Environment International 74 (2015) 42–53

Contents lists available at ScienceDirect

Environment International

j ourna l homepage: www.e lsev ie r .com/ locate /env int

Review

Sources and remediation techniques for mercury contaminated soil

Jingying Xu a,b,⁎, Andrea Garcia Bravo a, Anders Lagerkvist b, Stefan Bertilsson a,Rolf Sjöblom b, Jurate Kumpiene b

a Department of Ecology and Genetics, Limnology, University of Uppsala, Uppsala 75236, Swedenb Waste Science and Technology, Department of Civil, Environmental and Natural Resources Engineering, Luleå University of Technology, Luleå 97187, Sweden

⁎ Corresponding author at: Department of Ecology and

http://dx.doi.org/10.1016/j.envint.2014.09.0070160-4120/© 2014 Elsevier Ltd. All rights reserved.

a b s t r a c t

a r t i c l e i n f o

Article history:Received 15 October 2013Accepted 16 September 2014Available online xxxx

Keywords:Mercury contaminated soilMobilityRemediationSoil washingStabilisation/solidification

Mercury (Hg) in soils has increased by a factor of 3 to 10 in recent times mainly due to combustion of fossil fuelscombined with long-range atmospheric transport processes. Other sources as chlor-alkali plants, gold miningand cement production can also be significant, at least locally. This paper summarizes the natural and anthropo-genic sources that have contributed to the increase of Hg concentration in soil and reviews major remediationtechniques and their applications to control soil Hg contamination. The focus is on soil washing, stabilisation/solidification, thermal treatment and biological techniques; but also the factors that influence Hg mobilisationin soil and therefore are crucial for evaluating and optimizing remediation techniques are discussed. Further re-search on bioremediation is encouraged and future study should focus on the implementation of different reme-diation techniques under field conditions.

© 2014 Elsevier Ltd. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 432. Origins and transfers of Hg in soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

2.1. Sources of Hg . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 432.1.1. Natural sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 432.1.2. Anthropogenic sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 432.1.3. Re-emissions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

2.2. Factors affecting Hg mobilisation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 442.2.1. Hg carriers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 452.2.3. Redox potential and pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46

3. Technologies for the remediation of Hg contaminated soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 463.1. Soil washing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 463.2. Stabilisation/solidification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 473.3. Thermal treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 473.4. Biological techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48

3.4.1. Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 483.4.2. Bioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 49

4. Technology improvement . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 494.1. Soil washing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 49

4.1.1. Particle size . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 494.1.2. Liberation degree . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 494.1.3. Organic matter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 494.1.4. Extractant . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 49

4.2. Stabilisation/solidification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 494.2.1. Interferences . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 504.2.2. Binder types . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50

Genetics, Limnology, University of Uppsala, Uppsala 75236, Sweden.

43J. Xu et al. / Environment International 74 (2015) 42–53

5. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50

1. Introduction

Mercury (Hg) is a naturally occurring element that is commonly foundin the environment. During the post-industrial era, combustion of fossilfuels combined with long-range, atmospheric transport has increasedthe Hg in soils and sediments by a factor of 3 to 10 times (UNEP, 2013),and the global Hg emission into the atmosphere was reported to amountto 3000 t in 2005 (Branch, 2008).MostHg forms are highly toxic to highlyexposed humans, but even low exposure can seriously and adversely af-fect the central nervous system (Nance et al., 2012). The health risks over-all are greater for foetuses and young children than for adults (Holmes,2009). One process of major concern is the transformation of inorganicHg to methyl-Hg ([CH3Hg]+), a species more prone to bio-accumulatein organisms (USGS, 2000). Foraging behaviour can thus pass onHg to en-tire foodwebs and in this way threaten ecosystem health, and in particu-lar pose a serious threat to species at higher trophic levels (Gabriel, 2004).

To account for this environmental health-hazard, over 130 countriesrecently agreed to the United Nation's Minamata Convention for reduc-ing the emission and use of Hg (UNEP, 2013). The US EnvironmentalProtection Agency has furthermore developed regulations to controlHg emissions to air, water, or from wastes and products under certainFederal environmental statutes, such as the Clean Air Act, Clean WaterAct, and Resource Conservation and Recovery Act (U.S. EPA, 2013). InChina where Hg contamination is a major environmental health prob-lem, the China Council for International Cooperation on Environmentand Development carried out a special policy study of Hgmanagement,offering recommendations for priority actions to reduce Hg release anduse (Annual GeneralMeeting of China Council for International Cooper-ation on Environment and Development, 2011). Overall, there is agrowing awareness of the significance of Hg as a global environmentalcontaminant and the urgent need for remedies tominimise the negativeeffects on human health and ecosystem services.

Hg derived from both natural and anthropogenic sources is widelydistributed around the world (Frohne et al., 2012). It can be rapidlytransported away from a point source and subsequently enter the globalHg cycle, ultimately being wet or dry deposited in either aquatic orterrestrial ecosystems. Notably, Hg ismuchmore persistent in soils com-pared to lakes, oceans and other biomes (Padmavathiamma and Li,2007; Tangahu et al., 2011). In general, natural attenuation can occurin several ways: by means of biodegradation, dispersion, dilution, sorp-tion, volatilisation, reductive decay, chemical or biological stabilisation,and the transformation and destruction of contaminants (UnitedStates Environmental Protection Agency (US EPA), 1997a,b,c). Althoughsoil has a natural capacity to attenuate heavy metals through variousmechanisms, concentrations of heavy metals exceeding the attenuationcapacity will inevitably lead to soil pollution. Accordingly, remediationtechniques are needed to either remove Hg from the soil or to transformit into its most stable and least toxic forms in situ (Cui et al., 2011;Tangahu et al., 2011). Historically, thermal treatment, vitrification, soilwashing, biological techniques (e.g., phytoremediation), stabilisation/solidification and other techniques have been applied to counteract Hgcontamination in soil (Randall and Chattopadhyay, 2004; Richter andFlachberger, 2010; Rodríguez et al., 2012). The success of such actionsvaries, and in this paper we provide an overview of some current reme-dial techniques for treating Hg-contaminated soil as well as a brief ac-count of the principal sources of Hg pollution in this biome. Specialattention is furthermore given to the factors that affect Hg mobility insoil and in thisway influence the efficiency of selected treatments. Strat-egies and means of improving the techniques are also discussed.

2. Origins and transfers of Hg in soil

2.1. Sources of Hg

2.1.1. Natural sourcesThe average background contents ofHg indifferent types of soils from

all over theworld range between 0.58 and 1.8 mg/kg, and theworldwidemean content is estimated at 1.1 mg/kg. Higher Hg concentrations havebeen observed in Histosols and Cambisols (Kabata-Pendias, 2010).

Hg is released from many sources by means of a variety of naturalprocesses. This includes ubiquitous weathering of Hg-containing rocksin the Earth's crust, geothermal activity, or Hg emitted during episodicevents such as volcanic eruptions (AMAP/UNEP, 2013). Current Hgemissions to the atmosphere from natural sources are estimated atabout 80–600 t/year (Mason et al., 2012). Such contributions varyacross time and space depending on a number of factors, including thepresence of volcanic belts, the level of geothermal activity, geologicalformations (such as cinnabar deposits) and the frequency of naturalwildfires (Ferrara, 1998, 2000; Pirrone et al., 2001). After beingtransported some distance within the atmosphere, Hg returns to theearth's surface through wet and dry deposition. In this way, more than90% of the emitted Hg ends up in terrestrial ecosystem, with soilsbeing the largest recipient (Lindqvist et al., 1991).

2.1.2. Anthropogenic sourcesCurrent anthropogenic sources, which include numerous industrial

point sources, are estimated to release about 1960 t of Hg on an annualbasis (AMAP/UNEP, 2013). Figs. 1 and 2 summarize the global anthro-pogenic emissions of Hg to the air by region and sector. Themajor sourceregions are Asia and Africa (ca. 47.5% and 16.8% of the global total,respectively). The main sectors identified are artisanal and small-scale gold mining (ASGM), coal combustion, production of non-ferrousmetals (including copper, lead, zinc, aluminium and large-scale goldproduction), cement production, and disposal of wastes containing Hg(AMAP/UNEP, 2013; Mason et al., 2012). The emissions associatedwith ASGMoperations are significantly higher than previously reported,which is attributed mainly to new information on use of Hg in ASGM incertain regions (AMAP/UNEP, 2013).

There are also other Hg sources that need to be considered, such asdiscarded thermometers, batteries and fluorescent lamps that togetheraccount for as much as 40% of the Hg emissions in North America.Barometers used inweather stations, airports and airfields,wind tunnels,and engine manufacturing, as well as in installations offshore or onships also contribute to Hg release (Hutchison, 2003). In agriculturalsystems, Hg pollution originates from pesticides, fertilizers, sewagesludge and irrigation water (Hseu et al., 2010).

Overall, coal burning continues to increase, especially in Asia, but atthe same time improvements are taking place in other regions of theworld (Table 1). Cement production is another major contributor ofHg and has increased by almost 30% between 2005 and 2009, even ifthere are large differences between regions (USGS, 2012). The emis-sions fromnon-ferrousmetal production (Table 1) indicate apparent in-creasing trend in Asia and Central America.

2.1.3. Re-emissionsIn the context of the global Hg cycle, re-emission is defined as Hg

emissions that are derived from past natural and anthropogenic de-posits. Under conducive conditions, previously deposited Hg from theEarth's surfaces (soil, rocks, snow and ice, and surface waters) can be

0

200

400

600

800

1000

Africa Asia Central and South

America

CIS other European countries

European Union

Middle Eastern States

North America

Oceania Region undefined

Hg

Emis

sion

s (to

ns)

Fig. 1. Global anthropogenic emissions of Hg to the air by region, 2010 (modified from AMAP/UNEP, 2013). CIS: Commonwealth of Independent States, including Azerbaijan, Armenia,Belarus, Georgia, Kazakhstan, Kyrgyzstan, Moldova, Russia, Tajikistan, Turkmenistan, Uzbekistan and Ukraine.

44 J. Xu et al. / Environment International 74 (2015) 42–53

re-emitted to the atmosphere through alternative transport mecha-nism. Estimates of current annual re-emissions of Hg are in the range4000–6300 t/year (AMAP/UNEP, 2013; Mason et al., 2012), and alarge proportion of the re-emitted Hg might ultimately accumulate insurface soils.

2.2. Factors affecting Hg mobilisation

Factors that affect Hgmobility are crucial for evaluating andoptimiz-ing soil-remediation techniques. Hg removal from the soil is facilitated ifHg exists in an easily mobilised chemical form, while Hg stabilisationwithin the soil is favoured if Hg is strongly bound to the soil matrix.The mobility of Hg thus depends on its chemical speciation, which is a

C

Non-ferrous metal production (Al, Cu, Pb, Zn)(primary)

10%

Cement production 9%

Large-scale gold production 5%

Waste from Hg-containing products

5% Othe8%

Fig. 2. Distribution of global anthropogenic Hg emissions to air fro

function of several soil parameters and their interactions (Yin et al.,1997). Hg occurs in soil in various forms: (i) dissolved (free ion or solu-ble complex), (ii) non-specifically adsorbed (binding mainly as a resultof electrostatic forces), (iii) specifically adsorbed (strong binding owingto covalent or coordinative forces), (iv) chelated (bound to organic sub-stances) or (v) precipitated (as sulphide, carbonate, hydroxide, phos-phate, etc. (Schuster, 1991). Hg transformations, e.g., methylation anddemethylation, reduction and oxidation, might meanwhile modify Hgspeciation (Hu et al., 2013a,b) and thus mobilise/stabilise Hg from/insoils. Furthermore, geochemical parameters, such as pH and redox po-tential of the soil, strongly affect Hg mobility by changing its solubilityand the biological process affecting Hg transformations (Robles et al.,2014; Schuster, 1991).

Artisanal and small-scale gold mining 38%

oal combustion 25%

rs

m different sectors, 2010 (modified from AMAP/UNEP, 2013).

Table 1Regional emission (tons) from selected sectors, and changes from 2005 to 2010 (modified from AMAP/UNEP, 2013).

Coal combustion Cement production Non-ferrous metals Waste fromconsumer products

Ferrous metalproduction

Year 2005 2010 2005 2010 2005 2010 2005 2010 2005 2010Africa 40.6 43.1 8.6 11.3 38 32.7 7.3 6.6 0.4 0.4Asia 253.1 297.1 84.5 119.3 97.8 136.8 63.7 53.5 20.1 30.2Central America 10.3 3.5 3.3 3.2 8.3 11.7 4 3.4 0.3 0.3CIS and other European countries 35 26.9 4.7 4.7 42.9 42.2 8.3 7.1 7.8 7.5European Union 57.5 44.1 14.4 13.1 14.4 11.7 6.8 6 3 2.5Middle Eastern States 8 10.5 9.3 13.4 3.8 3.9 5.2 4.5 0.4 0.4North America 50.5 43.4 3.4 2.3 3.5 3.2 7.6 6.2 1.5 1.1Oceania 3.9 3.6 0.7 0.7 18.5 16.7 0.7 0.7 0.04 0.04South America 3.3 2.2 3.8 5.1 39.8 38.7 8.7 7.6 3.2 3Total 462.2 474.3 132.7 173 267.1 297.7 112.1 95.5 36.7 45.5

Table 2Common soil minerals (Kabata-Pendias, 2010; Sparks, 2003; Tan, 2011).

Mineral Mineral type Specific surface area (m2/g)

Montmorillonite Clay minerals 280–800Allophane Clay minerals 100–880Illite Clay minerals 65–100Goethite Oxides 41–81 (amorphous: 305 412)Mn oxide Oxides 32–300Chlorite Clay minerals 25–150

45J. Xu et al. / Environment International 74 (2015) 42–53

Therefore, in the following sections, the effects of (i) different car-riers, (ii) transformations and (iii) pH and redox potential on Hgstabilisation or mobilisation in soils will be described in detail.

2.2.1. Hg carriersUnder naturally occurring conditions Hg is primarily complexed

with Cl−, OH−, S2− and S-containing functional groups of organicligands (Burton, 2006; Keating et al., 1997; Loux, 1998; Schuster, 1991,Skyllberg et al., 2007). It has been shown that organic matter is thedominant factor controlling Hgmobility in acidic soils, whereasmineralcomponents influence Hg solubilitymore extensively in neutral to alka-line soils (Jackson, 1998; Kwaansa-Ansah et al., 2012; Schuster, 1991;Skyllberg, 2008; Yang et al., 2007; Yin et al., 1996).

2.2.1.1. Organic matter. Natural organic matter (NOM) interacts verystrongly with Hg, affecting its speciation, solubility, mobility and toxic-ity. Many organic compounds have high affinity for Hg by means offunctional groups, such as hydroxylic-, carboxylic-, aromatic- and espe-cially the S-containing ligands (Hintelmann, 1997; Karlsson andSkyllberg, 2003). The most stable organic compounds in soil are humicsubstances that contain a large number of functional groups known tointeract with Hg, such as OH, COOH and SH (Kabata-Pendias, 2010;Skyllberg et al., 2007). Skyllberg et al. (2005) showed that methyl-Hg([CH3Hg]+) was dominantly complexed by thiol ( RSH) in soil OMover the entire pH range in natural systems.

The presence of NOM can either increase or decrease Hg mobility insoil depending on whether or not the OM is mobile. Organic matterimmobilised soil particles may provide additional sorption sites, whichmay on the one hand enhance Hg immobilisation. For example, Yanget al. (2007) demonstrated that Hg emissions from the NOM-amendedsoil were apparently low and decreased further as the OM loading in-creased. On the other hand, NOMcan increaseHgmobility in soil throughthe formation of dissolved OM Hg complexes that move with the soilwater (Ravichandran, 2004; Skyllberg et al., 2000). For example, the pres-ence of soluble humic acids noticeably increased Hgmobility by transfer-ring Hg from solid phase to solution (Beone et al., 2009). The forms of Hgcan also influence its complexation to OM, as some studies suggest thatinorganic Hg binds more strongly and to a greater extent to soil NOMthan organomercurials do (Farrah, 1979; Skyllberg et al., 2003).

The stability and persistence of NOMalso influence the release of Hgfrom soils. Natural organic matter can be rapidly degraded by photo-chemical oxidation, or slowly broken down by aerobic or anaerobic ac-tivity of microbial communities and upon further degradation releasestrongly bound Hg. A recent study showed that the complexation of[CH3Hg]+ with thiol groups of well-defined organic molecules mightenhance the photolysis of [CH3Hg]+ (Fernández Gómez et al., 2013).

2.2.1.2. Minerals. Secondary soil minerals have a high surface area (up to800m2/g) and therefore have a high affinity for ions (Table 2). Claymin-erals, amorphous oxides and hydroxides of Fe, Mn and Al, as well as

amorphous iron sulphide (FeS) are significant inorganic sorbents forHg (Liao et al., 2009). Hydroxyl complexes constitute the active speciesfor Hg binding on oxides and hydroxides as a result of the strongHg\OH covalent bond (Schuster, 1991), whereas adsorption to FeS ismainly due to the precipitation of HgS in a reducing soil environment(Kabata-Pendias, 1993).

Hg may also be immobilised in the form of solid solution when Mnand Fe hydroxides/oxides, carbonates and apatites are formed (Olivaet al., 2012; Tiffreau et al., 1995). From an atomic radius point of view,Hg fits reasonably well into the two latter, whereas hydroxides/oxidesof Mn and Fe appear to be able to incorporate Hg complexes such asthosewithOH−or Cl−(Berndt et al., 2005; Serrano et al., 2012). ExtendedX-ray absorption fine structure is a promising tool for elucidating themolecular-level architecture of both surface and internal structuresas well as between short- and long-term binding effects (Kim et al.,2004; Serrano et al., 2012).

2.2.1.3. Chloride ions. Chlorides are regarded as one of the most mobilecomplexing agents for Hg (Kabata-Pendias, 2010). Chloride ions areeven able to outcompete OH− and organic ligands under certain condi-tions (Gabriel, 2004; Payne, 1964; Reimers and Krenkel, 1974; Yin et al.,1996, 1997). The chloride concentration needed for mobilising Hg byformation of HgCl2 was expected to increase with increasing pHowing to the competition with OH− (Schuster, 1991). Increased Cl−

concentration has also been seen to decrease the amount of Hg boundto organic matters (Reddy and Aiken, 2001; Reimers and Krenkel,1974). Reddy and Aiken (2001) conducted an experiment showingthat high Cl− concentrations caused Hg speciation to shift from pre-dominantly Hg-fulvic acid complexes to Hg–Cl complexes. Similar re-sults to the above-mentioned chloride effect on Hg mobilisation havealso been generated in modelling and simulations (Xu et al., 2014).However, some observations disagree, as some experiments haveshown a pH-dependent Hg dissolution regardless of the presence ofchlorides (Xu et al., 2014). One explanation for these contradictoryresults is that Hg–Cl in the latter study was directly bound to the soilOM at very high affinity.

2.2.2. Biological processes2.2.2.1. Methylation and demethylation. Although [CH3Hg]+ consti-

tutes less than 2% of total Hg in soil (Schlüter, 2000), it is the most

46 J. Xu et al. / Environment International 74 (2015) 42–53

toxic form of this element and can cause severe neurological damage tohumans and wildlife (Clarkson, 2003). The main source of humanexposure to [CH3Hg]+ is the consumption of Hg-containing fish andrice, and [CH3Hg]+ furthermore tends to be more biomagnified infood chains than other Hg species (Feng et al., 2007; Selin, 2009;Zhang et al., 2010). Methylation is mainly a biotic process where thetransformation of inorganic Hg (II) is controlled by either sulphate-reducing bacteria (SRB) (Pak and Bartha, 1998; Gilmour et al., 1992,1998; Tjerngren et al., 2012), iron-reducing bacteria (FeRB) (Fleminget al., 2006; Bravo et al., 2014), methanogens (Hamelin et al., 2011), orfirmicutes (Gilmour et al., 2013). Methylated Hg is usually associatedwith anionic ligands, and the order of its affinity for common ligandshas been reported: R S− N SH− N OH− N Cl− (Dyrssen, 1991; Gavis,1972). There are no reports of a direct impact of either redox potential(Eh) or pH on [CH3Hg]+ formations (Frohne et al., 2012; Hintelmann,1995). Instead, indirect effect of Eh and pH on Hg methylation throughtheir influence on the concentrations of dissolved organic carbon andionic strengths is considered important (Frohne et al., 2012;Wallschläger et al., 1998a,b). And most likely the composition of theresident microbial communities would also be important in this regard.

Bacteria are also responsible for the demethylation of [CH3Hg]+,which can decrease the quantity of [CH3Hg]+ available for accumu-lation through the food web (Barkay and Wagner-Dobler, 2005;Marvin-DiPasquale et al., 2000). Bacterially mediated [CH3Hg]+

demethylation occurs via either a reductive process that producesHg0 and CH4, or by an oxidative process that produces Hg2+ and CO2

(Barkay et al., 2003). The reductive process is catalysed by enzymes ofthe mer genetic system in Hg-resistant bacteria, and is often enhancedby increased concentrations of Hg2+. The oxidative process appears tobe a co-metabolism of [CH3Hg]+ and is similar to the use of othersmall organic compounds (Oremland et al., 1991).

2.2.2.2. Oxidation and reduction.Mercury undergoes both reductionand oxidation through biotic process in natural environments,resulting in changes in Hg speciation that in turn affects its biologicaluptake and methylation (Hsu-Kim et al., 2013). Several dissimilatorymetal-reducing bacteria, including Shewanella oneidensis MR-1 andGeobacter metallireducens GS-15, are known to reduce Hg (II) (Lovleyet al., 1993; Wiatrowski et al., 2006). Warner et al. (2003) reported anincrease in Hg(II) reduction under iron reducing conditions, resultingin inhibited Hg methylation. Similarly, reduction of Hg(II) was foundto occur on cells of G. sulfurreducens strain PCA under dark and anaero-bic conditions (Hu et al., 2013a,b), suggesting that coupled Hg(II)–cellinteractions could be important in controlling Hg species transfor-mation and bioavailability. On the other hand, three bacterial strainswithin the Deltaproteobacteria have been compared their ability tooxidise and methylate Hg(0) under dark, anaerobic conditions.D. desulphuricans ND132 can oxidise Hg(0) anaerobically and furtheruse it formethylation (Colombo et al., 2013);whilstD. alaskensisG20 canonly oxidise but not methylate Hg(0); meanwhile, G. sulphurreducensPCA oxidise andmethylates Hg(0)when supplementedwith thiol com-pounds such as cysteine. These findings suggest that methylating andnon-methylating bacteria may work synergistically to enhanceanaerobic Hg(0) oxidation and methylation (Hu et al., 2013a,b).

2.2.3. Redox potential and pHRedox potential might change Hg speciation, and anoxic conditions

may enhance the activity of sulphate-reducing bacteria, which facilitateHg methylation but also produce sulphides and polysulphides. In thepresence of sulphur/polysulphides, mildly reducing conditions cancause Hg to precipitate as sulphide. However, strongly reducing condi-tions may further increase Hg solubility, either by converting the Hg2+

in HgS0 to Hg0 (Benoit et al., 1999) or by the formation of stable and sol-uble HgS22− (Wollast et al., 1975). High solubility of Hg is only seen inwell-oxygenated environments (350 to 400 mV), which is commonsoils (Barrow, 1992; Farrah, 1979; Payne, 1968; Reimers and Krenkel,1974).

The pH has a fluctuant influence on Hg mobilisation from acidic toalkaline environment; the least Hg dissolution has been found aroundpH 3 while the most around either pH 5 or 11 (Xu et al., 2014; Yinet al., 1996; Farrah, 1979; Reimers and Krenkel, 1974; Wallschlägeret al., 1998a,b; Barrow, 1992; Lockwood and Chen, 1973). The effect ofpolysulphides on Hg solubility is highly pH dependent. The chemicalspeciation of Hg and [CH3Hg]+wasmodelled at pH 4.0 and 7.0 in a con-ceptual wetland soil/sedimentwith typical concentrations of thiols, sul-phides (nM to low mM range of total sulphide concentrations), Hg and[CH3Hg]+. At pH 7.0, thiols was outcompeted by Hg–polysulfide com-plexes when S(II) concentration exceeded 0.01 mM. If pH decreasedto 4.0, the competitiveness of polysulphides was much less, resultingin 50% complexation of Hg to thiols and 50% to bisulfides andpolysulphides (Skyllberg, 2008).

3. Technologies for the remediation of Hg contaminated soil

In a soil remediation context, trace elements cannot be degraded inthe same way as organic contaminants. Instead, the strategy is to relo-cate the element from one place (i.e., contaminated site) to another(i.e., landfill). Alternatively, they can be immobilised into stable formsin situ (Kumpiene et al., 2008;Mukherjee et al., 2004). Inmost cases, ex-traction is carried out to separateHg from soil or to lower the concentra-tion of bioavailable Hg in soil to acceptable levels or to reduce thevolume of contaminated soil. Immobilisation on the other hand, isbased on encapsulation and stabilisation of Hg in the soil to avoid nega-tive effects on humans or other organisms (Dermont, 2008; Petruzzelliet al., 2013; Wang et al., 2012a,b). Table 3 and the following chaptersfocus on the development stage, applicability, limitations, and sec-ondary waste of four major technologies for Hg-contaminated soil, asthey aremostwidely used and implemented in thefield. Other remedialtechniques, e.g., vitrification, nanotechnology and electro-remediationetc., are not further discussed here but can be found in earlier studieson this topic (Bonner et al., 1983; Dermont et al., 2008; Mulligan et al.,2001; US EPA, 2007; Wang et al., 2012a,b).

3.1. Soil washing

Soil washing is primarily a physical separation (PS) process inwhichwater is used to reduce Hg concentrations in soil. The principle is basedon the concept that most contaminants tend to bind to fine (clay andsilt) rather than coarse particles (sand and gravel) in soil (FRTR, 1995;Sierra et al., 2011; US EPA, 2007). Processes such as hydroclassification,gravity concentration, attrition scrubbing and frothflotation are typicallyapplied to separate particles in soil washing (Vik and Bardos, 2003). Hgbound to the fine particles is concentrated for further treatment, whilethe coarse-grained soil is left relatively clean and requires noadditional treatment. Washing-fluid from the process is either reusedin the process or disposed of (BioGenesis Enterprises and Roy F., 1999;Griffiths, 1995). Physical separation allows for recycling of the treatedsoil, and the process duration is typically short to medium (Dermontet al., 2008; Griffiths, 1995; Sierra et al., 2011). The methodologypresents several advantages: (1) the volume of soil to be further treatedor disposed of is considerably reduced; (2) the treatment systems areeasily modular, and some full-scale mobile units are available for on-siteremediation; (3) the technologies are well established in the mineral-processing industry; and (4) the operational costs are usually low com-pared to those of thermal treatment, etc. However, PS is difficult orunfeasible in the following cases:

• When Hg is strongly bound to soil particles owing to high levels ofinsoluble humic substances/clay minerals (Abumaizar and Smith,1999; Dermont, 2008; FRTR, 1995; Griffiths, 1995; Xu et al., 2014).

• When Hg is present in all particle-size fractions of contaminated soil(Xu et al., 2014).

• When the soils contain silt/clay content in excess of 30–50%.

Table 3Overview of selected Hg treatment technologies.

Technology Soil washing Stabilisation/solidification Thermal treatment Biological techniques

Description Physical separation concentrates Hginto a smaller volume; chemicalextraction solubilizes Hg in soilwith chemical reagents.

Stabilisation is to put Hg into stableand highly insoluble forms over wideranges of pH and oxidisingconditions; solidification is to trapthe stabilised Hg in a rigid anddurable matrix.

Using heat and reducedpressure to volatilize Hg,followed by condensing Hgvapours into liquid elementalHg.

Using Hg resistant microorganismsor plants to remove Hg from soil;using biosolids, microbes or plantsto sorb Hg and reduceits toxicity or bioavailability to theenvironment.

Developmentstatus

Full scale; not extensively used inthe United States; commonly usedin Europe.

Full scale; widely used in theUnited States; not extensively usedin Europe.

Full scale; specific applications. Pilot scale; in development phase;used more in the United Statesthan in Europe.

Applicability On/off site On/off site On/off site On siteLimitation Physical separation is difficult with

soils of high levels ofclay/insoluble humicsubstances/viscosity; for chemicalextraction, use of chemical agentsis both expensive and hazardous.

Increased volume of the treatedmaterial; compounds such as organicmatter can interfere; long-termmonitoring is needed.

Requires gas emission controland specialized facilities,pre-treatment to assist meltingis needed; the capital cost isvery high.

Requires more pilot studies toevaluate the efficiency. Remainingliability issues, includingmaintenance for an indefiniteperiod of time.

Secondarywaste

Sludge/water containing chemicalsand contaminants.

None Off-gas, waste water ifpre-treatment is included.

Contaminated plants; Hg emissionfrom microbial and plant volatilisation

References (Abumaizar and Smith, 1999;Dermont et al., 2008; FRTR, 1995)

(Guo et al., 2011; Mulligan et al., 2001;US EPA, 2007; Zhang and Bishop, 2002)

(Dermont, 2008; Richter andFlachberger, 2010; US EPA,1997b, 2007)

(Dermont, 2008; Kotrba, 2009;Padmavathiamma and Li, 2007;Peer et al., 2006; Tangahu et al., 2011)

47J. Xu et al. / Environment International 74 (2015) 42–53

• When differences in density or surface properties between Hg-bearingand clean particles are not significant.

Soil washing in which chemicals are used to remove Hg from soil iscalled chemical extraction (CE); and this method can be used in combi-nationwith PS. Chemical extraction solubilises Hgwith solvents such asacids, alkalis or chelating agents. Generally, the application of acids andalkali relies on the dissolution of Hg compounds or/and soil componentsthat sorb Hg, while chelating agents mobilise Hg by forming soluble Hgcomplexes (Dermont et al., 2008). The mixture of solids and liquids isseparated using hydrocyclones, and the solids are subsequently trans-ferred to a water-rinse system to remove reagents and contaminants.The extraction fluid and rinse water are mixedwith commercially avail-able precipitants, such as sodium hydroxide, lime or other formulations,and a flocculant to remove Hg. The precipitated solidsmay require addi-tional treatment ormay be disposed of in a landfill. This is one of the fewtreatments that permanently separate Hg from the soil, and the efficien-cy can be as high as 99% (FRTR, 1995, 2001; Twidwell and Thompson,2001; Universal Dynamics, 2004; US EPA, 1997b).

Generally, theprincipal advantage of CE over PS is that soil-sorbedHgthat is not soluble in water can be removed (Dermont et al., 2008; USEPA, 2007). However, large-scale application of CE presents severalobstacles:

• The use of chemical agents significantly increases processing costs,and the presence of certain chemicals in washing fluid complicateswater recycling and treatment.

• The processed soil may be inappropriate for re-vegetation or for directon-site disposal (Dermont et al., 2008; Guo et al., 2011).

• High levels of soil organic matter might strongly retain Hg in the soilthereby inhibiting the extraction (Bollen, 2011).

• Clay/silt content over 50% (of soil dry weight) may require longercontact time thus reducing the efficiency (Griffiths, 1995).

• High concentrations of cations of major elements, such as Fe and Ca,may interfere with the chelating process of Hg (FRTR, 2001).

• High heterogeneity of soils affects the formulations of extracting fluidand may require multiple processing steps (FRTR, 2001).

3.2. Stabilisation/solidification

Stabilisation is the process of convertingHg into chemical forms thatare stable and highly insoluble over wide ranges of pH and redoxconditions in soil (US EPA, 2007; Zhang and Bishop, 2002). Solidifica-tion is the encapsulation of the stabilised Hg forms in a rigid and

durable matrix—for example, in a solid monolith. Stabilisation/solid-ification (S/S) decreases Hg availability for biological uptake, as wellas Hg release to either surface or groundwater (Mulligan et al., 2001;Svensson, 2006). It is one of the most commonly used techniques be-cause of its compatibility with a wide range of contaminants and soiltypes (Mulligan et al., 2001). The process renders stable and sparinglyleachable Hg forms in the treated soil and 20 to 90% of the Hg can be se-questered (Dermont, 2008; US EPA, 1998; US EPA, 2002a). The technol-ogy is commercially available and generates a residue that typicallydoes not require further treatment prior to disposal (US EPA, 2007).

Ex situ S/S is relatively conventional while in situ S/S is an emergingtechnology still at the developmental stage. The main stabilising agentsused are mineral compounds, such as phosphates, lime, fly ashes andalumino-silicates. Stabilisation based on phosphate additives is themost thoroughly investigated technique for decreasing the bioavailabil-ity and solubility of Hg. However, few large-scale applications have beenperformed thus far (Dermont, 2008; Randall and Chattopadhyay, 2004;US EPA, 1997b) and more research is needed to evaluate the long-termperformance. The advantages and disadvantages of ex situ and in situS/S are presented in Table 4.

The applicability of S/S depends on themobility of Hg, which in turndepends on its oxidation state, the pH of the disposal environment andthe specific Hg compound present in the soil. The following factorsconstitute significant challenges:

• Typically, the leachability of heavy metals increases with decreasingpH. However, soluble compounds ofHg canalso format highly alkalineconditions (Randall and Chattopadhyay, 2004; Skyllberg, 2012;Svensson, 2006; US EPA, 2007; Xu et al., 2014).

• Interfering elements may decrease the process efficiency. E.g., chlo-rides form highly soluble species with Hg; dissolved organic matteris known to break insoluble Hg bonds, such as HgS (Schuster, 1991;Skyllberg et al., 2006).

• Insufficient mixing lowers the efficiency of process, especially for insitu S/S (Mulligan et al., 2001; Schuster, 1991).

3.3. Thermal treatment

Thermal treatment uses elevated temperatures to remove Hg fromsoil through volatilisation. High temperatures are usually appliedalong with reduced pressure to volatilise elemental Hg as well as Hgcompounds. Volatilisation is followed by a condensation of the Hgvapours into liquid elemental Hg. A typical thermal desorption unit for

Table 4Advantages and disadvantages of ex situ and in situ stabilisation/solidification for Hg-contaminated soil.

Advantages Disadvantages References

Ex situ Fast remediation of the site; nearly independent ofsite geology; process optimization can be easilycontrolled and final efficiency can be easily verified.

Involves exposure risks for workers and the environmentduring excavation; requires transport and fixed facilities;difficult with complex existing infrastructure;significantly increases disposal volume.

(Boyce and Almskog, 1999; Dermont, 2008;FRTR, 2001; Mulligan et al., 2001; US EPA, 2007)

In situ Minimises exposure to humans and the environment;requires no excavation, transport or fixed facilities;usually more cost-effective than ex situ.

Difficult to verify process efficiency; influenced bysite-specific conditions; relatively lower mixing efficiency;raises concerns about site's long-term integrity; limitsfuture reuse of the site.

(Dermont, 2008; FRTR, 1998; Ghosh et al., 2011;Svensson, 2006; US EPA, 2007)

48 J. Xu et al. / Environment International 74 (2015) 42–53

Hg removal operates at temperatures from 320 to 700 °C (US EPA,2007; Vik and Bardos, 2003). It is one of the fewmethods that are effec-tive for very high Hg concentrations (N260 mg/kg). It removes Hg froma soil matrix with an efficiency of 41 99%, typically reaching over 99%(Boyce and Almskog, 1999; US EPA, 2007). It also holds the potentialfor Hg recovery.

Temperatures and retention time are the main factors determiningthe decontamination levels that thermal treatment can achieve (Wanget al., 2012a,b). According to several experiments involving Hg thermaldesorption, greater efficiency is achieved at relatively high tempera-tures, e.g. from 460 to 700 °C (Busto, 2011; Richter and Flachberger,2010). To meet the concerns about the high-energy consumption asso-ciatedwith such temperatures, one proposal is to operate thermal treat-ment at a lower temperature for a longer time (Qu et al., 2004).

Some major disadvantages of this method as well as current mea-sures to overcome these disadvantages are as follows:

• Capital cost is very high because of the high-temperature process andthe specialised facilities required (Dermont, 2008; US EPA, 2003).Consequently, the process is usually performed under reduced pres-sure in order to lower the boiling points of Hg and its compounds(Comuzzi, 2011; US EPA, 2003). Recently, it has been proposed thatsolar energy be used instead of conventional non-renewable energysources (Wang et al., 2012a,b).

• Off-gas containing hazardous substances is generated and it requiresfurther treatment (Mulligan et al., 2001). Emissions of processinggases could be reduced by operating Hg desorption under a vacuum(US EPA, 2003, 2007);

• The presence of larger particles may impair heat transfer between theheating elements or the combustion gas and the medium; however,smaller particles may increase the particulate content of the off-gas(US EPA, 2002b).

3.4. Biological techniques

3.4.1. PhytoremediationPhytoremediation involving phytostabilisation, phytoextraction and

phytovolatilisation, relies on selected plants to clean up Hg contamina-tion in soil. It is also used as a final decontamination step, in conjunctionwith other treatment technologies (Padmavathiamma and Li, 2007;Peer et al., 2006; Tangahu et al., 2011). Phytostabilisation immobilisesHg in soil through absorption and accumulation of Hg in plant roots orthrough Hg precipitation in the root zone; the technique prevents Hgmigration by soil erosion and deflation (Peer et al., 2006). For example,it has been shown thatwillow species can accumulate bioavailableHg inthe root system thereby reducing the concentration of bioavailableHg inthe rhizosphere while leaving total Hg in the soil relatively unchanged(Wang et al., 2005). Recent studies demonstrated that Selenium (Se)plays an important role in limiting the bioaccessibility, absorption, andtranslocation/bioaccumulation of Hg, the formation of an Hg–Se insolu-ble complex in the rhizospheres and/or roots being the most plausiblemechanism (Zhang et al., 2012; Zhao et al., 2013).

Phytoextraction is the process of Hg uptake/absorption and translo-cation by plant roots into the above-ground parts (e.g., shoots) that can

then be harvested and burned (Pant et al., 2011; Petruzzelli et al., 2013).Chemicals such as ammonium thiosulfate [(NH4)2S2O3] and potassiumiodide (KI) can be used to assist in phytoextraction by promoting Hgbioavailability prior to plant accumulation (Pedron et al., 2011; Wanget al., 2014). Phytovolatilisation is a remediation strategy unique to Hgbecause of the relatively high volatility of this metal. Hg is taken up byplant roots, transported through the xylem, and finally released to theatmosphere from cellular tissues (Wang et al., 2012a,b). The emissionof Hg from the leaf tissues is then strongly affected by environmentalparameters such as light intensity and air temperature (Leonard et al.,1998). The main advantage of phytovolatilisation is the removal of Hgfrom soils without harvest and disposal. Although there could be someskepticism regarding the safety of this strategy, safety assessment stud-ies on Hg phytovolatilisation have indicated that the wide dispersionand dilution in the atmosphere outweighs the potential risks (Morenoet al., 2005).

Plant species is the essential factor determining the success ofphytoremediation for Hg. Candidate plants for phytostabilisation needto have extensive root systems and the toxicity of Hg to plant rootsand the survival rate of the plants being used also need to be takeninto consideration. Since no plant species have been identified as Hghyper-accumulators, plant species with high biomass and physiologicalmechanisms enhancing Hg uptake are crucial factors in phytoextraction(Wang et al., 2012a,b).

Phytoremediation has been widely accepted and is consideredthe most environmentally friendly and aesthetically attractivesoil-treatment option. It does not require environmentally damagingchemicals or costly heat treatment nor does it involve large-scale trans-port and expensive equipment (Petruzzelli et al., 2013). However, thefollowing issues need to be taken into account when phytoremediationis selected as the remedial method:

• The number of suitable plant species that can be used for Hg uptake islimited; moreover, the remediation effects are limited by the depth ofthe plants' root zones (Dermont, 2008; Peer et al., 2006; Tangahuet al., 2011). Hg does not belong to trace nutrient elements for plants,and so far not many suitable species (e.g., Pteris vittata and Sesbaniadrummondii) (Su et al., 2007; Venkatachalam et al., 2009) have beenidentified for this purpose. To address this, transgenic techniqueshave been introduced to enhance plant tolerance to Hg, yet theirtechno-economic perspective and environmental safety need to becarefully evaluated (Bizily, 1999; Kotrba, 2009).

• The high volatility of elemental Hgmay give rise to atmosphere pollu-tion, which in turn leads to secondary pollution. At the same time,biological exposure to Hg may occur as plants are used; Hg contami-nants may enter the food chain through herbivorous animals. Thereare also concerns about managing the biomass that contains Hgwhich may demand more efforts and thereby increase the cost (Peeret al., 2006; Weis and Weis, 2004).

• The process is long and at least several growing seasons are requiredto clean up a site, while excavation/disposal or incineration takesweeks to months.

• Very few studies have reported the use of plants for thephytovolatilisation of Hg (Wang et al., 2012a,b); phytovolatilisation

49J. Xu et al. / Environment International 74 (2015) 42–53

of Hg has caused public anxiety due to the secondary contaminationof the environment with elemental Hg (Wang et al., 2012a,b).

3.4.2. Bioremediation

Bioremediation consists of volatilisation and biosorption. The princi-ple of volatilisation includes the utilization of Hg-resistant bacteria,which carry the operons for binding, transport and detoxifying Hg(II)and organic Hg species to elemental Hg, thereby preventing foodchain accumulation (Brown et al. 2003; Permina et al. 2006; Sinha andKhare, 2012; Wagner-Dobler, 2013). On the other hand, microbialvolatilisation that is not due to the mer-operon encoded system hasalso been discovered.Wiatrowski et al. (2006) reported that the dissim-ilatory metal reducing bacterium S. oneidensis MR-1 reduces Hg(II) toHg(0) by an activity not related to the mer A mercuric reductase. Liveor dead microbial biomass from bacteria, fungi or algae, has been usedfor biosorption to promote the formation of less toxic and less solubleHg forms (US EPA 2004). These techniques involve immobilisationmechanisms relying on different physicochemical mechanisms, suchas adsorption, surface complexation, ion exchange, or surface precipita-tion (Le Cloirec and Andres, 2005; François et al., 2012).

Although awide range ofmicroorganismshas beendiscovered that areable to degrade organic Hg contaminants in soil, a number of challengeshave been identified, such as poor bioavailability of the Hg, presence ofother toxic compounds which might hinder the activity of Hg-tolerantmicrobes, inadequate supply of nutrients and insufficient biochemicalpotential for effective biodegradation (Krämer and Chardonnens,2001). Therefore, genetic engineering has been explored for more effi-cient biological treatment of Hg contaminated soils using geneticallymodifiedmicroorganisms (Bae et al., 2003; He et al., 2011). For example,the regulator mer R which exhibits high affinity for Hg was engineeredto be localized on the surface of Escherichia coli cells, enabling sixfold- higher Hg biosorption than the corresponding wild-type cells(Bae et al., 2001). However, up to date, no bioremediation field studiesfor Hg in pilot scale have been conducted (US EPA, 2007).

4. Technology improvement

Soil washing and stabilisation/solidification (S/S) have beenmost ex-tensively applied amongst the above-mentioned technologies (Dermont,2008; FRTR, 1995). Therefore, the following chapter will focus on thecritical aspects for the improvement of these two technologies.

4.1. Soil washing

The effectiveness and the cost of soilwashingdepends on several soilcharacteristics, such as particle size distribution, clay content, humiccontent, heterogeneity, etc. (Dermont et al., 2008; Sierra et al., 2011;US EPA, 2001).

4.1.1. Particle sizeThe particle size of feed soil is one of the most significant factors

affecting the applicability of physical separation (PS). Usually, PS thatuses hydroclassifiers and gravity concentrators (jig, shaking table andspiral) can be effectively applied to coarse-grained fraction (N63 μm).Therefore, if the silt/clay (b63 μm) content exceeds 30–50%, PS processcan be problematic (United States Environmental Protection Agency(US EPA), 1997a,b,c). Separation processes that incorporate froth flota-tion may be effective for treating relatively fine particles (b63 μm)(FRTR, 1995; PPG Canada Inc. and Biogenie SRDC Inc., 1993). Froth flota-tion is a physical–chemical technique that exploits differences in polar-ity and surface tension to separate metal-bearing particles from a soilmatrix. The principle is based on the surface tension of a particle'shydrophobic surfaces to the air bubbles injected into soil slurry. Infull-scale applications, froth flotation has been mostly applied incombination with hydroclassification and gravity concentration. As

an example, a Hg-contaminated site in New Jersey was remediatedusing screening and froth floatation, and the initial concentration of100 mg/kg was reduced to 1 mg/kg (FRTR, 1995).

4.1.2. Liberation degreeThe liberation degree refers to the release availability of Hg and its

compounds according to various associations with the soil particles,which is significant for predicting the applicability of physical separa-tion. It can be categorized as (i) incorporated into a mineral lattice,(ii) specifically bound, (iii) physically sorbed (non-specifically bound),or (iv) free from associations. Physical separation is difficult or unfeasi-ble when Hg is specifically bound on soil particles of all grain sizes, oreven incorporated in the mineral lattice (Dermont et al., 2008). Wetsieving is the easiest way to remove fine particles attached to coarsersoil constituents (Xu et al., 2014); attrition scrubbing is often used todisaggregate the small particles that are bound more strongly to coarseparticles or to remove the coating of a particle's surface (Kyllönen,2004). The typical attrition scrubbing aims at scouring and breaking,which are accomplished mostly through particle-to-particle attritionand via the interaction between paddles and soil particles (UnitedStates Environmental Protection Agency (US EPA) 1997a,b,c, 2007). In-stead of hydroclassification which functions by size differences, gravityconcentration is often the method of choice when the liberation degreeof the contaminant is low, exploiting gravity differences between con-taminated and clean soil. A Hg removal rate of 80 99% has been reachedfor sandy soil using gravity concentration (US EPA, 1995). Chemicalextraction can also facilitate soil washing if the Hg contaminant isspecifically bound or incorporated in a soil particle. Chelating agentsare used to release Hg specifically sorbed in soil, whilst acids/alkaliscan solubilise the entire mineral that incorporates Hg (Dermont et al.,2008).

4.1.3. Organic matterHigh levels of organic matter (OM) in soil tend to hinder the Hg

desorption, thereby limiting the effectiveness of soil washing (Sierraet al., 2011). The presence of soil OM interferes with both Hg distribu-tions in particle-size fractions andHgmobilisation by amobile chelatingagent, such as chloride (Xu et al., 2014). Salts of weak organic acids(e.g., citrate and tartarate) have been used to facilitate Hg mobilisationin soils rich in OM. Up to 92% of the Hg has been removed from a clayloam, with an OM-associated Hg fraction up to 60% (Subires-Munozet al., 2011;Wasay et al., 2001). Physicalmethods can also be integratedto enhance Hg washing efficiency for organic soils. In a full-scale reme-diation of dredged materials from the New York/New Jersey HarborEstuary, a 92% removal rate for Hgwas achieved by combining chemicalextraction and attrition scrubbing, with the total organic carbon rangingfrom 3 to 10% (w/w) (BioGenesis Enterprises and Roy F., 1999; Jones,2001).

4.1.4. ExtractantAcid leachate generated from acidogenesis at waste-collection

stations or landfills has been proposed as a substitute for extractingchemicals. Contaminants in such washing-fluids could subsequentlybe degraded in a methanogenic process as the extracted organic con-taminants are biodegradable (Lagerkvist et al., in manuscript). The useof acidogenic leachates to extract hazardous elements from soil (Hgand Zn, etc.) was found to be equally efficient as when 0.1 M EDTA(Ethylenediaminetetraacetic acid, widely used as an extractant forheavy metals) was used, although Hg extraction by either EDTA or theacidogenic leachate was sparing (Lagerkvist et al., in manuscript).

4.2. Stabilisation/solidification

Interfering elements as well as types of reagent and binders arethe factors that affect the efficiency and cost of Hg stabilisation/solidification (S/S) methods (US EPA, 2007; Zhang and Bishop, 2002).

50 J. Xu et al. / Environment International 74 (2015) 42–53

The excavation necessary for ex situ S/S further increases its cost, whilethe verification of the process efficiency becomes amajor considerationfor in situ S/S.

4.2.1. InterferencesCertain ions and soluble organic constituents in soil may impair

Hg stabilisation by forming mobile Hg species. Activated carbon (AC)powder has been suggested as a tool for solving this problem, as it cansimultaneously capture Hg and limit the interferences from dissolvedorganics. Furthermore, pre-treatment of AC by impregnating it withsulphide has been proposed to improve its adsorption capacity for Hg(Kwon, 2000). Zhang and Bishop (2002) conducted an experimentwith a Hg concentration of 2273 mg/kg and a chloride concentrationof 5940 mg/l. A low-cost AC powder impregnated with sulphides wasused to stabilise Hg, and ordinary Portland cementwas used afterwardsfor solidification. Toxicity characteristic leaching procedures and con-stant pH leaching tests showed that the S/S process successfullyimmobilised Hg in the cement matrix. Even under the attack of a highchloride concentration, little Hg leached from the matrix.

4.2.2. Binder typesConventional solidification using only Portland cement and fly ash is

very inefficient to encapsulate Hg because Hg does not form a low-solubility hydroxide (Barrow, 1992; Schuster, 1991). Therefore, bindertypes are critical and mixing the binder with stabilising agents, suchas sulphur, sodium sulphate or ferric–lignin derivatives, will benefitsolidification (Chattopadhyay and Randall, 2003; Zhang and Bishop,2002). The sulphur polymer S/S (SPSS) process has been used toconvert Hg compounds into the highly insoluble HgS form and tosimultaneously encapsulate the contaminated soil (Fuhrmann, 2002;US EPA, 2007). Chemically bonded phosphate ceramic (CBPC)techniques are similar to SPSS, immobilising Hg through both chemicalstabilisation and physical encapsulation. Addition of a small amount ofsodium sulphide or potassium sulphide to the binder to formHgS greatlyimproves the performance of CBPC after the formation of low-solubilityHg phosphate solids. A product of CBPC process can have a Hg load ashigh as 78% by weight, and an advantage of phosphate ceramic is itshigh physical stability (Randall and Chattopadhyay, 2004; Wagh,2013). The S/S of this phosphate ceramic can be further improved byadding ground granulated blast furnace slag which significantlyreduces Hg leaching and enhances the compressive strength of Hg-richCBPC matrixes (Lin et al., 1995; Wagh, 2004).

5. Conclusions

Natural processes as well as human activities contribute to Hg con-tamination in soil. Organic matter (OM), clay/minerals and complexa-tion ligands of soil are the principal factors influencing Hg mobility,which is crucial for evaluating and optimising remediation techniques.

Remediation techniques have been proposed to control Hg contam-ination in soils and to avoid adverse health effects. Amongst, thermaltreatment can be applied to soils with very high concentrations of Hg(up to 34,000 mg/kg) and its remedial efficiency is fairly high (up to99%). But altered soil properties after treatment (460–700 °C to achievegreater Hg removal) and very high capital and operational costs remainchallenging. Hence, extended high vacuum-rotary conditions andreliance on renewable energy sources have been proposed to counter-balance the disadvantages. Physical separation and chemical extractionare often used in combination in soil washing. Strategies as gravity con-centration, attrition scrubbing and froth flotation, etc., have been devel-oped to enhance treatment efficiency. Soils with various contaminantsmay requiremultiple washing steps and the formulations of extractantsmight be complex. Low liberation degree of Hg or high contents of clay/silt andOM remains challenging for soil washing. Although soil washingis commercially available, feasibility studies are needed for individualsoils prior to field implementation. Stabilisation/solidification (S/S)

can remediate sites with a wide range of mixed contaminants and soiltypes. Improved performance has been achieved using sulphur-containing additives in both stabilisation and solidification steps. Theinterference of a mobile complexant for Hg could be overcome bysorbing the complexant onto amedium that is also capable of stabilisingHg. Recently, in situ S/S technology has been considered as an alterna-tive to ex situ S/S because of its cost advantage and the option of sitere-vegetation. This could be a promising direction in the interest ofsaving landfill resources, but significant monitoring is required becauselittle is known about the long-term stability and integrity of thestabilised and solidified product.

Phytoremediation is environmentally friendly; nevertheless, suitableplant species for Hg immobilisation or uptake or propagation are few.They are additionally constrainedby soil fertility and climate regions. Che-lators such as potassium iodide (KI), sodium thiosulfate (Na2S2O3) andammonium thiocyanate (NH4SCN) have beenused to increaseHg solubil-ity in order to enhance the plant uptake of Hg from soil. Current researchis being conducted on enhancing toxicity tolerance to Hg using geneticmodification. However, ecological exposure of Hg through the use ofsuch plants as well as the food chain should be taken into account.

Future research should focus on the implementation of the differentremediation techniques under field conditions. As each Hg contaminat-ed site is unique, a specific and deep evaluation of the site must becarried out prior to the application of the selected technique. This willlead to specific modification and improvement of the above-mentionedremediation techniques.

Acknowledgements

This studywasfinancially supported by the Swedish Research Counciland the RagnSells AB (contract no. D0697801). The authors thank theproject Surte 2:38, Ale municipality, Sweden, for the delivery of soilsamples.

References

Abumaizar RJ, Smith EH. Heavy metal contaminants removal by soil washing. J HazardMater 1999;70(1–2):71–86.

AMAP/UNEP. Technical background report for the globalmercury assessment 2013. ArcticMonitoring and Assessment Programme, Oslo, Norway/UNEP ChemicalsBranch,Geneva, Switzerland; 2013.

Annual General Meeting of China Council for International Cooperation on Environ-ment and Development. Special policy study on mercury management in China,2011; 2011. [http://www.cciced.net/encciced/policyresearch/report/201205/P020120529368288424164.pdf].

Bae W, Mehra RK, Mulchandani A, Chen W. Genetic engineering of Escherichia coli forenhanced uptake and bioaccumulation of mercury. Appl Environ Microbiol 2001;67(11):5335–8.

Bae W, Wu CH, Kostal J, Mulchandani A, Chen W. Enhanced mercury biosorption bybacterial cells with surface-displayed MerR. Appl Environ Microbiol 2003;69(6):3176–80.

Barkay T, Wagner-Dobler I. Microbial transformations of mercury: potentials, challenges,and achievements in controlling mercury toxicity in the environment. Adv ApplMicrobiol 2005(57):1–52.

Barkay T, Miller SM, Summers AO. Bacterial mercury resistance from atoms to ecosystems.FEMS Microbiol Rev 2003;27(2‐3):355–84.

Barrow NJ. The effects of pH and chloride concentration on mercury sorption. I. Bygoethite. J Soil Sci 1992;43(2):295.

Benoit JM, Gilmour CC, Mason RP, Heyes A. Sulfide controls on mercury speciation andbioavailability to methylating bacteria in sediment pore waters. Environ Sci Technol1999;33(6):951–7.

Beone G, Trevisan M, Cattani I, Del Re A, Boccelli R, Zhang H. The role of natural purifiedhumic acids in modifying mercury accessibility in water and soil. J Environ Qual2009;38(2):493–501.

BerndtME, Engesser J, Berquó TS. Mercury chemistry andMössbauer spectroscopy of ironoxides during taconite processing on Minnesota's iron range. Proceedings Air QualityV, International Conference on Mercury, Trace Elements, SO3, and Particulate Matter.Washington, DC; 2005.

BioGenesis Enterprises, Inc., Roy F.Weston, Inc.. BioGenesisSM sedimentwashing technology.Full-scale, 40 Cy/Hr, sediment decontamination facility for the NY/NJ HarborFinalReport on the Pilot Demonstration Project; 1999. [December. http://www.bnl.gov/wrdadcon/publications/reports/BioGenesis_BNL.PDF].

Bizily SP. Phytoremediation of methylmercury pollution: MerB expression in arabidopsisthaliana confers resistance to organomercurials. Proc Natl Acad Sci U S A 1999;96(12):6808.

51J. Xu et al. / Environment International 74 (2015) 42–53

Bollen A. Mercury extraction from contaminated soils by L-cysteine: species dependencyand transformation processes. Water Air Soil Pollut 2011;219(1–4):175–89.

Bonner WF, Brouns R A, Buelt J L. U.S. Patent No. 4,376,598. Washington, DC: U.S. Patentand Trademark Office,1983.

Boyce PK, Almskog KE. Much ado about mercury: evaluation of treatment options formercury contaminated soil at BNL. Geology of Long Island and Metropolitan NewYork Conference; 1999.

Branch UC. The global atmospheric mercury assessment: sources, emissions and transport.Geneva: UNEP-Chemicals; 2008.

Bravo AG, Cosio C, Amouroux D, Zopfi J, Chevalley PA, Spangenberg JE, et al. Extremely ele-vatedmethylmercury levels inwater, sediment and organisms in a Romanian reservoiraffected by release of mercury from a chlor-alkali plant. Water Res 2014;49:391–405.

Brown NL, Stoyanov JV, Kidd SP, Hobman JL. TheMerR family of transcriptional regulators.FEMS Microbiol Rev 2003;27(2‐3):145–63.

Burton DT. Bioaccumulation of total mercury and monomethyl mercury in the earthwormEisenia fetida. Water Air Soil Pollut 2006;170(1–4):37–54.

Busto Y. Potential of thermal treatment for decontamination of mercury containingwastes from chlor-alkali industry. J Hazard Mater 2011;186(1):114–8.

Chattopadhyay S, Randall PM. Evaluation of chemically bonded phosphate ceramics formercury stabilisation of a mixed synthetic waste. US Environmental Proectectionagency, Office of Research and Development, National Risk Management ResearchLaboratory; 2003.

Clarkson TW. Human exposure to mercury: the threemodern dilemmas. J Trace Elem ExpMed 2003;16(4):321.

Colombo MJ, Ha J, Reinfelder JR, Barkay T, Yee N. Anaerobic oxidation of Hg (0) andmethylmercury formation byDesulfovibrio desulfuricansND132. GeochimCosmochimActa 2013;112:166–77.

Comuzzi C. Salt-assisted thermal desorption of mercury from contaminated dredgingsludge. J Hazard Mater 2011;193:177–82.

Cui WW,Wang XL, Duan JJ, Yang M, Zi W, Liu XC. Advances on remediation techniques ofheavy metal Cd and Hg in contaminated soil. Guizhou Agric Sci 2011;7:67.

Dermont G. Metal-contaminated soils: remediation practices and treatment technologies.Pract Period Hazard Toxic Radioact Waste Manag 2008;12(3):188–209.

Dermont G, Bergeron M, Mercier G, Richer-Lafleche M. Soil washing for metal removal: areview of physical/chemical technologies and field applications. J HazardMater 2008;152(1):1–31.

Dyrssen D. The sulphur–mercury (II) system in natural waters.Water Air Soil Pollut 1991;56(1):507.

Farrah H. pH effects in the adsorption of heavy metal ions by clays. Chem Geol 1979;25(4):317.

Federal Remediation Technologies Roundtable (FRTR.). Technology cost and performance—soil washing at the King of Prussia Technical Corporation Superfund Site[http://costperformance.org/profile.cfm?ID=125andCaseID=125] ; 1995.

Federal remediation technologies roundtable (FRTR). Guide to documenting and manag-ing cost and performance information for remediation projects; 1998. No. EPA 542-B-98-007.

Federal remediation technologies roundtable (FRTR.). Chemical extraction. Federalremediation technologies reference guide and screening manual, version 4.0;2001 [http://www.frtr.gov/matrix2/section4/4-15.html].

Feng X, Li P, Qiu G, Wang S, Li G, Shang L, et al. Human exposure to methylmercurythrough rice intake in mercury mining areas, Guizhou Province, China. Environ SciTechnol 2007;42(1):326–32.

Fernández-Gómez C, Drott A, Björn E, Díez S, Bayona JM, Tesfalidet S, et al. Towards Uni-versal Wavelength-Specific Photodegradation Rate Constants for Methyl Mercury inHumic Waters, Exemplified by a Boreal Lake-Wetland Gradient. Environ Sci Technol2013;47(12):6279–87.

Ferrara R. Atmospheric mercury sources in the Mt. Amiata area, Italy. Sci Total Environ1998;213(1–3):13–23.

Ferrara R. Volcanoes as emission sources of atmospheric mercury in the Mediterraneanbasin. Sci Total Environ 2000;259(1–3):115–21.

Fleming EJ, Mack EE, Green PG, Nelson DC. Mercury methylation from unexpectedsources: molybdate-inhibited freshwater sediments and an iron-reducing bacterium.Appl Environ Microbiol 2006;72(1):457–64.

Frohne T, Rinklebe J, Langer U, Du Laing G, Mothes S, Wennrich R. Biogeochemical factorsaffecting mercury methylation rate in two contaminated floodplain soils. Biogeosci-ences 2012;9(1):493–507.

François F, Lombard C, Guigner JM, Soreau P, Brian-Jaisson F, Martino G, et al. Isolation andcharacterization of environmental bacteria capable of extracellular biosorption ofmercury. Appl Environ Microbiol 2012;78(4):1097–106.

Fuhrmann M. Sulfur polymer solidification/stabilisation of elemental mercury waste.Waste Manag 2002;22(3):327.

Gabriel M. Principal biogeochemical factors affecting the speciation and transport ofmercury through the terrestrial environment. Environ Geochem Health 2004;26(3–4):421–34.

Gavis J. The cycling of mercury through the environment. Water Res 1972;6(9):989.Ghosh U, Luthy RG, Cornelissen G, Werner D, Menzie CA. In-situ sorbent amendments: a

new direction in contaminated sediment management. Environ Sci Technol 2011;45(4):1163–8.

Gilmour CC, Henry EA, Mitchell R. Sulfate stimulation of mercury methylation in freshwatersediments. Environ Sci Technol 1992;26(11):2281–7.

Gilmour CC, Riedel GS, Ederington MC, Bell JT, Gill GA, Stordal MC. Methylmercuryconcentrations and production rates across a trophic gradient in the northernEverglades. Biogeochemistry 1998;40(2–3):327–45.

Gilmour CC, Podar M, Bullock AL, Graham AM, Brown SD, Somenahally AC, et al. Mercurymethylation by novel microorganisms from new environments. Environ Sci Technol2013;47(20):11810–20.

Griffiths RA. Soil-washing technology and practice. J Hazard Mater 1995;40(2):175–89.Guo XX, Liu CQ, Zhu ZZ, Wang ZL, Li J. Evaluation methods for soil heavy metals contam-

ination: a review. Chin J Ecol 2011;30(5):889–96.Hamelin S, Amyot M, Barkay T, Wang Y, Planas D. Methanogens: principal methylators of

mercury in lake periphyton. Environ Sci Technol 2011;45(18):7693–700.He Z, Siripornadulsil S, Sayre RT, Traina SJ, Weavers LK. Removal of mercury from sediment

by ultrasound combined with biomass (transgenic Chlamydomonas reinhardtii).Chemosphere 2011;83(9):1249–54.

Hintelmann H. Levels of total mercury and methylmercury compounds in sediments ofthe polluted Elbe River: influence of seasonally and spatially varying environmentalfactors. Sci Total Environ 1995;166(1–3):1.

Hintelmann H. Measurement of complexation of methylmercury (II) compounds byfreshwater humic substances using equilibrium dialysis. Environ Sci Technol 1997;31(2):489.

Holmes P. Is low-level environmental mercury exposure of concern to human health? SciTotal Environ 2009;408(2):171–82.

Hseu ZY, Su SW, Lai HY, Guo HY, Chen TC, Chen ZS. Remediation techniques and heavymetal uptake by different rice varieties in metal-contaminated soils of Taiwan: newaspects for food safety regulation and sustainable agriculture. Soil Sci Plant Nutr2010;56(1):31–52.

Hsu-Kim H, Kucharzyk KH, Zhang T, Deshusses MA. Mechanisms regulating mercurybioavailability for methylating microorganisms in the aquatic environment: a criticalreview. Environ Sci Technol 2013;47(6):2441–56.

Hu H, Lin H, Zheng W, Rao B, Feng X, Liang L, et al. Mercury reduction and cell-surfaceadsorption by Geobacter sulfurreducens PCA. Environ Sci Technol 2013a;47(19):10922–30.

Hu H, Lin H, Zheng W, Tomanicek SJ, Johs A, Feng X, et al. Oxidation and methylation ofdissolved elemental mercury by anaerobic bacteria. Nat Geosci 2013b;6(9):751–4.

Hutchison AR. Mercury pollution and remediation: the chemist's response to a globalcrisis. J Chem Crystallogr 2003;33(8):631–45.

Jackson TA. Mercury in aquatic ecosystems. Metal metabolism in aquatic environments.US: Springer; 1998. p. 77–158.

Jones KW. Dredged material decontamination demonstration for the port of new York/New Jersey. J Hazard Mater 2001;85(1–2):127.

Kabata-Pendias A. Behavioural properties of trace metals in soils. Appl Geochem 1993;8:3–9.

Kabata-Pendias A. Trace elements in soils and plants. CRC Press; 2010.Karlsson T, Skyllberg U. Bonding of ppb levels of methyl mercury to reduced sulfur groups

in soil organic matter. Environ Sci Technol 2003;37(21):4912–8.Keating MH, Mahaffey KR, Schoeny R, Rice GE, Bullock OR. Mercury study report to

Congress. Volume 1. Executive summary (No. PB–98-124738/XAB; EPA–452/R-97/003). Environmental Protection Agency, Research Triangle Park, NC (United States),Office of Air Quality Planning and Standards; 1997.

Kim CS, Rytuba JJ, Brown Jr GE. EXAFS study of mercury (II) sorption to Fe-and Al-(hydr)oxides: II. Effects of chloride and sulfate. J Colloid Interface Sci 2004;270(1):9–20.

Kotrba P. Genetically modified plants in phytoremediation of heavy metal and metalloidsoil and sediment pollution. Biotechnol Adv 2009;27(6):799–810.

Krämer U, Chardonnens A. The use of transgenic plants in the bioremediation of soilscontaminated with trace elements. Appl Microbiol Biotechnol 2001;55(6):661–72.

Kumpiene J, Lagerkvist A, Maurice C. Stabilisation of As, Cr, Cu, Pb and Zn in soil usingamendments—a review. Waste Manag 2008;28(1):215–25.

Kwaansa-Ansah EE, Voegborlo RB, Adimado AA, Ephraim JH, Nriagu JO. Effect of pH, sul-phate concentration and total organic carbon on mercury accumulation in sedimentsin the Volta Lake at Yeji, Ghana. Bull Environ Contam Toxicol 2012;88(3):418–21.

Kwon S. Evaluation of two sulfur impregnationmethods on activated carbon and bentonitefor the production of elemental mercury sorbents. Environ Eng Sci 2000;17(6):303.

Kyllönen H. Ultrasonically aided mineral processing technique for remediation of soilcontaminated by heavy metals. Ultrason Sonochem 2004;11(3–4):211.

Le Cloirec P, Andres Y. Bioremediation of heavymetals using microorganisms. Bioremedi-ation of aquatic and terrestrial ecosystems; 2005. p. 97–140.

Leonard TL, Taylor GE, Gustin MS, Fernandez GC. Mercury and plants in contaminatedsoils: 2. Environmental and physiological factors governingmercury flux to the atmo-sphere. Environ Toxicol Chem 1998;17(10):2072–9.

Liao L, Selim HM, DeLaune RD. Mercury adsorption–desorption and transport in soils.J Environ Qual 2009;38(4):1608–16.

Lin SL, Lai JS, Chian ES. Modifications of sulfur polymer cement (SPC) stabilisation andsolidification (S/S) process. Waste Manag 1995;15(5):441–7.

Lindqvist O, Johansson K, Bringmark L, Timm B, Aastrup M, Andersson A, et al. Mercury inthe Swedish environment—recent research on causes, consequences and correctivemethods. Water Air Soil Pollut 1991;55(1-2):xi-261.

Lockwood RA, Chen KY. Adsorption of mercury (II) by hydrous manganese oxides. Envi-ron Sci Technol 1973;7(11):1028–34.

Loux NT. An assessment of mercury-species-dependent binding with natural organiccarbon. Chem Speciat Bioavailab 1998;10(4):127–36.

Lovley DR, Giovannoni SJ, White DC, Champine JE, Phillips EJP, Gorby YA, et al. Geobactermetallireducens gen. nov. sp. nov., a microorganism capable of coupling the completeoxidation of organic compounds to the reduction of iron and other metals. ArchMicrobiol 1993;159(4):336–44.

Marvin-DiPasquale M, Agee J, McGowan C, Oremland RS, Thomas M, Krabbenhoft D, et al.Methyl-mercury degradation pathways: a comparison among three mercury-impacted ecosystems. Environ Sci Technol 2000;34(23):4908–16.

Mason RP, Choi AL, Fitzgerald WF, Hammerschmidt CR, Lamborg CH, Soerensen AL, et al.Mercury biogeochemical cycling in the ocean and policy implications. Environ Res2012;119:101–17.

Moreno FN, Anderson CW, Stewart RB, Robinson BH. Mercury volatilisation andphytoextraction from base-metal mine tailings. Environ Pollut 2005;136(2):341–52.

52 J. Xu et al. / Environment International 74 (2015) 42–53

Mukherjee AB, Zevenhoven R, Brodersen J, Hylander LD, Bhattacharya P. Mercury inwaste in the European Union: sources, disposal methods and risks. Resour ConservRecycl 2004;42(2):155–82.

Mulligan CN, Yong RN, Gibbs BF. An evaluation of technologies for the heavy metal reme-diation of dredged sediments. J Hazard Mater 2001;85(1):145–63.

Nance P, Patterson J, Willis A, Foronda N, Dourson M. Human health risks from mercuryexposure from broken compact fluorescent lamps (CFLs). Regul Toxicol Pharmacol2012;62(3):542–52.

Oliva J, Cama J, Cortina J, Ayora C, De Pablo J. Biogenic hydroxyapatite (apatite II™) disso-lution kinetics andmetal removal from acidmine drainage. J HazardMater 2012;213:7–18.

Oremland RS, Culbertson CW, Winfrey MR. Methylmercury decomposition in sedimentsand bacterial cultures: involvement of methanogens and sulfate reducers in oxidativedemethylation. Appl Environ Microbiol 1991;57(1):130–7.

Padmavathiamma PK, Li LY. Phytoremediation technology: hyper-accumulation metals inplants. Water Air Soil Pollut 2007;184(1–4):105–26.

Pak KR, Bartha R. Mercury methylation by interspecies hydrogen and acetate transfer be-tween sulfidogens and methanogens. Appl Environ Microbiol 1998;64(6):1987–90.

Pant P, Allen M, Tansel B. Mercury contamination in the riparian zones along the east forkpoplar creek at oak ridge. Ecotoxicol Environ Saf 2011;74(3):467–72.

Payne R. The adsorption of hydroxyl ions at themercury–solution interface and the anodicdouble-layer capacity in fluoride solutions. J Electroanal Chem 1964;7(5):343.

Payne R. Specific adsorption of chloride ions at the mercury/aqueous solution interface.Trans Faraday Soc 1968;64:1638.

Pedron F, Petruzzelli G, Barbafieri M, Tassi E, Ambrosini P, Patata L. Mercury mobilizationin a contaminated industrial soil for phytoremediation. Commun Soil Sci Plant Anal2011;42(22):2767–77.

Peer WA, Baxter IR, Richards EL, Freeman JL, Murphy AS. Phytoremediation andhyperaccumulator plants. Molecular biology of metal homeostasis and detoxification.Berlin Heidelberg: Springer; 2006. p. 299–340.

Permina EA, Kazakov AE, Kalinina OV, GelfandMS. Comparative genomics of regulation ofheavy metal resistance in Eubacteria. BMC Microbiol 2006;6(1):49.

Petruzzelli G, Pedron F, Rosellini I, Barbafieri M. Phytoremediation towards the future:focus on bioavailable contaminants. Plant-based remediation processes. BerlinHeidelberg: Springer; 2013. p. 273–89.

Pirrone N, Costa P, Pacyna JM, Ferrara R. Mercury emissions to the atmosphere fromnatural and anthropogenic sources in the Mediterranean region. Atmos Environ2001;35(17):2997–3006.

PPG Canada Inc., Biogenie SRDC Inc.. Demonstration of a pilot unit to treat mercury-contaminated soil. Eco-technology innovation section. Montreal: EnvironmentCanada; 1993 [http://publications.gc.ca/collections/Collection/En1-17-9-1993E.pdf].

Qu LY, Fu SZ, Liu L, An YM, Li M. A study on the soil improvement polluted by mercury.J Guizhou Normal Univ (Nat Sci) 2004;2:13.

Randall P, Chattopadhyay S. Advances in encapsulation technologies for the managementof mercury-contaminated hazardous wastes. J Hazard Mater 2004;114(1):211–23.

Ravichandran M. Interactions between mercury and dissolved organic matter—a review.Chemosphere 2004;55(3):319–31.

Reddy MM, Aiken GR. Fulvic acid–sulfide ion competition for mercury ion binding in theFlorida Everglades. Water Air Soil Pollut 2001;132(1–2):89–104.

Reimers RS, Krenkel PA. Kinetics of mercury adsorption and desorption in sediments.J Water Pollut Control Fed 1974:352–65.

Richter RB, Flachberger H. Soil washing and thermal desorption: reliable techniques forremediatingmaterials contaminatedwithmercury. BHMBerg-UndHüttenmännischeMonatshefte 2010;155(12):571–7.

Robles I, Lakatos J, Scharek P, Planck Z, Hernández G, Solís S, et al. Characterization andremediation of soils and sediments polluted with mercury: occurrence, transforma-tions, environmental considerations and San Joaquin's Sierra Gorda Case; 2014.

Rodríguez O, Padilla I, Tayibi H, López-Delgado A. Concerns on liquid mercury andmercury-containing wastes: a review of the treatment technologies for the safestorage. J Environ Manag 2012;101:197–205.

Schlüter K. Review: evaporation of mercury from soils: an integration and synthesis ofcurrent knowledge. Environ Geol 2000;39(3–4):249.

Schuster E. The behaviour of mercury in the soil with special emphasis on complexationand adsorption processes a review of the literature. Water Air Soil Pollut 1991;56(1):667–80.

Selin NE. Global biogeochemical cycling of mercury: a review. Annu Rev Environ Resour2009;34(1):43.

Serrano S, Vlassopoulos D, Bessinger B, O'Day PA. Immobilisation of Hg (II) by co-precipitation in sulfate-cement systems. Environ Sci Technol 2012;46(12):6767–75.

Sierra C, Menéndez-Aguado JM, Afif E, Carrero M, Gallego JR. Feasibility study on the useof soil washing to remediate the As–Hg contamination at an ancient mining andmetallurgy area. J Hazard Mater 2011;196:93–100.

Sinha A, Khare SK. Mercury bioremediation by mercury accumulating Enterobacter sp.cells and its alginate immobilised application. Biodegradation 2012;23(1):25–34.

Skyllberg U. Competition among thiols and inorganic sulfides and polysulfides for Hg andMeHg in wetland soils and sediments under suboxic conditions: illumination ofcontroversies and implications for MeHg net production. Journal of GeophysicalResearch: Biogeosciences (2005–2012) 2008;113(G2).

Skyllberg U. Chemical speciation of mercury in soil and sediment. Environ Chem ToxicolMercury 2012:219–58.

Skyllberg U, Bloom PR, Qian J, Lin CM, Bleam WF. Complexation of mercury (II) in soilorganic matter: EXAFS evidence for linear two-coordination with reduced sulfurgroups. Environ Sci Technol 2006;40(13):4174–80.

Skyllberg U, Drott A, Lambertsson L, Björn E, Karlsson T, Johnson T, et al. Net methylmer-cury production as a basis for improved risk assessment of mercury-contaminatedsediments. AMBIO: A J Hum Environ 2007;36(6):437–42.

Skyllberg U, Qian J, FrechW. Combined XANES and EXAFS study on the bonding ofmethylmercury to thiol groups in soil and aquatic organic matter. Phys Scr 2005;2005(T115):894.

Skyllberg U, Qian J, FrechW, Xia K, BleamWF. Distribution of mercury, methyl mercury andorganic sulphur species in soil, soil solution and stream of a boreal forest catchment.Biogeochemistry 2003;64(1):53–76.

Skyllberg U, Xia K, Bloom PR, Nater EA, Bleam WF. Binding of mercury (II) to reducedsulfur in soil organic matter along upland-peat soil transects. J Environ Qual 2000;29(3):855–65.

Sparks DL. Environmental soil chemistry. Academic Press; 2003.Su Y, Shiyab S, Monts D. Phytoextraction and accumulation of mercury in selected plant

species grown in soil contaminated with different mercury compounds. WM 2007;2007.

Subires-Munoz JD, Garcia-Rubio A, Vereda-Alonso C, Gómez-Lahoz C, Rodríguez-MarotoJM, Garcia-Herruzo F, et al. Feasibility study of the use of different extractant agentsin the remediation of a mercury contaminated soil from Almaden. Sep Purif Technol2011;79(2):151–6.

Svensson M. Mercury immobilisation: a requirement for permanent disposal of mercurywaste in Sweden; 2006.

Tan KH. Principles of soil chemistry. CRC Press; 2011.Tangahu BV, Sheikh Abdullah SR, Basri H, Idris M, Anuar N, Mukhlisin M. A review on

heavy metals (As, Pb, and Hg) uptake by plants through phytoremediation. IntJ Chem Eng 2011;2011.

Tiffreau C, Lützenkirchen J, Behra P. Modeling the adsorption of mercury (II) on (hydr) ox-ides. I: amorphous iron oxide andα-quartz. J Colloid Interface Sci 1995;172(1):82–93.

Tjerngren I, Karlsson T, Björn E, Skyllberg U. Potential Hgmethylation andMeHg demeth-ylation rates related to the nutrient status of different boreal wetlands. Biogeochem-istry 2012;108(1–3):335–50.

Twidwell LG, Thompson RJ. Recovering and recycling Hg from chlor-alkali plant wastewatersludge. JOM 2001;53(1):15–7.

United States Environmental Protection Agency (US EPA). Soil washing at the King ofPrussia technical corporation superfund site, Winslow Township, New Jersey, costand performance report; 1995.

United States Environmental Protection Agency (US EPA). Technology alternatives for theremediation of soils contaminated with arsenic, cadmium, chromium, mercury, andlead; 1997a.

United States Environmental Protection Agency (US EPA). Recent developments for insitu treatment of metal contaminated soils; 1997c.

United States Environmental Protection Agency (US EPA). Technologies for immobilisinghigh mercury subcategory wastes; 1998.

United States Environmental Protection Agency (US EPA). A citizen's guide to soil washing;2001.

United States Environmental Protection Agency (US EPA). Cost and performance summaryreport: thermal desorption at the Lipari Landfill, Pitman, New jersey; 2002a.

United States Environmental Protection Agency (US EPA). Innovative technology summaryreport: the SepraDyne™-reduce system for recovery of mercury from mixed waste;2002b.

United States Environmental Protection Agency (US EPA). Cost and performance summaryreport: thermal desorption at site B, western United States; 2003.

United States Environmental Protection Agency (US EPA). Bioremediation of arsenic,chromium, lead, and mercury; 2004. Washington DC.

United States Environmental Protection Agency (US EPA). Treatment technologies formercury in soil, waste, and water; 2007.

United States Environmental Protection Agency (US EPA). Laws and Regulations; 2013[Online at: http://www.epa.gov/mercury/regs.htm#rcra].

United States Environmental Protection Agency (US EPA). Use of monitored naturalattenuation at Superfund, RCRA corrective action, and underground storage tanksites. Office of solid waste and emergency response directive 9200; 1997b. p. 4–17.

United States Geological Survey (USGS). Mercury in the environment. Online at:http://www.usgs.gov/themes/factsheet/146-00/, 2000.

United States Geological Survey (USGS). USGSminerals yearbook. Online at: http://minerals.usgs.gov/minerals/, 2012.

Universal Dynamics. The REMERC™ process. Online at: http://udl.com/systems/remerc_x.html, 2004.

Venkatachalam P, Srivastava AK, Raghothama KG, Sahi SV. Genes induced in response tomercury-ion-exposure in heavy metal hyperaccumulator Sesbania drummondii.Environ Sci Technol 2009;43(3):843–50.

Vik EA, Bardos P. Remediation of contaminated land technology implementation inEurope:[prepared by working group “Remediation technologies” of the concert-ed action “Contaminated Land Rehabilitation Network for EnvironmentalTechnologies”(CLARINET)…]. Umweltbundesamt; 2003.

Wagh A. Chemically bonded phosphate ceramics: twenty-first century materials with di-verse applications. Elsevier; 2004.

Wagh AS. Recent Progress in Chemically Bonded Phosphate Ceramics. InternationalScholarly Research Notices; 2013.

Wagner-Dobler I, editor. Bioremediation of mercury: current research and industrialapplications. Horizon Scientific Press; 2013.

Wallschläger D, Desai MVM, Spengler M, Wilken RD. Mercury speciation in floodplainsoils and sediments along a contaminated river transect. J Environ Qual 1998a;27(5):1034–44.

Wallschläger D, Desai MV, Spengler M, Windmöller CC, Wilken RD. How humicsubstances dominate mercury geochemistry in contaminated floodplain soils andsediments. J Environ Qual 1998b;27(5):1044–54.

Wang J, Feng X, Anderson CW,Wang H, Wang L. Thiosulphate-induced mercury accumu-lation by plants: metal uptake and transformation of mercury fractionation in soil-results from a field study. Plant Soil 2014;375(1–2):21–33.

53J. Xu et al. / Environment International 74 (2015) 42–53

Wang J, Feng X, Anderson CW,Wang H, Zheng L, Hu T. Implications of mercury speciationin thiosulfate treated plants. Environ Sci Technol 2012a;46(10):5361–8.

Wang J, Feng X, Anderson CW, Xing Y, Shang L. Remediation of mercury contaminatedsites—a review. J Hazard Mater 2012b;221:1–18.

Wang Y, Stauffer C, Keller C, Greger M. Changes in Hg fractionation in soil induced bywillow. Plant Soil 2005;275(1–2):67–75.

Warner KA, Roden EE, Bonzongo JC. Microbial mercury transformation in anoxic freshwatersediments under iron-reducing and other electron-accepting conditions. Environ SciTechnol 2003;37(10):2159–65.

Wasay SA, Barrington S, Tokunaga S. Organic acids for the in situ remediation of soilspolluted by heavy metals: soil flushing in columns. Water Air Soil Pollut 2001;127(1–4):301–14.

Weis JS, Weis P. Metal uptake, transport and release by wetland plants: implications forphytoremediation and restoration. Environ Int 2004;30(5):685–700.

Wiatrowski HA,Ward PM, Barkay T. Novel reduction ofmercury (II) bymercury-sensitivedissimilatory metal reducing bacteria. Environ Sci Technol 2006;40(21):6690–6.

Wollast R, Billen G, Mackenzie FT. Behavior of mercury in natural systems and its globalcycle. Ecological toxicology research. US: Springer; 1975. p. 145–66.

Xu J, Kleja DB, Biester H, Lagerkvist A, Kumpiene J. Influence of particle size distribution,organic carbon, pH and chlorides on washing of mercury contaminated soil.Chemosphere 2014;109:99–105.

Yang YK, Zhang C, Shi XJ, Lin T, Wang DY. Effect of organic matter and pH on mercuryrelease from soils. J Environ Sci 2007;19(11):1349–54.

Yin Y, Allen HE, Huang CP, Sparks DL, Sanders PF. Kinetics of mercury (II) adsorption anddesorption on soil. Environ Sci Technol 1997;31(2):496–503.

Yin Y, Allen HE, Li Y, Huang CP, Sanders PF. Adsorption of mercury (II) by soil: effects ofpH, chloride, and organic matter. J Environ Qual 1996;25(4):837–44.

ZhangH, Feng X, Larssen T, Qiu G, Vogt RD. In inland China, rice, rather than fish, is themajorpathway for methylmercury exposure. Environ Health Perspect 2010;118(9):1183.

Zhang H, Feng X, Zhu J, Sapkota A, Meng B, Yao H, et al. Selenium in soil inhibits mercuryuptake and translocation in rice (Oryza sativa L.). Environ Sci Technol 2012;46(18):10040–6.

Zhang J, Bishop PL. Stabilisation/solidification (S/S) of mercury-containing wastes usingreactivated carbon and portland cement. J Hazard Mater 2002;92(2):199–212.

Zhao J, Gao Y, Li YF, Hu Y, Peng X, Dong Y, et al. Selenium inhibits the phytotoxicity ofmercury in garlic (Allium sativum). Environ Res 2013;125:75–81.