Mercury distribution in the foliage and soil profiles of a subtropical forest: Process for mercury retention in soils

Journal of Geochemical Exploration 205 (2019) 106337

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Mercury distribution in the foliage and soil profiles of a subtropical forest: Process for mercury retention in soils

T

⁎

Buyun Dua,c, Jun Zhoua,b, , Lingli Zhoub, Xingjun Fana, Jing Zhoub a

College of Resource and Environment, Anhui Science and Technology University, Fengyang, Anhui 233100, China Key Laboratory of Soil Environment and Pollution Remediation, Institute of Soil Science, Chinese Academy of Sciences, Nanjing 210008, China c Nanjing Institute of Environmental Sciences, Ministry of Ecological Environment, No. 8 Jiang-wang-miao Street, Nanjing, Jiangsu 210042, China b

A R T I C LE I N FO

A B S T R A C T

Keywords: Litterfall Atmospheric deposition Altitudinal distribution Budget Storage Residence time

Forest canopy can exert great influence on the global mercury (Hg) accumulations and soil Hg pools. However, there exists a debate that forests are net sinks or sources of atmospheric Hg. The current study aimed to study whether forest soils in the southwestern China are net sinks for atmospheric Hg and the soil residence time of Hg in the subtropical forest soils. Foliage samples of fern and pine in sixty-six plots and six soil profiles were sampled at the Tieshanping Forest Park (TFP) in southwestern China. To study the process of Hg budget, the flux of Hg input and the output from the forest soils were estimated and the mean soil Hg residence time (MRT) were calculated using a simple two-box model. The Hg concentrations in foliage (98 ng g−1) was relatively higher than remote areas because the elevated atmospheric Hg concentrations. Mercury concentrations in the understory foliage were positively correlated with altitude, which was resulted from the altitudinal distribution of increasing soil Hg concentrations. The annual Hg deposition flux via litterfall was estimated at 42 μg m−2 and the total Hg input to the forests was approximately 108 μg m−2. The annual Hg retention in forest soil was about at 76 μg m−2. The MRT was longer in surface horizon (0–10 cm, 134 ± 14 yr) than that in the mineral horizon (20–80 cm, 595 ± 34 yr). Subtropical forests in southwestern China act as net sinks for atmospheric Hg and 69% of the total Hg deposition was sequestrated in the forest soils. Short residence time of Hg in surface soil enhanced the deposited Hg output from the forest, which may be the reason for the comparable soil Hg concentration with some other remote forests with lower Hg depositions.

1. Introduction There is growing concern about exposure to the toxic element mercury (Hg) from a public health aspect, because Hg and its organic compounds (e.g. methylmercury) can biomagnify through food webs, leading to health risks to wild animals, plants and humans (Bavec et al., 2018; Buch et al., 2018; Chen et al., 2017; Du et al., 2018). Atmospheric Hg is derived from both human activities (e.g. coal combustions, nonferrous metal smelting, mining and cement production) and natural source and process (e.g. volcano and rock weathering). However, unlike other heavy metals, over 90% of the Hg is presented as elemental Hg (Hg0) that can reside in the atmosphere for 0.5–1.5 years (Beckers and Rinklebe, 2017; Fu et al., 2015; Han et al., 2018), which allows it to reach remote areas via long-range transport and then deposit to surface environment, where it has no local source (Wright et al., 2016; Yu et al., 2014). Global Hg models estimated that anthropogenic emission of Hg

to the atmosphere was 2500 ± 500 Mg yr−1 and approximate 3600 ± 3200 Mg yr−1 of atmospheric Hg is deposited to terrestrial surfaces (Outridge et al., 2018). As well known, forest canopy can exert great impact on the global Hg accumulation and soil Hg emission (Agnan et al., 2016; Ericksen et al., 2006; Wang et al., 2009). Generally, Hg0 in the atmosphere is the major source of Hg in plant foliage. Foliage Hg uptake through stomatal and nonstomatal routes is enhanced with increasing Hg0 concentration in atmosphere (Fay and Gustin, 2007) and the highest uptake rates coincided with times of high photosynthetic activity (Laacouri et al., 2013). A recent study showed that terrestrial vegetation acted as a global Hg0 pump, and it contributed to seasonal variations of atmospheric Hg0 and decreased atmospheric Hg0 levels in the Northern Hemisphere during the past 20 years with the primary production of forest (Jiskra et al., 2018). The absorbed Hg0 by forest canopy ultimately deposited to the floor and the pathways of atmospheric Hg input

⁎ Corresponding author at: Key Laboratory of Soil Environment and Pollution Remediation, Institute of Soil Science, Chinese Academy of Sciences, Nanjing 210008, China. E-mail address: [email protected] (J. Zhou).

https://doi.org/10.1016/j.gexplo.2019.106337 Received 8 January 2019; Received in revised form 22 June 2019; Accepted 12 July 2019 Available online 16 July 2019 0375-6742/ © 2019 Elsevier B.V. All rights reserved.

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to forest floor include throughfall (precipitation interacting with canopy) deposition, litterfall deposition and dry deposition of Hg directly to the forest floor (Wright et al., 2016). Dry deposition flux via foliage uptake Hg in the atmosphere composes a large part of dry deposition (Gruba et al., 2019). Litterfall Hg depositions constituted about 30%–60% of total atmospheric deposition in forest ecosystems of North America (Wright et al., 2016). In southwestern China, the Hg deposition to forest from the contribution of litterfall has surpassed throughfall (Fu et al., 2010; Wang et al., 2009). Additionally, about 90% of Hg in organic soil is predominantly originated from litterfall decomposition in boreal forest, suggesting that the litterfall Hg depositions play a crucial role in the biogeochemical process in forest ecosystems. High Hg loading in forest areas may pose serious risk to the biota in the forest ecosystem. Hg concentrations in a terrestrial food web were investigated in high elevation forests of Vermont, which showed that Hg concentrations in the invertebrates ranged from 1000 to 2000 ng g−1 (Rimmer et al., 2010). Additionally, Cristol et al. (2008) showed that mean Hg concentrations in orthopterans, lepidopterans and spiders in Virginia upland habitats was about 310 ng g−1. High Hg loads in the forest may result in biota in forest ecosystems is not uniquely adapted to this environment. Additionally, the insects were usually the food of the birds and small mammals, which suggested that high Hg retention in the forest ecosystem could be an unrecognized source of potentially hazardous concentrations of Hg to insect and even predators such as birds, mammalian predators and potentially higher trophic levels (Herring et al., 2018; Zhou et al., 2016). Although Hg from atmospheric deposition is mostly accumulated in soil organic matter (SOM) after mineralization (Szopka et al., 2011), the deposited Hg can also be re-emitted back to the air (Agnan et al., 2016) or be leached into surface or underground runoff (Zhou et al., 2015). Earth's surface (particularly in East Asia) is an increasingly important source of Hg0 emissions, contributing up to half of the global source from nature. They estimated that the terrestrial Hg emissions was 607 Mg yr−1, but with a large uncertainty range of −513 to 1353 Mg yr−1, which was mainly from the contributions of leaf-atmosphere fluxes, questioning to what degree forest were net sinks or sources of atmospheric Hg0 (Agnan et al., 2016). The Hg from past deposition is mainly bond with recalcitrant SOM and is regarded as legacy Hg that has a long soil residence time (Dittman et al., 2010). The concentration patterns of Hg distribution in soil profiles may not directly reflect to present Hg depositions from atmosphere because of the accumulation of legacy Hg, differential depositions, SOM accumulation and loss, and the transportation and transformation of Hg and SOM (Obrist et al., 2011; Yu et al., 2014). Climate can significantly influence the SOM accumulation in the forest soils, such as low soil carbon stocks with cold and dry climate and high soil carbon stocks with warm and wet climate. Therefore, the warm and wet climate in subtropical forest may also influence the Hg accumulation in soils. Contamination of Hg is a major public health concern in the southwestern China because of the toxicant and elevated methylmercury in rice and wildlife (Du et al., 2016; Song et al., 2018; Xu et al., 2018). In the current study, we surveyed Hg in the foliage and soil profiles of the Tieshanping Forest Park (TFP) in southwestern China, providing the urban industry influence on the Hg pools and mass balance for this region. The objected of the current study is (1) to study the Hg distribution in foliage, litter and soil profile; (2) to estimate the Hg mass balance (input of atmospheric deposition, and outputs of surface and underground runoff and soil surface Hg emission); and (3) to quantify soil Hg residence time in the subtropical forest.

Acidification of Chinese Terrestrial Systems (IMPACTS) project monitoring site (29°38′ N, 104°41′ E). Some of the residents lived in the TFP, but most of forests have been protected and under conservation status for > 50 years after the clear cutting in 1962 (Zhou et al., 2017). The forest types are homogeneous, dominated by Masson pine (Pinus massoniana Lamb.) that accounts for 90% of the covering area, and some associated species, such as camphor (Cinnamomum camphora) and Schima (Schima superba Gardn. et Champ.). Dominated understory species are ferns, such as Old World forked fern (Dicranopteris dichotoma Bernh.) and Pteris (Pteris cretica L. var. nervosa (Thunb.) Ching et S. H. Wu). The main soil type is a Calcaric Alumic in World Reference Base for Soil Resources (WRB, 2014). The local climate is a typical Indian humid monsoon climate and annual mean air temperature and rainfall are 18.8 °C and 1200 mm (Yang et al., 2009). Chongqing is an important industrial city in southwestern China, and consumes large amount of energy. The industrial activity has released approximately 10.5 t Hg yr−1 to the atmosphere in 2010, with an annual average growth rate of 9.82% (Zhang et al., 2011; Zhang et al., 2015). The precipitation Hg concentrations and wet Hg deposition in the core were 31 ng L−1 and 29 μg m−2 yr−1, respectively (Wang et al., 2014; Wang et al., 2012). Mercury pollution was regarded as major environmental burdens in Chongqing (Ma et al., 2015; Zhou et al., 2018b). Large amounts of Hg also deposited to the surrounding areas of the city via atmospheric long-distance transport. Previous studies showed that atmospheric Hg concentrations and Hg depositions were > 3.5 ng m−3 and 16 μg m−2 yr−1, respectively, which were > 20 km away from the city (Ma et al., 2015; Wang et al., 2009; Zhou et al., 2018b). The higher Hg depositions resulted much high soil Hg concentrations in surrounding subtropical forests. 2.2. Sample collection The altitudes ranged from 200 to 550 m. The Hg concentrations in tree foliar increase with its growth (Kowalski and Frankowski, 2016), so all plots were sampled at the refoliation period (March 2014) to minimize the influence of temporal variability. Foliage of pine and Old World forked fern were collected at 66 plots within TFP (Fig. S1) and coordinates and altitudes were recorded by GPS. The fully developed undamaged needles were sampled from pine at heights of about 3 m above the ground using averruncator, and foliage were sampled from fern at heights of about 0.5 m above the forest floor by hand and then the fresh samples were stored in envelopes. In each sampling plot (5 m × 5 m), representative foliage samples were collected in triplicate from the pine and fern and then mixed together. Additionally, the corresponding litter, organic soils and mineral topsoils were also collected, the Hg concentrations of which were showed in the previous study (Zhou et al., 2015). Soil profiles were sampled from a total 6 plots with the depth of 80 cm and within the altitude of 302–528 m. Three soil profiles were sampled on the eastern slope and the remaining three soil profiles were sampled on the western slope. In each sampling site, representative samples were collected by auger in triplicate within a 5 m × 5 m plot and then mixed together. The litter horizon (about 1–4 cm) was collected by hand. The first 5 cm to the soil profiles represented the organic horizons. About the 5–10 cm and 10–20 cm of the profile represented the A1 and A2 horizons respectively, and the remaining of 20–80 cm represent the B horizon as we described previously (Zhou et al., 2016). All the sampling plots were selected randomly but were away from roads and buildings > 100 m. Plant samples were oven-dried at 60 °C and then ground to powder using a stainless steel blender before chemical analysis. Soils were air-dried, manually ground, and sieved to a size of 200 meshes to remove debris and rubble in the laboratory.

2. Materials and methods 2.1. Study area

2.3. Analytical methods The TFP is a 1.5 × 104 ha forest preserve (Fig. S1) that is a SinoNorwegian multidisciplinary Integrated Monitoring Program on

Mercury 2

analysis

was

performed

by

Lumex

RA-

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Litterfall Hg deposition constitutes a large portion of Hg dry deposition to forests (Wright et al., 2016). The foliage biomass was estimated by the empirically model relating to the tree diameter at breast height. Understory foliage biomass were determined by destructive harvests of 3 randomly local replicate plots (Zhou et al., 2016). Litterfall fluxes from pine was assumed to be 1/3 the needle biomass as Wang (2012) observed the average longevity time of needle was 3 years. Assuming the linear accumulation of Hg concentration in foliage during the growth (Gong et al., 2014), the litterfall Hg concentration was approximately calculated by multiplying the determined Hg concentration in leaf by a factor of 1.5 (Gong et al., 2014). The total Hg input (Finput) includes litterfall deposition (Flitterfall) and throughfall deposition (Fthroughfall), and is calculated as:

915 + multifunctional Hg analyzer equipped with a pyrolysis attachment (Lumex Ltd., Russia), which directly analyzed plant and soil and conforms to EPA Method 7473 (USEPA 1998). The analyzer was calibrated by a certified standard (citrus leaf standard, GBW10020) for plant analysis. For soils, calibration curves were calibrated using soil standard (IGGE IRMA China). Plant and soil were thermally decomposed in an atomizer chamber at 800 °C with aided catalytic action, and then the RA-915 + analyzer checked the elemental Hg. Soil pH was determined with a glass electrode pH meter in a mass ratio of 1:2.5 soil:water suspension. Briefly, pH meter was calibrated by standard buffer solution and 4 g soil was added to a beaker with 10 mL of C02-free distilled water. The mixture was stirred for half an hour intermittently and then measured by the pH meter after half an hour. Metal cores were used to measure soil bulk density (Falciglia et al., 2017). Normally, wet oxidation procedure, is a routine, relatively accurate, and popular method for the determination of SOM. Many studies showed that determination of SOM through sequential loss on ignition (LOI) was suitable for exploratory soil surveys and consistent well with the wet oxidation procedure (Nakhli et al., 2019; Salehi et al., 2011). Thus, SOM were determined according to LOI (Zhou et al., 2013). Briefly, according to the procedure, air-dried soil samples (AD) were oven-dried at 105 °C for > 12 h to obtain the dry weight (DW105). The oven-dried samples were burned at 550 °C for > 4 h to obtain the weight were DW550, and then LOI was calculated based on the following formula:

LOI550 = 100(DW105 − DW550)/AD

Finput = Flitterfall + Fthroughfall

2.5.2. Output of Hg The dominate output pathways of Hg from forest were runoff and soil-air Hg flux. The runoff fluxes (Frunoff) were calculated based on the runoff amounts and Hg concentrations.

Frunoff = vsurface runoff × c surface runoff + vunderground runoff × c underground runoff (5) where vsurface runoff and vunderground runoff are the surface runoff and underground runoff, which are assumed to be 25% rainfall amount (Liu, 2005) and 50% throughfall amount (Luo et al., 2015); csurface runoff and cunderground runoff were assumed to be 6 ng L−1 and 22 ng L−1, which was the Hg concentration determined in the surface runoff of TFP area studied by Wang et al. (2009) and Zhou et al. (2015). The total Hg output (Finput) was calculated by the sum of surface runoffs (Fsurface runoff), underground runoff (F underground runoff) and soilair Hg fluxes (Femission) as the following equation:

(1)

Quality assurance and control of the analytical processes were ensured with duplicates and certified reference materials. Detection limit of Hg was 0.5 ng g−1 for solid samples. Each sample was measured twice and the mean concentration was used in this study. Calibration was verified every ten samples during analysis. The standard deviations of five duplicated measurements of soil and plant were 2.1% and 3.9%, respectively. The recoveries of citrus leaf standard and soil standard ranged from 95.7 to 103.9% and 96.1 to 104.9%, respectively (Table S1). For soil pH determination, pH meter was verified by standard buffer solution with pH = 4.0 every five samples during analysis. If the percentage of deviation > 5%, the pH meter would be recalibrated. For soil SOM determination, each sample was measured twice to obtain the mean concentration.

Foutput = Fsurface runoff + Funderground runoff + Femission

R = Finput − Foutput

A simple two-box model was used to evaluate the residence time of Hg in the forest soil. We estimated an approximation of Hg residence time (MRT) by assuming the surface (0–10 cm) and mineral horizon (10–80 cm) Hg pools and fluxes were in steady state. Although it is unlikely due to interannual variations of atmospheric deposition rates (Luo et al., 2015; Wang et al., 2009; Zhou et al., 2018b) and forest maturation, it can approximately estimate the Hg retention in soils. Surface and mineral horizon MRT were calculated according the method of Richardson and Friedland (2015) with the following Eqs. (5) and (6), respectively:

n i=1

(2)

2.5. Estimation of Hg mass balance To study the process of Hg budget, the flux of Hg input (atmospheric deposition) and the flux of Hg output from the forest soils (air-to-air emission, leaching to surface and underground runoffs) were estimated in the current study.

MRTsurface horizons = Pool surface horizons/Finput

(8)

MRTmineral horizons = Poolmineral horizons/R

(9)

where Poolsurface horizons and Poolmineral horizons were Hg pool in corresponding horizons. The weathering input and plant uptake were assumed to be zero. The previous study in this area showed that the Hg pools in forest biomass is negligible compared to the Hg pools in soil profiles with the depth of 0–40 cm (< 1% of the total Hg pools) (Zhou et al., 2016). Additionally, Hg in above ground biomass was mostly from atmospheric accumulation (Fay and Gustin, 2007). Moreover, low Hg concentrations may exist in igneous and metamorphic parent materials. Their weathering contributions were hypothesized to be negligible (Juillerat et al., 2012). Mercury emission from the mineral horizon to atmosphere and down mineral soils was considered negligible as

2.5.1. Input of Hg Litterfall Hg deposition (Flitterfall, mg m−2 yr−1) was estimated according to the litterfall production flux (mlitterfall, g m−2 yr−1) and the litterfall Hg concentration (clitterfall, mg g−1). Thus, Flitterfall was calculated as follows:

Flitterfall = mlitterfall × clitterfall

(7)

2.6. Soil Hg residence time

The storage of Hg in each soil layer (j) were calculated based on the soil Hg concentration (ci, ng g−1) and corresponding bulk density (Bdi, g m−3) in each layer. The sum of the Hg storage in each soil layer was used to calculate the Hg pool (Msoil, mg m−2) in the depth (di, m) of 80 cm. Therefore, Msoil was estimated based on the equation:

∑ ci × Bdi/di

(6)

where Eemission was reported by Zhou et al. (2019). The Hg retention (R) was calculated according to the Hg mass balance in the forest as:

2.4. Soil Hg pool

Msoil =

(4)

(3) 3

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(a)

150 n=20

n=66

-1

Hg concentrations (ng g )

n=13

120

n=13

120

90

60

(b)

n=66

n=33

-1

Hg concentrations (ng g )

150

n=20

90

60

200-400 400-500 500-600 200-600 Altitude (m)

n=33

200-400 400-500 500-600 200-600 Altitude (m)

Fig. 1. The Hg concentrations in foliage of Old World forked fern (a) and needle of Masson pine in various altitude. The samples size for the altitude was also showed (n) and no significant differences between the three elevations were observed.

surface horizon and bottom horizon emission have been showed to be decoupled (Richardson and Friedland, 2015; Richardson et al., 2013).

high atmospheric Hg0 concentration in the current study area resulted in elevated Hg concentrations in foliage.

2.7. Statistical analysis

3.2. Vertical distribution and pools of Hg in soil profiles

Mean Hg concentrations and pools in foliage and soil horizons were compared using one-way analyses of variance (ANOVAs) to determine if differences in Hg concentrations and pools existed. All differences in means were significant at the p = 0.05 level, and all means are reported with ± standard deviations from the means. The correlations between elevation and foliage as well as soil Hg concentrations/pools were analyzed by Pearson's correlation tests using SPSS software (SPSS Inc. 16.0), and the correlation coefficient and p values are presented and significantly correlated at the level of 0.05.

Concentrations of Hg and SOM in the six soil profiles are shown in Fig. 2 and the Hg concentrations ranged from 43 to 209 ng g−1. The highest Hg and SOM concentrations were observed in the litter layers, whereas their concentrations decreased significantly in the deeper layers. Under the depth of 20 cm, Hg concentrations were relatively stable. In the current study, Hg concentrations in litter and surface soils (0–10 cm) were comparable to some forests in North America and subtropical forests in China (Table 1). However, the content of Hg in litter and surface soils were much higher compared to those in remote temperate forests of China. Previous study showed that forest canopy can trap atmospheric Hg through stomatal and nonstomatal routes and the foliage Hg concentrations are enhanced with increasing Hg0 concentration in atmosphere (Fay and Gustin, 2007). Additionally, high atmospheric Hg can also increase the Hg concentrations in throughfall, because a large portion of the particulate bound Hg (PBM) and reactive gaseous Hg (RGM) scavenged by foliage is believed to wash off foliage surfaces in throughfall (Rea et al., 2000). The direct atmospheric Hg0 depositions may unlikely resulted in the higher Hg concentrations in litter and organic soils, because our previous study showed that overall emission flux was observed between the forest floor (soil and litter) and air in this forest (Zhou et al., 2019). Therefore, the higher concentrations in forest litter and soil were resulted from the high atmospheric Hg concentrations in the study area (Zhou et al., 2016), which given rise to elevated Hg in foliage as we stated above and then resulted in elevated Hg loadings. Relatively lower Hg concentrations in deeper soil layers than those in surface soil layers were observed in many other areas, which showed that Hg concentrations decreased significantly from organic layers to mineral layers (Grigal, 2003; Gruba et al., 2019). Enrichment of Hg in the surface layers than mineral layers demonstrated the important role of SOM in the Hg accumulation of atmospheric depositions. Significant correlations between Hg concentrations and SOM contents in soil at all the sampling sites (Pearson's test, Fig. S2), demonstrating that Hg was bound up with the SOM. This result was in accordance well with previous studies that observed the affinity of atmospheric deposited Hg to accumulate in the surface soil horizon, especially those rich in SOM (Grigal, 2003; Gruba et al., 2019). Just like some other heavy metals, Hg preferentially bond to thiol, which is the compounds with soil SOM (Grigal, 2003). Soil Hg and SOM interacted strongly with each other and played a role in the transport and mobility (Gruba et al., 2014), which helped stabilize Hg in the organic soil horizons and inhibited transport to mineral soil horizons. Additionally, pH values can also

3. Results and discussion 3.1. Hg concentrations in needle/foliage Fig. 1 shows the Hg concentrations in the foliage of overstory (pine) and understory (fern) ranged from 63 to 139 ng g−1 and 68–130 ng g−1 with mean values of 98 and 97 ng g−1, respectively (Fig. 1). There was no significant difference between average understory leaf and overstory needle Hg concentrations across the elevation gradient (p > 0.05). The Hg concentrations of leaves and needles were much higher than those in North America and Europe (Table 1). The average leaf and needle Hg concentrations in the current study was 3–5 times higher than those in the forests of USA (25 ng g−1 in 14 forests of the USA, (Obrist et al., 2011); 16.3 ng g−1 in 4 forests of northeastern USA, (Yang et al., 2018)) and Slovakia (41–68 ng g−1) (Ollerova et al., 2010). When comparing the Hg concentrations in the current study with some other remote forests of China, the needle Hg concentrations were also 2–9 times higher than the reported data for Tibetan Plateau, Mt. Gongga and Mt. Dongling (Table 1). The higher concentrations in the study site were resulted from the TFP locating in the mega industrial city, Chongqing. Previous studies demonstrated that the Hg in the atmosphere is nearly the dominated contributor of Hg in foliage by both laboratory and field experiments (Fay and Gustin, 2007; McClenahen et al., 2013; Yang et al., 2018). Zhang et al. (2015) showed that emissions of Hg to atmosphere from anthropogenic sources were about 10.5 t in Chongqing in the year of 2010. The anthropogenic emission has not only resulted in higher atmospheric Hg0 concentrations (6.7 ng m−3) in core urban areas (Yang et al., 2009), but also elevated the atmospheric Hg0 concentration (3.5 ng m−3) in our study area (Zhou et al., 2016), which was much higher than those in forested sites in North America (atmospheric Hg0 generally < 2.0 ng m−3) and remote areas in China (atmospheric Hg0 generally < 3.0 ng m−3) (Fu et al., 2015). Therefore, 4

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McClenahen et al., 2013 Richardson and Friedland, 2015 Yang et al., 2018; Yu et al., 2014 Obrist et al., 2011 Bushey et al. (2008) Pokhrel et al. (2016) Zhou et al. (2018b) Gong et al., 2014 Fu et al., 2010 This study This study

References

25–93 25–93

8–32 38.2 ± 32.2

64 2–56 96–251

246–336 179 131.8–272.9 22–420 48.2–92.4 2.5–45.6 62–102 59.2 ± 21.8 120–260 159–209 159–209 2004–2005 2012 2015 2008 2006 2014 2014

2002, 2010 2012–2014 2015

3.3. Altitudinal distribution of Hg in soils and foliage Spatial distribution can provide information regarding the role of regional contributions and topographical effects to the observed atmospheric Hg concentrations and its loads in areas without local resource. In the current study, the Hg concentrations in litter and surface soils (0–10 cm) were significantly correlated with the altitude (Fig. 3), and the result was consistent with our previous study which showed the correlation between Hg concentrations in the organic and mineral topsoil horizon and altitude was significant. The reason may be that soil at higher altitude received higher Hg deposition because of higher annual precipitation (Evans and Hutchinson, 1996). Additionally, organic soil at higher altitude released lower Hg from runoff and evasion flux because of higher SOM content and lower temperature at higher altitude, which effectively inhibited the Hg loss from output of stream water and soil-air exchange flux (Zhou et al., 2015). Furthermore, high deposition of sulfur has been frequently reported (Li et al., 2014), which would be occurred at higher altitude and also accumulate higher Hg and formed as HgS in the organic horizon. The Hg concentrations in the mineral soils were relatively stable for all the six soil profiles and were not correlated with altitudes (Fig. 3), because the atmospheric deposited Hg was mainly accumulated in the surface soils as we descripted above, which has little impact on the concentrations in deeper soil layers. For the foliage Hg concentrations, the needle Hg of pine displayed no significant correlation with altitude (p > 0.05) and showed no statically significant difference among the three gradient altitudes (p > 0.05) (Fig. 1). Zhou et al. (2015) reported that the Hg concentrations in the litter horizon, organic horizon and mineral topsoil horizon of the same 66 plots. The result showed that the Hg concentrations in litter (mainly pine litter) on the forest floor were not significantly different and related to altitudes, which was consistent with our result of Hg in pine needles. Fay and Gustin (2007) have

530 4000 m 1100 351–4400 2650–3580 200–600 200–600

Remote Remote Remote Remote Remote Remote Remote Remote Remote Suburban Suburban 700–800 650–750 430–780

Hardwood and conifer Pine and fir Evergreen broad leaf Old World forked fern Masson pine

Oak forest Hardwood and conifer Hardwood and conifer Hardwood and conifer Birch

Laurel Ridge Green Mountains, White Mountains Four forests of northeastern USA 14 forests of the USA Huntington Wildlife Forest, USA Central Himalayas, Nepal Mt. Dongling, China Tibetan Plateau, China Mt. Gongga China Tieshanping Forest Park Tieshanping Forest Park

Organic layer/surface soil

Mineral layer

24–31 4–13 16.3 8–48 4.4–29.1 5.6–19.4 26–65.9 5.4–13.6 27.5 ± 3.6 98 ± 19 77 ± 15

Concentrations in foliage/needle Concentration in soils Study period Location type Altitude (m a.s.l.) Forest type Site

Table 1 The Hg concentrations in foliage and soils in this study and literatures.

influence the migration and transformation of Hg in forest soils. Higher Hg concentrations but lower pH observed in the surface soils than those in mineral soils, indicating that the role of pH may play an important role in the Hg accumulation in soils. TFP has been experiencing a longterm serious acid deposition, and the low mean pH was about 4.0–4.2 and the high frequency of acid rain was about 90.0% with rainfall pH < 5.6 (Li et al., 2014). Significant negative correlation between Hg concentrations and pH values in all the soil samples of the six soil profiles was observed (Pearson's test, Fig. S2b), because pH decrease could lead to increasing more radicals and the increased radicals can absorbed more Hg according to the suggestion in previous study (Gunda and Scanlon, 2013). Table S2 showed Hg pool for individual soil horizons and the total Hg pool ranged from 73 to 84 mg m−2 in the six soil profiles, with an average pool of 79 mg m−2, which was lower than the data 89 mg m−2 observed from coniferous forests in NE Bavaria (a thinner soil layer, down to 60 cm) (Schwesig and Matzner, 2000). However, the Hg pools in TFP was also comparable with the pool data in North America at the similar soil layers (0–20 cm) of our study (Obrist et al., 2009; Richardson et al., 2013), because the Hg concentrations in litter and soils were similar with our study (Friedli et al., 2007; Obrist et al., 2011). It is interesting that much higher Hg concentrations in foliage and depositions were observed in our study areas than those in North America, but the soil Hg concentrations and pools were similar with our study area. The reason may be that the vegetation restoration started at 1960s before clearcutting, so the forest was a relatively young secondary forest. Additionally, litter decomposition rates were rapid in TFP and little litter and soil SOM accumulated on the forest floor because of warm and moist climate (Zhou et al., 2018b), which resulted in the acceleration of litter and organic soil Hg loss during SOM decomposition. Additionally, the soil environmental capacity of Hg may be limited in the study area.

5

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0

40

Hg concentrations (ng g ) 50 100 150 200

0

80

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Hg concentrations (ng g ) 50 100 150 200

(f)

0

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pH OM THg

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(d)

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Hg concentrations (ng g ) 50 100 150 200 (c) 0

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Hg concentrations (ng g ) 50 100 150 200 (b)

(a)

40 60

pH OM THg

80

0 2 4 6 8-1 10 TOM concentrations (g kg )/pH

Altitude: 500-600

Fig. 2. Vertical distribution of Hg in soil profiles. Each figure represents one pit: (a) 302 m; (b) 398 m; (c) 413 m; (d) 437 m; (e) 520 m; (f) 525 m.

Litter Surface soil Mineral soil

-1

Hg concentration (ng g )

200

(p > 0.05) (Fig. 1). Fay and Gustin (2007) have also studied Hg accumulation in a shrub species (Artemisia tridentata (Sagebrush)), and found that higher soil Hg content (27.7 ± 6.3 μg g−1) significantly increased the foliage Hg concentrations of shrub than that in background soil (0.06 ± 0.02 μg g−1), although the atmospheric Hg concentrations were regarded as the dominant factor. According to the foliage corresponded organic soil and mineral topsoil Hg concentrations studied by Zhou et al. (2015), we have found that the foliage Hg of fern was also significantly correlated to the soil Hg of the two soil layers (Fig. 4b, c). Similar to the study of Fay and Gustin (2007), the altitudinal foliage Hg distribution for in fern may result from the altitudinal distribution of soil Hg concentrations. Some other studies also showed soil or foliage Hg concentrations significantly correlated with altitude, but the elevation range is relatively higher than that in the current study (Gong et al., 2014; Liu et al., 2019; Szopka et al., 2011). Therefore, some other factors may also influence the altitude Hg distribution, such as soil pH, SOM and growth state.

y=0.28x+43.48 2 r =0.86, p<0.01

150 y=0.08x+80.61 2 r =0.78, p<0.01

100 y=-0.01x+59.20 2 r =0.01, p>0.05

50 300

360

420 Altitude (m)

480

540

Fig. 3. Mercury concentrations in litter, surface soil (0–10 cm) and mineral soil Hg concentrations vs. altitude.

studied Hg accumulation in two tree species (Robinia pseudoacacia (Black locust) and Juniperus scopulorum (Juniper)) and found that atmospheric Hg concentration was the leading factor associated with Hg concentration in foliage. Additionally, the foliage of both species was not significantly different between the tree planted in high Hg (27.7 ± 6.3 μg g−1) and background (0.06 ± 0.02 μg g−1) soils (Fay and Gustin, 2007). The homogenous distribution of Hg in needles was possibly because of atmospheric Hg involved in foliage-level bioaccumulation, while atmospheric Hg may be well mixed and abundant. Thus, needle Hg distributed with little variations with respect to aspect. Unlike Hg in pine needles, we have found that the foliage Hg of fern showed significant correlation with altitude (Fig. 4a), indicating that higher altitude accumulated more Hg loadings, although there were no statically significant differences among the three gradient altitudes

3.4. Estimation of Hg mass balance According to the vegetative cycle of foliage, the litterfall Hg deposition was calculated by the litterfall biomass (overstory + understory) and corresponding Hg concentrations. Empirically region specific allometric relationships relating tree diameter at breast height (D) and needle biomass (B) were used (B = 0.0063 × D2.2014) (Zhang et al., 2006). The D was averaged was 23 cm and the pine density was 0.112 tree m−2 according to previous study (Zhang et al., 2006; Zhou et al., 2016), so the needle biomass is 1080 g m−2. Litterfall mass from pine was 238 g m−2, which was assumed to be 1/3 the needle biomass, because the average longevity time of needle was 3 years (Wang, 2012). The needle litterfall Hg concentration was 147 ng g−1, which was calculated by multiplying the determined Hg concentration in needle 6

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150 (c)

120 100 80 60 200

y=0.10x+53 r =0.16, p<0.01 500 600 2

300

400 Altitude (m)

Linear 95% prediction band

-1

Hg concen. in foliage (ng g )

Linear 95% prediction band

-1

Hg concen. in foliage (ng g )

150 (b)

Linear 95% prediction band

1

Hg concentrations (ng g )

(a) 140

120

90 y=0.12x+76 2 r =0.12, p<0.05

60 80

160 240 320 -1 Hg concen. in organic soil(ng g )

400

120

90

60 0

y=0.17x+85 2 r =0.06, p<0.05 30 60 90 120 150 180 -1 Hg concen. in mineral topsoil(ng g )

Fig. 4. The correlation between Hg concentrations in Old World forked fern and altitude (a), organic soil (b) and mineral topsoil (c).

(98 ng g−1) by a factor of 1.5 (Gong et al., 2014). The needle litterfall Hg deposition was 35 μg m−2 yr−1 based on the calculation the biomass (360 g m−2) and the needle Hg concentration (147 ng g−1). The understory leaf Hg concentration was 146 ng g−1, calculating by a factor of 1.5 of the fresh leaf Hg concentration, and the mass of understory leaf was about 49 g m−2 (Zhou et al., 2016); the understory leaf Hg deposition was 7 μg m−2 yr−1. Therefore, the total litterfall Hg deposition was 42 μg m−2, similar to our litterfall trap study in the altitude of 500 m (41 μg m−2) (Zhou et al., 2018b). Luo et al. (2015) collected throughfall samples by four throughfall (TF) collectors for one year from October 2009 and they showed that the throughfall Hg deposition was 68 μg m−2 yr−1. Thus, annual Hg input (litterfall + throughfall) to the forest was 108 μg m−2 according to the formula (4). Our study in the subtropical forest of TFP found that the annual soilair Hg exchange flux was 258 μg m−2 using dynamics flux chamber (Zhou et al., 2019). The Hg concentration in the stream water was about 3 ± 1 ng L−1 (Zhou et al., 2015), and the amount of surface runoff was 758 L m−2, which are assumed to be 25% rainfall amount (Liu, 2005); therefore, the annual output fluxes from stream water were estimated to be 2.4 μg m−2. Although underground runoff was another major pathway of Hg export from forest because of its amount and Hg concentration generally higher compared to the surface runoff in Chinese forest (Liu, 2005; Luo et al., 2015), there was no direct Hg instigations for underground runoff. Three studies measured Hg concentrations in soil solutions in TFP, which were in the range of 2–60 ng L−1 and averaged 22 ng L−1 (Zhou et al., 2018a). Although no direct measurements studied the export flux of Hg via underground runoff in forested area, the flux was roughly estimated to 68 μg m−2 yr−1 according to the Hg concentration in soil solution (22 ng L−1) and runoff volume (275 L) in TFP. The total output of Hg (surface runoff + underground runoff + soil-air Hg exchange flux) was 33 μg m−2 yr−1. Thus, the Hg retention in forest soil was 76 μg m−2 yr−1, accounting for 69% of the total Hg depositions (Fig. 5), suggesting that forest soils in the southwestern China are net sinks for atmospheric Hg. We hypothesized that the current deposition rates of Hg to soils were used to estimate the soil accumulation rate and the deposited Hg was mainly accumulated in the surface horizon (0–10 cm). According to the bulk density, the soil mass of the surface horizon (0–10 cm) was averaged about 1.16 × 102 kg m−2. Therefore, the soil Hg accumulation rate was about 0.86 ng g−1 yr−1 in the subtropical forest.

Fig. 5. Mercury pools in soil profiles and Hg budget at the TFP Masson pine forest.

forest of China (Fu et al., 2010; Luo et al., 2015). Secondly, Hg accumulation has been examined as strongly related to the biogeochemical cycles of carbon in forest soils (Richardson and Friedland, 2015). However, in the subtropical forest, soil organic matter turnover rates are rapid and little SOM accumulates because of warm and moist climate (Zhou et al., 2018b), simultaneously resulting in Hg loss during the carbon mineralization. The Hg MRT in the mineral soil (change to soil profile of 60 cm, MRT: 417 ± 28 yr) was comparable to that of boreal forest in North America (soil profile of 60 cm, 386 ± 57 yr) (Richardson and Friedland, 2015). The reason may be that, compared to background area of North America, although higher Hg deposition was observed, our study forest had similar Hg pools in mineral soils and our output fluxes from soil Hg volatilization and runoff were also elevated (Choi and Holsen, 2009; Ericksen et al., 2006).

3.5. Hg residence time in surface and mineral soils The calculated MRT in the mineral horizon (20–80 cm, 595 ± 34 yr) were much longer at surface soil horizon (0–10 cm, 134 ± 14 yr) due to significantly larger Hg pools at mineral horizons (p < 0.001). The surface horizon MRT was much smaller than that in boreal forest (90–386 yr) of northeastern United States (Demers et al., 2007; Richardson and Friedland, 2015; Yu et al., 2014). Firstly, the difference likely arose from large Hg deposition to soil in subtropical 7

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4. Conclusion

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The foliar Hg concentrations displayed much higher Hg concentrations, suggesting urban industry emitted large amount of atmospheric Hg to surrounding terrestrial ecosystem. The foliage Hg deposited with litterfall increased the atmosphere-to-forest input of Hg in TFP, which was much higher compared to remote forest in North America and China. Budget of Hg in the forest indicated that about 69% from the total atmospheric Hg deposition was resided in the soils. Therefore, it can be deduced that TFP was large sinks of atmospheric Hg. Short residence time of Hg in surface soils enhanced the deposited Hg output from the forest, which may the reason for the similar soil Hg concentrations in some other remote forests. The large Hg pools in the subtropical forest may resulted in animals in forest ecosystems is not uniquely adapted to this environment and may be highly vulnerable to increased Hg loads. Therefore, further detailed researches are required to investigate the Hg in food webs at forest catchment with highly elevated atmospheric Hg retentions. Acknowledgements This work was financially supported by the National Natural Science Foundation of China (41807385), National Key Research and Development Program of China (2018YFD0800302) and National Key Technology Research and Development Program of China (2015BAD05B01). We acknowledge two anonymous referees for their valuable comments on our manuscript. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.gexplo.2019.106337. References Agnan, Y., Le Dantec, T., Moore, C.W., Edwards, G.C., Obrist, D., 2016. New constraints on terrestrial surface atmosphere fluxes of gaseous elemental mercury using a global database. Environ. Sci. Technol. 50, 507–524. Bavec, S., Biester, H., Gosar, M., 2018. A risk assessment of human exposure to mercurycontaminated soil and household dust in the town of Idrija (Slovenia). J. Geochem. Explor. 187, 131–140. Beckers, F., Rinklebe, J., 2017. Cycling of mercury in the environment: sources, fate, and human health implications: a review. Crit. Rev. Environ. Sci. Technol. 47, 693–794. Buch, A.C., Sisinno, C.L.S., Correia, M.E.F., Silva-Filho, E.V., 2018. Food preference and ecotoxicological tests with millipedes in litter contaminated with mercury. Sci. Total Environ. 633, 1173–1182. Bushey, J.T., Nallana, A.G., Montesdeoca, M.R., Driscoll, C.T., 2008. Mercury dynamics of a northern hardwood canopy. Atmos. Environ. 42, 6905–6914. Chen, Y., Yin, Y., Shi, J., Liu, G., Hu, L., Liu, J., Cai, Y., Jiang, G., 2017. Analytical methods, formation, and dissolution of cinnabar and its impact on environmental cycle of mercury. Crit. Rev. Environ. Sci. Technol. 47, 2415–2447. Choi, H.-D., Holsen, T.M., 2009. Gaseous mercury emissions from unsterilized and sterilized soils: the effect of temperature and UV radiation. Environ. Pollut. 157, 1673–1678. Cristol, D.A., Brasso, R.L., Condon, A.M., Fovargue, R.E., Friedman, S.L., Hallinger, K.K., Monroe, A.P., White, A.E., 2008. The movement of aquatic Mercury through terrestrial food webs. Science 320, 335. Demers, J.D., Driscoll, C.T., Fahey, T.J., Yavitt, J.B., 2007. Mercury cycling in litter and soil in different forest types in the Adirondack region, New York, USA. Ecol. Appl. 17, 1341–1351. Dittman, J.A., Shanley, J.B., Driscoll, C.T., Aiken, G.R., Chalmers, A.T., Towse, J.E., Selvendiran, P., 2010. Mercury dynamics in relation to dissolved organic carbon concentration and quality during high flow events in three northeastern US streams. Water Resour. Res. 46. Du, B., Li, P., Feng, X., Qiu, G., Zhou, J., Maurice, L., 2016. Mercury exposure in children of the wanshan mercury mining area, Guizhou, China. Int. J. Environ. Res. Public Health 13. Du, B., Feng, X., Li, P., Yin, R., Yu, B., Sonke, J.E., Guinot, B., Anderson, C.W.N., Maurice, L., 2018. Use of mercury isotopes to quantify mercury exposure sources in inland populations, China. Environ. Sci. Technol. 52, 5407–5416. Ericksen, J.A., Gustin, M.S., Xin, M., Weisberg, P.J., Fernandez, G.C.J., 2006. Air-soil exchange of mercury from background soils in the United States. Sci. Total Environ. 366, 851–863. Evans, C.A., Hutchinson, T.C., 1996. Mercury accumulation in transplanted moss and lichens at high elevation sites in Quebec. Water Air Soil Pollut. 90, 475–488.

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