Geochemical effects of dissolved organic matter biodegradation on arsenic transport in groundwater systems

Journal of Geochemical Exploration 149 (2015) 8–21

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Geochemical effects of dissolved organic matter biodegradation on arsenic transport in groundwater systems Kunfu Pi a, Yanxin Wang a,⁎, Xianjun Xie a, Shuangbing Huang b, Qian Yu a, Mei Yu a a b

School of Environmental Studies & State Key Laboratory of Biogeology and Environmental Geology, China University of Geosciences, 430074 Wuhan, China Institute of Hydrogeology and Environmental Geology, Chinese Academy of Geological Sciences, 050061 Shijiazhuang, China

a r t i c l e

i n f o

Article history: Received 4 December 2013 Accepted 10 November 2014 Available online 25 November 2014 Keywords: Dissolved organic matter Arsenic Biogeochemistry Sulfate reduction

a b s t r a c t To understand the effects of natural dissolved organic matter (DOM) on microbially catalyzed As mobilization, the chemical properties and transformation of DOM as well as the hydrochemical characteristics of high As groundwater from the Datong Basin of northern China have been studied. Hydrochemical results indicated that As concentration in groundwater ranged from 18.0 to 182 μg/L with an average value of 58.4 μg/L. Dissolved Fe(II), sulfide and dissolved organic carbon (DOC) concentrations were high in the elevated As groundwater. The As contents in both the bulk sediments and extractable fractions show an obvious positive correlation with the Fe contents, demonstrating that As is strongly associated with Fe-bearing minerals via adsorption and/or coprecipitation. The DOM in the shallow groundwater mainly contained terrestrially and microbially derived humic substances, and most of the quinone-like humic substances were in the oxidized state. By contrast, the DOM in the deep groundwater was predominantly microbially derived and had a higher concentration of labile DOM, and the quinone-like humic substances were more reduced. The strong linear correlation between DOC − and HCO− 3 suggests that the high concentration of HCO3 may be related to the microbially mediated oxidation 2− using Fe(III) or SO as electron acceptor in the anoxic groundwater. The relationships of of DOM into HCO− 3 4 As species with Fe(II), DOC concentrations, redox index and SUVA values of DOM indicate that abundant labile DOM may serve as an important electron donor to promote the microbially mediated reductions of Fe(III) and As(V). Furthermore, the quinone-like humic substances act as electron shuttles oxides/hydroxides, SO2− 4 to transfer electrons from labile DOM to Fe(III), SO2− 4 , and As(V) adsorbed on the surfaces of Fe minerals, hence enhancing the mobilization of As (particularly As(III)) into groundwater. However, the observed negative correlations of dissolved sulfide with As and Fe(II) concentrations in groundwater demonstrate that dissolved As is most likely scavenged via co-precipitation with Fe(II) sulfides or As-bearing sulfides. Therefore, reductive dissolution of Fe(III) oxides/hydroxides coupled to microbial oxidation of natural DOM may serve as the primary geochemical process controlling As mobilization in groundwater systems, while the formation of Fe(II) sulfides due to microbially mediated sulfate reduction and DOM oxidization could sequester As from groundwater. © 2014 Elsevier B.V. All rights reserved.

1. Introduction Arsenic-contaminated groundwater has caused severe public health problems worldwide; endemic As poisoning among millions of residents in many countries where As concentrations approach mg/L level, such as Bangladesh, India, Cambodia, Vietnam, the USA and China, has been extensively documented (Berg et al., 2001, 2008; Camacho et al., 2011; Nickson et al., 1998; Polya et al., 2005; Smedley and Kinniburgh, 2002; Wang et al., 2009;). Mechanisms of As mobilization in groundwater systems have been studied by researchers in order to develop better decontamination procedures for water resource management (McArthur et al., 2001; Nickson et al., 2000; Polizzotto et al., 2005; Smedley and Kinniburgh, 2002; Stüben et al., 2003; Zheng et al., ⁎ Corresponding author. Tel.: +86 27 67883998; fax: +86 27 87481030. E-mail address: [email protected] (Y. Wang).

http://dx.doi.org/10.1016/j.gexplo.2014.11.005 0375-6742/© 2014 Elsevier B.V. All rights reserved.

2004). The reductive dissolution of Fe(III) oxides/hydroxides with the concurrent release of As initially bound to Fe minerals has been the most widely accepted mechanism of As mobilization in groundwater systems (Ahmed et al., 2004; Smedley and Kinniburgh, 2002). In recent years, microbially mediated processes have been found to play important roles in As mobilization (Drahota et al., 2013; Islam et al., 2004; Jiang et al., 2013; Lloyd and Oremland, 2006). For instance, Fereducing microbes can accept electrons from natural organic carbon or other organic substances to fuel their respiration, leading to reductive dissolution of Fe(III) oxides/hydroxides and subsequent release of As bound to the Fe(III) oxides/hydroxides (Islam et al., 2004; McArthur et al., 2004; Wang and Mulligan, 2006). Besides, studies revealed that biotic processes could alter the redox conditions of aquifers (Oremland and Stolz, 2003). In the anoxic environment, the biogeochemical cycling of Fe and As release were recognized to be strongly related with the reactivity of natural organic matter (NOM) (Fendorf et al.,

K. Pi et al. / Journal of Geochemical Exploration 149 (2015) 8–21

2010; McArthur et al., 2004; Rowland et al., 2007). For instance, peat dispersed throughout the aquifer sediments may act as a primary energy source for the relevant microbial communities, and the availability of labile NOM can greatly stimulate the indigenous microbial activities and facilitate As release into groundwater (McArthur et al., 2004; Mladenov et al., 2010). Therefore, the presence of NOM exerts significant influence on the microbial activities and becomes an important factor for As mobilization in aquifers (Fendorf et al., 2010; Rowland et al., 2006, 2007). Several aquifer systems with elevated As concentrations around the world have been reported to contain high level of sedimentary organic matter that can be microbially degraded (Mladenov et al., 2010; Reza et al., 2010, 2011; Varsányi and Kovács, 2006; Xie et al., 2012a). Although the impact of sedimentary organic matter on the mobilization of As has been documented in many previous studies (Berg et al., 2001; Harvey et al., 2002; Islam et al., 2004, 2005; Mladenov et al., 2010), the redox interactions between As and natural dissolved organic matter (DOM) in groundwater has been poorly understood (Bauer and Blodau, 2006; Rowland et al., 2007; Wang and Mulligan, 2006). DOM derived from sedimentary organic matter can also fuel the reductive dissolution of Fe(III) oxides/hydroxides via microbial mediation, releasing the bound As into groundwater (Islam et al., 2004; McArthur et al., 2004; Polizzotto et al., 2005; Tufano and Fendorf, 2008). During this process, the redox-active quinone-like humic substances (HS) of DOM may act as electron shuttles between labile DOM and Fe(III) oxides/ hydroxides (Bauer and Kappler, 2009; Jiang and Kappler, 2008; Jiang et al., 2009; Mladenov et al., 2010). Additionally, DOM is able to react directly with As to change the redox state of As (Palmer and von Wandruszka, 2010; Palmer et al., 2006). However, the effect of DOM has been mostly studied in the laboratory. Little work has been done to observe the influence of DOM on As mobilization under natural conditions. The process of Fe(III) reduction and As mobilization in natural scenarios is influenced by some environmental factors, such as the heterogeneity of the quinone-like moieties in DOM and the circumambient bacterial communities (Jiang and Kappler, 2008; Scott et al., 1998). Furthermore, As redox reactions (Jiang et al., 2009; Redman et al., 2002; Sharma et al., 2011) and the formation of Fe(II) sulfides (Burton et al., 2013; Jung et al., 2012; Kirk et al., 2004; Tufano and Fendorf, 2008) can be facilitated by DOM, which has been shown to affect As retention in groundwater. Therefore, despite the recognized impact of DOM on As mobilization, numerous questions remain to be answered regarding the specific geochemical processes, particularly the transformation of DOM during its interactions with As-bearing Fe minerals and SO2− 4 in groundwater systems. The Datong Basin is a typical Cenozoic sedimentary basin located in the arid/semi-arid region of northern China (Wang et al., 2009; Xie et al., 2012b). The basin contains Pliocene to Pleistocene and Holocene unconsolidated sediments with the thicknesses between 50 and 2500 m. The Quaternary sediments mainly consist of brown to black alluvial-pluvial, lacustrine, alluvial-lacustrine fine-grain sand, sandy silt, silt and silty clay which are commonly rich in sedimentary organic matter and As (Xie et al., 2012a). Similarly, our previous studies conducted in this area indicate that naturally occurring As is spatially associated with high sedimentary organic carbon levels (0.2 to 1.6 wt%), and biodegradation of organic matter in the sediments may promote the reduction of Fe(III) oxides/hydroxides and release of As into groundwater (Xie et al., 2012a, 2013a). Shallow groundwater (3 to 60 m below land surface) occurs in the Quaternary alluvial-pluvial and alluvial-lacustrine aquifers where high As concentrations have been commonly detected (Xie et al., 2008). The major hydrochemical type is Na-HCO3/NaHCO3-Cl, and the groundwater contains high levels of dissolved organic carbon (DOC) (from several 10 mg/L up to 100 mg/L), Fe(II) and hydrogen sulfide under reducing conditions. The interactions between Fe minerals and DOM are anticipated to have a powerful effect on the redox conditions and biogeochemical processes related to As enrichment in the aquifers. Understanding the chemical properties and

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transformation of DOM can reveal the dominant process controlling As transport. However, no systematic work has been conducted to characterize DOM in the high As groundwater systems and evaluate its effect on As mobilization at Datong Basin. Therefore, in this study, field investigation and laboratory experiments were carried out at the monitoring site of As-contaminated aquifers in central Datong Basin in order to: (1) characterize the chemical properties and distribution of DOM in the groundwater, (2) identify the relationship between DOM and the redox-active components of Fe, S and As in the groundwater and (3) elucidate the possible influence of DOM on the biogeochemical processes concerning As mobilization. 2. Material and methods 2.1. Sampling The sediment and groundwater samples were collected from a multi-level groundwater monitoring site adjacent to the Sanggan River (SY Site in Fig. 1a) in the central Datong Basin where Ascontaminated groundwater occurs. Hydrogeological survey of the groundwater monitoring site revealed that the lithology of the sediments was brown to black fine-grained sand, silt and clay. At wells No. 1-2, No. 2-2, No. 3-2, No. 4-2 and No. 5-2 (Fig. 1b), twelve undisturbed core samples were collected in August 2010. Upon retrieving from the borehole, the sediments were packed in PVC casings immediately and wax-sealed on site to eliminate air exposure. The collected samples were stored at 4 °C in the dark. The groundwater samples were collected from the multi-level monitoring wells that were installed in the boreholes and screened at the corresponding depths. Before sampling, the groundwater in each monitoring well was purged for 10 min with a peristaltic pump. The physical and chemical parameters including temperature, pH, EC and redox potential (ORP) were measured on site with the YSI 600XLM portable meter that was calibrated before use. The groundwater samples were filtered through a 0.45 μm MilliPore filter (HDPE) using a vacuum pump. Groundwater samples used for the As determination and spectroscopic characterization of DOM were acidified (pH = 2) with ultrapure 12 N HCl and wrapped in aluminum foil. The groundwater samples used for major cation and trace metal analysis were acidified (pH b 2) with ultra-pure HNO3; groundwater samples used for anion analysis were not acidified. The alkalinity of groundwater was determined with a Gran titration within 24 h. Moreover, a representative surface water sample (RW) was also collected near the monitoring site. The water samples were stored at 4 °C in the dark, and the measurements were conducted within one week. 2.2. Extraction experiment The fresh sediment samples were air-dried, ground, and sieved through a 200 gauge screen. After 0.25 g of the prepared sample was digested using ultra-pure concentrated HNO3 and HF at 180 °C for 24 h, Fe and As contents of the bulk sediment were measured using inductively coupled plasma atomic emission spectrometry (ICP-AES) and hydride generation atomic fluorescence spectrometry (HG-AFS), respectively. The sequential As extraction procedure is based on the method reported by Keon et al. (2001) with slight modifications. (1) The phosphate-extractable As (As weakly and strongly adsorbed to sediment surfaces) was extracted as follows: 0.50 g of a fresh sediment sample was homogeneously mixed with 20 mL 1 M NaH2PO4 (pH = 5) in a 50 mL HDPE centrifuge tube. After being continuously shaken for 24 h at 25 °C in an overhead shaker, the tube was centrifuged at 5000 rpm for 20 min before the supernatant was decanted. The centrifugate was washed twice with ultrapure water and centrifuged again. The wash water was combined with the extraction solution and

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K. Pi et al. / Journal of Geochemical Exploration 149 (2015) 8–21

Fig. 1. (a) The location of the groundwater monitoring site (SY Site), (b) Plan view of the site and the monitoring wells. In (b) the labeled dots are the monitoring wells, the contour lines are the groundwater table contours, and the solid red circles mark the boreholes in which sediment samples were collected.

operation was the same as the first step. The extractants were stored in the dark at 4 °C before As was determined. The sedimentary poorly crystalline Fe(II) and total Fe extractions were conducted by mixing 0.5 g of fresh sediment with 20

then filtered through a 0.45 μm MilliPore filter. (2) The HCl-extractable As (As bound to the poorly crystalline Fe minerals) was extracted as follows: after the first step, the cleaned sample residue in the tube was immersed in 20 mL of 0.5 M HCl and shaken for 12 h at 25 °C. The next

Table 1 Physicochemical parameters and chemical compositions of the groundwater samples from the SY monitoring site in central Datong Basin. Sample ID

Type

Well Depth (m)

pH

EC (μS/cm)

ORP (mV)

As (μg/L)

As(III) (μg/L)

Fe (mg/L)

Fe(II) (mg/L)

HS− (μg/L)

SO2− 4 (mg/L)

HCO− 3 (mg/L)

DOC (mg/L)

SUVA (L/mg/m)

FI

RI

1-2S 1-2M 1-2D 2-1S 2-1M 2-1D 2-2S 2-2M 2-2D 2-3S 2-3M 2-3D 2-4S 2-4M 2-4D 3-2S 3-2M 3-2D 4-2S 4-2D 5-2S 5-2M 5-2D RW

GW GW GW GW GW GW GW GW GW GW GW GW GW GW GW GW GW GW GW GW GW GW GW SW

10.55 14.50 19.65 9.40 15.80 19.85 10.70 16.55 19.35 9.10 15.20 19.50 9.80 15.15 19.80 10.50 14.60 20.40 11.05 19.90 10.00 15.60 19.90 n.d.

8.13 7.63 8.17 7.73 8.06 8.27 7.80 7.74 8.41 8.07 7.89 8.54 7.81 8.51 8.10 7.84 7.68 8.29 8.17 8.21 7.94 7.62 8.37 9.01

4971 5349 3394 5340 4405 4372 4689 2736 4565 5231 3255 5102 4930 7389 2379 2138 2459 4322 4686 4564 3788 3156 3402 810

8.7 −79.2 −24.0 −32.5 −168.5 −154.5 −44.5 −63.4 −128.0 40.7 26.2 −105.0 25.2 −70.8 −43.8 17.6 −131.3 −16.8 −29.1 −129.0 −182.0 −196.0 −221.2 -

35.9 52.5 42.9 35.8 38.2 140 41.8 29.4 182 45.3 32.8 122 32.2 46.0 47.4 33.4 18.0 105 64.0 91.1 27.6 33.7 46.3 2.40

19.3 23.9 12.5 18.0 20.6 63.4 18.7 12.7 79.9 22.2 14.3 61.8 15.1 22.5 14.3 12.3 12.0 57.5 24.5 53.8 15.1 21.0 26.0 n.d.

0.44 0.99 0.56 0.46 0.61 1.16 0.56 0.57 1.34 0.55 0.15 0.92 0.47 1.07 0.43 0.65 0.57 0.47 0.30 0.41 0.46 0.29 0.22 0.20

0.04 0.72 0.41 0.34 0.41 0.30 0.36 0.03 0.85 0.32 0.08 0.46 0.18 0.81 0.24 0.32 0.02 0.29 0.06 0.23 0.02 0.02 0.02 0.01

10 4 7 6 13 45 5 8 3 5 3 4 3 1 6 5 94 8 5 7 25 52 30 n.d.

433 574 568 1370 1439 132 308 579 89.0 566 851 238 2353 1882 308 551 169 196 220 202 551 292 55.2 258

762 691 410 613 595 851 750 388 857 746 384 884 508 650 505 314 615 863 825 868 546 817 825 214

64.31 67.90 45.31 55.58 41.71 92.68 60.42 46.61 115.7 56.92 51.82 127.4 58.05 76.97 72.06 42.62 75.62 123.5 68.32 94.42 79.28 113.8 120.8 40.80

0.664 0.479 0.426 0.308 0.587 0.571 0.662 0.300 0.474 0.692 0.230 0.403 0.236 0.182 0.355 0.256 0.603 0.422 0.735 0.557 0.211 0.550 0.472 0.723

1.60 1.81 1.78 1.45 1.82 1.79 1.62 1.65 1.78 1.40 1.65 1.77 1.44 1.64 1.59 1.43 1.81 1.56 1.41 1.57 1.82 1.76 1.75 1.30

0.20 0.35 0.66 0.20 0.62 0.99 0.98 0.03 0.97 0.14 0.18 0.25 0.16 0.56 0.34 0.61 -

Note: GW-Groundwater; SW-Surface water; “-”: not determined; n.d.: not detected. The capital letter at the end of each sample ID refers to shallow (S), medium (M) and deep (D) aquifer respectively at each multi-level monitoring well.

K. Pi et al. / Journal of Geochemical Exploration 149 (2015) 8–21

mL of 0.5 M HCl in a 50 mL HDPE centrifuge tube under N2. After 24 h of shaking at 25 °C, the tube was centrifuged at 5000 rpm for 20 min and the supernatant was decanted; afterward, the material was washed and centrifuged as described above. The Fe(II) and total Fe contents of the extractant were immediately measured with a spectrophotometer (DR2800, HACH). The Fe(II) content was determined by adding the Ferrozine reagent to a small sample aliquot. The total Fe was analyzed using the same method after treated with a reducing agent (10% hydroxylamine hydrochloride). The analytical error is within 5% for Fe(II) and total Fe using the Ferrozine method. 2.3. Analytical methods Total organic carbon (TOC) contents of the sediments were measured using an elemental analyzer (Vario EL cube, Elementar) after the samples were air-dried, ground to 200 mesh and cleaned of inorganic carbon with 0.1 M HCl. The major cation concentrations in the groundwater samples were analyzed using ICP-AES (IRIS Intrepid II XSP,

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Thermo Elemental). The anion contents were determined using ion chromatography (IC) (DX-120, Dionex). The total dissolved As was determined by HG-AFS (AFS-9130, Titan). The As species were firstly separated with LC-SAX anion-exchange resin (Supelclean) (500 mg sorbent of 40-μm particle size and 60 Å pore size) in the field laboratory (Le et al., 2000), acidified to pH b2 with ultra-pure HCl, and then determined using HG-AFS. After being preconditioned with 50% methanol and double-deionized water, the resin retained As(V) but not As(III) in groundwater samples. The dissolved Fe(II) and sulfide concentrations were measured with spectrophotometry (DR2800, HACH) using the Ferrozine and Methylene Blue methods respectively onsite at the time of sampling. The concentration of total dissolved Fe in groundwater samples was determined using the same method for dissolved Fe(II) after treatment with the reducing agent (10% hydroxylamine hydrochloride). The DOC concentration was determined for the acidified samples using a high-temperature catalytic combustion method with a TOC analyzer (TOC-V, Shimadzu), generating data with a standard deviation of ± 2%. The average analytical error of the ICP-AES, IC, HG-AFS and spectrophotometry analysis is generally below 5%.

Fig. 2. Depth profiles of the dissolved (a) As, As(III), (b) Fe, Fe(II), (c) sulfide concentrations and (d) ORP values in the groundwater.

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Fig. 3. Plots of the dissolved As concentration versus the (a) Fe concentration and the (b) Fe(II) concentration in the groundwater. Blue dashed ellipse in (b) marks the samples that had low As and Fe(II) concentrations.

Fig. 4. Plots of the (a) ORP, (b) SO2− 4 , (c) Fe(II) and (d) As concentration versus dissolved sulfide concentration in groundwater.

K. Pi et al. / Journal of Geochemical Exploration 149 (2015) 8–21

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Table 2 Lithology, TOC contents and chemical compositions of the bulk sediments and the extractable fractions. Sample ID*

Depth (m)

Lithology

1-2-9 1-2-14 1-2-20 2-2-9 2-2-15 2-2-21 3-2-10 3-2-18 4-2-12 4-2-16 5-2-6 5-2-15

10.60 14.55 19.70 10.80 16.60 19.80 10.60 20.10 10.90 20.00 10.10 20.00

Brown silty clay Brown silt Black silt Brown sandy clay Gray fine sand Black silty clay Brown silty clay Dark gray silt Gray silt Deep gray silt Brown fine sand Dark gray silt

As (mg kg−1)

Fe (g kg−1)

Bulk Sediment

Phosphate extractable

HCl extractable

Percent of (P + H) (%)**

Bulk Sediment

HCl extractable Fe(II)

HCl extractable Fe

Percent of HCl extractable Fe (II) (%) in Fe

Percent of HCl extractable Fe (%) in total Fe

25.19 24.46 29.17 24.78 22.24 29.84 31.33 23.21 32.34 20.63 25.79 21.56

3.94 3.13 5.50 3.70 2.05 9.83 6.30 5.50 7.93 2.87 2.37 2.58

3.67 3.15 6.57 4.21 1.88 7.05 5.47 4.70 6.61 3.04 5.66 2.25

30.21 25.67 41.38 31.92 17.67 56.57 37.57 43.95 44.96 28.65 31.14 22.40

29.69 19.66 27.33 29.08 21.36 30.41 35.12 18.70 25.46 18.69 20.17 25.61

0.76 0.59 6.78 0.69 1.13 2.67 0.89 2.91 2.33 2.88 1.39 2.84

5.36 2.34 14.22 4.48 2.87 20.98 5.97 8.12 8.21 7.17 3.42 8.16

14.18 25.21 47.68 15.40 39.37 12.73 14.91 35.84 28.38 40.17 40.64 34.80

18.05 11.90 52.03 15.41 13.44 68.99 17.00 43.42 32.25 38.36 16.96 31.86

TOC (wt%)

0.28 1.10 0.96 0.25 0.73 1.22 0.70 0.38 1.30 0.12 0.34 0.24

*: The number at the end of each sample ID refers to the serial number of core samples along the depth. ** P + H: Phosphate and HCl extractable.

The UV-vis absorbance of the DOM was measured via UV-vis spectrophotometer (U-3900, Hitachi) in a 1 cm path length cell, generating data with a 5% confidence interval. The specific UV absorbance (SUVA) was calculated by normalizing the UV absorbance (at 254 nm) to the DOC concentration. The 3D excitation emission matrix spectra (EEMs) were obtained by scanning DOM samples from 200-450 nm (excitation) at 5 nm intervals and from 300–550 nm (emission) at 2 nm intervals on a luminescence spectrometer (F-4600, Hitachi), following the method of Stedmon et al. (2003). The EEMs were fit to the parallel factor analysis (PARAFAC) modeling to produce independent components (Stedmon and Bro, 2008), and the relative amount of each component was calculated as a percentage of the total fluorescence. Additionally, the fluorescence index (FI) and redox index (RI) were calculated from EEMs and PARAFAC modeling results following the methods of Mladenov et al. (2010) to offer information regarding the redox state of the humic substances.

3. Results and discussion 3.1. Hydrochemistry At the monitoring site, the groundwater table was less than 4 m below the ground surface. The groundwater samples were collected from three different depths at each multi-level monitoring well, ranging from approximately 9–20 m below land surface (Table 1). Groundwater pH was near-neutral to moderately alkaline, varying from 7.62 to 8.54 with an average value of 8.04 (Table 1). Both the total As and As(III) concentrations in the groundwater samples exceed the WHO guideline limit value (10 μg/L) for drinking water (WHO, 1993). The total As concentration in the groundwater samples ranged from 18.0 to 182 μg/L with an average of 58.4 μg/L, while the As(III) concentrations changed from 12.0 to 79.9 μg/L with an average of 27.9 μg/L (Table 1). The percentages of As(III) in the total dissolved As varied from 29.2% to 66.9%, and 80% of those values exceeded contents of 40%, indicating that As(III) is the predominant species in groundwater. Higher As and As(III) concentrations occurred at approximately 20 m (Fig. 2a) where the groundwater also contained a higher DOC concentration (refer to Fig. 8a); these sites had aquifer sediments that consist of dark gray silt rich in NOM. The dissolved Fe concentrations in the groundwater samples varied from 0.15 to 1.34 mg/L and showed a vertical variation pattern similar to that of As (Fig. 2b), probably due to their associated geochemical behaviors. The Fe(II) concentrations varied from 0.02 to 0.85 mg/L, and comprised 3–76% of the total Fe contents. Notably, the high levels of dissolved Fe(II) indicate that Fe(II) is the dominant dissolved Fe species in groundwater under the strongly reducing condition of the anoxic

aquifers, which is clearly reflected by the observed negative ORP values (Fig. 2d). Under the reducing conditions, the Fe(III) oxides/hydroxides were microbially reduced in the presence of labile organic matter, as demonstrated by the Fe isotope data (Xie et al., 2013a). Consequently, Fe(II) would accumulate in the aqueous phase, accounting for a high proportion of total dissolved Fe (up to 76%). Meanwhile, the bound As would be mobilized because the decrease in sorption sites of Fe(III) oxides/ hydroxides and the lower affinity of neutral As(III) species onto Fe(III) oxides/hydroxides under alkaline condition (Dixit and Hering, 2003). Therefore, reduction of As-bearing Fe(III) oxides/hydroxides can lead to the increase in concentrations of dissolved Fe, Fe(II) and As. However, the distribution trend for As was different from that of Fe, Fe(II) and ORP values (Fig. 2). The As concentration is only moderately correlated with total dissolved Fe (r2 = 0.41, α = 0.05) and very poorly correlated with the dissolved Fe(II) concentration (r2 = 0.24, α = 0.05). As a matter of fact, two trends can be recognized between dissolved Fe and As from Fig. 3. First, there exhibits a mutual increase and the As shows a good positive correlation with both Fe and Fe(II) (Red line 1 and 3 in Fig. 3), which can be attributed to the concurrent release of Fe and As that is controlled by the reductive dissolution of Fe(III) oxides/hydroxides. Second, as indicated by the observed variation of As and Fe concentrations at the depth of 15 m (Fig. 2a), there are monotonous increases in Fe and Fe(II) concentrations, but As concentration keeps at a relatively low level (Red line 2 and 4 in Fig. 3). In this case, As should have been involved in other non-conservative processes. It is worth of noting that high concentration of DOC was detected in the high As groundwater (Table 1), and DOC showed a distribution pattern similar to those of Fe and As (refer to Figs. 2a, b and 8a). In the anoxic aquifers, Fe(III) and SO2− can serve as important electron acceptors in microbially me4 diated oxidation of DOM, producing high concentrations of dissolved Fe(II) and sulfide (Xie et al., 2013b). Some of the released Fe(II) may form secondary Fe(II)-bearing minerals, such as siderite and amorphous FeS in the anoxic aquifers (Matsunaga et al., 1993). These minerals can adsorb or co-precipitate with dissolved As, decreasing its concentration in the aqueous phase. Besides, As can react with dissolved sulfide to produce As-bearing sulfide precipitates in strongly reducing environment, resulting in As sequestration. Therefore, although As mobilization is primarily associated with the reductive dissolution of Fe(III) oxides/ hydroxides under reducing conditions, other geochemical processes, such as DOM biodegradation and microbially mediated SO2− 4 reduction, may have great impact on the fate of Fe and As in groundwater systems. Interestingly, dissolved sulfide was prevalent in most of groundwater samples from the monitoring site. The sulfide concentrations ranged from 1 to 94 μg/L, with an average value of 15 μg/L (HS− in Table 1). The negative correlations between dissolved sulfide and ORP values and SO2− 4 concentrations are evident (Fig. 4a and b). Formation of dissolved

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Fig. 6. Relationship between extractable poorly crystalline Fe and As contents in the sediments.

and As(III), releasing As into groundwater (Burton et al., 2013). By contrast, the prevailing negative correlation between As and sulfide concentrations suggests that such a decrease in As concentration may be related either to its precipitation as arsenopyrite or to its adsorption onto the Fe(II) sulfides. The Fe(II) sulfides were generally considered as the main sequester for As in aquifers (Bostick et al., 2004; Kirk et al., 2004; Lowers et al., 2007). In addition, other As sulfide minerals, such as realgar and orpiment, could form in strongly reducing environments where ORP was much more negative and the sulfide level remained high (such as in samples 3-2M and 5-2M, Table 1) (O'Day et al., 2004). As shown in Fig. 3b, the groundwater samples with elevated dissolved sulfide concentrations (in blue dashed circle) had abnormally low dissolved concentrations of As and Fe(II).

3.2. Geochemistry of sediments Fig. 5. Relationships between the (a) TOC, (b) Fe and As contents in the sediments.

sulfide via microbially mediated SO2− 4 reduction coupled to oxidation of organic matter was validated by dramatically enriched δ34S[SO4] values (up to +36.1‰) in the high As aquifers (Xie et al., 2009, 2013b). However, as shown in Fig. 4a, high sulfide concentrations were only detected under stronger reducing conditions (almost ORP b −150 mV), which indicates that SO24 − reduction is more difficult to occur, given the more negative redox potential value needed for SO2− reduction half4 reaction than for Fe(III) reduction (Christensen et al., 2000). There is a negative correlation between the dissolved sulfide and Fe(II) (Fig. 4c). Presumably, under the strongly reducing condition, dissolved sulfide resulting from active SO2− 4 reduction can react with Fe(II) from reduction of Fe(III) oxides/hydroxides to preferentially form Fe(II) sulfides in the aquifers (Kirk et al., 2004; Nickson et al., 2000; Pal et al., 2002). As a matter of fact, except for several groundwater samples with relatively high concentrations of dissolved As and sulfide, low concentrations of dissolved As were mostly found in groundwater samples with high sulfide concentrations, and As generally displayed a negative correlation with dissolved sulfide (Fig. 4d). The exceptional cases of cooccurring high concentrations of As and sulfide may be due to the reduction of Fe(III) to Fe(II) and As(V) by dissolved sulfide to thioarsenate

The results of geochemical analysis for the sediment samples are summarized in Table 2. The grain size of the typical sediments ranged from fine sand and silt to sandy, silty clay. The total As contents in sediments varied from 20.63 to 32.34 mg/kg with an average value of 25.88 mg/kg, far exceeding its ordinary level of 5–10 mg/kg in modern unconsolidated sediments (Smedley and Kinniburgh, 2002). Relatively high TOC levels (0.12% to 1.30%, average 0.64%) were detected in the gray and black sediment samples. The total Fe contents ranged from 18.69 to 35.12 g/kg, with an average of 25.11 g/kg. The positive correlation between TOC and total As contents in the bulk sediments (r2 = 0.50, α = 0.05) (Fig. 5a) indicate that As is closely associated with the sedimentary organic matter. The organic matter dispersed in the aquifer sediments is easily released into the aqueous phase and acts as an important source for labile DOM in the groundwater to fuel the microbial respiration (Mladenov et al., 2010; Reza et al., 2010). Furthermore, the organic matter bound to the surfaces of Fe minerals can function as electron donor to promote in-situ reductive dissolution of Fe(III) oxides/hydroxides, releasing DOM, Fe(II) and As into groundwater (Chen et al., 2006; Mladenov et al., 2010). The linear correlation between As and total Fe contents (r2 = 0.41, α = 0.05) (Fig. 5b) implies that Fe minerals, including Fe(III) oxides/hydroxides, should be the main sink for As, as further confirmed by the results of sequential extraction analysis.

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Fig. 7. Spectral characteristics of the four identified DOM components using PARAFAC modeling.

Notably, adsorbed and poorly crystalline Fe-bound As accounted for 17.67–56.57% of the total As in sediment samples (Table 2). These two As fractions (referred to as Extractable As) are highly vulnerable to

geochemical mobilization under alkaline and reducing conditions (Jung et al., 2012; Keon et al., 2001). The maximum As content in this pool was observed in sample 2-2-21, and the corresponding

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groundwater sample (2-2D in Table 1) contained the highest As concentration. The extractable Fe contents ranged from 2.34 to 20.98 g/kg, accounting for 11.90–68.99% of total Fe in bulk sediments; the poorly crystalline Fe phase made up a significant portion, or even the majority of total Fe in bulk sediments. Our previous extraction experiment with the sediments from Datong Basin demonstrated that As was predominantly present in adsorbed, amorphous and poorly crystalline Fe minerals (Xie et al., 2008). Similar to the results of bulk sediments, the positive linear relationship (r2 = 0.55, α = 0.05) between the extractable Fe and As (Fig. 6) suggests that As is primarily bound to Fe oxides/hydroxides via surface adsorption and/or co-deposition. The Fe(II) phase accounted for a significant portion of the poorly crystalline Fe phase (12.73–47.68%). Although this Fe(II) fraction most likely contains diverse secondary Fe(II) minerals, a considerable portion of the Fe(II) phase might be Fe(II) sulfides because high levels of dissolved sulfide were detected in the same aquifers. Additional clues to support this are the H2S odor produced during HCl extraction experiments as well as the results of our Fe isotope study (Xie et al., 2013a).

3.3. DOM in groundwater 3.3.1. Characteristics of DOM The groundwater was characterized by high DOC concentrations (up to 127.4 mg/L) (Table 1) which should be derived from NOM rich in the sediments. The four DOM components identified by the PARAFAC modeling are shown in Fig. 7. One terrestrial humic substance (THA, Component 1) corresponds to a derivative of fulvic acid and three microbially derived humic substances (MHA) include a reduced quinone (Component 2) and two oxidized quinone (Component 3 and 4) moieties (Cory and McKnight, 2005; Stedmon and Markager, 2005; Stedmon et al., 2003). The vertical distributions of the DOC concentration, SUVA, FI and RI values of the groundwater are shown in Fig. 8. Shallow groundwater samples had relatively low DOC concentrations and small RI values, but with significant variations in SUVA and FI values. The presence of both high and low FI values (1.82 and 1.40 respectively) indicates that there are two DOM sources: terrestrial and microbial (McKnight et al., 2001). Low FI and high SUVA values similar to those of surface water (1.30 and 0.723 respectively) suggest that

Fig. 8. Depth profiles of the (a) DOC concentration, as well as the (b) SUVA, (c) FI, and (d) RI values of DOM in the groundwater.

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DOM input from plant degradation in the unsaturated zone along with infiltration of irrigation recharge may be important since irrigation using surface water has been a common practice for agricultural development in this arid/semi-arid region (Xie et al., 2012b). However, high FI and low SUVA values reveal significant microbial influence over the DOM formation (McKnight et al., 2001). By contrast, the low RI and positive ORP values, as well as the dominance of the Component 3 fluorophores in those DOM samples, all indicate that a large portion of the microbially derived quinone-like HS should be in the oxidized state. For the deep groundwater samples from approximately 20 m below land surface, a different scenario was observed. Their higher DOC concentrations than those of the shallow groundwater should be due to DOM release from peat in sediments or other dispersed sedimentary organic matter in the aquifer that has accumulated under the sluggish groundwater flow at central part of the Datong Basin (Xie et al., 2012b). The microbial degradation of DOM became a dominant process, and thereby high FI and low SUVA values were commonly observed. Accordingly, the DOM at these depths overwhelmingly generated microbial derived fluorophores (Component 2 and 3). Moreover, the redox conditions of the aquifer systems became more reducing at the depth, as reflected by the lower ORP and higher RI values.

3.3.2. DOM transformation A positive correlation (r2 = 0.58, α = 0.05) between the HCO− 3 and DOC concentrations is evident (Fig. 9a). According to the results of carbon isotope study, DOM in high As groundwater systems was commonly proposed as a major source for HCO− 3 (Liu et al., 2009; McArthur et al., 2004; Meharg et al., 2006; Xie et al., 2013b). Under anaerobic conditions, labile DOM can be utilized by indigenous microorganisms as an electron donor to promote the microbial respiration (Mladenov et al., 2010). CO2 derived from the biodegradation of DOM would promote the dissolution of carbonate minerals in sediments to elevate HCO− 3 concentration in groundwater. During the biodegradation of DOM in the anoxic aquifers, the reducing organic carbon could simultaneously promote microbially mediated redox reactions of other elements, including Fe, S and As. Using DOM as an electron donor, the reduction of Fe(III) oxides/hydroxides and SO2− 4 would be favored, resulting in Fe(II) and sulfide enrichment in the aqueous phase. This is verified by the ubiquitous occurrence of microbially derived humic-like Components 2 and 3 in the groundwater samples. Therefore, elevated concentration of HCO− 3 in the high As groundwater can be related to DOM biodegradation. Additionally, As oxyanions might be desorbed into the aqueous phase due competitive sorption of HCO− 3 (Xie et al, 2008). In Fig. 9b, a poor correlation between As and DOC concentrations (r2 = 0.37, α = 0.05, blue line) is observed in groundwater samples. However, if the samples in the blue dashed ellipse (Fig. 9) were excluded, a good positive linear correlation (r2 = 0.72, α = 0.05, red line) can be obtained. The excluded samples contained lower As concentration but high sulfide and low dissolved Fe(II) concentrations, implying that their As concentrations may be lowered down due to co-precipitation with Fe(II) sulfides. This can be seen from the As(III) vs. RI plot (Fig. 10a): the samples in the blue dashed ellipse had higher RI values and low As(III) concentrations. The availability of labile DOM enhanced the microbially mediated SO2− reduction. Consequently, the oxidized 4 quinone-like humic substances were accumulated and tended to increase under reducing conditions. From the ORP vs. RI plot, it can be seen that the samples with the most negative ORP values had moderate RI values (blue dashed ellipse in Fig. 10b). Therefore, microbially mediated degradation of DOM is coupled with SO2− 4 reduction to promote As scavenging, particularly when easily biodegradable DOM contents and strongly reducing conditions are jointly available. By contrast, a good linear correlation between As and DOC concentrations in the remaining groundwater samples indicate that As enrichment should be closely related to the presence of DOM.

Fig. 9. Relationships between the concentrations of (a) HCO− 3 , (b) As and DOC in groundwater. Blue dashed ellipse in (b) marks the samples that had high DOC but low As concentrations.

Moreover, in reduction reactions, the electrons were likely to be transferred from the labile DOM to the As(V) adsorbed on Fe oxides/ hydroxides, leading to As(V) reduction into As(III) and subsequent desorption from the surfaces due to its weaker affinity (Palmer et al., 2006; Redman et al., 2002; Tongesayi and Smart, 2006). Since the redox potential of As(V) reduction is close to that of Fe(III) reduction (Christensen et al., 2000), As(V) reduction is expected to occur between Fe(III) and SO2− reduction. The groundwater samples with higher As 4 and As(III) concentrations also had higher RI values (Table 1), and dissolved As(III) concentrations and RI values are positively correlated (r2 = 0.88, α = 0.05, red line in Fig. 10a). Hence, the electron transfer from the labile DOM to the As(V) via shuttling of the quinone-like HS should have occurred in the reducing aquifers. And reductive desorption of As(V) could be a mechanism conducive to elaborate the weak correlation between As and Fe concentrations in groundwater. From the relationship between As concentrations and SUVA values of DOM (Fig. 10c), two contrary trends can be discerned. In the positive trend (blue dashed line 1 in Fig. 10c), SUVA values increased with elevated As concentrations, which might be due to the production of aromatic humic substances during biodegradation of labile DOM (Chin

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Fig. 10. Plots of (a) As(III) concentration versus RI value, (b) ORP versus RI, (c) As versus SUVA and (d) FI versus SUVA values in groundwater. Blue dashed ellipse in each plot marks the same groundwater samples.

et al., 1998; Weishaar et al., 2003). For the negative trend (blue dashed line 2 in Fig. 10c), groundwater with higher SUVA values had lower As concentrations, particularly the samples containing high sulfide concentrations (in the blue dashed ellipse). For the latter case, a greater degree of labile DOM biodegradation may have proceeded, making the residual HS more aromatic. Consequently, SO24 − was microbially reduced using DOM as electron donors, resulting in elevated dissolved sulfide concentration in groundwater. This conclusion is consistent with the relationships between As and DOC and RI values addressed above. Furthermore, those marked groundwater samples also had higher FI values (Fig. 10d), supporting the hypothesis that DOM have experienced active biodegradation. 3.4. The influence of DOM on arsenic transport As and Fe are two closely-related elements in geochemical evolution of high As groundwater systems. Arsenic adsorbed to or co-precipitated

with Fe(III) oxides/hydroxides is responsible for the initial As deposition in solid phase (Xie et al., 2008). However, the sediment samples with high As and relatively low Fe contents may be due to the sequestration of As by sulfide minerals (Kirk et al., 2004). Apart from the essential provenance of As enrichment, the reducing conditions and the relationships of dissolved As with Fe, DOC concentrations and RI values of DOM jointly provide evidences about the important effect of the coupling of reductive dissolution of Fe(III) oxides/hydroxides to biodegradation of DOM on As mobilization in the aquifers. Therefore, DOM plays a critical role in the fate of As (see conceptual model in Fig. 11). The labile DOM can serve as an energy source for the indigenous bacteria and facilitate microbially mediated reductive dissolution of Fe(III) oxides/hydroxides and reduction of SO2− 4 . Furthermore, the quinone-like humic substances can act as an electron shuttle between the labile DOM and Fe(III), SO2− and As(V) (Islam et al., 2004; 4 Mladenov et al., 2010; Tufano and Fendorf, 2008). The reductive desorption of As(V) from the surfaces of Fe(III) oxides/hydroxides further

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Fig. 11. Conceptual geochemical model for effects of DOM biodegradation on As transport at the study site. DOM serves as electron donor and quinone-like humic substances (HS) in DOM act as electron shuttle to promote the microbially mediated reductions of Fe(III) oxides/hydroxides (FeOOH(s)), As(V) and SO2− 4 , resulting in substantial As mobilization or some As sequestration into the solid phase.

enhances the release of As(III) into groundwater (Jiang et al., 2009). On the other hand, elevated sulfide concentration in groundwater due to SO2− 4 reduction immobilizes As via precipitation. Although both mobilization and immobilization processes are thermodynamically feasible in the aquifers, the heterogeneity and availability of the ambient DOM determines the extent of these two reactions, which may account for the variable spatial distribution of As concentration in groundwater (Jiang and Kappler, 2008). Specifically, the ubiquity of the elevated As can be attributed to the predominance of the reductive dissolution of Fe(III) oxides/hydroxides and coincident As mobilization processes. Therefore, DOM can not only trigger the enrichment of As in groundwater, but also promote the re-sequestration of As into the sediment. Given the ubiquitous presence of elevated As concentration in groundwater, it must be kept in mind that DOM may assist in As immobilization via the co-precipitation with Fe(II) sulfides or the formation of As sulfide minerals, particularly when abundant labile DOM and favorable sulfate-reducing bacteria are available in a SO24 −-rich and reducing environment. Thus, As mobilization and immobilization are closely related to the microbially mediated cycling of Fe, S and C in the aquifers, although details of the biogeochemical mechanism merit further exploration.

electron acceptors. Correspondingly, the electron donated by labile DOM promotes the microbially mediated reduction of Fe(III) and SO2− 4 . Electrons from DOM may not only be transferred to Fe(III) and SO2− under the shuttling of quinone-like humic substances, but also 4 to As(V) adsorbed on the surfaces of Fe minerals, facilitating As (particularly As(III)) mobilization and enrichment in the groundwater. The correlations of As species with Fe(II), DOC concentrations, RI and SUVA values of DOM indicate that As release is primarily controlled by microbially mediated reductive dissolution of Fe(III) oxides/hydroxides coupled to biodegradation of natural DOM. However, in the groundwater with abnormally low As and Fe(II) and high sulfide concentrations, As may be most likely scavenged via co-precipitation with Fe(II) sulfides and/or As sulfide minerals. Therefore, this study demonstrates that As transport in groundwater systems is closely associated with the microbially mediated cycling of Fe, S and C. Natural DOM may not only control the mobilization and enrichment of As in groundwater but also influence As sequestration under reducing conditions. Our study provides new approaches and perspectives to understand the biogeochemical processes involving Fe, As and S which are driven by DOM in aquifers under conditions similar to Datong Basin.

4. Conclusions

Acknowledgments

In the groundwater samples from the central Datong Basin aquifers, As concentration ranged from 18.0 to 182 μg/L with an average value of 58.4 μg/L. Dissolved Fe(II), sulfide and DOC concentrations were high in the elevated As groundwater. As(III) and Fe(II) were the predominant As and Fe species under the strongly reducing conditions The As and TOC contents in sediments were both high, with an average of 25.88 mg/kg and 0.64 wt% respectively. The total and extractable As contents exhibit positive correlations with those of Fe, indicating that solid As is strongly associated with Fe minerals via adsorption and/or coprecipitation in the sediments. The groundwater with elevated As was also characterized by high concentrations of labile DOM with DOC concentrations up to 127.4 mg/L. Shallow groundwater DOM primarily contained terrestrially and microbially derived humic substances, and most of the quinone-like humic substances were oxidized. By contrast, the deep groundwater DOM was overwhelmingly microbially derived, consisting of reduced quinone-like humic substances. The good linear correlation between − DOC and HCO− 3 concentrations suggests that high HCO3 concentrations may be related to biodegradation of DOM using Fe(III) and SO24 − as

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