Methylmercury production in soil in the water-level-fluctuating zone of the Three Gorges Reservoir, China: The key role of low-molecular-weight organic acids

Environmental Pollution 235 (2018) 186e196

Contents lists available at ScienceDirect

Environmental Pollution journal homepage: www.elsevier.com/locate/envpol

Methylmercury production in soil in the water-level-fluctuating zone of the Three Gorges Reservoir, China: The key role of low-molecularweight organic acids* Deliang Yin a, Yongmin Wang a, Tao Jiang a, c, Caiqing Qin a, Yuping Xiang a, Qiuyu Chen a, Jinping Xue a, Dingyong Wang a, b, * a b c

College of Resources and Environment, Southwest University, Chongqing 400715, China Chongqing Key Laboratory of Agricultural Resources and Environment, Chongqing 400715, China Department of Forest Ecology and Management, Swedish University of Agricultural Sciences, Umea, SE-90183, Sweden

a r t i c l e i n f o

a b s t r a c t

Article history: Received 13 September 2017 Received in revised form 1 December 2017 Accepted 20 December 2017

As important parts of dissolved organic matter, low-molecular-weight organic acids (LMWOAs) typically play important roles in desorbing Hg(II) from the soil solid-phase, which may directly or indirectly impact methylmercury (MeHg) production. However, the mechanism of these processes remains unclear. To better understand the effects of LMWOAs on Hg methylation in the soil, a field study was conducted to investigate the distribution of LMWOAs and their relationship with soil MeHg in a seasonally inundated area in the Three Gorges Reservoir (TGR), China. Meanwhile, laboratory simulation experiments were performed to determine the potential mechanism of LMWOAs in Hg methylation. The field investigation detected considerable amounts of LMWOAs in soil, among which tartaric acid and oxalic acid were dominant components. Among which, tartaric acid and oxalic acid were dominant components. Also, a seasonally and spatially heterogeneous distribution of LMWOAs in soil was observed. Notably, a significant positive relationship was found between MeHg concentrations and LMWOA pools in soil (r ¼ 0.969, p < .01), implying that LMWOAs could promote soil MeHg production. The simulation experiments confirmed that the MeHg levels in soil were largely elevated with the addition of LMWOAs, which occurred mainly in oxygen-deficient environment and was mediated by biotic factors. The soluble HgLMWOA complexes, which were formed by the enhanced desorption of Hg(II) from solid-phase, were mostly responsible for the elevated MeHg production in soil. Moreover, those LMWOAs with more carboxylic groups were believed to̊ enhance the net production of MeHg. The generated MeHg in sediment could diffuse into the overlying water, which thus poses a potential threat to the aquatic food web. Therefore, the enhanced Hg methylation caused by LMWOAs should be given more attention, especially in a seasonally inundated ecosystem, where the MeHg exposure is usually related to fishery activities. © 2017 Elsevier Ltd. All rights reserved.

Keywords: Low-molecular-weight organic acids Methylmercury Water-level-fluctuating zone Three Gorges Reservoir

1. Introduction Methylmercury (MeHg), an extremely toxic Hg species, can be easily accumulated and magnified in food chains in the aquatic ecosystem. The bioaccumulation factors of Hg in fish can reach up

* This paper has been recommended for acceptance by Dr. Harmon Sarah Michele. * Corresponding author. College of Resources and Environment, Southwest University, No. 2 Tiansheng Road, BeiBei District, Chongqing, 400715, China. E-mail address: [email protected] (D. Wang).

https://doi.org/10.1016/j.envpol.2017.12.072 0269-7491/© 2017 Elsevier Ltd. All rights reserved.

to 104 - 107 (Gustin et al., 2005), and fish consumption has been recognized to be the primary source of human exposure to Hg (Clarkson, 1998; Mergler et al., 2007). In the past few decades, elevated MeHg levels in water bodies and fish have been widely observed in newly-constructed reservoirs (Bodaly et al., 1997; Brigham et al., 2002; Kamman et al., 2005; Montgomery et al., 2000; Porvari, 1998). These elevated MeHg levels are primarily attributed to the influences of flooded soil and plants (Eckley et al., 2015; Hall et al., 2004, 2005; St Louis et al., 2004). Previous studies have reported that soil undergoing seasonal inundation is more conducive to the production of MeHg than is permanently

D. Yin et al. / Environmental Pollution 235 (2018) 186e196

inundated sediment (Eckley et al., 2015, 2017; Evers et al., 2007). This phenomenon is mainly attributed to the enhanced breakdown of organic matter (OM) in sediment experiencing water-level fluctuation, which can stimulate the activity of sulfate-reducing bacteria (SRB) (Eckley et al., 2015, 2017; Graham et al., 2012; Jones, 1998). These SRB are widely confirmed to be the principal Hg methylator, and depend greatly on anoxic conditions (Compeau and Bartha, 1985). Also, re-suspension of surface sediment particulates in the flooding process is another source of the elevated Hg levels in overlying water (Feng et al., 2011). The three Gorges Reservoir (TGR), one of the world's largest hydroelectric project, is a highly dynamic reservoir with its water level fluctuating between 145 m above sea level (a.s.l) in summer and 175 m a.s.l in winter, which results in a water-level-fluctuating zone (WLFZ) of approximately 350 km2 (Bao et al., 2015). In response to this disturbance of the water level, the plants in the WLFZ are gradually degraded after flooding, whereas they prosper when exposed to air. In the WLFZ, an above-ground biomass of plants of up to 1.50 kg m2 was found in our field investigation. These plants can release considerable amounts of dissolved OM (DOM) into the sediment, which can affect the solid-liquid distribution of inorganic Hg (Eckley et al., 2017; Haitzer et al., 2002; Hesterberg et al., 2001; Jing et al., 2007; Schartup et al., 2014). The biogeochemical cycling of Hg in soil and sediment is typically regulated by soil physicochemical and biogeochemical factors, such as microorganisms, temperature, sulfate, DOM, pH, and redox conditions (Boszke et al., 2003; Frohne et al., 2012; Shao et al., 2012; Ullrich et al., 2001). Of which, DOM seems to play a dual role in MeHg production as it can either bind Hg(II) to reduce its bioavailability or enhance the activity of Hg-methylatimg bacteria (Graham et al., 2012; Ullrich et al., 2001). Its varying effects on Hg bioavailability are possibly due to the fact that DOM is an extremely heterogeneous mixture containing abundant low to high molecular weight organic compounds exhibiting different solubility and reactivity (Liu et al., 2011; Ullrich et al., 2001). Recent studies have attributed the DOM-induced production of MeHg to the lowmolecular-weight OM (Siciliano et al., 2005; Yin et al., 2012). The degradation of OM, root secretion, and microbial metabolism can produce a large amount of low-molecular-weight organic acids (LMWOAs) in soil (Jones, 1998; Krzyszowska, 1996). In general, these LMWOAs, characterized as the essential components of DOM (Hagedorn et al., 2008; Sun et al., 2013), are believed to enhance the bioavailability and potential toxicity of Pb, Al, Cu, Zn, and Cd in soil systems (Bajda, 2011; Li et al., 2006; Liao et al., 2006; Qin et al., 2004; Wang and Mulligan, 2013). This is mostly attributed to the fact that LMWOAs contain multiple organic ligands (e.g., carboxyl groups), which can desorb the metal ions from the solid-phase to form soluble metal-LMWOA complexes (Krishnamurti et al., 1997; Li et al., 2005; Wang and Mulligan, 2013). The previous study has demonstrated that LMWOAs play positive roles in the desorption of soil-bound Hg(II) (Jing et al., 2007). Meanwhile, LMWOAs, recognized to be the bioavailable carbon, can greatly promote the activity of microorganisms in soil (Fischer et al., 2010; Gunina et al., 2014, 2017). Consequently, enhanced Hg methylation in the presence of LMWOAs may occur, which would pose a potential threat to aquatic organisms in the periodically flooded ecosystem. However, the distribution of LMWOAs and their potential effects on Hg methylation in soil experiencing water-level fluctuation are still unclear. To test the hypothesis that LMWOAs can mediate the methylation of Hg in the soil, this study was first to explore the temporal and spatial distribution of LMWOAs and their influences on MeHg production in the soil of the WLFZ; and then provided a mechanistic understanding of soil Hg methylation with the addition of LMWOAs. Considering that the MeHg bioaccumulation in aquatic food chains is affected by the water-level fluctuation (Bodaly et al.,

187

1997; Selch et al., 2007), a new indicator was provided in this study to predict the MeHg exposure risk in a seasonally flooded ecosystem. 2. Materials and methods 2.1. Study area and site description Geomorphologically, the WLFZ is distributed along both the mainstreams and tributaries of the Yangtze River, China. In the past few decades, the ecological and environmental safety of the TGR has greatly attracted wide public attention due to intensive nonpoint source pollution and rampant soil erosion (Zhang et al., 2009; Zhang and Lou, 2011). The WLFZ, however, is substantially shaped by both aquatic processes (e.g., oscillatory flow induced by waves during inundated periods) and terrestrial processes (e.g., rainfall erosion during non-flooded periods) (Bao et al., 2015). Additionally, the soil at different altitudes is generally subject to different flooding durations due to the operation of the TGR, which results in discernible differences in plant biomass and soil redox conditions. In this study, two representative WLFZ sites, Kaixian (KX, E108º29014.6”; N31º080 05.7”) and Zhongxian (ZX, 30 240 4800 N, 108 100 2600 E), were selected as the research areas. KX is located in a tributary while ZX is in the mainstream of the Yangtze River, China, both of which suffer from significant hydrological disturbance (Supplementary Information, SI, Fig. S1). 2.2. Sample collection and preparation Field sampling was carried out monthly from April to September (except for August) of 2016, in both ZX and KX sites, including three altitude ranges (155-165 m, 165-175 m, and >175 m), according to the changes of the water level due to hydrological adjustment. Samples were collected during the “dry period” of WLFZ. At each altitude, seven species of plants including Cynodon dactylon (L.) Pers., Bidens tripartita L., Xanthium sibiricum Patrin ex Widder. Polygonum lapathifolium L., Alternanthera philoxeroides (Mart.) Griseb., Setaria viridis (L.) Beauv. and Echinochloa crusgalli (L.) Beauv., were collected following the sample-handling protocol that any mechanical damage of plant tissues, such as root, stalk, and leaf, was not allowed in the sampling process. The corresponding bulk soil samples (about 15 cm away from the root surface), used for MeHg and LMWOAs analysis, were in-situ collected and preserved with liquid nitrogen (-196  C) (Chen et al., 2001). The plant roots were covered with undisturbed bulk soil (about 15 cm) prior to being cultivated, and then the bulk soil was gently shaken off back to the laboratory (<3 h). Samples of the soil adhering to the roots (a few millimeters away from the root surface) were collected using a water washing method (Milli-Q® deionized water) to represent rhizosphere soil (Xu et al., 2007). These samples were analysed for LMWOAs. Next, the fresh plants with complete roots were cultivated using deionized water in an artificial climate incubator to collect the root-excreted LMWOAs. The detailed description of the sample preparation, including bulk soil, rhizosphere soil and roots, is provided in the SI. For the simulation experiment, the bulk soil samples (0e20 cm) were collected from the 165-175 m a.s.l location at the ZX site to investigate the effects of LMWOAs on desorption and methylation of Hg(II). Soil samples were collected after removing plant residuals (e.g., roots), and then preserved by the aforementioned method. 2.3. Simulation experiment Soil samples for the simulation experiment were freeze-dried (-60  C), and ground into fine powder (through a 0.150 mm

188

D. Yin et al. / Environmental Pollution 235 (2018) 186e196

mesh). The essential physicochemical properties are listed in Table S1 of the SI. The OM contents were at a low level according to the classification standards of the National Soil Survey Office, China (Office of national soil survey, 1979). This possibly resulted from the moderate soil erosion of 35.08 t$ha1$a1 (unpublished data) in the TGR according to the standards for classification and gradation of soil erosion (The Ministry of Water Resources of the People's Republic of China, 2008). The soil total Hg (THg) concentration (45.09 ± 0.12 mg kg1) was slightly higher than the national background value of 38 mg kg1 (Feng and Qiu, 2008).

bottles (3 replicates  6 incubation intervals) were used throughout the incubation process. All the sterilized and unsterilized samples were placed in an artificial climate incubator at constant temperature (25  C), humidity (60%), and luminous intensity (0 mmol m2 s1). At selected time points (0, 6, 8, 9, 10, and 13 days), three replicates (bottles) were collected from each concentration gradient. The soil sample was freeze-dried (-60  C), and grounded into a fine powder (through a 0.150 mm mesh) for THg and MeHg analysis. 2.4. Solution extraction and chemical analysis

2.3.1. Desorption of Hg(II) Based on the results of field investigation, tartaric acid (TA) and oxalic acid (OA) were the predominant components in the detected LMWOAs (approximately 40% for each of TA and OA), while citric acid (CA) accounted for the lowest proportion, at 3%. Additionally, these three LMWOAs have different molecular structures from each other. Therefore, CA, OA and TA were selected for comparison, and each LMWOA was prepared at concentrations of 0, 1, 2, 4, 5, 6, and 8 mM. Next, 2 g of soil was weighted in a 50 mL polypropylene tube, followed by adding 20 mL of CA, OA or TA at the different concentrations. Triplicates (three different tubes) were conducted for each treatment, and thus, for each concentration gradient, a total of 21 tubes (7 incubation intervals) were needed in whole extraction process. These tubes were then shaken on a reciprocating shaker (220 rpm, 25  C) in dark conditions. At each selected time point (1, 5, 10, 30, 60, 90, and 120 min), three subsamples (tubes) from each concentration gradient were taken out and subsequently centrifuged at 4000 rpm for 30 min. Then, the supernatant in each tube was collected to determine the reactive Hg (RHg), which was operationally defined, and mainly represented the inorganic Hg species reduced directly by stannous-chloride (SnCl2) (Dalziel, 1995; Gill and Bruland, 1990). RHg has been recognized as a variety of dissolved inorganic Hg compound, mostly in Hg(II) state, available for microbial Hg(II)-methylation (Stumpner et al., 2013), and also as an important species of Hg in regulating the Hg dynamics in the aquatic systems (Fu et al., 2013; Jiang et al., 2017). 2.3.2. Methylation of Hg To determine the level of Hg methylation in the presence of LMWOAs, soil samples were divided into two aliquots. One was sterilized at 121  C for 120 min (i.e., the microbial community was eliminated), whereas the other was not (i.e., the microbial community was maintained). For each of the LMWOAs (i.e., CA, OA and TA), concentrations of 0, 2 and 8 mM were prepared using presterilized deionized water. Simultaneously, a mixture of CA, OA and TA with individual concentration of 2 mM was also prepared. Under aseptic conditions created with a UV-light, 5 g sterilized or unsterilized soil was weighed into a 100 mL borosilicate glass bottle (pre-cleaned at 500  C), followed by adding 3 mL prepared LMWOAs solution with different concentrations. Next, unsterilized soil samples were incubated under anaerobic and normal conditions, while sterilized samples were only incubated under anaerobic conditions. Specifically, the anaerobic conditions were created in an anaerobic incubator (YQX-1, Shanghai Yuejin Medical Instruments Co. Ltd, China) treated with pure N2, in which a frosted plug with Parafilm M® film were used to keep air and bacteria out of the culture bottles. To mimick the real environments that are not anaerobic but still air exposed (e.g., upland), a ‘normal’ condition for incubation was established. In this treatment, a breathable sealing film sterilized at 121  C for 120 min was used to allow normal air circulation, while avoid the potential invasion of external microorganisms. Triplicates (three different bottles) for each treatment were conducted, and thus, for each concentration gradient, a total of 18

The water-soluble low-molecular-weight organic acids, including CA, OA, TA, acetic acid (AA), propanedioic acid (PA), succinic acid (SA), and malic acid (MA) in the root exudates, the rhizosphere and the bulk soil, were determined using reversedphase high-performance liquid chromatography (HPLC, Shimadzu LC-20, Shimadzu, Osaka, Japan) featured with a diode array detector (van Hees et al., 1999). Soil THg concentrations were determined using a direct Hg analyzer (DMA-80, Milestone Inc. Shelton, CT), and soil MeHg concentrations were measured using a gas chromatography-cold vapor atomic fluorescence analyzer (GCCVAFS) (Model Ⅲ, Brooks Rand LLC, U.S.A). RHg in supernatants was detected using a CVAFS (Model 2500, Tekran Corporation, Inc. Canada). The elaborate description of the solution extraction, chemical analysis, and QA/QC is provided in the SI. 2.5. Statistical analysis Origin 9, Excel 2010 and SPSS 19.0 were used for statistical analysis. To test for differences, T-tests were used for data with a normal distribution, and Kruskal-Wallis (K-W) tests were used for data which were not normally distributed. For correlation analysis, Pearson's correlation coefficient (r) was used. In statistical hypothesis testing, the significance levels (p-value) were chosen at 0.05 and 0.01. A value of 0.05 is conventionally considered as “statistically significant”, and a value of 0.01 is “extremely significant” (Nuzzo, 2014). 3. Results and discussion 3.1. LMWOAs in root exudates, rhizosphere, and bulk soil The biochemical processes of LMWOAs in rhizosphere soil are highly dependent on the metabolic activity of plant roots. Higher levels of TA, OA, SA and AA in root exudates were observed compared with other LMWOAs (Fig. 1a). The distribution of rootexcreted LMWOAs derived from different plant species was provided in the SI (Fig. S2). LMWOA components in root exudates have been reported to be influenced by the plant species and ages (Grayston et al., 1997), however, except for SA and MA (K-W test, p < .05), the other LMWOAs showed no significant difference among the seven species of plants described above (K-W test, p > .05). This result indicated that the influence of plant species on the production of root LMWOAs from the WLFZ was limited. In addition, no significant influence on the distribution of LMWOAs in root exudates was observed among the sampling sites (K-W test, p > .05). The rhizosphere, a microzone only a few millimeters away from the root surface, supports high microbial activities owing to a continuous release of organic materials from the plant roots (Van
cura and Hovadík, 1965). In this study, the mean concentrations of TA and OA in rhizosphere soil were 19.76 ± 24.48 and 32.00 ± 46.34 mmol kg1 root dry weight (dw), respectively, which were significantly higher than the other LMWOA concentrations.

D. Yin et al. / Environmental Pollution 235 (2018) 186e196

189

Fig. 1. The distribution of LMWOAs in root exudates (a), rhizosphere soil (b), and bulk soil (c). The values are presented as the mean, with standard deviation as error bars, which are generated from all soil or plant samples collected at the three altitudes of 155e165 m, 165-175 m and >175 m in both ZX and KX. The extracted and washed LMWOAs per unit weight of root (dry weight) are used to represent the LMWOA concentrations in root exudates and rhizosphere soil, respectively. Abbreviations: OA ¼ oxalic acid; PA ¼ propanedioic acid; CA ¼ citric acid; TA ¼ tartaric acid; AA ¼ acetic acid; SA ¼ succinic acid; MA ¼ malic acid.

However, CA was distinguished as having the lowest mean concentrations of only 0.30 ± 0.48 mmol kg1 root dw. As depicted in Fig. 1a, b, the dominant/non-dominant LMWOAs in the rhizosphere soil showed great consistency with those in root exudates, indicating that the potential influence of root metabolism on the distribution of LMWOAs in the rhizosphere zone was possible. Moreover, regardless of the influence of plant species on the rootexcreted LMWOAs, in the rhizosphere soil, concentrations of OA, CA, TA, PA, and SA showed significant differences (K-W test, p < .01). This result indicated that the distribution patterns of most LMWOAs in the rhizosphere soil possibly depended on the rhizosphere environment, such as adsorption/desorption of LMWOAs on the inorganic mineral solid-phase (e.g., iron oxides) and the bac€derberg, 2003; terial activity and community composition (So Wieland et al., 2001). However, other influencing factors such as the sampling sites and seasons were not significant (K-W test, p > .05). The mean concentrations of LMWOAs in bulk soil gradually increased in the order of CA (23.64 ± 23.26 m
similar. Although the bulk soil was away from the rhizosphere zone, a potential influence derived from the metabolic activities of roots was still possible. Nevertheless, the plant species showed no significant influence on the distribution of LMWOAs in bulk soil (K-W test, p > .05). The distribution patterns of LMWOAs in the bulk soil in response to variation of altitude are described in the SI (Fig. S3). The results of a one-way ANOVA showed a significant difference of the total LMWOA concentrations in the soil among the three altitudes (p < .01). Meanwhile, the total LMWOA levels in the soil at 165175 m a.s.l were noticeably higher than those at 155-165 m a.s.l (Ttest, p < .05), which was very consistent with the distribution of plant biomass reported in a previous study (Liang et al., 2016). According to the operation of the TGR, the average annual flooding duration period of soil at 155-165 m a.s.l was longer than that at 165-175 m a.s.l (the gap was approximately 100 d), which resulted in a relatively high plant biomass growth/accumulation at 165175 m a.s.l. Therefore, the water-level fluctuation caused a significant change in the distribution of plants, which subsequently impacted the spatial-temporal distribution of soil LMWOAs. More importantly, the temporal variation of soil LMWOAs was statistically significant (K-W test, p < .01). In this study, the total LMWOA concentrations in bulk soil increased from 821.73 ± 705.58 mmol kg1 dw in April to

190

D. Yin et al. / Environmental Pollution 235 (2018) 186e196

2953.77 ± 1365.13mmol kg1 dw in July, with monthly increments, but subsequently decreased to 1345.65 ± 820.52 mmol kg1 dw in September, which showed a similar pattern with soil temperature (r ¼ 0.89, p < .05). This might be explained by the influence of temperature on the bio-degradation of OM. The rates of soil OM decomposition caused by microorganisms were accelerated with elevated soil temperature (Min et al., 2014). Hence, soil temperature could be a crucial factor controlling the production of LMWOAs. Although the ratio of LMWOAs to dissolved organic carbon is usually low in soil (van Hees et al., 1999), an understanding of the LMWOA pool (the summation of all detected LMWOAs from the bulk soil at each sampling site) may provide insight into the critical roles that LMWOAs play in the biogeochemical cycling of Hg in soil ecosystems. In regards to the potential role of LMWOAs in the environmental fate of contaminants, the LMWOA pool significantly increased from 0.06 ± 0.05 mg cm3 in April to 0.28 ± 0.16 mg cm3 in July with monthly increments, and then was reduced to 0.13 ± 0.08 mg cm3 in September (Fig. 2), possibly indicating that higher bioavailability of Hg may occur in summer due to the increasing LMWOA pool (Jing et al., 2007; Qin et al., 2015; You et al., 2016). In contrast, higher microbial degradation of MeHg was not observed in warm months (Fig. 2), although temperature increased, which further suggests that the net production of MeHg co-varied with LMWOA variation, which is discussed in detail in the following sections. 3.2. MeHg in field soil Covariance between the LMWOA pool and MeHg concentrations in field soil was established (Fig. 2), and the MeHg was correlated with LMWOA pool within the sampling period (r ¼ 0.97, p < .01), suggesting that the presence of LMWOAs could play a positive role in the net MeHg production. This result could be attributed to the elevated bioavailability of Hg caused by LMWOAs (Jing et al., 2007), which enhanced bonded Hg(II) release from the soil solid-phase and indirectly stimulated the subsequent Hg methylation of microorganisms (Ullrich et al., 2001). Soil temperature showed a significant positive relationship with the LMWOA pool (r ¼ 0.89, p < .05), indicating that moderately high temperature was likely to favor the production of LMWOAs. However, we found no significant relationship between temperature and MeHg concentrations in soil (p > .05), although they showed a similar distribution pattern of temperature vs. LMWOAs (Fig. 2), indicating that soil temperature might affect the production of MeHg in soil through an indirect

Fig. 2. Relationship between LMWOA pool and MeHg concentrations in the bulk soil. The values are presented as the mean, with standard deviation as error bars.

pathway (e.g., through microbial metabolism or other enzymatic processes). The considerable pools of LMWOAs in the soil, together with the fact that different organic acids have a different composition of ligands, are drawing our attention to looking further into how LMWOAs affect Hg methylation, and which one is the predominant factor in the TGR areas. 3.3. Hg methylation in soil with the addition of LMWOAs As illustrated in Fig. 3 and Fig. 4, Hg methylation in unsterilized and sterilized soils with the addition of LMWOAs was studied under either anaerobic or normal conditions. For unsterilized soil, according to the statistical results, net increases of MeHg in the deionized water treatment (i.e., CK group) were 358.56 ± 17.27 and 312.14 ± 24.34 ng kg1 after anaerobic and normal incubation, respectively. In comparison to the CK group, after anaerobic incubation, soil MeHg concentrations were significantly elevated when 2 mM and 8 mM CA, OA and TA were added (T-test, p < .01, Fig. 3a, b and c). When 2 mM LMWOA was applied under anaerobic conditions, the highest net increase of MeHg reached 837.29 ± 51.64 ng kg1 in the CA treatment, while it was only 627.91 ± 31.58 ng kg1 and 621.17 ± 13.29 ng kg1 in the OA and TA treatments, respectively. As expected, the net increases of MeHg increased as the LMWOA concentration was raised to 8 mM, and the incremental order of OA (668.04 ± 19.33 ng kg1) < TA (708.40 ± 28.99 ng kg1) < CA (1047.04 ± 64.31 ng kg1) occurred under anaerobic incubation. In fact, microbial community and redox conditions were believed to differ substantially between normal and anaerobic soil, which might affect the Hg methylation efficiency. Under normal conditions, although the MeHg concentrations still increased after incubation, only 2 and 8 mM CA treatments had significantly elevated MeHg production compared to the CK group (T-test, p < .05). The methylation potential of Hg was lower than that under anaerobic conditions in the presence of LMWOAs (Fig. 3a, b and c). These results indicated that Hg(II) was more effectively methylated to MeHg in an anaerobic environment, especially in the presence of LMWOAs. The low Hg methylation capacity in normal conditions could be ascribed to the inhibitory activity of the reduced microbial community (e.g., SRB) (Ullrich et al., 2001). Compared to the LMWOAs applied alone, the coexistence of OA, CA and TA exhibited no synergistic effects in soil Hg methylation under both anaerobic and normal conditions (Fig. 3d). Although the quantities of both organic ligands and bioavailable carbon were elevated with the addition of the mixture including the three different LMWOAs, the abundance of the microbial community was possibly deficient, and more likely to become one of the limiting factors controlling MeHg production. However, for sterilized soil, regardless of the LMWOA amounts added under anaerobic conditions, the net increase of MeHg was generally less than 70 ng kg1 (Fig. 4). The additions of CA, OA, and TA in sterilized soil were hardly expected to increase the production of MeHg as compared to those additions in unsterilized soil. This result provided further evidence that the elevated MeHg production caused by LMWOAs was mostly dependent on biotic processes. In conclusion, the LMWOA input was demonstrated to be an important factor enhancing the Hg methylation in the WLFZ soil, whereas increasing MeHg production in the unsterilized soil was limited when LMWOA concentrations were raised from 2 to 8 mM (T-test, p > .05) possibly due to the deficient abundance of the microbial community. Interestingly, low CA concentrations relative to OA and TA were found in bulk soil, whereas CA effectively increased MeHg production more than did the other LMWOAs, probably due to its higher number of carboxylic groups (Jing et al., 2007). Additionally, the LMWOAs with bigger molecular weight contained

D. Yin et al. / Environmental Pollution 235 (2018) 186e196

191

Fig. 3. MeHg concentrations in unsterilized soil as a function of incubation time with the addition of CA (a), OA (b), TA (c), and their mixture (d) in anaerobic and/or normal conditions. Solid and dotted lines represent the anaerobic and normal process, respectively. The CK represents the deionized water treatment. The values are presented as the mean, with standard deviation as error bars.

more negative charge, which could favor the desorption of soilbound Hg(II) (Jing et al., 2007). The molecular weight of CA, is largest, while OA is the smallest, and hence the stimulating effects of CA on Hg methylation were mostly significant, while OA had the lowest effect. Therefore, those LMWOAs with more carboxylic groups and bigger molecular weight had the higher capacity to convert inorganic Hg to MeHg.

3.4. Methylation capacity in soil MeHg normalized to THg (%MeHg), a net production of MeHg, was generally used to reflect the potential Hg methylation efficiency in an environmental system (Eckley et al., 2017; Sunderland et al., 2006). The initial %MeHg in the soil, used to understand MeHg production with the addition of LMWOAs, was 0.55 ± 0.02%, which was characterized as a low level of net Hg methylation. Under anaerobic conditions, the maximal value of %MeHg (% MeHgmax) in the unsterilized soil was 1.37 ± 0.05% in the CK group, whereas it was significantly elevated to 2.82 ± 0.13%, 2.30 ± 0.10% and 2.18 ± 0.08% in the presence of 2 mM CA, OA and TA, respectively (T-test, p < .01). Although the elevated %MeHgmax (3.25 ± 0.15% for CA, 2.48 ± 0.08% for OA, and 2.56 ± 0.1% for TA) was observed when LMWOA concentration was continually raised to

8 mM, the magnitude of increase was minor, without significant differences (T-test, p > .05). Furthermore, compared to the effects of LMWOAs (2 mM) applied alone, synergistic effects of CA, OA, and TA in Hg methylation did not seem to occur (%MeHgmax ¼ 3.02 ± 0.12%). Although the elevated %MeHg values in normal conditions were also observed when both 2 and 8 mM LMWOAs were added, they were generally lower than that in the anaerobic soil. D%MeHgmax, the difference between the %MeHgmax and the initial value of %MeHg, can be used to estimate an increase in Hg methylation. Both in anaerobic and normal conditions, the D% MeHgmax in unsterilized soil treated with CA was higher than that treated with OA and TA. The statistical results depicted in Table 1 confirmed that the soil Hg methylation in the WLFZ of ZX was enhanced mainly with the addition of LMWOAs. Considering that the soil D%MeHgmax under normal incubation was estimated to account for approximately 69.31 ± 6.99% of that under anaerobic conditions, the Hg methylation in soil in the WLFZ is possibly enhanced after flooding due to the formation of an oxygendeficient environment. However, for sterilized soil, although an anaerobic process was performed, the D%MeHgmax only accounted for 8.01 ± 1.24% of that in the unsterilized soil, which confirmed the limited Hg methylation in the absence of microorganisms. In

192

D. Yin et al. / Environmental Pollution 235 (2018) 186e196

Fig. 4. MeHg concentrations in sterilized soil as a function of incubation time with the addition of CA (a), OA (b), TA (c), and their mixture (d) in anaerobic condition. The CK represents the deionized water process. The values are presented as the mean, with standard deviation as error bars.

Table 1 Differences between the %MeHgmax and the initial values (D%MeHgmax) under anaerobic and/or normal incubation for unsterilized a and sterilized soil b. The CK represents the deionized water process. Mixture represents the coexisting system with 2 mM each of CA, OA, and TA. Values are presented as the mean, with standard deviation as error bars.

D%MeHgmax

Anaerobic Normal a Anaerobic

a

b

CK

Mixture

CA

OA

2 (mM)

8 (mM)

2 (mM)

8 (mM)

2 (mM)

8 (mM)

0.82 ± 0.01 0.70 ± 0.01 0.00 ± 0.00

2.48 ± 0.27 1.58 ± 0.21 0.16 ± 0.00

2.28 ± 0.36 1.48 ± 0.21 0.14 ± 0.01

2.72 ± 0.15 1.66 ± 0.21 0.22 ± 0.00

1.76 ± 0.06 1.23 ± 0.14 0.16 ± 0.01

1.94 ± 0.05 1.32 ± 0.14 0.18 ± 0.02

1.64 ± 0.03 1.34 ± 0.11 0.14 ± 0.01

2.02 ± 0.05 1.42 ± 0.11 0.17 ± 0.00

conclusion, both microorganisms (e.g., SRB) and nutrients (e.g., LMWOAs) in a low-oxygen condition played key roles in controlling the Hg methylation rates. 3.5. Kinetics of Hg methylation in soil In comparison with other fitting curves based on the regression analysis, the logistic regression curve perfectly fitted the kinetic process of MeHg production in the unsterilized soil in the presence of LMWOAs (Table 2) (coefficient of determination (R2) > 0.90). This provides evidence that MeHg production sequentially undergoes gentle, logarithmic, and recessionary/stationary periods, which is very consistent with microbial growth patterns (S-curve). Hg methylation rates were higher in anaerobic environments than

TA

those in normal conditions, which was further confirmed by the reaction rate constant (k) values. Furthermore, in the later period of anaerobic incubation, reduced MeHg concentrations were observed, while this trend was not apparent in the normal conditions. In the oxygen-deficient environment, the high competitive capacity of Hg-methylating anaerobes (e.g., SRB) for substrates € nheit, 1983) might be responsible for the net (Kristjansson and Scho accumulation of MeHg from day 0e10. Nevertheless, microbial degradation involving methanogens, recognized as the important demethylating process in anaerobic sediments, could convert MeHg to CH4/CO2 (Marvin-Dipasquale and Oremland, 1998; Oremland et al., 1991; Pak and Bartha, 1998), resulting in the decrease of MeHg in the later incubation period. In a normal environment, however, the methanogens were limited to low

D. Yin et al. / Environmental Pollution 235 (2018) 186e196

193

Table 2 Kinetics of MeHg production in soil with the addition of LMWOAs, where Ct is the MeHg concentration. The k and t represent the reaction rate constants and incubation time, respectively. Mixture indicates the co-existing CA, OA, and TA. The CK represents the deionized water process. Treatments

CK CA

Conditions

Anaerobic Normal Anaerobic Normal

OA

Anaerobic Normal

TA

Anaerobic Normal

Mixture **

Anaerobic Normal

Concentrations (mM)

0 0 2 8 2 8 2 8 2 8 2 8 2 8 2þ2þ2 2þ2þ2

Ct ¼ a= ð1 þ b  expðk  tÞÞ k

R2

0.30 0.21 0.40 0.43 0.23 0.30 0.42 0.41 0.22 0.24 0.40 0.46 0.21 0.21 0.44 0.29

0.97** 0.74** 0.93** 0.94** 0.93** 0.92** 0.89** 0.89** 0.93** 0.94** 0.90** 0.89** 0.92** 0.91** 0.92** 0.93**

Correlation is significant at the 0.01 level.

activity (Kumar et al., 2011), and then a plateau was reached from the 10th day.

3.6. Possible mechanism mediated by LMWOAs In general, ambient Hg methylation is typically a biotic process, whereas it is largely mediated by many abiotic factors, such as the

concentrations of bioavailable Hg(II), OM content, temperature, sulfide, and redox conditions (Li and Cai, 2013; Ullrich et al., 2001). RHg has been increasingly recognized as a variety of inorganic Hg compound in mainly the free Hg(II) state, which represents the Hg pool bioavailable for methylating microorganisms (Beaulieu et al., 2014; Cristina and Jardim, 2003; Stumpner et al., 2013). As depicted in Fig. 5, elevated RHg concentrations were found in water with

Fig. 5. Desorption of Hg(II) in soil with the addition of CA (a), OA (b), and TA (c). The CK represents the deionized water group. The values are presented as the mean, with standard deviation as error bars.

194

D. Yin et al. / Environmental Pollution 235 (2018) 186e196

Table 3 Desorption kinetics of Hg(II) from the solid phase, where Ct is the Hg(II) concentrations in supernatant; k and t represent the reaction rate constant and extraction time, respectively. The CK represents the deionized water process. Treatments

Concentrations (mM)

Time (min)

CK

0

0e120

CA

1 2 4 5 6 8

0e120 0e60 0e60 0e30 0e30 0e10

2 4 5 6 8

60e120 60e120 30e120 30e120 10e120

1 2 4 5 6 8

0e120 0e120 0e90 0e60 0e60 0e30

5 6 8

60e120 60e120 30e120

1 2 4 5 6 8

0e120 0e120 0e90 0e60 0e60 0e30

5 6 8

60e120 60e120 30e120

TA

OA

** *

,

Ct ¼ k  t

Ct ¼ a  expðk  tÞ þ b 2

k

R

0.07

0.99**

1.13

0.99**

k

R

0.02 0.02 0.02 0.07 0.06

0.95** 0.99** 0.90** 0.99* 0.99**

0.01 0.01 0.01 0.03 0.01 0.05 0.06 0.05

0.34

2

Ct ¼ b  tk k

R2

0.91 0.86 0.49 0.60 0.30

0.91* 0.97* 0.99** 0.96* 0.97**

0.27

0.98**

0.47

0.99**

0.99** 0.98** 0.99** 0.99** 0.99* 0.99*

0.99** 0.92* 0.01 0.02 0.02 0.01 0.01 0.06

0.93** 0.96** 0.98** 0.99** 0.99** 0.99**

0.01

0.99**

0.99**

Correlation is significant at the 0.01 and 0.05 level, respectively.

the addition of LMWOAs compared to those with deionized water, indicating that the inorganic Hg(II) levels largely increased due to the application of LMWOAs. As expected, the desorption rates of Hg(II) were accelerated at higher concentrations of LMWOAs (Fig. 5). Additionally, compared to TA and OA, the CA treatment showed the highest desorption amounts of Hg(II) from the soil, suggesting that CA had the most critical role in the Hg(II)-LMWOA interaction. Furthermore, a desorption experiment of reactive Hg(II) (i.e., RHg) by addition of LMWOAs was conducted to determine the capacity of LMWOAs to dissociate Hg(II) from the soil solid-phase, which had an important implication for the free inorganic Hg(II) available for microbial methylation. As shown in Table 3, the desorption process of Hg(II) was divided into two stages including a rapid desorption at the beginning and a slow desorption at a later period. The rising phase was perfectly described as an exponential function (R2 > 0.90), whereas the falling trend was ambiguous (Table 3). The dramatic variation of RHg levels was speculated to be the net result of combined effects, including ion-exchange and ligand complexation. The addition of excessive LMWOAs could provide enough hydrogen ions (Hþ) (Kortum et al., 1961; Rukshana et al., 2012), which could participate in ion-exchange with soilbound Hg(II). Importantly, the organic acidic ligands in LMWOAs, such as carboxyl groups, were expected to be involved in a complexation reaction with soil-bound Hg(II) and free Hg(II) to form soluble Hg-LMWOA complexes (Jing et al., 2007; Qin et al., 2004). The ion-exchange and ligand complexation were speculated to coincide. However, considering that a dramatic decrease of RHg concentrations occurred in the water phase within the

extraction time, soluble Hg(II)-LMWOA complexes were more likely to be the dominant species of Hg in the water-phase after the desorption reaction. In summary, LMWOAs possibly had two-fold roles in affecting soil Hg methylation: 1) the desorption of soil-bound Hg(II) and the formation of soluble Hg-LMWOA complexes could efficiently increase the bioavailability of Hg species (i.e., affect Hg speciation) (Jing et al., 2007; Li et al., 2006); and 2) LMWOAs, as a labile carbon source, were more available to the methylating bacteria, and thus could greatly promote the methylation capacity of a series of bacteria including SRB (i.e., affect microbial metabolism) (Compeau and Bartha, 1985; King et al., 2001; Liu et al., 2017). Therefore, the combined effects of Hg(II) desorption and the carbon source for microbial metabolism were responsible for the elevated MeHg production in the soil when LMWOAs were added in this study, especially in anaerobic conditions. 4. Conclusions The spatial-temporal distribution of different LMWOAs in the field soil of the WLFZ area was characterized, with oxalic acid and tartaric acid being the dominant species. As compared to the LMWOAs in the rhizosphere zone, the distribution patterns of LMWOAs in bulk soil greatly depend on seasonal variation. High soil temperature and aboveground biomass can favor the production of LMWOAs in the soil. Additionally, the methylation experiment illustrates the enhanced Hg methylation in soil after addition of LMWOAs, because of their roles in desorbing Hg(II) from the soil solid-phase, and in acting as a carbon source to be utilized by

D. Yin et al. / Environmental Pollution 235 (2018) 186e196

methylating bacteria. Furthermore, the results of this study suggest an important role of LMWOAs in soil that is subject to seasonal water-level fluctuation in the TGR area. Hence, the potential for MeHg production in soil of these areas should be re-evaluated during the whole water-level-fluctuating period. This includes 1) when exposed to air (i.e., the “dry period”), the root-excreted LMWOAs will promote MeHg production in the rhizosphere soil, which can partially diffuse into overlying water after flooding; and 2) during flooding period (i.e., the “wet period”), the LMWOAs derived from the decomposition of flooded plants may act as microbial energy sources and greatly favor microbial Hg methylation. Importantly, these results may partially help explain the elevated MeHg levels in newly-constructed reservoirs, especially in areas with flooding-drying cycles. Acknowledgments This research was sponsored by the National Basic Research Program of China (2013CB430004) and the National Natural Science Foundation of China (41373113 and 41403079). Dr. Tao Jiang would like to thank the Swedish Research Council (VR) program (621-2014-5370 and D697801) for generously sponsoring his research position at the Swedish University of Agricultural Sciences. He also thanks Dr. Jeffra Schaefer of Rutgers University for her valuable comments and suggestions in their private communication. Great appreciations to Dr. Carol Frost of Swedish University of Agricultural Sciences are necessary for her valuable helps in the English editing. Finally, all three anonymous reviewers are greatly appreciated for improving the manuscript quality. Appendix A. Supplementary data Supplementary data related to this article can be found at https://doi.org/10.1016/j.envpol.2017.12.072. References Bajda, T., 2011. Dissolution of mimetite Pb5(AsO4)3Cl in low-molecular-weight organic acids and EDTA. Chemosphere 83, 1493e1501. Bao, Y.H., Gao, P., He, X.B., 2015. The water-level fluctuation zone of Three Gorges Reservoir d a unique geomorphological unit. Earth Sci. Rev. 150, 14e24. Beaulieu, E., Marvin-Dipasquale, M.C., Alpers, C.N., Fleck, J., 2014. 'Reactive' Inorganic Mercury: a Critical Examination of Preservation and Storage Techniques (AGU Fall Meeting). Bodaly, R.A., St Louis, V.L., Paterson, M.J., Fudge, R.J.P., Hall, B.D., Rosenberg, D.M., Rudd, J.W.M., 1997. Bioaccumulation of mercury in the aquatic food chain in newly flooded areas. Met. Ions Biol. Syst. 34, 259e287. Boszke, L., Kowalski, A., Glosinska, G., Szarek, R., Siepak, J., 2003. Environmental factors affecting speciation of mercury in the bottom sediments; an overview. Pol. J. Environ. Stud. 12, 5e13. Brigham, M.E., Krabbenhoft, D.P., Olson, M.L., DeWild, J.F., 2002. Methylmercury in flood-control impoundments and natural waters of northwestern Minnesota, 1997e99. Water Air Soil Pollut. 138, 61e78. Chen, M.C., Wang, M.K., Chiu, C.Y., Huang, P.M., King, H.B., 2001. Determination of low molecular weight dicarboxylic acids and organic functional groups in rhizosphere and bulk soils of Tsuga and Yushania in a temperate rain forest. Plant Soil 231, 37e44. Clarkson, T.W., 1998. Human toxicology of mercury. J. Trace Elem. Exp. Med. 11, 303e317. Compeau, G.C., Bartha, R., 1985. Sulfate-reducing bacteria: principal methylators of mercury in anoxic estuarine sediment. Appl. Environ. Microbiol. 50, 498e502. Cristina, B.M., Jardim, W.F., 2003. Production of organic mercury from Hg0: experiments using microcosms. J. Braz. Chem. Soc. 14, 244e248. Dalziel, J.A., 1995. Reactive mercury in the eastern north-atlantic and southeast atlantic. Mar. Chem. 49, 307e314. Eckley, C.S., Luxton, T.P., Goetz, J., Mckernan, J., 2017. Water-level fluctuations influence sediment porewater chemistry and methylmercury production in a flood-control reservoir. Environ. Pollut. 222, 32e41. Eckley, C.S., Luxton, T.P., Mckernan, J.L., Goetz, J., Goulet, J., 2015. Influence of reservoir water level fluctuations on sediment methylmercury concentrations downstream of the historical Black Butte mercury mine, OR. Appl. Geochem. 61, 284e293. Evers, D.C., Han, Y.J., Driscoll, C.T., Kamman, N.C., Goodale, M.W., Lambert, K.F.,

195

Holsen, T.M., Chen, C.Y., Clair, T.A., Butler, T., 2007. Biological mercury hotspots in the northeastern United States and southeastern Canada. Bioscience 57, 29e43. Feng, X.B., Bai, W.Y., Shang, L.H., He, T.R., Qiu, G.L., Yan, H.Y., 2011. Mercury speciation and distribution in Aha Reservoir which was contaminated by coal mining activities in Guiyang, Guizhou, China. Appl. Geochem. 26, 213e221. Feng, X.B., Qiu, G.L., 2008. Mercury pollution in guizhou, southwestern China - an overview. Sci. Total Environ. 400, 227e237. Fischer, H., Ingwersen, J., Kuzyakov, Y., 2010. Microbial uptake of low-molecularweight organic substances out-competes sorption in soil. Eur. J. Soil Sci. 61, 504e513. Frohne, T., Rinklebe, J., Langer, U., Du Laing, G., Mothes, S., Wennrich, R., 2012. Biogeochemical factors affecting mercury methylation rate in two contaminated floodplain soils. Biogeosciences 9, 493e507. Fu, X.W., Feng, X.B., Yin, R.S., Zhang, H., 2013. Diurnal variations of total mercury, reactive mercury, and dissolved gaseous mercury concentrations and water/air mercury flux in warm and cold seasons from freshwaters of southwestern China. Environ. Toxicol. Chem. 32, 2256e2265. Gill, G.A., Bruland, K.W., 1990. Mercury speciation in surface freshwater systems in California and other areas. Environ. Sci. Technol. 24, 1392e1400. Graham, A.M., Aiken, G.R., Gilmour, C.C., 2012. Dissolved organic matter enhances microbial mercury methylation under sulfidic conditions. Environ. Sci. Technol. 46, 2715e2723. Grayston, S.J., Vaughan, D., Jones, D., 1997. Rhizosphere carbon flow in trees, in comparison with annual plants: the importance of root exudation and its impact on microbial activity and nutrient availability. Appl. Soil Ecol. 5, 29e56. Gunina, A., Dippold, M.A., Glaser, B., Kuzyakov, Y., 2014. Fate of low molecular weight organic substances in an arable soil: from microbial uptake to utilisation and stabilisation. Soil Biol. Biochem. 77, 304e313. Gunina, A., Smith, A.R., Kuzyakov, Y., Jones, D.L., 2017. Microbial uptake and utilization of low molecular weight organic substrates in soil depend on carbon oxidation state. Biogeochemistry 1e12. Gustin, M.S., Saito, L., Peacock, M., 2005. Anthro-pogenic impacts on mercury concentrations and ni-trogen and carbon isotope ratios in fish muscle tis-sue of the Truckee River watershed, Nevada. USA. Science of the Total Environment 347, 282e294. Hagedorn, F., van Hees, P.A.W., Handa, I.T., H€ attenschwiler, S., 2008. Elevated atmospheric CO2 fuels leaching of old dissolved organic matter at the alpine treeline. Global Biogeochem. Cycles 22, GB2004. Haitzer, M., Aiken, G.R., Ryan, J.N., 2002. Binding of mercury(II) to dissolved organic matter: the role of the mercury-to-DOM concentration ratio. Environ. Sci. Technol. 36, 3564e3570. Hall, B.D., St Louis, V.L., Rolfhus, K.R., Bodaly, R.A., Beaty, K.G., Paterson, M.J., Cherewyk, K.A., 2005. Impacts of reservoir creation on the biogeochemical cycling of methyl mercury and total mercury in boreal upland forests. Ecosystems 8, 248e266. Hall, B.D., St Louis, V.L., Bodaly, R.A., 2004. The stimulation of methylmercury production by decomposition of flooded birch leaves and jack pine needles. Biogeochemistry 68, 107e129. Hesterberg, D., Chou, J.W., Hutchison, K.J., Sayers, D.E., 2001. Bonding of Hg(II) to reduced organic sulfur in humic acid as affected by S/Hg ratio. Environ. Sci. Technol. 35, 2741e2745. Jiang, T., Chen, X.S., Wang, D.Y., Liang, J., Bai, W.Y., Zhang, C., Wang, Q.L., Wei, S.Q., 2017. Dynamics of dissolved organic matter (DOM) in a typical inland lake of the Three Gorges Reservoir area: fluorescent properties and their implications for dissolved mercury species. J. Environ. Manag. 418e429. Jing, Y.D., He, Z.L., Yang, X.E., 2007. Effects of pH, organic acids, and competitive cations on mercury desorption in soils. Chemosphere 69, 1662e1669. Jones, D.L., 1998. Organic acids in the rhizosphere e a critical review. Plant Soil 205, 25e44. Kamman, N.C., Burgess, N.M., Driscoll, C.T., Simonin, H.A., Goodale, W., Linehan, J., Estabrook, R., Hutcheson, M., Major, A., Scheuhammer, A.M., 2005. Mercury in freshwater fish of northeast North Americaea geographic perspective based on fish tissue monitoring databases. Ecotoxicology 14, 163e180. King, J.K., Kostka, J.E., Frischer, M.E., Saunders, F.M., Jahnke, R.A., 2001. A quantitative relationship that demonstrates mercury methylation rates in marine sediments are based on the community composition and activity of sulfate-reducing bacteria. Environ. Sci. Technol. 35, 2491e2496. Kortum, G., Vogel, W., Andrussow, K., 1961. Dissociation Constants of Organic Acids in Aqueous Solution. Plenum Press, New York. Krishnamurti, G.S.R., Cieslinski, G., Huang, P.M., VanRees, K.C.J., 1997. Kinetics of cadmium release from soils as influenced by organic acids: implication in cadmium availability. J. Environ. Qual. 26, 271e277. €nheit, P., 1983. Why do sulfate-reducing bacteria outcompete Kristjansson, J.K., Scho methanogenic bacteria for substrates? Oecologia 60, 264e266. Krzyszowska, A.J., 1996. Ion-chromatographic analysis of low molecular weight organic acids in spodosol forest floor solutions. Soil Science Society of America 60, 1565e1571. Kumar, S., Dagar, S.S., Mohanty, A.K., Sirohi, S.K., Puniya, M., Kuhad, R.C., Sangu, K.P.S., Griffith, G.W., Puniya, A.K., 2011. Enumeration of methanogens with a focus on fluorescence in situ hybridization. Naturwissenschaften 98, 457e472. Li, J.Y., Xu, R.K., Ji, G.L., 2005. Dissolution of aluminum in variable charged soils as affected by low-molecular-weight organic acids. Pedosphere 15, 484e490. Li, J.Y., Xu, R.K., Tiwari, D., Ji, G.L., 2006. Effect of low-molecular-weight organic acids

196

D. Yin et al. / Environmental Pollution 235 (2018) 186e196

on the distribution of mobilized Al between soil solution and solid phase. Appl. Geochem. 21, 1750e1759. Li, Y.B., Cai, Y., 2013. Progress in the study of mercury methylation and demethylation in aquatic environments. Sci. Bull. 58, 177e185. Liang, L., Wang, Y.M., Zhang, C., Yu, Y.W., An, S.W., Wang, D.Y., 2016. Effect of soil and dominant plants on mercury speciation in soil and water system of water-levelfluctuation zone in the three Gorges area. Environ. Sci. 37, 955e962 (in Chinese with English abstract). Liao, Y.C., Chien, S.W.C., Wang, M.C., Shen, Y., Hung, P.L., Das, B., 2006. Effect of transpiration on Pb uptake by lettuce and on water soluble low molecular weight organic acids in rhizosphere. Chemosphere 65, 343e351. Liu, G.L., Cai, Y., O'Driscoll, N., 2011. Environmental Chemistry and Toxicology of Mercury. John Wiley & Sons. Liu, J., Jiang, T., Huang, R., Wang, D.Y., Zhang, J.Z., Qian, S., Yin, D.L., Chen, H., 2017. A simulation study of inorganic sulfur cycling in the water level fluctuation zone of the Three Gorges Reservoir, China and the implications for mercury methylation. Chemosphere 166, 31e40. Marvin-Dipasquale, M.C., Oremland, R.S., 1998. Bacterial methylmercury degradation in Florida everglades peat sediment. Environ. Sci. Technol. 32, 2556e2563. Mergler, D., Anderson, H.A., Chan, L.H., Mahaffey, K.R., Murray, M., Sakamoto, M., Stern, A.H., 2007. Methylmercury exposure and health effects in humans: a worldwide concern. Ambio 36, 3e11. Min, K., Lehmeier, C.A., Ballantyne, F., Tatarko, A., Billings, S.A., 2014. Differential effects of pH on temperature sensitivity of organic carbon and nitrogen decay. Soil Biol. Biochem. 76, 193e200. Montgomery, S., Lucotte, M., Rheault, I., 2000. Temporal and spatial influences of flooding on dissolved mercury in boreal reservoirs. Sci. Total Environ. 260, 147e157. Nuzzo, R., 2014. Scientific method: statistical errors. Nature 506, 150e152. Office of National Soil Survey, 1979. Provisional Technical Regulations for the Second Soil Census in China. China Agricultural Press, Beijing (in Chinese). Oremland, R.S., Culbertson, C.W., Winfrey, M.R., 1991. Methylmercury decomposition in sediments and bacterial cultures: involvement of methanogens and sulfate reducers in oxidative demethylation. Appl. Environ. Microbiol. 57, 130e137. Pak, K., Bartha, R., 1998. Products of mercury demethylation by sulfidogens and methanogens. Bull. Environ. Contam. Toxicol. 61, 690e694. Porvari, P., 1998. Development of fish mercury concentrations in Finnish reservoirs from 1979 to 1994. Sci. Total Environ. 213, 279e290. Qin, C.Q., Liang, L., You, R., Deng, H., Wang, D.Y., 2015. Effects of citric acid on activation and methylation of mercury in the soils of water-level-fluctuating zone of the three Gorges.Reservoir. Environmental Science 36, 4494e4450. (in Chinese with English abstract). Qin, F., Shan, X.Q., Wei, B., 2004. Effects of low-molecular-weight organic acids and residence time on desorption of Cu, Cd, and Pb from soils. Chemosphere 57, 253e263. Rukshana, F., Butterly, C.R., Baldock, J.A., Xu, J.M., Tang, C., 2012. Model organic compounds differ in priming effects on alkalinity release in soils through carbon and nitrogen mineralisation. Soil Biol. Biochem. 51, 35e43. Schartup, A.T., Balcom, P.H., Mason, R.P., 2014. Sediment-Porewater partitioning, total sulfur and methylmercury production in estuaries. Environ. Sci. Technol. 48, 954e960. Selch, T.M., Hoagstrom, C.W., Weimer, E.J., 2007. Influence of fluctuating water levels on mercury concentrations in adult walleye. Bull. Environ. Contam.

Toxicol. 79, 36e40. Shao, D.D., Kang, Y., Wu, S.C., Wong, M.H., 2012. Effects of sulfate reducing bacteria and sulfate concentrations on mercury methylation in freshwater sediments. Sci. Total Environ. 424, 331e336. Siciliano, S.D., O'Driscoll, N.J., Tordon, R., Hill, J., Beauchamp, S., Lean, D.R.S., 2005. Abiotic production of methylmercury by solar radiation. Environ. Sci. Technol. 39, 1071e1077. €derberg, K., 2003. Bacterial Activity and Community Composition in the RhizoSo sphere. Influence of Plant Species, Root Age, Nitrogen Addition and Mycorrhizal Status. Lund University. St Louis, V.L., Rudd, J.W.M., Kelly, C.A., Bodaly, R.A., Paterson, M.J., Beaty, K.G., Hesslein, R.H., Heyes, A., Majewski, A.R., 2004. The rise and fall of mercury methylation in an experimental reservoir. Environ. Sci. Technol. 38, 1348e1358. Stumpner, E.B., Marvindipasquale, M.C., Alpers, C.N., Fleck, J., 2013. A Recommended Protocol for the Preservation and Storage of 'reactive' Inorganic Mercury in Sediment (AGU Fall Meeting). Sun, Q., Xie, H.B., Chen, J., Li, X., Wang, Z., Sheng, L., 2013. Molecular dynamics simulations on the interactions of low molecular weight natural organic acids with C60. Chemosphere 92, 429e434. Sunderland, E.M., Gobas, F.A.P.C., Branfireun, B.A., Heyes, A., 2006. Environmental controls on the speciation and distribution of mercury in coastal sediments. Mar. Chem. 102, 111e123. The Ministry of Water Resources of the People’s Republic of China, 2008. Standards for Classification and Gradation of Soil Erosion, SL 190e2007. China Water Power Press, Beijing (in Chinese). Ullrich, S.M., Tanton, T.W., Abdrashitova, S.A., 2001. Mercury in the aquatic environment: a review of factors affecting methylation. Crit. Rev. Environ. Sci. Technol. 31, 241e293. van Hees, P.A.W., Dahlen, J., Lundstrom, U.S., Boren, H., Allard, B., 1999. Determination of low molecular weight organic acids in soil solution by HPLC. Talanta 48, 173e179. Van
cura, V., Hovadík, A., 1965. Root exudates of plants. Plant Soil 57, 69e83. Wang, S.L., Mulligan, C.N., 2013. Effects of three low-molecular-weight organic acids (LMWOAs) and pH on the mobilization of arsenic and heavy metals (Cu, Pb, and Zn) from mine tailings. Environ. Geochem. Health 35, 111e118. Wieland, G., Neumann, R., Backhaus, H., 2001. Variation of microbial communities in soil, rhizosphere, and rhizoplane in response to crop species, soil type, and crop development. Appl. Environ. Microbiol. 67, 5849e5854. Xu, W.H., Liu, H., Ma, Q.F., Xiong, Z.T., 2007. Root exudates, rhizosphere Zn fractions, and Zn accumulation of ryegrass at different soil Zn levels. Pedosphere 17, 389e396. Yin, Y.G., Chen, B.W., Mao, Y.X., Wang, T., Liu, J.F., Cai, Y., Jiang, G.B., 2012. Possible alkylation of inorganic Hg(II) by photochemical processes in the environment. Chemosphere 88, 8e16. You, R., Liang, L., Qin, C.Q., Deng, H., Wang, D.Y., 2016. Effect of low molecular weight organic acids on the chemical speciation and activity of mercury in the soils of the water-level-fluctuating zone of the three Gorges reservoir. Environmental Science 37, 173e179 (in Chinese with English abstract). Zhang, J.X., Liu, Z.J., Sun, X.X., 2009. Changing landscape in the Three Gorges Reservoir Area of Yangtze River from 1977 to 2005: land use/land cover, vegetation cover changes estimated using multi-source satellite data. Int. J. Appl. Earth Obs. Geoinf. 11, 403e412. Zhang, Q., Lou, Z., 2011. The environmental changes and mitigation actions in the Three Gorges Reservoir region, China. Environ. Sci. Pol. 14, 1132e1138.