A Review of Proposed Mechanism Leading to as Release_ E.berset

8/6/2019 A Review of Proposed Mechanism Leading to as Release_ E.berset

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Groundwater in Bangladesh: A review of the proposed

Mechanisms leading to Arsenic Release

Estelle Berset

Department of Environmental Sciences

ETH Zurich

Mai 2007

Term Paper in Biogeochemistry and Pollutant Dynamics

Tutor : Prof. Dr. Ruben Kretzschmar


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Groundwater in Bangladesh: A review of the proposed Mechanisms leading to Arsenic Release

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Abstract

During the 1970s, millions of tube wells were installed in Bangladesh in order toprovide pathogen-free drinking water instead of the biologically contaminated surfacewater used before. At the same time, irrigation wells were also drilled across the

country. Today, Bangladesh faces the largest mass poisoning of a population inhistory because groundwater contains high levels of arsenic. Several researchgroups worldwide have formulated various hypotheses on mechanisms suitable toexplain the influence of particular hydrologic and biogeochemical conditions on thearsenic release. This paper presents a review of the current state of research onthese release mechanisms. The geochemical depth profile in Southern Bangladeshtypically consists of a Holocene aquifer with high dissolved arsenic concentrationsand a Pleistocene aquifer containing Fe oxyhydroxides, where dissolved arsenicconcentrations are low. As a first explanation for the arsenic mobilization, the sulfideoxidation hypothesis has been intensively discussed because of the low SO4

2-

concentration, its inverse correlation with the arsenic profile and low dissolved sulfurlevels. According to the Fe oxyhydroxides reduction hypothesis, the onset ofreducing conditions could lead to microbial dissolution of FeOOH coatings, resultingin the release of arsenic. The relevance of this theory for the Holocene aquifer hasbeen highly challenged, because geochemical conditions are all inconsistent withongoing of ferric-iron reduction at the depth where most of the tube wells withdrawwater. As Fe oxyhydroxides seems to be depleted from the Holocene aquifer,adsorption of arsenic is limited to phases which have a weak affinity for arsenic.Consequently, it could be easily liberated and transported during groundwatermovement. Agriculture may affect arsenic contamination in two ways. First, theapplication of P-fertilizer may induce the release of arsenic because phosphate

competes for the same sorption sites. Second, young inorganic carbon brought todepth by recent irrigation pumping may drive arsenic reduction and thereforeincrease its mobility. However, recent investigations pointed to the fact that most ofthe arsenic-rich groundwater is older than groundwater irrigation, implying only aminor relevance of re-infiltrated irrigation water for the arsenic release. Furtherstudies are necessary to investigate the link between the geochemistry of arsenicand the hydrodynamics of the aquifers.

1. Introduction

Bangladesh is a country sitting on eroded Himalayan sediments, transported anddeposited by the Ganges, Brahmaputra, and Meghna Rivers [Fig. 1]. The alluvial soildeposited by these rivers has created some of the most highly fertile plains of theworld. Straddling the Tropic of Cancer, Bangladeshs climate is tropical with a mildwinter from October to March and a hot, humid summer from March to June. A warmand humid monsoon season lasts from June to October and supplies most of thecountry's rainfall. Natural calamities, such as floods, tropical cyclones, tornadoes,and tidal bores occur almost every year, combined with the effects of deforestation,soil degradation, and erosion.


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Historically, surface water sources in Bangladesh have been contaminated withmircroorganisms, causing a significant burden of disease and mortality. Thepopulation suffered from acute gastrointestinal diseases (such as cholera) resultingfrom bacterial contamination of stagnant pond water. During the 1970s tube-wellswere installed to provide pathogen-free drinking water instead of the biologicallycontaminated surface water used before. In the 1980s the private sector was able tosupply and install millions more of them (Smith et al., 2000). As much as 6 to 10million drinking water wells have been installed in Bangladesh until now (Harvey etal., 2002). This transition to well water was readily adopted because of theconvenience of having a water supply in close proximity to homes and the ease ofdrilling in the regions high-yielding aquifers (Harvey et al., 2005). Water-bornediseases have since decreased, but recent incidences of arsenicosis and cancerhave resulted. At the same time as drinking wells were drilled, irrigation wells werealso installed across the country. Groundwater pumping for irrigation greatlyincreased food production enabling Bangladesh to become self-sufficient in food.

Thus, issues of groundwater quality and quantity have become vital for both thesupply of drinking water and the production of food in Bangladesh (Harvey et al.,2005). The wells consist of tubes that are 5 cm in diameter and are inserted into theground at depths of usually less than 200 m. At the time the wells were installed,arsenic was not recognized as a problem in water supplies, and therefore standardwater testing procedures did not include tests for arsenic (Smith et al., 2000).

The use of hazardous groundwater for drinking, cooking, and irrigation in Bangladeshhas led to the largest mass poisoning of a population in history because groundwaterhas been contaminated with naturally occurring inorganic arsenic (Charlet and Polya,2006). Of the 125 million inhabitants of the country, between 28 million and 35 million

are exposed to drinking water in which arsenic exceeds the Bangladesh standard of50 g/L As. If the WHO guideline value is used (10 g/L) these figures increase to 46

Fig. 1: Map of Bangladesh(From www.exportinfo.org)

Fig. 2: Regional trends in groundwaterAs concentrations in shallow wells(From BGS and DPHE, 2001)


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and 57 million people, respectively (BGS and DPHE, 2001). The arseniccontamination of groundwater was confirmed in 1993, when the BangladeshDepartment of Public Health Engineering tested wells in western Bangladesh afterextensive arsenic contamination was discovered in West Bengal nearly a decadebefore (Swartz et al., 2004). While mega-cities can handle the high arsenic content in

groundwater because of centralized adduction systems and treatment plants, smallvillages face an enormous problem because the efficiency of water treatment is hardto control for such a large number of wells (Charlet and Polya, 2006).

The two most common exposure routes of As to humans are ingestion andinhalation. Drinking of As-contaminated groundwater has been suggested as themain risk for humans. While water-borne diseases have decreased, recentincidences of skin lesions (arsenicosis) and cancer have resulted due to arsenicpoisoning (Polizzotto et al., 2006). Humans may ingest arsenic in water from wellsdrilled into arsenic-rich geologic strata, or in water contaminated by industrial oragricultural emissions (Adriano, 2001). Groundwater contamination by geogenic

arsenic is a widespread problem in south and southeast Asia, Bangladesh being themost severely affected country. The health effects of ingesting arsenic-contaminateddrinking water appear slowly. The most common early symptoms are gastrointestinal,including diarrhoea and abdominal pain. Peripheral neuropathy may also occur. Thelatency for arsenic-caused skin lesions is typically about 10 years, although theirrapidity of appearance seems to be dose dependent. The first cases of such skinlesions were identified in 1983. Today over 100000 people in Bangladesh havealready developed skin lesions. Only small numbers of cases of skin cancer havestarted to appear. Since the latency is typical more than 20 years, the fact that only asmall number have been found provides little reassurance about the future incidenceof skin cancer. In other countries, the main causes of death associated with chronicingestion of arsenic in drinking-water are internal cancers (bladder, kidney or lungcancer). It is reasonable to expect marked increases of mortality from internalcancers once the sufficient latency has been reached. In contrast to diseases likemalaria, cholera and tuberculosis, which require a more complex public healthresponse, the response to arsenic contamination is clear: provide arsenic-free water.Unlike other major health problems experienced in Bangladesh, arsenic-causeddiseases can be eradicated at relatively low costs. Although the precise extent of thecontamination is not known, this does not invalidate the need for an emergencyresponse (Smith et al., 2000).

Much of our understanding of the distribution of arsenic across Bangladeshsgroundwater comes from the work of the British Geological Survey. In 1998, thissurvey of 41 districts collected 2022 water samples. 25 % were found to have arsenicconcentrations above 50 g/L, the Bangladesh drinking-water standard. In addition, 9% were above 200 g/L, 1.8 % exceeded 500 g/L and 0.1 % exceeded 1000 g/L.The mean arsenic concentration found in the tube-wells sampled was 55 g/L andthe maximum concentration found was 1670 g/L. There were important differencesbetween shallow wells and deep wells (defined as greater than or equal to 150 mdepth), as well as between samples from recent (Holocene) alluvium and older(Pleistocene) alluvium. Arsenic was essentially confined to groundwaters from theshallow aquifer. Only 9 % of the sampled tube-wells were deep. The results of the

BGS survey showed clear differences in arsenic concentration in different parts ofBangladesh, with the greatest number of high-arsenic wells in the south and the


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south-east of the country [Fig. 2]. There is very pronounced short-range variation inthe arsenic concentration from well to well, which makes predicting the concentrationof arsenic in groundwater from unsampled wells difficult, even when theconcentrations in adjacent wells are known. Even in areas of generally low arsenicconcentrations, there are occasionally hot spots where a cluster of wells with

unusually high concentrations of arsenic are found (BGS and DPHE, 2001). TheBritish Geological Survey contributed to the determination of the extent of the arsenicpoisoning in Bangladesh, however it could not show clear causes for thiscontamination.

There is as yet no consensus amongst scientists about the precise cause of thearsenic problem in Bangladesh, and considerable debate remains about the keycontrols on arsenic concentrations in the groundwaters. Sediments typically containonly small levels of arsenic, and the cause and timing of arsenic mobilization fromthese sediments remain unclear. Today, several research groups worldwide aretrying to determine the influence of particular hydrologic and biogeochemical

conditions on this arsenic release, notably:

the controls on arsenic transfer from sediments to groundwater by sedimentmineralogy and chemistry, hydrology and microbiology

the controls on arsenic release and flushing rates by the nature of thepermeability structure in the host sediments

the relative importance of microbiological and inorganic release processes

the impact of surface-derived organic carbon on arsenic transfer from the hostsediments to the groundwaters and hence the impact of anthropogenicprocesses on arsenic release rates (Charlet and Polya, 2006).

Since arsenic contamination of the groundwater was confirmed in 1993, severalhypotheses on possible mechanisms of arsenic release have been formulated. Someof them point to an anthropogenic origin of contamination, other suggest an entirelynatural release of arsenic in groundwater. This paper is a review of the current stateof research findings on the mechanisms of arsenic release in the groundwater ofBangladesh.

2. Geochemistry of arsenic (As)

The main origin of As in soils is the parent material from which the soil is derived.Sedimentary rocks such as limestone (2.6 ppm) or sandstone (4.1 ppm) containmuch higher As average concentrations than igneous rocks (1.5 ppm). Inuncontaminated soils, As levels seldom exceed 10 ppm. But anthropogenic sourcesof As such as agricultural use, mining and smelting, and coal combustion haveelevated its background to very high levels. Agricultural soils treated with arsenicalscan reach As concentrations of 600 ppm or more (Adriano, 2001).


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Chemical speciationAs occurs both as inorganic [As(III) and As(V)] and as organic forms [Fig. 3]. As(V)(arsenate) and As(III) (arsenite) are the primary As species in soils and naturalwaters. Of these forms, As(III) is the most soluble and mobile, as well as the mosttoxic species. As(III) exists as arsenous oxide (As2O3), arsenious acid (HAsO2),

arsenite ions (H2AsO32-

, HAsO32-

, AsO33-

), arsenic trichloride (AsCl3), arsenic sulfide(AsS3), and arsine (AsH3). As(V) commonly occurs as arsenic pentoxide (As2O5),orthoarsenic acid (H3AsO4), metaarsenic acid (HAsO3), and arsenate ions (H2AsO4

-,HAsO4

2-, AsO43-). Inorganic arsenic in soil and water is subject to biological

transformation, yielding organo-arsenicals and other compounds. It can undergomicrobially mediated transformation, forming Monomethylarsonic acid (MMA),Dimethylarsenic acid (DMA) and Trimethylarsine (TMA). The pathway of As(V)methylation initially involves the reduction of As(V) to As(III), with the subsequentmethylation of As(III) to dimethyl arsine. Methylation is often enhanced by sulphate-reducing bacteria (Adriano, 2001). In reducing environments such as sediments,methanogenic bacteria can reduce As(V) to As(III) and methylate it to methylarsonic

acid(III) or dimethylarsinic acid. These compounds may further methylate or reduce,and volatilize to the atmosphere. These different chemical species influence thesorption affinity of As (Mahimairaja et al., 2005).

In well-aerated soils As(V) generally predominates over As(III), whereas in waters therelative proportion of these two species varies depending on a number of factors,including As sources, redox potential, pH and microbial activity. Alterations in theoxidation state of As, as influenced by redox potential and pH, greatly affect itssolubility in soils. Under anoxic conditions (Eh = -200 mV), soluble As increased 13-fold as compared to oxic conditions (Eh = 500 mV) (Mahimairaja et al., 2005).H3AsO3 is the dominant species under reducing conditions. Seasonal variations of Aswere found depending on the water status of the fields. Under flooding conditions,there is an increase in pH and in the fraction of As(III), whereas under uplandconditions, the pH decreases but the fraction of As(V) increases (Adriano, 2001).

Arsenic in soil can undergo transformations via abiotic or biotic processes. Severalstudies showed oxidation of As(III) to As(V), plausibly through the oxidation of As(III)by Mn(IV) and Fe(III). Thus the Mn(IV) and Fe(III) oxyhydroxides, present on thesurface of mineral and organic compounds and existing as discrete particles, mayplay a significant role in catalyzing the oxidation of As(III) through an electron transfermechanism. As(V) can become immobilized in sediments by coprecipitation with

hydrous Fe and Mn oxyhydroxides. As solubility is controlled by the formation ofMn3(AsO4)2, FeAsO4 and Ca(AsO4). The As release in many natural waters isderived naturally from the dissolution of a mineral phase, the most important primarysource being the sulfide minerals. These minerals oxidize rapidly on exposure to theatmosphere releasing As for partitioning between water and various secondaryminerals. Fe oxides play the most important role in adsorbing As species. Mn oxidesplay a role in the oxidation of As(III) to As(V) and also adsorb significant quantitiesalthough to a much lesser degree than the Fe oxides. The mobility of As can belimited by its coprecipitation with secondary sulfide minerals, and more generally byclays. Organic As species tend to be less strongly sorbed by minerals than inorganicspecies (Plant et al., 2004). Under the redox conditions that occur in aquatic systems,

As is stable in four oxidation states: +V, +III, 0, -III (Adriano, 2001).


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Fig. 3: Simplified transformation pathways of As in the environment (Adriano, 2001)

As sorption in soilThe fate of As depends on its speciation. Maybe the first reaction to occur in soils, isAs adsorption onto soil particles. Formation of inner-sphere complexes of As(V) withsoil constituents has been postulated as the main mechanism for the sorption. The

adsorption affinity of As species by sediments decreases in the following order: As(V)> As(III) > As(II) > DMA (Mahimairaja et al., 2005). The As sorbing inorganicconstituents of soils include Fe and Al oxides, clay minerals, and calcite. The extentof adsorption depends on several geochemical factors such as particle size, organicmatter content, nature of constituent minerals, pH, redox potential, and competingions (i.e., Cl-, NO3

-, SO42-, and HPO4

2-) (Adriano, 2001). In general, adsorption ofAs(V) decreases with increasing pH. This decrease is attributed to two interactingfactors: the increasing negative surface potential at the mineral-water interface andthe increasing concentration of negatively charged As(V) species present in the soilsolution. In contrast, adsorption of As(III) increases with increasing pH. As(V)sorption on amorphous Al and Fe oxides is characterized by an apparent sorptionmaximum at pH 4, whereas As(III) sorption maximum occurs in the pH range of 7 to8.5. At acidic pH, the soil components contain large amounts of positive charge, andadsorption of As(V) may become important (Mahimairaja et al., 2005). In acidic soils,Mn oxides play a dominant role in the adsorption of As. Arsenic mobility andbioavailability are greater in sandy than in clayey soils. Soils having higher silicateclay content retain more As than sandy soil with low clay content. The main reasonfor this phenomenon is that both hydrous Fe and Al oxides vary directly with the claycontent of the soil. The silicate clay minerals also generally adsorb more As(V) thanAs(III) (Mahimairaja et al., 2005).


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3. Case study of Bangladesh

3.1. Geochemical profiles with depth

The typical depth profile of dissolved As and other hydrogeochemical parameters insouthern Bangladesh was described by several research teams (BGS 2001, Harveyet al. 2002, Klump et al. 2006, Polizzotto et al. 2006, Swartz et al. 2006, van Geen etal. 2006). Fig. 4 shows a geochemical depth profile (recorded in the Munshiganjdistrict, 30 km south of Dhaka), which can be divided into four layers: (1) surficial claylayer (3 m thick, low permeability), (2) Holocene aquifer (100 m thick, gray sand), (3)aquitard of marine clay (40 m thick) and (4) Pleistocene aquifer (deep, light brown toorange sands). The Holocene aquifer, which is used for local drinking water suppliesand for irrigation, is hydraulically disconnected from the Pleistocene aquifer. It hashigh dissolved As concentrations, whereas the Pleistocene aquifer presumablycontains Fe oxihydroxides where dissolved As concentrations are low [Fig. 4]. Theaquifers are ~85 % quartz and feldspar by weight, with the remainder composed of

biotite, hornblende, magnetite and other minerals. The distribution of As among thesolid phase is similar throughout the deep (Pleistocene) and shallow (Holocene)sediments, with higher concentration in the clayish aquitards [Fig. 4B]. The depth of30 to 40 m is characterised by a peak of dissolved As concentration of more than 600g/L [> 90% As(III)] [Fig. 4A].

Fig. 4: Vertical profiles of subsurface geochemical characteristics. (A) Dissolved As levels peak. (B)Arsenic dissolved by the sequence of extractions that target different solid phases. (C) Dissolved Asand sulfate. (D) Dissolved As, ammonium and calcium. (E) DIC, DOC and methane.(Harvey et al., 2002)

Klump et al. (2006) analysed the vertical profile of electrical conductivity and 18O[Fig.5]. The Holocene aquifer down to a depth of 20-30 m is characterized by low

electrical conductivity and low 18O values. On the contrary, the Pleistocene showshigh electrical conductivity and high 18O values. Looking at the Holocene aquifer,


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both tracer profiles suggest the existence of two groundwater bodies mixing at 30 mdepth, where the As peak was found. Interestingly, many tube wells withdraw waterat the same depth.

Fig. 5: Depth profiles of As, electrical conductivity, 18

O,3H,

3He/

4He and

4Heter (Klump et al., 2006)

The As contamination in Bangladesh attracted the attention of several researchgroups who formulated different hypotheses on possible mechanisms of As

mobilization. In the following sections, the As release mechanisms proposed in theliterature are presented and discussed.

3.2. The sulfide oxidation hypothesis

The As-affected area of Bangladesh is in the region where sediment deposition tookplace during the quaternary period, i.e. 25000 to 80000 years ago (Younger Deltaicdeposition). This sediment contains As-rich pyrite (FeAsS) and covers almost theentire alluvial region of the river Ganges. The sulfide oxidation hypothesis proposesthat, due to high groundwater withdrawal rates coming from irrigation by shallow anddeep tube wells, the O2 entering the aquifer leads to oxidation of As-rich pyritefollowed by As mobilization from the vadose zone (Adriano, 2001). Insoluble As-

bearing minerals such as arsenopyrite are rapidly oxidized by exposure toatmosphere, releasing soluble As(III), SO4

2- (sulfate) and ferrous iron [Fe(II)] [Eq. (1)].The dissolution of these As-containing minerals is dependent on the availability of O2and the rate of oxidation of sulfide. The released As(III) is partially oxidized to As(V)by microbially mediated reactions (Mahimairaja et al., 2005).

FeAsS + 13 Fe3+ + 8 H2O 14 Fe2+ + SO4

2- + 13 H+ + H3AsO4(aq) (1)

Harvey et al. (2002) argued that As cannot be directly mobilized from sulfideminerals, because of the low concentration of SO4

2- and the negative correlationbetween As and SO4

2- [Fig. 4C]. Instead, low dissolved sulfur levels appear to limitthe precipitation of sulfides near the As peak. However, it remains possible thatsulfide minerals are oxidized at the land surface and dissolved by liberating As.


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Polizotto et al. (2005) found As-bearing sulfides representing the largest solid-phaseAs-fraction within the Holocene sediments, accounting for up to 60% of the total As.Pyrite was also observed in Holocene sediments, as well as acid volatile sulfides.

3.3. The Fe oxyhydroxides reduction hypothesis

Another theory says that As is released from aquifer sediments during microbialreductive dissolution of Fe oxyhydroxides. Dissolved As is presumably low in thePleistocene sediments because of the capacity of Fe oxyhydroxides to adsorb it(Harvey et al., 2005). Under oxidizing conditions, and in the presence of Fe, inorganicAs-species are predominantly retained in the solid phase through interaction withFeOOH coatings on soil particles. The onset of reducing conditions resulting fromorganic carbon oxidation could lead to the microbial dissolution of FeOOH coatings.The microbial reduction of FeOOH results in the release of Fe(II), As(III), and As(V)present on such coatings [Eq. 2] (Mahimairaja et al., 2005).

8 FeOOH-As(sorbed) + CH3COOH + 14 H2CO3 8 Fe2+ + As(dissolved) (2)

+ 16 HCO3- + 12 H2O

This hypothesis has been considered by numerous authors who performed detailedinvestigations in order to find out if the biogeochemical conditions in both aquifersmay support the presence of Fe oxyhydroxides. Despite the fact that in the solidphases As and Fe appear to be associated [Fig. 4B], Harvey et al. (2005) found noFe oxyhydroxides in the grey sediment of the Holocene aquifer. They argued thathigh concentrations of CH4 and H2 in strongly reducing water indicate thatgeochemical conditions are not conducive to the stability of Fe oxyhydroxides.Polizzotto et al. (2005) added that proxies for active bacterial metabolism,

concentration of dissolved electron acceptors (e.g. SO42-

) and their products (e.g.CH4) [Fig. 4E], and molecular H2, are all inconsistent with ongoing ferric-ironreduction at well depths of 30 to 40 m. The high concentrations of CH 4 and H2 inporewaters indicate that Fe(III) reduction is no longer occurring at well-depths, andonly methanogenesis is likely to occur. The analysis of Fe speciation by Polizzotto etal. (2006) showed that the Holocene aquifer is dominated by Fe-bearing silicatessuch as hornblende, biotite and illite, but significant proportions of Fe oxyhydroxideswere not reported. Due to the abundance of Fe oxyhydroxides and the lowconcentrations of Fe(II) in the Pleistocene aquifer, it seems likely that As isimmobilized by adsorption to Fe oxyhydroxides. Because reactive Fe(III)oxyhydroxides are predominantly depleted from the Holocene aquifer, adsorption of

aqueous As is limited to phases such as the silicate and carbonate minerals, whichhave a low surface area and a relatively weak affinity for As. So, As is more reactivethan would be expected if it were bound to Fe oxyhydroxides. Consequently, theweakly adsorbed phase of As within the Holocene aquifer can be easily liberated andtransported during groundwater movement. This may indicate why groundwater Asconcentrations are high (Polizzotto et al., 2005). Van Geen (2006) found theHolocene aquifer to contain a higher proportion of dissolved Fe(II) and to betherefore more reduced than the Pleistocene aquifer dominated by Fe(III). Thepresence of Fe(III) oxyhydroxides seems to favour the absorption of As. As noindication of reactive Fe(III) oxyhydroxides was found in the sediments, it can beassumed that Fe oxyhydroxides reduction is not a relevant mechanism for As

mobilization at well-depth. However, having found similar reducing conditions, levelsof reactive solid phase As, and concentrations of phosphate and silicate in the


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Holocene and in the Pleistocene aquifer, Swartz et al. (2004) suggested that thereductive dissolution of Fe(III) oxyhydroxides should be taken in consideration forpossible future effects on As mobility in the Pleistocene aquifer as well.

3.4. The agricultural impact on As mobilization

Agriculture may affect As contamination of the groundwater in two different ways.First, the addition of P-Fertilizer to As contaminated soil is known to increase Assolubility and mobility. Second, some evidence suggests that reinfiltrating irrigationwater increases As concentration in the aquifer.

Because of their geochemical similarity, arsenate (AsO43-) and phosphate (PO4

3) areassumed to react similarly in soil, forming insoluble compounds with Al, Fe, and Ca.Both arsenate and phosphate have comparable dissociation constants for their acidsand solubility products for their salts. Therefore, phosphate substantially suppressesAs sorption by the soil, because H2AsO4

- and H2PO4- ions compete for the same

sorption sites. Among the competing anions, H2PO4- suppresses As(V) sorption by

soil more significantly than Cl- (chloride), NO3- (nitrate) and SO42- (sulfate)(Mahimairaja et al., 2005). Because of the competiting nature of phosphate witharsenate, application of P-fertilizer on the surface of a lead-arsenate contaminatedsoil column significantly increases the amount of As leached or solubilized from thesoil (Adriano, 2001). Ion displacement by fertilizer-derived phosphate may result inthe partitioning of As in the aqueous phase. Phosphate also effectively competeswith As(III) sorption. Compared to the Fe(II) depth profile, Geen et al. (2006)observed that the depth concentration of P-extractable As better correlated withgroundwater As concentrations. Significant concentrations of P-extractable As wereonly found at sites where dissolved As concentrations are particularly high.

Harvey at al. (2002) characterized aqueous and solid phases of a depth profile, toshow that bioavailable dissolved organic carbon (DOC), brought to depth by recentirrigation pumping, may play a role in As mobilization. Arsenic mobility may be drivenby the reducing conditions induced by DOC, which concentration increases withincreasing As concentration to the peak depth [Fig. 4E]. The oxidation of DOC,producing NH4

+ (ammonium) and CO2 (with the latter inducing calcite dissolution),may explain the correlation of the dissolved NH4

+ and Ca2+ profiles with the dissolvedAs profile throughout the shallow aquifer [Fig. 4D]. The determination of theradiocarbon ages for DOC and DIC revealed that DOC is much older (3000 to 5000years) than DIC (about 700 years at 31 m depth) in the same water. The similar

profiles of DIC and DOC [Fig. 4E] may indicate the mobilization of old DOC (from Feoxyhydroxides) through geochemical perturbations caused by the invasion of youngwater transporting young DIC. The similar radiocarbon ages of CH4 and DIC suggestthat CH4 is derived from the DIC pool. The source of DOC responsible may besurface water (permanent ponds and rivers), which receive large inputs of organicwaste, sediments beneath these recharge areas, or irrigated rice fields. The fact thataquifer water levels rise before the monsoon confirms that surface water bodiescontaining high concentrations of DOC recharge the aquifer (Harvey et al., 2002).Polizzotto et al. (2005) add that As is also released from the Holocene sedimentsolids by simple addition of water, and thus labile DOC would not be necessary toinvoke rapid desorption of As. Van Geen et al. (2006) contradict the proposition of

Harvey et al. (2002) with their conclusion that local recharge supplies an excess of


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oxidants. This means that the oxidizing capacity of recharge water, initially containingO2 and NO3

-, exceeds the reducing capacity of DOC.

Comparing water residence times in the Holocene aquifer, Polizzotto et al. (2005)drew the conclusion that the groundwater residence time of ~80 yr without agriculture

decreased to < 40 yr because of irrigation. So, this increased groundwater flow mayexplain the current As contamination of groundwater, but only if the past groundwaterresidence times were sufficiently low to prevent any As-flushing. If As concentrationsremain elevated, there must be an upstream source. Therefore Polizzotto et al.(2006) propose that, contrary to general beliefs, As may not be released in placewithin the aquifer, but in the near-surface zone where variable redox conditions exist.Higher As concentrations in soils are most probably due to a combination of irrigationusing groundwater and annual deposition, bringing additional solid-phase As into thesoil system. Important depositional sources of As include As associated with Feoxyhydroxides and detrital sulfide minerals at the near surface (within the upper 5 m).Due to monsoons and extensive groundwater removal for dry-season irrigation, soils

are variably saturated and undergo seasonal redox cycles. The wet season ischaracterized by a reduction period, whereas the dry season indicates an oxidationperiod. As-bearing sulfides would yield a source of As during oxic cycles, whichwould then be subject to reductive dissolution during monsoonal and flooding events[Fig. 6]. During the oxidation period, Fe(III) oxyhydroxides would be formed; As in thesulfide matrix would undergo a transformation and be adsorbed to and coprecipitatedwith Fe oxyhydroxides. Arsenic released from Fe(III) minerals during the subsequentreducing period could be re-sequestered into authigenic sulfide minerals. If surfacesediments are the only source of As to groundwater, varying mechanisms of Asrelease may appear to be occurring in different locations: sulfide oxidation, sorptioncompetition by fertilizer-derived phosphate and microbial reductive processes(Polizzotto et al., 2006).

Fig. 6: Cycling and transport of As (Polizzotto et al., 2006)


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Irrigation strongly modified the groundwater flow, so that pumping became a moreimportant groundwater sink than river discharge. Klump et al. (2006) usedenvironmental tracer data to study the effect of groundwater pumping and to assessthe role of reinfiltrated irrigation water in the mobilization of As. Radiogenic noblegases like 3H (tritium), 3Hetri (stable daughter isotope of

3H), 3He and 4Heter (produced

by U and Th decay) are used in hydrology in order to study groundwater transport.3

Hcan be used to distinguish between pre-bomb groundwater (with virtually no 3H),which infiltrated before ~1955, and water containing significant amounts of bomb-derived 3H, indicating infiltration less than 50 years ago. 4Heter and

3H/3He are alsoused to determine groundwater ages. A peak of 3H was found at 20 m depth,resulting from the atmospheric 3H bomb peak [Fig. 5]. The groundwater is virtuallyfree of 3H below a depth of 35 m; this indicates that it was recharged before theatmospheric bomb peak and is at least 55 years old. The 3H/3He dating methodindicates that above 10 m, the groundwater age is low (< 5 years), whereas between15 and 30 m depth, the water ages are almost ~40 years. The youngest shallowwater is probably recharged from the adjacent ponds. 4Heter [Fig. 5] was hardly found

in the shallow groundwater. At a depth of about 30 m, the 4Heter concentrationsincrease, indicating an abrupt increase in groundwater age. These results and the14C data are therefore in line with the interpretation that, at 30 m depth, groundwaterrecharged during the last 5 decades is mixed with much older groundwater frombelow. Making the assumption that As is mobilized by water recharged after irrigationusing groundwater began, requires As-contaminated water to be younger thanirrigation. As most of the As-rich groundwater is older than irrigation usinggroundwater, which underwent a great increase across the country in the 1970s, thecontribution of re-infiltrated irrigation water to the As contamination is probably ofminor importance (Klump et al., 2006).

4. Conclusion

At present, the microbial reductive dissolution of Fe oxyhydroxides is the most widelyaccepted theory to explain the arsenic release from aquifer sediments in Bangladesh.This hypothesis has been supported by a considerable number of findings, i.e. thelow redox potential of the Holocene aquifer, its high dissolved Fe content, injection-withdrawal experiments and several indicators of biological activity (CH4, NH4

+, DIC).However, if the potential release of arsenic from surface sediments and its following

transport to depth could be confirmed, the mechanism of sulfide oxidation would gainagain in importance. Therefore, surface and near-surface biogeochemical processesas well as the associated arsenic, DOC and oxidants transport through the aquifershould be one focus of further research.

Even if the principal geochemical mechanisms of arsenic release are more or lessknown on a conceptual level, the processes resulting in the patchy distribution ofarsenic over the country and the occasionally high concentration hot spots are stillnot understood. Currently, there is no convincing correlation between the presence ofcompeting anions and the spatial pattern of dissolved arsenic. Drastic differences inthe arsenic concentrations from wells as close as 100 m to another may be the result

of water retrieval from separate hydraulic domains. Further studies are necessary toinvestigate the link between the geochemistry of arsenic and the hydrodynamics of


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the aquifers. Such investigations could also provide some information about apossible link between the arsenic release and the mixing of shallow and deepgroundwater due to irrigation wells.

5. References

Adriano, D.C. 2001. Arsenic, p. 219-261, In Trace Elements in TerrestrialEnvironments: Biogeochemistry, Bioavailability and Risks of Metalsed. Springer, New York.

BGS, and DPHE. 2001. Arsenic Contamination of Groundwater in BangladeshWC/00/19, Keyworth.

Charlet, L., and D.A. Polya. 2006. Arsenic in Shallow, Reducing Groundwaters inSouthern Asia: An Environmental Health Disaster. Elements 2:91-96.

Harvey, C.F., C.H. Swartz, A.B.M. Badruzzaman, N. Keon-Blute, W. Yu, M.A. Ali, J.

Jay, R. Beckie, V. Niedan, and D. Brabander. 2005. Groundwater ArsenicContamination on the Ganges Delta: Biogeochemistry, Hydrology, humanPerturbations, and human Suffering on a large Scale. Comptes RendusGeosciences 337:285-296.

Harvey, C.F., C.H. Swartz, A.B.M. Badruzzaman, N. Keon-Blute, W. Yu, M.A. Ali, J.Jay, R. Beckie, V. Niedan, D. Brabander, P.M. Oates, K.N. Ashfaque, S.Islam, H.F. Hemond, and M.F. Ahmed. 2002. Arsenic Mobility andGroundwater Extraction in Bangladesh 298:1602-1606.

Klump, S., R. Kipfer, O.A. Cirpka, C.F. Harvey, M.S. Brennwald, K.N. Ashfaque,A.B.M. Badruzzaman, S.J. Hug, and D.M. Imboden. 2006. GroundwaterDynamics and Arsenic Mobilization in Bangladesh Assessed Using NobleGases and Tritium. Environmental Science and Technology 40:243-250.

Mahimairaja, S., N.S. Bolan, D.C. Adriano, and B. Robinson. 2005. ArsenicContamination and its Risk Management in Complex Environmental Settings.Advances in Agronomy.

Plant, J.A., D.G. Kinniburgh, P.L. Smedley, F.M. Fordyce, and B.A. Klinck. 2004.Arsenic and Selenium, p. 17-66, InTreatise on Geochemistry, ed. E. Ltd., Vol.9.

Polizzotto, M.L., C.F. Harvey, S.R. Sutton, and S. Fendorf. 2005. Processesconducive to the Release and Transport of Arsenic into Aquifers ofBangladesh. PNAS 102:18819-18823.

Polizzotto, M.L., C.F. Harvey, G. Li, B. Badruzzman, A. Ali, M. Newville, S. Sutton,and S. Fendorf. 2006. Solid-phases and Desorption Processes of Arsenicwithin Bangladesh Sediments. Chemical Geology 228:97-111.

Smith, A.H., E.O. Lingas, and M. Rahman. 2000. Contamination of Drinking-Water byArsenic in Bangladesh: a public Health Emergency. Bulletin of the WorldHealth Organization 78:1093-1103.

Swartz, C.H., N.K. Blute, B. Badruzzman, A. Ali, D. Brabander, J. Jay, J. Besancon,S. Islam, H.F. Hemond, and C.F. Harvey. 2004. Mobility of Arsenic in aBangladesh Aquifer: Inferences from geochemical Profiles, leaching Data, andmineralogical Characterization. Geochimica et Cosmochimica Acta 68:4539-4557.

Front page: Tube well in Bangladesh. The New York Times

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