Environmental Engineering Reference
In-Depth Information
In this chapter, we review and synthesize the current
state of knowledge on inputs, outputs, and stores of THg
and MeHg in the terrestrial landscape, including wetlands,
as well as our current understanding of how Hg moves
through catchments and its transformations en route.
Wetlands have long been recognized as important methyl-
ation sites, but here we also present the emerging evidence
for methylation in terrestrial uplands. Following the litera-
ture, our treatment is weighted toward forested landscapes
in temperate and high latitudes, but we consider agricul-
tural, urban, and tropical landscapes to the extent possible.
In this context we also discuss the effect of disturbance
(such as forest harvesting and urbanization) on Hg and
MeHg fl uxes. Finally, we consider the fate of this legacy Hg
and its sensitivity to future changes in Hg emissions, land
use, and climate.
Much of this dry deposition occurs by forest canopy scav-
enging of Hg(0), which enters leaf stomata at a fairly con-
stant rate and binds to foliar tissue throughout the growing
season (Rea et al., 2002; Miller et al., 2005). The leaf cuti-
cle has been proposed as an alternative Hg(0) entry point
(Stamenkovic and Gustin, 2009; Converse et al., 2010). In
contrast, dry deposition of RGM and HgP occurs on external
foliar surfaces (Krabbenhoft et al., 2005; Miller et al., 2005).
Together, these mechanisms result in dry Hg deposition in
forests that exceeds wet Hg deposition, in some cases by sev-
eralfold (Munthe et al., 1995; Lee et al., 1998; Kolka et al.,
1999; St. Louis et al., 2001). Hg in litterfall appears to be pri-
marily a new, not recycled, input of atmospheric Hg (Bushey
et al., 2008; Graydon et al., 2009). Conversely, dry Hg deposi-
tion is lower in nonforested areas and smaller yet to water
surfaces (Miller et al., 2005). Operationally, dry deposition of
Hg(0) is commonly quantifi ed as the Hg in litterfall, while dry
deposition of Hg(II) is quantifi ed as the Hg in net throughfall
(throughfall Hg minus open precipitation Hg) (Driscoll et al.,
1994; Miller et al., 2005; Risch et al., 2012). Enrichment of Hg
in the snowpack under forest canopy shows that throughfall
is important even in winter (Nelson et al., 2010).
MeHg in wet deposition is on the order of 1% of THg
(Lee et al., 1998). In catchments without signifi cant inter-
nal net MeHg generation, this input may be quantitatively
suffi cient to account for stream MeHg output, though it is
unlikely that MeHg in deposition transits through a catch-
ment conservatively. Dry deposition of MeHg also appears
to be signifi cant; St. Louis et al. (2004) found a twofold
enrichment in annual MeHg under the forest canopy
relative to open spaces. Nearly all of the excess MeHg
was in litterfall, but it is not clear whether it is taken up
by stomata or binds to external foliar surfaces. Schwesig
and Matzner (2001) found that litterfall supplied only one
third of the annual input of THg, but more than half of the
annual input of MeHg.
Mercury Inputs
Hg is a global pollutant because atmospheric transport
and deposition effectively connect anthropogenic emis-
sions to the most remote areas of the globe. Some of this
anthropogenic Hg is deposited directly to water bodies, but
most falls on the surrounding terrestrial landscape, where
it is a potential source to freshwater ecosystems via run-
off. Moreover, deposition per unit area is greater to land
than to water surfaces because the forest canopy and other
vegetative surfaces tend to scavenge gaseous Hg from the
atmosphere more effectively than water surfaces (Allan and
Heyes, 1998; Miller et al., 2005).
Atmospheric deposition of Hg is treated thoroughly in
chapter 6. Here we present the basic concepts relevant to
terrestrial Hg cycling. Over the past decade, an important
debate has taken place in the literature about whether
freshly deposited atmospheric Hg (“new Hg”) is more labile
and biologically available than Hg that has been incorpo-
rated into soils and vegetation (“aged Hg”) (Krabbenhoft
et al., 2004). The implications are important, because if
only new Hg bio-accumulates, reductions in Hg emissions
would have an immediate ecologic benefi t.
Atmospheric Hg enters watersheds via wet deposition,
in rain or snow, or as dry deposition, by various processes,
when it is not precipitating. Most Hg emissions from both
natural and anthropogenic sources are in the form of gas-
eous Hg(0), which is relatively unreactive and has a 1-year
average residence time in the atmosphere. A smaller por-
tion is emitted as ionic Hg(II), either as reactive gaseous
mercury (RGM) or particulate Hg (HgP). Hg(II) has a much
shorter residence time; model results suggest that more
than 60% of the RGM and more than 20% of HgP depos-
its within 1000 km of its emission source (Cohen et al.,
2007). Wet deposition of Hg at a given site originates partly
from the globally distributed Hg(0) pool and partly from
regional emissions of Hg(II).
Atmospheric deposition of Hg, in contrast to most other
elements, occurs primarily as dryfall in most landscapes.
Mercury Stocks
The affi nity of Hg for organic matter governs its distribution
on the landscape. We consider here Hg stores in vegetation
and soils (forest fl oor and mineral soils), including peatlands.
Vegetation
In forested landscapes, a relatively small pool of Hg (~2%
of the total soil pool) is retained by living vegetation and
associated coarse woody debris (Grigal, 2003). This pool
is about 4 times the annual Hg deposition, suggesting an
average 4-year residence time of Hg in vegetation. How-
ever, more than half of this Hg is stored in tree boles and
has a much longer residence time, while ~15% is present
in foliage, with a much shorter residence time. In the
METALLICUS study, where isotopically distinct Hg inputs
could be directly tracked, 66% of new Hg input remained
in aboveground vegetation after 1 year (Hintelmann et al.,
