Spatial, vertical and temporal variation of arsenic in shallow aquifers of the Bengal Basin: Controlling geochemical processes

Chemical Geology 387 (2014) 157–169

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Spatial, vertical and temporal variation of arsenic in shallow aquifers of the Bengal Basin: Controlling geochemical processes Ashis Biswas a,b,⁎, Harald Neidhardt c,2, Amit K. Kundu b, Dipti Halder a,b,1, Debashis Chatterjee b, Zsolt Berner c, Gunnar Jacks a, Prosun Bhattacharya a a KTH-International Groundwater Arsenic Research Group, Department of Sustainable Development, Environmental Science and Engineering, KTH Royal Institute of Technology, Teknikringen 76, SE-100 44 Stockholm, Sweden b Department of Chemistry, University of Kalyani, 741235 Kalyani, West Bengal, India c Institute of Mineralogy and Geochemistry, Karlsruhe Institute of Technology, Adenauerring 20b, D-76131 Karlsruhe, Germany

a r t i c l e

i n f o

Article history: Received 31 March 2014 Received in revised form 19 August 2014 Accepted 21 August 2014 Available online 28 August 2014 Editor: Carla M. Koretsky Keywords: Bengal Basin Arsenic Spatial distribution Vertical distribution Temporal variation Redox cycling

a b s t r a c t A detailed understanding of the geochemical processes that regulate the spatial, temporal and vertical variation of dissolved arsenic (As) in shallow aquifers (b50 m) is a prerequisite for sustainable drinking water management in the Bengal Basin. The present study conducted at Chakdaha Block of the Nadia District, West Bengal, India, combined a high resolution hydrogeochemical monitoring study over 20 months from two sets of piezometers (2 × 5) to the sediment geochemistry at areas with high (average: 146 μg/L, n = 5) and relatively low (average: 53.3 μg/L, n = 10) dissolved As concentrations in groundwater. The determination of the isotopic composition of δ2H and δ18O in groundwater of the two sites indicated the recharge of evaporative surface − water to the aquifer. The concentrations of major aqueous solutes (Ca2+, Mg2+, Na+, K+, HCO− 3 and Cl ) and electrical conductivity were considerably higher in wells at the high As site compared to the low As site. Addition2 18 ally, at the high As site, the major ions, Fe, SO2− 4 , electrical conductivity, δ H and δ O showed markedly greater enrichment in the shallowest part (b24 m) of the aquifer compared to the deeper part, reflecting vertical layering of groundwater composition within the aquifer. The oxidation of pyrites has been attributed to the high rate of mineral dissolution resulting in such greater enrichments in this part of the aquifer. In addition, the anthropogenic input with recharge water possibly increased the concentrations of Cl − in this part of the aquifer. The vertical layering of groundwater was absent in the aquifer at the low As site. The absence of such layering and relatively low major ion concentrations and electrical conductivity could be linked to the enhanced aquifer flushing and decreased water–sediment interactions influenced by local-scale groundwater abstraction. The seasonal variations of As concentrations in groundwater were observed only in the shallowest part of the aquifers (b30 m). Furthermore, the As concentrations in groundwater at the uppermost part of the shallow aquifers (b 21 m) increased continuously over the monitoring period at both sites. This study supports the view that the reductive dissolution sorption reactions in the aquifer sediment enriches As in of Fe oxyhydroxides coupled with competitive PO3− 4 groundwater of the Bengal Basin. However, the additional Fe released by the weathering of silicate minerals, especially biotite, or the precipitation of Fe as secondary mineral phases such as siderite, vivianite and acid volatile sulfides may result in the decoupling of As and Fe enrichment in groundwater. The redox zonation within the aquifer possibly regulates the vertical distribution of As in the groundwater. © 2014 Elsevier B.V. All rights reserved.

1. Introduction

⁎ Corresponding author at: Department of Geological Sciences, University of Saskatchewan, 114 Science Place, Saskatoon, SK S7N 5E2, Canada. Tel.: + 1 306 966 5764; fax: +1 306 966 8593. E-mail addresses: [email protected], [email protected] (A. Biswas). 1 Present address: Department of Geological Sciences, University of Saskatchewan, 114 Science Place, Saskatoon, SK S7N 5E2, Canada. 2 Present address: Department Of Water Resources And Drinking Water, Eawag, Swiss Federal Institute Of Aquatic Science And Technology, Überlandstrasse 133, 8600 Dübendorf, Switzerland.

http://dx.doi.org/10.1016/j.chemgeo.2014.08.022 0009-2541/© 2014 Elsevier B.V. All rights reserved.

The natural occurrences of high arsenic (As) concentrations in shallow groundwater of the Bengal Basin have caused millions of people to be chronically exposed to As through drinking water (Chakraborti et al., 2008; Chatterjee et al., 2010). The extent of As exposure has been termed ‘the largest mass poisoning in human history’ (Smith et al., 2000). The current consensus is that the reductive dissolution of Fe oxyhydroxides coupled to the microbially catalyzed mineralization of organic matter releases As in groundwater of the Bengal Basin (Bhattacharya et al., 1997; Nickson et al., 1998; Harvey et al., 2002;

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Stüben et al., 2003; Ahmed et al., 2004; Akai et al., 2004; Islam et al., 2004; McArthur et al., 2004; Zheng et al., 2004; Nath et al., 2005; Charlet and Polya, 2006; Berg et al., 2008; Nath et al., 2008a). Nevertheless, the source of organic matter in the aquifers is still controversial (Harvey et al., 2002; Rowland et al., 2006; Sengupta et al., 2008; Neumann et al., 2010; Datta et al., 2011; McArthur et al., 2012; Planer-Friedrich et al., 2012; Lawson et al., 2013). Based on both laboratory and field studies, many researchers have also reported the decoupling of As and Fe release to the aqueous phase, which has created debate on the role of Fe oxyhydroxide reduction on the mobilization of As in groundwater (Horneman et al., 2004; van Geen et al., 2004, 2006; Burnol et al., 2007; Burnol and Charlet, 2010). Based on the observations of contrasting concentrations of As at similar levels of dissolved Fe in groundwaters, van Geen et al. (2008) have inferred that the reduction of Fe oxyhydroxides is possibly a necessary condition, but not always sufficient to mobilize As in groundwater. The microbially catalyzed reduction of Fe oxyhydroxides, especially of ferrihydrite, has even been reported to favor As retention through the formation of biogenic secondary phases such as goethite and magnetite, instead of releasing sorbed As into groundwater (Islam et al., 2005; Kocar et al., 2006; Tufano and Fendorf, 2008). An important aspect of As enrichment in shallow aquifers is its spatial distribution. It is reported that the concentration of As in groundwater may vary laterally from unsafe to safe within a range of 10–100 m, which often constrains the identification of safe regions and exact As mobilization mechanisms (van Geen et al., 2003). Attempts have been made to constrain the underlying reason(s) for such heterogeneity, but the outcomes are often contradictory. The studies by Harvey et al. (2005, 2006) have pointed out that because of very flat basinal topography, the 3D groundwater flow system is mostly dominated by the local scale flow and is influenced by massive groundwater withdrawal, with recharge and discharge areas often within the scale of tens of meters. This local scale flow system and the resulting transport of As determine the spatial and vertical heterogeneity of As distribution in the shallow aquifers. It is also proposed that the high rate of groundwater withdrawal can accelerate As mobilization by enhancing the transport of degradable organic matter with recharge water from the surface and thus preserving reducing conditions in the shallow aquifers (Harvey et al., 2002, 2005; Lawson et al., 2013). The study by van Geen et al. (2008) has emphasized the role of different flushing histories of the aquifers, determined by sub-surface geology, as the regulator for regional and local scale spatial distribution of As in groundwater. It is reported that the aquifers beneath the thick aquitard are more reducing because of

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limited recharge and are thus enriched with As. In contrast, the aquifers overlain by permeable sandy soils are regularly supplied with oxygen and other oxidants with the recharge water, limiting the mobilization of As in groundwater (Aziz et al., 2008; Weinman et al., 2008). The electromagnetic conductivity surveys performed by Métral et al. (2008) and Nath et al. (2010) at ~10 km west of the investigated area of the present study have shown the presence of sandy lenses within the surface aquitard and described their role as a passage of focused recharge of oxidant rich water causing spatially variable redox conditions in the underlying aquifer. Stute et al. (2007) have further proposed that the effect of enhanced drawdown of labile carbon from the surface due to massive groundwater extraction on the As mobilization processes in shallow aquifers might also be offset by the increased rate of groundwater flushing. Mukherjee et al. (2008), however, have suggested that the extent of overlapping redox zonation plays a key role in the spatial and vertical distribution of As in groundwater of the Bengal Basin. The seasonal variation of As concentrations in groundwater of the Bengal Basin has also been monitored. Again, the findings often differ regarding the time period, extent of change in As concentrations, depth of the aquifer that experiences maximum variation and the processes that regulate the seasonal cycling of As in aquifer (for review see Chatterjee et al., 1995; BGS and DPHE, 2001; Cheng et al., 2005, 2006; Ravenscroft et al., 2006; Sengupta et al., 2006; van Geen et al., 2007; Dhar et al., 2008; Bhattacharya et al., 2011; Farooq et al., 2011 and Planer-Friedrich et al., 2012). Furthermore, by monitoring As concentrations in shallow groundwater from West Bengal over 10 years, McArthur et al. (2010) have reported all possible scenarios such as decreasing, increasing and no change in As concentrations over time. The seasonal and temporal variation in groundwater As concentrations can have severe consequences with respect to potential As exposure by the people who are drinking presumably safe water. Thus, the processes regulating these variations are important to understand properly (Cheng et al., 2005). The above presented review of the recent findings indicates that despite considerable progress towards the understanding of As mobilization processes in the aquifers of the Bengal Basin, the findings are often contradictory and need further refinement. The objective of the present study was to further develop (or verify) the current understandings of the controlling geochemical processes that may regulate the spatial, vertical and temporal variation of groundwater As concentrations in shallow aquifers of the Bengal Basin. To meet the objective, a high resolution (once per 2 weeks) hydrogeochemical monitoring program spanning 20 months was conducted using two

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Longitude Fig. 1. Study area map showing the geographic location of (a) Chakdaha Block, Nadia, West Bengal in the Bengal Basin (modified from Neidhardt et al., 2013b) and (b) site 1 and site 2 at Chakdaha Block. The drilling locations of Biswas et al. (2014b) are included to display the proximity of the piezometer installation sites to these locations. The satellite images were acquired from Google Earth 6.0.2.

A. Biswas et al. / Chemical Geology 387 (2014) 157–169

sets of piezometer networks (2 × 5) installed at areas with high (average: 146 μg/L, n = 5) and relatively low (average: 53.3 μg/L, n = 10) groundwater As concentrations. Finally, the results of the hydrogeochemical monitoring program have been combined with the results of the geochemical characterization of the aquifer sediments from the two sites. 2. Materials and methods 2.1. Monitoring sites and piezometer installation Detailed information on monitoring site selection and piezometer installation have been provided in our previous publications (Biswas et al., 2011; Neidhardt et al., 2013a,b, 2014; Biswas et al., 2014a). The study was conducted at Chakdaha Block of the Nadia District, West Bengal, India, 60 km north of Kolkata City. Geographically, the study area represents the western stable shelf part of the central floodplain area of the Bengal Basin (Fig. 1a) (Mukherjee et al., 2009). Based on a hydrogeochemical survey over the entire Chakdaha Block, two sites at the villages of Sahispur (Site 1; 23°04′15.5″N, 088°36′33.5″E, altitude: ~10 m above sea level) and Chakudanga (Site 2; 23°04′58″N, 088°38′13″E, altitude: ~10 m above sea level), separated by ~3 km, with high and relatively low groundwater As concentrations, respectively, were selected for piezometer installation (Fig. 1b). At each site, five piezometers (designated as wells A, B, C, D and E) with different screening positions within the shallow aquifer (site 1—A: 12-21 m, B: 22-25 m, C: 26-29 m, D: 3033 m and E: 34-37 m; site—2: A: 12-21 m, B: 24-27 m, C: 30-33 m, D: 36-39 m and E: 42-45 m), were installed over an area of 25 m2, to collect groundwater samples at different depths. A continuous sediment core was collected for geochemical characterization at both sites during drilling of the deepest piezometers (well E). The closest irrigation wells were installed at a distance of 300 m and 75 m from sites 1 and 2, respectively, at a depth of 24 m. The irrigation well at site 2 was operating almost throughout the year with maximum intensity during the dry period for Boro rice cultivation; however, the irrigation well at site 1 was primarily operating during the dry period. At site 1, there was a pond within 1 m of the well installation site, while at site 2 the nearest pond was at a distance of 75 m. 2.2. Sampling and analysis of groundwater, pond water and rain water Groundwater samples were collected at 2-week intervals over a period of 20 months (December 2008 to August 2010). The regular monitoring at both sites was interrupted once in December 2009 when a pumping and an in-situ bio-stimulation experiment was conducted at sites 1 and 2, respectively (results have been presented in Neidhardt et al., 2013a and Neidhardt et al., 2014 respectively). The regular monitoring was started again in January 2010. Groundwater samples were collected for analysis of major cations, trace elements, anions, Fe speciation, As speciation, dissolved organic carbon (DOC), δ2H and δ18O. The piezometric head, pH, electrical conductivity (EC), oxidation reduction potential (ORP), temperature (T) and alkalinity (reported as HCO− 3 ) were measured in the field prior to groundwater sampling. Details of the groundwater sampling technique and preservation strategy are provided in Appendix A. The methodology and quality control for the laboratory analysis of major cations, trace elements, anions, Fe speciation and As speciation in groundwater samples have been reported in Biswas et al. (2014a). The concentration of DOC was determined with a TOC analyzer (Elementar, Hanau). The composition of δ2H and δ18O in groundwater samples was determined by an isotope-ratio mass spectrometer (IRMS, Delta V Advantage, Thermo Scientific) as documented in Neidhardt et al. (2013a). The precision of the analysis was determined by repeated analysis of the VSMOW reference standard (δ18O: ±0.03‰, δ2H: 0.55‰). Pond water samples from five ponds near the two sites were sampled once in 2012. Monthly cumulative rain water samples were also collected

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during the monsoon season in 2012 on the rooftop at the Department of Chemistry, University of Kalyani. The compositions of δ2H and δ18O in pond and rain water samples were analyzed at the Stable Isotope Laboratory of Stockholm University with a Liquid Water Isotope Analyzer (LWIA) following the principle of laser absorption spectroscopy (off-axis integrated cavity output spectroscopy). The precision of the analysis for δ2H and δ18O was 0.6‰ and 0.15‰, respectively. Daily rainfall data during the monitoring period was collected from the nearest (25 km from the study site) available weather monitoring site at the Directorate of Research, Bidhan Chandra Krishi Viswavidyalaya (BCKV), West Bengal, India. 2.3. Geochemical equilibrium speciation modeling The saturation indices (SI = log [IAP × K−1 T ], where IAP and KT represent the ion activity product and equilibrium solubility constant at 25 °C, respectively) of the major mineral phases, which may regulate the groundwater composition, were calculated for the individual groundwater monitoring samples using the geochemical code Visual MINTEQ ver. 3.0 (Gustafsson, 2011). 3. Results 3.1. Aquifer architecture and sediment geochemistry at the monitoring sites Detailed descriptions of lithology, geological evolution, mineralogy, elemental profiles and associations of As to the aquifer sediment at the two sites have been presented in our previous publication by Neidhardt et al. (2013a,b, 2014). Here we are providing only the information necessary to discuss the evolution of the hydrochemistry at the two study sites. Drilling at both sites revealed a typical architecture of an aquifer with As rich groundwater (Fig. A.1), where channel fill aquifer sands of dark gray color were overlain by a silty clay layer of overbank deposit forming the surface aquitard (Biswas et al., 2014b). The surface aquitard extended down to 3.35 m and 3.20 m below ground level (bgl) at site 1 and site 2, respectively. Underneath this aquitard, a continuous aquifer extended down to the drilling depth of 39.2 m at site 1 and 45.5 m at site 2. At site 1, a clay rich layer was encountered at a depth of 39.2 m and drilling was terminated, while no such layer was found at site 2 (Fig. A.1). The sediment geochemistry at both sites was also similar. The major minerals identified were detrital quartz, feldspars, carbonates (calcite and dolomite), mica (muscovite and phlogopite) and chlorite in the sandy sediments and smectite, kaolinite and likely illite in the silty clay sediments. Traces of magnetite, hematite, garnet (almandine), biotite (phlogopite), chloritoid, actinolite and minerals presumably of epidote type were identified as Fe-rich accessory minerals. The concentrations of As, Fe2O3, MnO2 and total organic carbon (TOC) were usually high at the top silty clay layer, especially around the water table fluctuation zone, and low in the aquifer sands (Fig. A.1). The results of sequential extractions (Fig. A.2) indicated that As was mainly bound to amorphous and crystalline Fe oxides in the aquifer sands of both sites. A major portion of this As was readily exchangeable with PO3− 4 . Significant associations of Fe and Mn to the secondary phases of acid volatile sulfides and carbonates were common in the aquifers of the two sites. Although the incorporation of As into these secondary phases was negligible at site 1, secondary phases constituted the second largest pool of As in the aquifer sands of site 2 (Fig. A.2). 3.2. Hydraulics of the aquifer at the monitoring sites The cumulative precipitation between two successive sampling events and the corresponding position of the piezometric head in the wells of the two sites over the monitoring period are shown in Fig. 2. With the exception of well C at site 1, the variation of piezometric head over the monitoring period was very similar in all of the wells at both sites, reflecting that the wells were installed within a single hydro-

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−4.65‰ to −1.34‰ with a median of −3.88‰, respectively, for site 1 and − 35.8‰ to − 3.88‰ with a median of − 29.6‰ and − 4.75‰ to 5.13‰ with a median of −4.41‰, respectively, for site 2. The groundwater samples from both sites are mostly clustered between the LMWL and the pond evaporation line providing the signature of evaporation before recharge (Fig. 3). In addition, an important feature to note from the depth distribution of δ2H and δ18O is that the groundwater samples from wells A and B at site 1 were significantly more enriched in 2H and 18O compared to the other three wells (Fig. 3), possibly reflecting the vertical stratification of two groundwater bodies within the aquifer. However, such vertical stratification of groundwater was absent at site 2, which is reflected by a similar depth distribution of δ2H and δ18O for all wells (Fig. 3). 3.4. Hydrochemistry in the wells of the monitoring sites

Well E

Fig. 2. Cumulative rainfall between two successive sampling events and variation in piezometric head in the wells at sites 1 and 2 over the monitoring period.

stratigraphic unit of the respective sites. In well C at site 1, the variation of piezometric head followed a similar trend to the other wells, but the piezometric head was comparatively higher (Fig. 2). It should be noted that half of the screen of well C was surrounded by the outer casing due to a fault during well installation, which possibly constrained the natural hydraulic equilibrium with the aquifer in this well, causing a difference in piezometric head. At both sites, the piezometric head declined continuously with the progress of the dry season in both years and was lowest in the month of April. The piezometric head increased throughout the monsoon season in 2009 and 2010 and was at maximum in October (Fig. 2). Although the piezometric head responded almost instantaneously to the precipitation events at the early stages of the monsoon season, there was a time lag between peak monsoon and the highest elevation of piezometric head at both sites (Fig. 2). It was further revealed that the maximum piezometric head during the wet period was similar (~1.0 m bgl) at both sites, while the minimum during the dry period was ~0.5 m higher for site 1 compared to site 2 (Fig. 2). It should also be noted that at both sites the minimum piezometric head was ~0.5 m lower during the dry period in 2010 (site 1: ~5.50 m bgl, site 2: ~4.70 m bgl) compared to 2009 (site 1: ~5.00 m bgl, site 2: ~4.30 m bgl) (Fig. 2). 3.3. Distribution of 2H and 18O in rain water, pond water and groundwater of the study areas The values of δ2H and δ18O in the monthly collected rain water samples varied between −24.6‰ to 14.0‰ and −4.38‰ to 0.46‰, respectively. The highest value of δ2H and δ18O represents the sample that was collected at the beginning of the monsoon. The isotopic compositions of rain waters measured in the present study were compiled with the available data of the whole western Bengal Basin, reported in Kumar et al. (2010), Sengupta and Sarkar (2006) and Mukherjee et al. (2007) to construct the local meteoric water line (LMWL) by a linear regression fit (Fig. 3). The resulting LMWL is defined by the equation of δ2H = 7.86 δ18O + 8.42 (R2 = 0.96), which differs very slightly from the lines of δ2H = 8.03 δ18O + 8.27 (Kumar et al., 2010), δ2H = 7.88 δ18O + 8.93 (Sengupta and Sarkar, 2006), and δ2H = 7.20 δ18O + 7.7 (Mukherjee et al., 2007). The LMWL also closely follows the GMWL constructed by Rozanski et al. (1993), but with a signature of slight evaporation (Fig. 3). The values of δ2H and δ18O in the collected pond water samples varied within − 34.1‰ to − 18.0‰ and − 4.54‰ to − 2.03‰, respectively, with the corresponding median value of −28.9‰ and −3.88‰. The linear regression fitting to these data defines the pond evaporation line of the area as δ2H = 6.02 δ18O − 5.87 (R2 = 0.94), indicating that pond water experienced strong evaporation after accumulation (Fig. 3). The values of δ2H and δ18O in the groundwater samples varied within − 32.2‰ to − 3.68‰ with a median of − 25.9‰ and

A statistical synthesis of the time series data for the wells at site 1 and site 2 has been given in Table A.1 and Table A.2, respectively. The depth profiles of different aqueous solutes and EC for the two sites have been constructed by taking the median values of time series data for each parameter (Figs. 4 and 5). 3.4.1. Depth profiles of major ions and electrical conductivity The groundwater samples at the two sites were of Ca–Mg–HCO3 type to Ca–HCO3 type with circumneutral pH (Table A.1 and Table A.2). The depth profiles of major ionic constituents followed the trends of δ2H and δ18O at both sites. At site 1, the concentrations of all the major ions were considerably higher in the shallowest wells (A and B) compared to the deeper wells (C, D and E) (Fig. 4), again reflecting the vertical stratification of groundwater within the aquifer. The vertical variation of concentrations was very low to negligible for most of the constituents in the wells at site 2 (Fig. 4). With the exception of K+ for the deepest three wells of site 2, the concentrations of all other ions were considerably higher in the wells of site 1 compared to site 2 (Fig. 4). The depth profile of EC in groundwater closely followed the distribution patterns of major ions at both sites (Fig. 4). 3.4.2. Depth profiles of As and other redox sensitive species The depth profiles of As and other redox sensitive species in groundwater of the two sites are shown in Fig. 5. At site 1, the As concentration increased with increasing depth, peaked in well C and then decreased again in the deeper part of the aquifer, producing a typical “bell shaped” (Harvey et al., 2005) vertical profile of As distribution within the shallow aquifer (Fig. 5). Interestingly, in well C, the concentration of dissolved As was highest where the concentration of Fe was lowest (Fig. 5). While the speciation of dissolved As and Fe in all the wells showed that As(III) and Fe(II) mostly dominated over As(V) and Fe(III), the prevalence of the lower oxidation state species was considerably less in well C (Fig. 5). The concentration of Mn was highest in well C, showing the opposite ver3− tical trend of Fe (Fig. 5). The concentrations of NH+ 4 and PO4 were relatively higher in the deeper parts of the aquifer (Fig. 5). The concentration of SO2− 4 was considerably higher in wells A and B, while it was close to the detection limit in the deeper parts of the aquifer (Fig. 5). The value of ORP did not change significantly over the depths of the aquifer (Fig. 5). At site 2, the concentration of dissolved As did not show significant variation with depth, besides the relatively low As concentration in well A (Fig. 5). Similar to site 1, the depth distributions of Fe and Mn were opposite to each other (Fig. 5). The speciation of dissolved As and Fe again showed that they were almost exclusively present in the lower oxidation states of As(III) and Fe(II) and there was no variation with depth (Fig. 5). 3− were similar to each other The vertical distributions of NH+ 4 and PO4 and closely followed the trend of As (Fig. 5). The concentration of SO2− 4 was mostly below the detection limit throughout the aquifer (Fig. 5). As at site 1, the value of ORP did not vary significantly over the depth of the aquifer (Fig. 5). At both sites, over the entire depth of the aquifer, the concentration of NO− 3 was mostly below the detection limit.

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Fig. 3. The variation of δ2H and δ18O in the wells at sites 1 and 2 and their comparison to the local meteoric water line (LMWL) and the pond evaporation line. The LMWL has been constructed by linear regression fit to the compiled rain water data set of the present study, Kumar et al. (2010), Mukherjee et al. (2007) and Sengupta and Sarkar (2006). The depth profiles of δ2H and δ18O in the groundwater at both sites have been constructed by taking the median values of time series data in a particular well. Error bars represent the standard deviation (SD) of the corresponding parameter over the monitoring period.

3.4.3. Temporal variation of As and other aqueous solutes over the monitoring period At site 1, the concentration of As varied considerably over the monitoring period in most wells (RSD ≥10%) (Table A.1 and Table A.2) and the extent of variation was substantially greater in the shallowest part of the aquifer (wells A and B), which is directly visible in Fig. 6. In well A, a general trend of continuous increase of As over the monitoring period was observed. However, no seasonal pattern was prominent in the temporal variation (Fig. 6). A high degree of seasonal variation of As concentrations was observed for well B (Fig. 6). From the beginning of monitoring (December 2008) to the end of the dry season (April 2009), As concentrations decreased continuously. In the following period up to November 2009, when the regular monitoring was stopped for the pumping experiment, As concentrations did not change considerably. In January 2010, when the monitoring was resumed, As concentrations increased to the value observed in December 2008. In the following dry period, the change in As concentrations was similar to the previous year. However, unlike the previous year, As concentrations began to increase from the start of the monsoon season, and when monitoring was stopped in July 2010, the As concentration returned to the concentration observed in January 2010 (Fig. 6). Arsenic(III) was always the predominant species and the prevalence did not change markedly over time (Fig. A.3). The seasonal change in PO3− 4 concentrations very closely followed the pattern of As (Fig. 6). However, in contrast to the seasonal increased during the variation of As, the concentrations of Fe and SO2− 4

dry period and decreased during the monsoon season, especially in 2010 (Fig. 6). Arsenic concentrations also varied cyclically in well C in 2010 (Fig. 6). Similar to well B, the As concentrations decreased when monitoring began in January 2010 after the pumping experiment, and eventually reached a minimum at the end of the dry season. Interestingly, in well C, the percentage of As(III) also greatly decreased reaching the value of 12.2% at the end of the dry season (Fig. A.3). The change in Fe concentrations resulted in the lowest concentration of 0.07 mg/L at the end of the dry season, which also strongly resembled the change in As concentrations (Fig. 6). Phosphate also showed similar behavior to As in this well (Fig. 6). No seasonal variation in As concentrations was observed in wells D and E over the monitoring period (Fig. 6). At site 2, a strong seasonal variation in As concentrations was observed in well A only (Fig. 6), where As concentrations varied between 17.6 and 132 μg/L with an RSD of 50.6% (Table A.2). In contrast to the variation observed in wells B and C at site 1, As concentrations in well A at site 2 notably increased during the dry season both in 2009 and 2010. After reaching a maximum, the As concentrations gradually approached the lowest concentration at the early stage of the monsoon season and no significant variation was observed in the following part of the monsoon (Fig. 6). It is worthwhile to mention that while As concentrations increased considerably during the early stage of the bio-stimulation experiment (Neidhardt et al., 2014), concentrations had returned to the baseline value before resuming regular monitoring in January 2010 (Fig. 6), which indicates that the peak in As concentrations during the

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Depth (m bgl)

20 25 30 35 40 45 Fig. 4. Depth profiles of major aqueous ionic constituents and electrical conductivity (EC) in the aquifer at site 1 (filled symbols) and site 2 (open symbols). The depth profiles have been constructed by taking the median values of time series data for each parameter in the wells at both sites. Error bars represent the standard deviation (SD) of the corresponding parameter over the monitoring period.

dry period of 2010 was not due to the bio-stimulation experiment. In this well also, the variation of As concentrations was accompanied by a similar variation of PO3− 4 concentrations (Fig. 6). However, in contrast to the variation of As, Cl− concentrations increased during the monsoon in both did not 2009 and 2010 (Fig. A.3). Other parameters such as Fe and SO2− 4 show characteristic seasonal variation over the monitoring period in this well (Fig. 6). No significant temporal variation in As and other aqueous solutes was observed in other wells at site 2 (Fig. 6 and Fig. A.3), except during the period of the bio-stimulation experiment, when the concentrations increased considerably (Neidhardt et al., 2014). Although As concentrations in all wells had returned to the baseline before resuming the monitoring in January 2010, other parameters, especially Fe, took multiple weeks to return to the baseline value (Fig. 6). It is further worth noting that the composition of δ18O did not change considerably in the wells at either site over the monitoring period, except in well A at site 1, where, like As, a general trend of continuous increase was observed (Fig. 6). 4. Discussion 4.1. Source of recharge to the shallow aquifers The origin of groundwater recharge for the As rich shallow aquifers of the Bengal Basin is still a controversial issue. The different possibilities that have been explored by previous researchers include: i) the recharge of groundwater by evaporation influenced surface water through ponds, wetlands, and agricultural paddy fields, where the rate of recharge is largely controlled by the scale of groundwater abstraction (Harvey et al., 2002, 2005; Neumann et al., 2010; Lawson et al., 2013) and ii) direct meteoric recharge, without having contribution from the evaporated surface water (Sengupta et al., 2008; Datta et al., 2011). By comparing the composition of δ2H and δ18O over the western Bengal Basin, Mukherjee et al. (2007) have suggested that the regional groundwater recharge is mostly

of meteoric origin, although the contribution of evaporated surface water to the recharge of shallow aquifers cannot be ruled out. The compositions of δ2H and δ18O in the monitoring samples from the two sites and their comparison to the LMWL and the pond evaporation line (Fig. 3) indicate that wetlands and constructed ponds do have some role in the groundwater recharge at both study sites. Previous studies have highlighted that such recharge through ponds and wetlands can contribute DOC in the shallowest parts of the aquifers (Harvey et al., 2002; Neumann et al., 2010; Lawson et al., 2013; Mailloux et al., 2013). Although the concentration of DOC in all wells showed high temporal variability at both sites (%RSD N40) (Table A.1 and Table A.2), the median concentration of DOC decreases from well A to well E (Fig. 5), which possibly indicates the import of DOC with recharge water from ponds and wetlands into the aquifer at the study sites. It is worth mentioning that Lawson et al. (2013) have identified the presence of DOC that is younger than the sedimentary organic carbon even at depth N 100 m in areas 10 km west of the present study sites. They have concluded that the drawdown of surface derived organic matter through the ponds and wetlands, due to groundwater extraction, is the origin of this younger DOC in the aquifer. By comparing with the findings of previous studies, it seems that the origin of groundwater recharge to the sub-surface shallow aquifers is site-specific and largely controlled by the local lithology. For example, by conducting a study in an area of Barasat-1 Block of North 24 Parganas, West Bengal, where the surface silty clay layer extends down 9–24 m bgl (McArthur et al., 2004), Sengupta et al. (2008) rejected the possibility of evaporated surface water recharging the aquifer. A similar conclusion is also made by Datta et al. (2011). The study area of Datta et al. (2011) is distributed over six blocks, in the district of Murshidabad, West Bengal. In five of six blocks, the surface silty clay layer extends down to more than 9 m and, consequently, groundwater samples collected from these blocks plotted along the LMWL. Interestingly, in the sixth block, where the surface silty clay layer is limited to just 3 m, a significant number of

A. Biswas et al. / Chemical Geology 387 (2014) 157–169

400 50

200

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% of Fe(II)

Fe (mg/L)

% of As(III)

As (µg/L) 0

163

100 0

5

10 50

75

100

15

Depth (m bgl)

20 25 30 35 40 45

0

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+

3-

Mn (µg/L)

NH 4 (mg/L)

PO 4 (mg/L) 12001.25

2.75

4.25 1.25

2.75

DOC (mg/L) 4.250.00

2.75

5.50

15

Depth (m bgl)

20 25 30 35 40 45

SO 42- (mg/L) 0.0

6.5

ORP (mV) 13.0 -20

80

180

15

Depth (m bgl)

20 25 30 35 40 45 Fig. 5. Depth profiles of dissolved As and other aqueous redox parameters in the aquifer at site 1 (filled symbols) and site 2 (open symbols). The depth profiles have been constructed by taking the median values of time series data for each parameter in the wells at two sites. Error bars represent the standard deviation (SD) of the corresponding parameter over the monitoring period.

the groundwater samples plotted along the local pond evaporation line (Datta et al., 2011), indicating the recharge of evaporative surface water. At both sites of the present study and the study area of Harvey et al. (2002) and Neumann et al. (2010) in Munshiganj District of Bangladesh, the surface silty clay layer extends down to 3 m, where similar to the sixth block of Datta et al. (2011), the signature of recharge from ponds and rice fields is evident. Furthermore, as previously mentioned, the sandy lenses within the surface aquitard may also act as a conduit for evaporative surface water recharge to the shallow aquifers (Métral et al., 2008; Nath et al., 2010). As the lithology varies largely over the Bengal Basin, any basin-scale generalization of the origin of groundwater recharge to the shallow aquifers could be an over simplification. 4.2. Evolution of major ion composition in groundwater Previous studies conducted on a regional scale have suggested that the chemical evolution of groundwater in the aquifers of the Bengal

Basin is primarily controlled by carbonate mineral dissolution, silicate weathering, and to some extent, ion exchange (Dowling et al., 2003; Mukherjee and Fryar, 2008; Biswas et al., 2012). The concentration of Ca2+, Mg2+ and HCO− 3 is mainly controlled by carbonate mineral dissolution, while the weathering of silicate minerals and ion exchange are responsible for Na+ and K+ in groundwater. The bulk mineralogical investigation of the two sites of the present study also indicated the abundance of carbonate and silicate minerals in the aquifer sediment, supporting their role in controlling the major ion composition of groundwater at both sites. As previously mentioned, the concentration of major ions and EC were high in the groundwater of wells A and B and decreased with increasing depth at site 1 (Fig. 4), reflecting that mineral dissolution was extensive in the upper most part of the aquifer. This poses a question regarding which factors are responsible for this high rate of mineral dissolution in the depth of wells A and B. In aquifers, the rate of carbonate and silicate mineral dissolution is controlled by the availability of mineral content in

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Site 2 400

300

300

200

200

100

100

Conc. of Fe (mg/L)

4.0

4.0

2.0

2.0

0.0

0.0

10.0

10.0

8.0

8.0

6.0

6.0

4.0

4.0

2.0

2.0

0.0

0.0

15.0

15.0

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5.0

0.0 0.0

-1.0

-1.0

-2.0

-2.0

-3.0

-3.0

-4.0

-4.0

-5.0

Dec 08

δ18O (VSMOW)

0.0 0.0

Conc. of SO42(mg/L)

Conc. of SO42(mg/L)

6.0

Conc. of Fe (mg/L)

δ18O (VSMOW)

0

6.0

Conc. of PO43(mg/L)

Conc. of PO43(mg/L)

0

Conc. of As (µg/L)

Conc. of As (µg/L)

Site 1 400

-5.0

May 09

Well A

Nov 09

Well B

June 10 Jan 09

Well C

June 09

Well D

Nov 09

June 10

Well E

18 2− Fig. 6. Temporal variation of As, PO3− 4 , Fe, SO4 , and δ O in the wells at sites 1 and 2 over the monitoring period of 20 months (December 2008 to August 2010). The blue shaded area and the white area in the figure represents the monsoon period and the dry period, respectively. Green and red vertical lines represent the time of the pumping experiment at site 1 and the bio-stimulation experiment at site 2, respectively.

the aquifer sediment and the supply of protons (H+) (Appelo and Postma, 2005). The mineralogical investigation indicated an abundance of carbonate and silicate minerals in the aquifer sediments, suggesting that the availability of H+ was the rate limiting process for mineral dissolution. Two different possibilities, the enrichment of PCO2 and pyrite oxidation (Appelo and Postma, 2005), can account for the increased supply of H+ to the depth of wells A and B at site 1. In a reducing aquifer, the enrichment of PCO2 due to the microbially catalyzed degradation of organic carbon is frequently reported (Harvey et al., 2002; Postma et al., 2007). The microbially catalyzed degradation of organic carbon also releases 3− NH+ 4 and PO4 , which are often considered to be bio-indicators of organic carbon degradation in aquifers (McArthur et al., 2001; Bhattacharya et al., 2002; Harvey et al., 2002; Bhattacharyya et al., 2003; McArthur et al., 3− 2004). However, at site 1, the concentrations of NH+ 4 and PO4 were relatively low in wells A and B compared to the deeper wells (Fig. 5). Such decoupling of the enrichment of major ions to that of NH+ 4 and

possibly invalidates the link between organic matter degradation PO3− 4 and enhanced mineral dissolution in the depth of wells A and B. Instead, and Fe in wells A and B the occurrence of high concentrations of SO2− 4 (Fig. 5) supports H+ availability being coupled to pyrite oxidation as follows (Appelo and Postma, 2005): + FeS2 + 7/2 O2 + H2O = Fe2+ + 2SO2− 4 + 2H

Fe2+ + ¼ O2 + 5/2 H2O = Fe(OH)3 + 2H+ Although the mineralogical investigation did not indicate the presence of detrital pyrite in the sediment, it could be formed authigenically during seasonal redox cycling of SO24 − as reported by Cheng et al. (2005). During the dry season with the progressive lowering of the piezometric head, it is possible for an O2 front to enter the aquifer and cause pyrite oxidation, producing SO24 − and H+. Following the monsoonal

A. Biswas et al. / Chemical Geology 387 (2014) 157–169

recharge, when anoxic conditions develop in the aquifer, SO24 − is reduced and authigenic pyrite is precipitated, which can again be oxidized in the next dry season. The time series data for ORP in the groundwater also reflects the transition to oxidizing conditions in the aquifer as the dry season progresses (Fig. A.3). The higher enrichment of Cl− in shallow wells A and B, compared to the deeper wells at site 1 (Fig. 4), discards the possibility that the diffusion of relict sea water into the aquifer from the underlain clay aquitard was the origin of such high Cl− concentrations. Furthermore, if meteoric recharge was the origin of Cl− in groundwater, Cl− should be balanced by the equivalent concentration of Na+. However, the molar ratio of Cl− to Na+ was predominantly N 1 in the groundwater of wells A (range: 0.64–1.82, median: 1.40) and B (0.23–1.99, 1.38), while the ratio was mostly b1 for other wells (C: 0.40–1.03, 0.69; D: 0.25–0.77, 0.65; E: 0.18–1.26, 0.92). We anticipate that the high Cl− concentrations in wells A and B were of anthropogenic origin (Jacks et al., 1999; Rajmohan and Elango, 2006; Biswas et al., 2012; McArthur et al., 2012). The recharge of water either through the pond at the immediate vicinity of the piezometer installation site, which also receives input of domestic waste waters and agricultural run-off from the nearby paddy field, or through the sandy outcrops within the surface aquitard at a nearby site could be the potential source of Cl− in these two wells. This is also consistent with the higher enrichment of 2H and 18O in wells A and B of site 1 (Fig. 3). The higher enrichment of 2H and 18O in wells A and B may also indicate the possibility that evaporation/evapotranspiration could increase the concentration of major ions and EC in these wells over the dry period. However, the temporal variation of δ18O did not show any specific enrichment pattern over the dry periods for these wells (Fig. 6). At site 2, the similar concentration of major ions and EC at different depths of the aquifer could be the result of a relatively faster rate of groundwater recharge at this site, making the aquifer homogeneous with respect to groundwater composition (Fig. 4). As mentioned in Section 2.1, a high yielding irrigation well was installed at a depth of 24 m and located approximately 75 m apart from the well installation site. It is very likely that the monitoring wells were within the influence zone of groundwater abstraction by this irrigation well which increased the rate of vertical recharge and hence decreased the groundwater residence time and water–sediment interactions substantially at this study site (Stute et al., 2007). The absence of any inter-bedded clay layer further allowed the aquifer system to be easily homogenized. However, at present, no groundwater dating is available from the study site to verify this hypothesis. An inter-bedded clay layer was also absent at site 1 (Fig. A.1). Nevertheless, the natural groundwater flow system was mostly unperturbed because of the absence of any high yielding irrigation well in the immediate vicinity of the piezometer nest at site 1, allowing the vertical layering of the groundwater composition in the studied aquifer to be maintained. The higher rate of recharge resulting in decreased water–sediment interactions at site 2 is consistent with the lower enrichment of major ions and EC in the groundwater of site 2 compared to site 1 (Fig. 6). 4.3. Arsenic mobilization and its vertical distribution in groundwater of the shallow aquifer Predominantly very low to negligible NO− 3 concentrations in groundwater over the monitoring period indicates that the overall redox status in the aquifer at both sites was reducing in nature, possibly at the stage of metal oxyhydroxides reduction; however, the redox status could considerably fluctuate at different seasons over the year. The coexistence of high Fe and Mn concentrations in the groundwater indicates the overlapping stages of Mn and Fe oxyhydroxides reduction in the aquifers of the two sites (Mukherjee et al., 2008). The outcome of the geochemical characterizations of the aquifer sediments, which suggested that Fe oxyhydroxides were the predominant hosts of As in the sediment of both sites and the positive association of Fe and As in groundwater of most of the wells at the two sites (Table 1), where

165

Table 1 3− in the Mutual correlation between As and Fe, As and Mn, As and PO3− 4 , and Fe and PO4 wells at site 1 and site 2. Correlations are statistically significant at the level p b 0.05. Site

Well ID

As vs. Fe

As vs. Mn

As vs. PO3− 4

Site 1

A B C D E A B C D E

0.23 −0.75 0.81 0.53 0.82 0.3 0.72 0.38 0.20 0.71

0.25 −0.64 0.28 0.71 0.25 −0.21 0.72 0.46 0.20 0.70

0.66 0.86 0.87 0.82 0.69 0.90 0.82 0.89 0.69 0.55

Site 2

Fe(II) and As(III) were the prevailing species of dissolved Fe and As, respectively (Table A.1 and A.2), support the view that As was primarily mobilized into the groundwater through the reductive dissolution of Fe oxyhydroxides (Bhattacharya et al., 1997; Nickson et al., 1998, 2000; Bhattacharya et al., 2002; Harvey et al., 2002; Stüben et al., 2003; McArthur et al., 2004). The enrichment of NH+ 4 in the groundwater of the two sites (site 1—range: 1.87–4.22 mg/L, median: 2.88 mg/L; site 2—range: 0.54–3.26 mg/L, median: 2.39 mg/L) indicates that the microbially catalyzed degradation of organic matter was closely related to the progress of these redox processes in the aquifers (Nickson et al., 2000; McArthur et al., 2001; Bhattacharya et al., 2002; Bhattacharyya et al., 2003; McArthur et al., 2004). The vertical distribution of As in the groundwater at site 1 could be related to the redox zonation in the aquifer, where the relatively high enrichment of NH+ 4 in the deeper part of the aquifer reflects stronger reducing conditions, while the relatively weaker reducing conditions in the shallowest part of the aquifer is concentrations (Fig. 5). reflected by the presence of high SO2− 4 At site 1, it is worth noting that although the groundwater in wells A and B was comparatively more enriched with dissolved Fe, the concentration of As was relatively low; however, in well C, the concentration of As was highest among the five wells, while the concentration of Fe was the lowest (Fig. 5), reflecting the decoupling of As and Fe release into the groundwater in these wells (Horneman et al., 2004; van Geen et al., 2004, 2006). It is likely that in wells A and B, the Fe oxyhydroxides reduction was not the sole process responsible for Fe enrichment. The weathering of silicate minerals, specifically of biotite-type, coupled to pyrite oxidation (as discussed in the previous section), might be the additional source of dissolved Fe in the groundwater (Seddique et al., 2008). This is also consistent with the results of sequential extractions which indicated that Fe was largely associated with silicate minerals in the aquifer sediments (Fig. A.2). It should be noted that during oxidation of pyrite, some Fe could be immobilized because of oxidation to Fe(III); however, it might also contribute Fe(II) to the groundwater when the oxidation was limited by the availability of O2 in the aquifer (Appelo and Postma, 2005). Thus the measured concentration of Fe in groundwater of wells A and B at site 1 was possibly the net result of Fe oxyhydroxides reduction, pyrite oxidation coupled to silicate weathering and authigenic pyrite precipitation. It seems that the consideration of the presence of Fe(II) in groundwater as the indicator for ongoing Fe oxyhydroxides reduction in the aquifer is not always justified. Previous studies have also accounted for the weathering of silicate minerals as the source of As in groundwater (Dowling et al., 2003; Charlet et al., 2005; Chakraborty et al., 2007; Seddique et al., 2008; Chakraborty et al., 2011). However, as indicated by the sequential extractions, the concentration of As in the crystal lattice of silicate minerals was negligible in the aquifer sediments (Fig. A.2). Therefore, the weathering of silicate minerals did not significantly contribute to the As enrichment in groundwater in our study sites. This provides one possible explanation for the occurrence of comparatively low concentrations of As in the Fe rich groundwater of wells A and B of site 1 (Fig. 5). Although in well C, the high groundwater As concentration

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was accompanied with a low Fe concentration, the variations of As and Fe were mutually correlated very strongly and positively (r = 0.81, n = 33) (Table 1). Despite high concentrations of Mn in this well, the correlation of As with Mn was not significant (r = 0.28, n = 33) (Table 1). This reflects that in well C, As was also mobilized by the reductive dissolution of Fe oxyhydroxides. As reported by Horneman et al. (2004) for the aquifer sediment of Araihazar, Bangladesh, possibly at the depth of well C at site 1 too, the reduction of Fe oxyhydroxides mostly proceeded with the formation of new authigenic Fe(II) or Fe(II/III) phases such as magnetite (identified as Fe accessory minerals in the aquifer sediments), rather than releasing Fe(II) in the groundwater, and resulting in high As/low Fe concentrations. In anoxic groundwater, the concentration of Fe(II) can also be limited by precipitation of secondary minerals such as siderite and vivianite (van Geen et al., 2004; Postma et al., 2007; von Brömssen et al., 2008). The calculated SI for these minerals reveals that their precipitation was thermodynamically favorable in the aquifers of both study sites (Fig. A.4). Furthermore, the extraction with 1 M HCl released the second largest amount of Fe from the sediment samples from both sites (Fig. A.2), indicating the significant incorporation of Fe into carbonate minerals and acid volatile sulfides. The precipitation of siderite and framboidal pyrite in the aquifers of West Bengal has already been reported (Pal et al., 2002; Nath et al., 2008b). Previous studies have reported that the precipitation of these secondary minerals can also immobilize As through adsorption (Islam et al., 2005; Lowers et al., 2007). However, although the leaching of sediments with 1 M HCl released the second largest amount of As for site 2, the amount was very low for site 1 (Fig. A.2). This demonstrates that in site 2, there was a possibility of co-precipitation of As with these minerals, while at site 1, the precipitation of these secondary minerals could be responsible for the decoupling of As and Fe enrichment in the groundwater. Precipitation of these minerals causing the decoupling of As and Fe in the aqueous phase has already been reported both in laboratory and field studies (van Geen et al., 2004; Burnol et al., 2007; Burnol and Charlet, 2010). 4.4. Processes regulating the seasonal variation of As in shallow aquifers The 20 month long monitoring of As concentrations in the groundwater of two piezometer nests reveals that the seasonal variation of As was only confined to the shallowest part of the aquifer at both study sites (Fig. 6). It is worth mentioning that the studies by Cheng et al. (2005), Dhar et al. (2008) and Planer-Friedrich et al. (2012) have also monitored As concentrations at 2–4 week intervals in different groups of wells in Bangladesh and have reported similar temporal variations primarily within the shallow aquifers (b30 m). In order to make a general conclusion on the seasonal variation of As in shallow groundwater of the entire Bengal Basin, the different possibilities explaining the observed temporal variation, suggested by these previous studies were explored at our two study sites. Two different processes can explain the observed seasonal variation of As in well B at site 1; however, it is difficult to determine which, if not both, was the predominant process. The first possibility is that the seasonal cycling of As was closely related to the cycling of SO24 − as discussed by Cheng et al. (2005). In the dry period, pyrite oxidation leads to the formation of Fe oxyhydroxides, which immobilize As through adsorption (Fig. 6). In the following period of the monsoon, the recharge of groundwater gradually reestablishes the anoxic conditions in the aquifer. The development of anoxic conditions causes two processes to occur simultaneously: i) remobilization of As through reductive dissolution of Fe oxyhydroxides and ii) precipitation of reduction. The result of these two siauthigenic pyrite through SO2− 4 multaneous processes is an increase in As concentrations and a decrease concentrations (Fig. 6). The precipitation of authigenic in Fe and SO2− 4 pyrite without sequestering aqueous As has already been reported in laboratory and field studies (Bostick et al., 2004; van Geen et al., 2004). However, it is not clear why reestablishing anoxic conditions

required more time following the monsoonal recharge in 2009 compared to 2010 (Fig. 6). One possibility is the development of stronger oxidizing conditions in the shallowest part of the aquifer in 2009 compared to 2010 due to the flushing of well A by compressed air after installation in November 2008. However, As concentrations in well A did not respond to the stronger oxidizing conditions in 2009. It is worth noting that As concentrations in well A showed a trend of continuous increase over the monitoring period (Fig. 6). This steady increase in As concentrations could be linked to the gradual development of anoxia because of continuous seepage of organic rich evaporative surface water at this aquifer depth, as reflected by similar continuous enrichment of δ18O over the monitoring period (Fig. 6). This gradual development of anoxia possibly overpowered the influence of O2 during the dry season in 2009 and 2010. Additionally, the evaporation/evapotranspiration could also play a role in the continuous increase in As concentrations and δ18O in well A. A similar continuous increase in As concentrations over the monitoring period of 3 years was also observed by Cheng et al. (2005) in a very shallow well (depth 8 m) in Bangladesh. The second plausible explanation for the observed seasonal variation of As concentrations in well B is the scavenging of As by colloidal aggregates formed in the presence of DOC at comparatively higher concentrations of major ions during the dry period. During the monsoon season, As was remobilized because of dispersion of the colloidal aggregates by dilution with recharge water (Planer-Friedrich et al., 2012). Although the present study did not attempt to quantify the amount of As scavenged by the colloidal aggregates, a study by Majumder et al. (2014) found organic colloidal aggregates as a potential scavenger of As in the aquifer of the study area. In well B, the decreasing trend of As with increasing trends for major cations and anions during the dry period (Fig. 6 and Fig. A.3) supports this hypothesis. However, this hypothesis again cannot explain why As concentrations did not increase during the wet period of 2009. Sometimes vertical shifting of groundwater layers has also been implicated for the seasonal variation of As in shallow aquifers (Cheng et al., 2005; Dhar et al., 2008). However, the fact that As concentrations in well B began to increase just at the onset of the monsoon in 2010 (Fig. 6) eliminates the possibility that As was increasing due to an upward shift of the As rich groundwater layer from well C to the screen position of well B. Because in that case, a time delay, which is necessary for a 1 m (the difference between two screen positions) upward shift of the groundwater layer, is expected between the beginning of the monsoon and the increase of As concentrations in well B. Furthermore, a similar increase in As concentrations in well B after the beginning of the monsoon was also expected in 2009. The fluctuation of redox conditions in the aquifer during the dry and wet periods and their control over the seasonal variation of As concentrations was more evident in well C at site 1, where the concentration of total As, abundance of As(III) and concentration of Fe decreased drastically during the dry period and began to increase just at the onset of the monsoon in 2010 (Fig. 6). It should be mentioned that As concentrations in well C did not show a similar seasonal variation in 2009 (Fig. 6). The decreasing trend of As during the dry period in 2010 might be the result of the development of relatively more oxidizing conditions in the aquifer because of greater drawdown (~0.5 m) of piezometric head in this year (Fig. 2). At site 2, the trend of As enrichment during the dry period in well A (Fig. 6) was possibly linked to groundwater abstraction by the nearby irrigation well for Boro rice cultivation. The increased groundwater abstraction might influence the As concentrations either by migrating an As plume from the nearby contaminated site (Charlet et al., 2007) or by perturbing the local redox equilibria through mixing of the organic rich water, and thus triggering As mobilization at this aquifer depth (Harvey et al., 2002). The concentration of As decreased at the onset of the monsoon for both years, possibly because of dilution with the recharge water (Planer-Friedrich et al., 2012). It is also worth noting that the seasonal minimum As concentration during the wet period in July 2009 (30.5 μg/L) and 2010 (45.8 μg/L) in well A was higher compared

A. Biswas et al. / Chemical Geology 387 (2014) 157–169

to the concentration (17.6 μg/L) in first sample taken in January 2009, indicating the net increase in the As concentration with time. Such continuous increase in As concentration suggests that similar to the well A at site 1, the actual As release processes were active at the depth of well A at site 2. mobilization in the aquifer 4.5. Coupling of As and PO3− 4 An interesting observation is that in all wells at both sites, dissolved As concentrations strongly and positively correlated with dissolved PO3− 4 concentrations (Table 1) and seasonal variations in As concentraconcentrations were closely followed by seasonal variations in PO3− 4 tions (Fig. 6). Considering the fact that major fractions of As in the aquifer sediments of both sites were readily exchangeable with PO3− 4 (Fig. A.2), the positive correlation of As and PO3− in groundwater indi4 cates that competitive sorption to the aquifer sediment could also play a role in the As enrichment processes in the aquifers (Manning and Goldberg, 1996; Acharyya et al., 1999; Jain and Loeppert, 2000; Gao and Mucci, 2001; Dixit and Hering, 2003; Zheng et al., 2005). It should be mentioned that in an accompanying investigation, Biswas et al. (2014a) modeled the observed temporal variation of As concentrations in the piezometers at the two sites. They found that in the absence of PO3− 4 , the modeled concentration of As in the aqueous phase decreases by 92.5% and concluded that the reductive dissolution of Fe oxyhydroxides followed by competitive ion exchange with the aquifer sediment is the complete process of As enrichment in groundwater of the Bengal Basin. 5. Summary and conclusion The combination of the high resolution hydrogeochemical monitoring study with the sediment geochemistry at the two sites in the Bengal Basin where the groundwater As concentrations are relatively contrasting indicates that the spatial, vertical and temporal variations of As in shallow groundwater are controlled by the interplay of a number of geochemical and hydrological processes which are often site-specific. The determination of the isotopic composition of δ2H and δ18O in groundwater samples from both study sites and their comparison to the LMWL and the pond evaporation line indicates the recharge of evaporative surface water to the aquifer from ponds and constructed wetlands. The present study also highlights that the origin of recharge to the As rich shallow aquifers is largely determined by the local lithology. The present study further supports the view that the reductive dissolution of Fe oxyhydroxides releases As into groundwater of the reducing aquifers. Additionally, by considering the modeling outcomes of the observed temporal variability of As in these piezometer nests by Biswas et al. (2014a) and the strong positive correlation of dissolved As and in groundwater, together with the fact that the major pool of As PO3− 4 in the aquifer sediments of the two study sites is readily exchangeable with PO34 −, this study suggests that competitive sorption reactions with the aquifer sediment couples to the reductive dissolution of Fe oxyhydroxides to enrich As in groundwater of the Bengal Basin. The different degree of sediment–water interactions and aquifer flushing, which is locally influenced by groundwater abstraction and recharge, can, to some extent, explain the spatial heterogeneity in As distribution in shallow aquifers. The vertical distribution of dissolved As in groundwater is likely the result of redox zonation within the aquifer, where the shallowest part is often weakly reducing to oxidizing and the deeper part is strongly reducing, resulting in relatively higher As concentrations in the deeper parts of the shallow aquifer. This study supports the idea that processes other than reductive dissolution of Fe oxyhydroxides, such as weathering of silicate minerals (especially biotite) may also release Fe into groundwater, resulting in the decoupling of As and Fe enrichment in the aquifer. Additionally, the precipitation of secondary mineral phases such as siderite and vivianite and incomplete reduction of Fe oxyhydroxides leading to the formation of secondary Fe(II) or

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Fe(II/III) phases like magnetite are also supposed to be responsible for such decoupling of As and Fe in shallow groundwater. The hydrogeochemical monitoring indicates that the seasonal variation of As and other aqueous solutes is very much limited to the upper portion of the shallow aquifer only and the types of variation are site-specific. Different processes such as seasonal cycling of redox conditions, seasonal aggregation and dispersion of colloids, which can potentially scavenge dissolved As, local scale groundwater abstraction for irrigation and monsoonal recharge may explain the observed seasonal variations in As concentrations. Acknowledgments We would like to acknowledge the German Research Foundation (DFG) and the German Federal Ministry for Economic Cooperation and Development (BMZ) (Stu 169/37-1) for providing the funding for this research. AB and DH are thankful to the Erasmus Mundus External Cooperation Window (EMECW-Action II) EURINDIA Program for providing them doctoral fellowship to carry out their research. We would also like to acknowledge Mr. Atul Chandra Mandal and Mr. Sadhan Ghosh for ensuring us the unlimited access to their courtyards for piezometer installation and the sampling campaign. We are also thankful to Mr. Jared Robertson at the Department of Geological Sciences, University of Saskatchewan and editor Dr. Carla M Koretsky for their help to revise the English language of the manuscript. The comments by the two reviewers and editor have also developed the article substantially. Appendix A. Supplementary data The aquifer architecture and sediment geochemistry of the two monitoring sites, the groundwater sampling technique and preservation strategy and statistical synthesis of the time series data for the wells at the two sites have been provided in Appendix A. Supplementary data related to this article can be found online at http://dx.doi.org/10. 1016/j.chemgeo.2014.08.022. References Acharyya, S.K., Chakraborty, P., Lahiri, S., Raymahashay, B.C., Guha, S., Bhowmik, A., 1999. Arsenic poisoning in the Ganges delta. Nature 401, 545-545. Ahmed, K.M., Bhattacharya, P., Hasan, M.A., Akhter, S.H., Alam, S.M.M., Bhuyian, M.A.H., 2004. Arsenic contamination in groundwater of alluvial aquifers in Bangladesh: an overview. Appl. Geochem. 19, 181–200. Akai, J., Izumi, K., Fukuhara, H., Masuda, H., Nakano, S., Yoshimura, T., 2004. Minerological and geomicrobiological investigation on groundwater arsenic enrichment in Bangladesh. Appl. Geochem. 19, 215–230. Appelo, C.A.J., Postma, D., 2005. Geochemistry, Groundwater and Pollution. A. A. Balkema, Rotterdam. Aziz, Z., van Geen, A., Versteeg, R., Horneman, A., Zheng, Y., Goodbred Jr., S.L., Steckler, M., Stute, M., Weinman, B., Gavrieli, I., Hoque, A.M., Ahmed, K.M., 2008. Impact of local recharge on arsenic concentrations in shallowaquifers inferred from the electromagnetic conductivity of soils in Araihazar, Bangladesh. Water Resour. Res. 44, W07416. Berg, M., Trang, P.T.K., Stengel, C., Buschmann, J., Viet, P.H., Dan, N.V., Giger, W., Stüben, D., 2008. Hydrological and sedimentary controls leading to arsenic contamination of groundwater in the Hanoi area, Vietnam: the impact of iron-arsenic ratios, peat, river bank deposits and excessive groundwater abstraction. Chem. Geol. 249, 91–112. BGS, DPHE, 2001. Arsenic Contamination of Groundwater in Bangladesh. In: Kinniburgh, D.G., Smedley, P.L. (Eds.), Final report. BGS Tech. Rep. WC/00/19. British Geological Survey, Keyworth (267 pp.). Bhattacharya, P., Chatterjee, D., Jacks, G., 1997. Occurrence of arsenic-contaminated groundwater in alluvial aquifers from delta plains, Eastern India: options for safe drinking water supply. Int. J. Water Resour. Dev. 13, 79–92. Bhattacharya, P., Jacks, G., Ahmed, K.M., Routh, J., Khan, A.A., 2002. Arsenic in groundwater of the Bengal delta plain aquifers in Bangladesh. Bull. Environ. Contam. Toxicol. 69, 538–545. Bhattacharya, P., Hossain, M., Rahman, S.N., Robinson, C., Nath, B., Rahman, M., Islam, M. M., von Brömssen, M., Ahmed, K.M., Chowdhury, D., Rahman, M., Persson, L.A., Vahter, M., 2011. Temporal and seasonal variability of arsenic in drinking water wells in Matlab, Southeastern Bangladesh: a preliminary evaluation on the basis of a 4 year study. J. Environ. Sci. Health A 46, 1177–1184. Bhattacharyya, R., Jana, J., Nath, B., Sahu, S.J., Chatterjee, D., Jacks, G., 2003. Groundwater As mobilization in the Bengal Delta Plain, the use of ferralite as a possible remedial measure—a case study. Appl. Geochem. 18, 1435–1451.

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