arsenic in groundwater: testing pollution mechanisms...

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1 This paper is reproduced courtesy of AGU, the American Geophysical Union, the copyright holders, from Water Resources Research, 37(1), 109-117. Arsenic in groundwater: testing pollution mechanisms for sedimentary aquifers in Bangladesh. J.M. McArthur, Geological Sciences, UCL, Gower Street, London WC1E 6BT, UK. [email protected] P. Ravenscroft, Mott MacDonald International, 122 Gulshan Avenue, Dhaka 1212, Bangladesh. S. Safiullah, Dept. Environmental Sciences, Jahangirnagar University, Savar, Dhaka, Bangladesh. M.F. Thirlwall, Department of Geology, RHUL, Egham, Surrey TW20 0EX, UK. Abstract In the deltaic plain of the Ganges-Meghna-Brahmaputra rivers, arsenic concentrations in groundwater commonly exceed regulatory limits (50 μg l -1 ) because FeOOH is microbially reduced and releases its sorbed load of arsenic to groundwater. Neither pyrite oxidation nor competitive exchange with fertilizer-phosphate contribute to arsenic pollution. The most intense reduction, and so severest pollution, is driven by microbial degradation of buried deposits of peat. Concentrations of ammonium up to 23 mg l -1 come from microbial fermentation of buried peat and organic waste in latrines. Concentrations of phosphorus of up to 5 mg l -1 come from the release of sorbed phosphorus when FeOOH is reductively dissolved, and from degradation of peat and organic waste from latrines. Calcium and barium in groundwater come from dissolution of detrital (and possibly pedogenic) carbonate, whilst magnesium is supplied by both carbonate dissolution and weathering of mica. The 87 Sr/ 86 Sr values of dissolved strontium define a two component mixing trend between monsoonal rainfall (0.711 ± 0.001) and detrital carbonate (0.735). 1. Introduction Aquifers less than 300 m deep (mostly < 100 m) provide Bangladesh and West Bengal with more than 90% of its drinking water. The groundwater contains more than 50 μg l -1 of arsenic in up to 1 000 000 water wells and adversely affects health, putting up to 20 million people at risk [Dhar et al., 1997; Ullah, 1998; Mandal et al., 1998; DPHE, 1999; http://bicn.com/acic/, 28/07/00]. We use new data for Bangladesh well waters, and literature data, to test three mechanisms invoked to explain arsenic release to this groundwater, i.e. reductive dissolution of FeOOH and release of sorbed arsenic to groundwater, oxidation of arsenical pyrite, and anion (competitive) exchange of sorbed arsenic with phosphate from fertilizer. We show that neither fertilizer-phosphate nor pyrite oxidation cause arsenic pollution (a term meaning the addition to the environment of a species in amounts sufficient to cause environmental harm). We postulate that the severity and distribution of arsenic pollution is controlled by the distribution of buried peat deposits, rather than the distribution of arsenic in aquifer sediments, as the former drives reduction of FeOOH. This postulate has wide applicability because the process of FeOOH reduction is generic and not limited by geography nor by time. 2. The Ganges-Meghna-Brahmaputra Delta Plain Arsenic Pollution The area to the west of the Meghna and north of the Ganges, is occupied by slightly elevated alluvial terraces of the Barind and Madhupur Tracts (Fig. 1), which are underlain by deposits of Lower Pleistocene age [Alam et al., 1990]. Aquifers beneath these areas are assigned to the Dupi Tila Formation. There are sharp lateral contrasts in age between the terraces and the Holocene

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Page 1: Arsenic in groundwater: testing pollution mechanisms …users.physics.harvard.edu/~wilson/arsenic/references/McArthur_et... · Arsenic in groundwater: testing pollution mechanisms

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This paper is reproduced courtesy of AGU, the American Geophysical Union, the copyright holders, from Water Resources Research, 37(1), 109-117. Arsenic in groundwater: testing pollution mechanisms for sedimentary aquifers in Bangladesh. J.M. McArthur, Geological Sciences, UCL, Gower Street, London WC1E 6BT, UK. [email protected] P. Ravenscroft, Mott MacDonald International, 122 Gulshan Avenue, Dhaka 1212, Bangladesh. S. Safiullah, Dept. Environmental Sciences, Jahangirnagar University, Savar, Dhaka, Bangladesh. M.F. Thirlwall, Department of Geology, RHUL, Egham, Surrey TW20 0EX, UK. Abstract

In the deltaic plain of the Ganges-Meghna-Brahmaputra rivers, arsenic concentrations in groundwater commonly exceed regulatory limits (50 µg l-1) because FeOOH is microbially reduced and releases its sorbed load of arsenic to groundwater. Neither pyrite oxidation nor competitive exchange with fertilizer-phosphate contribute to arsenic pollution. The most intense reduction, and so severest pollution, is driven by microbial degradation of buried deposits of peat. Concentrations of ammonium up to 23 mg l-1 come from microbial fermentation of buried peat and organic waste in latrines. Concentrations of phosphorus of up to 5 mg l-1 come from the release of sorbed phosphorus when FeOOH is reductively dissolved, and from degradation of peat and organic waste from latrines. Calcium and barium in groundwater come from dissolution of detrital (and possibly pedogenic) carbonate, whilst magnesium is supplied by both carbonate dissolution and weathering of mica. The 87Sr/86Sr values of dissolved strontium define a two component mixing trend between monsoonal rainfall (0.711 ± 0.001) and detrital carbonate (≤ 0.735). 1. Introduction Aquifers less than 300 m deep (mostly < 100 m) provide Bangladesh and West Bengal with more than 90% of its drinking water. The groundwater contains more than 50 µg l-1 of arsenic in up to 1 000 000 water wells and adversely affects health, putting up to 20 million people at risk [Dhar et al., 1997; Ullah, 1998; Mandal et al., 1998; DPHE, 1999; http://bicn.com/acic/, 28/07/00]. We use new data for Bangladesh well waters, and literature data, to test three mechanisms invoked to explain arsenic release to this groundwater, i.e. reductive dissolution of FeOOH and release of sorbed arsenic to groundwater, oxidation of arsenical pyrite, and anion (competitive) exchange of sorbed arsenic with phosphate from fertilizer. We show that neither fertilizer-phosphate nor pyrite oxidation cause arsenic pollution (a term meaning the addition to the environment of a species in amounts sufficient to cause environmental harm). We postulate that the severity and distribution of arsenic pollution is controlled by the distribution of buried peat deposits, rather than the distribution of arsenic in aquifer sediments, as the former drives reduction of FeOOH. This postulate has wide applicability because the process of FeOOH reduction is generic and not limited by geography nor by time. 2. The Ganges-Meghna-Brahmaputra Delta Plain Arsenic Pollution The area to the west of the Meghna and north of the Ganges, is occupied by slightly elevated alluvial terraces of the Barind and Madhupur Tracts (Fig. 1), which are underlain by deposits of Lower Pleistocene age [Alam et al., 1990]. Aquifers beneath these areas are assigned to the Dupi Tila Formation. There are sharp lateral contrasts in age between the terraces and the Holocene

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Fig. 1. Map of Bangladesh with circled areas showing study areas of DPHE [1999, 2000].CN = Chapai Nawabganj, F = Faridpur, L = Lakshmipur. Colouring shows the percentage of wells that exceed an arsenic concentration of 0.05 mg l-1, as estimated from Union averages of 18 471 data and based on the centre of each Union. Calculated using a fixed radius of 7.5 km, a 1.5 km grid, and 3125 Union centres. Unions are administrative areas. Cross hatched areas are old and elevatedterraces in which groundwater is free of arsenic pollution. Cross-hatched areas are elevated Madhupur and Barind Tracts.

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floodplains [Ravenscroft, in press], owing to the effect of river incision during the Pleistocene sea level low. Maximum incision occurred 18,000 years ago when world sea level was about 120 m below the present level. The main rivers may have cut down more than 100m along the axial courses [Umitsu, 1993], and formed a broad plain about 50 m below the present surface of the modern coastal plains [Goodbred and Kuehl, 1999, 2000]. Rapid sedimentary infilling resulted in

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Fig. 2. Percentage of wells in Bangladesh exceeding specified arsenic concentrations, shown as a function ofdepth (data from Regional Survey, DPHE 1999).

regional fining upward sequences. The alluvial infill ranges from coarse sand and gravel at the base and passes upwards through sand deposits, laid down by braided rivers, into more heterogeneous sand and silts, laid down by meandering streams. Extensive peat deposits accumulated during the mid-Holocene climatic optimum [Reimann, 1993; Umitsu, 1993]. Aquifers beneath the elevated alluvial terraces (Dupi Tila Formation) are almost free of arsenic pollution. In aquifers beneath the Holocene floodplains, within the alluvial and deltaic plains of the Ganges, Meghna, and Brahmaputra (in Bangladesh, Jamuna) rivers, concentrations of arsenic (Fig. 1) commonly exceed the Bangladesh drinking-water standard (50 µg l-1). The distribution of pollution is very patchy, being commonest in the southeast and northeast of Bangladesh. Limited data show that highest arsenic concentrations occur at depths of around 30 m [Frisbie et al., 1999; Karim et. al., 1997; Roy Chowdhury et al., 1999; Acharyya et al., 1999; AAN, 1999]. Using 2024 new data-pairs of well depth and arsenic concentration [DPHE, 1999], we have graphed, as a function of depth, the percentage of wells that exceed regulatory limits (Fig. 2) and so confirm that the highest percentage of contaminated wells occurs at depths between 28 and 45 m. Hand-dug wells are mostly < 5 m deep and usually unpolluted by arsenic. Below 45 m, a reduction occurs in the percentage of wells that are contaminated, but risk remains significant until well-depth exceeds 150 m.

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Water Composition We use data from Nickson et al. [2000], new data in Table 1, and published data from DPHE [1999, 2000] for two areas of Bangladesh, viz. Faridpur and Lakshmipur (Fig. 1). Analytical methods used to obtain DPHE data are given in DPHE [1999]. Our 87Sr/86Sr data (Table 1) were obtained on unfiltered, acidified, water samples using the method given in McArthur et al. [1991]. We do not use DPHE data for Nawabganj because those EC and bicarbonate data are suspect [McArthur et al., unpublished]. We use δ13C data from DPHE [1999] rather than the modified data in DPHE [2000], as we believe the former more accurately reflect aquifer values. When discussing chemical mechanisms, rather than arsenic distributions, we use data only for wells less than 100 m depth, as the severe arsenic pollution occurs at these depths (Fig. 2). Full data are available from http://www.bgs.ac.uk/arsenic/Bangladesh/home.htm. The waters in the Ganges-Meghna-Brahmaputra delta plain (GMBD) are anoxic, calcium-magnesium bicarbonate waters [DPHE, 2000]. Typically, they contain neither dissolved oxygen nor nitrate, which have been removed by reduction. Localised pollution adds nitrate and/or sulfate to a few wells and, in a few others, especially where sodium and chloride are high, sulfate may be remnant from marine connate water. Waters commonly contain concentrations of ammonium and phosphorus in the milligramme per litre range, and hundreds of microgrammes per litre of arsenic. Values of pH range from 6.4 to 7.6 (minimum 5.9 at Nawabganj; DPHE 2000). Concentrations of silica (as H4SiO4) reach 131 mg l-1. Free methane occurs in the aquifer [Ahmed et al., 1998]. Saturation indices, calculated with WATEQF embedded in NETPATH [Plummer et al., 1994], shows that most waters are at close to equilibrium with calcite and dolomite, with saturation indices for both ranging from +0.6 to -0.4 in Faridpur and from +1.2 to -1.2 in Lakshmipur. Manganese is mostly undersaturated with respect to rhodochrosite (SI from -1.4 to +0.6 at Faridpur and -0.6 to +0.2 at Lakshmipur). Water are mostly oversaturated with vivianite (SI mostly +2 to +3.5 at Faridpur and -0.4 to +4.2 at Lakshmipur) and siderite (SI +0.5 to +1.4 at Faridpur and +0.1 to +1.5 at Lakshmipur). Such oversaturation may reflect slow precipitation kinetics, or the stabilization of iron in solution by organic complexing. 3. Arsenic Pollution Mechanisms Three mechanisms have been invoked to explain arsenic pollution of groundwater in the GMBD: 1) arsenic is released by oxidation of arsenical pyrite in the alluvial sediments as aquifer drawdown permits atmospheric oxygen to invade the aquifer [Mallick and Rajagopal, 1996; Mandal et al., 1998; Roy Chowdhury et al., 1999]; 2) arsenic anions sorbed to aquifer minerals are displaced into solution by competitive exchange of phosphate anions derived from over-application of fertilizer to surface soils [Acharrya, 1999]; 3) anoxic conditions permit reduction of iron oxyhydroxides (FeOOH) and release of sorbed arsenic to solution [Bhattacharya et al., 1997; Nickson et al., 1998, 2000]. We discount pyrite oxidation as a mechanism for arsenic pollution, even though trace pyrite is present in the aquifer sediments [PHED, 1991; AAN, 1999; Nickson et al., 1998, 2000]. Measured sulfur concentrations in aquifer sediments represent both pyritic and organic sulfur but allow upper limits to be placed on pyrite abundance of 0.3% [Nickson et al., 2000], 0.02% [AAN, 1999], 0.1%

Table 1. Sr isotopic composition of Faridpur well waters. DPHE well nos. are those of DPHE [1999, 2000].

No. DPHE Sr 87Sr/ 86Sr No mg l-1

S1 BTS208 0.480 0.72107 S2 0.400 0.72103 S3 BTS243 0.330 0.71536 S4 BTS260 0.370 0.71603 S6 BTS258 0.470 0.72228 S7 BTS206 0.360 0.71798 S8 BTS241 0.420 0.72065 S9 BTS242 0.400 0.71903

S10 BTS214 0.300 0.71349 S11 0.554 0.72343 S12 0.486 0.72333 S14 0.497 0.72400 S15 0.591 0.72510

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[J.M. McArthur unpublished] and 0.06% [DPHE, 1999]. The presence of pyrite shows that it has not been oxidised and that it is a sink for, not a source of, arsenic in Bangladesh groundwater. Were pyrite to be oxidised, its arsenic would be sorbed to the resulting FeOOH [Mok and Wai, 1994; Savage et al., 2000], rather than be released to groundwater. Furthermore, Bangladesh groundwaters, which are anoxic, would contain iron and sulfate in the molar ratio of 0.5 were pyrite oxidation releasing arsenic; in reality, these constituents are mutually exclusive in solution

[DPHE, 2000], as are arsenic and sulfate, i.e. arsenic concentrations above 50 µg l-1 are found only where sulfate concentrations are less than 30 mg l-1 [DPHE, 1999, 2000]. Finally, arsenic pollution is uncommon in hand-dug wells [DPHE, 1999] which are shallowest and most exposed to atmospheric oxygen and so would be most polluted were arsenic derived from pyrite by oxidation. Arsenic pollution may be caused by the displacement of arsenic from sorption sites on aquifer minerals as a result of competitive (anion) exchange by fertilizer-phosphate, which may leach from soils after excessive use of fertilizer [e.g. Acharyya et al., 1999]. We reject this idea because the waters attain a bicarbonate concentration of at least 200 mg l-1 before phosphorus, arsenic, or iron, are found in significant amounts (Fig. 3). Waters lowest in bicarbonate are the youngest and least evolved, but they would contain most phosphorus (and so arsenic), were phosphorus supplied from surface application of fertilizer. Furthermore, concentrations of phosphorus increase with depth in both Faridpur and Lakshmipur (McArthur unpublished, based on DPHE, 2000). Finally, the areal distribution of phosphorus in aquifer waters [Davies and Exley, 1992, Frisbie et al., 1999] show that areas high in phosphorus are also arsenical; this coincidence implies that, if fertilizer-phosphate promotes arsenic release, the process operates only in some areas of Bangladesh, which seems unlikely. The arguments above suggest that competitive exchange with fertilizer phosphate neither worsens nor causes arsenic pollution. Nevertheless, concentrations of phosphorus in the mg l-1 range are released to groundwater from latrines and from the fermentation of buried peat deposits (see later sections).

Concentrations of arsenic co-vary with those of phosphorus for waters from Lakshmipur, but not for waters from Faridpur (Fig. 4b), suggesting that competitive exchange with phosphate generated in-situ may contribute to arsenic pollution. For reasons given later, we believe this contribution to be small.

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Fig. 3. a) Relation of bicarbonate to a) Fe2+, b) P, and c) As in Bangladesh groundwater. For the significance of line A, see text. Data from Nickson et al. [2000] are open circles, data from DPHE [2000] are from Faridpur (triangles) and Lakshmipur (squares).

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Reduction of FeOOH is common in nature and has been invoked previously to explain the presence of arsenic in anoxic surface waters [Aggett and O'Brien, 1985; Cullen and Reimer, 1989; Belzile and Tessier, 1990; Ahmann et al., 1997] and anoxic ground waters [Matisoff et al., 1982; Cullen and Reimer, 1989; Korte, 1991; Korte and Fernando, 1991; Bhattacharya et al., 1997; Nickson et al., 1998, 2000; refs. therein]. Reduction of FeOOH (stoichiometry in Equation 1)

8FeOOH + CH3COO– + 15H2CO3 → 8Fe2+ + 17HCO3– + 12H2O (1)

is driven by microbial metabolism of organic matter [Chapelle and Lovley, 1992; Nealson, 1997; Lovley, 1997; Banfield et al., 1998; Chapelle, 2000]. That FeOOH reduction is common and intense in GBMD aquifers is shown by several observations. Firstly, high concentrations of

dissolved iron have been reported by DPHE [2000; ≤ 24.8 mg l-1], by Nickson et al. [1998, 2000; ≤ 29.2 mg l-1, and by Safiullah [1998; ≤ 80 mg l-1]. Secondly, at concentrations above about 200 mg l-1 of bicarbonate, iron shows a weak correlation with bicarbonate (line A of Fig. 3a). The relation is not stoichiometric for reduction of FeOOH (Equation 1), but data fall on, or to the right of, line A, the slope of which (molar HCO3/Fe of 13) is within a factor of 2 of that (about 30) given for FeOOH reduction by Chapelle and Lovley [1992]. Samples enriched in bicarbonate relative to line A have possibly derived additional bicarbonate from other redox reactions, calcite dissolution and weathering of mica and feldspar, or have lost iron into precipitated phases. The data of Nickson et al. [2000] show a relation between arsenic and bicarbonate that was interpreted as evidence that arsenic was derived from reduction of FeOOH; arsenic and bicarbonate data of DPHE [2000] do not show such a co-variance (Fig. 3c). Concentrations of iron

and arsenic co-vary in aquifer sediments, with molar ratios of Fe/As (oxalate-extractable) of between 1500 and 6000 [DPHE, 1999] and Fe/As (diagenetically-available) ratios of 1800 [Nickson et al., 1998, 2000]. Nevertheless, concentrations of arsenic and iron do not co-vary in solution (Fig. 4a). This may be because, firstly, arsenic and iron may be sequestered differentially into diagenetic pyrite [ Moore et al., 1988; Rittle et al., 1995] and so not behave conservatively in solution. Secondly, dissolved iron may also be derived from weathering of biotite. Thirdly, the iron/arsenic ratio in dissolving FeOOH is variable. Finally, iron may be removed from solution into vivianite, siderite, or mixed-valency hydroxycarbonates [R. Loeppert, pers comm., 2000].

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Fig. 4. Relation of a) As to Fe and b) As to P, in Bangladesh groundwater for arsenic concentrations below 500µg l-1. Data from DPHE [2000] and Nickson et al. [2000]. Symbols as in Fig. 3.

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4. The Redox Driver The lateral and vertical differences in arsenic concentration in well water (Figs. 1, 2) cannot arise from variations in the abundance of arsenic in aquifer sediments: these are micaceous quartzo-feldspathic sands and are not unusual in their concentrations of arsenic, which are commonly in the range between 1 and 30 mg kg-1 [Nickson et al., 1998, 2000; AAN, 1999; DPHE, 1999]. Arsenic at these concentrations is present as a dispersed element sorbed to dispersed FeOOH. Higher concentrations of arsenic, e.g. 196 ppm of Roy Chowdhury et al., [1999], are

uncommon and occur where (rare) localised pyrite has formed during burial diagenesis and scavenged arsenic from solution [Moore et al., 1988; Rittle et al., 1995; AAN, 1999]. Arsenic in Bangladesh sediments will not be released from FeOOH unless organic matter is present to drive microbial reduction (or release phosphate for competitive exchange), so we postulate that it is the distribution of organic matter, particularly peat, in the aquifer sediments that is the primary control on arsenic pollution. Peat beds are common beneath the Old Meghna Estuarine Floodplain in Greater Comilla [Ahmed et al., 1998], in Sylhet, and in the Gopalganj-Khulna Peat Basins [Reimann, 1993]. Many wells in the area around Faridpur may be screened in waterlogged peat [Safiullah, 1998] and the aquifer in Lakshmipur contains peat [DPHE, 1999]. Peat is often found in geotechnical borings (piston samples), although it is rarely recorded during rotary drilling for water wells because such drilling masks its presence unless the peat is very thick. One indicator of peat is the TOC content of some aquifer sediment; a sample from a depth of 2.1 m at Gopalganj (100 km SW of Dhaka) contained 6% TOC [Nickson et. al., 1998] and sediment from a depth of 75 feet (23m) at Tepakhola (Faridpur) had 7.8% TOC [Safiullah, 1998]. Further indicators of buried peat are the co-variance (Fig. 5) of the concentrations of iron, phosphorus, ammonium, and δ13C of dissolved inorganic carbon (DIC), which suggests all are controlled by a master process, which we take to be the microbial metabolisation of buried peat. Complexing moities derived from fermentation of peat (e.g. short-chain carboxylic acids and methylated amines) will drive redox reactions and ammonium production [Bergman et al., 1999]. Furthermore, methane is common in

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groundwater [Ahmed et al., 1998; Hoque et al., in press], in places in amounts sufficient to impede pumping of groundwater and to provide domestic fuel. Where methanogenesis is not seen directly, the chemical signature of methanogenic-CO2 is visible as low pH (≥ 5.9; DPHE, 2000) and high

pCO2 of 10-0.7 to 10-1.5 atm. We postulate also that decreasing pH with increasing bicarbonate (in Faridpur) and dissolved H4SiO4 (Fig. 6) reflects the reaction of methanogenic-CO2 with carbonates, micas and feldspars in the aquifer. Concentrations of H4SiO4 (≤ 131 mg l-1) approach

saturation values for amorphous silica (195 mg l-1, Parkhurst, 1995) and suggest active weathering is occurring. Values of δ13C (DIC) range up to +10‰ [DPHE, 1999], an upper limit for methanogenic-CO2 [Whiticar, 1999]. In Faridpur wells, values of δ13C (DIC) decrease as the calcium concentration increases (Fig. 7) because methanogenic-CO2 (δ13C of +5 to +10‰) dissolves (and equilibrates with) detrital calcite (δ13C of 0 to -6‰; Quade et al., 1997; Singh et al., 1998) and, possibly, pedogenic calcite, which would be more 13C-depleted (cf. δ13C values to -12‰ in pedogenic carbonates of the Siwalik Group; Quade et al., 1997). Isotopic lightening of ground water may result from oxidation in-situ of δ13C-depleted methane, but the importance of this mechanism cannot be established

with current data. The δ13C (DIC) values of Lakshmipur ground waters scatter and show no trend. That calcite dissolution, and subordinate mica weathering, is an important control on the calcium and magnesium concentration in Bangladesh well water is shown by good correlation between Ca and Mg for many waters (Fig. 8a), the good correlation between Ca and 87Sr/86Sr (Fig. 8b), and an isotopic mixing trend for strontium that defines two end-members with 87Sr/86Sr values of about 0.711 and 0.735 (Fig. 8c). These values are close to those of monsoonal rain in Bangladesh (0.710 to 0.712; Galy et al., 1999) and modern detrital carbonate (≤ 0.735; Quade et al., 1997; Singh et al., 1998). The coliform count of Bangladesh wells [Hoque, 1998] co-varies with ammonium concentrations(Fig. 9). Latrines occur within 2 metres of wells, possibly allowing pollution into

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Fig. 7. Relation of Ca to δ13C (DIC) in Bangladesh groundwater. Symbols as Fig. 3. Arrow shows trend for carbonate dissolution from methanogenic-CO2, M, to detrital/pedogenic carbonate, P.

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Fig. 8. Relation of Ca to a) Mg b) 87Sr/86Sr and c) 87Sr/86Sr to 1/Sr, inBangladesh groundwater. Symbols as Fig. 3. In a) arrow A shows trend for micaweathering and arrow B shows trend for carbonate dissolution.

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wells via insecure casing. This source will also supply phosphorus to groundwater. That another source of ammonium, and so phosphorus, exists is shown by the fact that wells with a faecal coliform counts of zero have ammonium concentrations up to 6.6 mg l-1 (Fig. 9; Hoque, 1998) and

the fact that latrines are found throughout the country, but phosphorus enrichment parallels the distribution of arsenic enrichment and is concentrated mostly in northeast and southeast Bangladesh (Fig. 1). This other source of ammonium and phosphorus must be buried peat. A few wells contain amounts of ammonium and phosphorus that reflect the maximum ratio likely to be found in common wetland vegetation (Fig. 5; molar N/P of 16, Redfield et al., 1963; Bedford et al., 1999) suggesting that both come from this source. At

lower concentrations, molar N/P values « 16 indicate a source of additional phosphorus, which we take to be phosphorus sorbed to FeOOH and released during its reductive dissolution; likely vegetative sources have N/P ratios > 16 [Bedford et al., 1999; Richardson et al., 1999]. From Fig. 5, we estimate that more than 70% of phosphorus comes from reduction of FeOOH at phosphorus concentrations below 3 mg l-1. It seems likely that most arsenic also is derived this way, rather than by competitive exchange with phosphate derived from organic matter. In Hungary, arsenic-polluted wells contain methane, ammonium concentrations between 1 and 5 mg l-1, and often high concentrations of iron [M. Csanady, pers. comm., 2000]. The similarity with Bangladesh ground waters may indicate a common pollution driver - burial and degradation of peat deposits. Haskoning [1981] noted that ammonium was a minor nuisance in deep (> 200m) production wells at Khulna, southwestern Bangladesh, so some deep wells in Bangladesh may be susceptible to arsenic pollution, not because of leakage of polluted water from overlying aquifers, but because in-situ degradation of organic matter drives FeOOH reduction and release of arsenic. The arguments presented above suggest that the areal distribution of arsenic pollution corresponds closely to the areal distribution of buried peat. The geographic distribution of arsenic pollution shows some concordance with the distribution of paludal basins recorded by Goodbred and Kuehl [2000]. Peat deposits are, and were, formed in waterlogged areas, rather than active river-channel deposits, a fact that helps to define today’s areal pattern of pollution. Umitsu [1987, 1993, pers. comm. 1998] proposed that much peatland development occurred in the GMBD during a climatic/sea-level optimum some 5 000 years BP. The high number of polluted wells with depths of 28-45 m may result from their being screened near the depth of this major peat horizon. As peat must have formed at other times, other peat layers, at other depths and of differing ages, might explain why arsenic pollution also peaks at depths of 55, 75, 100, and 130 m (Fig. 2).

0

2

4

6

8

10

0 20 40 60 80

Faecal Coliform Count

NH

4 m

g l-1

Fig. 9. Relation between NH4+ and faecal coliform count in Bangladesh wells.

Data from national survey of Hoque [1998]. Wells with a coliform count of zerocontain up to 6.6 mg l-1 of ammonium.

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5. Implications Arsenic pollution by oxidation of arsenical pyrite is a mechanism that is valid for oxic environments, typically surface waters. It may apply to the subsurface where high-permeability allows polluted surface water access to the subsurface, as in Zimapán, Mexico [Armienta et al., 1997]. It may apply where oxic conditions invade a previously anoxic environment hosting sulfide ore, for example in northeastern Wisconsin [Schreiber et al., 2000], where a commercially prospective sulfide ore-body up to 3 metres thick is exposed to oxic conditions by water-level drawdown and in domestic boreholes. Oxidation of the ore results in pollution of groundwater by high concentrations of arsenic (≤ 15 000 µg l-1), sulfate (≤ 618 mg l-1), iron (≤ 160 mg l-1) and acidity (pH ≥ 2.1) [Schreiber et al., 2000; A. Weissbach pers. comm., 2000]. Where arsenic pollution occurs in most subsurface, and most anoxic, environments, the pyrite oxidation model is inappropriate and a different model is needed. Reduction of FeOOH (invoked before for ground water e.g. Matisoff et al., 1982; Cullen and Reimer, 1989; Korte, 1991; Bhattacharya et al., 1997; Nickson et al., 1998, 2000; refs. therein) will serve in most instances. As the process is generic and not site specific it should be examined (not necessarily accepted) wherever naturally-occurring arsenic pollution occurs in groundwater, such as in Argentina [Nicolli et al., 1989], Taiwan [Chen et al., 1994], China [Wang and Huang, 1994; Sun et al., 2000], Hungary, and the USA [Welch et al., 2000]. It is likely that any fluvial or deltaic basin that has hosted marshland and swamp will be prone to severe arsenic contamination of borehole water. In many areas of the world, agriculture and urbanization occur on lowland coastal plains in a setting similar in type, although not always in scale, to that in Bangladesh. Such areas might be afflicted by arsenic contamination, if not pollution, and it should be looked for. Vulnerable regions include the deltas of the Mekong, Red, Irrawaddy, and Chao Phraya rivers. 6. Conclusions Neither pyrite oxidation, nor competitive exchange of fertilizer-phosphate for sorbed arsenic, cause arsenic pollution of groundwater in the Ganges-Meghna-Brahmaputra deltaic plain. Indeed, pyrite in Bangladesh aquifers is a sink for, not a source of, arsenic. Pollution by arsenic occurs because FeOOH is microbially reduced and releases its sorbed load of arsenic to groundwater. The reduction is driven by microbial metabolism of buried peat deposits. Dissolved phosphorus comes mainly from FeOOH, as it is reductively dissolved, with subordinate amounts being contributed by degradation of human organic waste in latrines and fermentation of buried peat deposits. Dissolved ammonium in the aquifer derives predominantly from microbial fermentation of buried peat deposits, but significant amounts are contributed by unsewered sanitation. Ammonium ion is not, therefore, an infallible indicator of faecal contamination of groundwater. Reduction of FeOOH, and release of sorbed arsenic, serves as a generic model for arsenic contamination of aquifers where waters are anoxic, particularly where organic matter is abundant, e.g. in deltaic or fluvial areas that supported peatland during climatic optimums. Acknowledgements. JMMcA thanks the Friends of UCL for travel funds to visit Bangladesh and West Bengal during the preparation of this paper. We thank Bill Cullen, W. Berry Lyons, R. Loeppert and an anonymous reviewer for constructive reviews that helped us improve the manuscript. The 87Sr/86Sr measurements were made by JMMcA in the Radiogenic Isotope Laboratory at RHUL, which is supported, in part, by the University of London as an intercollegiate facility. PR thanks the Department of Public Health Engineering, Government of Bangladesh, for permission to use its data.

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