Mercury contaminated sediment sites—An evaluation of remedial options

Environmental Research 125 (2013) 131–149

Contents lists available at ScienceDirect

Environmental Research

0013-93

http://d

n Corr

E-m

pran105

Sandip.C

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, USAb Tetra Tech, Inc., 250 West Court Street, Suite 200W, Cincinnati, OH 45202, USA

a r t i c l e i n f o

Available online 12 March 2013

Keywords:

Mercury

Sediment

Remediation

Partitioning coefficients

Modeling

51/$ - see front matter Published by Elsevier

x.doi.org/10.1016/j.envres.2013.01.007

esponding author.

ail addresses: [email protected],

[email protected] (P.M. Randall),

[email protected] (S. Chattopadh

a b s t r a c t

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.

1. Introduction

Mercury accumulates in sediment globally from many physi-cal, chemical, biological, geological and anthropogenic environ-mental processes (U.S.EPA, 1997, 2006; Benoit et al., 1999b; Bragaet 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 fromabandoned Hg mines, gold-mining activities (Ebinghaus et al.,1998; Meech et al., 1998; Veiga and Meech, 1999; Telmer andVeiga, 2009; Cordy et al., 2011; Drace et al., 2012; Krisnayanti et al.,2012), ore refining, and products and processes such as recycledmercury processing or the chlor-alkali industry (Randall et al., 2006;Ullrich et al., 2007; Reis et al., 2009; Gluszcz et al., 2012; Ilyushchenkoet al., 2012). With artisanal and small scale gold mining, Telmer andVeiga (2009) estimates that approximately 1000 metric tons/yr of Hgwas released from at least 70 countries. Approximately, 350 metric

Inc.

yay).

tons/yr of this amount is directly emitted to the atmosphere while theremainder, 650 metric tons/yr, is released into the hydrosphere (i.e.rivers, lakes, soils, tailings). Indirect (non-point source) Hg contam-ination is largely attributed to atmospheric deposition (wet and dry)originating from coal-fired power plants. Global mercury emissionsfrom coal-fired power plants were estimated at approximately 850metric tons/yr (Pirrone et al., 2010). Other indirect sources to theaquatic environment can be attributed to runoff to water bodies orleaching from groundwater flows in the upper soil layers.

At large contaminated sediment sites, engineers and scientistsface many challenges primarily due to the large volumes ofsediment that are typically involved. Usually, the remediationtimeframes and spatial scales are in many ways unprecedented.The complexities and high costs associated with characterizationand cleanup are magnified by evolving regulatory requirementsand the difficulties inherent in tracking the contaminants inaquatic environments. Remedial strategies often require unex-pected adjustments in response to new knowledge about siteconditions or advances in technology (such as improved dredge orcap design or in situ sorption materials and treatments). Regula-tors and engineers adapt continuously to evolving conditions andenvironmental responses. Depending on site specific conditions,

www.elsevier.com/locate/envres

www.elsevier.com/locate/envres

http://dx.doi.org/10.1016/j.envres.2013.01.007



http://crossmark.dyndns.org/dialog/?doi=10.1016/j.envres.2013.01.007&domain=pdf



mailto:[email protected]




P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149132

effective management of Hg contaminated sites includes moreadaptive site investigation, remedy selection, and remedy imple-mentation. In a remedial investigation and screening of poten-tial alternatives, practitioners may consider various approachesincluding: (a) dredging and excavation with sediment dewateringand handling, (b) sediment treatment of dredged materials byphysical, chemical and biological processes, (c) in-situ/ex-situsubaqueous capping in combination with dredging), (d) in-situ/ex-situ capping treatments that contain contaminants by chemi-cal and biological processes, (e) containment in contained dis-posal facilities (CDFs), contained aquatic disposal (CAD), andlandfills, (f) monitored natural recovery (MNR), (g) phytoreme-diation and (h) combination of above mentioned options. Hgcontaminated sites sometimes implement a suite of remedialapproaches to clean-up the site. For example, at the Lavaca BayPoint Comfort site (TX, USA), Alcoa spent approximately $110million to implement several remedial options (i.e. dredging,capping, MNR, disposal in a CDF, long-term monitoring) in andaround the bay (U.S.EPA, 2012).

There are several U.S. agencies (i.e. U.S. Army Corps ofEngineers (USACE), National Oceanic and Atmospheric Adminis-tration (NOAA), U.S. Geological Survey (USGS), and the Depart-ment of the Navy) that manage contaminated sediment programsincluding Hg contaminated sediments sites. U.S. EPA’s Great LakesNational Program Office estimates that 76 million cubic yards ofcontaminated sediments in the Great Lakes require remediationat an approximate cost of $1.6 to $4.4 billion (U.S.EPA, 2006). TheDepartment of the Navy has estimated that there are more than200 contaminated sediment sites which they manage with aprojected remediation cost to cleanup of $1.3 billion dollars(Blake et al., 2007). Although the U.S. EPA has historicallyemphasized that no presumptive remedy exists for sediments,most removal actions have included dredging (e.g. 56 of the 63sediment sites in 2006) (Huibregtse, 2006; U.S.EPA, 2006). How-ever, remedial actions recently have included reactive thin-layercapping, phytoremediation, and other remedial alternatives toreduce the resuspension and mobility of contaminants, carbonfootprint, and other factors. Moreover, selection of remedialoptions is dependent on site-specific conditions that constituteacceptable levels of effectiveness and performance.

In Hg contaminated site cleanup, most remedial technologiesfocus on highly contaminated areas and are not suitable forremediating vast, diffuse, Hg contaminants at low concentrations.Speciation of Hg is an important consideration that concerns theidentification and quantification of specific chemical forms ofHg and is a critical determinant of its mobility, reactivity, andpotential 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 expectedthat the speciation and transport properties of Hg might changeas the result of implementing a remedial action. However, theeffectiveness of many remediation practices and long-term relia-bility has not been adequately assessed (Degetto et al., 1997).

Fish advisories on contaminated water bodies are plentiful inthe U.S. because of the inorganic Hg(II) that is converted to MeHgand thus, moves up the food chain. In the U.S. fish advisories aredue to five (5) bioaccumulative chemical contaminants: mercury,polychlorinated biphenyls (PCBs), chlordane, dioxins, and dichlor-odiphenyltrichloroethane (DDT). In 2010, the EPA reported morethan 4598 fish advisories with 81% due to Hg (U.S.EPA, 2011). Theaccumulation of Hg in the food chain depends primarily on theconcentration of MeHg, rather than total Hg, in water. It has beenreported in the literature that only a minor fraction of Hg innatural water is in the form of MeHg; however, MeHg concentra-tions generally in surface water are extremely low, near thedetection limit of the currently available techniques (o50

femto-molar)(Kraepiel et al., 2003). To protect aquatic life, ascientific benchmark or reference point called the sedimentquality guidelines (SQG) was developed (U.S.EPA, 1989; Longand 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 Councilof 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 forobserving adverse biological effects in aquatic systems for che-mical contaminants (i.e. metals and metalloids, organic compounds,polycyclic aromatic hydrocarbons (PAHs), organochlorine pesticides,and others).

NOAA annually collects and analyzes sediment samples fromsites located in coastal marine and estuarine environmentsthroughout the U.S. They evaluated a wide variety of marinesediment toxicity studies that were conducted in laboratories andin the field for the effects of sediment concentrations on benthicorganisms. They established effects range-low (ERL) and effectsrange-medium (ERM) concentrations for each constituent evalu-ated. ERL and ERM values are those concentrations above whichadverse biological effects were seen in 10% and 50%, respectively.ERL and ERM values together define the concentration ranges thatwere (1) rarely and (2) frequently associated with adverse effects.Long et al. (1995) reported ERL and ERM for total Hg as 0.15 mgper kilogram (mg/kg) and 0.71 mg/kg dry weight basis, respec-tively. ERL is not a threshold below which sediment toxicity isimpossible and above which it is likely. Rather, an ERL is simplya low point on a continuum of bulk chemical concentrations insediment that roughly relate to sediment toxicity (Beckvar et al.,1996; O’Connor, 2004). Another criterion limit is the apparenteffects threshold (AET) values derived from a correlation of theweight of evidence from multiple matched chemical and biologi-cal effect data sets (laboratory toxicity testing on field sedimentsamples). The AET value for a particular contaminant is defined asthe sediment concentration above which an adverse biologicaleffect is always statistically observed (U.S.EPA, 1989). For exam-ple, 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’sLavaca Bay Point Comfort site, Texas USA. For PAHs, the ERL andERM 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 fresh-water and marine sediments: a threshold effect level (TEL) and aprobable effect level (PEL). These two values were adopted for theassessment of sediment quality in Quebec and were developedusing a nationally-approved protocol (Canadian Council of Min-isters of the Environment (CCME), 2003). The Hg TEL and PELvalues for freshwater sediment are 0.17 mg/kg and 0.49 mg/kg,respectively; and the same for marine sediments are 0.13 mg/kgand 0.70 mg/kg, respectively (Smith et al., 1996; Canadian Councilof Ministers of the Environment (CCME), 2003; EnvironmentCanada, 2007).

Massachusetts Department of Environmental Protection (MDEP)assessed the screening criteria by adopting consensus-based thresh-old 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 Envir-onmental Protection (MDEP), 2002). The TECs are intended toidentify contaminant concentrations below which harmful effectson sediment-dwelling organisms are not expected. These concen-trations may not necessarily be protective of higher level organismsexposed to bioaccumulating chemicals. These consensus-based TECvalues were chosen because they incorporate a large data set,

Table 1Sediment 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-basedTEC

Arsenic 8.5 70 35 9.79

Cadmium 1.2 9.6 3 0.99

Chromium 81 370 62 43.4

Copper 34 270 390 31.6

Lead 46.7 218 400 35.8

Hg 0.15 0.71 0.41 0.18

Nickel 20.9 51.6 110 22.7

Zinc 150 410 410 121

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

Forms of Hg Water (lg/L) Oil (lg/L)

Hg0 50 2000

XHgX NA infinite

HgCl2 70,000,000 410,000

HgS 10 Very low, o10

HgO 50,000 low

CH3HgCl Very high 1,000,000

P.M. Randall, S. Chattopadhyay / Environmental Research 125 (2013) 131–149 133

provide an estimate of central tendency that is not unduly affectedby extreme values, and incorporate sediment quality guidelines thatrepresent a number of approaches for developing sediment bench-marks. A list of these consensus-based TECs is provided in Table 1.

It is well known that sediments are reservoirs for toxiccompounds. Mercury discharged into the hydrosphere rapidlybecomes associated with particulate matter and incorporated inbottom sediments. At the water–sediment interface, it is impor-tant to understand the mechanisms and environmental variablesthat drive or constrain methylation dynamics. Although diage-netic processes in the sediments can modify and redistribute themercury between solid and solution phases, immobilization bysedimentation dominates for most elemental contaminants suchas 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 andmechanisms that influence the aqueous phase and the solid phaseMeHg concentrations.

2. Biogeochemistry of mercury

2.1. Mercury speciation

The distribution, mobility and biological availability of chemicalelements depend not simply on their concentrations but, critically,on the forms in which they occur in natural systems. This possiblemobility and bioavailability are the result of the reactivity of tracemetals (especially Hg) in sediment, in other words, their localizationin different sediment components, which is usually called speciationor the different physico-chemical forms or oxidation states of thesame element (Bermond et al., 1998). The mobility and availabilityof Hg in aquatic environments is influenced by various processesincluding 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

(NOM), and other substrates that can settle out and effectivelyremove Hg from the mobile aqueous phase (Choe et al., 2003;Zheng et al., 2012). The aqueous speciation and coordination of Hghave been well-documented (Benes and Halvic, 1979; Dyrssen andWedborg, 1991; Hurley et al., 1998b; Bloom and Lasorsa, 1999;Horvat et al., 2003; Kim et al., 2004b; Skyllberg et al., 2006; Kotniket al., 2007; Balcom et al., 2008; Skyllberg and Drott, 2010; Gibsonet al., 2011; Matsuyama et al., 2011; O’Driscoll et al., 2011; Liu et al.,2012b; Mladenova et al., 2012; Wang et al., 2012b). The oxidationstates of Hg in aqueous systems are 0, þ1, and þ2. In typical aeratedwater, Hg(II) is most stable. Aqueous Hg(II) speciation and coordina-tion in the absence of other strongly complex ligands is largelydictated by hydrolysis reactions. At low pH, the hexaqua ion(Hg(H2O)6

2þ) is octahedrally coordinated by water molecules, withHg–O bond lengths of 2.34–2.41 A (Kim et al., 2004b). As the pH israised and the extent of hydrolysis increases to HgOHþ 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 forHg(II) to form mononuclear linear, double-coordinated complexes, asalso occurs in halides, oxyanions, and certain solids. The stability ofHg(OH)2 complexes in the pH range of natural water (5 to 9) (Kimet al., 2004b).

Metal speciation in aquatic environments is affected by inor-ganic and organic ligands present in water. The relative impor-tance of each ligand for metal complexation will depend on theconcentration of the metal and the ligand, and the bindingstrength (conditional stability constants) for the metal–ligandcomplex. Among the inorganic ligands, hydroxide, chloride, andsulfide are considered important in controlling the speciation ofHg in water (Ravichandran, 1999). In the absence of any sig-nificant chelators, Hg—hydroxide complexes (Hg(OH)2, HgOHþ)are likely to be the important species in most freshwaters (Stummand Morgan, 1995). Hg—chloride complexes (HgCl2, HgCl4

2� ,HgCl3

�) are thought to be important at low pH and/or highchloride concentrations. In sediment and aquatic environmentscontaining dissolved sulfide (including some oxic surface waterswhere nanomolar levels of sulfide and thiols have been detected),Hg is hypothesized to form Hg—sulfide species (Dyrssen andWedborg, 1991; Hudson et al., 1994). Among the organic ligands,sulfur-containing ligands (e.g. cysteine, mercaptoacetate) bindto 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 partsper trillion in several physical and chemical forms in oxic surfacefreshwaters. The partitioning of Hg between the dissolved, col-loidal and particulate phases varies widely spatially, seasonallyand by depth in the water column. Some of this variation seemsto be related to temporal changes in living particulate matter,mostly phytoplankton and bacteria (Hurley et al., 1991). The con-centration of particulate Hg per unit particle weight is relativelyconstant reflecting perhaps sorption equilibrium between dissolvedand particulate phases (Meili, 1991). The exact chemical form ofparticulate Hg is unknown, although most of it is probably tightlybound in suspended organic matter. Sorption of Hg to oxyhydroxidesmay also be important in lakes. The commonly observed enrichmentof MeHg and Hg(II) in anoxic waters of lakes may result from thesedimentation of Hg-laden oxyhydroxides of iron and manganesefrom the epilimnion and their dissolution in the anoxic hypolimnionlayers (Meili, 1991).

2.2. Presence of natural organic matter (NOM)

Natural organic matter (NOM) consists of redox reactive butchemically heterogeneous organic matter substances that existubiquitously in aquatic environments (Chen et al., 2002; Zheng


et al., 2012). NOM is known to bind trace metals strongly, aff-ecting their speciation, solubility, mobility and toxicity (Buffle,1988). There is increasing evidence that dissolved organic matter(DOM) interacts very strongly with Hg, affecting its speciationand bioavailability in aquatic environments (Loux, 1998; Lamborget al., 2003; Ravichandran, 2004; Miller et al., 2007; Schusteret al., 2008; Hill et al., 2009; Henneberry et al., 2011; Grahamet al., 2012). Strong interactions between Hg and DOM have alsobeen indicated by positive correlation between their concentra-tions in many natural waters. NOM interacts with Hg in severaldifferent ways, affecting the transport, transformation and bioa-vailability of Hg. Major mechanisms by which NOM may adsorbonto mineral surfaces involve: (a) anion exchange (electrostaticinteraction), (b) ligand exchange surface complexation, (c) hydro-phobic interaction, (d) entropic effect, (e) hydrogen bonding, and(f) cation bridging (Gu et al., 1994). In addition to its recognizedmetal binding capabilities (e.g. Hg(II), Fe(III)), (Zhou et al., 2005),NOM can mediate electron transfer (Gu et al., 2011) and is one ofthe most important electron shuttles in natural environments(Van der Zee and Cervantes, 2009). For example, NOM has beenshown to accept electrons from microorganisms and then transferthem to electron acceptors such as Fe(III) and iron oxide minerals(Uchimiya and Stone, 2009). NOM forms extremely strong com-plexes between Hg(II) and reduced sulfur (Skyllberg et al., 2006).Strong complexation facilitates the mobility of Hg from sedi-ments (Wallschlager et al., 1996) into streams (Mierle andIngram, 1991), lakes (Driscoll et al., 1995), and groundwater(Krabbenhoft and Babiarz, 1992). This enhanced mobility resultsin increased water column concentrations of Hg in otherwisepristine lakes and streams. Complexation also affects the parti-tioning of Hg to suspended solids in the water column and thesequestration of Hg to sediments.

Dissolved organic matter (DOM) is also known to promote(Weber, 1993) or inhibit (Miskimmin et al., 1992) the formationof toxic and bioaccumulative MeHg species. DOM plays animportant role in bioaccumulation and biomagnifications of Hg(Cormack, 2001). Contaminants in sediment are taken up bybenthic organisms in a process called bioaccumulation. Whenanimals higher in the food chain feed on contaminated organisms,the toxins are taken into their bodies, moving up the food chain inincreasing concentrations in a process known as biomagnifica-tion. Complexation with DOM limits Hg(II) availability to methy-lating bacteria and CH3Hgþ availability for bioaccumulation(Barkay et al., 1997). Earthworms applied in laboratory tests atgold mining sites showed high bioavailability in organic soils andlow bioavailability in lateritic soil and clay-rich sediments(Hinton and Veiga, 2009). Humic and fulvic acid fractions ofDOM are also capable of reducing ionic Hg to the volatileelemental Hg (Alberts et al., 1974), increasing the flux of Hg fromwater and soil to the atmosphere. Although the role of humicmatter in the methylation of Hg remains unclear, it seems thatorganic carbon can enhance methylation by stimulating theactivity of heterotrophic microorganisms, or through direct abio-tic methylation of Hg by humic or fulvic substances. Also, DOMenhances the formation of Hg0 from Hg(II) in photochemicalreactions (Ravichandran, 2004), which could reduce the avail-ability 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 ‘‘softsphere’’ of highly polarizable electrons in its outer shell. Soft metals(like Hg) show a pronounced preference for ligands of sulfur, the lesselectronegative 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 itstoxicity in humans is attributed to the high affinity of Hg for sulfur-containing proteins such as metallothionein and glutathione. Stronginteractions between Hg and organic matter found in sediment andaquatic environments are attributed to the binding of Hg with sulfur-containing functional groups in organic matter.

Sulfur is a minor constituent in DOM, ranging from approxi-mately 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 þ6. The stabilityconstant for Hg(II) complexation with an oxidized sulfur ligand,SO4

2� , is 101.3, whereas, the stability constant for Hg(II) com-plexation with a reduced sulfur ligand, S2� , is 1052.4. Thereduced sulfur sites are expected to be important for Hg binding.Generally, hydrophobic acid fractions of DOM (which includes thehumic and fulvic acid fractions) had significantly higher reducedsulfur content than the low molecular weight hydrophilic acidfractions. Even if it is assumed that only a small fraction (about 2%as suggested by Amirbahman et al. (2002)) of the reduced sulfuris available for binding with Hg in natural systems, the strongbinding sites in organic matter far exceed the amount of Hgavailable in natural aquatic systems.

Since binding of Hg to DOM under natural conditions iscontrolled by a small fraction of DOM molecules containingreactive thiol functional groups (Haitzer et al., 2002), a positivecorrelation may not always exist between Hg and dissolvedorganic carbon (DOC) concentration (Hurley et al., 1998b). InHurley’s study, a strong relationship between filtered Hg speciesand DOC was evident for wetlands draining in Wisconsin, but notin the Everglades suggesting either differences in the binding sitesof the two regions or non-organic complexation in the Everglades.In general, positive correlation between Hg and DOC concentra-tions could be expected in cases where Hg is released and co-transported with the organic matter (Wallschlager et al., 1996).But, in systems where water column Hg is primarily derived fromdirect atmospheric sources, correlation between Hg and DOC mayor may not be present. In both cases, significant differences can beexpected in the reactivity of DOM with Hg depending on thestructural 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 coordinatedwith Hg. Ravichandran (2004) reported that the organic matterand mercaptoacetic acid (HS–CH2–COOH), a thiol-containingcompound, caused a dramatic increase in Hg release from theFlorida Everglades (up to 35 mM total dissolved Hg) from redcinnabar (HgS), a relatively insoluble solid (Ksp¼10–36.8). DOMalso inhibited the precipitation of metacinnabar (black HgS), avery insoluble solid (Ksp¼10–36.4), at an initial Hg concentrationof 5�10�88 M (Ravichandran, 2004). In contrast to sulfur-containing ligands, oxygen-containing ligands, such as acetic acidand EDTA, dissolve very little or no Hg from cinnabar.

Hg speciation models calculated as a function of sulfideconcentrations and pH suggest that HgSaq

0 , Hg(S2H)� , Hg(SH)20,

and HgSsol are likely to be the most important species (Hurleyet al., 1994). Sulfur cycling in aquatic sediments involves bothreductive and oxidative processes (Jorgensen, 1990) and theyoften play a significant role in forming metal complexes. Thestability constants for Hg–organic sulfur (HgRSþ) complexes aremuch lower than for inorganic sulfide. The stability constants forthese complexes are (Dyrssen and Wedborg, 1991; Benoit et al.,1999a) indicated below.

Hg2þþHS�2HgSaq

0þHþ K ¼ 1026:5

Hg2þþ2HS�2Hg S2Hð Þ

�þHþ K ¼ 1032:0


Hg2þþ2HS�2Hg SHð Þ2

0 K ¼ 1037:5

Hg2þþRS�2HgRSþ K ¼ 1022:1

However, the Hg–DOM binding constants in natural environmentsare reported to be much higher. The binding constants for the HgRSþ

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 withinorganic sulfides. Differences between the above values may beattributed to the Hg/DOM ratio in these studies, wherein Hg may bebound by a single thiol (RSH) group, by bidentate aromatic andaliphatic thiols and phenols or carboxyls (Xia et al., 1999; Hesterberget al., 2001; Drexel et al., 2002). These high stability constants indicatethat organic matter can easily outcompete sulfide for the complexa-tion of Hg in anoxic environments.

2.4. Effect of chloride and sulfate

The sorption of Hg onto particles can be significantly affected bythe presence of complexing ligands, like chloride and sulfate that arepresent in freshwater or seawater. These ligands affect the sorption ofHg due to several possible processes including (a) formation of stable,non-sorbing metal–ligand aqueous complexes; (b) formation ofmetal–ligand ternary surface complexes, which at high metal andligand concentrations can lead to surface precipitation; (c) comp-etitive ligand sorption to particle surfaces, effectively blocking themore reactive sorption sites at the surface; and (d) reduction ofpositive charges at particle surfaces, thus lowering the electrostaticrepulsion of cations by surfaces (considering ligands are anionsand pH levels are below the point of zero charge, pHpzc, of themineral particles). Kim et al. (2004a) reported that presence ofchloride and sulfate reduced or increased sorption of Hg(II) ongoethites (a-FeOOH), g-alumina (g-Al2O3), and bayerite (b-Al[OH]3),which are useful surrogates for the natural sediments. Hg(II) sorbsstrongly 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 tothe Al (O,OH)6 octahedra that compose the bayerite structure. Hg(II)sorbs weakly to g-alumina due to the conversion of the hydrated g-alumina surface to a secondary bayerite-like phase. Over the chlorideconcentration range of 10�5 M to 10�2 M, a lowering in Hg sorptionon a-FeOOH, g-Al2O3, and b-Al(OH)3 was from 0.42 mmol/m2 to0.07 mmol/m2, 0.06 mmol/m2 to 0.006 mmol/m2, and 0.55 mmol/m2 to0.39 mmol/m2, respectively. This reduction in Hg(II) sorption isprimarily due to the formation of stable, non-sorbing aqueous HgCl2complexes in solution, limiting the amount of free Hg(II) available tosorb. At higher chloride concentrations (Cl�Z10�3 M) and a pH of 6,the large proportion of unsorbed, aqueous Hg(II) facilitated reduc-tion of Hg(II) to Hg(I) and the formation of Hg2Cl2 (s) (calomel) orHg2Cl2 (aq) species. Sulfate, in contrast, enhanced Hg(II) sorption overthe sulfate concentration range 10�5 M to 0.9 M, increasing Hgsurface coverage on a-FeOOH, g-Al2O3, and b-Al(OH)3 from0.39 mmol/m2 to 0.45 mmol/m2, 0.11 to 0.38 mmol/m2, and 0.36 to3.33 mmol/m2, respectively. This effect might be due to the sorptionor accumulation of sulfate ions at the substrate interface, effectivelyreducing the positive surface charge that electrostatically inhibitsHg(II) sorption.

2.5. Methylation and bioaccumulation

In the aquatic environment (including water, sediment, and biotaphases), most of the mercury is in the inorganic and organic forms ofdivalent Hg(II) compounds with Hg(0) being a considerable fraction ofthe 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 ofmonomethylmercury (CH3Hg). Thus, conversion of ionic Hg to MeHg

is an important link in the bioaccumulation of Hg in fish andultimately its toxicity to humans and wildlife. MeHg production insediment and aquatic systems is not a simple function of total Hgconcentration in the system. MeHg formation is influenced by anumber of environmental factors including temperature, pH, redoxpotential, 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 itstransport are shown in Fig. 1.

Dissolved Hg is distributed among several chemical forms:elemental Hg (Hgaq

0 ), which is volatile but relatively non-reactive,a number of mercuric species (Hg[II]), and organic Hg, mainlyMeHg, dimethyl (Me2Hg), and some ethyl (EtHg) Hg. In general,and particularly in stratified systems, concentrations of Hg0 arehigher near the air–water interface, whereas levels of Hg andMeHg are higher near sediments.

Hg methylation is mainly a microbial mediated process, withabiotic methylation likely to be important in organic-rich lakes(Ullrich et al., 2001). Bacteria assimilate Hg through passivediffusion of neutrally-charged species (Barkay et al., 1997) aswell as by active uptake of both charged and uncharged Hg (Kellyet al., 2003). Wetland sediments commonly have a loweroxidation–reduction potential (ORP), or redox potential (Eh),thereby promoting the reduction of Hg(II) to Hg(I) or Hg0. Oneof the key pathways, ORP influences Hg speciation through itseffect on sulfur chemistry. Decreases in ORP promote microbial-mediated sulfur-reduction, which in turn promotes Hg methyla-tion. Furthermore, the accumulation of reduced sulfur, primarilyas dissolved sulfide, will precipitate inorganic Hg as a highlyinsoluble HgS mineral, cinnabar (red coloration) or meta-cinnabar(black and slightly more soluble). Increases in dissolved sulfideconcentrations result in decreases in Hg methylation ratesbecause inorganic Hg is removed as a sparingly soluble solid(Gilmour et al., 1992). It is important to note that an excess ofsulfide can lead to formation of soluble Hg–S complexes.

Wetlands typically have very high concentrations of organicmatter (OM) due to the slow rate of OM oxidative degradationoccurring in this environment. The OM may either act as a sorbentor provide high concentrations of dissolved ligands that form verystrong complexes to Hg(II) (Cheam and Gamble, 1974; Wall-schlager et al., 1996, 1998a, 1998b). In microbial methylation,complexation with DOM, commonly measured as DOC, plays animportant role in Hg bioavailability. DOM is believed to limit theamount of inorganic Hg available for uptake by methylatingbacteria because DOM molecules are generally too large to crossthe cell membrane of bacteria (Golding et al., 2002; Kelly et al.,2003). Further, there is competition between various ions forbinding sites. Under acidic conditions, Hþ may compete withHg2þ for binding sites on DOM and limit its sequestration, leavingmore inorganic Hg available for uptake or methylation by organ-isms. The presence of other ions such as Naþ and Kþ may alsocompete with Hg2þ for binding to negatively charged functionalgroups on DOM (Kidd, 2012). In sulfate-limited environmentswhere microbes may be utilizing organic matter as an energysource, DOM may have a stimulating effect on microbial growthand thus enhance methylation rates in the water column andsediments (Watras et al., 1995). It may be hypothesized thatwhen OM is largely labile and readily biodegradable, it maypromote methylation by stimulating microbial growth; whenthe OM is relatively recalcitrant and consists of high molecularweight humic and fulvic acids, then it may contribute to abioticmethylation. Lean et al. (2004) reportedly observed that most ofthe MeHg in the south-central lakes of Ontario (Kegimkujik Park,Nova Scotia) is not dissolved or bound to small particulatematerial but bound to DOM (less than 1 kD).

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 Hgin surface waters, Hg(II), is not present as the free ion Hg(II) butshould be complexed in variable amounts to hydroxide(Hg(OH)þ , Hg(OH)2, Hg(OH)3

�), and to chloride (HgClþ , HgClOH,HgCl2, HgCl3

� , HgCl42�) ions depending on the pH and the chloride

concentration. Even in oxic surface waters, some or much of theHg(II) might be bound to sulfides. In addition, an unknownfraction of Hg(II) is likely bound to humic acids, the assemblageof poorly defined organic compounds that constitute 50 to 90% ofthe DOC in natural waters. According to Meili (1991) nearly 95%of inorganic oxidized Hg in lakes is bound to DOM. Through itsbinding to DOM, Hg can be mobilized from the drainage basin andtransported to lakes (Morel et al., 1998). The reactions of ionic Hgare relatively fast, and it is thought that the various species ofHg(II), including those in the particulate phase, are at equilibriumwith each other. In the organometallic species of Hg, the carbon-to-metal bonds are stable in water because they are partlycovalent and the hydrolysis reaction, which is thermodynamicallyfavorable (and makes the organometallic species of most other metalsunstable), is kinetically hindered. As a result, the dimethylmercuryspecies, Me2Hg (CH3HgCH3), is non-reactive. The monomethyl spe-cies, MeHg, is usually present as chloro- and hydroxocomplexes(CH3HgCl andCH3HgOH) in oxic water.

Once MeHg is formed, DOM facilitates its solubility, and thusincreasing the water column concentration, and transports throughcomplexation (Miskimmin, 1991). At the same time, complexationwith DOM also tends to limit its uptake in biota (Driscoll et al., 1995).Apart from DOM, concentration and bioaccumulation of MeHg in fishis also affected by pH, temperature, redox potential, concentrations ofaluminum and calcium, fish age and food source, and other factors(Watras et al., 1995). Temperature and season influence the avail-ability and accumulation of Hg in addition to the factors alreadydiscussed. Changes in temperature can affect Hg concentrations inorganisms either directly by affecting metabolic rate and therebyexposure, or indirectly by influencing the methylation of Hg andtherefore enhancing availability. Rates of MeHg or inorganic mercuryuptake increase with increasing aqueous concentrations and/orincreasing temperature in the water for some species, such asphytoplankton, gastropods, fish (Rodgers and Beamish, 1981;

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

�
Transalkylation reaction by other methylated metals like leadand 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 influenceMeHg concentrations in aquatic systems. An inverse relationship


between low lake pH and elevated Hg levels in fish has beenobserved in several regions of the northern hemisphere. Richmanet al. (1988) have reviewed some of the mechanisms proposed toaccount for the elevated Hg levels in biota and acid-stressed lakes.Generally, low pH conditions facilitate the release of heavy metalsand particulate matter from sediments. But the data on thepartitioning and mobility of Hg has been somewhat contradictory(Ullrich et al., 2001). The solubility and mobility of Hg and MeHgis pH dependent, however, even pH dependency may oversimplifythe many factors that regulate Hg uptake by fish and the mobilityof Hg and MeHg. Some studies show mobility of Hg is higherin the acidic pH range (Duarte et al., 1991), but Schindler et al.(1980) reported that lake water acidification caused a higherproportion of Hg to bind to particulates, thereby decreasing thesolubility of Hg in the water column. Ramlal et al. (1985) reportedthat the amount of dissolved Hg in sediment porewater wasfound to decrease with decreasing pH. Further, metallic Hg wasreported to be readily soluble at low pH with organic acids(Meech et al., 1998). Despite contradictory evidence on the effectof pH on inorganic Hg solubility, studies show that MeHg is moresoluble under low pH conditions. As indicated, this may be dueto various geochemical factors including decreased availabilityof binding sites on organic matter and redox conditions (Ullrichet al., 2001; Kidd, 2012). Furthermore, the pH may effect theproduction of dimethylHg. Neutral or acidic pH conditions appearto favor the production of MeHg over dimethylHg (Beijer andJernelov, 1979; Ullrich et al., 2001) and alkaline pH favors theformation of dimethylHg (NOAA, 1996). Methylation may occur atalkaline conditions, however, the favorable pH range for inorganicHg methylation was reported to be between pH 2 and pH 5.5(Falter, 1999). Acidic pH was also reported to enhance themethylation of Hg in the Carson River-Lahontan Reservoir system(Bonzongo et al., 1996). Kelly et al. (2003) studied the effect ofincreasing hydrogen ion (Hþ) concentrations on the uptake ofHg(II) by an aquatic bacterium. Even small changes in pH (7.3 to6.3) resulted in large increases in Hg(II) uptake, in defined media.The increased rate of bioaccumulation was directly proportionalto the concentration of Hþ . Lowering the pH of Hg solutionsmixed together with natural dissolved organic carbon, or withwhole lake water, also increased bacterial uptake of Hg(II). Usingboth defined inorganic solutions and lake water, uptake of Hg(II)was faster at lower pH, and the increased rate of uptake was notrelated to changes in neutral Hg species such as HgCl2 or Hg(OH)2.Rather, uptake of both charged and uncharged Hg(II) speciesappeared to increase as Hþ increased, indicating a facilitatedbacterial Hg(II) uptake process that responds to pH. Hg(II) uptakerate by bacteria (for example, Vibrio anguillarum) under aerobic oranaerobic conditions is controlled by the collective concentrationof a number of available Hg(II) species, both charged anduncharged, which indicates that a cell-mediated process isimportant in determining how much Hg(II) enters the cell. Inaddition to the bacterial Hg(II) uptake process, an increase inbioavailable Hg concentration at reduced pH conditions is due tothe Hþ outcompeting Hg(II) for binding sites on DOM particles,or because Hþ may displace Hg(II) by protonating sulfhydrylmoieties that bind Hg(II) to DOM (Benoit et al., 1999a) or byreplacing Hg(II) on negatively charged functional groups surfacessuch as clay minerals.

2.7. Interactions with minerals

Numerous studies have been conducted to examine Hg(II)sorption and release (desorption) from natural and syntheticparticles, 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

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 productionof MeHg (Stein et al., 1996). The treatment and removal of Hgfrom sediments are necessary for control of methylation andbioaccumulation. As discrete particles and/or as coatings on othermineral surfaces in natural systems, especially in well-weatheredsoil and sediments with low natural organic matter, crystallineand amorphous alumina play significant roles (Sposito, 1996;Kasprzyk-Hordern, 2004). Because of their chemical propertiesand physical structure, aluminum (hydr)-oxides are efficient sinksfor many contaminants including cations of Hg, Pb, Zn, Cd, and Sr(Coston et al., 1995; Sarkar et al., 2000). In addition to Hgspeciation, surface characteristics of aluminosilicates (surfacearea, porosity, pore size distribution) can have a significantimpact on the fate of these contaminants. However, desorptionof heavy metals from sediments and aluminosilicates can bemuch slower and/or nonreversible (Yin et al., 1997; Gao et al.,2003), which may lead to significant challenges due to the longertime needed for the cleanup (He et al., 2005). Furthermore,hydrous ferric oxides (sometimes referred to as iron oxyhydr-oxides), can act as sorbents to complex mercury species on thesurface whereby mercury is retained in the solution and trans-ported as sorbed species (Rytuba, 2000; Lowry et al., 2004; Liuet al., 2012a).

2.8. Colloids and suspended materials

The distribution of Hg species between the particulate, colloi-dal, and dissolved phase affects the toxicity, transport, and bio-uptake of Hg in water and sediment systems (Chattopadhyay andChattopadhyay, 2003). Among these size classes, the colloidalphase has been inferred to play several key roles in the biogeo-chemistry of Hg: (a) regulating the concentration of dissolvedmetal ion and neutral complexes in solution as the binding of freemetal ion reduces acute toxicity; regulating neutral complexes inorder to affect Hg transport across bacterial walls (Leppard andBurnison, 1983; Benoit et al., 1999a); (b) downstream transportvector due to the relatively large surface area of colloids;(c) uptaking of MeHg by bacteria, fungi, zooplankton, and mol-lusks either by direct consumption or the free ion activity model(Hessen et al., 1990; Guo et al., 2001). The phase distributionof both Hg and MeHg in freshwater may differ from that inmarine environments because freshwater is generally lowerin ionic strength, higher in alkalinity, and higher in DOC(Chattopadhyay, 2005). Stordal et al. (1996) found that a majorportion of the filtered fraction (12 to 93%) was associated with the0.4 mm to 1 kDa size fraction and that the colloidal-phase Hgconcentration was correlated with carbon content. Comparingmarine water and freshwater results, it is also reported thatcolloid coagulation in high salinity water was shown to be amajor removal mechanism for Hg. Guentzel et al. (1996) reportedHg in the 0.4 mm to 1 kDa fraction was a large portion of thefiltered phase (37 to 88%) in coastal marine water in theOchlockonee River Estuary, in Florida. They presented evidencethat thiol functional groups associated with organic carbon wereimportant in the partitioning of Hg in the colloidal phase. Theirstudy also reported the first colloidal-phase MeHg concentrationsin marine environments. Babiarz et al. (2001) studied partitioningof Hg and MeHg in 15 freshwater systems located in the upperMidwest (Minnesota, Michigan, and Wisconsin) and the SouthernUnited States (Georgia and Florida). Though they reported thatthe correlation between Hg and organic carbon in the colloidalphases was not statistically significant (r2r0.14; pZ0.07), MeHgin the colloidal phase and dissolved phase were correlated with


the concentration of organic carbon as:

MeHgC ¼ 0:15� OCCþ0:0018 ðr2 ¼ 0:54, po0:01Þ

MeHgD ¼ 0:006� OCDþ0:0458 ðr2 ¼ 0:23, p¼ 0:02Þ

where MeHg concentrations in colloidal phase and dissolvedorganic carbon phase are indicated as MeHgC and MeHgD,respectively.

2.9. Salinity

A negative correlation between the rate of MeHg formationand conductivity (salinity) in estuarine sediments has beenreported (NOAA, 1996). The rate of MeHg formation is lower inmore saline environments because the bicarbonate component ofseawater slows methylation of Hg(II) under both aerobic andanaerobic conditions (Compeau and Bartha, 1983). The release ofreactive 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 sulfideconcentrations as a result of rapid sulfate reduction in salinewater compared to sulfate-limited freshwater environments.Gilmour and Henry (1991) reported that the percentage of totalHg that is MeHg is higher in freshwater sediments (up to 37%) andwater (up to 25% in aerobic water and 58% in anoxic bottomwater) than in estuarine and marine water (o5%) and associatedsediments (o5%). Dissolved reactive Hg (inorganic species) formsthe majority of the total Hg in open oceans (Bloom and Crecelius,1983; Gill and Fitzgerald, 1987). The study conducted by Babiarzet al. (2001) did not show strong trends in Hg concentration (ng/g)with suspended particulate matter, conductivity, organic carbon inthe o0.4 mm fraction, pH, or percent organic matter, but MeHgconcentration (ng/g) was correlated with conductivity (mS/cm�1) ofthe riverine water as

½MeHg� ¼ 14:6þ0:0:2952½conductivity� ðr2 ¼ 0:21, p¼ 0:03Þ

2.10. Methylation dynamics in estuarine and coastal environment

Merritt and Amirbahman (2009) reviewed the various pro-posed mechanisms to hypothesize the influence of inorganicHg(II) methylation rate and MeHg accumulation in estuarineand coastal marine environments. The examined whether themethylation of Hg is controlled or limited by the metabolicactivity of bacteria (i.e. sulfate-reducing type), the availability oftotal Hg or the geochemical speciation of Hg and/or the processesresponsible for the methylation/demethylation processes. Theseprocesses may be influenced by depth-dependent balancebetween MeHg production and consumption in estuary andmarine environments.

Field and research data from the Lavaca Bay Superfund site (i.e.a stratified estuarine environment) has been well published in theliterature. Mason et al. (1998) reported development of porewatersampling 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 calcu-lated sediment accumulation rates as high as approximately2 cm/yr at near-shore sites near the Alcoa facility. These authorsalso predicted Hg concentrations in surface sediments to decreaseexponentially with a recovery halftime of 472 years. Sager(2002) reported that monitoring of Hg in organisms since 1977shows a gradual downward trend in Hg in crabs and finfish.Hammerschmidt and Fitzgerald (2004) reported various geo-chemical controlling factors on the production and distributionof 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 reportedthe average log Kd for inorganic Hg and MeHg as 4.8970.43 and2.7070.78, respectively. Particulate Hg concentration in surfacewater reportedly varied from 350 ng/g to 1610 ng/g and iscontrolled by physical mixing of polluted fluvial particulates withrelatively unpolluted marine particulates (Baeyens et al., 1998).Dissolved Hg species show large seasonal variations essentiallycontrolled by the redox conditions and bacterial and phytoplank-ton activities. It has been observed that in oxidizing conditionsin freshwater, inorganic Hg (Hg(II)) is predominately present asHg(OH) 2 and HgOHCl, and in the reducing conditions the sulfur-based Hg species (such as HgS) are primarily present (Kannan andFalandysz, 1998), but in marine systems, Hg complexes withchloride (Cl�) to form soluble HgCl4

2� . Hg also complexes with thehumic and strong complexing ligands of organic substances(Leermakers et al., 1995) present in the aquatic systems but thisphenomenon is subdued in the marine waters because of thepresence of abundant Cl� ions (Leermakers et al., 1995; Morelet al., 1998). Thus, in Lavaca Bay, Hg predominately exists in Hg0,Hg(OH)2 and HgOHCl formed in the oxidizing conditions, and asHgS in the reducing environment. In addition, organomercuryspecies, monomethylmercury (CH3Hgþ) and dimethylmercury([CH3]2Hg) were present. The transformation of Hg between theelemental form, the ionic form and organomercury is controlledby biotic and abiotic processes in aerobic and anaerobic environ-ments. MeHg is generally a product of methylation of inorganicHg carried out by the sulfate-reducing bacteria in the aquaticenvironments 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 inter-face of the sediment system or in the first 5 cm of the sedimentlayer (Bloom et al., 1999). MeHg concentrations were high in themarshes of Lavaca Bay. These marshes are highly productiveenvironments with detrital carbon (plant litter) which drive themicrobial process promoting methylation of the bioavailable Hg.The higher Hg levels in these areas indicated that these were theareas where Hg was being transferred into the food chains asmost of the higher tropic level organisms feed on the lowertrophic level organisms found in these areas. The lower organismson the lower levels of the food chain pick up their Hg from thesediments of these areas.

3. Remediation options

3.1. Dredging

Environmental dredging is of special interest because it can beexpensive and technically challenging to implement. Dredgingitself may create exposures (for example, through the resuspen-sion of buried Hg contaminants), but it removes persistentcontaminants (and their associated potential for transport andrisk) from the aquatic environment permanently. Whether todredge contaminated sediments has proved to be one of the mostcontroversial aspects of decision-making for sediment remedia-tion sites (Barbosa and de Almeida, 2001; Blazquez et al., 2001;Doody and Cushing, 2002; Bridges et al., 2006; NRC, 2007). One ofthe examples of dredging was in Japan’s Minamata Bay, where Hgconcentrations as high as 600 mg/kg were detected in settledsediment (Hosokawa, 1993). High concentration was observedat the inner part of the bay. MeHg in most sediment waso0.005 mg/kg and 0.03 mg/kg at maximum. Sediment was dom-inantly soft silty or silty clay and contained rich sulfides. TheJapanese Environmental Protection Agency established standardsfor removal of contaminated bottom sediments for Hg; thestandards took into consideration release rate, dilution rate by


the tide, and biological accumulation with safety factor. A dred-ging removal standard of 25 mg Hg/kg was established. In Octo-ber 1977, the Minamata Bay project was initiated and completedin March 1990 (Kudo et al., 1998, 2000; Tomiyasu et al., 2006).The total amount of dredged sediment removed was about1.5 million tons of sediments at an estimated cost of $500 million(Kudo et al., 1998). The depth of sediment, which contained Hggreater 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 byHosokawa (1993), monitoring data showed that at most samplingpoints, Hg concentrations were found to be below 5 mg/kg afterdredging. The maximum concentration of Hg was 8.75 mg/kg.

Despite the results seen at Minamata Bay, dredging activitiesmay cause adverse environmental threats if they are not wellplanned and implemented (Nichols et al., 1990; Schultz et al.,1995; Van den Berg et al., 2001). Dredging-induced sedimentre-suspension is a major environmental concern. Given no sig-nificant disturbance, buried Hg and other metals are generallysorbed by sediment, and can generally be regarded as safelyseparated from the overlying water. Activities such as dredging,shipping, and natural occurrences, such as storms and tides, canremobilize Hg that was sorbed by sediment (Van den Berg et al.,2001). Sunderland et al. (2004) defined the active zone of thesediments as sediments that can potentially exchange Hg withthe water column and buried sediments through re-suspension,diffusion and burial. This thickness of the active layer is a functionof the depth of biological mixing and the depth of physicalmixing/continual reworking (Boudreau, 2000). Sunderland et al.(2006) found that re-suspension and mixing of tidal sedimentsenhanced MeHg production. Bloom and Lasorsa (1999) conductedlaboratory testing to simulate ocean dredging. They reported thatabout 5% of MeHg and less than 1% of total Hg can be releasedfrom contaminated sediment as a result of dredging. It is alsonoteworthy that sediment pore water, which usually containshigh concentrations of Hg, can readily release Hg into the over-lying water (Gilmour et al., 1992).

After comparing various dredging techniques, it was suggestedthat a combination of mechanical and hydraulic dredging pro-duces the least sediment resuspension (Hauge et al., 1998; Wanget al., 2004). Mathematical models were developed to estimatedredging costs, efficiency, and environmental effects (Hayes et al.,2000; Blazquez et al., 2001). During dredging, oxygen in overlyingwater can enter buried anoxic sediment and possibly oxidizeand release contaminants (Vale et al., 1998). Under undisturbedconditions, the formation of MeHg is restricted primarily to theuppermost 3 cm to 15 cm of benthic sediment (Sunderland et al.,2004). Sediment and porewater Hg speciation show that theupper 10 cm of sediment are dynamic regions of MeHg formationand 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 afterdredging, some buried sediment is mixed with surface sediment,or water, which can produce an environment with high sulfateand organic matter concentrations that favor the production ofMeHg (Bloom and Lasorsa, 1999). Studies conducted in the UnitedKingdom (UK) showed that water discharged from dredging siteshad higher concentrations of organic matter that favors theproduction of MeHg (Newell et al., 1999).

Furthermore, dredging of the contaminated sediment is onlya 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 directdisposal are more common alternatives (Wang et al., 2004). Thetwo most widely used disposal sites are land and sea (Barbosa andde Almeida, 2001). The disposal of dredged sediments poses apotential threat to the surrounding environment. Increased

turbidity is usually observed at dredge disposal sites (Nicholset al., 1990). The leakage of Hg into groundwater systems fromdisposal sites is another concern. In Georgia, the lower SavannahRiver showed elevated concentrations of some metals (includingHg) in living organisms close to an upland dredge disposal site(Winger et al., 2000). Contaminated, dredged sediment confine-ment is widely used to prevent potential adverse environmentaleffects from dredge disposal. Adjusting pH to an optimal level(based on laboratory and pilot tests) or adding sorbent materialsis a common method to immobilize heavy metals. For example,adding materials containing iron is quite effective in immobilizingmercury (e.g. FeS has been shown to exchange its Fe(II) withHg(II) to form HgS(s) (Xiong et al., 2009).

Dredging can be very effective in cleaning up heavily Hg-contaminated sediment; however it has disadvantages that needto be carefully addressed such as sediment resuspension. and cost(i.e. cost of remediation including environmental dredging, couldbe as high as $1409/m3 (Doody and Cushing, 2002)). A dredgingdemonstration project conducted by Alcoa evaluated a hydrauliccutterhead technology (10-in. hydraulic dredge with a 14-inchsuction pipe and 36-inch cutterhead) to control Hg residuals. Thedredging cost was $251,000 (disposal costs not included) for9500 yd3 of sediment (3–12% solid)(Alcoa, 2000). Six acres ofvery soft plastic clay sediment was dredged in 20 days. Results ofthe study showed the following key observations: (a) hydraulicdredging could be readily implemented at this site; (b) offsitetransport of Hg on tidal flows moving through and around thecurtained-off dredging unit were minimal; (c) a large mass ofmercury was removed (2300 lbs) with 60,000 yd3 to 80,000 yd3 ofsediment and placed in a confined disposal facility; (d) increasedHg concentrations in oysters above the historical observed back-ground in the wider bay did not occur. The dredging operationswere conducted with multiple passes with sampling betweenpasses to define the residual present after each pass. There was anotable increase in residual concentration between passes 2 and3, apparently reflecting exposure of more highly contaminatedsediment. Overall, the pass-to-pass concentration changes werenot statistically significant. The pilot study was judged to besuccessful and the data collected were important in the evalua-tion 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 orproper isolating materials to cover and separate the contaminatedsediments from the water column. A cap can reduce contaminationrisk by one or multiple activities: (a) physical isolation of thecontaminated sediment from the aquatic environment; (b) stabiliza-tion/erosion protection of contaminated sediment; and (c) chemicalisolation/reduction of the movement of dissolved and colloidallytransported contaminants into the water.

In situ capping is on-site placement of proper coveringmaterial over contaminated sediment in aquatic systems. Labora-tory treatability studies suggest that in situ capping can beeffective in reducing the impact of Hg contamination in aquaticsystems. In ex-situ capping, contaminated sediment is dredgedand relocated to another site, where one or multiple isolatinglayers 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 cap-ping and dredged material capping or ex situ capping, whichinvolves removal of sediments and placement at a subaqueoussite, followed by placement of a cap. Dredged material capping isa disposal alternative that has been used for sediments dredgedfrom navigation projects, and may also be suitable for disposal


of sediments and treatment residues from remediation projects.There are two forms of dredged material capping: (a) level bottomcapping, where a mound of dredged material is capped, and(b) contained aquatic disposal (CAD) in which dredged materialis placed in a depression or other area that provides lateralconfinement 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 morecomplex design with geotextiles, liners and multiple layers withcombinations of passive or reactive materials. A variation in capscould involve the removal of contaminated sediments to somedepth, followed by capping the remaining sediments in-place.This is suitable where capping alone is not feasible because ofhydraulic or navigational restrictions on the waterway depth.Experimental tests show that capping material composed of amixture of sand and finer particles (silty sands per ASTMclassification) can sorb Hg and other heavy metals (Moo-Younget al., 2001). Moo-Young et al. (2001) showed that cappingmaterials can sorb 99.9% of the Hg from sediment, whichcontained Hg between 200 and 500 micrograms per liter (mg/L).This test showed that a capping layer can be a good barrierbetween Hg-contaminated sediment and the overlying water.

In-situ capping field studies were conducted in HamiltonHarbor, Canada, which has high concentrations of zinc, copper,Hg, and other metals. A cap, approximately 35 cm thick andcomposed mostly of sand, was placed in the system to containpolluted sediment (Azcue et al., 1998). After one year of in situcapping, a field study investigated the effectiveness of the cap. Ingeneral, Hg concentrations were found to be low (o5 mg/kg) inthe capping layer, while the concentration of Hg in the originalsediment ranged between 0.43 g/kg and 0.96 g/kg (Azcue et al.,1998). Johnson et al. (2010) conducted simulated tests using 1-cmsand and found no significant Hg migration into the cap over aperiod of approximately 8 months. The total Hg flux in theoverlying water was undetectable for the capped case comparedto about 10�3 ng/m2/s from exposed uncapped sediment. Theseresults suggest that a passive sand (though do not bind significantamount of Hg on its inert structure) may provide a temporarybarrier to contain Hg in the native sediment. In general, more Hgis found in sediment than in water because Hg compounds areattracted to and may attach to the small grains and particles(including decayed plants and animals) that make up the sedi-ment. Hines et al. (2004) measured and reported Hg and MeHg at1 cm to 2 cm resolution in sediment pore water and sedimentcores from Spring Lake in Minnesota. Total Hg accumulation inthe sediment was 21.4 mg/m2/year, two orders of magnitudegreater than the accumulation of MeHg (0.20 mg/m2/year). Thehighest solid phase concentrations of MeHg were observedpersistently at the sediment surface and declined sharply withdepth. Pore water profiles showed a small diffusive flux of MeHgfrom sediment to water (5 ng/m2/month). Springtime pore waterconcentrations of MeHg were relatively low (approximately0.5 ng/L) and increased by late summer to early fall (1.5–2.2 ng/L), showing a correlation with maxima in sulfate reducing activityat 5 cm and 15 cm. Advective transport carrying MeHg deeperinto the sediment was evident in summer and fall. The percent ofHg present as MeHg was highest in the water column above thesediment (10%) and decreased with sediment depth in both thesolid and pore water phases.

The major advantages of in situ capping are low cost, extensivesuitability to a wide range of contaminants, and low adverseenvironmental effects (Azcue et al., 1998; Palermo, 1998). As in-situ capping is not a treatment process, long-term environmentaleffects, including possible remobilization of contaminated sedi-ment, need to be carefully considered by performance monitoringat regular time intervals after the installation of the capped layer.

However, there is a possibility that buried Hg may pass throughthe capping layer and enter into the overlying water due tovarious reasons (hydrodynamic flows, consolidation, transforma-tion, diffusion, etc.). Hydrodynamic currents caused by humanactivities or natural processes, such as shipping, tide, and ground-water flow may scour the capping layer and release Hg into thewater. Groundwater flow through the cap material can reduce theefficiency of capping significantly (Liu et al., 2001). The movementof benthic organisms may also facilitate the remobilization ofburied Hg. Sediment consolidation, due to gravity, can move Hgfrom buried sediment into the capping layer. This sedimentconsolidation may be a more important factor in the transfer ofHg from buried sediment into the capping layer than moleculardiffusion of Hg (Moo-Young et al., 2001).

Although a pilot test conducted in a Canadian harbor suggested nosignificant sediment resuspension due to capping (Hamblin et al.,2000), there is always the possibility of resuspension of settledsediment due to the placement of the capping layer. Such resuspen-sion can be the cause for transforming some of the inorganic Hg intoorganic Hg (MeHg). MeHg can escape into the overlying water moreeasily than inorganic Hg (Wang et al., 2004).

Site characterization is the preliminary and crucial stepto decide whether a contaminated aquatic system is suitable forcapping. In general, aquatic environments with low hydrody-namic flows, such as lakes and bays, are good candidates forcapping (Thoma et al., 1993). The type of capping material thatcan be used depends on the hydrodynamic and geotechnicalconditions, and target contaminants. Sand-size and other finematerials are good for quiescent environments (Palermo, 1998).For erosive systems, coarser materials should be considered(Palermo, 1998). Gavaskar and Chattopadhyay (2008) proposedthe idea of using reactive material as an in-situ cap using variousnatural minerals that effectively sorb contaminants and non-toxicto the benthic organisms. Zeolite is a good candidate for applyingin-situ capping with active barrier systems (ABS) (Jacobs andForstner, 1999). ABS usually is a reactive geochemical barrierlayer that can actively block the contaminant release from thesediment entering into the overlying water, without the hydrauliccontact between the sediment and the overlying water beingdisturbed. In-situ capping with reactive materials sorbs targetconstituents from the sediment and prevents the release of targetcontaminants into the overlying water more effectively thanpassive material alone. The cost of passive and reactive materialsdepend on the type of material, purity, size, delivery, source,material processing needs, and means of application. McDonoughet al. (2007) reported cap placement costs for large scale site(�1000 acre) at about $25/yd2, excluding the material cost. Thebreak-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 mightnot be representative due to small project footprint compared tolarge scale sites.

3.3. Monitored natural recovery

Monitored natural recovery (MNR) is a remedial technologyfor contaminated sediments that typically uses ongoing,naturally-occurring processes to contain, destroy, or reduce thebioavailability or toxicity of contaminants in sediment. Theseprocesses may include physical, biological, and chemical mechan-isms that act together to reduce risks posed by contaminants. Thekey factors that dictate the selection of MNR as a remedialtechnology are the concentrations of constituents of concernand whether they pose an unacceptable risk, any ongoing


degradation/transformation, or dispersion of contaminant, andthe establishment of a cleanup level that MNR is expected to meetwithin a particular timeframe. The sites, which are ecologicallysensitive in nature and where Hg in low concentrations isstrongly bound to the sediments, are reasonable candidates forusing MNR as the remedial technology. This would involvemonitoring for Hg movement in the aqueous phase. Detailedspectroscopic study of the nature of the Hg in solid phases and theenvironmental conditions conducive to their dissolution is neces-sary to define the necessary safeguards to impose on the site forsuccessfully implementing MNR. As it is generally considered thatthe solid phases holding Hg are themselves sensitive to the stateof oxidation-reduction (Fe(III)-oxide or sulfide phases), institu-tional controls would have to be imposed to safeguard the sitefrom extreme fluxes of ORP. This may involve protecting the sitefrom extremely oxidizing conditions, which may result fromwater being directed away from or drained from the site. Suchconditions may promote the dissolution of sulfide precipitatesand the degradation of organic matter. Conversely, institutionalcontrols may involve protecting the site from extremely reducingconditions, which may result from sustained flooding conditionsthat may cause Fe(III)-oxide phases to dissolve.

The two primary advantages of MNR are its relatively lowimplementation cost and its non-invasive nature that does notneed construction/infrastructure. Although costs associated withcharacterization and/or modeling to evaluate natural recovery canbe extensive, the primary cost associated with implementingMNR is monitoring. The other advantages of MNR over activeremedial methods include no sediment re-suspension, and nochange in benthic conditions (Garbaciak et al., 1998). The keylimitations of MNR may be the potential risk of re-exposure ordispersion of buried Hg if the sediment bed is disturbed by strongnatural or man-made forces and uncertainties in predictingvarious situations, like future sedimentation rates in dynamicenvironments, rate of contaminant flux through stable sediment,or rate of natural recovery. Contaminated systems in naturalattenuation should be regularly monitored to ensure environ-mental safety.

Experiments and field studies demonstrate possible naturalattenuation of Hg contamination by reduction, demethylation,and volatilization. Species-specific enriched stable isotopes havebeen used by Martin-Doimeadios et al. (2004) to study Hgtransformations (methylation, demethylation and volatilization)in estuarine sediments under different environmental conditions(both biotic and abiotic and oxic and anoxic). They reportedthat MeHg levels in sediments are controlled by competingand simultaneous methylation and demethylation reactions.Korthals and Winfrey (1987) used radioactive tracers to under-stand seasonal and spatial variations in Lake Clara, an oligo-trophic, acid-susceptible seepage lake located in Lincoln County,Wisconsin. They demonstrated the usefulness of simultaneousmeasurement of Hg methylation and demethylation. The net rateof MeHg production is significantly affected by the amount ofdemethylation, but also by environmental parameters such astemperature and anoxic conditions. The measured MeHg concen-trations reflect the resultant of site-specific complex processes.Hintelmann et al. (2000) calculated the half-life of MeHg in lakesediment and the reported value is 1.7 days. Marvin-DiPasqualeet al. (2003) reported a Hg methylation rate in San Pablo Bay(northern San Francisco Bay) from below the surface 0 cm to 4 cmhorizon sediment profile as 0.2–1.1 ng g�1 wet sediment day�1.Since the site-specific characteristics are likely to dictate the fateof Hg in the natural environment, significant losses of inorganicHg from surface water occurs through photoreduction and micro-bial reduction. Photoreduction is generally thought to contributeto rapid recycling of Hg at the air–water interface and not

necessarily have a significant influence on archived Hg in thesediment compartment of the aquatic systems. At low Hg con-centrations (low picomolar range), photoreduction is more effec-tive than microbial reduction (Amyot et al., 1997). Morel et al.(1998) reported that at high Hg concentrations (over 50 pico-mole), microbial reduction is more effective and in deep anoxicenvironments, certain bacteria in the presence of humic sub-stances reduce Hg(II). Microbial demethylation of MeHg wasobserved in contaminated sediment (Oremland et al., 1995;Marvin-Dipasquale and Oremland, 1998). Sulfate-reducing bac-teria and methanogenic bacteria are probable agents in microbialdemethylation (Oremland et al., 1995; Marvin-Dipasquale andOremland, 1998). Total Hg concentration and organic substancecontent are important factors in microbial demethylation(Marvin-DiPasquale et al., 2000). A demethylation rate rangedfrom 0.02 ng/g to 0.5 ng/g (dry sediment) per day in a field study(Marvin-Dipasquale and Oremland, 1998). Photodegradation ofMeHg can also happen in surface waters. Photodegradation ofMeHg seems to be a first-order reaction with respect to MeHgconcentration and sunlight intensity (Seller et al., 1996). Inaquatic systems, Hg0 volatilization plays an important role inthe natural attenuation of Hg contamination (Amyot et al., 1997).Hg0 is probably the end-product of some reduction processes ofMeHg and Hg(II) (Seller et al., 1996). Due to its high volatility, Hg0

produced by the reduction of MeHg and Hg(II) may evaporate intothe atmosphere. This evaporation is a major natural attenuationof Hg in some aquatic systems. Flushing can contribute to thenatural attenuation of Hg. Bloom et al. (2004) conducted biogeo-chemical assessment of the Venice Lagoon, which is a large(549 km2), shallow (E1.0 m), enclosed embayment located onthe northwestern Adriatic Sea. The lagoon was contaminated withmore than 50,000 kg of Hg from chlor-alkali plants which oper-ated from 1951 to 1986. Reportedly, �2000 kg of Hg per yearcycled through the lagoon, most of it as a result of the resuspen-sion of historically contaminated bottom sediments. These sedi-ments were resuspended by both wave action and anthropogenicactivities, and then transported to the Adriatic Sea by tidalflushing (Bloom et al., 2004).

Garbaciak et al. (1998) reported field experiments performedin the Whatcom Waterway at Bellingham, Washington, usingnatural attenuation of Hg-contaminated aquatic systems. In the1960s, the Hg concentration in the surface sediment was approxi-mately 4.5 mg/kg. After source control and natural attenuation,Hg concentration in the surface sediment was reduced to about0.5 mg/kg. Garbaciak et al. (1998) also defined enhanced naturalattenuation as natural decontamination, accelerated by humaninfluences. Garbaciak et al. (1998) reported the result of enhancednatural attenuation of Hg-contaminated at the Eagle Harbor site,in Washington. A thin, clean sediment cap (6 cm) was placed onthe contaminated sediment to enhance the burial and separationeffects, because the natural sedimentation process was too slow.These authors reported that compared to thick capping, thisenhanced natural attenuation method of thin capping did notchange the benthic environment significantly.

Other natural processes may include phytoremediation (i.e.phytostabilization). Phytostabilization is the use of plant roots toprevent metal movement that occur in the roots or within theroot neighborhood. Researchers demonstrated that the salt marshplant Juncus Maritumus has a high capacity to stabilize mercuryin sediment (Anjum et al., 2011; Marques et al., 2011).Chattopadhyay et al. (2012) investigated the potential for waterhyacinths (Eichhornia crassipes) to assimilate Hg and MeHg intothe plant biomass over a 68 day hydroponic study. Hg and MeHgwere found to concentrate preferentially in the roots of theE. crassipes with little translocation in the shoots or leaves ofthe plant which was consistent with other macrophytes. The use


of plants to stabilize Hg should take into consideration thetoxicity 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 immobiliza-tion techniques to detoxify Hg in soil and/or sediment (Wanget al., 2012a).

Due to the strong persistence of Hg in the environment, it maytake a long time for heavily contaminated aquatic systems to fullyrecover through natural attenuation processes.

4. Partitioning of mercury on sediments and other substratesand application of models

Partitioning most likely plays a dominant role in the distribu-tion of Hg species between the particulate, colloidal and dissolvedphase controlling the toxicity, fate, transport and bio-uptake of Hg(Stumm and Morgan, 1995). Inorganic Hg that is transported fromsoils and sediments to lakes is predominantly bound to dissolvedor suspended organic matter (Mierle and Ingram, 1991;Hintelmann and Harris, 2004). The sorption capacity of Hg ondifferent natural sorbents (illite, kaolinite, silica, and calcite) hasbeen extensively studied under various test conditions (Gagnonand Fisher, 1997; De Diego et al., 2001). A list of potentialsorbents of Hg is summarized in Table 3. Other studies havefocused on determining partitioning constants of Hg(II) betweenwater and sediments using natural particles (Stordal et al., 1996;Turner et al., 2001) or directly from field studies (Leermakerset al., 1995). Addition of Fe(II) ions in the presence of phlogo-pite (yellow to dark brown mica, general chemical formulaKMg3[Si3Al]O10[F,OH]2) particles can enhance the reduction ofHg(II) (Charlet et al., 2002). The distribution coefficients Kdobtained from sediment samples (log Kd ranges from 4.5 to 6)greatly differ from one substrate to another, most likely due to thenature 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 forHg(II) and MeHg are comparable, but usually slightly lower forMeHg. Generally, the presence of organic matter enhances thesorption of Hg(II) to mineral surfaces (Gagnon and Fisher, 1997;Turner et al., 2001). Some studies investigated the partitioningbetween water and living biota, such as freshwater alga (Mileset 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 isscant. Often, fast sorption of Hg onto particles is observed withstrong binding of Hg to the particles (Gagnon and Fisher, 1997; LeRoux et al., 2001). It is speculated that Hg is initially sorbed andsubsequently migrates into the soil lattice or is covered by organic

Table 3Hg sorption capacities by selected sorbents.

Sorbent

Montmorillonite

Ferrihydrite

Goethite (a-FeOOH)

gamma-alumina (g-Al2O3)

Bayerite (b-Al(OH)3)

Natural zeolites (clinoptilolite)

Activated carbon

Furfural-based carbon

Bauxite

Yellow tuff (soft porous rock usually formed by compaction and cementation of volc

Pozzolana (a type of slag that may be either natural – i.e., volcanic – or artificial, from

Estuarine sediments

biofilms (Mikac et al., 1999). This theory is supported by otherstudies, where newly added Hg was shown to be more availablefor methylation than ambient Hg (Hintelmann and Harris, 2004).Hintelmann and Harris (2004) concluded that the added materialis much more available for species transformation reactions whileambient Hg may be more strongly bound to particles or evenincorporated into the solid matrix and not freely available forligand exchange reactions. It is postulated that strong bindingsites are occupied and saturated first. New Hg species (Hg(II) andCH3Hgþ) entering the system will then initially associate withweaker sites and the time needed for the new Hg to find its highaffinity sites and to equilibrate with the already present Hg (andother metal ions) is uncertain. Other studies have developed thecolloidal pumping model (Stordal et al., 1996; Babiarz et al., 2001)postulating that Hg is initially complexed by colloids (with-ino24 h), which subsequently sorb or coagulate onto particles(within days). Nevertheless, sorbed Hg will be in a dynamicequilibrium with dissolved Hg, the equilibrium being shifted fartoward the solid phase (characterized by very large Kd values).Hintelmann and Harris (2004) conducted studies using stableisotopes of Hg (200HgCl2 and Me199HgCl) and suspension offreshwater sediments to determine the kinetics of Hg and MeHgsorption onto sediment particles and the subsequent rate ofdesorption. These results indicated that equilibrium for sorptionof Hg(II) and MeHg is reached between 1 h and 1 day. The initialdesorption is apparently instantaneous (equilibration takes lessthan 30 min) with no further desorption measurable in thefollowing two days. This study concluded that Hg partitioningbetween water and sediments is dependent on the solid phaseconcentration implying that a fraction of Hg(II) binds strongly toparticles. Strong sorption sites become saturated with increasinglevels 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 ratioof the concentration sorbed to suspended sediment in the watercolumn to the dissolved phase water concentration at equilibrium(Lyon et al., 1997). The total benthic sediment concentration iscomposed of dissolved chemical plus chemical sorbed to thebenthic sediment. In the literature, there are limited measureddata under realistic conditions available. For divalent Hg (Hg[II]),Moore and Rarnamodomy (1984) reported Kd as a range of1380 L/kg to 188,000 L/kg. Glass et al. (1990) reported a value of118,000 L/kg, and Robinson and Shuman (1989) reported a rangeof 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 magni-tude, the log Kd for seston (suspended matter) was in the range of5.5 to 6.0, which corresponds to a range from 316,000 L/kg to1,000,000 L/kg. Babiarz et al. (2001) results show that log Kd forHg ranged from 3.9 to 6.4 with a median of 5.0 and for MeHg

Sorption capacity Reference

296 to 346 mmol/kg Cruz-Guzman et al. (2003)

501 to 577 mmol/kg Cruz-Guzman et al. (2003)

0.39 to 0.42 mmol/m2 Kim et al. (2004a)



1.21 meq/g Chojnacki et al. (2004)

65 mg/g Babel and Kurniawan (2003)

132 to 174 mg/g Budinova et al. (2003)

0.8 to 5.29 L/kg Gavaskar and Chattopadhyay

(2008)

anic ash or dust) 0.18 mg/g at 3000 mg/L Di Natale et al. (2006)

a blast furnace) 0.8 mg/g at 1000 mg/L Di Natale et al. (2006)

log Kd¼4.3 to 6.00 mL/g Turner et al. (2001)


ranged from 3.7 to 6.3, again with a median of 5.0. Lyon et al.(1997) calculated the benthic sediment partition coefficients forHg(II) and MeHg based on the data available in the literature (seeTable 4). It should be noted that partitioning coefficients for Hgspecies 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 significantcontribution to the MeHg stored in water bodies. MeHg concen-trations in fish in Swedish lakes were explained in terms of thefluxes of MeHg into the water bodies from the measured directdeposition rates and runoff/erosion loads from the watershed.However, for lakes with minimal Hg input from the watershed, itsuggested that bioavailable MeHg was created within the lakeitself. Lyon et al. (1997) concluded that for lakes with appreciableinput from the watershed, MeHg in the water body could be dueto a combination of in-lake net methylation and input fromdeposition and/or runoff/erosion.

Many processes influence the fate of contaminants in bottomsediments. Contaminants can be transported into the overlying watercolumn by advective and diffusive mechanisms. Mixing and rework-ing of the upper layer of contaminated sediment by benthicorganisms continually exposes contaminated sediment to thesediment–water interface where it can be released to the watercolumn (Johnson et al., 2010). Bioaccumulation of contaminants bybenthic organisms in direct contact with contaminated sedimentsmay result in movement of contaminants into the food chain.Sediment resuspension, caused by natural and man-made erosiveforces, can greatly increase the exposure of contaminants to the watercolumn and result in the transportation of large quantities ofsediment contaminants downstream (Brannon et al., 1987).

Partitioning of Hg on various substrates including cap materi-als can limit bioavailability through three primary functions: (a)physical isolation of the contaminated sediment from the benthicenvironment, (b) stabilization of contaminated sediments, pre-venting re-suspension and transport to other sites, and (c)reduction of the flux of dissolved contaminants into the watercolumn. Johnson et al. (2010) provided a detailed evaluation ofchemical flux through a cap to assess the effectiveness ofchemical containment. One dimensional advective-diffusivemodel can be applied once cap design objectives with respect toflux are determined, a specific capping material has been selectedand characterized and a minimum cap thickness has beendetermined based on components for isolation, bioturbation,erosion, consolidation, and operational considerations.

Mathematical models can be used to help understand theimportant processes and interactions that affect water quality(Veiga and Meech, 1995). These models can be used in makingdecisions regarding pollution control strategies by evaluatingtheir effectiveness on water quality improvement and performingcost-benefit analysis. Models are extensively used by water resource

Table 4Concentrations of Hg and partitioning coefficients.

Description Total Hg (HgT)concentration in aquaticsediment (ng/g/dry/wt)

Estimconc

Min. Max. Min

80 study lakes, MN (Sorensen et al., 1990) 34 753 0.7

25 study lakes in Sweden (Meili, 1991) 150 460 1.6

Little Rock Lake, WI (Wiener et al., 1990) 10 170 0.3

Savannah river site, Aiken, SC (Kaplan et al., 2002) 20 9450 NA

Fox river (Hurley et al., 1998a) 970 7400 1.8 (

1.0 (

NA¼not available.

planners, water quality managers, remedial project managers, envir-onmental engineers and scientists to evaluate effectiveness of variouscontrol strategies. The success in utilization of models in variousenvironmental applications has resulted in wide acceptance ofmodels as an objective evaluation tool. However, some models areoversimplification of a complex problem. Models are only approx-imate representations of the complex natural processes and due totime and budget constraints involve many assumptions made by themodel developer and user. Certain simplifications considered for oneapplication might not be valid for other applications due to theuniqueness of a problem and counter-intuitive results may beproduced (AWWA, 2001).

Based on functionality, suspended solids and sediments (SSAS),and nutrient water quality models can be broadly categorized intothree groups: (a) loading models, which simulate field scale hydro-logic processes and determine the generation and transport of SSASand 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, circula-tion, and stratification processes), and water quality models tosimulate the movement of SSAS in the water column and determinethe fate and transport of contaminants, nutrients; and (c) eutrop-hication/ecological models, which relate to biomass production,sediment flux, growth in the water body to contaminant and/ornutrient loading, and photosynthesis.

Utilities of mathematical models are: (a) constrain, synthesize,and interpret data, (b) quantify effects of different transportprocesses, (c) make quantitative predictions, and (d) developinsights about processes that affect sediment stability. However,the limitations of the mathematical models are: (a) data collec-tion to support model development and calibration, and (b) levelof uncertainty in results may be unacceptable to stakeholders anddecision-makers. Uncertainty bounds on predictions are the keyissues regarding model reliability and utility at a particular site.However, sediment stability studies conducted at a variety ofsites demonstrate that useful models can be developed providedthat sufficient site-specific data are available, and an experiencedmodeling team conducts the study (Ziegler, 2002).

Simulation of Hg transport and transformation in aquaticsystems is complex, involving hydrodynamic and sediment pro-cesses and Hg transport and transformation processes. Consider-able site-specific data are needed to calibrate and validate Hgtransport and transformation models. Based on the type ofaquatic systems, Wang et al. (2004) conducted a literature reviewon Hg transport and transformation models. Models were cate-gorized in three types of systems: (1) river, (2) lake, and(3) coastal. Limited numbers of Hg transport and transformationmodels are available in the literature. Only a few models link themodeling tool with contamination remediation and predict theremedial results in benthic sediment and overlying water.

ated Hg(II) surface waterentration (ng/L)

Calculated benthicsediment kd for Hg(II)(L/kg)

Calculated benthicsediment kd for MeHg(L/kg)

. Max. Min. Max. Min. Max.

5.8 5,700 990,000 650 110,000

6.5 23,000 290,000 2,600 32,000

0.6 16,000 560,000 1,800 63,000

2000 4,704 5,725 NA NA

unfiltered)

filtered)

182 (unfiltered)

1.6 (filtered)

275,422 912,010 43,651 151,356


The speciation of Hg is an important consideration in anyrepresentation of Hg fate and transport. Depending upon itsspeciation, Hg may be either available or unavailable for biogeo-chemical reactions. Based upon a series of chemical procedures,four predominant species of Hg within an aquatic system havebeen identified into which the aquatic Hg pool may reasonably besubdivided for modeling: Hg[II], MeHg, Hg0, and Inert Hg.

Feyte et al. (2012) showed applying thermodynamic and kineticmodeling to field measurements of total Hg and MeHg and ancillaryparameters in sediments and porewaters can help understand theMeHg cycling and dynamics in sediments. Another researcher (Bale,2000) described rates of transformation among four fundamental Hgspecies to calculate the partitioning of Hg species between thedissolved and particulate-bound phases. The various physical andbiochemical rate processes included for the Hg cycle were: atmo-spheric deposition, diffusive exchange with the atmosphere, biogeo-chemical transformation of Hg species, bio-uptake by aquatic speciesin 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 andtransport developed by Bale (2000) was designed for incorporationinto an advection–diffusion equation, and may be applied to aquaticsystems with complex morphology and hydrodynamics. Bale (2000)demonstrated that this model, calibrated at three sites and validatedat four sites within Clear Lake, California, can reasonably simulatetotal Hg and MeHg as functions of sediment Hg.

The major uncertainty in Hg fate and transport modelingat this time is really predicting: (a) bioavailability of Hg(II) tomethylating microbes, which appears to be mainly a function ofHg-sulfide speciation and Hg-DOC complexes; and (b) identifica-tion and quantification of factors affecting activity of methylatingmicrobes (temperature, organic carbon, sulfate, etc.). Robustgoverning equations that describe these controls are not yetavailable. Thermodynamic speciation models have some valuebut do not capture some of the biologically mediated controls onHg bioavailability and uptake.

Kim et al. (2004c) conducted a simulation of the fate of Hg inaquatic systems by modifying WASPs as part of a remedial investiga-tion of Onondaga Lake (New York). Remediation strategies included:dredging, capping and natural attenuation. Their model predictionsfor the water column generally agreed with the measured valuesreported in the literature for Onondaga Lake. The authors estimatedthe remobilization of sediment based on cutterhead suction dredgingprocesses with a rate of 15,000 m3 of sediment per hour andsediments from the sediment–water interface up to 20 cm deep wereremoved during 20 working days with an estimated removedsediment volume of 2.4�106 m3.

Sensitivity analyses of the model were conducted for deter-mining the impact of transport mechanisms and speciationmechanisms. The simulation results concluded that advection,sorption and settling were important mechanisms of Hg transportin the water column. In the benthic sediment, settling of Hg fromthe 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 ofHg speciation. During remedial design, care should be taken bycollecting performance evaluation parameters as these simula-tions might have uncertainties in the prediction of Hg cyclingconsidering the complexity of Hg speciation and transport.

5. Conclusions

Though considerable progress has been made worldwide toremediate mercury contaminated sediments; significant challenges

still remain. The fate and transport of Hg in the environment isgreatly influenced by the speciation of Hg, its biogeochemicalinteractions with surrounding species, concentrations and state ofspecies that can be bioavailable and other site-specific environmentalconditions. Advances in sample collection and analytical techniques,scientific understanding and engineering approaches with computermodeling provide more cost effective and smarter methods of Hg-contaminated cleanup problems. The remedial decision should befocused on multiple site-specific parameters instead of total concen-tration of contaminant of concern. Innovative sustainable technolo-gies (phytoremediation, thin layer reactive cap, application of naturalmaterials), conventional remediation technologies (dredging, capping,and MNA), novel analytical techniques (biomarker fingerprinting,radioisotope labeling, genetic profiling), and modeling tools areavailable to analyze long term fate and transport of Hg and perfor-mance monitoring of cleanup technology. Significant challenges stillexist in effectively combining technical knowledge and assessmentand management frameworks. The technical complexity associatedwith Hg-contaminated sediment remediation requires that we seekand develop meaningful and reasonable understanding of the site-specific physical and chemical characteristics with fate and transportof Hg at each site. Source control, contaminated sediment remedia-tion, long-term performance monitoring, or their combinations areavailable options for cleaning up Hg—contaminated sites.

Acknowledgments

We would like to thank the anonymous reviewers whoreviewed the manuscript and made suggestions for its improve-ment. The U.S. Environmental Protection Agency through itsOffice of Research and Development performed the researchdescribed here. This research has not been subjected to Agencyreview and therefore, does not necessarily reflect the views of theAgency. Mention of trade names and products should not beinterpreted as conveying official EPA approval, endorsement, orrecommendation.

References

Alberts, J.J., Schindler, J.E., Miller, R.W., Nutter, D.E., 1974. Elemental mercuryevolution mediated by humic acid. Science 184, 895–897.

Alcoa, Treatability Dredge Study for the Alcoa (Point Comfort)/Lavaca Bay Super-fund Site. Draft. 2000.

Amirbahman, A., Reid, A.L., Haines, T.A., Kahl, J.S., Arnold, C., 2002. Associationof 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 gaseousmercury 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 accumula-tion and partitioning in halimione portulacoides and Juncus maritimus col-lected 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 ofsediment and porewater after one year of subaqueous capping of contami-nated sediments in Hamilton Harbour, Canada. Water Sci. Technol. 37,323–329.

Babel, S., Kurniawan, T.A., 2003. Low-cost adsorbents for heavy metals uptakefrom 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 colloidalphase in freshwaters. Environ. Sci. Technol. 35, 4773–4782.

Baeyens, W., Meuleman, C., Muhaya, B., Leermakers, M., 1998. Behaviour andspeciation of mercury in the Scheldt estuary (water, sediments and benthicorganisms). Hydrobiologia 366, 63–79.


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 thewaters 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 inthe state of Rio de Janeiro, Brazil. J. Hazard. Mater. 85, 15–38.

Barkay, T., Gillman, M., Turner, R.R., 1997. Effects of dissolved organic carbonand 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 insediment quality weight-of-evidence (WOE) assessments. Hum. Ecol. RiskAssess. 8, 1517–1547.

Baumgarten, G., R.P.M. Panel Discussion: Sediment Remediation. In: Presented atthe Technical Support Project Meeting. 2001.

Beckvar, N., Field, J., Salazar, S., Hoff, R., Contaminants in Aquatic Habitats atHazardous Waste Sites: Mercury. NOAA Technical Memorandum NOS ORCA100. Seattle: Hazardous Materials Response and Assessment Division, NationalOceanic 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-Hol-land Biomedical Press, pp. 175–178.

Benoit, J.M., Gilmour, C.C., Mason, R.P., Heyes, A., 1999a. Sulfide controls onmercury speciation and bioavailability to methylating bacteria in sedimentpore waters (33, pg 951, 1999). Environ. Sci. Technol. 33, 1780.

Benoit, J.M., Mason, R.P., Gilmour, C.C., 1999b. Estimation of mercury-sulfidespeciation in sediment pore waters using octanol-water partitioning andimplications for availability to methylating bacteria. Environ. Toxicol. Chem.18, 2138–2141.

Bermond, A., Ghestem, J.-P., Yousfi, I., 1998. Kinetic approach to the chemicalspeciation 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 AssessingSediment 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 dredgingoperations for sediment remediation. J. Waterw. Port Coast. Ocean Eng. -ASCE127, 299–307.

Bloom, N.S., Crecelius, E.A., 1983. Determination of mercury in seawater at sub-nanogram per liter levels. Mar. Chem. 14, 49–59.

Bloom, N.S., Effler, S.W., 1990. Seasonal variability in the mercury speciation ofOnondaga 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 ofmethyl mercury as a result of marine sediment dredging activities. Sci. TotalEnviron. 237–238, 379–385.

Bloom, N.S., Moretto, L.M., Scopece, P., Ugo, P., 2004. Seasonal cycling of mercuryand 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 themethylmercury 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 insurface 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. Environmentalfactors 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 contam-ination and bioaccumulation in aquatic ecosystems. Rev. Environ. Contam.Toxicol. 164, 69–92.

Brannon, J.M., Hoeppel, R.E., Gunnison, D., 1987. Capping contaminated dredgedmaterial. 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 furfuraladsorbent. Ind. Eng. Chem. Res. 42, 2223–2229.

Buffle, J., 1988. Complexation Reactions in Aquatic Systems: an AnalyticalApproach, England. Ellis Horwood, Chichester, West Sussex.

Callister, S.M., Winfrey, M.R., 1986. Microbial methylation of mercury in upperwisconsin 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 qualityguidelines for the protection of aquatic life: guidance on the site-specific

application of water quality guidelines in Canada: procedures for derivingnumerical water quality objectives. Canadian Council of Ministers of theEnvironment, 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 mercuryby 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. KluwerAcademic Publishers, Dordrecht; Boston.

Chattopadhyay, S., Fimmen, R.L., Yates, B.J., Lal, V., Randall, P., 2012. Phytoreme-diation of mercury- and methyl mercury-contaminated sediments by waterhyacinth (Eichhornia Crassipes). Int. J. Phytorem. 14, 142–161.

Cheam, V., Gamble, D.S., 1974. Metal-fulvic acid chelation equilibrium in aqueousnano3 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 characteriza-tion of the structural and functional properties of natural organic matterfractions. Chemosphere 48, 59–68.

Choe, K.Y., Gill, G.A., Lehman, R., 2003. Distribution of particulate, colloidal, anddissolved 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 ofnatural 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 methylationin periphyton of the Florida Everglades. Limnol. Oceanogr. 44, 1815–1825.

Coates, J.A., Delfino, J.J., 1993. Sources of uncertainty in the application of theequilibrium 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 andstability 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 artisanalgold mining in Antioquia, Colombia: the world’s highest per capita mercurypollution. Sci. Total Environ. 410, 154–160.

Cormack, R., 2001. Sediment Quality Guideline Options for the State of Alaska. AlaskaDepartment of Environmental Conservation and OASIS/Bristol Joint Venture.

Coston, J.A., Fuller, C.C., Davis, J.A., 1995. Pb2þ and Zn2þ adsorption by a naturalaluminum-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 insediments 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 soilcolloids. 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 suspendedkaolin: 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 andrestoration 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 forremediation of contaminated sediment. In: Lehr, J., et al. (Eds.), Handbook ofComplex 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: utiliza-tion 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 weakbinding 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 ofmercury 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 fromcontaminated 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 Remedia-tion. Springer.

Environment Canada, Criteria for the Assessment of Sediment Quality in Quebecand Application Frameworks: Prevention, Dredging and Remediation. 2007.

Falter, R., 1999. Experimental study on the unintentional abiotic methylation ofinorganic mercury during analysis: Part 1: Localisation of the compoundseffecting the abiotic mercury methylation. Chemosphere 39, 1051–1073.


Feyte, S., Gobeil, C., Tessier, A., Cossa, D., 2012. Mercury dynamics in lakesediments. 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 matterin lake sediments. Appl. Geochem. 25, 984–995.

Gagnon, C., Fisher, N.S., 1997. Bioavailability of sediment-bound methyl andinorganic mercury to a marine bivalve. Environ. Sci. Technol. 31, 993–998.

Gao, Y., Kan, A.T., Tomson, M.B., 2003. Critical evaluation of desorption phenomenaof heavy metals from natural sediments. Environ. Sci. Technol. 37, 5566–5573.

Garbaciak, S., Spadaro, P., Thornburg, T., Fox, R., 1998. Sequential risk mitigationand the role of natural recovery in contaminated sediment projects. Water Sci.Technol. 37, 331–336.

Gavaskar, A., Chattopadhyay, S., Treatment of Environmental Pollutants withMineral Ores. US Patent 7,396,470. 2008.

Gibson, B.D., Ptacek, C.J., Lindsay, M.B.J., Blowes, D.W., 2011. Examining mechan-isms 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 systemsin California and other areas. Environ. Sci. Technol. 24, 1392–1400.

Gill, G.A., Fitzgerald, W.F., 1987. Picomolar mercury measurements in seawaterand other materials using stannous chloride reduction and 2-stage goldamalgamation with gas-phase detection. Mar. Chem. 20, 227–243.

Gilmour, C.C., Henry, E.A., 1991. Mercury methylation in aquatic systems affectedby acid deposition. Environ. Pollut. 71, 131–169.

Gilmour, C.C., Henry, E.A., Mitchell, R., 1992. Sulfate stimulation of mercurymethylation 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 identifica-tion 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 electrolysisindustry. In: Wagner-Dobler, I. (Ed.), Bioremediation of Mercury: CurrentResearch 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 Escherichiacoli under Anaerobic and Aerobic conditions. Limnol. Oceanogr. 47, 967–975.

Graham, A.M., Aiken, G.R., Gilmour, C.C., 2012. Dissolved organic matter enhancesmicrobial 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 reductionand 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 anddesorption 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 associatedwith 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 organicmatter 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 dissolvedorganic 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 – BioinorganicMedicine. Marcel Dekker. Basel, Switzerland, pp. 749–754.

Hamblin, P.F., Zhu, D.Z., Chiocchio, F., He, C., Charlton, M.N., 2000. Monitoringsuspended sediment plumes by optical and acoustical methods with applica-tion to sand capping. Can. J. Civ. Eng. 27, 125–137.

Hammerschmidt, C.R., Fitzgerald, W.F., 2004. Geochemical controls on the produc-tion 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 ofcontaminated sediments in Oslo Harbour, Norway. Water Sci. Technol. 37,299–305.

Hayes, D.F., Crockett, T.R., Ward, T.J., Averett, D., 2000. Sediment resuspensionduring 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 ofmercury 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 fromsurface waters following coagulation of dissolved organic matter with metal-based salts. Sci. Total Environ. 409, 631–637.

Hessen, D.O., Andersen, T., Lyche, A., 1990. Carbon metabolism in a HumicLake—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) toreduced 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 methylmercuryassociated 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 profilesand fluxes of mercury and methylmercury in a small seepage lake in northernMinnesota. 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 todetermine the adsorption and desorption dynamics of Hg(II) and MeHg tosediments. Mar. Chem. 90, 165–173.

Hintelmann, H., Keppel-Jones, K., Evans, R.D., 2000. Constants of mercury methy-lation and demethylation rates in sediments and comparison of tracer andambient mercury availability. Environ. Toxicol. Chem. 19, 2204–2211.

Hinton, J.J., Veiga, M.M., 2009. Using earthworms to assess Hg distribution andbioavailability in gold mining soils. Soil Sediment Contam. 18, 512–524.

Horvat, M., Kotnik, J., Logar, M., Fajon, V., Zvonaric, T., Pirrone, N., 2003. Speciationof 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 andtransport 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 MercuryAcross 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 north-ern Florida Everglades. Biogeochemistry 40, 293–310.

Hurley, J.P., Watras, C.J., Bloom, N.S., 1991. Mercury cycling in a northernWisconsin 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 ofseason and water parameters. Sci. Total Environ. 261, 9–20.

Ilyushchenko, M., Panichkin, V., Randall, P., Kamberov, R., 2012. Former chlor-alkali 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 contaminatedsediments with active barrier systems (ABS) using natural and modifiedzeolites. Water Res. 33, 2083–2087.

Johnson, N.W., Reible, D.D., Katz, L.E., 2010. Biogeochemical changes and mercurymethylation 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 thiosul-fate. Limnol. Oceanogr. 35, 1329–1342.

Kannan, K., Falandysz, J., 1998. Speciation and concentrations of mercury in certaincoastal marine sediments. Water Air Soil Pollut. 103, 129–136.

Kaplan, D.I., Knox, A.S., Myers, J., 2002. Mercury geochemistry in wetland and itsimplications for in situ remediation. J. Environ. Eng. -ASCE 128, 723–732.

Kasprzyk-Hordern, B., 2004. Chemistry of alumina, reactions in aqueous solutionand 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 anaquatic 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 concentrationa good predictor of methyl mercury concentration in aquatic systems. WaterAir Soil Pollut. 80, 715–724.

Kidd, K., 2012. Bioaccumulation and biomagnification of mercury through foodwebs. In: Liu, G., et al. (Eds.), Environmental Chemistry and Toxicology ofMercury. John Wiley & Sons, Hoboken, NJ.

Kim, C.S., Rytuba, J., Brown, G.E., 2004a. EXAFS study of mercury(II) sorption toFe- and Al-(hydr)oxides—II. Effects of chloride and sulfate. J. Colloid InterfaceSci. 270, 9–20.

Kim, C.S., Rytuba, J.J., Brown, G.E., 2004b. EXAFS study of mercury(II) sorption toFe- 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 modelapproach for evaluating effects of remedial actions on mercury speciation andtransport in a lake system. Sci. Total Environ. 327, 1–15.

Korthals, E.T., Winfrey, M.R., 1987. Seasonal and spatial variations in mercurymethylation 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 insurface 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 aquaticmercury cycling. Water Resour. Res. 28, 3119–3128.

Kraepiel, A.M.L., Keller, K., Chin, H.B., Malcolm, E.G., Morel, F.M.M., 2003. Sourcesand 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 environmentalmercury discharge at a four-year-old artisanal gold mining area on LombokIsland, Indonesia. J. Environ. Monit. 14, 2598–2607.


Kudo, A., Fujikawa, Y., Mitui, M., Sugahara, M., Tao, G., Zheng, J., Sasaki, T.,Miyahara, S., Muramatsu, T., 2000. History of mercury migration fromMinamata 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 dissolvedorganic 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 ofmercury 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 inAquatic 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 thescheldt estuary. Water Air Soil Pollut. 80, 641–652.

Leppard, G., Burnison, B., 1983. Bioavailability: trace element associations withcolloids and an emerging interest in colloidal organic fibrils. In: Leppard, G.(Ed.), Trace Element Speciation in Surface Waters and Its Ecological Implica-tions. Plenum Press, New York, pp. 105–122.

Liao, L., Selim, H.M., DeLaune, R.D., 2009. Mercury adsorption–desorption andtransport 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 metal-contaminated marine sediment under conditions of submarine GroundwaterDischarge. Environ. Sci. Technol. 35, 2334–2340.

Liu, G., Li, Y., Cai, Y., 2012a. Adsorption of Mercury on Solids in the AquaticEnvironment. In: Liu, G., et al. (Eds.), Environmental Chemistry and Toxicologyof Mercury. John Wiley & Sons, Hoboken, NJ.

Liu, J., Feng, X., Zhu, W., Zhang, X., Yin, R., 2012b. Spatial distribution andspeciation 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 marinesediments 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 thetoxicity of sediment-associated trace metals with simultaneously extractedtrace metal: acid-volatile sulfide concentrations and dry weight-normalizedconcentrations: a critical comparison. Environ. Toxicol. Chem. 17, 972–974.

Long, E.R., Macdonald, D.D., Smith, S.L., Calder, F.D., 1995. Incidence of adversebiological effects within ranges of chemical concentrations in marine andestuarine sediments. Environ. Manage. 19, 81–97.

Long, E.R., Morgan, L.G., 1990. The Potential for Biological Effects of Sediments-Sorbed Contaminants Tested in the National Status and Trends Program, NOAATechnical Memorandum NOS OMA 52. National Oceanic and AtmosphericAdministration, Seattle, WA.

Loux, N.T., 1998. An assessment of mercury-species-dependent binding withnatural organic carbon. Chem. Speciation Bioavailability 10, 127–136.

Lowry, G.V., Shaw, S., Kim, C.S., Rytuba, J.J., Brown, G.E., 2004. Macroscopic andmicroscopic observations of particle-facilitated mercury transport from newidria 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 andbenthic sediment partition coefficients for mercury. Chemosphere 35,791–808.

MacDonald, D.D., 1994. Approach to the assessment of sediment quality in Floridacoastal waters. Prepared for the Florida Department of Environmental Protec-tion, vol. I. MacDonald Environmental Sciences, Ltd., Ladysmith, BC.

MacDonald, D.D., Ingersoll, C.G., Berger, T.A., 2000. Development and evaluation ofconsensus-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 andsequestration in salt marshes sediments: an ecosystem service provided byJuncus 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 andvolatilization pathways in estuarine sediment slurries using species-specificenriched 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 cyclingof 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 degrada-tion in Florida Everglades peat sediment. Environ. Sci. Technol. 32, 2556–2563.

Mason, R., Bloom, N., Cappellino, S., Gill, G., Benoit, J., Dobbs, C., 1998. Investigationof 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 MinamataBay, Japan. Water Air Soil Pollut. 218, 399–412.

McCready, S., Birch, G.F., Long, E.R., 2006a. Metallic and organic contaminants insediments of Sydney Harbour, Australia and vicinity—a chemical dataset forevaluating sediment quality guidelines. Environ. Int. 32, 455–465.

McCready, S., Birch, G.F., Long, E.R., Spyrakis, G., Greely, C.R., 2006b. An evaluationof 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. Predictiveabilities 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 inthe Anacostia River. Soil Sediment Contam. 16, 313–322.

Massachusetts Department of Environmental Protection (MDEP), Freshwatersediment screening Benchmarks for use under the Massachusetts ContingencyPlan. Massachusetts Department of Environmental Protection,TechnicalUpdate—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 goldmining activities in darkwater ecosystems. Ambio 27, 92–98.

Meili, M., 1991. The coupling of mercury and organic matter in the biogeochemicalcycle—towards a mechanistic model for the boreal forest zone. Water Air SoilPollut. 56, 333–347.

Merritt, K.A., Amirbahman, A., 2009. Mercury methylation dynamics in estuarineand 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 ofmercury from Watersheds. Water Air Soil Pollut. 56, 349–357.

Mikac, N., Niessen, S., Ouddane, B., Wartel, M., 1999. Speciation of mercury insediments 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 monomethyl-mercury 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 dissolvedorganic 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 organic-carbon, Ph, and microbial respiration rates on mercury methylation anddemethylation 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 determi-nation 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 inducedconvective transport through capped sediment. J. Hazard. Mater. 85, 53–72.

Moore, J.W., Rarnamodomy, S., 1984. Heavy Metal in Natural Waters—AppliedMonitoring in Impact Assessment. Springer-Verlag, New York.

Morel, F.M.M., Kraepiel, A.M.L., Amyot, M., 1998. The chemical cycle and bioaccu-mulation of mercury. Annu. Rev. Ecol. Syst. 29, 543–566.

Newell, R.C., Hitchcock, D.R., Seiderer, L.J., 1999. Organic enrichment associatedwith outwash from marine aggregates dredging: a probable explanation forsurface sheens and enhanced benthic production in the vicinity of dredgingoperations. Mar. Pollut. Bull. 38, 809–818.

Nichols, M., Diaz, R.J., Schaffner, L.C., 1990. Effects of hopper dredging andsediment 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 Effective-ness, The National Research Council of the National Academies. NationalAcademies Press, Washington, D.C.

O’Connor, T.P., 2004. The sediment quality guideline, ERL, is not a chemicalconcentration 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 anddistribution in coastal wetlands and tidal mudflats: relationships with sulphurspeciation and organic carbon. Water Air Soil Pollut. 220, 313–326.

Oremland, R.S., Miller, L.G., Dowdle, P., Connell, T., Barkay, T., 1995. Methylmer-cury oxidative-degradation potentials in contaminated and pristine sedimentsof the Carson River, Nevada. Appl. Environ. Microbiol. 61, 2745–2753.

Palermo, M.R., 1998. Design considerations for in-situ capping of contaminatedsediments. Water Sci. Technol. 37, 315–321.

Parks, J.W., Lutz, A., Sutton, J.A., 1989. Water column methylmercury in theWabigoon-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 mercuryemissions to the atmosphere from anthropogenic and natural sources. Atmos.Chem. Phys. 10, 5951–5964.

http://www.mass.gov/dep/cleanup/laws/sedscrn.doc

http://www.mass.gov/dep/cleanup/laws/sedscrn.doc


Ramlal, P.S., Rudd, J.W.M., Furutani, A., Xun, L.Y., 1985. The effect of pH on methylmercury 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: MercuryPollution near a Chemical Plant in Northern Kazakhstan. EM Magazine. Air &Waste Management Association, Pittsburgh, PA.

Ravichandran, M., 1999. Interactions between Mercury and Dissolved OrganicMatter in the Florida Everglades. University of Colorado, Boulder, pp. 206.

Ravichandran, M., 2004. Interactions between mercury and dissolved organicmatter—a review. Chemosphere 55, 319–331.

Reis, A.T., Rodrigues, S.M., Araujo, C., Coelho, J.P., Pereira, E., Duarte, A.C., 2009.Mercury contamination in the vicinity of a chlor-alkali plant and potentialrisks to local population. Sci. Total Environ. 407, 2689–2700.

Richman, L.A., Wren, C.D., Stokes, P.M., 1988. Facts and fallacies concerningmercury 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 watersusing 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 byrainbow-trout (Salmo-gairdneri) in relation to oxygen-consumption andmethylmercury 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 environ-mental impact. Sci. Total Environ. 260, 57–71.

Sager, D.R., 2002. Long-term variation in mercury concentrations in estuarineorganisms 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 fromradionuclide and Hg distributions. Environ. Sci. Technol. 33, 378–391.

Sarkar, D., Essington, M.E., Misra, K.C., 2000. Adsorption of mercury(II) bykaolinite. Soil Sci. Soc. Am. J. 64, 1968–1975.

Schindler, D.W., Hesslein, R.H., Wagemann, R., 1980. Effects of acidification onmobilization of heavy-metals and radionuclides from the sediments of a fresh-water lake. Can. J. Fish. Aquat. Sci. 37, 373–377.

Schultz, T., Korhonen, P., Virtanen, M., 1995. Mercury model used for assessmentof 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 carbondynamics during runoff episodes from a northeastern USA watershed. Water AirSoil Pollut. 187, 89–108.

Seller, P., Kelly, C.A., Rudd, J.W.M., MacHutchon, A.R., 1996. Photodegradation ofmethylmercury in lakes. Nature 380, 694–697.

Skyllberg, U., Bloom, P.R., Qian, J., Lin, C.M., Bleam, W.F., 2006. Complexation ofmercury(II) in soil organic matter: EXAFS evidence for linear two-coordinationwith reduced sulfur groups. Environ. Sci. Technol. 40, 4174–4180.

Skyllberg, U., Drott, A., 2010. Competition between disordered iron sulfide andnatural 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 ofmercury(II) to reduced sulfur in soil organic matter along upland-peat soiltransects. J. Environ. Qual. 29, 855–865.

Smith, S.L., MacDonald, D.D., Keenleyside, K.A., Gaudet, C.L., 1996. The develop-ment and implementation of Canadian sediment quality guidelines. In:Munawar, M., Dave, G. (Eds.), Development and Progress in Sediment QualityAssessment: Rationale, Challenges, Techniques, and Strategies. SPB AcademicPublishing, Amsterdam.

Sorensen, J.A., Glass, G.E., Schmidt, K.W., Huber, J.K., Rapp, G.R., 1990. Airbornemercury deposition and watershed characteristics in relation to mercuryconcentrations in water, sediments, plankton, and fish of 80 NorthernMinnesota 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 transfor-mation 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 thecoagulation 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-mixedestuarine sediments. Mar. Chem. 90, 91–105.

Sunderland, E.M., Gobas, F.A.P.C., Branfireun, B.A., Heyes, A., 2006. Environmentalcontrols 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 useand pollution. Ambio 36, 45–61.

Telmer, K.H., Veiga, M.M., 2009. World emissions of mercury from artisanal andsmall 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 cadmiumand 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 cappingcontaminated sediments in-situ.2. Mathematics of diffusion adsorption in thecapping layer. Environ. Sci. Technol. 27, 2412–2419.

Tiffreau, C., Lutzenkirchen, J., Behra, P., 1995. Modeling the adsorption ofmercury(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 ofinorganic mercury in estuaries. Environ. Sci. Technol. 35, 4648–4654.

U.S.EPA, Ambient Water Quality Criteria for Mercury—1984. Office of WaterWashington, DC, 1985.

U.S.EPA, Report of the Sediment Criteria Subcommittee Evaluation of the ApparentEffects Threshold (AET) Approach for Assessing Sediment Quality. U.S. Envir-onmental Protection Agency, Office of the Administrator, Science AdvisoryBoard, Washington, DC. 1989.

U.S.EPA, Mercury Study Report to Congress, Office of Air Quality Planning andStandards and Office of Research and Development, U.S. EnvironmentalProtection Agency, EPA/452/R-96/001A-H. 1997.

U.S.EPA, TRIM: total risk integrated methodology, TRIM.FaTE, technical suportdocument, volume II: description of chemical transport and transformationalgorithms.(EPA-453/R-02-011b)—Appendix A: Derivation of Mercury-Specific Algorithms, North Carolina, 2002.

U.S.EPA, Evaluation Report-EPA Can Better Implement Its Strategy for ManagingContaminated Sediments. Report No. 2006-P-00016,March 15, 2006. In: Officeof the Inspector General (Ed.), 2006.

U.S.EPA, National Listing of Fish Advisories, 2010, EPA-820-F-11-014, November2011. 2011.

U.S.EPA, 2012. Alcoa/Lavaca Bay, Calhoun County, Texas EPA Region 6 Congres-sional 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 onbiogeochemical cycles. Chemosphere 77, 451–458.

Ullrich, S.M., Ilyushchenko, M.A., Kamberov, I.M., Tanton, T.W., 2007. Mercurycontamination in the vicinity of a derelict chlor-alkali plant. Part I: Sedimentand water contamination of Lake Balkyldak and the River Irtysh. Sci. TotalEnviron. 381, 1–16.

Ullrich, S.M., Tanton, T.W., Abdrashitova, S.A., 2001. Mercury in the aquaticenvironment: a review of factors affecting methylation. Crit. Rev. Environ.Sci. Technol. 31, 241–293.

UNEP, Mercury in the Aquatic Environment:Sources, Releases, Transport andMonitoring, published by UNEP’s Division of Technology, Industry andEconomics (DTIE) Chemicals Branch, prepared by The Joint Group of Expertson 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 opera-tions 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 shuttleson the redox (bio) transformation of contaminants: a review. Biotechnol. Adv.27, 256–277.

Veiga, M., Meech, J., 1999. Reduction of mercury emissions from gold miningactivities 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 inthe Amazon. Water Air Soil Pollut. 80, 123–132.

Wallschlager, D., Desai, M.V.M., Spengler, M., Wilken, R.D., 1998a. Mercuryspeciation in floodplain soils and sediments along a contaminated rivertransect. 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 contami-nated floodplain soils and sediments. J. Environ. Qual. (27), 1044–1054.

Wallschlager, D., Desai, M.V.M., Wilken, R.D., 1996. The role of humic substancesin 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. Remediationof 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 andremediation for mercury contamination in aquatic systems—a literaturereview. Environ. Pollut. 131, 323–336.

Wang, S., Zhang, M., Li, B., Xing, D., Wang, X., Wei, C., Jia, Y., 2012b. Comparison ofmercury speciation and distribution in the water column and sedimentsbetween the algal type zone and the macrophytic type zone in a hypereu-trophic lake (Dianchi Lake) in Southwestern China. Sci. Total Environ. 417,204–213.

http://www.epa.gov/region6/6sf/pdffiles/0601752.pdf

http://www.epa.gov/region6/6sf/pdffiles/0601752.pdf


Watras, C.J., Morrison, K.A., Host, J.S., Bloom, N.S., 1995. Concentration of mercuryspecies in relationship to other site-specific factors in the surface waters ofNorthern 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 Assess-ment. 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 andbioavailability 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 indredge material from the lower Savannah River. Arch. Environ. Contam.Toxicol. 38, 128–136.

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) byreduced sulfur in soil humic substances. Environ. Sci. Technol. 33, 257–261.

Xiong, Z., He, F., Zhao, D., Barnett, M.O., 2009. Immobilization of mercury insediment using stabilized iron sulfide nanoparticles. Water Res. 43,5171–5179.

Yang, Y., Liang, L., Wang, D., 2008. Effect of dissolved organic matter on adsorptionand 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. Kineticsof 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 reducednatural organic matter in anoxic environments. Environ. Sci. Technol. 46,292–299.

Zhou, P., Yan, H., Gu, B.H., 2005. Competitive complexation of metal ions withhumic substances. Chemosphere 58, 1327–1337.

Ziegler, C.K., 2002. Evaluating sediment stability at sites with historic contamina-tion. Environ. Manage. 29, 409–427.

http://www.dnr.state.md.us/coastalbays/publications/Chapter5.2.pdf

http://www.dnr.state.md.us/coastalbays/publications/Chapter5.2.pdf

Date post:	19-Dec-2016
Category:	Documents
Upload:	sandip
View:	215 times
Download:	2 times

Mercury contaminated sediment sites—An evaluation of remedial options

Documents