Removal of Zn2+ and Pb2+ ions from aqueous solution using sulphonated waste polystyrene

Journal of Environmental Chemical Engineering 3 (2015) 2528–2537

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Removal of Zn2+ and Pb2+ ions from aqueous solution using sulphonated waste polystyrene Deborah Ruziwaa , Nhamo Chaukurab,* , Willis Gwenzic , Innocent Pumured a

Harare Polytechnic, PO Box CY 407, Causeway, Harare, Zimbabwe Department of Polymer Science & Engineering, Harare Institute of Technology, PO Box BE 277, Belvedere, Harare, Zimbabwe Department of Soil Science & Agricultural Engineering, University of Zimbabwe, Box MP 167, Mt. Pleasant, Harare, Zimbabwe d School of Environmental, Physical and Applied Sciences, University of Central Missourri, 108 W South St., Warrensburg, MO 64093, USA b c

A R T I C L E I N F O

A B S T R A C T

Article history: Received 24 April 2015 Accepted 8 August 2015 Available online 14 August 2015

We report the use of sulphonated waste polystyrene (SWPS) in the removal of heavy metal ions from water. Waste polystyrene (WPS) comprising of high impact polystyrene (HIPS) and expanded polystyrene (EPS) collected from dump sites in Harare were activated through sulphonation to produce a cationexchange resin. The presence of the sulphonic group was studied with fourier transform infrared spectroscopy (FTIR). Flame atomic absorption spectrometry (FAAS) was used to determine the concentrations of residual metals (Zn2+, Pb2+) after batch adsorption experiments. Sulphonated HIPS reduced Zn2+ from 80 to 38.3 mg/L compared to 10–1.6 mg/L for sulphonated EPS. Similarly, sulphonated HIPS reduced Pb2+ from 100 to 33 mg/L compared to 80–50.3 mg/L for sulphonated EPS. The adsorption data followed both the Langmuir and Freundlich isotherms and pseudo-second order kinetics. Maximum adsorption capacities as quantified by the Langmuir parameter qmax for HIPS was 5.01 mg/g, EPS 0.38 mg/ g for Zn2+ and HIPS 6.80 mg/g, EPS 0.68 mg/g for Pb2+. The data were analysed using pseudo first order and pseudo second order Lagergren equation and the adsorption kinetics of the metals Pb2+ and Zn2+ was found to follow the pseudo second order kinetic model. Interpretation of the sorption data in terms of separation factor (SF) suggested that the removal of Pb2+ and Zn2+ from water mainly occurred through chemisorption. ã 2015 Elsevier Ltd. All rights reserved.

Keywords: Adsorption isotherm Adsorption kinetics Cation-exchange Recycling Sulphonation Waste polystyrene

Introduction There is a rapid growth in plastics consumption throughout the world with polystyrene as one of the major plastics that is being used for producing a variety of materials for a wide range of applications. The forms of polystyrene widely used in industry are general purpose polystyrene (GPPS), expanded polystyrene (EPS), high impact polystyrene (HIPS), and syndiotactic polystyrene (SPS). Most polystyrene applications are for products with a limited life-cycle and then the vast majority of these are discarded. Owing to its low density, polystyrene can be easily scattered by the wind in the environment, creating a visible nuisance. Polystyrene is nonbiodegradable, implying that it persist in the environment and accumulates in landfills resulting in reduced landfill capacity.

* Corresponding author. Tel.: +263 782209671; fax: +263 4741422. E-mail address: [email protected] (N. Chaukura). http://dx.doi.org/10.1016/j.jece.2015.08.006 2213-3437/ ã 2015 Elsevier Ltd. All rights reserved.

Moreover, in most developing countries, landfills and incineration facilities are still lacking, thus most of the waste polystyrene is disposed of through open burning. Both open-burning and incineration emits toxic gases (e.g. PAHS and dioxins) and contribution to greenhouses gases causing climate change [1–3]. Therefore, it is desirable to develop an effective method that uses the waste materials [4]. There is clearly a need to move towards recycling for environmental, as well as economic reasons. Recycling of polystyrene is now a well-established practice. However, in most countries there is a restriction on the use of recycled polystyrene in food applications [5]. This work therefore, focuses on converting EPS and HIPS to an ion exchange resin (SWPS) through sulphonation [6–9]. These resins have a wide range of applications including water and wastewater treatment, remediation of contaminated soils and the purifications of industrial products through bioseparation and chromatographic applications [7]. The objective of the current study is to use kinetic and adsorption studies to assess the efficacy of SWPS in removing Zn2+ and Pb2+ through batch experiments.

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Experimental

Application testing using batch experiments

Sulphonation of WPS

Pb2+ and Zn2+ adsorption Batch adsorption experiments were conducted at pH 7 at six initial concentrations of Zn2+ and Pb2+. For each adsorbent, SWPS (0.25 g) was weighed into a 100 mL beaker, and 50 mL of each solution was added. The mixture was shaken with a platform shaker for 2 h at 200 rpm and at room temperature. The samples were diluted 10-fold for Pb2+ and 20-fold for Zn2+ before being filtered and the filtrate was analysed for Zn2+ and Pb2+ using a calibrated AAS (Model: Savant AA, Melbourne, Australia).

Coffee cups, meat trays, fast food containers and packaging of computers were crushed with and passed through a 4.75 mm diameter sieve. A portion of the material (5 g) was measured into a conical flask containing 95% sulphuric acid (100 mL). The mixture was reacted under agitation for 1 h at 70
C. Thereafter, the slurry was filtered with a funnel and washed with 250 mL portions of distilled water. After the sixth wash, the pH was measured with a pH meter (Model: Li-702, Pachikula-Haryana, India) to ensure the residual sulphuric acid has been removed from the resin. The sulphonated resin was then dried in an oven (Model: BJPX Spring, Shandong, China) at 40
C for 30 min.

Preparation of Pb2+ and Zn2+ solutions All chemicals were obtained from Sigma Aldrich and used without further purification unless specified. Six Pb2+ and Zn2+ solutions of concentrations 5, 10, 20, 40, 80, and 100 mg/L were prepared by diluting the initial Pb2+ and Zn2+ stock solutions (prepared from Pb(NO3)2 and ZnCl2 respectively) of 1000 mg/L.

Adsorption kinetics. To determine the effect of retention time on adsorption, batch experiments were conducted at pH 7 at four initial concentrations of Zn2+ and Pb2+. For each adsorbent, SWPS (0.25 g) was weighed into a 100 mL beaker, and 50 mL of each metal ion solution was added. The mixture was shaken with a platform shaker for 30 min, 1 h, 2 h and 3 h, at 200 rpm and at 25
C. The samples were filtered through a 45 mm Whatman filter paper and the filtrate analysed for Zn2+ and Pb2+ using an AAS. The results of the kinetics study for a system with qe (mg/g), qt (mg/g) as adsorption capacities at equilibrium and at time t, respectively, and rate constant of pseudo first order adsorption K1 (L/min), were tested with Lagergren’s pseudo first-order model given by Eq. (1) [12]: K1t 2:303

ð1Þ

Characterisation of SWPS

logðqe  qt Þ ¼ logqe 

The resulting SWPS was characterised for BET surface area (SBET), pore volume and pore size using an automated N2 adsorption analyser (Model: TriStar 3000 V6.08 A, Micromeritics, Norcross, USA), and CHNS microanalysis using an elemental analyser (Model: SDCHN435, HunanSundy, China). Background lead and zinc levels in the adsorbents were determined after digestion with aqua regia using AAS. The degree of sulphonation of the polymer was evaluated by titration. SWPS (0.3 g) was dissolved in about 30 mL of a toluene/methanol (9:1, v/v) [9]. A solution 0.02 M NaOH in methanol was used to titrate the polymer solution with phenolphthalein as indicator. The cation exchange capacity (CEC) of SWPS was determined by measuring the concentration of H+ that was exchanged with Na+ when acid-form of SWPS equilibrated with NaCl solution [9]. Dry SWPS (0.5 g) in the H+ form was placed in 100 mL of NaCl (0.2 M) solution and shaken occasionally for 2 h using a platform shaker (Model: Thermolyne, M50825, MA, USA) at a speed of 150 rpm. The amount of H+ released by the polymer was then determined by titration with 0.01 M NaOH. The sulphonic acid and the sulphonate groups in the SWPS were confirmed by FTIR (Model: Infra 3000 FTIR, WOF-520, Analytical Technologies, Mumbai, India). The spectra were recorded with characteristic peaks in the wave numbers ranging from 400 to 4000 cm1. Tablets of KBr pressed with polystyrene sulphonate were analysed in triplicate. 24 scans were performed using a resolution of 4 cm1.

The value of the constants qe,k1 and correlation co-efficient (R2) were obtained from the linear plot of log (qe  qt) against t. The pseudo second-order model for a system with the rate constant K2 (g/mg/min) is expressed in Eq. (2) [12]:

Effect of initial pH The adsorption procedure by Dong et al. [10] with an equilibration time of 1 h was used. The pH values investigated were concentrated in the acidic region because precipitation of Zn2 + and Pb2+ occurs under alkaline conditions [11]. Residual Pb2+ and Zn2+ after adsorption by sulphonated HIPS and sulphonated EPS were determined using AAS. A reactor with no adsorbent was used as a blank at each pH.

t 1 t ¼ þ qt K 2 q2e qe

ð2Þ

The values of K2, R2 and qe were calculated from the plots of t/qt versus t. The slopes and intercepts of the plots were used to determine qe and K2, respectively [12]. The sum of squares of error (SSE) and the coefficient of determination (R2), were used to test the validity of each kinetic model [13]. A high correlation coefficient and low SSE indicate high goodness of fit between the experimental data and the model [14]. Adsorption isotherms. Batch adsorption experiments were conducted at pH 7 at six initial concentrations of Zn2+ and Pb2+. Linearised Langmuir (Eq. (3)) and Freundlich (Eq. (5)) models were used to characterise the sorption data [15]: Cf 1 C1 ¼ þ C ads Q b qmax

ð3Þ

where Cf = equilibrium concentration of metal in solution, Cads is the number of ions sorbed onto the adsorbent, Q and b are Langmuir constants related to sorption capacity and sorption energy respectively. Maximum sorption capacity (qmax) represents monolayer coverage of sorbent with sorbate and b represents enthalpy of sorption and should vary with temperature. If the obtained chart is linear the constant Q will be the slope and b will be the y-intercept [16]. The separation factor parameter, SF, for an initial metal concentration C0 (mg/L), and Langmuir constant KL was calculated from Eq. (4): SF ¼

1 1 þ K L Co

ð4Þ

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SF > 1 indicates unfavourable interaction between the adsorbate and adsorbent exists and in this case adsorption may not be possible. SF = 1, indicates a linear relationship between adsorbate and adsorbent exits, and the mechanism characterising the interactive process is a mixture of physisorption. SF = 0 indicates an indefinite sticking between sorbate and adsorbent and chemisorption predominates. The linearised form of the Freundlich isotherm equation for an equilibrium concentration Ce (mg/L), and concentration of adsorbed metal Cads (mg/L) [17] is shown in Eq. (5).

lower by about 16.0%. Although the BET surface area (SBET) of sulphonated HIPS (2.20  103 m2/g) is lower than that of sulphonated EPS (7.69  102 m2/g), the pore volumes and pore sizes are higher at 1.74  102 cm3/g and 3.15  105 Å, respectively. The higher pore volume and pore size could possibly account for the increased propensity of HIPS to adsorbing the bigger Pb2+ compared to Zn2+ ions whose ionic radii are 1.19 and 0.74 Å, respectively [19]. The low SBET values as compared to biosorbents [20,21] suggest that the adsorption mechanism might not be significantly influenced by surface area.

1 logC ads ¼ log K þ logC e n

FTIR spectra of SWPS

ð5Þ

where K is a parameter related to temperature and n is a characteristic constant for the adsorbent under study. Linearised relationships are obtained upon plotting log Cads versus log C from which n and K can be determined [16]. Values of the adsorption parameters were deduced from the slopes and intercepts of the plots together with the correlation coefficients (R2) after linear regression [13]. The values of n and Kf are the slope and curvature of the isotherm respectively. Kf is a measure of the adsorption capacity and n the intensity of the adsorption process. The Freundlich isotherm is convex for 1/n < 1, linear for 1/n = 1 and concave for 1/n > 1. The lower the value of 1/n, the higher the affinity and heterogeneity of the adsorbent sites. Low Kf values indicate unfavourable adsorption [13]. Statistical data analysis Experimental data were tested for the analysis of variance (ANOVA) assumptions before testing for the effects of initial solution pH, initial concentration of metal ions and adsorbents on adsorption. A probability level, p of 0.05, was used for all statistical comparisons. The statistical software IBM1 SPSS Statistics version 19 was used for statistical analysis. Results and discussion Physicochemical characterisation Measurements on the background concentrations of Pb2+ and Zn in the HIPS and EPS feedstocks gave 0.00 mg/g for each of these metals. This was expected since neither these metals nor their salts are routinely added during polymerisation or to the virgin materials during processing. Both sulphonated HIPS and sulphonated EPS were alkaline with pH values of 9.45 and 9.95, respectively (Table 1). Activated carbons have pH in the region 3.9– 4.3 [18]. The respective EC values were remarkably different at 376.7 and 1493.3 mS/m respectively. This suggests sulphonated EPS has a higher concentration of inorganics, this could account for the lower adsorption capacity for both Pb2+ and Zn2+. Both these EC values were however, lower than those of biosorbents derived from water hyacinth and sawdust which were in the ranges 103– 173.2 and 2.3–3.9 mS/m respectively [18]. The water content of sulphonated HIPS is slightly higher (about 5.79%) than that of sulphonated EPS. The C and H content of sulphonated HIPS are lower than that of sulphonated EPS. Consistent with the higher degree of sulphonation for the latter material, the S content was 2+

H out of plane The peaks at around 750 cm1 are for the C normal vibrations of the phenyl groups that are found on the polystyrene. The peaks at 700 cm1 are due to the presence of isotactic polystyrene in both HIPS and EPS materials. Significant differences in the FTIR spectra of sulphonated and unsulphonated material were observed. In both cases, the S O vibration appears as a very broad band at approximately 1100–1350 cm1. OH stretching of the SO3H bands were also observed from 3000 to 3500 cm1 (Fig. 1(a) and (b)). This agrees with observations by Brandao et al. [22]. New peaks (not on WPS spectra) due to the S OH bond appeared between 650 and 700 cm1[23]. Another new band appeared at 1035 cm1 upon sulphonation indicating the presence of sulphonic groups in the SWPS. The band resulted from the symmetric stretching vibration of the sulphonic groups [24]. Another band was observed at 1129 cm1 and resulted from the sulphonated anion attached to the phenyl group. Similar results were observed by Bekri-Abbes [9] and Martins et al. [25]. New peaks were also observed at wavenumbers approximately from 800 to 830 cm1. These absorbances show the C S bonding of sulphonic groups to the aromatic group of the polystyrene [26]. This is because of out of plane deformation bands assigned to substituted aromatic ring. The peak at 975 cm1 disappeared after sulphonation, probably because of the out of plane deformation of the CH¼CH stretch. Overall, comparison of the FTIR spectra for WPS and SWPS confirmed that sulphonation occurred as evidenced by the presence of the sulphonic group on the SWPS. Cation exchange capacity and degree of sulphonation CEC and DOS blanks for WPS were carried out and negative results, in comparison to SWPS, were obtained, confirming the sulphonic groups improved the CEC properties of WPS. The CEC and DOS for sulphonated EPS were higher than that of sulphonated HIPS. This could be related to the structural morphology of sulphonated EPS as compared to sulphonated HIPS which is more compact and a bit rigid. The resin that produced the maximum DOS was EPS (DOS = 26.2%) and sulphonated HIPS (DOS = 24.1%) at the same temperature and time. Bekri-Abbes et al. [27] obtained best values of DOS for samples treated at 40
C for 4 h as 16% and for 60
C for 1.5 h as 17%. The CEC for the sulphonated EPS was 1.85 mEq/g whereas that of the sulphonated HIPS was 2.3 mEq/g. These values are comparable to the 3.1 mEq/g obtained by Unnikrishnan et al. [28] which is similar to that of the commercial resin Amberlite IR (3–5 mEq/g). Kim et al. [29] also produced two sulphonated membranes with CEC of 2.95 and 3.2

Table 1 Physicochemical characteristics of the adsorbents. Adsorbent

pH

EC (mS m1)

(%) water

SBET (m2/g)

Pore volume (cm3/g)

Pore size (Å)

(%) C

(%) H

(%) N

(%) S

Sulphonated HIPS Sulphonated EPS

9.45 9.95

376.7 1493.3

3.80 3.58

0.0022 0.0769

0.01738 0.00567

3.15 x 105 2.95 x 103

79.23 79.85

9.24 9.4

0.00 0.00

3.26 3.88

D. Ruziwa et al. / Journal of Environmental Chemical Engineering 3 (2015) 2528–2537

Fig. 1. FTIR spectra of WPS (dotted line) and SWPS (bold line) for EPS ((a), (c) and (e)), and for HIPS ((b), (d) and (f))

Fig. 2. Relationship between cation exchange capacity (CEC) and degree of sulphonation (DOS). Errors indicate standard error of the mean.

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mEq/g. As expected, a positive relationship was observed between CEC and DOS (Fig. 2). The more effective the sulphonation process, the more efficient the SWPS will be in removing cations. The higher CEC and DOS observed for sulphonated EPS than sulphonated HIPS could be related to the structural morphologies of the two resins. For instance, sulphonated HIPS is more compact and rigid than sulphonated EPS. Effect of initial solution pH The relationship between pH and adsorption was nonlinear for both adsorbents (Fig. 3). The percentages of Pb2+ and Zn2+ adsorption for both adsorbents increased with increasing initial solution pH up to a maximum at pH 7 for both adsorbents, hence all other adsorption and kinetic studies were carried out at pH 7. The maximum percentage adsorption for sulphonated HIPS for Pb2+ and Zn2+ were 67.00% and 52.13% respectively, and that for sulphonated EPS were 37.13% and 84.00% for Pb2+ and Zn2+ respectively. Blank runs at the different pH values indicate no or insignificant adsorption (0.00– 0.008% adsorption) by the reactor. This is to be expected since the reactor was made of borosilicate glass which is relatively inert under the experimental conditions. The influence of pH on Zn2+ and Pb2+ adsorption can be attributed to the altering of solubility and speciation of the ions, and to the nature of adsorption sites on adsorbents. At pH values lower than pH 5 there is minimum removal due to competition between Pb2+ and Zn2+ ions with the H+ ions for the sulphonic adsorption sites on the adsorbents. The H+ ions being highly concentrated and mobile are adsorbed preferentially and reduce the population of available adsorption sites [20]. Above pH 7. Pb2+ and Zn2+ ions are expected to precipitate out of solution [11]. Generally, the pH of a solution affects the ionisation and speciation of the adsorbate and the overall surface charge of the adsorbent [30]. Solution pH is an important parameter in the adsorption process, and the initial pH of the solution has more influence than the final pH. The change of pH affects the adsorptive process by influencing the dissociation of functional groups and the active sites on the adsorbent, which alters reaction kinetics and equilibrium characteristics of the absorption process. Even though the sorption is affected by pH, the % Pb2+ removed did not vary much with pH and therefore HIPS might be applicable to remove significant amounts of Pb2+ ions in solutions over a wide range of pH values in acid/alkaline mine drainages. Effect of initial concentration Adsorption by both sulphonated HIPS and sulphonated EPS generally increased linearly and approached a maximum at an

initial concentration ranging from 30 to 40 mg/L (Fig. 4). For Zn2+, the equilibrium adsorption increased linearly with initial concentration up to about 40 mg/L at which the equilibrium Zn2+ adsorption is about 5000 mg/L. At this point it attained a maximum and thereafter decreased. This seems to suggest that adsorption sites on the adsorbent get saturated at 40 mg/L, beyond which it would require frequent regeneration. The current study did not investigate the stability of the adsorbents for repeated use, hence parameters like eluent type and concentration were not reported. This is the subject of future research with our research group. These results also suggest sulphonated HIPS is a more efficient adsorbent than sulphonated EPS. While EPS is a mainly made up of styrene, HIPS is a graft copolymer of styrene and butadiene. The paraffinic chain introduced by the butadiene moiety allows for free rotation about C C s-bonds and hence provides more efficient packing of the HIPS repeat unit. This results in a higher density of HIPS compared to EPS. It is likely that the polymer chains orientate themselves so that the phenyl group is more exposed to sulphonation during the activation process. Although the Pb HIPS graph seems to be still increasing beyond an initial concentration of 80 mg/L, the decrease in slope at 40 mg/L suggests saturation is beginning to set in. The Pb EPS graph however, shows a bimodal variation suggestive of two different adsorption mechanisms. These could be adsorption due to the surface and intraparticle diffusion causing the peak at an initial concentration of 20 mg/L, followed by another adsorption peak at 40 mg/L due to the sulphonic functional groups. One important variable in metal ion adsorption is the effect of initial concentration of the metal since this determines the maximum uptake capacity of the polymer. Bekri-Abbes et al. [9] measured the amount of metal ions absorbed from solution onto the sulphonated EPS absorbent after sufficient contact to reach equilibrium. Percentage adsorption of Cd2+ and Pb2+ decreased with increasing metal concentration in aqueous solutions. The results indicated that energetically less favourable sites become involved with increasing metal concentrations in the aqueous solution. Adsorption isotherms The adsorption data followed both Langmuir and Freundlich isotherm models suggesting multilayer adsorption or the occurrence of both surface and intra-particle diffusion [18]. WPS showed no heavy metal adsorption capacity. Based on the Langmuir model, maximum Zn2+ adsorption capacity (qmax) decreased in the order: sulphonated HIPS (5.01 mg/g) > sulphonated EPS (0.38 mg/g). For Pb2+ the qmax decreased as sulphonated

Fig. 3. Effect of initial solution pH on % removal of (a) Pb2+ and (b) Zn2+ in aqueous solution by sulphonated HIPS and sulphonated HIPS. Errors bars indicate standard error of the mean.

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Fig. 4. Effects of initial concentration (Co) on equilibrium adsorption (qe) for Zn2+ and Pb2+ by SWPS. Error bars indicate standard error of the mean.

Fig. 5. Linearised Langmuir (a) and Freundlich (b) adsorption isotherms for adsorption of Pb2+ and Zn2+ on SWPS. Error bars indicate standard error of the mean.

Table 2 Langmuir and Freundlich parameters for adsorption of (a) Zn2+ and (b) Pb2+. (a) Langmuir and Freundlich parameters for adsorption of Zn2+ Material

HIPS EPS

Langmuir model

Freundlich model

qmax (mg/g)

b (L/mg)

R2

Kf

n

R2

5.0128 0.3847

0.0634 0.3675

0.9544 0.5223

0.5358 0.1247

0.6513 0.2282

0.9064 0.7826

(b) Langmuir and Freundlich parameters for adsorption of Pb2+ Material

HIPS EPS

Langmuir model

Freundlich model

qmax (mg/g)

b (L/mg)

R2

Kf

n

R2

6.8049 0.6776

0.121 0.2315

0.9867 0.7018

1.0177 1.261

1.3452 1.2672

0.949 0.7898

HIPS (6.80 mg/g) > sulphonated EPS (0.68 mg/g). Both Zn2+ (R2 = 0.95 and 0.91) and Pb2+ (R2 = 0.99 and 0.95, p < 0.05) adsorption on sulphonated HIPS and sulphonated EPS obeyed the Langmuir and Freundlich models (Fig. 5, Table 2).

Adsorption data following the Langmuir model indicate homogeneous monolayer or surface adsorption of Zn2+ and Pb2+ on the SWPS resins, while those that followed the Freundlich isotherm model suggest multilayer adsorption or the occurrence of both surface and intra-particle diffusion. Similar behaviour has been reported for sulphonated biochar [18]. This confirms the insertion of sulphonic groups, which improve the adsorption capacity of the waste polystyrene. The high concentration of sulphonic groups renders the SWPS strongly hydrophilic and consequently, the interaction with the metal ions by electrostatic attraction is facilitated with the increase of ions solvation [13]. Langmuir isotherms indicate high qmax values for both Zn2+ and Pb2 + adsorption by HIPS. Mahamadi and Madocha [31] suggested that high qmax values confirm stronger bonding affinity. The degrees of correlation for adsorption of Zn2+ and Pb2+ by sulphonated HIPS are also very high (R2 = 0.90–0.99, p < 0.05) suggesting a single surface reaction with constant activation energy as the predominant sorption step [13]. Results for adsorption of Zn2+ and Pb2+ by sulphonated EPS show acceptable adsorption capacity although they are lower than those of the sulphonated HIPS which are more favourable as indicated by the

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higher R2 values. Kf values for adsorption of Pb2+ were high for both sulphonated HIPS and sulphonated EPS. Both high values for Kf and n > 1 indicate favourable adsorption [13,32]. Equilibrium adsorption isotherms are of fundamental importance in the study of adsorption systems since they indicate how metal ions partition themselves between the media and liquid phase with increasing concentration at equilibrium. SF < 1 at all concentrations (Fig. 6) indicating that adsorption is favourable for the removal of both Pb2+ and Zn2+[31]. The results also confirm that the adsorption process becomes increasingly favourable as the initial metal concentration increased. Table 3 compares Pb2+ and Zn2+ adsorption capacities of different adsorbent materials. The adsorption efficiencies of the materials under study are at least an order of magnitude lower than the other adsorbents. This difference could be due to all the other adsorbents being derived from pyrolysed biomass, a natural polymer, whereas SWPS is derived from a synthetic polymeric material. Pyrolysed biomass has numerous inherent functional groups and increased porosity and surface area conferring the material with a more complex and efficient adsorption mechanism [18]. Despite the lower adsorption capacities of sulphonated EPS and sulphonated HIPS, the removal efficiencies (HIPS: 51.1% and 67.0% for Zn2+ and Pb2+ respectively, and EPS: 84.0 and 29.7% for Zn2+ and Pb2+ respectively) are still high enough to make the materials useful. The major advantage is that the feedstocks for these two materials are waste plastics whose conversion to a useful material also helps clean up the aesthetic environment.

Adsorption kinetics The batch reactor attained equilibrium in 90 min on average. Adsorption kinetics describe metal ion uptake rate controls with residence time of adsorbate uptake at solid and solution interface [33]. The Lagergren model for both Pb2+ and Zn2+ adsorption by sulphonated HIPS and sulphonated EPS does not properly fit the pseudo first-order kinetic data which gave regression R2 < 0.9 (p < 0.05). The pseudo first order rate constant k1 ranged between 2.3  104 and 9.2  104 L/min. The k2 for Pb2+ adsorption on sulphonated EPS decreased from 2.7  101 to 7.2  103 at 5– 100 mg/L. Equilibrium sorption capacity (qe) for Pb2+ by sulphonated EPS increased from 0.37 to 6.46 mg/g with increasing initial metal concentration (Fig. 7). The rate constant k1 ranged between 2.1 103 and 7.1 103 L/min. The k2 for Zn2+ also decreased from 2.77  101 to 4.73  103 with increasing initial metal concentration. Equilibrium sorption capacity (qe) for Zn2+ increased from 1.00 to 7.46 mg/g with increasing initial metal concentration. The results showed clearly that k1 is independent of initial concentration. Similar results have been presented by Abechi et al. [14]. The values of the initial adsorption rate, h increased from 1.5  103 to 6.9  103 mg/g min, indicating that adsorption was faster. The rate constant k1 ranged between 2.3  104 and 9.2  104 L/min for adsorption of Pb2+ and for Zn2+ the rate constant k1 ranged between 2.07  103 and 7.10  103 L/min. In Fig. 7(b), the k2 for Pb2+ by sulphonated EPS also decreased from 2.72  101 to 7.189  103 as the initial concentration increases.

Fig. 6. Adsorption favourability for Zn2+ and Pb2+ on SWPS. Error bars indicate standard error of the mean.

Table 3 Comparison of the maximum adsorption capacities, pH and equilibration times for different adsorbents. Adsorbent

Microwaved olive stone Conventional olive stone Sulphonated sawdust Sawdust biochar Sulphonated HIPS Sulphonated EPS

Adsorption capacity (mg g-1) Pb2+

Zn2+

23.47 22.37 – – 6.80 0.68

15.08 11.14 21.01 7.36 5.01 0.38

pH

Equilibration time (min)

Reference

5.0 5.0 6.0 3.0 7.0 7.0

180 – 120 120 90 90

[21,20] [30,36] [18] [18] Present work Present work

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Fig. 7. Lagergren kinetic plots for sulphonated EPS (a) Pb2+ (first order), (b) Pb2+ (second order), (c) Zn2+ (first order), and (d) Zn2+ (second order). Error bars represent standard error of the mean.

Fig. 8. Lagergren kinetic plots for sulphonated HIPS (a) Pb2+ (first order), (b) Pb2+ (second order), (c) Zn2+ (first order), and (d) Zn2+ (second order). Error bars represent standard error of the mean.

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However, the kinetic data gave regression R2 > 0.9 (p < 0.05), for pseudo second-order for both Pb2+ and Zn2+ adsorption by sulphonated HIPS and sulphonated EPS (Fig. 7(b) and (d), and Fig. 8(b) and (d)) implying that the pseudo second-order model provided the best fit for the adsorption data [33]. For both adsorbents, the Langergen model fits well with the whole range of contact time and both metal ions. When this happens the ratelimiting step appears to be controlled by chemisorption [34]. The pseudo first order rate constant k1 ranged between 6.9  103 L/ min and 1.8  103 L/min. Initial adsorption rate for Pb2+ was higher that of Zn2+ and this is an indication that adsorption of Pb2+ was faster. Shahmohammadi-Kalalag et al. [35] obtained similar results and suggested that, in multi-component aqueous solutions of metal ions, the metal with the higher initial adsorption rate may be selectively removed before other metals. Using microwave irradiated olive stone as an adsorbent, Alslaibi et al. [36] found the adsorption capacity of Pb2+ was higher than that of Zn2+. The electronegativities of Pb2+ and Zn2+ are 2.33 and 1.65, respectively [19]. This confers Pb2+ with a higher affinity for the adsorbent. The ionic radii of Pb2+ and Zn2+ are 1.19 and 0.74 Å respectively. Thus Zn2+, being the smaller cation, is expected to have greater access to the adsorption sites than Pb2+. However, it seems from the results that the higher electronegativity of Pb2+ overrides its relatively large ionic radius, thus this could be the dominant factor that describes how the divalent cationic species are attaching to the sulphonated waste polystyrene. This phenomenon may find use in heavy metal separation through hydrometallurgy or similar applications. The pseudo second-order rate constant k2 for Pb2+ decreased from 6.67  102 to 3.58  103 L/min as the initial concentration increases. Increasing the Pb2+ and Zn2+ concentration seems to reduce the diffusion of the metal ions in the boundary layer and to enhance diffusion in the SWPS resin. Equilibrium sorption capacity (qe) for Pb2+ increased from 2.73 to 31.06 mg/g as the initial concentration increased, and the h values increased from 0.50 to 3.45 mg/g min. This is an indication that adsorption was fast. Equilibrium sorption capacity (qe) of Zn2+ increased from 0.82 to 12.59 mg/g. The values for h increased from 0.10 to 2.63 mg/ g min. This is an indication that adsorption is fast. Similar results were obtained by Abechi et al. [14]. The correlation coefficient (R2, p < 0.05) for the pseudo first-order ranged between 0.59 and 0.84 for Pb2+ adsorption by sulphonated HIPS, 0.60–0.89 for Pb2+ adsorption by sulphonated EPS and from 0.81 to 0.89 for Zn2+ adsorption by sulphonated HIPS, 0.48–0.86 adsorption by sulphonated EPS. These results were 14–52% lower (p = 0.01) than those of the pseudo second-order which ranged from 0.99 to 1.00 mg/g for Pb2+ adsorption by sulphonated HIPS, about 1.0 for Zn2+ adsorption by sulphonated HIPS, 0.992–0.999 for Pb2+ adsorption by sulphonated EPS and 0.98–0.99 for Zn2+ adsorption by sulphonated EPS (p < 0.05). The k2 for Pb2+ and for Zn2+ decreased from 2.72  101 to 7.19  103 and from 2.77  101 to 4.73  103 L/min, respectively as concentration increased. Equilibrium sorption capacity (qe) for Pb2+ increased from 0.39 to 6.46 mg/g as the initial concentration increased, and that of Zn2+ increased from 1.00 to 7.46 mg/g. Conclusion The current study activated WPS by sulphonation and investigated the capacity of the sulphonated HIPS and EPS to remove heavy metals from aqueous solutions. FTIR confirmed successful sulphonation as evidenced by increase in CEC and adsorption capacity. Application of the SWPS showed that both sulphonated HIPS and sulphonated EPS were effective in the removal of Zn2+ and Pb2+ in solution by between 37 and 84%. The adsorption of Zn2+ and Pb2+ on both sulphonated EPS and HIPS

were governed by both Langmuir and Freundlich isotherms, and obeyed pseudo-second order kinetics. Overall, the findings indicate that SWPS is effective in removing Zn2+ and Pb2+ from aqueous solutions. Possible application of the SWPS include metal ion separation in hydrometallurgy, treatment of wastewater and industrial effluent such as acid/alkaline mine drainage. Previous researchers have sulphonated waste HIPS and carried out application tests without comparing the relative adsorption efficiencies of sulphonated HIPS and sulphonated EPS. This work has gone further to fill this gap and provide a kinetic treatment of Pb2+ and Zn2+ adsorption on both these materials. Further research should include investigating the regeneration capacity of the adsorbents, pilot application in contaminant remediation, and development and evaluation of composite resins incorporating biochar and clay.

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