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Historical (1850–2010) mercury stable isotope inventory from anthropogenic sources to the atmosphere


Ruoyu Sun ,

CAS Key Laboratory of Crust-Mantle Materials and Environment, School of Earth and Space Sciences, University of Science and Technology of China, Hefei, Anhui, China; Observatoire Midi-Pyrénées, Laboratoire Géosciences Environnement Toulouse, CNRS/IRD/Université de Toulouse, France; Water Quality Centre, Trent University, Peterborough, Ontario, Canada de Toulouse, France, FR
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David G. Streets,

Energy Systems Division, Argonne National Laboratory, Argonne, Illinois, United States, US
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Hannah M. Horowitz,

Department of Earth and Planetary Sciences, Harvard University, Cambridge, Massachusetts, United States, US
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Helen M. Amos,

School of Engineering and Applied Science, Harvard University, Cambridge, Massachusetts, United States, US
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Guijian Liu,

CAS Key Laboratory of Crust-Mantle Materials and Environment, School of Earth and Space Sciences, University of Science and Technology of China, Hefei, Anhui, China, CN
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Vincent Perrot,

Observatoire Midi-Pyrénées, Laboratoire Géosciences Environnement Toulouse, CNRS/IRD/Université de Toulouse, France, FR
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Jean-Paul Toutain,

Observatoire Midi-Pyrénées, Laboratoire Géosciences Environnement Toulouse, CNRS/IRD/Université de Toulouse, France
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Holger Hintelmann,

Water Quality Centre, Trent University, Peterborough, Ontario, Canada, CA
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Elsie M. Sunderland,

School of Engineering and Applied Science, Harvard University, Cambridge, Massachusetts, United States, US
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Jeroen E. Sonke

Observatoire Midi-Pyrénées, Laboratoire Géosciences Environnement Toulouse, CNRS/IRD/Université de Toulouse, France, FR
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Mercury (Hg) stable isotopes provide a new tool to trace the biogeochemical cycle of Hg. An inventory of the isotopic composition of historical anthropogenic Hg emissions is important to understand sources and post-emission transformations of Hg. We build on existing global inventories of anthropogenic Hg emissions to the atmosphere to develop the first corresponding historical Hg isotope inventories for total Hg (THg) and three Hg species: gaseous elemental Hg (GEM), gaseous oxidized Hg (GOM) and particulate-bound Hg (PBM). We compile δ202Hg and Δ199Hg of major Hg emissions source materials. Where possible, δ202Hg and Δ199Hg values in emissions are corrected for the mass dependent Hg isotope fractionation during industrial processing. The framework and Hg isotope inventories can be updated and improved as new data become available. Simulated THg emissions from all sectors between 1850s and 2010s generally show an increasing trend (−1.1‰ to −0.7‰) for δ202Hg, and a stable trend (−0.02‰ to −0.04‰) for Δ199Hg. Δ200Hg are near-zero in source materials and therefore emissions. The δ202Hg trend generally reflects a shift of historically dominant Hg emissions from 19th century Hg mining and liquid Hg0 uses in Au/Ag refining to 20th century coal combustion and non-ferrous metal production. The historical δ202Hg and Δ199Hg curves of GEM closely follow those of THg. The δ202Hg curves of GOM and PBM show no trends. Δ199Hg values for both GOM and PBM decrease from the 1850s to 1950s by ∼0.1‰, and then gradually rebound towards the 2010s. Our updated δ202Hg values (−0.76 ± 0.11 ‰, 1SD, n=9) of bulk emissions from passively degassing volcanoes overlap with δ202Hg of present-day anthropogenic THg emissions.
Knowledge Domain: Atmospheric Science Earth & Environmental Science Ecology
How to Cite: Sun, R., Streets, D.G., Horowitz, H.M., Amos, H.M., Liu, G., Perrot, V., Toutain, J.-P., Hintelmann, H., Sunderland, E.M. and Sonke, J.E., 2016. Historical (1850–2010) mercury stable isotope inventory from anthropogenic sources to the atmosphere. Elem Sci Anth, 4, p.000091. DOI:
 Published on 12 Feb 2016
 Accepted on 21 Jan 2016            Submitted on 20 Oct 2015
Domain Editor-in-Chief: Joel D. Blum; Department of Earth & Environmental Sciences, University of Michigan, Ann Arbor, Michigan, United States
Guest Editor: Robert Mason; University of Connecticut, United States

1. Introduction

Mercury (Hg) is a toxic element. Transfer of Hg from Earth’s lithosphere to the atmosphere and other surface environments is continuous by natural processes, and has been accelerated by human activities dating back to antiquity (Goldwater, 1972; Nriagu, 1979). Natural atmospheric Hg emissions from volcanoes, crustal weathering and hydrothermal activity are thought to be 1–2 orders of magnitude smaller than modern anthropogenic Hg emissions (Amos et al., 2015; Bagnato et al., 2014; UNEP, 2013). Prior to the 1850s, anthropogenic Hg releases mainly came from primary Hg mining and use of Hg as amalgamation agent for silver extraction in the Spanish colonial Americas (Camargo, 2002; Hagan et al., 2011; Robins and Hagan, 2012). Following the industrial revolution, anthropogenic Hg releases further increased due to large-scale gold mining, non-ferrous metal production and combustion of fossil fuels (Streets et al., 2011). Streets et al. (2011) estimated that 215 Gg of cumulative anthropogenic Hg has been directly emitted into the atmosphere since the 1850s. This inventory was recently updated in Horowitz et al., (2014) by incorporating Hg releases from previously unquantified commercial Hg uses. The authors showed that an additional 540 Gg of Hg has been released to the atmosphere (20%), water (30%) and terrestrial (50%) reservoirs (Horowitz et al., 2014).

Released Hg will undergo complex redox reactions within and among atmospheric, terrestrial and aqueous reservoirs and is ultimately sequestered in the Earth’s lithosphere when it is incorporated in marine sediments (Amos et al., 2014; Andren and Nriagu, 1979; Mason et al., 1994). Substantial legacy Hg has accumulated in Earth’s surface reservoirs, significantly amplifying Hg loading to biogeochemically active pools (Amos et al., 2013). Natural Hg archives such as sediments, ice cores and peat cores (Faïn et al., 2009; Fitzgerald et al., 2004; Lamborg et al., 2002; Martínez-Cortizas et al., 1999; Schuster et al., 2002), and box models of global Hg cycling (Amos et al., 2013, 2015) suggest that present-day atmospheric Hg deposition has increased by a factor of 3–5 since the industrial period and by a factor of 17–27 since 3000 BC. Substantial debate exists on the historical Hg emission sources, e.g., Spanish-American mining in the 16th–18th century, North-American mining in the late 19th century, and other global change factors that have led to the increase in atmospheric Hg deposition (Amos et al., 2015; Engstrom et al., 2014; Krabbenhoft and Sunderland, 2013).

Hg stable isotope signatures have the potential to differentiate natural and anthropogenic Hg sources, and identify and quantify Hg transformations (Blum et al., 2014; Sonke, 2011; Sonke and Blum, 2013; Yin et al., 2014b). More than 10‰ variations in mass dependent Hg isotope fractionation (MDF, indicated by δ202Hg) and mass independent fractionation (MIF, indicated by Δ199Hg, Δ201Hg) of odd Hg isotopes have been documented in natural samples such as crustal rocks (Smith et al., 2008), cinnabar (Gehrke et al., 2011; Gray et al., 2013; Hintelmann and Lu, 2003; Smith et al., 2008; Stetson et al., 2009; Wiederhold et al., 2013), coal (Biswas et al., 2008; Sun et al., 2016; Yin et al., 2014a), and non-ferrous metals (Smith, 2010; Sonke et al., 2010; Yin et al., 2016). Different physical, chemical and biological Hg transformation processes induce characteristic MDF and MIF signs and magnitudes (Bergquist and Blum, 2007; Chandan et al., 2015; Ghosh et al., 2013; Jiskra et al., 2012; Kritee et al., 2013; Perrot et al., 2015; Rodríguez-González et al., 2009; Rose et al., 2015; Smith et al., 2015; Wiederhold et al., 2010; Zheng and Hintelmann, 2010a, 2010b). Historical Hg isotope composition of anthropogenic emissions can aid our understanding of the Hg source-receptor relationships and the redox cycling of Hg after it is emitted from sources. Such an inventory is the first step in incorporating Hg isotopes into state-of-the-science global Hg cycling models. This would in turn help constrain current poorly-known Hg fluxes and transformations such as volcanic Hg emissions, Hg wet/dry deposition, atmospheric Hg0 oxidation and HgII reduction, marine and terrestrial Hg re-emissions (Engstrom et al., 2014; Holmes et al., 2010; Huang and Gustin, 2015; Lindberg et al., 2007; Obrist et al., 2014; Pongprueksa et al., 2008; Pyle and Mather, 2003; Streets et al., 2011).

In addition to Hg emissions from different sources, accurate estimates of Hg isotope composition of anthropogenic emissions rely on two key factors: Hg isotope variation ranges of source Hg materials, and Hg isotope shifts between source Hg and emitted Hg. Over one decade of research has generated a large set of Hg isotope data in primary source materials (e.g., coal, cinnabar), and has advanced our understanding of several important Hg isotope fractionation processes during processing/combustion of source materials (e.g., coal combustion, cinnabar roasting) (Blum et al., 2014; Gray et al., 2013; Hintelmann and Lu, 2003; Sun et al., 2014a). Nevertheless, not all Hg emissions sources are covered and Hg isotope MDF during industrial processing is complex.

The goal of this work is to include Hg isotopes into the historically anthropogenic Hg emission inventory. We first compile the Hg isotope composition of source materials used by main human activities. Where possible, we estimate Hg isotope MDF during processing/combustion of source materials, following the framework of Sun et al. (2014a). A Monte Carlo approach is used to quantify the uncertainties of Hg isotope composition of anthropogenic emissions, following Streets et al. (2011). Finally, we provide new Hg isotope observations on bulk volcanic emissions that help better discriminate the natural and anthropogenic Hg emission end-members.

2. Methods and data

2.1 Methods description

The vast majority of published Hg isotope data use delta notation with 198Hg as the denominator relative to Hg standard solution NIST 3133:


MIF signatures are defined using capital delta notation as:


where XXX is Hg isotope mass number, (202Hg/198Hg) sample is the measured isotope ratio of the sample, and (202Hg/198Hg) NIST 3133 is the average isotope ratio of the bracketing NIST 3133 Hg standard solution during measurement. The mass dependent scaling factor βxxx is 0.2520 for 199Hg, 0.5024 for 200Hg, 0.7520 for 201Hg and 1.4930 for 204Hg (Blum and Bergquist, 2007). In this study, we use δ202Hg, Δ199Hg, and Δ200Hg as the tracers of Hg isotope MDF, odd Hg isotope MIF, and even Hg isotope MIF, respectively.

Based on historical (1850s–2010s, with decadal resolution) Hg inventories (Horowitz et al., 2014; Streets et al., 2011), emission sectors are broadly divided into two categories: ‘by-product’ and ‘intentional Hg use’, which respectively contribute 112 Gt and 213 Gt cumulative anthropogenic emissions to the atmosphere since 1850s. For all the individual emission sectors in each category, their atmospheric total Hg (THg) emissions are partitioned as gaseous elemental Hg (GEM), gaseous oxidized Hg (GOM), and particulate bound Hg (PBM). δ202Hgit and Δ199Hgit values of atmospheric speciated Hg emissions at a decadal year (t) are estimated as:


in which, the superscript i, s and r represents Hg species (i.e., GEM, GOM and PBM), sectors and regions, respectively. The Hg emissions in ‘by-product’ sectors (copper, zinc and lead smelting; iron and steel manufacturing; liquid Hg0 production; cement manufacturing; combustion of coal and oil; large-scale gold mining without Hg amalgamation) (Figure S1A) are divided into 17 world regions aggregated into 5 technological groupings based on levels of regional development. Emissions from the ‘intentional Hg use’ sectors (artisanal and small-scale gold mining (ASGM); silver mining; large-scale gold mining with Hg amalgamation; chlor-alkali production; and other 12 sectors that use liquid Hg0 in processes and products) (Figure S1B) only discriminate between developed and developing world regions. For all the ‘by-product’ sectors and selected ‘intentional Hg use’ sectors (large-scale gold mining, silver mining, ASGM and chlor-alkali production), their Hg species emission profiles are taken from Streets et al. (2011). For the remaining ‘intentional Hg use’ sectors of Horowitz et al. (2014), their Hg species emission profiles are calculated by weight-averaging emitted Hg species ratios at each emission stage of a complete life cycle. Mercury isotope composition in various source materials (δ202Hgs,rt and Δ199Hgs,rt) are summarized from published literature (Figure 1), and the shifts of MDF and MIF in emitted Hg species (ε202Hgi,s,rt and E199Hgi,s,rt) are estimated using the best-available information on Hg isotope fractionation during processing/combustion of source materials.

doi: 10.12952/journal.elementa.000091.f001.
Figure 1.  

Summary of δ202Hg and Δ199Hg in various source materials.

Box plots of published δ202Hg (A) and Δ199Hg (B) in various source materials. The horizontal lines at the bottom, middle and top of each boxplot are the lower quartile (below which 25% lowest values are found), median and upper quartile (above which 25% highest values are found), respectively. The box height (the difference between lower quartile and upper quartile) is defined as interquartile range (IQR). The data points either greater than the upper quartile + 1.5 IQR or less than the lower quartile-1.5 IQR are considered to be extreme values. Data sources: coal (Biswas et al., 2008; Lefticariu et al., 2011; Sherman et al., 2012; Sun et al., 2013, 2014a, 2014b; Yin et al., 2014a); oil (Blum et al., 2012); cinnabar (Cooke et al., 2013; Foucher et al., 2009; Gehrke et al., 2011; Hintelmann and Lu, 2003; Smith et al., 2005, 2008; Stetson et al., 2009; Wiederhold et al., 2013; Yin et al., 2013); Commercial liquid Hg0 (Blum and Bergquist, 2007; Estrade et al., 2009; Foucher et al., 2009; Laffont et al., 2011; Mead et al., 2013; Sonke et al., 2008); Zn ores (Smith, 2010; Sonke et al., 2010); Au ores (Smith, 2010); silicate rocks (Blum and Anbar, 2010; North, 2011; Sun et al., 2014b; Zhang et al., 2013, 2014); limestone (sampled from Huainan Basin, China, this study); volcanic gases (this study; Zambardi et al., 2009).

2.2 Source material mercury isotope composition and fractionation

Figure 1 summarizes the reported δ202Hg and Δ199Hg in source materials (e.g., fossil fuels, non-ferrous metal ores and crustal rocks). Δ200Hg in all source materials is insignificant (<±0.1‰). There is approximately 5.5‰ (−3.9‰ to 1.6‰) variation in δ202Hg in all source materials, with 80% of the values distributed between −1.84‰ and −0.06‰ (Figure 1). This large variation in δ202Hg of source materials reflects the wide variability of Hg sources and/or large MDF during biogeochemical processes including hydrothermalism, volcanism, diagenesis and metamorphosis. Δ199Hg in all source materials varies by 1.0‰ (−0.6 to 0.4‰) with 80% of the values distributed between −0.25‰ and 0.11‰. Most of the Hg in ores and rocks is derived from Earth’s crust and mantle, and has insignificant Δ199Hg (<±0.1‰) (Figure 1).

Most anthropogenic Hg emissions to the atmosphere are associated with high-temperature processing of source materials, e.g., combustion of fossil fuels, roasting/heating of gold/silver-Hg amalgam and cinnabar, roasting and smelting of ferrous and non-ferrous ores and concentrates. During the thermal processing of source materials, Hg in source materials is transformed into GEM, GOM and PBM in generated flue gases by a series of physicochemical reactions (oxidation, reduction, adsorption, diffusion etc.). These phase and redox transformations of Hg in high-temperature facilities are expected to induce significant Hg isotope MDF, but no photochemical MIF (Sherman et al., 2012; Sonke et al., 2010; Sun et al., 2013; Wiederhold et al., 2013). Therefore, we only address the Hg isotope MDF shifts of emitted Hg species relative to source materials. MDF and MIF of Hg isotopes during post-emission of volcanic and industrial plumes are also not further discussed here. Below we provide a detailed overview of Hg isotope composition in various source materials and associated Hg isotope MDF of emitted Hg species.

2.2.1. By-product sectors

Combustion of fossil fuels. More than 200 coal samples from historical and modern world coal-producing/consuming regions covering nearly all coal-forming periods and coal ranks have been analyzed for their Hg isotope composition (Biswas et al., 2008; Lefticariu et al., 2011; Sherman et al., 2012; Sun et al., 2013, 2014a, 2014b; Yin et al., 2014a). In contrast, only a limited number of oil sand samples from Athabasca, Canada have been reported (Blum et al., 2012). World coal deposits show large variations in both δ202Hg (−3.90‰ to 0.77‰) and Δ199Hg (−0.63‰ to 0.34‰), with an average value of −1.16±0.79‰ (1SD, n=216) and −0.11±0.18‰ (1SD, n=216), respectively. These coal samples are mainly from coal basins in China, former USSR, South Africa, Europe, India, Indonesia and Mongolia. Most of them are statistically distinguishable based on the means of δ202Hg, Δ199Hg or both (Sun et al., 2014a). The very limited Athabasca oil samples show moderate variation of δ202Hg (−1.68‰ to −1.22‰, n=5) and large variation of Δ199Hg (−0.39‰ to −0.02‰, n=5) (Blum et al., 2012) (Figure 1). We consider that more isotope data on oil are needed before this sector can be included in the isotope emission inventory.

Sun et al. (2014a) developed a speciated Hg isotope fractionation model for pulverized coal-fired boilers (the worldwide typical boiler type) by treating Hg isotope fractionation associated with Hg transformations in flue gases as two independent and consecutive Rayleigh fractionation systems: homogeneous gaseous oxidation of GEM to GOM, and heterogeneous gas-particle conversion of GOM to PBM. This model assumes Hg in coal is completely thermally reduced into GEM in the combustion zone of a boiler and Hg isotopes are only fractionated between the furnace outlet and inlet of air pollution control devices (APCD). Using a generalized GEM:GOM:PBM ratio of 0.15:0.52:0.33 in the flue gases before APCD, this model predicts a shift of 1.0±0.2‰ (1SD) for GEM, 0.05±0.2‰ (1SD) for GOM and −0.5±0.3‰ (1SD) for PBM, relative to δ202Hg of feed coal. However, the generalized pre-APCD GEM:GOM:PBM ratio of Sun et al. (2014a) does not differentiate the various sector/fuel/technology Hg emission scenarios that are characterized by different Hg species emission profiles. According to Streets et al. (2011), 69 out of 144 sector/fuel/technology combinations for combustion sources are designated to coal combustion, and all these combinations have their corresponding Hg species emission characteristics. Here, we improve the previous estimates on isotope composition of speciated Hg from coal combustion as follows: 1) APCDs are divided into five types: cyclone, scrubber, fiber filter (FF), electrostatic precipitator (ESP), flue gas desulfurization (FGD). It is assumed the former four types can only remove PBM, while the last type only removes GOM (Sun et al., 2014a). Each type of APCD is assumed to have a constant Hg removal efficiency (cyclone=1%, scrubber=6.5%, FF/ESP=30.6%, FGD=40%, Streets et al., 2011). 2) For the coal combustion sectors installed with APCDs, Hg species ratios in flue gases before APCDs are calculated based on emitted Hg species ratios and Hg removal efficiencies of APCDs. For the coal combustion sectors without APCDs, the emitted Hg species ratios are used directly to represent Hg species ratios before emission.

Ferrous (iron/steel) and non-ferrous metals (Cu, Pb, Zn, Au) production. In the present study, Hg emissions from large-scale Au production are quantified separately as ‘by-product’ (without Hg amalgamation) and ‘intentional Hg use’ (with Hg amalgamation) sectors, with the former from the trace Hg impurity in gold ores and the latter from the liquid Hg0 used for amalgamation. Up until now, only Hg isotope composition in zinc ores (sphalerite, from North America, China, South Africa, Spain, Australia and Congo, n=124) (Smith, 2010; Sonke et al., 2010; Yin et al., 2016) and gold ores (Carlin type, from USA, n=9) (Smith, 2010) have been reported. Due to lack of data, we assume Cu and Pb ores have the same Hg isotope composition as Zn ores based on their close geological association, commonly occurring together as hydrothermal sulfide ores. They possibly have common Hg sources. We further justify this by the similar median δ202Hg (and Δ199Hg) of well documented cinnabar (−0.61‰) and Zn ores (−0.47‰) (Table S1). No data are available for iron ore processing by the iron/steel sector, and we therefore omit this relatively small (<1% historically) sector in the Hg isotope inventory.

During high temperature smelting, roasting or sintering of ferrous and non-ferrous metal ores, most Hg will be volatilized into flue gases (Wu et al., 2012; Zhang et al., 2012). Sonke et al. (2010) suggested >99% of Hg in raw sphalerite ore is volatilized during smelting processes, and the refinery slags (δ202Hg=−0.24 ± 0.71‰, 1SD, n=4) are only slightly enriched in heavier Hg isotopes as compared to raw sphalerite ores (δ202Hg=−0.76 ± 1.25‰, 1SD, n=7). Based on mass balance, the volatilized bulk GEM fraction should therefore have nearly the same isotope composition as raw sphalerite ores. Subsequent partitioning of bulk GEM to GOM and PBM during cooling of flue gases and by-product removal by APCDs are expected to fractionate Hg isotopes, but no studies have been performed to determine the signs and magnitudes. At current stage, we assume all emitted Hg species conserve the Hg isotope composition of sphalerite ores.

Cement production. Following UNEP/AMAP (2013), we assume Hg emissions from cement production are exclusively from limestone, although other raw materials (e.g., clay, sandstone) contribute small amounts of Hg. We note that cement Hg emissions from combustion of fueled coals have been separately accounted for in industrial sectors of coal combustion. Here we measured Hg isotope composition in bioclastic limestone used in some Chinese areas for cement production. δ202Hg in these limestone samples varies from −2.47‰ to −0.37‰, with an average value of −1.64±0.61‰ (1SD, n=11) that is very similar to the one limestone sample (δ202Hg=−1.43±0.10‰, 2SD) reported by Wang et al. (2015). Limestone is significantly depleted in heavier Hg isotopes compare to other crustal rocks (Figure 1).

Cement production involves clinker formation at very high temperatures (up to 2000 °C in the cement kiln) by roasting a mixture of limestone and silicate rocks (Wang et al., 2014), which should theoretically induce minimal Hg isotope fractionation between flue gases and raw materials. In addition, for some cement plants with particulate matter recycling and raw material preheating, more than 90% of the input Hg is emitted into atmosphere (Wang et al., 2014). We therefore assume Hg emissions have the same Hg isotope composition as limestone in Figure 1. This assumption was also validated in a typical cement plant, Sichuan Province, China where the emitted THg was calculated to have nearly the same Hg isotope composition as raw materials (Wang et al., 2015).

Liquid Hg0 production. Hg-bearing ores (predominantly cinnabar, hexagonal α-HgS) have been extensively mined for liquid Hg0 production. It is estimated that nearly one million tons of liquid Hg0 has been produced over the last 500 years (Hylander and Meili, 2003; Streets et al., 2011). Hg isotope composition in cinnabar ores from primary Hg mining areas producing >90% world historical liquid Hg0 have been reported, including the world’s largest Spanish Almaden mine, the world’s second largest Slovenian Idrija mine, the largest three Hg production districts of the USA (California Coast Ranges, Nevada, and Texas), the Central-South America Hg mining districts, and the Hg production center in China (Foucher et al., 2009; Gehrke et al., 2011; Gray et al., 2013; Hintelmann and Lu, 2003; Smith et al., 2005, 2008; Stetson et al., 2009; Wiederhold et al., 2013; Yin et al., 2013).

During liquid Hg0 production, crushed cinnabar ores are sealed in retort or rotary furnaces and roasted at high temperatures (600–850 °C) to convert the solid ore HgS phases into Hg0 vapor, which is subsequently condensed as liquid Hg0. This conversion is not complete. Historically, <1% to 40% of Hg is estimated to have escaped into the ambient atmosphere as GEM, and <10% of Hg is retained in the residual calcine (Hylander and Meili, 2003, 2005; Streets et al., 2011). Significant MDF of Hg isotopes has been observed between cinnabar and its roasting products (calcine, escaped GEM, liquid Hg0) in the field and in controlled laboratory experiments (Stetson et al., 2009; Wiederhold et al., 2013; Yin et al., 2013). Following Gray et al, (2013), we use an isotope mass balance equation to constrain the δ202Hg values of emitted GEM during HgS roasting and Hg vapor condensation:

fcalcine × δ202Hgcalcine + fHgL0 × δ202HgHgL0 + fGEM × δ202HgGEM = δ202Hgcinnabar

In this equation, f’s represent the Hg mass fractions (%) of roasting products of cinnabar, summing to unity. The <10% residual Hg pool in calcine is commonly enriched in heavier Hg isotopes relative to cinnabar (Figure S2) (Stetson et al., 2009; Wiederhold et al., 2013; Yin et al., 2013). A comparison of mean δ202Hg between cinnabar ores and roasted calcine wastes at five individual Hg mines from US, Spain, and China shows a 0.1-1.9‰ (mean=0.8‰) positive shift of δ202Hg from cinnabar to calcine (Figure S2). This amounts to a <0.05‰ negative shift, on average, of δ202Hg in the generated bulk GEM that is subsequently cooled and condensed into liquid Hg0 at the condenser column/coils. Hg isotope fractionation at high temperatures is theoretically predicted to be minimal (Schauble, 2007). We therefore assume, as have others (Wiederhold et al., 2013; Yin et al., 2013), that the bulk GEM produced by cinnabar reduction approximately conserves the Hg isotope composition of cinnabar. Therefore, equation 5 can be approximated as:

fHgL0 × δ202HgHgL0 + fGEM × δ202HgGEM = δ202Hgcinnabar

During condensation of bulk GEM into liquid Hg0, the fraction of GEM unintentionally lost to the atmosphere may carry a different isotope signature relative to cinnabar and condensed liquid Hg0. Through a laboratory cinnabar retorting experiment, Gray et al, (2013) showed that on average, the condensed liquid Hg0 is shifted by +0.52±0.13‰ (1SD, n=3) and the GEM fraction lost is shifted by −0.79±0.44‰ (1SD, n=3) in δ202Hg relative to cinnabar ore (−0.36±0.10‰, 1SD, n=3). A direct comparison of δ202Hg in world cinnabar (−0.66±0.73‰, 1SD, n=210) (Cooke et al., 2013; Foucher et al., 2009; Gehrke et al., 2011; Hintelmann and Lu, 2003; Smith et al., 2005, 2008; Stetson et al., 2009; Wiederhold et al., 2013; Yin et al., 2013) and commercial liquid Hg0 (−0.38±0.34‰, 1SD, n=13) (Blum and Bergquist, 2007; Estrade et al., 2009; Foucher et al., 2009; Laffont et al., 2011; Mead et al., 2013; Sonke et al., 2008) also shows a slight enrichment of heavier Hg isotopes in liquid Hg0. It appears that the emitted GEM and condensed liquid Hg0 are complementary Hg pools, and their difference in δ202Hg is controlled by the extent to which Hg isotopes fractionate between GEM and condensed liquid Hg0. Hg0 condensation takes place from a saturated Hg0 vapor on a cooled surface. We therefore assume that these conditions are more similar to Hg0 liquid/vapor equilibrium Hg isotope fractionation than to kinetic isotope fractionation into a vacuum (Estrade et al., 2009). Equilibrium Hg0 liquid/vapor isotope fractionation enriches the vapor in the lighter Hg isotopes relative to the liquid (Estrade et al., 2009; Ghosh et al., 2013), which can be expressed as:

δ202HgHgL0  δ202HgGEM = 202Hgliqvap

The per mil fractionation factor ε202Hgliq-vap has an uncertainty range of 0.9–1.4‰ (mean=1.1‰), as experimentally determined by Estrade et al. (2009) and Ghosh et al. (2013). The observed δ202Hg difference (i.e., 1.31±0.31‰, 1SD) between GEM (reactant) and liquid Hg0 (product) during cinnabar retorting of Gray et al. (2013) is within the range of experimentally observed ε202Hgliq-vap. As discussed above, the δ202Hg difference between global cinnabar and commercial liquid Hg0 is ∼0.28‰. To achieve this shift, the Equations 6 and 7 would give a fGEM value of 20% when using ε202Hgliq-vap of 1.4‰. This is in the range of previously estimated fGEM (<1–40%) for the 20th century (Hylander and Meili, 2003; Streets et al., 2011). The fraction of emitted GEM (fGEM), can be calculated as a function of historical Hg emission factor, EFHg, of Hg production in each region:

EFHg = fGEMfHgL0 = fGEM1fGEM

Figure S3A shows EFHg in five world regions between 1900s and 2010s (Streets et al., 2011) and the calculated fractions of GEM emitted to the atmosphere (fGEM). Before 1900, EFHg and fGEM are identical to values in 1900 (Streets et al., 2011). Using a constant ε202Hgliq-vap of 1.4‰ and the calculated fGEM, we simulate the evolution of δ202Hg in emitted GEM and condensed liquid Hg0 in Figure S3B. It shows that the variations of δ202Hg in both produced liquid Hg0 and emitted GEM vary over a very limited range (<0.3‰), depending on the fGEM for each region. The predicted δ202Hg of liquid Hg0 is within the range measured for commercial liquid Hg0 (Blum and Bergquist, 2007; Estrade et al., 2009; Foucher et al., 2009; Laffont et al., 2011; Mead et al., 2013; Sonke et al., 2008).

Small Hg MIF (Δ199Hg up to 0.14‰, and Δ201Hg up to 0.09‰) due to the nuclear volume effect was observed during laboratory experiments of liquid Hg0 in equilibrium with its vapor (Estrade et al., 2009; Ghosh et al., 2013). This could potentially lead to small and opposite Hg MIF, < ±0.1‰, in emitted GEM and condensed liquid Hg0 during cinnabar retorting. Such small effects are close to measurement uncertainty, and were consequently not detected during the cinnabar retorting experiment of Gray et al. (2013). We therefore omit cinnabar retorting Hg MIF in the Hg isotope emission inventory, but suggest that further work in this direction is worthwhile.

2.2.2. Intentional Hg use sectors

Historically produced liquid Hg0 (cumulative 720 Gg in 1850–2010) was primarily used in Au/Ag mining in the late 19th to early 20th century, and then in commercial products and industrial processes (Horowitz et al., 2014; Streets et al., 2011). The Au/Ag mining uses liquid Hg0 to amalgamate Au/Ag ores. Mercury is evaporated by roasting the Au/Ag–Hg amalgam at temperatures >1000 °C to obtain Au/Ag (Robins and Hagan, 2012; Velásquez-López et al., 2010). This process involves a high-temperature, near-complete physical transformation of liquid Hg0 to GEM, and should induce insignificant Hg MDF (Laffont et al., 2011). We assume that the bulk Hg evaporated from Au/Ag-Hg amalgam conserves the Hg isotope composition of liquid Hg0 used for amalgamation. Liquid Hg0 and its derivatives are also widely used in commercial products (e.g., lamps, battery, medical devices, measuring devices), industrial processes (e.g., chlor-alkali production, vinyl chloride monomer, and pulp and paper) and agricultural fungicides and pesticides. Emissions of Hg to the atmosphere can occur during initial production, storage stages and final disposal processes (e.g., dumping, incineration) of Hg-containing products (Horowitz et al., 2014). Within one decade, most of the used liquid Hg0 in products and processes is thought to be volatilized to the atmosphere (Horowitz et al., 2014). Analogous to liquid Hg0 used in Au/Ag mining, the bulk Hg evaporated from these sectors also presumably conserve the Hg isotope composition of liquid Hg0. We use the measured isotope composition of commercial liquid Hg0 to represent those of intentional uses of liquid Hg0.

2.2.3. Uncertainty analysis

Following the previous uncertainty estimates of Hg emission inventories (Streets et al., 2011; Wu et al., 2010), we use the Monte Carlo stochastic simulator embedded in the Crystal BallTM software to quantify the uncertainties of Hg isotope inventory from anthropogenic emissions. Each set of the key input variables (isotope composition, emission-source isotope shifts) in individual decadal years and regions is best-fitted using probability distribution functions incorporated into the Crystal BallTM framework. When the number of data points of a variable is equal to or larger than 15, the distribution function that best describes the data characteristics of the variable is automatically fitted by Crystal BallTM. For the variables with limited data points, we use simple distribution functions (commonly triangle or normal) to describe these variables (Subramanyan et al., 2008; Wu et al., 2010). For example, both δ202Hg and Δ199Hg values for world coals (n=216) are automatically fitted by Weibull distributions. Due to limited coal samples in India (n=12) and OECD Europe (n=3), normal and triangle distribution functions are used, respectively, to represent their data scatter. A list of assumed distribution functions of Hg isotope composition of source materials with their main descriptive parameters are shown in Table S1. These distributions are truncated at both sides using minimum and maximum values of corresponding variables, and then are combined with previously estimated distributions of Hg emission inventories (Horowitz et al., 2014; Streets et al., 2011) to simulate the isotope inventories of anthropogenic THg and speciated Hg emissions. The number of stochastic simulation trials is set at 10 000 to obtain the reliable forecasted results that are expressed in the forms of statistical probability distributions (e.g., median value, P50) bounded by 10% (P10)–90% (P90) confidence interval).

Some parameters are not quantified for uncertainties at present. Examples include Hg isotope shifts of emitted GEM relative to raw Hg ores during liquid Hg0 production, the historical variations of specific APCDs removal efficiencies during coal combustion, and the ratios of GEM:GOM:PBM in emission sources. Currently, Hg isotope data are lacking for the oil combustion and iron/steel metallurgy sectors, as well as for the roasting and smelting of Cu and Pb ores. These aspects were therefore omitted from the Hg isotope emission estimates. The uncertainty associated with these omissions is limited as the cumulative Hg emissions from oil combustion, Fe/steel and Cu/Pb/Zn production only represent <8% of total anthropogenic Hg emissions since 1850, compared to >92% from the coal combustion, liquid Hg0 production and the uses of liquid Hg0 in Au/Ag mining and industrial processes/products.

3. Results and discussion

3.1 Estimated Hg isotope signatures of anthropogenic emissions

Figure 2 shows the estimated median values (P50) of δ202Hg and Δ199Hg for THg emissions from ‘by-product’ sectors and all sectors (‘by-product’ + ‘intentional Hg use’) between 1850s and 2010s, which are bounded by 80% confidence intervals (P10–P90). The corresponding speciated Hg (GEM, GOM and PBM) emissions are shown in Figures 3A–C. Table 1 lists the estimated δ202Hg and Δ199Hg for THg and speciated Hg emissions from ‘by-product’ sectors and all sectors in the 2010s. A full dataset of δ202Hg and Δ199Hg for THg, GEM, GOM, PBM and oxidized HgII (weighted mean of GOM and PBM) is tabulated in Table S2. Δ200Hg of all historical THg and speciated Hg emissions are <±0.1‰.

doi: 10.12952/journal.elementa.000091.f002.
Figure 2.  

Historical variations of δ202Hg and Δ199Hg in anthropogenic THg emissions.

Historical (1850s to 2010s) variations of δ202Hg and Δ199Hg for THg emitted from ‘by-product’ sectors and all sectors (‘by-product’ + ‘intentional Hg use’), bounding within 90% confidence levels (P10 to P90).


Table 1.

δ202Hg and Δ199Hg values of speciated Hg and THg emissions for both ‘by-product’ sectors and all sectors (‘by-product’ + ‘intentional Hg use’) in the 2010sa

By-product sectors All sectors
Mean -0.79 -0.84 -1.41 -0.88 -0.59 -0.77 -1.17 -0.69
P50 -0.80 -0.86 -1.40 -0.90 -0.58 -0.79 -1.15 -0.70
P10 -1.04 -1.43 -1.90 -1.28 -0.75 -1.27 -1.58 -0.94
P90 -0.54 -0.19 -0.95 -0.47 -0.43 -0.24 -0.82 -0.44
By-product sectors All sectors
Mean -0.04 -0.08 -0.06 -0.06 -0.02 -0.06 -0.04 -0.04
P50 -0.04 -0.08 -0.06 -0.06 -0.02 -0.07 -0.05 -0.04
P10 -0.07 -0.14 -0.12 -0.11 -0.04 -0.12 -0.09 -0.07
P90 -0.01 -0.01 0.01 -0.01 -0.01 -0.01 0.01 -0.01

3.1.1. By-product sectors

The isotope inventory of THg emissions from by-product sectors show an increasing trend in δ202Hg from −2.1‰ in 1850s to −0.9‰ in 2010s (Figure 2). This curve can be broadly divided into three phases: 1) 1850s-1930s with δ202Hg slightly increasing from −2.1‰ to −1.3‰, representing a transition of dominant Hg emission sectors from liquid Hg0 production and use to coal combustion and non-ferrous metals smelting; 2) 1930s-1960s with δ202Hg stabilized between −1.3‰ and −1.5‰, representing an insignificant variation in Hg emission amounts from different sectors; 3) 1960s–2010s with δ202Hg increasing from −1.5‰ and −0.9‰, representing a rapid increase of Hg emissions from coal combustion (Figure S1A and Table S1). δ202Hg values of GEM (−2.1‰ to −0.8%) (Figure 3A), GOM (−2.1‰ to −0.8‰) (Figure 3B) and PBM (−2.0‰ to −1.4‰) (Figure 3C) of by-product sectors from 1850s to 2010s generally follow the same increasing trends as THg. However, as compared to THg, PBM in recent decades is as large as 0.5‰ lower in δ202Hg. This is because δ202Hg of PBM predicted by the coal combustion MDF model is significantly lower than THg (Sun et al., 2014a). In contrast to δ202Hg, Δ199Hg of THg (Figure 2) decreases from −0.03‰ to −0.10‰ in the first 100 years of the industrial period (1850s-1950s) due to dominant Hg emissions from the combustion of European/North American coals that have significantly negative Δ199Hg (0.4‰ to −0.1‰, Table S1). Δ199Hg of THg then gradually rebounds to −0.06‰ in 2010s due to increasing Hg emissions from the combustion of Asian coals (e.g., China, India) and the smelting of non-ferrous metal ores which both are characterized by near-zero Δ199Hg (Figure S1A and Table S1). The temporal (1850s to 2010s) variation of Δ199Hg in GEM (−0.02‰ to −0.04% for 1850s-1950s, then stabilizes between −0.04% and −0.03% until 2010s) (Figure 3A), GOM (−0.01‰ to −0.13‰ for 1850s–1950s, then increases to −0.07% in 2010s) (Figure 3B) and PBM of by-product sectors (−0.08‰ to −0.16‰ for 1850s–1950s, then increases to −0.05% in 2010s) (Figure 3C) are basically similar to that of THg.

doi: 10.12952/journal.elementa.000091.f003.
Figure 3.  

Historical variations of δ202Hg and Δ199Hg in speciated Hg emissions.

Historical (1850s to 2010s) variations of δ202Hg and Δ199Hg for GEM (A), GOM (B) and PBM (C) emitted from ‘by-product’ sectors and all sectors (by-product’ + ‘intentional Hg use’), bounding within 90% confidence levels (P10 to P90).

3.1.2. All sectors

Adding Hg emissions from intentional uses of commercial liquid Hg0202Hg=0.4‰; Δ199Hg=0.02‰, Table S1) positively shifts the δ202Hg and Δ199Hg of by-product THg emissions by 0.2–1.2‰ and 0.00-0.05‰, respectively. From the 1850s to 2010s, this results in a variation of −1.1‰ to −0.7‰ in δ202Hg, and −0.02‰ to −0.04‰ in Δ199Hg for THg emissions from all sectors (Figure 2). The distinctly increasing δ202Hg and decreasing Δ199Hg trends for by-product sectors are not clearly seen for all sectors (Figure 2), due to large proportions (80%–50% for 1850s–2010s) of THg emissions from intentional uses of commercial liquid Hg0 (Figure S1). The δ202Hg (−1.1 to −0.6‰) and Δ199Hg (−0.02 to −0.01‰) curves of GEM (Figure 3A) from all sectors closely follow those of THg as most (∼90%) of Hg in the ‘intentional Hg use’ sectors is emitted as GEM. δ202Hg of GOM varies within a rather limited range of −1.1‰ to −0.8‰ without clear trend (Figure 3B), while δ202Hg of PBM shows a decreasing trend from −1.2‰ in the 1850s to −1.5‰ in the 1980s before rising to −1.1‰ in the 2010s (Figure 3C). Δ199Hg trends for both GOM (−0.02‰ to −0.10‰ for 1850s-1950s, then increases to −0.06% in 2010s) and PBM (−0.05‰ to −0.13‰ for 1850s-1950s, then increases to −0.04% in 2010s) from all sectors are similar to by-product sectors, as only ∼10% fraction of THg in ‘intentional Hg use’ sectors is emitted as GOM and PBM.

In addition to the uncertainties of previous Hg emission inventories (Horowitz et al., 2014; Streets et al., 2011), the expanded uncertainties of the Hg isotope inventories include variability in Hg isotope composition of source materials, and uncertainty in Hg isotope MDF between speciated Hg emissions and source materials. Sensitivity analysis of Hg isotope inventory shows that the largest contributors to the variance of δ202Hg of THg emissions from all sectors are the historical Hg emission inventory from 1850s to 1920s, δ202Hg of liquid Hg0 used in industrial processes/products between 1930s and 1970s, and δ202Hg of coal after 1980s (Figure S4A). In addition, Hg isotope MDF during coal combustion (ε202Hg_Coal) accounts for as much as 10-25% of the variance of δ202Hg since the 1980s. For the variance of Δ199Hg of THg emissions from all sectors, the largest contributors are Δ199Hg of liquid Hg0 uses before the 1940s, and Δ199Hg of coal after the 1950s (Figure S4B). Hg emissions from Cu/Zn/Pb smelting; cement production, and large-scale gold mining without Hg amalgamation only contribute ∼10% of variances in δ202Hg and Δ199Hg.

3.2 Comparison with Hg isotope composition of natural volcanic emissions

Primary natural Hg emissions to the atmosphere result from soil degassing and volcanic activity. The Hg flux from passively degassing volcanic activity is relatively well constrained at 76 ± 31 Mg y-1 (Bagnato et al., 2014). No Hg flux measurements exist for explosive volcanism. The Hg flux from explosive volcanic activity has been suggested to be larger, 600 Mg y-1 (Pyle and Mather, 2003) based on a crude extrapolation of registered Hg deposition in a single ice core archive (Schuster et al., 2002). Global box model constraints on primary natural Hg emissions, including both volcanic emissions and soil degassing, suggest that the sum of these is unlikely to be >300 Mg y-1 (Amos et al., 2015). We therefore suggest that the passive degassing volcanic flux of 76 ± 31 Mg y-1 is a realistic estimate of annual total volcanic Hg emissions.

A single study examined the Hg isotope composition of fumaroles at the passively degassing Vulcano, Italy (Zambardi et al., 2009). A mean δ202Hg for bulk fumarole THg emissions of −0.74 ± 0.18‰ (2SD, n=4), together with insignificant MIF was observed. Here we include additional observations from five fumaroles at the passively degassing Merapi and Papandayan volcanoes, Indonesia. Bulk fumarole sample collection, processing and Hg isotope analysis were done in 2007–2008 following Zambardi et al. (2009) and results are summarized in Table 2. We find overall similar Hg isotope composition, based on nine fumaroles at the three studied volcanoes, with a mean δ202Hg and Δ199Hg of −0.76 ± 0.22 ‰ (2SD, n=9) and 0.05 ± 0.06‰ (2SD, n=9), respectively. These values are very similar to those of geological cinnabar and silicate rocks (Figure 1). Compared to volcanic processes, hydrothermal processes can cause as large as 6‰ variation of δ202Hg in fossil and active hydrothermal systems due to vitalization of Hg0 vapor and active Hg redox reactions (Smith et al., 2005, 2008). The limited observations of δ202Hg and Δ199Hg of bulk volcanic THg emissions overlap our estimated present-day anthropogenic THg emissions (Table 1). This may limit discerning natural from anthropogenic Hg emissions on a global scale, depending on whether bulk volcanic Hg emissions are further fractionated within volcanic plumes. A single observation on Hg speciation in an aged fumarole plume at Vulcano (Italy) suggests that GEM (δ202Hg, −1.7‰) and PBM (δ202Hg, −0.1‰) can indeed be fractionated mass dependently relative to bulk THg emitted (δ202Hg, −0.7‰) (Zambardi et al., 2009). Hg isotope fractionation in volcanic plumes is however not sufficiently understood to predict the δ202Hg of volcanic GEM emissions at this point.


Table 2.

Summary of Hg isotope composition from bulk fumarole THg emissions at passively degassing volcanos

Fumarole ID Volcano name δ202Hg Δ199Hg Δ200Hg Δ201Hg References
C2 Merapi -0.81 0.08 0.02 0.07 This study
C4 Merapi -0.58 0.00 0.13 -0.03 This study
C5 Merapi -0.82 0.04 0.04 0.02 This study
C9 Merapi -0.82 0.09 0.06 0.09 This study
C193 Papandayan -0.83 0.07 -0.01 -0.02 This study
FA Vulcano -0.27 0.01 -0.04 0.04 (Zambardi et al., 2009)
F0 Vulcano -0.79 0.08 0.05 0.01 (Zambardi et al., 2009)
F5 Vulcano -1.09 -0.05 -0.03 -0.10 (Zambardi et al., 2009)
F11 Vulcano -0.79 0.13 0.03 0.02 (Zambardi et al., 2009)
Mean±2SD -0.76±0.22 0.05±0.06 0.03±0.05 0.01±0.06  

4. Concluding remarks

In this study, we have constructed isotope inventories of total Hg and speciated Hg (GEM, GOM and PBM) emissions for both ‘by-product’ and ‘intentional Hg use’ sectors with a decadal-resolution from the 1850s to 2010s. More work is needed to constrain Hg isotope composition of primary source materials and Hg isotope shifts in MDF and MIF between emitted Hg species and source materials. In particular, data on Fe, Cu and Pb ore concentrates are lacking. Further experimental and observational work on Hg isotope MDF during coal combustion, metal refining and liquid Hg0 use is needed to confirm the presented conceptual models.

High-precision Hg isotope measurement of total gaseous Hg (TGM, GEM+GOM) and precipitation Hg has been reported for several N-American and European regions (Chen et al., 2012; Demers et al., 2013, 2015; Donovan et al., 2013; Fu et al., 2014; Gratz et al., 2010; Sherman et al., 2010, 2012, 2015; Wang et al., 2015). Global TGM is enriched in the heavy isotopes (δ202Hg=0.51 ± 0.41 ‰, 1SD) (Demers et al., 2013, 2015; Fu et al., 2014; Gratz et al., 2010; Sherman et al., 2010) compared to our model estimated GEM emissions (Table 1). It has been shown that foliar uptake of GEM enriches foliage in the light Hg isotopes by −2.9‰ in δ202Hg, driving residual GEM to higher δ202Hg values (Demers et al., 2013). This process is possibly responsible for the δ202Hg shift between GEM emissions and global TGM observations. As we start to understand Hg isotope fractionation during atmospheric Hg transport and deposition, our high-resolution Hg isotope MDF and MIF inventories may help explain the observed isotope composition of atmospheric Hg and natural receptors and archives that directly receive atmospheric Hg. This would bridge the gap of our understanding on Hg isotope differences between sources and receptors. Furthermore, our inventories are expected to be embedded into state-of-the-science biogeochemical cycle models of Hg to understand the complex interplay between primary natural and anthropogenic Hg emissions and between primary and re-emitted natural and anthropogenic Hg.

Data accessibility statement

The Hg emission inventories of ‘by-product’ and ‘intentional Hg use’ sectors are publicly available from research group of Biogeochemistry of Global Contaminants in Harvard University: The stable Hg isotope (δ202Hg and Δ199Hg) emission inventories of ‘by-product’ and ‘intentional Hg use’ sectors for THg, GEM, GOM, PBM and oxidized HgII (weighted mean of GOM and PBM) are available in Table S2 of the supplemental material.


© 2016 Sun et al. This is an open-access article distributed under the terms of the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited.

Supplemental material

Figure S1.
Historical emission inventories of THg.
File Type: JPG
File Size: 2.18 MB

Figure S2.
Comparison of δ202Hg and Δ199Hg between cinnabar ores and retorted calcine.
File Type: JPG
File Size: 2.12 MB

Figure S3.
Historical variation of Hg emission factors, GEM emitted fractions, and δ202Hg in liquid Hg0 and emitted GEM.
File Type: JPG
File Size: 1.89 MB

Figure S4.
Variance analysis of δ202Hg and Δ199Hg in THg emissions.
File Type: JPG
File Size: 1.58 MB

Table S1.
The distribution functions and associated parameters for Hg isotope composition of source materials.
File Type: DOC
File Size: 0.08 MB

Table S2.
Historical (1850s to 2010s) δ202Hg and Δ199Hg variations for THg, GEM, GOM, PBM and oxidized HgII (weighted mean of GOM and PBM) emissions from anthropogenic sources.
File Type: XLSX
File Size: 0.03 MB


Contributed to conception and design: RS, JES, EMS, HMA

Contributed to acquisition of data: RS, DGS, HMH, VP, JPT

Contributed to analysis and interpretation of data: All authors

Drafted and/or revised the article: RS, JES, EMS, HMA

Approved the submitted version for publication: All authors

Competing interests

The authors do not have any competing interests.

Funding information

The study is supported by the National Basic Research Program of China (973 Program, 2014CB238903) and the China Postdoctoral Science (Special) Foundation (2014M551821; 2015T80668) and the Anhui Provincial Natural Science Foundation (1608085QD73). Additional supports are from the French Agence Nationale de Recherche (ANR-09-JCJC-0035-01) and European Research Council (ERC-2010-StG_20091028) to JES, and from the Natural Science and Engineering Research Council of Canada Discovery Grant to HH.


  1. Amos HM , Jacob DJ , Kocman D , Horowitz HM , Zhang Y , et al.. 2014. Global Biogeochemical Implications of Mercury Discharges from Rivers and Sediment Burial. Environ Sci Technol 48(16): 9514–9522. doi: 10.1021/es502134t.

  2. Amos HM, Jacob DJ, Streets DG, Sunderland EM. 2013. Legacy impacts of all-time anthropogenic emissions on the global mercury cycle. Global Biogeochem Cycles 27(2): 410–421. doi: 10.1002/gbc.20040.

  3. Amos HM , Sonke JE , Obrist D , Robins N , Hagan N , et al.. 2015. Observational and Modeling Constraints on Global Anthropogenic Enrichment of Mercury. Environ Sci Technol 49(7): 4036–4047. doi: 10.1021/es5058665.

  4. Andren MO, Nriagu JO. 1979. The global cycle of mercury, in NriaguJO ed. , ed., Biochemistry of mercury in the environment . Amsterdam: Elsevier: 1–15.

  5. Bagnato E , Tamburello G , Avard G , Martinez-Cruz M , Enrico M , et al. 2014. Mercury fluxes from volcanic and geothermal sources: An update. Geol Soc London, Spec Publ 410. doi: 10.1144/sp410.2.

  6. Bergquist BA, Blum JD. 2007. Mass-Dependent and -Independent Fractionation of Hg Isotopes by Photoreduction in Aquatic Systems. Science 318(5849): 417–420. doi: 10.1126/science.1148050.

  7. Biswas A, Blum JD, Bergquist BA, Keeler GJ, Xie Z. 2008. Natural Mercury Isotope Variation in Coal Deposits and Organic Soils. Environ Sci Technol 42(22): 8303–8309. doi: 10.1021/es801444b.

  8. Blum JD, Anbar AD. 2010. Mercury isotopes in the late Archean Mount McRae Shale. Geochim Cosmochim Acta 74(12): A98–A98.

  9. Blum JD, Bergquist BA. 2007. Reporting of variations in the natural isotopic composition of mercury. Anal Bioanal Chem 388(2): 353–359. doi: 10.1007/s00216-007-1236-9.

  10. Blum JD , Johnson MW , Gleason JD , Demers JD , Landis MS , et al. 2012. Mercury Concentration and Isotopic Composition of Epiphytic Tree Lichens in the Athabasca Oil Sands Region, in PercyKE ed. , ed., Developments in Environmental Science . Amsterdam: Elsevier: 373–390.

  11. Blum JD, Sherman LS, Johnson MW. 2014. Mercury Isotopes in Earth and Environmental Sciences. Annu Rev Earth Planet Sci 42(1): 249–269. doi: 10.1146/annurev-earth-050212-124107.

  12. Camargo JA. 2002. Contribution of Spanish-American silver mines (1570–1820) to the present high mercury concentrations in the global environment: A review. Chemosphere 48(1): 51–57. doi: 10.1016/s0045-6535(02)00047-4.

  13. Chandan P, Ghosh S, Bergquist BA. 2015. Mercury Isotope Fractionation during Aqueous Photoreduction of Monomethylmercury in the Presence of Dissolved Organic Matter. Environ Sci Technol 49(1): 259–267. doi: 10.1021/es5034553.

  14. Chen J, Hintelmann H, Feng X, Dimock B. 2012. Unusual fractionation of both odd and even mercury isotopes in precipitation from Peterborough, ON, Canada. Geochim Cosmochim Acta 90: 33–46. doi: 10.1016/j.gca.2012.05.005.

  15. Cooke CA , Hintelmann H , Ague JJ , Burger R , Biester H , et al.. 2013. Use and Legacy of Mercury in the Andes. Environ Sci Technol 47(9): 4181–4188. doi: 10.1021/es3048027.

  16. Demers JD, Blum JD, Zak DR. 2013. Mercury isotopes in a forested ecosystem: Implications for air-surface exchange dynamics and the global mercury cycle. Global Biogeochem Cycles 27(1): 222–238. doi: 10.1002/gbc.20021.

  17. Demers JD, Sherman LS, Blum JD, Marsik FJ, Dvonch JT. 2015. Coupling atmospheric mercury isotope ratios and meteorology to identify sources of mercury impacting a coastal urban-industrial region near Pensacola, Florida, USA. Global Biogeochem Cycles 29(10): 1689–1705. doi: 10.1002/2015GB005146.

  18. Donovan PM, Blum JD, Yee D, Gehrke GE, Singer MB. 2013. An isotopic record of mercury in San Francisco Bay sediment. Chem Geol 349–350: 87–98. doi: 10.1016/j.chemgeo.2013.04.017.

  19. Engstrom DR , Fitzgerald WF , Cooke CA , Lamborg CH , Drevnick PE , et al.. 2014. Atmospheric Hg Emissions from Preindustrial Gold and Silver Extraction in the Americas: A Reevaluation from Lake-Sediment Archives. Environ Sci Technol 48(12): 6533–6543. doi: 10.1021/es405558e.

  20. Estrade N, Carignan J, Sonke JE, Donard OFX. 2009. Mercury isotope fractionation during liquid-vapor evaporation experiments. Geochim Cosmochim Acta 73(10): 2693–2711. doi: 10.1016/j.gca.2009.01.024.

  21. Faïn X , Ferrari CP , Dommergue A , Albert MR , Battle M , et al.. 2009. Polar firn air reveals large-scale impact of anthropogenic mercury emissions during the 1970s. P Natl Acad Sci USA 106(38): 16114–16119. doi: 10.1073/pnas.0905117106.

  22. Feng X , Foucher D , Hintelmann H , Yan H , He T , et al.. 2010. Tracing Mercury Contamination Sources in Sediments Using Mercury Isotope Compositions. Environ Sci Technol 44(9): 3363–3368. doi: 10.1021/es9039488.

  23. Fitzgerald WF , Engstrom DR , Lamborg CH , Tseng C-M , Balcom PH , et al.. 2004. Modern and Historic Atmospheric Mercury Fluxes in Northern Alaska: Global Sources and Arctic Depletion. Environ Sci Technol 39(2): 557–568. doi: 10.1021/es049128x.

  24. Foucher D, Ogrinc N, Hintelmann H. 2009. Tracing Mercury Contamination from the Idrija Mining Region (Slovenia) to the Gulf of Trieste Using Hg Isotope Ratio Measurements. Environ Sci Technol 43(1): 33–39. doi: 10.1021/es801772b.

  25. Fu X, Heimburger L-E, Sonke JE. 2014. Collection of atmospheric gaseous mercury for stable isotope analysis using iodine- and chlorine-impregnated activated carbon traps. J Anal At Spectrom 29(5): 841–852. doi: 10.1039/C3JA50356A.

  26. Gehrke GE, Blum JD, Marvin-DiPasquale M. 2011. Sources of mercury to San Francisco Bay surface sediment as revealed by mercury stable isotopes. Geochim Cosmochim Acta 75(3): 691–705. doi: 10.1016/j.gca.2010.11.012.

  27. Ghosh S, Schauble EA, Lacrampe Couloume G, Blum JD, Bergquist BA. 2013. Estimation of nuclear volume dependent fractionation of mercury isotopes in equilibrium liquid–vapor evaporation experiments. Chem Geol 336: 5–12. doi: 10.1016/j.chemgeo.2012.01.008.

  28. Goldwater L. 1972. Mercury: A history of quicksilver . Baltimore, MD: York Press.

  29. Gratz LE, Keeler GJ, Blum JD, Sherman LS. 2010. Isotopic Composition and Fractionation of Mercury in Great Lakes Precipitation and Ambient Air. Environ Sci Technol 44(20): 7764–7770. doi: 10.1021/es100383w.

  30. Gray JE, Pribil MJ, Higueras PL. 2013. Mercury isotope fractionation during ore retorting in the Almadén mining district, Spain. Chem Geol 357: 150–157. doi: 10.1016/j.chemgeo.2013.08.036.

  31. Hagan N , Robins N , Hsu-Kim H , Halabi S , Morris M , et al.. 2011. Estimating historical atmospheric mercury concentrations from silver mining and their legacies in present-day surface soil in Potosi, Bolivia. Atmos Environ 45(40): 7619–7626. doi: 10.1016/j.atmosenv.2010.10.009.

  32. Hintelmann H, Lu S. 2003. High precision isotope ratio measurements of mercury isotopes in cinnabar ores using multi-collector inductively coupled plasma mass spectrometry. Analyst 128(6): 635–639. doi: 10.1039/B300451A.

  33. Holmes CD , Jacob DJ , Corbitt ES , Mao J , Yang X , et al.. 2010. Global atmospheric model for mercury including oxidation by bromine atoms. Atmos Chem Phys 10(24): 12037–12057. doi: 10.5194/acp-10-12037-2010.

  34. Horowitz HM, Jacob DJ, Amos HM, Streets DG, Sunderland EM. 2014. Historical Mercury Releases from Commercial Products: Global Environmental Implications. Environ Sci Technol 48(17): 10242–10250. doi: 10.1021/es501337j.

  35. Huang J, Gustin MS. 2015. Use of Passive Sampling Methods and Models to Understand Sources of Mercury Deposition to High Elevation Sites in the Western United States. Environ Sci Technol 49(1): 432–441. doi: 10.1021/es502836w.

  36. Hylander LD, Meili M. 2003. 500 years of mercury production: Global annual inventory by region until 2000 and associated emissions. Sci Total Environ 304(1–3): 13–27. doi: 10.1016/S0048-9697(02)00553-3.

  37. Hylander LD, Meili M. 2005. The Rise and Fall of Mercury: Converting a Resource to Refuse After 500 Years of Mining and Pollution. Crit Rev Env Sci Technol 35(1): 1–36. doi: 10.1080/10643380490492485.

  38. Jiskra M, Wiederhold JG, Bourdon B, Kretzschmar R. 2012. Solution Speciation Controls Mercury Isotope Fractionation of Hg(II) Sorption to Goethite. Environ Sci Technol 46(12): 6654–6662. doi: 10.1021/es3008112.

  39. Krabbenhoft DP, Sunderland EM. 2013. Global Change and Mercury. Science 341(6153): 1457–1458. doi: 10.1126/science.1242838.

  40. Kritee K, Blum JD, Reinfelder JR, Barkay T. 2013. Microbial stable isotope fractionation of mercury: A synthesis of present understanding and future directions. Chem Geol 336: 13–25. doi: 10.1016/j.chemgeo.2012.08.017.

  41. Laffont L , Sonke JE , Maurice L , Monrroy SL , Chincheros J , et al.. 2011. Hg Speciation and Stable Isotope Signatures in Human Hair As a Tracer for Dietary and Occupational Exposure to Mercury. Environ Sci Technol 45(23): 9910–9916. doi: 10.1021/es202353m.

  42. Lamborg CH , Fitzgerald WF , Damman AWH , Benoit JM , Balcom PH , et al.. 2002. Modern and historic atmospheric mercury fluxes in both hemispheres: Global and regional mercury cycling implications. Global Biogeochem Cycles 16(4): 1104. doi: 10.1029/2001gb001847.

  43. Lefticariu L, Blum JD, Gleason JD. 2011. Mercury Isotopic Evidence for Multiple Mercury Sources in Coal from the Illinois Basin. Environ Sci Technol 45(4): 1724–1729. doi: 10.1021/es102875n.

  44. Lindberg S , Bullock R , Ebinghaus R , Engstrom D , Feng X , et al.. 2007. A Synthesis of Progress and Uncertainties in Attributing the Sources of Mercury in Deposition. AMBIO: A Journal of the Human Environment 36(1): 19–33. doi: 10.1579/0044-7447(2007)36[19:asopau];2.

  45. Martínez-Cortizas A, Pontevedra-Pombal X, García-Rodeja E, Nóvoa-Muñoz JC, Shotyk W. 1999. Mercury in a Spanish Peat Bog: Archive of Climate Change and Atmospheric Metal Deposition. Science 284(5416): 939–942. doi: 10.1126/science.284.5416.939.

  46. Mason RP, Fitzgerald WF, Morel FMM. 1994. The biogeochemical cycling of element mercury - anthropogenic influences Geochim Cosmochim Acta 58(15): 3191–3198. doi: 10.1016/0016-7037(94)90046-9.

  47. Mead C, Lyons JR, Johnson TM, Anbar AD. 2013. Unique Hg Stable Isotope Signatures of Compact Fluorescent Lamp-Sourced Hg. Environ Sci Technol 47(6): 2542–2547. doi: 10.1021/es303940p.

  48. North S. 2011. Mercury concentrations and isotopic signature of the Alpine and Otago Schists, New Zealand. Ann Arbor, Michigan: University Michgen, Earth and Environmental Sciences.

  49. Nriagu JO. 1979. Production and uses of mercury, in NriaguJO ed. , ed., Biogeochemistry of Mercury in the Environment. Amsterdam: Elsevier: 23–39.

  50. Obrist D, Pokharel AK, Moore C. 2014. Vertical Profile Measurements of Soil Air Suggest Immobilization of Gaseous Elemental Mercury in Mineral Soil. Environ Sci Technol 48(4): 2242–2252. doi: 10.1021/es4048297.

  51. Perrot V , Bridou R , Pedrero Z , Guyoneaud R , Monperrus M , et al.. 2015. Identical Hg Isotope Mass Dependent Fractionation Signature during Methylation by Sulfate-Reducing Bacteria in Sulfate and Sulfate-Free Environment. Environ Sci Technol 49(3): 1365–1373. doi: 10.1021/es5033376.

  52. Pongprueksa P , Lin CJ , Lindberg SE , Jang C , Braverman T , et al.. 2008. Scientific uncertainties in atmospheric mercury models III: Boundary and initial conditions, model grid resolution, and Hg(II) reduction mechanism. Atmos Environ 42(8): 1828–1845. doi: 10.1016/j.atmosenv.2007.11.020.

  53. Pyle DM, Mather TA. 2003. The importance of volcanic emissions for the global atmospheric mercury cycle. Atmos Environ 37(36): 5115–5124. doi: 10.1016/j.atmosenv.2003.07.011.

  54. Robins NA, Hagan NA. 2012. Mercury Production and Use in Colonial Andean Silver Production: Emissions and Health Implications. Environ Health Perspect 120(5): 627–631. doi: 10.1289/ehp.1104192.

  55. Rodríguez-González P , Epov VN , Bridou R , Tessier E , Guyoneaud R , et al.. 2009. Species-Specific Stable Isotope Fractionation of Mercury during Hg(II) Methylation by an Anaerobic Bacteria (Desulfobulbus propionicus) under Dark Conditions. Environ Sci Technol 43(24): 9183–9188. doi: 10.1021/es902206j.

  56. Rose CH, Ghosh S, Blum JD, Bergquist BA. 2015. Effects of ultraviolet radiation on mercury isotope fractionation during photo-reduction for inorganic and organic mercury species. Chem Geol 405: 102–111. doi: 10.1016/j.chemgeo.2015.02.025.

  57. Schauble EA. 2007. Role of nuclear volume in driving equilibrium stable isotope fractionation of mercury, thallium, and other very heavy elements. Geochim Cosmochim Acta 71(9): 2170–2189. doi: 10.1016/j.gca.2007.02.004.

  58. Schuster PF , Krabbenhoft DP , Naftz DL , Cecil LD , Olson ML , et al.. 2002. Atmospheric Mercury Deposition during the Last 270 Years:  A Glacial Ice Core Record of Natural and Anthropogenic Sources. Environ Sci Technol 36(11): 2303–2310. doi: 10.1021/es0157503.

  59. Sherman LS, Blum JD, Dvonch JT, Gratz LE, Landis MS. 2015. The use of Pb, Sr, and Hg isotopes in Great Lakes precipitation as a tool for pollution source attribution. Sci Total Environ 502: 362–374. doi: 10.1016/j.scitotenv.2014.09.034.

  60. Sherman LS , Blum JD , Johnson KP , Keeler GJ , Barres JA , et al.. 2010. Mass-independent fractionation of mercury isotopes in Arctic snow driven by sunlight. Nat Geo 3(3): 173–177.

  61. Sherman LS, Blum JD, Keeler GJ, Demers JD, Dvonch JT. 2012. Investigation of Local Mercury Deposition from a Coal-Fired Power Plant Using Mercury Isotopes. Environ Sci Technol 46: 382–390. doi: 10.1021/es202793c.

  62. Smith CN. 2010. Isotopic geochemistry of mercury in active and fossil hydrothermal systems. Ann Arbor, Michigan: University of Michigan, Geological Sciences.

  63. Smith CN, Kesler SE, Blum JD, Rytuba JJ. 2008. Isotope geochemistry of mercury in source rocks, mineral deposits and spring deposits of the California Coast Ranges, USA. Earth Planet Sci Lett 269(3–4): 399–407. doi: 10.1016/j.epsl.2008.02.029.

  64. Smith CN, Kesler SE, Klaue B, Blum JD. 2005. Mercury isotope fractionation in fossil hydrothermal systems. Geology 33(10): 825–828. doi: 10.1130/g21863.1.

  65. Smith RS, Wiederhold JG, Kretzschmar R. 2015. Mercury Isotope Fractionation during Precipitation of Metacinnabar (β-HgS) and Montroydite (HgO). Environ Sci Technol 49(7): 4325–4334. doi: 10.1021/acs.est.5b00409.

  66. Sonke JE. 2011. A global model of mass independent mercury stable isotope fractionation. Geochim Cosmochim Acta 75(16): 4577–4590. doi: 10.1016/j.gca.2011.05.027.

  67. Sonke JE, Blum JD. 2013. Advances in mercury stable isotope biogeochemistry. Chem Geol 336: 1–4. doi: 10.1016/j.chemgeo.2012.10.035.

  68. Sonke JE , Schäfer J , Chmeleff J , Audry S , Blanc G , et al. 2010. Sedimentary mercury stable isotope records of atmospheric and riverine pollution from two major European heavy metal refineries. Chem Geol 279(3–4): 90–100. doi: 10.1016/j.chemgeo.2010.09.017.

  69. Sonke JE, Zambardi T, Toutain J-P. 2008. Indirect gold trap-MC-ICP-MS coupling for Hg stable isotope analysis using a syringe injection interface. J Anal At Spectrom 23(4): 569–573. doi: 10.1039/B718181G.

  70. Stetson SJ, Gray JE, Wanty RB, Macalady DL. 2009. Isotopic Variability of Mercury in Ore, Mine-Waste Calcine, and Leachates of Mine-Waste Calcine from Areas Mined for Mercury. Environ Sci Technol 43(19): 7331–7336. doi: 10.1021/es9006993.

  71. Streets DG , Devane MK , Lu Z , Bond TC , Sunderland EM , et al.. 2011. All-Time Releases of Mercury to the Atmosphere from Human Activities. Environ Sci Technol 45(24): 10485–10491. doi: 10.1021/es202765m.

  72. Subramanyan K, Wu Y, Diwekar U, Wang M. 2008. New stochastic simulation capability applied to the GREET model. Int J Life Cycle Assess 13(3): 278–285. doi: 10.1065/lca2007.07.354.

  73. Sun R , Heimbürger L-E , Sonke JE , Liu G , Amouroux D , et al.. 2013. Mercury stable isotope fractionation in six utility boilers of two large coal-fired power plants. Chem Geol 336: 103–111. doi: 10.1016/j.chemgeo.2012.10.055.

  74. Sun R , Sonke JE , Heimbürger L-E , Belkin HE , Liu G , et al.. 2014a. Mercury Stable Isotope Signatures of World Coal Deposits and Historical Coal Combustion Emissions. Environ Sci Technol 48(13): 7660–7668. doi: 10.1021/es501208a.

  75. Sun R, Sonke JE, Liu G. 2016. Biogeochemical controls on mercury stable isotope compositions of world coal deposits: A review. Earth-Science Rev 152: 1–13. doi: 10.1016/j.earscirev.2015.11.005.

  76. Sun R, Sonke JE, Liu G, Zheng L, Wu D. 2014b. Variations in the stable isotope composition of mercury in coal-bearing sequences: Indications for its provenance and geochemical processes. Int J Coal Geol 133: 13–23. doi: 10.1016/j.coal.2014.09.001.

  77. UNEP. 2013. Global Mercury Assessment 2013: Sources, Emissions, Releases and Environmental Transport. Geneva, Switzerland: UNEP Chemicals Branch.

  78. Velásquez-López PC, Veiga MM, Hall K. 2010. Mercury balance in amalgamation in artisanal and small-scale gold mining: Identifying strategies for reducing environmental pollution in Portovelo-Zaruma, Ecuador. J Clean Prod 18(3): 226–232. doi: 10.1016/j.jclepro.2009.10.010.

  79. Wang F , Wang S , Zhang L , Yang H , Wu Q , et al.. 2014. Mercury enrichment and its effects on atmospheric emissions in cement plants of China. Atmos Environ 92: 421–428. doi: 10.1016/j.atmosenv.2014.04.029.

  80. Wang Z , Chen J , Feng X , Hintelmann H , Yuan S , et al.. 2015. Mass-dependent and mass-independent fractionation of mercury isotopes in precipitation from Guiyang, SW China. C R Geosci . doi: 10.1016/j.crte.2015.02.006.

  81. Wiederhold JG , Cramer CJ , Daniel K , Infante I , Bourdon B , et al.. 2010. Equilibrium Mercury Isotope Fractionation between Dissolved Hg(II) Species and Thiol-Bound Hg. Environ Sci Technol 44(11): 4191–4197. doi: 10.1021/es100205t.

  82. Wiederhold JG , Smith RS , Siebner H , Jew AD , Brown GE , et al.. 2013. Mercury Isotope Signatures as Tracers for Hg Cycling at the New Idria Hg Mine. Environ Sci Technol 47(12): 6137–6145. doi: 10.1021/es305245z.

  83. Wu QR , Wang SX , Zhang L , Song JX , Yang H , et al.. 2012. Update of mercury emissions from China’s primary zinc, lead and copper smelters, 2000–2010. Atmos Chem Phys 12: 11153–11163. doi: 10.5194/acp-12-11153-2012.

  84. Wu Y, Streets DG, Wang SX, Hao JM. 2010. Uncertainties in estimating mercury emissions from coal-fired power plants in China. Atmos Chem Phys 9(6): 23565–23588. doi: 10.5194/acp-10-2937-2010.

  85. Yin R, Feng X, Chen J. 2014a. Mercury Stable Isotopic Compositions in Coals from Major Coal Producing Fields in China and Their Geochemical and Environmental Implications. Environ Sci Technol 48(10): 5565–5574. doi: 10.1021/es500322n.

  86. Yin R , Feng X , Hurley JP , Krabbenhoft DP , Lepak RF , et al.. 2016. Mercury Isotopes as Proxies to Identify Sources and Environmental Impacts of Mercury in Sphalerites. Sci Rep 6: 18686. doi: 10.1038/srep18686.

  87. Yin R, Feng X, Li X, Yu B, Du B. 2014b. Trends and advances in mercury stable isotopes as a geochemical tracer. Trends Environ Anal Chem 2: 1–10. doi: 10.1016/j.teac.2014.03.001.

  88. Yin R , Feng X , Wang J , Li P , Liu J , et al.. 2013. Mercury speciation and mercury isotope fractionation during ore roasting process and their implication to source identification of downstream sediment in the Wanshan mercury mining area, SW China. Chem Geol 336: 72–79. doi: 10.1016/j.chemgeo.2012.04.030.

  89. Zambardi T, Sonke JE, Toutain JP, Sortino F, Shinohara H. 2009. Mercury emissions and stable isotopic compositions at Vulcano Island (Italy). Earth Planet Sci Lett 277(1–2): 236–243. doi: 10.1016/j.epsl.2008.10.023.

  90. Zhang H , Yin R , Feng X , Sommar J , Anderson CWN , et al.. 2013. Atmospheric mercury inputs in montane soils increase with elevation: Evidence from mercury isotope signatures. Sci Rep 3(3322): 1–8. doi: 10.1038/srep03322.

  91. Zhang L , Liu Y , Guo L , Yang D , Fang Z , et al.. 2014. Isotope geochemistry of mercury and its relation to earthquake in the Wenchuan Earthquake Fault Scientific Drilling Project Hole-1 (WFSD-1). Tectonophysics 619–620: 79–85. doi: 10.1016/j.tecto.2013.08.025.

  92. Zhang L , Wang S , Wu Q , Meng Y , Yang H , et al.. 2012. Were mercury emission factors for Chinese non-ferrous metal smelters overestimated? Evidence from onsite measurements in six smelters. Environ Pollut 171: 109–117. doi: 10.1016/j.envpol.2012.07.036.

  93. Zheng W, Hintelmann H. 2010a. Isotope Fractionation of Mercury during Its Photochemical Reduction by Low-Molecular-Weight Organic Compounds. J Phys Chem A 114(12): 4246–4253. doi: 10.1021/jp9111348.

  94. Zheng W, Hintelmann H. 2010b. Nuclear Field Shift Effect in Isotope Fractionation of Mercury during Abiotic Reduction in the Absence of Light. J Phys Chem A 114(12): 4238–4245. doi: 10.1021/jp910353y.

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