Investigation of the sorption behavior of Cd(II) on GMZ bentonite as affected by solution chemistry

Chemical Engineering Journal 166 (2011) 1010–1016

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Chemical Engineering Journal journal homepage: www.elsevier.com/locate/cej

Investigation of the sorption behavior of Cd(II) on GMZ bentonite as affected by solution chemistry Donglin Zhao a,b , Shaohua Chen a , Shubin Yang b , Xin Yang b , Shitong Yang b,∗ a b

School of Materials and Chemical Engineering, Anhui University of Architecture, Hefei 230601, PR China Key Laboratory of Novel Thin Film Solar Cells, Institute of Plasma Physics, Chinese Academy of Sciences, P.O. Box 1126, Hefei 230031, PR China

a r t i c l e

i n f o

Article history: Received 25 September 2010 Received in revised form 22 November 2010 Accepted 23 November 2010 Keywords: Cd(II) GMZ bentonite Sorption Coexisting electrolyte ions Humic substances

a b s t r a c t Clay minerals have been studied extensively due to their strong sorption and complexation ability towards various environmental pollutants. In this study, the removal of cadmium from wastewaters by GMZ bentonite was studied as a function of various solution chemistry conditions such as pH, ionic strength, coexisting electrolyte ions, humic substances and temperature under ambient conditions. The results indicated that the sorption of Cd(II) on GMZ bentonite was strongly dependent on pH and ionic strength. Langmuir and Freundlich models were used to simulate the sorption isotherms of Cd(II) at three different temperatures of 298, 318 and 338 K. The thermodynamic parameters (H◦ , S◦ and G◦ ) calculated from the temperature dependent sorption isotherms indicated that the sorption process of Cd(II) on GMZ bentonite was endothermic and spontaneous. At low pH, the sorption of Cd(II) was dominated by outer-sphere surface complexation and ion exchange with Na+ /H+ on GMZ bentonite surfaces, whereas inner-sphere surface complexation was the main sorption mechanism at high pH. From the experimental results, one can conclude that GMZ bentonite may have good potentialities for the disposal of cadmium bearing wastewaters. © 2010 Elsevier B.V. All rights reserved.

1. Introduction In recent years, the levels of toxic metal ions in surface waters have increased dramatically due to the pollution caused by industrial and agricultural wastewater discharges. Among these heavy metal ions, cadmium is a non-essential and highly toxic heavy metal element that can be released into the environment by various ways such as metallurgical alloying, photograph development, textile printing industries, ceramics, alkaline batteries and electroplating, metal plating, pigment works [1]. As cadmium is present stably in the environment with great difficulty for microbiological degradation, continuous accumulating of cadmium concentration can cause severe health hazards on human including muscle cramp, diarrhea, hypertension, yellow coloration of teeth (cadmium ring formation), loss of calcium from bones, loss of sense of smell after inhalation, kidney failure following oral ingestion, nausea, vomiting, chest pain, damage of bone marrow and reduction of red blood cells. For the sake of ecosystem stability and public health, it is of great imminence to eliminate cadmium from polluted water bodies or at least decrease the concentration of cadmium to the permissible limits before its discharge to the environment.

∗ Corresponding author. Tel.: +86 551 5591368. E-mail address: [email protected] (S. Yang). 1385-8947/$ – see front matter © 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.cej.2010.11.092

Sorption, diffusion and precipitation reactions at clay–water interfaces are closely related with the migration, transformation, accumulation and biological effectiveness of heavy metal ions in soil and aquatic environment. In recent years, the sorption behaviors of metal ions in clay–water systems have been studied extensively and many mechanisms have been postulated such as ion exchange, outer-sphere/inner-sphere surface complexation, diffusion into particle micropores and precipitation [2–5]. Most of the studies presented on this subject were restricted to the sorption of single metal ions in the absence of other electrolyte ions. However, the metal of interest in wastewater is usually found to be in a matrix containing various electrolyte ions. The interaction between these charged particles leads to the formation of an electrostatic field with certain intensity around them, which will further modify the behavior of heavy metal ions towards the sorbent materials. The metal ions of interest in wastewater are usually found to be in a matrix containing various inorganic ions. Cations in the multi-component system may compete for the binding sites of solid particles owing to the difference in their affinities on sorbent. Besides, some anions can form insoluble or soluble complexes with the metal ions, displacing with difficulty in the presence of sorbent [6,7]. Therefore, to better understand the practical application of clays as potential sorbents in sewage disposal, it is of great importance to investigate the sorption behavior of heavy metal ions on clays in the presence of various coexisting electrolyte ions.

D. Zhao et al. / Chemical Engineering Journal 166 (2011) 1010–1016

2. Experimental 2.1. Materials and reagents The sample of GMZ bentonite was obtained from Gaomiaozi county (Inner Mongolia, China). It was converted into Na–bentonite by treating with 1.0 M NaCl, then washing with doubly distilled water until it was free from chloride ions. Then the sample was dried and ground to 53 ␮m. In the following sections, we call the GMZ Na–bentonite simply GMZ bentonite. The N2 -BET surface area of the sample was 29.5 m2 /g and the zero point charge (pHzpc ) was measured to be 6.3 ± 0.1 by using potentiometric titration method at different ionic strengths [15,16]. Humic acid (HA) and fulvic acid (FA) were extracted from the soil of Hua-Jia county (Gansu province, China) and has been characterized in detail [17,18]. The concentrations of functional groups of HA and FA were determined by fitting the potentiometric titration data using FITEQL 3.1 and the surface site densities were calculated to be 6.46 × 10−3 and 2.71 × 10−2 mol/g for HA and FA, respectively [17]. All chemicals used in the experiments were purchased in analytical purity and used without any purification. Milli-Q water was used to prepare all the solutions in the experiments. 2.2. Experimental procedure All the experiments were carried out by using batch technique in polyethylene centrifuge tubes under ambient conditions. Except for the sorption isotherms, all the other experiments were carried out in an initial Cd(II) concentration of 8.90 × 10−5 mol/L. The stock solutions of GMZ bentonite and NaNO3 were pre-equilibrated for 24 h and then Cd(II) stock solution was added to achieve the desired concentrations of these three components. The pH of each test solu-

100

I=0.001 M NaNO 3 I=0.01 M NaNO 3 I=0.1 M NaNO3

90 80

Sorption (%)

Mineral surfaces are often bound to a large extent of natural organic materials such as humic substances (HS), which is well known to exert strong effect on the sorption properties of metal ions in water and soil systems [8–10]. The influence of HS on the sorption of metal ions to minerals is dependent on the nature of minerals, nature of metal ions, nature of HS, solution pH, etc. The presence of HS can affect the behavior of metal ions by the formation of metal–HS complexes in solution or on solid surfaces, and thus plays an important role in the migration, transformation, sequestration and biological effectiveness of metal ions in the natural environment [11]. Hence, it is essential to study the effect of HS on the sorption behaviors of heavy metal ions on clay minerals. Bentonite, with various outstanding physicochemical properties, i.e., large specific area, low permeability, low cost, strong adsorptive affinity for organic and inorganic pollutants, high cation exchange capacity, accessibility and ubiquitous presence in most soils, has been reported as a main candidate in the decontamination of water bodies polluted by detrimental metal ions [12–14]. In China, the bentonite in Gaomiaozi county (herein abbreviated as GMZ bentonite) has been selected as the candidate of backfill material for nuclear waste repositories and hazardous chemicals. However, according to our literature survey, few studies were focused on the uptake of Cd(II) on GMZ bentonite as effected by humic substances and coexisting electrolyte ions. Hence, the GMZ bentonite was used in this study as a sorbent for the removal of Cd(II) from wastewater. The basic objectives of this paper are: (1) to study the effect of different environmental parameters on Cd(II) sorption, such as pH, ionic strength, coexisting electrolyte ions, humic substances and temperature by using batch technique; and (2) to presume the sorption mechanism of Cd(II) on GMZ bentonite and to estimate the possible application of GMZ bentonite in wastewater disposal.

1011

70 60 50 40 30 20 10 0 2

3

4

5

6

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8

9

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pH Fig. 1. Effect of ionic strength on Cd(II) sorption to GMZ bentonite as a function of pH values. T = 298 ± 2 K, C(Cd)initial = 8.90 × 10−5 mol/L, m/V = 0.5 g/L.

tion was adjusted to desired values by adding negligible volumes of 0.01 or 0.1 M HNO3 or NaOH. The suspensions were oscillated for 24 h and then centrifuged at 9000 rpm (7788 × g) for 30 min to separate the solid from liquid phases. It was necessary to note that the sorption of Cd(II) on the tube wall was negligible according sorption tests performed in the absence of GMZ bentonite. The concentration of Cd(II) was analyzed by spectrophotometry at wavelength of 578.4 nm by using Cd xylenol orange complex. The percentage of Cd(II) adsorbed on GMZ bentonite (sorption% = (C0 − Ce )/C0 × 100%) and distribution coefficient (Kd (mL/g) = (C0 − Ce )/Ce V/m) were derived from the difference of initial concentration (C0 ), the final concentration (Ce ) in supernatant after centrifugation, the mass of GMZ bentonite (m) and the volume of the suspension (V). All experimental data were the average of triplicate determinations and the relative errors were about 5%. 3. Results and discussion 3.1. Effect of ionic strength The effect of ionic strength on Cd(II) sorption on GMZ bentonite as a function of pH values is shown in Fig. 1. One can see that the sorption of Cd(II) is strongly affected by ionic strength at pH < 7.5 and no distinct effect can be found at pH > 7.5. The ionic strength can influence the interface potential and thickness of double electrode layer, and further affect the binding of the adsorbed species. Outer-sphere surface complexes are expected to be more impressionable to ionic strength variations than inner-sphere complexes as the background electrolyte ions are placed in the same plane for outer-sphere surface complexes [8]. Based on the theory mentioned above, one can deduce that cation exchange or outer-sphere surface complexation mainly contribute to Cd(II) sorption on GMZ bentonite at pH < 7.5, while inner-sphere surface complexation is the main sorption mechanism of Cd(II) on GMZ bentonite at pH > 7.5 [19,20]. As can be seen from Fig. 1, the sorption of Cd(II) on GMZ bentonite decreases with increasing ionic strength at pH < 7.5. This phenomenon may be attributed to the reasons stated as follows: (1) the increase of ionic strength decreases the activity of Cd(II) ions and then limits their transfer to GMZ bentonite surfaces [21]; (2) Cd(II) ions can form electrical double layer complexes with GMZ bentonite, which favors the sorption when the concentration of the competing salt is decreased. This may indicate that the sorption

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Cd

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pH Fig. 2. Sorption of Cd(II) to GMZ bentonite as a function of pH values. T = 298 ± 2 K, C(Cd)initial = 8.90 × 10−5 mol/L, m/V = 0.5 g/L, I = 0.01 M NaNO3 .

interaction between the functional groups of GMZ bentonite (i.e., ≡SOH and ≡SONa) and Cd(II) ions is mainly an ionic interaction in nature, which is in agreement with the ion exchange mechanism [22]; (3) ionic strength effects can impact particle aggregation by affecting electrostatic interactions. Increased ionic strength has been shown to reduce electrostatic repulsion and thereby increases particle aggregation of GMZ bentonite, which reduces the amount of available binding sites and thereby decreases the sorption of Cd(II) on GMZ bentonite [23].

cant impact on the removal efficiency of GMZ bentonite towards Cd(II). The relative proportion of Cd(II) species are calculated from the stability constants (pK1 = 7.9, pK2 = 10.6 and pK3 = 14.3) [25] and the results demonstrate that Cd(II) presents in the form of Cd2+ , Cd(OH)+ , Cd(OH)2 0 and Cd(OH)3 − at various pH values (inset in Fig. 2). At pH < 6.0, the predominant Cd(II) species is Cd2+ and the removal of Cd(II) is mainly accomplished by sorption reaction. The sorption of Cd(II) that takes place at pH < 6.0 (region I) can be attributed to the competition between Cd2+ and H+ /Na+ on the surface ion exchange sites. It is necessary to make certain whether the formation of Cd(OH)2 (s) precipitation contributes to the rapid increase of Cd(II) sorption on GMZ bentonite in pH range 6.0–9.0. The precipitation curve of Cd(II) calculated from the precipitation constant of Cd(OH)2 (s) (Ksp = 2.50 × 10−14 ) and the initial Cd(II) concentration (i.e., 8.90 × 10−5 mol/L) is also shown in Fig. 2. One can see that Cd(II) begins to form precipitation at pH ∼ 9.1 in the absence of GMZ bentonite. However, more than 90% Cd(II) is adsorbed on GMZ bentonite at pH 9.0. Thereby, it is impossible to form precipitation because of the very low concentration of Cd(II) remained in solution (herein, ∼8.90 × 10−6 mol/L). Therefore, the abrupt increase of Cd(II) sorption on GMZ bentonite in the pH range 6.0–9.0 (region II) is not attributed to the formation of Cd(OH)2 (s). The fast increasing of Cd(II) sorption may be attributed to the formation of inner-sphere surface complexes. The mechanism can be tentatively explained by the following reactions: I) The hydrolysis of Cd2+ in solution: Cd2+ + nH2 O → Cd(OH)(H2 O)n−1 + + H+

(1)

II) To form complexes with hydrolyzed species: 3.2. Effect of pH In 0.01 M NaNO3 electrolyte solution, the sorption of Cd(II) on GMZ bentonite as a function of pH in can be divided into three regions (Fig. 2): (1) the uptake of Cd(II) increases gradually from zero to ∼25% in the pH range of 2.5–6.0 (region I); (2) in region II, from pH 6.0–9.0, the sorption of Cd(II) on GMZ bentonite increases sharply from ∼25% to a maximum value of ∼97%; and (3) at pH > 9.0, the removal of Cd(II) maintains the high level at ∼97% (region III). The sorption edges spread over three pH units implies the formation of multifarious surface complexes and represents different sorption mechanisms [24]. Two mechanisms are considered in the sorption experiments: cation exchange and inner-sphere complexation. The cation exchange process happens at the planar sites, and the formation of inner-sphere complexes happens at the clay particle edges where the metal ions are retained by Si–O− and Al–O− groups. The increase of Cd(II) sorption on GMZ bentonite with increasing solution pH may be attributed to the surface properties of GMZ bentonite in terms of surface charge and dissociation of functional groups. As is mentioned above, the pHzpc of GMZ bentonite is about 6.3. In the pH range lower than the pHzpc , the surface charge of GMZ bentonite is positive due to the protonation reaction on the surfaces (i.e., SOH + H+ ⇔ SOH2 + ). The electrostatic repulsion occurred between Cd(II) ions and the edge groups with positive charge (SOH2 + ) on GMZ bentonite surface leads to the low sorption percentage of Cd(II) at pH < 6.0. At pH > pHzpc , the surface of GMZ bentonite becomes negatively charged due to the deprotonation process (i.e., SOH ⇔ SO− + H+ ) and electrostatic repulsion decreases with raising pH due to the reduction of positive charge density on the sorption edges, resulting in the electrostatic attraction of Cd(II) to the deprotonated surface of GMZ bentonite. Moreover, as more surface functional groups are dissociated at high pH values than that at low pH values, more sorption sites are available for binding Cd(II) ions. The exact speciation of Cd(II) complexes that predominates at a particular solution pH may also have a signifi-

S–ONa + Cd(OH)(H2 O)n−1 + → 2 S–ONa + Cd(OH)(H2 O)n−1 + →

S–OCdOH + Na+ + (n − 1)H2 O (S–O)2 Cd(OH)− + 2Na+ + (n − 1)H2 O

(2) (3)

The above-mentioned inner-sphere surface complexation reactions are a tentative interpretation for the sorption mechanism of Cd(II) on GMZ bentonite at high pH values. However, it is difficult to distinguish the specific form of inner-sphere surface complexes (i.e., monodentate or bidentate surface complexes) by macroscopic experiment results. Hence, in our further study, we will carry out a more detailed investigation of the specific complex forms of Cd(II) on GMZ bentonite in molecular level by using EXAFS technique. Depending on the pH and Cd(II) initial concentration, the hydrolytic actions of Cd(II) may generate various species such as Cd(OH)+ , Cd(OH)2 (s), and Cd(OH)3 − at higher pH values (pH > 9.0) and thus the removal of Cd(II) in this pH range is possibly accomplished by simultaneous precipitation of Cd(OH)2 (s) and sorption of Cd(OH)+ and Cd(OH)3 − on GMZ bentonite. 3.3. Effect of solid content The dependence of Cd(II) sorption on GMZ bentonite at different solid contents was studied at T = 298 K and pH = 6.5 ± 0.1. As can be seen from Fig. 3, the sorption percentage of Cd(II) increases rapidly with increasing GMZ bentonite contents at m/V < 1.0 g/L, and then maintains unchanged with the increase of solid content at m/V > 1.0 g/L. With increasing solid contents, the number of functional groups at GMZ bentonite surfaces increases, thereby, more exchangeable surface sites are available to form complexes with Cd(II). One can also see from Fig. 3 that the sorption capacity of Cd(II) on GMZ bentonite (qe ) decreases gradually with the increase of solid content. At low solid content, Cd(II) can easily access the sorption sites of GMZ bentonite and qe is high. At high solid content, a large amount of GMZ bentonite effectively reduces the unsaturation of the sorption sites and results in the overcrowding of solid particles,

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which correspondingly decreases the sorption capacity of Cd(II) on GMZ bentonite [26]. Besides, higher solid content increases the probability of collision between solid particles and therefore creates particle aggregation, causing a decrease in the total surface area and an increase in diffusional path length, both of which contribute to the decrease in the sorption capacity of Cd(II) on GMZ bentonite. 3.4. Effects of coexisting electrolyte ions Fig. 4 shows the sorption of Cd(II) on GMZ bentonite as affected by different electrolyte ions, viz., NO3 − , Cl− , CO3 2− , SO4 2− , Na+ , K+ , Mg2+ and Ca2+ , respectively. The reason to choose these ions is based on their widespread distribution in natural water and industrial wastewaters. To facilitate the comparison of results, all data in the present study were obtained with sodium salt for anions and nitrate for cations. As shown in Fig. 4A, the sorption of Cd(II) on GMZ bentonite is greatly influenced by electrolyte anions. Taking the sorption of Cd(II) in NO3 − electrolyte solution as a point of reference, one can see that the presence of Cl− and SO4 2− enhances Cd(II) sorption at pH < 6.0, while reduces Cd(II) sorption at pH > 6.0. In contrast, the presence of CO3 2− enhances Cd(II) sorption at pH < 9.0 and has no obvious influence on Cd(II) sorption at pH > 9.0. Briefly, electrolyte anions may either enhance or inhibit the sorption of metal ions on solid particles through processes including ternary complex formation (solid particle–metal ion–electrolyte anion or solid particle–electrolyte anion–metal ion) [27], site competition [28], formation of solution complexes, formation of surface precipitation and/or alteration of surface charge [29]. Specifically, the phenomenon that Cd(II) uptake at pH < 6.0 is enhanced in the presence of Cl− is to some extent attributed to the formation of complex between Cd(II) ions and Cl− (e.g., CdCl+ ), which decreases the electropositivity of Cd(II) ions and thereby promotes the sorption of positively Cd(II) ions on GMZ bentonite due to the decrease of electrostatic repulsion. Enhanced affinity of Cd–Cl− complex relative to the free Cd2+ ions for the GMZ bentonite surface is attributable at least in part to reducing the electrostatic barrier that Cd(II) must overcome when adsorbed to the positive GMZ bentonite surfaces (pH < pHzpc ∼6.3). In SO4 2− electrolyte solution, the enhancement of Cd(II) sorption on GMZ bentonite at pH < 6.0 may be attributed to the formation of ternary GMZ bentonite–Cd–SO4 2− or GMZ bentonite–SO4 2− –Cd complexes. Besides, the idiocratic sorption of SO4 2− on GMZ bentonite enhances the electronegativity of GMZ bentonite surface and thereby promotes the sorption of positively charged Cd(II) ions due to electrostatic attraction.

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Fig. 3. Sorption of Cd(II) on GMZ bentonite as a function of solid content. T = 298 ± 2 K, pH = 6.5 ± 0.1, C(Cd)initial = 8.90 × 10−5 mol/L, I = 0.01 M NaNO3 .

70 60 50 40 30 20 10 0 -10 2

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pH

Fig. 4. Influence of coexisting electrolyte anions (A) and cations (B) ions on the sorption of Cd(II) on GMZ bentonite as a function of pH values. T = 298 ± 2 K, C(Cd)initial = 8.90 × 10−5 mol/L, m/V = 0.5 g/L.

At high pH values, Cl− and SO4 2− are difficult to be adsorbed on the negatively charged surfaces of GMZ bentonite due to electrostatic repulsion. The competition between Cl− /SO4 2− and GMZ bentonite increases the formation of Cd(II)–Cl− /SO4 2 stable complexes in solution, which competitively diminishes the extent of Cd(II) sorption on GMZ bentonite. Compared with Cl− and SO4 2− , CO3 2− has a different influence on the sorption curve of Cd(II) on GMZ bentonite. Briefly, Cd(II) ions can form various complexes with CO3 2− such as Cd(CO3 )2 2− , CdHCO3 − and CdCO3 at different pH values [30], affecting the migration and transformation of Cd(II) ions in aquatic systems. Meanwhile, CO3 2− can be adsorbed on GMZ bentonite surface via complexation or chelating reactions, which will further influence the sorption of Cd(II) ions on GMZ bentonite. Based on the above-mentioned theories, at pH < 7.0, the enhancement of Cd(II) sorption in CO3 2− electrolyte solution may be attributed to the formation of ternary GMZ bentonite–Cd–CO3 2− or GMZ bentonite–CO3 2− –Cd complex. However, the specific form of the ternary complexes is difficult to be determined from the macroscopic experiment data and further investigation is needed to obtain the microstructure information. In the pH range of 7.0–9.0, the sorption percentage of Cd(II) on GMZ bentonite rises to ∼100%, which may result from the formation of CdCO3 (s) precipitation. From Fig. 4A, one can also see that the enhancement force of different anions towards Cd(II) sorption at low pH values is in the sequence of Cl− < SO4 2− < CO3 2− . The difference is proportional to

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Fig. 5. Effect of pH on sorption of Cd(II) on bare and HA/FA bound GMZ bentonite. T = 298 ± 2 K, C(Cd)initial = 8.90 × 10−5 mol/L, C[HA/FA] = 10 mg/L, m/V = 0.5 g/L, I = 0.01 M NaNO3 .

their complex stability constants with Cd(II) ions (i.e., Cl− , 1.98; SO4 2− , 2.4; CO3 2− , 4.02) [6]. As can be seen from Fig. 4B, the sorption of Cd(II) on GMZ bentonite is strongly influenced by the electrolyte cations. The competition effects of Ca2+ and Mg2+ towards Cd(II) sorption are somewhat important that those of Na+ and K+ . This phenomenon can be easily interpreted by the competition of different cations with Cd(II) for binding on the surface functional groups of GMZ bentonite. The sorption ability of the four cations on GMZ bentonite is in the sequence of Na+ < K+ < Mg2+ < Ca2+ , and the higher valence ion is much easier to be adsorbed by GMZ bentonite, thereby the influences of Ca2+ and Mg2+ on Cd(II) sorption are stronger than those of Na+ and K+ . Generally, the cations may follow an order of increasing selectivity for binding to the functional groups of solid surfaces as follows: alkali metal cations < H+ < alkaline earth cations < transitional group monovalent cations (e.g. Ag+ ) < transitional group divalent cations (e.g., Cd2+ and Pb2+ ) < trivalent cations (e.g., Eu3+ ) [31]. Thus, the influence of alkali metal and alkaline earth cations on the sorption of bivalent Cd(II) should be weak. However, in the pH range of 2–7.5, the influences of Na+ , K+ , Mg2+ and Ca2+ on Cd(II) sorption are a little drastic. In the experiments, the concentration of the background electrolyte ions (1.00 × 10−2 mol/L) is much higher than that of Cd(II) (8.90 × 10−5 mol/L). Before the addition of Cd(II) ions, GMZ bentonite has been pre-equilibrated with the background electrolyte ions for 24 h. The sorption of Cd(II) on GMZ bentonite is attributed to ion-exchange reaction of Cd(II) ions with the coexisting cations under the experimental conditions. Thereby, it is reasonable that the coexisting electrolyte cations can affect Cd(II) sorption. 3.5. Effect of humic substances The pH dependence of Cd(II) sorption on GMZ bentonite in the absence and presence of HA/FA (Fig. 5) shows that the presence of HA/FA enhances the sorption of Cd(II) on GMZ bentonite at pH < 8, while inhibits Cd(II) sorption at pH > 8. In previous studies, HA and FA were reported to have negative zeta potentials at pH > 2 [32,33]. Therefore, at low pH values, the negatively charged HA/FA can be easily adsorbed on the positively charged surfaces of GMZ bentonite due to electrostatic attraction. Actual sorption of HA/FA on GMZ bentonite surface modifies both the long-range electrostatic properties of the aqueous-mineral interface as well as the con-

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1.8x10

Ce (mol/L) Fig. 6. Sorption isotherm, Langmuir (solid lines) and Freundlich (dashed lines) fits of Cd(II) sorption on GMZ bentonite at three different temperatures. pH = 6.5 ± 0.1, C(Cd)initial = 1.78 × 10−5 to 2.14 × 10−4 mol/L, m/V = 0.5 g/L, I = 0.01 M NaNO3 .

centration and molecular characteristics of specific metal-binding sites present, which results in a more favorable electrostatic environment for Cd(II) sorption and enhances the formation of type B “ligand-bridging” ternary surface complexes where the Cd(II) ions are complexed by humic molecules that are simultaneously adsorbed on GMZ bentonite surface [34]. However, at high pH values, the negatively charged HA/FA is difficult to be adsorbed on the negatively charged surfaces of GMZ bentonite due to electrostatic repulsion. The competition between the soluble HA/FA and GMZ bentonite with Cd(II) increases the formation of strong HA/FACd complexes in solution, which diminishes the extent of Cd(II) sorption on HA/FA-GMZ bentonite hybrids [35]. From Fig. 5, one can also see that the influence of FA on the sorption of Cd(II) is stronger than that of HA at the same mass concentrations of HA and FA. Although the HA and FA samples were extracted from the same soil sample and both of them contain functional groups such as hydroxyl, carboxyl, amine and phenolic, the proportions and configurations of these functional groups and surface site densities are different. In a previous literature, the acidic functional group contents of HA and FA isolated from LSM, unamended and amended soils were measured by a current potentiometric titration method and the results showed that FA had bigger total acidity, larger site density of both carboxylic- and phenolic-type groups than HA [36]. As mentioned above, the surface site density of FA (i.e., 2.71 × 10−2 mol/g) is higher than that of HA (i.e., 6.46 × 10−3 mol/g), which means that FA can provide more available surface sites for binding Cd(II). Furthermore, the functional groups of FA such as –OH and –COOH would be ionized as pH increased, leading to the disappearing of these hydrogen-bond donors of FA and the increase of FA solubility (i.e., the decrease of hydrophobic effects) [33]. Therefore, FA has a stronger effect on Cd(II) sorption than HA in the whole pH range. 3.6. Sorption isotherms and thermodynamic studies Fig. 6 shows the sorption isotherms of Cd(II) on GMZ bentonite at three different temperatures (viz. 298, 318 and 338 K). It is clear that the sorption isotherm is the highest at T = 338 K and is the lowest at T = 298 K. The result indicates that high temperature is advantageous for Cd(II) sorption on GMZ bentonite. Several factors may account for the increase of Cd(II) sorption on GMZ bentonite with increasing temperature. Increased diffusion rate of Cd(II) into the

D. Zhao et al. / Chemical Engineering Journal 166 (2011) 1010–1016 Table 2 Values of thermodynamic parameters for Cd(II) sorption on GMZ bentonite.

Table 1 The parameters for the two isotherm models at different temperatures. Correlation parameters Langmuir qmax (mol/g) b (L/mol) CC (R2 ) Freundlich kF (mol1−n Ln /g) n CC (R2 )

T = 298 K

T = 318 K

T = 338 K

9.79 × 10−5 1.61 × 104 0.991

1.04 × 10−4 2.59 × 104 0.992

1.17 × 10−4 4.45 × 104 0.990

5.75 × 10−3 0.496 0.982

3.21 × 10−3 0.414 0.974

1.97 × 10−3 0.336 0.970

GMZ bentonite pores due to increased temperature may account for the observed behavior [37]. The increase in temperature may lead to the increase in proportion and activity of Cd(II) ions in solution, the affinity of Cd(II) ions to the surface, or the potential charge of GMZ bentonite surface [38]. For isotherms modeling, herein, Langmuir and Freundlich isotherm equations are conducted to simulate the sorption isotherms and to establish the relationship between the amount of Cd(II) adsorbed on GMZ bentonite and the concentration of Cd(II) remained in solution. The Langmuir model assumes that sorption occurs in a monolayer with all sorption sites identical and energetically equivalent. Its form can be described by the following equation [39]: qe =

bqmax Ce 1 + bCe

(4)

where Ce is the equilibrium concentration of metal ions remained in the solution (mol/L); qe is the amount of metal ions adsorbed on per weight unit of solid after equilibrium (mol/g); qmax , the maximum sorption capacity, is the amount of sorbate at complete monolayer coverage (mol/g), and b (L/mol) is a constant that relates to the heat of sorption. The Freundlich isotherm model is an exponential equation that represents properly the sorption data at low and intermediate concentrations on heterogeneous surfaces. The equation is represented by the following equation [40]: qe = kF Cen

1015

(5)

where kF (mol1−n Ln /g) represents the sorption capacity when the equilibrium concentration of metal ion equals to 1, and n represents the degree of dependence of sorption with equilibrium concentration. The experimental data of Cd(II) sorption are regressively simulated with the Langmuir and Freundlich models and the results are given in Fig. 6. The relative values calculated from the three models are listed in Table 1. As can be seen from Fig. 9, the two isotherm equations fit the sorption isotherms of Cd(II) on GMZ bentonite well, which is supported by the good correlation coefficients (all > 0.90) in Table 1. It can be concluded from the R2 values that Langmuir model simulates the experimental data better than Freundlich model. The fact that the sorption data of Cd(II) according with Langmuir isotherm indicates that the binding energy on the whole surface of GMZ bentonite is uniform. In other words, the whole surface has identical sorption activity and therefore the adsorbed Cd(II) ions do not interact or compete with each other, and they are adsorbed by forming an almost complete monolayer coverage of the GMZ bentonite particles. This phenomenon also indicates that chemosorption is the principal uptake mechanism in sorption process [41]. Moreover, GMZ bentonite has a finite specific surface and sorption capacity, thus the sorption could be better described by Langmuir model rather than by Freundlich model, as an exponentially increasing sorption was assumed in the Freundlich model. At all temperatures, the value of qe was found to be smaller than qmax , which confirms that Cd(II) sorption on GMZ bentonite is by a

T (K)

G◦ (kJ/mol)

S◦ (J/mol/K)

H◦ (kJ/mol)

298 318 338

−16.87 −19.35 −21.44

114.25 114.25 114.25

17.18 16.98 17.18

monolayer type in which the surface of GMZ bentonite is not saturated. The values of qmax obtained from the Langmuir model for Cd(II) sorption on GMZ bentonite are the highest at T = 338 K and the lowest at T = 298 K, which indicates that the sorption process is enhanced with increasing temperature. The value of n calculated from the Freundlich model is from unity, indicating that a nonlinear sorption of Cd(II) takes place on GMZ bentonite surfaces. The thermodynamic parameters (H◦ , S◦ , and G◦ ) for Cd(II) sorption on GMZ bentonite can be determined from the temperature-dependent sorption. Free energy change (G◦ ) is calculated from the relationship: G◦ = −RT ln K ◦

(6)

K◦

where is the sorption equilibrium constant. Values of ln K◦ obtained by plotting ln Kd versus qe for sorption of Cd(II) on GMZ bentonite and extrapolating qe to zero are 6.63 (T = 298 K), 6.95 (T = 318 K) and 7.64 (T = 338 K), respectively. Standard entropy change (S◦ ) is calculated using the equation:



∂G◦ ∂T



= −S ◦

(7)

P

The average standard enthalpy change (H◦ ) is then calculated from the expression: (8)H ◦ = G ◦ + TS ◦ The values obtained from Eqs. (6)–(8) are tabulated in Table 2. A positive value of the standard enthalpy change indicates that the sorption is endothermic. Several factors may account for the observed phenomenon. Firstly, the sorption of Cd(II) requires a diffusion procedure, which is an endothermic process. Secondly, Cd(II) is solved well in water, and the hydration sheath of Cd(II) has to be destroyed before its sorption on GMZ bentonite. This dehydration process needs energy, and it is favored at high temperature [42]. The implicit assumption here is that after sorption the environment of Cd(II) ions is less aqueous than it is in solution. The removal of water from ions is essentially an endothermic process, and it appears that the endothermicity of the desolvation process exceeds that of the enthalpy of sorption by a considerable extent. Furthermore, the sorption is mainly via chemical ion-exchange reactions at pH = 6.5 and the exchange of Cd(II) ions with H+ /Na+ need to break firstly the O–H/O–Na bond, which is also an endothermic process. The Gibbs free energy change (G◦ ) is negative as expected for a spontaneous process under the conditions applied. The value of G◦ becomes more negative with the increase of temperature, which indicates that the reaction is more favorable at higher temperatures. At high temperature, Cd(II) ions are readily desolvated and hence the sorption becomes more favorable. The positive value of entropy change (S◦ ) implies some structural changes in Cd(II) ions and GMZ bentonite during the sorption process, which leads to an increase in the disorderness of the solid–solution system during the uptake of Cd(II) on GMZ bentonite. 4. Conclusions In this study, batch technique was adopted to investigate the sorption of Cd(II) from aqueous solutions on GMZ bentonite as a function of various environmental factors such as pH, ionic strength, solid content, coexisting electrolyte ions, humic substances and temperature under ambient conditions. The sorption

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of Cd(II) is dependent on ionic strength at low pH values, and independent of ionic strength at high pH values. Results obtained from this study show the presence of different electrolyte ions can enhance or inhibit the sorption of Cd(II) on GMZ bentonite in various degrees. The various results are attributed to the different complex abilities of these electrolyte ions with Cd(II) ions and the difference in their affinities on the binding sites of GMZ bentonite. The thermodynamic analysis derived from temperature dependent sorption isotherms suggests that the sorption process of Cd(II) on GMZ bentonite is spontaneous and endothermic. By integrating all the above-mentioned analysis results together, one can conclude that the sorption of Cd(II) on GMZ bentonite is dominated by ion exchange or outer-sphere surface complexation at low pH values, and by inner-sphere surface complexation at high pH values. Considering the low permeability, low cost, accessibility, ubiquitous presence in most soils and large-scale applications of GMZ bentonite, one can conclude that this material has a great application potential for cost-effective disposal of Cd(II)-contaminated wastewaters. Based on the results noted for Cd(II), GMZ bentonite may also be suitable for the removal of other heavy metals and radionuclides. Hence, more investigations on the sorption property of GMZ bentonite towards various pollutants are ongoing in our laboratory so as to determine the possibility of using GMZ bentonite as a high-efficiency material for wastewater disposal. Acknowledgements Financial supports from the National Natural Science Foundation of China (20907055; 20971126), the Knowledge Innovation Program of CAS and Special Foundation for High-level Waste Disposal (2007-840) are acknowledged. References [1] M. Mohapatra, S. Anand, Studies on sorption of Cd(II) on Tata chromite mine overburden, J. Hazard. Mater. 148 (2007) 553–559. [2] G.D. Sheng, S.W. Wang, J. Hu, Y. Lu, J.X. Li, Y.H. Dong, X.K. Wang, Adsorption of Pb(II) on diatomite as affected via aqueous solution chemistry and temperature, Colloid Surf. A 339 (2009) 159–166. [3] L.R. Van Loon, B. Baeyens, M.H. Bradbury, Diffusion and retention of sodium and strontium in Opalinus clay: comparison of sorption data from diffusion and batch sorption measurements, and geochemical calculations, Appl. Geochem. 20 (2005) 2351–2363. [4] X.Y. Gu, L.J. Evans, Surface complexation modelling of Cd(II), Cu(II), Ni(II), Pb(II) and Zn(II) adsorption onto kaolinite, Geochim. Cosmochim. Acta 72 (2008) 267–276. [5] D.G. Strawn, D.L. Sparks, The use of XAFS to distinguish between inner- and outer-sphere lead adsorption complexes on montmorillonite, J. Colloid Interface Sci. 216 (1999) 257–269. [6] H. Benaissa, B. Benguella, Effect of anions and cations on cadmium sorption kinetics from aqueous solutions by chitin: experimental studies and modeling, Environ. Pollut. 130 (2004) 157–163. [7] K.A. Matis, A.I. Zouboulis, I.C. Hancock, Waste microbial biomass for cadmium ion removal: application of flotation for downstream separation, Bioresour. Technol. 49 (1994) 253–259. [8] C.L. Chen, X.K. Wang, Sorption of Th (IV) to silica as a function of pH, humic/fulvic acid, ionic strength, electrolyte type, Appl. Radiat. Isot. 65 (2007) 155–163. [9] S.T. Yang, D.L. Zhao, H. Zhang, S.S. Lu, L. Chen, X.J. Yu, Impact of environmental conditions on the sorption behavior of Pb(II) in Na–bentonite suspensions, J. Hazard. Mater. 183 (2010) 632–640. [10] P. Reiller, F. Casanova, V. Moulin, Influence of addition order and contact time on thorium(IV) retention by hematite in the presence of humic acids, Environ. Sci. Technol. 39 (2005) 1641–1648. [11] P. Zhou, H. Yan, B.H. Gu, Competitive complexation of metal ions with humic substances, Chemosphere 58 (2005) 1327–1337. [12] L.Z. Zhu, J.F. Ma, Simultaneous removal of acid dye and cationic surfactant from water by bentonite in one-step process, Chem. Eng. J. 139 (2008) 503–509. [13] M.H. Al-Qunaibit, W.K. Mekhemer, A.A. Zaghloul, The adsorption of Cu(II) ions on bentonite—a kinetic study, J. Colloid Interface Sci. 283 (2005) 316– 321.

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