Mercury contaminated sediment sites—An evaluation of remedial options

Environmental Research 125 (2013) 131–149

Contents lists available at ScienceDirect

Environmental Research journal homepage: www.elsevier.com/locate/envres

Mercury contaminated sediment sites—An evaluation of remedial options Paul M. Randall a,n, Sandip Chattopadhyay b a

U.S. Environmental Protection Agency, Office of Research and Development, National Risk Management Research Laboratory, 26 West Martin Luther King Drive, Cincinnati, OH 45268, USA Tetra Tech, Inc., 250 West Court Street, Suite 200W, Cincinnati, OH 45202, USA

b

a r t i c l e i n f o

abstract

Available online 12 March 2013

Mercury (Hg) is a naturally-occurring element that is ubiquitous in the aquatic environment. Though efforts have been made in recent years to decrease Hg emissions, historically-emitted Hg can be retained in the sediments of aquatic bodies where they may be slowly converted to methylmercury (MeHg). Consequently, Hg in historically-contaminated sediments can result in high levels of significant exposure for aquatic species, wildlife and human populations consuming fish. Even if source control of contaminated wastewater is achievable, it may take a very long time, perhaps decades, for Hg-contaminated aquatic systems to reach relatively safe Hg levels in both water and surface sediment naturally. It may take even longer if Hg is present at higher concentration levels in deep sediment. Hg contaminated sediment results from previous releases or ongoing contributions from sources that are difficult to identify. Due to human activities or physical, chemical, or biological processes (e.g. hydrodynamic flows, bioturbation, molecular diffusion, and chemical transformation), the buried Hg can be remobilized into the overlying water. Hg speciation in the water column and sediments critically affect the reactivity (i.e. conversion of inorganic Hg(II) to MeHg), transport, and its exposure to living organisms. Also, geochemical conditions affect the activity of methylating bacteria and its availability for methylation. This review paper discusses remedial considerations (e.g. key chemical factors in fate and transport of Hg, source characterization and control, environmental management procedures, remediation options, modeling tools) and includes practical case studies for cleaning up Hg-contaminated sediment sites. Published by Elsevier Inc.

Keywords: Mercury Sediment Remediation Partitioning coefficients Modeling

1. Introduction Mercury accumulates in sediment globally from many physical, chemical, biological, geological and anthropogenic environmental processes (U.S.EPA, 1997, 2006; Benoit et al., 1999b; Braga et al., 2000; Hylander et al., 2000; Ullrich et al., 2001; Huibregtse, 2006; Sunderland et al., 2006; Swain et al., 2007; UNEP, 2011). Direct (point source) Hg contamination is usually the result from abandoned Hg mines, gold-mining activities (Ebinghaus et al., 1998; Meech et al., 1998; Veiga and Meech, 1999; Telmer and Veiga, 2009; Cordy et al., 2011; Drace et al., 2012; Krisnayanti et al., 2012), ore refining, and products and processes such as recycled mercury processing or the chlor-alkali industry (Randall et al., 2006; Ullrich et al., 2007; Reis et al., 2009; Gluszcz et al., 2012; Ilyushchenko et al., 2012). With artisanal and small scale gold mining, Telmer and Veiga (2009) estimates that approximately 1000 metric tons/yr of Hg was released from at least 70 countries. Approximately, 350 metric

n

Corresponding author. E-mail addresses: [email protected], [email protected] (P.M. Randall), [email protected] (S. Chattopadhyay). 0013-9351/$ - see front matter Published by Elsevier Inc. http://dx.doi.org/10.1016/j.envres.2013.01.007

tons/yr of this amount is directly emitted to the atmosphere while the remainder, 650 metric tons/yr, is released into the hydrosphere (i.e. rivers, lakes, soils, tailings). Indirect (non-point source) Hg contamination is largely attributed to atmospheric deposition (wet and dry) originating from coal-fired power plants. Global mercury emissions from coal-fired power plants were estimated at approximately 850 metric tons/yr (Pirrone et al., 2010). Other indirect sources to the aquatic environment can be attributed to runoff to water bodies or leaching from groundwater flows in the upper soil layers. At large contaminated sediment sites, engineers and scientists face many challenges primarily due to the large volumes of sediment that are typically involved. Usually, the remediation timeframes and spatial scales are in many ways unprecedented. The complexities and high costs associated with characterization and cleanup are magnified by evolving regulatory requirements and the difficulties inherent in tracking the contaminants in aquatic environments. Remedial strategies often require unexpected adjustments in response to new knowledge about site conditions or advances in technology (such as improved dredge or cap design or in situ sorption materials and treatments). Regulators and engineers adapt continuously to evolving conditions and environmental responses. Depending on site specific conditions,

132

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

effective management of Hg contaminated sites includes more adaptive site investigation, remedy selection, and remedy implementation. In a remedial investigation and screening of potential alternatives, practitioners may consider various approaches including: (a) dredging and excavation with sediment dewatering and handling, (b) sediment treatment of dredged materials by physical, chemical and biological processes, (c) in-situ/ex-situ subaqueous capping in combination with dredging), (d) in-situ/ ex-situ capping treatments that contain contaminants by chemical and biological processes, (e) containment in contained disposal facilities (CDFs), contained aquatic disposal (CAD), and landfills, (f) monitored natural recovery (MNR), (g) phytoremediation and (h) combination of above mentioned options. Hg contaminated sites sometimes implement a suite of remedial approaches to clean-up the site. For example, at the Lavaca Bay Point Comfort site (TX, USA), Alcoa spent approximately $110 million to implement several remedial options (i.e. dredging, capping, MNR, disposal in a CDF, long-term monitoring) in and around the bay (U.S.EPA, 2012). There are several U.S. agencies (i.e. U.S. Army Corps of Engineers (USACE), National Oceanic and Atmospheric Administration (NOAA), U.S. Geological Survey (USGS), and the Department of the Navy) that manage contaminated sediment programs including Hg contaminated sediments sites. U.S. EPA’s Great Lakes National Program Office estimates that 76 million cubic yards of contaminated sediments in the Great Lakes require remediation at an approximate cost of $1.6 to $4.4 billion (U.S.EPA, 2006). The Department of the Navy has estimated that there are more than 200 contaminated sediment sites which they manage with a projected remediation cost to cleanup of $1.3 billion dollars (Blake et al., 2007). Although the U.S. EPA has historically emphasized that no presumptive remedy exists for sediments, most removal actions have included dredging (e.g. 56 of the 63 sediment sites in 2006) (Huibregtse, 2006; U.S.EPA, 2006). However, remedial actions recently have included reactive thin-layer capping, phytoremediation, and other remedial alternatives to reduce the resuspension and mobility of contaminants, carbon footprint, and other factors. Moreover, selection of remedial options is dependent on site-specific conditions that constitute acceptable levels of effectiveness and performance. In Hg contaminated site cleanup, most remedial technologies focus on highly contaminated areas and are not suitable for remediating vast, diffuse, Hg contaminants at low concentrations. Speciation of Hg is an important consideration that concerns the identification and quantification of specific chemical forms of Hg and is a critical determinant of its mobility, reactivity, and potential bioavailability in the impacted sediment-water systems. Since each remedial action can result in a change in the physical, chemical and biological conditions of the sediment, it is expected that the speciation and transport properties of Hg might change as the result of implementing a remedial action. However, the effectiveness of many remediation practices and long-term reliability has not been adequately assessed (Degetto et al., 1997). Fish advisories on contaminated water bodies are plentiful in the U.S. because of the inorganic Hg(II) that is converted to MeHg and thus, moves up the food chain. In the U.S. fish advisories are due to five (5) bioaccumulative chemical contaminants: mercury, polychlorinated biphenyls (PCBs), chlordane, dioxins, and dichlorodiphenyltrichloroethane (DDT). In 2010, the EPA reported more than 4598 fish advisories with 81% due to Hg (U.S.EPA, 2011). The accumulation of Hg in the food chain depends primarily on the concentration of MeHg, rather than total Hg, in water. It has been reported in the literature that only a minor fraction of Hg in natural water is in the form of MeHg; however, MeHg concentrations generally in surface water are extremely low, near the detection limit of the currently available techniques ( o50

femto-molar)(Kraepiel et al., 2003). To protect aquatic life, a scientific benchmark or reference point called the sediment quality guidelines (SQG) was developed (U.S.EPA, 1989; Long and Morgan, 1990; Coates and Delfino, 1993; MacDonald, 1994; Chapman, 1995; Long et al., 1995, 1998a, 1998b; Carr et al., 1996; Smith et al., 1996; Long and MacDonald, 1998; MacDonald et al., 2000; Anderson et al., 2001; Batley et al., 2002; Canadian Council of Ministers of the Environment (CCME), 2003; O’Connor, 2004; McCready et al., 2006a, 2006b, 2006c; Environment Canada, 2007). SQGs attempt to foresee and assess the potential for observing adverse biological effects in aquatic systems for chemical contaminants (i.e. metals and metalloids, organic compounds, polycyclic aromatic hydrocarbons (PAHs), organochlorine pesticides, and others). NOAA annually collects and analyzes sediment samples from sites located in coastal marine and estuarine environments throughout the U.S. They evaluated a wide variety of marine sediment toxicity studies that were conducted in laboratories and in the field for the effects of sediment concentrations on benthic organisms. They established effects range-low (ERL) and effects range-medium (ERM) concentrations for each constituent evaluated. ERL and ERM values are those concentrations above which adverse biological effects were seen in 10% and 50%, respectively. ERL and ERM values together define the concentration ranges that were (1) rarely and (2) frequently associated with adverse effects. Long et al. (1995) reported ERL and ERM for total Hg as 0.15 mg per kilogram (mg/kg) and 0.71 mg/kg dry weight basis, respectively. ERL is not a threshold below which sediment toxicity is impossible and above which it is likely. Rather, an ERL is simply a low point on a continuum of bulk chemical concentrations in sediment that roughly relate to sediment toxicity (Beckvar et al., 1996; O’Connor, 2004). Another criterion limit is the apparent effects threshold (AET) values derived from a correlation of the weight of evidence from multiple matched chemical and biological effect data sets (laboratory toxicity testing on field sediment samples). The AET value for a particular contaminant is defined as the sediment concentration above which an adverse biological effect is always statistically observed (U.S.EPA, 1989). For example, the ERL for Hg is 0.15 mg/kg of sediments, ERM is 0.71 mg/kg, and the AET is 2.1 mg/kg (Baumgarten and Panel, 2001) in Alcoa’s Lavaca Bay Point Comfort site, Texas USA. For PAHs, the ERL and ERM in sediments are 4.02 mg/kg and 44.79 mg/kg, respectively. Similar criteria were adopted by Canada to protect aquatic life. The Canadian Council of Ministers of the Environment (CCME) derived two reference values for some 30 substances in freshwater and marine sediments: a threshold effect level (TEL) and a probable effect level (PEL). These two values were adopted for the assessment of sediment quality in Quebec and were developed using a nationally-approved protocol (Canadian Council of Ministers of the Environment (CCME), 2003). The Hg TEL and PEL values for freshwater sediment are 0.17 mg/kg and 0.49 mg/kg, respectively; and the same for marine sediments are 0.13 mg/kg and 0.70 mg/kg, respectively (Smith et al., 1996; Canadian Council of Ministers of the Environment (CCME), 2003; Environment Canada, 2007). Massachusetts Department of Environmental Protection (MDEP) assessed the screening criteria by adopting consensus-based threshold effect concentrations (TECs) for the 28 chemicals, including Hg, to determine risk to benthic organisms in freshwater sediment (MacDonald et al., 2000; Massachusetts Department of Environmental Protection (MDEP), 2002). The TECs are intended to identify contaminant concentrations below which harmful effects on sediment-dwelling organisms are not expected. These concentrations may not necessarily be protective of higher level organisms exposed to bioaccumulating chemicals. These consensus-based TEC values were chosen because they incorporate a large data set,

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

Table 1 Sediment quality guidelines for selected metals (Massachusetts Department of Environmental Protection (MDEP), 2002; Wells and Hill, 2004). Substance (metals)

ERL (mg/kg)

ERM (mg/kg)

AET (mg/kg)

Consensus-based TEC

Arsenic Cadmium Chromium Copper Lead Hg Nickel Zinc

8.5 1.2 81 34 46.7 0.15 20.9 150

70 9.6 370 270 218 0.71 51.6 410

35 3 62 390 400 0.41 110 410

9.79 0.99 43.4 31.6 35.8 0.18 22.7 121

Table 2 Approximate solubility of mercury compounds at 25 1C (Wilhelm, 1999). Forms of Hg

Water (lg/L)

Oil (lg/L)

Hg0 XHgX HgCl2 HgS HgO CH3HgCl

50 NA 70,000,000 10 50,000 Very high

2000 infinite 410,000 Very low, o 10 low 1,000,000

provide an estimate of central tendency that is not unduly affected by extreme values, and incorporate sediment quality guidelines that represent a number of approaches for developing sediment benchmarks. A list of these consensus-based TECs is provided in Table 1. It is well known that sediments are reservoirs for toxic compounds. Mercury discharged into the hydrosphere rapidly becomes associated with particulate matter and incorporated in bottom sediments. At the water–sediment interface, it is important to understand the mechanisms and environmental variables that drive or constrain methylation dynamics. Although diagenetic processes in the sediments can modify and redistribute the mercury between solid and solution phases, immobilization by sedimentation dominates for most elemental contaminants such as Hg. Understanding the chemistry and these physical, chemical, and biological processes are important to all site remediations. Next, we explore the chemistry of Hg methylation dynamics, geochemical factors that influence methylation dynamics and mechanisms that influence the aqueous phase and the solid phase MeHg concentrations.

2. Biogeochemistry of mercury 2.1. Mercury speciation The distribution, mobility and biological availability of chemical elements depend not simply on their concentrations but, critically, on the forms in which they occur in natural systems. This possible mobility and bioavailability are the result of the reactivity of trace metals (especially Hg) in sediment, in other words, their localization in different sediment components, which is usually called speciation or the different physico-chemical forms or oxidation states of the same element (Bermond et al., 1998). The mobility and availability of Hg in aquatic environments is influenced by various processes including the thermodynamic solubility of Hg and Hg compounds (see Table 2). Hg can become associated with streambed sediments, suspended particles, precipitated matter, natural organic matter

133

(NOM), and other substrates that can settle out and effectively remove Hg from the mobile aqueous phase (Choe et al., 2003; Zheng et al., 2012). The aqueous speciation and coordination of Hg have been well-documented (Benes and Halvic, 1979; Dyrssen and Wedborg, 1991; Hurley et al., 1998b; Bloom and Lasorsa, 1999; Horvat et al., 2003; Kim et al., 2004b; Skyllberg et al., 2006; Kotnik et al., 2007; Balcom et al., 2008; Skyllberg and Drott, 2010; Gibson et al., 2011; Matsuyama et al., 2011; O’Driscoll et al., 2011; Liu et al., 2012b; Mladenova et al., 2012; Wang et al., 2012b). The oxidation states of Hg in aqueous systems are 0, þ 1, and þ2. In typical aerated water, Hg(II) is most stable. Aqueous Hg(II) speciation and coordination in the absence of other strongly complex ligands is largely dictated by hydrolysis reactions. At low pH, the hexaqua ion (Hg(H2O)26 þ ) is octahedrally coordinated by water molecules, with Hg–O bond lengths of 2.34–2.41 A˚ (Kim et al., 2004b). As the pH is raised and the extent of hydrolysis increases to HgOH þ and Hg(OH)2, ˚ two of the Hg–O bonds are shortened to distances of 2.00–2.10 A, ˚ while the remaining bonds are lengthened to approximately 2.50 A. The distorted octahedral coordination is indicative of the tendency for Hg(II) to form mononuclear linear, double-coordinated complexes, as also occurs in halides, oxyanions, and certain solids. The stability of Hg(OH)2 complexes in the pH range of natural water (5 to 9) (Kim et al., 2004b). Metal speciation in aquatic environments is affected by inorganic and organic ligands present in water. The relative importance of each ligand for metal complexation will depend on the concentration of the metal and the ligand, and the binding strength (conditional stability constants) for the metal–ligand complex. Among the inorganic ligands, hydroxide, chloride, and sulfide are considered important in controlling the speciation of Hg in water (Ravichandran, 1999). In the absence of any significant chelators, Hg—hydroxide complexes (Hg(OH)2, HgOH þ ) are likely to be the important species in most freshwaters (Stumm and Morgan, 1995). Hg—chloride complexes (HgCl2, HgCl24  , HgCl3 ) are thought to be important at low pH and/or high chloride concentrations. In sediment and aquatic environments containing dissolved sulfide (including some oxic surface waters where nanomolar levels of sulfide and thiols have been detected), Hg is hypothesized to form Hg—sulfide species (Dyrssen and Wedborg, 1991; Hudson et al., 1994). Among the organic ligands, sulfur-containing ligands (e.g. cysteine, mercaptoacetate) bind to Hg much more strongly than oxygen-containing ligands (e.g. acetate, citrate, ethylene dinitrilo-tetra-acetic acid [EDTA]). Natural Hg can be present at concentrations of 1 to 20 parts per trillion in several physical and chemical forms in oxic surface freshwaters. The partitioning of Hg between the dissolved, colloidal and particulate phases varies widely spatially, seasonally and by depth in the water column. Some of this variation seems to be related to temporal changes in living particulate matter, mostly phytoplankton and bacteria (Hurley et al., 1991). The concentration of particulate Hg per unit particle weight is relatively constant reflecting perhaps sorption equilibrium between dissolved and particulate phases (Meili, 1991). The exact chemical form of particulate Hg is unknown, although most of it is probably tightly bound in suspended organic matter. Sorption of Hg to oxyhydroxides may also be important in lakes. The commonly observed enrichment of MeHg and Hg(II) in anoxic waters of lakes may result from the sedimentation of Hg-laden oxyhydroxides of iron and manganese from the epilimnion and their dissolution in the anoxic hypolimnion layers (Meili, 1991). 2.2. Presence of natural organic matter (NOM) Natural organic matter (NOM) consists of redox reactive but chemically heterogeneous organic matter substances that exist ubiquitously in aquatic environments (Chen et al., 2002; Zheng

134

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

et al., 2012). NOM is known to bind trace metals strongly, affecting their speciation, solubility, mobility and toxicity (Buffle, 1988). There is increasing evidence that dissolved organic matter (DOM) interacts very strongly with Hg, affecting its speciation and bioavailability in aquatic environments (Loux, 1998; Lamborg et al., 2003; Ravichandran, 2004; Miller et al., 2007; Schuster et al., 2008; Hill et al., 2009; Henneberry et al., 2011; Graham et al., 2012). Strong interactions between Hg and DOM have also been indicated by positive correlation between their concentrations in many natural waters. NOM interacts with Hg in several different ways, affecting the transport, transformation and bioavailability of Hg. Major mechanisms by which NOM may adsorb onto mineral surfaces involve: (a) anion exchange (electrostatic interaction), (b) ligand exchange surface complexation, (c) hydrophobic interaction, (d) entropic effect, (e) hydrogen bonding, and (f) cation bridging (Gu et al., 1994). In addition to its recognized metal binding capabilities (e.g. Hg(II), Fe(III)), (Zhou et al., 2005), NOM can mediate electron transfer (Gu et al., 2011) and is one of the most important electron shuttles in natural environments (Van der Zee and Cervantes, 2009). For example, NOM has been shown to accept electrons from microorganisms and then transfer them to electron acceptors such as Fe(III) and iron oxide minerals (Uchimiya and Stone, 2009). NOM forms extremely strong complexes between Hg(II) and reduced sulfur (Skyllberg et al., 2006). Strong complexation facilitates the mobility of Hg from sediments (Wallschlager et al., 1996) into streams (Mierle and Ingram, 1991), lakes (Driscoll et al., 1995), and groundwater (Krabbenhoft and Babiarz, 1992). This enhanced mobility results in increased water column concentrations of Hg in otherwise pristine lakes and streams. Complexation also affects the partitioning of Hg to suspended solids in the water column and the sequestration of Hg to sediments. Dissolved organic matter (DOM) is also known to promote (Weber, 1993) or inhibit (Miskimmin et al., 1992) the formation of toxic and bioaccumulative MeHg species. DOM plays an important role in bioaccumulation and biomagnifications of Hg (Cormack, 2001). Contaminants in sediment are taken up by benthic organisms in a process called bioaccumulation. When animals higher in the food chain feed on contaminated organisms, the toxins are taken into their bodies, moving up the food chain in increasing concentrations in a process known as biomagnification. Complexation with DOM limits Hg(II) availability to methylating bacteria and CH3Hg þ availability for bioaccumulation (Barkay et al., 1997). Earthworms applied in laboratory tests at gold mining sites showed high bioavailability in organic soils and low bioavailability in lateritic soil and clay-rich sediments (Hinton and Veiga, 2009). Humic and fulvic acid fractions of DOM are also capable of reducing ionic Hg to the volatile elemental Hg (Alberts et al., 1974), increasing the flux of Hg from water and soil to the atmosphere. Although the role of humic matter in the methylation of Hg remains unclear, it seems that organic carbon can enhance methylation by stimulating the activity of heterotrophic microorganisms, or through direct abiotic methylation of Hg by humic or fulvic substances. Also, DOM enhances the formation of Hg0 from Hg(II) in photochemical reactions (Ravichandran, 2004), which could reduce the availability of Hg for methylation and bioaccumulation. 2.3. Sulfidic environments Based on the preference of a cation for complexation with ligands, Hg is classified as a B-type metal cation, characterized by a ‘‘soft sphere’’ of highly polarizable electrons in its outer shell. Soft metals (like Hg) show a pronounced preference for ligands of sulfur, the less electronegative halides, and nitrogen over ligands containing oxygen (Stumm and Morgan, 1995). From an ecological standpoint, Halbach

(1995) concluded that the bioaccumulation of Hg in fish and its toxicity in humans is attributed to the high affinity of Hg for sulfurcontaining proteins such as metallothionein and glutathione. Strong interactions between Hg and organic matter found in sediment and aquatic environments are attributed to the binding of Hg with sulfurcontaining functional groups in organic matter. Sulfur is a minor constituent in DOM, ranging from approximately 0.5% to 2.0% by weight. Sulfur in DOM occurs as reduced (e.g. sulfide, thiol) or as oxidized species (e.g. sulfonate, sulfate), with oxidation states ranging from 2 to þ 6. The stability constant for Hg(II) complexation with an oxidized sulfur ligand, SO24  , is 101.3, whereas, the stability constant for Hg(II) complexation with a reduced sulfur ligand, S2  , is 1052.4. The reduced sulfur sites are expected to be important for Hg binding. Generally, hydrophobic acid fractions of DOM (which includes the humic and fulvic acid fractions) had significantly higher reduced sulfur content than the low molecular weight hydrophilic acid fractions. Even if it is assumed that only a small fraction (about 2% as suggested by Amirbahman et al. (2002)) of the reduced sulfur is available for binding with Hg in natural systems, the strong binding sites in organic matter far exceed the amount of Hg available in natural aquatic systems. Since binding of Hg to DOM under natural conditions is controlled by a small fraction of DOM molecules containing reactive thiol functional groups (Haitzer et al., 2002), a positive correlation may not always exist between Hg and dissolved organic carbon (DOC) concentration (Hurley et al., 1998b). In Hurley’s study, a strong relationship between filtered Hg species and DOC was evident for wetlands draining in Wisconsin, but not in the Everglades suggesting either differences in the binding sites of the two regions or non-organic complexation in the Everglades. In general, positive correlation between Hg and DOC concentrations could be expected in cases where Hg is released and cotransported with the organic matter (Wallschlager et al., 1996). But, in systems where water column Hg is primarily derived from direct atmospheric sources, correlation between Hg and DOC may or may not be present. In both cases, significant differences can be expected in the reactivity of DOM with Hg depending on the structural and chemical characteristics of DOM (Babiarz et al., 2001) and the presence of other competing ions in water. As the S/Hg ratio increased in sediments, Hesterberg et al. (2001) concluded that multiple sulfur ligands were coordinated with Hg. Ravichandran (2004) reported that the organic matter and mercaptoacetic acid (HS–CH2–COOH), a thiol-containing compound, caused a dramatic increase in Hg release from the Florida Everglades (up to 35 mM total dissolved Hg) from red cinnabar (HgS), a relatively insoluble solid (Ksp ¼10–36.8). DOM also inhibited the precipitation of metacinnabar (black HgS), a very insoluble solid (Ksp ¼10–36.4), at an initial Hg concentration of 5  10  88 M (Ravichandran, 2004). In contrast to sulfurcontaining ligands, oxygen-containing ligands, such as acetic acid and EDTA, dissolve very little or no Hg from cinnabar. Hg speciation models calculated as a function of sulfide concentrations and pH suggest that HgS0aq, Hg(S2H)  , Hg(SH)02, and HgSsol are likely to be the most important species (Hurley et al., 1994). Sulfur cycling in aquatic sediments involves both reductive and oxidative processes (Jorgensen, 1990) and they often play a significant role in forming metal complexes. The stability constants for Hg–organic sulfur (HgRS þ ) complexes are much lower than for inorganic sulfide. The stability constants for these complexes are (Dyrssen and Wedborg, 1991; Benoit et al., 1999a) indicated below. Hg2 þ þHS 2HgSaq 0 þ H þ

K ¼ 1026:5

Hg2 þ þ 2HS 2HgðS2 HÞ þ H þ K ¼ 1032:0

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

Hg2 þ þ 2HS 2HgðSHÞ2 0 2þ

Hg

þ RS 2HgRS 

þ

K ¼ 1037:5 K ¼ 1022:1

However, the Hg–DOM binding constants in natural environments are reported to be much higher. The binding constants for the HgRS þ complex was determined at 1025.8–1027.2 by Drexel et al. (2002), 1028.5 by Haitzer et al. (2002), and 1031.6–1032.2 by Skyllberg et al. (2000). These values are higher than for Hg complexation with inorganic sulfides. Differences between the above values may be attributed to the Hg/DOM ratio in these studies, wherein Hg may be bound by a single thiol (RSH) group, by bidentate aromatic and aliphatic thiols and phenols or carboxyls (Xia et al., 1999; Hesterberg et al., 2001; Drexel et al., 2002). These high stability constants indicate that organic matter can easily outcompete sulfide for the complexation of Hg in anoxic environments. 2.4. Effect of chloride and sulfate The sorption of Hg onto particles can be significantly affected by the presence of complexing ligands, like chloride and sulfate that are present in freshwater or seawater. These ligands affect the sorption of Hg due to several possible processes including (a) formation of stable, non-sorbing metal–ligand aqueous complexes; (b) formation of metal–ligand ternary surface complexes, which at high metal and ligand concentrations can lead to surface precipitation; (c) competitive ligand sorption to particle surfaces, effectively blocking the more reactive sorption sites at the surface; and (d) reduction of positive charges at particle surfaces, thus lowering the electrostatic repulsion of cations by surfaces (considering ligands are anions and pH levels are below the point of zero charge, pHpzc, of the mineral particles). Kim et al. (2004a) reported that presence of chloride and sulfate reduced or increased sorption of Hg(II) on goethites (a-FeOOH), g-alumina (g-Al2O3), and bayerite (b-Al[OH]3), which are useful surrogates for the natural sediments. Hg(II) sorbs strongly as a bidentate corner-sharing surface complex to the Fe (O,OH)6 octahedra of the goethite structure and as a monodebtate, corner-sharing bidentate, and edge-sharing bidentate complexes to the Al (O,OH)6 octahedra that compose the bayerite structure. Hg(II) sorbs weakly to g-alumina due to the conversion of the hydrated galumina surface to a secondary bayerite-like phase. Over the chloride concentration range of 10  5 M to 10  2 M, a lowering in Hg sorption on a-FeOOH, g-Al2O3, and b-Al(OH)3 was from 0.42 mmol/m2 to 0.07 mmol/m2, 0.06 mmol/m2 to 0.006 mmol/m2, and 0.55 mmol/m2 to 0.39 mmol/m2, respectively. This reduction in Hg(II) sorption is primarily due to the formation of stable, non-sorbing aqueous HgCl2 complexes in solution, limiting the amount of free Hg(II) available to sorb. At higher chloride concentrations (Cl  Z10  3 M) and a pH of 6, the large proportion of unsorbed, aqueous Hg(II) facilitated reduction of Hg(II) to Hg(I) and the formation of Hg2Cl2 (s) (calomel) or Hg2Cl2 (aq) species. Sulfate, in contrast, enhanced Hg(II) sorption over the sulfate concentration range 10  5 M to 0.9 M, increasing Hg surface coverage on a-FeOOH, g-Al2O3, and b-Al(OH)3 from 0.39 mmol/m2 to 0.45 mmol/m2, 0.11 to 0.38 mmol/m2, and 0.36 to 3.33 mmol/m2, respectively. This effect might be due to the sorption or accumulation of sulfate ions at the substrate interface, effectively reducing the positive surface charge that electrostatically inhibits Hg(II) sorption. 2.5. Methylation and bioaccumulation In the aquatic environment (including water, sediment, and biota phases), most of the mercury is in the inorganic and organic forms of divalent Hg(II) compounds with Hg(0) being a considerable fraction of the dissolved Hg in the water phase (U.S.EPA, 1997; Ullrich et al., 2001) whereas in most fish species, 495% of Hg is in the form of monomethylmercury (CH3Hg). Thus, conversion of ionic Hg to MeHg

135

is an important link in the bioaccumulation of Hg in fish and ultimately its toxicity to humans and wildlife. MeHg production in sediment and aquatic systems is not a simple function of total Hg concentration in the system. MeHg formation is influenced by a number of environmental factors including temperature, pH, redox potential, activity and structure of bacterial community, speciation, age, and the presence of inorganic and organic complexing agents (Ullrich et al., 2001). These factors also interact with each other. Generation of MeHg in anoxic sediment and water systems, and its transport are shown in Fig. 1. Dissolved Hg is distributed among several chemical forms: elemental Hg (Hg0aq), which is volatile but relatively non-reactive, a number of mercuric species (Hg[II]), and organic Hg, mainly MeHg, dimethyl (Me2Hg), and some ethyl (EtHg) Hg. In general, and particularly in stratified systems, concentrations of Hg0 are higher near the air–water interface, whereas levels of Hg and MeHg are higher near sediments. Hg methylation is mainly a microbial mediated process, with abiotic methylation likely to be important in organic-rich lakes (Ullrich et al., 2001). Bacteria assimilate Hg through passive diffusion of neutrally-charged species (Barkay et al., 1997) as well as by active uptake of both charged and uncharged Hg (Kelly et al., 2003). Wetland sediments commonly have a lower oxidation–reduction potential (ORP), or redox potential (Eh), thereby promoting the reduction of Hg(II) to Hg(I) or Hg0. One of the key pathways, ORP influences Hg speciation through its effect on sulfur chemistry. Decreases in ORP promote microbialmediated sulfur-reduction, which in turn promotes Hg methylation. Furthermore, the accumulation of reduced sulfur, primarily as dissolved sulfide, will precipitate inorganic Hg as a highly insoluble HgS mineral, cinnabar (red coloration) or meta-cinnabar (black and slightly more soluble). Increases in dissolved sulfide concentrations result in decreases in Hg methylation rates because inorganic Hg is removed as a sparingly soluble solid (Gilmour et al., 1992). It is important to note that an excess of sulfide can lead to formation of soluble Hg–S complexes. Wetlands typically have very high concentrations of organic matter (OM) due to the slow rate of OM oxidative degradation occurring in this environment. The OM may either act as a sorbent or provide high concentrations of dissolved ligands that form very strong complexes to Hg(II) (Cheam and Gamble, 1974; Wallschlager et al., 1996, 1998a, 1998b). In microbial methylation, complexation with DOM, commonly measured as DOC, plays an important role in Hg bioavailability. DOM is believed to limit the amount of inorganic Hg available for uptake by methylating bacteria because DOM molecules are generally too large to cross the cell membrane of bacteria (Golding et al., 2002; Kelly et al., 2003). Further, there is competition between various ions for binding sites. Under acidic conditions, H þ may compete with Hg2 þ for binding sites on DOM and limit its sequestration, leaving more inorganic Hg available for uptake or methylation by organisms. The presence of other ions such as Na þ and K þ may also compete with Hg2 þ for binding to negatively charged functional groups on DOM (Kidd, 2012). In sulfate-limited environments where microbes may be utilizing organic matter as an energy source, DOM may have a stimulating effect on microbial growth and thus enhance methylation rates in the water column and sediments (Watras et al., 1995). It may be hypothesized that when OM is largely labile and readily biodegradable, it may promote methylation by stimulating microbial growth; when the OM is relatively recalcitrant and consists of high molecular weight humic and fulvic acids, then it may contribute to abiotic methylation. Lean et al. (2004) reportedly observed that most of the MeHg in the south-central lakes of Ontario (Kegimkujik Park, Nova Scotia) is not dissolved or bound to small particulate material but bound to DOM (less than 1 kD).

136

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

Fig. 1. Generation of MeHg in anoxic sediment and water systems, and transportation by diffusion and advection (Morel et al., 1998).

According to Stumm and Morgan (1995), the divalent Hg in surface waters, Hg(II), is not present as the free ion Hg(II) but should be complexed in variable amounts to hydroxide (Hg(OH) þ , Hg(OH)2, Hg(OH)3 ), and to chloride (HgCl þ , HgClOH, HgCl2, HgCl3 , HgCl24  ) ions depending on the pH and the chloride concentration. Even in oxic surface waters, some or much of the Hg(II) might be bound to sulfides. In addition, an unknown fraction of Hg(II) is likely bound to humic acids, the assemblage of poorly defined organic compounds that constitute 50 to 90% of the DOC in natural waters. According to Meili (1991) nearly 95% of inorganic oxidized Hg in lakes is bound to DOM. Through its binding to DOM, Hg can be mobilized from the drainage basin and transported to lakes (Morel et al., 1998). The reactions of ionic Hg are relatively fast, and it is thought that the various species of Hg(II), including those in the particulate phase, are at equilibrium with each other. In the organometallic species of Hg, the carbonto-metal bonds are stable in water because they are partly covalent and the hydrolysis reaction, which is thermodynamically favorable (and makes the organometallic species of most other metals unstable), is kinetically hindered. As a result, the dimethylmercury species, Me2Hg (CH3HgCH3), is non-reactive. The monomethyl species, MeHg, is usually present as chloro- and hydroxocomplexes (CH3HgCl andCH3HgOH) in oxic water. Once MeHg is formed, DOM facilitates its solubility, and thus increasing the water column concentration, and transports through complexation (Miskimmin, 1991). At the same time, complexation with DOM also tends to limit its uptake in biota (Driscoll et al., 1995). Apart from DOM, concentration and bioaccumulation of MeHg in fish is also affected by pH, temperature, redox potential, concentrations of aluminum and calcium, fish age and food source, and other factors (Watras et al., 1995). Temperature and season influence the availability and accumulation of Hg in addition to the factors already discussed. Changes in temperature can affect Hg concentrations in organisms either directly by affecting metabolic rate and thereby exposure, or indirectly by influencing the methylation of Hg and therefore enhancing availability. Rates of MeHg or inorganic mercury uptake increase with increasing aqueous concentrations and/or increasing temperature in the water for some species, such as phytoplankton, gastropods, fish (Rodgers and Beamish, 1981;

Tessier et al., 1994). A rise in temperature (and a corresponding rise in respiratory volume) can increase the rate of uptake via the gills (U.S.EPA, 1985). The abiotic methylation increases with an increase in temperature (Hudson et al., 1994; U.S.EPA, 2002; Boszke et al., 2003). An increase in the reaction temperature from 5 1C to 40 1C doubled the MeHg yield, as did the doubling of the spike concentration of the Hg(II) (Rogers, 1977). Biological productivity of methylating microbes is affected by seasonal changes in temperature, nutrient supply, oxygen supply, and hydrodynamics (changes in suspended sediment concentrations and flow rates). MeHg concentrations varied seasonally by an order of magnitude at most sites studied (Parks et al., 1989). Methylation may tend to increase during the summer months when biological productivity and temperature are high and decrease during winter months when biological productivity and temperature are low (Callister and Winfrey, 1986; Kelly et al., 1995). Although the potential MeHg production is greatest during the summer, actual production may not peak during this time (Kelly et al., 1995). In Onondaga Lake, New York, the Hg species in the water column varied temporally (Robertson et al., 1987; Bloom and Effler, 1990). Total Hg concentrations may also vary seasonally due to physical factors such as winter storms re-suspending Hg-contaminated sediments (Gill and Bruland, 1990). Various abiotic reactions that could be responsible for abiotic Hg methylation are indicated below.

 Transalkylation reaction by other methylated metals like lead and arsenic.

 Methylation by released methylcobalamine from bacteria.  Methylation by separate compounds due to cellular compo

nents like S-adenosylmethionine, 3-dimethylsulfone-propionate (DMSP), methyl iodide, homocysteine, dimethylsulfide. Methylation by humic and fulvic acids and degradation products.

2.6. Effect of pH There are many ways in which pH changes may influence MeHg concentrations in aquatic systems. An inverse relationship

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

between low lake pH and elevated Hg levels in fish has been observed in several regions of the northern hemisphere. Richman et al. (1988) have reviewed some of the mechanisms proposed to account for the elevated Hg levels in biota and acid-stressed lakes. Generally, low pH conditions facilitate the release of heavy metals and particulate matter from sediments. But the data on the partitioning and mobility of Hg has been somewhat contradictory (Ullrich et al., 2001). The solubility and mobility of Hg and MeHg is pH dependent, however, even pH dependency may oversimplify the many factors that regulate Hg uptake by fish and the mobility of Hg and MeHg. Some studies show mobility of Hg is higher in the acidic pH range (Duarte et al., 1991), but Schindler et al. (1980) reported that lake water acidification caused a higher proportion of Hg to bind to particulates, thereby decreasing the solubility of Hg in the water column. Ramlal et al. (1985) reported that the amount of dissolved Hg in sediment porewater was found to decrease with decreasing pH. Further, metallic Hg was reported to be readily soluble at low pH with organic acids (Meech et al., 1998). Despite contradictory evidence on the effect of pH on inorganic Hg solubility, studies show that MeHg is more soluble under low pH conditions. As indicated, this may be due to various geochemical factors including decreased availability of binding sites on organic matter and redox conditions (Ullrich et al., 2001; Kidd, 2012). Furthermore, the pH may effect the production of dimethylHg. Neutral or acidic pH conditions appear to favor the production of MeHg over dimethylHg (Beijer and Jernelov, 1979; Ullrich et al., 2001) and alkaline pH favors the formation of dimethylHg (NOAA, 1996). Methylation may occur at alkaline conditions, however, the favorable pH range for inorganic Hg methylation was reported to be between pH 2 and pH 5.5 (Falter, 1999). Acidic pH was also reported to enhance the methylation of Hg in the Carson River-Lahontan Reservoir system (Bonzongo et al., 1996). Kelly et al. (2003) studied the effect of increasing hydrogen ion (H þ ) concentrations on the uptake of Hg(II) by an aquatic bacterium. Even small changes in pH (7.3 to 6.3) resulted in large increases in Hg(II) uptake, in defined media. The increased rate of bioaccumulation was directly proportional to the concentration of H þ . Lowering the pH of Hg solutions mixed together with natural dissolved organic carbon, or with whole lake water, also increased bacterial uptake of Hg(II). Using both defined inorganic solutions and lake water, uptake of Hg(II) was faster at lower pH, and the increased rate of uptake was not related to changes in neutral Hg species such as HgCl2 or Hg(OH)2. Rather, uptake of both charged and uncharged Hg(II) species appeared to increase as H þ increased, indicating a facilitated bacterial Hg(II) uptake process that responds to pH. Hg(II) uptake rate by bacteria (for example, Vibrio anguillarum) under aerobic or anaerobic conditions is controlled by the collective concentration of a number of available Hg(II) species, both charged and uncharged, which indicates that a cell-mediated process is important in determining how much Hg(II) enters the cell. In addition to the bacterial Hg(II) uptake process, an increase in bioavailable Hg concentration at reduced pH conditions is due to the H þ outcompeting Hg(II) for binding sites on DOM particles, or because H þ may displace Hg(II) by protonating sulfhydryl moieties that bind Hg(II) to DOM (Benoit et al., 1999a) or by replacing Hg(II) on negatively charged functional groups surfaces such as clay minerals. 2.7. Interactions with minerals Numerous studies have been conducted to examine Hg(II) sorption and release (desorption) from natural and synthetic particles, including soils, clay minerals such as kaolinite, metal (hydr)oxides, and metal sulfides (Gu et al., 1994; Tiffreau et al., 1995; Yin et al., 1997; Bonnissel-Gissinger et al., 1999; Sarkar

137

et al., 2000; He et al., 2005; Yang et al., 2008; Liao et al., 2009; Feyte et al., 2010). Complexation and sorption of the precursor, Hg(II), by ligands and sediments may inhibit the production of MeHg (Stein et al., 1996). The treatment and removal of Hg from sediments are necessary for control of methylation and bioaccumulation. As discrete particles and/or as coatings on other mineral surfaces in natural systems, especially in well-weathered soil and sediments with low natural organic matter, crystalline and amorphous alumina play significant roles (Sposito, 1996; Kasprzyk-Hordern, 2004). Because of their chemical properties and physical structure, aluminum (hydr)-oxides are efficient sinks for many contaminants including cations of Hg, Pb, Zn, Cd, and Sr (Coston et al., 1995; Sarkar et al., 2000). In addition to Hg speciation, surface characteristics of aluminosilicates (surface area, porosity, pore size distribution) can have a significant impact on the fate of these contaminants. However, desorption of heavy metals from sediments and aluminosilicates can be much slower and/or nonreversible (Yin et al., 1997; Gao et al., 2003), which may lead to significant challenges due to the longer time needed for the cleanup (He et al., 2005). Furthermore, hydrous ferric oxides (sometimes referred to as iron oxyhydroxides), can act as sorbents to complex mercury species on the surface whereby mercury is retained in the solution and transported as sorbed species (Rytuba, 2000; Lowry et al., 2004; Liu et al., 2012a).

2.8. Colloids and suspended materials The distribution of Hg species between the particulate, colloidal, and dissolved phase affects the toxicity, transport, and biouptake of Hg in water and sediment systems (Chattopadhyay and Chattopadhyay, 2003). Among these size classes, the colloidal phase has been inferred to play several key roles in the biogeochemistry of Hg: (a) regulating the concentration of dissolved metal ion and neutral complexes in solution as the binding of free metal ion reduces acute toxicity; regulating neutral complexes in order to affect Hg transport across bacterial walls (Leppard and Burnison, 1983; Benoit et al., 1999a); (b) downstream transport vector due to the relatively large surface area of colloids; (c) uptaking of MeHg by bacteria, fungi, zooplankton, and mollusks either by direct consumption or the free ion activity model (Hessen et al., 1990; Guo et al., 2001). The phase distribution of both Hg and MeHg in freshwater may differ from that in marine environments because freshwater is generally lower in ionic strength, higher in alkalinity, and higher in DOC (Chattopadhyay, 2005). Stordal et al. (1996) found that a major portion of the filtered fraction (12 to 93%) was associated with the 0.4 mm to 1 kDa size fraction and that the colloidal-phase Hg concentration was correlated with carbon content. Comparing marine water and freshwater results, it is also reported that colloid coagulation in high salinity water was shown to be a major removal mechanism for Hg. Guentzel et al. (1996) reported Hg in the 0.4 mm to 1 kDa fraction was a large portion of the filtered phase (37 to 88%) in coastal marine water in the Ochlockonee River Estuary, in Florida. They presented evidence that thiol functional groups associated with organic carbon were important in the partitioning of Hg in the colloidal phase. Their study also reported the first colloidal-phase MeHg concentrations in marine environments. Babiarz et al. (2001) studied partitioning of Hg and MeHg in 15 freshwater systems located in the upper Midwest (Minnesota, Michigan, and Wisconsin) and the Southern United States (Georgia and Florida). Though they reported that the correlation between Hg and organic carbon in the colloidal phases was not statistically significant (r2 r0.14; pZ0.07), MeHg in the colloidal phase and dissolved phase were correlated with

138

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

the concentration of organic carbon as: MeHgC ¼ 0:15  OCC þ0:0018 MeHgD ¼ 0:006  OCD þ 0:0458

2

ðr ¼ 0:54, 2

ðr ¼ 0:23,

p o0:01Þ p ¼ 0:02Þ

where MeHg concentrations in colloidal phase and dissolved organic carbon phase are indicated as MeHgC and MeHgD, respectively. 2.9. Salinity A negative correlation between the rate of MeHg formation and conductivity (salinity) in estuarine sediments has been reported (NOAA, 1996). The rate of MeHg formation is lower in more saline environments because the bicarbonate component of seawater slows methylation of Hg(II) under both aerobic and anaerobic conditions (Compeau and Bartha, 1983). The release of reactive Hg(II) and Hg0 is slowed when chloride ions bind to Hg, thereby inhibiting MeHg formation (Craig and Moreton, 1985). Salinity also affects methylation due to the high porewater sulfide concentrations as a result of rapid sulfate reduction in saline water compared to sulfate-limited freshwater environments. Gilmour and Henry (1991) reported that the percentage of total Hg that is MeHg is higher in freshwater sediments (up to 37%) and water (up to 25% in aerobic water and 58% in anoxic bottom water) than in estuarine and marine water ( o5%) and associated sediments ( o5%). Dissolved reactive Hg (inorganic species) forms the majority of the total Hg in open oceans (Bloom and Crecelius, 1983; Gill and Fitzgerald, 1987). The study conducted by Babiarz et al. (2001) did not show strong trends in Hg concentration (ng/g) with suspended particulate matter, conductivity, organic carbon in the o0.4 mm fraction, pH, or percent organic matter, but MeHg concentration (ng/g) was correlated with conductivity (mS/cm  1) of the riverine water as ½MeHg ¼ 14:6 þ 0:0:2952½conductivity

ðr 2 ¼ 0:21,

p ¼ 0:03Þ

2.10. Methylation dynamics in estuarine and coastal environment Merritt and Amirbahman (2009) reviewed the various proposed mechanisms to hypothesize the influence of inorganic Hg(II) methylation rate and MeHg accumulation in estuarine and coastal marine environments. The examined whether the methylation of Hg is controlled or limited by the metabolic activity of bacteria (i.e. sulfate-reducing type), the availability of total Hg or the geochemical speciation of Hg and/or the processes responsible for the methylation/demethylation processes. These processes may be influenced by depth-dependent balance between MeHg production and consumption in estuary and marine environments. Field and research data from the Lavaca Bay Superfund site (i.e. a stratified estuarine environment) has been well published in the literature. Mason et al. (1998) reported development of porewater sampling methods for Hg and MeHg from 15 sites in Lavaca Bay. Santschi et al. (1999) reported radionuclide, Hg concentrations, radiography, and grain size data from sediment cores and calculated sediment accumulation rates as high as approximately 2 cm/yr at near-shore sites near the Alcoa facility. These authors also predicted Hg concentrations in surface sediments to decrease exponentially with a recovery halftime of 4 72 years. Sager (2002) reported that monitoring of Hg in organisms since 1977 shows a gradual downward trend in Hg in crabs and finfish. Hammerschmidt and Fitzgerald (2004) reported various geochemical controlling factors on the production and distribution of MeHg in the near-shore marine sediments. Bloom and Lasorsa (1999) described the speciation and cycling of Hg in Lavaca Bay

and calculated the distribution coefficients (log Kd). They reported the average log Kd for inorganic Hg and MeHg as 4.89 70.43 and 2.7070.78, respectively. Particulate Hg concentration in surface water reportedly varied from 350 ng/g to 1610 ng/g and is controlled by physical mixing of polluted fluvial particulates with relatively unpolluted marine particulates (Baeyens et al., 1998). Dissolved Hg species show large seasonal variations essentially controlled by the redox conditions and bacterial and phytoplankton activities. It has been observed that in oxidizing conditions in freshwater, inorganic Hg (Hg(II)) is predominately present as Hg(OH) 2 and HgOHCl, and in the reducing conditions the sulfurbased Hg species (such as HgS) are primarily present (Kannan and Falandysz, 1998), but in marine systems, Hg complexes with chloride (Cl  ) to form soluble HgCl24  . Hg also complexes with the humic and strong complexing ligands of organic substances (Leermakers et al., 1995) present in the aquatic systems but this phenomenon is subdued in the marine waters because of the presence of abundant Cl  ions (Leermakers et al., 1995; Morel et al., 1998). Thus, in Lavaca Bay, Hg predominately exists in Hg0, Hg(OH)2 and HgOHCl formed in the oxidizing conditions, and as HgS in the reducing environment. In addition, organomercury species, monomethylmercury (CH3Hg þ ) and dimethylmercury ([CH3]2Hg) were present. The transformation of Hg between the elemental form, the ionic form and organomercury is controlled by biotic and abiotic processes in aerobic and anaerobic environments. MeHg is generally a product of methylation of inorganic Hg carried out by the sulfate-reducing bacteria in the aquatic environments rather than abiotic processes (Bloom et al., 1999). A stratified estuarine environment, such as Lavaca Bay, showed higher concentrations of MeHg at the oxic/anoxic interface of the sediment system or in the first 5 cm of the sediment layer (Bloom et al., 1999). MeHg concentrations were high in the marshes of Lavaca Bay. These marshes are highly productive environments with detrital carbon (plant litter) which drive the microbial process promoting methylation of the bioavailable Hg. The higher Hg levels in these areas indicated that these were the areas where Hg was being transferred into the food chains as most of the higher tropic level organisms feed on the lower trophic level organisms found in these areas. The lower organisms on the lower levels of the food chain pick up their Hg from the sediments of these areas.

3. Remediation options 3.1. Dredging Environmental dredging is of special interest because it can be expensive and technically challenging to implement. Dredging itself may create exposures (for example, through the resuspension of buried Hg contaminants), but it removes persistent contaminants (and their associated potential for transport and risk) from the aquatic environment permanently. Whether to dredge contaminated sediments has proved to be one of the most controversial aspects of decision-making for sediment remediation sites (Barbosa and de Almeida, 2001; Blazquez et al., 2001; Doody and Cushing, 2002; Bridges et al., 2006; NRC, 2007). One of the examples of dredging was in Japan’s Minamata Bay, where Hg concentrations as high as 600 mg/kg were detected in settled sediment (Hosokawa, 1993). High concentration was observed at the inner part of the bay. MeHg in most sediment was o0.005 mg/kg and 0.03 mg/kg at maximum. Sediment was dominantly soft silty or silty clay and contained rich sulfides. The Japanese Environmental Protection Agency established standards for removal of contaminated bottom sediments for Hg; the standards took into consideration release rate, dilution rate by

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

the tide, and biological accumulation with safety factor. A dredging removal standard of 25 mg Hg/kg was established. In October 1977, the Minamata Bay project was initiated and completed in March 1990 (Kudo et al., 1998, 2000; Tomiyasu et al., 2006). The total amount of dredged sediment removed was about 1.5 million tons of sediments at an estimated cost of $500 million (Kudo et al., 1998). The depth of sediment, which contained Hg greater than 25 mg/kg, was 1 to 2 m in the inner part of the bay, falling to 0 to 0.1 m in off-shore areas. Based on the report by Hosokawa (1993), monitoring data showed that at most sampling points, Hg concentrations were found to be below 5 mg/kg after dredging. The maximum concentration of Hg was 8.75 mg/kg. Despite the results seen at Minamata Bay, dredging activities may cause adverse environmental threats if they are not well planned and implemented (Nichols et al., 1990; Schultz et al., 1995; Van den Berg et al., 2001). Dredging-induced sediment re-suspension is a major environmental concern. Given no significant disturbance, buried Hg and other metals are generally sorbed by sediment, and can generally be regarded as safely separated from the overlying water. Activities such as dredging, shipping, and natural occurrences, such as storms and tides, can remobilize Hg that was sorbed by sediment (Van den Berg et al., 2001). Sunderland et al. (2004) defined the active zone of the sediments as sediments that can potentially exchange Hg with the water column and buried sediments through re-suspension, diffusion and burial. This thickness of the active layer is a function of the depth of biological mixing and the depth of physical mixing/continual reworking (Boudreau, 2000). Sunderland et al. (2006) found that re-suspension and mixing of tidal sediments enhanced MeHg production. Bloom and Lasorsa (1999) conducted laboratory testing to simulate ocean dredging. They reported that about 5% of MeHg and less than 1% of total Hg can be released from contaminated sediment as a result of dredging. It is also noteworthy that sediment pore water, which usually contains high concentrations of Hg, can readily release Hg into the overlying water (Gilmour et al., 1992). After comparing various dredging techniques, it was suggested that a combination of mechanical and hydraulic dredging produces the least sediment resuspension (Hauge et al., 1998; Wang et al., 2004). Mathematical models were developed to estimate dredging costs, efficiency, and environmental effects (Hayes et al., 2000; Blazquez et al., 2001). During dredging, oxygen in overlying water can enter buried anoxic sediment and possibly oxidize and release contaminants (Vale et al., 1998). Under undisturbed conditions, the formation of MeHg is restricted primarily to the uppermost 3 cm to 15 cm of benthic sediment (Sunderland et al., 2004). Sediment and porewater Hg speciation show that the upper 10 cm of sediment are dynamic regions of MeHg formation and diagenesis (Gilmour et al., 1992; Bloom and Lasorsa, 1999) and is usually insignificant below 10 cm (Leermakers et al., 1993; Bloom et al., 1999). However, it has been observed that after dredging, some buried sediment is mixed with surface sediment, or water, which can produce an environment with high sulfate and organic matter concentrations that favor the production of MeHg (Bloom and Lasorsa, 1999). Studies conducted in the United Kingdom (UK) showed that water discharged from dredging sites had higher concentrations of organic matter that favors the production of MeHg (Newell et al., 1999). Furthermore, dredging of the contaminated sediment is only a temporary solution to the problem (Barbosa and de Almeida, 2001). The treatment of dredged sediment is usually costly. Therefore, confinement (disposal followed by capping) and direct disposal are more common alternatives (Wang et al., 2004). The two most widely used disposal sites are land and sea (Barbosa and de Almeida, 2001). The disposal of dredged sediments poses a potential threat to the surrounding environment. Increased

139

turbidity is usually observed at dredge disposal sites (Nichols et al., 1990). The leakage of Hg into groundwater systems from disposal sites is another concern. In Georgia, the lower Savannah River showed elevated concentrations of some metals (including Hg) in living organisms close to an upland dredge disposal site (Winger et al., 2000). Contaminated, dredged sediment confinement is widely used to prevent potential adverse environmental effects from dredge disposal. Adjusting pH to an optimal level (based on laboratory and pilot tests) or adding sorbent materials is a common method to immobilize heavy metals. For example, adding materials containing iron is quite effective in immobilizing mercury (e.g. FeS has been shown to exchange its Fe(II) with Hg(II) to form HgS(s) (Xiong et al., 2009). Dredging can be very effective in cleaning up heavily Hgcontaminated sediment; however it has disadvantages that need to be carefully addressed such as sediment resuspension. and cost (i.e. cost of remediation including environmental dredging, could be as high as $1409/m3 (Doody and Cushing, 2002)). A dredging demonstration project conducted by Alcoa evaluated a hydraulic cutterhead technology (10-in. hydraulic dredge with a 14-inch suction pipe and 36-inch cutterhead) to control Hg residuals. The dredging cost was $251,000 (disposal costs not included) for 9500 yd3 of sediment (3–12% solid)(Alcoa, 2000). Six acres of very soft plastic clay sediment was dredged in 20 days. Results of the study showed the following key observations: (a) hydraulic dredging could be readily implemented at this site; (b) offsite transport of Hg on tidal flows moving through and around the curtained-off dredging unit were minimal; (c) a large mass of mercury was removed (2300 lbs) with 60,000 yd3 to 80,000 yd3 of sediment and placed in a confined disposal facility; (d) increased Hg concentrations in oysters above the historical observed background in the wider bay did not occur. The dredging operations were conducted with multiple passes with sampling between passes to define the residual present after each pass. There was a notable increase in residual concentration between passes 2 and 3, apparently reflecting exposure of more highly contaminated sediment. Overall, the pass-to-pass concentration changes were not statistically significant. The pilot study was judged to be successful and the data collected were important in the evaluation of the role of dredging at this site.

3.2. In-situ and ex-situ subaqueous capping Capping refers to the process of placing a subaqueous covering or proper isolating materials to cover and separate the contaminated sediments from the water column. A cap can reduce contamination risk by one or multiple activities: (a) physical isolation of the contaminated sediment from the aquatic environment; (b) stabilization/erosion protection of contaminated sediment; and (c) chemical isolation/reduction of the movement of dissolved and colloidally transported contaminants into the water. In situ capping is on-site placement of proper covering material over contaminated sediment in aquatic systems. Laboratory treatability studies suggest that in situ capping can be effective in reducing the impact of Hg contamination in aquatic systems. In ex-situ capping, contaminated sediment is dredged and relocated to another site, where one or multiple isolating layers are placed over the sediment (Palermo, 1998; Liu et al., 2001). Ex-situ capping is a combination of dredging and capping. Important distinctions should be made between in situ capping and dredged material capping or ex situ capping, which involves removal of sediments and placement at a subaqueous site, followed by placement of a cap. Dredged material capping is a disposal alternative that has been used for sediments dredged from navigation projects, and may also be suitable for disposal

140

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

of sediments and treatment residues from remediation projects. There are two forms of dredged material capping: (a) level bottom capping, where a mound of dredged material is capped, and (b) contained aquatic disposal (CAD) in which dredged material is placed in a depression or other area that provides lateral confinement prior to placement of the cap. An isolation cap may be constructed of clean sediments, sand, gravel, natural/synthetic reactive material or may involve a more complex design with geotextiles, liners and multiple layers with combinations of passive or reactive materials. A variation in caps could involve the removal of contaminated sediments to some depth, followed by capping the remaining sediments in-place. This is suitable where capping alone is not feasible because of hydraulic or navigational restrictions on the waterway depth. Experimental tests show that capping material composed of a mixture of sand and finer particles (silty sands per ASTM classification) can sorb Hg and other heavy metals (Moo-Young et al., 2001). Moo-Young et al. (2001) showed that capping materials can sorb 99.9% of the Hg from sediment, which contained Hg between 200 and 500 micrograms per liter (mg/L). This test showed that a capping layer can be a good barrier between Hg-contaminated sediment and the overlying water. In-situ capping field studies were conducted in Hamilton Harbor, Canada, which has high concentrations of zinc, copper, Hg, and other metals. A cap, approximately 35 cm thick and composed mostly of sand, was placed in the system to contain polluted sediment (Azcue et al., 1998). After one year of in situ capping, a field study investigated the effectiveness of the cap. In general, Hg concentrations were found to be low ( o5 mg/kg) in the capping layer, while the concentration of Hg in the original sediment ranged between 0.43 g/kg and 0.96 g/kg (Azcue et al., 1998). Johnson et al. (2010) conducted simulated tests using 1-cm sand and found no significant Hg migration into the cap over a period of approximately 8 months. The total Hg flux in the overlying water was undetectable for the capped case compared to about 10  3 ng/m2/s from exposed uncapped sediment. These results suggest that a passive sand (though do not bind significant amount of Hg on its inert structure) may provide a temporary barrier to contain Hg in the native sediment. In general, more Hg is found in sediment than in water because Hg compounds are attracted to and may attach to the small grains and particles (including decayed plants and animals) that make up the sediment. Hines et al. (2004) measured and reported Hg and MeHg at 1 cm to 2 cm resolution in sediment pore water and sediment cores from Spring Lake in Minnesota. Total Hg accumulation in the sediment was 21.4 mg/m2/year, two orders of magnitude greater than the accumulation of MeHg (0.20 mg/m2/year). The highest solid phase concentrations of MeHg were observed persistently at the sediment surface and declined sharply with depth. Pore water profiles showed a small diffusive flux of MeHg from sediment to water (5 ng/m2/month). Springtime pore water concentrations of MeHg were relatively low (approximately 0.5 ng/L) and increased by late summer to early fall (1.5–2.2 ng/ L), showing a correlation with maxima in sulfate reducing activity at 5 cm and 15 cm. Advective transport carrying MeHg deeper into the sediment was evident in summer and fall. The percent of Hg present as MeHg was highest in the water column above the sediment (10%) and decreased with sediment depth in both the solid and pore water phases. The major advantages of in situ capping are low cost, extensive suitability to a wide range of contaminants, and low adverse environmental effects (Azcue et al., 1998; Palermo, 1998). As insitu capping is not a treatment process, long-term environmental effects, including possible remobilization of contaminated sediment, need to be carefully considered by performance monitoring at regular time intervals after the installation of the capped layer.

However, there is a possibility that buried Hg may pass through the capping layer and enter into the overlying water due to various reasons (hydrodynamic flows, consolidation, transformation, diffusion, etc.). Hydrodynamic currents caused by human activities or natural processes, such as shipping, tide, and groundwater flow may scour the capping layer and release Hg into the water. Groundwater flow through the cap material can reduce the efficiency of capping significantly (Liu et al., 2001). The movement of benthic organisms may also facilitate the remobilization of buried Hg. Sediment consolidation, due to gravity, can move Hg from buried sediment into the capping layer. This sediment consolidation may be a more important factor in the transfer of Hg from buried sediment into the capping layer than molecular diffusion of Hg (Moo-Young et al., 2001). Although a pilot test conducted in a Canadian harbor suggested no significant sediment resuspension due to capping (Hamblin et al., 2000), there is always the possibility of resuspension of settled sediment due to the placement of the capping layer. Such resuspension can be the cause for transforming some of the inorganic Hg into organic Hg (MeHg). MeHg can escape into the overlying water more easily than inorganic Hg (Wang et al., 2004). Site characterization is the preliminary and crucial step to decide whether a contaminated aquatic system is suitable for capping. In general, aquatic environments with low hydrodynamic flows, such as lakes and bays, are good candidates for capping (Thoma et al., 1993). The type of capping material that can be used depends on the hydrodynamic and geotechnical conditions, and target contaminants. Sand-size and other fine materials are good for quiescent environments (Palermo, 1998). For erosive systems, coarser materials should be considered (Palermo, 1998). Gavaskar and Chattopadhyay (2008) proposed the idea of using reactive material as an in-situ cap using various natural minerals that effectively sorb contaminants and non-toxic to the benthic organisms. Zeolite is a good candidate for applying in-situ capping with active barrier systems (ABS) (Jacobs and Forstner, 1999). ABS usually is a reactive geochemical barrier layer that can actively block the contaminant release from the sediment entering into the overlying water, without the hydraulic contact between the sediment and the overlying water being disturbed. In-situ capping with reactive materials sorbs target constituents from the sediment and prevents the release of target contaminants into the overlying water more effectively than passive material alone. The cost of passive and reactive materials depend on the type of material, purity, size, delivery, source, material processing needs, and means of application. McDonough et al. (2007) reported cap placement costs for large scale site (  1000 acre) at about $25/yd2, excluding the material cost. The break-up of cap placement costs are approximately as follows: (a) mobilization/demobilization $1/yd2, (b) cap placement$10/ yd2, (c) project management $2/yd2, (d) monitoring$10/yd2, and (e) miscellaneous (site preparation, construction management, design and permit) $ 2/yd2. In addition, construction costs might not be representative due to small project footprint compared to large scale sites. 3.3. Monitored natural recovery Monitored natural recovery (MNR) is a remedial technology for contaminated sediments that typically uses ongoing, naturally-occurring processes to contain, destroy, or reduce the bioavailability or toxicity of contaminants in sediment. These processes may include physical, biological, and chemical mechanisms that act together to reduce risks posed by contaminants. The key factors that dictate the selection of MNR as a remedial technology are the concentrations of constituents of concern and whether they pose an unacceptable risk, any ongoing

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

degradation/transformation, or dispersion of contaminant, and the establishment of a cleanup level that MNR is expected to meet within a particular timeframe. The sites, which are ecologically sensitive in nature and where Hg in low concentrations is strongly bound to the sediments, are reasonable candidates for using MNR as the remedial technology. This would involve monitoring for Hg movement in the aqueous phase. Detailed spectroscopic study of the nature of the Hg in solid phases and the environmental conditions conducive to their dissolution is necessary to define the necessary safeguards to impose on the site for successfully implementing MNR. As it is generally considered that the solid phases holding Hg are themselves sensitive to the state of oxidation-reduction (Fe(III)-oxide or sulfide phases), institutional controls would have to be imposed to safeguard the site from extreme fluxes of ORP. This may involve protecting the site from extremely oxidizing conditions, which may result from water being directed away from or drained from the site. Such conditions may promote the dissolution of sulfide precipitates and the degradation of organic matter. Conversely, institutional controls may involve protecting the site from extremely reducing conditions, which may result from sustained flooding conditions that may cause Fe(III)-oxide phases to dissolve. The two primary advantages of MNR are its relatively low implementation cost and its non-invasive nature that does not need construction/infrastructure. Although costs associated with characterization and/or modeling to evaluate natural recovery can be extensive, the primary cost associated with implementing MNR is monitoring. The other advantages of MNR over active remedial methods include no sediment re-suspension, and no change in benthic conditions (Garbaciak et al., 1998). The key limitations of MNR may be the potential risk of re-exposure or dispersion of buried Hg if the sediment bed is disturbed by strong natural or man-made forces and uncertainties in predicting various situations, like future sedimentation rates in dynamic environments, rate of contaminant flux through stable sediment, or rate of natural recovery. Contaminated systems in natural attenuation should be regularly monitored to ensure environmental safety. Experiments and field studies demonstrate possible natural attenuation of Hg contamination by reduction, demethylation, and volatilization. Species-specific enriched stable isotopes have been used by Martin-Doimeadios et al. (2004) to study Hg transformations (methylation, demethylation and volatilization) in estuarine sediments under different environmental conditions (both biotic and abiotic and oxic and anoxic). They reported that MeHg levels in sediments are controlled by competing and simultaneous methylation and demethylation reactions. Korthals and Winfrey (1987) used radioactive tracers to understand seasonal and spatial variations in Lake Clara, an oligotrophic, acid-susceptible seepage lake located in Lincoln County, Wisconsin. They demonstrated the usefulness of simultaneous measurement of Hg methylation and demethylation. The net rate of MeHg production is significantly affected by the amount of demethylation, but also by environmental parameters such as temperature and anoxic conditions. The measured MeHg concentrations reflect the resultant of site-specific complex processes. Hintelmann et al. (2000) calculated the half-life of MeHg in lake sediment and the reported value is 1.7 days. Marvin-DiPasquale et al. (2003) reported a Hg methylation rate in San Pablo Bay (northern San Francisco Bay) from below the surface 0 cm to 4 cm horizon sediment profile as 0.2–1.1 ng g  1 wet sediment day  1. Since the site-specific characteristics are likely to dictate the fate of Hg in the natural environment, significant losses of inorganic Hg from surface water occurs through photoreduction and microbial reduction. Photoreduction is generally thought to contribute to rapid recycling of Hg at the air–water interface and not

141

necessarily have a significant influence on archived Hg in the sediment compartment of the aquatic systems. At low Hg concentrations (low picomolar range), photoreduction is more effective than microbial reduction (Amyot et al., 1997). Morel et al. (1998) reported that at high Hg concentrations (over 50 picomole), microbial reduction is more effective and in deep anoxic environments, certain bacteria in the presence of humic substances reduce Hg(II). Microbial demethylation of MeHg was observed in contaminated sediment (Oremland et al., 1995; Marvin-Dipasquale and Oremland, 1998). Sulfate-reducing bacteria and methanogenic bacteria are probable agents in microbial demethylation (Oremland et al., 1995; Marvin-Dipasquale and Oremland, 1998). Total Hg concentration and organic substance content are important factors in microbial demethylation (Marvin-DiPasquale et al., 2000). A demethylation rate ranged from 0.02 ng/g to 0.5 ng/g (dry sediment) per day in a field study (Marvin-Dipasquale and Oremland, 1998). Photodegradation of MeHg can also happen in surface waters. Photodegradation of MeHg seems to be a first-order reaction with respect to MeHg concentration and sunlight intensity (Seller et al., 1996). In aquatic systems, Hg0 volatilization plays an important role in the natural attenuation of Hg contamination (Amyot et al., 1997). Hg0 is probably the end-product of some reduction processes of MeHg and Hg(II) (Seller et al., 1996). Due to its high volatility, Hg0 produced by the reduction of MeHg and Hg(II) may evaporate into the atmosphere. This evaporation is a major natural attenuation of Hg in some aquatic systems. Flushing can contribute to the natural attenuation of Hg. Bloom et al. (2004) conducted biogeochemical assessment of the Venice Lagoon, which is a large (549 km2), shallow ( E1.0 m), enclosed embayment located on the northwestern Adriatic Sea. The lagoon was contaminated with more than 50,000 kg of Hg from chlor-alkali plants which operated from 1951 to 1986. Reportedly, 2000 kg of Hg per year cycled through the lagoon, most of it as a result of the resuspension of historically contaminated bottom sediments. These sediments were resuspended by both wave action and anthropogenic activities, and then transported to the Adriatic Sea by tidal flushing (Bloom et al., 2004). Garbaciak et al. (1998) reported field experiments performed in the Whatcom Waterway at Bellingham, Washington, using natural attenuation of Hg-contaminated aquatic systems. In the 1960s, the Hg concentration in the surface sediment was approximately 4.5 mg/kg. After source control and natural attenuation, Hg concentration in the surface sediment was reduced to about 0.5 mg/kg. Garbaciak et al. (1998) also defined enhanced natural attenuation as natural decontamination, accelerated by human influences. Garbaciak et al. (1998) reported the result of enhanced natural attenuation of Hg-contaminated at the Eagle Harbor site, in Washington. A thin, clean sediment cap (6 cm) was placed on the contaminated sediment to enhance the burial and separation effects, because the natural sedimentation process was too slow. These authors reported that compared to thick capping, this enhanced natural attenuation method of thin capping did not change the benthic environment significantly. Other natural processes may include phytoremediation (i.e. phytostabilization). Phytostabilization is the use of plant roots to prevent metal movement that occur in the roots or within the root neighborhood. Researchers demonstrated that the salt marsh plant Juncus Maritumus has a high capacity to stabilize mercury in sediment (Anjum et al., 2011; Marques et al., 2011). Chattopadhyay et al. (2012) investigated the potential for water hyacinths (Eichhornia crassipes) to assimilate Hg and MeHg into the plant biomass over a 68 day hydroponic study. Hg and MeHg were found to concentrate preferentially in the roots of the E. crassipes with little translocation in the shoots or leaves of the plant which was consistent with other macrophytes. The use

142

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

of plants to stabilize Hg should take into consideration the toxicity of mercury to plant roots, the survival rate of the plant, and the adaptability of the plant to site-specific conditions. Phytostabilization may be effectively combined with immobilization techniques to detoxify Hg in soil and/or sediment (Wang et al., 2012a). Due to the strong persistence of Hg in the environment, it may take a long time for heavily contaminated aquatic systems to fully recover through natural attenuation processes.

4. Partitioning of mercury on sediments and other substrates and application of models Partitioning most likely plays a dominant role in the distribution of Hg species between the particulate, colloidal and dissolved phase controlling the toxicity, fate, transport and bio-uptake of Hg (Stumm and Morgan, 1995). Inorganic Hg that is transported from soils and sediments to lakes is predominantly bound to dissolved or suspended organic matter (Mierle and Ingram, 1991; Hintelmann and Harris, 2004). The sorption capacity of Hg on different natural sorbents (illite, kaolinite, silica, and calcite) has been extensively studied under various test conditions (Gagnon and Fisher, 1997; De Diego et al., 2001). A list of potential sorbents of Hg is summarized in Table 3. Other studies have focused on determining partitioning constants of Hg(II) between water and sediments using natural particles (Stordal et al., 1996; Turner et al., 2001) or directly from field studies (Leermakers et al., 1995). Addition of Fe(II) ions in the presence of phlogopite (yellow to dark brown mica, general chemical formula KMg3[Si3Al]O10[F,OH]2) particles can enhance the reduction of Hg(II) (Charlet et al., 2002). The distribution coefficients Kd obtained from sediment samples (log Kd ranges from 4.5 to 6) greatly differ from one substrate to another, most likely due to the nature and abundance of respective binding sites. Nevertheless, the magnitude of log Kd exemplifies the strong affinity of Hg(II) and MeHg to sediment and suspended particles. The Kd values for Hg(II) and MeHg are comparable, but usually slightly lower for MeHg. Generally, the presence of organic matter enhances the sorption of Hg(II) to mineral surfaces (Gagnon and Fisher, 1997; Turner et al., 2001). Some studies investigated the partitioning between water and living biota, such as freshwater alga (Miles et al., 2001), bacteria (Hintelmann et al., 1993), periphyton (Cleckner et al., 1999) and phytoplankton (Watras et al., 1995). Detailed information on desorption of Hg species from surfaces is scant. Often, fast sorption of Hg onto particles is observed with strong binding of Hg to the particles (Gagnon and Fisher, 1997; Le Roux et al., 2001). It is speculated that Hg is initially sorbed and subsequently migrates into the soil lattice or is covered by organic

biofilms (Mikac et al., 1999). This theory is supported by other studies, where newly added Hg was shown to be more available for methylation than ambient Hg (Hintelmann and Harris, 2004). Hintelmann and Harris (2004) concluded that the added material is much more available for species transformation reactions while ambient Hg may be more strongly bound to particles or even incorporated into the solid matrix and not freely available for ligand exchange reactions. It is postulated that strong binding sites are occupied and saturated first. New Hg species (Hg(II) and CH3Hg þ ) entering the system will then initially associate with weaker sites and the time needed for the new Hg to find its high affinity sites and to equilibrate with the already present Hg (and other metal ions) is uncertain. Other studies have developed the colloidal pumping model (Stordal et al., 1996; Babiarz et al., 2001) postulating that Hg is initially complexed by colloids (withino24 h), which subsequently sorb or coagulate onto particles (within days). Nevertheless, sorbed Hg will be in a dynamic equilibrium with dissolved Hg, the equilibrium being shifted far toward the solid phase (characterized by very large Kd values). Hintelmann and Harris (2004) conducted studies using stable isotopes of Hg (200HgCl2 and Me199HgCl) and suspension of freshwater sediments to determine the kinetics of Hg and MeHg sorption onto sediment particles and the subsequent rate of desorption. These results indicated that equilibrium for sorption of Hg(II) and MeHg is reached between 1 h and 1 day. The initial desorption is apparently instantaneous (equilibration takes less than 30 min) with no further desorption measurable in the following two days. This study concluded that Hg partitioning between water and sediments is dependent on the solid phase concentration implying that a fraction of Hg(II) binds strongly to particles. Strong sorption sites become saturated with increasing levels of Hg(II). Weaker binding sites start to dominate Hg(II) binding resulting in greater partitioning of Hg(II) into the water. The suspended sediment partition coefficient, Kd, is the ratio of the concentration sorbed to suspended sediment in the water column to the dissolved phase water concentration at equilibrium (Lyon et al., 1997). The total benthic sediment concentration is composed of dissolved chemical plus chemical sorbed to the benthic sediment. In the literature, there are limited measured data under realistic conditions available. For divalent Hg (Hg[II]), Moore and Rarnamodomy (1984) reported Kd as a range of 1380 L/kg to 188,000 L/kg. Glass et al. (1990) reported a value of 118,000 L/kg, and Robinson and Shuman (1989) reported a range of 86,800 L/kg to 113,000 L/kg. For MeHg, Bloom et al. (1991) indicated that regardless of pH, for over three orders of magnitude, the log Kd for seston (suspended matter) was in the range of 5.5 to 6.0, which corresponds to a range from 316,000 L/kg to 1,000,000 L/kg. Babiarz et al. (2001) results show that log Kd for Hg ranged from 3.9 to 6.4 with a median of 5.0 and for MeHg

Table 3 Hg sorption capacities by selected sorbents. Sorbent

Sorption capacity

Montmorillonite Ferrihydrite Goethite (a-FeOOH) gamma-alumina (g-Al2O3) Bayerite (b-Al(OH)3) Natural zeolites (clinoptilolite) Activated carbon Furfural-based carbon Bauxite

296 to 346 mmol/kg 501 to 577 mmol/kg 0.39 to 0.42 mmol/m2 0.04 to 0.13 mmol/m2 0.39 to 0.44 mmol/m2 1.21 meq/g 65 mg/g 132 to 174 mg/g 0.8 to 5.29 L/kg

Reference

Cruz-Guzman et al. (2003) Cruz-Guzman et al. (2003) Kim et al. (2004a) Kim et al. (2004a) Kim et al. (2004a) Chojnacki et al. (2004) Babel and Kurniawan (2003) Budinova et al. (2003) Gavaskar and Chattopadhyay (2008) Yellow tuff (soft porous rock usually formed by compaction and cementation of volcanic ash or dust) 0.18 mg/g at 3000 mg/L Di Natale et al. (2006) Pozzolana (a type of slag that may be either natural – i.e., volcanic – or artificial, from a blast furnace) 0.8 mg/g at 1000 mg/L Di Natale et al. (2006) Estuarine sediments log Kd¼ 4.3 to 6.00 mL/g Turner et al. (2001)

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

143

planners, water quality managers, remedial project managers, environmental engineers and scientists to evaluate effectiveness of various control strategies. The success in utilization of models in various environmental applications has resulted in wide acceptance of models as an objective evaluation tool. However, some models are oversimplification of a complex problem. Models are only approximate representations of the complex natural processes and due to time and budget constraints involve many assumptions made by the model developer and user. Certain simplifications considered for one application might not be valid for other applications due to the uniqueness of a problem and counter-intuitive results may be produced (AWWA, 2001). Based on functionality, suspended solids and sediments (SSAS), and nutrient water quality models can be broadly categorized into three groups: (a) loading models, which simulate field scale hydrologic processes and determine the generation and transport of SSAS and nutrients from source in the upper lands to the receiving water; (b) receiving water models, which includes hydrodynamic models (hydraulics of water quality models for transport, deposition, circulation, and stratification processes), and water quality models to simulate the movement of SSAS in the water column and determine the fate and transport of contaminants, nutrients; and (c) eutrophication/ecological models, which relate to biomass production, sediment flux, growth in the water body to contaminant and/or nutrient loading, and photosynthesis. Utilities of mathematical models are: (a) constrain, synthesize, and interpret data, (b) quantify effects of different transport processes, (c) make quantitative predictions, and (d) develop insights about processes that affect sediment stability. However, the limitations of the mathematical models are: (a) data collection to support model development and calibration, and (b) level of uncertainty in results may be unacceptable to stakeholders and decision-makers. Uncertainty bounds on predictions are the key issues regarding model reliability and utility at a particular site. However, sediment stability studies conducted at a variety of sites demonstrate that useful models can be developed provided that sufficient site-specific data are available, and an experienced modeling team conducts the study (Ziegler, 2002). Simulation of Hg transport and transformation in aquatic systems is complex, involving hydrodynamic and sediment processes and Hg transport and transformation processes. Considerable site-specific data are needed to calibrate and validate Hg transport and transformation models. Based on the type of aquatic systems, Wang et al. (2004) conducted a literature review on Hg transport and transformation models. Models were categorized in three types of systems: (1) river, (2) lake, and (3) coastal. Limited numbers of Hg transport and transformation models are available in the literature. Only a few models link the modeling tool with contamination remediation and predict the remedial results in benthic sediment and overlying water.

ranged from 3.7 to 6.3, again with a median of 5.0. Lyon et al. (1997) calculated the benthic sediment partition coefficients for Hg(II) and MeHg based on the data available in the literature (see Table 4). It should be noted that partitioning coefficients for Hg species are dependent on considerable site-specific variability, and therefore judgment should always be utilized as appropriate. The loads of MeHg from surface runoff/erosion are a significant contribution to the MeHg stored in water bodies. MeHg concentrations in fish in Swedish lakes were explained in terms of the fluxes of MeHg into the water bodies from the measured direct deposition rates and runoff/erosion loads from the watershed. However, for lakes with minimal Hg input from the watershed, it suggested that bioavailable MeHg was created within the lake itself. Lyon et al. (1997) concluded that for lakes with appreciable input from the watershed, MeHg in the water body could be due to a combination of in-lake net methylation and input from deposition and/or runoff/erosion. Many processes influence the fate of contaminants in bottom sediments. Contaminants can be transported into the overlying water column by advective and diffusive mechanisms. Mixing and reworking of the upper layer of contaminated sediment by benthic organisms continually exposes contaminated sediment to the sediment–water interface where it can be released to the water column (Johnson et al., 2010). Bioaccumulation of contaminants by benthic organisms in direct contact with contaminated sediments may result in movement of contaminants into the food chain. Sediment resuspension, caused by natural and man-made erosive forces, can greatly increase the exposure of contaminants to the water column and result in the transportation of large quantities of sediment contaminants downstream (Brannon et al., 1987). Partitioning of Hg on various substrates including cap materials can limit bioavailability through three primary functions: (a) physical isolation of the contaminated sediment from the benthic environment, (b) stabilization of contaminated sediments, preventing re-suspension and transport to other sites, and (c) reduction of the flux of dissolved contaminants into the water column. Johnson et al. (2010) provided a detailed evaluation of chemical flux through a cap to assess the effectiveness of chemical containment. One dimensional advective-diffusive model can be applied once cap design objectives with respect to flux are determined, a specific capping material has been selected and characterized and a minimum cap thickness has been determined based on components for isolation, bioturbation, erosion, consolidation, and operational considerations. Mathematical models can be used to help understand the important processes and interactions that affect water quality (Veiga and Meech, 1995). These models can be used in making decisions regarding pollution control strategies by evaluating their effectiveness on water quality improvement and performing cost-benefit analysis. Models are extensively used by water resource

Table 4 Concentrations of Hg and partitioning coefficients. Description

Total Hg (HgT) concentration in aquatic sediment (ng/g/dry/wt)

Estimated Hg(II) surface water Calculated benthic concentration (ng/L) sediment kd for Hg(II) (L/kg)

Calculated benthic sediment kd for MeHg (L/kg)

Min.

Max.

Min.

753 460 170 9450 7400

0.7 5.8 1.6 6.5 0.3 0.6 NA 2000 1.8 (unfiltered) 182 (unfiltered) 1.0 (filtered) 1.6 (filtered)

80 study lakes, MN (Sorensen et al., 1990) 34 25 study lakes in Sweden (Meili, 1991) 150 Little Rock Lake, WI (Wiener et al., 1990) 10 Savannah river site, Aiken, SC (Kaplan et al., 2002) 20 Fox river (Hurley et al., 1998a) 970

NA¼ not available.

Max.

Min.

Max.

Min.

Max.

5,700 23,000 16,000 4,704 275,422

990,000 290,000 560,000 5,725 912,010

650 2,600 1,800 NA 43,651

110,000 32,000 63,000 NA 151,356

144

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

The speciation of Hg is an important consideration in any representation of Hg fate and transport. Depending upon its speciation, Hg may be either available or unavailable for biogeochemical reactions. Based upon a series of chemical procedures, four predominant species of Hg within an aquatic system have been identified into which the aquatic Hg pool may reasonably be subdivided for modeling: Hg[II], MeHg, Hg0, and Inert Hg. Feyte et al. (2012) showed applying thermodynamic and kinetic modeling to field measurements of total Hg and MeHg and ancillary parameters in sediments and porewaters can help understand the MeHg cycling and dynamics in sediments. Another researcher (Bale, 2000) described rates of transformation among four fundamental Hg species to calculate the partitioning of Hg species between the dissolved and particulate-bound phases. The various physical and biochemical rate processes included for the Hg cycle were: atmospheric deposition, diffusive exchange with the atmosphere, biogeochemical transformation of Hg species, bio-uptake by aquatic species in the food chain, sorption and desorption of Hg to particulates, diffusive exchange between the water column and sediment bed, deposition and re-suspension of suspended sediment and bound Hg, and burial of contaminated sediments. The model of Hg fate and transport developed by Bale (2000) was designed for incorporation into an advection–diffusion equation, and may be applied to aquatic systems with complex morphology and hydrodynamics. Bale (2000) demonstrated that this model, calibrated at three sites and validated at four sites within Clear Lake, California, can reasonably simulate total Hg and MeHg as functions of sediment Hg. The major uncertainty in Hg fate and transport modeling at this time is really predicting: (a) bioavailability of Hg(II) to methylating microbes, which appears to be mainly a function of Hg-sulfide speciation and Hg-DOC complexes; and (b) identification and quantification of factors affecting activity of methylating microbes (temperature, organic carbon, sulfate, etc.). Robust governing equations that describe these controls are not yet available. Thermodynamic speciation models have some value but do not capture some of the biologically mediated controls on Hg bioavailability and uptake. Kim et al. (2004c) conducted a simulation of the fate of Hg in aquatic systems by modifying WASPs as part of a remedial investigation of Onondaga Lake (New York). Remediation strategies included: dredging, capping and natural attenuation. Their model predictions for the water column generally agreed with the measured values reported in the literature for Onondaga Lake. The authors estimated the remobilization of sediment based on cutterhead suction dredging processes with a rate of 15,000 m3 of sediment per hour and sediments from the sediment–water interface up to 20 cm deep were removed during 20 working days with an estimated removed sediment volume of 2.4  106 m3. Sensitivity analyses of the model were conducted for determining the impact of transport mechanisms and speciation mechanisms. The simulation results concluded that advection, sorption and settling were important mechanisms of Hg transport in the water column. In the benthic sediment, settling of Hg from the water column was the most important input source of Hg. Both in the water column and the benthic sediment, reduction, methylation and demethylation were important mechanisms of Hg speciation. During remedial design, care should be taken by collecting performance evaluation parameters as these simulations might have uncertainties in the prediction of Hg cycling considering the complexity of Hg speciation and transport.

5. Conclusions Though considerable progress has been made worldwide to remediate mercury contaminated sediments; significant challenges

still remain. The fate and transport of Hg in the environment is greatly influenced by the speciation of Hg, its biogeochemical interactions with surrounding species, concentrations and state of species that can be bioavailable and other site-specific environmental conditions. Advances in sample collection and analytical techniques, scientific understanding and engineering approaches with computer modeling provide more cost effective and smarter methods of Hgcontaminated cleanup problems. The remedial decision should be focused on multiple site-specific parameters instead of total concentration of contaminant of concern. Innovative sustainable technologies (phytoremediation, thin layer reactive cap, application of natural materials), conventional remediation technologies (dredging, capping, and MNA), novel analytical techniques (biomarker fingerprinting, radioisotope labeling, genetic profiling), and modeling tools are available to analyze long term fate and transport of Hg and performance monitoring of cleanup technology. Significant challenges still exist in effectively combining technical knowledge and assessment and management frameworks. The technical complexity associated with Hg-contaminated sediment remediation requires that we seek and develop meaningful and reasonable understanding of the sitespecific physical and chemical characteristics with fate and transport of Hg at each site. Source control, contaminated sediment remediation, long-term performance monitoring, or their combinations are available options for cleaning up Hg—contaminated sites.

Acknowledgments We would like to thank the anonymous reviewers who reviewed the manuscript and made suggestions for its improvement. The U.S. Environmental Protection Agency through its Office of Research and Development performed the research described here. This research has not been subjected to Agency review and therefore, does not necessarily reflect the views of the Agency. Mention of trade names and products should not be interpreted as conveying official EPA approval, endorsement, or recommendation.

References Alberts, J.J., Schindler, J.E., Miller, R.W., Nutter, D.E., 1974. Elemental mercury evolution mediated by humic acid. Science 184, 895–897. Alcoa, Treatability Dredge Study for the Alcoa (Point Comfort)/Lavaca Bay Superfund Site. Draft. 2000. Amirbahman, A., Reid, A.L., Haines, T.A., Kahl, J.S., Arnold, C., 2002. Association of methylmercury with dissolved humic acids. Environ. Sci. Technol. 36, 690–695. Amyot, M., Gill, G.A., Morel, F.M.M., 1997. Production and loss of dissolved gaseous mercury in coastal seawater. Environ. Sci. Technol. 31, 3606–3611. Anderson, B.S., Hunt, J.W., Phillips, B.M., Fairey, R., Roberts, C.A., Oakden, J.M., Puckett, H.M., Stephenson, M., Tjeerdema, R.S., Long, E.R., Wilson, C.J., Lyons, J.M., 2001. Sediment quality in Los Angeles Harbor, USA: a triad assessment. Environ. Toxicol. Chem. 20, 359–370. Anjum, N.A., Ahmad, I., Valega, M., Pacheco, M., Figueira, E., Duarte, A.C., Pereira, E., 2011. Impact of seasonal fluctuations on the sediment-mercury, its accumulation and partitioning in halimione portulacoides and Juncus maritimus collected from Ria de Aveiro Coastal Lagoon (Portugal). Water Air Soil Poll. 222, 1–15. AWWA, 2001. Water Resource Planning Manual (M50). In: Maddaus, W.O. (Ed.), American Water Works Association. Denver, CO. Azcue, J.M., Zeman, A.J., Mudroch, A., Rosa, F., Patterson, T., 1998. Assessment of sediment and porewater after one year of subaqueous capping of contaminated sediments in Hamilton Harbour, Canada. Water Sci. Technol. 37, 323–329. Babel, S., Kurniawan, T.A., 2003. Low-cost adsorbents for heavy metals uptake from contaminated water: a review. J. Hazard. Mater. 97, 219–243. Babiarz, C.L., Hurley, J.P., Hoffmann, S.R., Andren, A.W., Shafer, M.M., Armstrong, D.E., 2001. Partitioning of total mercury and methylmercury to the colloidal phase in freshwaters. Environ. Sci. Technol. 35, 4773–4782. Baeyens, W., Meuleman, C., Muhaya, B., Leermakers, M., 1998. Behaviour and speciation of mercury in the Scheldt estuary (water, sediments and benthic organisms). Hydrobiologia 366, 63–79.

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

Balcom, P.H., Hammerschmidt, C.R., Fitzgerald, W.F., Lamborg, C.H., O’Connor, J.S., 2008. Seasonal distributions and cycling of mercury and methylmercury in the waters of New York/New Jersey Harbor Estuary. Mar. Chem. 109, 1–17. Bale, A.E., 2000. Modeling aquatic mercury fate in Clear Lake, Calif. J. Environ. Eng. -ASCE 126, 153–163. Barbosa, M.C., de Almeida, M.D.S., 2001. Dredging and disposal of fine sediments in the state of Rio de Janeiro, Brazil. J. Hazard. Mater. 85, 15–38. Barkay, T., Gillman, M., Turner, R.R., 1997. Effects of dissolved organic carbon and salinity on bioavailability of mercury. Appl. Environ. Microbiol. 63, 4267–4271. Batley, G.E., Burton, G.A., Chapman, P.M., Forbes, V.E., 2002. Uncertainties in sediment quality weight-of-evidence (WOE) assessments. Hum. Ecol. Risk Assess. 8, 1517–1547. Baumgarten, G., R.P.M. Panel Discussion: Sediment Remediation. In: Presented at the Technical Support Project Meeting. 2001. Beckvar, N., Field, J., Salazar, S., Hoff, R., Contaminants in Aquatic Habitats at Hazardous Waste Sites: Mercury. NOAA Technical Memorandum NOS ORCA 100. Seattle: Hazardous Materials Response and Assessment Division, National Oceanic and Atmospheric Administration, 1996. Beijer, K., Jernelov, A., 1979. Methylation of mercury in aquatic environments. In: Nriagu, J.O. (Ed.), Biogeochemistry of Mercury in the Environment. Elsevier/ North-Holland Biomedical Press, New York, pp. 203–210. Benes, P., Halvic, B., 1979. Speciation of mercury in natural waters. In: Nriagu, J.O. (Ed.), The Biochemistry of mercury in the environment. Elsevier/North-Holland Biomedical Press, pp. 175–178. Benoit, J.M., Gilmour, C.C., Mason, R.P., Heyes, A., 1999a. Sulfide controls on mercury speciation and bioavailability to methylating bacteria in sediment pore waters (33, pg 951, 1999). Environ. Sci. Technol. 33, 1780. Benoit, J.M., Mason, R.P., Gilmour, C.C., 1999b. Estimation of mercury-sulfide speciation in sediment pore waters using octanol-water partitioning and implications for availability to methylating bacteria. Environ. Toxicol. Chem. 18, 2138–2141. Bermond, A., Ghestem, J.-P., Yousfi, I., 1998. Kinetic approach to the chemical speciation of trace metals in soils[dagger]. Analyst 123, 785–789. Blake, A.C., Chadwick, D.B., White, P.J., Jones, C.A., User’s Guide for Assessing Sediment Transport at Navy Facilities. In: U.S. Navy, S.S.C. (Ed.), San Diego, CA, 2007. Blazquez, C.A., Adams, T.M., Keillor, P., 2001. Optimization of mechanical dredging operations for sediment remediation. J. Waterw. Port Coast. Ocean Eng. -ASCE 127, 299–307. Bloom, N.S., Crecelius, E.A., 1983. Determination of mercury in seawater at subnanogram per liter levels. Mar. Chem. 14, 49–59. Bloom, N.S., Effler, S.W., 1990. Seasonal variability in the mercury speciation of Onondaga Lake (New York). Water Air Soil Poll. 53, 251–265. Bloom, N.S., Gill, G.A., Cappellino, S., Dobbs, C., McShea, L., Driscoll, C., Mason, R., Rudd, J., 1999. Speciation and cycling of mercury in Lavaca Bay, Texas, sediments. Environ. Sci. Technol. 33, 7–13. Bloom, N.S., Lasorsa, B.K., 1999. Changes in mercury speciation and the release of methyl mercury as a result of marine sediment dredging activities. Sci. Total Environ. 237–238, 379–385. Bloom, N.S., Moretto, L.M., Scopece, P., Ugo, P., 2004. Seasonal cycling of mercury and monomethyl mercury in the Venice Lagoon (Italy). Mar. Chem. 91, 85–99. Bloom, N.S., Watras, C.J., Hurley, J.P., 1991. Impact of acidification on the methylmercury cycle of remote seepage lakes. Water Air Soil Poll. 56, 477–491. Bonnissel-Gissinger, P., Alnot, M., Lickes, J.P., Ehrhardt, J.J., Behra, P., 1999. Modeling the adsorption of mercury(II) on (hydr)oxides II: alpha-FeOOH (goethite) and amorphous silica. J. Colloid Interface Sci. 215, 313–322. Bonzongo, J.C., Heim, K.J., Warwick, J.J., Lyons, W.B., 1996. Mercury levels in surface waters of the Carson River Lahontan Reservoir system, Nevada: influence of historic mining activities. Environ. Pollut. 92, 193–201. Boszke, L., Kowalski, A., Glosinska, G., Szarek, R., Siepak, J., 2003. Environmental factors affecting speciation of mercury in the bottom sediments: an overview. Pol. J. Environ. Stud. 12, 5–13. Boudreau, B.P., 2000. The mathematics of early diagenesis: from worms to waves. Rev. Geophys. 38, 389–416. Braga, M.C.B., Shaw, G., Lester, J.N., 2000. Mercury modeling to predict contamination and bioaccumulation in aquatic ecosystems. Rev. Environ. Contam. Toxicol. 164, 69–92. Brannon, J.M., Hoeppel, R.E., Gunnison, D., 1987. Capping contaminated dredged material. Mar. Pollut. Bull. 18, 175–179. Bridges, T.S., Apitz, S.E., Evison, L., Keckler, K., Logan, M., Nadeau, S., Wenning, R.J., 2006. Risk-based decision making to manage contaminated sediments. Integr. Environ. Assess. Manage. 2, 8–51. Budinova, T., Savova, D., Petrov, N., Razvigorova, M., Minkova, V., Ciliz, N., Apak, E., Ekinci, E., 2003. Mercury adsorption by different modifications of furfural adsorbent. Ind. Eng. Chem. Res. 42, 2223–2229. Buffle, J., 1988. Complexation Reactions in Aquatic Systems: an Analytical Approach, England. Ellis Horwood, Chichester, West Sussex. Callister, S.M., Winfrey, M.R., 1986. Microbial methylation of mercury in upper wisconsin river sediments. Water Air Soil Pollut. 29, 453–465. Carr, S.R., Chapman, D.C., Long, E.R., Windom, H.L., Thursby, G., Sloane, G.M., Wolfe, D.A., 1996. Sediment quality assessment studies of Tampa bay, Florida. Environ. Toxicol. Chem. 15, 1218–1231. Canadian Council of Ministers of the Environment (CCME), Canadian water quality guidelines for the protection of aquatic life: guidance on the site-specific

145

application of water quality guidelines in Canada: procedures for deriving numerical water quality objectives. Canadian Council of Ministers of the Environment, Winnipeg, 2003. Chapman, P.M., 1995. Sediment quality assessment: status and outlook. J. Aquat. Ecosyst. Health 4, 183–194. Charlet, L., Peretyashko, T., Grimaldi, M., Bosbach, D., 2002. Reduction of mercury by surface Fe(II) and the formation of Hg degrees in hydromorphic soils. Geochim. Cosmochim. Acta 66 A130-A130. Chattopadhyay, S., 2005. Grain incidents and other mercury tragedies: forms, fate, and effects. In: Philip, W. (Ed.), Encyclopedia of Toxicology, Second ed. Elsevier, New York, pp. 464–469 -in-Chief:. Chattopadhyay, S., Chattopadhyay, D., 2003. Colloidal properties of sediments. In: Middleton, G.V. (Ed.), Encyclopedia of Sediments & Sedimentary Rocks. Kluwer Academic Publishers, Dordrecht; Boston. Chattopadhyay, S., Fimmen, R.L., Yates, B.J., Lal, V., Randall, P., 2012. Phytoremediation of mercury- and methyl mercury-contaminated sediments by water hyacinth (Eichhornia Crassipes). Int. J. Phytorem. 14, 142–161. Cheam, V., Gamble, D.S., 1974. Metal-fulvic acid chelation equilibrium in aqueous nano3 solution - Hg(II), Cd(II), and Cu(II) fulvate complexes. Can. J. Soil Sci. 54, 413–417. Chen, J., Gu, B.H., LeBoeuf, E.J., Pan, H.J., Dai, S., 2002. Spectroscopic characterization of the structural and functional properties of natural organic matter fractions. Chemosphere 48, 59–68. Choe, K.Y., Gill, G.A., Lehman, R., 2003. Distribution of particulate, colloidal, and dissolved mercury in San Francisco Bay estuary. 1. Total mercury. Limnol. Oceanogr. 48, 1535–1546. Chojnacki, A., Chojnacka, K., Hoffmann, J., Gorecki, H., 2004. The application of natural zeolites for mercury removal: from laboratory tests to industrial scale. Miner. Eng. 17, 933–937. Cleckner, L.B., Gilmour, C.C., Hurley, J.P., Krabbenhoft, D.P., 1999. Mercury methylation in periphyton of the Florida Everglades. Limnol. Oceanogr. 44, 1815–1825. Coates, J.A., Delfino, J.J., 1993. Sources of uncertainty in the application of the equilibrium partitioning approach to sediment quality assessment. Water Sci. Technol. 28, 317–328. Compeau, G., Bartha, R., 1983. Effects of sea salt anions on the formation and stability of methylmercury. Bull. Environ. Contam. Toxicol. 31, 486–493. Cordy, P., Veiga, M.M., Salih, I., Al-Saadi, S., Console, S., Garcia, O., Alberto Mesa, L., Velasquez-Lopez, P.C., Roeser, M., 2011. Mercury contamination from artisanal gold mining in Antioquia, Colombia: the world’s highest per capita mercury pollution. Sci. Total Environ. 410, 154–160. Cormack, R., 2001. Sediment Quality Guideline Options for the State of Alaska. Alaska Department of Environmental Conservation and OASIS/Bristol Joint Venture. Coston, J.A., Fuller, C.C., Davis, J.A., 1995. Pb2þ and Zn2 þ adsorption by a natural aluminum-bearing and iron-bearing surface coating on an aquifer sand. Geochim. Cosmochim. Acta 59, 3535–3547. Craig, P.J., Moreton, P.A., 1985. The role of speciation in mercury methylation in sediments and water. Environ. Pollut. B 10, 141–158. Cruz-Guzman, M., Celis, R., Hermosin, M.C., Leone, P., Negre, M., Cornejo, J., 2003. Sorption–desorption of lead(II) and mercury(II) by model associations of soil colloids. Soil Sci. Soc. Am. J. 67, 1378–1387. De Diego, A., Tseng, C.M., Dimov, N., Amouroux, D., Donard, O.F.X., 2001. Adsorption of aqueous inorganic mercury and methylmercury on suspended kaolin: influence of sodium chloride, fulvic acid and particle content. Appl. Organomet. Chem. 15, 490–498. Degetto, S., Schintu, M., Contu, A., Sbrignadello, G., 1997. Santa Gilla lagoon (Italy): a mercury sediment pollution case study. Contamination assessment and restoration of the site. Sci. Total Environ. 204, 49–56. Di Natale, F., Lancia, A., Molino, A., Di Natale, M., Karatza, D., Musmarra, D., 2006. Capture of mercury ions by natural and industrial materials. J. Hazard. Mater. 132, 220–225. Doody, J.P., Cushing, B.S., 2002. An evaluation of environmental dredging for remediation of contaminated sediment. In: Lehr, J., et al. (Eds.), Handbook of Complex Environmental Remediation Problems. McGraw-Hill, New York. Drace, K., Kiefer, A.M., Veiga, M.M., Williams, M.K., Ascari, B., Knapper, K.A., Logan, K.M., Breslin, V.M., Skidmore, A., Bolt, D.A., Geist, G., Reidy, L., Cizdziel, J.V., 2012. Mercury-free, small-scale artisanal gold mining in Mozambique: utilization of magnets to isolate gold at clean tech mine. J. Cleaner Prod. 32, 88–95. Drexel, R.T., Haitzer, M., Ryan, J.N., Aiken, G.R., Nagy, K.L., 2002. Mercury(II) sorption to two Florida Everglades peats: evidence for strong and weak binding and competition by dissolved organic matter released from the peat. Environ. Sci. Technol. 36, 4058–4064. Driscoll, C.T., Blette, V., Yan, C., Schofield, C.L., Munson, R., Holsapple, J., 1995. The role of dissolved organic-carbon in the chemistry and bioavailability of mercury in remote Adirondack Lakes. Water Air Soil Pollut. 80, 499–508. Duarte, A.C., Pereira, M.E., Oliveira, J.P., Hall, A., 1991. Mercury desorption from contaminated sediments. Water Air Soil Pollut. 56, 77–82. Dyrssen, D., Wedborg, M., 1991. The sulfur–mercury(II) system in natural-waters. Water Air Soil Pollut. 56, 507–519. Ebinghaus, R., Turner, R.R., de Lacerda, L.D., Vasiliev, O., Salomons, W., 1998. Mercury Contaminated Sites: Characterization, Risk Assessment and Remediation. Springer. Environment Canada, Criteria for the Assessment of Sediment Quality in Quebec and Application Frameworks: Prevention, Dredging and Remediation. 2007. Falter, R., 1999. Experimental study on the unintentional abiotic methylation of inorganic mercury during analysis: Part 1: Localisation of the compounds effecting the abiotic mercury methylation. Chemosphere 39, 1051–1073.

146

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

Feyte, S., Gobeil, C., Tessier, A., Cossa, D., 2012. Mercury dynamics in lake sediments. Geochim. Cosmochim. Acta 82, 92–112. Feyte, S., Tessier, A., Gobeil, C., Cossa, D., 2010. In situ adsorption of mercury, methylmercury and other elements by iron oxyhydroxides and organic matter in lake sediments. Appl. Geochem. 25, 984–995. Gagnon, C., Fisher, N.S., 1997. Bioavailability of sediment-bound methyl and inorganic mercury to a marine bivalve. Environ. Sci. Technol. 31, 993–998. Gao, Y., Kan, A.T., Tomson, M.B., 2003. Critical evaluation of desorption phenomena of heavy metals from natural sediments. Environ. Sci. Technol. 37, 5566–5573. Garbaciak, S., Spadaro, P., Thornburg, T., Fox, R., 1998. Sequential risk mitigation and the role of natural recovery in contaminated sediment projects. Water Sci. Technol. 37, 331–336. Gavaskar, A., Chattopadhyay, S., Treatment of Environmental Pollutants with Mineral Ores. US Patent 7,396,470. 2008. Gibson, B.D., Ptacek, C.J., Lindsay, M.B.J., Blowes, D.W., 2011. Examining mechanisms of groundwater Hg(II) treatment by reactive materials: an EXAFS study. Environ. Sci. Technol. 45, 10415–10421. Gill, G.A., Bruland, K.W., 1990. Mercury speciation in surface fresh-water systems in California and other areas. Environ. Sci. Technol. 24, 1392–1400. Gill, G.A., Fitzgerald, W.F., 1987. Picomolar mercury measurements in seawater and other materials using stannous chloride reduction and 2-stage gold amalgamation with gas-phase detection. Mar. Chem. 20, 227–243. Gilmour, C.C., Henry, E.A., 1991. Mercury methylation in aquatic systems affected by acid deposition. Environ. Pollut. 71, 131–169. Gilmour, C.C., Henry, E.A., Mitchell, R., 1992. Sulfate stimulation of mercury methylation in fresh-water sediments. Environ. Sci. Technol. 26, 2281–2287. Glass, G.E., Sorensen, J.A., Schmidt, K.W., Rapp, G.R., 1990. New source identification of mercury contamination in the great-lakes. Environ. Sci. Technol. 24, 1059–1069. ¨ Gluszcz, P., Furch, K., Ledakowicz, S., 2012. Mercury in the chlor-alkali electrolysis ¨ industry. In: Wagner-Dobler, I. (Ed.), Bioremediation of Mercury: Current Research and Industrial Applications. Caister Academic Press, Norwich, UK. Golding, G.R., Kelly, C.A., Sparling, R., Loewen, P.C., Rudd, J.W.M., Barkay, T., 2002. Evidence of facilitated uptake of Hg(II) by Vibrio anguillarum and Escherichia coli under Anaerobic and Aerobic conditions. Limnol. Oceanogr. 47, 967–975. Graham, A.M., Aiken, G.R., Gilmour, C.C., 2012. Dissolved organic matter enhances microbial mercury methylation under sulfidic conditions. Environ. Sci. Technol. 46, 2715–2723. Gu, B., Bian, Y., Miller, C.L., Dong, W., Jiang, X., Liang, L., 2011. Mercury reduction and complexation by natural organic matter in anoxic environments. Proc. Nat. Acad. Sci. 108, 1479–1483. Gu, B., Schmitt, J., Chen, Z., Liang, L., McCarthy, J.F., 1994. Adsorption and desorption of natural organic matter on iron oxide: mechanisms and models. Environ. Sci. Technol. 28, 38–46. Guentzel, J.L., Powell, R.T., Landing, W.M., Mason, R.P., 1996. Mercury associated with colloidal material in an estuarine and an open-ocean environment. Mar. Chem. 55, 177–188. Guo, L.D., Hunt, B.J., Santschi, P.H., Ray, S.M., 2001. Effect of dissolved organic matter on the uptake of trace metals by American oysters. Environ. Sci. Technol. 35, 885–893. Haitzer, M., Aiken, G.R., Ryan, J.N., 2002. Binding of mercury(II) to dissolved organic matter: the role of the mercury-to-DOM concentration ratio. Environ. Sci. Technol. 36, 3564–3570. Halbach, S., 1995. Toxicity of detrimental metal ions. In: Berthon, G. (Ed.), Handbook of Metal – Ligand Interactions in Biological Fluids – Bioinorganic Medicine. Marcel Dekker. Basel, Switzerland, pp. 749–754. Hamblin, P.F., Zhu, D.Z., Chiocchio, F., He, C., Charlton, M.N., 2000. Monitoring suspended sediment plumes by optical and acoustical methods with application to sand capping. Can. J. Civ. Eng. 27, 125–137. Hammerschmidt, C.R., Fitzgerald, W.F., 2004. Geochemical controls on the production and distribution of methylmercury in near-shore marine sediments. Environ. Sci. Technol. 38, 1487–1495. Hauge, A., Konieczny, R.M., Halvorsen, P.O., Eikum, A., 1998. Remediation of contaminated sediments in Oslo Harbour, Norway. Water Sci. Technol. 37, 299–305. Hayes, D.F., Crockett, T.R., Ward, T.J., Averett, D., 2000. Sediment resuspension during cutterhead dredging operations. J. Waterw. Port Coast. Ocean Eng. ASCE 126, 153–161. He, Z.Q., Traina, S.J., Bigham, J.M., Weavers, L.K., 2005. Sonolytic desorption of mercury from aluminum oxide. Environ. Sci. Technol. 39, 1037–1044. Henneberry, Y.K., Kraus, T.E.C., Fleck, J.A., Krabbenhoft, D.P., Bachand, P.M., Horwath, W.R., 2011. Removal of inorganic mercury and methylmercury from surface waters following coagulation of dissolved organic matter with metalbased salts. Sci. Total Environ. 409, 631–637. Hessen, D.O., Andersen, T., Lyche, A., 1990. Carbon metabolism in a Humic Lake—pool sizes and cycling through zooplankton. Limnol. Oceanogr. 35, 84–99. Hesterberg, D., Chou, J.W., Hutchison, K.J., Sayers, D.E., 2001. Bonding of Hg(II) to reduced organic, sulfur in humic acid as affected by S/Hg ratio. Environ. Sci. Technol. 35, 2741–2745. Hill, J.R., O’Driscoll, N.J., Lean, D.R.S., 2009. Size distribution of methylmercury associated with particulate and dissolved organic matter in freshwaters. Sci. Total Environ. 408, 408–414. Hines, N.A., Brezonik, P.L., Engstrom, D.R., 2004. Sediment and porewater profiles and fluxes of mercury and methylmercury in a small seepage lake in northern Minnesota. Environ. Sci. Technol. 38, 6610–6617.

Hintelmann, H., Ebinghaus, R., Wilken, R.D., 1993. Accumulation of mercury(II) and methylmercury by microbial biofilms. Water Res. 27, 237–242. Hintelmann, H., Harris, R., 2004. Application of multiple stable mercury isotopes to determine the adsorption and desorption dynamics of Hg(II) and MeHg to sediments. Mar. Chem. 90, 165–173. Hintelmann, H., Keppel-Jones, K., Evans, R.D., 2000. Constants of mercury methylation and demethylation rates in sediments and comparison of tracer and ambient mercury availability. Environ. Toxicol. Chem. 19, 2204–2211. Hinton, J.J., Veiga, M.M., 2009. Using earthworms to assess Hg distribution and bioavailability in gold mining soils. Soil Sediment Contam. 18, 512–524. Horvat, M., Kotnik, J., Logar, M., Fajon, V., Zvonaric, T., Pirrone, N., 2003. Speciation of mercury in surface and deep-sea waters in the Mediterranean Sea. Atmos. Environ. 37, S93–S108. Hosokawa, Y., 1993. Remediation work for mercury Contaminated Bay— experiences of Minamata Bay Project, Japan. Water Sci. Technol. 28, 339–348. Hudson, R.J.M., Gherini, S.A., Watras, C.J., Porcella, D.B., 1994. Modeling the biogeochemical cycle of mercury in Lakes. In: Watras, C.J., Huckabee, J.W. (Eds.), Mercury Pollution: Integration and Synthesis. Lewis, Boca Raton, Florida, pp. 473–523. Huibregtse, K., 2006. Sediment management: should it be on your radar screen? Pollut. Eng. 38 (10), 26–30. Hurley, J.P., Cowell, S.E., Shafer, M.M., Hughes, P.E., 1998a. Partitioning and transport of total and methyl mercury in the Lower Fox River, Wisconsin. Environ. Sci. Technol. 32, 1424–1432. Hurley, J.P., Krabbenhoft, D.P., Babiarz, C.L., Andren, A.W., 1994. Cycling of Mercury Across The Sediment–Water Interface In Seepage Lakes. In: Baker, L.A. (Ed.), Environmental Chemistry of Lakes and Reservoirs, pp. 425–449. Hurley, J.P., Krabbenhoft, D.P., Cleckner, L.B., Olson, M.L., Aiken, G.R., Rawlik, P.S., 1998b. System controls on the aqueous distribution of mercury in the northern Florida Everglades. Biogeochemistry 40, 293–310. Hurley, J.P., Watras, C.J., Bloom, N.S., 1991. Mercury cycling in a northern Wisconsin Seepage Lake—the role of particulate matter in vertical transport. Water Air Soil Pollut. 56, 543–551. Hylander, L.D., Pinto, F.N., Guimaraes, J.R.D., Meili, M., Oliveira, L.J., Silva, E.D.E., 2000. Fish mercury concentration in the Alto Pantanal, Brazil: influence of season and water parameters. Sci. Total Environ. 261, 9–20. Ilyushchenko, M., Panichkin, V., Randall, P., Kamberov, R., 2012. Former chloralkali factory in Pavlodar, Kazakhstan: mercury pollution, treatment options, ¨ and results of post-demercurization monitoring. In: Wagner-Dobler, I. (Ed.), Bioremediation of Mercury: Current Research and Industrial Applications. Caister Academic Press, Norwich, UK. Jacobs, P.H., Forstner, U., 1999. Concept of subaqueous capping of contaminated sediments with active barrier systems (ABS) using natural and modified zeolites. Water Res. 33, 2083–2087. Johnson, N.W., Reible, D.D., Katz, L.E., 2010. Biogeochemical changes and mercury methylation beneath an in-situ sediment cap. Environ. Sci. Technol. 44, 7280–7286. Jorgensen, B.B., 1990. The sulfur cycle of fresh-water sediments—role of thiosulfate. Limnol. Oceanogr. 35, 1329–1342. Kannan, K., Falandysz, J., 1998. Speciation and concentrations of mercury in certain coastal marine sediments. Water Air Soil Pollut. 103, 129–136. Kaplan, D.I., Knox, A.S., Myers, J., 2002. Mercury geochemistry in wetland and its implications for in situ remediation. J. Environ. Eng. -ASCE 128, 723–732. Kasprzyk-Hordern, B., 2004. Chemistry of alumina, reactions in aqueous solution and its application in water treatment. Adv. Colloid Interface Sci. 110, 19–48. Kelly, C.A., Rudd, J.W.M., Holoka, M.H., 2003. Effect of pH on mercury uptake by an aquatic bacterium: implications for Hg cycling. Environ. Sci. Technol. 37, 2941–2946. Kelly, C.A., Rudd, J.W.M., Louis, V.L., Heyes, A., 1995. Is total mercury concentration a good predictor of methyl mercury concentration in aquatic systems. Water Air Soil Pollut. 80, 715–724. Kidd, K., 2012. Bioaccumulation and biomagnification of mercury through food webs. In: Liu, G., et al. (Eds.), Environmental Chemistry and Toxicology of Mercury. John Wiley & Sons, Hoboken, NJ. Kim, C.S., Rytuba, J., Brown, G.E., 2004a. EXAFS study of mercury(II) sorption to Fe- and Al-(hydr)oxides—II. Effects of chloride and sulfate. J. Colloid Interface Sci. 270, 9–20. Kim, C.S., Rytuba, J.J., Brown, G.E., 2004b. EXAFS study of mercury(II) sorption to Fe- and Al-(hydr)oxides I. Effects of pH. J. Colloid Interface Sci. 271, 1–15. Kim, D., Wang, Q.R., Sorial, G.A., Dionysiou, D.D., Timberlake, D., 2004c. A model approach for evaluating effects of remedial actions on mercury speciation and transport in a lake system. Sci. Total Environ. 327, 1–15. Korthals, E.T., Winfrey, M.R., 1987. Seasonal and spatial variations in mercury methylation and demethylation in an Oligotrophic Lake. Appl. Environ. Microbiol. 53, 2397–2404. Kotnik, J., Horvat, M., Tessier, E., Ogrinc, N., Monperrus, M., Amouroux, D., Fajon, V., Gibicar, D., Zizek, S., Sprovieri, F., Pirrone, N., 2007. Mercury speciation in surface and deep waters of the Mediterranean Sea. Mar. Chem. 107, 13–30. Krabbenhoft, D.P., Babiarz, C.L., 1992. The role of groundwater transport in aquatic mercury cycling. Water Resour. Res. 28, 3119–3128. Kraepiel, A.M.L., Keller, K., Chin, H.B., Malcolm, E.G., Morel, F.M.M., 2003. Sources and variations of mercury in tuna. Environ. Sci. Technol. 37, 5551–5558. Krisnayanti, B.D., Anderson, C.W.N., Utomo, W.H., Feng, X., Handayanto, E., Mudarisna, N., Ikram, H., Khususiah, 2012. Assessment of environmental mercury discharge at a four-year-old artisanal gold mining area on Lombok Island, Indonesia. J. Environ. Monit. 14, 2598–2607.

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

Kudo, A., Fujikawa, Y., Mitui, M., Sugahara, M., Tao, G., Zheng, J., Sasaki, T., Miyahara, S., Muramatsu, T., 2000. History of mercury migration from Minamata Bay to the Yatsushiro Sea. Water Sci. Technol. 42, 177–184. Kudo, A., Fujikawa, Y., Miyahara, S., Zheng, J., Takigami, H., Sugahara, M., Muramatsu, T., 1998. Lessons from Minamata mercury pollution, Japan—after a continuous 22 years of observation. Water Sci. Technol. 38, 187–193. Lamborg, C.H., Tseng, C.M., Fitzgerald, W.F., Balcom, P.H., Hammerschmidt, C.R., 2003. Determination of the mercury complexation characteristics of dissolved organic matter in natural waters with ‘‘reducible Hg’’ titrations. Environ. Sci. Technol. 37, 3316–3322. Le Roux, S.M., Turner, A., Millward, G.E., Ebdon, L., Appriou, P., 2001. Partitioning of mercury onto suspended sediments in estuaries. J. Environ. Monit. 3, 37–42. Lean, D., Dillon, P., Hintelmann, H., Amyot, M., Evans, D., Mazumder, A., Arp, P., Lucotte, M., 2004. Methylmercury Photodegradation and Mercury Recycling in Aquatic Ecosystems. Leermakers, M., Elskens, M., Panutrakul, S., Monteny, F., Baeyens, W., 1993. Geochemistry of mercury in an intertidal flat of the Scheldt estuary. Neth. J. Aquatic Ecol. 27, 267–277. Leermakers, M., Meuleman, C., Baeyens, W., 1995. Mercury speciation in the scheldt estuary. Water Air Soil Pollut. 80, 641–652. Leppard, G., Burnison, B., 1983. Bioavailability: trace element associations with colloids and an emerging interest in colloidal organic fibrils. In: Leppard, G. (Ed.), Trace Element Speciation in Surface Waters and Its Ecological Implications. Plenum Press, New York, pp. 105–122. Liao, L., Selim, H.M., DeLaune, R.D., 2009. Mercury adsorption–desorption and transport in soils. J. Environ. Qual. 38, 1608–1616. Liu, C., Jay, J.A., Ika, R., Shine, J.P., Ford, T.E., 2001. Capping efficiency for metalcontaminated marine sediment under conditions of submarine Groundwater Discharge. Environ. Sci. Technol. 35, 2334–2340. Liu, G., Li, Y., Cai, Y., 2012a. Adsorption of Mercury on Solids in the Aquatic Environment. In: Liu, G., et al. (Eds.), Environmental Chemistry and Toxicology of Mercury. John Wiley & Sons, Hoboken, NJ. Liu, J., Feng, X., Zhu, W., Zhang, X., Yin, R., 2012b. Spatial distribution and speciation of mercury and methyl mercury in the surface water of East River (Dongjiang) tributary of Pearl River Delta, South China. Environ. Sci. Pollut. Res. 19, 105–112. Long, E.R., Field, L.J., MacDonald, D.D., 1998a. Predicting toxicity in marine sediments with numerical sediment quality guidelines. Environ. Toxicol. Chem. 17, 714–727. Long, E.R., MacDonald, D.D., 1998. Recommended uses of empirically derived, sediment quality guidelines for marine and estuarine ecosystems. Hum. Ecol. Risk Assess. 4, 1019–1039. Long, E.R., MacDonald, D.D., Cubbage, J.C., Ingersoll, C.G., 1998b. Predicting the toxicity of sediment-associated trace metals with simultaneously extracted trace metal: acid-volatile sulfide concentrations and dry weight-normalized concentrations: a critical comparison. Environ. Toxicol. Chem. 17, 972–974. Long, E.R., Macdonald, D.D., Smith, S.L., Calder, F.D., 1995. Incidence of adverse biological effects within ranges of chemical concentrations in marine and estuarine sediments. Environ. Manage. 19, 81–97. Long, E.R., Morgan, L.G., 1990. The Potential for Biological Effects of SedimentsSorbed Contaminants Tested in the National Status and Trends Program, NOAA Technical Memorandum NOS OMA 52. National Oceanic and Atmospheric Administration, Seattle, WA. Loux, N.T., 1998. An assessment of mercury-species-dependent binding with natural organic carbon. Chem. Speciation Bioavailability 10, 127–136. Lowry, G.V., Shaw, S., Kim, C.S., Rytuba, J.J., Brown, G.E., 2004. Macroscopic and microscopic observations of particle-facilitated mercury transport from new idria and sulphur bank mercury mine tailings. Environ. Sci. Technol. 38, 5101–5111. Lyon, B.F., Ambrose, R., Rice, G., Maxwell, C.J., 1997. Calculation of soil–water and benthic sediment partition coefficients for mercury. Chemosphere 35, 791–808. MacDonald, D.D., 1994. Approach to the assessment of sediment quality in Florida coastal waters. Prepared for the Florida Department of Environmental Protection, vol. I. MacDonald Environmental Sciences, Ltd., Ladysmith, BC. MacDonald, D.D., Ingersoll, C.G., Berger, T.A., 2000. Development and evaluation of consensus-based sediment quality guidelines for freshwater ecosystems. Arch. Environ. Contam. Toxicol. 39, 20–31. Marques, B., Lillebo, A.I., Pereira, E., Duarte, A.C., 2011. Mercury cycling and sequestration in salt marshes sediments: an ecosystem service provided by Juncus maritimus and Scirpus maritimus. Environ. Pollut. 159, 1869–1876. Martin-Doimeadios, R.C., Tessier, E., Amouroux, D., Guyoneaud, R., Duran, R., Caumette, P., Donard, O.F.X., 2004. Mercury methylation/demethylation and volatilization pathways in estuarine sediment slurries using species-specific enriched stable isotopes. Mar. Chem. 90, 107–123. Marvin-DiPasquale, M., Agee, J., McGowan, C., Oremland, R.S., Thomas, M., Krabbenhoft, D., Gilmour, C.C., 2000. Methyl-mercury degradation pathways: a comparison among three mercury-impacted ecosystems. Environ. Sci. Technol. 34, 4908–4916. Marvin-DiPasquale, M.C., Agee, J.L., Bouse, R.M., Jaffe, B.E., 2003. Microbial cycling of mercury in contaminated pelagic and wetland sediments of San Pablo Bay, California. Environ. Geol. 43, 260–267. Marvin-Dipasquale, M.C., Oremland, R.S., 1998. Bacterial methylmercury degradation in Florida Everglades peat sediment. Environ. Sci. Technol. 32, 2556–2563.

147

Mason, R., Bloom, N., Cappellino, S., Gill, G., Benoit, J., Dobbs, C., 1998. Investigation of porewater sampling methods for mercury and methylmercury. Environ. Sci. Technol. 32, 4031–4040. Matsuyama, A., Eguchi, T., Sonoda, I., Tada, A., Yano, S., Tai, A., Marumoto, K., Tomiyasu, T., Akagi, H., 2011. Mercury speciation in the water of Minamata Bay, Japan. Water Air Soil Pollut. 218, 399–412. McCready, S., Birch, G.F., Long, E.R., 2006a. Metallic and organic contaminants in sediments of Sydney Harbour, Australia and vicinity—a chemical dataset for evaluating sediment quality guidelines. Environ. Int. 32, 455–465. McCready, S., Birch, G.F., Long, E.R., Spyrakis, G., Greely, C.R., 2006b. An evaluation of australian sediment quality guidelines. Arch. Environ. Contam. Toxicol. 50, 306–315. McCready, S., Birch, G.F., Long, E.R., Spyrakis, G., Greely, C.R., 2006c. Predictive abilities of numerical sediment quality guidelines in Sydney Harbour, Australia, and vicinity. Environ. Int. 32, 638–649. McDonough, K.M., Murphy, P., Olsta, J., Zhu, Y.W., Reible, D., Lowry, G.V., 2007. Development and placement of a sorbent-amended thin layer sediment cap in the Anacostia River. Soil Sediment Contam. 16, 313–322. Massachusetts Department of Environmental Protection (MDEP), Freshwater sediment screening Benchmarks for use under the Massachusetts Contingency Plan. Massachusetts Department of Environmental Protection,Technical Update—May. Available at: /http://www.mass.gov/dep/cleanup/laws/sedscrn. docS., 2002. Meech, J.A., Veiga, M.M., Tromans, D., 1998. Reactivity of mercury from gold mining activities in darkwater ecosystems. Ambio 27, 92–98. Meili, M., 1991. The coupling of mercury and organic matter in the biogeochemical cycle—towards a mechanistic model for the boreal forest zone. Water Air Soil Pollut. 56, 333–347. Merritt, K.A., Amirbahman, A., 2009. Mercury methylation dynamics in estuarine and coastal marine environments—a critical review. Earth Sci. Rev. 96, 54–66. Mierle, G., Ingram, R., 1991. The role of humic substances in the mobilization of mercury from Watersheds. Water Air Soil Pollut. 56, 349–357. Mikac, N., Niessen, S., Ouddane, B., Wartel, M., 1999. Speciation of mercury in sediments of the Seine estuary (France). Appl. Organomet. Chem. 13, 715–725. Miles, C.J., Moye, H.A., Phlips, E.J., Sargent, B., 2001. Partitioning of monomethylmercury between freshwater algae and water. Environ. Sci. Technol. 35, 4277–4282. Miller, C.L., Mason, R.P., Gilmour, C.C., Heyes, A., 2007. Influence of dissolved organic matter on the complexation of mercury under sulfidic conditions. Environ. Toxicol. Chem. 26, 624–633. Miskimmin, B.M., 1991. Effect of natural levels of dissolved prganic-carbon (Doc) on methyl mercury formation and sediment water partitioning. Bull. Environ. Contam. Toxicol. 47, 743–750. Miskimmin, B.M., Rudd, J.W.M., Kelly, C.A., 1992. Influence of dissolved organiccarbon, Ph, and microbial respiration rates on mercury methylation and demethylation in lake water. Can. J. Fish. Aquat. Sci. 49, 17–22. Mladenova, E.K., Dakova, I.G., Tsalev, D.L., Karadjova, I.B., 2012. Mercury determination and speciation analysis in surface waters. Cent. Eur. J. Chem. 10, 1175–1182. Moo-Young, H., Myers, T., Tardy, B., Ledbetter, R., Vanadit-Ellis, W., Sellasie, K., 2001. Determination of the environmental impact of consolidation induced convective transport through capped sediment. J. Hazard. Mater. 85, 53–72. Moore, J.W., Rarnamodomy, S., 1984. Heavy Metal in Natural Waters—Applied Monitoring in Impact Assessment. Springer-Verlag, New York. Morel, F.M.M., Kraepiel, A.M.L., Amyot, M., 1998. The chemical cycle and bioaccumulation of mercury. Annu. Rev. Ecol. Syst. 29, 543–566. Newell, R.C., Hitchcock, D.R., Seiderer, L.J., 1999. Organic enrichment associated with outwash from marine aggregates dredging: a probable explanation for surface sheens and enhanced benthic production in the vicinity of dredging operations. Mar. Pollut. Bull. 38, 809–818. Nichols, M., Diaz, R.J., Schaffner, L.C., 1990. Effects of hopper dredging and sediment dispersion, Chesapeake Bay. Environ. Geol. Water Sci. 15, 31–43. NOAA, Contaminants in Aquatic Habitats at Hazardous Waste Sites: Mercury. Technical Memorandum NOS ORCA 100, 1996. NRC, 2007. Sediment Dredging at Superfund Megasites: Assessing the Effectiveness, The National Research Council of the National Academies. National Academies Press, Washington, D.C. O’Connor, T.P., 2004. The sediment quality guideline, ERL, is not a chemical concentration at the threshold of sediment toxicity. Mar. Pollut. Bull. 49, 383–385. O’Driscoll, N.J., Canario, J., Crowell, N., Webster, T., 2011. Mercury speciation and distribution in coastal wetlands and tidal mudflats: relationships with sulphur speciation and organic carbon. Water Air Soil Pollut. 220, 313–326. Oremland, R.S., Miller, L.G., Dowdle, P., Connell, T., Barkay, T., 1995. Methylmercury oxidative-degradation potentials in contaminated and pristine sediments of the Carson River, Nevada. Appl. Environ. Microbiol. 61, 2745–2753. Palermo, M.R., 1998. Design considerations for in-situ capping of contaminated sediments. Water Sci. Technol. 37, 315–321. Parks, J.W., Lutz, A., Sutton, J.A., 1989. Water column methylmercury in the Wabigoon-English River–Lake System—factors controlling concentrations, speciation, and net production. Can. J. Fish. Aquat. Sci. 46, 2184–2202. Pirrone, N., Cinnirella, S., Feng, X., Finkelman, R.B., Friedli, H.R., Leaner, J., Mason, R., Mukherjee, A.B., Stracher, G.B., Streets, D.G., Telmer, K., 2010. Global mercury emissions to the atmosphere from anthropogenic and natural sources. Atmos. Chem. Phys. 10, 5951–5964.

148

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

Ramlal, P.S., Rudd, J.W.M., Furutani, A., Xun, L.Y., 1985. The effect of pH on methyl mercury production and decomposition in lake-sediments. Can. J. Fish. Aquat. Sci. 42, 685–692. Randall, P., Ilyushchenko, M., Lapshin, E., Kuzmenko, L., 2006. Case Study: Mercury Pollution near a Chemical Plant in Northern Kazakhstan. EM Magazine. Air & Waste Management Association, Pittsburgh, PA. Ravichandran, M., 1999. Interactions between Mercury and Dissolved Organic Matter in the Florida Everglades. University of Colorado, Boulder, pp. 206. Ravichandran, M., 2004. Interactions between mercury and dissolved organic matter—a review. Chemosphere 55, 319–331. Reis, A.T., Rodrigues, S.M., Arau´jo, C., Coelho, J.P., Pereira, E., Duarte, A.C., 2009. Mercury contamination in the vicinity of a chlor-alkali plant and potential risks to local population. Sci. Total Environ. 407, 2689–2700. Richman, L.A., Wren, C.D., Stokes, P.M., 1988. Facts and fallacies concerning mercury uptake by fish in acid stressed lakes. Water Air Soil Pollut. 37, 465–473. Robertson, D.E., Sklarew, D.S., Olsen, K.B., Bloom, N.S., Crecelius, E.A., Apts, C.W. Measurement of Bioavailable Mercury Species in Fresh Water and Sediments: Final Report. 1987. Robinson, K.G., Shuman, M.S., 1989. Determination of mercury in surface waters using an optimized cold vapor spectrophotometric technique. Int. J. Environ. Anal. Chem. 36, 111–123. Rodgers, D.W., Beamish, F.W.H., 1981. Uptake of waterborne methylmercury by rainbow-trout (Salmo-gairdneri) in relation to oxygen-consumption and methylmercury concentration. Can. J. Fish. Aquat. Sci. 38, 1309–1315. Rogers, R.D., 1977. Abiological methylation of mercury in soil. J. Environ. Qual. 6, 463–467. Rytuba, J.J., 2000. Mercury mine drainage and processes that control its environmental impact. Sci. Total Environ. 260, 57–71. Sager, D.R., 2002. Long-term variation in mercury concentrations in estuarine organisms with changes in releases into Lavaca Bay, Texas. Mar. Pollut. Bull. 44, 807–815. Santschi, P.H., Allison, M.A., Asbill, S., Perlet, A.B., Cappellino, S., Dobs, C., McShea, L., 1999. Sediment transport and Hg recovery in Lavaca Bay, as evaluated from radionuclide and Hg distributions. Environ. Sci. Technol. 33, 378–391. Sarkar, D., Essington, M.E., Misra, K.C., 2000. Adsorption of mercury(II) by kaolinite. Soil Sci. Soc. Am. J. 64, 1968–1975. Schindler, D.W., Hesslein, R.H., Wagemann, R., 1980. Effects of acidification on mobilization of heavy-metals and radionuclides from the sediments of a freshwater lake. Can. J. Fish. Aquat. Sci. 37, 373–377. Schultz, T., Korhonen, P., Virtanen, M., 1995. Mercury model used for assessment of dredging impacts. Water Air Soil Pollut. 80, 1171–1180. Schuster, P.F., Shanley, J.B., Marvin-Dipasquale, M., Reddy, M.M., Aiken, G.R., Roth, D.A., Taylor, H.E., Krabbenhoft, D.P., DeWild, J.F., 2008. Mercury and organic carbon dynamics during runoff episodes from a northeastern USA watershed. Water Air Soil Pollut. 187, 89–108. Seller, P., Kelly, C.A., Rudd, J.W.M., MacHutchon, A.R., 1996. Photodegradation of methylmercury in lakes. Nature 380, 694–697. Skyllberg, U., Bloom, P.R., Qian, J., Lin, C.M., Bleam, W.F., 2006. Complexation of mercury(II) in soil organic matter: EXAFS evidence for linear two-coordination with reduced sulfur groups. Environ. Sci. Technol. 40, 4174–4180. Skyllberg, U., Drott, A., 2010. Competition between disordered iron sulfide and natural organic matter associated thiols for mercury(II)-an EXAFS study. Environ. Sci. Technol. 44, 1254–1259. Skyllberg, U., Xia, K., Bloom, P.R., Nater, E.A., Bleam, W.F., 2000. Binding of mercury(II) to reduced sulfur in soil organic matter along upland-peat soil transects. J. Environ. Qual. 29, 855–865. Smith, S.L., MacDonald, D.D., Keenleyside, K.A., Gaudet, C.L., 1996. The development and implementation of Canadian sediment quality guidelines. In: Munawar, M., Dave, G. (Eds.), Development and Progress in Sediment Quality Assessment: Rationale, Challenges, Techniques, and Strategies. SPB Academic Publishing, Amsterdam. Sorensen, J.A., Glass, G.E., Schmidt, K.W., Huber, J.K., Rapp, G.R., 1990. Airborne mercury deposition and watershed characteristics in relation to mercury concentrations in water, sediments, plankton, and fish of 80 Northern Minnesota Lakes. Environ. Sci. Technol. 24, 1716–1727. Sposito, G., 1996. The Environmental Chemistry of Aluminum. Lewis Publishers, Boca Raton, Florida. Stein, E.D., Cohen, Y., Winer, A.M., 1996. Environmental distribution and transformation of mercury compounds. Crit. Rev. Environ. Sci. Technol. 26, 1–43. Stordal, M.C., Santschi, P.H., Gill, G.A., 1996. Colloidal pumping: evidence for the coagulation process using natural colloids tagged with Hg-203. Environ. Sci. Technol. 30, 3335–3340. Stumm, W., Morgan, J.J., 1995. Aquatic Chemistry. John Wiley & Sons, New York. Sunderland, E.M., Gobas, F., Heyes, A., Branfireun, B.A., Bayer, A.K., Cranston, R.E., Parsons, M.B., 2004. Speciation and bioavailability of mercury in well-mixed estuarine sediments. Mar. Chem. 90, 91–105. Sunderland, E.M., Gobas, F.A.P.C., Branfireun, B.A., Heyes, A., 2006. Environmental controls on the speciation and distribution of mercury in coastal sediments. Mar. Chem. 102, 111–123. Swain, E.B., Jakus, P.M., Rice, G., Lupi, F., Maxson, P.A., Pacyna, J.M., Penn, A., Spiegel, S.J., Veiga, M.M., 2007. Socioeconomic consequences of mercury use and pollution. Ambio 36, 45–61. Telmer, K.H., Veiga, M.M., 2009. World emissions of mercury from artisanal and small scale gold mining. In: Mason, R., Pirrone, N. (Eds.), Mercury Fate and

Transport in the Global Atmosphere: Emissions, Measurements and Models. Springer, US, Dordrecht; New York, pp. 131–172. Tessier, L., Vaillancourt, G., Pazdernik, L., 1994. Temperature effects on cadmium and mercury kinetics in fresh-water mollusks under laboratory conditions. Arch. Environ. Contam. Toxicol. 26, 179–184. Thoma, G.J., Reible, D.D., Valsaraj, K.T., Thibodeaux, L.J., 1993. Efficiency of capping contaminated sediments in-situ.2. Mathematics of diffusion adsorption in the capping layer. Environ. Sci. Technol. 27, 2412–2419. Tiffreau, C., Lutzenkirchen, J., Behra, P., 1995. Modeling the adsorption of mercury(II) on (hydr)oxides.1. Amorphous iron-oxide and alpha-quartz. J. Colloid Interface Sci. 172, 82–93. Tomiyasu, T., Matsuyama, A., Eguchi, T., Fuchigami, Y., Oki, K., Horvat, M., Rajar, R., Akagi, H., 2006. Spatial variations of mercury in sediment of Minamata Bay, Japan. Sci. Total Environ. 368, 283–290. Turner, A., Millward, G.E., Le Roux, S.M., 2001. Sediment-water partitioning of inorganic mercury in estuaries. Environ. Sci. Technol. 35, 4648–4654. U.S.EPA, Ambient Water Quality Criteria for Mercury—1984. Office of Water Washington, DC, 1985. U.S.EPA, Report of the Sediment Criteria Subcommittee Evaluation of the Apparent Effects Threshold (AET) Approach for Assessing Sediment Quality. U.S. Environmental Protection Agency, Office of the Administrator, Science Advisory Board, Washington, DC. 1989. U.S.EPA, Mercury Study Report to Congress, Office of Air Quality Planning and Standards and Office of Research and Development, U.S. Environmental Protection Agency, EPA/452/R-96/001A-H. 1997. U.S.EPA, TRIM: total risk integrated methodology, TRIM.FaTE, technical suport document, volume II: description of chemical transport and transformation algorithms.(EPA-453/R-02-011b)—Appendix A: Derivation of MercurySpecific Algorithms, North Carolina, 2002. U.S.EPA, Evaluation Report-EPA Can Better Implement Its Strategy for Managing Contaminated Sediments. Report No. 2006-P-00016,March 15, 2006. In: Office of the Inspector General (Ed.), 2006. U.S.EPA, National Listing of Fish Advisories, 2010, EPA-820-F-11-014, November 2011. 2011. U.S.EPA, 2012. Alcoa/Lavaca Bay, Calhoun County, Texas EPA Region 6 Congressional District 14. EPA Publication Date: March 14, 2012. Available at: /http:// www.epa.gov/region6/6sf/pdffiles/0601752.pdfS. Uchimiya, M., Stone, A.T., 2009. Reversible redox chemistry of quinones: impact on biogeochemical cycles. Chemosphere 77, 451–458. Ullrich, S.M., Ilyushchenko, M.A., Kamberov, I.M., Tanton, T.W., 2007. Mercury contamination in the vicinity of a derelict chlor-alkali plant. Part I: Sediment and water contamination of Lake Balkyldak and the River Irtysh. Sci. Total Environ. 381, 1–16. Ullrich, S.M., Tanton, T.W., Abdrashitova, S.A., 2001. Mercury in the aquatic environment: a review of factors affecting methylation. Crit. Rev. Environ. Sci. Technol. 31, 241–293. UNEP, Mercury in the Aquatic Environment:Sources, Releases, Transport and Monitoring, published by UNEP’s Division of Technology, Industry and Economics (DTIE) Chemicals Branch, prepared by The Joint Group of Experts on the Scientific Aspects of the Marine Environmental Protection (GESAMP), Working Group 37, Geneva, Switzerland. 2011. Vale, C., Ferreira, A.M., Micaelo, C., Caetano, M., Pereira, E., Madureira, M.J., Ramalhosa, E., 1998. Mobility of contaminants in relation to dredging operations in a mesotidal estuary (Tagus estuary, Portugal). Water Sci. Technol. 37, 25–31. Van den Berg, G.A., Meijers, G.G.A., Van der Heijdt, L.M., Zwolsman, J.J.G., 2001. Dredging-related mobilisation of trace metals: a case study in the Netherlands. Water Res. 35, 1979–1986. Van der Zee, F.R., Cervantes, F.J., 2009. Impact and application of electron shuttles on the redox (bio) transformation of contaminants: a review. Biotechnol. Adv. 27, 256–277. Veiga, M., Meech, J., 1999. Reduction of mercury emissions from gold mining activities and remedial procedures for polluted sites. In: Azcue, J.M. (Ed.), Environmental Impacts of Mining Activities. Springer-Verlag, pp. 143–162. Veiga, M.M., Meech, J.A., 1995. HgEx—a heuristic system on mercury pollution in the Amazon. Water Air Soil Pollut. 80, 123–132. Wallschlager, D., Desai, M.V.M., Spengler, M., Wilken, R.D., 1998a. Mercury speciation in floodplain soils and sediments along a contaminated river transect. J. Environ. Qual. (27), 1034–1044. Wallschlager, D., Desai, M.V.M., Spengler, M., Windmoller, C.C., Wilken, R.D., 1998b. How humic substances dominate mercury geochemistry in contaminated floodplain soils and sediments. J. Environ. Qual. (27), 1044–1054. Wallschlager, D., Desai, M.V.M., Wilken, R.D., 1996. The role of humic substances in the aqueous mobilization of mercury from contaminated floodplain soils. Water Air Soil Pollut. (90), 507–520. Wang, J.X., Feng, X.B., Anderson, C.W.N., Xing, Y., Shang, L.H., 2012a. Remediation of mercury contaminated sites—a review. J. Hazard. Mater. 221, 1–18. Wang, Q.R., Kim, D., Dionysiou, D.D., Sorial, G.A., Timberlake, D., 2004. Sources and remediation for mercury contamination in aquatic systems—a literature review. Environ. Pollut. 131, 323–336. Wang, S., Zhang, M., Li, B., Xing, D., Wang, X., Wei, C., Jia, Y., 2012b. Comparison of mercury speciation and distribution in the water column and sediments between the algal type zone and the macrophytic type zone in a hypereutrophic lake (Dianchi Lake) in Southwestern China. Sci. Total Environ. 417, 204–213.

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149

Watras, C.J., Morrison, K.A., Host, J.S., Bloom, N.S., 1995. Concentration of mercury species in relationship to other site-specific factors in the surface waters of Northern Wisconsin Lakes. Limnol. Oceanogr. 40, 556–565. Weber, J.H., 1993. Review of possible paths for abiotic methylation of mercury(II) in the aquatic environment. Chemosphere 26, 2063–2077. Wells, D., Hill, J., A synthesis of sediment chemical contaminant studies in the Maryland Coastal Bays. Maryland’s Coastal Bays—Ecosystem Health Assessment. Available at: /http://www.dnr.state.md.us/coastalbays/publications/ Chapter5.2.pdfS., 2004. Wiener, J.G., Fitzgerald, W.F., Watras, C.J., Rada, R.G., 1990. Partitioning and bioavailability of mercury in an experimentally acidified Wisconsin Lake. Environ. Toxicol. Chem. 9, 909–918. Wilhelm, S.M., 1999. Generation and disposal of petroleum processing waste that contains mercury. Environ. Prog. 18, 130–143. Winger, P.V., Lasier, P.J., White, D.H., Seginak, J.T., 2000. Effects of contaminants in dredge material from the lower Savannah River. Arch. Environ. Contam. Toxicol. 38, 128–136.

149

Xia, K., Skyllberg, U.L., Bleam, W.F., Bloom, P.R., Nater, E.A., Helmke, P.A., 1999. X-ray absorption spectroscopic evidence for the complexation of Hg(II) by reduced sulfur in soil humic substances. Environ. Sci. Technol. 33, 257–261. Xiong, Z., He, F., Zhao, D., Barnett, M.O., 2009. Immobilization of mercury in sediment using stabilized iron sulfide nanoparticles. Water Res. 43, 5171–5179. Yang, Y., Liang, L., Wang, D., 2008. Effect of dissolved organic matter on adsorption and desorption of mercury by soils. J. Environ. Sci. —China 20, 1097–1102. Yin, Y.J., Allen, H.E., Huang, C.P., Sparks, D.L., Sanders, P.F., 1997. Kinetics of mercury(II) adsorption and desorption on soil. Environ. Sci. Technol. 31, 496–503. Zheng, W., Liang, L., Gu, B., 2012. Mercury reduction and oxidation by reduced natural organic matter in anoxic environments. Environ. Sci. Technol. 46, 292–299. Zhou, P., Yan, H., Gu, B.H., 2005. Competitive complexation of metal ions with humic substances. Chemosphere 58, 1327–1337. Ziegler, C.K., 2002. Evaluating sediment stability at sites with historic contamination. Environ. Manage. 29, 409–427.