Adsorption of lead onto formaldehyde or sulphuric acid treated acorn waste: Equilibrium and kinetic studies

Biochemical Engineering Journal 37 (2007) 192–200

Adsorption of lead onto formaldehyde or sulphuric acid treated acorn waste: Equilibrium and kinetic studies a , Mahmut Ozacar b,∗ , ˙I. Ayhan S ¨ ¨ Ahmet Ornek ¸ engil c a

Institute of Sciences and Technology, Sakarya University, 54040 Sakarya, Turkey Department of Chemistry, Science & Arts Faculty, Sakarya University, 54100 Sakarya, Turkey c Department of Environmental Engineering, Engineering Faculty, Sakarya University, 54040 Sakarya, Turkey b

Received 23 May 2006; received in revised form 2 April 2007; accepted 25 April 2007

Abstract The adsorption of lead onto formaldehyde or sulphuric acid treated acorn waste was studied using a batch sorber. The aim of this study was to understand the mechanisms that govern lead removal and find a suitable equilibrium isotherm and kinetic model for the lead removal in a batch reactor. The experimental isotherm data were analyzed using the Langmuir, Freundlich and Temkin equations. The equilibrium data fit well the Langmuir isotherm. The experimental data were analyzed using four adsorption kinetic models – the pseudo first- and second-order equations, intraparticle diffusion equation and the Elovich equation – to determine the best fit equation for the adsorption of lead ions onto formaldehyde or sulphuric acid treated acorn waste. The rate constants, equilibrium capacities and related correlation coefficients for each kinetic model were calculated and discussed. Also, predicted qt values from the kinetic equations were compared with the experimental data. Results show that the pseudo second-order equation provides the best correlation for the adsorption process, whereas the Elovich equation also fits the experimental data well. © 2007 Elsevier B.V. All rights reserved. Keywords: Acorn waste; Lead; Isotherm; Adsorption kinetics; Elovich equation; Pseudo second-order equation

1. Introduction Many industries, including mining, metal electroplating, painting and coating, smelting, petrochemical, plumbing and battery manufacturing, discharge lead into the environment without adequate purification in some cases [1–3]. Untreated effluents from these industries have an adverse impact on the environment and aquatic life. Lead can threaten human life due to its toxicity, accumulation in food chains and persistence in nature. It is a general metabolic poison and enzyme inhibitor and can accumulate in bones, brain, kidney and muscles. Long-term drinking water containing high level of lead can cause serious disorders, such as anemia, kidney disease mental retardation and nervous system damages [2,4]. Unlike organic contaminants, lead is non-biodegradable, and therefore, must be removed from wastewater. Conventional methods for heavy metal removal are chemical precipitation, coagulation, flotation, ion exchange, filtration, electrochemical treatment, evaporation, membrane fil-

∗

Corresponding author. Fax: +90 264 3460371. ¨ E-mail address: [email protected] (M. Ozacar).

1369-703X/$ – see front matter © 2007 Elsevier B.V. All rights reserved. doi:10.1016/j.bej.2007.04.011

tration, reverse osmosis and solvent extraction. However, these methods are much less efficient for concentrations lower than about 100 ppm, for which they can be prohibitively expensive and can even fail to achieve legal limits [5,6]. For these low concentrations it is preferable to use sorption techniques; and numerous low cost alternative adsorbents have been proposed in the last decade. These low-cost adsorbent materials recently studied can be classified into four categories. (1) Natural minerals and similar materials like coal, peat, clays (bentonite, kaolinite, montmorillonite, sepiolite, etc.) alunite, perlite, boxite, red mud, hydrous ferric oxide, etc. (2) Industrial wastes like fly ash, biogas slurry, chrome sludge, furnace slag, etc. (3) Agricultural wastes like coconut shell, banana pith, orange peel, soya cake, olive cake, hazelnut shell, rice husk and in many cases their carbonized products. (4) Forest wastes like barks, leaves and sawdust of certain timber trees and in many cases their carbonized products. These forest wastes are inexpensive, abundant, and contain polyphenolic compounds (generically termed tannins) that

¨ A. Ornek et al. / Biochemical Engineering Journal 37 (2007) 192–200

under appropriate conditions of pH and temperature are capable of sorbing significant amounts of metal cations from solution [6]. Furthermore, prior treatment to bring about cross linking and/or functionalization can improve both the immobilization of watersoluble substances and sorption capacity of these compounds [6–8]. Acorn waste (AW) is a by-product of the tannin production of Salihli–Manisa (Turkey) that is currently used only as cooking fuel. AW is an effective adsorbent because it contains hydrolyzable tannin. At the center of a hydrolyzable tannin molecule, there is a polyol carbohydrate (usually d-glucose). The hydroxyl groups of the carbohydrate are partially or totally esterified with phenolic groups such as gallic acid (in gallotannins) or ellagic acid (in ellagitannins). The tannin compounds containing polyphenol groups have a high affinity for heavy metal ions such as lead, copper, cadmium, cobalt, zinc, etc. The polyhydroxy polyphenol groups of tannin can be thought the active species in the adsorption process. Ion exchange takes place as metal cations displace adjacent phenolic hydroxyl groups, forming a chelate [6]. One problem with tannin-containing materials such as bark, sawdust, various agricultural and forest by-products is discoloration of the water from soluble phenols [9]. In the literature certain pretreatments, such as acidified formaldehyde [10] and acid, base and formaldehyde [11,12] have been shown to eliminate the bleeding of colored compounds without appreciably affecting capacity. While pretreatment slightly increases cost, some pretreatment may be necessary for controlling color. When AW is used as an adsorbent without any pretreatment, it gives a color to the water. Therefore, the untreated AW was not used in the adsorption process. To prevent the coloration problem, AW was treated with formaldehyde or sulphuric acid before using in the adsorption process [13]. The treated AW can be reused, but the regeneration brings extra cost. The adsorbent is already cheap, thus it is not necessary to regenerate to be able to reuse. Therefore, the desorption experiments were not studied. In this work, the sorption of Pb(II) ions by acorn waste treated with formaldehyde or sulphuric acid were studied, and the influences of adsorbent dose, contact time, pH and lead concentration on the adsorption system were investigated.

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2.2. Sulphuric acid treated acorn waste (STAW) One part of AW was treated with two part of 4 mol/L H2 SO4 at 150 ◦ C for 3 h. After treatment, the reaction product was washed with distilled water and soaked in 1% sodium bicarbonate solution overnight to remove residual acid and washed with distilled water again until a neutral pH. The material was dried in an oven at 105 ◦ C for 24 h and used for the study. 2.3. Adsorption experiments The lead solutions were prepared by dissolving the Pb(NO3 )2 (Merck) in appropriate amounts in distilled water. In the determination of equilibrium adsorption isotherm and effect of operating parameters on the adsorption, 1 g FTAW or STAW and 100 mL of the chosen desired concentration of Pb(NO3 )2 solutions were transferred in 250 mL flask, and shaken on a horizontal bench shaker (N¨uve SL 250), operating at 200 rpm, for 180 min (the time required for equilibrium to be reached between Pb2+ adsorbed and Pb2+ in solution) using a bath to control the temperature at 298 ± 2 K. The experiments were performed at the initial pH 5 of lead solution except those in which the effect of the solution pH was investigated. The pH of the solutions was adjusted with HCl or NaOH solution by using a pH meter. At the end of the adsorption period, the samples (5 mL) were taken and centrifuged for 15 min at 5000 rpm and then analyzed using AAS equipped with an auto-sampler (Shimadzu AA6701F). The amount of adsorption at equilibrium, qe (mg/g), was computed as follows: qe =

(C0 − Ce )V W

(1)

where C0 and Ce are the initial and equilibrium solution concentrations (mg/L), respectively, V the volume of the solution (L) and W the weight of adsorbent used (g). Batch adsorption kinetic experiments were carried out by agitating 1 g of FTAW or STAW with 1 L of Pb(NO3 )2 solutions of desired concentration in a magnetic stirrer operating 500 rpm. At predecided intervals of time, samples were taken, and their concentrations were determined. 3. Results and discussion

2. Materials and methods 3.1. Effect of particle size 2.1. Formaldehyde treated acorn waste (FTAW) Oak tree AW was obtained AR-TU chemical Co. SalihliManisa, T¨urkiye. Acorn waste was dried in sunlight until all the moisture evaporated and ground to a fine powder. The AW powder of 90–212 ␮m size was used for adsorption experiments. To polymerize and immobilize the color and water-soluble substances the ground AW was treated with 1% formaldehyde in the ratio of 1:5 (AW:formaldehyde, w/v) at 50 ◦ C for 4 h. The treated AW was filtered out with a Buchner funnel, washed with distilled water to remove free formaldehyde and activated at 80 ◦ C in an air oven for 24 h. The material was placed in an airtight container for further use.

The effect of particle size on the adsorption of Pb2+ was studied using both FTAW and STAW as adsorbents. Fig. 1 shows the experimental results obtained from a series of experiments performed, using different acorn waste particle size ranges. As shown in Fig. 1, the adsorption capacity for Pb2+ increased with decrease in the particle size. At a particle size range of 90–212 ␮m the values of saturation capacities are 93.1 and 96.8 mg/g Pb2+ for FTAW and STAW, respectively. Decrease in the particle size would lead to increase in surface area and the increase in the adsorption opportunity at the outer surface of the adsorbent. Besides adsorption at the outer

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¨ A. Ornek et al. / Biochemical Engineering Journal 37 (2007) 192–200

Fig. 1. Effect of particle size on adsorption of lead by FTAW and STAW. Conditions: 100 mg/L concentration, 1 g/100 mL dose, 180 min agitation and pH 5.

surface of the adsorbent there is also the possibility of intraparticle diffusion from the outer surface into the pores of the material. The diffusional resistance to mass transfer is greater for large particles. Because of various factors, such as diffusional path length or mass transfer resistance, contact time, and blockage sections of the particle may not be utilized for adsorption, therefore, the adsorption capacity of large particles may be low [14–16]. 3.2. Effect of FTAW and STAW doses Fig. 2 shows the adsorption of Pb2+ as a function of FTAW and STAW dosages. It is apparent that by increasing the adsorbent dose the amount of adsorbed Pb2+ increases but adsorption density, the amount adsorbed per unit mass, decreases. It is readily understood that the number of available adsorption sites increases by increasing the adsorbent dose and it, therefore, results in the increase of the amount of adsorbed Pb2+ . The decrease in adsorption density with increase in the adsorbent dose is mainly because of unsaturation of adsorption sites through the adsorption process [16,17]. Another reason may be due to the particle interaction, such as aggregation, resulting from high adsorbent dose. Such aggregation would lead to decrease in total surface area of the adsorbent and an increase in diffusional path length [16].

Fig. 2. Effect of adsorbent dose on adsorption of lead by FTAW and STAW. (AA: Amount adsorbed, AD: Adsorption density) Conditions: 100 mg/L concentration, 90–212 ␮m particle size, 180 min agitation and pH 5.

Fig. 3. Effect of pH on adsorption of lead by FTAW and STAW. Conditions: 100 mg/L concentration, 90–212 ␮m particle size, 1 g/100 mL dose and 180 min agitation.

3.3. Effect of pH The pH of a suspension is one of the most important parameters in the adsorption investigations. Fig. 3 shows the effect of pH on the lead adsorption by FTAW and STAW. As can be seen from Fig. 3, the pH value of the lead solution plays an important role in the whole adsorption process and particularly on the adsorption capacity of the adsorbent. Solution pH would affect both aqueous chemistry and surface binding sites of the adsorbents. Moreover, a change in pH also results in a change in the charge profile of the adsorbate species which consequently influences the interactions between the adsorbate species and adsorbent. The adsorption of Pb(II)was observed at an optimum pH of 5. Below and above this pH, a decreasing trend in adsorption was observed. When the initial pH of the adsorption medium was adjusted to a higher value of pH 6, lead precipitation (Pb(OH)2 ) was observed due to the existence of OH− ions in the adsorption medium. The adsorbed amount of Pb2+ increased from 22.7 to 93.1 mg/g for FTAW and from 23.6 to 96.8 mg/g for STAW at an initial lead concentration of 100 mg/L, when the initial pH value of the system varied from pH 2 to 5. However, the sorption capacity decreased slightly when the initial pH value changed from pH 5 to 6. The amount of lead adsorption onto STAW and FTAW was expected to be verified by the pH value of the solution in the experiments, because ion exchange is one of the adsorption processes. This trend has generally been observed in sorption studies such as the sorption of Pb(II) by tree fern [18] and the sorption of Pb(II) on cone biomass of Pinus sylvestris [19]. In order to interpret the adsorption behavior of Pb(II) ions on an adsorbent surface, a knowledge of Pb(II) speciation and the sorbent surface characteristics is essential [20]. The major components of the polymeric material in AW are phenolic compounds based on tannin. These kinds of material possess the capability of capturing heavy metal ions. It can be speculated that lignin, tannins or other phenolic compounds are the active ion exchange compounds and that active sites are the phenolic groups of those compounds. They have multiple adjacent polyhydroxyphenyl groups in their chemical structure which have extremely high affinity for heavy metal ions [16,21,22].

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Based on the electron donating nature of the O-containing phenol and carboxyl groups in AW and the electron-accepting nature of heavy metal ions, the ion exchange mechanism could be preferentially considered. For instance, a divalent heavy metal ion such as Pb2+ may attach itself to two adjacent hydroxyl groups which can donate two pare of electrons to the metal ion, forming four coordination number compounds and releasing two hydrogen ions into solution. It is then readily understood that the equilibrium is quite dependent on pH of the aqueous solution. At lower pH, the H+ ions compete with Pb2+ cations for the exchange sites on the AW, thereby partially releasing the latter. The Pb2+ cations are completely released under circumstances of extreme acidic conditions. In most cases, the adsorbed amount of metal ions increased with an increase in pH up to a certain value and then decreased with further increase of pH. In a certain pH range for Pb2+ , there may be number of species present in solution, such as Pb2+ , Pb(OH)+ , Pb(OH)2 , etc. At lower pH, the positive charged lead ion species may compete with H+ and be adsorbed at the surface of the AW by ion exchange mechanism. With an increase in pH, lead ion species, mainly neutral, may be adsorbed by hydrogen bonding mechanism along with ion exchange. These mechanisms are shown in the following equations [16]: Ion exchange 2(R-COH) + Pb2+ → (R-CO)2 Pb + 2H+ R-COH + Pb(OH)+ → R-COPb(OH) + H+ Hydrogen bonding 2R-COH + Pb(OH)2 → (R-COH)2 Pb(OH)2 where R is the matrix of the AW. The chemical bonding results from the sharing of a free electron pair between the surface oxygen atom and metal atom or the formation of an O–Pb(II) bond. An increase in pH shows an increase in adsorption up to 5 in which the surface of AW is negatively charged and the sorbate species are also still positively charged. The increasing electrostatic attraction between positively charged sorbate species [Pb2+ and Pb(OH)+ ] and negative surface sites will lead to increased adsorption of Pb(II) on FTAW and STAW. This same pH phenomenon was observed by earlier workers when examining metal adsorption on activated carbon [20]. 3.4. Effect of contact time and initial Pb2+ concentration Effects of contact time and initial Pb2+ concentration on adsorption of lead by FTAW and STAW are presented in Fig. 4. The amount of Pb2+ adsorbed increased with increase in contact time and reached equilibrium after 150 min for the Pb2+ concentrations 30–300 mg/L used in this study. The equilibrium time is independent of initial lead concentration. But in the first 30 min, the initial rate of adsorption was greater for higher initial lead concentration. Because the diffusion of Pb2+ ions through the solution to the surface of adsorbents is affected by the lead concentration, since agitation speed is constant. An increase of the lead concentration accelerates the diffusion of Pb2+ from the

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Fig. 4. Adsorption kinetics of lead on FTAW and STAW at different initial concentrations. Conditions: 90–212 ␮m particle size, 1 g/L dose, 180 min agitation and pH 5.

lead solution onto adsorbents due to the increase in the driving force of the concentration gradient [23,24]. Hence, the amount of adsorbed lead at equilibrium increased from 13.9 to 95 mg/g for FTAW and from 14.1 to 99.3 mg/g for STAW as the lead concentration was increased from 30 to 300 mg/L. 3.5. Equilibrium studies Fig. 5 shows the equilibrium adsorption isotherm of lead by FTAW and STAW. The isotherm rises sharply in the initial stages for low Ce and qe values. This indicates that there are plenty of readily accessible sites. Eventually a plateau is reached, indicating that the adsorbent is saturated at this level. The decrease in the curvature of the isotherm, tending to a monolayer, considerably increasing the Ce values for a small increase in qe, is possibly due to the less active sites being available at the end of the adsorption process and/or the difficulty of the edge ions in penetrating the adsorbent, lead ions partially covering the surface sites. The equilibrium of a solute separated between the liquid and solid phases is described by various models of sorption isotherms such as Langmuir, Freundlich and Temkin models. These models suggest a monolayer sorption, with lateral interactions between the adsorbed molecules or ions in the case of the Freundlich models: the energetic distribution of sites is heterogeneous, due to the diversity of sorption sites or the diverse nature of the metal ions adsorbed, free or hydrolyzed species. The Langmuir model supposes a monolayer sorption with a homogenous distribution of sorption sites and sorption energies, without interactions between the adsorbed molecules or ions. A basic assumption

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The values of the Langmuir constants aL , KL and Q0 with the correlation coefficients are listed in Table 1 for the lead-FTAW and lead-STAW systems and the theoretical Langmuir isotherm is plotted in Fig. 5 together with the experimental data points. The monolayer saturation capacities of FTAW and STAW, Q0 , are 164.2 and 124.1 mg/g, respectively. The value of the correlation coefficient is higher than the other two isotherms values. In all cases, the Langmuir equation represents the best fit of experimental data than the other isotherm equations (Fig. 5). Assuming the batch adsorption to be a single-staged equilibrium operation, the separation process can be defined mathematically using these isotherm constants to estimate the residual concentration of lead or amount of adsorbent for desired purification. The lead adsorption capacities of some adsorbents are given in Table 2. It can be seen from Table 2 that the capacities of FTAW and STAW are lower than chitosan nanoparticles and condensed tannin gel, and the capacity of STAW is comparable to activated phosphate. The adsorption capacities of FTAW and STAW are significantly higher than most of the adsorbents reported in the literature.

Fig. 5. Equilibrium isotherms of lead on FTAW and STAW. Conditions: 90–212 ␮m particle size, 1 g/L dose, 180 min agitation and pH 5.

of Temkin isotherm is that the heat of adsorption of all the molecules or ions in the layer decreases linearly with coverage due to sorbate/sorbate interactions. In order to optimize the design of a sorption system to remove lead from effluents, it is important to establish the most appropriate correlation for the equilibrium curve. Three isotherm equations have been tested in the present study, namely, Langmuir, Freundlich, and Temkin. 3.5.1. The Longmuir isotherm The widely used Langmuir isotherm [25–27] has found successful application in many real sorption processes and its linear form is expressed as: 1 aL 1 1 = + qe KL K L Ce

(2)

where qe (mg/g) and Ce (mg/L) are the amount of adsorbed lead per unit weight of adsorbent and unadsorbed lead concentration in solution at equilibrium, respectively. The KL (L/g) and aL (L/mg) are the Langmuir equilibrium constants and the KL /aL gives the theoretical monolayer saturation capacity, Q0 . Therefore, a plot of 1/qe versus 1/Ce gives a straight line of slope 1/KL and intercepts aL /KL .

3.5.2. The Freundlich isotherm The well-known Freundlich isotherm [23,42] is often used for heterogeneous surface energy systems. A linear form of the Freundlich equation is given as: log qe = log KF +

1 log Ce n

(3)

where KF is the Freundlich constant and n the Freundlich exponent. KF and n can be determined from the linear plot of log qe versus log Ce . The values of the Freundlich constants together with the correlation coefficient are presented in Table 1 and the theoretical Freundlich equation is shown in Fig. 5. The value of correlation coefficient is higher than the Temkin value but lower than the Langmuir value. Therefore, the Freundlich equation only represents a better fit of experimental data than the Temkin equation, especially up to 135 mg/L Ce values, but not in the case of the Langmuir equation (Fig. 5). 3.5.3. The Temkin isotherm The Temkin isotherm [43–45] has been used in many sorption processes. A linear form of the Temkin isotherm can be expressed as: qe =

RT RT ln A + ln Ce b b

(4)

where RT/b = B. The adsorption data can be analyzed according to Eq. (4). Therefore a plot of qe versus ln Ce enables one to determine the constants A and b. The values of the Temkin

Table 1 Langmuir, Freundlich and Temkin isotherm constants for lead adsorption Adsorbent

FTAW STAW

Langmuir

Freundlich

KL (L/g)

aL (L/mg)

Q0 (mg/g)

r2

0.821 0.993

0.005 0.008

164.2 124.1

0.999 0.999

KF (L/g) 1.771 2.476

Temkin n

r2

B

A (L/g)

r2

1.374 1.495

0.988 0.987

27.16 26.27

0.076 0.091

0.978 0.980

¨ A. Ornek et al. / Biochemical Engineering Journal 37 (2007) 192–200 Table 2 Monolayer adsorption capacities (Q0 in mg/g) in the literature for adsorption of Pb2+ on various adsorbents Adsorbent

Q0 (mg/g)

Reference

Chitosan nanoparticles Condansed tannin gel FTAW Activated phosphate STAW Chitosan Natural phosphate Condansed tannin resin Modified rice husk Peat Sawdust activated carbon (SDAC) Zeolite PHEMA/chitosan membranes Gelidium algae Commercial activated carbon (CAC) Activated carbon (Sorbo-Norit) Bacteria modified activated carbon (Sorbo-Norit) Modified peat-resin particles Algal waste Live biomass Bacteria modified activated carbon (Merck) Humic acid Activated carbon (Merck) Coir Carbon nanotubes Jute Sawdust Groundnut shells Cone biomass of Pinus sylvestris Goethite Montmorillonite Sawdust Waste tea leaves

398 190 164.2 155 124.1 115.5 (0.558 mmol/g) 115 114.9 108 103.1 93.36 (0.451 mmol/g) 70.58 68.81 64 54.65 (0.264 mmol/g)

[28] [29] This study [5] This study [30] [5] [3] [31] [32] [20] [33] [34] [35] [20]

54.10 54.10

[36] [36]

47.39 44 35.69 26,40

[37] [35] [38] [36]

22,70 21.50 19.87 (0,096 mmol/g) 17.44 17.18 (0.083 mmol/g) 12.63 (0.061 mmol/g) 12.21 (0.059 mmol/g) 11.38 11.04 10.40 3.19 2.096

[39] [36] [40] [2] [40] [40] [40] [19] [39] [39] [22] [41]

constants A and B are listed in Table 1 and the theoretical plot of this isotherm is shown in Fig. 5. The correlation coefficient is also listed in Table 1 and is lower than the other two isotherm value. In all cases, the Temkin equation represents the poorest fit of experimental data than the other isotherm equations (Fig. 5). 3.6. Kinetic studies In order to examine the mechanism of adsorption process such as mass transfer and chemical reaction, a suitable kinetic model is needed to analyze the rate data. Many models such as homogeneous surface diffusion model, pore diffusion model, and heterogeneous diffusion model (also known as pore and diffusion model) have been extensively applied in batch reactors to describe the transport of adsorbates inside the adsorbent particles [46,47]. Any kinetic or mass transfer representation is likely to be global. From a system design viewpoint, a lumped analysis of kinetic data is hence sufficient for practical operations.

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3.6.1. Pseudo first-order equation The sorption kinetics may be described by a pseudo first-order equation [14,26,48,49]. The linear pseudo first-order equation is the folowing: log(q1 − qt ) = log q1 −

k1 t 2.303

(5)

where q1 and qt are the amounts of lead adsorbed at equilibrium and at time t (mg/g), respectively, and k1 is the equilibrium rate constant of pseudo first-order adsorption, (1/min). The slopes and intercepts of plots of log(q1 − qt ) versus t were used to determine the first-order rate constant k1 and equilibrium adsorption density q1 . However, the experimental data deviated considerably from the theoretical data. A comparison of the results with the correlation coefficients is shown in Table 3. The correlation coefficients for the first-order kinetic model obtained at all the studied concentrations were low. Also the theoretical q1 values found from the first-order kinetic model did not give reasonable values. This suggests that this adsorption system is not a first-order reaction. 3.6.2. Pseudo second-order equation The adsorption kinetics may also be described by a pseudo second-order equation [50–56]. The linear pseudo second-order equation is the following: t 1 1 = + t qt q2 k2 q22

(6)

where k2 is the equilibrium rate constant of pseudo second-order adsorption (g/mg min). The slopes and intercepts of plots t/qt versus t were used to calculate the second-order rate constants k2 and q2 . The straight lines in plot of t/qt versus t show good agreement of experimental data with the second-order kinetic model for different initial lead concentrations. Table 3 lists the computed results obtained from the second-order kinetic model. The correlation coefficients for the second-order kinetic model obtained were greater than 0.994 for all concentrations. The calculated q2 values also agree very well with the experimental data. These indicate that the adsorption system studied belongs to the second-order kinetic model. 3.6.3. Intraparticle diffusion equation Because Eqs. (5) and (6) can not identify the diffusion mechanisms, the intraparticle diffusion model was also tested [43,47,57]. The rate parameters for intraparticle diffusion (kint ) at different initial concentrations are determined using the following equation. qt = kint t 1/2

(7)

where kint is the intraparticle diffusion rate constant, (mg/gmin1/2 ). Such plots may present a multilinearity [28,38], indicating that two or more steps take place. The first, sharper portion is the external surface adsorption or instantaneous adsorption stage. The second portion is the gradual adsorption stage, where intraparticle diffusion is rate-controlled. The third portion is the final equilibrium stage where intraparticle

1.775 2.598 3.082 4.536 5.329 5.898 9.608 0.982 0.968 0.975 0.965 0.992 0.983 0.987 14.1 22.5 29.3 43.9 54.7 62 99.3 30 50 70 100 150 200 300

FTAW

STAW

16.57 18.35 23.84 37.68 43.92 47.12 81.51

3.29 3.20 2.65 4.03 2.72 3.27 3.25

0.993 0.992 0.995 0.994 0.978 0.994 0.995

17,67 25.00 32.68 47.62 59.88 67.11 108.7

1.48 2.30 1.52 1.62 0.966 1.13 0.619

0.994 0.998 0.997 0.998 0.999 0.999 0.999

1.064 3.365 4.044 10.43 9.472 15.85 21.20

0.263 0.200 0.156 0.115 0.088 0.084 0.050

0.990 0.997 0.996 0.969 0.988 0.988 0.984 1.827 2.200 2.951 4.319 5.300 5.838 8.390 0.980 0.979 0.981 0.952 0.981 0.991 0.980 0.995 0.997 0.998 0.999 0.999 0.999 0.999 1.92 1.81 1.49 2.10 1.14 0.974 0.965 16.72 24.27 31.55 41.49 54.95 64.10 101 0.986 0.988 0.997 0.951 0.994 0.995 0.949 3.06 2.86 2.49 2.70 2.56 2.51 3.85 13.9 21.4 28.2 38.8 50.4 59.1 95 30 50 70 100 150 200 300

14.18 19.94 22.80 21.60 35.49 42.36 76.95

k2 q2 (mg/g) k1 q1 (mg/g)

1.203 2.621 3.675 9.806 9.096 11.18 41.49

0.276 0.207 0.160 0.131 0.096 0.084 0.061

2 rint

kint [mg/(g min1/2 )] β α

Intraparticle diffusion equation

rE2

The Elovich equation

r22 [g/(mg min)] × 104

Second-order kinetic equation

r12 (1/min) × 102

First-order kinetic equation qe,exp (mg/g) C0 (mg/L) Adsorbent

Table 3 Kinetic parameters for the adsorption of lead onto FTAW and STAW

0.996 0.992 0.993 0.997 0.981 0.986 0.989

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diffusion starts to slow down due to extremely low adsorbate concentrations in the solution. The external surface adsorption (stage 1) was absent in the plot of linear Intraparticle diffusion equation. Because stage 1 is completed before 5 min, and then the stage of intraparticle diffusion control (stage 2) is attained and continues from 5 to 60 min. Finally, final equilibrium adsorption (stage 3) starts after 60 min. The lead is slowly transported via intraparticle diffusion into the particles and is finally retained in the micropores. In general, the slope of the line in stage 2 is called as intraparticle diffusion rate constant, kint . The rate parameters, kint , together with the correlation coefficients are also listed in Table 3. 3.6.4. The Elovich equation The linear Elovic equation is given as follows [53,58,59]: qt =

1 1 ln(αβ) + ln t β β

(8)

where α is the initial sorption rate (mg/g min), and the parameter β is related to the extent of surface coverage and activation energy for chemisorption (g/mg). In this case, a linear relationship was obtained between Pb2+ adsorbed, qt , and ln t over the whole adsorption period, with correlation coefficients between 0.952 and 0.992 for all the lines (Table 3). Also Table 3 lists the kinetic constants obtained from the Elovich equation. In the case of using the Elovich equation, the correlation coefficients are lower than those of the pseudo second-order equation, but it may be used to describe the kinetics of adsorption of Pb2+ onto FTAW and STAW. Although the Elovich equation does not provide any mechanistic evidence, it has proved suitable for highly heterogeneous systems of which the adsorption of Pb2+ onto FTAW and STAW is undoubtedly such a case. A comparison of calculated and measured results for 100 mg/L initial lead concentration is shown in Fig. 6. As can be seen from Fig. 6, the pseudo second-order equation provides the best correlation for all of the sorption process, whereas the Elovich equation also fits the experimental data well. The pseudo first-order and intraparticle diffusion equations do not give a good fit to the experimental data for the adsorption of lead. Two kinetic models have been used extensively to describe the sorption of dyes, metals and other pollutants onto various adsorbents. The main disadvantage of these models are that the linear first-order equation does not give theoretical q1 values which agree with experimental qe values, and the plots of intraparticle diffusion model are only linear over the initial 60 min of the adsorption process. This suggest that the adsorption systems studied belong to the second-order kinetic model, based on the assumption that the rate limiting step may be chemical sorption or chemisorption involving valency forces through sharing or exchange of electrons between adsorbent and adsorbate. The agreement of the Elovich equation with experimental data may be explained as below. The previous successful application of the Elovich equation to heterogeneous catalyst surfaces helps to explain its success in predicting the adsorption of lead on FTAW and STAW. The general explanation for this form of kinetic law involves

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mechanism being the rate controlling step. The adsorption of lead onto FTAW and STAW can also be successfully interpreted by the Elovich equation. It may be concluded that FTAW and STAW may be used as low-cost, natural and abundant source for the removal of lead and they may be an alternative to more costly materials. References

Fig. 6. Comparison between the measured and modeled time profiles for adsorption of lead (100 mg/L initial lead concentration) onto FTAW and STAW.

a variation of the energetics of chemisorption with the active sites are heterogenous AW and therefore, exhibit different activation energies for chemisorption [60]. Because the cell walls of AW mainly consist of cellulose and lignin, and many hydroxyl groups, such as tannins or other phenolic compounds [59]. 4. Conclusion The results of present investigation show that FTAW and STAW have considerable potential for the removal of lead from aqueous solution over a wide range of concentration. The adsorbed amounts of lead increased with decreasing particle size of FTAW and STAW due to the increasing in the surface area. A decrease in the pH of solutions leads to a significant increase in the adsorption capacities of lead on the FTAW and STAW. The adsorbed amounts of lead increased with increase in contact time and reached the equilibrium after 120 min. The equilibrium time is independent of initial lead concentration. The equilibrium data have been analyzed using Langmuir, Freundlich and Temkin isotherms. The Langmuir isotherm was demonstrated to provide the best correlation for the adsorption of lead onto FTAW and STAW. The kinetics of adsorption of lead onto FTAW and STAW was studied by using pseudo first- and second-order equations, intraparticle diffusion equation and the Elovich equation. For two systems examined, the pseudo second-order kinetic model provided the best correlation of the experimental data. The pseudo second-order equation is based on the adsorption capacity on the solid phase and is in agreement with a chemisorption

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