aqueous mercury adsorption by activated carbon

7/24/2019 Aqueous mercury adsorption by activated carbon

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Review

Aqueous mercury adsorption by activated carbons

Pejman Hadi a, Ming-Ho To a, Chi-Wai Hui a, Carol Sze Ki Lin b,Gordon McKay a,c,*

a Chemical and Biomolecular Engineering Department, Hong Kong University of Science and Technology,

Clear Water Bay Road, Hong Kongb School of Energy and Environment, City University of Hong Kong, Tat Chee Avenue, Kowloon, Hong Kongc

Division of Sustainable Development, College of Science, Engineering and Technology, Hamad Bin KhalifaUniversity, Qatar Foundation, Doha, Qatar

a r t i c l e i n f o

Article history:

Received 14 October 2014

Received in revised form

19 December 2014

Accepted 9 January 2015

Available online 21 January 2015

Keywords:Activated carbon

Adsorption

Mercury

Porous structure

Sulfur functional groups

a b s t r a c t

Due to serious public health threats resulting from mercury pollution and its rapid dis-

tribution in our food chain through the contamination of water bodies, stringent regula-

tions have been enacted on mercury-laden wastewater discharge. Activated carbons have

been widely used in the removal of mercuric ions from aqueous effluents. The surface and

textural characteristics of activated carbons are the two decisive factors in their efficiency

in mercury removal from wastewater. Herein, the structural properties and binding affinity

of mercuric ions from effluents have been presented. Also, specific attention has been

directed to the effect of sulfur-containing functional moieties on enhancing the mercuryadsorption. It has been demonstrated that surface area, pore size, pore size distribution

and surface functional groups should collectively be taken into consideration in designing

the optimal mercury removal process. Moreover, the mercury adsorption mechanism has

been addressed using equilibrium adsorption isotherm, thermodynamic and kinetic

studies. Further recommendations have been proposed with the aim of increasing the

mercury removal efficiency using carbon activation processes with lower energy input,

while achieving similar or even higher efficiencies.

2015 Elsevier Ltd. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38

2. Preparation of activated carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39

3. Effect of treatment techniques on mercury removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40

3.1. Physical and chemical activation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40

3.2. Sulfurization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

* Corresponding author. Chemical and Biomolecular Engineering Department, Hong Kong University of Science and Technology, ClearWater Bay Road, Hong Kong. Tel.:852 23588412; fax: 852 23580054.

E-mail address:[email protected](G. McKay).

Available online atwww.sciencedirect.com

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4. Effect of adsorption parameters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 44

4.1. Equilibrium contact time . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 44

4.2. Initial concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45

4.3. pH value . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45

4.4. Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45

4.5. Adsorbent dosage and particle size . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46

5. Mercury affinity to various functional groups . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46

6. Equilibrium adsorption isotherms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 507. Conclusion and future perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 51

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

1. Introduction

Mercury is categorized as an extremely toxic substance whose

health hazards have primarily been associated with inhala-

tion of mercury vapor or ingestion of organic mercury viaaquatic organisms causing mercury poisoning, widely known

as Minamata disease (Harada, 1982). The absorption of this

hazardous substance into the bloodstream, its distribution to

the entire tissues and its bioaccumulation in the receptive

sites result in many well-recognized adverse effects, such as

potent neurotoxicity, blood vessel congestion and kidney

damages (Kidd and Batchelar, 2012).

The presence of mercury in the environment as a result of

naturogenic sources, such as geothermal eruptions and

seismic activities, represents only a small fraction of the total

annual mercury emission. The anthropogenic activities are

the major contributors for mercury pollution in the

ecosystem, among which the fossil fuel combustion plays themost dominant role. An overarching estimate of anthropo-

genic input of 16 elements into the environment by Nriagu

and Pacyna has revealed that mercury releases into the at-

mosphere and aquatic bodies are in the same range

(Ebinghaus et al., 1999; Nriagu and Pacyna, 1988). The

anthropogenic sources of mercurycan be divided intoprimary

and secondary sources. The former involves the mobilization

and release of mercury of geological origin to the environ-

ment, such as mining, industrial processing of ores or fossil

fuel combustion and more specifically coal. The latter is

associated with the direct use of mercury in industrial pro-

cesses, including vinyl chloride monomer production as

catalyst, batteries as cathodes and chlor-alkali production aswell (Pacyna et al., 2010). Improper discharge of effluents and

exhaust emissions from both primary and secondary sources

accounts for the environmental concerns over this toxic

compound. Despite the constraints in the usage of certain

toxic compounds by the Restriction of Hazardous Substances

Directive (RoHS), some substances, including mercury, have

been granted exemption to narrowly-defined applications due

to the technical/scientific impracticality of the substance

prohibition (United Nations Environment Programme, 2010).

Hence, it can be evidently remarked that complete elimina-

tion of mercury from the ecosystem is unrealistic due to the

uncontrollable discharge of mercury by primary anthropo-

genic sources and the inevitable, though limited, utilization of

mercury in industrial processes (secondary anthropogenic

sources).

Accordingly, as a result of the catastrophic impacts of

mercury presence in the ecosystem and its potential fatal

health consequences, many stringent regulations and di-

rectives have been enacted regarding the control of the mer-cury discharge into the environment (Sunderlan and Chmura,

2000). Therefore, many researchers have been engaged with

the removal of mercury from aqueous media. Numerous

techniques employed for this purpose include precipitation

(Blue et al., 2010; Hutchison et al., 2008; Matlock et al., 2001 ),

coagulation (Henneberry et al., 2011; Nanseu-Njiki et al., 2009),

cementation (Ku et al., 2002; Lo and Yu, 1988), ultrafiltration

(Barron-Zambrano et al., 2002; Han et al., 2014; Uludag et al.,

1997), solvent extraction (Huebra et al., 2003; Sevdic et al.,

1980), photocatalysis (Botta et al., 2002; de la Fourniere et al.,

2007; Lopez-Mu~noz et al., 2011), adsorption (Aguado et al.,

2005; Antochshuk and Jaroniec, 2002; Bandaru et al., 2013;

Cui et al., 2013; Di Natale et al., 2011; Li et al., 2011, 2013;Mondal et al., 2013), ion exchange (Anirudhan et al., 2008;

Chiarle et al., 2000; Gash et al., 1998; Lloyd-Jones et al., 2004 )

or a combination of these methods (Barron-Zambrano et al.,

2004; Byrne and Mazyck, 2009). However, the challenges con-

cerning some of these methods include high energy demand

for process operation, large amounts of chemicals used, high

operation and/or capital costs, removal inefficiency and

unselectivity (Ismaiel et al., 2013; Li et al., 2011; Taurozzi et al.,

2013; Wajima and Sugawara, 2011). Among all these removal

techniques, adsorption has been found to be a very promising

method for the removal of heavy metals from wastewater

streams owing to the ease of operation, heavy metal removal

efficiency, high adsorption rate, selective removal and theavailability of a wide range of adsorbent materials (Hadi et al.,

2014a, 2013a, 2013b, 2013c, 2013d; Ismaiel et al., 2013; Nabais

et al., 2006; Ramadan et al., 2010; Xu et al., 2014). Among all

types of adsorbent materials including extracellular bio-

polymers (Inbaraj et al., 2009), cellulosic materials (Takagai

et al., 2011), zeolites and aluminosilicates (Liu et al., 2013;

Somerset et al., 2008), nanomaterials (Bandaru et al., 2013)

and activated carbons (Anoop Krishnan and Anirudhan, 2002;

Di Natale et al., 2011; Ranganathan, 2003), the latter has gained

considerable attention both in research-based studies and

practical industrial applications. Also, some studies have

argued the high cost of activated carbon materials is still an

important problem and have resolved the issue by using

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waste-derived activated carbons to diminish the production

cost of this adsorbent (Kadirvelu et al., 2004; Mohan et al.,

2001; Rao et al., 2009; Zabihi et al., 2010; Zhang et al., 2005).

The current work aims at reviewing the adsorptive removal

of aqueous mercury from effluents by activated carbons. The

focus of this paper is to present a comprehensive overview of

the activated carbon preparation by chemical and physical

methods, their modification techniques, their mercuryremoval efficiencies and the effect of various parameters,

such as pH, initial concentration, activated carbon amount,

adsorbent particle size and temperature, on mercury uptake.

These factors are of the utmost significance, as any change in

these parameters may considerably change the mercury

removal efficiency of an adsorbent. Therefore, a general

knowledge of the effect of these parameters is critical in

designing the appropriate mercury-laden wastewater treat-

ment facilities.

2. Preparation of activated carbon

Activated carbons have long been used as absorbent for the

removal of various pollutants. Textural properties of activated

carbons and functional groups on their surface are two of the

principal characteristics which should be enhanced by certain

modification processes in order to make them exhibit high

pollutant removal efficiencies.

Physical and chemical activation techniques are the most

commonly-used approaches to develop high internal porosity

and desired pore size and also to introduce certain functional

groups onto the adsorbent surface (Hadi et al., 2015).

The physical activation mainly involves two steps,

carbonization and activation. Carbonization includes a heattreatment of a carbonaceous precursor at moderate temper-

atures, mainly below 700 C, in an inert atmosphere to pyro-

lize the precursor material into a low surface area char.

Carbonization leads to the partial evolution of the volatile

matter from the carbonaceous precursor, enriching the pro-

duced char in carbon content and developing the preliminary

branched porous structure. For a carbonization temperature

up to 700 C, the volume of the micropores gradually in-

creases, while furtherincrease in the temperature reduces the

pore volume of the material. The produced char is subse-

quently activated at elevated temperatures, usually above

700 C, under a partial oxidizing atmosphere (primarily steam

or carbon dioxide as gasifying agents) (Hadi et al., 2015). Theaim of the activation stage is to produce a highly porous

structure from the weakly-developed porous char. The acti-

vation process entails the reaction of the oxidizing agent with

tar decomposition products blocking the pores to volatilize

carbon oxide, which, in turn, opens the closed pores, widens

the existing small pores and forms new pores. The resulting

activated product at this stage has high pore volume and

surface area, therefore enhancing its ability for the removal of

pollutants by capturing absorbate molecules in its porous

network (Budinova et al., 2006; Hadi et al., 2014b). The

simplified gasification reactions of the oxidizing agents with

carbon have been illustrated according to the following stoi-

chiometric equations:

C CO2/2CO DH 159 KJ mol1 (R1)

C H2O/CO H2 DH 117 KJ mol1 (R2)

The chemical activation technique is a one-step process in

which the carbonization and activation processes occur

simultaneously. During chemical activation, the source ma-

terial is submerged in a dehydrating compound (such asphosphoric acid, zinc chloride, alkaline hydroxides or alkaline

carbonates) resulting in the diffusion of the chemical reagent

into the particles and its incorporation into the carbon struc-

ture. The resultant slurry is then heatedup to temperatures in

the rangeof 400e600 C under an inert atmosphere. This leads

to the depolymerization of cellulose, hemicellulose or lignin

catalyzed by the chemical reagent, followed by dehydration

and condensation leading to the formation of more aromatic

and reactive products. Also, in some cases, the alkaline metals

are intercalated between the graphene layers while creating

some porosity by the oxidation of carbon into carbon oxides

(Marsh and Reinoso, 2006). This intercalation inhibits the

contraction of the carbonaceous precursor during the heattreatment.

Energy required for chemical activation process is consid-

erably lower than that of physical activation method. Signifi-

cantly lower activation temperatures, shorter reaction time

and employment of a single-step treatment in the case of

chemical activation process account for this low energy con-

sumption. Moreover, it has been reported that, in most cases,

both carbon yield and specific surface area for the materials

produced by chemical activation are higher than those of the

physical activation method. These two reasons account for

the more extensive application of the chemical activation

method compared with the physical one (Macia-Agulloet al.,

2004).Although, commercial activated carbons have been widely

used in various industries to remove mercury from waste-

water streams, their use has been recently challenged by their

high cost (Di Natale et al., 2006). Hence, much research has

recently been devoted to the exploration of alternative waste-

based carbon sources as precursor for activated carbon pro-

duction. Toles et al. found that it is highly profitable to pro-

duce almond shell-based granular activated carbon (GAC) at

US$20/ton compared with two comparable commercial GACs

at about US$3.3/kg using it as a basis for economic feasibility

study (Toles et al., 2000). Considering such a great financial

incentive, numerous researches have been conducted on

certain inexpensive waste precursors to produce cost-effective activated carbons with high pollutant removal effi-

ciencies. Among certain precursor materials to be used in

activated carbon production, high mercury adsorption ca-

pacities have been reported using coirpith (Namasivayam and

Kadirvelu, 1999), furfural (Yardim et al., 2003), walnut shell

(Zabihi et al., 2010), agricultural solid waste (Kadirvelu et al.,

2003), silk cotton hull (Roberts and Rowland, 1973), sago

waste(Kadirvelu et al., 2004), coconut tree sawdust, maize cob

and banana pith (Kadirvelu et al., 2003) while moderate to low

adsorption capacities have been observed for activated car-

bons derived from pazzolana and yellow tuff (Di Natale et al.,

2006), Fullers earth (Oubagaranadin et al., 2007), Ceiba-

pentandra hulls, Phaseolus aureus hulls and Cicerarietinum

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(Rao et al., 2009). Hence, it can be inferred that the type of

precursor is an influential factor in determining the adsorp-

tion efficiency of the activated carbons.

3. Effect of treatment techniques on mercuryremoval

3.1. Physical and chemical activation

The nature of precursor materials and their activation

methods directly influences the surface area, pore size and

functional groups on the surface of the prepared activated

carbons which can, in turn, have an influence on the mercury

removal efficiency.Tables 1 and 2show the effect of the pre-

cursor type and activation conditions on mercury adsorption

capacities of the prepared activated carbons.

A suitable activation procedure leads to the production of a

high surface area adsorbent material and thus, reasonably

high mercury removal efficiency. Budinova et al. have

demonstrated the effect of steam activation and air oxidationon the specific surface area of a carbonaceous biomass waste.

Sample activated with steam showed significantly higher

surface area than the one oxidized in air atmosphere. This

resulted in a higher mercury adsorption capacity for the

steam-activated adsorbent as compared with the air-oxidized

adsorbent (Budinova et al., 2003). A more detailed study of the

activated carbon preparation conditions under various at-

mospheres has been reported elsewhere (Budinova et al.,

2006). Water vapor has been shown to penetrate into the

carbon structure, react with the carbon at the internal surface

of the carbon and extract the carbon from the pore walls,

resulting in the enlargement of the existing pores and creation

of new pores (as shown in the reaction R2) and thereby anincrease in the adsorption capacity of the produced adsorbent.

It has been demonstrated that the pore size and pore size

distribution of activated carbons can be manipulated by using

different activating atmospheres. Molina-Sabio et al. have

shown that steam-activated carbons exhibit larger mesopores

compared to carbons activated under a CO2 atmosphere

(Molina-Sabio et al., 1996). This will, undoubtedly, influence

the adsorption behavior of these adsorbents towards various

adsorbates. However, no comprehensive study has been

conducted on the effect of activation atmosphere, which, in

turn, results in a change in the surface functionalities of the

adsorbents, to maximize mercury removal from effluents.

In addition to the activation conditions, both surface areaand functional groups are predominantly affected by the na-

ture of the precursor material used. As shown in Table 1,

Ekinci et al. have found that, under similar activation condi-

tions, the surface areas range from 460 to 720 m2:g1 for coal-

based activated carbons and 1000 and 1110 m2:g1 for apricot

stone and furfural-based activated carbons, respectively

(Ekinci et al., 2002). The difference in the textural properties of

these produced activated carbons is reflected in their mercury

adsorption capacities. Furthermore, Rao et al. compared the

mercury removal efficiencies of activated carbons prepared

from three different carbonaceous precursors under similar

activation conditions and observed huge differences in their

surface areas and adsorption capacities (Rao et al., 2009).Table1e

Effectoftheprecursortype

andphysicalactivationconditionson

themercuryremovalefficienciesof

theactivatedcarbons.

Carbonizationconditions

Activationconditions

SBET

qm

Vt

Vmicro

Vmeso

Dp

Ref.

T(C)

T(h)

HRa

(Cmin

1)

Atm.

T(C)

T(h)

HR(Cmin

1)

Atm

.

(m2

g

1)

(mgg

1)

(cm

3

g

1)

(cm

3

g

1)

(cm

3

g

1)

(nm)

200

2

750

2

20

H2O

521

25.8

8

(Raoeta

l.,

2009)

325

23.6

6

280

22.8

8

600

1

3

N2

761

0.6

81

0.2

97

0.3

21

2.3

2

(Budinovaetal.,

2006)

750

2

10

H2O

1100

174

0.8

9

0.4

25

0.1

1

(Budinovaetal.,

2003)

400

H2O

480

134

0.6

4

0.2

1

0.0

85

750

2

10

H2O

600

92

(Ekincie

tal.,

2002)

460

56

720

105

520

37

1000

153

1100

174

500e950

10

N2

(Inbaraj

andSulochana,

2006)

890

9

5

N2

&

CO2

1528

0.6

4

(Macia-A

gulloetal.,

2004)

890

22.5

5

N2

&

CO2

2487

0.8

6

800

1

5

N2

900

5

15

N2

&

CO2

997

0.3

7

(Nabaisetal.,

2006)

900

5

N2

702

140

2.7

22.5

(Choian

dJang,

2008)

850

1

N2

&

H2O

870

4.9

3

0.4

28

0.3

28

0.0

04

1.9

67

(Luetal.,

2014)

a

HRdenotesheatingrate.

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Table 2 e Effect of the precursor type and chemical activation conditions on the mercury removal efficiencies of the activated carbo

Carbonization/Activationconditions

Chemical treatment SBET qm Vt Vmicro

T (C) T (h) HRa

(C min1)Atm. Chem. Conc. Imp. Rb Imp.tc (h) (m2 g1) (mg g1) (cm3 g1) (cm3 g1)

600/600 1/1 3 H2O/N2 1290 0.939 0.471

700 2 H2O 1360 160 1.026 0.485

600 10 N2 H2SO4 Conc. 1:20 1050 154 0.827 0.38

H2SO4 Conc. 10

10

10

10

10

200/450 1 Air H2O2 629 18.1

H2SO4 & (NH4)2S2O8 50% 1:200 0.5 625 55.6

600/800 1 N2/H2O H2SO4 Conc. 1100 174 0.425

H2SO4 & (NH4)2S2O8 Conc. 3: 5 & 1:33 12 592 154

400/700 0.5/1 NaOH 20% (W/V) 2 379.4 52.7

N2 ZnCl2 98% 1:0.5 780 151.5

1:1 803 100.9

H2SO4 Conc. 1:1.8 24 208.1 109.9

K2CO3 33-75% 1:1 2 1260 129 0.492

KOH 1:2 1 1090 0.49

1:4 1 1635 0.78

1:6 1 2225 0.9

1:8 1 2420 0.94

NaOH 1:2 1 1130 0.51

1:4 1 2000 0.81

1:6 1 2541 0.96

1:8 1 3033 1.02

800/900 1/5 5/15 N2 & CO2 S 1:3 2 848 410 0.33

H2S 1 905 450 0.33

900 5 N2 Pyrrole 1:3 413 541 1.52

43:100 346 682 1.04

11:20 158 441 0.54

900 N2 922 301 0.87 0.37 900 H2S 785 351 0.78 0.31

200 SO2, H2S & N2 764 351 0.72 0.3

200/400 2/1 10 H2O, SO2 & H2S 500.5 227.3 0.43

H2O & SO2 506.5 222.2 0.48 0.04

H2O,&H2S 530.2 217.4 0.48 0.06

H2O 536.5 208.3 0.52 0.13

800 0.5 N2 K2S 1:3.6 30.0 235.7

850 0.5 30.0 243.9
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Table 2 e (continued)

Carbonization/Activationconditions

Chemical treatment SBET qm Vt Vmicro

T (C) T (h) HRa

(C min1)Atm. Chem. Conc. Imp. Rb Imp.tc (h) (m2 g1) (mg g1) (cm3 g1) (cm3 g1)

900 0.5 28.0 254.4

1000 0.5 30.0 224.1

H2O2 3:10 168 836.0 100 0.23

H2 & CS2 778.0 170 0.18H2O2 & He 950.0 110

H2O2, He & CS2 776.0 175 0.22

He 880.0 110

2 50 N2 S 2:1 645.0 467

2:1 594.0 507

2:1

2:1 1070.0 827

2:1 1045.0 427

700 4 HCl 1 N 4 311.0 9.32

H2SO4 13 M 1:5 0.33 66.0 303.03

13 M 1:5 0.33 384

H2SO4 12 M 1:9 0.5 19.0 385

12 M 1:9 0.5 526

HNO3, SOCl2 &

C2H4(NH2)2

5 M, 5%

&0.05 M

3: 20 & 3:8 7 120

H2O2 30% 1:2 829.0 5 0.413 0.318

2:5 825.0 4.2 0.408 0.315

600 10 N2 SO2 1:25 0.5 720.9 122.8

600 1:25 1 772.5 129.8

600 1:25 2 776.4 130.5

600 1:25 3 790.7 131.6

600 1:50 1 773.7 128.2

600 2:25 1 518.5 114.8

600 1:25 1 757.2 125.7

500 1:25 1 764.1 135.9

700 1:25 1 751.3 184.2

800 1:25 1 1087.0 196.8

900 1:25 1 1057.0 207.8

TOMATS 1:0.3 48 107.7 83.3

HNO3 0.01 M 72 136.7 315.8 HSCH2COOH,

(CH3CO)2O & H2SO4

13 193.7 694.9

a HR denotes heating rate.b Imp. R denotes impregnation ratio.c Imp. t denotes impregnation time.
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Nonetheless, although specific surface area is one of the fac-

tors affecting the mercury adsorption, it is not the only

influential parameter. Zabihi et al. prepared two activated

carbons with surface areas of 780 and 803 m2:g1, while their

mercury adsorption capacities were 151 and 101 m2:g1,

respectively (Zabihi et al., 2009). Also, the surface areas of the

activated carbons prepared by Roa et al. ranged from 280 to

521 m2:g1, whereas their mercury adsorption capacitiesunder similar adsorption conditions did not exhibit a signifi-

cant difference (Rao et al., 2009). Wang et al. demonstrated

that an activated carbon with a surface area of 1896 m2:g1

had a much smaller adsorption capacity than an activated

carbon with a surface area of 1070 m2:g1 (160 vs 827 m2:g1,

respectively) (Wang et al., 2009). Since the size of the solvated

mercury is much larger than the nitrogen molecules, the pore

size and pore size distribution of the produced activated car-

bons besides their specific surface area is of significance.

Hence, it can be concluded that other factors, such as surface

functional groups, pore size and pore size distribution also

have considerable effects besides surface area in mercury

removal. Nevertheless, few studies have simultaneouslyinvestigated the effects of pore structure and functional

groups on mercury removal.

Considering the surface areas listed inTables 1 and 2, it is

noticeable that the chemical activation technique has greater

performance in pore formation at appropriate chemical re-

agent to adsorbent ratio. Comparing the surface areas ob-

tained by chemical and physical activation of coal tar pitch

carbon fibers, it is obvious that although high surface area

activated carbon (2487 m2:g1) can be obtained by physical

activation, it is not an economical option in terms of energy

consumption due to prolonged activation time at a high

temperature (22 h at 890 C) causing an excessive carbon burn-

off (94%). On the contrary, chemical activation of this materialusing an alkaline solution as activating reagent(KOH or NaOH)

at an impregnation ratio of 6:1 (w/w) yields activated carbon

with similar pore width and higher surface area. Other ad-

vantages of the chemically activated carbon are higher prod-

uct yield (60% and 27%, respectively), lower activation

temperature (750 C) and shorter activation time (1 h). The

highestsurface area (3033 m2:g1) can be obtainedusing NaOH

as activating agent at an impregnation ratio of 8:1 (w/w)

(Macia-Agulloet al., 2004).

In order to enhance mercury adsorption, several authors

have studied the combination of chemical and physical acti-

vation techniques. Budinova et al. confirmed that when the

H3PO4-impregnated carbonaceous sample was treated understeam atmosphere, both the surface area and iodine number

were considerably higher than the samples pyrolyzed under

nitrogen atmosphere. They also revealed that the concentra-

tion of the chemical reagent used for impregnation has a

significant effect on pore development. When the concentra-

tion of phosphoric acid was increased from 20% to 50%, the

mercury adsorption capacity of the activated carbon was

enhanced considerably (Budinova et al., 2006; Yardim et al.,

2003). Although no in-depth reason was provided for this

phenomenon, we believe that increasing the acid concentra-

tion increases the rate of the pyrolytic decomposition of the

precursor and enhances the density of the cross-linked

structure due to the catalytic effect of the phosphoric acid

and thus results in the modification of the textural properties

of the activated carbon.

3.2. Sulfurization

The binding ability of the carbonaceous compound surfaces

with sulfur-containing functional groups is well-recognized

(Cai and Jia, 2010; Hsi et al., 2001; Korpiel and Vidic, 1997;Vitolo and Pini, 1999; Wang et al., 2009). It has been widely

verified that sulfurization of activated carbons results in

enhanced adsorption capacity and selectivity towards mer-

cury. Therefore, the application of sulfur-functionalized acti-

vated carbons in the removal of mercury has become a

common practice. A variety of techniques have been

employed to immobilize sulfur on the surface of adsorbents,

including treatment with carbon disulfide (CS2), sodium sul-

fide (Na2S), hydrogen sulfide (H2S), sulfur dioxide (SO2) or

sulfur powder, with the aim of increasing their mercury up-

takes (Feng et al., 2006a; Fouladi Tajar et al., 2009; Mohan et al.,

2001; Vitolo and Pini, 1999; Wajima et al., 2009; Zhang et al.,

2003). Nabais et al. have applied two modification tech-niques, namely impregnation with elemental sulfur and using

hydrogen sulfide gas as modifying agent. Both of these ad-

sorbents exhibited higher adsorption capacities than the un-

sulfurized activated carbons (Nabais et al., 2006). Asasian

et al. found a 50% increase in mercury adsorption capacity by

sulfurizing the activated carbon with 4% sulfur dioxide gas

stream (Asasian et al., 2014). Wang et al. have studied the ef-

fect of the impregnation of activated carbon with elemental

sulfur. Elemental sulfur not only can directly deposit on the

adsorbent surface and interact with mercury, but also can

react with the adsorbent surface and lead to the formation of

new functional groups to enhance mercury adsorption (Wang

et al., 2009). They found that the mercury adsorption capacityof the unmodified and modified activated carbons were

190 mg:g1 and 820 mg:g1, respectively. The chemical reac-

tion between elemental sulfur and the surface of the adsor-

bent leads to the formation of disulfide, thiophene, sulfoxide

and sulfone groups that have more affinity to mercuric ions

and can enhance the overall mercury adsorption capacity and

selectivity (Cai and Jia, 2010; Wang et al., 2009). Mohan et al.

have observed a doubled mercury uptake after soaking the

adsorbent in carbon disulfide (Mohan et al., 2001). Despite the

well-acknowledged sulfur effect on mercury sequestration,

the processof sulfurbinding onto the activated carbon surface

and consequently, the mechanism of mercury adsorption

onto sulfur-containing moieties have not been satisfactorilyexploited and established. An in-depth understanding of the

activation and mercury adsorption mechanisms will assist in

designing a proper activation/functionalization procedure in

order to achieve high mercury abatement. Pillay et al. have

investigated the activation and adsorption mechanisms using

Raman spectroscopy as an analytical tool to monitor the

changes in the functional groups before and after the

adsorption process (Pillay et al., 2013). They verified the

presence of S]CeS bonds (at 475 cm1, 495 cm1 and

503 cm1) associated with thiol and thioester groups after

treating the virgin carbon nanotube with phosphorus penta-

sulfide. Subsequent mercury adsorption revealed significantly

diminished intensities of the bands corresponding to the thiol

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groups and appearance of new band assigned to Hg(SH)2and

Hg2(SH)2bonds on the used adsorbent material. In addition to

interaction with thiol moieties, weak chemisorption between

the mercury and hydroxyl groups were also noticed by the

reduction of hydroxyl peak intensity at 620 cm1 and the

formation of new peak at 550 cm1 corresponding to HgeO

bond, indicating strong binding of mercury ions to the thiol

groups rather than oxygen functional groups. Furthermore,Nabais et al. have identified the presence of SeS, C]S, CeS

and SeO bonds by FT-IR analysis after sulfurization, but

changes in the intensities of these peaks after the mercury

adsorption have not been provided (Nabais et al., 2006).

Due to the affinity of sulfur functional groups with mer-

cury, higher amounts of sulfur moieties will theoretically be

advantageous in mercury removal. Several researchers have

reported the direct linear relationship between mercury up-

take and sulfur content (Cai and Jia, 2010; Pillay et al., 2013).

However, Wang et al. have ruled out this hypothesis and have

demonstrated that the activated carbon with lower sulfur

amount (22%) on its surface had a higher mercury uptake than

the adsorbent with higher sulfur content (34%) (Wang et al.,2009), but no justification has been provided in the paper.

However, the aggregation of sulfur within the large pores,

rather than the uniform distribution of sulfur on the activated

carbon, may account for this phenomenon. Also, Nabais et al.

have compared several sulfur introduction methods and

identified that mixing of activated carbon fibre (ACF) with

solid sulfur at a ratio of 1:3 (w/w), followed by treatment at

600e800 C resulted in the production of an adsorbent with

higher sulfur content compared with the introduction of sul-

fur to ACF via gas stream H2S. This may be reasonable due to

the melting, recrystallization and deposition of the solid sul-

furon the activated carbon surface at such high temperatures.

However, subsequent mercury adsorption tests showed thathigher mercury uptake was obtained by the latter method

which elucidated the importance of the type of sulfur func-

tional groups on the carbon surface besides its quantity

(Nabais et al., 2006). This may also occur due to the aggrega-

tion of the sulfur on the activated carbon when solid sulfur is

used as the surface modifying agent which diminishes the

effect of sulfur functional groups in mercury removal.

Furthermore, despite similar sulfur contents of K2S-impreg-

nated coal samples were prepared at three distinct tempera-

tures (800e1000 C), whereas the activated carbon sample

prepared at 900 C exhibited the highest and fastest adsorp-

tion (Wajima and Sugawara, 2011). This indicates that the

sulfur content on the adsorbent surface, type of sulfur func-tional groups and porous structure of the activated carbons

collectively influence their mercury adsorption efficiency.

In addition to higher capacity, exceptional affinity between

mercury and sulfur has been demonstrated in a multi-

component system of mercury, cadmium and lead where

highly-selective adsorption towards mercury was achieved

(Gomez-Serrano et al., 1998). The superior adsorption of

mercury on sulfur-grafted adsorbent is believed to originate

from the Pearson acid-base concept in which the hard acids

prefer to coordinate with hard bases and soft acids react in a

higher rate with soft bases. Accordingly, the soft acid mercury

species in the solution, such as HgCl2, (HgCl2)2, Hg(OH)2 and

HgOHCl tend to predominantly react with sulfur groups (soft

bases) on the adsorbent surface (Cai and Jia, 2010). This phe-

nomenon has also been confirmed by comparing the inter-

action of HgX2(where X is a halide) with C4H8O and C4H8S. It

has been reported that HgX2interacts weakly with C4H8O, but

much stronger with C4H8S (Farhangi and Graddon, 1973;

Fisher and Drago, 1975). Vazquezet al. have related the

higher affinity towards mercury in a multi-component system

of cadmium, zinc and mercury to the higher electronegativityof mercury (Vazquez et al., 2002).

4. Effect of adsorption parameters

4.1. Equilibrium contact time

Equilibrium contact time is the period of time required for the

adsorption and desorption processes to reach equilibrium.

When the equilibrium is reached, the amount of adsorption

from the solution to the adsorbent surface equals the amount

of desorption from the adsorbent surface to the solution and

no further increase in the uptake occurs. The adsorptionprocess involves several steps including mass transfer from

bulk fluid phase to the particle surface across the boundary

layer, adsorption on the surface of the adsorbent and diffusion

within the pores (Wang et al., 2011). Depending on which of

these steps is the rate determining stepand also depending on

the boundary layer thickness and diffusion rate, the contact

time required to reach equilibrium will be different.

Adsorption of mercury has been shown to comply with a

general trend in which the mercury uptake rate is very fast at

the beginning because of the large number of vacant func-

tional group sites on the surface of activated carbon available

for the mercury ions.As the sitesare occupied in the course of

time, the uptake rate is gradually slowed down until a plateauis reached upon equilibrium (Zabihi et al., 2009). Shorter

contact time required by the adsorbent to reach equilibrium is

economically more favorable in industry.

Kadirvelu et al. have suggested that the rate of adsorption

depends on several factors such as the type of precursor used

for adsorbent production, pore size and pore size distribution

and concentration of functional groups (Kadirvelu et al., 2004).

Namasivayam and Periasamyhave reported that the activated

carbon from bicarbonate-treated peanut hull (BPHC) exhibited

7 times higher adsorption rate compared with commercial

activated carbon. They ascribed this high adsorption rate to

higher porosity and ion exchange ability of BPHC resulting in

less adsorption time required to acquire a certain mercuryremoval percentage and thus more cost effectiveness

(Namasivayam and Periasamy, 1993). Also, rapid mercury

adsorption of less than 20 min to reach equilibrium was re-

ported for activated carbons prepared from antibiotic waste

and rice husk ash as precursors (Budinova et al., 2008; Feng

et al., 2004). It can be hypothesized that as the pore size in-

creases up to a certain extent, the diffusion path is reduced

and the adsorption rate increases. Also, higher concentration

of adsorption sites will increase the probability of contact

between mercury molecules and functional moieties and

therefore increase the uptake rate. Hence, it is believed that

more abundant adsorption sites with an optimum pore size

for mercury will increase the rate of mercuryadsorption. More

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studies are necessary to be conducted to prove these hy-

potheses regarding the relationship between mercury

adsorption rate and textural and surface properties of the

adsorbents.

It has been further demonstrated that as the initial con-

centration of mercury increases, the Lagergren rate constant

decreases and thus, longer time is required to achieve equi-

librium (Namasivayam and Kadirvelu, 1999; Namasivayamand Periasamy, 1993). This could be due to the saturation of

sites presenton the exterior of adsorbent surface by adsorbate

at an initial stage of adsorption. Further adsorption can only

occur by the diffusion of the mercury ions into the pores and

adsorption in the interior surface of the pores which requires

relatively longer contact time (Hameed, 2007). Hence, in

modeling the mercury adsorption kinetics, a combination of

pseudo-type models with diffusion models is worthy of

consideration to elucidate the adsorption mechanism. How-

ever, for certain applications, these models have been shown

to be inconclusive (Plazinski et al., 2009).

4.2. Initial concentration

In general, the mercury adsorption experiments display a

direct relationship between the metal uptake and initial con-

centration of the metal ions present in the solution up to a

certain limiting initial concentration and inverse relationship

between the removal percentage and initial metal concen-

tration. An apparent distinction has to be drawn between

removal percent and adsorption capacity. The former term

does not reflect the efficiency of the material in mercury

removal at various initial concentrations and adsorbent dos-

ages, whereas the latter takes into account the adsorbent

dosage and reveals the genuine mercury adsorption efficiency

of the material at different initial concentrations. Several au-thors have reported complete mercury removal (Mohan et al.,

2001; Rao et al., 2009; Wahi et al., 2009). But when the initial

concentration and adsorbent amount are taken into consid-

eration, the adsorption capacity is found very small in some

cases. Therefore, reporting mercury removal percentage is

highly discouraged due to misleading results (Hadi et al.,

2015). As the initial concentration of mercury in the solution

increases, the percent removal of the adsorbate decreases

because of the presence of more mercury ions and limited

adsorption sites on the adsorbent materials. On the other

hand, at low mercury concentrations, the adsorption capacity

of the adsorbent material is low, while it increases by

increasing the initial concentration. This has been related tothe fact that at low mercury concentrations, the adsorption

sites are not completely occupied (Budinova et al., 2008, 2003;

Zabihi et al., 2009), whereas increasing the initial concentra-

tion of mercury results in higher collision probability between

the adsorbate molecules and adsorbent active sites, higher

occupation of active sites and thus higher adsorption capacity

(Zabihi et al., 2010). When the initial mercury concentration is

sufficiently increased, the adsorption capacity reaches a

plateau and does not increase anymore by increasing the

initial mercury concentration. This has been attributed to the

full occupation of the active sites on the surface of the

adsorbent at a certain initial concentration above which no

more adsorption enhancement can be achieved. Inbaraj and

Sulochana have observed a similar trend and have suggested

that this effect is caused by an increase in the driving force

offered by concentration gradient at high mercury concen-

trations (Inbaraj and Sulochana, 2006).

4.3. pH value

Adsorption of mercury is a highly pH dependent process. As

the pH value of the solution increases, more mercury uptake

occurs. The increased adsorption of mercury ion has been

shown to be related to the species of mercury present in the

solution at various pH values and their solubility. Higher pH

values of the solution results in the presence of more soluble

mercuric species which, in turn, promotes the effective con-

tact between the adsorbate molecules and the adsorbent

materials thus enhancing the possibility of the mercury up-

take by the porous adsorbent particles (Adams, 1991; Lopes

et al., 2010; Namasivayam and Periasamy, 1993). Moreover,

lower pH values increase the solubility of the mercuric ions

and thus their subsequent desorption from the activated

carbon surface into the solution. Therefore, the relative

attraction between the adsorbent and adsorbate is lower than

between the adsorbate and the solvent phase at lower pH

values, leading to the lower adsorption of the mercuric ions.

Solution acidity also plays an important role in the ioni-

zation of the functional groups on the adsorbent surface. In

acidic environment, high concentration of hydronium ion

(H3O) in the solution drives the equilibrium ionization reac-

tion (reaction R4) to the left and prevents the formation of

ionized functional groups, thereby hampering the ion ex-

change reaction between metal ions and adsorbent surface

functional groups (reaction R5). When the pH level of the so-

lution increases to above 4, the hydronium ion concentration

in the solution decreases. This shifts the equilibrium reaction

R4 to the right resulting in the availability of more ionized

functional groups for ion exchange and therefore an increase

in the metal uptake (Eligwe et al., 1999).

Adsorbent COOH4Adsorbent COO Haq (R4)

Adsorbent COO Mnaq4Adsorbent COOM (R5)

It is noteworthy that, the surface properties of adsorbent

can significantly affect the adsorption of mercury. The

adsorbent can be positively or negatively charged depending

on its point of zero charge (PZC). When the pH of the medium

is lower than the PZC, the adsorbent surface becomes posi-tively charged leading to electrostatic repulsion of the mer-

cury ions and the adsorbent surface and reduction in mercury

adsorption (Budinova et al., 2008; Rao et al., 2009).

4.4. Temperature

Many researchers have shown that increasing the tempera-

ture results in higher mercury uptake due to the endothermic

nature of this process. Inbaraj and Sulochana have used the

thermodynamic parameters to study the effect of temperature

on the mercury adsorption behavior of fruit shell-based acti-

vatedcarbon and found a decrease in Gibbs free energy,DG, as

well as a positive enthalpy value, DH, by raising the

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temperature which revealed that the adsorption process is

endothermic (Inbaraj and Sulochana, 2006). Also, Giles et al.

believe that high temperatures increase the mobility of the

mercuric ion and widen the pore on the sorbent surface

leading to enhanced intra-particle diffusion rate (Giles et al.,

1974). However, no justification has been provided regarding

the pore widening phenomenon. The effect could not be due

to physical widening, but apparent widening because of anincrease in compressibility of the particles with increasing

temperature. Since porous structure of carbonaceous mate-

rials usually forms at very high temperatures, it is particularly

implausible to alter the pore sizes at adsorption temperatures

as low as 20e80 C.

On the contrary,Mohan et al.have reported the exothermic

nature of the mercury adsorption using activated carbon

derived from fertilizer waste. This is confirmed with the

negative enthalpy value, DH, and an increase in Gibbs free

energy DG by an increase in temperature. They have related

this behavior to the physical adsorption mechanism of mer-

cury by the adsorbent material. Physical adsorption caused by

the van der Waals forces between the adsorbent surface andthe adsorbate molecules is typically favored at low tempera-

tures. This explains the higher adsorption capacity of the

fertilizer-based adsorbent for mercury at low temperatures

(Mohan et al., 2001).

4.5. Adsorbent dosage and particle size

Well-documented researches have proven that an increase in

the dosage of adsorbent at a constant pH and adsorbate con-

centration has positive effect on the removal of pollutants

from wastewater (Gupta et al., 2003; Namasivayam et al.,

2001). Although many researchers have reported that the

mercury removal percent increases as the adsorbent dosage isincreased, as discussed in preceding sections, removal per-

centage is an entirely relative term changing by initial con-

centration of mercury and adsorbent dosage and thus it is not

appropriate to evaluate the efficiency of an adsorbent using

this parameter. Percentage mercury removal increased from

40% to nearly 100% when the C. pentandrahull adsorbent dose

increased from 25 to 200 mg (equal to 0.5 g:L1 and 4 g:L1,

respectively) (Rao et al., 2009). Increasing the dose of Indian

almond fruit shell from 0.05 to 5 g:L1 also led to a maximum

mercury removal of 99.5% (Inbaraj and Sulochana, 2006).

Similar mercury adsorption trends have been reported using

activated carbon from sago waste and commercial activated

carbon (Kadirvelu et al., 2004). Typically, an increase in theadsorbent dosage results in the availability of higher surface

area and larger number of functional groups for ion exchange

in the system and leads to more chemisorption and/or phys-

isorption as well as higher rate of adsorbate removal (Wahi

et al., 2009). It is suggested that more tangible adsorption ca-

pacities should be reported in this context instead of simply

quoting the percent removal.

On the other hand, although the removal percentage in-

creases by increasing adsorbent load, the mercury adsorption

capacity of the adsorbent has been shown to steadily

decrease. This has been related to the decrease in the avail-

ability of mercury ions in aqueous phase per adsorbent site

and unsaturation of the adsorbent surface active sites. Rao

et al. used three types of adsorbents for mercury removal and

all of the adsorbents exhibited a decrease in the adsorption

capacity and an increase in the mercury removal percentage

by increasing the adsorbent dosage (Rao et al., 2009).

In addition, particle size also plays an important role in

altering the rate and capacity of mercury adsorption. It has

been demonstrated that when the size of the adsorbent par-

ticles decreased from 1.25e2.5mmto0.21e1 mm,the mercuryadsorption capacity showed a two-fold increase (430 mg:g1

versus 815 mg:g1, respectively) (Mckay et al., 1989; Peniche-

Covas et al., 1992). Similarly, Kadirvelu et al. have also

demonstrated that a stepwise decrease of activated carbon

particle size produced from sago waste (750e500 mm,

500e250 mm and 250e125 mm) resulted in an increase in

mercury adsorption (85%, 90% and 93% removal, respectively)

(Kadirvelu et al., 2004). Similar results have also been reported

by Mohan et al. (Mohan et al., 2001) and Feng et al. (Feng et al.,

2004). It has been shown that reducing the adsorbent particle

size increases the effective surface area and enhances the

availability of adsorption sites (Kara et al., 2007). Also, the

diffusion path becomes shorter and the adsorbate moleculescan more easily penetrate into the internal pores of the

adsorbent (Gupta et al., 2011).

5. Mercury affinity to various functionalgroups

The adsorption of adsorbate by activated carbon can be cate-

gorized into chemical and physical adsorption. Briefly, phys-

ical adsorption is mediated by the weak van der Waal

interaction between the adsorbate and adsorbent, while

chemical adsorption is governed by the bonding between the

functional groups on the adsorbent surface and adsorbate.Weak van der Waal interaction have beenproven inefficient in

promoting mercury adsorption, however surface functional

groups, specially the oxygen containing groups of the adsor-

bent, exhibit a key role on the adsorption of mercury (Sun

et al., 2011). The behavior of enhanced mercury adsorption

by oxygen-containing functional groups has been explained

by the Lewis characteristic of Hg (II) which can be bonded to

the basic functional groups of the adsorbent surface (Nabais

et al., 2006). In the aqueous medium, the oxygen-containing

functional groups on the surface of the adsorbent tend to

lose their protons and become ionized, thus leading to un-

balanced charge on the adsorbent surface where ion exchange

with the mercuric ion can occur (Sun et al., 2011). This ac-counts for the critical effect of pH level of the mercury-laden

solution on the adsorption of mercuric ions. As discussed in

the preceding sections, changing the pH level significantly

changes the adsorbent surface charge, thus resulting in a

considerable difference in the adsorption capacity of the

adsorbent. Moreover, electron lone pairs on nitrogen-

containing functional groups can interact with mercury ions

and assist in their removal (Zhu et al., 2009).

Although functional groups of the activated carbons have

long been consideredto be crucial in chemisorption, the effect

of oxygen-containing functional group quantity on mercury

adsorption capacity and rate has not been comprehensively

examined. It is also noteworthy that despite an inverse

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relationship between the percentage of oxygen functional

groups and the total surface area of the adsorbent material,

both of which are regarded positive factors for mercury

adsorption capacity, no trade-off graph between these two

crucial parameters has been provided to optimize the effi-

ciency of adsorption.

As functional groups largely determine the surface prop-

erties and thus the intensity of ion exchange, manipulation ofthe surface functional group have been of great interest. Sul-

fur group has been widely reported to promote mercury

adsorption. Rao et al. have observed an increase in the mer-

cury adsorption capacity of activated carbon with the intro-

duction of sulfur groups on the activated carbon surface and

ascribed the higher removal efficiency of the sulfur-

containing activated carbon to the interaction of various

Hg(II) species, such as HgCl2, (HgCl2)2, Hg(OH)2 and HgOHCl,

with surface sulfur groups. The following redox reaction has

been proposed as the adsorption mechanism of the activated

carbon for mercury (Rao et al., 2009).

2Hg2

SO23 2OH

4Hg

22 SO

24 H2O (R6)

This reaction is in good agreement with the effect of pH

value on the adsorption capacity, where increasing the pH

level of the solution increases the hydroxide ion content,

driving the reaction to theright side, and thus leads to a higher

mercury adsorption capacity.

Enhancement in the removal of mercury has also been

carried out by grafting thiol group onto the surface of the

activated coke, where the adsorption capacity has been

increased from 315.8 mg:g1 for the unmodified material to

694.9 mg:g1 for the modified material (Li et al., 2013). Anoop

Krishnan and Anirudhan have verified the effect of sulfur

modification by H2S and SO2 on the adsorption capacity of

mercury (Anoop Krishnan and Anirudhan, 2002). They

observed that irrespective of the type of modifying agent

(either H2S or SO2), the mercury adsorption capacity of the

activated carbon increases. This can be due to the similar

types of sulfur functional groups doped on the adsorbent

surface by gas surface modification. These results are

different from the gas-phase adsorption of mercury which canbe related to the different mechanism in gas-phase and

aqueous-phase mercury adsorption (Feng et al., 2006b).

Studies concerning the effect of activation parameters on

the type and quantity of the functional groups are listed in

Table 3. Toles et al. have identified that the type of precursor

has minor effect on the functional groups of the produced

activated carbons and implicated the importance of activation

temperature in the formation of the functional groups (Toles

et al., 1999). The oxygen-containing functional groups can be

formed by exposing the carbonaceous precursor to oxygen at

temperatures between 200 and 700 C (Bansal et al., 1988).

Also, more carbonyl groups can be formed by oxidizing the

activated carbon at 400 C, but subsequent oxidization of theactivated carbon destroys the carbonyl group and produces

more phenol, lactones and carboxylic acid group (Toles et al.,

1999). Furthermore, it has been observed that the carboxylic

groups begin to decompose at 200e500 C and all the acidsites

are destroyed at 700 C(Budinova et al., 2008). Although such

manipulation can be carried out on the surface functional

groups of activatedcarbons, it can be criticized that there is no

comprehensive study to compare the effects of various func-

tional groups on mercury removal.

The type of the activating agent has also been considered

effective in the manipulation of the surface functional groups

Table 3 e Acid-base neutralization capacity (meq/g) of the activated carbon adsorbents.

Precursor material Name Base uptake Acid uptake Reference

NaHCO3 Na2CO3 NaOH EtONa HCL

Mengen 0.092 0.120 0.184 1.900 1.120 (Ekinci et al., 2002)

Seyitomer e 0.120 0.250 2.040 2.670

Some 0.100 0.110 0.183 1.570 1.120

Bulluca e 0.110 0.320 1.900 2.010

Apricot Stones 0.130 0.210 0.360 1.350 0.842

Furfural 0.120 0.160 0.230 1.500 0.600

Furfural Carbon A 0.120 0.160 0.230 1.500 0.600 (Budinova et al., 2003)

Mixture of steam pyrolysis

tar and furfural (30:70)

Carbon B 0.030 0.080 0.250 1.300 0.560

Air oxidized Furfural Carbon C 1.900 1.930 3.660 6.340 e

Woody biomass birch N600-1 0.744 0.126 0.480 2.234 0.083 (Budinova et al., 2006)

NS600-1 0.124 0.034 0.572 2.530 1.100

S700-2 e 0.123 0.422 2.355 0.902

Walnut shell Walnut shell 0.450 0.490 0.390 0.520 0.520 (Zabihi et al., 2009)

Carbon A 0.540 0.480 0.350 0.420 0.420

Carbon B 0.720 0.420 0.300 0.290 0.290

Coconut activated carbon AC 1.097 0.627 0.561 0.495 e (Lu et al., 2014)

AC1 1.174 0.594 0.693 0.341

AC1-1 1.295 0.341 0.726 0.572

AC1-2 1.328 0.099 0.869 0.594

Sago 1.200 1.800 0.900 1.600 1.100 (Kadirvelu et al., 2004)

Furfural 0.120 0.160 0.230 1.500 0.600 (Yardim et al., 2003)

Antibiotic waste BDLa BDL 0.230 2.300 1.300 (Budinova et al., 2008)

a

Below detection limit.

w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 3 7 e5 5 47
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45 472 0.0090 0.9800 15.9 0.560 0.940

60 578 0.0080 0.8900 15 0.600 0.890

Sulfur impregnation of (AC) with

4%(v/v) SO2at 700 C for 60 min

30 523 0.0220 0.9600 71.2 0.330 0.990

45 496 0.0740 0.9200 129.2 0.250 0.980

60 510 0.0860 0.9500 148.2 0.230 0.960

Steam activation of bagasse

pith (SA-C)

Langmuir 30 172.4 0.0072 e e

40 181.8 0.008

50 200 0.0086

60 208.3 0.0106

Steam activation of bagassepith in presence of SO2

(SAeSO2eC)

30 185.2 0.019540 204.1 0.0202

50 208.3 0.0229

60 222.2 0.0262


pith in presence of H2S

(SAeH2SeC)

30 181.8 0.0113

40 200 0.0123

50 204.1 0.0164

60 217.4 0.0229


pith in presence of SO2 & H2S

(SAeSO2eH2SeC)

30 188.7 0.0281

40 208.3 0.0273

50 212.8 0.0367

60 227.3 0.0480

Rice husk ash Langmuir and Freundlich 15 9.3 0.0115 0.9868 0.42 0.493 0.973

30 6.7 0.0158 0.9900 0.54 0.469 0.965

Sulfuric acid treated with rice

husk (dry sorbent)

Langmuir and Freundlich 25 303 0.0052 0.9990 7.1 0.5579 0.981

35 336.7 0.0107 0.9992 28.7 0.377 0.98745 384.6 0.0219 0.9988 48.9 0.3407 0.986

Sulfuric acid treated with rice

husk (wet sorbent)

25 227.3 0.0052 0.9991 8.7 0.4607 0.952

35 270.3 0.0088 0.9998 15.0 0.428 0.928

45 303 0.0129 0.9993 18.1 0.448 0.959

Sulfuric acid treated with

flax shave (dry sorbent)

Langmuir 45 416 0.0805 0.9980

Sulfuric acid treated with

flax shave (wet sorbent)

344 0.0468 0.9990

Commercial activated carbon (AC) Langmuir and Freundlich 10 4.1 0.2083 0.9522 0.8 0.974 0.956

25 3.5 0.1053 0.9827 0.3 1.262 0.971

50 3 0.1330 0.9828 0.4 1.150 0.970

Steam activated AC (AC-1) 10 4.9 0.6870 0.9889 2.8 0.790 0.975

25 4.5 0.1540 0.9914 0.6 1.248 0.982

50 4.1 0.3862 0.9866 1.4 0.750 0.952

AC1 was oxidized with H2O2at 1:2 (m:v)

(AC1-1)

10 5.0 0.2329 0.9828 1.1 0.914 0.986

25 5 0.2000 0.9914 1 0.858 0.984

50 4.6 0.2618 0.9949 1.6 0.857 0.985

AC1 was oxidized with H2O2at 2:5 (m:v)

(AC1-2)

10 5.2 0.7795 0.9835 3.4 0.823 0.976

25 5.1 0.4491 0.9841 2.3 1.031 0.979

50 4.6 0.9287 0.9521 2.7 0.720 0.980

PSAC grafted with TOMATS Langmuir and Freundlich 20 76.9 0.0588 0.9920 5.5 0.630 0.916

25 76.9 0.0778 0.9940 6.9 0.590 0.906

30 83.3 0.1100 0.9930 9.4 0.560 0.892

35 83.3 0.1250 0.9910 10.2 0.570 0.898

Activated coke (AC) Langmuir and Freundlich 25 315.8 0.0500 0.9770 12.2 0.714 0.980

Thiol-functionalized activated coke (SH-AC) 694.9 0.0600 0.9840 71.1 0.526 0.954


14/19

for better mercury uptake. Nabais et al. have demonstrated

that the type of sulfur doped onto the adsorbent surface is a

more critical factor than its quantity. They observed that the

increase in the sulfur content of the activated carbon using

solid sulfur as the modifying agent did not improve mercury

uptake, whereas the modification of the adsorbent by H2S

resulted in a considerable increase in mercury adsorption.

They related this phenomenon to better accessibility of thesulfurto mercury by gas modification compared with the solid

modification (Nabais et al., 2006).

The activation atmosphere can also affect the surface

functionality of activated carbon. Budinovaet al. have found

that activated carbons pyrolyzed under nitrogen atmosphere

have high carboxylic group on the material surface, but

consecutive pyrolysis and steam activation results in a sig-

nificant drop in the content of carboxylic and lactone groups

and formation of more hydroxyl and carbonyl groups

(Budinova et al., 2006). It has also been highlighted that acti-

vation in the presence of air and water vapor results in an

increase and decrease of oxygen content of the final modified

product, respectively (Budinova et al., 2003).Besides physical activation, chemical activation can also

alter the functional groups on the activated carbon surface.

Most activated carbons contain varying amounts of functional

groups such aseOH, eCH]O and COOH without any treat-

ment. When activated carbons are treated with oxidizing

agent such as HNO3, H2O2, or (NH4)2S2O8, chemical reaction

occurs between the activating agent and the adsorbent sur-

face which alters the surface functionality and pKa of the

activated carbons as well as their porous structure and

adsorption capacity (Bandosz et al., 1993; Montagnaro and

Santoro, 2009).

X-ray photoelectron spectroscopy has demonstrated that

that oxygen- and nitrogen-containing functional groups actas electron donors during mercury adsorption and it has

been hypothesized that chemical coordination of mercury

with these functional groups are accountable for mercury

adsorption (Zhu et al., 2009). Therefore, higher oxygen- and

nitrogen-containing functional groups favor the mercury

adsorption. Zhu et al. studied the effect of activating agent

and chemical activation time on the surface functionality of

the activated carbons and detected the formation of hy-

droxyl, carboxylic and carboxylic anhydride group by nitric

acid treatment (Zhu et al., 2009). When the contact time be-

tween activated carbon and the nitric acid increases, signif-

icant amount of phenolic group forms while the content of

lactone group is reduced. Increase in the concentration ofnitric acid also leads to the formation of higher carboxylic

acid and carbonyl group content (Xianglan et al., 2011).

Xianglan et al. have investigated the use of hydrogen

peroxide as modifying agent and found that the lactone

moieties are decomposed into carbonyl and phenol groups by

increasing the chemical concentration and reaction time

(Xianglan et al., 2011). Danish et al. have explored the effect

of two chemical agents and have found that phosphoric acid-

treated activated carbon contains higher amount of acidic

functional group as compared with using zinc chloride

(Danish et al., 2013). However, when acid treatment becomes

more intense, no further oxygen functional groups, crucial

for the adsorption capacity of the activated carbons, are

detected. Thus altering the type and concentration of the

activating agent is an important factor to be studied for

adsorption purposes. Ahmad et al. used 2 Mhydrochloric acid

to treat cocoa shell for 2 h at a relatively high temperature

and found that, despite the high surface area obtained, the

oxygen functional group was not detected. It has been hy-

pothesized that the oxygen attached to the minerals are

removed by intense acid treatment (Ahmad et al., 2013).

6. Equilibrium adsorption isotherms

Langmuir and Freundlich isotherm models have been mostly

applied to describe the equilibrium adsorption of mercury (II)

on the adsorbent.

The Langmuir isotherm model was originally developed to

describe gasesolid adsorption onto adsorbent. The model

assumes irreversible homogeneous monolayer adsorption

and each adsorbate being adsorbed only to one adsorption

site. It also assumes that all the adsorption sites are identical.Therefore, the affinity of each adsorbate to adsorbent is

equivalent resulting in constant enthalpies and sorption

activation energy, without lateral interaction and steric hin-

drance between the adsorbed molecules on the adjacent sites

(Langmuir, 1918). The mathematical expression for this model

can be represented as:

qe qmaLCe1 aLCe

(1)

whereqeis the adsorption capacity of the adsorbent (mg:g1);

qm is the maximum adsorption capacity of the adsorbent

(mg:g1);Ce is the equilibrium concentration of the adsorbate

in the solution (mg:L1

); and aL is the Langmuir constant. Amathematical expression developed by Webber and Chakk-

ravorti has been used to test the favourability of the adsorp-

tion process (Weber and Chakravorti, 1974):

RL 1

1 aLC0(2)

where RL, called separation factor, is a dimensionless term

and Co is the adsorbate initial concentration (mg:L1). The

separation factor (RL) indicates the adsorption nature to be

unfavorable (RL >1), linear (RL 1), favorable (0


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adsorption process is favorable and the adsorbent surface is

considered heterogeneous (Jodeh et al., 2014).

Although both the Langmuir and Freundlich isotherm

models are popular tools in the prediction of equilibrium

adsorption isotherm, these models are sometimes criticized

due to their limitations of oversimplified assumptions for the

former and lack of fundamental basis for the latter isotherm

model. In addition, other hybrid forms of the Langmuir andFreundlich adsorption models are also established. Redlich-

Peterson and Sips isotherm models are the modified

isotherm models that incorporate the features of both the

Langmuir and Freundlich equations.

The Redlich-Peterson model can be applied either in ho-

mogeneous and heterogeneous adsorption (Redlich and

Peterson, 1959):

qe qmaRCe

1 aRCbe

(4)

whereaR is the Redlich-Peterson isotherm constant and b is

the Redlich-Peterson isotherm exponent, which lies between

0 and 1. The two extremes of the exponent b transform thisequation to the Henry's law and Langmuir equations when its

value is the lowest and highest, respectively. For any other

exponent values, this equation can be considered as an

incorporation of the features of the Langmuir and Freundlich

models.

Sips isotherm model is applied for the prediction of het-

erogeneous adsorption system and assumes the occurrence of

dissociative adsorption (Sips, 1948). At low adsorbate con-

centration, the isotherm can be reduced to Freundlich

isotherm, while at high concentration, it complies with ho-

mogenous adsorption characteristic of the Langmuir isotherm

(Diaz et al., 2007). The Sips isotherm is still criticized not to

follow the Henry's law at low adsorbent concentration. TheSips isotherm model is expressed through the following

equation:

qe qmaSCbse1 aSC

bse

(5)

where as is the Sips isotherm model constant andbsisthe Sips

isotherm model exponent.

Table 4 summarizes the Langmuir and Freundlich con-

stants obtained for the removal of aqueous mercury by

various adsorbent materials. Pena-Rodriguez et al. have

tested the mercury adsorption behavior of three adsorbents

obtained from calcined mussel shell, finely ground musselshell and coarsely ground mussel shell and identified that

the KF value obtained from Freundlich isotherm model is

not closely correlated with the surface area, given that the

highest KF value corresponds to the calcined shell with

surface area lower than that of finely ground mussel ( Pe~na-

Rodrguez et al., 2013). Cai and Jia showed the S-shaped

relationship between SBET and mercury adsorption and

suggested that the high porosity contributed by micropore

might not be accessible to mercury and its species, such as

HgCl2 and HgOHCl, instead a better linear correlation was

found between the adsorption capacity and mesopore sur-

face area (Cai and Jia, 2010). This phenomenon suggested

that the surface area alone is not sufficient to reflect the

adsorption capacity of the adsorbent, but several other

factors such as pore size, surface structure and adsorbate

species will also interfere with the overall adsorption

capacity.

Cai and Jia have found a positive correlation between the

mercury adsorption capacity of activated carbons and its

sulfur content (Cai and Jia, 2010). Wang et al. have also

determined that although sulfur impregnation results in sig-nificant reduction in SBET of the activated carbons, a stag-

gering enhancement in the mercury adsorption capacity was

achieved (Wang et al., 2009). The strong affinity between

mercury and sulfur has also been reported by Asasian et al.

(Asasian et al., 2014).

Further information could be deduced regarding the

mechanism of mercury adsorption by comparing the

isotherm shapes as a means of fitting mercury adsorption

isotherms, and correlating these with Giles isotherm classifi-

cation (Giles et al., 1960). To date, no authors have attempted

to do this for aqueous mercury adsorption.

7. Conclusion and future perspectives

Removal of mercury from wastewater using activated carbons

has been shown to be very promising if proper combination of

properties is possessed by the adsorbent materials. The ef-

fects of surface area and functional groups of the activated

carbons on mercury uptake have been examined in numerous

studies. In this study, the generic misconception that higher

surface area of an adsorbent leads to higher mercury

adsorption has been criticized. Herein, it has been demon-

strated that a combination of medium-to-high surface area

with well-functionalized surface properties are collectivelycrucial in enhancing the mercury removal. Despite these

findings, very limited research has been carried out on

simultaneous optimization of surface and textural charac-

teristics of activated carbons. Hence, further study is neces-

sary for such optimization to save the high amount of energy

required to obtain unnecessary very high surface area acti-

vated carbons.

Furthermore, although a lot of studies are concerned with

sulfurization of activated carbons with the aim of higher

mercury uptake, the mechanism of sulfurization process,

including the type and quantity of sulfur-containing moieties

doped onto the activated carbon surfaces and their function-

alization process, and consequently the mercury adsorptionmechanism are not thoroughly examined. A profound insight

into the activation and adsorption mechanisms will assist in

designing a proper adsorbent-adsorbate system for optimal

mercury abatement from effluents.

In addition, the mercury adsorption is believed to take

place in two stages; initially the surface active sites are

involved in the adsorption process and when the surface sites

are less available, the mercuric ions have to diffuse into the

pores. Therefore, a combination of pseudo and diffusion

models has to be considered for modeling the mercury

adsorption kinetic results. Nevertheless, this modeling strat-

egy has been disregarded.

Further recommendations can be presented as follows:

w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 3 7 e5 5 51


16/19

Mercury forms complexes in real wastewater systems and

is rarely found in ionic form. Therefore, study of the effi-

ciency of the activated carbon adsorbents using industrial

wastewater seems to be of high priority.

The presence of other metallic compounds in the effluent

will unequivocally affect the mercury removal efficiency of

the activated carbon samples. Therefore, more detailed

study of the multi-component adsorption systems have tobe carried out.

The regeneration of the activated carbon samples has to be

conducted for economic feasibility enhancement. Unfor-

tunately, in adsorption processes, the regeneration is

overlooked.

If not regenerated, due to its hazardous nature, mercury-

loaded activated carbon needs to be stabilized or vitrified

and then disposed of in hazardous landfill. Nonetheless,

this landfilled carbon is not reusable, therefore empha-

sizing the importance of the production cost of the acti-

vated carbon.

Typically, the focus of the research in mercury adsorption

systems is lab-scale batch adsorption studies. However,the studies should not be confined only to these lab-scale

experiments and column studies have to be performed

for better understanding the industrial-scale operation

challenges.

r e f e r e n c e s

Adams, M.D., 1991. The Mechanisms of Adsorption of Hg ( CN ) 2and HgC12on to Activated Carbon, vol. 26, pp. 201e210.

Adamson, A.W., Gast, A.P., 1967. Physical Chemistry of Surfaces.Aguado, J., Arsuaga, J.M., Arencibia, A., 2005. Adsorption of

aqueous Mercury(II) on propylthiol-functionalizedmesoporous silica obtained by cocondensation. Ind. Eng.Chem. Res. 44, 3665e3671.

Ahmad, F., Daud, W.M.A.W., Ahmad, M.A., Radzi, R., 2013. Theeffects of acid leaching on porosity and surface functionalgroups of cocoa (Theobroma cacao)-shell based activatedcarbon. Chem. Eng. Res. Des. 91, 1028e1038.

Anirudhan, T.S., Divya, L., Ramachandran, M., 2008. Mercury(II)removal from aqueous solutions and wastewaters using anovel cation exchanger derived from coconut coir pith and itsrecovery. J. Hazard. Mater. 157, 620e627.

Anoop Krishnan, K., Anirudhan, T.S., 2002. Removal ofmercury(II) from aqueous solutions and chlor-alkali industryeffluent by steam activated and sulphurised activated carbonsprepared from bagasse pith: kinetics and equilibrium studies.

J. Hazard. Mater. 92, 161e183.Antochshuk, V., Jaroniec, M., 2002. 1-Allyl-3-propylthiourea

modified mesoporous silica for mercury removal. Chem.Commun. 258e259.

Asasian, N., Kaghazchi, T., Faramarzi, A., Hakimi-Siboni, A.,Asadi-Kesheh, R., Kavand, M., Mohtashami, S.-A., 2014.Enhanced mercury adsorption capacity by sulfurization ofactivated carbon with SO2in a bubbling fluidized bed reactor.

J. Taiwan Inst. Chem. Eng. 45, 1588e1596.Bandaru, N.M., Reta, N., Dalal, H., Ellis, A.V., Shapter, J.,

Voelcker, N.H., 2013. Enhanced adsorption of mercury ions onthiol derivatized single wall carbon nanotubes. J. Hazard.Mater. 261, 534e541.

Bandosz, T.J., Jagieuo, J., Schwarz, J.A., 1993. Effect of SurfaceChemical Groups on Energetic Heterogeneity of Activated

Carbons, pp. 2518e2522.

Bansal, R.C.C., Donnet, J.-B.J.-B., Stoeckli, F., 1988. Active Carbon.Macel Dekker, New York.

Barron-Zambrano, J., Laborie, S., Viers, P., Rakib, M., Durand, G.,2002. Mercury removal from aqueous solutions bycomplexationdultrafiltration. Desalination 144, 201e206.

Barron-Zambrano, J., Laborie, S., Viers, P., Rakib, M., Durand, G.,2004. Mercury removal and recovery from aqueous solutionsby coupled complexationeultrafiltration and electrolysis. J.

Memb. Sci. 229, 179e186.Blue, L.Y., Jana, P., Atwood, D.A., 2010. Aqueous mercury

precipitation with the synthetic dithiolate, BDTH2. Fuel 89,1326e1330.

Botta, S.G., Rodriguez, D.J., Leyva, A.G., Litter, M.I., 2002. Featuresof the transformation of Hg II by heterogeneousphotocatalysis over TiO2. Catal. Today 76, 247e258.

Budinova, T., Ekinci, E., Yardim, F., Grimm, a, Bjornbom, E.,Minkova, V., Goranova, M., 2006. Characterization andapplication of activated carbon produced by H3PO4and watervapor activation. Fuel Process. Technol. 87, 899e905.

Budinova, T., Petrov, N., Parra, J., Baloutzov, V., 2008. Use of anactivated carbon from antibiotic waste for the removal ofHg(II) from aqueous solution. J. Environ. Manage 88, 165e172.

Budinova, T., Savova, D., Petrov, N., Razvigorova, M., Minkova, V.,

Ciliz, N., Apak, E., Ekinci, E., 2003. Mercury adsorption bydifferent modifications of furfural adsorbent. Ind. Eng. Chem.Res. 42, 2223e2229.

Byrne, H.E., Mazyck, D.W., 2009. Removal of trace level aqueousmercury by adsorption and photocatalysis on silica-titaniacomposites. J. Hazard. Mater. 170, 915e919.

Cai, J.H., Jia, C.Q., 2010. Mercury removal from aqueous solutionusing coke-derived sulfur-impregnated activated carbons. Ind.Eng. Chem. Res. 49, 2716e2721.

Chiarle, S., Ratto, M., Rovatti, M., 2000. Mercury removal fromwater by ion exchange resins adsorption. Water Res. 34,2971e2978.

Choi, M., Jang, J., 2008. Heavy metal ion adsorption ontopolypyrrole-impregnated porous carbon. J. Colloid InterfaceSci. 325, 287e289.

Cox, M., El-shafey, E.I., Pichugin, A.A., Appleton, Q., 2000. Removalof Mercury ( II ) from Aqueous Solution on a CarbonaceousSorbent Prepared from Flax Shive, vol. 435, pp. 427e435.

Cui, H., Qian, Y., Li, Q., Wei, Z., Zhai, J., 2013. Fast removal of Hg(II)ions from aqueous solution by amine-modified attapulgite.Appl. Clay Sci. 72, 84e90.

Danish, M., Hashim, R., Ibrahim, M.N.M., Sulaiman, O., 2013.Effect of acidic activating agents on surface area and surfacefunctional groups of activated carbons produced from Acaciamangium wood. J. Anal. Appl. Pyrolysis 104, 418e425.

De la Fourniere, E.M., Leyva, A.G., Gautier, E.A., Litter, M.I., 2007.Treatment of phenylmercury salts by heterogeneousphotocatalysis over TiO(2). Chemosphere 69, 682e688.

Di Natale, F., Erto, A., Lancia, A., Musmarra, D., 2011. Mercuryadsorption on granular activated carbon in aqueous solutions

containing nitrates and chlorides. J. Hazard. Mater. 192,1842e1850.

Di Natale, F., Lancia, a, Molino, a, Di Natale, M., Karatza, D.,Musmarra, D., 2006. Capture of mercury ions by natural andindustrial materials. J. Hazard. Mater. 132, 220e225.

Diaz, E., Ordonez, S., Vega, a, 2007. Adsorption of volatile organiccompounds onto carbon nanotubes, carbon nanofibers, andhigh-surface-area graphites. J. Colloid Interf. Sci. 305, 7e16.

Ebinghaus, R., Turner, R., de Lacerda, L., Vasiliev, O.,Salomons, W., 1999. Mercury Contaminated Sites:Characterization, Risk Assessment and Remediation.

Ekinci, E., Budinova, T., Yardim, F., Petrov, N., Razvigorova, M.,Minkova, V., 2002. Removal of mercury ion from aqueoussolution by activated carbons obtained from biomass andcoals. Fuel Process. Technol. 77e78, 437e443.

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