Arsenic in groundwater_ testing pollution mechanisms for sedimentary aquifers in.pdf

7/28/2019 Arsenic in groundwater_ testing pollution mechanisms for sedimentary aquifers in.pdf

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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.

AbstractIn the deltaic plain of the Ganges-Meghna-Brahmaputra rivers, arsenic concentrations in

groundwater commonly exceed regulatory limits (50 g 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. IntroductionAquifers 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 g 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 postulatethat 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 PlainArsenic 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 [Alamet al., 1990]. Aquifers beneath these areas are assigned to the DupiTila 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 elevated

terraces 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

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 deltaicplains of the Ganges, Meghna, and Brahmaputra (in Bangladesh, Jamuna) rivers, concentrations of

arsenic (Fig. 1) commonly exceed the Bangladesh drinking-water standard (50 g 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

[Frisbieet al., 1999;Karimet. al., 1997;Roy Chowdhuryet 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.

0%

25%

50%

75%

100%

0 50 100 150 200 250 300 350

Depth, mbgl

PercentageExceedence

10 g l-1

50 g l-1

250 g l-1

Fig. 2. Percentage of wells in Bangladesh exceeding specified arsenic concentrations, shown as a function of

depth (data from Regional Survey, DPHE 1999).


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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].

Our87

Sr/86

Sr data (Table 1) were obtained on unfiltered,acidified, water samples using the method given in McArthuret

al. [1991]. We do not use DPHE data for Nawabganj because

those EC and bicarbonate data are suspect [McArthur et al.,

unpublished]. We use 13C 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 MechanismsThree 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 andRajagopal, 1996;

Mandal et al.,1998;Roy Chowdhuryet 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 [Bhattacharyaet al., 1997;

Nicksonet 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;Nicksonet al., 1998, 2000]. Measuredsulfur concentrations in aquifer sediments represent both pyritic and organic sulfur but allow upper

limits to be placed on pyrite abundance of 0.3% [Nicksonet 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 Sr87

Sr/86

Sr

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 0.71798S8 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 g 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 onaquifer 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, theprocess 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 generatedin-situ may contribute to arsenic pollution. For reasons given later, we believe this contribution to

be small.

0

10

20

30

40

0 200 400 600 800 1000 1200

HCO3 mg l-1

Fe

mgl-1

a

0

1

2

3

4

5

0 500 1000HCO3 mg l

-1

P

mgl-1

b

11

0

200

400

0 500 1000

HCO3 mg l-1

Asgl-1

c

Fig. 3. a) Relation of bicarbonate to a) Fe

2+, b) P,

and c) As in Bangladesh groundwater. For the

significance of line A, see text. Data from Nickson etal. [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 [Matisoffet 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, ironshows 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 bicarbonaterelative 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 betweenarsenic and bicarbonate that was

interpreted as evidence that

arsenic was derived fromreduction 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

[Nicksonet 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;Rittleet 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 solutioninto vivianite, siderite, or mixed-valency hydroxycarbonates [R. Loeppert,pers comm., 2000].

0

100

200

300

400

500

0 5 10 15 20 25

Fe mg l-1

Asugl-1

a

0

100

200

300

400

500

0 1 2 3 4 5

P mg l-1

As

ugl-1

b

Fig. 4. Relation of a) As to Fe and b) As to P, in Bangladesh

groundwater for arsenic concentrations below 500g l-1

. Data from

DPHE [2000] and Nickson et al. [2000]. Symbols as in Fig. 3.


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4. The Redox DriverThe 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 [Nicksonet al., 1998, 2000;AAN, 1999;DPHE,

1999]. Arsenic at these concentrations is present as a dispersed element sorbed to dispersedFeOOH. 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;Rittleet 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, particularlypeat, 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 13C 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 ammoniumproduction [Bergman et al., 1999].

Furthermore, methane is common in

0

1

2

3

4

5

0 5 10 15 20 25

Fe mg l-1

P

mgl-

1

a

0

1

2

3

4

5

0 5 10 15 20

NH 4-N mg l-1

P

mgl-1

b

-25

-20

-15

-10

-5

0

5

10

0 1 2 3 4

P mg l-1

13

C(DIC)

c

Fig. 5. Relation between 13

C (DIC), P, Fe and NH4 in

Bangladesh groundwater. Data from DPHE [1999] for13

C

(DIC), otherwise DPHE [2000]. Symbols as Fig. 3. In b)

large arrow represents N/P ratio of 16 for degradingorganic matter; small arrow shows departure from this N/P

ratio as reductive dissolution of FeOOH adds additional Pto groundwater.


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groundwater [Ahmedet al., 1998;Hoqueet 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

13C (DIC) range up to +10 [DPHE,

1999], an upper limit for methanogenic-

CO2 [Whiticar, 1999].

In Faridpur wells, values of 13C (DIC)

decrease as the calcium concentration

increases (Fig. 7) because methanogenic-

CO2 (13C of +5 to +10) dissolves (and

equilibrates with) detrital calcite (13C of 0

to -6; Quade et al., 1997; Singh et al.,

1998) and, possibly, pedogenic calcite,

which would be more 13C-depleted (cf.

13C values to -12 in pedogenic

carbonates of the Siwalik Group; Quadeetal., 1997). Isotopic lightening of ground

water may result from oxidation in-situ of

13C-depleted methane, but the importance

of this mechanism cannot be established

with current data. The 13C (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; Quadeet al., 1997; Singhetal., 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

6.0

6.5

7.0

7.5

8.0

0 50 100 150

H4SiO4 mg l-1

pH

a

6.0

6.5

7.0

7.5

8.0

0 250 500 750 1000

HCO3 mg l-1

pH

b

Fig. 6. Relation of pH to a) H4SiO4 and b) bicarbonate in Bangladesh groundwater. Data from DPHE [2000].Symbols as Fig. 3.

Fig. 7. Relation of Ca to 13

C (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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0

20

40

60

80

0 50 100 150 200

Ca mg l-1

Mg

mgl-

1

A

B

a

0.710

0.715

0.720

0.725

0.730

0 50 100 150

Ca mg l-1

87Sr/86Sr

b

monsoonrain

Det / Ped

carbonate

0.710

0.715

0.720

0.725

0.730

1 2 3 4

1 / Sr

87Sr/

86Sr

Det / Ped

carbonate

monsoon

rain

c

Fig. 8. Relation of Ca to a) Mg b)87

Sr/86

Sr and c)87

Sr/86

Sr to 1/Sr, in

Bangladesh groundwater. Symbols as Fig. 3. In a) arrow A shows trend for mica

weathering 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 phosphorusenrichment 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 ofammonium 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 besusceptible 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 todays 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, mightexplain 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

NH4

mgl-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 zero

contain up to 6.6 mg l-1

of ammonium.


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5. ImplicationsArsenic 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 Zimapn, Mexico [Armientaet 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 commerciallyprospective 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 g 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;

Bhattacharyaet 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 [Welchet 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. ConclusionsNeither 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 aquiferswhere 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 Bengalduring 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. The87

Sr/86

Sr 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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