This article by Hadi, To, Hui, Ki Lin, McKay describes methods of mercury (aq) adsorption through activated carbon for t...
water research 73 (2015) 37 55 e
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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 Kong b School of Energy Energy and Environme Environment, nt, City University University of Hong Kong, Tat Chee Avenue, Avenue, Kowloon, Kowloon, Hong Kong c Division of Sustainable Development, College of Science, Engineering and Technology, Hamad Bin Khalifa University, Qatar Foundation, Doha, Qatar
a r t i c l e
i n f o
Article history:
Received Received 14 October October 2014 Received Received in revised form 19 December 2014 Accepted Accepted 9 January 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 distribution tribution in our food chain through the contamination contamination of water bodies, stringent regulations 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 mercury adsorption. 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 addressed using equilibrium equilibrium adsorption adsorption isotherm, isotherm, thermodynami thermodynamicc and kinetic 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 Elsevier Ltd. All rights reserved.
Contents 1. 2. 3.
Introduction Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38 Preparation Preparation of activated carbon carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39 Effect of treatment treatment techniques on mercury removal removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40 3.1. Physical Physical and chemical activation activation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40 3.2. Sulfuriza Sulfurization tion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43
* Corresponding author. Chemical and Biomolecular Engineering Department, Hong Kong University of Science and Technology, Clear Water Bay Road, Hong Kong. Tel.: þ 852 23588412; fax: þ 852 23580054. 23580054. E-mail address:
[email protected] address:
[email protected] (G. (G. McKay). http://dx.doi.org/10.1016/j.watres.2015.01.018 0043-1354/© 2015 Elsevier Ltd. All rights reserved.
38 4.
5. 6. 7.
1.
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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 Mercury affinity to various functional groups . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46 Equilibrium adsorption isotherms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50 Conclusion and future perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 51 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52
Introduction
Mercury is categorized as an extremely toxic substance whose health hazards have primarily been associated with inhalation of mercury vapor or ingestion of organic mercury via aquatic 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 the most dominant role. An overarching estimate of anthropogenic input of 16 elements into the environment by Nriagu and Pacyna has revealed that mercury releases into the atmosphere 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 environment, 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 processes, including vinyl chloride monomer production as catalyst, batteries as cathodes and chlor-alkali production as well (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 elimination of mercury from the ecosystem is unrealistic due to the uncontrollable discharge of mercury by primary anthropogenic 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 directives have been enacted regarding the control of the mercury 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-Munoz 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 concerning 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 the availability 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 biopolymers (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
~
39
water research 73 (2015) 37 55 e
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 mercury removal 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 treatment 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 heat treatment of a carbonaceous precursor at moderate temperatures, mainly below 700 C, in an inert atmosphere to pyrolize 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 produced 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 increases, while furtherincrease in the temperature reduces the pore volume of the material. The produced char is subsequently 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). The aim of the activation stage is to produce a highly porous structure from the weakly-developed porous char. The activation 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 stoichiometric equations:
C þ CO2
2CO
/
C þ H2 O
DH ¼ þ159
CO þ H2
/
KJ mol1
DH ¼ þ117
(R1)
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 material is submerged in a dehydrating compound (such as phosphoric 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 structure. The resultant slurry is then heatedup to temperatures in the rangeof 400 600 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 heat treatment. Energy required for chemical activation process is considerably lower than that of physical activation method. Significantly 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 consumption. 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-Agullo et al., 2004). Although, commercial activated carbons have been widely used in various industries to remove mercury from wastewater 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 wastebased carbon sources as precursor for activated carbon production. Toles et al. found that it is highly profitable to produce 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 costeffective activated carbons with high pollutant removal efficiencies. Among certain precursor materials to be used in activated carbon production, high mercury adsorption capacities 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 carbons derived from pazzolana and yellow tuff (Di Natale et al., 2006), Fullers earth (Oubagaranadin et al., 2007), Ceibapentandra hulls, Phaseolus aureus hulls and Cicerarietinum e
40
water research 73 (2015) 37 55 e
. f e R
) 9 0 0 2 , l . a t e o a R (
) p m D ( n
) ) 6 3 0 0 0 0 2 2 , , l . l . a a t t e e a a v v o o n i n i d d u u B ( B (
) 2 0 0 2 , l . a t e i c n i k E (
(Rao et al., 2009). Hence, it can be inferred that the type of precursor is an influential factor in determining the adsorption efficiency of the activated carbons.
) 6 0 0 ) 2 4 , a 0 n 0 a 2 , h l . c a o t l u e S l o d l n u a g j A a a r i a c b a n I ( M (
) ) 8 6 0 0 0 0 2 , ) 2 g 4 , 1 l . n a 0 a J t d 2 e n l . , s a a i a i t b o e a h u N ( C ( L (
3. Effect of treatment techniques on mercury removal 3.1.
7 6 5 . 9 . 2 1 2
2 3 . 2
. s ) n o o 1 5 4 s g b 2 1 8 0 r e 3 1 0 0 m . . . . a 0 0 0 0 c V m c d ( e t a v o ) i t r 7 5 8 c g c i 9 2 1 2 2 4 2 3 a m . . . . 0 0 0 0 e V m c h ( t f o ) s 1 8 e t g i 8 9 4 4 6 7 2 c V 6 8 6 6 8 3 7 . . . . . . . 4 . n 0 0 0 0 0 0 2 0 m e i c ( c fi f ) e 8 6 8 3 l 8 9 a m . 6 . 8 . . v q g 5 3 2 4 4 2 6 5 7 3 4 0 4 g 2 2 2 7 3 9 5 0 3 5 7 4 o 1 1 1 1 1 1 m m ( e r ) y r T 1 5 0 1 0 0 0 0 0 0 0 0 8 7 7 2 0 u E g 2 2 8 6 0 8 0 6 2 2 0 0 2 8 9 0 7 c B 5 3 2 7 1 4 6 4 7 5 0 1 5 4 9 7 8 r S 1 1 1 1 2 e m ( m e h . O O O O t C C C H m n t O & & & & o O O A s H N H H N N N N n o s i n ) t i i o d t n n i d i o c n m 0 3 0 5 5 5 1 1 n o c C 2 o ( i t n o R a i v t i a H t v c i a t ) 5 . l c h 2 1 2 9 2 5 1 A ( a 2 T c i s ) y C 0 h 0 0 0 0 0 0 0 0 5 0 9 9 0 5 p ( 5 7 6 7 4 8 8 9 8 d T n a . e O m p t H N N N y A t r o ) s r u n i c e r s m 0 0 5 1 1 p n C o e i ( h t i t R f d H o n o t c c e n ) f ( 2 o h 2 1 5 f i E t T a e z 0 1 i n 5 e o ) 9 l C b e b r ( 0 0 0 0 0 a a 0 5 0 0 0 T C T 2 7 5 8 9 1
3
1
3
1
3
1
1
2
2
2
2
2
2
2
2
2
2
2
2
2
1
2
2
2
2
1
a
. e t a r g n i t a e h s e t o n e d R H a
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 2 show the effect of the precursor 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 oxidation on 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 atmospheres 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 an increase 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 area and functional groups are predominantly affected by the nature of the precursor material used. As shown in Table 1, Ekinci et al. have found that, under similar activation conditions, the surface areas range from 460 to 720 m2 :g 1 for coalbased activated carbons and 1000 and 1110 m2 :g 1 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).
Table 2
e
Effect of the precursor type and chemical activation conditions on the mercury removal efficiencies of the activated carbons.
Carbonization/Activation conditions T ( C)
T (h)
600/600 1/1 700 2 600
Chemical treatment
HRa ( C min1) 3 10
200/450 1
Atm. H2O/N2 H2O N2
H2SO4 H2SO4
Air
600/800 1
Conc.
Conc. Conc.
H2O2 H2SO4 & (NH4)2S2O8 50% H2SO4 Conc. H2SO4 & (NH4)2S2O8 Conc.
N2 /H2O
400/700 0.5/1 N2
Chem.
NaOH ZnCl2 H2SO4 K2CO3 KOH
900
5
N2 & CO 2 N2
900 900 200 200/400 2/1
800 850
5/15
10
0.5 0.5
1:200 3: 5 & 1:33
33-75%
S H2S Pyrrole
1:1 1:2 1:4 1:6 1:8 1:2 1:4 1:6 1:8 1:3
1290 1360 1050
0.5 12 2
24 2 1 1 1 1 1 1 1 1 2 1
1:3 43:100 11:20
N2 H2S SO2, H 2S & N 2 H2O, SO2 & H 2S H2O & SO 2 H2O,& H 2S H2O N2 K2S
qm
V t
V micro
V meso
Dp
Ref.
Imp.tc (h) (m2 g 1) (mg g 1) (cm3 g 1) (cm3 g 1) (cm3 g 1) (nm)
1:20
20% (W/V) 98% 1:0.5 1:1 Conc. 1:1.8
NaOH
800/900 1/5
Imp. Rb
SBET
1:3.6
629 625 1100 592 379.4 780 803 208.1 1260 1090 1635 2225 2420 1130 2000 2541 3033 848 905 413 346 158 922 785 764 500.5 506.5 530.2 536.5 30.0 30.0
160 154 10 10 10 10 10 18.1 55.6 174 154
0.939 1.026 0.827
0.471 0.485 0.38
0.425
0.486 0.541 0.19
0.11
1.4 1.28
(Budinova et al., 2003) (Kadirvelu et al., 2003)
>7.5
52.7 151.5 100.9 109.9 129
410 450 541 682 441 301 351 351 227.3 222.2 217.4 208.3 235.7 243.9
(Mohan et al., 2001) (Kadirvelu et al., 2004) (Yardim et al., 2003) (Namasivayam and Kadirvelu, 1999) (Wahi et al., 2009) (Zabihi et al., 2009) (Namasivayam et al., 1993) (Budinova et al., 2008) (Macia-Agullo et al., 2004)
0.492 0.49 0.78 0.9 0.94 0.51 0.81 0.96 1.02 0.33 0.33 1.52 1.04 0.54 0.87 0.78 0.72 0.43 0.48 0.48 0.52
(Budinova et al., 2006)
e
5 5
(Nabais et al., 2006) (Choi and Jang, 2008)
0.37 0.31 0.3 0.04 0.06 0.13
0.16 0.13 0.16 0.33 0.29 0.28 0.2
(Gomez-Serrano et al., 1998) (Anoop Krishnan and Anirudhan, 2002)
(Wajima and Sugawara, 2011) (continued on next page )
Table 2
e
Carbonization/Activation conditions T (h)
900 1000
0.5 0.5
2
700
Chemical treatment
HRa ( C min1)
50
Atm.
N2
4
Chem.
HCl H2SO4
HNO3, SOCl2 & C2H4(NH2)2 H2O2 10
N2
Conc.
H2O2 H2 & CS2 H2O2 & He H2O2, He & CS 2 He S
H2SO4
600 600 600 600 600 600 600 500 700 800 900
4 1
4 2
( continued )
T ( C)
w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 3 7
3:10
1N 13 M 13 M 12 M 12 M 5 M, 5% & 0.05 M 30%
0 01 M
1:5 1:5 1:9 1:9 3: 20 & 3:8 1:2 2:5 1:25 1:25 1:25 1:25 1:50 2:25 1:25 1:25 1:25 1:25 1:25 1:0.3
qm
V t
V micro
V meso
Dp
Ref.
Imp.tc (h) (m2 g 1) (mg g 1) (cm3 g 1) (cm3 g 1) (cm3 g 1) (nm)
168
2:1 2:1 2:1 2:1 2:1
SO2
TOMATS HNO
Imp. Rb
SBET
4 0.33 0.33 0.5 0.5 7
0.5 1 2 3 1 1 1 1 1 1 1 48 72
28.0 30.0 836.0 778.0 950.0 776.0 880.0 645.0 594.0
254.4 224.1 100 170 110 175 110 467 507
1070.0 1045.0 311.0 66.0
827 427 9.32 303.03 384 385 526 120
19.0
829.0 825.0 720.9 772.5 776.4 790.7 773.7 518.5 757.2 764.1 751.3 1087.0 1057.0 107.7 136 7
5 4.2 122.8 129.8 130.5 131.6 128.2 114.8 125.7 135.9 184.2 196.8 207.8 83.3 315 8
0.23 0.18
(Lopez-Gonzalez et al., 1982)
0.22 (Wang et al., 2009)
(Feng et al., 2004) (El-Shafey, 2010) (Cox et al., 2000) (Zhu et al., 2009) 0.413 0.408
0.318 0.315
0.01 0.013
1.994 (Lu et al., 2014) 1.98 (Asasian et al., 2014)
w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 3 7 e
5 5
(Ismaiel et al., 2013) (Li et al., 2013)
Table 2
e
4 2
( continued )
Carbonization/Activation conditions T ( C)
T (h)
900 1000
0.5 0.5
2
700
Chemical treatment
HRa ( C min1)
50
Atm.
N2
4
Chem.
H2O2 H2 & CS2 H2O2 & He H2O2, He & CS 2 He S
HCl H2SO4 H2SO4 HNO3, SOCl2 & C2H4(NH2)2 H2O2
600 600 600 600 600 600 600 500 700 800 900
10
N2
Conc.
3:10
1N 13 M 13 M 12 M 12 M 5 M, 5% & 0.05 M 30%
SO2
b c
1:5 1:5 1:9 1:9 3: 20 & 3:8 1:2 2:5 1:25 1:25 1:25 1:25 1:50 2:25 1:25 1:25 1:25 1:25 1:25 1:0.3
qm
V t
V micro
V meso
Dp
Ref.
Imp.tc (h) (m2 g 1) (mg g 1) (cm3 g 1) (cm3 g 1) (cm3 g 1) (nm)
168
2:1 2:1 2:1 2:1 2:1
TOMATS HNO3 0.01 M HSCH2COOH, (CH3CO)2O & H 2SO4 a
Imp. Rb
SBET
4 0.33 0.33 0.5 0.5 7
0.5 1 2 3 1 1 1 1 1 1 1 48 72 13
28.0 30.0 836.0 778.0 950.0 776.0 880.0 645.0 594.0
254.4 224.1 100 170 110 175 110 467 507
1070.0 1045.0 311.0 66.0
827 427 9.32 303.03 384 385 526 120
19.0
829.0 825.0 720.9 772.5 776.4 790.7 773.7 518.5 757.2 764.1 751.3 1087.0 1057.0 107.7 136.7 193.7
0.23 0.18
(Lopez-Gonzalez et al., 1982)
0.22 (Wang et al., 2009)
5 4.2 122.8 129.8 130.5 131.6 128.2 114.8 125.7 135.9 184.2 196.8 207.8 83.3 315.8 694.9
w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 3 7
(Feng et al., 2004) (El-Shafey, 2010) (Cox et al., 2000) (Zhu et al., 2009) 0.413 0.408
0.318 0.315
0.01 0.013
1.994 (Lu et al., 2014) 1.98 (Asasian et al., 2014)
e
5 5
(Ismaiel et al., 2013) (Li et al., 2013)
HR denotes heating rate. Imp. R denotes impregnation ratio. Imp. t denotes impregnation time.
water research 73 (2015) 37 55 e
Nonetheless, although specific surface area is one of the factors 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 m 2 :g 1 , while their mercury adsorption capacities were 151 and 101 m2 :g 1 , respectively (Zabihi et al., 2009). Also, the surface areas of the activated carbons prepared by Roa et al. ranged from 280 to 521 m2 :g 1 , whereas their mercury adsorption capacities under similar adsorption conditions did not exhibit a significant difference (Rao et al., 2009). Wang et al. demonstrated that an activated carbon with a surface area of 1896 m2 :g 1 had a much smaller adsorption capacity than an activated carbon with a surface area of 1070 m 2 :g 1 (160 vs 827 m2 :g 1 , 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 carbons 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 simultaneously investigated the effects of pore structure and functional groups on mercury removal. Considering the surface areas listed in Tables 1 and 2, it is
43
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 mercury. Therefore, the application of sulfur-functionalized activated 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 sulfide (Na2S), hydrogen sulfide (H2S), sulfur dioxide (SO2) or sulfur powder, with the aim of increasing their mercury uptakes (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 techniques, namely impregnation with elemental sulfur and using hydrogen sulfide gas as modifying agent. Both of these adsorbents exhibited higher adsorption capacities than the un-
43
water research 73 (2015) 37 55 e
Nonetheless, although specific surface area is one of the factors 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 m 2 :g 1 , while their mercury adsorption capacities were 151 and 101 m2 :g 1 , respectively (Zabihi et al., 2009). Also, the surface areas of the activated carbons prepared by Roa et al. ranged from 280 to 521 m2 :g 1 , whereas their mercury adsorption capacities under similar adsorption conditions did not exhibit a significant difference (Rao et al., 2009). Wang et al. demonstrated that an activated carbon with a surface area of 1896 m2 :g 1 had a much smaller adsorption capacity than an activated carbon with a surface area of 1070 m 2 :g 1 (160 vs 827 m2 :g 1 , 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 carbons 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 simultaneously investigated the effects of pore structure and functional groups on mercury removal. Considering the surface areas listed in Tables 1 and 2, it is noticeable that the chemical activation technique has greater performance in pore formation at appropriate chemical reagent to adsorbent ratio. Comparing the surface areas obtained by chemical and physical activation of coal tar pitch carbon fibers, it is obvious that although high surface area activated carbon (2487 m2 :g 1 ) 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 burnoff (94%). On the contrary, chemical activation of this material using 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 advantages of the chemically activated carbon are higher product yield (60% and 27%, respectively), lower activation temperature (750 C) and shorter activation time (1 h). The highestsurface area (3033 m2 :g 1 ) can be obtainedusing NaOH as activating agent at an impregnation ratio of 8:1 ( w/w) (Macia-Agullo et al., 2004). In order to enhance mercury adsorption, several authors have studied the combination of chemical and physical activation techniques. Budinova et al. confirmed that when the H3PO4-impregnated carbonaceous sample was treated under steam atmosphere, both the surface area and iodine number were considerably higher than the samples pyrolyzed under nitrogen atmosphere. They also revealed that the concentration of the chemical reagent used for impregnation has a significant effect on pore development. When the concentration 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 concentration 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 mercury. Therefore, the application of sulfur-functionalized activated 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 sulfide (Na2S), hydrogen sulfide (H2S), sulfur dioxide (SO2) or sulfur powder, with the aim of increasing their mercury uptakes (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 techniques, namely impregnation with elemental sulfur and using hydrogen sulfide gas as modifying agent. Both of these adsorbents exhibited higher adsorption capacities than the unsulfurized 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 effect 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 capacity of the unmodified and modified activated carbons were 190 mg :g 1 and 820 mg :g 1 , respectively. The chemical reaction between elemental sulfur and the surface of the adsorbent 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 satisfactorily exploited 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 C S bonds (at 475 cm1 , 495 cm1 and 503 cm1 ) associated with thiol and thioester groups after treating the virgin carbon nanotube with phosphorus pentasulfide. Subsequent mercury adsorption revealed significantly diminished intensities of the bands corresponding to the thiol ]
e
44
water research 73 (2015) 37 55 e
groups and appearance of new band assigned to Hg(SH)2 and Hg 2(SH)2 bonds 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 cm 1 corresponding to Hg O 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 S S, C S, C S and S O 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 mercury, higher amounts of sulfur moieties will theoretically be advantageous in mercury removal. Several researchers have reported the direct linear relationship between mercury uptake 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 600 800 C resulted in the production of an adsorbent with higher sulfur content compared with the introduction of sulfur to ACF via gas stream H2S. This may be reasonable due to the melting, recrystallization and deposition of the solid sulfuron the activated carbon surface at such high temperatures. However, subsequent mercury adsorption tests showed that higher mercury uptake was obtained by the latter method which elucidated the importance of the type of sulfur functional groups on the carbon surface besides its quantity (Nabais et al., 2006). This may also occur due to the aggregation 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-impregnated coal samples were prepared at three distinct temperatures (800 1000 C), whereas the activated carbon sample prepared at 900 C exhibited the highest and fastest adsorption (Wajima and Sugawara, 2011). This indicates that the sulfur content on the adsorbent surface, type of sulfur functional 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 multicomponent 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 e
e
]
e
bases) on the adsorbent surface (Cai and Jia, 2010). This phenomenon has also been confirmed by comparing the interaction of HgX2 (where X is a halide) with C4H8O and C4H8S. It has been reported that HgX2 interacts 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 electronegativity of mercury (Vazquez et al., 2002).
e
e
e
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 adsorption process 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 functional 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 plateau is 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 mercury removal percentage and thus more cost effectiveness (Namasivayam and Periasamy, 1993). Also, rapid mercury adsorption of less than 20 min to reach equilibrium was reported 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 increases 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
45
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studies are necessary to be conducted to prove these hypotheses regarding the relationship between mercury adsorption rate and textural and surface properties of the adsorbents. It has been further demonstrated that as the initial concentration of mercury increases, the Lagergren rate constant decreases and thus, longer time is required to achieve equilibrium (Namasivayam and Kadirvelu, 1999; Namasivayam and 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. However, 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 concentration 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 concentration. 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 dosages, whereas the latter takes into account the adsorbent dosage and reveals the genuine mercury adsorption efficiency of the material at different initial concentrations. Several authors 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 consideration, 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 to the 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 concentration 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 concentrations (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 contact between the adsorbate molecules and the adsorbent materials thus enhancing the possibility of the mercury uptake 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 ionization of the functional groups on the adsorbent surface. In acidic environment, high concentration of hydronium ion (H3Oþ) in the solution drives the equilibrium ionization reaction (reaction R4) to the left and prevents the formation of ionized functional groups, thereby hampering the ion exchange reaction between metal ions and adsorbent surface functional groups (reaction R5). When the pH level of the solution 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 COOH
Adsorbent COO þ HþðaqÞ
4
þ Adsorbent COO þ Mnðaq Þ
Adsorbent COO… M
4
(R4) (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 positively charged leading to electrostatic repulsion of the mercury 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 temperature 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 activatedcarbon 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 an increase in compressibility of the particles with increasing temperature. Since porous structure of carbonaceous materials usually forms at very high temperatures, it is particularly implausible to alter the pore sizes at adsorption temperatures as low as 20 80 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 mercury by the adsorbent material. Physical adsorption caused by the van der Waals forces between the adsorbent surface and the adsorbate molecules is typically favored at low temperatures. This explains the higher adsorption capacity of the fertilizer-based adsorbent for mercury at low temperatures (Mohan et al., 2001). e
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 concentration 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 is increased, as discussed in preceding sections, removal percentage is an entirely relative term changing by initial concentration 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. pentandra hull 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 the adsorbent 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 physisorption as well as higher rate of adsorbate removal (Wahi et al., 2009). It is suggested that more tangible adsorption capacities should be reported in this context instead of simply quoting the percent removal. On the other hand, although the removal percentage increases 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 availability 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 particles decreased from 1.25 2.5mmto0.21 1 mm,the mercury adsorption capacity showed a two-fold increase (430 mg :g 1 versus 815 mg :g 1 , respectively) (Mckay et al., 1989; PenicheCovas et al., 1992). Similarly, Kadirvelu et al. have also demonstrated that a stepwise decrease of activated carbon particle size produced from sago waste (750 500 mm, 500 250 mm and 250 125 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 molecules can more easily penetrate into the internal pores of the adsorbent (Gupta et al., 2011). e
e
e
e
e
5. Mercury affinity to various functional groups The adsorption of adsorbate by activated carbon can be categorized into chemical and physical adsorption. Briefly, physical 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 adsorbent, 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 unbalanced charge on the adsorbent surface where ion exchange with the mercuric ion can occur (Sun et al., 2011). This accounts 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 nitrogencontaining 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 efficiency of adsorption. As functional groups largely determine the surface properties and thus the intensity of ion exchange, manipulation of the surface functional group have been of great interest. Sulfur group has been widely reported to promote mercury adsorption. Rao et al. have observed an increase in the mercury adsorption capacity of activated carbon with the introduction of sulfur groups on the activated carbon surface and ascribed the higher removal efficiency of the sulfurcontaining 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). 2Hg 2þ þ SO23 þ 2OH
Hg 22þ þ SO24 þ H2 O
(R6)
4
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 :g 1 for the unmodified material to 694.9 mg :g 1 for the modified material (Li et al., 2013). Anoop Krishnan and Anirudhan have verified the effect of sulfur
Table 3
e
e
Acid-base neutralization capacity (meq/g) of the activated carbon adsorbents.
Precursor material Mengen Seyitomer Some Bulluca Apricot Stones Furfural Furfural Mixture of steam pyrolysis tar and furfural (30:70) Air oxidized Furfural Woody biomass birch
Walnut shell
Coconut activated carbon
Sago Furfural Antibiotic waste a
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 can be 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 the activated 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 200 500 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 functional groups on mercury removal. The type of the activating agent has also been considered effective in the manipulation of the surface functional groups
Below detection limit.
Name
Base uptake NaHCO3
Na2CO3
NaOH
EtONa
HCL
0.092
0.120 0.120 0.110 0.110 0.210 0.160 0.160 0.080
0.184 0.250 0.183 0.320 0.360 0.230 0.230 0.250
1.900 2.040 1.570 1.900 1.350 1.500 1.500 1.300
1.120 2.670 1.120 2.010 0.842 0.600 0.600 0.560
1.930 0.126 0.034 0.123 0.490 0.480 0.420 0.627 0.594 0.341 0.099 1.800 0.160 BDL
3.660 0.480 0.572 0.422 0.390 0.350 0.300 0.561 0.693 0.726 0.869 0.900 0.230 0.230
6.340 2.234 2.530 2.355 0.520 0.420 0.290 0.495 0.341 0.572 0.594 1.600 1.500 2.300
e
0.100 e
Carbon A Carbon B Carbon C N600-1 NS600-1 S700-2 Walnut shell Carbon A Carbon B AC AC1 AC1-1 AC1-2
Acid uptake
0.130 0.120 0.120 0.030 1.900 0.744 0.124 e
0.450 0.540 0.720 1.097 1.174 1.295 1.328 1.200 0.120 BDLa
Reference (Ekinci et al., 2002)
(Budinova et al., 2003)
e
0.083 1.100 0.902 0.520 0.420 0.290 e
1.100 0.600 1.300
(Budinova et al., 2006)
(Zabihi et al., 2009)
(Lu et al., 2014)
(Kadirvelu et al., 2004) (Yardim et al., 2003) (Budinova et al., 2008)
Sulfur impregnation of (AC) with 4%(v/v) SO2 at 700 C for 60 min Steam activation of bagasse pith (SA-C)
Langmuir
Steam activation of bagasse pith in presence of SO 2 (SA SO2 C) e
e
Steam activation of bagasse pith in presence of H 2S (SA H2S C) e
e
Steam activation of bagasse pith in presence of SO 2 & H2S (SA SO2 H2S C) e
e
e
Rice husk ash
Langmuir and Freundlich
Sulfuric acid treated with rice husk (dry sorbent)
Langmuir and Freundlich
Sulfuric acid treated with rice husk (wet sorbent) Sulfuric acid treated with flax shave (dry sorbent) Sulfuric acid treated with flax shave (wet sorbent) Commercial activated carbon (AC)
Steam activated AC (AC-1)
AC1 was oxidized with H O
Langmuir
45 60 30 45 60 30 40 50 60 30 40 50 60 30 40 50 60 30 40 50 60 15 30 25 35 45 25 35 45 45
472 578 523 496 510 172.4 181.8 200 208.3 185.2 204.1 208.3 222.2 181.8 200 204.1 217.4 188.7 208.3 212.8 227.3 9.3 6.7 303 336.7 384.6 227.3 270.3 303 416
0.0090 0.0080 0.0220 0.0740 0.0860 0.0072 0.008 0.0086 0.0106 0.0195 0.0202 0.0229 0.0262 0.0113 0.0123 0.0164 0.0229 0.0281 0.0273 0.0367 0.0480 0.0115 0.0158 0.0052 0.0107 0.0219 0.0052 0.0088 0.0129 0.0805
0.9800 0.8900 0.9600 0.9200 0.9500
15.9 15 71.2 129.2 148.2
e
0.9868 0.9900 0.9990 0.9992 0.9988 0.9991 0.9998 0.9993 0.9980
0.560 0.600 0.330 0.250 0.230
0.940 0.890 0.990 0.980 0.960
e
0.42 0.54 7.1 28.7 48.9 8.7 15.0 18.1
0.493 0.469 0.5579 0.377 0.3407 0.4607 0.428 0.448
751.3
e
0.973 0.965 0.9812 0.987 0.986 0.9523 0.928 0.959
(Anoop Krishnan and Anirudhan, 2002)
311
(Feng et al., 2004)
66
(El-Shafey, 2010)
e
19
(Cox et al., 2000)
w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 3 7 e
344 Langmuir and Freundlich
10 25 50 10 25 50 10
4.1 3.5 3 4.9 4.5 4.1 50
0.0468
0.9990
0.2083 0.1053 0.1330 0.6870 0.1540 0.3862 0 2329
0.9522 0.9827 0.9828 0.9889 0.9914 0.9866 0 9828
5 5
e
0.8 0.3 0.4 2.8 0.6 1.4 11
0.974 1.262 1.150 0.790 1.248 0.750 0 914
0.956 0.971 0.970 0.975 0.982 0.952 0 986
797
870
829
(Lu et al., 2014)
Sulfur impregnation of (AC) with 4%(v/v) SO2 at 700 C for 60 min Steam activation of bagasse pith (SA-C)
Langmuir
Steam activation of bagasse pith in presence of SO 2 (SA SO2 C) e
e
Steam activation of bagasse pith in presence of H 2S (SA H2S C) e
e
Steam activation of bagasse pith in presence of SO 2 & H2S (SA SO2 H2S C) e
e
e
Rice husk ash
Langmuir and Freundlich
Sulfuric acid treated with rice husk (dry sorbent)
Langmuir and Freundlich
Sulfuric acid treated with rice husk (wet sorbent) Sulfuric acid treated with flax shave (dry sorbent) Sulfuric acid treated with flax shave (wet sorbent) Commercial activated carbon (AC)
Langmuir
Activated coke (AC) Thiol-functionalized activated coke (SH-AC)
50
472 578 523 496 510 172.4 181.8 200 208.3 185.2 204.1 208.3 222.2 181.8 200 204.1 217.4 188.7 208.3 212.8 227.3 9.3 6.7 303 336.7 384.6 227.3 270.3 303 416
0.0090 0.0080 0.0220 0.0740 0.0860 0.0072 0.008 0.0086 0.0106 0.0195 0.0202 0.0229 0.0262 0.0113 0.0123 0.0164 0.0229 0.0281 0.0273 0.0367 0.0480 0.0115 0.0158 0.0052 0.0107 0.0219 0.0052 0.0088 0.0129 0.0805
0.9800 0.8900 0.9600 0.9200 0.9500
15.9 15 71.2 129.2 148.2
e
0.9868 0.9900 0.9990 0.9992 0.9988 0.9991 0.9998 0.9993 0.9980
0.560 0.600 0.330 0.250 0.230
0.940 0.890 0.990 0.980 0.960
e
0.42 0.54 7.1 28.7 48.9 8.7 15.0 18.1
0.493 0.469 0.5579 0.377 0.3407 0.4607 0.428 0.448
751.3
e
0.973 0.965 0.9812 0.987 0.986 0.9523 0.928 0.959
(Anoop Krishnan and Anirudhan, 2002)
311
(Feng et al., 2004)
66
(El-Shafey, 2010)
e
19
(Cox et al., 2000)
w a t e r r e s e a r c h 7 3 ( 2 0 1 5 ) 3 7 e
Langmuir and Freundlich
Steam activated AC (AC-1)
AC1 was oxidized with H 2O2 at 1:2 (m:v) (AC1-1) AC1 was oxidized with H 2O2 at 2:5 (m:v) (AC1-2) PSAC grafted with TOMATS
45 60 30 45 60 30 40 50 60 30 40 50 60 30 40 50 60 30 40 50 60 15 30 25 35 45 25 35 45 45
Langmuir and Freundlich
Langmuir and Freundlich
10 25 50 10 25 50 10 25 50 10 25 50 20 25 30 35 25
344
0.0468
0.9990
4.1 3.5 3 4.9 4.5 4.1 5.0 5 4.6 5.2 5.1 4.6 76.9 76.9 83.3 83.3 315.8 694.9
0.2083 0.1053 0.1330 0.6870 0.1540 0.3862 0.2329 0.2000 0.2618 0.7795 0.4491 0.9287 0.0588 0.0778 0.1100 0.1250 0.0500 0.0600
0.9522 0.9827 0.9828 0.9889 0.9914 0.9866 0.9828 0.9914 0.9949 0.9835 0.9841 0.9521 0.9920 0.9940 0.9930 0.9910 0.9770 0.9840
5 5
e
0.8 0.3 0.4 2.8 0.6 1.4 1.1 1 1.6 3.4 2.3 2.7 5.5 6.9 9.4 10.2 12.2 71.1
0.974 1.262 1.150 0.790 1.248 0.750 0.914 0.858 0.857 0.823 1.031 0.720 0.630 0.590 0.560 0.570 0.714 0.526
0.956 0.971 0.970 0.975 0.982 0.952 0.986 0.984 0.985 0.976 0.979 0.980 0.916 0.906 0.892 0.898 0.980 0.954
797
(Lu et al., 2014)
870
829
825
107.7
(Ismaiel et al., 2013)
136.7 193.7
(Li et al., 2013)
water research 73 (2015) 37 55
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 the sulfurto 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 significant 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 activation 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
e
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 M hydrochloric 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 hypothesized 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 gas solid 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 hine
4 9
50
water research 73 (2015) 37 55 e
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 the sulfurto 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 significant 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 activation 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 as OH, CH O and COOH without any treatment. 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 surface 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 act as 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 hydroxyl, carboxylic and carboxylic anhydride group by nitric acid treatment (Zhu et al., 2009). When the contact time between activated carbon and the nitric acid increases, significant amount of phenolic group forms while the content of lactone group is reduced. Increase in the concentration of nitric 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 acidtreated 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 e
e
]
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 M hydrochloric 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 hypothesized 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 gas solid 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 hindrance between the adsorbed molecules on the adjacent sites (Langmuir, 1918). The mathematical expression for this model can be represented as: e
qe ¼
qm aL Ce
1 þ aL Ce
(1)
where q e is the adsorption capacity of the adsorbent (mg :g 1 ); qm is the maximum adsorption capacity of the adsorbent (mg :g 1 ); C e is the equilibrium concentration of the adsorbate in the solution (mg :L1 ); and a L is the Langmuir constant. A mathematical expression developed by Webber and Chakkravorti has been used to test the favourability of the adsorption process (Weber and Chakravorti, 1974): RL ¼
1 1 þ aL C0
(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 ¼ 1), favorable (0 < R L < 1), or irreversible (RL ¼ 0). The Freundlich model, on the other hand, has been used to fit a heterogeneous surface with non-identical adsorption sites (Freundlich, 1906). This empirical model can be applied to multilayer adsorption with uneven distribution of heat and affinities when adsorption occurs over the heterogeneous surface (Adamson andGast, 1967). Although Freundlich model is a widely-used empirical isotherm model to fit the experimental adsorption results, it has been criticized for its inability to comply with the Henry's law at low adsorbate concentrations (Ho et al., 2002). The expression of the Freundlich isotherm can be described as: 1
qe ¼ K F Cne
(3)
where K F is the Freundlich constant and 1=n is the heterogeneity factor. When the value of n is between 1 and 10, the
water research 73 (2015) 37 55 e
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 and Freundlich adsorption models are also established. RedlichPeterson 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 homogeneous and heterogeneous adsorption (Redlich and Peterson, 1959): qe ¼
qm aR Ce 1 þ aR Cbe
(4)
where a R 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 this equation 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 heterogeneous adsorption system and assumes the occurrence of dissociative adsorption (Sips, 1948). At low adsorbate concentration, the isotherm can be reduced to Freundlich isotherm, while at high concentration, it complies with homogenous 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. The Sips isotherm model is expressed through the following equation: qe ¼
qm aS Cbe s 1 þ aS Cbe s
(5)
where as is the Sips isotherm model constant and bs isthe Sips isotherm model exponent. Table 4 summarizes the Langmuir and Freundlich constants 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 mussel shell 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 ( PenaRodrı´ guez 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 surface area (Cai and Jia, 2010). This phenomenon suggested that the surface area alone is not sufficient to reflect the ~
51
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 significant reduction in SBET of the activated carbons, a staggering 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 classification (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 effects 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 demonstrated that a combination of medium-to-high surface area with well-functionalized surface properties are collectively crucial in enhancing the mercury removal. Despite these findings, very limited research has been carried out on simultaneous optimization of surface and textural characteristics of activated carbons. Hence, further study is necessary for such optimization to save the high amount of energy required to obtain unnecessary very high surface area activated 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 functionalization process, and consequently the mercury adsorption mechanism 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 strategy has been disregarded. Further recommendations can be presented as follows:
52
water research 73 (2015) 37 55 e
Mercury forms complexes in real wastewater systems and
is rarely found in ionic form. Therefore, study of the efficiency 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 to be carried out. The regeneration of the activated carbon samples has to be conducted for economic feasibility enhancement. Unfortunately, in adsorption processes, the regeneration is overlooked. If not regenerated, due to its hazardous nature, mercuryloaded activated carbon needs to be stabilized or vitrified and then disposed of in hazardous landfill. Nonetheless, this landfilled carbon is not reusable, therefore emphasizing the importance of the production cost of the activated 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
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