the south-eastern USA in the range 0.15-0.45 ug low(in the region of 0.1-0.8 ug I-I but can range up to I-I( Table 1) ca2 Hg 1-l; Table 1). They vary according to the com Relatively high concentrations of naturally-occurring position of the surface recharge, the contribution from As can occur in some areas as a result of inputs from baseflow and the bedrock lithology Concentrations at geothermal sources or high-As groundwaters. Arsenic he low end of the range have been found in rivers concentrations in river waters from geothermal areas draining As-poor bedrocks. Seyler and Martin(1991) have been reported typically at around 10-70 ug I-I found average river concentrations as low as 0.13 ug l-I (e. g. western USA and New Zealand: McLaren and in the Krka region of Y ugoslavia where the bedrock is Kim, 1995; Robinson et al, 1995: Nimick et al., 1998; As-poor karstic limestone (Table 1). Lenvik et al. (1978) Table 1), although higher concentrations have been

with industrial sources of As, atmospheric precipitation contributes little As to surface and groundwater bodies. 2.2.2. River water Baseline concentrations of As in river waters are also low (in the region of 0.1–0.8 mg l
1 but can range up to ca. 2 mg l
1 ; Table 1). They vary according to the com￾position of the surface recharge, the contribution from baseflow and the bedrock lithology. Concentrations at the low end of the range have been found in rivers draining As-poor bedrocks. Seyler and Martin (1991) found average river concentrations as low as 0.13 mg l
1 in the Krka region of Yugoslavia where the bedrock is As-poor karstic limestone (Table 1). Lenvik et al. (1978) also found low average concentrations of about 0.25 mg l
1 in rivers draining basement rocks in Norway, the lowest being in catchments on Precambrian rocks. Waslenchuk (1979) found concentrations in river waters from the south-eastern USA in the range 0.15–0.45 mg l
1 (Table 1). Relatively high concentrations of naturally-occurring As can occur in some areas as a result of inputs from geothermal sources or high-As groundwaters. Arsenic concentrations in river waters from geothermal areas have been reported typically at around 10–70 mg l
1 (e.g. western USA and New Zealand; McLaren and Kim, 1995; Robinson et al., 1995; Nimick et al., 1998; Table 1), although higher concentrations have been Table 1(continued) Water body and location As concentration average or range (mg l
1 ) Reference Rhoˆne Estuary, France 2.2 (1.1–3.8) Seyler and Martin (1990) Krka Estuary, Yugoslavia 0.13–1.8 Seyler and Martin (1991) Mining and industry influenced Loire Estuary, France Tamar Estuary, UK Schelde Estuary, Belgium up to 16 2.7–8.8 1.8–4.9 Seyler and Martin (1990) Howard et al. (1988) Andreae and Andreae (1989) Seawater Deep Pacific and Atlantic Coastal Malaysia Coastal Spain Coastal Australia 1.0–1.8 1.0 (0.7–1.8) 1.5 (0.5–3.7) 1.3 (1.1–1.6) Cullen and Reimer (1989) Yusof et al. (1994) Navarro et al. (1993) Maher (1985) Groundwater Baseline UK <0.5–10 Edmunds et al. (1989) As-rich provinces (e.g. Bengal Basin, Argentina, Mexico, northern China, Taiwan, Hungary) 10–5000 Das et al. (1995); BGS and DPHE (2001); Nicolli et al. (1989); Smedley et al. (2001a); Del Razo et al. (1990); Luo et al. (1997); Hsu et al. (1997); Varsa´nyi et al. (1991) Mining-contaminated groundwaters 50–10,000 Wilson and Hawkins (1978);Welch et al. (1988); Williams et al. (1996) Geothermal water <10–50,000 Baur and Onishi (1969); White et al., (1963), Ellis and Mahon (1977) Arsenical herbicide plant, Texas 408,000 Kuhlmeier (1997a,b) Mine drainage Various, USA <1–34,000 Plumlee et al. (1999) Iron Mountain up to 850,000 Nordstrom and Alpers (1999) Ural Mountains 400,000 Gelova (1977) Sediment porewater Baseline, Swedish Estuary 1.3–166 Widerlund and Ingri (1995) Baseline, clays, Saskatchewan, Canada 3.2–99 Yan et al. (2000) Baseline, Amazon shelf sediments up to 300 Sullivan and Aller (1996) Mining-contam’d, British Columbia 50–360 Azcue et al. (1994) Tailings impoundment, Ontario, Canada 300–100,000 McCreadie et al. (2000) Oilfield and related brine Ellis Pool, Alberta, Canada 230 White et al. (1963) Searles Lake brine, California up to 243,000 White et al. (1963) P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568 523

P L. Smedley, D G. Kinniburgh/Applied Geochemistry 17(2002 )517-568 n htrations up to 370 Hg I-l in Madison River water most of the catchment was in the range 0.75-38 HeA B- found. Nimick et al. (1998) for example found As con- sewage. However, the concentration of As in water fro yoming and Montana) as a result of geothermal and not significantly different from baseline concentra inputs from the Yellowstone geothermal system. wilkie tions. Durum et al. (1971)reported As concentrations in nd Hering(1998)also found concentrations in the 727 samples of surface waters from the United States range 85-153 ug I-I in Hot Creek(tributary of the While 79% of the samples had As concentrations below Owens River, California). Some river waters affected by the(rather high) detection limit of 10 ug l-l, the highest geothermal activity show distinct seasonal variations in observed concentration, 1 100 ug I-l, was found in Sugar As concentration Concentrations in the madison river Creek. South Carolina. downstream of an industrial have been noted to be highest during low-flow condi- complex tions. This has been attributed to a greater contribution Arsenic can also be derived from mine wastes and mill of geothermal water during times of low flow and dilu- tailings. Azcue and Nriagu (1995) found baseline con tion from spring runoff at times of high fow(Nimick et centrations in the Moira River, Ontario of 0.7 ug aL., 1998). In the Waikato river system of New Zealand, upstream of the influence of tailings from gold-mine As maxima were found in the summer months. These workings. Downstream, concentrations increased to 23 increases were linked to temperature-controlled micro- Hg l-l. Azcue et al. (1994)found concentrations up to bial reduction of As(v) to As(Ill with consequent 556 ug l-(average 17. 5 ug l-)in streams adjacent to increased mobility of As(lID(McLaren and Kim, 1995) posits in British Columbia. Williams et al. Increased concentrations are also reported in some (1996)and Smedley et al.(1996) noted high As con river waters from arid areas where the surface water centrations(typically around 200-300 ug 1-)in surface dominated by river baseflow. The resulting surface waters affected respectively by Sn- and Au-mining activ- waters often have a high pH and alkalinity. For example, ities. Though often involving notable increases above in surface waters from the loa river basin of northern baseline concentrations such anomalies tend to be rela Chile( antofagasta area, Atacama desert), Caceres et al. tively localised around the pollution source, principally ( 1992) found concentrations of naturally-occurring As because of the strong adsorption affinity of oxide miner ranging between 190 and 21, 800 ug I-. The high As als, especially Fe oxide, for As under oxidising, neutral concentrations correlated well with salinity. While geo- to mildly acidic conditior hermal inputs are likely to have had an importan impact on the chemical compositions of the river waters 2. 2.3. Lake water in this area(Section 5.5), evaporative concentration of Concentrations of As in lake waters are typica baseflow-dominated river water is also likely to be to or lower than those found in river water important in the arid conditions. Increased As con- concentrations have been found at <l ug I- centrations(up to 114 ug l-)have also been reported in (Azcue and Nriagu, 1995; Azcue et al., 1995). As with river waters from central Argentina where regional river waters, increased concentrations are found in lake groundwater-As concentrations(and pH, alkalinity) are waters affected by geothermal water and by mining high(Lerda and Prosperi, 1996). activity. Ranges of typically 100-500 ug I-I have been Although bedrock inevitably has an influence on reported in some mining areas and up to 1000 ug l-in iver-water As concentrations, concentrations in rivers geothermal areas ( Table 1). Arsenic concentrations in ith more typical pH and alkalinity values(c pH 5-7, mining-affected lake waters are not always high how alkalinity <100 mg I-I as HCO3)do not show the extre- ever, as removal from solution can be achieved effec mely high concentrations found in groundwaters because tively by adsorption onto Fe oxides under neutral to of oxidation and adsorption of As species onto the river mildly acidic conditions. Azcue et al.(1994), for exam- sediments as well as dilution by surface recharge and run- ple, found As concentrations in Canadian lake waters off. Arsenic concentrations in seven river water samples affected by mining effluent similar to those not affected from Bangladesh have been reported in the range <0.5-2.7 by mining effluent, in each case about 0.3 ug I ug I- but with one sample having a high concentration of High As concentrations are also found in some alk 29 ug I(BGS and DPHE, 2001). The highest value line closed-basin lakes as a result of extreme evapora observed is significantly above world-average baseline tion and/or geothermal inputs. Mono Lake in the concentrations (Table 1)but is much lower than some of California, USA, for example, has concentrations of he values found in the groundwaters dissolved As of 10,000-20,000 ug I-, with pH values in Significant increases in As concentrations of river the range 9.5-10 as a result of inputs from geothermal waters may also occur as a result of pollution from springs and the weathering of volcanic rocks followed industrial or sewage effluents. Andreae and Andreae by evaporation(Maest et al., 1992) ( 1989)found concentrations up to 30 ug I-in water There is also much evidence for stratification of as from the River Zenne, Belgium which is affected by concentrations in some lake waters as a result of varying inputs from urban and industrial sources, particularl redox conditions(Aggett and o Brien, 1985). Azcue and

found. Nimick et al. (1998) for example found As con￾centrations up to 370 mg l
1 in Madison River water (Wyoming and Montana) as a result of geothermal inputs from the Yellowstone geothermal system. Wilkie and Hering (1998) also found concentrations in the range 85–153 mg l
1 in Hot Creek (tributary of the Owens River, California). Some river waters affected by geothermal activity show distinct seasonal variations in As concentration. Concentrations in the Madison River have been noted to be highest during low-flow condi￾tions. This has been attributed to a greater contribution of geothermal water during times of low flow and dilu￾tion from spring runoff at times of high flow (Nimick et al., 1998). In the Waikato river system of New Zealand, As maxima were found in the summer months. These increases were linked to temperature-controlled micro￾bial reduction of As(V) to As(III) with consequent increased mobility of As(III) (McLaren and Kim, 1995). Increased concentrations are also reported in some river waters from arid areas where the surface water is dominated by river baseflow. The resulting surface waters often have a high pH and alkalinity. For example, in surface waters from the Loa River Basin of northern Chile (Antofagasta area, Atacama desert), Ca´ceres et al. (1992) found concentrations of naturally-occurring As ranging between 190 and 21,800 mg l
1 . The high As concentrations correlated well with salinity. While geo￾thermal inputs are likely to have had an important impact on the chemical compositions of the river waters in this area (Section 5.5), evaporative concentration of baseflow-dominated river water is also likely to be important in the arid conditions. Increased As con￾centrations (up to 114 mg l
1 ) have also been reported in river waters from central Argentina where regional groundwater-As concentrations (and pH, alkalinity) are high (Lerda and Prosperi, 1996). Although bedrock inevitably has an influence on river-water As concentrations, concentrations in rivers with more typical pH and alkalinity values (c. pH 5–7, alkalinity <100 mg l
1 as HCO3) do not show the extre￾mely high concentrations found in groundwaters because of oxidation and adsorption of As species onto the river sediments as well as dilution by surface recharge and run￾off. Arsenic concentrations in seven river water samples from Bangladesh have been reported in the range <0.5–2.7 mg l
1 but with one sample having a high concentration of 29 mg l
1 (BGS and DPHE, 2001). The highest value observed is significantly above world-average baseline concentrations (Table 1) but is much lower than some of the values found in the groundwaters. Significant increases in As concentrations of river waters may also occur as a result of pollution from industrial or sewage effluents. Andreae and Andreae (1989) found concentrations up to 30 mg l
1 in water from the River Zenne, Belgium which is affected by inputs from urban and industrial sources, particularly sewage. However, the concentration of As in water from most of the catchment was in the range 0.75–3.8 mg l
1 and not significantly different from baseline concentra￾tions. Durum et al. (1971) reported As concentrations in 727 samples of surface waters from the United States. While 79% of the samples had As concentrations below the (rather high) detection limit of 10 mg l
1 , the highest observed concentration, 1100 mg l
1 , was found in Sugar Creek, South Carolina, downstream of an industrial complex. Arsenic can also be derived from mine wastes and mill tailings. Azcue and Nriagu (1995) found baseline con￾centrations in the Moira River, Ontario of 0.7 mg l
1 upstream of the influence of tailings from gold-mine workings. Downstream, concentrations increased to 23 mg l
1 . Azcue et al. (1994) found concentrations up to 556 mg l
1 (average 17.5 mg l
1 ) in streams adjacent to tailings deposits in British Columbia. Williams et al. (1996) and Smedley et al. (1996) noted high As con￾centrations (typically around 200–300 mg l
1 ) in surface waters affected respectively by Sn- and Au-mining activ￾ities. Though often involving notable increases above baseline concentrations, such anomalies tend to be rela￾tively localised around the pollution source, principally because of the strong adsorption affinity of oxide miner￾als, especially Fe oxide, for As under oxidising, neutral to mildly acidic conditions. 2.2.3. Lake water Concentrations of As in lake waters are typically close to or lower than those found in river water. Baseline concentrations have been found at <1 mg l
1 in Canada (Azcue and Nriagu, 1995; Azcue et al., 1995). As with river waters, increased concentrations are found in lake waters affected by geothermal water and by mining activity. Ranges of typically 100–500 mg l
1 have been reported in some mining areas and up to 1000 mg l
1 in geothermal areas (Table 1). Arsenic concentrations in mining-affected lake waters are not always high how￾ever, as removal from solution can be achieved effec￾tively by adsorption onto Fe oxides under neutral to mildly acidic conditions. Azcue et al. (1994), for exam￾ple, found As concentrations in Canadian lake waters affected by mining effluent similar to those not affected by mining effluent, in each case about 0.3 mg l
1 . High As concentrations are also found in some alka￾line closed-basin lakes as a result of extreme evapora￾tion and/or geothermal inputs. Mono Lake in the California, USA, for example, has concentrations of dissolved As of 10,000–20,000 mg l
1 , with pH values in the range 9.5–10 as a result of inputs from geothermal springs and the weathering of volcanic rocks followed by evaporation (Maest et al., 1992). There is also much evidence for stratification of As concentrations in some lake waters as a result of varying redox conditions (Aggett and O’Brien, 1985). Azcue and 524 P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568

P L. Smedley, D G. Kinniburgh/Applied Geochemistry 17(2002 )517-568 Nriagu(1995)found that concentrations increased witI depth(up to 10 m) in lake waters from Ontario, prob- quoted in the literature show a very large range from ably because of an increasing ratio of As(lln to As(V <0.5 to 5000 ug l-I(i.e. four orders of magnitude). This with depth and an influx of mining-contaminated sedi- range occurs under natural conditions. High concentra ment porewaters at the sediment-water interface. The tions of As are found in groundwater in a variety of concentrations were higher in summer when the pro- environments. These include both oxidising(under condi- portion of As(lIn) was observed to be higher. Depleted tions of high pH) and reducing aquifers and areas affected O, levels in the bottom lake waters as a result of biolo- by geothermal, mining and industrial activity. Evapora- gical productivity during the summer months are a tive concentration can also increase concentrations sub- ely cause of the higher As stantially. Most high-As groundwater provinces are the lake waters result of natural occurrences of As. Cases of mining- induced As pollution are numerous in the literature but 2. 2. 4. Seawater and estuaries tend to be localised. Cases of industrially-induced As Average As concentrations in open seawater usually pollution (including that from agriculture) may be show little variation and are typically around 1.5 ug l-l severe locally (Table 1) but occurrences are relatively (Table 1). Concentrations in estuarine water are more rare. Arsenic occurrences in groundwater are discussed variable as a result of varying river inputs and salinity or more fully in Section 5 than 4 Hg I- under natural conditions. Peterson and 2.2.6. Mine drainage Carpenter (1983)found concentrations between 1. 2-2.5 Under the extremely acid conditions of some acid ug I- in waters from Saanich Inlet, British Columbia. mine drainage, whicl have negative pH values Values less than 2 ug I-I were found in Oslofjord, Nor- (Nordstrom et al., 2000), high concentrations of a wid way(Abdullah et al., 1995: Table 1) Concentrations are range of solutes are found, including Fe and As. The commonly higher when riverine inputs are affected by highest reported As concentration of 850,000 ug 1-I is industrial or mining effluent(e. g. Tamar, Schelde, Loire from an acid seep in the richmond mine at Iron Estuaries; Table D)or by geothermal water. Unlike some Mountain, California(Nordstrom and Alpers, 1999). In other trace elements such as B, saline intrusion of sea- a compilation of some 180 samples of mine drainage water into an aquifer is unlikely to lead to a significant from the USA, Plumlee et al. (1999)reported concentra increase of As in the afected groundwater tions ranging from detection limits(<l ug I-I or more Arsenate shares many chemical characteristics with to 340,000 ug l-, again the highest values being from phosphate and hence in oxic marine and estuarine waters. the richmond mine. Gelova(1977) also reported an A depletions in phosphate in biologically productive surface concentration of 400,000 ug I- from the Ural Moun- waters are mirrored by depletions in arsenate. Arsenate tains. Dissolved As in acid mine waters is rapidly concentration minima often coincide with photosynthetic removed as the Fe is oxidised and precipitated and the axima evidenced by high concentrations of chlor As scavenged through adsorption. At Iron Mountail ophyll a( Cullen and reimer, 1989). Several studies have an efficient neutralization plant removes the As and noted variations in the behaviour of As during estuarine metals for safe disposal mixing. Some have reported conservative behaviour. In the unpolluted Krka Estuary of Yugoslavia, Seyler and 2. 2.7. Sediment porewaters Martin(1991)observed a linear increase in total As with Some high concentrations of As have been found in increasing salinity ranging from 0.13 Hg I- in fresh porewaters extracted from unconsolidated sediments waters to 1.8 ug I- offshore (i.e. seawater value). How- and often form sharp contrasts to the concentrations ever, other studies have observed non-conservative observed in overlying surface waters (e.g. Belzile and behaviour (departures from simple mixing) in estuaries Tessier, 1990). Widerlund and Ingri (1995) found con due to processes such as diffusion from sediment pore centrations in the range 1.3-166 ug I-I in porewaters waters, coprecipitation with Fe oxides or anthropogenic from the Kalix River estuary of northern Sweden. Yan inputs(e.g Andreae et al., 1983: Andreae and Andreae, et al. (2000) found As concentrations in the range 3. 2-99 1989). The flocculation of Fe oxides at the freshwater- ug I-i in porewaters from clay sediments in Saskatch- saline interface is an important consequence of increases ewan, Canada(Table 1). Increased concentrations have in pH and salinity. This can lead to major decreases in been found in porewaters affected by geothermal inputs. the As fux to the oceans( Cullen and Reimer, 1989) Aggett and Kriegman(1988) found As concentrations u to 6430 ug I- in anoxic porewaters from New Zealand 2.2.5. Groundwater Even higher concentrations can be found in porewaters Background concentrations of As in groundwater are from sediments affected by mining contamination(tail in most countries less than 10 ug l-(e.g. Edmunds et ings, mineral-rich deposits). McCreadie et al.(2000) al., 1989 for the UK: Welch et al, 2000 for the USA) reported As concentrations up to 100,000 ug I-I in

Nriagu (1995) found that concentrations increased with depth (up to 10 m) in lake waters from Ontario, prob￾ably because of an increasing ratio of As(III) to As(V) with depth and an influx of mining-contaminated sedi￾ment porewaters at the sediment-water interface. The concentrations were higher in summer when the pro￾portion of As(III) was observed to be higher. Depleted O2 levels in the bottom lake waters as a result of biolo￾gical productivity during the summer months are a likely cause of the higher As concentrations in the deeper lake waters. 2.2.4. Seawater and estuaries Average As concentrations in open seawater usually show little variation and are typically around 1.5 mg l
1 (Table 1). Concentrations in estuarine water are more variable as a result of varying river inputs and salinity or redox gradients but are also usually low, at typically less than 4 mg l
1 under natural conditions. Peterson and Carpenter (1983) found concentrations between 1.2–2.5 mg l
1 in waters from Saanich Inlet, British Columbia. Values less than 2 mg l
1 were found in Oslofjord, Nor￾way (Abdullah et al., 1995; Table 1). Concentrations are commonly higher when riverine inputs are affected by industrial or mining effluent (e.g. Tamar, Schelde, Loire Estuaries; Table 1) or by geothermal water. Unlike some other trace elements such as B, saline intrusion of sea￾water into an aquifer is unlikely to lead to a significant increase of As in the affected groundwater. Arsenate shares many chemical characteristics with phosphate and hence in oxic marine and estuarine waters, depletions in phosphate in biologically productive surface waters are mirrored by depletions in arsenate. Arsenate concentration minima often coincide with photosynthetic maxima evidenced by high concentrations of chlor￾ophyll a (Cullen and Reimer, 1989). Several studies have noted variations in the behaviour of As during estuarine mixing. Some have reported conservative behaviour. In the unpolluted Krka Estuary of Yugoslavia, Seyler and Martin (1991) observed a linear increase in total As with increasing salinity ranging from 0.13 mg l
1 in fresh waters to 1.8 mg l
1 offshore (i.e. seawater value). How￾ever, other studies have observed non-conservative behaviour (departures from simple mixing) in estuaries due to processes such as diffusion from sediment pore￾waters, coprecipitation with Fe oxides or anthropogenic inputs (e.g. Andreae et al., 1983; Andreae and Andreae, 1989). The flocculation of Fe oxides at the freshwater￾saline interface is an important consequence of increases in pH and salinity. This can lead to major decreases in the As flux to the oceans (Cullen and Reimer, 1989). 2.2.5. Groundwater Background concentrations of As in groundwater are in most countries less than 10 mg l
1 (e.g. Edmunds et al., 1989 for the UK; Welch et al., 2000 for the USA) and sometimes substantially lower. However, values quoted in the literature show a very large range from <0.5 to 5000 mg l
1 (i.e. four orders of magnitude). This range occurs under natural conditions. High concentra￾tions of As are found in groundwater in a variety of environments. These include both oxidising (under condi￾tions of high pH) and reducing aquifers and areas affected by geothermal, mining and industrial activity. Evapora￾tive concentration can also increase concentrations sub￾stantially. Most high-As groundwater provinces are the result of natural occurrences of As. Cases of mining￾induced As pollution are numerous in the literature but tend to be localised. Cases of industrially-induced As pollution (including that from agriculture) may be severe locally (Table 1) but occurrences are relatively rare. Arsenic occurrences in groundwater are discussed more fully in Section 5. 2.2.6. Mine drainage Under the extremely acid conditions of some acid mine drainage, which can have negative pH values (Nordstrom et al., 2000), high concentrations of a wide range of solutes are found, including Fe and As. The highest reported As concentration of 850,000 mg l
1 is from an acid seep in the Richmond mine at Iron Mountain, California (Nordstrom and Alpers, 1999). In a compilation of some 180 samples of mine drainage from the USA, Plumlee et al. (1999) reported concentra￾tions ranging from detection limits (<1 mg l
1 or more) to 340,000 mg l
1 , again the highest values being from the Richmond mine. Gelova (1977) also reported an As concentration of 400,000 mg l
1 from the Ural Moun￾tains. Dissolved As in acid mine waters is rapidly removed as the Fe is oxidised and precipitated and the As scavenged through adsorption. At Iron Mountain, an efficient neutralization plant removes the As and metals for safe disposal. 2.2.7. Sediment porewaters Some high concentrations of As have been found in porewaters extracted from unconsolidated sediments and often form sharp contrasts to the concentrations observed in overlying surface waters (e.g. Belzile and Tessier, 1990). Widerlund and Ingri (1995) found con￾centrations in the range 1.3–166 mg l
1 in porewaters from the Kalix River estuary of northern Sweden. Yan et al. (2000) found As concentrations in the range 3.2–99 mg l
1 in porewaters from clay sediments in Saskatch￾ewan, Canada (Table 1). Increased concentrations have been found in porewaters affected by geothermal inputs. Aggett and Kriegman (1988) found As concentrations up to 6430 mg l
1 in anoxic porewaters from New Zealand. Even higher concentrations can be found in porewaters from sediments affected by mining contamination (tail￾ings, mineral-rich deposits). McCreadie et al. (2000) reported As concentrations up to 100,000 mg l
1 in P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568 525

P L. Smedley, D G. Kinniburgh/Applied Geochemistry 17(2002 )517-568 porewaters extracted from tailings in Ontario(Table 1). species for later laboratory analysis. Alternatively, pre In such cases, high porewater As concentrations are servation with HCl and ascorbic acid has been success most likely to be linked to the strong redox gradients ful although this may destroy monomethylarsonic acid that occur below the sediment-water interface, often (MMAA)if present. over depth scales of centimeters. Burial of fresh organic In rain water, oxidation states of As present will vary matter and the slow diffusion of o, through the sedi according to the source. This is likely to be dominantly ment leads to reducing conditions just below the sed As(In2O3 when derived from smelters, coal burning ment-water interface. This encourages the reduction of and volcanic sources, although organic species may be As(V) and desorption from Fe and Mn oxides, as well as derived by volatilization from soils, arsine(As(-IlDH3) reductive dissolution of these minerals. There is much may derive from landfills and reducing soils such as evidence for cycling of As between shallow sediment peats, and arsenate may be derived from marine aero- porewaters and overlying surface waters in response to sols. Reduced forms will undergo oxidation by O2 in the temporal variations in redox conditions. atmosphere and reactions with atmospheric SO2 or O3 Sullivan and Aller(1996)carried out an elegant study are likely( Cullen and reimer, 1989). of the cycling of As in shallow sediments from the off- In oxic seawater, the As is typically dominated by shore shelf of the Amazon situated far from population As(V), though some As(lln) is invariably present and centres. They measured porewater As and Fe con- becomes of increasing importance in anoxic bottom entration profiles as well as sediment As and Fe(l waters. Ratios of As(V)/As(Ill are typically in the concentrations. There was frequently a well-correlated ange 10-100 in open seawater(Andreae, 1979; Peterson peak in dissolved As and Fe concentrations some 50- and Carpenter, 1983; Pettine et al., 1992). Arsenic(V) 150 cm beneath the surface, with As concentrations in should exist mainly as HAsO and H,Asoa in the ph the peak averaging about 135 ug I-I and reaching a range of seawater (pH around 8.2; Figs. I and 2)and aximum of 300 ug l-l, much greater than from marine As(lID) mainly as the neutral species H3AsO3 Relatively coastal environments. The dissolved As/Fe molar ratio high proportions of H3AsO3 are found in surface ocean varied but was typically about 1: 300. Dissolved As var- waters( Cullen and Reimer, 1989; Cutter et al., 2001) ied inversely with easily-leachable(6 M HCI) As in the These coincide with zones of primary productivity. sediment and increased directly with solid-phase Fe(In). Increases in organic As species have also been recorded In these sediments. Fe oxides were believed to be in these zones as a result of methylation reactions by much more important source of As than Mn oxides phytoplankton. The relative proportions of As species are more variable 2. 2.8. Oilfield and other brines n estuarine waters because of variable redox and salinity. Only limited data are available for As in oilfield and and terrestrial inputs(Howard et al, 1988; Abdullah et al other brines, but some published accounts suggest that 1995). However, they are still dominated by As(v) concentrations can be very high. White et al. (1963) Andreae and Andreae(1989)found As(V/As(lIn ratios reported a dissolved As concentration of 230 ug I-in a varying between 5-50 in the Schelde Estuary of Belgiu Na-HCO3 groundwater from a 1000 m deep oilfield well with the lowest ratios in anoxic zones where inputs of from Ellis Pool, Alberta, Canada. They also reported a industrial effluent had an impact. Increased proportions concentration of 5800 ug I-I As in a Na-Cl-dominated of As(ln also result from inputs of mine effluent 可 n Searles Lal6n1s rom Tisakurt variations in As fornia, have As concentrations up to 243 mg 1-l(Na 119 aries(Riedel, 1993). In seasonally anoxic estuarine I-l White et al., 1963: Table 1) waters, variations in the relative proportions of As(n) and As(V) can be large. Peterson and Carpenter(1983) 2.3. Distribution of arsenic species in water bodies found a distinct crossover in the proportions of the two species with increasing depth in response to the onset of Many studies of As speciation in natural waters have anoxic conditions in the estuarine waters of Saanich been carried out. Most attempt to separate the inorganic Inlet of British Columbia. Arsenic(IIn) represented only species into As(lID) and As(V), usually by chromato- 5%(0.10 ug l-)of the dissolved As above the redox graphic separation or by making use of the relatively front but 87%(1.58 ug l-)below it. In marine and low reduction of As(V) by Na borohydride. Some studies estuarine waters, organic forms are usually less abun lso measure the organic As species. The sampling and dant but are nonetheless often detected(e.g. Riedel, analytical techniques required are not trivial and not yet 1993: Howard et al., 1999). Concentrations of these will ell-established (Edwards et aL., 1998). Both oxidatio depend on abundance and species of biota present and f As(lID and reduction of As(v) may occur during sto- on temperature. rage(Hall et al., 1999). Separation of species may be car- In lake and river waters, As(V) is also generally the ried out in the field to avoid the problem of preserving dominant species(e.g Seyler and Martin, 1990; Pettine

porewaters extracted from tailings in Ontario (Table 1). In such cases, high porewater As concentrations are most likely to be linked to the strong redox gradients that occur below the sediment-water interface, often over depth scales of centimeters. Burial of fresh organic matter and the slow diffusion of O2 through the sedi￾ment leads to reducing conditions just below the sedi￾ment-water interface. This encourages the reduction of As(V) and desorption from Fe and Mn oxides, as well as reductive dissolution of these minerals. There is much evidence for cycling of As between shallow sediment porewaters and overlying surface waters in response to temporal variations in redox conditions. Sullivan and Aller (1996) carried out an elegant study of the cycling of As in shallow sediments from the off- shore shelf of the Amazon situated far from population centres. They measured porewater As and Fe con￾centration profiles as well as sediment As and Fe(II) concentrations. There was frequently a well-correlated peak in dissolved As and Fe concentrations some 50– 150 cm beneath the surface, with As concentrations in the peak averaging about 135 mg l
1 and reaching a maximum of 300 mg l
1 , much greater than from marine coastal environments. The dissolved As/Fe molar ratio varied but was typically about 1:300. Dissolved As var￾ied inversely with easily-leachable (6 M HCl) As in the sediment and increased directly with solid-phase Fe(II). In these sediments, Fe oxides were believed to be a much more important source of As than Mn oxides. 2.2.8. Oilfield and other brines Only limited data are available for As in oilfield and other brines, but some published accounts suggest that concentrations can be very high. White et al. (1963) reported a dissolved As concentration of 230 mg l
1 in a Na–HCO3 groundwater from a 1000 m deep oilfield well from Ellis Pool, Alberta, Canada. They also reported a concentration of 5800 mg l
1 As in a Na–Cl-dominated brine from Tisaku¨rt, Hungary. Composite brines from the interstices of salt deposits from Searles Lake, Cali￾fornia, have As concentrations up to 243 mg l
1 (Na 119 g l
1 ; White et al., 1963; Table 1). 2.3. Distribution of arsenic species in water bodies Many studies of As speciation in natural waters have been carried out. Most attempt to separate the inorganic species into As(III) and As(V), usually by chromato￾graphic separation or by making use of the relatively slow reduction of As(V) by Na borohydride. Some studies also measure the organic As species. The sampling and analytical techniques required are not trivial and not yet well-established (Edwards et al., 1998). Both oxidation of As(III) and reduction of As(V) may occur during sto￾rage (Hall et al., 1999). Separation of species may be car￾ried out in the field to avoid the problem of preserving species for later laboratory analysis. Alternatively, pre￾servation with HCl and ascorbic acid has been success￾ful although this may destroy monomethylarsonic acid (MMAA) if present. In rain water, oxidation states of As present will vary according to the source. This is likely to be dominantly As(III)2O3 when derived from smelters, coal burning and volcanic sources, although organic species may be derived by volatilization from soils, arsine (As(-III)H3) may derive from landfills and reducing soils such as peats, and arsenate may be derived from marine aero￾sols. Reduced forms will undergo oxidation by O2 in the atmosphere and reactions with atmospheric SO2 or O3 are likely (Cullen and Reimer, 1989). In oxic seawater, the As is typically dominated by As(V), though some As(III) is invariably present and becomes of increasing importance in anoxic bottom waters. Ratios of As(V)/As(III) are typically in the range 10–100 in open seawater (Andreae, 1979; Peterson and Carpenter, 1983; Pettine et al., 1992). Arsenic(V) should exist mainly as HAsO4 2- and H2AsO4 - in the pH range of seawater (pH around 8.2; Figs. 1and 2) and As(III) mainly as the neutral species H3AsO3. Relatively high proportions of H3AsO3 are found in surface ocean waters (Cullen and Reimer, 1989; Cutter et al., 2001). These coincide with zones of primary productivity. Increases in organic As species have also been recorded in these zones as a result of methylation reactions by phytoplankton. The relative proportions of As species are more variable in estuarine waters because of variable redox and salinity, and terrestrial inputs (Howard et al., 1988; Abdullah et al., 1995). However, they are still dominated by As(V). Andreae and Andreae (1989) found As(V)/As(III) ratios varying between 5–50 in the Schelde Estuary of Belgium with the lowest ratios in anoxic zones where inputs of industrial effluent had an impact. Increased proportions of As(III) also result from inputs of mine effluent (Klumpp and Peterson, 1979). Seasonal variations in As concentration and speciation have been noted in estu￾aries (Riedel, 1993). In seasonally anoxic estuarine waters, variations in the relative proportions of As(III) and As(V) can be large. Peterson and Carpenter (1983) found a distinct crossover in the proportions of the two species with increasing depth in response to the onset of anoxic conditions in the estuarine waters of Saanich Inlet of British Columbia. Arsenic(III) represented only 5% (0.10 mg l
1 ) of the dissolved As above the redox front but 87% (1.58 mg l
1 ) below it. In marine and estuarine waters, organic forms are usually less abun￾dant but are nonetheless often detected (e.g. Riedel, 1993; Howard et al., 1999). Concentrations of these will depend on abundance and species of biota present and on temperature. In lake and river waters, As(V) is also generally the dominant species (e.g. Seyler and Martin, 1990; Pettine 526 P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568