Anthropogenic influences on dissolved organic matter transport in high arsenic groundwater: Insights from stable carbon isotope analysis and electrospray ionization Fourier transform ion cyclotron resonance mass spectrometry

Journal Pre-proofs Anthropogenic Influences on Dissolved Organic Matter Transport in High Arsenic Groundwater: Insights from Stable Carbon Isotope Analysis and Electrospray Ionization Fourier Transform Ion Cyclotron Resonance Mass Spectrometry Kai Yu, Yanhua Duan, Yiqun Gan, Yanan Zhang, Ke Zhao PII: DOI: Reference:

S0048-9697(19)35154-X https://doi.org/10.1016/j.scitotenv.2019.135162 STOTEN 135162

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Science of the Total Environment

Received Date: Revised Date: Accepted Date:

13 August 2019 2 October 2019 23 October 2019

Please cite this article as: K. Yu, Y. Duan, Y. Gan, Y. Zhang, K. Zhao, Anthropogenic Influences on Dissolved Organic Matter Transport in High Arsenic Groundwater: Insights from Stable Carbon Isotope Analysis and Electrospray Ionization Fourier Transform Ion Cyclotron Resonance Mass Spectrometry, Science of the Total Environment (2019), doi: https://doi.org/10.1016/j.scitotenv.2019.135162

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Anthropogenic Influences on Dissolved Organic Matter Transport in High Arsenic Groundwater: Insights from Stable Carbon Isotope Analysis and Electrospray Ionization Fourier Transform Ion Cyclotron Resonance Mass Spectrometry Kai Yua, Yanhua Duanb, Yiqun Ganb*, Yanan Zhangb, Ke Zhaob a

State Environmental Protection Key Laboratory of Integrated Surface Water-Groundwater Pollution Control, School of Environmental Science & Engineering, Southern University of Science and Technology, Shenzhen 518055, China

b

School of Environmental Studies, China University of Geosciences, Wuhan 430074, China

* Corresponding author at: School of Environmental Studies, China University of Geosciences, Wuhan 430074, China. E-mail address: [email protected]. Abstract In East and Southeast Asia, the health of over 100 million people is threatened by the consumption of groundwater containing high concentrations of arsenic (> 10 μg L−1), which is released from sediments through reductive dissolution of arsenic-bearing iron/manganese oxides. Dissolved organic matter (DOM) is known to play a crucial role in the process of arsenic mobilization in shallow aquifers, and its availability and reactivity are key factors controlling the variation of arsenic concentrations in groundwater. However, it is unclear how human activities influence the transport of DOM and how the transportation affects the DOM molecular properties in high arsenic groundwater. This study provides insights on the sources and molecular compositions of DOM in groundwater from the Jianghan Plain, central China, a newly discovered area with seasonal fluctuations in arsenic concentrations in shallow groundwater. Monitoring of water levels and stable carbon isotope compositions in groundwater from different depths and canal water over a year indicated that terrestrial DOM was the dominant source, 1

accounting for 54.2%–85.5% of groundwater DOM. Electrospray ionization combined with ultrahigh-resolution Fourier transform ion cyclotron resonance mass spectrometry revealed that canal water infiltration transferred aliphatic, tannin-like and leached aromatic DOM from sediments into groundwater. Therefore, groundwater recharge through irrigation using canal water not only inputs terrestrial DOM, but also accelerates the release of sedimentary DOM. Furthermore, carboxylic-rich alicyclic molecule (CRAM)-like DOM that is derived from biomolecules has the highest proportion (60.1%–65.5%) among the identified DOM structures. And, it might be reused in biochemical processes during arsenic mobilization, suggesting a third source of groundwater DOM in addition to canal water and sediments. The findings in this study advance the understanding on transport processes and molecular properties of DOM in high arsenic groundwater under extensive anthropogenic influences. Keywords: high arsenic groundwater; dissolved organic matter; stable carbon isotope; Fourier transform ion cyclotron resonance mass spectrometry 1.

Introduction High arsenic (As) concentration in groundwater is a common environmental issue that threatens the

health of over 100 million people in East and Southeast Asia. The availability of dissolved organic matter (DOM) is a key factor governing As mobilization in aquifers (Harvey et al., 2002; Mailloux et al., 2013; Mladenov et al., 2015; Neumann et al., 2010). Labile DOM fuels microbial activities, liberating As from sediments through reductive dissolution of As-bearing iron (Fe) oxides (Jia et al., 2018; Mladenov et al., 2010). It is also an electron donor in the reduction of As-bearing oxides, and specific DOM species (e.g., quinones) can shuttle electrons and accelerate the reductive dissolution of As-bearing minerals (Jiang and Kappler, 2008; Klupfel et al., 2014; Kulkarni et al., 2018a; Lovley et al., 2

1996; Mladenov et al., 2015; Olk et al., 2019). Furthermore, DOM can release As directly through competitive sorption (Bauer and Blodau, 2006; Redman et al., 2002) and complexation (Liu and Cai, 2013; Fakour and Lin, 2014; Ren et al., 2017; Sharma et al., 2010). Complexation between As and DOM can occur via Fe-bridges in ternary As-Fe-DOM complexes (Liu et al., 2011; Sundman et al., 2014) or through interaction with oxygen-containing functional groups such as carboxylic groups (Biswas et al., 2019) to enhance the mobility of As. Interpretation of the sources and transport processes of DOM is of great importance for understanding the patterns of As contamination and the mechanisms of As concentration fluctuation in groundwater. Studies in West Bengal, India, within the Ganges Delta showed that the presence of organic-rich clay or peat lenses can accelerate As mobilization through reductive dissolution of Fe oxides (Datta et al., 2011; Huang et al., 2015; McArthur et al., 2004; Meharg et al., 2006; Sengupta et al., 2008). Other field and experiment studies showed that sedimentary organic matter can supply sufficient DOM that is necessary for microbial mediated reductive dissolution of As-bearing oxides (Kulkarni et al., 2017; Kulkarni et al., 2018b; Neumann et al., 2014; Schittich et al., 2018; Vega et al., 2017). However, some studies in the Munshiganj district, also within the Ganges Delta, did not find peaty strata in shallow aquifers (Harvey et al., 2002; Harvey et al., 2006; Neumann et al., 2010). Instead, they demonstrated that groundwater extraction and irrigation greatly altered groundwater flow paths, driving organic matter from anoxic pond sediments into shallow aquifers (Harvey et al., 2002; Harvey et al., 2006; Neumann et al., 2010). In Vietnam, surface water contaminated by human waste was demonstrated to be another source of organic matter in groundwater, and it may increase groundwater pollution by As (McArthur et al., 2012). In Cambodia, although groundwater flows were not disturbed by anthropogenic activities, pond-derived DOM can be transported to depths of up to 50 3

m, releasing As from aquifer sediments (Benner et al., 2008; Lawson et al., 2013; Polizzotto et al., 2008). The reactivity of DOM is another factor controlling the behavior of groundwater As. Incubation experiments with sediments have demonstrated that the input of biologically degradable DOM from pond water drove arsenic mobilization in shallow groundwater in the Ganges Delta, whereas recharge from rice fields contained mainly recalcitrant DOM (Neumann et al., 2010). DOM reactivity is largely determined by its chemical composition, and the detailed composition and structure of DOM can provide valuable information relating to its reactivity and alteration during transport (Chen et al., 2014; Hedges et al., 1992). Recent studies using stable isotope analysis and excitation–emission matrix fluorescence spectrometry have shown that short-chain n-alkanes/n-alkanoic acids and terrestrially derived low molecular weight humic-like substance were vulnerable to biodegradation (Li et al., 2019; Zhou et al., 2018). However, more detailed information regarding the fundamental properties of DOM at the molecular scale is needed to understand the influence of DOM reactivity on As behavior. In the present study, we investigated DOM properties and groundwater As concentrations over a year with the goal of elucidating the influence of human activities on DOM transport and corresponding groundwater As behavior. Our study area, the Jianghan Plain, central China, was heavily impacted by human activities and showed elevated As concentrations in shallow groundwater with apparent seasonal fluctuations (Duan et al., 2015; Schaefer et al., 2016; Yu et al., 2018). Water and sediment samples from two sites influenced by different human activities were taken and the groundwater flow patterns were analyzed using hydraulic gradients and geochemical indicators. Stable carbon isotope analysis of different type of waters and sediments was used to trace the origins of DOM in groundwater at different depths. Electrospray ionization (ESI) combined with ultrahigh-resolution 4

Fourier transform ion cyclotron resonance mass spectrometry (FTICR-MS), a promising approach that can provided a molecular fingerprint for DOM (Chen et al., 2014; D’Andrilli et al., 2010; Flerus et al., 2012; Hertkorn et al., 2006; Kujawinski et al., 2004; Reemtsma et al., 2008; Sleighter and Hatcher, 2008; Stubbins et al., 2010), was used to characterize DOM at the molecular level. The findings of the present study directly improve our understanding of the molecular properties of DOM in shallow groundwater under extensive anthropogenic influences, and the effect of DOM transport on seasonal variations of groundwater As. 2.

Materials and methods

2.1 Field site The field site locates in the southeast of the Jianghan Plain (Fig. 1), which is an approximately 55 000 km2 alluvial basin built of sediments from the Yangtze River. The study area has a subtropical climate with annual precipitation of 1269 mm (Duan et al., 2015). The monsoon in the Jianghan Plain is between April and August with 100–200 mm of precipitation per month (Duan et al., 2015). Two sites (SY05 and SY07) impacted by different human activities were studied (Fig. 1). Site SY05 was located in a paddy field (Fig. 1), which was irrigated from April to September annually. The water for irrigation was drawn from an adjacent canal (SY05-C in Fig.1). Site SY07 was located in a corn field (no irrigation), which was near a canal (SY07-C) and a chemical plant producing titanium compounds (Fig. 1). The plant pumped groundwater irregularly and therefore influences the groundwater flow in shallow aquifer (Schaefer et al., 2016; Yu et al., 2018). At each site, two wells with a 1-m screen on each well centered at 10 and 25 m, respectively, were installed in November 2011. The drilling and well construction were detailed in previous studies (Duan et al., 2015; Schaefer et al., 2016).

5

Fig. 1. Map of the study sites in the Jianghan Plain, central China. Monitoring wells and canal water sampling sites are indicated by circles in different colors on a Google Earth satellite image. 2.2 Sample collection During drilling, sediment samples were collected between 0 and 30 m at SY05 (n = 13) and SY07 (n = 9), respectively, and the lithological log was shown in Fig. 6. Sediments were stored in 17.5-cm stainless steel tubes (ø 3.8 cm), which were capped with Teflon lids on both ends. Then, the sample tubes were sealed with paraffin, vacuum-sealed in polyethylene bags on site, and preserved in anoxic gas-tight boxes at 4 °C until required for analysis. Sediments were air-dried in a 65 °C oven and ground to a particle size of < 0.15 mm before analysis. Water in each well and canal (SY05-C and SY07-C, Fig. 1) were sampled monthly from April 2014 to March 2015. All water samples were filtered through 0.22-μm membranes on site to remove microorganisms, preventing samples from further alteration during transportation and preservation. For the As concentration measurements, water samples were stored in 50-mL high-density polyethylene (HDPE) bottles wrapped with aluminum foil, and acidified with concentrated hydrochloric acid to pH < 2. Water samples for measuring anions (SO42– and Cl–) were stored in 50-ml HDPE bottles. Water samples for measuring cations (K+, Na+, Ca2+ and Mg2+) were stored in 50-ml HDPE bottles and acidified to pH < 2 using concentrated nitric acid. Water samples for measuring alkalinity were stored in 100-ml HDPE bottles. Water samples for measuring dissolved organic carbon (DOC), which represents the carbon content of DOM, and its stable carbon isotope values (δ13CDOC) were stored in 40-mL amber glass bottles and acidified with 85% H3PO4 to pH < 2. Samples for FTICR-MS analysis were collected in duplicate 40-mL amber glass bottles in April and October 2014. All water samples were transported and stored at 4 °C before analysis. 6

2.3 Analytical methods Groundwater levels were measured semimonthly from April 2014 to March 2015 and canal water levels were measured monthly from May 2014 to March 2015. Water levels were measured using beeper tape and calibrated to the elevations (see (Schaefer et al., 2016) for calculation details). The electrical conductivity (EC), temperature and pH of water were measured during sampling using a HQ40D portable meter (HACH, Loveland, USA). The Fe(II) and sulfide concentrations in groundwater were also measured on site using a portable spectrophotometer (HACH 2800, HACH). Alkalinity was measured within 24 h by standard hydrochloric acid titration. Concentrations of As in water samples were measured using an atomic fluorescence spectrometer (AFS-930, Beijing Titan Instruments Co., Ltd, China). Anions and cations were measured, respectively, using ion chromatograph (IC, ICS-1100, Thermo Fisher Scientific, Waltham, USA) and inductively coupled plasma atomic emission spectrometer (ICP-AES, Optima 8000, Perkin Elmer, Waltham, USA). The DOC concentration of the water samples and total organic carbon (TOC) concentration of the sediments were measured using a total organic carbon analyzer (multi N/C 3100, Analytik Jena, Germany). The TOC concentration in sediments were presented as mass fractions (%). The δ13CDOC values of water samples were measured using a gas chromatograph (Trace GC Ultra, Thermo Fisher Scientific, USA) coupled with an isotope ratio mass spectrometer (MAT 253, Thermo Fisher Scientific) using an established method (Yu et al., 2015). The stable carbon isotope values of sedimentary organic carbon (δ13CSOC) were analyzed using a Flash 2000 elemental analyzer (Thermo Fisher Scientific) coupled with an isotope ratio mass spectrometer (MAT 253, Thermo Fisher Scientific) following an established method (Gandhi et al., 2004). The stable carbon isotope values are expressed in per mil: 𝛿13C =

(

Rsample ― RVPDB RVPDB

) × 1000 7

(1)

where R is the abundance ratio of 13C to 12C; and VPDB represents Vienna Pee Dee Belemnite, which is the international standard for stable carbon isotope analysis (Craig, 1957). The precisions of the δ13CDOC and δ13CSOC values were ± 0.5‰ and ± 0.3‰, respectively. 2.4 Analysis of DOM compositions Samples for FTICR-MS analysis were concentrated by solid phase extraction with ~50% DOC recovery. The sample volume of used for solid phase extraction was determined according to the initial DOC concentrations. All samples were concentrated to a DOC concentration of ~20 mg L–1. Ultrapure water was used to adjust the concentrated samples if the DOC concentration was higher than 20 mg L–1. The FTICR-MS analysis was conducted in Environmental Molecular Sciences Laboratory of the Pacific Northwest National Laboratory at the U.S. Department of Energy (Richland, WA). The molecular properties of DOM were characterized using a 12 Tesla Bruker SolariX FTICR mass spectrometer with an ESI source in negative ionization mode. Samples were injected into the ESI source through a syringe pump (3.0 μL min–1). The needle voltage and Q1 were set to +4.4 kV and 50 m/z, respectively. The ions observed in the spectra were singly charged. Raw spectra were converted to peak locations using the DataAnalysis software. Formularity software (Tolić et al., 2017) was used to assign the chemical formulas. Peaks with signal to noise ratio > 7 and mass measurement error < 1 ppm were selected for formula assignments. Only C, H, O, N, S and P were considered during formula assignments. A list of criteria established by previous studies (Kujawinski and Behn, 2006; Koch and Dittmar, 2006; Stubbins et al., 2010) was used to screen the formulas to eliminate those that unlikely exist in natural organic matter. If a peak at an m/z value has no matched chemical formula, it was left unassigned. The DOM with CHO formulas were further classified into aromatic, carboxylic-rich alicyclic molecule (CRAM)-like, aliphatic, tannin-like, and 8

carbohydrate-like compounds according to the double bond equivalents (DBEs), which was the sum of unsaturation plus rings in a molecule, and the aromaticity index (AI), which was an unambiguous indicator for the identification of aromatic structures (Chen et al., 2014; Hertkorn et al., 2006; Koch and Dittmar, 2006; Perdue, 1984; Sleighter and Hatcher, 2007). The DBE and AI were calculated using the equations DBE = 1 + C − 0.5H and AI = (1 + C − O − 0.5H) / ( C – O ) (Chen et al., 2014; Koch and Dittmar, 2006), where C, H, and O represent the number of C, H, and O atoms in the molecule. 3.

Results and discussion

3.1 Water chemistry and groundwater flow Results of major constituents in canal and groundwater were illustrated in a piper diagram (Fig. 2). Cations in canal and groundwater samples were dominated by Ca2+ and Mg2+ (>80%), whereas Na+ and K+ accounted for less than 20% (Fig. 2). Bicarbonate was the dominating anion (>80%) in groundwater at the depths of both 10 and 25 m (Fig. 2). In canal water, Cl– and HCO3– took 20%–80% of anions, and SO42– was less than 20% in all water samples (Fig. 2). Fig. 2. Piper diagram illustrating the major constituents in canal and groundwater. At both sites, groundwater levels at a depth of 10 m were higher than those at a depth of 25 m over the year (Fig. 3a and d). This indicated that downward flow from 10 m to 25 m prevailed in the shallow aquifer over the whole monitoring period. Because of paddy field irrigation, the groundwater flow pattern at SY05 was different to that at SY07. Water levels of SY05-C (the canal near SY05, Fig. 1) were between those of 10-m and 25-m groundwater in the monsoon and lower than that of 25-m groundwater in the dry season, with only one exception in January 2015 (Fig. 3a). This suggests that 10-m groundwater at SY05 was likely discharged to the canal over the whole monitoring period, whereas the 25-m groundwater discharged to the canal only in dry season. The constant high levels of 9

10-m groundwater at SY05 in the monsoon were likely caused by paddy field irrigation using canal water. Therefore, irrigation was a key factor controlling the groundwater flow at SY05. In the monsoon, the canal both recharged the groundwater (via irrigation) and received groundwater. In the dry season, the canal only received groundwater because there was no irrigation for recharge. At SY07, water levels of SY07-C were apparently higher than those of groundwater at both depths (i.e., 10 and 25 m) during the monsoon (Fig. 3d). Similar to SY05, the water level of canal SY07-C decreased sharply in November and December 2014 (Fig. 3d). The inversion of the water level difference between SY07-C and the 10-m groundwater may cause reversal of groundwater flow between the canal and shallow groundwater. For the 25-m groundwater, the canal was a continuous recharge source with a relatively low recharge rate in the dry season because of a minimal hydraulic gradient. The variation trends of EC in canal water and 10-m groundwater at SY05 were opposite in the first 3-4 months of the irrigation period and were consistent with each other in the following months (Fig. 3b). The time lag of EC changes between canal water (used for irrigation) and 10-m groundwater was approximately 3-4 months, which indicated that there was rapid delivery of irrigation water to the shallow aquifer. A study in Bangladesh, where rice fields are ubiquitous, found that irrigation water could infiltrate deeply via bunds (the raised soil boundaries of the paddy fields) because they have higher permeability than the field pan (Neumann et al., 2011). At SY07, EC increased dramatically from 654 to 6570 μS cm−1 from April to September 2014 (Fig. 3f), which could be caused by disposal of wastewater with a high chloride content from the chemical plant (Fig. 1). The EC in 10-m groundwater at SY07 also increased and remained at high levels, with a range of 789–1085 μS cm−1 (Fig. 3f), but was lower than the EC in the canal water. However, when the water flow between SY07-C and the 10-m groundwater reversed, EC in the canal water fall rapidly to 600 μS cm−1. These 10

results suggested that shallow groundwater at SY07 likely has a good hydraulic link with the nearby canal. At SY05 and SY07, canal water temperatures fluctuated seasonally between 6.9 and 33.4 °C (Fig. 3c and g). The groundwater temperatures followed a similar seasonal trend with smaller fluctuations (15.7–24.2 °C) (Fig. 3c and g). At both sites, temperature fluctuations in the 10-m groundwater were greater than those in 25-m groundwater (Fig. 3c and g). This indicated that 10-m groundwater exchanged more with canal water than that at 25 m, which was consistent with the water level and EC results. Fig. 3. Temporal variations of water levels, EC, water temperatures and pH of canal water and groundwater at SY05 and SY07. In all subplots, 10-m and 25-m groundwater are indicated by blue squares and red circles, respectively. SY05-C (purple triangles) and SY07-C (green triangles) represent the canal water at SY05 and SY07, respectively. 3.2 Seasonal fluctuations of groundwater As Groundwater As concentrations at SY05 and SY07 ranged from 2 to 546 μg L−1 (average 121 μg L−1) and were higher than the World Health Organization (2011) drinking water limit (10 μg L−1) most of the time (Fig. 4a and e). The As concentrations in canal water were constantly low (< 5 μg L−1, Fig. 4a and e). Groundwater As concentrations at both sites fluctuated seasonally (Fig. 4a and e). A similar seasonal trend for groundwater As has been observed at other sites in the study area and the seasonal fluctuation was attributed to the fluctuation of redox conditions in groundwater caused by seasonal incursions of oxic surface water (Schaefer et al., 2017; Schaefer et al., 2016; Yu et al., 2018). Studies on the sediment column in this area have revealed that As was mainly absorbed on Fe(III) oxides because of the high As affinity on the Fe(III) mineral surface (Duan et al., 2017; Schaefer et al., 2017). 11

Groundwater As was mobilized from sediments through reductive dissolution of As-bearing Fe(III) oxides (Fendorf et al., 2010; Islam et al., 2004; Nickson et al., 2000; Tufano and Fendorf, 2008). Therefore, Fe(II) concentrations in groundwater also showed seasonal trends (Fig. 4b and f). The Fe(II) fluctuations were more gradual than those of As in groundwater (Fig. 4a, b, e and f), which could be explained by the formation of iron sulfide precipitates (Duan et al., 2019). The reductive dissolution of Fe(III) oxides and SO42– produces Fe(II) and S2−, which can coprecipitate as iron sulfide (Kirk et al., 2004; Kirk et al., 2010; Liu et al., 2019; Lowers et al., 2007). In this study, SO42– concentrations in groundwater ranged from <0.01 to 15.79 mg L–1, which were lower than those in canal water (15.68–46.92 mg L–1, Fig. 4c and g). Sulfate concentrations in groundwater increased in monsoon, indicating the input of SO42– from canal water (Fig. 4c and g). The decrease of SO42– in dry season in groundwater was likely caused by the SO42– reduction (Fig. 4c and g). Sulfide were also found in groundwater at SY05 and SY07, which may form insoluble iron sulfide with Fe(II) (Fig. 4d and h). When the monsoon came and irrigation began, the oxic recharge water precipitated the dissolved As through formation of Fe(III) oxides, which decreased the groundwater As concentrations (Biswas et al., 2014; Duan et al., 2019; Kulkarni et al., 2018b; Majumder et al., 2016; Schaefer et al., 2017). Fig. 4. Temporal variations of As and SO42– concentrations in groundwater and canal water at SY05 and SY07. Temporal variations of Fe(II) and sulfide concentrations in groundwater at SY05 and SY07. In all subplots, 10-m and 25-m groundwater are indicated by blue squares and red circles, respectively. SY05-C (purple triangles) and SY07-C (green triangles) represent the canal water at SY05 and SY07, respectively. 3.3 DOM transport in shallow groundwater

12

At SY05, DOC concentrations in the canal ranged from 0.23 to 24.10 mg L−1 (average 11.42 mg L−1, Fig. 5a). The DOC concentrations in 10-m groundwater were similar to those in the canal (range 1.35–45.33 mg L−1 and average 12.90 mg L−1, Fig. 5a). However, DOC in 10-m groundwater and canal water showed opposite variation trends (Fig. 5a), suggesting that canal water from irrigation was not the only source of groundwater DOM. The DOC concentrations in the 25-m groundwater were low and stable (range 0.83–4.48 mg L−1 and average 3.77 mg L−1, Fig. 5a). Sedimentary organic carbon was another potential source of groundwater DOM. The TOC contents of sediments from depths of 0 to 30 m ranged from 0.44 to 1.35 wt.% with the highest value appearing at a depth of approximately 22 m (Fig. 6a). Fig. 5. Temporal variations of DOC concentrations, δ13CDOC values and HCO3− concentrations in canal water and groundwater at SY05 and SY07. In all subplots, 10-m and 25-m groundwater are indicated by blue squares and red circles, respectively. SY05-C (purple triangles) and SY07-C (green triangles) represent the canal water at SY05 and SY07, respectively. The δ13CDOC values in canal water at SY05 ranged between −38.29‰ and −28.14‰ (average −31.84‰, Fig. 5b). The δ13CSOC values in the sediments at SY05 ranged from −25.89‰ to −20.85‰ (average −23.62‰, Fig. 6a), and were apparently richer in

13C

than that of groundwater DOC.

Groundwater δ13CDOC values at SY05 ranged from −35.55‰ to −25.87‰ (average −28.55‰, Fig. 5b), and showed no overlap with the δ13CSOC values in the sediments. This indicated that groundwater DOC was mainly sourced from canal water through irrigation. The highest δ13CDOC value (−25.87‰) in groundwater was higher than that in canal water (−28.14‰), indicating that sedimentary organic carbon also contributed to the groundwater DOC. The contributions of canal water and sedimentary organic carbon to groundwater DOC could be calculated using a mixing model based on mass balance: 13

𝛿13CGW = 𝑓CW𝛿13CCW + 𝑓Sed𝛿13CSed, 1 = 𝑓CW + 𝑓Sed,

(2) (3)

where δ13C is the stable carbon isotope value of DOC, and f is the proportion that canal DOC or sedimentary organic carbon contributes to groundwater DOC. The subscripts GW, CW, and Sed represent groundwater, canal water, and sediment, respectively. In the calculations, the δ13CDOC values of groundwater and canal water in the same season (monsoon or dry period) were averaged. Also, average δ13CSOC values of sediments above the depth of the groundwater were used to calculate the contribution of sedimentary organic carbon to groundwater DOC. The results were shown in Table 1. Fig. 6. Lithological logs, sediment TOC concentrations, and δ13CSOC values at SY05 and SY07. The TOC concentrations are shown as percentages of dry mass of the sediments. The measurement precision (± 0.3‰) of the δ13CSOC values is shown as error bars. The contribution of canal water DOC to 10-m groundwater DOC was 57.6% in the monsoon, which was 15.4% higher than that of sedimentary organic carbon (Table 1). The difference decreased to 8.4% in the dry period (Table 1). For groundwater at a depth of 25 m, the contribution of canal water DOC was 61% higher than that of sedimentary organic carbon (Table 1). This difference decreased greatly to 8.8% in dry period, which was similar to the situation for 10-m groundwater (Table 1). These results suggested that irrigation using canal water leached sedimentary organic carbon at depths from 0 to 10 m in the monsoon. However, this leaching effect may be weaker during infiltration from 10 to 25 m as indicated by the low contribution of sedimentary organic carbon at the depth of 25 m. This may be caused by the presence of clay layer between the depths of 15 and 20 m (Fig. 6a). Table 1. Contributions of canal water DOC and sedimentary organic carbon to groundwater DOC at different depths and in different seasons 14

At SY07, DOC concentrations in the canal water (SY07-C) varied from 2.84 to 35.27 mg L−1 (average 15.14 mg L−1, Fig. 5d). In 10-m groundwater, DOC concentrations ranged from 1.17 to 24.09 mg L−1 (average 8.56 mg L−1, Fig. 5d). The DOC concentrations in 25-m groundwater were slightly lower than those in 10-m groundwater most of the time (Fig. 5d). Unlike at SY05, the variation trends of groundwater DOC at depths of 10 and 25 m were similar to that of canal water DOC (Fig. 4c). This suggested that organic carbon from canal water and groundwater may influence each other because of the changing groundwater flow. Moreover, groundwater pumping by the nearby chemical plant was likely to accelerate the groundwater–surface water exchange (Schaefer et al., 2016). The δ13CDOC values in canal water were between −34.91‰ and −28.70‰ (average −32.05‰, Fig. 5e). The δ13CSOC values in sediments at SY07 ranged from −24.51‰ to −16.73‰ (average −22.69‰, Fig. 6b), and showed no overlap with canal water. In groundwater, the δ13CDOC values ranged from −32.9‰ to −25.09‰ (average −29.97‰, Fig. 5e). These values were between those in the canal water and sediments. The contributions of canal water and sedimentary organic carbon to groundwater DOC can be calculated using the mixing model showed in equation (2) and (3). The contributions of canal water DOC to groundwater DOC were apparently higher than those from sediments in the monsoon, and the difference ranged from 66.4% to 71% (Table 1). In the dry period, the difference decreased to 18.4% in 10-m groundwater and 51.8% in 25-m groundwater (Table 1), which was likely caused by reversal of groundwater flow between the canal and groundwater. 3.4 Molecular compositions of DOM in groundwater The availability of DOM in groundwater not only depended on the total abundance but also was influenced by the DOM reactivity. Only labile DOM can be used in microbe-mediated reductive dissolution of As-bearing Fe oxides (Mladenov et al., 2010; Neumann et al., 2010; Rowland et al., 15

2007). FTICR-MS is one of the most promising tools to molecularly identify uncharacterized DOM. In this study, groundwater samples with the largest seasonal fluctuations of As concentrations at each site (10-m groundwater at SY05 and 25-m groundwater at SY07, Fig. 4a and e) in the different seasons (i.e., monsoon and dry) were selected for FTICR-MS analysis. Spectra of the four groundwater samples showed approximately 3 000–6 000 peaks in the m/z range 100–1 000 (Fig. 7). The fractions of different formulas (CHO, CHON, CHOS, CHOP, CHONS, CHONP, CHOSP, and CHONSP) were shown in Fig. 6a. Obviously, the CHO formulas comprised the largest fraction (29.1%– 45.4%, Fig. 7a) in all samples. Together, CHO, CHON, and CHOS accounted for 45.1%–67.3% of the identified formulas (Fig. 6a). In 10-m groundwater at SY05 and 25-m groundwater at SY07, the As concentrations increased from 2 to 98 μg L−1 and from 313 to 416 μg L−1, respectively (Fig. 4a and e). During the same period, the number of CHO, CHON, and CHOS formulas decreased by 443, 394, and 197, respectively, in 10-m groundwater at SY05 (Fig. 7a). In 25-m groundwater at SY07, only the number of CHO formulas decreased, with a small change of 85 (Fig. 7a), and the CHON and CHOS formula numbers increased by 39 and 115, respectively (Fig. 7a). These results suggest that DOM with CHO formulas decreased most among all the DOM formulas as groundwater As increased at both sites from April to October. Therefore, DOM with CHO formulas was likely the key class consumed in the process of As mobilization. Fig. 7. Fractions of different (a) formulas of DOM and (b) compound classes of CHO molecular formulas identified by FTICR-MS. The x-axis shows the number of identified peaks. The sampling sites (SY05 or SY07) and depths (10 or 25 m) are shown on the y-axis. The months (April or October) of sampling are shown on the figure.

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The fractions of different classes of CHO formulas were presented in Fig. 7b. In the studied groundwater samples, the CRAM-like CHO formula was the most abundant component, accounting for 60.1%–65.5% (calculated using peak number) of the identified peaks. This result was consistent with the measurements in other studies, where CRAM-like compounds were important components of marine humic substances (Chen et al., 2014; Hedges et al., 1992) and freshwater DOM (Leenheer et al., 2003). Although CRAM-like compounds were thought to be resistant to degradation (Flerus et al., 2012; Hertkorn et al., 2006), the number of CRAM-like compounds decreased from 1 333 to 1 092 in 10-m groundwater at SY05 from April to October (Fig. 7b). The CRAM-like compounds consist of a complex mixture of carboxylated and fused alicyclic structures that are ultimately derived from biomolecules (Hertkorn et al., 2006). In this study, the apparent decrease of CRAM-like compounds and increase of groundwater As concentrations in the same period suggested that the CRAM-like compounds might be reused in biogeochemical processes, including As mobilization. This might be a third organic carbon source for groundwater in addition to the canal water and sediment. The CRAM-like DOM might involve in the complexation between As and DOM because they constitute a strong ligand for metal binding (Hertkorn et al., 2006). Also, the multiple coordination of CRAM-like DOM across cations could promote aggregation of colloid, which may affect the bioavailability of nutrients and trace metals including As (Chin et al., 1998). The aliphatic DOM, representing 9.7%–13.1% of the identified peaks in our samples, is a product of photochemical reactions such as long-term photobleaching (Stubbins et al., 2010). Therefore, aliphatic DOM input to the groundwater samples likely occurred during surface water infiltration. The number of aliphatic formulas decreased from 152 to 106 in 25-m groundwater at SY07 from April to October (Fig. 7b), which was consistent with the pattern of groundwater flow reversal from October to 17

December (Fig. 3e) and decrease in surface water recharge. The decrease of aliphatic DOM (214 to 174) in 10-m groundwater at SY05 (Fig. 7b) during the same period may be caused by the ceasing of irrigation. Tannin-like compounds are another class of photoresistant DOM and are the most abundant secondary metabolites produced by plants, representing up to 20% of the dry mass in plants (Adamczyk et al., 2017; Barbehenn and Constabel, 2011). Therefore, tannin-like compounds are abundant in terrestrial water (Helms et al., 2014; Maie et al., 2007). At our study sites, tannin-like DOM accounted for 12.9%–13.7% of the identified DOM peaks in the groundwater samples (Fig. 7b). Similar to aliphatic DOM, tannin-like DOM decreased by 20–69 peaks over the monitoring period (Fig. 7b), which was likely caused by lower input from surface water and biodegradation in groundwater. Aromatic DOM is photoreactive because the aromatic rings are photolabile, and it is expected to decrease by > 90% upon irradiation (Helms et al., 2014; Stubbins et al., 2010). Therefore, aromatic DOM was not found in many studies on fresh and seawater DOM. (Chen et al., 2014; Spencer et al., 2009). However, aromatic compounds can still be found in surface waters if they have short water residence times (Kellerman et al., 2015). In our study site, aromatic DOM was found in all the groundwater samples (Fig. 7b), indicating that sediments might be the main source of aromatic DOM in groundwater. The release of aromatic DOM from sediments were also found in Bangladesh (Mladenov et al., 2010) and West Bengal, India (Kulkarni et al., 2018b). At SY05, the aromatic DOM peaks in 10-m groundwater decreased from April to October (Fig. 7b). The simultaneous decrease in As and aromatic DOM in shallow groundwater were also observed in West Bengal, India and it was attributed to dilution with less aromatic recharge water from the surface or reduction during reductive dissolution of Fe(III) oxides (Kulkarni et al., 2018b; Mladenov et al., 2010). 18

Carbohydrate-like compounds are the most biodegradable component among the five classes of DOM (Carlson, 2002) and reportedly accelerate As mobilization in shallow aquifers (Rowland et al., 2009; Rowland et al., 2007; van Dongen et al., 2008). However, in the present study, only one peak was detected for carbohydrate-like formulas in 10-m groundwater at SY05 and 25-m groundwater at SY07 in October (Fig. 6b). One of the possible reasons might be the content of carbohydrate-like compounds tends to be underestimated because their ionization efficiency is suppressed compared with other components (Chen et al., 2011; Shen and Perreault, 1998; Stubbins et al., 2010). Nevertheless, the extremely low content of carbohydrate-like DOM may indicate that microbial activities that consume carbohydrate-like DOM were prevalent in the shallow aquifer in our study sites, leading to low stability of carbohydrate-like DOM. The high concentrations of HCO3– in groundwater (Fig. 5c and f), which is the production of microbial metabolism, was likely reflected the microbial consumption of organic matter including the carbohydrate-like DOM. 4.

Conclusions This study expands our knowledge of anthropogenic influences on the transport and molecular

properties of DOM in groundwater containing elevated concentrations of As. Seasonal hydraulic variations of canal water levels and irrigation were key factors controlling local groundwater flow patterns. The stable carbon isotope analysis demonstrated that canal water was an important source for DOM in shallow groundwater, especially in the monsoon when hydraulic gradients between canal water and groundwater were relatively high. Furthermore, infiltration of canal water inputted aliphatic and tannin-like DOM and leached aromatic DOM from sediments into groundwater. Therefore, groundwater recharge from canal water both inputted terrestrial DOM and accelerated the release of sedimentary DOM. Additionally, CRAM-like DOM, derived from biomolecules, might be reused in 19

biochemical processes during As mobilization. This insight is valuable for increasing our understanding of DOM availability and carbon cycles in groundwater. However, we certainly realize that the identified DOM classes may be incomplete and the classification method of DOM is rough, and further research is required to assess the reactivity of different types of DOM in the biochemical reactions regarding As mobilization. Acknowledgements This study is funded by the National Natural Science Foundation of China (No. 41472217, 41702275, 41702274), Ministry of Science and Technology of People’s Republic of China (No. 2014DFA20720) and China Scholarship Council (201406410035). References Adamczyk, B., Simon, J., Kitunen, V., Adamczyk, S., Smolander, A., 2017. Tannins and their complex interaction with different organic nitrogen compounds and enzymes: old paradigms versus recent advances. Chemistryopen 6, 610-614. Barbehenn, R.V. and Constabel, C.P., 2011. Tannins in plant-herbivore interactions. Phytochem. 72, 1551-1565. Bauer, M. and Blodau, C., 2006. Mobilization of arsenic by dissolved organic matter from iron oxides, soils and sediments. Sci. Total Environ. 354, 179-190. Benner, S.G., Polizzotto, M.L., Kocar, B.D., Ganguly, S., Phan, K., Ouch, K., Sampson, M., Fendorf, S., 2008. Groundwater flow in an arsenic-contaminated aquifer, Mekong Delta, Cambodia. Appl. Geochem. 23, 3072-3087.

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Table 1. Contributions of canal water DOC and sedimentary organic carbon to groundwater DOC at different depths and in different seasons. SY05 10 m

SY05 25 m

SY07 10 m

SY07 25 m

fS

fS

fS

fS

ed

Monsoon (April to August) Dry Period (September to March)

fC W

ed

fC W

ed

fC W

ed

fC W

4

5

1

8

1

8

1

8

2.4%

7.6%

9.5%

0.5%

4.5%

5.5%

6.8%

3.2%

4

5

4

5

4

5

2

7

5.8%

4.2%

5.6%

4.4%

0.8%

9.2%

4.1%

5.9%

40

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

Terrestrial DOM account for 54.2%–85.5% of groundwater DOM.



Irrigation using canal water inputs aliphatic and tannin-like DOM to groundwater.



Recharge from irrigation leaches aromatic DOM from sediments into groundwater.



Reuse of CRAM-like DOM might be an additional DOM source to groundwater.

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