Adsorption of 2,4,5-trichlorophenol by organo-montmorillonites from aqueous solutions: Kinetics and equilibrium studies

Adsorption of 2,4,5-trichlorophenol by organo-montmorillonites from aqueous solutions: Kinetics and equilibrium studies

Chemical Engineering Journal 170 (2011) 120–126 Contents lists available at ScienceDirect Chemical Engineering Journal journal homepage: www.elsevie...

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Chemical Engineering Journal 170 (2011) 120–126

Contents lists available at ScienceDirect

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

Adsorption of 2,4,5-trichlorophenol by organo-montmorillonites from aqueous solutions: Kinetics and equilibrium studies Hassina Zaghouane-Boudiaf ∗ , Mokhtar Boutahala Laboratoire de Génie des Procédés Chimiques (L.G.P.C), Département de Génie des Procédés, Faculté de Technologie, Université Ferhat Abbas, Sétif 19000, Algeria

a r t i c l e

i n f o

Article history: Received 22 December 2010 Received in revised form 7 March 2011 Accepted 8 March 2011 Keywords: Adsorption Organo-clays 2,4,5-Trichlorophenol Kinetics Thermodynamic

a b s t r a c t Two montmorillonites modified with organic surfactant hexadecyltrimethylammonium bromide via ion exchange were used as adsorbents to remove 2,4,5-trichlorophenol (2,4,5-TCP) from aqueous solution in a batch system. Due to their organophilic nature, exchanged montmorillonites are able to adsorb 2,4,5-TCP at a very high extents. The maximum capacity at 20 ◦ C and pH 4 was 368 and 303 mg/g for organo-montmorillonite (MtC16) and acid-activated-organo-montmorillonite (AMtC16) respectively. Experiments were showed that lower pH increased the amount of adsorbed TCP which reached a maximum at pH 4. The adsorption kinetics was found to follow the pseudo-second-order kinetic model. The non-linear Langmuir model provided the best correlation of experimental data. Isotherms were also used to obtain the thermodynamic parameters. The negative values of G◦ and H◦ indicated the spontaneous and exothermal nature of the processes. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Chlorophenols are mainly produced in chemical industries, such as petroleum refineries, plastics, pharmaceuticals and pesticides. Phenols give an unpleasant taste and odour to drinking water and can exert negative effects on different biological processes. Most of these compounds are known or suspected to be human carcinogens. They are weak acids and permeate human skin by in vitro and are readily absorbed by gastro-intestinal tract. The acute toxicity of chlorophenols includes increased respiratory rate, vomiting, and nausea [1]. They create complicated problems to water bodies such as death of aquatic life in inland water bodies, inhibition of the normal activities of microbial population in wastewater treatment plant. Toxicity generally increases with the degree of chlorination. The 2,4,5-trichlorophenol (TCP) is mainly used as a fungicide in the paper industry and as a precursor in the herbicide industry. Low levels of TCP were found in drinking water in several places in the world, and it presents a hazard both to human health and to the environment. The 2,4,5-TCP was listed by the EPA [2] as one of 31 high-priority pollutants. Various physicochemical methods have been proposed for the treatment of waters and wastewaters containing phenolic wastes. It is now widely recognized that sorption processes provide a feasible method for the removal of chlorophenols from waters and wastewaters. Adsorption on activated carbons is a well-known pro-

∗ Corresponding author. E-mail address: [email protected] (H. Zaghouane-Boudiaf). 1385-8947/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.cej.2011.03.039

cess for organic contaminants removal [3–6]. The porous nature of this adsorbent material and its high internal surface area are favourable properties for adsorption. However, high cost and recovering problems of activated carbon particles from treated water are of disadvantage. Some studies illustrated the importance of low cost materials modified by chemical or physical processes as adsorbents for removal of organic pollutants [7–9] from waters and wastewaters. Recently organo-clays have been characterized for their great ability to remove phenols and chlorophenols [10–13] and it showed that they were a powerful sorbent. The aim of this work is to study adsorption of the 2,4,5-TCP on organo-montmorillonite and acid-activated-organomontmorillonite from aqueous solutions. This adsorption study has not been reported in literature. We have been interested in the Namontmorillonite and acid-activated-montmorillonite intercalated with hexadecyltrimethylammonium (HDTMA+ ) via ion exchange. As the montmorillonite is progressively intercalated by a surfactant, the surface properties of the clay change considerably, and so does the ability of organo-montmorillonites to remove organic contaminants from water. To understand the adsorption mechanism, the corresponding kinetics and equilibrium models were studied. 2. Materials and experimental methods 2.1. Materials The bentonite used in this study was from Hammam Boughrara (West Algeria). Its chemical composition was found to be as follows: 69.4% SiO2 , 1.1% MgO, 14.7% Al2 O3 , 0.8% K2 O, 0,3% CaO, 1.2%

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2.4. Kinetic studies

Fig. 1. Structure of 2,4,5-trichlorophenol.

Fe2 O3 , 0.5% Na2 O, 0.2% TiO2 , 0.05% As and 11% loss of ignition [14]. The cation exchange capacity (CEC) was 0.90 meq/g [15]. It had a specific surface area of 80 m2 /g [15]. The 2,4,5-trichlorophenol (TCP) supplied by Sigma–Aldrich Chemicals was used as an adsorbate. 2,4,5-Trichlorophenol has a molecular weight of 197.45 g/mol and linear formula: Cl3 C6 H2 OH. Its solubility in water at 20 ◦ C is 900 mg/L and it has a pKa between 6.7 and 6.94. The chemical structure of TCP used is shown in Fig. 1. All others reactants were also purchased from Sigma–Aldrich Chemicals. 2.2. Preparation of samples and TCP solutions The NaMt was prepared with a procedure similar to that of reported by Khalaf et al. [14]: 30 g of crude bentonite were mixed with 1 L of 1 M NaCl solution and stirred for 24 h. After three successive treatments, the homoionic bentonite was dialyzed in deionized water until it was free of chloride. Then it was separated by centrifugation to eliminate all other solid phases (quartz, cristobalite, etc.) [16]. The montmorillonite fraction (<2 ␮m) was recovered by decantation and dried at 80 ◦ C. The NaMt was treated under mechanical stirring with 1 M H2 SO4 solution at 90 ◦ C for 4 h. The mass ratio of the bentonite to the acid solution was 1:1. After activation, the solid was washed by distilled water until SO4 2− free and dried at 80 ◦ C. The derivate is an acid-activated montmorillonite noted AMt. The NaMt and AMt were treated with the cationic surfactant HDTMA+ with the molecular formula C19 H42 N+ for the purpose of sorption enhancement. Surfactant-modified montmorillonite was prepared by adding amounts of the cationic surfactant equivalent to 100% of the value CEC of montmorillonite. The surfactant was dissolved in 1 L of distilled water at 80 ◦ C and stirred for 3 h. A total of 10 g of sample (NaMt and AMt) were added separately to the 1 L surfactant solution. The dispersions were stirred for 3 h at 80 ◦ C. The separated organo-montmorillonites were washed with distilled water. Washing was repeated until the supernatant solution was free of chloride ions, as indicated by the AgNO3 test. The organo-montmorillonites were oven-dried at 80 ◦ C until the water was completely evaporated. The derivates are MtC16 (organo-montmorillonite) and AMtC16 (organo-acidactivated montmorillonite) respectively. The TCP solutions were prepared in the neutral distilled water from an initial solution of 500 mg/L of TCP for all experiments. Their pH was adjusted by adding either 0.01 N HCl or 0.01 N NaOH for varying from 2 to 8 and used in experiments. 2.3. The determination of pHpzc Batch equilibrium experiments were used to estimate zero point charge (pHpzc). 50 mL of 0.01 M NaCl solution was poured into several erlenmeyer flasks. The pH of solution for each flask was adjusted to a value between 2 and 12 by addition of 0.1 M HCl or 0.1 M NaOH solution. Then, 0.15 g of adsorbent was added to the flasks and the dispersion was stirred for 48 h. After this time the final pH was measured. A plot of the final pHf as a function of the initial pHi provides pHpzc of the adsorbents by the plateau of constant pH to the ordinate (figure not showed).

Effect of initial pH (2, 4, 6 and 8) of TCP solution, temperature (20, 30 and 40 ◦ C) and adsorbent mass (20, 30 and 40 mg) on adsorption was investigated. The initial pH was adjusted and measured at the beginning and the end of each experiment. 20 mg of organo-montmorillonite were dispersed in 50 mL of TCP solution at a concentration of 100 mg/L and stirred with speed of 100 rpm. After each time of contact a sample was removed and centrifuged. The 2,4,5-TCP concentrations were determined with a Shimadzu UV-1700 UV–visible spectrophotometer at 290 nm for dispersions at pH 2, 4 and 6, the determination was done at 310 nm for dispersions at pH 8. The amount of TCP adsorbed was derived from the initial and final concentrations of 2,4,5-TCP. All experiments were run in triplicate to ensure reproducibility.

2.5. Adsorption isotherms A constant volume of 2,4,5-TCP solution (50 mL) with varying initial concentrations of TCP (20–300 mg/L) were mixed with a constant mass of samples (20 mg). The dispersions were shaken at constant temperature of 20 ± 1 ◦ C, the stirring speed was 100 rpm; the dispersions were maintained at constant pH 4 for 120 min. Most of the adsorption data can be expressed by the Langmiur, Freundlich and Redlich–Peterson (R–P) models. The Langmuir [17] model assumes uniform energies of adsorption onto the surface isotherm. The maximum amount of the adsorbate per unit weight of the adsorbent forms a complete monolayer on the surface. The Freundlich isotherm model is an empirical and exponential equation. It assumes that as the adsorbate concentration increases, the concentration of adsorbate on the adsorbent surface also increases. The model is valid for adsorption that occurs on the heterogeneous surfaces [18]. The Redlich–Peterson [19] isotherm is also an empirical isotherm incorporating three parameters. It combines elements from both the Langmuir and Freundlich equations, and the mechanism of adsorption is a hybrid and does not follow ideal monolayer adsorption.

3. Results and discussion 3.1. Effect of stirring time and solution pH on TCP adsorption The effect of stirring time and pH solution on the adsorption of the 2,4,5-TCP on the organo-montmorillonites is shown in Fig. 2. All the other parameters like temperature, mass of adsorbents, initial concentration and stirring speed were kept constant. The plots show that the adsorption of TCP increases with an increase in stirring time and attains equilibrium. The contact time required for the adsorption of TCP onto AMtC16 was very short (20–30 min) for all pH. This result is interesting because equilibrium time is one of the important parameters for economical wastewater treatment applications. The same result was found by Hameed in adsorption of the 2,4,6-trichlorophenol onto activated clay [20]. The equilibrium time for the adsorption of TCP onto MtC16 was 30–40 min. It might be that in the first the TCP was adsorbed on the surface clay and then it entered in the sheet of the clay [20,21]. In this case a long time is needed. The adsorption curves are single, smooth and continuous leading to saturation. Based on this result, the contact time was fixed at 120 min for the rest of the batch experiments to make sure that equilibrium was reached in all cases. It is also clear that the adsorption depends on the initial concentrations of the solution, the adsorbent mass, the temperature and the stirring speed for effective removal of TCP.

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160

A

140

B

140 120 120 100

80 60 pH

40

2

qt (mg/g)

qt (mg/g)

100

80 60 40

20

4

20

0

6

0

pH 2 4 6

8

8 -20

-20 0

10

20

30

40

50

60

0

10

20

t (min)

30

40

50

60

t (min)

Fig. 2. Fit with pseudo second order: effect of contact time on the uptake of TCP on (A) MtC16 and (B) AMtC16 at different pH at 20 ◦ C (C0 = 100 mg/L, V = 50 mL, adsorbent mass = 20 mg, agit. speed = 100 rpm).

Fig. 2 shows also that the TCP amount adsorbed was found to decrease with increase in solution pH for the MtC16 and AMtC16. The highest amount of TCP adsorbed was achieved at pH 4 for the two samples. This is because at pH 4, the total surface of the organo-clays with acidic pHpzc 5–6 has more positively charged sites. On the other hand when pH (solution) < pKa (TCP pKa = 6.8), the concentration of unionized chlorophenol species were high and decreases with increasing pH. This does not favour any repulsion between the adsorbent surface and adsorbate and the dispersion interactions predominated [22]. However, at basic pH, when pH (solution) > pKa , the ionised species of TCP were higher than the unionised species (it was found to increase with pH), the 2,4,5-TCP uptake was lower due to the electrostatic repulsions between the negative surface charge (pHpzc < pH (solution)) and the chlorophenolate anions and between chlorophenolate–chlorophenolate anions in the solution [22]. Similar results were reported in the adsorption of phenol, 2,4,6-trichlorophenol and pentachlorophenol on coir pith carbon [23], in the adsorption of 2,4,6-TCP on coconut shell-based activated carbon [24], on activated clay [20] and on coconut husk-based activated carbon [25] processes. The aluminol groups on bentonite surface were protonated in the form of AlOH2+ when pH (solution) < pHpzc and the ionic species of 2,4,5-TCP at pH 2 are lower than that at pH 4. The 2,4,5-TCP is adsorbed on MtC16 and AMtC16 by Van der Waals attractions to alkyl chains [26] (partition process) and interactions between protonated aluminol groups (AlOH2+ ) in bentonite and anionique species of TCP (which are higher at pH 4 than that at pH 2) also take place (ion exchange) [27]. The result was that adsorption capacity at pH 4 was greater than that at pH 2.

3.2. Effect of adsorbent mass The adsorbent mass was varied from 20 to 40 mg keeping the TCP concentration constant (100 mg/L), the temperature 20 ± 1 ◦ C and pH 4. The percentage of TCP sorption increased from 58 to 87% and from 55 to 78% for MtC16 and AMtC16 respectively. This is attributed to the availability of more sorption sites due to higher amount of the adsorbent. The same results were observed by Koyuncu [28], Al-Asheh et al. [29], Yapar et al. [30]. The uptake capacity of the 2,4,5-TCP decreased from 145 to 108 mg/g and from 138 to 101 mg/g for MtC16 and AMtC16 respectively when adsorbent mass varied from 20 to 40 mg (figure not showed). The most important factor which contributes to this adsorbent mass effect is that adsorption sites remain unsaturated during the adsorption

reaction. This is due to the fact that as the dosage of adsorbent is increased, there is less commensurate increase in adsorption resulting from the lower adsorptive capacity utilization of the adsorbent [31,32]. This is also attributed to the effect of the interactions between the particles such as aggregation leading to an increase in diffusion path length and a decrease in total surface area of the adsorbent [33]. 3.3. Effect of temperature on TCP adsorption The uptake of TCP by organo-montmorillonites decreases with increasing temperature. The amount of TCP adsorbed per unit mass of adsorbent decreased from 145 to 121 mg/g and from 138 to 113 mg/g for MtC16 and AMtC16 respectively when temperature increases from 293 to 313 K (figure not showed). These results indicate that the adsorption of TCP on organo-montmorillonites is controlled by an exothermal process. A similar trend was also observed by Hameed [20] with 2,4,6-trichlorophenol adsorption by activated clay and Bilgili [34] with 4-chlorophenol adsorption onto XAD-4 resin. This is partly due to a weakening of the attractive forces between TCP and adsorbent sites and when temperature increases, solubility of TCP increases and its adsorption decreases. 3.4. Kinetic modelling Generally, three steps are involved during the process of adsorption by porous-adsorbent particles: external mass transfer, intraparticle transport and chemisorption. The adsorption kinetic was modelled as pseudo-first-order [35], pseudo-second-order [36] and intraparticle diffusion process [37]. The linear plot of ln(qe − qt ) versus t (figure not shown) and t/qt versus t (Fig. 3) give the values of the rate constant of the pseudo-first-order (k1 ) and of the pseudo-second-order k2 respectively and correlation coefficient R2 . The results are in Table 1 which shows that correlation coefficient values for the pseudosecond order kinetic model were over 0.99 for all tests, indicating the applicability of the model to describe the adsorption process. The experimental qe values agree well with the calculated values obtained from the pseudo-second order. The second-order rate constants were used to calculate the initial sorption rate given by h = k2 qe 2 . The calculated h values are plotted against temperature in Fig. 4. The initial sorption rate, h, was found to decrease with increase in temperature. This is because the process is exothermal. The h values for AMtC16 are higher than that of MtC16. This effect is due to more available positively adsorption sites on AMtC16 sur-

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A

1,0

pH

1,0

B pH 2

2 4

0,8

123

0,8

4 6

6 8

8

0,6

t/qt

t/qt

0,6

0,4 0,4 0,2 0,2 0,0 0,0 0

10

20

30

40

50

0

60

10

t (min)

20

30

40

50

60

t (min)

Fig. 3. Pseudo-second order plots of TCP adsorption on (A) MtC16 and (B) AMtC16 at different pH at 20 ◦ C (C0 = 100 mg/L, adsorbent mass = 20 mg, V = 50 mL, agit. speed = 100 rpm). Table 1 Kinetic models parameters obtained in adsorption of 2,4,5-TCP on MtC16 and AMtC16. Pseudo-first-order pH

m

T

qe,exp

qe,cal

k1 (10−2 )

Pseudo-second-order R2

k2 (10−3 )

qe,cal

R2

MtC16 2 20 4 20 6 20 8 20 4 20 4 20 4 30 4 40

293 293 293 293 303 313 293 293

132 45 114 58 129 121 120 100

84 66 127 25 80 150 124 108

9.23 10.74 13.13 12.14 0.12 0.12 0.12 0.07

0.819 0.826 0.957 0.918 0.900 0.968 0.949 0.950

140.6 153.8 133.3 59.8 140.8 139.3 130.7 113.9

1.84 2.21 0.91 8.50 1.64 1.00 2.35 2.03

0.997 0.998 0.988 0.999 0.992 0.986 0.999 0.999

AMtC16 2 20 4 20 6 20 8 20 4 20 4 20 4 30 4 40

293 293 293 293 303 313 293 293

110 138 94 61 130 113 120 101

81 59 25 26 80 108 58 39

42.80 12.03 13.00 16.10 9.00 11.45 8.60 8.70

0.986 0.873 0.817 0.949 0.962 0.950 0.935 0.845

111.2 140.6 94.9 61.5 135.7 121.7 123.2 102.6

23.00 6.82 20.00 21.11 2.73 2.00 4.52 7.11

0.999 0.999 0.999 0.999 0.988 0.999 0.999 0.999

m (mg), T (K), qe (mg g−1 ), k1 (min−1 ), k2 (g(mg min)−1 ).

100

MtC16 AMtC16

90

h (mg/g.min)

80 70 60 50 40 30 20 290

295

300

305

310

315

Temperature (K) Fig. 4. Initial sorption rate versus temperature for adsorption of TCP at 20 ◦ C on prepared organo-samples.

Table 2 Intraparticle diffusion model: constants for adsorption of TCP on prepared organomontmorillonites at 20 ◦ C. Adsorbent mass (mg)

ki

C

R2

MtC16 20 30 40

1.60 13.10 10.22

133.0 54.1 42.0

0.985 0.951 0.872

AMtC16 20 30 40

12.17 7.63 4.45

83.5 75.0 74.0

0.943 0.966 0.903

ki (mg g−1 min−0.5 ), C (mg g−1 ).

face which are created by H2 SO4 activation. It was also observed that the process of adsorption of TCP on AMtC16 is very quick in the beginning and equilibrium time was very short. 3.5. Adsorption mechanism The adsorbate species are most probably transported from the bulk of the solution into the solid phase through intraparticle diffusion/transport process, which is often the rate-limiting step in many adsorption processes, especially in a rapidly stirred batch reactor [32]. Since the TCP is probably transported from its aqueous solution to the organo-montmorillonites by intraparticle diffusion. Because the above kinetic models were not able to identify the diffusion mechanism, the intraparticle diffusion is another kinetic model used to study the rate of TCP adsorption onto organomontmorillonites. Fig. 5 shows that plots of qt against t0.5 consist of two or three separate linear regions. It has been suggested that the first one can be attributed to the instantaneous adsorption or external surface adsorption, the second to intraparticle diffusion and the third portion referred to the final equilibrium stage for which the intraparticle diffusion started to slow down due to the extremely low adsorbate concentration left in the solution [38]. As seen from Table 2, the constant C was found to decrease with the increase of adsorbent mass from 20 to 40 mg, which reflect decrease of the thickness of the boundary layer and hence increase of the chance of the external mass transfer [8]. Fig. 4 also showed the linear lines of the second and the third stages did not pass through the origin. It shows that intraparticle diffusion was not the only rate limiting mechanism in the adsorption process. Some other mechanism along with intraparticle diffusion is also involved.

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160 140 140 120 120

qt (mg/g)

qt (mg/g)

100 100 80 60

20

60 AMtC16 mass 20

40

MtC16 mass 20 30 40

40

80

30

20

40

0

0

-20 0

1

2

3

4

5

6

7

-1

8

0

1

t0,5 (min)0,5

2

3

4

5

6

7

8

t0,5 (min)0,5

Fig. 5. Plot of intraparticle diffusion model for adsorption of TCP on organo-montmorillonites at 20 ◦ C.

Table 3 Langmuir, Freundlich and Redlich–Peterson isotherm model constants and correlation coefficients for adsorption of TCP on prepared organo-montmorillonites at 20 ◦ C.

300 250

Isotherms

AMtC16

Langmuir

qm = 368 mg/g KL = 0.023 L/mg R2 = 0.992

qm = 303 mg/g KL = 0.018 L/mg R2 = 0.998

Freundlich

KF = 15 mg/g 1/n = 0.14 R2 = 0.988

KF = 13 mg/g 1/n = 0.13 R2 = 0.989

Redlich–Peterson

A = 12 L/g B = 0.080 L/mgˇ ˇ = 0.8 R2 = 0.991

A = 4.18 L/g B = 0.0018 L/mgˇ ˇ = 1.4 R2 = 0.996

qe (mg/g)

200 150 100 Experimental points (MtC16) Experimental points (AMtC16) Langmuir model Redlich-Peterson model Freundlich model

50 0 0

50

100

150

Parameters MtC16

200

Ce (mg/L) Fig. 6. Equilibrium adsorption isotherms of TCP on organo-montmorillonites at 20 ◦ C fitted to Langmuir, Freundlich and Redlich–Peterson models (adsorbent mass = 20 mg, pH 4, V = 50 mL, agit. speed = 100 rpm).

It may be concluded that surface adsorption and intraparticle diffusion were concurrently operating during the TCP and organomontmorillonites interactions [8]. 3.6. Equilibrium isotherms Fig. 6 showed the equilibrium isotherms of adsorption of TCP on MtC16 and AMtC16 at 20 ◦ C. It indicated that the amount of adsorbed TCP increased with increasing it equilibrium concentration. The parameters obtained in non-linear forms of Langmuir, Freundlich and Redlich–Peterson (R–P) isotherms are listed in Table 3. It showed that the model which gives a good fit to the experimental data is the Langmuir model. The values of maximum adsorption capacity determined using Langmuir expression are higher than the experimental adsorbed amounts and correspond to the adsorption isotherms plateaus. The 2,4,5-TCP isotherms exhibited classical type I characteristics and, as such, obeyed the Langmuir equation. The high levels of adsorption indicate that the 2,4,5-TCP attracted sufficiently to the hydrophobic surfaces of organomontmorillonites, it must have penetrated into interlayer surfaces [30]. In water solvent, as the number of chlorines on the phenol

structure was increased, phenolics compounds was not interacted with water (hydrogen bonding), or interacted weakly and was attracted sufficiently to the hydrophobic surfaces of organoclays. There are two types of interaction between polar organic compounds and organo-clays, adsorption and partition, associated with the interfacial behavior of organic compounds in the system organo-clays/water, and are related to the organic matter and mineral surface, respectively, of the medium. Surface adsorption may include ion exchange, protonation, hydrogen bonding, and coordination/ion-dipole reactions with clays [39,40]. Partition involves interactions between organic matter and NOCs. When a large organic molecule, for example HDTMAB, interacts with clays, the partition process plays an important role. For this mechanism, the adsorption of organic contaminants is functionally and conceptually similar to the dissolution of organic contaminants in a bulk-phase organic solvent such as octanol. This mechanism results from the change of surface properties from hydrophilic to hydrophobic because the long tails of organic molecules form a hydrophobic phase in the interlayer of the organo-clays [41,42]. One of the essential characteristics of the Langmuir equation could be expressed by dimensionless constant called equilibrium parameter, RL [43]: RL =

1 1 + KL .C0

(1)

where C0 is the highest concentration (mg/L), KL is the Langmuir constant (L/mg). The value of RL indicates the type of isotherm to be irreversible (RL = 0), favourable (0 < RL < 1), linear (RL = 1) or

H. Zaghouane-Boudiaf, M. Boutahala / Chemical Engineering Journal 170 (2011) 120–126 Table 4 Thermodynamic parameters for the adsorption of 2,4,5-TCP on AMtC16 and MtC16. G◦ (kJ/mol)

T (K)

H◦ (kJ/mol)

S◦ (J/mol K)

MtC16 293 303 313

−3.0 −2.6 −2.2

−14.7

−40.0

AMtC16 293 303 313

−2.8 −2.4 −2.0

−15.3

−42.5

unfavourable (RL > 1). The dimensionless separation factors calculated for TCP adsorption at 20 ◦ C are: RL = 0.14 for adsorption of TCP on MtC16 and RL = 0.16 for adsorption of TCP on AMtC16. RL values were less than 1 and greater than zero indicating favourable adsorption for the two processes. Thermodynamic parameters for the adsorption were calculated using the equation: ln KD =

S ◦ H ◦ − R RT

(2)

where KD is the distribution coefficient of adsorbent and equal to qe /Ce , H◦ (kJ/mol), S◦ (J/mol) are respectively, the standard enthalpy and the standard entropy, T the absolute temperature and R is the gas constant. The plot of ln KD versus 1/T (figure not showed) is linear with the slope and the intercept giving values of H◦ and S◦ . The Gibbs free energy change G◦ , indicating the spontaneity of the adsorption process, was calculated as: G◦ = H ◦ − TS ◦

(3)

The thermodynamic parameters obtained are in Table 4. The negative value of G◦ and H◦ indicated the spontaneous and exothermal nature of the process. The change of Gibbs free energy and the enthalpy for the physical adsorption is generally in the range of nil to −30 and −42 kJ/mol respectively. For the chemical adsorption G◦ and H◦ are in the range of −80 to −400 kJ/mol and −42 to −125 kJ/mol respectively. The values of G◦ and H◦ in this study showed that the adsorption of 2,4,5-TCP onto organomontmorillonites could be considered as a physical adsorption enhanced by the electrostatic effect.

4. Conclusion The results of this investigation showed that the two organomontmorillonites have a suitable adsorption capacity for the removal of TCP from aqueous solutions. They are good and low-cost adsorbents. The high levels of adsorption showed that the prepared samples have an excellent affinity towards TCP. The adsorption for the two systems was found to be dependent on pH, contact time and temperature. The maximum capacity at 20 ◦ C and pH 4 was 368 mg/g for organo-montmorillonite (MtC16) and 303 mg/g for acid-activated-organo-montmorillonite (AMtC16). The equilibrium isotherms of 2,4,5-TCP were best represented by the non-linear Langmuir model. The data indicate that the adsorption kinetics follow the pseudo-second-order rate. The negative values of G◦ and H◦ suggested that the processes of adsorption are spontaneous and exothermal. As a valuable raw material, these organo-montmorillonites may have environmental applications in treating water for the removal of organic pollutants and could be employed as low-cost adsorbents as alternatives to commercial activated carbon.

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