Precipitation Hg concentrations and deposition fluxes
Concentrations of Hg in precipitation and corresponding precipitation depth
are presented in Fig. 2. A large variation in precipitation Hg concentrations
was observed at all sampling sites, with the maximum concentrations up to an
order of magnitude higher than the minimum concentrations. The VWM Hg
concentrations in precipitation at the remote sites varied from 3.7 to
7.7 ng L-1 (mean: 5.6 ± 2.0 ng L-1, Table 1), with the
highest VWM Hg concentration observed at the BYBLK site and the lowest at the
MAL and MDM sites. The VWM Hg concentration in precipitation at the urban
site of GY was 11.9 ± 6.1 ng L-1, which was 1.5 to 3.2 times
higher than the values at remote sites (Table 1). We acknowledge that, due to
the lack of long-term simultaneous observations, the variation in VWM Hg
concentrations in precipitation among the sampling sites may have
uncertainties. In the present study, precipitation samples at the urban and
rural sites were collected throughout 1 to 3 years, and there might exist
interannual variations in VWM Hg concentrations at each sampling site. For
example, the maximum annual VWM Hg concentration (5.1 ng L-1) at MAL
was observed during June 2013–May 2014, which was approximately 1.8 times
higher than that (2.9 ng L-1) during June 2011–May 2012. At MCB,
annual VWM Hg concentrations were highest (8.1 ng L-1) during August
2012–July 2013 and lowest (6.0 ng L-1) during August 2013–July 2014.
VWM Hg concentrations in precipitation at all sites showed a clear season
trend, with lower concentrations in the summer wet season and higher
concentrations in the winter dry season (Fig. 3). This pattern is consistent
with previous observations in rural and urban areas of China (Huang et al.,
2012, 2013; Ma et al., 2016). Higher Hg concentrations in precipitation
during the winter dry season were potentially due to elevated wintertime
atmospheric PBM concentrations in China (Fu et al., 2008; Zhang et al., 2013;
Xu et al., 2014; Zhu et al., 2014), which could be incorporated into wet
deposition via scavenging processes below cloud. Lower VWM Hg concentrations
in precipitation during the summer wet season were mostly associated with
higher precipitation amounts at the sampling sites, suggesting increasing
amounts of precipitation would dilute the Hg concentrations in samples that
were scavenged from the boundary layer during the onset of the precipitation
(Gratz et al., 2009; Yuan et al., 2015).
Annual fluxes of Hg in precipitation at the sampling sites varied from 2.0 to
12.6 µg m-2 yr-1 (mean:
5.9 ± 3.6 µg m-2 yr-1, Table 1). Wet deposition
fluxes showed a clear urban–rural difference, with the annual deposition
flux at the urban site of GY elevated by a factor of 1.8 to 6.3 compared to
the values at rural sites. This could be partly due to the elevated VWM Hg
concentration in precipitation at the GY site. Wet deposition fluxes at rural
sites also showed a clear regional difference. The annual wet deposition
fluxes of Hg in the subtropical zones in southwestern and eastern China
(i.e., MAL, MLG, and MDM) were relatively higher (by a factor of 1.1 to 1.3)
than that at the MCB site in the temperate zone in northeastern China, and
much higher (by a factor of 3 to 3.6) than that at the MWLG and BYBLK sites,
which were located in the arid/semi-arid zones in northwestern China. This
regional variation could not be explained by the difference of VWM Hg
concentrations in precipitation because the correlation between annual wet
deposition fluxes of Hg and VWM Hg concentrations in precipitation is not
significant (p > 0.05). Instead, annual wet deposition
fluxes of Hg were positively correlated with annual precipitation depth at
the remote sites (r2=0.86, p < 0.01). This suggests
that precipitation depth had a greater influence on the regional variation of
wet deposition fluxes of Hg at remote sites of China than VWM Hg
concentrations, which is in agreement with previous studies in North America
(Risch et al., 2012; L. Zhang et al., 2012).
The VWM Hg concentrations in precipitation at the remote sites of this study
were overall consistent with previous observations in China (Fig. 4). For
example, VWM Hg concentrations in precipitation and wet deposition fluxes of
Hg at the Nam Co and SET stations of the Tibetan Plateau and in Mt. Simian,
southwestern China, ranged from 4.0 to 10.9 ng L-1 and from 1.8 to
15.4 µg m-2 yr-1, respectively (Huang et al., 2012,
2015; Ma et al., 2016). However, the VWM Hg concentration at the GY site was
1.0–4.4 times lower than the levels (12.3–52.9 ng L-1) observed in
other urban areas of China, and the wet deposition fluxes of Hg at the GY
site were consequently lower than those
(14.0–56.5 µg m-2 yr-1) in the urban areas of China,
with the exception of the flux observed in Lhasa of the Tibetan Plateau
(flux: 8.2 µg m-2 yr-1) (Wang et al., 2009, 2012; Huang
et al., 2013; Xu et al., 2014; Zhu et al., 2014).
Litterfall Hg concentrations and deposition fluxes
Average Hg concentrations in litterfall at the MCB, MDM, MLG, and MAL sites
were 47.0 ± 19.0, 42.3 ± 5.6, 91.1 ± 29.4, and 56.9±4.4 ng g-1, respectively (mean: 59.3 ± 22.0 ng g-1,
Table 2). Concentrations of Hg in litterfall could be affected by many
factors, including atmospheric Hg concentrations, tree species, a “leaf
maintenance” period, and environmental factors (Lindberg and Stratton, 1998;
Frescholtz et al., 2003; Ericksen and Gustin, 2004; Millhollen et al., 2006;
Poissant et al., 2008). The variation in litterfall Hg concentrations
observed in different collectors (corresponding to sampling of different tree
species) at each sampling site was insignificant (p values for all
> 0.05, Table 2). Annual mean atmospheric TGM at the MCB, MDM,
MLG, and MAL sites was 1.73 ± 0.48, 3.31 ± 1.44, 2.80±1.51,
and 2.09 ± 0.63 ng m-3 (Fu et al., 2015b), respectively, which
was not significantly correlated with the Hg concentrations in litterfall
samples (p = 0.87). The lack of significant correlation might be
partly attributed to the fact that mean atmospheric TGM concentration during
the “leaf maintenance” period for each tree species was not entirely equal
to the annual means of atmospheric TGM at the sampling sites because of its
seasonal variations (Fu et al., 2015b). In addition, other factors, including
tree species, a “leaf maintenance” period, and regional environmental
factors, should also play a more important role in litterfall Hg
concentrations. Hg concentrations in litterfall at the MDM site were found to
increase from July to December (correlation slope =
7.0 ± 0.7 ng g-1 mon-1, r2=0.78,
p < 0.01; litterfall was not collected during
January–June due to little production of litterfall biomass). In contrast,
significant monthly variation in Hg concentrations in litterfall samples at
the MCB and MAL sites was not found (p values for both
> 0.05).
Annual fluxes of Hg through litterfall at the four sampling sites ranged from
22.8 to 62.8 µg m-2 yr-1 (mean of
37.0 µg m-2 yr-1, Table 2). The litterfall fluxes of Hg
showed a clear regional distribution pattern, with the fluxes decreasing with
latitude. The highest flux (62.8 µg m-2 yr-1) was
observed at the MAL site in the southern subtropical zone in southwestern
China, followed by the MLG site (flux:
39.5 µg m-2 yr-1) in the middle subtropical zone, the
MDM site (flux: 23.1 µg m-2 yr-1) in the northern
subtropical zone, and the MCB site (flux:
23.1 µg m-2 yr-1) in the middle temperate zone (Zheng
et al., 2010). The relatively higher litterfall flux of Hg at the MAL and MLG
sites could be explained by either higher annual biomass of litterfall or
higher Hg concentrations in litterfall samples (Table 2). Deposition fluxes
of Hg through litterfall in this study were comparable to those
(35.5–42.9 µg m-2 yr-1) measured in Mt. Gongga and
Mt. Simian, southwestern China (Fu et al., 2010b; Ma et al., 2016), but
substantially lower than that (220 µg m-2 yr-1)
measured at Tieshanping, which was close to Chongqing, southwestern China
(Wang et al., 2009).
Annual VWM Hg concentrations in precipitation and wet deposition
fluxes of Hg in this study and the literature (Huang et al., 2012, 2013,
2015; Xu et al., 2014; Zhu et al., 2014; Ma et al., 2016). Black and red
values above the columns correspond to VWM concentration and wet deposition
of Hg, respectively.
Relative contribution of wet and litterfall deposition to total
Hg deposition in forests
Ratios of annual mean litterfall deposition flux to annual wet deposition
flux of Hg at the four sampling sites ranged from 3.9 to 8.7 (mean: 5.8±2.3). The ratios were overall consistent with the previous observations (2.8
to 7.6) in China (Wang et al., 2009; Fu et al., 2010b; Ma et al., 2016). On
the other hand, the observed ratios in China were much greater than those
observed in North America and Europe. Rich et al. (2012) collected litterfall
at 23 remote sites in the eastern USA and found that the mean ratio of
litterfall Hg deposition to Hg wet deposition was 1.3 (ranged from 0.4 to
2.6), which was 3.0 to 6.7 times lower compared to the ratios observed in
China. In Europe, ratios of litterfall Hg deposition to wet Hg deposition
were in the range of 0.4 to 2.6 (mean: 1.2 ± 0.8; n=5) (Iverfeldt,
1991; Munthe et al., 1995; Lee et al., 2000; Schwesig and Matzner, 2000).
Comparison of (a) wet deposition flux of Hg; (b)
litterfall deposition of Hg; (c) atmospheric total gaseous mercury
(TGM)/gaseous elemental mercury (GEM) concentrations; (d)
atmospheric particulate bound mercury (PBM) concentrations; and (e)
atmospheric gaseous oxidized mercury (GOM) concentrations between China and
North America and Europe. Note that atmospheric PBM in Europe is referred to
as total particulate bound mercury and in the remaining regions is referred
to as particulate bound mercury on particles with an aerodynamic diameter of
< 2.5 µm. Data are from this study, the literature, and
references therein (EMEP; Munthe et al., 1995, 2003; Lee et al., 2000;
Schwesig and Matzner, 2000; St. Louis et al., 2001; Pirrone et al., 2003;
Swartzendruber et al., 2006; Wang, 2006; Yatavelli et al., 2006; Demers et
al., 2007; Lindberg et al., 2007; Valente et al., 2007; Bushey et al., 2008;
Choi et al., 2008; Larssen et al., 2008; Li et al., 2008; Fain et al., 2009;
Peterson et al., 2009; Prestbo and Gay, 2009; Song et al., 2009; Wang et al.,
2009; Engle et al., 2010; Sprovieri et al., 2010; Liu et al., 2011; Fisher
and Wolfe, 2012; Fu et al., 2012b; Juillerat et al., 2012; Risch et al.,
2012; X. T. Zhang et al., 2012; Chen et al., 2013; Zhu et al., 2014; Fu et
al., 2015b, 2016; Ma et al., 2016).
Correlations between wet deposition fluxes of Hg and (a)
volume-weighted mean (VWM) Hg concentrations in precipitation, (b)
atmospheric total gaseous mercury (TGM)/gaseous elemental mercury (GEM)
concentrations, (c) atmospheric particulate bound mercury (PBM)
concentrations, and (d) atmospheric gaseous oxidized mercury (GOM)
concentrations in China. Data are from this study and the literature (Wang,
2006; Fu et al., 2011, 2015b; Huang et al., 2012; Wang et al., 2012; Zhu et
al., 2012, 2014; Huang et al., 2013; Xu et al., 2014, 2015).
Hg in litterfall biomass has been suggested to be mostly from uptake of
atmospheric TGM, and therefore litterfall deposition could be a good
indicator of TGM dry deposition to forest ecosystems (Frescholtz et al.,
2003; Gustin, 2012; L. Zhang et al., 2012). In addition to litterfall and wet
deposition, dry deposition of PBM and GOM to the forest floor and other
surfaces could also contribute to the total Hg deposition to a forest. Given
the measured atmospheric PBM and GOM concentrations (Fu et al., 2015b; Yu et
al., 2015), the dry deposition fluxes of PBM and GOM at the MCB, MDM, and MAL
sites were estimated to be 3.0, 9.6, and
2.3 µg m-2 yr-1, respectively, using the average dry
deposition velocities of PBM and GOM over forests modeled by L. Zhang et
al. (2012). At the MLG site, annual dry deposition flux of PBM and GOM was
estimated to be 4.4 µg m-2 yr-1 using the comparison of
precipitation and throughfall data collected side by side (St. Louis et al.,
2001; Fu et al., 2010a; Gustin, 2012). The importance of litterfall in the
total deposition of Hg has been highlighted by many previous studies (St.
Louis et al., 2001; Lindberg et al., 2007; Risch et al., 2012; L. Zhang et
al., 2012). In this study, we estimate that litterfall deposition represented
60–87 % (mean: 74.5 ± 11.4 %) of total Hg deposition to the
four studied forests, which were much higher compared to those (mean:
46.2 ± 12.5 %) over rural forests in North America and Europe
(Munthe et al., 1995; Rea et al., 1996; Grigal et al., 2000; Lee et al.,
2000; St. Louis et al., 2001; L. Zhang et al., 2012), whereas the wet
deposition played a minor role (mean: 13.9±3.5 %) in the total Hg
deposition budget. Therefore, Hg deposition through litterfall played a
predominant role in the total Hg deposition budget in forest ecosystems in
China.
Comparison with observations in other regions worldwide
Figure 5 shows the comparison of wet deposition and litterfall fluxes of Hg
as well as TGM/GEM, PBM, and GOM concentrations in China, North America, and
Europe. The mean wet deposition flux of Hg at remote sites in China was 5.6±4.2 µg m-2 yr-1 (Fig. 5a, data from this study and
the literature, Huang et al., 2012, 2015; Ma et al., 2016), which is 4.4
times lower than the mean
(24.8 ± 17.8 µg m-2 yr-1) at urban sites of China
(Fig. 5a, data are from this study and the literature, Wang et al., 2009,
2012; Huang et al., 2013; Xu et al., 2014; Zhu et al., 2014). The mean wet
deposition fluxes of Hg in North America and Europe were 9.5 ± 4.2 and
6.8 ± 3.2 µg m-2 yr-1 (Fig. 5a), respectively
(EMEP; Prestbo and Gay, 2009). In contrast to the observations in China, the
urban–rural variation in the wet deposition fluxes of Hg was insignificant
in North America (L. Zhang et al., 2012).
The observations from this study and the literature suggested that wet
deposition fluxes of Hg in urban areas of China were highly elevated (by a
factor of 2.6 to 3.6) compared to North America and Europe. In China, wet
deposition fluxes of Hg were significantly correlated with VWM Hg
concentrations in precipitation (r2=0.87, p < 0.01,
Fig. 6a), whereas no significant correlation existed between wet deposition
fluxes of Hg and annual precipitation depth (r2=0.02,
p = 0.65). Elevated wet deposition fluxes of Hg at urban sites
of China were associated with the elevated VWM Hg concentrations in
precipitation (Fig. 6a). Wet deposition fluxes of Hg in China were also
positively correlated with ground-level TGM/GEM, PBM, and GOM concentrations
(Fig. 6b, c, and d). Wet deposition of Hg has been suggested to result from
the scavenging of PBM and GOM in cloud (i.e., rainout) and below cloud (i.e.,
washout) (Seigneur et al., 2004; Lin et al., 2006). In North America, a
modeling study suggests that scavenging of GOM in and below cloud contributed
mostly (∼ 89 %) to wet deposition of Hg, with ∼ 41 %
contributed by washout (Selin and Jacob, 2008). In China, ground-based
measurements of GOM in urban areas found that the mean GOM concentrations
(means: 47.9 pg m-3, Fig. 5e, Fu et al., 2011; Xu et al., 2015) were
5.4 times greater than the mean (8.9 pg m-3) in North America
(Swartzendruber et al., 2006; Yatavelli et al., 2006; Valente et al., 2007;
Fain et al., 2009; Peterson et al., 2009; Song et al., 2009; Engle et al.,
2010; L. Zhang et al., 2012). Scavenging of GOM in the continental boundary
layer (i.e., washout) would therefore contribute to the elevated wet
deposition fluxes of Hg at urban sites of China. It should be noted that PBM
concentrations were also highly elevated (mean: 239±102 pg m-3)
in the urban areas of China, which were 10–20 times greater than the levels
observed in North America and Europe, and approximately 5 times greater than
the mean GOM concentrations at the same locations (Fig. 5d and e). Lee et
al. (2001) estimated that washout of PBM contributed approximately
1.0 µg m-2 yr-1 to the total Hg deposition in the
United Kingdom at a background PBM concentration of 10 pg m-3. Given
the mean PBM concentration in urban areas of China, the mean flux of washout
of PBM below cloud is roughly estimated to be
24 µg m-2 yr-1, which explains approximately 90 %
of the mean wet Hg deposition flux in the urban areas of China. Since
scavenging of PBM below cloud also depends on other factors, including the
vertical distribution of PBM, intensity of precipitation, and cloud base
height (Tanner et al., 1997; Hicks, 2005; Brooks et al., 2014), the estimate
may have large uncertainties. Nevertheless, the estimate is in agreement with
the measured fraction of particulate mercury (Hgp) in wet
deposition flux of Hg at an urban site in China. Huang et al. (2013) found
that ∼ 86 % of the annual wet deposition of Hg in Lhasa of the
Tibetan Plateau was associated with Hgp, much higher than that at
a rural site in the Tibetan Plateau (55 %, Huang et al., 2015) as well as
at rural and urban sites in North America (26–63 %, Burke et al., 1995;
Lamborg et al., 1995; Poissant and Pilote, 1998). These suggest the
scavenging of PBM below cloud was an important contributor to the elevated
wet deposition fluxes of Hg at urban sites of China.
On the other hand, mean wet deposition flux of Hg
(5.6 ± 4.2 µg m-2 yr-1) at the rural sites of
China was relatively lower (by a factor of 1.2 to 1.7) compared to those
measured in North America (9.5±4.2 µg m-2 yr-1)
and Europe (6.8 ± 3.2 µg m-2 yr-1) (EMEP; Prestbo
and Gay, 2009). This regional pattern is different from model results that
predicted higher wet deposition in China because of large anthropogenic Hg
emissions (Bergan et al., 1999; Dastoor and Larocque, 2004). There are
several possible explanations for the lower wet deposition fluxes of Hg
observed in the rural areas of China. Wet deposition fluxes of Hg at the
rural sites of China were mostly observed in arid, semi-arid, and sub-humid
climate zones in northwestern and northeastern China (i.e., MWLG, MCB, and
BYBLK in this study, Nam Co, and SET stations), where the precipitation depth
is generally low (260–975 mm, data from this study and the literature,
Huang et al., 2012, 2015) and anthropogenic Hg sources are scarce
(X. T. Zhang et al., 2015). The remaining four rural sites (i.e., MDM, MLG,
and MAL in this study, and Mt. Simian, Ma et al., 2016) were all located in
mountaintop forests. Although the observations at some of these sites showed
elevated PBM concentrations (mean: 31–154 pg m-3) (Fu et al., 2015b;
Yu et al., 2015), washout of PBM below cloud was not expected to contribute
significantly to Hg in precipitation because of low cloud base heights (Ray
et al., 2006). In addition, observations of GOM at high-altitude sites in
China (i.e., MWLG, MAL, and Shangri-La) showed mean concentrations of
2–8 pg m-3 (Fu et al., 2012a, b; H. Zhang et al., 2015),
significantly lower than those (20–87 pg m-3) measured at
high-altitude sites in North America and Europe (Swartzendruber et al., 2006;
Fain et al., 2009; Weiss-Penzias et al., 2009; Fu et al., 2016). Relatively
lower GOM concentrations at the high-altitude sites in China were possibly
due to the elevated atmospheric particulate matters in China that facilitate
the partitioning of GOM to the particulate phase (Slemr et al., 2009;
Swartzendruber et al., 2009; van Donkelaar et al., 2010; Amos et al., 2012;
Zhang et al., 2013). Since the scavenging of GOM in the free troposphere and
continental boundary layer is an important source of wet deposition of Hg
(Selin and Jacob, 2008), the lower GOM concentrations in the rural areas of
China could be responsible for the lower wet deposition fluxes of Hg observed
in the rural areas of China.
Scatterplot of (a) Hg concentrations in litterfall and
litterfall fluxes of Hg, and (b) litterfall biomasses and litterfall
fluxes of Hg for the global observations. Data are from this study and the
literature (Iverfeldt, 1991; Rea et al., 1996; Roulet et al., 1998; Fostier
et al., 2000; Grigal et al., 2000; Schwesig and Matzner, 2000; St. Louis et
al., 2001; Mélières et al., 2003; Magarelli and Fostier, 2005;
Sheehan et al., 2006; Silva-Filho et al., 2006; Demers et al., 2007; Wangberg
et al., 2007; Bushey et al., 2008; Larssen et al., 2008; Wang et al., 2009;
Fu et al., 2010b; Fisher and Wolfe, 2012; Juillerat et al., 2012; Risch et
al., 2012; Teixeira et al., 2012; Benoit et al., 2013; Ma et al., 2016).
Annual fluxes of Hg through litterfall at the rural sites in this and
previous studies in China ranged from 22.8 to
62.8 µg m-2 yr-1 (mean:
37.8 ± 14.8 µg m-2 yr-1, n=6, Fig. 5b; data
are from this study and the literature, Fu et al., 2010b; Ma et al., 2016).
Hg fluxes through litterfall in the rural areas of China were 1.4–4.7 times
higher than the means observed in North America
(13.3 ± 5.8 µg m-2 yr-1) and Europe
(16.5 ± 8.7 µg m-2 yr-1) (Munthe et al., 1995;
Rea et al., 1996; Lee et al., 2000; Schwesig and Matzner, 2000; St. Louis et
al., 2001; Lindberg et al., 2007; Larssen et al., 2008; Fisher and Wolfe,
2012; Juillerat et al., 2012; Risch et al., 2012), but approximately 2.2
times lower than those (mean:
84.4 ± 49.0 µg m-2 yr-1) measured in South
America (Roulet et al., 1998; Fostier et al., 2003; Mélières et al.,
2003; Magarelli and Fostier, 2005; Silva-Filho et al., 2006; Teixeira et al.,
2012). Global Hg fluxes through litterfall were positively correlated with
both Hg concentrations in litterfall (r2=0.69,
p < 0.01) and litterfall biomass production (r2=0.70, p < 0.01) (Fig. 7). Forward stepwise multiple
regression analysis suggests that litterfall biomasses and Hg concentrations
in litterfall explained 69.2 and 25.4 % of the regional variations in
litterfall Hg fluxes, respectively. Productions of litterfall biomasses at
the rural sites of China ranged from 434 to 1100 g m-2 yr-1
(mean: 661 ± 307 g m-2 yr-1, n=6) and were
approximately 2 times higher than that in North America and Europe (Munthe et
al., 1995; Rea et al., 1996; Lee et al., 2000; Schwesig and Matzner, 2000;
St. Louis et al., 2001; Lindberg et al., 2007; Larssen et al., 2008; Fisher
and Wolfe, 2012; Juillerat et al., 2012; Risch et al., 2012), which is the
dominant factor causing the difference in litterfall Hg fluxes between China
and North America/Europe. It is worth noting that most (five out of the six)
observations at the rural sites of China were made in subtropical moist
forests, where the litterfall biomass productions are larger than those in
the temperate and boreal forests in North America and Europe (Xiong and
Nilsson, 1997; Running et al., 2004; Wang et al., 2008). Additionally, mean
Hg concentrations in litterfall at the rural sites of China
(63.3 ± 29.0 ng g-1) were elevated by a factor of 1.4 compared
to that (44.0 ± 10.4 ng g-1) in North America and Europe (Munthe
et al., 1995; Rea et al., 1996; Lee et al., 2000; Schwesig and Matzner, 2000;
St. Louis et al., 2001; Lindberg et al., 2007; Larssen et al., 2008; Fisher
and Wolfe, 2012; Juillerat et al., 2012; Risch et al., 2012). This could be
partly attributed to the elevated TGM concentrations (Fig. 5c) and longer
“leaf maintenance” period at most rural sites in China (Frescholtz et al.,
2003; Poissant et al., 2008; Gustin, 2012; Fu et al., 2015b).