Atmospheric gaseous elemental mercury (GEM) concentrations and mercury depositions at a high-altitude mountain peak in south China

China is regarded as the largest contributor of mercury (Hg) to the global atmospheric Hg budget. However, concentration levels and depositions of atmospheric Hg in China are poorly known. Continuous measurements of atmospheric gaseous elemental mercury (GEM) were carried out from May 2008 to May 2009 at the summit of Mt. Leigong in south China. Simultaneously, deposition fluxes of THg and MeHg in precipitation, throughfall and litterfall were also studied. Atmospheric GEM concentrations averaged 2.80 ±1.51 ng m−3, which was highly elevated compared to global background values but much lower than semirural and industrial/urban areas in China. Sources identification indicates that both regional industrial emissions and long range transport of Hg from central, south and southwest China were corresponded to the elevated GEM level. Seasonal and diurnal variations of GEM were observed, which reflected variations in source intensity, deposition processes and meteorological factors. Precipitation and throughfall deposition fluxes of THg and MeHg in Mt. Leigong were comparable or lower compared to those reported in Europe and North America, whereas litterfall deposition fluxes of THg and MeHg were higher compared to Europe and North America. This highlights the importance of vegetation to Hg atmospheric cycling. In th remote forest ecosystem of China, deposition of GEM via uptake of foliage followed by litterfall was very important for the depletion of atmospheric Hg. EleCorrespondence to: X. Feng (fengxinbin@vip.skleg.cn) vated GEM level in ambient air may accelerate the foliar uptake of Hg through air which may partly explain the elevated litterfall deposition fluxes of Hg observed in Mt. Leigong.


Introduction
Mercury (Hg), especially its organic forms (e.g. methylmercury (MeHg) and dimethylmercury (DMeHg)), is a highly toxic pollutant that poses a serious threat to human health and wildlife (National Research Council, 2000). Atmospheric Hg consists of three chemical and physical forms, including gaseous elemental Hg (GEM or Hg 0 ), divalent reactive gaseous Hg (RGM), and particulate Hg (PHg). Unlike other heavy metals, which tend to exist in the atmosphere in the particulate phase, Hg exists mainly (>95%) in the gaseous phase (total gaseous mercury (TGM), TGM=GEM+RGM) (Schroeder and Munthe, 1998;Poissant et al., 2005;Gabriel et al., 2005;Aspmo et al., 2006;Valente et al., 2007). RGM and PHg are more reactive and readily scavenged via wet and dry deposition. However, GEM, the predominant form of atmospheric Hg (generally constitutes more than 90% the total Hg in atmosphere), is very stable in atmosphere with a residence time between 6 month and 2 years (Schroeder and Munthe, 1998). This enables Hg to undergo a long range transport and makes it well-mixed in a global scale. As such, long range transport of Hg in the atmosphere has been identified as the predominant source of Hg in pristine ecosystem in remote areas.
Published by Copernicus Publications on behalf of the European Geosciences Union.  (Shetty et al., 2008, and reference therein) and sampling site.
Wet and dry depositions are very important pathways for scavenging of atmospheric Hg. Because PHg and RGM have the significantly high surface reactivity and water solubility and GEM is very stable in atmosphere, dry and wet deposition of Hg in atmosphere is largely dominated by RGM and PHg. In recently studies, however, dry deposition of atmospheric GEM to forest canopies is increasingly recognized as an important sink for atmospheric Hg. For example, Zhang et al. (2005) reported that the deposition flux of GEM to leaf surface constituted over 99% of total atmospheric Hg loss to vegetation, while St. Louis et al. (2001) and Graydon et al. (2009) found litterfall deposition of Hg constituted about 60% of the total Hg deposition in the forest of experimental lake area (ELA) in Canada . Because Hg adsorbed by root of plant could hardly be translocated from roots to leaf due to the barrier effect of root zone, Hg in leaves should be considered to come from atmosphere (Ericksen et al., 2003;Greger et al., 2005;Selvendiran et al., 2008;Bushey et al., 2008). Therefore, it is reasonable to believe that deposition of GEM to vegetation followed by litterfall deposition is an important sink of atmospheric Hg. Methyl mercury (MeHg) deposition via wet and dry deposition generally constitutes a small portion of total Hg deposition St. Louis et al., 2001). Sources of MeHg in precipitation include capture of MeHg and/or oxidation of dimethyl mercury to MeHg, but the extents of both processes are typically low in the atmosphere (Brosset et al., 1995;Lee et al., 2003).
Since the industrial revolution, global Hg emissions have increased significantly (Fitzgerald, 1995;Mason and Sheu, 2002). In China, many attempts have been made to decrease Hg emissions from coal combustion, smelting activities, cement production, etc. However, China is still regarded as the biggest emission source of atmospheric Hg in the world (Pacyna et al., 2006;Street et al., 2005;Wu et al., 2007). Anthropogenic Hg emissions in China show a clear regional distribution pattern. As shown in Fig. 1, central, east and south China are major atmospheric Hg source regions, due to higher population density, proximity to industrial sources and generally higher energy consumption (Street et al., 2005;Wu et al., 2007).
To understand the regional budget of atmospheric Hg and the chemical and physical processes in the atmosphere, it is important to determine spatial and long-tem temporal variability of atmospheric Hg concentrations and deposition fluxes. Numerous studies with regard to atmospheric Hg have been carried out at many sites in North America and Europe (e.g., Poissant et al., 2005;Valente et al., 2007;Sigler et al., 2009;Rutter et al., 2009). However, to the best of our knowledge, only a few long-term monitoring studies of atmospheric Hg and deposition fluxes have been performed in semi-rural and urban/industrial areas of China. Fu et al. (2008a) and Wan et al. (2009a) reported atmospheric TGM concentrations in two semi-rural areas of China (Mt. Gongga in southwest China and Mt. Changbai in northeast China, respectively) were approximately two times higher than the common background values in North America and Europe (1.5-2.0 ng m −3 , Travnikov, 2005; Kim et al., 2005;Valente et al., 2007). In addition, studies in semi-rural and urban areas of China also showed extremely high Hg deposition fluxes (Guo et al., 2008;. These results suggested that many urbanized areas of China are exposed to atmospheric Hg contaminations due to regional anthropogenic emissions. However, there are still limitations to fully describe temporal and spatial distributions of Hg in China and its relationship to global atmospheric Hg cycling. Hence, there is a great need to conduct long-term continuous measurements of atmospheric Hg and deposition fluxes in remote areas of China. In this study, we present atmospheric GEM data derived from year-long measurements along with precipitation, throughfall and litterfall deposition Hg fluxes at a highaltitude mountain peak in remote area of south China. The major goals of this study are three-fold: 1) to characterize the regional background level of atmospheric GEM as well as deposition fluxes of Hg in south China; 2) to evaluate the regional sources and long range transport affecting the GEM concentrations; 3) to discuss the deposition and sink of atmospheric Hg in the forest ecosystem in China.

Site description
The sampling site was located at the summit of Mt. Leigong (26.39 • N,108.20 • E, 2178 m above sea level), which is the highest mountain in southeast Guizhou province in southwest China (Fig. 1). Mt. Leigong is an isolated peak with an elevation of about 1000 m against the surrounding landmass. The surrounding areas are naturally preserved semitropical evergreen broadleaf forests and semi-tropical deciduous broadleaf and coniferous forests. Mt. Leigong has a subtropical climate, with distinct rainy (May to October) and dry (November to April) seasons. Annual mean air temperature and precipitation depth at the peak of Mt. Leigong are 9 • C and 1400-1700 mm, respectively. Misty weather prevails at the peak of Mt. Leigong, and the number of days with cloud generally exceeds 300 days per year.
The sampling site was relatively isolated from human activities; however, several industrial areas and population centers, which might contribute to significant atmospheric Hg release, are located to the west of the sampling site ( Fig. 1). Guiyang, the capital of Guizhou province, is located about 160 km to the west of the sampling site. The nearest population center is Leishan County (Population: 33 000), which is located 20 km to the southwest but at an elevation of 1300 m below the sampling site. Kaili city, the capital city of the Southeast Guizhou Miao-Dong Autonomous Prefecture, is the biggest population center (population: 520 000) and industrial area in the surrounding area of Mt. Leigong located about 35 km to the northwest of the sampling site.

Measurements of atmospheric GEM
Real time continuous (every 10 min) measurements of GEM were made between 9 May 2008 and 18 May 2009 using an automated Hg vapor analyzer (Tekran 2537A) . Its technique is based on the collection of TGM (GEM+RGM) on gold traps, followed by thermal desorption, and detection of Hg 0 by cold vapor atomic fluorescence spectrometry (λ=253.7 nm). The instrument features two cartridges which trap gaseous Hg on to gold absorbents. While one cartridge is adsorbing Hg during sampling period, the other is being desorbed thermally and analyzed subsequently for TGM. The functions of each cartridge are then reversed, allowing continuous sampling of ambient air. PHg in ambient air was removed using a 45 mm diameter Teflon filter (pore size 0.2 µm). In this study, the measured TGM concentration was probably dominated by GEM because GEM generally has a concentration level at least two order of magnitude higher than RGM especially in remote areas Poissant et al., 2005;Valente et al., 2007;Fu et al., 2008b). Previous study by Swartzendruber et al. (2006) and Faïn et al. (2009) reported the intrusion of RGM from free troposphere could highly increase atmosphere RGM concentrations at high elevation mountain peaks occasionally. However, RGM from free troposphere would be probably scavenged quickly by the frequent could contact at Mt. Leigong and might not increase RGM concentrations significantly. Moreover, RGM in ambient air was likely removed when passing the sampling tube, which should have very high humidity in it and was installed with a soda lime before entering the Tekran instrument. Therefore, the atmospheric Hg measured herein was referred to as GEM. Precision (Relative standard deviations) of the sampling system is better than 2% and the absolute detection limit is about 0.1 pg (Tekran, 2002). A Teflon sampling tube with its inlet 8 m above the ground was employed at the sampling site. To mitigate the influence of low atmospheric pressure on the pump's strain, a low sampling rate of 0.75 l min −1 (at standard temperature and pressure) was used during the whole sampling period. The data quality of Tekran Model 2537A was controlled via periodic internal recalibration with a 25 h interval, and the internal permeation source was calibrated every 2 months (after the field measurement study, the external check on the permeation source were within 95.8% (n=5) of expected values).

Sampling method and analysis of precipitation and throughfall
Precipitation samples were collected from May 2008 to May 2009 at an open-air site near the atmospheric GEM sampling site at the peak of Mt. Leigong. Simultaneously, throughfall samples were also collected from a Cuculidae forest located within 30 m from the precipitation sampling site. Precipitation and throughfall samples were collected by using a sampler with an acid-washed borosilicate glass bottle and a borosilicate glass wide-mouthed (15 cm in diameter) jar supported in a PVC housing system (developed from Oslo and Paris Commission 1998). Collectors were set out manually just prior or within 15 min of the beginning of a precipitation event. Just following the end of a precipitation event, collectors were sealed using Polyethylene bags to prevent contamination of Hg dry deposition to the collectors. Precipitation and throughfall samples were kept in the collectors over one week during the sampling. Each week, samples were transferred carefully to acid-cleaned Teflon sample bottles (volume: 250 mL) and preserved by adding tracemetal grade HCl (to 5‰ of total sample volume). To ensure clean operation, polyethylene gloves were used throughout the setup and collection processes. Teflon bottles with samples were individually sealed into three successive polyethylene bags and rapidly brought to the laboratory and stored in a refrigerator until analysis. Before each of the new sampling cycle, the sampling collectors were rigorously rinsed by Milli-Q water or replaced by new collectors as necessary.
In this study, both total mercury (THg) and methylmercury (MeHg) concentrations in precipitation and throughfall samples were determined following US EPA Method 1631 (US EPA, 1999) and Method 1630 (US EPA, 2001), respectively. THg was analyzed by BrCl oxidation followed by SnCl 2 reduction, and dual amalgamation combined with CVAFS detection (US EPA, 1999), while MeHg was determined by using distillation, aqueous phase ethylation and GC separation followed by pyrolysis and CVAFS detection (US EPA, 2001). The detection limits of THg and MeHg were 0.15 ng L −1 and 0.03 ng L −1 , respectively, which were determined by three times the standard deviation of blanks. Field blanks (n=10) were determined by rinsing the whole sampling collectors with Milli-Q water and then collecting the rinsing water into the 250-mL Teflon bottles as was made for samples to ensure that there was no contamination by sampling collectors, sampling Teflon bottles, and HCl preservative. The overall average THg and MeHg concentrations of field blanks were 0.32 and 0.011 ng L −1 , respectively. Precision and accuracy test for the analytical method was made using recoveries on duplicate samples (n=12). The recoveries of THg and MeHg were in the ranges of 96-111% and 95-120%, respectively.

Sampling method and analysis of litterfall
Three typical forests (Cinnamomum camphora (L.) Presl forest, Rhododendron simsii Planch forest, and Fargesia spathacea Franch forest) located at the peak of Mt. Leigong were selected to collect litterfall samples by using three 0.25 m 2 litterfall collectors (St. Louis et al., 2001). Litterfall samples were collected monthly, packed into paper bags and air-dried in a clean environment near the sampling site. Monthly litterfall samples from each site were completely combined to analyze Hg concentrations in litterfall and calculate annual mass flux of each species.
Air-dried litterfall samples were ground to a fine powder in a pre-cleaned food blender and stored in a clean environment in the laboratory until analysis. Between grinding, the blender was extensively cleaned with Mili-Q water and ethanol to prevent any cross contaminations. THg concentrations in litterfall samples were determined by acid digestion followed by oxidation, purge and trap, and cold vapor atomic absorption spectrophotomety (CVAS). Approximately 0.2 g sample was digested in 10 mL of freshly mixed HNO 3 /H 2 SO 4 (4:1 v/v) for 6 h at 95 • C in a water bath. The digested solution was then diluted by adding Mili-Q water to a volume of 50 mL and analyzed for THg. For MeHg analysis, approximately 0.2 g of ground sample was digested for 3 h at 75 • C in polyethylene bottles containing 5 mL of 25% KOH in methanol (Liang et al., 1996). After cooling to room temperature, MeHg was extracted with methylene chloride and back-extracted from the solvent phase into water, and then the aqueous phase was ethylated for determination of MeHg (Liang et al., 1995(Liang et al., , 1996. Quality assurance and quality control were conducted using duplicates, method blanks, matrix spikes, and certified reference material (Tort-2, lobster reference material was used since reference material for plants was not available in our lab). The analytical detection limits were 4 ng g −1 for THg and 0.2 ng g −1 for MeHg in samples, respectively. Recoveries on matrix spikes of MeHg in samples were in the range of 78-119%. The relative percentage difference was <15% in duplicate samples. An average MeHg concentration of 171±15 ng g −1 (n=3) was obtained from Tort-2 which was comparable to the certified value of 152±13 ng g −1 .

Meteorological data and backward trajectory calculation
In this study, wind direction and wind speed were measured using a portable weather station located within 5 m from the TGM sampling site with a time resolution of 30 min. Precipitation depths were measured daily using standard graduated rain gauge, while throughfall depths were obtained by comparing the collected volume with precipitation. Precipitation depth was measured in an open area near its sampling site. The annual precipitation depth from May 2008 to May 2009 was 1533 mm. Throughfall depth was 1182 mm and constituted about 77% (ranging between 60% and 98%) of direct precipitation depth. There was no snow event throughout the sampling campaign because of relative high air temperature and little rainfall in the winter. In order to identify the long range transport of Hg and its influences on the atmospheric Hg distribution at the study site, three-day back trajectories were calculated using Gridded Meteorological Data combined with the Geographic Information System (GIS) based software, Trajstat, from HYSPLIT ; http://www.arl.noaa.gov/ ready/hysplit4.html). The Global Data Assimilation System (GDAS) meteorological data archives of the Air Resource Laboratory, National Oceanic and Atmospheric Administration (NOAA), which are available on line at (ftp: //arlftp.arlhq.noaa.gov/pub/archives/gdas1/) were used as the input. All the back trajectories ended at the sampling site (26.39 • N, 108.20 • E) with an arrival height of 500 m above the ground. The back trajectories were calculated at 6-h interval (00:00, 06:00, 12:00, 18:00 UTC; i.e. 08:00, 14:00, 20:00, 02:00 local time), and then used for further analysis. Figure 2 shows the time series of 10-min averaged GEM concentrations in ambient air for the entire study period (9 May 2008 to 18 May 2009). The distribution of GEM was characterized by significant variations throughout the sampling campaign. GEM concentrations followed a lognormal distribution pattern (Fig. 3), and values between 1.5-4.5 ng m −3 accounted for approximately 85% of the total frequency. However, episodes with extremely high GEM concentrations were abundant (0.4% of values were greater than 10 ng m −3 ), and 62% of these events was observed in the daytime. These episodes were probably attributed to encounters of air masses from industrial and urbanized areas which probably related to heavily atmospheric Hg pollutions (Kim and Kim, 2001;Liu et al., 2002;Feng et al., 2004). As shown in Sect. 3.4, the sampling site was mainly affected by air masses from regional boundary layer in the daytime and air masses in the free troposphere probably related to long range transport in the night. Considering the site was isolated from industrial actitivities, most of these events were caused by strong Hg sources from industrial and urban areas in the surrounding area in Mt. Leigong whereas others were related to distant sources.

General characteristics of GEM in atmosphere
The annual geometric mean of GEM concentrations at the study site was 2.80±1.51 ng m −3 (median=3.03 ng m −3 ) and ranged from 0.41 to 23.9 ng m −3 , which are much higher than those observed from different remote areas in Europe and North America (generally lower than 2.0 ng m −3 , Travnikov, 2005; Kim et al., 2005;Valente et al. 2007). On the other hand, average GEM concentrations in Mt. Leigong are comparable or lower than those observed in remote areas in China and other Asian countries. The mean TGM concentrations in Mt. Gongga in southwest China and Mt. Changbai in northeast China were 3.98±1.62 ng m −3 and 3.58±1.78 ng m −3 , respectively (Fu et al., 2008a;Wan et al., 2009a). The level of GEM in Mt. Leigong was within the range (0 to 6.3 ng m −3 ) measured in the upper troposphere and lower stratosphere around China, Korea and Japan (Friedli et al., 2004), but significantly lower than the levels reported in a Global Atmospheric Watch (GAW) station located on An-Myun Island, Korea (4.61±2.21 ng m −3 , Nguyen et al., 2007). Higher GEM levels in China and other countries in East Asia were probably attributed to the large emissions of Hg in these areas in which industries areas and population centers were often densely distributed.  Generally, levels of atmospheric GEM in remote areas are closely related to regional atmospheric Hg emissions. The highly elevated GEM concentrations in rural areas of China probably indicate significant emissions from coal combustion, non-ferrous smelting, cement production, etc. (Pacyna et al., 2006;Street et al., 2005;Wu et al., 2007). However, regional levels of atmospheric GEM in China are not consistent with regional atmospheric Hg emission inventories (Street et al., 2005;Wu et al., 2007). Mt. Changbai and Mt. Gongga are located in northeast China and quaternary section of the eastern Qinghai-Tibet plateau and its transit zone to Sichuan province, respectively, which are within the low atmospheric Hg emission regions in China (Fig. 1). However, these two areas showed relatively higher GEM concentrations compared to Mt. Leigong. The major reason for the discrepancy between regional GEM levels and Hg emission inventories is that the sampling sites at Mt. Changbai and Mt. Gongga were affected by regional and local atmospheric Hg emissions (Fu et al., 2008a;Wan et al., 2009a), which probably overestimated the regional backgrounds of atmospheric Hg. In Mt. Gongga, the sampling site was effected by both local domestic Hg emissions and Hg emissions from Zinc and Lead smelting factories in Shimian city located within 50 km from the study site (Fu et al., 2009), while the sampling site in Mt. Changbai was probably contaminated by local sources because our subsequent measurements in the other site in this area exhibited much lower GEM concentrations (mean: 1.59±0.68 ng m −3 , from October 2008 to September 2009, unpublished data).

Deposition fluxes of Hg in precipitation, throughfall and litterfall
Totally, we collected 32 weekly samples throughout the sampling campaign. We intended to monitor sampling on routine basis; however, interruptions were inevitable for certain periods. The samples during 23-29 June, 18-24 August    (Fig. 4), and the overall volumeweighted mean concentration was 4.0 ng L −1 . By comparison, THg concentrations in all throughfall samples throughout the period were higher than precipitation levels (Fig. 4). The annual volume-weighted mean THg concentration for throughfall was 8.9 ng L −1 with a range of 2.8-32.5 ng L −1 . In general, vegetation is a sink for atmospheric Hg (Erichsen et al., 2003;Bushey et al., 2008). Zhang et al. (2005) and Poissant et al. (2008) found that foliage was a sink of all atmospheric Hg species and deposition of atmospheric Hg to foliar surfaces are enhanced as atmospheric Hg concentrations increased. Generally, both stomatal and nonstomatal pathways are important ways for atmospheric Hg to enter vegetation leaves (Stamenkovic and Gustin, 2009). However, not all atmospheric Hg deposited to the foliar surface is assimilated and fixed by foliage, and most of the PHg and RGM are probably washed off the leaf surface or reduced and then reemitted to the atmosphere (Rea et al., 2001). Therefore, elevated THg concentration in throughfall was mostly attributed to the dry deposition of PHg and RGM to vegetation followed by washout of throughfall. MeHg concentrations in precipitation ranged from below the detection level to 0.15 ng L −1 with a volume-weighted concentration of 0.04 ng L −1 . Like THg, MeHg in throughfall was elevated compared to direct precipitation; MeHg in throughfall ranged from 0.02-0.40 ng L −1 and the volumeweighted mean was 0.10 ng L −1 . This was probably due to dry deposition of MeHg to foliage. The average fractions of THg present as MeHg in precipitation and throughfall were 1% and 1.1%, respectively, which were comparable or lower than values reported for Europe and North America (Porvair and Verta, 2003;Hall et al., 2005;Witt et al., 2009), as well as an upland forest in Mt. Gongga in southwest China (Fu et al., 2010).
Concentrations of THg and MeHg in litterfall were highly variable and ranged from 57-110 and 0.44-0.80 ng g −1 , respectively. The distribution of THg in litter types was consistent with MeHg. The lowest THg and MeHg concentrations were measured in Rhododendron simsii Planch, whereas the concentrations levels in the other two litter types were comparable. The fractions of Hg present as MeHg in throughfall samples ranged from 0.65 to 0.77%, which are lower than those of direct precipitation and throughfall. This suggests that uptake of MeHg from precipitation by foliage is relatively small.
THg and MeHg concentrations in precipitation and throughfall in the study area were considerably lower than those reported in rural and semi-rural areas of China (Guo et al., 2008;Wan et al., 2009b;Fu et al., 2010). This is because the sampling site was isolated from sites near human activities. However, it is interesting that THg and MeHg concentrations at Mt. Leigong were comparable or lower than those reported in North America and Europe (Schwesig and Matzner, 2000;St. Louis et al., 2001;Hall et al., 2005;Choi et al., 2008;Witt et al., 2009). This is quite different from atmospheric GEM concentrations, which were elevated compared to sites in Europe and North America. This result suggests that atmospheric GEM had less immediate effect on direct wet deposition and throughfall compared to PHg and RGM. Since the study site was not a durably impacted receptor for direct anthropogenic Hg emission sources, the likelihood for steadily elevated RGM and PHg levels here are limited. These airborne fractions are transients in atmospheric boundary layer and have a limited role in the corresponding long-range transport. Although not measured, we may therefore speculate that the contribution of RGM and PHg to airborne Hg levels and deposition fluxes at Mt. Leigong is small. On the other hand, GEM could undergo long range transport from emission sources and cause elevated levels at the sampling site. This may explain the low precipitation THg concentrations and high atmospheric GEM concentrations in Mt. Leigong.
Besides, precipitation THg concentration at the peak of Mt. Leigong was also significantly lower compared to a nearby site (within 10 km) (19.5 ng L −1 ,  located at a lower elevation (1400 m above sea level), indicates that there is a significant difference in the source of rain because of mountain climate or that wash out of atmospheric Hg during a rain event constitutes a major portion of wet deposition compared to cloud process. THg concentrations in litterfall at Mt. Leigong were much higher than those from North America and Europe (Schwesig and Matzner, 2000;St.Louis et al., 2001;Hall and St.Louis et al., 2004;Sheehan et al., 2006), which was consistent with results for atmospheric GEM. This may suggest that Hg content in vegetation leaf was significantly related to atmospheric GEM concentration.
Fluxes of THg and MeHg in precipitation and throughfall were estimated using the volume-weighted concentrations and depths of precipitation and throughfall, and litterfall deposition fluxes of THg and MeHg were estimated using the litterfall Hg concentrations and litter mass fluxes. The annual THg deposition flux was 6.1 µg m −2 yr −1 for direct precipitation, 10.5 µg m −2 yr −1 for throughfall and 39.5 µg m −2 yr −1 for litterfall. The annual MeHg deposition flux was 0.06 µg m −2 yr −1 for direct precipitation, 0.12 µg m −2 yr −1 for throughfall and 0.28 µg m −2 yr −1 for litterfall.
litterfall deposition flux of THg in the forest area of Mt. Leigong was much higher than values observed from forests in North America and Europe (Grigal et al., 2000;Larssent et al., 2008;Graydon et al., 2009). We attributed this to the high GEM concentrations in the study area. As suggested by Zhang et al. (2005) and Poissant et al. (2008), atmospheric Hg 0 is almost the exclusive source of Hg in vegetation leaf. Highly elevated litterfall deposition fluxes suggests that in remote forest areas of China, deposition of atmospheric Hg 0 via uptake by vegetation leaf was the major pathway for the depletion of atmospheric Hg. More-  over, the great deposition via litterfall also constituted a very important source of Hg in forest ecosystem of China. Previous study by Obrist (2007) suggested that Hg deposited to forest might be probably reemitted to atmosphere during carbon mineralization. Our study in the upland forest of Mt. Gongga (2500 m above sea level) found that emission of Hg from litterfall (the sampling site was covered by 1-2 cm undecomposed leaf litter. Mean flux was 0.5 ng m −2 h −1 in August 2006) was much lower compared to bare soil (Fu et al., 2008c). Using this data, the annual emission flux of litterfall in Mt. Leigong was calculated to be 4.4 µg m −2 yr −1 , (which was probably overestimated because evasion of Hg from litter in cold season was probably lower than the values in warm season), indicating that the reemission during litter decomposition was very small and most of the Hg in litter probably retained in the forest soil.

Seasonal variation of atmospheric GEM and Hg deposition fluxes
Monthly geometric mean GEM concentrations in Mt. Leigong are shown in Fig. 5a. GEM concentrations showed a clear seasonal cycle with high concentrations in cold seasons and low concentrations in warm seasons. The lowest monthly geometric mean GEM concentration (1.52±1.06 ng m −3 ) was observed in July 2008, and then increased consistently and reached the highest monthly value (4.42±0.95 ng m −3 ) in January 2008, which was nearly threefold higher than that in July 2008. The seasonal variation of GEM concentrations was in the descending order of winter, spring, autumn and summer (Table 2). Low summer concentrations and elevated winter concentrations were reported for south and southwest China (Feng et al., 2004;Fu et al., 2008a). A previous study in the Mt. Gongga area suggested enhanced coal and biomass  combustion played a significant role in elevated TGM concentrations in cold season (Fu et al., 2008a(Fu et al., , 2009. Enhanced coal and biomass combustion in cold seasons is generally driven by the need for residential heating in China. Compared to industrial and urban areas in which the coal consumption is dominated by industrial and power plant activities which are often evenly distributed in each month, the increase of coal consumption in Mt. Leigong area in cold seasons could be more significant because domestic use of coal is dominant. This was probably one of most important reasons for the highly elevated GEM concentrations in Mt. Leigong in cold seasons. Moreover, the seasonal variability may also be due to long range transport of atmospheric Hg during monsoons. In cold seasons, the wind field in the study area was dominated by winter monsoons, which originated from or passed through central China (Fig. 6a), one of the most Hg-polluted regions in China (Street et al., 2005;Wu et al., 2007). The air masses would probably capture a large amount of Hg during transport and cause elevated GEM concentrations in cold seasons. In the warm season, air masses originating from the ocean dominated, although sometimes the inland air masses occasionally affected the study site (Fig. 6b). Marine air masses in warm seasons likely diluted the atmospheric Hg in the study area. Precipitation deposition fluxes of THg and MeHg also showed clear seasonal variations (Fig. 5b). The seasonal distribution pattern for both THg and MeHg fluxes were in the descending order: summer > spring > autumn > winter ( Table 2). The highest monthly THg wet deposition flux of 1573 ng m −2 mon −1 was observed in August 2008,

Diurnal variation of atmospheric GEM
GEM concentrations in Mt. Leigong exhibited a noticeable diurnal pattern as shown in Fig. 7. GEM concentrations decreased consistently from mid-night (at 00:00 in China central time (CCT)) to 2.5-2.6 ng m −3 at 05:30, and remained at this level for about 2 h until 07:30. After sunrise, GEM concentrations increased quickly to 2.9-3.0 ng m −3 at 12:00 and remained constant until 16:00. After that, GEM concentrations decreased consistently to the lowest values before sunrise. Totally, the annual nighttime geomean GEM concentration (2.72 ng m −3 ) was 6% lower compared to that in the daytime (2.89 ng m −3 ).
The diurnal distribution of GEM in Mt. Leigong gives insight into the interplay of regional emissions and long range transport of Hg influencing the GEM concentrations. Since the sampling site was located at the peak of Mt. Leigong, mountain valley breezes played a significant role in the wind direction. During the daytime, the air mass in the low altitude area of Mt. Leigong is heated because of increasing solar radiation, resulting in an upslope flow that brings air from the boundary layer to the mountain peak; while at night, air adjacent to mountain peak cools faster than air in the low altitude area, causing a reversal flow (downslope), which enables transport of air masses from the free troposphere to the sampling site. It is clear that air masses from the boundary layer and the free troposphere both showed elevated GEM concentrations compared to background values in Northern Hemisphere (Travnikov, 2005;Kim et al., 2005;Valente et al., 2007). Relatively higher GEM concentrations in air masses from boundary layer indicated a large amount of atmospheric Hg emission in the surrounding areas of Mt. Leiong. The elevated GEM concentrations in the free troposphere were probably caused by the updraft, which carried polluted air masses from the boundary layer. On the other hand, since the air masses in the free troposphere have a higher velocity (Fig. 7), the highly Hg polluted air in south China could probably be transported long distances and caused pollution problems in other regions. Figure 8a shows wind roses at the sampling site, which indicated wind directions at the study site were dominantly from northeast, southeast and southwest, which may reflect the three predominant monsoons in China including Southeast monsoon, Southwest monsoon and Asian winter monsoon. Distributions of GEM concentrations in each of the wind direction were quite different from each other. In general, northeast wind and southwest wind exhibited higher GEM events, while southeast wind exhibited lower GEM events. Wind dependence of atmospheric GEM was probably attributed to an interplay of regional sources and long range transport of Hg.

Regional sources
As discussed in Sect. 3.4, the sampling site was dominantly affected by the valley wind during the daytime, which transported regional air masses from the boundary layer. To gain a better understanding of the influences of regional sources on GEM distributions in Mt. Leigong, daytime GEM dependence on wind direction is shown in Fig. 8b. In general, high mean GEM concentrations were observed mostly from west, and this indicates the influence from regional sources in this direction. Due to high coal Hg content, non-ferrous smelting activities and artisanal Hg mining activities, Guizhou is re- garded as one of the highest atmospheric Hg source regions in China (Street et al., 2005;Wu et al., 2007;Feng and Qiu, 2008;Li et al., 2009). Kaili (located 35 km northwest of Mt. Leiong) is the largest city in the study area, and may be affected by elevated Hg emissions (Liu et al., 2002;Feng et al., 2004). Besides, several smelting factories were located around the city. Therefore, it is an important regional source for the study area. Leishan (20 km southwest of Mt. Leigong) is the nearest population centre of the study site, which is also an important source region influencing the study site. Moreover, the densely populated areas and industrial areas were located to the west of the sampling sites, which definitely contributed to the elevated concentrations in this direction. Except for the northeast direction which was probably affected by the artisanal Hg mining activities in Tongren city , air flow from the east and south showed low GEM levels. The lowest mean GEM concentrations were observed from the southeast direction, which was probably because this area is more naturally preserved and less populated.

Air mass back trajectories analysis for long range transport
In order to gain an insight to the influence of long range transport on distribution of GEM in Mt. Leigong, we calculated 3-day air mass back trajectories using the Gridded Meteorological Data combined with the free software Trajstat from HYSPLIT website ; http://www.arl.noaa. gov/ready/hysplit4.html). The 3-day back trajectories arriving at the study site over the study period were grouped into four clusters, which are shown in Fig. 9. Cluster 1 consists of air masses originating from the continental inland areas of central Asia, passing over the north and central China.
Cluster 2 shows air masses originated from south China. Air masses in cluster 3 were originated from South China Sea, then passed over Guangxi province. Cluster 4 shows air masses originated from southwest China and then passed over Guizhou province. Contributions of the above four types of air masses were quite different throughout the year. Cluster 1 had the highest  frequency (50.8%) of all the four groups, while cluster 4 contributed least (4.6%) to the total air masses (Fig. 9). Frequencies of cluster 2 and 3 were 22.1% and 22.5%, respectively. Since Mt. Leigong was mainly affected by regional sources during daytime because of the prevalent valley breeze, nighttime measurements were used in the calculation of mean GEM concentrations for clusters. For the four types of air masses, cluster 1 was related to the highest GEM concentrations (3.43 ng m −3 ). Air masses in cluster 1 passed over the central China plain region, which is the most densely populated and heavily Hg polluted area in China due to industrial and domestic coal combustion, smelting industries, cement production, biomass burning, etc. For example, Henan and Hunan provinces were the first and fourth biggest Hg source provinces in China (Wu et al., 2007), respectively, which together accounted for about 15% of the total Hg emissions. Highly elevated GEM concentrations, together with the highest occurrence, may be an important reason for elevated GEM level in Mt. Leigong. Mean GEM concentration in cluster 2 was 2.72 ng m −3 , which was considerably lower compared to cluster 1. This is probably because the air masses originated or passed over the border area of the three provinces of Guangdong, Guangxi and Hunan, which is generally less populated and less developed. Air masses in cluster 3 showed the lowest mean GEM concentration of 2.03 ng m −3 , which was probably attributed to the oceanic origins of these air masses. Air masses in cluster 4 were also heavily polluted with regard to Hg, with a mean concentration of 3.35 ng m −3 , slightly lower compared to cluster 1 result. This result might be explained by the elevated Hg emissions in Guizhou province, the second biggest atmospheric Hg source region in China.
In general, GEM distribution in the study area was controlled by Hg emissions in the regional boundary layer and Hg levels in the free troposphere which were probably related to well-mixed long range transport air masses. As discussed in Sect. 3.4, the sampling site was affected by the alternation of mountain breeze and valley breeze, which were probably related to long range transport and regional Hg sources, respectively. Using the lowest GEM concentrations in the night and highest concentrations in the daytime, we could simply speculate that the levels of GEM in regional boundary layer and well-mixed long range transport air masses in the free troposphere (assuming to be the regional base-case level) were about 2.94 (mean GEM concentration between 12:00 and 16:00 in the daytime) and 2.56 ng m −3 (mean GEM concentration between 05:30 and 07:30 in the night), respectively. the high GEM concentration (0.38 ng m −3 higher than regional base-case concentration) in regional boundary layer indicates that regional Hg emissions played an important role in Mt. Leigong. However, it should be pointed out that the GEM concentration in well-mixed long range transported air masses in the free troposphere was elevated by 1.0 ng m −3 compared to background values in the Northern Hemisphere (assuming to be 1.5 ng m −3 ). Therefore, we suggest that long range transport played a more significant role in Mt. Leigong compared to regional sources.

Summary and conclusions
Measurements of GEM in ambient air and Hg deposition fluxes were investigated at Mt. Leigong, a high-altitude peak in south China from May 2008 to May 2009. Atmospheric GEM levels were highly elevated compared to background values observed in the Northern Hemisphere (1.5-2.0 ng m −3 ) with an overall geometric mean concentration of 2.80 ng m −3 . A distinct seasonal distribution pattern was observed with GEM concentrations with higher levels in winter and lower levels in summer, and this was probably attributed to seasonal variations in anthropogenic emission sources, meteorological conditions and atmospheric scavenging processes (transformation and deposition). Diurnal variations in GEM concentrations were observed with higher concentrations in the daytime and lower levels at night and were related to mountain valley breeze circulation. The prevalent valley wind during the daytime carried polluted air masses from regional boundary layer and increased GEM levels at the sampling site, while during the night the sampling site was mainly infused with mountain wind which carried fresh air from free troposphere.
Annual means of THg and MeHg concentrations in precipitation were 4.0 and 0.04 ng L −1 , respectively. THg and MeHg concentrations in throughfall were more than twofold higher than precipitation, with the annual means of 8.9 and 0.10 ng L −1 , respectively. Precipitation deposition fluxes of THg and MeHg in Mt. Gonggga were 6.1 µg m −2 yr −1 and MeHg 0.06 µg m −2 yr −1 , which were comparable or lower compared to those reported in Europe and North America. Deposition fluxes of THg and MeHg were 10.5 µg m −2 yr −1 and MeHg 0.12 µg m −2 yr −1 for throughfall and 39.5 µg m −2 yr −1 and 0.28 µg m −2 yr −1 for litterfall, respectively.
The study site was affected by both regional emissions and long range transport of Hg. Regional emissions of Hg included coal combustion and smelting activities which were generally located in the west of Mt. Leigong. Our results indicate Mt. Leigong may be affected by both continental inland monsoon and Southeast monsoon, which carried Hg polluted air masses from central China and Guizhou province, whereas air masses from south China were generally related to low atmospheric Hg concentrations because of oceanic flow.