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Sedimentary records of mercury stable isotopes in Lake Michigan


Runsheng Yin ,

Environmental Chemistry and Technology Program, University of Wisconsin-Madison, Madison, Wisconsin, United States; State Key Laboratory of Ore Deposit Geochemistry, Institute of Geochemistry, Chinese Academy of Sciences, Guiyang, China, CN
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Ryan F. Lepak,

Environmental Chemistry and Technology Program, University of Wisconsin-Madison, Madison, Wisconsin, United States, US
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David P. Krabbenhoft,

U.S. Geological Survey, Middleton, Wisconsin, United States, US
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James P. Hurley

Environmental Chemistry and Technology Program, University of Wisconsin-Madison, Madison, Wisconsin, United States; Department of Civil and Environmental Engineering, University of Wisconsin-Madison, Madison, Wisconsin, United States; University of Wisconsin Aquatic Sciences Center, Wisconsin, United States, US
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Mercury (Hg) concentrations and Hg isotopic composition were investigated in three sediment cores in Lake Michigan (LM). Two cores were collected from Green Bay, a region heavily impacted by Hg contamination and one core from an offshore region of LM absent of direct point source Hg. Historical trends of Hg influxes suggest increased Hg deposition began in the 1890s in Green Bay and in the early 1800’s in offshore LM. Recently deposited sediment reflecting more anthropogenic influence shows similar δ202Hg values (-1.0 to -0.5‰) for all three cores however, deep core sediments, reflecting pre-industrial eras, show much lower δ202Hg values (-1.7 to -1.2‰). Using a binary mixing model based on δ202Hg signatures, the proportion of anthropogenic Hg was estimated. Model output confirms that Green Bay is more contaminated by local point source than the offshore LM. An increase in positive Δ199Hg values (-0.02 to +0.27‰) was observed from inner Green Bay to the offshore of LM, which may indicate increased input of atmospheric Hg and decreased watershed inputs along this transect. Overall, this study suggests that sedimentary Hg isotopes maybe a useful tracer in understanding Hg sources and history of Hg contamination in large lakes.
Knowledge Domain: Earth & Environmental Science
How to Cite: Yin, R., Lepak, R.F., Krabbenhoft, D.P. and Hurley, J.P., 2016. Sedimentary records of mercury stable isotopes in Lake Michigan. Elem Sci Anth, 4, p.000086. DOI:
 Published on 27 Jan 2016
 Accepted on 17 Dec 2015            Submitted on 30 Sep 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 from both natural and anthropogenic sources can enter aquatic ecosystems from point sources, watersheds and atmospheric deposition (Mason and Sheu, 2002; Fitzgerald et al., 2007). Upon entering aquatic ecosystems, Hg may undergo complex reactions and transformations including oxidation, reduction, evasion to the atmosphere as gaseous Hg(0), binding to organic matter in the water column, and adsorbing to particles and depositing to sediments (Fitzgerald et al., 2007). Mercury can be methylated under anoxic and suboxic conditions by sulfate reducing bacteria (e.g. deep water column and sediments), producing methyl mercury (MeHg) (Benoit et al., 2003; Blum et al., 2013; Jonsson et al., 2014), a highly toxic organic compound that is easily bioaccumulated in aquatic food webs (U.S. EPA, 1997).

Sediment Hg profiles, coupled with high resolution age dating (e.g. 210Pb and 137Cs), has been broadly used to evaluate historical changes and to predict future trends in Hg deposition (Yang et al., 2010; Engstrom et al., 2014; Pizzuto, 2014). Historical reconstruction has shown an increase Hg deposition since global industrialization that began roughly in the mid-1800s. Estimates predict the global amount of Hg in the atmosphere has increased by a factor of 3 to 5 since the preindustrial era (Yang et al., 2010; Engstrom et al., 2014; Pizzuto, 2014). As Hg emission sources are numerous and emission pathways are highly variable (Pirrone et al., 2010), it is often challenging to differentiate and quantify the historical contribution of specific sources both on local and global scales.

Mercury stable isotope geochemistry allow for tracking ambient Hg sources and processes and the fate of Hg in the environment by utilizing the seven stable isotopes of Hg (196Hg, 198Hg, 199Hg, 200Hg, 201Hg, 202Hg and 204Hg) (Bergquist and Blum, 2009; Blum et al., 2014; Yin et al., 2014a). Technological advances in the past two decades [particularly multiple collector inductively coupled plasma mass spectrometry (MC-ICP-MS)] have enabled high precision analysis of Hg isotope ratios (Foucher and Hintelmann, 2006; Blum and Bergquist, 2007; Yin et al., 2010a). Laboratory experiments have observed both mass dependent fractionation (MDF) and mass independent fractionation (MIF) of Hg isotopes. MDF (reported as δ202Hg) occurs during various physical, chemical, and biological processes. The utilization of MDF improves our understanding of transport, transformation, and bioaccumulation of Hg (Bergquist and Blum, 2009; Yin et al., 2010b, 2014a; Sonke, 2011; Hintelmann, 2012; Kwon et al., 2012, 2013; Blum et al., 2014). MIF of odd Hg isotopes (reported as Δ199Hg and Δ201Hg), the combined result of nuclear volume effect (NVE) and magnetic isotope effect (MIE), can provide additional insight to Hg biogeochemical cycling (Schauble, 2007; Buchachenko et al., 2007). Laboratory experiments quantifying elemental Hg0 volatilization (Estrade et al., 2009; Ghosh et al., 2013), equilibrium Hg-thiol complexation (Wiederhold et al., 2010), and dark aqueous Hg(II) reduction (Zheng and Hintelmann, 2010a) have demonstrated that NVE results in Δ199Hg:Δ201Hg of ∼1.6. Whereas MIE, a byproduct of photochemical reactions of aqueous Hg species and most prevalent mechanism resulting in MIF, results in Δ199Hg:Δ201Hg ranging from 1.00 to 1.30 (Bergquist and Blum, 2007; Zheng and Hintelmann, 2009, 2010b; Rose et al., 2015; Chandan et al., 2014). Large variations (>10‰) of both δ202Hg and Δ199Hg values have been reported in environmentally relevant matrices such as rocks (Smith et al., 2008), mineral deposits (Hintelmann and Lu, 2003; Smith et al., 2005, 2008; Stetson et al., 2009; Sonke et al., 2010), coals (Biswas et al., 2008; Lefticariu et al., 2011; Sun et al., 2014; Yin et al., 2014b), soils (Biswas et al., 2008; Zhang et al., 2013; Feng et al., 2013), sediments (Gehrke et al., 2011; Foucher and Hintelmann, 2009; Feng et al., 2010; Liu et al., 2011; Bartov et al., 2013; Donovan et al., 2013; Ma et al., 2013; Yin et al., 2013a, 2015), water (Štrok et al., 2014), air (Demers et al., 2013; Rolison et al., 2013), precipitation (Gratz et al., 2010; Chen et al., 2012; Demers et al., 2013; Sherman et al., 2011, 2015) and biological samples (Bergquist and Blum, 2007; Blum et al., 2013; Demers et al., 2013; Yin et al., 2013b). With well-defined source signatures, researchers may estimate the relative contribution of sources of Hg through the utilization of binary and/or triple mixing models (Foucher and Hintelmann, 2009; Feng et al., 2010; Liu et al., 2011; Bartov et al., 2013; Donovan et al., 2013; Ma et al., 2013; Yin et al., 2013a; Lepak et al., 2016).

Determination of isotopic Hg signatures in sediment profiles have proven useful to better understand historical Hg loading to a variety of systems (Jackson et al., 2004; Feng et al., 2010; Sonke et al., 2010; Cooke et al., 2013; Donovan et al., 2013; Ma et al., 2013; Mil-Homens et al., 2013; Gray et al., 2013, 2015; Balogh et al., 2015). However, this task may be more complicated when attempting historical reconstruction for large lakes receiving loads from many tributaries with varied source signatures, contamination histories, and subject to significant changes in nutrient loading and benthos re-engineering by invasive species (Hecky et al., 2004; Rossmann, 2009). Source signatures of Hg isotopes may be altered following their release into the environment through continuous processing by various external physical, chemical and biological influences (Jackson, 2013; Donovan et al., 2014; Yin et al., 2015). For these reasons, the usefulness of Hg isotopes as tracers in large lakes is still unclear and therefore warrants further investigation.

Lake Michigan (LM), the fifth largest freshwater lake in the world, has received Hg loading from local industry through tributaries and from atmospheric deposition (regional to global) (Hartig and Law, 1994; Mason and Sullivan, 1997; Landis and Keeler, 2002; Hurley et al., 1998a, 1998b; Jeremiason et al., 2009; Rossmann, 2002, 2009; Rossmann and Edgington, 2000). Most point-source discharges have been eliminated as a result of the Clean Water Act (CWA) in the early 1970s (Adler et al., 1993), however residual contamination exists in rivers and nearshore sediments (Hurley et al. 1998a; Rossmann, 2002). In our study, both concentrations and isotopic Hg composition were investigated in sediment cores collected from a heavily impacted embayment of LM (Green Bay), and an offshore depositional basin of northern LM, not influenced by direct point sources. The objectives of this study were to: (1) reconstruct rates of input of Hg to northern LM and Green Bay; (2) compare sources of Hg to these depositional zones. Utilizing Hg stable isotopes allows us to better understand sources and processing pathways that lead to distinct depositional patterns in contrasting sedimentary zones of LM.

2. Materials and methods

2.1 Study area and sampling

LM (Figure 1), the third largest of the Laurentian Great Lakes, is the only Great Lake located entirely within the United States. Atmospheric deposition has been reported to be the primary pathway for Hg input to LM, contributing approximately 84% of total annual input in 1994 to 1995 (Landis and Keeler, 2002; U.S. EPA, 2004). However on a local scale, atmospheric Hg may not be the primary source of Hg especially in regions such as bays and nearshore zones influenced by rivers contaminated with Hg. The Fox River, for example, has a long history of industrialization since the first paper mill was built in 1865, and has been listed as an Area of Concern by the USEPA (Christensen and Chien, 1981; Hermanson et al., 1991; Hartig and Law, 1994). Mercury has been used at during paper production and there are numerous pathways for Hg to enter the river, including direct wastewater discharge (Wisconsin Department of Natural Resources, 1997). Water inputs from the Fox River to Green Bay account for one-third of the entire LM drainage basin (Christensen and Chien, 1981) and of the eleven major tributaries (comprising approximately 55% of the LM drainage basin) selected to assess Hg inputs to LM (Hurley et al., 1998a; 1998b), the Fox River contained the highest mean Hg concentrations (Fox River: 27.9 ng L-1, >90% in the particulate phase; other tributaries: 1.05 to 10.3 ng L-1). Elevated THg concentrations (up to 7400 ng g-1) were measured in the sediments of the Fox River (Redman, 1993) and Hurley et al. (1998a) measured a mean of 970 ng g-1 THg and a maximum of 4190 ng g-1 on particles collected at the river mouth in 1994.

doi: 10.12952/journal.elementa.000086.f001.
Figure 1.  

Study area and sampling locations.

White markers indicate sites that were studied previously forage dating and mass sedimentation rates [Station 21 and 48: Klump et al. (1997); LM 41: Song et al. (2005)]. Colored markers indicate sites within our study; marker remain site specific throughout the figures.

In the fall 2013, sediment cores were collected from three sites using a custom-designed polycarbonate tube multicorer on board the USEPA R/V Lake Guardian. Sites MI-116 and MI-50 are located in depositional zones of Green Bay, and site MI-112 is located in the offshore region of the Northern Basin (Figure 1). Four cores were collected at each site with polycarbonate core barrels pre-cleaned following method by Hammerschmidt et al. (2011). One of the four cores at MI-50 and MI-112 was field sectioned using a clean polycarbonate slicer at 1.0 cm intervals from the surface to 5 cm, and then at 2.0 cm intervals to the base of the core. The core from MI-116 was frozen in the field, delivered to USGS Wisconsin Mercury Research Lab (WMRL), and then sectioned at 0.5 cm intervals from the surface to the base of the core using a stainless steel band saw. All samples were freeze-dried and homogenized by a ball mill prior to THg and Hg isotopic analyses. Moisture content (water %) and loss on ignition of carbon for each slice was measured (Table S1, Supplemental Information Text S1).

2.2 Total mercury concentration and mercury isotopic composition analysis

THg in sediments was analyzed at the USGS-Wisconsin Mercury Research Laboratory with a direct combustion system (Nippon MA-2) and atomic absorbance detection, based on U.S. EPA Method 7473 (SW-846). The method detection limit was 1.4 ng g-1 dry weight. Recoveries of standard reference material (IAEA SL 1) were within 90 to 100%, and the coefficients of variation of triplicate analyses were less than 10%. Between 0.025 and 0.5 g of ground sample was digested (95 °C, 1 hour) in a 5 mL aqua regia (HCl:HNO3 = 3:1, v:v). Certified reference materials (NIST-2711 and MESS-1) and a secondary solution (UM-Almadén; in aqua regia) were similarly prepared in 10% of the sample count. Isotopic compositions of sample digests were measured on a Neptune Plus MC-ICP-MS, located at the Wisconsin State Laboratory of Hygiene (WSLH). Details of the method used for MC-ICP-MS analysis are given in Supplemental Information Text S1. Following the convention recommended by Blum and Bergquist (2007), Hg-MDF is expressed in δ202Hg notation in units of permil (‰) referenced to the NIST-3133 Hg standard (analyzed before and after each sample):

δ202Hg() = [(202Hg/198Hgsample)/(202Hg/198Hgstandard)1]×1000

Hg-MIF is reported in Δ notation (ΔxxxHg), and it describes difference between the measured δxxxHg and the theoretically predicted δxxxHg value, in units of permil (‰), using the following formula:

ΔxxxHg  δxxxHg  δ202Hg×β

β is equal to 0.2520 for 199Hg, 0.5024 for 200Hg, and 0.7520 for 201Hg (Blum and Bergquist, 2007).

UM-Almadén solutions at varying concentrations (0.3, 0.5 and 1 ng g-1), matching specific sample concentrations of Hg, were measured in every 10 samples. Data uncertainties reported in this study reflect the larger values of either the external precision of the replication of the UM-Almadén or the measurement uncertainty of the sample. The overall average and uncertainty of the UM-Almadén measurements (δ202Hg: -0.52±0.09‰; Δ199Hg: -0.01±0.05‰; Δ200Hg: 0.00±0.03‰; Δ201Hg: 0.00±0.05‰, 2σ, n = 9) agreed well with previous studies (Blum and Bergquist, 2007). Measurements of UM-Almadén with different Hg concentrations showed no statistical differences in isotopic composition (Table S3). Measurements of replicate digests of NIST-2711 (δ202Hg: -0.13±0.11‰; Δ199Hg: -0.16±0.09‰; Δ200Hg: -0.02±0.03‰; Δ201Hg: -0.16±0.04‰, 2σ, n = 3) and MESS-1 (δ202Hg -1.85±0.10‰; Δ199Hg 0.01±0.06‰; Δ200Hg: 0.01±0.03‰; Δ201Hg 0.02±0.05‰; 2σ, n = 3) also agreed well with previous studies (Biswas et al., 2008; Donovan et al., 2013; Yin et al., 2014b).

3. Results and discussion

3.1 Sediment age calculation

Spatial variation in sedimentation rate has been reported in both Green Bay and offshore of LM. These studies suggest the sedimentation in a particular site is generally constant, based upon log-linear relationships between excess 210Pb activities and the accumulated masses (Klump et al., 1997; Song et al., 2005; Zhu and Hits, 2005; Rossmann, 2002, 2009; Rossmann and Edgington, 2000). Based on the assumption of constant mass sedimentation (g cm-2 yr-1) and dry density measurements (g cm-3) interpolated from loss on ignition and wet density of the sediment core intervals, the dry deposition rate (cm yr-1) of each interval was calculated (Kadlec and Robbins, 1984). With this information, the sediment ages were estimated using the following:

TiN = Ti0(diN*0.5)(RiρiN)Σ1NdiN1(RiρiN1)

where i represents the name of each core (i: MI-116, MI-50 and MI 112); N represents the number of slices removed from the surface slice of each core [e.g., MI-116 (1 ≤N≤ 58), MI-50 (1 ≤N≤ 17) and MI 112 (1 ≤N≤ 13)]; TiN, diN and ρiN represent the age, thickness (cm) and density (g cm-3) of each slice (N) in each core (i) (summarized in Table S1), respectively. Ti0 and Ri represent the collection time (fall 2013) and the mass sedimentation rate (g cm-2 yr-1) of each core. The mass sedimentation rates used in this study for sites MI-116 and MI-50 were previously reported by Klump et al. (1997) as station 21 (0.052 g cm-2 yr-1) and station 48 (0.022 g cm-2 yr-1), respectively (Figure 1). MI-112 is located within the Algoma central basin, with a mean sedimentation rate of 0.040 g cm-2 yr-1 reported for this basin (LM 41; Song et al., 2005). Based on equation 3, the range of dates in cores MI-116, MI-50 and MI 112 were estimated to be from 1873 to 2013, 1799 to 2012 and 1734 to 2012, respectively (Figure 2, Table S1).

doi: 10.12952/journal.elementa.000086.f002.
Figure 2.  

THg (A, dry weight), Hg influx (B), δ202Hg (C), Δ199Hg (D) and Δ200Hg (E) in sediment profiles of Lake Michigan.

Gray solid lines represent measurement uncertainties.

3.2 Historical records of total mercury concentrations and mercury influxes

THg concentration profiles of the three cores show elevated THg in the uppermost sections (Figure 2A), consistent with lake-wide cores by Rossmann (2009). The oldest slices of MI-116, MI-50, and MI-112 have THg concentrations of 36, 35, and 17 ng g-1, respectively. This is comparable with the geochemical background of Hg in sediments of Great Lakes (∼25 ng g-1) (Marvin et al., 2004). The greatest THg concentrations are in recently deposited surface layers of the 1970’s (Figure 2A). Influxes of Hg (mass sedimentation rate × THg) were calculated to account for the differences in sedimentation rate in each of the sediment cores (Figure 2B). Fluxes to sediment are comparable with that reported by previous studies (Pirrone et al., 1998; Rossmann and Edgington, 2000; Rossmann, 2002). Influx of Hg in MI-112 in 1998 was 6.0 ng cm-2 yr-1, similar to that reported for Algoma Central (5.2 ng cm-2 yr-1) from 1994 to 1996 (Rossmann, 2002). Influxes of Hg in the oldest sections of sediment cores are comparable among deposition regions: MI-116 (1.8 ng cm-2 yr-1); MI-50 (0.8 ng cm-2 yr-1); and, MI-112 (0.7 ng cm-2 yr-1).These values are comparable to the preindustrial Hg influx (0.8 ng cm-2 yr-1) reported for sediment cores in southern LM (Pirrone et al., 1998). Hg influx profiles peak around the mid-1900s and then decrease after the 1970’s, potentially coinciding with the Clean Water Act (Figure 2B) (Adler et al., 1993). The peak Hg influxes in MI-116, MI-50, and MI-112 are about 3100%, 1600%, and 1400% enhanced relative to preindustrial backgrounds, respectively.

3.3 Historical records of δ202Hg and Δ199Hg

Our sediment profiles (Figure 2C) represented here show large variations in δ202Hg in MI-116 (δ202Hg: -1.75±0.11 to -0.52±0.14‰, 2σ), MI-50 (δ202Hg: -1.73±0.17 to -0.63±0.14‰, 2σ) and MI-112 (δ202Hg: -1.29±0.17 to -0.56±0.15‰, 2σ). Variable odd Hg-MIF was also observed (Figure 2D). MI-116 (Δ199Hg: -0.02±0.05‰ to 0.10±0.05‰, 2σ) appears absent of significant Hg-MIF before anthropogenic influences were introduced, while MI-50 (Δ199Hg: 0.05±0.05 to 0.18±0.05‰, 2σ) and MI-112 (Δ199Hg: 0.09±0.05 to 0.27±0.05‰, 2σ) show small but significant Hg-MIF. All sediments in LM yielded a Δ199Hg:Δ201Hg of 0.93±0.08 (2σ) (Figure 3), which is consistent with the photo-reduction of Hg(II) reported by Bergquist and Blum (2007). Evidences for stability of Hg isotope ratios in sediments over time have been document by Bartov et al. (2013). Loss of Hg during microbial Hg reduction may cause changes in Hg isotopic compositions over time (Kritee et al., 2013). However, with sediments containing organic matter at concentrations greater than Hg, such as LM sediments, exceptionally strong complexes of Hg(II) with organic matter may form (Morel et al., 1998). High organic matter content has been shown to inhibit Hg(II) reduction to Hg(0) (Mauclair et al., 2008; Gu et al., 2011). The variation of Hg isotopes in sediments then, is largely the result of varied Hg sources both spatially (Foucher and Hintelmann, 2009; Feng et al., 2010; Liu et al., 2011; Bartov et al., 2013; Donovan et al., 2013; Ma et al., 2013; Yin et al., 2013a; Lepak et al., 2016).

doi: 10.12952/journal.elementa.000086.f003.
Figure 3.  

Mass-independent fractionation of Hg isotopes in sediments of Lake Michigan.

Slope error was calculated by Williamson-York regression (Cantrell, 2008); Gray hashed line indicates aqueous Hg(II) photoreduction (Bergquist and Blum, 2007). Gray solid lines represent measurement uncertainties.

Historically, sources of Hg to LM have included direct discharge of anthropogenic Hg, watershed loading of soil Hg and precipitation (Mason and Sullivan, 1997; Landis and Keeler, 2002; Hurley et al., 1998a, 1998b; Jeremiason et al., 2009; Rossmann, 2002, 2009; Rossmann and Edgington, 2000). Intriguingly, the transition of δ202Hg from the deep layers is always positive in direction and very similar in magnitude for all three cores (Figure 2C), similar to previous observations on sediment cores from lakes and coastal areas (Donovan et al., 2013; Gray et al., 2013, 2015; Balogh et al., 2015). The deepest layer of core MI-112, presumably a pre-industrial period, has a negative δ202Hg value of ∼ -1.29±0.17(2σ) and positive Δ199Hg value of ∼ +0.16±0.05 (2σ) (Figure 4A-B). These values are similar to previous data on deep layers of coastal cores (δ202Hg: ∼ -1.0‰; Δ199Hg: ∼ +0.2‰)(Donovan et al., 2013; Balogh et al., 2015). Atmospheric deposition is an important input of Hg to Great Lakes (Landis and Keeler, 2002) and sea waters (Mason and Sheu, 2002). Precipitation scavenging oxidized Hg species (e.g., HgIIg and Hgp) in the atmosphere is most likely the primary carrier for atmospheric Hg signals in offshore lake sediments (Shanley et al., 2015). Previous studies have reported mean δ202Hg of approximately -0.5‰ and positive Δ199Hg (∼ +0.4‰) and Δ200Hg (mean: ∼0.25‰) in precipitation collected in the Great Lakes region (Gratz et al., 2010; Chen et al., 2012; Demers et al., 2013; Sherman et al., 2011, 2015). Atmospheric Hg(II) is susceptible to photoreduction in precipitation water droplets (Gratz et al., 2010), which may at least partially explain the positive pre-industrial Δ199Hg in deep layer sediments of MI-112 and other coastal cores. Precipitation-derived Hg appears to be adsorbed to particulates and settle to the benthos. Adsorption of aqueous Hg(II) by particles containing thiol groups (Wiederhold et al., 2010), goethite (Jiskra et al., 2012) and sulfides (Foucher et al., 2013; Smith et al., 2015) may result in negative δ202Hg (-0.4 to -0.6‰) shifts in the solid Hg phase. Precipitation Hg with δ202Hg value of ∼ -0.5‰ plus a negative shift of -0.4 to -0.6‰ during adsorption, may also result in similar δ202Hg to deep layer sediments in MI-112 and coastal cores.These observations suggest that precipitation is main input of Hg to deep layers of MI-112.

doi: 10.12952/journal.elementa.000086.f004.
Figure 4.  

1/THg versus δ202Hg (A) and Δ199Hg (B) in sediments of Lake Michigan.

Red solid lines represent MI-112; green solid lines represent MI-50, and blue solid line represents MI-116. Measurement uncertainties are presented as gray solid lines.

The deepest layer sediments of MI-116 and MI-50 show δ202Hg values of -1.75±0.11‰(2σ) and -1.73±0.17‰(2σ), respectively, which are lower than that of the deepest layers in MI-112, but higher than previous data on the deep layer sediments from other lakes (δ202Hg: -3 to -2‰) (Figure 4A, Gray et al., 2013, 2015). The negative δ202Hg in deep layers of other lakes are similar to that reported for terrestrial soils (δ202Hg: -2.0±0.6‰, σ, n = 48) (Biswas et al., 2008; Zhang et al., 2013; Demers et al., 2013; Jiskra et al., 2015). Mercury strongly bound to soil matrices maybe subjected to only minimal aqueous processing and therefore result in insignificant isotope fractionation during sedimentation (Foucher and Hintelmann, 2009; Liu et al., 2011). Deep layer sediments from other lakes have shown negative Δ199Hg values similar to ours (-0.3 to -0.1‰) (Figure 4B; Gray et al., 2013, 2015). Watershed soil particles are likely the primary source of Hg to these lakes as Hg in terrestrial soils has shown to be characterized by negative Δ199Hg values (-0.2±0.2‰, σ, n = 48) (Biswas et al., 2008; Zhang et al., 2013; Demers et al., 2013; Jiskra et al., 2015; Feng et al., 2010; Liu et al., 2011). If we consider physical mixing of sources containing positive and negative MIF, the lack of significant MIF (either positive or negative) in the pre-industrial sediment of MI-116 may indicate similar proportions of Hg from watersheds and precipitation were introduced (if similar magnitudes of MIF from each source are assumed). In general, Δ199Hg values in deep layers decreased from MI-112, MI-50 to MI-116 (Figure 2D), suggesting increase of Hg inputs by watershed soils.

The uppermost sections of the MI-116 core show much higher δ202Hg (-1 to 0‰) and the absence of Hg-MIF (Δ199Hg ∼ 0) (Figures 2C–D and 4A–B). These values are similar to previous data on major anthropogenic Hg sources, which are in general characterized by higher δ202Hg values ranging from -1 to 0‰ and insignificant Hg-MIF (Sonke et al., 2010; Liu et al., 2011; Blum et al., 2014). Although there are many industries that may release Hg into LM, Hg used in industry is primarily derived from mining, and therefore should exhibit similar signatures. Previous results on Hg ores other metal deposits exhibited δ202Hg of ∼ -0.6±0.4‰ (σ) and Δ199Hg close to 0 (Sonke et al., 2010; Liu et al., 2011; Blum et al., 2014; Hintelmann and Lu, 2003; Smith et al., 2005, 2008; Stetson et al., 2009; Laffont et al., 2011; Yin et al., 2013a). Sediments with high THg concentrations is predominately derived from anthropogenic sources as deposition typically occurs near the source, waters are relatively turbid (thereby reducing photochemical processes) or the Hg from these sources are less susceptible to photo-reduction (Foucher and Hintelmann, 2009; Feng et al., 2010; Gehrke et al., 2011; Sonke et al., 2010; Donovan et al., 2013; Cooke et al., 2013; Ma et al., 2013; Mil-Homens et al., 2013; Gray et al., 2013, 2015; Balogh et al., 2015; Yin et al., 2013a; Lepak et al., 2016). Uppermost sections of MI-50 and MI-112 have similar δ202Hg values with MI-116, but show slightly positive Δ199Hg. Δ199Hg generally decreases with increased THg concentrations, which may indicate a dilution by anthropogenic Hg sources. Overall, we suggest that anthropogenic Hg sources are important inputs of Hg in top layer sediments of LM, certainly within Green Bay.

3.4 Quantifying the anthropogenic Hg input using δ202Hg values

Evaluation of mixing of different sources can be achieved when plotting δ202Hg and 1/THg values, when using binary and ternary mixing models of Hg isotopes to trace and quantify source apportionment of Hg in sediments (Foucher and Hintelmann, 2009; Liu et al., 2011; Donovan et al., 2013; Ma et al., 2013; Yin et al., 2013a). In our study, δ202Hg and 1/THg form a linear correlation, suggesting physical mixing of anthropogenic Hg and geochemical background sources of Hg (Figure 4A). Like other small inland water bodies, the background Hg of MI-116 and MI-50 in Green Bay may represent mixing of precipitation and watershed soils, whereas that of MI-112 in offshore LM may be more similar to coastal regions of the ocean where primary Hg influences are mainly atmospherically derived (as mentioned in Section 3.3). For these reasons, we believe the δ202Hg and 1/THg mixing lines appear to fall on two different slopes in Figure 4. With the mixing lines shown in Figure 4A, it is possible to estimate the influence of both anthropogenic and background signals using an end-member approach and the following binary mixing model (Equations 5–7):

Fant. + Fbac. = 1
Fant.* δ202Hgant. + Fbac.*.δ202Hgbac. = δ202Hgsample
Fant. = (δ202Hgsample  δ202Hgbac.)/(δ202Hgant.  δ202Hgbac.)

where Fant. and Fbac. represent fractions of the anthropogenic and background Hg end-members, respectively. δ202Hgant. and δ202Hgbac. represent the δ202Hg of anthropogenic and background Hg end-members, respectively, and δ202Hgsample represents the δ202Hg of a given sample. Values of δ202Hg for the samples at the base of the each core (MI-116: -1.75‰; MI-50: -1.73‰; MI-112: -1.29‰) were used to represent the background end-member for each core in this study. A core slice with the highest δ202Hg (-0.52‰) from core MI-116 was chosen to represent the isotopic signature of the anthropogenic end-member. This was chosen because it correlates with the highest concentration of THg at about 1930. With this information the variation in Fant of each sediment core was evaluated (Figure 5).

doi: 10.12952/journal.elementa.000086.f005.
Figure 5.  

Distribution of Fant. in sediment profiles of Lake Michigan.

Model output indicates two time periods when the onset of anthropogenic Hg pollution in LM was evident. The first indication of Hg pollution starts in the early 1800s in MI-112 and MI-116, and most likely represents the influence of the global industrial revolution; the second indication of Hg pollution starts in the 1880s in both MI-116 and MI-50, and likely reflects the influence of local industry which started in 1865 when the first paper mill was built. Post 1920’s, all three sites display relatively high degrees of anthropogenic influence ranging from 50 to 100%. In addition, a peak of Fant. was observed in the 1850s in MI-112. Schuster et al. (2002) observed a similar episodic Hg accumulation event in an ice core from a Wyoming glacier which was correlated to California Gold Rush (1850 to 1878) at such a distance from the source. It is not clear whether this peak is related to the Gold Rush, but sediment cores from other lakes in the Midwest have also recorded episodic events during that time period although that study did not attempt to interpret the origin of the peaks (Engstrom and Swain, 1997). We also observed a decreasing trend of Fant. following the 1970s in offshore core MI-112, which may be the result of recent remediation efforts, as a result of the CWA.

3.5 Mass independent fractionation of 200Hg

Due to the lack of Hg isotope data in precipitation throughout time, further estimates of the contribution of Hg from precipitation and watershed soils was not conducted in our study. Mercury isotopic composition in modern precipitation show large variations between pristine and industrial-urban areas (Gratz et al., 2010; Chen et al., 2012; Demers et al., 2013; Sherman et al., 2011, 2015). While variable, modern precipitation has shown average Δ200Hg values of ∼ +0.20‰, whereas most other environmental matrices have shown Δ200Hg values close to zero. Δ200Hg has been used as a tracer for precipitation Hg (Lepak et al., 2016). In this study, the MI-116 profile shows insignificant MIF of 200Hg compared to analytical uncertainty for Δ200Hg (±0.04‰, 2σ) (Figure 2E). Somewhat interestingly, many samples in MI-50 and MI-112 have small but significant positive Δ200Hg (> +0.04‰). These samples are located in the deepest layers (prior-1850s) and the uppermost slices (post-1970s). The absence of 200Hg MIF in deep layers of MI-116 suggests that precipitation-derived Hg is diluted by watershed soils, whereas the small positive Δ200Hg in the deep layers (prior-1850s) of MI-50 and MI-112 may suggest that the proportion of atmospherically derived Hg is larger. Increased inputs of anthropogenic Hg may signify an absence of 200Hg MIF in sediments from 1850s to 1970s in MI-50 and MI-112, when industrialization and Hg loading to this system was the greatest. The slightly positive Δ200Hg in top layers of MI-50 and MI-112 is apparent.

4. Conclusions

Our results on sedimentary Hg influx profiles confirm previous studies on the timing and extent of Hg pollution in LM. Our isotopic studies, coupled with a binary mixing model, enabled us to differentiate between the historical contribution from anthropogenic Hg sources and background inputs. Isotopic analyses using Hg-MIF further allowed us to compare the extent of Hg(II) photo-reduction processes in the water column among sites with different Hg sources. Hg isotopes can be a useful tracer to reveal sources and fate of Hg in large aquatic ecosystems, such as the Great Lakes.

Data accessibility statement

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© 2016 Yin 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

Text S1.
Experimental methods
File Type: DOC
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Table S1.
THg, Hg isotopic composition, age, bulk density, evaluated results for the LM sediment cores
File Type: DOC
File Size: 0.04 MB

Table S2.
Operating parameters of the Neptune Plus MC-ICP-MS
File Type: DOC
File Size: 0.02 MB

Table S3.
Isotopic composition of Hg for UM-Almadén
File Type: DOC
File Size: 0.02 MB


Contributed to conception and design: R-S Y, JPH, DPK

Contributed to acquisition of data: R-S Y, RFL

Contributed to analysis and interpretation of data: R-S Y, JPH, DPK, RFL

Drafted and/or revised the article: R-S Y, JPH, DPK, RFL

Approved the submitted version for publication: R-S Y, JPH, DPK, RFL

Competing interests

The authors have no competing interests.

Funding information

RFL was supported by the University of Wisconsin-Madison Graduate School. Sediment work was supported by USEPA Great Lakes Restoration Initiative, project GL-00E01139. Although the research described in this article has been partly funded by the USEPA, it has not been subjected to the agency’s required peer and policy review and therefore, does not necessarily reflect the views of the agency and no official endorsement should be inferred.


We would like to thank Michael Tate, Jacob Ogorek, John DeWild, Morgan Maglio and Charlie Thompson from Wisconsin Mercury Research Lab, U.S. Geological Survey, and Middleton WI for the assistance in sample collection and THg analysis. We also thank professor Xinbin Feng from Institute of Geochemistry, Chinese Academy of Sciences to revise this paper. Two anonymous reviewers and editors Robert P. Mason and Joel D. Blum are acknowledged for their constructive comments and useful suggestions that have largely improved the quality of this paper.


  1. Adler RW, Landman JC, Cameron DM. 1993. The Clean Water Act 20 years later . Island Press.

  2. Balogh SJ , Tsui MTK , Blum JD , et al. 2015. Tracking the Fate of Mercury in the Fish and Bottom Sediments of Minamata Bay, Japan, Using Stable Mercury Isotopes. Environ Sci Technol 49(9): 5399–5406.

  3. Bartov G , Deonarine A , Johnson TM , et al. 2013. Environmental impacts of the Tennessee Valley Authority Kingston coal ash spill. 1. Source apportionment using mercury stable isotopes. Environ Sci Technol 47(4): 2092–2099.

  4. Benoit JM , Gilmour CC , Heyes A , et al. 2003. Geochemical and biological controls over methylmercury production and degradation in aquatic ecosystems, in CaiY, BraidsOC, eds., Biogeochemistry of Environmentally Important Trace Elements . Washington DC: American Chemical Society: pp 262–297.

  5. Bergquist BA, Blum JD. 2007. Mass-dependent and-independent fractionation of Hg isotopes by photoreduction in aquatic systems. Science 318(5849): 417–420.

  6. Bergquist BA, Blum JD. 2009. The odds and evens of mercury isotopes: Applications of mass-dependent and mass-independent isotope fractionation. Elements 5(6): 353–357.

  7. Biswas A , Blum JD , Bergquist BA , et al. 2008. Natural mercury isotope variation in coal deposits and organic soils. Environ Sci Technol 42(22): 8303–8309.

  8. Blum JD, Bergquist BA. 2007. Reporting of variations in the natural isotopic composition of mercury. Anal Bioanal Chem 388(2): 353–359.

  9. Blum JD , Popp BN , Drazen JC , et al. 2013. Methylmercury production below the mixed layer in the North Pacific Ocean. Nat Geosci 6(10): 879–884.

  10. Blum JD, Sherman LS, Johnson MW. 2014. Mercury isotopes in earth and environmental sciences. Annu Rev Earth Pl Sc 42: 249–269.

  11. Buchachenko AL , Ivanov VL , Roznyatovskii VA , et al. 2007. Magnetic isotope effect for mercury nuclei in photolysis of bis (p-trifluoromethylbenzyl) mercury. Dokl Phys Chem 13(1): 39–41.

  12. Cantrell CA. 2008. Technical Note: Review of methods for linear least-squares fitting of data and application to atmospheric chemistry problems. Atmos Chem Phys 8(17): 5477–5487.

  13. Chandan P, Ghosh S, Bergquist BA. 2014. Mercury isotope fractionation during aqueous photoreduction of monomethylmercury in the presence of dissolved organic matter. Environ Sci Technol 49(1): 259–267.

  14. Chen J , Hintelmann H , Feng X , et al. 2012. Unusual fractionation of both odd and even mercury isotopes in precipitation from Peterborough, ON, Canada. Geochim Cosmochim Acta 90: 33–46.

  15. Christensen ER, Chien NK. 1981. Fluxes of arsenic, lead, zinc, and cadmium to Green Bay and Lake Michigan sediments. Environ Sci Technol 15(5): 553–558.

  16. Cooke CA , Hintelmann H , Ague JJ , et al. 2013. Use and legacy of mercury in the Andes. Environ Sci Technol 47(9): 4181–4188.

  17. 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 Cy 27(1): 222–238.

  18. Donovan PM , Blum JD , Demers JD , et al. 2014. Identification of multiple mercury sources to stream sediments near Oak Ridge, TN, USA. Environ Sci Technol 48(7): 3666–3674.

  19. Donovan PM , Blum JD , Yee D , et al. 2013. An isotopic record of mercury in San Francisco Bay sediment. Chem Geol 349: 87–98.

  20. Engstrom DR , Fitzgerald WF , Cooke CA , 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.

  21. Engstrom DR, Swain EB. 1997. Recent declines in atmospheric mercury deposition in the upper Midwest. Environ Sci Technol 31(4): 960–967.

  22. Estrade N , Carignan J , Sonke JE , et al. 2009. Mercury isotope fractionation during liquid–vapor evaporation experiments. Geochim Cosmochim Acta 73(10): 2693–2711.

  23. Feng X , Foucher D , Hintelmann H , et al. 2010. Tracing mercury contamination sources in sediments using mercury isotope compositions. Environ Sci Technol 44(9): 3363–3368.

  24. Feng X , Yin R , Yu B , et al. 2013. Mercury isotope variations in surface soils in different contaminated areas in Guizhou Province, China. Chin Sci Bull 58(2): 249–255.

  25. Fitzgerald WF, Lamborg CH, Hammerschmidt CR. 2007. Marine biogeochemical cycling of mercury. Chem Rev 107(2): 641–662.

  26. Foucher D, Hintelmann H. 2006. High-precision measurement of mercury isotope ratios in sediments using cold–vapor generation multi-collector inductively coupled plasma mass spectrometry. Anal Bioanal Chem 384 (7–8): 1470–1478.

  27. Foucher D, 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.

  28. Foucher D , Hintelmann H , Al TA , et al. 2013. Mercury isotope fractionation in waters and sediments of the Murray Brook mine watershed (New Brunswick, Canada): Tracing mercury contamination and transformation. Chem Geol 336: 87–95.

  29. 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.

  30. Ghosh S , Schauble EA , Couloume GL , et al. 2013. Estimation of nuclear volume dependent fractionation of mercury isotopes in equilibrium liquid–vapor evaporation experiments. Chem Geol 336: 5–12.

  31. Gratz LE , Keeler GJ , Blum JD , et al. 2010. Isotopic composition and fractionation of mercury in Great Lakes precipitation and ambient air. Environ Sci Technol 44(20): 7764–7770.

  32. Gray JE , Pribil MJ , Van Metre PC , et al. 2013. Identification of contamination in a lake sediment core using Hg and Pb isotopic compositions, Lake Ballinger, Washington, USA. Appl Geochem 29: 1–12.

  33. Gray JE , Van Metre PC , Pribil MJ , et al. 2015. Tracing historical trends of Hg in the Mississippi River using Hg concentrations and Hg isotopic compositions in a lake sediment core, Lake Whittington, Mississippi, USA. Chem Geol 395: 80–87.

  34. Gu B , Bian Y , Miller CL , et al. 2011. Mercury reduction and complexation by natural organic matter in anoxic environments. P Natl Acad Sci USA 108(4): 1479–1483.

  35. Hammerschmidt CR , Bowman KL , Tabatchnick MD , et al. 2011. Storage bottle material and cleaning for determination of total mercury in seawater. Limnol Oceanogr: Methods 9(10): 426–431.

  36. Hartig JH, Law NL. 1994. Progress in Great Lakes Remedial Action Plans: Implementing the Ecosystem Approach in Great Lakes Areas of Concern . Chicago, IL: U.S. Environmental Protection Agency.

  37. Hecky RE , Smith REH , Barton DR , et al. 2004. The nearshore phosphorus shunt: A consequence of ecosystem engineering by dreissenids in the Laurentian Great Lakes. Can J Fish Aquat Sci 61(7): 1285–1293.

  38. Hermanson MH , Christensen ER , Buser DJ , et al. 1991. Polychlorinated biphenyls in dated sediment cores from Green Bay and Lake Michigan. J Great Lakes Res 17(1): 94–108.

  39. Hintelmann H. 2012. Use of stable isotopes in mercury research . Berkeley: University of California Press: 55–71.

  40. Hintelmann H, Lu SY. 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.

  41. Hurley JP , Cowell SE , Shafer MM , et al. 1998a. Partitioning and transport of total and methyl mercury in the lower Fox River, Wisconsin. Environ Sci Technol 32(10): 1424–1432.

  42. Hurley JP , Cowell SE , Shafer MM , et al. 1998b. Tributary loading of mercury to Lake Michigan: Importance of seasonal events and phase partitioning. Sci Total Environ 213(1): 129–137.

  43. Jackson TA. 2013. Mass-dependent and mass-independent variations in the isotope composition of mercury in a sediment core from Lake Ontario as related to pollution history and biogeochemical processes. Chem Geol 355: 88–102.

  44. Jackson TA, Muir DC, Vincent WF. 2004. Historical variations in the stable isotope composition of mercury in Arctic lake sediments. Environ Sci Technol 38(10): 2813–2821.

  45. Jeremiason JD , Kanne LA , Lacoe TA , et al. 2009. A comparison of mercury cycling in Lakes Michigan and Superior. J Great Lakes Res 35(3): 329–336.

  46. Jiskra M , Wiederhold JG , Bourdon B , et al. 2012. Solution speciation controls mercury isotope fractionation of Hg (II) sorption to goethite. Environ Sci Technol 46(12): 6654–6662.

  47. Jiskra M , Wiederhold JG , Skyllberg U , et al. 2015. Mercury deposition and re-emission pathways in boreal forest soils investigated with Hg isotope signatures. Environ Sci Technol 49(12): 7188–7196

  48. Jonsson S , Skyllberg U , Nilsson MB , et al. 2014. Differentiated availability of geochemical mercury pools controls methylmercury levels in estuarine sediment and biota. Nat Com . doi: 10.1038/ncomms5624.

  49. Kadlec RH, Robbins JA. 1984. Sedimentation and sediment accretion in Michigan coastal wetlands (USA). Chem Geol 44(1): 119–150.

  50. Klump JV , Edgington DN , Sager PE , et al. 1997. Sedimentary phosphorus cycling and a phosphorus mass balance for the Green Bay (Lake Michigan) ecosystem. Can J Fish Aquat Sci 54(1): 10–26.

  51. Kritee K , Blum JD , Reinfelder JR , et al. 2013. Microbial stable isotope fractionation of mercury: A synthesis of present understanding and future directions. Chem Geol 336: 13–25.

  52. Kwon SY , Blum JD , Carvan MJ , et al. 2012. Absence of fractionation of mercury isotopes during trophic transfer of methylmercury to freshwater fish in captivity. Environ Sci Technol 46: 7527–7534.

  53. Kwon SY , Blum JD , Chirby MA , et al. 2013. Application of mercury isotopes for tracing trophic transfer and internal distribution of mercury in marine fish feeding experiments. Environ Toxicol Chem 23: 2322–2330.

  54. Laffont L , Sonke JE , Maurice L , et al. 2011. Anomalous mercury isotopic compositions of fish and human hair in the Bolivian Amazon. Environ Sci Technol 43: 8985–8990.

  55. Landis MS, Keeler GJ. 2002. Atmospheric mercury deposition to Lake Michigan during the Lake Michigan mass balance study. Environ Sci Technol 36(21): 4518–4524.

  56. 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.

  57. Lepak RF , Yin R , Krabbenhoft DP , et al. 2016. Use of Stable Isotope Signatures to Determine Mercury Sources in the Great Lakes. Environ Sci Technol Let . doi: 10.1021/acs.estlett.5b00277.

  58. Liu J , Feng X , Yin R , et al. 2011. Mercury distributions and mercury isotope signatures in sediments of Dongjiang, the Pearl River Delta, China. Chem Geol 287(1): 81–89.

  59. Ma J , Hintelmann H , Kirk JL , et al. 2013. Mercury concentrations and mercury isotope composition in lake sediment cores from the vicinity of a metal smelting facility in Flin Flon, Manitoba. Chem Geol 336: 96–102.

  60. Marvin C , Painter S , Williams D , et al. 2004. Spatial and temporal trends in surface water and sediment contamination in the Laurentian Great Lakes. Environ Pollut 129(1): 131–144.

  61. Mason RP, Sheu GR. 2002. Role of the ocean in the global mercury cycle. Global Biogeochem Cy 16(4): 40–41.

  62. Mason RP, Sullivan KA. 1997. Mercury in Lake Michigan. Environ Sci Technol 31(3): 942–947.

  63. Mauclair C, Layshock J, Carpi A. 2008. Quantifying the effect of humic matter on the emission of mercury from artificial soil surfaces. Appl Geochem 23(3): 594–601.

  64. Mil-Homens M , Blum J , Canário J , et al. 2013. Tracing anthropogenic Hg and Pb input using stable Hg and Pb isotope ratios in sediments of the central Portuguese Margin. Chem Geol 336: 62–71.

  65. Morel FMM, Kraepiel AML, Amyot M. 1998. The chemical cycle and bioaccumulation of mercury. Annual Review of Ecology and Systematics 543–566.

  66. Pirrone N , Allegrini I , Keeler GJ , et al. 1998. Historical atmospheric mercury emissions and depositions in North America compared to mercury accumulations in sedimentary records. Atmos Environ 32(5): 929–940.

  67. Pirrone N , Cinnirella S , Feng X , et al. 2010. Global mercury emissions to the atmosphere from anthropogenic and natural sources. Atmos Chem Phys 10(13): 5951–5964.

  68. Pizzuto JE. 2014. Long-term storage and transport length scale of fine sediment: Analysis of a mercury release into a river. Geophys Res Lett 41(16): 5875–5882.

  69. Redman S. 1993. Preliminary Evaluation of Mercury in the Sediments of the Fox River and Lower Green Bay. Planning and Policy Section, Bureau of Water Resources Management . Madison, WI: Wisconsin Department of Natural Resources.

  70. Rolison JM , Landing WM , Luke W , et al. 2013. Isotopic composition of species-specific atmospheric Hg in a coastal environment. Chem Geol 336: 37–49.

  71. Rose CH , Ghosh S , Blum JD , et al. 2015. Effects of ultraviolet radiation on mercury isotope fractionation during photo-reduction for inorganic and organic mercury species. Chem Geol 405: 102–111.

  72. Rossmann R. 2002. Lake Michigan 1994–1996 surficial sediment mercury. J Great Lakes Res 28(1): 65–76.

  73. Rossmann R. 2009. Protocol to reconstruct historical contaminant loading to large lakes: The Lake Michigan sediment record of mercury. Environ Sci Technol 44(3): 935–940.

  74. Rossmann R, Edgington DN. 2000. Mercury in Green Bay, Lake Michigan surficial sediments collected between 1987 and 1990. J Great Lakes Res 26(3): 323–339.

  75. 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.

  76. Schuster PF , Krabbenhoft DP , Naftz DL , 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.

  77. Shanley JB , Engle M , Scholl MA , et al. 2015. High mercury wet deposition at a “clean air” site in Puerto Rico. Environ Sci Technol 49(20): 12474–12482.

  78. Sherman LS , Blum JD , Dvonch JT , et al. 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.

  79. Sherman LS , Blum JD , Keeler GJ , et al. 2011. Investigation of local mercury deposition from a coal-fired power plant using mercury isotopes. Environ Sci Technol 46(1): 382–390.

  80. Smith CN , Kesler SE , Blum JD , et al. 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): 399–407.

  81. Smith CN , Kesler SE , Klaue B , et al. 2005. Mercury isotope fractionation in fossil hydrothermal systems. Geology 33(10): 825–828.

  82. 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.

  83. Song W , Li A , Ford JC , et al. 2005. Polybrominated diphenyl ethers in the sediments of the Great Lakes. 2. Lakes Michigan and Huron. Environ Sci Technol 39(10): 3474–3479.

  84. Sonke JE, 2011. A global model of mass independent mercury stable isotope fractionation. Geochim Cosmochim Acta 75(16): 457–4590.

  85. Sonke JE . Schäfer J , Chmeleff J , et al. 2010. Sedimentary mercury stable isotope records of atmospheric and riverine pollution from two major European heavy metal refineries. Chem Geol 279(3): 90–100.

  86. Stetson SJ , Gray JE , Wanty RB , et al. 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.

  87. Štrok M, Hintelmann H, Dimock B. 2014. Development of pre-concentration procedure for the determination of Hg isotope ratios in seawater samples. Anal Chim Acta 851: 57–63.

  88. Sun R , Sonke JE , Heimbürger LE , et al. 2014. Mercury stable isotope signatures of world coal deposits and historical coal combustion emissions. Environ Sci Technol 48(13): 7660–7668.

  89. U.S. EPA. 1997. Mercury study report to Congress. Volume 5. Health effects of mercury and mercury compounds. EPA-452/R-97/007 .

  90. U.S. EPA. 2004. Results of the Lake Michigan Mass Balance Study: Mercury Data Report. EPA-905/R-01/012 .

  91. Wiederhold JG , Cramer CJ , Daniel K , et al. 2010. Equilibrium mercury isotope fractionation between dissolved Hg (II) species and thiol-bound Hg. Environ Sci Technol 44(11): 4191–4197.

  92. Wisconsin Department of Natural Resources. 1997. Draft Wisconsin Mercury Source book.

  93. Yang H , Battarbee RW , Turner SD , et al. 2010. Historical reconstruction of mercury pollution across the Tibetan Plateau using lake sediments. Environ Sci Technol 44(8): 2918–2924.

  94. Yin R , Feng X , Chen B , et al. 2015. Identifying the sources and processes of mercury in subtropical estuarine and ocean sediments using Hg isotopic composition. Environ Sci Technol 49(3): 1347–1355.

  95. Yin R, Feng X, Chen J. 2014b. 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.

  96. Yin R , Feng X , Foucher D , et al. 2010a. High precision determination of mercury isotope ratios using online mercury vapor generation system coupled with multicollector inductively coupled plasma-mass spectrometer. Chin J Anal Chem 38(7): 929–934.

  97. Yin R , Feng X , Li X , et al. 2014a. Trends and advances in mercury stable isotopes as a geochemical tracer. TrEAC 2: 1–10.

  98. Yin R, Feng X, Meng B. 2013b. Stable mercury isotope variation in rice plants (Oryza sativa L.) from the Wanshan mercury mining district, SW China. Environ Sci Technol 47(5): 2238–2245.

  99. Yin R, Feng X, Shi W. 2010b. Application of the stable-isotope system to the study of sources and fate of Hg in the environment: A review. Appl Geochem 25(10): 1467–1477.

  100. Yin R , Feng X , Wang J , et al. 2013a. 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.

  101. Zhang H , Yin R , Feng X , et al.. 2013. Atmospheric mercury inputs in montane soils increase with elevation: Evidence from mercury isotope signatures. Sci Rep . doi: 10.1038/srep03322.

  102. Zheng W, Hintelmann H. 2009. Mercury isotope fractionation during photoreduction in natural water is controlled by its Hg/DOC ratio. Geochim Cosmochim Acta 73(22): 6704–6715.

  103. Zheng W, Hintelmann H. 2010a. 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.

  104. Zheng W, Hintelmann H. 2010b. Isotope fractionation of mercury during its photochemical reduction by low-molecular-weight organic compounds. J Phys Chem A 114(12): 4246–4253.

  105. Zhu LY, Hits RA. 2005. Brominated flame retardants in sediment cores from Lakes Michigan and Erie. Environ Sci Technol 39(10): 3488–3494.

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