Graft copolymerization of cellulose acetate for removal and recovery of lead ions from wastewater

International Journal of Biological Macromolecules 79 (2015) 913–922

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Graft copolymerization of cellulose acetate for removal and recovery of lead ions from wastewater Nourelhoda A. Abdelwahab a,∗ , Nabila S. Ammar b , Hanan S. Ibrahim b a b

Polymers and Pigments Department, National Research Centre, 33 El Bohouth Street, Dokki-Giza P.O. 12622, Egypt Water Pollution Research Department, National Research Centre, 33 El Bohouth Street, Dokki-Giza P.O. 12622, Egypt

a r t i c l e

i n f o

Article history: Received 22 January 2015 Received in revised form 16 April 2015 Accepted 10 May 2015 Available online 23 May 2015 Keywords: Cellulose acetate Graft copolymerization Lead recovery Isotherm models Kinetic models

a b s t r a c t In this study, cellulose acetate (CA) was modified by grafting with an equimolar binary mixture of acrylic acid (AA) and acrylamide (AAm) via radical polymerization technique using benzoyl peroxide as an initiator. Comparative studies between CA powder and modified CA [CA-g-(AA-co-AAm)] were investigated for Pb(II) ions removal and recovery from wastewater. The main operating conditions such as pH, concentration of Pb(II) ions and sorbent dose were also studied. Kinetic modeling has been studied and lead uptake capacity was calculated using the Langmuir, Freundlich and Dubinin–Kaganer–Radushkevich (DKR) models. The maximum sorption capacity (qemax ) for Pb ions was only 9.4 mg/g for unmodified CA, while, it was reached to 66.67 mg/g by using modified CA. Spectroscopic analysis (FTIR), SEM, EDX and XRD analysis were investigated for CA and modified CA before and after recovery of lead ions from wastewater. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Contamination of various water resources with heavy metals is considered as a serious problem because of their harmful and toxic effects on human health and living organisms. So, it is of great interest to remove heavy metals from wastewater [1]. Lead is a widespread heavy metal and has wide applications in different fields depending on its various physical and chemical properties, such as ductility, softness, malleability, poor conductivity and resistance to corrosion. Unfortunately, lead is highly toxic to the human body and causes damage to the nervous, reproductive and blood circulation systems, even at trace level, such as nephrotoxicity, neurotoxicity and adverse effects on the hematological and cardiovascular systems [2,3]. So, it is important to remove lead from wastewaters effectively before their discharge into the environment. Several techniques have been developed, such as chemical precipitation, ion-exchange, adsorption,coagulation, flotation and electrochemical methods to eliminate heavy metals from wastewaters [4–6]. Recently, many approaches have been considered for the development of most effective, cheapest and efficient technologies, both to reduce the amount of wastewater produced and to improve its quality. In recent years, low-cost adsorbents that have

∗ Corresponding author. Tel.: +20 1223488928. E-mail address: nor [email protected] (N.A. Abdelwahab). http://dx.doi.org/10.1016/j.ijbiomac.2015.05.022 0141-8130/© 2015 Elsevier B.V. All rights reserved.

metal-binding capacities have become one of the alternative treatments for the removal of heavy metals from wastewater [7]. The adsorbents may be of mineral, organic or biological origin, zeolites, industrial by-products, agricultural wastes, biomass and polymeric materials, which have been potentially used as an effective and recyclable sorbents for the selective removal of Pb(II) ions from aqueous media [8,9]. Cellulose acetate (CA) is one of the most economically applicable polymers due to its high hydrophilicity, good toughness, high biocompatibility, good resistance to chlorine and solvent and cheapness [10]. Polyvinylchloride-blend-cellulose acetate/iron oxide nanoparticles nanocomposite membrane was used to remove lead ion from wastewater [11]. Specifically, acrylamide-modified CA and polyethylene glycol-modified CA were used to remove Cs(I) and Eu(III) from aqueous waste solutions [12,13]. Poly(methacrylic acid)-modified CA membrane have high selectivity for adsorption for Hg (II) [14]. The surface of electrospun CA microfibers was modified by grafting poly(methacrylic acid) using Ce(IV) for initiated polymerization [15,16]. On the other hand, it was found that grafted CA with functional groups, such as COOH, SO3 H and NH2 groups can bond efficiently with heavy metal ions through surface complexation mechanisms [17]. To the best of our knowledge, no work was made on the grafting of CA with equimolar binary mixture of acrylic acid and acrylamide. Therefore, the main objective of this study is to apply the optimized grafted CA as sorbent for lead removal and recovery from

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wastewater and this was made in comparison with ungrafted CA. Isothermal and kinetic models were studied. The optimized grafted CA was characterized before and after loading with lead ions using FT-IR spectroscopy, scanning electron microscope (SEM), energy dispersive X-ray (EDX) and X-ray diffraction (XRD).

2.3.4. X-ray diffraction (XRD) Powder X-ray diffraction (XRD) patterns of CA, lead-loaded CA, modified CA and lead-loaded modified CA were recorded on EMPYREAN diffractometer in the range of 2 = 5◦ -50◦ and operates at 45 kV with Cu K␣ ( = 1.5405 A◦ ) radiation.

2. Materials and Experimental Methods

2.4. Uptake of lead ions by modified cellulose acetate

2.1. Materials

Sorption kinetics and isotherm models of the modified cellulose acetate for Pb (II) ions removal were studied using batch adsorption process. Each of the batch adsorption studies was carried out by contacting the adsorbent with the lead ions at room temperature (25 ± 0.1 ◦ C) in a glass bottles. Before addition of sorbent, the pH values of the metal ion solutions were adjusted by Thermo Scientific pH Orion VERSA STAR pH meter. A series of experiments were also conducted in order to determine the effect of pH (2.0–5.0), contact time (90 min), sorbent dosage (0.25–2 g/L) and initial ion concentration (10–100 mg/L) on the adsorption of Pb (II). Each experiment was conducted in a mechanical shaker at 200 rpm. All samples were filtered through filter paper (No. 42) and the Pb (II) ion concentration was determined in the filtrate according to APHA [18] by Atomic Absorption Spectrometer (Varian-SpectrAA (220)) with graphite furnace accessory and equipped with deuterium arc background corrector (Varian Australia, Pty Ltd, Manufacturing site). Precision of the metal measurement was determined by analyzing (in triplicate) the metal ion concentration of all samples. To distinguish between possible metal precipitation and actual metal sorption, controls were used without adsorbent materials. The quantitative mean of triplicate experiments was used for calculations of the percentage of lead ions removal by modified cellulose acetate CA-g-(AA-co-AAm) during the series of batch investigations using the following Eq. (3) expressed as:

Cellulose acetate (Molecular weight, 100,000 g/mol) was purchased from Fluka. However, acrylic acid, acrylamide and benzoyl peroxide were obtained from Aldrich. All solvents are of analytical grade and used as obtained. Stock standard solution (1000 mg/L) was prepared using Pb(NO3 )2 of analytical grade from Merck. Preparation of the synthetic solutions was then prepared by dilution of the stock standard solution. pH control was achieved by using 0.1 M HNO3 and 0.1 M NaOH. 2.2. Grafting procedure and grafting parameters In a typical experiment, certain weighed amount of cellulose acetate (CA) powder was mixed with a definite weight of benzoyl peroxide (BPO) (initiator) in toluene. The mixture was immersed in a water bath at 60 ◦ C under mechanical stirring to generate free radicals on the CA backbone. Equimolar binary comonomer mixture of acrylic acid (AA) and acrylamide (AAm) was added to the CA macroradicals. The reaction was carried out at different monomer/initiator molar ratios ranging from 2 to 10 under temperature range (30–100 ◦ C) and with continuous stirring for different time intervals from 1 to 4 h. The grafted CA [CA-g-(AAco-AAm)] was then filtered and washed with distilled water to remove unreacted monomers. The free homopolymers and copolymers were removed by washing the grafted polymer with toluene and dimethylformamide. CA-g-(AA-co-AAm) was then dried at 60 ◦ C till constant weight and ground with mortar to fine particles. Various reaction parameters such as reaction time, temperature, monomer/initiator molar ratio were optimized for graft copolymerization. The grafting parameters as percent grafting (%G) and percent grafting efficiency (%E) were calculated from the following Eqs. (1) and (2): %G =

%E =

weight of grafted CA − weight of ungrafted CA × 100 weight of ungrafted CA

(1)

weight of grafted CA − weight of ungrafted CA × 100 weight of comonomers(acrylic acid and acrylamide) (2)

2.3. Characterization of graft copolymer 2.3.1. FT-IR Spectroscopy The FT-IR spectra were measured in KBr pellets using a Nexus 670 FT-IR spectrophotometer, Nicolet, USA, in the range of 4000–400 cm−1 . 2.3.2. Scanning electron microscope (SEM) SEM images at different magnifications were obtained using JEOL JXA-840A Electron Probe Microanalyzer on gold coated specimens. The SEM observation was carried out with secondary electron imaging and the acceleration of the electron beam at 10 kV. 2.3.3. Energy dispersive X-ray (EDX) EDX spectra for lead-loaded samples were obtained using Quanta FEG250 instrument.

Removal(%) =

Co − Cf × 100 C0

(3)

where Co and Cf are the initial and final concentrations (mg/L) of lead ions in solution, respectively. Percent of the relative standard deviation of results was calculated and data were not used if the value of standard deviation for any sample was > 3%. 2.5. Reuse of grafted CA The grafted CA loaded with Pb(II) were used to verify the efficiency of several reuse of grafted CA for Pb (II) ion removal. In a typical batch sorption process, 0.05 g of sorbent was added to 100 mL of 75 mg/L Pb(II) ion solution at the optimum operating conditions, the mixture was filtered through a filter paper (No. 42). The lead-loaded sorbent was washed with distilled water several times then dried until reaching constant weight. The lead-loaded sorbent was placed in 100 mL of 0.1 M HCl and stirred for 2 h to elute Pb(II) ions at room temperature. After filtration, the sorbent was washed with deionized water until reaching pH 5.0, then dried in a vacuum oven at 50 ◦ C till reaching constant weight before another use. The same cycle was repeated for six times. 3. Results and discussions 3.1. Preparation and optimization of grafted CA Cellulose acetate (CA) is a stiff polymer due to hydrogen bonds between the glucose rings of neighbouring chains. For grafting CA, benzoyl peroxide (BPO) emerges as the best performing initiator for a radical polymerization-based self-healing system [19]. BPO was dissolved in toluene as an organic solvent and also as a

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Table 1 Percent grafting (%G) and percent grafting efficiency (%E) values for CA-g-(AA-co-AAm). Temperature, ◦ C

Parameters

Monomer: initiator molar ratio

Time, hrs

10

4

2

1

2

3

4

30

60

80

100

%G %E

133.5 124.3

82.45 79.4

99.3 89.1

88.4 88.1

115 114.3

135.2 124.3

135.2 124.3

101 101.5

112 111.34

135.2 124.3

120 118.3

Scheme 1. Chemical structure of CA and CA-g-(AA-co-AAm).

wetting agent for CA. Several factors that affecting grafting copolymerization of AA/AAm as a binary mixture onto CA, such as monomer-to-initiator molar ratio, reaction time and reaction temperature were investigated and the results of %G and %E were represented in Table 1. It was found that maximum %G (133.5) and %E (124.3) were achieved at monomer: initiator molar ratio 10, temperature 80 ◦ C and duration period 3 h. So, the optimized CA-g-(AA-co-AAm) will be selected for the removal and recovery of Pb(II) from synthetic wastewater. The chemical structure of CA and CA-g-(AA-co-AAm) were shown in Scheme 1. 3.2. Characterization 3.2.1. FT-IR spectra The modification of CA by grafting with binary mixture of AA and AAm was proved by FT-IR spectroscopy. The FT-IR spectra of CA, modified CA, Pb2+ -loaded-CA and Pb2+ -loaded-modified CA are shown in Fig. 1(a–d). The most characteristic peaks for CA are 3460 cm−1 corresponding to OH group, 2930 cm−1 attributed to asymmetric and symmetric C H stretching, 1758 cm−1 due to carbonyl group of COOR group, 1377 cm−1 resulted from methyl group and 1060 cm−1 for cyclic ether bonds [20]. On the other hand, CA-g-(AA-co-AAm) showed a characteristic peak for carbonyl of amide group in the acrylamide moiety at 1671 cm−1 . Moreover, the peak corresponding to OH group becomes broader and shifted to 3440 cm−1 due to the grafting process, also the intensity of the peak at 1758 cm−1 was reduced sharply and shifted to 1740 cm−1 and this peak is attributed to carbonyl group of COOR moiety and COOH of acrylic acid of the modified CA, in the same time. All of these changes confirm successful grafting of the equimolar binary mixture (AA/AAm) into CA as shown in Fig. 1a and 1c. After loading CA and CA-g-(AA-co-AAm) with lead (Fig. 1b and 1d), it was observed that there is some changes in the frequency and intensity of the peaks characteristic to hydroxyl ( OH), amide carbonyl ( CONH2 ) and carbonyl of COOR and COOH groups. For lead-loaded CA, it was noticed that the peaks were shifted and became sharper, whereas lead-loaded grafted CA showed change in the peaks frequency by shifting, but there is no change in the intensity of the peaks was observed. 3.2.2. Scanning electron microscope (SEM) SEM observations of CA, modified CA, Pb2+ -loaded-CA and Pb2+ loaded-modified CA are represented in Fig. 2 a−f. Fig. 2a and 2b

showed morphology of unmodified CA with magnifications 500× and 1000×, respectively, where the morphology of unmodified CA seems to be short thin fibers with different lengths. The SEM micrographs of modified CA with magnifications, 500× and 1000× are given in Fig. 2c and d, respectively, where, the SEM micrographs (Fig. 2a−d) showed the significant difference between unmodified and modified CA, the morphology of modified CA showed remarkable increase in the fiber thickness and the shape of fibers became rod-like structure due to coating with copolymer after grafting. The morphology of both lead-loaded CA and lead-loaded modified CA is represented in Fig. 2e and 2f, it was observed that the bulk deposition from lead ions seems clear on the surface of both CA and modified CA fibers. For lead-loaded CA, the lead deposited on the surface of the individual fibers, whereas for lead-loaded modified CA, the deposition of lead clearly seems to be on the surface of the fibers and also between individual fibers to adhere them to each other and to make them as a cemented mass, the deposition of lead on modified CA is higher than CA and this is in agreement with the results obtained later from sorption experiments in which modified CA showed higher adsorption capacity for lead ion than that for unmodified CA. 3.2.3. Energy dispersive X-ray (EDX) Fig. 3 showed EDX spectra of Pb2+ -loaded CA and Pb2+ -loaded modified CA pellets. Fig. 3a showed the percentage and peaks location of carbon, oxygen and lead in Pb2+ -loaded CA. On the other hand, Fig. 3b represented peaks and percentage of carbon, nitrogen, oxygen and lead. Additionally, the morphology of each sample was included in the figure. The peaks represented at 0.27 keV is assigned to carbon, at 0.52 keV is due to oxygen, at 0.39 keV is attributed to nitrogen and between 1.7 and 3.4 keV are characteristic for lead, at this region (1.7−3.4 keV), the irradiation resulted from the electrons of orbital M, whereas the peaks between 10.2 and 12 keV arising from orbital L electrons. The results elucidated higher lead concentration for lead-loaded modified CA than for lead-loaded CA. Carbon can be found mainly in the CA backbone, polyacrylic acid and polyacrylamide graft moieties, nitrogen refers to amide group of polyacrylamide, whereas oxygen presents in CA backbone, COOH and CONH2 of polyacrylic acid and polyacrylamide grafts. 3.2.4. X-ray diffraction (XRD) The XRD patterns of CA, lead-loaded CA, modified CA and leadloaded modified CA are represented in Fig. 4a−d. As shown in Fig. 4a and 4c, CA and modified CA exhibited some crystalline behavior due

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Fig. 1. FT-IR spectra of (a) CA, (b) Pb2+ -loaded-CA, (c) modified CA and (d) Pb2+ -loaded-modified CA.

Fig. 2. SEM micrographs of CA with (a) 500 magnification, (b)1000 magnification, modified CA with (c) 500 magnification, (d) 1000 magnification, (e) Pb2+ -loaded-CA and (f) Pb2+ -loaded-modified CA.

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little competition between Pb2+ cations and protons compared with lower pH values, which leads to high Pb+2 uptake [22]. The mass balance for Pb2+ uptake per gram of sorbent material was determined. If C0 and Cf are the initial and final Pb2+ concentration (mg/L or meq/L), respectively, V is the suspension volume (L) and m is the mass of sorbent material (g), the equilibrium Pb2+ uptake qt (mg/g or meq/g) can be calculated as: qt =

Fig. 3. EDX of (a) lead loaded-CA and (b) lead loaded-modified CA.

to hydrogen bonding between polymeric chains which arising from OH groups, additionally, grafted chains introduces some regularity, but after loading CA and modified CA with lead ion through removal process, the crystalline structures turned amorphous and this may be attributed to destruction of inter- and intra-molecular hydrogen bonding between CA and modified CA chains [21]. Moreover, lead-loaded modified CA appeared more amorphous than lead-loaded CA and this is a good evidence for high removal efficiency achieved by modified CA. 3.3. Comparison between CA and modified CA as sorbents A comparison between Pb2+ uptake capacities of CA and modified CA was performed. Pb2+ uptake capacities were determined by contacting 0.05 g of the CA or modified CA with a synthetic solution of Pb (II) ions (9.5 mg/L) in a mechanical shaker at 200 rpm for 2 h. pH value of 5.0 ± 0.1 was maintained during the experiment by adding 0.1 M NaOH or 0.1 M HCl. At pH 5.0 ± 0.1, there is a

(C0 − Cf )V m

(4)

A comparison between uptake capacities of CA and modified CA showed that modified CA has higher Pb2+ uptake value of 16.4 mg/g or 0.158 meq/g with removal efficiency reached 86.3% than that for CA which showed an uptake value of 9.4 mg/g or 0.091 meq/g with removal efficiency only 49.5%. The active binding sites for Pb2+ in CA are supposed to be hydroxyl and carbonyl groups only, while modified CA has shown great promise in improving the cation exchange capacity due to additional functional groups where CA-g-(AA-coAAm) has three-dimensional cross-linked polymer networks with a lot of functional groups including OH, CONH2 , COOH and COOR groups, which can be considered as the sorption sites for ion exchange and chelation [23]. Therefore, further studies will be accomplished on the modified CA. Depending on the low sorption capacity of CA, surface modification of CA by grafting with acrylic acid/acrylamide mixture aims to add two new function groups, including carboxylic acid group ( COOH) which present in acrylic acid and amide group ( CONH2 ) which present in acrylamide, in order to increase sorption capacity of CA and achieve high removal efficiency for lead ions from wastewater. It was reported previously that poly(acrylic acid-co-acrylamide) has dual function of ion exchange and chelating absorption toward heavy metal ions. COOH group acts as ion exchanger and chelating agent while CONH2 group acts as chelating agent only, So, COOH group is more effective in the removal process [24]. Moreover, it was observed that acidic groups are more effective than basic groups for adsorption of heavy metals [25]. 3.4. Effect of pH Studying the effect of pH on the uptake of Pb(II) from the aqueous solution using modified CA was carried out in the pH range from 1.0 to 5.0 with the initial Pb(II) concentration of 10 mg/L. Table 2 accentuates the residual, uptake concentration (mg/L),

Fig. 4. XRD of CA, lead-loaded CA, modified CA and lead-loaded-modified CA.

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Table 2 Lead ions removal efficiency using 0.5 g/L sorbent weight of grafted CA as a function of pH solution. pH

Initial conc. (mg/L)

Residual conc. (mg/L)

Uptake conc. (mg/L)

Uptake capacity (mg/g)

Percent of Removal

1 2 3 4.5 5

10

8.61 8.5 6.25 1.3 1.3

1.39 1.5 3.75 8.7 8.7

2.78 3 7.5 17.4 17.4

9.36 10.5 30.6 86.3 86.3

Fig. 5. Effect of contact time on the removal of Pb2+ by sorption onto grafted CA at an initial concentration of lead ion 10 mg/L, at pH ranged from 4.5 to 5.0 and 0.5 g/L adsorbent weight.

uptake capacity (mg/g) and the removal percent of Pb(II) using 0.5 g/L of grafted CA. It was observed that the uptake capacity and removal efficiency of Pb(II) increased progressively when the pH increased from 1.0 to 3.0, and increased sharply at pH range from 3.0 to 4.5, while remains constant from pH 4.5 to 5.0. This behavior may be attributed to the ionization of COOH group into H+ and free COO− . Where at very low pH, the concentration of the free COO− on the surface of the modified CA decreased, while at pH range from 3 to 4.5, the removal efficiency sharply increases due to considerable increase of free COO− which is in favor of the electrostatic attraction between Pb2+ and COO− [26–31]. Thus, all studies will be conducted at optimum pH range (4.5−5.0) to ensure the presence of the divalent form of Pb ions and also to conserve the chemical permanence of modified CA with its maximum removal efficiency reaching 86.3%. 3.5. Effect of contact time Treatment of wastewater is significantly affected by the reaction equilibrium time. Fig. 5 illustrates the uptake of Pb(II) ions onto the modified CA, the maximum removal efficiency (86.3%) was achieved rapidly when the reaction equilibrium time attained within 90 min and beyond this time, the removal efficiency values remained constant. This means that, there was a progressive increase in the quantity of Pb(II) ions bound as the contact time increased and equilibrium was reached within 90 min, where, the sorption sites became saturated to maximum uptake capacity onto the modified CA [32].

Fig. 6. (a) Pseudo-first order, (b) pseudo-second order and (c) intra-particle diffusion models for the sorption of Pb(II) by grafted CA at 25±0.2 ◦ C, 0.5 g/L of sorbent weight and at pH ranged from 4.5 to 5.0.

present on the adsorbent surface area, so, after certain time and certain concentration, all active sites are occupied and there is no availability for adsorption [34]. The rate of kinetic reactions of Pb(II) ions sorption onto modified CA was examined using pseudo-first order, pseudo-second order and Intra-particle diffusion models. The conformity between experimental studies and the data of the more applicable models are expected by the higher values of correlation coefficients. 3.6.1. Pseudo-first order model The data of kinetics were simplified with the Lagergren firstorder model [35], which expressed the adsorption rate based on the capacity of adsorption. Generally, the integral form of the pseudofirst order model was articulated by Eq. (5): log[qe − qt ] = log[qe ] −

3.6. Kinetic modeling for sorption process The presence of the active site and the ease of Pb(II) ions access onto the active sites without sterical hindrance are the main factors affecting the characteristic of kinetics which is greatly determined by the matrices of the modified CA. Also, the sorption mechanism depends on the physical and chemical properties of the sorbent as well as on the process of mass transfer [33]. The improvement in Pb(II) removal efficiency until certain time and certain concentration of Pb(II) ions are affected greatly by available active sites

 k  1 2.303

t

(5)

where qe and qt (meq/g) are the amounts of adsorbed Pb(II) ions on the sorbent at equilibrium and at any time t, respectively; and k1 is the Lagergren rate constant of the first-order sorption (min−1 ). The model of first-order is based on the hypothesis that the rate is proportional to the number of free sites. Where, a plot of log (qe -qt ) versus t should provide a linear relationship from which k1 and predicted qe can be determined from the slope and intercept of the plot, respectively, as represented in Fig. 6a. The variation in the rate should be proportional to the first power of concentration for

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uptake of strict surface. However, the relationship between initial solute concentration and rate of uptake will not be linear when pore diffusion limits for the process of uptake. It was noticed that the first-order model doesn’t give a practical estimate of qe for uptake of Pb(II) ions, where the experimental values of qe (0.1583 meq/g) were higher than the fitted value of qe (0.085 meq/g) for modified CA as shown in Table 3. This probably due to estimation of the amount of binding sites qe , which was determined from the y-intercept (t = 0). The intercept is strongly affected by the short term of lead ions uptake by modified CA which is usually lower than the equilibrium uptake. Therefore, the uptake of Pb(II) ions onto modified CA is not a first-order reaction even when the correlation coefficient R2 is relatively high (R2 = 0.914). 3.6.2. Pseudo-secondorder model The pseudo-second order model is based on the hypothesis that the rate of uptake is proportional to the square of the number of unoccupied sites [36]. The linearized form of the Eq. (6) is expressed as: t 1 t = + 2 qt q K2 ∗ qe e

(6)

where k2 is the rate constant of second-order sorption (g/meq min). Fig. 6b shows the linearized second-order plot of t/qt against t according to Eq. (6) that led to the determination of the second-order rate constants (k2 ) (0.8904) and qe from the slope and y-intercept. The qe values (0.166 meq/g) were approximately close to the experimental result (0.1583 meq/g), which indicated the suitability of pseudo-second-order model as shown in Table 3. The R2 value of the correlation coefficient indicated that the uptake data for Pb(II) onto modified CA best fit the pseudo-second-order model (R2 = 0.999). The basic hypothesis subsequent to the pseudosecond-order model designated that chemisorptions played the most important role and may control the sorption process [37]. This means that the chemisorption rate of reaction would be proportional to the square number of metal concentration free sites that expressed to (qe -qt )2 in the second-order model [38], which indicated by the following equation: M + 2B ↔ B2 M

r = K[M][B]2

where M is the divalent metal ion (Pb 2+ ) which is binding to three free binding sites of B (modified CA) 3.6.3. Intra-particle diffusion model Diffusion mechanisms during the uptake process cannot be identified with pseudo-first-order and/or pseudo-second-order models. The intraparticle diffusion model has been used to explain the process of uptake that occurred on a porous sorbent. A plot of the amount of lead ion uptake by modified CA (qt in mg/g) and the square root of the time, gives the rate constant. It is calculated by using the intra-particle diffusion model given as Eq. (7) [39,40]. qt = ki t 0.5 + Ci

(7)

where, qt is the amount of lead ions uptake by sorbent (mg/g) at time t, ki is the intra-particle diffusion constant (mg/g min0.5 ) that indicates improvement in the rate of uptake and Ci is the intercept of the line (mg/g) which equal to intraparticle diffusion constant. It is directly proportional to the boundary layer thickness. Fig. 6c shows that the intra-particle diffusion of Pb(II) within modified CA occurred in three stages. Several authors [41–43] clarified the reason for these three stages, where the linear step is corresponding to a quite fast uptake of lead ions by modified CA and highly favorable chemical mechanism such as ion-exchange. The line in the initial stage doesn’t pass through the origin; this means that uptake is dominated by transfer of ions by external diffusion through the surface boundary layer (film diffusion) than for the intra-particle

Fig. 7. Effect of sorbent dose on the removal of lead ions by grafted CA at 25 ± 0.2 ◦ C, initial lead ion concentration 10 mg/L and at pH ranged from 4.5 to 5.0.

diffusion process. The second stage is the gradual sorption stage, where, the intraparticle diffusion starts to hinder (slow down) due to extremely low lead ion concentrations in the solution until final equilibrium third stage.

3.7. Effect of sorbent dose As the process of metal uptake by sorbent is mainly a surface fact, therefore, efficiency of uptake can be considerably affected by surface area and available active sites due to the mass amount of sorbent [44]. Fig. 7 shows the percentage of lead ions removal at different amounts of modified CA ranging from 0.25 to 2 g/L. At a sorbent dose of 0.5 g/L, the lead ions concentration on the sorbent surface and in their solution becomes in equilibrium with each other, therefore, there is no significant removal of Pb(II) ions by increasing of sorbent dose [45]. However, the incremental decrease of removal efficiency with increasing of the sorbent dose is mainly due to the aggregation and overlapping of active sites at higher sorbent masses which lead to decrease in the effective surface area required for sorption [44,46], interference between available binding sites with respect to active sites and insufficiency of Pb(II) ions in the solution [47].

3.8. Effect of lead ions concentration and isothermal modeling 3.8.1. Effect of lead ions concentration Table 4 shows the effect of different lead ions concentration ranged from 10 to 100 mgL−1 on the residual, uptake concentration (mg/L), uptake capacity and the removal percent of Pb(II) using modified CA under optimum operating conditions at pH ranged from 4.5 to 5.0, contact time (90 min) and sorbent dose (0.5 gm/L). From Table 4, it was noted that by increasing the Pb(II) ions concentration, the uptake capacity of the modified CA increases until reached the saturation point (60 mg/g) with a maximum equilibrium uptake capacity, 60 mg/g for Pb ions then the uptake capacity did not notably change with increasing of Pb(II) ions concentration. In addition, uptake amount of Pb(II) increases with increasing the initial Pb(II) concentration until reached constant value (30 mg/L). While, the removal percent of Pb(II) decreases with increasing the initial Pb(II) concentration. These results pointed that energetically less-favorable sites become involved with increasing Pb(II) concentration in the aqueous solution.

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Table 3 Summary of pseudo-first- and second-order rate constants and equilibrium for Pb(II) uptake using modified CA. Experimental Pb(II) Uptake

Pseudo-first-order kinetics

qe Expe. (meq/g)

qe (meq/g)

k1 (min−1 )

R2

Pseudo-second-order kinetics qe (meq/g)

k2 (g mg−1 min−1 )

R2

0.158301

0.085

0.03915

0.914

0.168

0.855

0.999

3.8.2. Isotherm models Langmuir, Freundlich and Dubinin–Kaganer–Radushkevich (DKR) sorption isotherms are the mainly used models for understanding the mechanism of Pb(II) ions sorption on grafted CA and can be used for the design purposes [33,48–50]. The Linear and simplified form of Langmuir equation is given by Eq. (8): 1 Ce Ce = + qe Kqmax qmax

(8)

where Ce is the equilibrium concentration of Pb(II) ions in solution (mg/L), qe is the amount sorbed at equilibrium onto grafted CA (mg/g), K and qmax are Langmuir constants related to sorption energy and sorption capacity, respectively. K represents enthalpy of sorption and should vary with temperature. qmax is the maximum sorption capacity of lead ions per unit mass of sorbent when all binding sites are occupied. Where, monolayer sorption onto a surface with a finite number of identical sites are the hypothesis of Langmuir isotherm [51]. The Simplified Freundlich equation is given by: log qe = log Kf +

1 log Ce n

(9)

where kf and n are the Freundlich constants and are related to the sorption capacity of the grafted CA and its intensity. Equilibrium data are the main aspect for selection between Langmuir and Freundlich isotherms [52]. Isotherm parameters of Langmuir and Freundlich are listed in Table 5. Langmuir isotherm assumes monolayer adsorption, the R2 values for Lead ions was 0.977, which revealed the extremely good applicability of the Langmuir model to these sorptions. The R2 values propose that the data are not good represented by Freundlich isotherm, which describes the sorption characteristics for the heterogeneous surface [53]. The experimental equilibrium sorption data were best fitted to Langmuir isotherm with sorption capacity of 66.67 mg Pb(II)/g for grafted CA as calculated from Fig. 8a and 8b. It is abroad model than the Langmuir isotherm, because it does not suppose a homogeneous surface or constant sorption potential [54,55]. The linear equation of DKR isotherm can be simplified to Eq. (10) ln qe = ln Xm − ˇε2

(10)

where Xm is the maximum amount of lead ions that can be sorbed onto unit weight of sorbents (grafted CA) i.e. sorption capacity

Fig. 8. (a) Langmuir, (b) Freundlich and (c) DKR sorption isotherms of Lead ion onto grafted CA.

(mol/g), ˇ is the constant related to the sorption energy (mol2 /J2 ); and ε is the Polanyi potential, which is equal to: ε = RT ln(1 +

1 ) Ce

(11)

where R is the gas constant (8.314 J/mol K), and T the absolute temperature (K) [56,57]. Fig. 8c represents a plot of ln qe versus ε2 , the slope gives the ˇ value (mol2 /J2 ) and the intercept yields the sorption capacity Xm (mol/g) of grafted CA. The sorption space in the vicinity of a solid surface is characterized by a series of

Table 4 Effect of initial Pb (II) ions concentration on sorption capacity of grafted CA at optimum operating condition. Initial conc. (mg/L)

Residual conc. (mg/L)

Uptake conc. (mg/L)

Uptake capacity (mg/g)

Percent of removal

10 25 50 75 85 100

1.37 6.75 28.5 45 55 70

8.63 18.25 21.5 30 30 30

17.26 36.5 43 60 60 60

86.3 73 43 40 35.29 30

N.A. Abdelwahab et al. / International Journal of Biological Macromolecules 79 (2015) 913–922

921

Table 5 Summary of isotherm models parameters for Pb (II) uptake on modified CA. Langmuir model

Freundlich model

Dubinin–Kaganer–Radushkevich (DKR)

K L/mg

qmax (mg/g)

R2

Kf

n

R2

Xm (mol/g)

ˇ (mol2 /j2 )

E,KJ/mol

R2

31.39

66.67

0.977

16.98

3.23

0.95

8.22 × 10−4

−0.255 × 10−8

14.01

0.96

Table 6 Six repetitive cycles of loading and unloading of Pb(II) ions on grafted CA at optimum operating condition. Q (mg/g)

Cremoved

60 60 60 60 59.78 59.7674

30 29.4 28.8 27.6 26.9 25.7

V(ml) 100 98 96 92 90 86

W(g/L)

Cycles

0.5 0.49 0.48 0.46 0.45 0.43

1 2 3 4 5 6

of Pb(II) from modified CA at optimum operating condition with constant uptake capacity of 60 mg/g. Scheme 2. Ion exchange mechanism for the modified CA.

4. Conclusions equipotential surfaces having the same sorption potential and the free energy of sorption is the free energy change when one mole of ions is transferred to the surface of sorbents from infinity in the solution. Therefore, the sorption energy can be calculated using Eq. (12): E=



1

−2ˇ

(12)

Table 5 shows the DKR constants, which is calculated from the slope of the line. It shows that the E value is 14.01 kJ/mol for Pb(II) ions onto grafted CA. The E value is positive which indicates that the sorption process is endothermic and the magnitude of E can be related to the reaction mechanism. If E is in the range of 8–16 kJ/mole, sorption is governed by ion-exchange mechanism [58–60]. The grafted CA has amide, hydroxyl and carboxylic acid groups, which act as a potential ion-exchange/complexing groups that are capable of retaining lead ions [61]. A possible mechanism of ion exchange could be considered as Pb2+ attaching itself to adjacent functional groups which could donate two pairs of electrons to lead ions, forming chelated compounds and releasing two hydrogen ions into solution [62–64]. In ion-exchange mechanism, deprotonation of the functional groups of the graft copolymer (amide, hydroxyl or carboxylic acid groups) occurs firstly then the attachment of the metal cation to the reactive anion through complexation effect and this was illustrated in Scheme 2. 3.9. Elution of lead ions and reuse of modified CA Regeneration of modified CA is a key factor in the economic costs of the treatment processes [65]. Therefore, elution of the Pb(II) loaded onto modified CA was performed in a batch experimental setup using 0.1 M HCl. The acid elution could disrupt the coordinate spheres of the chelated Pb(II) ions and released from the polymer surface into the desorption media. As shown in Table 6 for the same sample of modified CA, the sorption/desorption behavior of Pb(II) from modified CA is stable for six cycles, taking into consideration the amount of the sorbent and the volume of Pb(II) solution in accordance with the loss of the sorbent dose during each cycle to be with comparable measure. The Cremoved almost equals to the Crecovered for the six repetitive cycles of loading and unloading

Cellulose acetate CA was modified by grafting with an equimolar binary mixture of acrylic acid (AA) and acrylamide (AAm) via free radical polymerization technique using benzoyl peroxide as an initiator. The optimum degree of grafting was achieved at monomer-to-initiator molar ratio 1, temperature 80 ◦ C and time 3 h. The Langmuir isotherm and pseudo-second-order kinetic model provide the best correlation with the experimental data for lead ions uptake by the modified CA. The models suppose that as the CA-g-(AA-co-AAm) surface, containing reactive OH groups, COOH groups as well as CONH2 groups which bind lead-ions, is homogenous and the operating mechanism is chemisorption involving valency forces through sharing or exchange of electrons between lead and modified CA. Also, the homogenous surface of modified CA provides multi-sites to the lead ions uptake. The reuse of modified CA as a novel material for removal of Pb(II) from wastewater was maintained almost constant even after six cycles. Acknowledgments The authors gratefully acknowledge the central laboratory, National Research Centre for its technical support. References [1] M. Hua, S. Zhang, B. Pan, W. Zhang, L. Lv, Q. Zhang, Heavy metal removal from water/wastewater by nanosized metal oxides: a review, J. Hazard. Mater. 211–212 (211) (2012) 317–320. [2] J.G. Lestón, J. Méndez, E. Pásaro, B. Laffon, Genotoxic effects of lead: an updated review, Environ. Int. 36 (2010) 623–636. ˜ ˜ [3] J. Cruz-Olivares, C. Pérez-Alonso, C. Barrera-Díaz, F. Urena-Nu nez, M.C. Chaparro-Mercado, B. Bilyeu, Modeling of lead(II) biosorption by residue of allspice in a fixed-bed column, Chem. Eng. J. 228 (2013) 21–27. [4] F. Fu, Q. Wang, Removal of heavy metal ions from wastewaters: a review, J. Environ. Manag. 92 (2011) 407–418. [5] M.A. Hashim, S. Mukhopadhyay, J.N. Sahu, B. Sengupta, Remediation technologies for heavy metal contaminated groundwater, J. Environ. Manag. 92 (2011) 2355–2388. [6] J.F. Peng, Y.H. Song, P. Yuan, X.Y. Cui, G.L. Qiu, The remediation of heavy metals contaminated sediment, J. Hazard. Mater. 161 (2009) 633–640. [7] W.C. Leung, M.F. Wong, H. Chua, W. Lo, C.K. Leung, Removal and recovery of heavy metals by bacteria isolated from activated sludge treating industrial effluents and municipal wastewater, Water Sci. Technol. 41 (2000) 233–240. [8] T.A. Kurniawan, G.Y.S. Chan, W.H. Lo, S. Babel, Comparisons of low-cost adsorbents for treating wastewaters laden with heavy metals, Sci. Total Environ. 366 (2005) 409–426. [9] D. Yuehua, G. Zhanqi, L. Benzhi, H. Xiaobin, W. Zhongbo, S. Cheng, Selective removal of lead from aqueous solutions by ethylenediamine modified Attapulgite, Chem. Eng. J. 223 (2013) 91–98.

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