P L. Smedley, D.G. Kinniburgh/ Applied Geochemistry 17(2002)517-568 Table 4(continued) As concentration average and/ No of Reference r range(mg kg-) analyses lining-contaminated reservoir sediment, Montana 100-800 Moore et al. (1988) Mine tailings. British Colombia 9030396-2000 Soils and tailings- contaminated soil. ul Kavanagh et al. (1997) aminated soil, Montana to 1100 Nagorski and Moore (1999) Industrially polluted inter-tidal sediments, USA 0.38-1260 Davis et al. (1997) Soils below chemicals factory, USA Sewage sludge 9.8(24396 Zhu and Tabatabai (1995) sediments are enriched in As relative to igneous rocks. are also relatively enriched in As, with values up to ca Sands and sandstones tend to have the lowest con- 400 mg kg-I having been reported centrations. reflecting the low As concentrations of thei dominant minerals: quartz and feldspars. Average sand 3. 2. 4. Unconsolidated sediments stone As concentrations are around 4 mg kg-I(Table 4 Concentrations of as in unconsolidated sediments are although Ure and Berrow (1982)gave a lower average not notably different from those in their indurated equiva- value of 1 mg kg- lents. Muds and clays usually have higher concentrations Argillaceous deposits have a broader range and than sands and carbonates. values are typically 3-10 mg igher average As concentrations than sandstones, with kg-, depending on texture and mineralogy (Table 4) a typical average of around 13 mg kg- Table 4: Ure and Elevated concentrations tend to reflect the amounts of Berrow, 1982). The higher values reflect the larger pro- pyrite or Fe oxides present. Increases are also common portion of sulphide minerals, oxides, organic matter and in mineralised areas. Placer deposits in streams can have clays. Black shales have As concentrations at the high very high concentrations as a result of the abundance of end of the range, principally because of their enhanced sulphide minerals. Average As concentrations for pyrite content Data given in Table 4 suggest that mar- stream sediments in England and Wales are in the range ine argillaceous deposits have higher concentrations 5-8 mg kg(AGRG, 1978). Similar concentrations than non-marine deposits. This may also be a reflection have also been found in river sediments where ground of the grain-size distributions, with potential for a water-As concentrations are high: Datta and Sub- higher proportion of fine material in offshore pelagic rumanian (1997) found concentrations in sediments sediments as well as systematic differences in sulphur from the River Ganges averaging 2.0 mg kg-(range and pyrite contents. Marine shales tend to contain 1. 2-2.6 mg kg-). from the Brahmaputra River aver higher S concentrations. Sediment provenance is also a aging 2.8 mg kg-I(range 1. 4-5.9 mg kg-)and from the likely factor. Particularly high As concentrations have Meghna River averaging 3.5 mg kg-I(range 1.3-5.6 mg been determined for shales from mid-ocean settings kg) (Mid-Atlantic Ridge average 174 mg kg-I Table 4) Cook et al. (1995) found concentrations in lake sedi variable but often high. Samples of organic-rich sh"re ments ranging between 0.9 and 44 mg kg-I(median 5.5 Concentrations in coals and bituminous deposits mg kg)but noted that the highest concentrations were (Kupferschiefer) from Germany have As concentrations present up to a few kilometres down-slope of miner of 100-900 mg kg(Riedel and Eikmann, 1986: alised areas. The upper baseline concentration for these Table 4). Some coal samples have been found with sediments is likely to be around 13 mg kg(90th per- extremely high concentrations up to 35,000 mg kg- centile). They also found concentrations in glacial till of (Belkin et al., 2000), although generally low concentra 1.9-170 mg kg(median 9.2 mg kg Table 4) and tions of 2.5-17 mg kg-I were reported by Palmer and noted the highest concentrations down-ice of mine Klizas(1997). Carbonate rocks typically have low con- alised areas (upper baseline, 90th percentile, 22 centrations, reflecting the low concentrations of the kg-). Relative As enrichments have been observed in constituent minerals(ca 3 mg kg; Table 4) reducing sediments in both nearshore and continental- Some of the highest observed As concentrations are shelf deposits(Peterson and Carpenter, 1986; Legeleux found in ironstones and Fe-rich rocks. James(1966) et aL., 1994). Legeleux et al. (1994)noted concentrations collated data for ironstones from various parts of the increasing with depth(up to 30 cm)in continental shelf world and reported As concentrations up to 800 mg sediments as a result of the generation of increasingly g-l in a chamosite- limonite oolite from the former reducing conditions. Concentrations varied between USSR. Boyle and Jonasson(1973) gave concentrations for sites, but generally increased with depth in the range Fe-rich rocks up to 2900 mg kg-l(table 4). Phosphorites 2.3-8.2 mg kg-(Table 4)
sediments are enriched in As relative to igneous rocks. Sands and sandstones tend to have the lowest concentrations, reflecting the low As concentrations of their dominant minerals: quartz and feldspars. Average sandstone As concentrations are around 4 mg kg1 (Table 4) although Ure and Berrow (1982) gave a lower average value of 1mg kg1 . Argillaceous deposits have a broader range and higher average As concentrations than sandstones, with a typical average of around 13 mg kg1 (Table 4; Ure and Berrow, 1982). The higher values reflect the larger proportion of sulphide minerals, oxides, organic matter and clays. Black shales have As concentrations at the high end of the range, principally because of their enhanced pyrite content. Data given in Table 4 suggest that marine argillaceous deposits have higher concentrations than non-marine deposits. This may also be a reflection of the grain-size distributions, with potential for a higher proportion of fine material in offshore pelagic sediments as well as systematic differences in sulphur and pyrite contents. Marine shales tend to contain higher S concentrations. Sediment provenance is also a likely factor. Particularly high As concentrations have been determined for shales from mid-ocean settings (Mid-Atlantic Ridge average 174 mg kg1 ; Table 4). Concentrations in coals and bituminous deposits are variable but often high. Samples of organic-rich shale (Kupferschiefer) from Germany have As concentrations of 100–900 mg kg1 (Riedel and Eikmann, 1986; Table 4). Some coal samples have been found with extremely high concentrations up to 35,000 mg kg1 (Belkin et al., 2000), although generally low concentrations of 2.5–17 mg kg1 were reported by Palmer and Klizas (1997). Carbonate rocks typically have low concentrations, reflecting the low concentrations of the constituent minerals (ca. 3 mg kg1 ; Table 4). Some of the highest observed As concentrations are found in ironstones and Fe-rich rocks. James (1966) collated data for ironstones from various parts of the world and reported As concentrations up to 800 mg kg1 in a chamosite-limonite oolite from the former USSR. Boyle and Jonasson (1973) gave concentrations for Fe-rich rocks up to 2900 mg kg1 (Table 4). Phosphorites are also relatively enriched in As, with values up to ca. 400 mg kg1 having been reported. 3.2.4. Unconsolidated sediments Concentrations of As in unconsolidated sediments are not notably different from those in their indurated equivalents. Muds and clays usually have higher concentrations than sands and carbonates. Values are typically 3–10 mg kg1 , depending on texture and mineralogy (Table 4). Elevated concentrations tend to reflect the amounts of pyrite or Fe oxides present. Increases are also common in mineralised areas. Placer deposits in streams can have very high concentrations as a result of the abundance of sulphide minerals. Average As concentrations for stream sediments in England and Wales are in the range 5–8 mg kg1 (AGRG, 1978). Similar concentrations have also been found in river sediments where groundwater-As concentrations are high: Datta and Subramanian (1997) found concentrations in sediments from the River Ganges averaging 2.0 mg kg1 (range 1.2–2.6 mg kg1 ), from the Brahmaputra River averaging 2.8 mg kg1 (range 1.4–5.9 mg kg1 ) and from the Meghna River averaging 3.5 mg kg1 (range 1.3–5.6 mg kg1 ). Cook et al. (1995) found concentrations in lake sediments ranging between 0.9 and 44 mg kg1 (median 5.5 mg kg1 ) but noted that the highest concentrations were present up to a few kilometres down-slope of mineralised areas. The upper baseline concentration for these sediments is likely to be around 13 mg kg1 (90th percentile). They also found concentrations in glacial till of 1.9–170 mg kg1 (median 9.2 mg kg1 ; Table 4) and noted the highest concentrations down-ice of mineralised areas (upper baseline, 90th percentile, 22 mg kg1 ). Relative As enrichments have been observed in reducing sediments in both nearshore and continentalshelf deposits (Peterson and Carpenter, 1986; Legeleux et al., 1994). Legeleux et al. (1994) noted concentrations increasing with depth (up to 30 cm) in continental shelf sediments as a result of the generation of increasingly reducing conditions. Concentrations varied between sites, but generally increased with depth in the range 2.3–8.2 mg kg1 (Table 4). Table 4 (continued) Rock/sediment type As concentration average and/ or range (mg kg1 ) No of analyses Reference Mining-contaminated reservoir sediment, Montana 100–800 Moore et al. (1988) Mine tailings, British Colombia 903 (396–2000) Azcue et al. (1995) Soils and tailings-contaminated soil, UK 120–52,600 86 Kavanagh et al. (1997) Tailings-contaminated soil, Montana up to 1100 Nagorski and Moore (1999) Industrially polluted inter-tidal sediments, USA 0.38–1260 Davis et al. (1997) Soils below chemicals factory, USA 1.3–4770 Hale et al. (1997) Sewage sludge 9.8 (2.4–39.6) Zhu and Tabatabai (1995) 532 P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568
P L. Smedley, D G. Kinniburgh/Applied Geochemistry 17(2002 )517-568 3. 2.5. Soils the mineral type(Krause and Ettel, 1989). There is some Baseline concentrations of As in soils are generally of onfusion in the analysis of these solubility relationships the order of 5-10 mg kg. boyle and Jonasson(1973) between congruent dissolution, incongruent dissolution quoted an average baseline concentration in world soils and sorption/desorption reactions. Secondary arsenolite of 7.2 mg kg-I(Table 4)and Shacklette et al.(1974) (As O3)is also relatively soluble. Arsenic bound to Fe quoted an average of 7. 4 mg kg(901 samples) for oxides is relatively immobile, particularly under oxidising American soils. Ure and Berrow (1982)gave a higher conditions average value of 11.3 mg kg. Peats and bog soils can have higher concentrations (average 13 mg kg 3.3. The atmosphere Table 4), principally because of increased prevalence of sulphide mineral phases under the reduced conditions The concentrations of As in the atmosphere are Acid sulphate soils which are generated by the oxidation usually low but as noted above, are increased by inputs of pyrite in sulphide-rich terrains such as pyritic shales, from smelting and other industrial operations, fossil mineral veins and dewatered mangrove swamps can also fuel combustion and volcanic activity. Concentrations be relatively enriched in As. Dudas(1984) found As amounting to around 10--10-3 ug m-3 have been concentrations up to 45 mg kg-I in the b horizons of recorded in unpolluted areas, increasing to 0.003-0.18 acid sulphate soils derived from the weathering of pyr ug m-3 in urban areas and greater than 1 ug m close ite-rich shales in Canada. Concentrations in the over- o industrial plants (WHO, 2001). Much of the lying leached (eluvial, E) horizons were low (1.5-8.0 mg atmospheric As is particulate. Total As deposition rates kg-')as a result of volatilisation or leaching of As to have been calculated in the range <1-1000 ug m-2a lower levels. Gustafsson and Tin(1994) found similarly depending on the relative proportions of wet and dry evated concentrations(up to 41 mg kg-)in acid sul- deposition and proximity to contamination sources phate soils from the Mekong delta of vietnam. (Schroeder et al., 1987). Values in the range 38-266 ug Although the dominant source of As in soils is geo- m-2 a-l(29-55% as dry deposition)were estimated for gical, and hence dependent to some extent on the the mid-Atlantic coast(Scudlark and Church, 1988) concentration in the parent rock material, additional Airborne As is transferred to water bodies by wet or dry inputs may be derived locally from industrial sources deposition and may therefore increase the aqueous con such as smelting and fossil-fuel combustion products entration slightly. However, there is little evidence to and agricultural sources such as pesticides and phos- suggest that atmospheric As poses a real health threat phate fertilisers. Ure and Berrow (1982)quoted con- for drinking-water sources. Atmospheric As arising centrations in the range 366-732 mg kg in orchard from coal burning has been invoked as a major cause of soils as a result of the historical application of arsenical lung cancer in parts of China( Guizhou Province), but pesticides to fruit crops. threat is from direct inhalation of domestic coal- smoke and especially from consumption of foods dried 3. 2.6. Contaminated surficial deposits over domestic coal fires, rather than from drinking water Arsenic concentrations much higher than baseline affected by atmospheric inputs( Finkelman et al., 1999) values have been found in sediments and soils con- laminated by the products of mining activity, including mine tailings and effluent Concentrations in tailings piles 4. Mineral-water interactions and tailings-contaminated soils can reach up to several thousand mg kg(Table 4). The high concentration 4. Controls on arsenic mobilisation reflect not only increased abundance of primary As-rich sulphide minerals, but also secondary Fe arsenates and As with most trace metals. the concentration of as in Fe oxides formed as reaction products of the original ore natural waters is probably normally controlled by some minerals. The primary sulphide minerals are susceptible form of solid-solution interaction. This is most clearly to oxidation in the tailings pile and the secondary the case for soil solutions. interstitial waters and minerals have varying solubility in oxidising conditions groundwaters where the solid/solution ratio is large but in groundwaters and surface waters. Scorodite(FeA- it is also often true in open bodies of water(oceans, sO4. 2H20)is a common sulphide oxidation product and lakes and reservoirs) where the concentration of solid its solubility is considered to control As concentrations particles is small but still significant. In these open bodies in such oxidising sulphide environments. Scorodite is the particles can be of mineral and biological origin. It is metastable under most groundwater conditions and likely that in most soils and aquifers, mineral-As inter tends to dissolve incongruently, forming Fe oxides and actions are likely to dominate over organic matter-As releasing As into solution(Robins, 1987; Krause and interactions, although organic matter may interact to Ettel, 1989). In practice, a wide range of Fe-As mineral some extent through its reactions with the surfaces of olubility relationships are found which in part relate to minerals. Knowing the types of interaction involved is
3.2.5. Soils Baseline concentrations of As in soils are generally of the order of 5–10 mg kg1 . Boyle and Jonasson (1973) quoted an average baseline concentration in world soils of 7.2 mg kg1 (Table 4) and Shacklette et al. (1974) quoted an average of 7.4 mg kg1 (901samples) for American soils. Ure and Berrow (1982) gave a higher average value of 11.3 mg kg1 . Peats and bog soils can have higher concentrations (average 13 mg kg1 ; Table 4), principally because of increased prevalence of sulphide mineral phases under the reduced conditions. Acid sulphate soils which are generated by the oxidation of pyrite in sulphide-rich terrains such as pyritic shales, mineral veins and dewatered mangrove swamps can also be relatively enriched in As. Dudas (1984) found As concentrations up to 45 mg kg1 in the B horizons of acid sulphate soils derived from the weathering of pyrite-rich shales in Canada. Concentrations in the overlying leached (eluvial, E) horizons were low (1.5–8.0 mg kg1 ) as a result of volatilisation or leaching of As to lower levels. Gustafsson and Tin (1994) found similarly elevated concentrations (up to 41mg kg1 ) in acid sulphate soils from the Mekong delta of Vietnam. Although the dominant source of As in soils is geological, and hence dependent to some extent on the concentration in the parent rock material, additional inputs may be derived locally from industrial sources such as smelting and fossil-fuel combustion products and agricultural sources such as pesticides and phosphate fertilisers. Ure and Berrow (1982) quoted concentrations in the range 366–732 mg kg1 in orchard soils as a result of the historical application of arsenical pesticides to fruit crops. 3.2.6. Contaminated surficial deposits Arsenic concentrations much higher than baseline values have been found in sediments and soils contaminated by the products of mining activity, including mine tailings and effluent. Concentrations in tailings piles and tailings-contaminated soils can reach up to several thousand mg kg1 (Table 4). The high concentrations reflect not only increased abundance of primary As-rich sulphide minerals, but also secondary Fe arsenates and Fe oxides formed as reaction products of the original ore minerals. The primary sulphide minerals are susceptible to oxidation in the tailings pile and the secondary minerals have varying solubility in oxidising conditions in groundwaters and surface waters. Scorodite (FeAsO4 .2H2O) is a common sulphide oxidation product and its solubility is considered to control As concentrations in such oxidising sulphide environments. Scorodite is metastable under most groundwater conditions and tends to dissolve incongruently, forming Fe oxides and releasing As into solution (Robins, 1987; Krause and Ettel, 1989). In practice, a wide range of Fe–As mineral solubility relationships are found which in part relate to the mineral type (Krause and Ettel, 1989). There is some confusion in the analysis of these solubility relationships between congruent dissolution, incongruent dissolution and sorption/desorption reactions. Secondary arsenolite (As2O3) is also relatively soluble. Arsenic bound to Fe oxides is relatively immobile, particularly under oxidising conditions. 3.3. The atmosphere The concentrations of As in the atmosphere are usually low but as noted above, are increased by inputs from smelting and other industrial operations, fossilfuel combustion and volcanic activity. Concentrations amounting to around 105 –103 mg m3 have been recorded in unpolluted areas, increasing to 0.003–0.18 mg m3 in urban areas and greater than 1 mg m3 close to industrial plants (WHO, 2001). Much of the atmospheric As is particulate. Total As deposition rates have been calculated in the range <1–1000 mg m2 a1 depending on the relative proportions of wet and dry deposition and proximity to contamination sources (Schroeder et al., 1987). Values in the range 38–266 mg m2 a1 (29–55% as dry deposition) were estimated for the mid-Atlantic coast (Scudlark and Church, 1988). Airborne As is transferred to water bodies by wet or dry deposition and may therefore increase the aqueous concentration slightly. However, there is little evidence to suggest that atmospheric As poses a real health threat for drinking-water sources. Atmospheric As arising from coal burning has been invoked as a major cause of lung cancer in parts of China (Guizhou Province), but the threat is from direct inhalation of domestic coal-fire smoke and especially from consumption of foods dried over domestic coal fires, rather than from drinking water affected by atmospheric inputs (Finkelman et al., 1999). 4. Mineral–water interactions 4.1. Controls on arsenic mobilisation As with most trace metals, the concentration of As in natural waters is probably normally controlled by some form of solid-solution interaction. This is most clearly the case for soil solutions, interstitial waters and groundwaters where the solid/solution ratio is large but it is also often true in open bodies of water (oceans, lakes and reservoirs) where the concentration of solid particles is small but still significant. In these open bodies, the particles can be of mineral and biological origin. It is likely that in most soils and aquifers, mineral–As interactions are likely to dominate over organic matter–As interactions, although organic matter may interact to some extent through its reactions with the surfaces of minerals. Knowing the types of interaction involved is P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568 533
P L. Smedley, D G. Kinniburgh/Applied Geochemistry 17(2002 )517-568 important because this will govern the response of As to mineral components, Fe oxides are probably the most changes in water chemistry. It will also determine the important adsorbents in sandy aquifers because of their modelling approach required for making predictions greater abundance and the strong binding affinity about possible future changes and for understanding Nevertheless, Al oxides can also be expected to play a past changes in As concentration significant role when present in quantity(Hingston et The importance of oxides in controlling the con- al, 1971; Manning and Goldberg, 1997b). Experience ration of As in natural waters has been appreciated from water treatment suggests that below pH 7.5, Al for a long time (Livesey and Huang, 1981; Matisoffet al, hydroxides are about as effective as Fe hydroxides(on 1982)and there has been a wide range of studies to molar basis) for adsorbing As(V) but that Fe salts ar measure the adsorption isotherms on natural and synthetic more efficient at higher pH and for adsorbing As(Ill oxide minerals and to establish the sorption processes at (Edwards, 1994) the molecular scale (Table 5). Even so, importan The interactions of as with Fe oxides have been uncertainties still remain in relation to the interactions tudied in considerable detail in the laboratory and of As(lIn) and As(V) at environmental concentrations therefore provide the best insight into the likely beha and in the presence of other interacting ions. Fre- vior of As-mineral interactions in aquifers. However, quently, the element which correlates best with As in most of these laboratory studies, particularly the older sediments is Fe. This is also the basis for the use of Fe studies have been undertaken at rather high as salts(as well as Al and Mn salts) in water treatment for centrations(Hingston et al., 1971)and there is a paucity the removal of As and other elements(e.g. Edwards, of reliable adsorption data at the low ug I-I level of 1994). The As content of residual sludges from water relevance to natural waters. In addition, there is uncer- treatment can be in the range 1000-10, 000 mg kg tainty over the extent to which the Fe oxides most (Forstner and Haase, 1998: Driehaus et al., 1998). Clays commonly studied in the laboratory reflect the Fe oxides can also adsorb As(Ill) and As(V)(Manning and found in the field. Field data for As(V)adsorption to Goldberg, 1997b) but their role in sediments in terms of natural 'diagenetic Fe oxides(captured in a lake with As binding is unclear at presen vertically-installed Teflon sheets) closely paralleled the It is difficult to study mineral-water interactions laboratory data of Pierce and Moore(1982) which was directly in aquifers. Most studies, including those with a included in the dzombak and morel (1990)data base(De bearing on As in groundwater, have been undertaken either Vitre et al., 1991). However, it was considerably greater in soils, or in lake or ocean sediments and usually from than that calculated using Hingston et al.s(1971)data quite shallow depths. There is much to be learnt from the for As(V)adsorption on goethite, highlighting the high studies of soils and sediments since the same general prin- affinity for As of freshly-formedamorphous' Fe oxides. areas where cross-fertilization of ideas can occur is in acid dissolution of a synthetic ferrihydrite containing understanding the behaviour of Fe oxides in reducing soils sorbed As(V)and concluded that the dissolution was and sediments and the influence of this on the release of as incongruent (i.e. Fe and As were not released in the Matisoff et al. ( 1982) related reductive dissolution of Fe same proportion as found in the bulk mineral) and that oxides to the possible release of As in groundwater from an the initial As released was probably sorbed on the sur- alluvial aquifer in NE Ohio Korte(1991)and Korte and face of the very small ferrihydrite particles. The same is Fernando(1991)also speculated that t desorption of As likely to happen during reductive dissolution. The from Fe oxides could occur in reducing, alluvial sediments adsorbed As also slowed down the acid dissolution of and that this could lead to high -As groundwaters. the ferrihydrite 4. 2. Arsenic associations in sediments 43. Reduced sediments and the role of iron oxides The major minerals binding As(as both arsenate and A well-known sequence of reduction reactions occurs arsenite)in sediments are the metal oxides, particularly when lakes, fjords, soils, sediments and aquifers become those of Fe, Al and Mn(De Vitre et al., 1991; Sullivan anaerobic(Berner, 1981; Stumm and Morgan, 1995; nd Aller, 1996). About 50% of the Fe in freshwater Langmuir, 1997). The processes causing changes in F sediments is in the form of Fe oxides and about 20% of redox chemistry are particularly important since they the Fe is 'reactive Fe. Clays also adsorb As because can directly affect the mobility of As. One of the princi the oxide- like character of their edges. The extent of pal causes of high As concentrations in subsurface As(V) sorption to, and coprecipitation on, carbonate waters is the reductive dissolution of hydrous Fe oxides minerals is unknown but if it behaves like phosphate, it and or the release of adsorbed or combined As. Thi is likely to be strongly retained by these minerals and sequence begins with the consumption of o, and an this may limit As concentrations in groundwaters from increase in dissolved CO2 from the decomposition of limestone aquifers (Millero et al., 2001). Of these organic matter. Next, NO decreases by reduction to
important because this will govern the response of As to changes in water chemistry. It will also determine the modelling approach required for making predictions about possible future changes and for understanding past changes in As concentrations. The importance of oxides in controlling the concentration of As in natural waters has been appreciated for a long time (Livesey and Huang, 1981; Matisoff et al., 1982) and there has been a wide range of studies to measure the adsorption isotherms on natural and synthetic oxide minerals and to establish the sorption processes at the molecular scale (Table 5). Even so, important uncertainties still remain in relation to the interactions of As(III) and As(V) at environmental concentrations and in the presence of other interacting ions. Frequently, the element which correlates best with As in sediments is Fe. This is also the basis for the use of Fe salts (as well as Al and Mn salts) in water treatment for the removal of As and other elements (e.g. Edwards, 1994). The As content of residual sludges from water treatment can be in the range 1000–10,000 mg kg1 (Forstner and Haase, 1998; Driehaus et al., 1998). Clays can also adsorb As(III) and As(V) (Manning and Goldberg, 1997b) but their role in sediments in terms of As binding is unclear at present. It is difficult to study mineral-water interactions directly in aquifers. Most studies, including those with a bearing on As in groundwater, have been undertaken either in soils, or in lake or ocean sediments and usually from quite shallow depths. There is much to be learnt from the studies of soils and sediments since the same general principles are expected to apply. One of the most important areas where cross-fertilization of ideas can occur is in understanding the behaviour of Fe oxides in reducing soils and sediments and the influence of this on the release of As. Matisoff et al. (1982) related reductive dissolution of Fe oxides to the possible release of As in groundwater from an alluvial aquifer in NE Ohio. Korte (1991) and Korte and Fernando (1991) also speculated that desorption of As from Fe oxides could occur in reducing, alluvial sediments and that this could lead to high-As groundwaters. 4.2. Arsenic associations in sediments The major minerals binding As (as both arsenate and arsenite) in sediments are the metal oxides, particularly those of Fe, Al and Mn (De Vitre et al., 1991; Sullivan and Aller, 1996). About 50% of the Fe in freshwater sediments is in the form of Fe oxides and about 20% of the Fe is ‘reactive’ Fe. Clays also adsorb As because of the oxide-like character of their edges. The extent of As(V) sorption to, and coprecipitation on, carbonate minerals is unknown but if it behaves like phosphate, it is likely to be strongly retained by these minerals and this may limit As concentrations in groundwaters from limestone aquifers (Millero et al., 2001). Of these mineral components, Fe oxides are probably the most important adsorbents in sandy aquifers because of their greater abundance and the strong binding affinity. Nevertheless, Al oxides can also be expected to play a significant role when present in quantity (Hingston et al., 1971; Manning and Goldberg, 1997b). Experience from water treatment suggests that below pH 7.5, Al hydroxides are about as effective as Fe hydroxides (on a molar basis) for adsorbing As(V) but that Fe salts are more efficient at higher pH and for adsorbing As(III) (Edwards, 1994). The interactions of As with Fe oxides have been studied in considerable detail in the laboratory and therefore provide the best insight into the likely behavior of As-mineral interactions in aquifers. However, most of these laboratory studies, particularly the older studies, have been undertaken at rather high As concentrations (Hingston et al., 1971) and there is a paucity of reliable adsorption data at the low mg l1 level of relevance to natural waters. In addition, there is uncertainty over the extent to which the Fe oxides most commonly studied in the laboratory reflect the Fe oxides found in the field. Field data for As(V) adsorption to natural ‘diagenetic’ Fe oxides (captured in a lake with vertically-installed Teflon sheets) closely paralleled the laboratory data of Pierce and Moore (1982) which was included in the Dzombak and Morel (1990) database (De Vitre et al., 1991). However, it was considerably greater than that calculated using Hingston et al.’s (1971) data for As(V) adsorption on goethite, highlighting the high affinity for As of freshly-formed ‘amorphous’ Fe oxides. Paige et al. (1997) measured the As/Fe ratios during the acid dissolution of a synthetic ferrihydrite containing sorbed As(V) and concluded that the dissolution was incongruent (i.e. Fe and As were not released in the same proportion as found in the bulk mineral) and that the initial As released was probably sorbed on the surface of the very small ferrihydrite particles. The same is likely to happen during reductive dissolution. The adsorbed As also slowed down the acid dissolution of the ferrihydrite. 4.3. Reduced sediments and the role of iron oxides A well-known sequence of reduction reactions occurs when lakes, fjords, soils, sediments and aquifers become anaerobic (Berner, 1981; Stumm and Morgan, 1995; Langmuir, 1997). The processes causing changes in Fe redox chemistry are particularly important since they can directly affect the mobility of As. One of the principal causes of high As concentrations in subsurface waters is the reductive dissolution of hydrous Fe oxides and/or the release of adsorbed or combined As. This sequence begins with the consumption of O2 and an increase in dissolved CO2 from the decomposition of organic matter. Next, NO3 - decreases by reduction to 534 P.L. Smedley, D.G. Kinniburgh / Applied Geochemistry 17 (2002) 517–568