WATER RESOURCES RESEARCH, VOL. 37, NO. 1, PAGES 109 –117, JANUARY 2001
Arsenic in groundwater: Testing pollution mechanisms
for sedimentary aquifers in Bangladesh
J. M. McArthur,1 P. Ravenscroft,2 S. Safiulla,3 and M. F. Thirlwall4
Abstract. In 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, while 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
2.
Aquifers ⬍300 m deep (most ⬍100 m) provide Bangladesh
and West Bengal with ⬎90% of its drinking water. The
groundwater contains ⬎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; Department of Public Health Engineering (DPHE), 1999]
(see also Bangladesh international community news home
page at http://www.bicn.com/acic/). 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 or by time.
1
Geological Sciences, University College London, London, England,
United Kingdom.
2
Mott MacDonald International, Dhaka, Bangladesh.
3
Department of Environmental Sciences, Jahangirnagar University,
Dhaka, Bangladesh.
4
Department of Geology, Royal Holloway University of London,
Egham, England, United Kingdom.
Copyright 2001 by the American Geophysical Union.
Paper number 2000WR900270.
0043-1397/01/2000WR900270$09.00
2.1.
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 (Figure 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 floodplains [Ravenscroft, 2000] 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 ⬃120 m below the present level. The main rivers may
have cut down ⬎100 m along the axial courses [Umitsu, 1993]
and formed a broad plain ⬃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 upward 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 (Figure 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 ⬃30 m [Frisbie et al., 1999; Karim et al., 1997; Roy
Chowdhury et al., 1999; Acharyya et al., 1999; Asian Arsenic
Network (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 (Figure 2) and so confirm that the highest percentage of contaminated wells occurs at depths between 28 and
109
110
McARTHUR ET AL.: ARSENIC IN GROUNDWATER
Figure 1. Map of Bangladesh with circled areas showing study areas of DPHE [1999, 2000]. CN, Chapai
Nawabganj; F, Faridpur; L, Lakshmipur. Shading 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 center of
each union, calculated using a fixed radius of 7.5 km, a 1.5 km grid, and 3125 union centers. Unions are
administrative areas. A color version is available from J. M. McArthur.
McARTHUR ET AL.: ARSENIC IN GROUNDWATER
45 m. Hand-dug wells are mostly ⬍5 m deep and are 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.
2.2.
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, namely, Faridpur and Lakshmipur (Figure 1). Analytical methods used to obtain DPHE data are given by DPHE
[1999]. Our 87Sr/86Sr data (Table 1) were obtained on unfiltered, acidified water samples using the method given by
McArthur et al. [1991]. We do not use DPHE data for Nawabganj because those EC and bicarbonate data are suspect (J. M.
McArthur et al., unpublished data, 2000). We use ␦13C data
from DPHE [1999] rather than the modified data by 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 ⬍100 m depth as
the most severe arsenic pollution occurs at these depths (Figure 2). Full data for wells are available in downloadable format
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. Localized
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 milligram per liter range, and hundreds of micrograms
per liter 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], show that most waters are
at close to equilibrium with calcite and dolomite, with saturation indices (SI) 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).
Waters are mostly oversaturated with vivianite (SI mostly ⫹2
Figure 2. Percentage of wells in Bangladesh exceeding specified arsenic concentrations, shown as a function of depth (data
from regional survey [DPHE, 1999]).
111
Table 1. Strontium Isotopic Composition of Well Waters
From Faridpur
Number
S1
S2
S3
S4
S6
S7
S8
S9
S10
S11
S12
S14
S15
DPHE Numbera
Sr, mg L⫺1
BTS208
0.480
0.400
0.330
0.370
0.470
0.360
0.420
0.400
0.300
0.554
0.486
0.497
0.591
BTS243
BTS260
BTS258
BTS206
BTS241
BTS242
BTS214
87
Sr/86Sr
0.72107
0.72103
0.71536
0.71603
0.72228
0.71798
0.72065
0.71903
0.71349
0.72343
0.72333
0.72400
0.72510
a
Department of Public Health Engineering (DPHE) well numbers
are those of DPHE [1999, 2000].
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 overapplication 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 [Public Health Engineering Department (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% (J. M. McArthur, unpublished data, 2000) and 0.06%
[DPHE, 1999]. The presence of pyrite shows that it has not
been oxidized and that it is a sink for, not a source of, arsenic
in Bangladesh groundwater. Were pyrite to be oxidized, 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 ⬍30
mg L⫺1 [DPHE, 1999, 2000]. Finally, arsenic pollution is uncommon in hand-dug wells [DPHE, 1999] which, being shallowest, are the most exposed to atmospheric oxygen and the
ones that would be most polluted were arsenic derived from
pyrite by oxidation.
112
McARTHUR ET AL.: ARSENIC IN GROUNDWATER
worsens nor causes arsenic pollution. Nevertheless, concentrations of phosphorus in the milligram per liter range are released to groundwater from latrines and from the fermentation
of buried peat deposits (see section 4). Concentrations of arsenic covary with those of phosphorus for waters from Lakshmipur (Figure 4b) but not for waters from Faridpur (Figure
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.
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, and references therein]. Reduction of FeOOH,
8FeOOH ⫹ CH3COO⫺ ⫹ 15H2CO3 3
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.
First, high concentrations of dissolved iron have been reported
by DPHE [2000] (ⱕ24.8 mg L⫺1), by Nickson et al. [1998, 2000]
(ⱕ29.2 mg L⫺1), and by Safiullah [1998] (ⱕ80 mg L⫺1). Second, at concentrations above ⬃200 mg L⫺1 of bicarbonate,
iron shows a weak correlation with bicarbonate (line A of
Figure 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
Figure 3. 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, and
data from DPHE [2000] are from Faridpur (triangles) and
Lakshmipur (squares).
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 (Figure 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, phosphorus concentrations increase with depth in both Faridpur and
Lakshmipur (J. M. McArthur (unpublished data, 2000) 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
Figure 4. Relation of (a) As to Fe and (b) As to P, in Bangladesh groundwater for arsenic concentrations below 500 ␮g
L⫺1. Data are from DPHE [2000] and Nickson et al. [2000].
Symbols are as in Figure 3.
McARTHUR ET AL.: ARSENIC IN GROUNDWATER
113
of 2 of that (⬃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 (see
below).
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 covariance (Figure 3c). Concentrations of iron and arsenic covary in
aquifer sediments, with molar ratios of Fe/As (oxalateextractable) 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 covary in solution (Figure 4a). This may be because,
first, arsenic and iron may be sequestered differentially into
diagenetic pyrite [Moore et al., 1988; Rittle et al., 1995] and so
do not behave conservatively in solution. Second, dissolved
iron may also be derived from weathering of biotite. Third, the
iron/arsenic ratio in dissolving FeOOH is variable. Finally, iron
may be removed from solution into vivianite, siderite, or mixed
valency oxides or hydroxycarbonates (R. Loeppert, personal
communication, 2000).
4.
Redox Driver
The lateral and vertical differences in arsenic concentration
in well water (Figures 1 and 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) localized
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 total organic
carbon (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 23 m at Tepakhola (Faridpur municipality) contained
7.8% TOC [Safiullah, 1998].
A further indicator of buried peat is the covariance (Figure
5) of the concentrations of iron, phosphorus, ammonium, and
␦13C of dissolved inorganic carbon (DIC), which suggests all
Figure 5. Relation between ␦13C (dissolved inorganic carbon
(DIC)), P, Fe, and NH4 in Bangladesh groundwater. Data are
from DPHE [1999] for ␦13C (DIC); otherwise, they are from
DPHE [2000]. Symbols are as in Figure 3. In Figure 5b the
large arrow represents N/P ratio of 16 for degrading organic
matter; the small arrow shows departure from this N/P ratio as
reductive dissolution of FeOOH adds additional phosphorus
to groundwater.
are controlled by a master process, which we take to be the
microbial metabolization 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 groundwater [Ahmed et al., 1998; Hoque
et al., 2000], 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 (Figure 6) reflects the reaction of methanogenic CO2 with carbonates, micas, and feldspars in the aquifer.
Concentrations of H4SiO4 (ⱕ131 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)
114
McARTHUR ET AL.: ARSENIC IN GROUNDWATER
That calcite dissolution and subordinate mica weathering
are important controls on the calcium and magnesium concentration in Bangladesh well water is shown by the good correlation between Ca and Mg for many waters (Figure 8a), the
good correlation between Ca and 87Sr/86Sr (Figure 8b), and an
isotopic mixing trend for strontium that defines two endmembers with 87Sr/86Sr values of ⬃0.711 and 0.735 (Figure 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 (ⱕ0.735 [Quade et al., 1997; Singh et al., 1998]).
The coliform count of Bangladesh wells [Hoque, 1998] covaries with ammonium concentrations (Figure 9). Latrines
have been recorded within 2 m of wells, possibly allowing
pollution access to 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 (Figure 9 [Hoque, 1998])
and the fact that latrines are found throughout the country but
phosphorus enrichment parallels the distribution of arsenic
Figure 6. Relation of pH to (a) H4SiO4 and (b) bicarbonate
in Bangladesh groundwater. Data are from DPHE [2000]. Symbols are as in Figure 3.
range up to ⫹10‰ [DPHE, 1999], a common upper limit for
methanogenic CO2 [Whiticar, 1999].
In Faridpur wells, values of ␦13C (DIC) decrease as the
calcium concentration increases (Figure 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 (compare ␦13C 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 13C-depleted methane, but the importance
of this mechanism cannot be established with current data. The
␦13C (DIC) values of Lakshmipur ground waters scatter considerably and show no trend.
Figure 7. Relation of Ca to ␦13C (DIC) in Bangladesh
groundwater. Symbols are as in Figure 3. Arrow shows trend
for carbonate dissolution from methanogenic CO2, M, to detrital/pedogenic carbonate, P.
Figure 8. Relation of Ca to (a) Mg and (b) 87Sr/86Sr and (c)
87
Sr/86Sr to 1/Sr in Bangladesh groundwater. Symbols are as in
Figure 3. In Figure 8a, arrow A shows trend for mica weathering, and arrow B shows trend for carbonate dissolution.
McARTHUR ET AL.: ARSENIC IN GROUNDWATER
enrichment and is concentrated mostly in the northeast and
southeast of Bangladesh (Figure 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 (Figure 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 Figure 5 we estimate
that ⬎70% of phosphorus comes from reduction of FeOOH at
phosphorus concentrations below 3 mg L⫺1. It therefore 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, personal communication,
2000). The similarity with Bangladesh groundwaters may indicate a common pollution driver: burial and degradation of peat
deposits. Haskoning [1981] noted that ammonium was a minor
nuisance in deep (⬎200 m) 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 may drive reduction of FeOOH 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,
personal communication, 1998] proposed that much peatland
development occurred in the GMBD during a climatic/sea
level optimum some 5000 years B.P. 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 (Figure 2).
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 m 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 (ⱕ15 000 ␮g L⫺1), sulfate (⬍618 mg L⫺1), iron (⬍160
mg L⫺1), and acidity (pH ⱖ2.1) [Schreiber et al., 2000; A.
Weissbach, personal communication, 2000].
Where arsenic pollution occurs in most subsurface and most
115
Figure 9. Relation between NH⫹
4 and faecal coliform count
in Bangladesh wells. Data are from national survey by Hoque
[1998]. Zero colliform wells contain up to 6.6 mg L⫺1 of ammonium.
anoxic environments, the pyrite oxidation model is inappropriate, and a different model is needed. Reduction of FeOOH
(invoked before for groundwater [e.g., Matisoff et al., 1982;
Cullen and Reimer, 1989; Korte, 1991; Bhattacharya et al., 1997;
Nickson et al., 1998, 2000, and references 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.,
1996], China [Wang and Huang, 1994; Sun et al., 2000], Hungary, and the United States [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 seen in Bangladesh.
It might be expected that such areas would 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 serve 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.
116
McARTHUR ET AL.: ARSENIC IN GROUNDWATER
Acknowledgments. J.M.M. 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 J.M.M. in
the Radiogenic Isotope Laboratory at RHUL, which is supported, in
part, by the University of London as an intercollegiate facility. P.R.
thanks the Department of Public Health Engineering, Government of
Bangladesh, for permission to use its data.
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