A review of global environmental mercury processes in response to human and natural perturbations: Changes of emissions, climate, and land use

We review recent progress in our understanding of the global cycling of mercury (Hg), including best estimates of Hg concentrations and pool sizes in major environmental compartments and exchange processes within and between these reservoirs. Recent advances include the availability of new global datasets covering areas of the world where environmental Hg data were previously lacking; integration of these data into global and regional models is continually improving estimates of global Hg cycling. New analytical techniques, such as Hg stable isotope characterization, provide novel constraints of sources and transformation processes. The major global Hg reservoirs that are, and continue to be, affected by anthropogenic activities include the atmosphere (4.4–5.3 Gt), terrestrial environments (particularly soils: 250–1000 Gg), and aquatic ecosystems (e.g., oceans: 270–450 Gg). Declines in anthropogenic Hg emissions between 1990 and 2010 have led to declines in atmospheric Hg0 concentrations and HgII wet deposition in Europe and the US (− 1.5 to − 2.2% per year). Smaller atmospheric Hg0 declines (− 0.2% per year) have been reported in high northern latitudes, but not in the southern hemisphere, while increasing atmospheric Hg loads are still reported in East Asia. New observations and updated models now suggest high concentrations of oxidized HgII in the tropical and subtropical free troposphere where deep convection can scavenge these HgII reservoirs. As a result, up to 50% of total global wet HgII deposition has been predicted to occur to tropical oceans. Ocean Hg0 evasion is a large source of present-day atmospheric Hg (approximately 2900 Mg/year; range 1900–4200 Mg/year). Enhanced seawater Hg0 levels suggest enhanced Hg0 ocean evasion in the intertropical convergence zone, which may be linked to high HgII deposition. Estimates of gaseous Hg0 emissions to the atmosphere over land, long considered a critical Hg source, have been revised downward, and most terrestrial environments now are considered net sinks of atmospheric Hg due to substantial Hg uptake by plants. Litterfall deposition by plants is now estimated at 1020–1230 Mg/year globally. Stable isotope analysis and direct flux measurements provide evidence that in many ecosystems Hg0 deposition via plant inputs dominates, accounting for 57–94% of Hg in soils. Of global aquatic Hg releases, around 50% are estimated to occur in China and India, where Hg drains into the West Pacific and North Indian Oceans. A first inventory of global freshwater Hg suggests that inland freshwater Hg releases may be dominated by artisanal and small-scale gold mining (ASGM; approximately 880 Mg/year), industrial and wastewater releases (220 Mg/year), and terrestrial mobilization (170–300 Mg/year). For pelagic ocean regions, the dominant source of Hg is atmospheric deposition; an exception is the Arctic Ocean, where riverine and coastal erosion is likely the dominant source. Ocean water Hg concentrations in the North Atlantic appear to have declined during the last several decades but have increased since the mid-1980s in the Pacific due to enhanced atmospheric deposition from the Asian continent. Finally, we provide examples of ongoing and anticipated changes in Hg cycling due to emission, climate, and land use changes. It is anticipated that future emissions changes will be strongly dependent on ASGM, as well as energy use scenarios and technology requirements implemented under the Minamata Convention. We predict that land use and climate change impacts on Hg cycling will be large and inherently linked to changes in ecosystem function and global atmospheric and ocean circulations. Our ability to predict multiple and simultaneous changes in future Hg global cycling and human exposure is rapidly developing but requires further enhancement.

Abstract We review recent progress in our understanding of the global cycling of mercury (Hg), including best estimates of Hg concentrations and pool sizes in major environmental compartments and exchange processes within and between these reservoirs. Recent advances include the availability of new global datasets covering areas of the world where environmental Hg data were previously lacking; integration of these data into global and regional models is continually improving estimates of global Hg cycling. New analytical techniques, such as Hg stable isotope characterization, provide novel constraints of sources and transformation processes. The major global Hg reservoirs that are, and continue to be, affected by anthropogenic activities include the atmosphere (4.4-5.3 Gt), terrestrial environments (particularly soils: 250-1000 Gg), and aquatic ecosystems (e.g., oceans: 270-450 Gg). Declines in anthropogenic Hg emissions between 1990 and 2010 have led to declines in atmospheric Hg 0 concentrations and Hg II wet deposition in Europe and the US (-1.5 to -2.2% per year). Smaller atmospheric Hg 0 declines (-0.2% per year) have been reported in high northern latitudes, but not in the southern hemisphere, while increasing atmospheric Hg loads are still reported in East Asia. New observations and updated models now suggest high concentrations of oxidized Hg II in the tropical and subtropical free troposphere where deep convection can scavenge these Hg II reservoirs. As a result, up to 50% of total global wet Hg II deposition has been predicted to occur to tropical oceans. Ocean Hg 0 evasion is a large source of present-day atmospheric Hg (approximately 2900 Mg/year; range 1900-4200 Mg/year). Enhanced seawater Hg 0 levels suggest enhanced Hg 0 ocean evasion in the intertropical convergence zone, which may be linked to high Hg II deposition. Estimates of gaseous Hg 0 emissions to the atmosphere over land, long considered a critical Hg source, have been revised downward, and most terrestrial environments now are considered net sinks of atmospheric Hg due to substantial Hg uptake by plants. Litterfall deposition by plants is now estimated at 1020-1230 Mg/year globally. Stable isotope analysis and direct flux measurements provide evidence that in many ecosystems Hg 0 deposition via plant inputs dominates, accounting for 57-94% of Hg in soils. Of global aquatic Hg releases, around 50% are estimated to occur in China and India, where Hg drains into the West Pacific and North Indian Oceans. A first inventory of global freshwater Hg suggests that inland freshwater Hg releases may be dominated by artisanal and small-scale gold mining (ASGM; approximately 880 Mg/year), industrial and wastewater releases (220 Mg/year), and terrestrial mobilization (170-300 Mg/year). For pelagic ocean regions, the dominant source of Hg is atmospheric deposition; an exception is the Arctic Ocean, where riverine and coastal erosion is likely the dominant source. Ocean water Hg concentrations in the North Atlantic appear to have declined during the last several decades but have increased since the mid-1980s in the Pacific due to enhanced atmospheric deposition from the Asian continent. Finally, we provide examples of ongoing and anticipated changes in Hg cycling due to emission, climate, and land use changes. It is anticipated that future emissions changes will be strongly dependent on ASGM, as well as energy use scenarios and technology requirements implemented under the Minamata Convention. We predict that land use and climate change impacts on Hg cycling will be large and inherently linked to changes in ecosystem function and global atmospheric and ocean circulations. Our ability to predict multiple and simultaneous changes in future Hg global cycling and human exposure is rapidly developing but requires further enhancement.

INTRODUCTION
Our understanding of the critical processes driving global mercury (Hg) cycling, in particular those that affect largescale exchange of Hg among major environmental compartments, has advanced substantially over the past decade. Progress has been driven by major advances in three interconnected areas: new data, new models, and new analytical tools and techniques. In this paper, we summarize the state of knowledge of the major global Hg reservoirs in the Earth system: the atmosphere, terrestrial ecosystems, and aquatic ecosystems. We describe the constraints on processes that control Hg exchanges between these reservoirs, and the relative influences of policy, land use, climate change, and anthropogenic disturbances on Hg cycling (Fig. 1).
Analyses of newly available data in the context of advances in modeling capabilities and novel analysis techniques have improved our understanding of fundamental processes relevant to Hg cycling. In the past decade, new data have become available from areas of the world where they previously were lacking, including Asia, the tropics, and the southern hemisphere. Environmental models are increasingly used for synthesizing global observations and describing the mechanisms driving Hg speciation, cycling, and bioavailability. Global three-dimensional (3D) models of Hg in the atmosphere (Dastoor and Larocque 2004;Selin et al. 2007;Jung et al. 2009;Travnikov and Ilyin 2009;Holmes et al. 2010;Durnford et al. 2012;Bieser et al. 2017;Horowitz et al. 2017), terrestrial ecosystems (Smith-Downey et al. 2010), and oceans (Zhang et al. , 2015bSemeniuk and Dastoor 2017) have improved our understanding of Hg processes. A major advance has been the development of a hierarchy of modeling tools that collapse the necessary detail from global simulations into more computationally feasible geochemical box models, enabling fully coupled simulations of the interactions among the land, atmosphere, and oceans over millennial time scales (Amos et al. 2013(Amos et al. , 2015. When combined with information on the cumulative history of human Hg release from antiquity to the present, this modeling approach has revealed a much greater contribution of human activity to the global Hg cycle than previously recognized (Streets et al. 2011(Streets et al. , 2017. The last 10 years has also seen rapid Fig. 1 Overview of global Hg cycling and impacts of policies and global change. Yellow numbers are estimated pool sizes in global reservoirs, and arrows indicate exchange processes between major environmental reservoirs. Pool sizes and exchanges are strongly modulated by anthropogenic emissions and releases (blue arrows) which transform in the atmosphere and deposit to aquatic and terrestrial ecosystems (green arrows). The Hg cycle will continue to experience disturbances due to changes in anthropogenic emissions and will be increasingly affected by land use and climate change, which remobilize (orange arrows) Hg that has accumulated in environmental reservoirs from previous Hg emissions and releases Ambio 2018, 47:116-140 development in Hg stable isotope biogeochemistry, providing a valuable tool to quantify Hg sources and study transformation processes. Incorporation of Hg stable isotopes in global models has the potential to constrain the relative importance of specific sources and processes (Sonke 2011;Sun et al. 2016a) (See Box 1).
These recent advances have proven particularly valuable for investigating the anticipated impacts of human and natural perturbations on global Hg cycling. Changes in anthropogenic emissions are ongoing and will continue into the future, including strong shifts in global source areas compared to current emission patterns (Giang et al. 2015). Accelerating land use and climate change are expected to have significant effects on global, regional, and local Hg cycles, with unexpected feedbacks and nonlinear impacts on Hg exposure. Models have been applied to assess the impact of regulatory interventions, such as emission controls (Selin et al. 2018), on specific outcomes and to evaluate policy efforts to mitigate Hg pollution (Selin 2014). An increasing number of studies are now available documenting such changes.
Here, we review the major insights gained from scientific advances over the past decade on global Hg cycling and Hg exchanges within and among the environmental compartments of the atmosphere, terrestrial ecosystems, and aquatic ecosystems. We then synthesize this knowledge to assess the impacts of human activities, including those affected by Hg-specific and other environmental policies, on the future of global Hg cycling. Further detailed discussion of factors affecting aquatic Hg loading, Hg methylation and demethylation processes, and exposure of wildlife and humans to Hg in the context of environmental change and disturbances are provided by Eagles-Smith et al. (2018) and Hsu- Kim et al. (2018), while relevant scientific insights for global policy are described in Selin et al. (2018).

RECENT ADVANCES IN UNDERSTANDING CRITICAL HG CYCLING PROCESSES OF GLOBAL IMPORTANCE
In this section, we discuss critical processes of importance for Hg cycling within (Fig. 2, blue arrows) and between (Fig. 2, red arrows) major environmental compartments (atmosphere, terrestrial, and aquatic [freshwater and ocean]).

Atmosphere
Atmospheric total gaseous mercury (TGM) concentrations have been measured since the late 1970s, with more reliable observations emerging since approximately 1990 (Slemr et al. 2003;Lindberg et al. 2007;Temme et al. 2007;Sprovieri et al. 2010;Slemr et al. 2011;Sprovieri et al. 2016). Mean TGM concentrations in background areas in the northern hemisphere, the tropics, and the southern hemisphere are 1.3-1.6, 1.1-1.3, and 0.8-1.1 ng/ m 3 , respectively (Slemr et al. 2011;Sprovieri et al. 2016). Many observations show that in the past three decades global background TGM concentrations have declined, particularly in North America, Europe, and over the North Atlantic Ocean (Slemr et al. 2003;Lindberg et al. 2007;Temme et al. 2007;Slemr et al. 2008;Ebinghaus et al. 2011;Slemr et al. 2011;Soerensen et al. 2012;Cole et al. 2013;Weigelt et al. 2015;Sprovieri et al. 2016), with rates of declines between -1.5 and -2.2% per year. Because anthropogenic emission inventories did not support such declines, proposed drivers for declines include decreased Hg concentrations and subsequent evasion in the upper North Atlantic Ocean and/or changes in terrestrial surfaceatmosphere fluxes (Slemr et al. 2011;Soerensen et al. 2012). Others have shown that declining point-source emissions of divalent Hg (Hg II ) in the northern hemisphere and from commercial products (Horowitz et al. 2014) were sufficient to reproduce observed declining trends in TGM from 1995 to present (Zhang et al. 2016d). TGM concentrations have decreased at lower rates in parts of the Arctic (approximately 0 to -0.2% per year) (Cole et al. 2013). Several studies have suggested that the recent reversal in global TGM trends in China and India are due to increasing emissions from coal-fired power plants in addition to artisanal and small-scale gold mining (ASGM) activity (Slemr et al. 2014;Martin et al. 2017;Streets et al. 2017). At Cape Point in South Africa, TGM has increased during the last decade, although a clear explanation for this trend is lacking (Martin et al. 2017).
The global atmospheric Hg reservoir is estimated to be between 4400 and 5300 Mg, and is enriched by more than an order of magnitude relative to natural levels and by approximately three-to fivefold relative to 1850 levels (Amos et al. 2013;Engstrom et al. 2014;Horowitz et al. 2017;Streets et al. 2017). The anthropogenic contribution to atmospheric Hg has been inferred from observed Hg deposition in polar firn air, peat bogs, and lake sediments, all of which reveal large-scale impacts of anthropogenic emissions, particularly in the second half of the twentieth century (Biester et al. 2007;Faïn et al. 2009;Farmer et al. 2009;Outridge et al. 2011;Goodsite et al. 2013;Engstrom et al. 2014;Amos et al. 2015;Blais et al. 2015;Drevnick et al. 2016). Combining deposition records with Hg stable isotope analysis now allows differentiation between atmospheric Hg II and Hg 0 deposition, and reconstruction of past Hg 0 levels (Enrico et al. 2016(Enrico et al. , 2017. Enrico et al. (2017) suggest that the maximum twentieth-century atmospheric Hg 0 concentration was 15 times larger than natural background at various locations, which agrees well with modeling analyses by Amos et al. (2015). Reconstructions suggest that historical atmospheric Hg 0 concentrations peaked around 1970 at 3-4 ng/m 3 and then declined to the current levels of 1.3-1.6 ng/m 3 (Faïn et al. 2009;Enrico et al. 2017), consistent with atmospheric Hg 0 concentration measurements of the last few decades. Lake sediment archives, reflecting total atmospheric Hg deposition, show similar time trends, with maximum Hg accumulation rates around 1970-1990(Engstrom et al. 2014Amos et al. 2015;Drevnick et al. 2016).
Direct anthropogenic Hg emissions account for approximately 30% of total annual inputs (i.e., total anthropogenic and natural emissions plus re-emission) to the atmosphere UNEP 2013;Streets et al. 2017). While several different emission inventories exist for the present day, few are comparable across the past decades (Muntean et al. 2014) or address all-time human emissions (Streets et al. 2011;Horowitz et al. 2014;Streets et al. 2017 Streets et al. (2011) found stationary fossil fuel combustion (SFFC), mainly coal combustion, to be the leading source of Hg (800-900 Mg/year), followed by ASGM (400 Mg/year). By assuming a higher activity level of unregulated and illegal ASGM in some countries, UNEP's (2013) global Hg assessment (AMAP/UNEP 2013) estimated ASGM emissions of more than 700 Mg/ year. SFFC accounted for less than 500 Mg/year of Hg in the UNEP inventory, while the EDGARv4 inventory (Muntean et al. 2014) estimated lower contributions from SFFC and ASGM. Given the global importance of ASGM releases, this source warrants further constraints. Nationallevel emission inventories rely on local field emission tests, with an increasing amount of data emerging from newly Fig. 2 Critical processes of global importance for Hg cycling, including fluxes between major environmental compartments. Perturbations of Hg processes and fluxes show anticipated impacts due to changes in emission, climate, and land use. A detailed discussion of the relevant processes is found in the ''Recent advances in understanding critical Hg cycling processes of global importance'' section, and major disturbances to global Hg cycling are discussed in the ''Anticipated impacts from human and natural perturbations, including emission changes and climate change, on global Hg cycling'' section industrialized countries. For example, Zhang et al. (2015a) estimated anthropogenic Hg emissions in China to be 538 Mg for 2010, somewhat lower than a previous estimate of 609 Mg/year . The discrepancy was attributed to better Hg removal efficiencies of air pollution control devices for nonferrous metal smelting (Wang et al. 2010;Zhang et al. 2012). Emissions in India were recently estimated at approximately 240 Mg/year Burger Chakraborty et al. 2013). National anthropogenic Hg emission estimates are available for several other countries (e.g., Kim et al. 2010;Masekoameng et al. 2010;Nansai et al. 2012;Nelson et al. 2012;US EPA 2015). As per the UNEP emission inventory, the eight highest emitting countries are (in order) China, India, Indonesia, Columbia, South Africa, Russia, Ghana, and the U.S. These countries emit a total of 1095 Mg/year of Hg, or 56% of the global anthropogenic Hg emissions, to the atmosphere (AMAP/UNEP 2013; UNEP 2013). More accurate emission inventories are expected in the future, as parties of the UNEP Minamata Convention on Hg improve national Hg use and emission inventories.
Uncertainties remain in global anthropogenic Hg emission inventories, mainly because of a lack of local field test data for key sectors such as cement production, iron and steel production, waste incineration, and gold production, and a lack of accurate activity-level data for ASGM and intentional Hg use. Horowitz et al. (2014) found that more than 400 Mg/year of atmospheric Hg emission was from the intentional use of Hg in products and processes (excluding gold production), previously unaccounted for in inventories. Top-down constraints and inverse modeling from atmospheric observations are currently being explored as additional tools to better constrain emissions (Song et al. 2015Denzler et al. 2017).
Hg speciation profiles of emissions are crucial to assess the environmental impacts of atmospheric emissions ). Streets et al. (2005) reported overall Hg 0 :Hg gas II :Hg particulate II ratios of 20:78:2 and 80:15:5 for Chinese coal-fired power plants (CFPPs) and cement plants (CPs), respectively, based on North American and European field test data. A recent study by Zhang et al. (2015a) updated the speciation profiles to 79:21:0 and 34:65:1 for Chinese CFPPs and CPs, respectively, based on onsite measurements. Such changes in speciation profiles for large point sources lead to large variations in Hg transport distances and local deposition fluxes.
Hg 0 oxidation to Hg II is considered a key step in removing Hg from the atmosphere (Selin et al. 2008;Lyman and Jaffe 2011;Subir et al. 2011), although increasing evidence shows that oxidized Hg II deposition is dominant only in oceans, while Hg 0 deposition is dominant in many terrestrial ecosystems (see below). A recent controversy has arisen around the techniques and artifacts in measuring atmospheric Hg oxidation and atmospheric Hg II concentrations. We refer to published reviews on atmospheric Hg and Hg II measurement methodologies (Lyman et al. 2010b;McClure et al. 2014a, b;Jaffe et al. 2014;Gustin et al. 2015;Angot et al. 2016;Mao et al. 2016;Marusczak et al. 2017). Earlier work suggested a primary role for the oxidants OH and O 3 in Hg 0 oxidation, but observations and theoretical calculations now suggest that the Br radical is likely the dominant oxidant (Hynes et al. 2009;Gratz et al. 2015;Shah et al. 2016). Current thinking is that the mechanism is a two-step process, where Hg 0 reacts with the Br radical to form HgBr, and HgBr reacts with multiple potential oxidants (Br,I,OH,BrO,IO,NO 2 ,etc.) to generate different Hg II species (Goodsite et al. 2004;Dibble et al. 2012;Wang et al. 2014;Ye et al. 2016;Horowitz et al. 2017;Jiao and Dibble 2017). Recent aircraft observations showed substantial atmospheric Hg II concentrations and detectable BrO in the subtropics, suggesting that subtropical anticyclones are significant global Hg II sources (Gratz et al. 2015;Shah et al. 2016). Photolytic formation of halogen radicals and Hg oxidation may occur at low light conditions, and recent Antarctic wintertime Hg oxidation may suggest a possible dark oxidation process (Nerentorp Mastromonaco et al. 2016). Hg oxidation chemistry has recently been updated in the GEOS-Chem model by Horowitz et al. (2017), which describes incloud Hg II photoreduction as potentially important in Hg redox chemistry Horowitz et al. 2017). A recent model comparison (Travnikov et al. 2017), not including the updated chemistry of Horowitz et al., showed that models with a Br oxidation mechanism reproduced the observed near-surface seasonal variations in the Hg II fraction, but did not capture seasonal variations in wet deposition observed at monitoring sites in North America and Europe. These findings suggest that more complex Hg oxidation chemistry and multi-oxidation pathways may be occurring in different regions of the atmosphere.

Terrestrial
In the terrestrial environment, the largest Hg pools are located in soils (Grigal 2003;Obrist 2012). Using global soil carbon inventories, Hg pools have been estimated at 300 Gg (Hararuk et al. 2013) and 240 Gg (Smith-Downey et al. 2010). The latter study estimated a 20% modern-day increase in overall soil storage and an even greater increase of Hg associated with the labile soil carbon fraction (Smith-Downey et al. 2010). Newer simulations using additional observational constraints (Streets et al. 2011;Horowitz et al. 2014) estimate higher present-day organic soil Hg pools (250-1000 Gg with a best estimate of 500 Gg) and propose that anthropogenic activities have doubled the Hg stored in organic soils (Amos et al. 2013(Amos et al. , 2015. Hg in organic-rich upper soils is predominantly from atmospheric deposition (Grigal et al. 2000;Schwesig and Matzner 2000;Guedron et al. 2006;Peña-Rodríguez et al. 2012;Demers et al. 2013;Peña-Rodríguez et al. 2014;Jiskra et al. 2015;Enrico et al. 2016;Zheng et al. 2016;Obrist et al. 2017;Wang et al. 2017). Atmospheric Hg deposits are retained in humus-rich upper soils and bound to organic matter (Meili 1991;Grigal 2003;Obrist et al. 2011;Jiskra et al. 2015). In mineral horizons, a significant component of Hg also stems from release from natural geologic sources.
New spatial soil datasets demonstrate that landscape Hg distribution is correlated to soil organic matter, latitude, annual precipitation, leaf area index, and vegetation greenness (Obrist et al. 2011;Richardson et al. 2013;Navrátil et al. 2014Navrátil et al. , 2016Obrist et al. 2016). Obrist et al. (2016) explained soil Hg variability across 1911 sites in the western U.S., largely by vegetation patterns, with high soil Hg accumulation in productive forests and 2.5 times lower soil concentrations in unproductive deserts and scrublands. Such patterns are due to a dominance of plant-funneled atmospheric Hg 0 deposition in many terrestrial ecosystems (see below), or by higher re-emission of Hg from deserts compared to forests (Eckley et al. 2016). Soil Hg accumulation and retention, however, are also determined by soil morphology and genesis as well as soil properties, including soil organic matter stability, content, texture, and pH (Obrist et al. 2011;Richardson and Friedland 2015;Navrátil et al. 2016). Landscape gradients with the highest inorganic Hg concentrations in forested watersheds have also been reported in river and lake sediments (Fleck et al. 2016), highlighting strong connectivity between upland Hg deposition and aquatic loading. This is consistent with earlier studies that linked high biological Hg hotspots in the eastern U.S. to watersheds with high forest densities (Driscoll et al. 2007).
Elevation gradients in Hg, largely associated with shifts in vegetation type, have also been observed, with vegetation and organic soil Hg concentrations increasing by up to fourfold with increasing elevation Townsend et al. 2014;Blackwell and Driscoll 2015;Wang et al. 2017). Blackwell and Driscoll (2015) reported a shift in deposition patterns along an alpine forest gradient, from deposition dominated by litterfall (i.e., uptake of Hg by plant foliage and transfer to ecosystems after senescence/ leaf shedding) to throughfall deposition (i.e., Hg uptake to plant surfaces and wash-off by precipitation) and deposition via cloudwater. They also noted that cloudwater accounted for up to 71% of total deposition at the highest altitudes. Hg stable isotope studies are inconsistent, reporting both higher contributions of Hg II at high-elevation sites Zheng et al. 2016) and lower contributions of Hg II with increasing altitude .

Aquatic
Oceans contain a substantial fraction of the global Hg reservoir and strongly affect atmospheric concentrations through air-sea exchange (Soerensen et al. 2010;Zhang et al. 2015b). Sunderland and Mason (2007) estimated the total ocean Hg to be 350 Gg with a 90% confidence limit of 270-450 Gg. More recent work Zhang et al. 2015b) suggests a reservoir of 260-280 Gg. Lamborg et al. (2014) estimated the upper ocean (top 1000 m) reservoir, based on observations from the GEO-TRACES cruise series, to be 63 Mg, which is on the low end of the range reported by Sunderland and Mason (2007) (63-120 Mg; 90% confidence interval), and may indicate either declining concentrations in some basins or improvements in analytical techniques (Lamborg et al. 2012). In the pelagic marine environment, trends in total Hg (THg) concentration are difficult to infer because of large inter-laboratory variability in measurements (Lamborg et al. 2012). Concentrations in the North Atlantic appear to be declining during the last several decades, from mean vertical profile concentrations of above 5 pM in the 1990s to consistently below 1 pM on recent cruises ). In the North Pacific, however, Sunderland et al. (2009) suggest that the THg concentration in the North Pacific Intermediate Water (NPIW) mass has increased since the mid-1980s due to enhanced atmospheric deposition from the Asian continent. Other cruises that sampled different locations (Hammerschmidt and Bowman 2012;Munson et al. 2015) did not see statistically different THg concentrations than those measured on earlier cruises, but current data coverage is sparse.
In freshwater ecosystems, predominant Hg sources include direct release from Hg-containing effluents, river runoff that contains atmospheric Hg deposits that accumulated in terrestrial environments, and direct atmospheric Hg deposition. The latter is particularly important in lakes with large surface area-to-volume ratios and small catchment-tolake surface areas. The relative importance of atmospheric and watershed Hg sources varies [see the ''Atmosphereaquatic interactions'' section and Hsu- Kim et al. (2018)] depending on the degree of development and land use change, hydrology, and dissolved organic carbon (DOC) content and composition (Engstrom and Swain 1997;Knightes et al. 2009;Lepak et al. 2015). Kocman et al. (2017) recently developed a first inventory of Hg inputs to freshwater ecosystems and estimated that 800-2200 Mg/ year of Hg enters freshwater ecosystems. This is lower than the amount of Hg entering coastal ecosystems estimated by Amos et al. (2014) (5500 ± 2700 Mg/year), in part because it does not include natural mobilization from terrestrial ecosystems. Earlier estimates of global riverine Hg release to the ocean were between 1000 and 2000 Mg/year (Cossa et al. 1997;Sunderland and Mason 2007). ASGM is the largest direct Hg source globally to both land and water (approximately 880 Mg/year); although the proportion released to aquatic ecosystems is not certain, it is estimated to be 50% (Kocman et al. 2017). Streets et al. (2017) suggest that approximately 40% of combined Hg releases to land and water globally are sequestered at the release site rather than traveling in rivers to the ocean. In India and China, industrial Hg sources (from the chlor-alkali industry, Hg and large-scale gold production, nonferrous metal production, and Hg-containing waste) release approximately 86 Mg/year to lakes and rivers, accounting for 51% of global aquatic releases from these sources (Kocman et al. 2017).
Methylmercury (MeHg) production is a critical process that occurs within aquatic ecosystems. The largest source of MeHg to freshwater lakes and wetlands is in situ microbial production, with surface sediments producing larger amounts than the water column. However, the relative contributions of water column versus sediment productions depend on specific characteristics of lakes such as stratification, depth of the anoxic hypolimnion, and organic carbon content of sediments. The activity of methylating microbes is controlled by temperature, redox conditions, pH, and the presence of suitable electron donors (e.g., organic carbon) and acceptors (e.g., sulfate, Fe III , methane). The primary controls on inorganic Hg bioavailability include DOC, sulfur, and Hg II concentrations and speciation (Benoit et al. 1999;Boening 2000;Ullrich et al. 2001;Gilmour et al. 2013;Hsu-Kim et al. 2013). A breakthrough paper by Parks et al. (2013) identified a two-gene cluster (hgcA and hgcB) in microbes involved in Hg methylation. The hgcA gene encodes a putative corrinoid protein capable of transferring a methyl group to Hg II , and the HgcB protein returns HgcA to a redox state that enables it to receive a new methyl group. Subsequent work identified these genes in many organisms, including sulfate and iron-reducing bacteria and methanogens Yu et al. 2013;Podar et al. 2015). Several studies have developed hgcA probes to screen aquatic sediments for Hg-methylating microorganisms (Bae et al. 2014;Schaefer et al. 2014;Bravo et al. 2015;Du et al. 2017) and to demonstrate that different methylating microbes inhabit various niches, such as regions of wetlands with varying sulfate concentrations (Bae et al. 2014;Schaefer et al. 2014). Recent development of a broad range of probes spanning all known hgcAB genes (Christensen et al. 2016) may allow future quantification of the methylating potential in environmental samples.
One conundrum that remains is how to explain high levels of MeHg in marine seawater where MeHg production is still poorly understood. Gionfriddo et al. (2016) found that the micro-aerophilic, nitrite-oxidizing bacterium Nitrospina, present in Antarctic sea ice and brine, contains genes that are slight rearrangements of the hgcAB genes, suggesting that diverse microbial communities may be capable of methylation. Fitzgerald et al. (2007) suggested that most marine MeHg is produced in ocean margin sediments, but a variety of studies across most major ocean basins have since produced strong evidence for water column MeHg production (Cossa et al. 2009;Sunderland et al. 2009;Heimbürger et al. 2010;Cossa et al. 2011;Lehnherr et al. 2011;Blum et al. 2013). In the last several years, the abundance of ocean water concentration measurements has increased dramatically with new data from the North Pacific (Kim et al. 2017); North Atlantic (Bowman et al. 2014); Equatorial, South Atlantic, and Pacific (Munson et al. 2015;Bowman et al. 2016); Arctic (Heimburger et al. 2015); and Antarctic Oceans (Gionfriddo et al. 2016). In the coastal marine environment, Schartup et al. (2015) used enriched Hg isotope incubation to detect active Hg methylation for the first time in oxic estuarine seawater. Methylation was facilitated by the presence of labile terrestrial DOC and a shift in ionic strength and microbial activity that accompanies the transition into saline waters in estuaries. Similarly, Ortiz et al. (2015) measured active methylation in laboratory experiments with marine snow aggregates.
Marine species are responsible for a large fraction of human exposure to MeHg; for example, Sunderland (2007) found that for the U.S. more than 90% of MeHg exposure is from marine and estuarine species. Thus, understanding MeHg dynamics in marine environments is particularly important for public health (Gribble et al. 2016). Advances have been made in understanding MeHg uptake to aquatic food webs (see Eagles-Smith et al. 2018). The photochemical demethylation of MeHg has been shown experimentally to result in a large enrichment of odd-massnumbered Hg isotopes (D 199 Hg; Fig. 3, Box 1) in the residual MeHg pool that accumulated through the food web (Bergquist and . Odd-mass-independent fractionation (MIF) anomalies in aquatic organisms were used as source tracers to distinguish between MeHg derived from sediments with small odd-MIF isotopes and MeHg from the open ocean with large odd-MIF isotopes (Senn et al. 2010;Blum et al. 2013). Odd-MIF anomalies in biota have been hypothesized to serve as source tracers and as a proxy for ecological parameters such as foraging depth; for example, the largest odd-MIF anomalies are found in fish feeding in surface waters where photochemical demethylation is most active . The extent of odd-MIF anomaly is reduced in aquatic organisms where light penetration in the aquatic water column is inhibited [e.g., by sea ice (Point et al. 2011;Masbou et al. 2015)] or with high DOC (Sherman and Blum 2013). Thus, odd-MIF anomalies can track changes in climatological/environmental parameters such as sea-ice cover. Field measurements and experimental data suggest that DOC controls MeHg availability for uptake in freshwater. French et al. (2014) observed a water column DOC threshold of 8.6-8.8 mg/L, above which MeHg uptake was inhibited in Arctic lakes, while Jonsson et al. (2014) showed that in estuarine microcosms DOC-bound MeHg was 5-250 times more available for bioaccumulation into lower food web organisms than newly produced MeHg.

Atmosphere-terrestrial interactions
Large uncertainties in terrestrial Hg 0 sinks and source strengths, particularly over background areas and vegetated areas, make atmosphere-terrestrial interactions one of least constrained processes in the global Hg cycle. Major emissions from terrestrial ecosystems include biomass burning and volatilization of Hg 0 from natural and Hgenriched soils, while deposition from the atmosphere includes dry Hg 0 deposition to plants and soils and wet Hg II deposition.
Recent    2012; Kolka et al. 2017). Wildfire studies consistently show that fires mobilize Hg primarily from biomass, litter, and organic soils (Friedli et al. 2001(Friedli et al. , 2003Engle et al. 2006;Biswas et al. 2007;Obrist 2007;Wiedinmyer and Friedli 2007;Biswas et al. 2008;Burke et al. 2010;Woodruff and Cannon 2010), although Hg released from upper mineral soils can comprise up to 10% of total emissions (Biswas et al. 2007;Burke et al. 2010;Melendez-Perez et al. 2014). Significant post-fire soil evasion of Hg 0 has been reported to occur years after fires, leading to an additional 50% in Hg emissions (Melendez-Perez et al. 2014), an effect which is not included in wildfire emission estimates.
A relatively well-constrained component of the atmosphere-terrestrial exchange is wet deposition of Hg II , at least over land where extensive monitoring networks exist, including the Mercury Deposition Network (MDN) for the U.S. and Canada (Prestbo and Gay 2009;Weiss-Penzias et al. 2016;Sprovieri et al. 2017), the European Monitoring and Evaluation Programme (EMEP) for Europe (Tørseth et al. 2012), and a new Asia-Pacific Mercury Monitoring Network (2017) and the Global Mercury Observation System (GMOS) for other parts of the world. At background sites, mean atmospheric Hg wet deposition globally decreased from 10-15 lg/m 2 /year in the early 1990s to the current level of 5-10 lg/m 2 /year (Zhang et al. 2016d). Declines in atmospheric Hg II wet deposition are reported from Europe and the U.S. with average deposition declines of -1.5% per year and -2.3% per year between 1990 and 2013, respectively (Zhang et al. 2016d), although increasing wet Hg II deposition has been reported in East Asia. Mean background Hg wet deposition fluxes in the northern hemisphere, the tropics, and the southern hemisphere are approximately 5-10, 2-8, and 1-5 lg/m 2 /year, respectively (Prestbo and Gay 2009;Fu et al. 2016a;Sprovieri et al. 2016Sprovieri et al. , 2017. There is, however, large spatial and temporal variability in wet Hg II deposition (Butler et al. 2008;Prestbo and Gay 2009;Risch et al. 2012b;Zhang and Jaeglé 2013;Cole et al. 2014;Muntean et al. 2014;Sprovieri et al. 2016).
Estimated median dry deposition fluxes of Hg gaseous II and Hg particulate II in Asia and North America are 10.7 and 6.1 lg/ m 2 /year, respectively. A particular challenge with dry Hg II deposition is that measurements and modeling schemes have uncertainty factors greater than two (Wright et al. 2016), in addition to large measurement uncertainties (see the ''Atmosphere'' section). Atmospheric dry deposition flux measurements can be broadly divided into micrometeorological methods based on gradient measurements, dynamic gas flux chambers, and surrogate surface passive Box 1 Use of stable isotopes to constrain Hg sources and processes The seven stable isotopes of Hg (196, 198, 199, 200, 201, 202, and 204) are subject to different mass-dependent (MDF) and mass-independent (MIF) isotope fractionation mechanisms. The isotopic signature of Hg in a sample carries information on its initial source and the transformation processes it underwent. Here we provide a simplified three-dimensional view of Hg stable isotope systematics and focus on exemplary processes and sources that have pronounced identifiable effects on Hg isotope signatures in environmental samples. For a more comprehensive introduction to Hg isotope fractionation and a compilation of Hg isotope variation in natural samples, we refer to the reviews of Blum et al. (2014) and Blum and Johnson (2017).
MDF (mass-dependent fractionation) is reported as d 202 Hg in the ''small delta'' notation, representing the 202/198 Hg ratio of a sample relative to the NIST SRM3133 international standard as defined by Blum and Bergquist (2007). Primary natural and anthropogenic Hg emission sources are generally associated with small negative d 202 Hg signatures, whereas atmospheric Hg pools have positive d 202 Hg values (Blum et al. 2014;Sun et al. 2016b). All biotic and abiotic transformation processes enrich Hg stable isotopes to different extents by MDF. One of the largest MDF fractionations observed is associated with preferential uptake of light atmospheric Hg 0 by foliage (arrow 1 in Fig. 3) (Demers et al. 2013;Enrico et al. 2016;Yu et al. 2016).
Odd-MIF (mass-independent fractionation of odd-mass-number Hg isotopes) is reported as D 199 Hg and D 201 Hg in the ''capital delta'' notation, representing the anomalies of the 199/198 Hg and 201/198 Hg ratios from the MDF line, respectively (Blum and Bergquist 2007). Two independent mechanisms have been experimentally shown to cause odd-MIF: (1) magnetic isotope effects (MIE) (Buchachenko 2001(Buchachenko , 2009 and (2)  samplers, which have been more widely used in recent years (Lyman et al. 2009;Zhang et al. 2009;Lyman et al. 2010a;Gustin et al. 2011;Pirrone et al. 2013;Gustin et al. 2015;Huang and Gustin 2015). To model deposition, resistance approaches are used in chemical transport models that apply dry deposition velocities for Hg II and measured or estimated Hg II concentrations (Lin et al. 2006;Zhang et al. 2009). For simulations of net Hg 0 deposition, bidirectional soil-vegetation-air exchange models are needed (Bash et al. 2007;Zhang et al. 2009); current models are not well constrained. Models have been used to estimate total Hg deposition compared with existing measurements and to attribute deposition to sources in the context of policy. Different models show similar global Hg deposition patterns, especially near source regions, and can reproduce major features of observed wet deposition (Travnikov et al. 2017). However, in some regions, there are substantial differences in simulated attributions of Hg deposition to local versus international sources. Atmospheric Hg 0 emissions from land surfaces was previously estimated between 1600 and 2900 Mg/year (Lindberg et al. 2007;Selin et al. 2008 et al. 2014, 2015); one global atmospheric model simulation  showed reduced global soil evasions from 2200 to 1360 Mg/year, which is still in the higher range based on flux measurements by Agnan et al. (2016).
Over land, Hg II deposition is minor compared to Hg 0 deposition, and the dominant pathway of atmospheric deposition to terrestrial ecosystems is through litterfall. Globally, this deposition pathway is estimated at 1020-1230 Mg (Obrist 2007;Risch et al. 2012a;Fu et al. 2016a, b;Wang et al. 2016;Wright et al. 2016;Zhang et al. 2016c). In addition, throughfall deposition may contribute an additional 50% of litterfall deposition , which would result in a global deposition flux of approximately 560 Mg/year, although this needs further observational confirmation. In addition, further plant depositions occur via woody tissues, including by tree blowdown (Mitchell et al. 2012). Given that Hg contained in woody biomass can exceed Hg contained in foliage biomass by 67-100% Yang et al. 2017), woody tissue deposition is likely an important, yet not well-constrained, source of deposition as well (Obrist 2007). Plant contributions (litterfall plus throughfall wet deposition) dominate deposition loads in forests globally (70-85%), of which approximately 75% (with the exception of South American rainforests) is litterfall . Wright et al. (2016) and Wang et al. (2016) show large continental differences with high litterfall fluxes in the Amazon and in China that are greater than fivefold those in North America and Europe. This suggests that deposition fluxes decrease spatially from tropical to temperate/boreal regions, with 30% of total deposition in the temperate and boreal regions and 70% of total deposition in the tropical/subtropical regions. Unfortunately, there is a lack of plant and associated Hg litterfall and throughfall deposition data outside of forest ecosystems, such that estimates for global grasslands, savannas, and shrublands remain highly uncertain. Hg stable isotope approaches now suggest that Hg 0 deposited via plants comprises the majority of Hg deposition to soils in central North America (57-94%) (Demers et al. 2013;Zheng et al. 2016), Alaskan tundra soils (71%) , central European peat soils (79%) (Enrico et al. 2016), boreal forest soils in North Sweden (90%) (Jiskra et al. 2015), and high mountain soils in Tibet . This is likely also true for other areas of the world, for example, the tropics, where some of the highest litterfall Hg deposition is observed . Recent field studies also show evidence for an active Hg 0 sink in soils (Sigler and Lee 2006;Moore and Castro 2012;Obrist et al. 2014Obrist et al. , 2017 and possibly direct deposition of Hg 0 to forest litter . The finding that litterfall is the dominant terrestrial Hg 0 source fundamentally reshapes our understanding of Hg deposition to terrestrial landscapes and Hg mobility, demonstrating the need for monitoring strategies to measure net ecosystem Hg 0 deposition, including direct exchange/deposition measurements. The dominant role of foliar Hg 0 uptake also results in a specific ''terrestrial'' Hg isotope signature that can be used to trace terrestrial contributions and transfer to aquatic ecosystems (Tsui et al. 2012;Kwon et al. 2015;Jiskra et al. 2017), aquatic biota (Li et al. 2016), and lake and ocean sediments (Grasby et al. 2015;Lepak et al. 2015;Enrico et al. 2016;Araujo et al. 2017;Gleason et al. 2017). Finally, wholeecosystem flux measurements using micrometeorological techniques can further confirm the dominant Hg 0 deposition source to terrestrial ecosystems. The longest data set spans 2 years in the Arctic tundra ) and shows that Hg 0 deposition (6.5 ± 0.7 lg/m 2 /year) dominates other deposition processes (71% of total deposition), in agreement with Hg stable isotope analysis. Previous studies over grasslands also reported a net annual dry Hg 0 deposition [17.5 lg/m 2 /year (Fritsche et al. 2008a) and 3.3 lg/m 2 /year (Castro and Moore 2016)] and shorter-term measurements showed a net Hg 0 deposition during vegetation periods (Lee et al. 2000;Bash and Miller 2008;Fritsche et al. 2008b;Bash and Miller 2009;Sommar et al. 2016). Other studies, however, have reported inconsistent results, with net terrestrial Hg 0 emissions even in the presence of vegetation (Lindberg et al. 1998;Cobos et al. 2002;Obrist et al. 2005;Cobbett and Van Heyst 2007;Baya and Van Heyst 2010).

Atmosphere-aquatic interactions
A new global model suggests that approximately 50% of total global wet Hg II deposition may occur to tropical oceans within 30°northern and southern latitudes due to a high abundance of free tropospheric Hg II reservoirs scavenged by precipitation in these regions . These important global hotspot regions for Hg II deposition need to be confirmed by direct deposition measurements.
Air-sea exchange of Hg 0 is critical to extending the lifetime of anthropogenic Hg in the atmosphere, terrestrial ecosystems, and the ocean Soerensen et al. 2010). The magnitude of net Hg 0 evasion from the ocean is approximately 2900 Mg/year (range 1900-4200 Mg/year) Streets et al. 2017), compared to approximately 2000 Mg/year from primary anthropogenic emissions; thus, the ocean is considered the largest source of atmospheric Hg globally. There is, however, substantial uncertainty in the magnitude of air-sea exchange from different ocean regions. Simultaneous measurements of dissolved Hg 0 in seawater and atmospheric concentrations are used to assess this source, with recent latitudinal gradients in both the Atlantic and Pacific Oceans showing distinct spikes in concentrations around the Intertropical Convergence Zone (ITCZ). These spikes are thought to result from deep convection and intense precipitation in these regions that increase atmospheric Hg II deposition (Soerensen et al. 2014). Using high-frequency measurements of atmospheric and aquatic Hg 0 concentrations, Soerensen et al. (2013) inferred that terrestrial DOC can inhibit Hg II reduction and ambient Hg 0 production in seawater. Lamborg et al. (2016) measured concentrations of Hg on settling organic particulate matter in the Atlantic Ocean and suggested that partitioning to particles may be up to 1000 times greater than assumed in some models. This would effectively lower the magnitude of the global evasion flux of Hg 0 from the ocean because scavenging of Hg from surface waters would reduce the pool of Hg II available for reduction and conversion to Hg 0 .

Terrestrial-aquatic interactions
In watersheds of the Americas, with historic Hg and precious metal mining, runoff from contaminated sites dominates aquatic Hg loads and is substantial (44 Mg/year) (Kocman et al. 2017). In contrast, in developed regions, including Europe and North America, catchment inputs and atmospheric deposition are generally the primary Hg sources, with background terrestrial environments accounting for approximately 230 Mg/year (170-300 Mg) of riverine loads. Hydrology and DOC mobility are the major controls on delivery of catchment Hg to aquatic ecosystems, with episodic high-flow periods from storms and/or snowmelt often representing the majority of annual catchment inputs .
Recent Hg stable isotope analyses have been able to differentiate among inorganic Hg sources to freshwater. The waters of remote Canadian lakes share the positive D 200 Hg signature observed in precipitation, suggesting that atmospheric deposition to these lakes is the major Hg source (42 ± 26%) . Similarly, analyses of sediment of the Laurentian Great Lakes indicate that atmospheric Hg II deposition is the dominant Hg source to Lakes Huron, Superior, and Michigan, while terrestrial catchment and industrial effluent are the dominant Hg sources to Lakes Erie and Ontario (Jackson et al. 2015;Lepak et al. 2015). Contrasting results have been reported for a boreal forest catchment in Sweden, where Hg stable isotope analyses suggest that Hg in forest runoff originated from the deposition of Hg 0 through foliar uptake, rather than from precipitation ). Analyses of dated lake sediment cores and catchment soils using geochemical tracers can also be used to tease apart changes in atmospheric and terrestrial Hg II inputs over time (Fitzgerald et al. 2005;Perry et al. 2005;Landers et al. 2008;Muir et al. 2009;Kirk et al. 2011;Hermanns and Biester 2013). Recent analyses of Hg stable isotopes in dated lake sediment cores have also been used to examine temporal changes in Hg sources (Sonke et al. 2010;Gehrke et al. 2011;Cooke et al. 2013;Donovan et al. 2013).
For pelagic ocean regions, the dominant source of Hg is atmospheric deposition (Soerensen et al. 2010). Rivers comprise only a small fraction of Hg input to most ocean basins because the majority of Hg is in the particle phase and settles in ocean margins. Zhang et al. (2015b) showed that the fraction of riverine Hg exported to the open ocean is 6-25% for all basins except the Arctic. The large shelf and relatively smaller surface area of the Arctic Ocean make riverine inputs of Hg a more important source. For example, modeling studies have used atmospheric observations to infer a large missing Hg source from rivers and coastal erosion (Fisher et al. 2012;Zhang et al. 2015b). Other work using measured Hg:DOC ratios and a synthesis of available observations suggests a riverine Hg source to the Arctic of approximately 50 Mg/year (Dastoor et al. 2015;Zhang et al. 2015b), which is slightly larger than atmospheric deposition of Hg (approximately 45 Mg/year).

ANTICIPATED IMPACTS FROM HUMAN AND NATURAL PERTURBATIONS, INCLUDING EMISSION CHANGES AND CLIMATE CHANGE, ON GLOBAL HG CYCLING
As illustrated in Fig. 2, drivers of perturbations on the global Hg cycle include emission, climate, and land use changes.

Atmosphere
Critical factors affecting future anthropogenic emissions include energy use and Hg emission control strategies [e.g., in response to the Minamata Convention, see Selin et al. 2018 andGiang et al. 2015] and mitigation measures for global climate change. One of the largest anthropogenic sources of Hg, ASGM emissions, will critically affect future Hg emissions, but is also associated with the largest uncertainties in future emission estimates. Streets et al. (2009) Giang et al. (2015) suggested 90 and 150 Mg/year of avoided power sector emissions for China and India in 2050, respectively, due to Minamata Convention control technology requirements. Declining Hg emissions in China through 2050 could, however, be offset by increasing Indian emissions (Giang et al. 2015), leading to a shift in source areas globally toward the tropics. Rafaj et al. (2013) evaluated the cobenefit of global greenhouse gas mitigation efforts and found that anthropogenic Hg emissions under a global climate mitigation regime would be reduced in 2050 by 45% compared to the case without climate measures. Pacyna et al. (2016) estimated that a best-case scenario of Hg emission decreases by 2050 could result in a decrease in Hg of up to 50% in the northern hemisphere and 35% in the southern hemisphere. Lei et al. (2014) compared the influence of projected emission changes and climate change on U.S. atmospheric Hg levels in 2050. They found that anthropogenic emissions would contribute 32-53% of projected changes in Hg air concentration, while climate and natural emission changes would account for 47-68%. Of the effects of climate change, influences via secondary emissions and ecosystem structure (see the ''Atmosphere-terrestrial interactions'' section) will likely exert the largest changes on atmospheric Hg pools. Although less important, atmospheric processes can also be directly impacted by climate change; for example, the oxidation rate constant of the Hg-Br reaction is sensitive to temperature. Zhang et al. (2016a) estimated a 5-7% increase in northern hemisphere atmospheric surface Hg 0 concentrations by 2050 because of reduced oxidation and increased in-cloud reduction. Changes in precipitation patterns can change spatial distribution, magnitude, and seasonal variation of Hg II deposition. Megaritis et al. (2014) projected increases in Hg II deposition due to climate change to the eastern U.S. of 6% during summer and 4% during winter by 2050, with significant spatial variability. Zhang et al. (2016a) predicted that land use and land cover changes may lead to increases in Hg 0 dry deposition fluxes under future climate. Hansen et al. (2015) projected climate-induced decreases in Hg deposition to the Arctic, while Zhang et al. (2016a) estimated increases in northern high latitudes. Large-scale meteorological patterns such as El Niño-Southern Oscillation may also influence interannual variability of atmospheric Hg concentrations (Slemr et al. 2016).

Terrestrial
Major impacts on global Hg cycling are expected due to shifts in global biomes, hydrology, fire patterns, water table depth, soil moisture, and redox conditions. In addition, forestry practices and deforestation can affect watershed Hg processes and MeHg exposures and biomagnification, although responses can be variable between different watersheds (discussed in detail in Hsu- Kim et al. (2018)). Based on the importance of vegetation for atmospheric Hg deposition, predicted shifts in biomes will directly impact terrestrial and aquatic Hg distribution and impacts. Modeling studies have predicted a strong sensitivity of Hg 0 dry deposition to changes in vegetation (Krabbenhoft and Sunderland 2013;Zhang et al. 2016a). Zhang et al. (2016a) projected that by 2050 the annual mean Hg 0 dry deposition flux over land will increase by 20% in northern mid-latitudes, driven by a combination of increased atmospheric Hg 0 exposure and increased vegetation and foliage density induced by CO 2 fertilization. Regionally decreased Hg 0 dry deposition may occur in South Asia and Africa, for example, where projected increases of agricultural land area will lead to losses of leafy areas. Strong shifts in soil Hg accumulation have been predicted as well. Hararuk et al. (2013) suggested that, along with changes in soil organic carbon, soil Hg sequestration in the U.S. will increase by 2100, mainly induced by changes in precipitation patterns. They also predicted that increases in CO 2 concentrations to 700 ppm would stimulate both soil carbon and soil Hg accruals, while increased air temperatures would have small negative effects on both. The combined effects of increased CO 2 , increased temperature, and increased or decreased precipitation will likely lead to pronounced regional differences in soil Hg changes. Anticipated shifts in coniferous versus deciduous forest abundance will also likely impact terrestrial Hg distribution. Richardson and Friedland (2015) suggested that anticipated losses of 2.2 million ha of coniferous forests will lead to 27 Mg less Hg sequestered in organic horizons across this region, either as a function of reduced atmospheric Hg deposition or increased release into watersheds. These anticipated effects of climate change and ecosystem properties on Hg cycling can be seen clearly in archive studies. In a remote lake in Patagonia, for example, pre-anthropogenic changes in sediment loads up to a factor of four were found, comparable to recent anthropogenic forcing (Hermanns et al. 2012), and were attributed to past changes in deposition related to climate-induced shifts in terrestrial organic matter flux and DOM leaching from catchment soils. Similarly, Rydberg et al. (2015) documented how a climate-induced change from deciduous to coniferous forests around 5700 years BP led to reduced soil erosion and increased transport of DOC-bound Hg from a catchment in the area of Germany, resulting in increased Hg concentrations and accumulation rates.
Other ecosystem disturbances are expected to affect Hg cycling, but in unknown directions. For example, changes in wildfire frequency and abundance will impact atmospheric Hg emissions and watershed fluxes in a nonlinear fashion given the complex relationships between Hg emissions and fire intensities, pre-fire Hg accumulation, and post-fire Hg mobilization. Wildfire Hg emissions now likely include substantial emissions of anthropogenic legacy Hg accumulated in litter and soils, and climate and land use change impacts on fire frequency and activity need to be considered as part of human impacts (Westerling et al. 2006). A recent modeling study (Kumar et al. 2017) estimated a 14% increase in wildfire-related Hg emission due to climate change. Corresponding land use changes may amplify or alleviate this effect, e.g., due to reductions in forest fires by agricultural land expansion in Africa or increased forest abundance in North America and Asia. In addition, land management practices such as burning of agricultural fields, grassland, and brush fires contribute to Hg emissions, although little information is available on their global contribution. Land use changes such as shifts in land management and forestry practices also have the potential to mobilize terrestrial Hg pools, via increased erosion, changes in hydrological pathways, and changes in yields (Kocman et al. 2017;Kronberg et al. 2016). Kocman et al. (2017) summarized studies that show pronounced Hg leaching when soils are converted to agricultural land and predicted that global deforestation of 13 million ha/yr in the 2000s may have yielded up to 260 Mg Hg per year entering local freshwaters.

Aquatic systems
Both oceans and freshwater ecosystems will be affected by changing Hg emissions and climate-induced alterations. Modeling studies show that response times for lakes can range from a few years to many decades (Harris et al. 2007;Knightes et al. 2009). Recovery of freshwater ecosystems to decreased atmospheric Hg emissions is expected when reduced atmospheric Hg loadings translates into a substantial decrease in DOC-bound MeHg and inorganic Hg catchment runoff Chiasson-Gould et al. 2014;French et al. 2014;Jonsson et al. 2014;Lepak et al. 2015;Ndu et al. 2015). A recent long-term mass balance study in New York state suggested that Hg inputs to lakes from forested catchments could be very responsive to decreased Hg emissions (Gerson and Driscoll 2016). They reported that a 25% reduction in atmospheric Hg 0 concentration resulted in a 40% reduction in litterfall Hg fluxes, which in turn decreased inorganic Hg and MeHg stream inputs (45 and 29%, respectively) and lake concentrations over a 10-year period (Gerson and Driscoll 2016).
A common disturbance of freshwater Hg cycling results from widespread reservoir creation, which generally amplifies MeHg production; reservoir effects are discussed in detail in Eagles-Smith et al. (2018) and Hsu- Kim et al. (2018). Most studies of climate change impacts on freshwater Hg cycling have been conducted in sensitive ecosystems experiencing accelerated changes (e.g., highaltitude and high-latitude lakes, ponds, and peatlands). In the western Canadian Arctic, where approximately 15% of the water bodies have shoreline retrogressive thaw slump lakes, lake sediment cores demonstrated increased sedimentation of inorganics and decreased sediment concentrations of DOC, THg, and MeHg (Deison et al. 2012). In contrast, a survey of MeHg accumulation in invertebrates in tundra lakes suggested that lakes with low DOC are experiencing increased DOC input because of humification and these lakes may see a two-to threefold increase in Hg bioaccumulation as the DOC concentration approaches a threshold of approximately 8 mg/L (MacMillan et al. 2015). Jonsson et al. (2017) predicted a three-to sixfold increase in MeHg concentration in estuarine zooplankton in association with large inputs of terrestrial DOC intended to mimic potential future changes in runoff. Peat mesocosm experiments suggest that changes in hydrological regimes and shifts in vascular plant communities may have a significant impact on Hg cycling in peatlands (Haynes 2017). For example, lower, more variable water tables and the removal of Ericaceae shrubs significantly enhanced inorganic Hg and MeHg mobility in peat pore waters and MeHg export from snowmelt, likely from enhanced peat decomposition and internal regeneration of electron acceptors related to water table changes .
The effect on Hg of large-scale changes in the marine environment expected from climate change was explored by Krabbenhoft and Sunderland (2013). Increased seawater temperatures may enhance organic matter remineralization and the propensity for MeHg production in some regions of the ocean. Rapid changes in sea-ice cover and seasonality in the polar oceans are likely to exert a major impact on air-sea exchange of Hg in these regions, and further work may be needed to understand how temperature and sea-ice dynamics alter Hg dynamics in the Arctic (Angot et al. 2016). Changes in microbial community structure and ocean productivity will propagate through marine food webs in potentially unexpected ways, altering bioaccumulation. Fisher et al. (2013) used 30 years of satellite data to quantify the impacts of interannual variability in meteorology on inorganic Hg dynamics in the Arctic. They concluded that the dominant climate mode of the future in the Arctic may result in a lower reservoir of Hg in the Arctic Ocean because of enhanced air-sea exchange. This work did not consider the effects of a changing (melting) terrestrial landscape, which may dramatically increase Hg input to the Arctic from the terrestrial environment. Soerensen et al. (2016) suggested that rivers are likely to become more dominant sources of Hg and MeHg in the future Arctic Ocean. Stern et al. (2012) reviewed a variety of changes that are occurring and expected to occur in the Arctic Ocean in the future, and suggested that future effects are likely to be large; they also highlighted current limitations in understanding multiple, simultaneously occurring impacts on Hg cycling and bioaccumulation. Finally, recent modeling suggests that the response times of marine fish tissue burdens to changes in Hg input will depend on the locations of MeHg production (Li et al. 2016). Benthic sediment will respond much more slowly than the upper Ocean and estuarine surface waters Schartup et al. 2015;Soerensen et al. 2016). Amos et al. (2013) predicted that legacy Hg will be present in global ecosystems for periods ranging from decades to millennia, suggesting a large time lag in response to changes in anthropogenic emissions and climate.

CONCLUSIONS AND REMAINING UNCERTAINTIES
We reviewed how our understanding of Hg global cycling has advanced in the last decade, focusing on environmental reservoirs and processes within and between these reservoirs. With emerging large global datasets, in combination with improved models and analytical techniques, new constraints are possible on the magnitude of reservoirs and fluxes. In the atmosphere, while Hg 0 concentrations and wet deposition are decreasing in some parts of Europe and North America, other regions such as Asia show increasing trends. Updated knowledge about the presence and chemistry of Hg II , especially in the free troposphere, suggests a source of global Hg to the tropics. In terrestrial ecosystems, updated information gained from new measurements and models has changed our understanding of Hg exchange between terrestrial systems and the ocean. Previous estimates of a large terrestrial source of Hg to the atmosphere have been revised downward, and new data suggest that Hg 0 uptake is dominant in natural systems. In aquatic systems, recent research has shown the potential importance of direct Hg releases to water in both freshwater and marine systems. In oceans, new constraints are available on the total amount of Hg in ocean systems and its distribution and trends, with declines in the Atlantic and possible increases in the Pacific.
Global changes, including climate, land use, and Hg emissions changes, are likely to profoundly alter Hg cycling. New analyses have indicated the direction of these possible changes. Global policy such as that emerging from the Minamata Convention may avoid significant increases in Hg concentrations associated with future economic growth; however, legacy emissions will continue to affect the global cycle for decades to centuries. Projections of future conditions are becoming more quantitative, supported by increasingly large and globally linked datasets, new tools, and better models. Global changes are also expected to affect concentrations of MeHg in potentially nonlinear ways, including changes in methylation and food web bioaccumulation [see Eagles-Smith et al. (2018) and Hsu-Kim et al. (2018)].
There remain uncertainties that limit our ability to predict future global cycling and therefore human exposure, especially in the context of multiple and simultaneous changes. While data availability has increased, several regions of the globe, especially the tropics and the southern hemisphere, remain understudied. In the atmosphere, uncertainties in atmospheric chemistry and in measurements of Hg II challenge our quantitative understanding of oxidation pathways. Advances and wider applications in isotope techniques are needed to better fingerprint specific sources and mechanisms. In terrestrial ecosystems, timeextended measurements of Hg 0 uptake, now known to be a significant input process, are largely lacking. Finally, the global Hg cycle will also depend on the choices made by governments both now and in the future, including policy efforts to mitigate emissions of Hg and control climate change.
Acknowledgements This paper was written as part of the synthesis effort for the 13th International Conference on Mercury as a Global Pollutant, Providence, Rhode Island, July 16-21, 2017. We thank Dan Jaffe, Jeroen E. Sonke, Xinbin Feng, Robert Mason, David Krabbenhoft, Alexandra Steffen, and Jason Stowe for valuable inputs, technical reviews, and comments on earlier drafts of this paper. Grant support was provided by the Dartmouth College Toxic Metals Superfund Research Program through funds from the National Institute of Environmental Health Sciences of the National Institutes of Health under Award Numbers R13 ES028077-01 and P42ES007373. Support was also provided by US National Science Foundation (NSF) award PLR #1304305) and NSF CNH #1313755 and by a H2020 Marie Sklodowska-Curie Grant Agreement #657195. The content of this paper is solely the responsibility of the authors and does not necessarily represent the official views of the funding agencies.
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