Gworek et al.

Environ Sci Eur (2020) 32:128

https://doi.org/10.1186/s12302-020-00401-x

REVIEW Open Access

Mercury in the terrestrial environment:

a review
Barbara Gworek1, Wojciech Dmuchowski1 and Aneta H. Baczewska‑Dąbrowska2*

Abstract
Background: Environmental contamination by mercury is and will continue to be a serious risk for human health.
Pollution of the terrestrial environment is particularly important as it is a place of human life and food production. This
publication presents a review of the literature on issues related to mercury pollution of the terrestrial environment:
soil and plants and their transformations.
Results: Different forms of atmospheric Hg may be deposited on surfaces by way of wet and dry processes. These
forms may be sequestered within terrestrial compartments or emitted back into the atmosphere, and the relative
importance of these processes is dependent on the form of Hg, the surface chemistry, and the environmental condi‑
tions. On the land surface, Hg deposition mainly occurs in the oxidized form (­ Hg2+), and its transformations are associ‑
ated primarily with the oxidation–reduction potential of the environment and the biological and chemical processes
of methylation. The deposition of Hg pollutants on the ground with low vegetation is as 3–5 times lower than that
in forests. The estimation of Hg emissions from soil and plants, which occur mainly in the H ­ g0 form, is very difficult.
Generally, the largest amounts of Hg are emitted from tropical regions, followed by the temperate zone, and the low‑
est levels are from the polar regions. Areas with vegetation can be ranked according to the size of the emissions as fol‑
lows: forests > other areas (tundra, savannas, and chaparral) > agricultural areas > grassland ecosystems; areas of land
devoid of vegetation emit more Hg than those with plants. In areas with high pollution, such as areas near Hg mines,
the Hg content in soil and plants is much higher than in other areas.
Conclusions: Mercury is recognized as a toxic, persistent, and mobile contaminant; it does not degrade in the envi‑
ronment and becomes mobile because of the volatility of the element and several of its compounds. Atmospheric
contamination by mercury continues to be one of the most important environmental problems in the modern world.
The general conclusions were drawn from a review of the literature and presented in this paper.
Keywords: Mercury, Deposition and emissions from land, Soil pollution, Content in plants

Background Over the last few decades, considerable scientific

Mercury is recognized as a toxic, persistent, and mobile knowledge has been developed on the sources and emis-
contaminant; it does not degrade in the environment and sions of mercury, its pathways and cycling through the
becomes mobile because of the volatility of the element environment, human exposure, and impacts on the envi-
and several of its compounds. Moreover, mercury has the ronment and human health [2]. Hg is the only element
ability to be transported within air masses over very long in the periodic table to have its own environmental con-
distances [1]. vention, i.e., the Minamata Convention on Mercury, thus
highlighting the importance of the Hg pollution issue [3].
An improved understanding of the global mercury
*Correspondence: a.baczewska‑dabrowska@obpan.pl (Hg) cycle is important for our capacity to predict
2
Polish Academy of Sciences Botanical Garden – Center for Biological
Diversity Conservation in Powsin, Prawdziwka 2 St, 02‑973 Warsaw, Poland how regulatory efforts to reduce current emissions
Full list of author information is available at the end of the article to air, water and land will affect Hg concentrations in

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Gworek et al. Environ Sci Eur (2020) 32:128 Page 2 of 19

environmental compartments, biota and humans. Hg is turbations caused by major anthropogenic drivers of
released into the environment through human activi- environmental change?
ties and via natural sources and processes, such as vol- 2. What is the relative risk of mercury exposure to
canoes and rock weathering. Following its release, Hg human health and wildlife in the context of human
is transported and recycled between the major environ- welfare?
mental compartments, i.e., air, soil and water, until it is 3. How can technological development contribute to
eventually removed from the system through burial in the reduction in mercury exposure and improvement
coastal and deep ocean sediments, lake sediments, and of environmental responsibility?
subsurface soils [2, 4]. 4. How can scientific knowledge contribute to the
Hg is considered to be a peculiar chemical element implementation and effectiveness evaluation of the
because it displays particularly strong chemical and Minamata Convention and other regulatory agree-
biological activity as well as variability in form (liq- ments, and what is the importance of integrating and
uid and gaseous). Hg compounds with very different implementing emerging and future mercury research
chemical and physical properties are included in vari- into the policy making?
ous cycles of its natural circulation [5, 6]. Hg is a glob-
ally distributed pollutant due to characteristics such as
low melting and boiling points, conversions between
chemical forms and participation in biological cycles. Methods
As a result of anthropogenic emissions, the global While developing this paper, the method recommended
atmospheric Hg deposition rate is approximately three by Liberati et al. [14] was used to some extent. However,
times higher than that in preindustrial times and has the strict use of this method was not possible due to the
increased by a factor of 2–10 in and around the most specifics of our publication. The problems of pollution
industrialized regions [7]. and Hg transformation in the environment caused by sig-
Hg-contaminated land environments pose a risk to nificant threats to human health represent the subject of
global public health, with Hg being listed as one of the research by numerous research teams around the world.
‘ten leading chemicals of concern’ [8]. In 2013, the United However, many publications do not meet the criterion of
Nations (UN) introduced the ‘Minamata Convention quality of research results. Therefore, the following crite-
on Mercury’ [9], which aims for a more global effort for ria were used to eliminate publications:
managing the risk presented by Hg to human health and
the environment. The Minamata Convention has as of i. Accurate descriptions of the research methods
today 123 parties and the convention entered into force were lacking;
16 August 2017. This concerted action, if successful, will ii. Methods that did not guarantee the quality of the
have great implications for public health for decades to results were used;
come; however, there are many hurdles on the way to iii. The latest publications were selected, but historical
achieving this goal [10, 11]. works were not omitted;
Our understanding of the critical processes driv- iv. Publications from various countries and regions
ing global Hg cycling, particularly those that affect the were cited.
large-scale exchange of Hg among major environmental
compartments, has advanced substantially over the past
decade. This progress has been driven by major advances Hg emission
in three interconnected areas: new data, new models, and The advances achieved over the last decade for the
new analytical tools and techniques [12]. assessment of Hg emissions from major man-made and
The main task for improving our knowledge of mercury natural sources have contributed to improvements in the
sources, fates, impacts, and emission control options assessments of the impacts of atmospheric deposition of
was defined at the 14th ICMGP Conference in 2019 [13] Hg on the terrestrial environment [15]. The assessment
as “Bridging knowledge on global mercury with envi- of Hg emissions poses serious methodological problems.
ronmental responsibility, human welfare and policy In estimating these impacts, state institutions mainly
response”. The main four questions that require urgent focus on inventories of their sources, while international
answers within this subject are defined as follows: organizations apply different models and use emission
factors and statistical data of industrial production and
1. How is mercury biogeochemical cycling changing on consumption of Hg-containing materials. It is particu-
global, regional, and local scales in response to per- larly difficult to distinguish natural and anthropogenic
emissions from re-emissions from the land and oceans
Gworek et al. Environ Sci Eur (2020) 32:128 Page 3 of 19

[16, 17]. Hg emission sources include both natural pro- The updated global mercury budget in 2018 [2] is shown
cesses unfolding in the biosphere and anthropogenic in Table 1. The annual deposits were 480 tons higher
sources. In 2008, the following classification was adopted than the emissions, indicating considerable enrichment
in a UNEP report [18], which distinguished three emis- in the environment. In general [23, 24], ­Hg0 emissions
sion sources: from undisturbed nongeologically thermally enriched
areas are < 1 ng m−2 ­h−1, whereas fluxes from contami-
• Current emissions from natural sources; nated sites can be several orders of magnitude higher
• Current emissions from anthropogenic sources; at > 5000 ng m−2 ­h−1. For contaminated sites that cover
• Re-emissions from historical deposits from natural large spatial areas (such as mining operations), the annual
and anthropogenic sources. emissions from the entire surface area have been shown
to range from 19 to 105 kg year−1 from active industrial
Environmental archives offer an opportunity to recon- gold mines and 51 kg year−1 from a large abandoned Hg
struct temporal trends in atmospheric Hg deposition at mining area.
various timescales. Lake sediment, peat, ice cores, tree Various models are applied to estimate Hg emission
rings, and Hg stable isotope measurements are offering levels into the atmosphere. However, the emission levels
new insights into historical Hg cycling. Preindustrial Hg determined by using these models differ substantially.
deposition has been studied over decadal to millennial Travnikov et al. [25] compared the global emission levels
timescales extending as far back as the Late Pleistocene. from natural and anthropogenic sources determined by
Exploitation of mercury deposits (mainly cinnabar) first using four models [26–29]. The models differed signifi-
began during the Mid- to Late-Holocene in South Amer- cantly in their estimations of global total emissions, with
ica, Europe, and Asia, but increased dramatically during values ranging from 4000 Mg year−1 to 9230 Mg year−1,
the Colonial era (1532–1900) for silver production [19, of which natural emissions and re-emission ranged from
20]. Artisanal gold mining is now thought responsible 45 to 66% of the total emissions. A later summary of the
for over half the global stream flux of Hg, followed by the four models [30] estimated that global anthropogenic
burning of coal [2, 21]. emissions were 1870 Mg year−1, although the global
Fluxes of Hg to the air occur via the volatilization of natural emissions and re-emissions were already sig-
­Hg0 as well as through wind entrainment of Hg bound nificantly different from each other at 3995 Mg year−1
to dust particles (often referred to as fugitive dust emis- to 8600 Mg year−1 [31]. The proportions between the
sions). The relative magnitude of these two types of Hg amounts of natural emissions and anthropogenic emis-
emission sources varies depending on site-specific condi- sions are not precisely determined. The ratios vary by
tions. However, due to the long atmospheric lifetime of authors and encompass a relatively wide range from 0.8
­Hg0, these emissions mostly contribute to the global pool to 1.8 [32–34]. The relatively broad emission ranges pre-
of Hg, whereas the fugitive dust emissions impacts are sented in various studies are caused by the following fac-
more local [22]. tors: volatility of Hg compounds, distribution of sources,
The updated global mercury budget shows the impact low levels of Hg concentrations in the air, concentrations
of human activities on the mercury cycle and the result-
ing increase in mercury accumulated in soil and oceans.
Table 2 Annual emissions of mercury by selected world
region [11] (Mg/year)
Table 1 The updated global mercury budget [2] (Mg/year) 2000 2015 2000–2015 2010–2015
growth (%/year) growth (%/year)
Emission:
i) Natural (geogenic) 500; USA 127.7 42.7 − 66.6 − 10.2
ii) Re-emission/re-mobilization (natural and legacy): OECD Europe 106.2 38.0 − 64.2 − 5.8
    Biomass burning 600 Canada 12.7 8.3 − 34.6 − 3.2
    From soil and vegetation 1000 Eastern Europe 49.4 33.6 − 32.0 − 1.3
iii) Anthropogenic 2500 (2000–3000) Central America 33.5 34.3 2.4 5.4
iv) Ocean, net evasion (gaseous elemental Hg) 3400 (2900–4000) South Asia 120.8 191.6 58.6 4.6
Total 7000 Eastern Africa 19.1 72.9 281.7 4.0
Deposition: East Asia 532.7 1012.3 90.0 2.6
i) To land/freshwater 3600 South America 239.0 275.8 15.4 0.7
ii) To ocean 3880 Southeast Asia 224.6 187.5 − 16.5 2.2
Total 7480 Global total 1964.1 2389.8 21.7 1.8
Gworek et al. Environ Sci Eur (2020) 32:128 Page 4 of 19

that are much lower than levels of other basic pollutants to primary natural sources, whereas the re-emission of
and difficulty of determination [35, 36]. previously deposited Hg on vegetation, land or water
Annual mercury emissions in selected regions of the surfaces is primarily related to land use changes, bio-
world in 2000–2015 are shown in Table 2. Despite the mass burning, meteorological conditions and gaseous Hg
many applicable global, regional and national programs exchange mechanisms at the air–water/topsoil/snow–ice
and conventions aimed at reducing Hg emissions, the pack interfaces [15, 38].
global THg emissions in the years 2000–2015 increased A characteristic feature of natural Hg emissions com-
by 1.8% [11]. Emission reductions only took place in pared to anthropogenic sources is their distributed
North America and Europe without the former USSR, nature and wide range of distribution. Hg from natural
and the largest increase was in Central America (5.4%) sources is introduced on a global scale into the atmos-
and South Asia (4.0%). phere, and Hg from anthropogenic sources is deposited
Emission values [11] for the aggregated source catego- mainly locally and regionally; therefore, determining
ries and their composition are shown in Table 3. Three emission levels and applying effective control methods
categories had the highest absolute value in 2015: gold are difficult. Natural Hg emission processes also include
ASMG (775.1 Mg/year), coal combustion (558.3 Mg/ re-emission of Hg previously deposited from the atmos-
year) and cement production (206.3 Mg/year). From phere in the process of wet and dry deposition from
2010 to 2015, the largest increase in emissions was dem- both natural and anthropogenic sources, which increases
onstrated by cement production (6.3%), gold production the difficulty of estimating Hg emissions from natural
at a large scale (5.2%) and industrial metal extraction sources [34, 35, 39].
(4.6%). The reductions [37] were in the following fields: Annual global Hg emissions from natural sources on
dental (− 5.6%) and electrical equipment (− 5.2%). This land are estimated by various authors and cover a wide
finding is consistent with the continued expansion of value range, e.g., total emissions of 1600–2500 Mg year−1,
the global economy, but is less than the growth in world including re-emissions of 790–2000 Mg year−1. Accord-
GDP over the same period (5.7%/year), suggesting con- ing to most authors, re-emissions were higher than pri-
tinuous improvement in pollutant emissions per unit of mary emissions [40–42].
production.
Volcanoes
Mercury emissions from natural sources
Volcanoes and geothermal activities are important
The estimate of Hg emissions from natural sources sources of Hg pollution in terrestrial environments. Hg
includes contributions from primary natural sources and is emitted from volcanoes primarily as gaseous H ­ g0,
the re-emission processes of historically deposited Hg and the Hg/SO2 ratio is generally adopted to estimate
over land and sea surfaces. Hg emitted from volcanoes, Hg emissions. The annual average of Hg released to the
geothermal sources and topsoil enriched in Hg pertains atmosphere without episodic strong eruptions for vol-
canoes and geothermal activities is ~ 75–112 Mg year−1
of Hg, accounting for approximately < 2% of the con-
Table 3 Annual emissions of mercury by selected source tribution from natural sources [15, 43–45]. In volcanic
category [11] (Mg/year) plumes, Hg is present both in the gas phase as elemen-
Source 2000 2015 2000–2015 2010–2015 tal ­Hg0 and reactive ­HgII and in the particle phase as ­Hgp
category growth (%/ growth (%/year) forms. The proportions of these species are highly vari-
year)
able. ­HgII and H
­ gp typically amount to < 5% of THg, with
Gold ASMG 583.7 775.1 32.8 1.3 Hg° as the most abundant form [46–48]. Hg levels in vol-
Coal combustion 359.9 558.3 55.1 0.7 canic ash nanoparticles (36 ± 4 mg kg−1) are dramatically
Cement produc‑ 74.3 206.3 177.7 6.3 higher than their bulk concentrations (0.08 mg kg−1) [49].
tion Many areas of geothermal activity have long been associ-
Waste burning 201.9 165.6 − 18.0 2.1 ated with elevated levels of Hg in the soil and air in places
Municipal waste 202.6 140.6 − 30.6 2.5 such as Hawaii, Iceland, western parts of the United
Gold, large-scale 82.7 112.1 35.6 5.2 States and New Zealand [43, 50].
Zinc smelting 81.7 103.6 26.8 1.4 The average annual global Hg emissions estimate from
Copper smelting 53.7 70.0 30.4 3.2 biomass burning (emissions from wildfires: forests,
Dental 49.4 16.1 − 67.4 − 5.6 savannas and grasslands) for 1997–2006 was 675 (± 240)
Electrical equip‑ 118.7 62.0 − 94.8 − 5.2 Mg year −1, which accounts for 8% of all current anthro-
ment
pogenic and natural emissions. The largest Hg emissions
Global total 1964.1 2389.8 21.7 1.8
are from tropical and boreal Asia, followed by Africa
Gworek et al. Environ Sci Eur (2020) 32:128 Page 5 of 19

and South America [51, 52]. The important factor for conservative estimation of atmospheric mercury dry
Hg concentrations in forest soils is the time since stand- deposition (the portion that is retained in leaves). Mer-
replacing fires have occurred, and high soil burn sever- cury in throughfall also includes a portion of previously
ity has the potential to reduce the concentrations of Hg dry-deposited mercury (the portion that is washed off
in burned soils for tens to hundreds of years [53, 54]. In from the canopy). Concurrent measurements of litterfall,
a specific emission source in Nisyros Island (Greece), throughfall, and open-space wet deposition measure-
Hg concentrations in fumarolic gases in Nisyros Island ments can be used to estimate dry deposition on seasonal
(Greece) ranged from 10,500 to 46,300 ng/m3, while Hg or longer time scales, whereby dry deposition is approxi-
concentrations in the air ranged from high background mated as litterfall plus throughfall minus open-space wet
values in the Lakki Plain caldera (10–36 ng/m3) up to deposition [64, 65].
7100 ng/m3 in the fumarolic areas [55]. On the land surface, Hg deposition is mainly in the oxi-
dized form (­ Hg2+), and its transformations are associated
Exchange of mercury between atmospheric and terrestrial primarily with the oxidation–reduction potential of the
ecosystems environment and with the biological and chemical pro-
Air-, soil- or vegetation-covered exchange fluxes are an cesses of methylation. For soils in which oxidizing condi-
important part of global and regional biogeochemical tions predominate, the H ­ g22+ forms dominate,
­ g2+ and H
cycles [56]. Much of the ­HgII deposited in precipitation or and in soils with reducing conditions, Hg and sulfur com-
taken up by plants is reduced to ­Hg0 and may be released pounds are mainly present. Methyl-Hg compounds are
back to the atmosphere. Recent vegetation and soil Hg most commonly found in soils with transient conditions
studies suggested that vegetation H ­ g0 uptake dominates [5].
(50–80%) Hg net deposition at terrestrial sites [57–59]. For GEM (­ Hg0), the residence time is estimated at 6 to
The different forms of atmospheric Hg may be depos- 18 months, while GOM (gaseous mercury in oxidized
ited on surfaces by way of wet and dry processes. These form) and TPM (total particulate mercury) are quickly
forms may be sequestered within terrestrial compart- removed from the air through wet and dry deposition,
ments or emitted back to the atmosphere, with the rela- and their residence times are estimated to be hours or
tive importance of these processes being dependent on days at most [66, 67]. Given the long time for removal
the form of Hg, surface chemistry, and environmental from the air, GEM can be transported over large dis-
conditions. Many models assume that the net GEM (gas- tances [16]. The particular Hg species are characterized
eous elemental mercury) exchange with soil surfaces is by different dry deposition rates, which also determine
zero; however, as discussed below, some components their residence times. The dry residence times of the dif-
are assimilated into foliage over the growing season and ferent Hg species form the following series [68]:
accumulate in soils [59]. Smith-Downey et al. [60] esti- GEM 0.19 cm s−1 < TPM 2:1 cm s−1 < GOM 7:6 cm s−1.
mated that evasion of Hg linked to the decomposition According to Marsik et al. [69], the dry deposition rates
of soil organic carbon pools and subsequent liberation of GOM and GEM are much higher during daytime than
of ­HgII sorbed to soil organic matter is over 700 t/y, thus nighttime. As with Lindberg et al. [68], these authors
reflecting the large pool of Hg stored in terrestrial ecosys- explain this fact by the closure of plant stomata at night.
tems globally (over 240 kgt). In total, this study estimated The deposition rates also depend on the type of surface
that 56% of Hg deposited to terrestrial ecosystems is ree- [70]. Caffrey et al. [71] determined that the deposition
mitted. Similarly, Graydon et al. [61] found that 45–70% rates of particulate air pollutants on the ground with low
of isotopically labeled ­HgII wet-deposited to a forested vegetation were 3–5 times lower than those in forests.
watershed had been reemitted to the atmosphere after a Deposition is also affected by the weather conditions,
year. air humidity, insolation and atmospheric precipitation.
Litter deposition is the predominant source of Hg in Research by Converse et al. [72] in an uncontaminated
soil. Forest litter horizons show significant increases in high-elevation wetland meadow in Shenandoah National
mass-dependent fractionation (MDF) during decomposi- Park, Virginia (USA) showed the highest Hg deposition
tion concurrent with augmented total Hg mass, and this occurred in spring (4.8 ng m−2 ­h−1), with a decrease
relationship is most significant at high-elevation sites [62, occurring in summer (2.5 ng m−2 ­h−1) to near zero flux
63]. Measurements of mercury in litterfall and through- in fall (0.3 ng m−2 ­h−1), followed by an increase in win-
fall have been increasingly used to provide knowledge ter emissions (4.1 ng m−2 ­h−1). These studies also suggest
of mercury deposition over forest canopies. The major- that stomatal processes are not the dominant mechanism
ity of mercury in litterfall is considered to be from the for ecosystem-level GEM exchange. Table 4 shows a sum-
stomatal uptake of H ­ g0 and can be used as a rough and mary of biome-level Hg depositions and soil Hg turnover
Gworek et al. Environ Sci Eur (2020) 32:128 Page 6 of 19

Table 4 Summary of biome-level Hg deposition and soil Table 5 Summary of mercury fluxes from terrestrial
Hg turnover times [60] regions [38]
Biome Mean Hg deposition Mean soil Hg turnover Region Evasion Ratioa
(g m−2) ­timea (years) (average) (%)
(Mg year−1)
Preindustrial Present Preindustrial Present day
day Forest 342 7,5
Tropical 0.9 3.7 234 126 Tundra/grassland/savannah/prairie/chaparral 448 9,9
forest Desert/metalliferous/non-vegetated zones 546 12,0
Temperate 0.8 2.9 250 151 Agricultural areas 128 2,8
forest Evasion after mercury depletion ­eventsb 200 4,4
Boreal 0.5 1.5 998 560 Total 1664 –
forest
Volcanoes and geothermal areas 90 –
Grassland 1.0 3.5 522 269
Biomass burning 675 –
Tundra 0.3 0.8 1108 702
a
Calculated over the total evasion from natural sources which sum
Desert 0.5 1.4 2387 1748
4532 Mg year−1
a
With respect to respiration b
Friedli [42] and Mason [68] distinguish as a natural source of mercury emission
to the atmospheric air of regions where there are episodes of sudden decreases
in mercury concentrations in the air by deposition to the ground and then
reemission. These phenomena occur mainly in the Arctic regions and Antarctica,
times [60]. The concentration of Hg in soils is therefore a and the emission from this source is estimated at 200 Mg year−1
function of the deposition rate and carbon turnover time.
High soil concentrations in desert ecosystems are driven
by a combination of higher deposition and extremely i. The largest amounts of Hg are emitted from tropi-
slow Hg turnover. Tropical and temperate lifetimes are cal regions (45%), followed by the temperate zones
similar despite the faster carbon turnover in tropical sys- (41%), with the lowest emissions from the polar
tems due to the relative balance between Hg provided by regions (8%), and emissions from volcanoes and
wet deposition and leaf uptake. geothermal areas account for 5%;
Emissions from soils have the form of GEM and depend ii. Areas with vegetation can be ranked according to
on many factors [73–77]: the size of their emissions as follows: forests > other
areas (tundra, savannas, and chaparral) > agricul-
• The properties of soils, e.g., Hg content, the contents tural areas > grassland ecosystems;
of organic compounds, and saturation; iii. Land areas devoid of vegetation emit more Hg than
• The concentrations of oxidants, mainly ozone, in the do areas with plants.
air;
• The weather conditions, e.g., solar radiation, temper- Deforestation can increase GEM emissions due to
ature, humidity and winds. higher solar radiation and increased temperature at the
soil surface [83, 84].
Soil Hg fluxes are significantly lower in dark conditions The overall background soil Hg flux in the United
than light conditions for all sites except grassland [64]. States is estimated to be 0.9 ± 0.2 ng/m2/h [78], and in
It is most difficult to estimate Hg emissions from areas with significant Hg pollution, soil emissions are
plants, and these emissions mainly occur in the form of much larger. In the canton of Valais, Switzerland, elemen-
­Hg0 [78–80]. Ericksen et al. [81] suggested the following tal Hg ­(Hg0) is undetectable in soil, although substantial
hierarchy of environmental parameters that influence Hg ­Hg0 emissions were found to occur (20–1392 ng m−2 ­h−1)
flux: [85].
Soil moisture > light > air concentration > relative Urban areas are of particular concern with respect to
humidity > temperature. the global Hg cycle due to the following [86]:
Table 5 shows a summary of total Hg (THg) fluxes from
terrestrial regions [38], and Table 6 shows the average i. Frequently high terrestrial Hg concentrations and
fluxes (or, in some cases, the range of fluxes) for various the physically and chemically diverse nature of
ecosystems measured by a number of investigators. urban surface covers;
When analyzing the data in Table 6, the following gen- ii. Highly variable time series concentrations of ambi-
eralizations can be made [38, 82]: ent atmospheric Hg as a result of regional and local
emissions;
iii. Urban meteorology (i.e., heat island effect).
Gworek et al. Environ Sci Eur (2020) 32:128 Page 7 of 19

Table 6 Average fluxes, or in some cases the range of fluxes, for various ecosystems measured by a number
of investigators
Species/ecosystem Fluxa References
nmol m−2 ­month−1

Ground-level forest floor Sweden 1.4–1.7 Hanson et al. [82]

Lindberg et al. [68]
Model estimates hard wood forest Max 4 Bash et al. [196]
Model estimates forest soil 4 Bash et al. [196]
Agricultural crops Max 11 Bash et al. [196]
Temperate forest 5.0 Rea et al. [197]
Deforested site 50 Magarelli and Fostier [198]
Desert soils 3.6–9.8 Magarelli and Fostier [198]
High Hg regions Max 1500 Gustin and Lindberg [199]
Maple 20 Hanson et al. [82]
Oak 16.4 Hanson et al. [82]
Spruce 6.1 Hanson et al. [82]
Prairie grass 12.5 Obrist et al. [53]
Typha sp. 60 Gustin et al. [200]
Average global soil 1.5 Selin et al. [201]
a
Values have all been converted to a common flux unit of nmol m−2 ­month−1. Results from the older literature are combined in estimates given in various review
papers

In the city of Tuscaloosa, Alabama (USA), Hg fluxes on exchange fluxes. High light transmittance can enhance
bare undisturbed soil surfaces were as follows (median) soil TGM emission rates and increase the magnitude of
[86]: diurnal variations in soil–air TGM exchange fluxes. The
estimated annual average soil–air TGM exchange flux
• Residential site—4.45 ng.m−2 ­h−1; was 5.46 ± 21.69 ng m−2 ­h−1 in corn–wheat rotation
• Industrial site—1.40 ng.m−2 ­h−1; cropland with 30 cm row spacing [89, 90].
• Commercial site—2.14 ng.m−2 ­h−1; The bidirectional exchange of Hg between the atmos-
• Mixed land use site—0.87 ng.m−2 ­h−1. phere and terrestrial surfaces is better understood
because of advancements in research that are primarily
Areas of land devoid of vegetation emit more Hg than associated with the interpretation from Hg isotopes, and
those with plants. The annual averaged fluxes in the the latest estimates place land surface Hg re-emission at
subtropical forest zones in China from soil in the for- values lower than previously thought [91].
ests were 14.2 ng m−2h−1, and for open-air sites, they
were 20.7 ng m−2 ­h−1 [87]. Soil Hg fluxes were sig- Methylmercury
nificantly lower in dark conditions than in light con- High doses of organic compounds of Hg, particularly
ditions. In grassland sites, the mean soil Hg flux was methyl-Hg, can be fatal to humans and wildlife, and even
0.6 ± 0.9 ng m−2 ­h−1 in darkness, 1.0 ± 0.7 ng m−2 ­h−1 in relatively low doses can seriously affect the nervous sys-
light, and 0.9 ± 0.7 ng m−2 ­h−1 overall [64]. tem of organisms. Hg has also been linked to harmful
Cropland is an important component of terrestrial eco- effects on the cardiovascular, immune and reproductive
systems. It is estimated that 33% of natural-source atmos- systems. Methyl-Hg passes through both the placenta
pheric Hg comes from the emissions at cropland surfaces and blood–brain barrier; therefore, the exposure of
[88]. The emission of Hg from cropland soil greatly affects women of child-bearing age and of children to methyl-
the global Hg cycle. Combinations of different crop cul- mercury is of great concern [1].
tivars and planting densities will result in different light Methyl-Hg can be both biotically and abiotically pro-
transmittance under canopies, which directly affects the duced in the environment. Methylation of Hg tends to
solar and heat radiation flux received by the soil surface occur in environments with low oxygen levels, low pH,
below crops. In turn, this might lead to differences in the Hg bioavailability, temperature, redox potential and high
soil–air total gaseous mercury (TGM) exchange under levels of dissolved organic compounds and in environ-
different cropping patterns. The light transmittance ments favored by sulfate-reducing bacteria, which are
under the canopy was the key control on soil–air TGM largely responsible for methylation. These conditions
Gworek et al. Environ Sci Eur (2020) 32:128 Page 8 of 19

are found primarily in deep sea environments, coastal In the Wuchuan Hg mining areas (Guizhou, China),
marine sediments, and some freshwater lakes as well as soil samples present THg values ranging from 0.33 to
soils. These conditions are also characteristic for paddy 320 mg kg−1 and methyl-Hg values ranging from 0.69
soil [92, 93]. Organic Hg is much more toxic to living to 20 ng g−1. The rice grain samples contain elevated
organisms than inorganic Hg [4, 94, 95]. The content of methyl-Hg concentrations ranging from 4.2 to 18 ng g−1,
methyl-Hg in soils and plants is significantly lower than while corn grain contained only 0.5–2.0 ng g−1 [92].
that of THg; however, due to its much higher toxicity, Research carried out in areas with coal-fired power
methyl-Hg is particularly dangerous for living organisms. plants in Hunan (China) [109] shows that in the soil
In soils in the coniferous boreal forests of Sweden, the samples, THg varied from 0.068 to 0.220 mg kg−1 (mean
background level of pollution from methyl-Hg accounted value of 0.130 ± 0040 mg kg−1), and methyl-Hg ranged
for 0.35–0.59% of THg [96]. In the Idrija Hg mining from 0.30 to 3.5 μg kg−1 (mean 1.6 ± 1.0 μg kg−1). In rice
area of Slovenia, a heavily polluted region, methyl-Hg samples, the ­Hg(II) concentrations varied from 0.002 to
accounted for 0.003% of THg, and its background pres- 0.022 mg kg−1 (mean 0.057 mg kg−1), and methyl-Hg
ence in controls was 0.17% [97]. The content of methyl- concentrations varied from 1.7 to 3.8 ng gg−1 (mean of
Hg relative to THG was 1.9% in the roots of rice under 2.4 ± 0.72 ng ­g−1). Meng et al. [98] showed that rice had
background conditions and 0.55% in the leaves; in areas high affinity for methyl-Hg and that the concentrations
of Hg mining sites, the concentrations were 0.07% in the in rice seeds may be 2 to 3 orders of magnitude higher
roots and 0.01% in the leaves; and in areas with artisanal in Hg mining sites than in other local edible crop plants.
Hg mining sites, the concentrations were 0.63% in the Freshly deposited Hg is more likely to methylate and be
roots and 0.02% in the leaves [98]. Methyl-Hg in heav- incorporated in rice than stored Hg [91].
ily contaminated soil in the Rhône Valley (Switzerland) Forest fires cause a significant reduction in mercury
accounted for < 0.8% of THg [99]. content in soil. Burned soils in northwestern Ontario
A number of factors that control microbial activity (Canada) had 82% less methyl-Hg than fresh soils [110].
and/or the geochemical speciation of inorganic ­ Hg2+ Current climate change has had a significant impact
govern MeHg formation in the environment [100]. on Hg transformation processes, especially in the Arc-
Microorganisms that live in soil can transform inorganic tic. The very large mass of mainly natural Hg found in
­Hg(II) species into H
­ g0 by using the enzyme Hg reductase, northern permafrost deposits, which is projected to be
which is found in various bacteria, such as Pseudomonas released with further climate warming, may profoundly
sp., Staphylococcus aureus, Thiobacillus and many others affect biotic Hg levels around the Northern Hemisphere,
[101]. Increases in temperature might lead to increases especially because large amounts of organic carbon,
in biological activity as well as higher H­ g2+ methylation which may stimulate Hg methylation rates, will be simul-
rates [102]. taneously released [111].
The direct conversion of insoluble HgS species to
MeHg in anaerobic soils is generally believed to be low, Mercury in soil
although this condition can change when environmen- Mercury has a relatively long half-life in surface soils
tal conditions favor HgS complexation [103]. The redox because of its recycling between the surface environment
potential also seems to be a key factor, as suboxic and and atmosphere. Permanent removal of anthropogenic
mildly reducing conditions seem to promote high H ­ g2+ Hg from the biologically active part of the environment
methylation rates, while anoxic and strongly reducing will only occur once it is buried in mineral soils [4]. Soil
conditions might lead to elevated sulfide concentrations plays an important role in biogeochemical Hg circulation
that eventually prevent ­Hg2+ from being available for because it accumulates this element and is a source for
methylation of some methylating bacteria, including SRB other environmental components. Hg occurs naturally in
(sulfate-reducing bacteria, e.g., Desulfobacter sp.), and soils from geologic sources [12] or as the result of natu-
some that control the availability of ­Hg2+ for methylation ral events such as forest fires and volcanic eruptions [49].
(e.g., Deltaproteobacteria or Clostridia) [104, 105]. The total amount worldwide of Hg accumulated in the
S plays a major role in influencing ­Hg2+ methylation by soils of terrestrial environments is estimated at 200–300
directly affecting the activity of some methylating bacte- Gg [112–114]. Smith-Downey et al. [60] suggested that
ria, such as SRB, and controlling the availability of H­ g2+ organically bound Hg in preindustrial soils is 200 Gg and
for methylation [106, 107]. that a 20% increase in organically bound soil Hg (to 240
The paddy soils in Hg mining areas have a high meth- Gg) has occurred from preindustrial steady-state condi-
ylation ability and may eventually result in heavily bio- tions to the present day.
logical effects on the local residents through the food In the 2013 Technical Background Report for the
chains, such as rice containing high methyl-Hg [108]. Global Mercury Assessment [4], based on a global model
Gworek et al. Environ Sci Eur (2020) 32:128 Page 9 of 19

and budget developed by Mason et al. [114], human Table 7 presents examples of the concentrations of
activities were estimated to cumulatively increase atmos- Hg in the soils in the vicinity of industrial emission
pheric Hg concentrations by 300–500% over the past sources according to different authors, The highest Hg
century. Because of the naturally high Hg amount present contents were found in soils near Hg mines: Almaden
in soil, the average Hg increase was only 20% in surface in Spain, with 2000 years of mining and ore processing
organic soil and negligible in mineral soils. The revola- (< 8889 mg kg−1) [126]; Idrija in Slovenia, with 500 years
tilization of “legacy Hg” (i.e., Hg from historical sources of mining activity (< 2759 mg kg−1) [97, 127]; and in
of pollution) from soil and ocean and its long residence Alaska (5326 mg kg−1) [125]. Chlor-alkali plants are also
time in those compartments contribute to maintaining an important source of environmental Hg pollution.
atmospheric Hg concentrations and deposition rates at Bernaus et al. [128] estimated that Hg levels in the soil
higher levels than those supported by current primary around a chlor-alkali plant in the Netherlands were as
emissions [115]. Recent estimates of the anthropogenic high as 1150 mg kg−1.
and natural Hg contents in global soils (organic layers) According to Richardson and Moore [129], in the
(data in kilotons) included 182 natural and 89 anthropo- urban environment, the diversity of Hg content in soils is
genic sources based on Amos et al. [116] and 130 natural relatively high because of the diversity of land functions
and 20 anthropogenic sources according to the AMAP/ in towns. Urban soils were found to accumulate higher
UN Environment [4]. concentrations and pools of Hg than their rural montane
All results for Hg soil content, which are presented in counterparts across New York and southern New Eng-
the next part of this publication, pertain to the topsoil land, which highlights the importance of soils in urban
layer. Generally, the average background concentration systems for sequestering Hg and preventing its move-
of Hg in soil ranges from 0.03 to 0.1 mg kg−1, with an ment towards riparian and aquatic ecosystems, where
average value of 0.06 mg kg−1, whereas Hg-contaminated it can bioaccumulate. Moreover, soil Hg concentrations
sites often have soil concentrations that are 2- to 4-orders were poorly correlated with pH, loss-on-ignition, and
of magnitude higher [117, 118]. Kabata Pendias and Pen- clay content. Instead, proximity to local industrial and
dias [5] defined a narrower range of 0.05–0.3 mg kg−1, agricultural sources proved a significant influence on Hg
although some volcanic and organic soils, especially in accumulation.
Canada, may contain higher values, and in the vicinity of The lowest median results were determined for soils
industrial emission sources, the values can be extremely in Changchun, China, at 0.018 mg kg−1, with a range
high. Obrist et al. [119] showed that a dataset with more of 0.012–0.036 mg kg−1 [130] and in Oslo, Norway
than 1900 randomly selected sampling points across the (0.06 mg kg−1, with a range of 0.01–2.3 mg kg−1) [131],
western USA indicated median Hg concentrations of while the highest concentrations were in Palermo, Italy
0.019 mg kg−1 and an average value of 24 mg kg−1, with (median value of 1.85 mg kg−1, with a range of 0.004–
only 1% of soil samples exceeding background values 2.61 mg kg−1) [132], and Glasgow, Scotland (1.2 mg kg−1,
(e.g., > 0.10 mg kg−1). with a range of 0.312–5.2 mg kg−1) [133]. Of note is the
The LUCAS Topsoil Survey of the European Union reduction of 270% (median from 0.68 mg kg−1 to 0.37)
organization collected over 23,000 topsoil samples in Hg pollution from 1987 to 2009 for soils from Beijing,
(upper 20 cm) from land in all European Union countries China [134].
(28) except for Croatia [120]. The average for European In agricultural soils, pollution by Hg was relatively
topsoil Hg concentrations was 0.04 mg kg−1, with a range low, as indicated by the low median and average values
of 0–159 mg kg−1. Studies have identified highly pol- at usually below 0.1 mg kg−1 (e.g., Scandinavia—Ottesen
luted, isolated sites, and the larger historical and recent et al. [135]; Poland—Loska et al. [136]; Iran—Ahmadi
industrial and Hg mining areas show elevated concentra- et al. [137]). However, the ranges of the results were rela-
tions of Hg. Historically, mining for gold and Hg led to tively wide, and the maximum values often exceeded
high Hg concentrations in these mining areas, which may 1 mg kg−1, which may indicate a threat to food produc-
explain the high Hg concentrations in some samples from tion due to the need to protect human health in some
Central Italy, Northwest England and Eastern Slovakia. areas with higher Hg soil levels. In Europe, pastures were
Moreover, the natural/background Hg level was 0.08 mg/ slightly more polluted with Hg than plowed fields [136].
kg in Brazil [121], 0.05 mg/kg in India [122], 0.23 mg/kg Soils in forest environments contained low levels of Hg.
in New Zealand [123], 0.11 mg/kg in the Norwegian Arc- Average and median values did not exceed 1 mg kg−1,
tic [124] and 0.4 mg/kg in Paris [125]. Most soil Hg was although compared with agricultural soils, the maxi-
found as soil matrix-bound divalent Hg (­HgII), whereas mum values were also lower than 1 mg kg−1. Mineral
elemental Hg (­ Hg0) was undetectable in soils [85, 99]. forest soils contained less Hg than organic ones (USA—
Woodruff and Cannon, [54] Czech Republic—Navrátil
Gworek et al. Environ Sci Eur (2020) 32:128 Page 10 of 19

Table 7 Some examples of surface soil layer contamination with mercury in regions of important sources of emissions
by various authors
Country Location Source of pollution Period Total Hg Mean/median Author
mg kg−1

USA Alaska Hg-mining 0.05–5326 Bailey et al. [125]

USA Texas Hg-mining 3.8–11 Gray et al. [202]
China Wuchuan Hg-mining 2003 0.33–320 Qiu et al. [92]
China Wanshan Hg-mining 2002 5.1–790 Qiu et al. [108]
China Xiaoqinling Gold mining 0.04–61.2 Mean 2.75 Wu et al. [203]
Slovenia Idrija Hg-mining 1991–97 1734–2759 Mean 2456 Gnamuš et al. [97]
Hg-mining 2000–01 24–1055 Median 47 Gosar et al. [127]
Slovenia Podljubelj Hg-mining 2003–04 0.35–244 Median 3.7 Teršič et al. [204]
Slovakia Rudnany Hg-mining 9.1–54.3 Banásová [205]
Spain Almaden Hg-mining 6–8889 Mean 604 Higueras et al. [126]
Spain Almaden Hg-mining 1340–4830 Dago et al. [157]
Spain Caunedo Old Hg-mining 0.09–50.0 Mean 13.1 Boente et al. [206]
Italy Vallalta Old Hg-mining 6–21 Wahsha et al. [207]
Turkey Halıköy Hg mining 2004 0.10–33 Mean 5.5 Gemici et al. [208]
China Tongguan Artisanal gold mining 0.69–23.7 Mean 2.91 Xiao et al. [209]
China Wanshan Hg-mining, artisanal gold 2012 0.5–187 Mean 31.0 Yin et al. [210]
China Guizhou Acetic acid 2016 1.09–3.71 Li et al. [93]
Italy Mt. Amiata, Hg mining, volcano-geothermal 2.4–68 Chiarantini et al. [211]
France Vosges Mountains Chlor-alkali 2002 0.16–3.99 Hissler and Probst [212]
France Grenoble Chlor-alkali 1.3–10 Grangeon et al. [213]
Kazakhstan Pavlodar Chlor-alkali 0.93–22.3 Ullrich et al. [214]
Spain Flix Chlor-alkali 0.04–12.9 Mean 0.77 Esbrí et al. [215]
Germany Chlor-alkali 0.5–4.2 Mean 1.6 Biester et al. [216]
Netherlands Chlor-alkali 2004 4.3–1150 Bernaus et al. [128]
Portugal Estarreja Chlor-alkali 0.010–91 Mean 5.4 Reis et al. [217]
China Huludao Chlor-alkali Zn-smelting 2006–08 0.05–14.6 Zheng et al. [218]
China An Ning Chlor-alkali 0.09–1.30 Mean 0.40 Song et al. [219]
China Kunming Chlor-alkali polyvinyl chloride 0.15–4.79 Zhu et al. [220]
Portugal Caveira Sulfide mine 1.1–76.5 Reis et al. [221]
China Lianyuan Coal main, steel industry 2015 1.20–3601 Mean 178 Liang et al. [222]
China Zhuzhou Zn/Pb smelter 2012 0.62–2.61 Mean 1.54 Wu et al. [223]
China Chongqing Thermometer factory 0.06–0.88 Wang et al. [224]
Czech Rep. Bohemian non-ferrous metal 2017 6.49 Navrátil et al. [225]
Switzerland Canton of Valais Industrial region 2017 0.2–390 Osterwalder et al. [85]
Poland Warsaw Thermometer factory 2005 122–393 Mean 147 Boszke et al. [140]
Pakistan Karachi/Lahore Highway 61.5–144 Mean 90.7 Khan et al. [226]
Germany Rhine-Westph. Floodplain soil 2017 31.2 Beckers et al. [227]
Italy Etna Volcano 0.1–0.4 Bonanno et al. [228]
European Union 23,000 samples 2009–12 0–1.59 Mean 0.04 Tóth et al. [120]

et al. [138]; Sweden—Åkerblom et al. [139]. According The global distributions of soil Hg storage and emis-
to Obrist et al. [119], soil Hg concentrations significantly sions for both preindustrial and present-day simulations
differed among land covers following the order: in different biomes are shown in Table 4 [60]. The rela-
Forested upland > planted/cultivated > herbaceous tively low soil Hg concentrations in boreal and arctic eco-
upland/shrubland > barren soils. systems are driven by extremely low deposition. The high
Concentrations in forests were an average of 2.5 times soil concentrations in desert ecosystems are driven by a
higher than those in barren locations. combination of higher deposition and extremely slow Hg
Gworek et al. Environ Sci Eur (2020) 32:128 Page 11 of 19

turnover. The concentration of Hg in soils is therefore a Mercury in plants

function of the deposition rate and carbon turnover time. Vegetation affects environmental factors at the ground
Physical and chemical properties of the soil affect the surface by reducing solar radiation, temperature, and
Hg cycle in the environment. The soil aggregate size frac- wind velocity and serves as a surface for Hg uptake [84].
tions have significant effects on the Hg content in soil. Many studies have recognized the essential role of ter-
The concentrations of Hg and other heavy metals in soils restrial plants in the biogeochemical cycling of Hg (e.g.,
and sediments generally tend to increase with decreas- Gustin et al. [149]; Fantozzi et al. [150]; Mazur et al.
ing grain size, which is due to the propensity of metals [151]).
to bind with finer particles [140]. Generally, higher values Approximately 80% of total Hg accumulated in the
of Hg in soil are found in the fraction at < 63 µm [141]. aboveground biomass is found in the leaves, and approxi-
In the Amazonian areas without anthropogenic sources, mately 1% of that Hg is methylated. The concentrations
the fine fraction (< 53 μm) of podzolized soils had higher of Hg in aspen tissue grown in high-Hg soil increases in
Hg contents than clayey soils [142]. In a temperate forest the following order [152]:
podzol, Hg mean values increased as the aggregate sizes Stems < branches < petioles < roots < leaves.
decrease, as follows: Research conducted by Leonard et al. [153] in Nevada
Clay (170 ng g−1) > fine silt (130 ng g−1) > coarse (USA) in an area with high levels of Hg contamination
silt (80 ng g−1) > fine sand (32 ng g−1) > coarse sand revealed that for the plant species Lepidium latifolium,
(14 ng g−1). 70% of the Hg taken up by the roots during the growing
Total Hg enrichment in clay-sized aggregates were 2 to season was emitted to the atmosphere.
11 times higher than the values shown by the bulk soil The main source of Hg in leaves comes from air pol-
(< 2 mm) [143]. In a heavily polluted area near the Wan- lution with ­Hg0 and not from soil contamination [149,
shan Hg mine (China), the fine soil aggregate size frac- 154, 155]. The studies by Fleck et al. [156] of Pinus res-
tions < 231 μm showed higher total Hg concentrations inosa have shown that neither woody tissue Hg nor any
and higher soil organic matter content than did the larger amount of Hg in the soil or forest floor were closely
aggregate size fractions (231 to 2000 μm) [144]. related to foliar levels, while for some relationships, the
Humic acid influences Hg transport and transforma- opposite was true. The authors interpret these data as
tion in soil–plant systems, especially for soils having low indicating that Hg in plant tissues is derived directly from
clay content. Humic acid reduces the amount of available the atmosphere and not from the soil. It is estimated that
Hg in soil and prevents Hg from being transported into in highly contaminated soils, generally less than 2% of the
plants or leached from the soil. Leaching can result in Hg Hg present is available for plants [157]. Total leaf con-
leaking into natural water systems under normal environ- centrations of Hg varied among species and were most
mental conditions. In practice, humic acid can be used closely correlated with the number of stomates per sam-
to control Hg transportation into food chains from soil ple, thus supporting the hypothesis that stomatal uptake
heavily polluted by Hg [145]. of atmospheric Hg (most likely ­Hg0) is a potential uptake
The chemical and mineralogical properties of soil affect pathway [158]. Research by Arnold et al. [159] also indi-
oxidation and retention of atmospheric Hg. Abiotic Hg cated the importance of the nonstomatal pathway for the
oxidation occurs because organic matter has -SH groups, uptake of total gaseous Hg (TGM).
which have a high affinity for Hg ions, and Hg oxida- Plants growing beyond the influence of high Hg emis-
tion is favored in the presence of compounds with high sions contained less than 100 ng g−1 THg. Plants growing
affinities for the Hg ion [121]. A microbial contribution in the vicinity of factories are large emitters of Hg, such
to Hg oxidation was first proposed by Smith et al. [146], as those around Hg mining sites (e.g., Moreno-Imenez
who demonstrated that typical soil bacteria (Bacillus and et al. [160]; Qian et al. [161]) and chlor-alkali mining
Streptomyces) can oxidize elemental Hg to ­Hg2+ through sites [162]. Au mining sites [163, 164] may also contain
enzymatic paths. Recent studies have shown that Hg can extremely high Hg contents. Mushrooms have been iden-
also be oxidized by anaerobic bacteria [147, 148]. The soil tified as organisms that accumulate more Hg than other
microbial community is very sensitive to Hg concentra- plants [165]. A synthesis of published vegetation Hg data
tions, and this sensitivity is influenced not only by soil from the western United States showed that aboveground
properties, but also by the plant species growing in the biomass concentrations followed the order [119]:
soil. A level of 0.36.mg kg−1 of Hg in soils is proposed to Leaves (26 μg kg−1) ~ branches (26 μg kg−1) > bark
be a critical concentration above which plant and soil (16 μg kg−1) > bole wood (1 μg kg−1).
organisms will be affected [121]. Hg concentrations in leaves were monitored from the
emergence to senescence and showed a strong positive
correlation with leaf age [155, 158, 166].
Gworek et al. Environ Sci Eur (2020) 32:128 Page 12 of 19

Toxic effects on plants The significant toxic effect of Hg on plants is the gener-
Hg does not have any beneficial effects on organisms and ation of reactive oxygen species (ROS) [178], e.g., super-
is thus regarded as the “main threat” since it is very harm- oxide anion radicals, H ­ 2O2, and hydroxyl radicals (OH.)
ful to both plants and animals; pollutes the air, water [179, 180]. Detoxification mechanisms to combat Hg-
and soil; and is toxic [167]. Mercury has toxic effects on induced oxidative stress include enzymatic antioxidants
plants, even at low concentrations, and leads to growth and some nonenzymatic antioxidants, such as the follow-
retardation [168] and many other adverse effects [105]. ing: glutathione [181], phytochelatin [182], salicylic acids
Hg in plants is strongly bound to sulfhydryl/thiol [183], ascorbic acid [184], selenium, [185], proline [186]
groups of proteins and forms SHgS. Hg toxicity in plants and tocopherols [187]. This process is correlated with the
occurs via its binding to SH groups of proteins, displace- disruption of biomembrane lipids and cellular metabo-
ment of essential elements and disruption of the pro- lism, resulting in plant injury [188].
tein structure [169]. This biochemical property probably Increasing levels of mercury species in the soil exert a
determines the toxic effects on plants [5, 170, 171]. Stud- wide range of adverse effects on the growth and metabo-
ies of the toxic effects of Hg on soil organisms and native lism of plants [167, 189, 190], such as reduced photosyn-
plants in fields are limited. The effects of Hg are usually thesis, transpiration, water uptake, chlorophyll synthesis
examined in sterile and much-simplified laboratory con- [188, 191, 192] and increased lipid peroxidation (Cho and
ditions, which may differ from field conditions to varying Park [179]). A high Hg content in plants affects the activ-
degrees [172]. ity of most enzymes. The total activity of stress indicators
The field study of Moreno-Jiménez et al. [160] was con- such as superoxide dismutase (SOD), peroxidase (POD)
ducted in the mining district of Almadén (Spain), which and ascorbate peroxides (APX) increased after Hg treat-
is a cinnabar (HgS) enriched zone, from which one-third ment, but the vast majority of enzymes were inhibited
of the total Hg produced worldwide is extracted. Mining at higher concentrations (e.g., Manikandan et al. [193];
activity began more than 2000 years ago, and no other Mahbub et al. [194]; Zhou et al. [195]).
region in the world has been influenced by Hg for such
a long period. The region is considered to be one of the
regions most polluted by Hg in the world. Hg concentra- Conclusions and commentary
tions in the field plants Rumex induratus and Marrubium Atmospheric contamination by mercury continues to
vulgare grown in these soils can be considered phyto- be one of the most important environmental problems
toxic, although no symptoms of Hg toxicity have been in the modern world. The following general conclusions
observed in any of the studied plant species. In most con- can be drawn from this review of the literature and are
taminated soils and mine tailings, Hg is not readily avail- accompanied by the authors’ critical commentary:
able for plant uptake [173].
The absorption of organic and inorganic Hg from soil • Models differ significantly in their estima-
by plants is low, and there is a barrier to Hg transloca- tions of global total Hg emissions—from 4000 to
tion from plant roots to tops. Thus, large increases in 9230 Mg year−1—of which natural emissions and re-
soil Hg levels produce only modest increases in plant emissions ranged from 45 to 66%.
Hg levels by direct uptake from soil [172]. In terres- • Many factors contribute to such large differences in
trial vegetation, Hg in the aboveground biomass origi- the assessments of the level of global emissions: (i)
nates primarily from the atmosphere, whereas Hg in the methodological difficulties exist in assessing re-emis-
roots comes from the soil [67, 174]. The research con- sion from heavily polluted areas under the influence
ducted by Lomonte et al. [175] suggested the existence of contemporary and historical emissions and areas
of Hg stress-activated defense mechanisms in plants and with background pollution; (ii) the transformation
hypothesized that these mechanisms were likely the rea- of various forms of Hg depends on many difficult to
son for the increased production of sulfur compounds in evaluate processes, which makes estimating emis-
the tested plant species, which stimulated their growth. sions difficult; and (iii) unusual phenomena associ-
Hg has very limited solubility in soil, low availability for ated with the transformation of various Hg forms
plant uptake and no known biological function, which (e.g., mercury depletion events (MDEs), which con-
may explain why Hg-hyperaccumulating plants have not sist of episodes of sudden drops in total gaseous mer-
yet been identified, meaning that a method for Hg phy- cury concentrations in the air in the Antarctic and
toremediation in soils contaminated with Hg has not yet Arctic, can occur.
been developed [175]. However, studies suggesting the • Despite the many applicable global, regional and
use of transgenic plants for phytoremediation have been national programs and conventions aimed at reduc-
published recently [176, 177]. ing Hg emissions, global total Hg emissions in the
Gworek et al. Environ Sci Eur (2020) 32:128 Page 13 of 19

years 2000–2015 increased by 1.8%. In many coun- • A large number of scientific publications have been
tries, including those with a high national income, devoted to the problem of Hg environmental pollu-
there is a lack of understanding by society at large tion. However, these studies face many difficulties: (i)
and politicians about the need to reduce emissions. analytical difficulties exist, which are caused by very
• The proportions between the amounts of natu- low Hg contents in all elements of the environment;
ral emissions and anthropogenic emissions have (ii) the need to determine a specific form of Hg pro-
not been precisely determined. This ratio, which is hibits providing results in total Hg; and (iii) the form
dependent on the authors, has been estimated over a in which Hg occurs depends on many environmen-
relatively wide range of 0.8–1.8. tal factors that must be accurately recognized and
• Annual global Hg emissions from natural sources on described. Unfortunately, many scientific publica-
land are estimated by various authors over a wide tions do not meet these requirements.
range, with total emissions of 1600–2500 Mg year−1,
including re-emissions of 790–2000 Mg year−1. The
low share of Hg taken up by plants from the soil is
beneficial from perspective of protecting food against
Abbreviations
contamination, although it also limits the possibility Hg: Mercury; UN: United Nations; UNEP: United Nations Environment Pro‑
of using plants in the phytoremediation of contami- gramme; GEM: Gaseous elemental mercury ­Hg0; GOM: Gaseous mercury in
nated soils. oxidized form; TPM: Total particulate mercury; TGM: Total gaseous Hg; ROS:
Reactive oxygen species; OH: Hydroxyl radicals; SOD: Superoxide dismutase;
• On the land surface, Hg deposition is mainly in the POD: Peroxidase; APX: Ascorbate peroxides.
oxidized form ­ (Hg2+), and its transformations are
associated primarily with the oxidation–reduction Acknowledgements
Not applicable.
potential of the environment and the biological and
chemical processes of methylation. Authors’ contributions
• The main source of Hg in plant leaves comes from WD has been responsible for the concept of the manuscript and drafted the
manuscript. BG and AH B-D helped to further elaborate the manuscript. All
air pollution with ­Hg0 and not from soil contamina- authors improved the final manuscript. All authors read and approved the
tion. It is very difficult to estimate Hg emissions from final manuscript.
plants, which mainly occur in the form of H ­ g0.
Funding
• Methyl-Hg can be produced both biotically and abi- The work was financed from the Own Research Fund Institute of Environment
otically in the environment. Methylation of Hg tends Protection - National Research Institute.
to occur in environments with low oxygen levels, low
Availability of data and materials
pH, Hg bioavailability, temperature, redox potential Not applicable; presented information is based on previously published data
and high dissolved organic compound levels, and only.
environments favored by SRB are largely responsible
Ethics approval and consent to participate
for this methylation. Not applicable.
• Rice growing conditions mean that the Hg meth-
ylation process is extremely intensive. Consequently, Consent for publication
Not applicable.
rice may contain significantly more Hg than other
crops, which is particularly dangerous because rice in Competing interests
many regions of the world is the basis for feeding the The authors declare that they have no competing interests.
population. Author details
• Hg has very limited solubility in soil and low avail- 1
Institute of Environmental Protection – National Research Institute, Warsaw,
ability for plant uptake, and it does not have any Poland. 2 Polish Academy of Sciences Botanical Garden – Center for Biological
Diversity Conservation in Powsin, Prawdziwka 2 St, 02‑973 Warsaw, Poland.
known biological function. These factors may explain
why Hg-hyperaccumulating plants have not yet been Received: 9 March 2020 Accepted: 14 September 2020
identified, meaning that an effective phytoreme-
diation methods for soil contaminated with Hg has
not yet been developed, which may explain why Hg
hyperaccumulator plants have not found practical References
1. Pacyna JM (2020) Recent advances in mercury research. Sci Total Envi‑
use in phytoremediation of contaminated soils. Such
ron 738:139955. https​://doi.org/10.1016/j.scito​tenv.2020.13995​5
work may be more applicable in wetland environ- 2. Environment UN (2019) Global mercury assessment 2018. UN Environ‑
ments. The possibility of using contaminated soil in ment Programme, Chemical and Health Branch Geneva, Switzerland
3. Bank MS (2020) The mercury science-policy interface: history, evolu‑
phytoremediation by transgenic plants is promising
tion and progress of the Minamata Convention. Sci Total Environ
and a future research direction. 722:137832. https​://doi.org/10.1016/j.scito​tenv.2020.13783​2
Gworek et al. Environ Sci Eur (2020) 32:128 Page 14 of 19

4. AMAP, UN Environment (2019) Technical background report for the 24. Agnan Y, Le Dantec T, Moore CW, Edwards GC, Obrist D (2016) New con‑
global mercury assessment 2018. Arctic Monitoring and Assessment straints on terrestrial surface–atmosphere fluxes of gaseous elemental
Programme, Oslo, Norway/UN Environment Programme, Chemicals mercury using a global database. Environ Sci Technol 50:507–524. https​
and Health Branch, Geneva, Switzerland ://doi.org/10.1021/acs.est.5b040​13
5. Kabata-Pendias A, Pendias H (2001) Trace elements in soils and plants, 25. Travnikov O, Dastoor A, Bullock R, Christensen JH (2008) Modeling
3rd edn. CRC Press, Boca Raton. https​://doi.org/10.1201/97814​20039​ atmospheric transport and deposition. In: AMAP/UNEP, technical back‑
900 ground report to the global atmospheric mercury assessment. Arctic
6. Newman MC (2014) Fundamentals of ecotoxicology: the science of pol‑ Monitoring and Assessment Programme, UNEP Chemical Branch, pp
lution, 4th edn. CRC Press, Boca Raton. https​://doi.org/10.1201/b1765​8 79–107
7. Hylander LD, Meili M (2003) 500 years of mercury production: global 26. Dastoor AP, Larocque Y (2004) Global circulation of atmospheric
annual inventory by region until 2000 and associated emissions. Sci mercury: a modelling study. Atmos Environ 38:147–161. https​://doi.
Total Environ 304:13–27. https​://doi.org/10.1016/S0048​-9697(02)00553​ org/10.1016/j.atmos​env.2003.08.037
-3 27. Henze DK, Hakami A, Seinfeld JH (2007) Development of the adjoint
8. WHO (2017) Ten chemicals of major health concern. http://www.who. of GEOS-Chem. Atmos Chem Phys 79:2413–2433. https​://resol​ver.calte​
int/ipcs/asses​sment​/publi​c_healt​h/chemi​cals_phc/en/index​.html ch.edu/Calte​chAUT​HORS:HENac​p07
9. Minamata Convention on Mercury. (2020). http://www.mercu​r ycon​ 28. Seigneur C, Vijayaraghavan K, Lohman K, Levin L (2009) The AER/EPRI
venti​on.org/Porta​ls/11/docum​ents/conve​ntion​Text/Minam​ata global chemical transport model for mercury (CTM-HG). In: Pirrone N,
10. Selin H, Keane SE, Wang S, Selin NE, Davis K, Bally D (2018) Linking Mason R (eds) Mercury fate and transports in the global atmosphere.
science and policy to support the implementation of the Minamata Springer, New York, pp 589–602
Convention on Mercury. Ambio 47:198–215. https​://doi.org/10.1007/ 29. Travnikov O, Ilyin I (2009) The EMEP/MSC-E mercury modeling
s1328​0-017-1003-x system. In: Pirrone N, Mason R (eds) Mercury fate and transports in
11. Streets DG, Horowitz HM, Lu Z, Levin L, Thackray CP, Sunderland the global atmosphere. Springer, New York, pp 571–587. https​://doi.
EM (2019) Global and regional trends in mercury emissions and org/10.1007/978-0-387-93958​-2_20
concentrations. 2010–2015. Atmos Environ 201:417–427. https​://doi. 30. Travnikov O, Angot H, Artaxo P, Bencardino M, Bieser J, D’Amore F et al
org/10.1016/j.atmos​env.2018.12.031 (2017) Multi-model study of mercury dispersion in the atmosphere:
12. Obrist D, Kirk JL, Zhang L, Sunderland EM, Jiskra M, Selin NE (2018) atmospheric processes and model evaluation. Atmos Chem Phys. https​
A review of global environmental mercury processes in response ://doi.org/10.5194/acp-17-5271-201
to human and natural perturbations: changes of emissions, climate, 31. De Simone F, Gencarelli CN, Hedgecock IM, Pirrone N (2014) Global
and land use. Ambio 47:116–140. https​://doi.org/10.1007/s1328​ atmospheric cycle of mercury: a model study on the impact of
0-017-1004-9 oxidation mechanisms. Environ Sci Pollut R 21:4110–4123. https​://doi.
13. ICMGP 2019) 14th international conference on mercury as a global pol‑ org/10.1007/s1135​6-013-2451-x
lutant, 8–13 September 2019 Cracow, Poland. http://www.mercu​r y201​ 32. Nriagu JO, Pacyna JM (1998) Quantitative assessment of worldwide
9krak​ow.com contamination of air, water and soils by trace metals. Nature 333:134–
14. Liberati A, Altman DG, Tetzlaff J, Mulrow C, Gøtzsche PC, Ioannidis JP 139. https​://doi.org/10.1038/33313​4a0
et al (2009) The PRISMA statement for reporting systematic reviews 33. Shetty SK, Lin CJ, Streets DG, Jang C (2008) Model estimate of mercury
and meta-analyses of studies that evaluate health care interventions: emission from natural sources in East Asia. Atmos Environ 42:8674–
explanation and elaboration. J Clin Epidemiol 62:1–34. https​://doi. 8685. https​://doi.org/10.1016/j.atmos​env.2008.08.026
org/10.1016/j.jclin​epi.2009.06.006 34. Liu G, Cai Y, O’Driscoll N, Feng X, Jiang G (2012) Overview of mercury
15. Pirrone N, Cinnirella S, Feng X, Finkelman RB, Friedli HR, Leaner J et al in the environment. In: Liu G, Cai Y, O’Driscoll N (eds) Environmental
(2010) Global mercury emissions to the atmosphere from anthropo‑ chemistry and toxicology of mercury. Wiley, Hoboken, pp 1–12
genic and natural sources. Atmos Chem and Phys 10:5951–5964. https​ 35. Gustin MS, Lindberg SE, Austin K, Coolbaugh M, Vette A, Zhang H (2000)
://doi.org/10.5194/acp-10-5951-2010 Assessing the contribution of natural sources to regional atmospheric
16. Pacyna EG, Pacyna JM, Sundsetha K, Munthec J, Kindbomc K, Wilsond mercury budgets. Sci Total Environ 259:61–71. https​://doi.org/10.1016/
S et al (2010) Global emission of mercury to the atmosphere from S0048​-9697(00)00556​-8
anthropogenic sources in 2005 and projections to 2020. Atmos Environ 36. Gustin MS, Lindberg SE, Weisberg PJ (2008) An update on the natural
44:2487–2499. https​://doi.org/10.1016/j.atmos​env.2009.06.009 sources and sinks of atmospheric mercury. Appl Geoch 23:482–493.
17. Gworek B, Dmuchowski W, Baczewska AH, Brągoszewska P, Bemowska- https​://doi.org/10.1016/j.apgeo​chem.2007.12.010
Kałabun O, Wrzosek-Jakubowska J (2017) Air contamination by mercury, 37. Streets DG, Horowitz HM, Jacob DJ, Lu Z, Levin L, Ter Schure AF et al
emissions and transformations—a review. Water Air Soil Poll 228:123. (2017) Total mercury released to the environment by human activi‑
https​://doi.org/10.1007/s1127​0-017-3311-y ties. Environ Sci Technol 51:5969–5977. https​://doi.org/10.1021/acs.
18. UNEP Chemical Branch (2008) The global atmospheric mercury assess‑ est.7b004​51
ment: sources, emissions and transport. UNEP-Chemicals, Geneva 38. Mason RP (2009) Mercury emissions from natural processes and their
19. Cooke CA, Martínez-Cortizas A, Bindler R, Gustin MS (2020) Environ‑ importance in the global mercury cycle. In: Pirrone N, Mason R (eds)
mental archives of atmospheric Hg deposition—a review. Sci Total Mercury fate and transports in the global atmosphere. Springer, New
Environ 709:134800. https​://doi.org/10.1016/j.scito​tenv.2019.13480​0 York, pp 173–191. https​://doi.org/10.1007/978-0-387-93958​-2_7
20. Miszczak E, Stefaniak S, Michczyński A, Steinnes E, Twardowska I (2020) 39. Wang Y, Greger M (2004) Clonal differences in mercury tolerance, accu‑
A novel approach to peatlands as archives of total cumulative spatial mulation, and distribution in willow. J Environ Qual 33:1779–1785. https​
pollution loads from atmospheric deposition of airborne elements ://doi.org/10.2134/jeq20​04.1779
complementary to EMEP data: priority pollutants (Pb, Cd, Hg). Sci Total 40. Bergan T, Gallardo L, Rohde H (1999) Mercury in the global tropo‑
Environ 705:135776. https​://doi.org/10.1016/j.scito​tenv.2019.13577​6 sphere—a three-dimensional model study. Atmos Environ 33:1575–
21. Zhu C, Tian H, Hao J (2020) Global anthropogenic atmospheric emission 1585. https​://doi.org/10.1016/S1352​-2310(98)00370​-7
inventory of twelve typical hazardous trace elements, 1995–2012. Atmos 41. Mason RP, Sheu GR (2002) Role of the ocean in the global mercury cycle.
Environ 220:117061. https​://doi.org/10.1016/j.atmos​env.2019.11706​1 Global Biogeoch Cy 16:1093. https​://doi.org/10.1029/2001G​B0014​40
22. Eckley C, Blanchard P, McLennan D, Mintz R, Sekela M (2015) Soil–air 42. Lamborg CH, Fitzgerald WF, O’Donnell J, Torgersen T (2002) A
mercury flux near a large industrial emission source before and after non-steady-state compartmental model of global-scale mercury
closure (Flin Flon, Manitoba, Canada). Environ Sci Techno 49:9750–9757. biogeochemistry with interhemispheric atmospheric gradients.
https​://doi.org/10.1021/acs.est.5b019​95 Geochim Cosmochim Ac 66:1105–1118. https​://doi.org/10.1016/S0016​
23. Eckley CS, Gilmour CC, Janssen S, Luxton TP, Randall PM, Whalin L et al -7037(01)00841​-9
(2020) The assessment and remediation of mercury contaminated sites: 43. Nriagu J, Becker C (2003) Volcanic emissions of mercury to the atmos‑
a review of current approaches. Sci Total Environ 707:136031. https​:// phere: global and regional inventories. Sci Total Environ 304:3–12. https​
doi.org/10.1016/j.scito​tenv.2019.13603​1 ://doi.org/10.1016/S0048​-9697(02)00552​-1
Gworek et al. Environ Sci Eur (2020) 32:128 Page 15 of 19

44. Pyle DM, Macher RA (2003) The importance of volcanic emissions for 62. Zheng W, Obrist D, Weis D, Bergquist BA (2016) Mercury isotope
the global atmospheric mercury cycle. Atmos Environ 37:5115–5124. compositions across North American forests. Global Biogeoch Cy
https​://doi.org/10.1016/j.atmos​env.2003.07.011 30:1475–1492. https​://doi.org/10.1002/2015G​B0053​23
45. Bagnato E, Aiuppa A, Parello F, Allard P, Shinohara H, Liuzzo M et al 63. Sommar J, Osterwalder S, Zhu W (2020) Recent advances in under‑
(2011) New clues on the contribution of Earth’s volcanism to the global standing and measurement of Hg in the environment: surface-atmos‑
mercury cycle. B Volcanol 73:497–510. https​://doi.org/10.1007/s0044​ phere exchange of gaseous elemental mercury ­(Hg0). Sci Total Environ.
5-010-0419-y https​://doi.org/10.1016/j.scito​tenv.2020.13764​8
46. Bagnato E, Aiuppa A, Parello F, Calabrese S, D’Alessandro W, Mather 64. Wright LP, Zhang L, Marsik FJ (2016) Overview of mercury dry deposi‑
TA et al (2007) Degassing of gaseous (elemental and reactive) and tion, litterfall, and throughfall studies. Atmos Chem Phys 16:13399.
particulate mercury from Mount Etna volcano (Southern Italy). Atmos https​://doi.org/10.5194/acp-16-13399​-201
Environ 41:377–7388. https​://doi.org/10.1016/j.atmos​env.2007.05.060 65. Lyman SN, Cheng I, Gratz LE, Weiss-Penzias P, Zhang L (2020) An
47. Witt MLI, Mather TA, Pyle DM, Aiuppa A, Bagnato E, Tsanev VI (2008) updated review of atmospheric mercury. Sci Total Environ 707:135575.
Mercury and halogen emissions from Masaya and Telica volcanoes. https​://doi.org/10.1016/j.scito​tenv.2019.13557​5
J Geophys Res Solid Earth, Nicaragua. https​://doi.org/10.1029/2007J​ 66. Skov H, Bullock R, Christensen JH, Sørensen LL (2008). Atmospheric
B0054​01 pathways. In: AMAP/UNEP, technical background report to the global
48. Martin RS, Witt MLI, Pyle DM, Mather TA, Watt SFL, Bagnato E et al atmospheric mercury assessment. Arctic Monitoring and Assessment
(2011) Rapid oxidation of mercury (Hg) at volcanic vents: insights Programme, UNEP Chemical Branch, pp 64–72
from high temperature thermodynamic models of Mt Etna’s emis‑ 67. Selin NE, Jacob DJ, Park RJ, Yantosca RM, Strode S, Jaeglé L et al (2007)
sions. Chem Geol 283:279–286. https​://doi.org/10.1016/j.chemg​ Chemical cycling and deposition of atmospheric mercury: global
eo.2011.01.027 constraints from observations. J Geophys Res 112:D02308. https​://doi.
49. Ermolin MS, Fedotov PS, Malik NA, Karandashev VK (2018) Nanoparti‑ org/10.1029/2006J​D0074​50
cles of volcanic ash as a carrier for toxic elements on the global scale. 68. Lindberg SE, Dong W, Meyers T (2002) Transpiration of gaseous
Chemosphere 200:16–22. https​://doi.org/10.1016/j.chemo​spher​ elemental mercury through vegetation in a subtropical wetland in
e.2018.02.089 Florida. Atmos Environ 36:5207–5219. https​://doi.org/10.1016/S1352​
50. Coolbaugh M, Gustin M, Rytuba J (2002) Annual emissions of -2310(02)00586​-1
mercury to the atmosphere from natural sources in Nevada and 69. Marsik FJ, Keeler GJ, Landis MS (2007) The dry-deposition of speci‑
California. Environ Geol 42:338–349. https​://doi.org/10.1007/s0025​ ated mercury to the Florida Everglades: measurements and mod‑
4-002-0557-4 eling. Atmos Environ 41:136–149. https​://doi.org/10.1016/j.atmos​
51. Cinnirella S, Pirrone N (2006) Spatial and temporal distributions of env.2006.07.032
mercury emissions from forest fires in Mediterranean region and Rus‑ 70. Hanson PJ, Lindberg SE, Tabberer TA, Owens JA, Kim KH (1995) Foliar
sian federation. Atmos Environ 40:7346–7361. https​://doi.org/10.1016/j. exchange of mercury vapor: evidence for a compensation point. Water
atmos​env.2006.06.051 Air Soil Poll 80:373–382. https​://doi.org/10.1007/BF011​89687​
52. Friedli HR, Arellano AF, Cinnirella S, Pirrone N (2009) Mercury emissions 71. Caffrey PF, Ondov JM, Zufall MJ, Davidson CI (1998) Determination of
from global biomass burning: spatial and temporal distribution. In: size-dependent dry particle deposition velocities with multiple intrinsic
Mason R, Pirrone N (eds) Mercury fate and transport in the global elemental tracers. Environ Sci Technol 32:1615–1622. https​://doi.
atmosphere. Springer, Boston. https​://doi.org/10.1007/978-0-387-93958​ org/10.1021/es970​644f
-2_8 72. Converse AD, Riscassi AL, Scanlon TM (2010) Seasonal variability in
53. Obrist D, Gustin MS, Arnone JA III, Johnson DW, Schorran DE, Verburg gaseous mercury fluxes measured in a high-elevation meadow. Atmos
PS (2005) Measurements of gaseous elemental mercury fluxes over Environ 44:2176–2185. https​://doi.org/10.1016/j.atmos​env.2010.03.024
intact tallgrass prairie monoliths during one full year. Atmos Environ 73. Carpi A, Fostier AH, Orta OR, dos Santos JC, Gittings M (2014) Gaseous
39:957–965. https​://doi.org/10.1016/j.atmos​env.2004.09.081 mercury Carp emissions from soil following forest loss and land use
54. Woodruff LG, Cannon WF (2010) Immediate and long-term fire effects changes: field experiments in the United States and Brazil. Atmos
on total mercury in forests soils of northeastern Minnesota. Environ Sci Environ 96:423–429. https​://doi.org/10.1016/j.atmos​env.2014.08.004
Technol 44:5371–5376. https​://doi.org/10.1021/es100​544d 74. Zhang H, Lindberg SE, Marsik FJ, Keeler GJ (2001) Mercury air/sur‑
55. Gagliano AL, Calabrese S, Daskalopoulou K, Cabassi J, Capecchiacci F, face exchange kinetics of background soils of the Tahquamenon
Tassi F et al (2019) Degassing and cycling of mercury at nisyros volcano River watershed in the Michigan Upper Peninsula. Water Air Soil Poll
(Greece). Geofluids. https​://doi.org/10.1155/2019/47835​14 126:151–169
56. Pierce AM, Moore CW, Wohlfahrt G, Hörtnagl L, Kljun N, Obrist D (2015) 75. Ferrari CP, Dommerguea A, Veysseyrea A, Planchona F, Boutrona CF
Eddy covariance flux measurements of gaseous elemental mercury (2002) Mercury speciation in the French seasonal snow cover. Sci Total
using cavity ring-down spectroscopy. Environ Sci Technol 49:1559– Environ 287:61–69. https​://doi.org/10.1016/S0048​-9697(01)00999​-8
1568. https​://doi.org/10.1021/es505​080z 76. Gustin MS, Biester H, Kim CS (2002) Investigation of the light-enhanced
57. Obrist D, Johnson DW, Lindberg SE (2009) Mercury concentrations and emission of mercury from naturally enriched substrates. Atmos Environ
pools in four Sierra Nevada forest sites, and relationships to organic car‑ 36:3241–3254. https​://doi.org/10.1016/S1352​-2310(02)00329​-1
bon and nitrogen. Biogeosciences 6:765–777. https​://doi.org/10.1016/j. 77. Poissant L, Pilote M, Xu X, Zhang H, Beauvais C (2004) Atmospheric
atmos​env.2004.09.081 mercury speciation and deposition in the Bay St. Francois wetlands. J
58. Jiskra M, Sonke JE, Obrist D, Bieser J, Ebinghaus R, Myhre CL et al (2018) Geophys Res 109:1–14. https​://doi.org/10.1029/2003J​D0043​64
A vegetation control on seasonal variations in global atmospheric mer‑ 78. Gustin MS (2003) Are mercury emissions from geologic sources
cury concentrations. Nat Geosci 11:244–250. https​://doi.org/10.1038/ significant? A status report. Sci Total Environ. 304:153–167. https​://doi.
s4156​1-018-0078-8 org/10.1016/S0048​-9697(02)00565​-X
59. Gustin MS (2012) Exchange of mercury between the atmosphere and 79. Gustin MS, Lindberg SE (2005) Terrestial Hg Fluxes: is the next exchange
terrestrial ecosystems. In: Liu G, Cai Y, O’Driscoll N (eds) Environmental up, down, or neither? In: Pirrone N, Mahaffey KR (eds) Dynamics of
chemistry and toxicology of mercury. Wiley, Hoboken, pp 423–451 mercury pollution on regional and global scales. Springer, Boston, pp
60. Smith-Downey NV, Sunderland EM, Jacob DJ (2010) Anthropogenic 241–259. https​://doi.org/10.1007/0-387-24494​-8_11
impacts on global storage and emissions of mercury from terrestrial 80. Stamenkovic J, Gustin MS (2007) Evaluation of use of EcoCELL technol‑
soils: insights from a new global model. Biogeosci, J Geophys Res. https​ ogy for quantifying total gaseous mercury fluxes over background
://doi.org/10.1029/2009J​G0011​24 substrates. Atmos Environ 41:3702–3712. https​://doi.org/10.1016/j.
61. Graydon JA, St. Louis VL, Lindberg SE, Sandilands KA, Rudd JW, Kelly atmos​env.2006.12.037
CA et al (2012) The role of terrestrial vegetation in atmospheric 81. Ericksen JA, Gustin MS, Xin M, Weisberg PJ, Fernandez GCJ (2006)
Hg deposition: pools and fluxes of spike and ambient Hg from the Air–soil exchange of mercury from background soils in the United
METAALICUS experiment. Global Biogeochem Cy 26:GB1022. https​:// States. Sci Total Environ 366:851–863. https​://doi.org/10.1016/j.scito​
doi.org/10.1029/2011G​B0040​31 tenv.2005.08.019
Gworek et al. Environ Sci Eur (2020) 32:128 Page 16 of 19

82. Mason RP, Pirrone N, Hedgecock I, Suzuki N, Levin L (2010) Conceptual Valais. Switzerland. Soil Systems 2:44. https​://doi.org/10.3390/soils​ystem​
overview. In: Pirrone N, Keating T (eds) Hemispheric transport of air s2030​044
pollution—part B. United Nations Publication, New York, pp 1–19. https​ 100. Lima FRD, Martins GC, Silva AO, Vasques ICF, Engelhardt MM, Cândido
://doi.org/10.18356​/38ccc​958-en GS et al (2019) Critical mercury concentration in tropical soils: impact
83. Carpi A, Lindberg SE (1997) Sunlight-mediated emission of elemental on plants and soil biological attributes. Sci Total Environ 666:472–479.
mercury from soil amended with municipal sewage sludge. Environ Sci https​://doi.org/10.1016/j.scito​tenv.2019.02.216
Technol 31:2085–2091. https​://doi.org/10.1021/Es960​910+ 101. Gonzalez-Raymat H, Liu G, Liriano C, Li Y, Yin Y, Shi J et al (2017) Elemen‑
84. Zhu W, Lin CJ, Wang X, Sommar J, Fu X, Feng X (2016) Global observa‑ tal mercury: its unique properties affect its behavior and fate in the
tions and modeling of atmosphere-surface exchange of elemental environment. Environ Pollut 229:69–86. https​://doi.org/10.1016/j.envpo​
mercury: a critical review. Atmos Chem Phys 16:4451–4480. https​://doi. l.2017.04.101
org/10.5194/acp-16-4451-2016 102. Xu J, Buck M, Eklöf K, Ahmed OO, Schaefer JK, Bishop K et al (2019)
85. Osterwalder S, Huang JH, Shetaya WH, Agnan Y, Frossard A, Frey B et al Mercury methylating microbial communities of boreal forest soils. Sci
(2019) Mercury emission from industrially contaminated soils in rela‑ Rep 9:518. https​://doi.org/10.1038/s4159​8-018-37383​-z
tion to chemical, microbial, and meteorological factors. Environ Pollut 103. Ling Q, Guo Y, Liang Y, Yin Y, Cai Y (2020) Microbial uptake of HgS
250:944–952. https​://doi.org/10.1016/j.envpo​l.2019.03.093 nanoparticles and its effect on mercury methylation. Environ Chem
86. Gabriel MC, Williamson DG, Brooks S, Zhang H, Lindberg S (2005) Spatial 2:292–300. https​://doi.org/10.7524/j.issn.0254-6108.20191​12601​
variability of mercury emissions from soils in a southeastern US urban 104. Bigham GN, Murray KJ, Masue-Slowey Y, Henry EA (2017) Biogeochemi‑
environment. Environ Geol 48:955–964. https​://doi.org/10.1007/s0025​ cal controls on methylmercury in soils and sediments: implications for
4-005-0043-x site management. Integr Environ Assess Manag 13:249–263. https​://doi.
87. Ma M, Wang D, Sun R, Shen Y, Huang L (2013) Gaseous mercury emis‑ org/10.1002/ieam.1822
sions from subtropical forested and open field soils in a national nature 105. Shahid M, Khalid S, Bibi I, Bundschuh J, Niazi NK, Dumat C (2020)
reserve, southwest. China Atmos Environ 64:116–123. https​://doi. A critical review of mercury speciation, bioavailability, toxicity and
org/10.1016/j.atmos​env.2012.09.038 detoxification in soil-plant environment: ecotoxicology and health risk
88. Wang X, Lin CJ, Yuan W, Sommar J, Zhu W, Feng X (2016) Emission-dom‑ assessment. Sci Total Environ 711:134749. https​://doi.org/10.1016/j.scito​
inated gas exchange of elemental mercury vapor over natural surfaces tenv.2019.13474​9
in China. Atmos Chem Phys 16:11125–11143. https​://doi.org/10.5194/ 106. Zhao L, Qiu G, Anderson CW, Meng B, Wang D, Shang L et al (2016)
acp-16-11125​-2016 Mercury methylation in rice paddies and its possible controlling factors
89. Sommar J, Zhu W, Shang L, Lin CJ, Feng XB (2015) Seasonal variations in the Hg mining area, Guizhou province, Southwest China. Environ
in metallic mercury (Hg 0) vapor exchange over biannual wheat- Pollut 215:1–9. https​://doi.org/10.1016/j.envpo​l.2016.05.001
corn rotation cropland in the North China Plain. Biogeosci Discuss 107. Eklöf K, Bishop K, Bertilsson S, Björn E, Buck M, Skyllberg U et al (2018)
12:16105–16158. https​://doi.org/10.5194/bgd-12-16105​-2015 Formation of mercury methylation hotspots as a consequence of
90. Gao Y, Wang Z, Zhang X, Wang C (2020) Observation and estimation of forestry operations. Sci Total Environ 613:1069–1078. https​://doi.
mercury exchange fluxes from soil under different crop cultivars and org/10.1016/j.scito​tenv.2017.09.151
planting densities in North China Plain. Environ Pollut 259:113833. https​ 108. Qiu G, Feng X, Wang S, Shang L (2005) Mercury and methylmercury
://doi.org/10.1016/j.envpo​l.2019.11383​3 in riparian soil, sediments, mine-waste calcines, and moss from aban‑
91. Bishop K, Shanley JB, Riscassi A, de Wit HA, Eklöf K, Meng B et al (2020) doned Hg mines in east Guizhou province, southwestern China. Appl
Recent advances in understanding and measurement of mercury in the Geoch 20:627–638. https​://doi.org/10.1016/j.apgeo​chem.2004.09.006
environment: terrestrial Hg cycling. Sci Total Environ 721:137647. https​ 109. Xu X, Meng B, Zhang C, Feng X, Gu C, Guo J et al (2017) The local
://doi.org/10.1016/j.scito​tenv.2020.13764​7 impact of a coal-fired power plant on inorganic mercury and methyl-
92. Qiu G, Feng X, Wang S, Shang L (2006) Environmental contamination mercury distribution in rice (Oryza sativa L.). Environ Pollut 223:11–18.
of mercury from Hg-mining areas in Wuchuan, northeastern Guizhou, https​://doi.org/10.1016/j.envpo​l.2016.11.042
China. Environ Pollut 142:549–558. https​://doi.org/10.1016/j.envpo​ 110. Mailman M, Bodaly RA (2005) Total mercury, methyl mercury, and car‑
l.2005.10.015 bon in fresh and burned plants and soil in Northwestern Ontario. Envi‑
93. Li Q, Tang L, Qiu G, Liu C (2020) Total mercury and methylmercury in ron Pollut 138:161–166. https​://doi.org/10.1016/j.envpo​l.2005.02.005
the soil and vegetation of a riparian zone along a mercury-impacted 111. Schuster PF, Schaefer KM, Aiken GR, Antweiler RC, Dewild JF, Gryziec JD
reservoir. Sci Total Environ 738:139794. https​://doi.org/10.1016/j.scito​ et al (2018) Permafrost stores a globally significant amount of mercury.
tenv.2020.13979​4 Geophys Res Lett 45:1463–1471. https​://doi.org/10.1002/2017G​L0755​
94. Goulet RR, Holmes J, Tessier A, Wang F, Siciliano SD, Page B et al (2007) 71
Mercury methylation in sediments of a riverine marsh: the role of redox 112. Hararuk O, Obrist D, Luo Y (2013) Modelling the sensitivity of soil
conditions sulfur chemistry and microbial communities. Geochim mercury storage to climate-induced changes in soil carbon pools. Bio‑
Cosmochim Ac. 71:3396–3406 geosciences 10:2393–2407. https​://doi.org/10.5194/bg-10-2393-2013
95. Windham-Myers L, Marvin-DiPasquale M, Kakouros E, Agee JL, Kieu LH, 113. Amos HM, Sonke JE, Obrist D, Robins N, Hagan N, Horowitz HM, Mason
Stricker CA et al (2014) Mercury cycling in agricultural and managed RP et al (2015) Observational and modeling constraints on global
wetlands of California, USA: seasonal influences of vegetation on mer‑ anthropogenic enrichment of mercury. Environ Sci Technol 49:4036–
cury methylation, storage, and transport. Sci Total Environ 484:308–318. 4047. https​://doi.org/10.1021/es505​8665
https​://doi.org/10.1016/j.scito​tenv.2013.05.027 114. Mason RP, Choi AL, Fitzgerald WF, Hammerschmidt CR, Lamborg CH,
96. Kronberg RM, Jiskra M, Wiederhold JG, Bjorn E, Skyllberg U (2016) Soerensen AL et al (2012) Mercury biogeochemical cycling in the
Methyl mercury formation in hillslope soils of boreal forests: the role of ocean and policy implications. Environ Res 119:101–117. https​://doi.
forest harvest and anaerobic microbes. Environ Sci Technol 50:9177– org/10.1016/j.envre​s.2012.03.013
9186. https​://doi.org/10.1021/acs.est.6b007​62 115. Fitzgerald WF, Lamborg CH (2014) Geochemistry of mercury in the
97. Gnamuš A, Byrne AR, Horvat M (2000) Mercury in the soil-plant-deer- environment. In: Treatise on Geochemistry, 2nd edition
predator food chain of a temperate forest in Slovenia. Environ Sci 116. Amos HM, Jacob DJ, Streets DG, Sunderland EM (2013) Legacy impacts
Technol 34:3337–3345. https​://doi.org/10.1021/es991​419w of all-time anthropogenic emissions on the global mercury cycle.
98. Meng B, Feng X, Qiu G, Cai Y, Wang D, Li P et al (2010) Distribution pat‑ Global Biogeoch Cy 27:410–421. https​://doi.org/10.1002/gbc.20040​
terns of inorganic mercury and methylmercury in tissues of rice (Oryza 117. Wang J, Feng X, Anderson CW, Xing Y, Shang L (2012) Remediation of
sativa L.) plants and possible bioaccumulation pathways. J Agr Food mercury contaminated sites–a review. J Hazard Mater 221:1–18. https​://
Chem 58:4951–4958. https​://doi.org/10.1021/jf904​557x doi.org/10.1016/j.jhazm​at.2012.04.035
99. Gilli R, Karlen C, Weber M, Rüegg J, Barmettler K, Biester H et al (2018) 118. Li S, Jia Z (2018) Heavy metals in soils from a representative rapidly
Speciation and mobility of mercury in soils contaminated by legacy developing megacity (SW China): levels, source identification and
emissions from a chemical factory in the Rhône valley in canton of apportionment. CATENA 163:414–423. https​://doi.org/10.1016/j.caten​
a.2017.12.035
Gworek et al. Environ Sci Eur (2020) 32:128 Page 17 of 19

119. Obrist D, Pearson C, Webster J, Kane T, Lin CJ, Aiken GR et al (2016) A 138. Navrátil T, Shanley J, Rohovec J, Hojdová M, Penížek V, Buchtová J (2014)
synthesis of terrestrial mercury in the western United States: spatial dis‑ Distribution and pools of mercury in Czech forest soils. Water Air Soil
tribution defined by land cover and plant productivity. Sci Total Environ Poll 225:1829. https​://doi.org/10.1007/s1127​0-013-1829-1
568:522–535. https​://doi.org/10.1016/j.scito​tenv.2015.11.104 139. Åkerblom S, Meili M, Bringmark L, Johansson K, Kleja DB, Bergkvist B
120. Tóth G, Hermann T, Szatmári G, Pásztor L (2016) Maps of heavy metals in (2008) Partitioning of Hg between solid and dissolved organic matter in
the soils of the European Union and proposed priority areas for detailed the humus layer of boreal forests. Water Air Soil Poll 189:239–252. https​
assessment. Sci Total Environ 565:1054–1062. https​://doi.org/10.1016/j. ://doi.org/10.1007/s1127​0-007-9571-1
scito​tenv.2016.05.115 140. Boszke L, Kowalski A, Siepak J (2004) Grain size partitioning of mercury
121. Lima LC, Egreja Filho FB, Linhares LA, Windmoller CC, Yoshida MI (2015) in sediments of the middle Odra River (Germany/Poland). Water Air Soil
Accumulation and oxidation of elemental mercury in tropical soils. Poll 159:125–138. https​://doi.org/10.1023/B:WATE.00000​49171​.22781​
Chemosphere 134:181–191. https​://doi.org/10.1016/j.chemo​spher​ .bd
e.2015.04.020 141. Fiorentino JC, Enzweiler J, Angélica RS (2011) Geochemistry of mercury
122. Sharma R, Ramteke S, Patel KS, Kumar S, Sarangi B, Agrawal SG et al along a soil profile compared to other elements and to the parental
(2015) Contamination of lead and mercury in coal basin of India. J rock: evidence of external input. Water Air Soil Poll 221:63–75. https​://
Environ Prot 6:1430–1441. https​://doi.org/10.4236/jep.2015.61212​4 doi.org/10.1007/s1127​0-011-0769-x
123. Taylor M (2014) Soil quality monitoring in the Waikato Region 2012. 142. do Valle CM, Santana GP, Augusti R, Egreja Filho FB, Windmöller CC
WRC Technical Report 2013/49, Waikato Regional Council, Hamilton, (2005) Speciation and quantification of mercury in Oxisol, Ultisol, and
New Zealand. http://www.waika​toreg​ion.govt.nz/tr201​349/. Accessed 3 Spodosol from Amazon (Manaus, Brazil). Chemosphere 58:779–792.
Sept 2014 https​://doi.org/10.1016/j.chemo​spher​e.2004.09.005
124. Halbach K, Mikkelsen Ø, Berg T, Steinnes E (2017) The presence of 143. Armesto AG, Bibián-Núñez L, Campillo-Cora C, Pontevedra-Pombal X,
mercury and other trace metals in surface soils in the Norwegian Arctic. Arias-Estévez M, Nóvoa-Muñoz JC (2018) Total mercury distribution
Chemosphere 188:567–574. https​://doi.org/10.1016/j.chemo​spher​ among soil aggregate size fractions in a temperate forest podzol. Span
e.2017.09.012 J Soil Sci 8:190–202. https​://doi.org/10.3232/SJSS.2018.V8.N1.05
125. Bailey EA, Gray JE, Theodorakos PM (2002) Mercury in vegetation and 144. Yin R, Gu C, Feng X, Hurley JP, Krabbenhoft DP, Lepak RF et al (2016)
soils at abandoned mercury mines in southwestern Alaska. USA. Geo‑ Distribution and geochemical speciation of soil mercury in Wanshan
chem Explor Environ A 2:275–285. https​://doi.org/10.1144/1467-78730​ Hg mine: effects of cultivation. Geoderma 272:32–38. https​://doi.
2-032 org/10.1016/j.jes.2018.04.028
126. Higueras P, Oyarzun R, Biester H, Lillo J, Lorenzo S (2003) A first insight 145. Wang DY, Qing CL, Guo TY, Guo YJ (1997) Effects of humic acid on
into mercury distribution and speciation in soils from the Almadén transport and transformation of mercury in soil-plant systems. Water Air
mining district, Spain. J Geochem Explor 80(1):95–104. https​://doi. Soil Poll 95:35–43. https​://doi.org/10.1007/BF024​06154​
org/10.1016/S0375​-6742(03)00185​-7 146. Smith T, Pitts K, McGarvey JA, Summers AO (1998) Bacterial oxidation of
127. Gosar M, Šajn R, Biester H (2006) Binding of mercury in soils and mercury metal vapor, Hg(0). Appl Environ Microbiol 64:1328–1332
attic dust in the Idrija mercury mine area (Slovenia). Sci Total Environ 147. Colombo MJ, Ha J, Reinfelder JR, Barkay T, Yee N (2013) Anaerobic
369:150–162. https​://doi.org/10.1016/j.scito​tenv.2006.05.006 oxidation of Hg (0) and methylmercury formation by Desulfovibrio
128. Bernaus A, Gaona X, van Ree D, Valiente M (2006) Determination of desulfuricans ND132. Geochim Cosmochim Ac 112:166–177. https​://doi.
mercury in polluted soils surrounding a chlor-alkali plant: direct specia‑ org/10.1016/j.gca.2013.03.001
tion by X-ray absorption spectroscopy techniques and preliminary 148. Colombo MJ, Ha J, Reinfelder JR, Barkay T, Yee N (2014) Oxidation of
geochemical characterisation of the area. Anal Chim Acta 565:73–80. Hg(0) to Hg(II) by diverse anaerobic bacteria. Chem Geol 363:334–340.
https​://doi.org/10.1016/j.aca.2006.02.020 https​://doi.org/10.1016/j.chemg​eo.2013.11.020
129. Richardson JB, Moore L (2020) A tale of three cities: mercury in urban 149. Gustin MS, Ericksen JA, Schorran DE, Johnson DW, Lindberg SE, Cole‑
deciduous foliage and soils across land-uses in Poughkeepsie NY, Hart‑ man JS (2004) Application of controlled mesocosms for understanding
ford CT, and Springfield MA USA. Sci Total Environ 715:136869. https​:// mercury air- soil—plant exchange. Environ Sci Technol 38:6044–6050.
doi.org/10.1016/j.scito​tenv.2020.13686​9 https​://doi.org/10.1021/es048​7933
130. Cheng H, Li M, Zhao C, Li K, Peng M, Qin A, Cheng X (2014) Overview of 150. Fantozzi L, Ferrara R, Dini F, Tamburello L, Pirrone N, Sprovieri F (2013)
trace metals in the urban soil of 31 metropolises in China. J Geochem Study on the reduction of atmospheric mercury emissions from mine
Explor 139:31–52. https​://doi.org/10.1016/j.gexpl​o.2013.08.012 waste enriched soils through native grass cover in the Mt. Amiata
131. Tijhuis L, Brattli B, Sæther OM (2002) A geochemical survey of topsoil in region of Italy. Environ Res 125:69–74. https​://doi.org/10.1016/j.envre​
the city of Oslo, Norway. Environ Geochem Hlth 24:67–94. https​://doi. s.2013.02.004
org/10.1023/A:10139​79700​212 151. Mazur M, Mitchell CPJ, Eckley CS, Eggert SL, Kolka RK, Sebestyen SD
132. Manta DS, Angelone M, Bellanca A, Neri R, Sprovieri M (2002) Heavy et al (2014) Gaseous mercury fluxes from forest soils in response to
metals in urban soils: a case study from the city of Palermo (Sicily), forest harvesting intensity: a field manipulation experiment. Sci Total
Italy. Sci Total Environ 300:229–243. https​://doi.org/10.1016/S0048​ Environ 496:678–687. https​://doi.org/10.1016/j.scito​tenv.2014.06.058
-9697(02)00273​-5 152. Ericksen JA, Gustin MS, Schorran DE, Johnson DW, Lindberg SE,
133. Rodrigues S, Pereira ME, Duarte AC, Ajmone-Marsan F, Davidson CM, Coleman JS (2003) Accumulation of atmospheric mercury in forest
Grčman H et al (2006) Mercury in urban soils: a comparison of local foliage. Atmos Environ 37:1613–1622. https​://doi.org/10.1016/S1352​
spatial variability in six European cities. Sci Total Environ 368:926–936. -2310(03)00008​-6
https​://doi.org/10.1016/j.scito​tenv.2006.04.008 153. Leonard TL, Taylor GE Jr, Gustin MS, Fernandez GC (1998) Mercury and
134. Cheng H, Zhao C, Liu F, Yang K, Liu Y, Li M et al (2013) Mercury drop plants in contaminated soils: 1. Uptake, partitioning, and emission
trend in urban soils in Beijing, China, since 1987. J Geochem Explor to the atmosphere. Environ Toxicol Chem 17:2063–2071. https​://doi.
124:195–202. https​://doi.org/10.1016/j.gexpl​o.2012.09.007 org/10.1002/etc.56201​71024​
135. Ottesen RT, Birke M, Finne TE, Gosar M, Locutura J, Reimann C et al 154. Frescholtz TF, Gustin MS, Schorran DE, Fernandez GC (2003) Assessing
(2013) Mercury in European agricultural and grazing land soils. Appl the source of mercury in foliar tissue of quaking aspen. Environ Toxicol
Geoch 33:1–12. https​://doi.org/10.1016/j.apgeo​chem.2012.12.013 Chem 22:2114–2119. https​://doi.org/10.1002/etc.56202​20922​
136. Loska K, Wiechuła D, Korus I (2004) Metal contamination of farming soils 155. Assad M, Parelle J, Cazaux D, Gimbert F, Chalot M, Tatin-Froux F (2016)
affected by industry. Environ Int 30:159–165. https​://doi.org/10.1016/ Mercury uptake into poplar leaves. Chemosphere 146:1–7. https​://doi.
S0160​-4120(03)00157​-0 org/10.1016/j.chemo​spher​e.2015.11.103
137. Ahmadi M, Akhbarizadeh R, Haghighifard NJ, Barzegar G, Jorfi S (2019) 156. Fleck JA, Grigal DF, Nater EA (1999) Mercury uptake by trees: an
Geochemical determination and pollution assessment of heavy metals observational experiment. Water Air Soil Poll. 115:513–523. https​://doi.
in agricultural soils of south western of Iran. J Environ Health Sci Eng. org/10.1023/A:10051​94608​598
https​://doi.org/10.1007/s4020​1-019-00379​-6 157. Dago A, González I, Ariño C, Martínez-Coronado A, Higueras P, Díaz-
Cruz JM et al (2014) Evaluation of mercury stress in plants from the
Gworek et al. Environ Sci Eur (2020) 32:128 Page 18 of 19

Almadén mining district by analysis of phytochelatins and their Hg 177. Ahmed I, Sebastain A, Prasad MNV, Kirti PB (2019) Emerging trends
complexes. Environ Sci Technol 8:6256–6263. https​://doi.org/10.1021/ in transgenic technology for phytoremediation of toxic metals and
es405​619y metalloids. In: Prasad MNV (ed) Transgenic plant technology for
158. Laacouri A, Nater EA, Kolka RK (2013) Distribution and uptake dynamics remediation of toxic metals and metalloids. Academic Press, Cam‑
of mercury in leaves of common deciduous tree species in Minnesota, bridge, pp 43–62
USA. Environ Sci Technol 47:10462–10470. https​://doi.org/10.1021/ 178. Kim Y-O, Bae H-J, Cho E, Kang H (2017) Exogenous glutathione
es401​357z enhances mercury tolerance by inhibiting mercury entry into plant
159. Arnold J, Gustin MS, Weisberg PJ (2018) Evidence for nonstomatal cells. Front Plant Sci 8:683. https​://doi.org/10.3389/fpls.2017.00683​
uptake of Hg by aspen and translocation of Hg from foliage to tree 179. Cho UH, Park JO (2000) Mercury-induced oxidative stress in
rings in Austrian pine. Environ Sci Technol 52:1174–1182. https​://doi. tomato seedlings. Plant Sci 156:1–9. https​://doi.org/10.1016/S0168​
org/10.1021/acs.est.7b044​68 -9452(00)00227​-2
160. Moreno-Jiménez E, Gamarra R, Carpena-Ruiz RO, Millán R, Peñalosa 180. Israr M, Sahi SV (2006) Antioxidative responses to mercury in the cell
JM, Esteban E (2006) Mercury bioaccumulation and phytotoxicity cultures of Sesbania drummondii. Plant Physiol Bioch 44:590–595. https​
in two wild plant species of Almadén area. Chemosphere 63:1969– ://doi.org/10.1016/j.plaph​y.2006.09.021
1973. https​://doi.org/10.1016/j.chemo​spher​e.2005.09.043 181. Zhang T, Lu Q, Su C, Yang Y, Hu D, Xu Q (2017) Mercury induced oxida‑
161. Qian X, Wu Y, Zhou H, Xu X, Xu Z, Shang L et al (2018) Total mercury tive stress, DNA damage, and activation of antioxidative system and
and methylmercury accumulation in wild plants grown at waste‑ Hsp70 induction in duckweed (Lemna minor). Ecotoxicol Environ Saf
lands composed of mine tailings: insights into potential candidates 143:46–56. https​://doi.org/10.1016/j.ecoen​v.2017.04.058
for phytoremediation. Environ Pollut 239:757–767. https​://doi. 182. Gómez-Jacinto V, García-Barrera T, Gómez-Ariza JL, Garbayo-Nores
org/10.1016/j.envpo​l.2018.04.105 I, Vílchez-Lobato C (2015) Elucidation of the defence mechanism in
162. De Temmerman L, Waegeneers N, Claeys N, Roekens E (2009) Com‑ microalgae Chlorella sorokiniana under mercury exposure. Identifica‑
parison of concentrations of mercury in ambient air to its accumula‑ tion of Hg–phytochelatins. Chem Biol Interact 238:82–90. https​://doi.
tion by leafy vegetables: an important step in terrestrial food chain org/10.1016/j.cbi.2015.06.013
analysis. Environ Poll 157:1337–1341. https​://doi.org/10.1016/j.envpo​ 183. Wani AB, Chadar H, Wani AH, Singh S, Upadhyay N (2017) Salicylic acid
l.2008.11.035 to decrease plant stress. Environ Chem Lett 15:101–123. https​://doi.
163. Egler SG, Rodrigues-Filho S, Villas-Bôas RC, Beinhoff C (2006) Evalu‑ org/10.1007/s1031​1-016-0584-0
ation of mercury pollution in cultivated and wild plants from two 184. Kováčik J, Rotková G, Bujdoš M, Babula P, Peterková V, Matúš P (2017)
small communities of the Tapajós gold mining reserve, Pará State. Ascorbic acid protects Coccomyxa subellipsoidea against metal toxicity
Brazil. Sci Total Environ 368(1):424–433 through modulation of ROS/NO balance and metal uptake. J Hazar Mat
164. Svoboda L, Havlíčková B, Kalač P (2006) Contents of cadmium, 339:200–207. https​://doi.org/10.1016/j.jhazm​at.2017.06.035
mercury and lead in edible mushrooms growing in a historical silver- 185. Zhou X, Gao A, Zhang C, Xu W, Shi X (2017) Exogenous selenium allevi‑
mining area. Food Chem 96:580–585. https​://doi.org/10.1016/j.foodc​ ates mercury toxicity by preventing oxidative stress in rice (Oryza sativa)
hem.2005.03.012 seedlings. Int J Agric Biol 19:1593–1600
165. Falandysz J, Mędyk M, Treu R (2018) Bio-concentration potential 186. Seneviratne M, Rajakaruna N, Rizwan M, Madawala HMSP, Yong Sik
and associations of heavy metals in Amanita muscaria (L.) Lam. from Ok YS, Vithanage M (2019) Heavy metal-induced oxidative stress
northern regions of Poland. Environ Sci Pollut R. 25(25):25190–25206. on seed germination and seedling development: a critical review.
https​://doi.org/10.1007/s1135​6-018-2603-0 Environ Geochem Health 41:1813–1831. https​://doi.org/10.1007/s1065​
166. Millhollen AG, Gustin MS, Obrist D (2006) Foliar mercury accumu‑ 3-017-0005-8
lation and exchange for three tree species. Environ Sci Technol 187. Hasanuzzaman M, Hossain MA, da Silva JAT, Fujita M (2012) Plant
40:6001–6006. https​://doi.org/10.1021/es060​9194 response and tolerance to abiotic oxidative stress: antioxidant
167. Asati A, Pichhode M, Nikhil K (2016) Effect of heavy metals on plants: defense is a key factor. In: Venkateswarlu B, Shanker A, Shanker
an overview. Int J Appl Innov Eng Mgmt 5:56–66 C, Maheswari M (eds) Crop stress and its management: perspec‑
168. Ahammad SJ, Sumithra S, Senthilkumar P (2018) Mercury uptake and tives and strategies. Springer, Dordrecht, pp 261–315. https​://doi.
translocation by indigenous plants. Rasayan J Chem. 11:1–12. https​:// org/10.1007/978-94-007-2220-0
doi.org/10.7324/RJC.2018.11117​26 188. Cargnelutti D, Tabaldi LA, Spanevello RM, de Oliveira Jucoski G, Battisti
169. Safari F, Akramian M, Salehi-Arjmand H, Khadivi A (2019) Physiological V, Redin M et al (2006) Mercury toxicity induces oxidative stress in
and molecular mechanisms underlying salicylic acid-mitigated mer‑ growing cucumber seedlings. Chemosphere 65:999–1006. https​://doi.
cury toxicity in lemon balm (Melissa officinalis L.). Ecotoxicol Environ org/10.1016/j.chemo​spher​e.2006.03.037
Saf 183:109542. https​://doi.org/10.1016/j.ecoen​v.2019.10954​2 189. Patra M, Bhowmik N, Bandopadhyay B, Sharma A (2004) Comparison
170. Zhou ZS, Wang SJ, Yang ZM (2008) Biological detection and analysis of mercury, lead and arsenic with respect to genotoxic effects on plant
of mercury toxicity to alfalfa (Medicago sativa) plants. Chemosphere systems and the development of genetic tolerance. Environ Exp Bot
70:1500–1509. https​://doi.org/10.1016/j.chemo​spher​e.2007.08.028 52:199–223. https​://doi.org/10.1016/j.envex​pbot.2004.02.009
171. Azevedo R, Rodriguez E (2012) Phytotoxicity of mercury in plants: a 190. Nagajyoti PC, Lee KD, Sreekanth TVM (2010) Heavy metals, occurrence
review. J Bot. https​://doi.org/10.1155/2012/84861​4 and toxicity for plants: a review. Environ Chem Lett 8:199–216. https​://
172. Patra M, Sharma A (2000) Mercury toxicity in plants. Bot Rev 66:379– doi.org/10.1007/s1031​1-010-0297-8
422. https​://doi.org/10.1007/BF028​68923​ 191. Marrugo-Negrete J, Durango-Hernández J, Pinedo-Hernández J, Enam‑
173. Moreno FN, Anderson CW, Stewart RB, Robinson BH (2004) Phy‑ orado-Montes G, Díez S (2016) Mercury uptake and effects on growth
toremediation of mercury-contaminated mine tailings by induced in Jatropha curcas. J Environ Sci 48:120–125. https​://doi.org/10.1016/j.
plant-mercury accumulation. Environ Pra 6:65–175. https​://doi. jes.2015.10.036
org/10.1017/S1466​04660​40002​74 192. Teixeira DC, Lacerda LD, Silva-Filho EV (2018) Foliar mercury content
174. Obrist D (2007) Atmospheric mercury pollution due to losses of from tropical trees and its correlation with physiological parameters
terrestrial carbon pools? Biogeochemistry 85:119–123. https​://doi. in situ. Environ Pollut 242:1050–1057. https​://doi.org/10.1016/j.envpo​
org/10.1007/s1053​3-007-9108-0 l.2018.07.120
175. Lomonte C, Doronila AI, Gregory D, Baker AJ, Kolev SD (2010) Phy‑ 193. Manikandan R, Sahi SV, Venkatachalam P (2015) Impact assessment of
totoxicity of biosolids and screening of selected plant species with mercury accumulation and biochemical and molecular response of
potential for mercury phytoextraction. J Hazard Mater 173:494–501. Mentha arvensis: a potential hyperaccumulator plant. Sci World J. https​
https​://doi.org/10.1016/j.jhazm​at.2009.08.112 ://doi.org/10.1155/2015/71521​7
176. Fasani E, Manara A, Martini F, Furini A, DalCorso G (2018) The poten‑ 194. Mahbub KR, Krishnan K, Naidu R, Andrews S, Megharaj M (2017)
tial of genetic engineering of plants for the remediation of soils Mercury toxicity to terrestrial biota. Ecol Indic 74:451–462. https​://doi.
contaminated with heavy metals. Plant, Cell Environ 41:1201–1232. org/10.1016/j.atmos​env.2005.07.067
https​://doi.org/10.1111/pce.12963​
Gworek et al. Environ Sci Eur (2020) 32:128 Page 19 of 19

195. Zhou ZS, Huang SQ, Guo K, Mehta SK, Zhang PC, Zhi MY et al (2007) role of the organic matter. Sci Total Environ 361:163–178. https​://doi.
Metabolic adaptations to mercury-induced oxidative stress in roots of org/10.1016/j.scito​tenv.2005.05.023
Medicago sativa L. J Inorg Biochem 101:1–9. https​://doi.org/10.1016/j. 213. Grangeon S, Guédron S, Asta J, Sarret G, Charlet L (2012) Lichen and soil
jinor​gbio.2006.05.011 as indicators of an atmospheric mercury contamination in the vicinity
196. Bash JO, Miller DR, Meyer TH, Bresnahan PA (2004) Northeast United of a chlor-alkali plant (Grenoble, France). Ecol Indic 13:178–183. https​://
States and Southeast Canada natural mercury emissions estimated doi.org/10.1016/j.ecoli​nd.2011.05.024
with a surface emission model. Atmos Environ 38:5683–5692. https​:// 214. Ullrich SM, Ilyushchenko MA, Kamberov IM, Tanton TW (2007) Mercury
doi.org/10.1016/j.atmos​env.2004.05.058 contamination in the vicinity of a derelict chlor-alkali plant. Part I:
197. Rea AW, Lindberg SE, Scherbatskoy T, Keeler GJ (2002) Mercury accumu‑ sediment and water contamination of Lake Balkyldak and the River
lation in foliage over time in two northern mixed-hardwood forests. Irtysh. Sci Total Environ 381:1–16. https​://doi.org/10.1016/j.scito​
Water Air Soil Pollut 133:49–67. https​://doi.org/10.1023/A:10129​19731​ tenv.2007.02.033
598 215. Esbrí JM, López-Berdonces MA, Fernández-Calderón S, Higueras P, Díez
198. Magarelli G, Fostier AH (2005) Influence of deforestation on the mer‑ S (2015) Atmospheric mercury pollution around a chlor-alkali plant
cury air/soil exchange in the Negro River Basin, Amazon. Atmos Environ in Flix (NE Spain): an integrated analysis. Environ Sci Pollut R 22:4842–
39:7518–7528. https​://doi.org/10.1016/j.atmos​env.2005.07.067 4850. https​://doi.org/10.1007/s1135​6-014-3305-x
199. Selin MS, Lindberg SE (2005) Terrestrial Hg fluxes: is the next exchange 216. Biester H, Müller G, Schöler HF (2002) Binding and mobility of mercury
up, down, or neither? In: Pirrone N, Mahaffey KR (eds) Dynamics of in soils contaminated by emissions from chlor-alkali plants. Sci Total
mercury pollution on regional and global scales. Springer, Boston, pp Environ 284:191–203. https​://doi.org/10.1016/S0048​-9697(01)00885​-3
241–259. https​://doi.org/10.1007/0-387-24494​-8_11 217. Reis AT, Rodrigues SM, Araújo C, Coelho JP, Pereira E, Duarte AC (2009)
200. Gustin MS, Engle M, Ericksen J, Lyman S, Stamenkovic J, Xin M (2006) Mercury contamination in the vicinity of a chlor-alkali plant and poten‑
Mercury exchange between the atmosphere and low mercury contain‑ tial risks to local population. Sci Total Environ 407:2689–2700. https​://
ing substrates. Appl Geoch 21:1913–1923. https​://doi.org/10.1016/j. doi.org/10.1016/j.scito​tenv.2008.10.065
apgeo​chem.2006.08.007 218. Zheng N, Liu J, Wang Q, Liang Z (2011) Mercury contamination due
201. Selin NE, Jacob DJ, Yantosca RM, Strode S, Jaeglé L, Sunderland to zinc smelting and chlor-alkali production in NE China. Appl Geoch
EM (2008) Global 3D land, ocean, atmosphere model for mercury: 26(2):188–193. https​://doi.org/10.1016/j.apgeo​chem.2010.11.018
present day versus preindustrial cycles and anthropogenic enrich‑ 219. Song Z, Li P, Ding L, Li Z, Zhu W, He T, Feng X (2018) Environmental
ment factors for deposition. Global Biogeoch Cy 22:1–13. https​://doi. mercury pollution by an abandoned chlor-alkali plant in Southwest
org/10.1029/2007G​B0030​40 China. J Geochem Explor 194:81–87. https​://doi.org/10.1016/j.gexpl​
202. Gray JE, Theodorakos PM, Fey DL, Krabbenhoft DP (2015) Mercury con‑ o.2018.07.017
centrations and distribution in soil, water, mine waste leachates, and 220. Zhu W, Li Z, Li P, Yu B, Lin CJ, Sommar J et al (2018) Re-emission of
air in and around mercury mines in the Big Bend region, Texas. Environ legacy mercury from soil adjacent to closed point sources of Hg
Geochem Hlth. 37:35–48. https​://doi.org/10.1007/s1065​3-014-9628-1 emission. Environ Pollut 242:718–727. https​://doi.org/10.1016/j.envpo​
203. Wu G, Kang H, Zhang X, Shao H, Chu L, Ruan C (2010) A critical review l.2018.07.002
on the bio-removal of hazardous heavy metals from contaminated 221. Reis AT, Rodrigues SM, Davidson CM, Pereira E, Duarte AC (2010)
soils: issues, progress, eco-environmental concerns and opportunities. J Extractability and mobility of mercury from agricultural soils sur‑
Hazard Mater 174:1–8. https​://doi.org/10.1016/j.jhazm​at.2009.09.113 rounding industrial and mining contaminated areas. Chemosphere
204. Teršič T, Gosar M, Šajn R (2009) Impact of mining activities on soils and 81:1369–1377. https​://doi.org/10.1016/j.chemo​spher​e.2010.09.030
sediments at the historical mining area in Podljubelj, NW Slovenia. J 222. Liang J, Feng C, Zeng G, Gao X, Zhong M, Li X et al (2017) Spatial
Geochem Explor 100:1–10. https​://doi.org/10.1016/j.gexpl​o.2008.02.005 distribution and source identification of heavy metals in surface soils
205. Banásová V (1999) Vegetation on contaminated sites near an Hg mine in a typical coal mine city, Lianyuan, China. Environ Pollut 225:681–690.
and smelter. In: Ebinghaus RR, Turner RR, de Lacedra LD, Vasiljev O, Salo‑ https​://doi.org/10.1016/j.envpo​l.2017.03.057
mons W (eds) Mercury contaminated sites. Springer, Berlin, pp 321–335. 223. Wu Q, Wang S, Wang L, Liu F, Lin CJ, Zhang L et al (2014) Spatial
https​://doi.org/10.1007/978-3-662-03754​-6 distribution and accumulation of Hg in soil surrounding a Zn/Pb
206. Boente C, Albuquerque MTD, Gerassis S, Rodríguez-Valdés E, Gallego JR smelter. Sci Total Environ 496:668–677. https​://doi.org/10.1016/j.scito​
(2019) A coupled multivariate statistics, geostatistical and machine- tenv.2014.02.067
learning approach to address soil pollution in a prototypical Hg-mining 224. Wang D, Shi X, Wei S (2003) Accumulation and transformation of
site in a natural reserve. Chemosphere 218:767–777. https​://doi. atmospheric mercury in soil. Sci Total Environ 304:209–214. https​://doi.
org/10.1016/j.chemo​spher​e.2018.11.172 org/10.1016/S0048​-9697(02)00569​-7
207. Wahsha M, Maleci L, Bini C (2019) The impact of former mining activity 225. Navrátil T, Burns DA, Nováková T, Kaňa J, Rohovec J, Roll M et al (2018)
on soils and plants in the vicinity of an old mercury mine (Vallalta, Stability of mercury concentration measurements in archived soil and
Belluno, NE Italy). Geochem-Explor Envir Anal 19:171–175. https​://doi. peat samples. Chemosphere 208:707–711. https​://doi.org/10.1016/j.
org/10.1144/geoch​em201​8-040 chemo​spher​e.2018.06.033
208. Gemici Ü, Tarcan G, Somay AM, Akar T (2009) Factors controlling the 226. Khan MN, Wasim AA, Sarwar A, Rasheed MF (2011) Assessment of heavy
element distribution in farming soils and water around the abandoned metal toxicants in the roadside soil along the N-5, National Highway,
Halıköy mercury mine (Beydağ, Turkey). Appl Geoch 24:1908–1917. Pakistan. Environ Monit Assess 182:587–595. https​://doi.org/10.1007/
https​://doi.org/10.1016/j.apgeo​chem.2009.07.004 s1066​1-011-1899-8
209. Xiao R, Wang S, Li R, Wang JJ, Zhang Z (2017) Soil heavy metal con‑ 227. Beckers F, Awad YM, Beiyuan J, Abrigata J, Mothes S, Tsang DC et al
tamination and health risks associated with artisanal gold mining in (2019) Impact of biochar on mobilization, methylation, and ethylation
Tongguan, Shaanxi, China. Ecotox Environ Safe 141:17–24. https​://doi. of mercury under dynamic redox conditions in a contaminated flood‑
org/10.1016/j.ecoen​v.2017.03.002 plain soil. Environ Inter 127:276–290. https​://doi.org/10.1016/j.envin​
210. Yin D, He T, Yin R, Zeng L (2018) Effects of soil properties on production t.2019.03.040
and bioaccumulation of methylmercury in rice paddies at a mercury 228. Bonanno G, Giudice RL, Pavone P (2012) Trace element biomonitoring
mining area, China. J Environ Sci 68:194–205. https​://doi.org/10.1016/j. using mosses in urban areas affected by mud volcanoes around Mt.
jes.2018.04.028 Etna. The case of the Salinelle, Italy. Monit Assess 184:5181–5188. https​
211. Chiarantini L, Rimondi V, Benvenuti M, Beutel MW, Costagliola P, ://doi.org/10.1007/s1066​1-011-2332-z
Gonnelli C et al (2016) Black pine (Pinus nigra) barks as biomonitors of
airborne mercury pollution. Sci Total Environ 569:105–113. https​://doi.
org/10.1016/j.scito​tenv.2016.06.029 Publisher’s Note
212. Hissler C, Probst JL (2006) Impact of mercury atmospheric deposition Springer Nature remains neutral with regard to jurisdictional claims in pub‑
on soils and streams in a mountainous catchment (Vosges, France) lished maps and institutional affiliations.
polluted by chlor-alkali industrial activity: the important trapping