Soil mercury accumulation, spatial distribution and its source identification in an industrial area of the Yangtze Delta, China

Ecotoxicology and Environmental Safety 163 (2018) 230–237

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Soil mercury accumulation, spatial distribution and its source identification in an industrial area of the Yangtze Delta, China

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Yanxia Zhanga, Mei Wangb, Biao Huanga, , Mohammad Saleem Akhtarc, Wenyou Hua, Enze Xiea a

Key Laboratory of Soil Environment and Pollution Remediation, Institute of Soil Science, Chinese Academy of Sciences, Nanjing 210008, China College of Water Sciences, Beijing Normal University, Beijing 100875, China c Institute of Soil Science, PMAS-Arid Agriculture University, Rawalpindi 4463, Pakistan b

A R T I C LE I N FO

A B S T R A C T

Keywords: Industrial emissions Soil Hg isotopic composition Prevailing wind effect Atmospheric deposition Yangtze River Delta

Understanding soil mercury (Hg) accumulation, spatial distribution, and its sources is crucial for effective regulation of Hg emissions. We chose a study area covering approximately 100 km2 representing one of the rapid growing industrial towns of the Yangtze River Delta (YRD), China, to explore soil Hg accumulation. In surface soil, total Hg ranged from 310 to 3760 μg/kg, and 53% samples exceeded the most generous Chinese soil critical value (1500 µg/kg). Hg concentration in rice ranged from 10 to 40 µg/kg, and 43% samples exceeded the regulatory critical value (20 µg/kg). Total Hg concentrations in soil profiles gradually decreased, reaching background levels up to 60 cm profile depth. Meanwhile, proportions of mobile, semi-mobile and non-mobile Hg to total Hg at every soil depth were similar, leading us to deduce that soil Hg has accumulated in this area over a long period. Total and bioavailable Hg in topsoil exhibited the highest concentrations in the center of the study area, and radially decreased towards the periphery, which might be explained by the distribution of industry and the prevailing wind. To trace the Hg sources, we selected soil and atmospheric dust samples for isotope analysis. Hg isotopic composition of surface soil (δ202Hg = −0.29 ± 0.10‰ and Δ199Hg = 0.03 ± 0.03‰) was close to that of atmospheric dust (δ202Hg = −0.54 ± 0.10‰ and Δ199Hg = 0.03 ± 0.05‰), but considerably different from Hg isotopic composition in subsoil (δ202Hg = −0.90 ± 0.09‰ and Δ199Hg = −0.04 ± 0.04‰). Thus, we speculated that atmospheric deposition could change Hg isotopic composition in topsoil. Our findings suggest that when Hg atmospheric dust deposition changes Hg levels in surface soil, soil remediation, and crop safety might be compromised.

1. Introduction Mercury (Hg) has been listed as one of the top ten chemicals of public health concern by the World Health Organization, posing a serious threat for ecosystem and human health, such as nervous, digestive and immune systems, lungs, kidneys, and skin and eyes, at local and global scales (World Health Organization WHO, 2016). Modern global Hg contamination is dominated by anthropogenic influences, such as fossil fuel combustion, smelting, electroplating, etc. (Sun et al., 2014; United Nations Environmental Programme UNEP, 2013), which is more serious in developing countries because of their accelerated industrialization in the last few decades (Feng et al., 2006; Liang et al., 2013; Sun et al., 2014; United Nations Environmental Programme UNEP, 2013). Prior studies have extended the research area on Hg accumulation into soil environment (Feng et al., 2010, 2004, 2006; Ottesen et al.,

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2013; Wiederhold et al., 2013). Due to cumulative effects and long-term interactions of heavy metals, accumulation of Hg in soil poses a threat to crops which can transfer to humans through food chain (Chang et al., 2014). Different from Hg accumulation in mining soils (Feng et al., 2010, 2004, 2006; Wiederhold et al., 2013), Hg transferring from agricultural soil to crops depends not only on its accumulation but also on soil properties such as soil pH, oxidation-reduction potential (Eh) (Jing et al., 2007; Randall et al., 2004), soil organic matters (SOM) (Qian et al., 2009), and residue time (Grimaldi et al., 2008). Therefore, it is of practical significance to identify and understand the accumulation, spatial distribution, sources of Hg in soil, and the correlations in soil and crops to effectively regulate emissions and reduce safety threat of Hg to food and populations. Hg accumulation in agricultural soil in China is getting increased attention, such as in the Yangtze River Delta (YRD) (Huang et al., 2011; Xu et al., 2014) and the Pearl River Delta (Chang et al., 2014) in the

Corresponding author. E-mail address: [email protected] (B. Huang).

https://doi.org/10.1016/j.ecoenv.2018.07.055 Received 6 February 2018; Received in revised form 12 July 2018; Accepted 14 July 2018 0147-6513/ © 2018 Elsevier Inc. All rights reserved.

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southeast of China. As one of the fastest economic development regions in China, the YRD (covering Jiangsu, Shanghai, and Zhejiang provinces), Hg accumulation in soil has occurred because of industrial emissions of Hg-containing chemicals and uncontrolled agricultural inputs over the last three decades (Huang et al., 2011; Shao et al., 2006; Shen et al., 2010; Xu et al., 2014). For example, Xu et al. (2014) pointed out that Hg pollution covered approximately 38% area of the YRD, and the sources of Hg pollution may be attributed to atmosphere deposition from industrial emissions and historical use of Hg containing pesticides. Shen et al. (2010) revealed that Hg pollution in surface soil in Suzhou, Jiangsu province exceeded national/local background levels of Hg (150/289 μg/kg). However, the existing studies mainly focused on the current levels of Hg accumulation in soils in the YRD area and its spatial distribution, less on exploring sources of Hg emissions. Furthermore, Huang et al. (2011) found that the highest total Hg and bioavailable Hg (HCl-Hg) concentrations in soil of Zhangjiagang, Jiangsu were shown in the southwest of the wastewater outlet within about 50 m away from the outlet, and total Hg and HCl-Hg concentrations decreased as the distance from the outlet increased. However, the degree of influence on Hg distribution by wastewater was limited compared to atmospheric pathways Feng et al., 2006). Thus, it is important to further consider atmospheric influence on the Hg spatial distribution pattern in soil. Source identification of soil Hg is crucial for regional soil remediation and management of emissions. With the improvement of analytical methods, Hg stable isotopic analysis is becoming an effective method to identify sources of Hg in the environment (Wiederhold et al., 2013; Yin et al., 2013; Yu et al., 2016). Hg possesses seven stable isotopes which can be systematically fractionated during specific transformation reactions, and Hg isotope compositions vary significantly between different source materials (Sherman et al., 2012; Sun et al., 2014; Wiederhold et al., 2013; Yin et al., 2013). With the development of multi-collector inductively coupled plasma mass spectrometry (MCICP-MS) (Blum and Bergquist, 2007), high-precision analyses of many “non-traditional” stable isotope systems including Hg have become feasible (Lauretta et al., 2001; Perrot et al., 2010; Sun et al., 2014; Wiederhold et al., 2013). Thus, it is appropriate to adopt this method to identify the source of agricultural soil Hg. In order to cover the shortage on the research of Hg accumulation in the YRD area, we chose a typical industrial town in this area to determine two objectives: 1) assess levels of total Hg in soil and rice and bioavailable Hg in soil, and 2) explore sources of Hg based on spatial pattern of soil Hg and Hg isotopic ratios. Sources of Hg pollution were traced in the vicinity of a typical industrial town in the YRD based on its spatial distribution of Hg concentration in soil, whereas the effect of atmospheric disposition was discerned through the isotopic analysis. The contribution of this paper adds our knowledge on identifying Hg pollution sources in soil, which will provide a basis for evaluating regional soil Hg risks and food safety.

Fig. 1. Sampling map in the study area.

2.2. Sample collection and preparation We considered the spatial homogeneity of samples, layout of factories and residential area, and town center, etc. when setting collecting sites in this area. The land use types of soil mainly involved rice field and industrial land. A total of 76 topsoil (0 – 20 cm) samples were collected throughout the study area, and the coordinates of each sample was recorded by GPS. Each soil sample was a composite of five subsamples within 100 m2 collected in a sealed polyethylene bag. A total of 56 rice grain samples were collected directly from the rice field when the soil samples in the same sites were collected at the same time (Fig. 1). Two typical soil profiles were sampled in the farmland close to typical industrial areas (Z07 and Z11), for determining vertical distributions of Hg. A stainless steel core sampler (3 cm inner diameter) was dug 100 cm into the soil at five locations with 100 m2, and composited separately at the depths of 0–10 cm, 10–20 cm, 20–40 cm, 40–60 cm and 60–100 cm. The soil was air-dried at room temperature, homogenized, and grinded to pass through a 2 mm mesh nylon sieve for measuring soil pH and bioavailable Hg. For SOM and total Hg determination, the soil was further grinded to pass through a 0.15 mm mesh. The rice grain was thoroughly washed and rinsed with deionized water, and then sucked by a filter paper. The grain was dried in an air-circulating oven at 60 ℃ until a constant weight, and then was manually de-husked and grinded by using a stainless steel grinding machine to pass a 0.25 mm sieve. To trace the influence of atmospheric deposition on Hg in soil, three atmospheric dust samplers were installed at locations: (i) close to industries (DI), (ii) in the town center (TC) and (iii) in the rural area (WT). The dust sampler consisted of a collection barrel made of stainless steel, 30 cm high and 15 cm inner diameter, bucket lined with a plastic bag and filled up with a layer of glass balls (diameter 1.2 cm) under the plastic lining. The barrel was positioned on a metallic stand 1.5 m above ground level. The dust was allowed to accumulate for 30 ± 2 day and quantitatively washed in a glass beaker using distilled deionized water. The suspension volume was reduced to 10–20 mL on a hot plate below 60 °C. The suspension was transferred to Teflon baker, and dried in an oven until constant weight for further analysis. As general characteristics, soil pH was measured in 1:2.5 soil to water ratio mixture (Lu, 2000). SOM was analyzed by wet oxidation using excess dichromate, and the unconsumed dichromate is then backtitrated with an iron sulfate solution (Walkley-Black method) (Nelson and Sommers, 1982). Mercury isotope analyses were carried for only selected soil and dust samples. Five surface soil from farmlands, three soil depths from each of the two profiles (0 – 10 cm, 10 – 20 cm, and 60

2. Materials and methods 2.1. Study area The study area covers approximately 100 km2 in Jiangsu province between longitude 31° 24′ N to 31° 30′ N and latitude 120° 26′ E to 120° 35′ E (Fig. 1). This area has a sub-tropical monsoon climate (prevailing southeast wind in summer and northwest wind in winter), with annual mean rainfall of 1040 mm and an average (lowest – highest) temperature of 17.6 (– 8.3 to 39.4) °C. The soil was developed from lacustrine deposits in the Holocene Epoch and have a neutral and weak acidic reaction and a clay loam texture according to the Soil Chronicle of Zhangjiagang County in Jiangsu province (Soil Survey Office of Zhangjiagang County SSOZJG, 1984). The distributions of industry, residence, and farmland around 2010 are shown in Fig. 1. 231

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– 100 cm), and two atmospheric dust samples (one close to rural residential area and the other close to industrial area) were chosen (Fig. 1).

2.4. Mercury concentration analysis Total Hg in soil, rice grain, dust digest, and the HCl-Hg and sequential extracts were measured using atomic fluorescence spectrometry by an atomic fluorescence spectrometer, AFS-160A, Beijing Rayleigh Analytical Instrument, China. For quality control, reagent blank and an external reference material were incorporated in each batch of digestion and analysis. The external reference soil GBW 07403 and rice GBW 07604, poplar leaves were from the Institute of Geophysical and Geochemical Exploration, Chinese Academy of Geological Sciences. The limit of detection of the instrument was 0.001 ng/mL. The precision from replicate analysis was more than 95%.

2.3. Sample digestion and extraction Soil and dust samples were digested in concentrated acid mixture for total Hg determination. Half gram of 0.15 mm sieved soil/dust was added to a digestion flask containing 10 mL concentrated HNO3, HCl and H2O mixture of 1:3:4 (v/v/v) ratio, and the suspension was heated at 95 ℃ in a water bath for 2 h (Stevens et al., 2003). Total Hg in rice was digested by high-pressure digestion. Total Hg in rice was digested by high-pressure digestion. 0.5 g grinded rice was put into a polytetrafluoroethylene tank with 7 mL HNO3, overnight, mixed by 3 mL H2O2, covered and tighten with a steel-less sleeve, put it into a constant temperature drying box at 120 ℃ for 2.5 h, cooled to room temperature, and then moved to a colorimetric tube for further analysis (Ni et al., 2008). For bioavailable Hg determination, one gram soil was extracted with 10 mL 1 mol/L HCl solution shaken on an orbit shaker running snugly for 30 min, and then suspension was centrifuged at 3200 rpm for five min (Huang et al., 2011). Prior studies have pointed out that heavy metals migrate from more mobile fractions (e.g., water soluble fractions) to refractory ones (e.g., strongly complex and residual fractions) over time, resulting in more stable association of heavy metals with soil (Jalali and Khanlari, 2008; Ma et al., 2015; Zhong et al., 2013). Meanwhile, the extractability, bioavailability, and toxicity of heavy metals decrease with time extension (Wang et al., 2017). Thus, to investigate if the Hg in soil is current or historical emissions, we did Hg sequential extraction for Hg fractions in each soil depth of soil profiles. The Hg fraction analysis was analyzed based on a method of sequential extraction of Hg in previous studies (Fernandez-Martinez et al., 2005, 2006; Huang et al., 2011). The different operationally forms of extraction were defined as mobile Hg (M-Hg), semi-mobile Hg (SM-Hg) and non-mobile Hg (NM-Hg). This method provides detailed information about the mobility of mercury in the soil samples. Based on Han et al. (2003), the alkyl mercury species and soluble inorganic species contribute to the major portion of M-Hg. Alkyl mercury species, such as methylmercury (II), are more mobile than inorganic mercury species, and thus are more toxic and bio-accumulated. Soluble inorganic mercury species, such as mercuric chloride, are easier to transport by natural processes than the other inorganic mercury species. SM-Hg includes mainly elemental Hg and mercury-metal amalgams with less toxic, which are less mobile in environmental processes than soluble inorganic mercury species. NM-Hg mainly includes mercuric sulfide and mercurous chloride, which are chemically stable in soil over geologic time periods, thus, are the least toxic mercury species. One gram soil was added to 2.5 mL extracting solution containing 2% HCl and 10% ethanol (v/v) solutions pre-mixed in equal proportion for the M-Hg. The suspension was stirred on a vortex mixer for 1.5 min and centrifuged. The supernatant was saved for analysis. For the SMHg, the residue was washed with distilled deionized water until a chloride free test. Then, 5 mL 33% (v/v) HNO3 was added, and suspension was agitated on a vortex mixer for one min, heated at 95 ℃ in a water bath for 20 min and centrifuged at 3200 rpm upon cooling. The tubes remained lidded during heating. The extraction procedure was repeated for one more time and the supernatant was combined before analysis. Finally, for the NM-Hg, the residue was washed and dissolved by concentrated HCl and HNO3 acids added in 1:6 (v/v) ratio to the suspension with 7 mL water. The suspension heated at 95 ℃ for 20 min and centrifuged upon cooling to save the supernatant. The extraction was repeated once again and the supernatants were combined. For all the three forms, the extracts were filtered through 0.45 µm cellulose paper before analysis.

2.5. Mercury isotopic analysis Mercury isotopes can exhibit both mass-dependent fractionation (MDF) and mass-independent fractionation (MIF) (Bergquist and Blum, 2009). Hg-MDF (reported as δ202Hg) has been demonstrated to ubiquitously occur in the environment as a result of chemical (Bergquist and Blum, 2007; Zheng and Hintelmann, 2009), physical (Estrade et al., 2009; Sherman et al., 2012), and biological processes. MIF of Hg isotopes can be caused by magnetic isotope effects (Buchachenko, 2009), kinetic radical-pair mechanisms in photochemical processes (Sherman et al., 2012), and nuclear volume effects caused by the nonlinear increase of nuclear charge radii with mass (Schauble, 2007; Sherman et al., 2012; Wiederhold et al., 2013). The MIF affects both odd (e.g., 199 Hg and 201Hg) and even (e.g., 200Hg and 204Hg) isotopes (Bergquist and Blum, 2007; Wiederhold et al., 2013). Hg isotope compositions, MDF, were reported in delta notation (δ) in units of per mil (‰) referenced to the NIST SRM 3133 Hg standard (Blum and Bergquist, 2007; Feng et al., 2010), and the Eq. (1) is following: xxx xxx Hg Hg ⎛⎡ ⎤ ⎞ δ xxxHg (‰) = ⎜ ⎢ ⎜⎛ 198 ⎟⎞ sample / ⎜⎛ 198 ⎟⎞ ⎥−1⎟ × 1000, Hg Hg ⎠ ⎝ ⎠SRM 3133⎦ ⎠ ⎝⎣⎝

(1)

where XXX is the isotope of Hg between 199 and 202, (XXXHg/198Hg)sample is the measured isotope ratio of a sample, and (XXXHg/198Hg)NIST 3133 is the average isotope ratio of the reference standard solution (NIST SRM 3133). MIF values were indicated by capital delta (Δ) notation (‰), which shows the differences between the measured values of δ199Hg, δ200Hg, δ201Hg and those predicted from δ202Hg using the Eq. (2).

ΔxxxHg = δ xxxHg − (δ 202Hg × β ),

(2)

where β is equal to 0.252 for Hg, 0.502 for Hg, and 0.752 for 201 Hg (Blum and Bergquist, 2007; Feng et al., 2010; Sherman et al., 2012). Mercury isotopic analysis was performed by MC-ICP-MS using a Neptune mass spectrometer equipped with nine Faraday cups (Thermo Fisher Scientific, Germany) in the Institute of Geochemistry, Chinese Academy of Sciences. Details of overall instrumental setup, and parameters and analytical conditions were related to Foucher and Hintelmann (2006) and Feng et al. (2010). According to Feng et al. (2010), digests of samples chosen for Hg isotopic analysis were diluted to 1 – 5 ng/mL based on the total Hg concentration measured in Section 2.3 before the MC-IPC-MS analysis. We estimated the maximum sample analytical uncertainty of a given isotope ratio as two standard deviation ( ± 2 SD) of the measurement of the ratio in procedural standards. Hg in reference material UM-Almaden was measured the same way as other samples in each analytical session. When the calculated 2 SD was smaller than that of the replicate analysis of reference material UM-Almaden, the uncertainty associated to UM-Almaden was used instead (Blum and Bergquist, 2007; Feng et al., 2010). 199

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exceeded the most generous critical value (1500 µg/kg) (Ministry of Environmental Protection of the People's Republic of China (MEPC), 1995). Hg concentrations in rice grain samples ranged between 11 and 41 μg/kg, with a mean of 20 μg/kg (n = 56) (Table 1), close to the values in rice from other areas in the YRD (Fu et al., 2008; Hang et al., 2009), and much lower than the values reported in rice from a Hg mine area (mean 300 μg/kg) (Qiu et al., 2008) (Table 1). Hg concentration in 42.9% of our rice grain samples exceeded the maximum admitted limit value in Chinese National Standards for Food Safety (GB 2762-2017) (20 μg/kg) (National Health and Family Planning Commission of the People's Republic of China (NHFPCC), 2017). There were no statistically significant correlations between pH and concentrations of total Hg, HCl-Hg, and rice Hg, respectively. Total Hg in topsoil showed a significant positive correlation with SOM (R=0.532**) at the 0.01 level. HCl-Hg had a positive correlation with total Hg (R=0.742**), and rice Hg represented positive correlations with total Hg (R=0.334 *) and HCl-Hg (R=0.289*) at the 0.05 level, respectively (Table S2). Hg sequential extraction analysis of Hg in each depth of soil profiles showed that concentrations of SM-Hg were the largest of the three fractions at every soil depth, of which proportions to total Hg concentrations were high to 69 – 77% at each soil depth. Concentrations of M-Hg were the lowest (0.3 – 12.4% at every soil depth) (Fig. 2b). Although M-Hg fraction represents the lowest share to total Hg, this fraction contains the highest bioavailable Hg which is the easiest fraction to be taken up by plants (Fernandez-Martinez et al., 2005).

2.6. Descriptive and spatial statistics Descriptive statistics on soil concentration of total Hg, HCl-Hg and rice Hg was obtained. The correlation among Hg concentration in soil as total and HCl extractable, rice, and soil properties were determined. Log transformed Hg data obeys normal distribution by using Kolmogorov-Smirnov test. Spatial distribution of total and HCl extractable Hg was determined through ordinary kriging interpolation by using the open-source R statistical computing environment (Venables and Smith, 2009) with gstat package and sp package (Pebesma, 2004). Furthermore, we adopted the validation method ‘3-fold cross validation’ (Mahmood and Khan, 2009) (Supplementary methodology) to compare the simulating results and the corresponding measurement concentrations of total Hg and HCl-Hg. 3. Results 3.1. Hg concentrations in soil and rice grain samples Values of topsoil pH in our study area ranged from 4.55 to 6.56 (mean = 5.36, n = 76) with a coefficient of variation (CV) of 7%, which showed that this area is in a weak acidic environment. SOM was found to be from 21.53 to 44.84 g/kg (mean = 33.57 g/kg, n = 76), with a CV of 15%. pH in topsoil was significantly lower compared to that in subsoil, in contrast, SOM in topsoil was significantly increased compared to that in subsoil (Table S1). The trend of variation of pH and SOM in soil profiles from the top to bottom were consistent with the results from a former study (Huang et al., 2011). The topsoil samples exhibited a wide range of Hg concentration, from 310 to 3760 μg/kg (mean = 1345.39 μg/kg, median = 963.50 μg/kg, n = 76), with a large CV (64%). HCl-Hg concentrations varied from 15 to 136 μg/kg (mean = 51.38 μg/kg, median = 43.13 μg/kg, n = 76) with a CV of 49%, accounting for 3–5% of the total Hg in topsoil. Total Hg concentrations in soil profiles were the highest in topsoil of 0 – 10 cm depth, with a mean of 3208 μg/kg, gradually decreasing with soil depth increasing up to 60 cm (146 μg/kg) (Figs. 1 and 2a). The mean Hg concentration in topsoil exceeded the background value in Jiangsu (289 μg/kg) and the national background level (150 μg/kg), higher than the levels in soil in other regions of the YRD, such as Jiaxing (mean 199 μg/kg) (Xu et al., 2014) and Zhangjiagang (mean 140 μg/kg) (Huang et al., 2011) (Table 1). Total Hg in our study area was also higher than the levels in urban soil reported in Beijing (range 22–1388, mean 372 μg/kg) (Cheng et al., 2013) and European cities (range 3–3120, mean 30 μg/kg) (Ottesen et al., 2013). Total Hg in all soil samples exceeded the second most stringent critical value of Hg concentration in the Chinese Environmental Quality Standards of Soil (EQS) (GB 15618–1995) (300 µg/kg, pH < 6.5), 52.6% of which

3.2. Topsoil Hg spatial distribution The results of validation of simulating Hg concentration showed that R2 values of total Hg and HCl-Hg were 74% and 58% (over 50%) compared to the measurement of Hg concentration in soil, respectively (Figs. S1 and S2). Total and bioavailable Hg in topsoil exhibited higher concentrations in the center of the study area, and radially decreased towards the periphery, with significant Hg concentration zones from the northeast to southwest, similar to the spatial distribution of Hg accumulation in other cities such as Beijing (Zhang et al., 2006). Total Hg concentration in topsoil was high (3500 μg/kg) in the center, decreasing to 500 μg/kg on the periphery of the total Hg concentration zone (Figs. 1 and 3a). Similarly, HCl-Hg concentration was nearly 100 μg/kg in the center, lowering to 40 μg/kg on the periphery of the HCl-Hg concentration zone (Figs. 1 and 3b). As most of the industries distribute on northwestern and southeastern transects (Fig. 1), and the southeast wind is the predominant wind direction in this area (Table S3), we analyzed characteristics of Hg concentration in topsoil samples downwind from the southeast wind (Fig. 4a). Hg concentrations in topsoil samples showed peak values at Fig. 2. (a) Total Hg concentration in different soil depths and (b) proportions of Hg fractions ), semi-mobile (non-mobile Hg (NM-Hg ), and mobile (M-Hg )) in soil (SM-Hg profiles.

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Table 1 Descriptive statistics of Hg concentration in surface soil and rice and coefficient of Variation (CV) in this study area and other areas in literatures. Sites

Concentration (μg/kg)

Yangtze River Delta Yangtze River Delta Jiangsu background value National background value Yangtze River Delta Wuzhong, Jiangsu Suzhou, Jiangsu Changshu, Jiangsu Zhangjiagang, Jiangsu Jiaxing, Zhejiang Beijing Europe Yangtze River Delta Changshu, Jiangsu Taizhou, Zhejiang Hg mine, Guizhou

Soil total Hg Soil HCl-Hg Soil total Hg Soil total Hg Soil total Hg Soil total Hg Soil total Hg Soil total Hg Soil total Hg Soil total Hg Soil total Hg Soil total Hg Rice Hg Rice Hg Rice Hg Rice Hg

Reference Min

Max

Median

Mean

CV

310 15

3760 136

964 43

1345 51 289 150 150

64 49

380 0 109 4 42 27 3 11 – 16 10.3

1720 760 14800 1510 654 1388 1560 41 60 68 1120

172 248 19

140 555 140 199 372 30 20 15 22 300

136 70

50

31

This study This study Shen et al. (2010) Shen et al. (2010) Zhang, et al. (2012) Wang, et al. (2003) Shen et al. (2010) Hang et al. (2009) Huang et al. (2011) Xu et al. (2014) Cheng et al. (2013) Ottesen et al. (2013) This study Hang et al. (2009) Fu et al. (2008) Qiu et al. (2017)

Table S4. Subsoil samples from farmlands were characterized by negative δ202Hg (mean = −0.90‰, 2 SD = 0.09‰, n = 2) (much lower than the value in topsoil) and slightly negative Δ199Hg values (mean = −0.04‰, 2 SD = 0.04‰, n = 2). Given that subsoil experienced the lowest anthropogenic influences, Hg isotopic composition in subsoil is supposed to be close to the background levels of parent materials. Topsoil samples were characterized by slightly negative δ202Hg (mean = −0.29‰, 2 SD = 0.10‰, n = 9) and positive Δ199Hg values (mean = 0.03‰, 2 SD = 0.03‰, n = 9), which were different from the Hg isotopic composition in subsoil samples (Fig. 6). That is to say, the Hg isotopic composition in topsoil has changed compared to those in subsoil. Values of δ202Hg and Δ199Hg in topsoil samples from industries were −0.37 ± 0.08‰ and 0.02 ± 0.03‰ (n = 4), respectively, and those from farmlands were −0.23 ± 0.12‰ and 0.03 ± 0.03‰ (n = 5), respectively. There were no significant differences for δ202Hg and Δ199Hg in topsoil samples (P > 0.05) between the industries and farmlands. Values of δ202Hg and Δ199Hg in topsoil were very close to the values in atmospheric dust samples from the industrial area (DI) (δ202Hg: −0.38 ± 0.15‰ and Δ199Hg: −0.004 ± 0.05‰). Values of δ202Hg and Δ199Hg in atmospheric dust samples from the rural area (WT, away from factories) were −0.70 ± 0.04‰ and 0.06 ± 0.04‰ respectively, and the value of Δ199Hg was higher compared to that in subsoil. Thus, we speculated that atmospheric deposition has changed the Hg isotopic composition in topsoil, and the source of Hg was supposed to be from local industries.

the distance of 1–4 km away from industrial areas in a southeastern direction (Fig. 4b). For example, Hg concentration in topsoil in an industry area (D47) was 1776 μg/kg, but the peak value in topsoil in the southeast direction (D30) was 2.2 km away from D47, with a value of 3760 μg/kg, and then Hg concentration decreased as the distance between the industrial area and farmland increased. 3.3. Hg in atmospheric dust Hg emissions in atmospheric dust samples from the rural area (WT) (mean = 6.46 g/km2/30 day, n = 10) were higher than the levels in samples from the industrial area (DI) (mean = 4.00 g/km2/30 day, n = 10) and the town center (TC) (mean = 4.94 g/km2/30 day, n = 10) (Fig. 5a). Hg concentrations in samples from the WT, DI, and TC were 321, 264, and 237 μg/kg respectively, based on average values from a year monitoring period (Fig. 5b). Hg concentrations in atmospheric dust in this study area were found to be lower than the values in Beijing (range 53–1378 μg/kg, mean 340 μg/kg (Li et al., 2016; Zhang et al., 2006)) and Baoji city in Shaanxi province (range 480–2320 μg/ kg, mean 1110 μg/kg (Lu et al., 2009)), higher than the level in Huainan city in Anhui province (range 20–560 μg/kg, mean 160 μg/kg (Zheng et al., 2015)). 3.4. Hg isotopic analysis Hg isotopic compositions in all selected samples are presented in

Fig. 3. Distribution of soil Hg contaminated area in the study area. (a) Total Hg, (b) Bioavailable Hg (HCl-Hg). 234

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Fig. 4. Hg concentrations in topsoil samples in the direction of southeast wind. (a) Spatial distribution, (b) Relationships of Hg concentrations in topsoil samples between the industries and farmland area.

4. Discussions 4.1. Industrial emissions causing Hg accumulation for a long term Mercury was seriously accumulated in topsoil with a large CV in our study area, and a downward migration (i.e., down to 60 cm soil depth) of Hg in soil has happened. Generally, heavy metals in soil exist as a mobile fraction at a short term (Han et al., 2006; Ma et al., 2015), which means the proportion of M-Hg to total Hg in topsoil can be higher than that in subsoil when soil Hg accumulates at a short term. However, the proportion of M-Hg to total Hg in topsoil (0.3%) was even lower than that in subsoil (12.4%) (Fig. 2b). That is to say, the M-Hg in topsoil could have transformed into the SM-Hg or the NM-Hg, which means topsoil Hg has accumulated for a long period, and a new balance of Hg fractions in topsoil has formed in this area. As one of the fastest economic development regions in China, the YRD experienced rapid urbanization and industrialization over the last few decades. Numbers of energy-intensive industries such as chemicals, textile, and electrical industry have been built since the 1980s’ and 1990s’ (Huang et al., 2006, 2011). Those industries were meant to be the main sources of Hg in this area (Yin, 2009), which is also consistent with the results from the Hg isotope analysis for selected samples.

Fig. 6. δ202Hg (‰) versus Δ199Hg (‰) measured in selected sampling sites.

Fig. 5. Hg concentration in atmospheric dust from different sampling sites. (a) Amount of Hg emissions, (b) Hg concentration. 235

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Fortunately, with an improvement of environmental consciousness and technical innovation, many old polluted factories have transformed to less polluted industries or shut down after the 2000s’ in this area. Meanwhile, plenty of new residential and commercial buildings occupied the polluted land, which partially cut off sources of Hg emissions and reduce the human direct exposure to Hg. This might be part of reasons for lower Hg concentration in atmospheric dust in this area compared to those areas with high Hg deposition (Li et al., 2016; Zhang et al., 2006). However, as we analyzed, Hg accumulation in topsoil is still at a high level as a whole in this area, which would increase the concentration of bioavailable Hg in topsoil and rice absorption of Hg, because of significant positive correlations between soil total Hg, HCl-Hg, and rice Hg, increasing the risk of rice safety. Therefore, new agricultural techniques, such as crop rotation with a weaker absorbency of hazardous elements (Huang et al., 2011) and new soil remediation techniques (e.g., thermal desorption and biological techniques (Sierra et al., 2016; Xu et al., 2015)), might be developed, avoiding Hg exposure for a long run. In addition, the local industry layout can be adjusted to some degree, and more clean industries should be introduced to further reduce Hg emissions.

deposition changed the Hg isotopic composition in topsoil, based on the Hg isotopic composition analysis. Hg isotopic analysis for other sources should be considered to identify sources of Hg accumulation in soil and provide more scientific supports to reduce emissions.

4.2. Prevailing wind influencing on spatial distribution

Bergquist, B.A., Blum, J.D., 2007. Mass-dependent and-independent fractionation of Hg isotopes by photoreduction in aquatic systems. Science 318, 417–420. Bergquist, B.A., Blum, J.D., 2009. The odds and evens of mercury isotopes: applications of mass-dependent and mass-independent isotope fractionation. Elements 5, 353–357. Blum, J.D., Bergquist, B.A., 2007. Reporting of variations in the natural isotopic composition of mercury. Anal. Bioanal. Chem. 388, 353–359. Buchachenko, A.L., 2009. Mercury isotope effects in the environmental chemistry and biochemistry of mercury-containing compounds. Russ. Chem. Rev. 78, 319. Chang, C., et al., 2014. Accumulation of heavy metals in leaf vegetables from agricultural soils and associated potential health risks in the Pearl River Delta, South China. Environ. Monit. Assess. 186, 1547–1560. Cheng, H., et al., 2013. Mercury drop trend in urban soils in Beijing, China, since 1987. J. Geochem. Explor. 124, 195–202. Estrade, N., et al., 2009. Mercury isotope fractionation during liquid–vapor evaporation experiments. Geochim. Et. Cosmochim. Acta 73, 2693–2711. Feng, X., et al., 2004. A preliminary study on mercury contamination to the environment from artisanal zinc smelting using indigenous methods in Hezhang county, Guizhou, China—Part 1: mercury emission from zinc smelting and its influences on the surface waters. Atmos. Environ. 38, 6223–6230. Feng, X., et al., 2006. A preliminary study on mercury contamination to the environment from artisanal zinc smelting using indigenous methods in Hezhang County, Guizhou, China: Part 2. Mercury contaminations to soil and crop. Sci. Total Environ. 368, 47–55. Feng, X., et al., 2010. Tracing mercury contamination sources in sediments using Mercury isotope compositions. Environ. Sci. Technol. 44, 3363–3368. Fernandez-Martinez, R., et al., 2005. Distribution and mobility of mercury in soils from an old mining area in Mieres, Asturias (Spain). Sci. Total Environ. 346, 200–212. Fernandez-Martinez, R., et al., 2006. Physicochemical characterization and mercury speciation of particle-size soil fractions from an abandoned mining area in Mieres, Asturias (Spain). Environ. Pollut. 142, 217–226. Foucher, D., Hintelmann, H., 2006. High-precision measurement of mercury isotope ratios in sediments using cold-vapor generation multi-collector inductively coupled plasma mass spectrometry. Anal. Bioanal. Chem. 384, 1470–1478. Fu, J., et al., 2008. High levels of heavy metals in rice (Oryza sativa L.) from a typical Ewaste recycling area in southeast China and its potential risk to human health. Chemosphere 71, 1269–1275. Gratz, L.E., et al., 2010. Isotopic composition and fractionation of mercury in Great lakes precipitation and ambient air. Environ. Sci. Technol. 44, 7764–7770. Grimaldi, C., et al., 2008. Mercury distribution in tropical soil profiles related to origin of mercury and soil processes. Sci. Total Environ. 401, 121–129. Han, F.X., et al., 2006. Binding, distribution, and plant uptake of mercury in a soil from Oak Ridge, Tennessee, USA. Sci. Total Environ. 368, 753–768. Han, Y., et al., 2003. Speciation of mercury in soil and sediment by selective solvent and acid extraction. Anal. Bioanal. Chem. 375, 428–436. Hang, X., et al., 2009. Risk assessment of potentially toxic element pollution in soils and rice (Oryza sativa) in a typical area of the Yangtze River Delta. Environ. Pollut. 157, 2542–2549. Huang, B., et al., 2006. Environmental assessment of small-scale vegetable farming systems in peri-urban areas of the Yangtze River Delta region, China. Agric. Ecosyst. Environ. 112, 391–402. Huang, B., et al., 2011. Accumulation, transfer, and environmental risk of soil mercury in a rapidly industrializing region of the Yangtze River Delta, China. J. Soils Sediment. 11, 607–618. Jalali, M., Khanlari, Z.V., 2008. Effect of aging process on the fractionation of heavy metals in some calcareous soils of Iran. Geoderma 143, 26–40. Jing, Y., et al., 2007. Effects of pH, organic acids, and competitive cations on mercury desorption in soils. Chemosphere 69, 1662–1669.

Acknowledgments The authors are grateful for the funding from the National Natural Science Foundation of China (41473073), the National Science-technology Support Plan Projects (2015BAD05B04), and the Key Science and Technology Demonstration Project of Jiangsu Province (BE2016812). We are also grateful for the institute of Geochemistry, Chinese Academy of Sciences (IGCAS) to process the experiment analysis of Hg isotope. Appendix A. Supporting information Supplementary data associated with this article can be found in the online version at doi:10.1016/j.ecoenv.2018.07.055. References

This area belongs to a sub-tropical monsoon climate, prevailing northwest-southeast wind that was vertical to the distribution zones of Hg and HCl-Hg concentration in topsoil. Hg can be transferred regionally and globally, and different types of Hg are rapidly scavenged in atmosphere and deposit to surface environments (Sherman et al., 2012). Therefore, Hg from the industries in southeastern direction could be transferred by the southeast wind, and deposit in the same direction away from the southeastern emission sources. Similarly, Hg emissions from the industries in northwestern direction could be transferred by the northwest wind, and then deposit in the same direction away from the northwestern emission sources. As the southeast wind is the predominant wind direction (Table S3), it is reasonable that Hg concentration changing from the southeast to northwest was more obvious (Fig. 4). Thus, the distribution zones of radial Hg and HCl-Hg concentration can form in the direction of northeast – southwest. In addition, the natural geochemical factor was hard to form such spatial distribution of soil Hg, especially when topography is relatively flat and soil parent material is homogenous in this area. Therefore, prevailing wind could be an important factor to form the distribution zones of Hg and HCl-Hg concentration in topsoil. Given that atmospheric deposition has changed Hg isotopic composition in topsoil, it is important to explore Hg isotope in soil/dust samples from specific industries such as coal-fired utility boilers and electroplating industries to further decide which type of industries are main sources of Hg emissions. In addition, as Gratz et al. (2010) report, the isotopic Hg composition in any particular samples is the result of mixing of Hg from different sources combined with the effects of insource and post-emission fractionation. Thus, the Hg isotopic analysis in other potential sources, such as Hg-containing pesticides and fertilizers, might be considered as well. 5. Conclusions Hg accumulation in surface soil was very serious in this study area, extending to 60 cm soil depth, even though many polluted enterprises removed after the 2000s’. A new balance of Hg fractions of M-Hg, SMHg, and NM-Hg in topsoil has formed due to a long term Hg accumulation. HCl-Hg in topsoil and rice Hg concentration increased with the increased of total Hg concentration. Total Hg and HCl-Hg in topsoil exhibited higher concentration in the center of the study area, radially decreased from the center to around with significant Hg concentration zones in the vertical direction of prevailing wind. Atmospheric 236

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Environ. Sci. Technol. 33, 87–93 (In Chinese). Sherman, L.S., et al., 2012. Investigation of local mercury deposition from a coal-fired power plant using mercury isotopes. Environ. Sci. Technol. 46, 382–390. Sierra, M.J., et al., 2016. Sustainable remediation of mercury contaminated soils by thermal desorption. Environ. Sci. Pollut. Res. 23, 4898–4907. Soil Survey Office of Zhangjiagang County (SSOZJG), 1984. Soil Chronicle of Zhangjiagang County in Jiangsu Province. Agricultural office of Suzhou city, Soil Survey Office in Jiangsu Province, Suzhou, China. Stevens, D., et al., 2003. CSIRO Land and Water’s methods manual. ACIAR and CSIRO Land and Water. Sun, R., et al., 2014. Mercury stable isotope signatures of world coal deposits and historical coal combustion emissions. Environ. Sci. Technol. 48, 7660–7668. United Nations Environmental Programme (UNEP), 2013. Global mercury assessment. UNEP Chemicals Geneva. Venables, W., Smith, D., 2009. The R development core team: an introduction to R, notes on R: a programming environment for data analysis and graphics version 2.10. 1. Bristol: Network Theory. Wang, D., et al., 2017. Selenate redistribution during aging in different Chinese soils and the dominant influential factors. Chemosphere 182, 284–292. Wiederhold, J.G., et al., 2013. Mercury isotope signatures as tracers for Hg cycling at the new Idria Hg mine. Environ. Sci. Technol. 47, 6137–6145. World Health Organization (WHO), 2016. Mercury and health. 〈http://www.who.int/ mediacentre/factsheets/fs361/en/〉. Xu, J., et al., 2015. Sources and remediation techniques for mercury contaminated soil. Environ. Int. 74, 42–53. Xu, X., et al., 2014. Sources of heavy metal pollution in agricultural soils of a rapidly industrializing area in the Yangtze Delta of China. Ecotoxicol. Environ. Saf. 108, 161–167. Yin, R., et al., 2013. Stable mercury isotope variation in rice plants (Oryza sativa L.) from the Wanshan mercury mining district, SW China. Environ. Sci. Technol. 47, 2238–2245. Yin, Y., 2009. Key periods in the development of China pesticide industry-For China 60 anniversary. Chem. Ind. 7, 005. Yu, B., et al., 2016. Isotopic composition of atmospheric mercury in China: new evidence for sources and transformation processes in air and in vegetation. Environ. Sci. Technol. 50, 9262–9269. Zhang, X., et al., 2006. Mercury in the topsoil and dust of Beijing city. Sci. Total Environ. 368, 713–722. Zheng, L., et al., 2015. Distribution and health risk assessment of mercury in urban street dust from coal energy dominant Huainan City, China. Environ. Sci. Pollut. Res. 22, 9316–9322. Zheng, W., Hintelmann, H., 2009. Mercury isotope fractionation during photoreduction in natural water is controlled by its Hg/DOC ratio. Geochim. Et. Cosmochim. Acta 73, 6704–6715. Zhong, H., et al., 2013. Influence of contact time and sediment composition on the bioavailability of Cd in sediments. Environ. Pollut. 173, 11–16.

Lauretta, D.S., et al., 2001. Mercury abundances and isotopic compositions in the Murchison (CM) and Allende (CV) carbonaceous chondrites. Geochim. Et. Cosmochim. Acta 65, 2807–2818. Li, Y., et al., 2016. A comparison of metal distribution in surface dust and soil among super city, town, and rural area. Environ. Sci. Pollut. Res. Int. 23, 7849–7860. Liang, S., et al., 2013. Socioeconomic drivers of mercury mmissions in China from 1992 to 2007. Environ. Sci. Technol. 47, 3234–3240. Lu, R., 2000. Analytical Methods of Soil and Agro-chemistry. China Agricultural Science and Technology Press, Beijing (In Chinese). Lu, X., et al., 2009. Contamination assessment of mercury and arsenic in roadway dust from Baoji, China. Atmos. Environ. 43, 2489–2496. Ma, L., et al., 2015. Effects of metal-soil contact time on the extraction of mercury from soils. Bull. Environ. Contam. Toxicol. 94, 399–406. Mahmood, Z., Khan, S., 2009. On the use of k-fold cross-validation to choose cutoff values and assess the performance of predictive models in stepwise regression. Int. J. Biostat. 5. Ministry of Environmental Protection of the People's Republic of China (MEPC), 1995. Environmental Quality Standard for Soils (GB 15618-1995). Ministry of Environmental Protectioin of the People's Republic of China, Beijing. National Health and Family Planning Commission of the People's Republic of China (NHFPCC), 2017. Food Security Standards and Pollutants Limitations in China (GB 2762–2017). National Health and Family Planning Commision ofPeople's Republic of China, Beijing. Nelson, D.W., Sommers, L.E., 1982. Total Carbon, Organic Carbon, and Organic Mattermethods of Soil Analysis. American Society of Agronomy, Wisconsin, US. Ni, X., et al., 2008. High pressure digestion Hg AFS determination of Hg in rice. Mod. Food Sci. Technol. 24, 1188–1190. Ottesen, R.T., et al., 2013. Mercury in European agricultural and grazing land soils. Appl. Geochem. 33, 1–12. Pebesma, E.J., 2004. Multivariable geostatistics in S: the gstat package. Comput. Geosci. 30, 683–691. Perrot, V., et al., 2010. Tracing sources and bioaccumulation of mercury in fish of lake Baikal- Angara river using Hg isotopic composition. Environ. Sci. Technol. 44, 8030–8037. Qian, J., et al., 2009. Distribution of mercury pollution and its source in the soils and vegetables in Guilin Area, China. Bull. Environ. Contam. Toxicol. 83, 920–925. Qiu, G., et al., 2008. Methylmercury accumulation in rice (oryza sativa l.) grown at abandoned mercury mines in Guizhou, China. Agric. Food Chem. 56, 2465–2468. Randall, P., et al., 2004. Influence of pH and oxidation-reduction potential (Eh) on the dissolution of mercury-containing mine wastes from the sulphur Bank Mercury Mine. Trans.-Soc. Min. Metall. Explor. Inc. 316, 93. Schauble, E.A., 2007. Role of nuclear volume in driving equilibrium stable isotope fractionation of mercury, thallium, and other very heavy elements. Geochim. Et. Cosmochim. Acta 71, 2170–2189. Shao, X., et al., 2006. Effect of industrial distribution on soil heavy metal accumulation in a typical area of the Yangtze River delta. Acta Pedol. Sin. 43, 404 (In Chinese). Shen, J., et al., 2010. Study on heavy metal pollution in agricultural soils of Suzhou.

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