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Water SA

versión On-line ISSN 1816-7950
versión impresa ISSN 0378-4738

Water SA vol.45 no.3 Pretoria jul. 2019


The actual concentrations of Cr(VI) formed during the pH investigation were well below the 50 μg·m£-1 drinking water standard. However, it would be misleading to extrapolate these bench-top experimental results directly to practical ozonation plant operations, since many parameters are likely to differ (e.g. solid loading, contact time, pH, O3 generator efficiency, agitation). It is, however, apparent from the results presented here that Cr(VI) could be formed through ozonation if Cr-containing materials are suspended in the water.

Influence of ozonation contact time

Rajagopaul et al. (2008) reported that the contact times during ozonation in operational plants are in the order of 3-6 minutes. However, contact time could even be longer, depending on the objective of the treatment and the compounds present (Beltrán, 2003). In order to establish the effect of ozonation contact time on Cr(VI) formation, contact times of 6, 12, 24 and 48 min were investigated at 2 pH values, i.e., pH 7, representing a neutral water system, and pH 10, since it was the optimum observed pH for Cr(VI) formation (Fig. 2). These results (Fig. 3) indicate an almost linear increase in Cr(VI) concentrations as a function of contact time, for the experimental conditions investigated.



Temperature effect

It is unlikely that ozonation will be applied at extremely high or extremely low water temperatures. Therefore, the effect of water temperature was investigated at 10, 20, 30 and 40°C. This temperature range is representative of normal and maybe slightly heated water, which might be applicable to some industrial waste waters (e.g. combution off-gas venturi scrubber water). As indicated in Fig. 4, increased water temperatures led to increased Cr(VI) formation. Aqueous O3 decomposition studies have indicated that higher temperatures lead to increased rates of ozone decomposition (Beltrán, 2003; Sotelo et al., 1987), hence higher concentrations of the more aggressive oxidation radicals, as discussed earlier.



Effect of solid loading

As expected, higher solid loading of the Cr-containing materials led to higher Cr(VI) concentrations (Fig. 5). These results represent relatively high solid loadings, which might never be achieved in water treatment applications. However, these high solid loadings assisted in identifying a trend.



Effect of gaseous O3 concentrations

The maximum O3 concentration in the O2 gas stream was limited by the ability of the O3 generator utilised in this study. All results presented in the previous sections of this paper were conducted at the maximum setting of the O3 generator utilised, i.e, achieving 0.0058 mg-m£-1 O3 in O2 gas. In order to assess the effect of O3 gaseous concentration on the formation of Cr(VI) during aqueous ozonation, two lower settings of the O3 generator were tested. These results are shown in Fig. 6. It is evident that lower gas O3 concentrations resulted in lower Cr(VI) formation.



Mechanism of Cr-liberation

In order to obtain insight into the actual liberation mechanism of Cr(VI) from the solid Cr-containing materials during ozonation, SEM elemental maps were generated for polished sections of the Cr-containing materials exposed to ozonation. Figures 7a and 7b indicate ozonated slag and ÜG2 ore particles, respectively. The colour distributions (colours representing different elements, as indicated on the images) of the two materials utilised clearly show that the slag is much more heterogeneous than the ÜG2 with regard to elemental distribution. This was expected, since the slag is a re-crystallised waste product, while the ÜG2 ore is more homogeneous. However, in neither of these two ozonated materials can any enrichment of a specific element be observed on the surface (outside) of the particles. For comparison, Fig. 7c is included, which indicates chromite ore particles which were treated in a different manner (not discussed in this paper), resulting in the enrichment of iron on the surface of the particles. The absence of enrichment of any elements on the surface of the ozonated slag and ÜG2 ore particles therefore indicates that ozonation did not selectively extract chromium or any other element. It is therefore most probable that entire surfaces of particles were eroded due to the strong oxidising conditions, resulting in the release and subsequent oxidation of chromium. Other elements present in the case study materials were therefore also likely to be liberated; however, these were not quantified since this was beyond the scope of the investigation.


The experimental differences observed between the two case study materials, i.e. UG2 ore and slag, could be due to several reasons. Chromite ore is likely to be more resistant to O3 erosion than the slag, due to the well-defined spinel crystal structure of the ore (Gu and Wills, 1988) and the less well-defined crystalline structure of the slag. Additionally Fe reduces at a lower temperature than Cr during the ferrochrome pyrometallurgical production process (Beukes et al., 2010). This results in the higher Cr:Fe ratio observed for the slag, if compared with the ore (Table 1). In the spinel crystalline structure of the ore most of the Fe occurs as Fe(II). Ore particle erosion by O3 oxidation, as indicated by the SEM elemental maps (Fig. 7), could therefore lead to the release of higher concentrations of Fe(II) than what would be expected for the slag O3 erosion. Fe(II) could consume O3 during its conversion to Fe(III) and Fe(II) is also a well-known reducing agent for Cr(VI). However, the possible release of Fe(II) from the ore spinel structure was not verified in this study and could be considered a future perspective in the clarification of the exact mechanism of this reaction system.



The experimental conditions employed in this study cannot be related directly to ozonation in drinking water or industrial wastewater treatment plants, since parameters (e.g. solid loading, ozone concentrations, mixing efficiencies) are likely to be different. Cr(VI) concentrations formed during the experimental ozonation conditions investigated can therefore not be used as a guide to predict possible Cr(VI) formation. However, the results clearly indicate that Cr(VI) can be formed in situ during ozonation of water with non-Cr(VI) Cr-containing materials in suspension. pH seems to be the most important parameter influencing the formation of Cr(VI), with higher pH levels favouring Cr(VI) formation. This can be attributed to the increased rate of aqueous O3 decomposition occurring at pH > 4, resulting in higher concentrations of hydroxyl radicals that are stronger oxidants than aqueous O3. Other parameters, such as contact time, water temperature, solid loading, ozone concentration and the characteristics of the Cr-containing material also have an influence on Cr(VI) formation.

The results indicate the importance of removing suspended particulates from water prior to ozonation. Although dissolved Cr(III) oxidation was not specifically investigated in this study, it can be assumed that dissolved Cr(III) would be more easily oxidised than the relatively inert chromite ore utilised as one of the case study Cr-containing materials. Although most Cr(III) compounds are precipitated out of solution at pH levels relevant to drinking water and wastewater treatment plants, some Cr(III) species are soluble (Bartlett, 1991).



The authors thank Prof Quentin Campbell and Prof Marthie Coetzee for the use of the particle size analyser and the pulveriser, respectively.



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Received 14 October 2011; accepted in revised form 11 July 2012.



* To whom all correspondence should be addressed. ffi +27 18 299 2337; fax: +27 18 299 2350; e-mail:


Comparative study of EVA-Cloisite® 20A and heat-treated EVA-Cloisite® 20A on heavy-metal adsorption properties



Derrick S Dlamini; Ajay K Mishra*; Bhekie B Mamba

Department of Applied Chemistry, University of Johannesburg, PO Box 17011, Doornfontein 2028, Johannesburg, South Africa




Ethylene vinyl acetate (EVA)/ Cloisite® 20A (C20A) composite fabricated via the melt-blending method was used for the development of a heavy-metal adsorbent through acid and heat treatment. Heat-treated composites were produced at 400°C to 1 000°C in air and N2 atmospheres. The materials were characterised through TGA, FT-IR, contact angle and Zetasizer. Treating EVA/C20A composites with H2SO, at 130°C reduced the contact angle from 99.73° to 30.40°. The acid-function-alised composite was tested for the removal of Pb2+ and an adsorption capacity of 49 mg-g-1 was recorded while the heat-treated composite exhibited an adsorption capacity of 153 mg-g-1.

Keywords: ash, EVA, bentonite, activation, adsorption




Heavy-metal pollution is known to cause instability, disorder, harm or discomfort to living organisms (Al-Attar, 2011). Several studies have shown that heavy metals such as lead, zinc, cadmium, chromium and copper can be very toxic even at low concentrations (Periasamy and Namasivayam, 1996). Industrial and agricultural activities are a major source of heavy-metal pollution worldwide. In South Africa, for instance, industrial, mining and agricultural activities are considered as the driving force of the country's economy. Due to the rapid development of agriculture, industrial and traffic activities large amounts of heavy-metal pollutants are discharged to the local environment (Bai et al., 2011; Liua et al., 2011. The aggressive development of agriculture and industrial sectors makes it difficult to combat heavy-metal pollution. It has been reported that landfill is the best available technology especially for developing countries for the disposal of solid waste; however, heavy-metal pollution by landfill leachate is still possible (Longa et al., 2011). Heavy metals may also end up in sewage effluents (Bystrzejewski et al., 2011).

Removal of heavy-metal ions from sewage effluents and other water resources is essential to ensure environmental and human safety. Several techniques like reverse osmosis, nano-filtration, ion exchange and adsorption have been used in the removal of heavy metals from water. The use of activated carbon (AC) as an adsorbent is a simple and economically viable method of pollutant removal (Vargas et al., 2011). According to Tongpoothorn et al. (2011) activated carbon is a widely-used adsorbent because of its extremely high surface areas, micro-pore volumes, large adsorption capacities, fast adsorption kinetics, and relative ease of regeneration.

In this study, a heavy-metal adsorbent was derived from EVA/C20A composites that had been synthesised via the melt-blending method. The melt-blending method has been gaining significant attention in the fabrication of polymeric composites with improved mechanical properties, relative to those produced by alternative composite fabrication strategies (solution blending and in situ polymerisation). Polymer/ clay composites have been the centre of research over the past decade because of their potentially large application area. To ensure environmental safety, the materials should be discarded safely after use. Experiments in our laboratory have shown that EVA/C20A composites synthesised via the melt-blending method are hydrophobic and non-biodegradable in garden soil and compost. Therefore, we have derived a heavy-metal adsorbent from the EVA/C20A composite through acid and heat treatments of the non-biodegradable composites. The treatment methods were compared. Both procedures were kept as simple as possible to minimise costs. Bench-scale experiments on the removal of Pb2+ from aqueous solution were undertaken to test the potential application of the novel adsorbent in heavy-metal removal from water. Equilibrium and kinetic models were extensively applied on the adsorption results in order to establish whether the adsorbent is suitable for heavy-metal removal. The results indicate that composites used for different purposes may still be reused in adsorption technology after acid or heat treatment.




Ethylene vinyl acetate (EVA) with 9% vinyl acetate (VA) was supplied by Plastamid, South Africa. The density of EVA was 0.930 g-cm-3 and the melting point was 95°C. Cloisite® 20A (C20A), a natural montmorillonite clay modified with dimethyl dihydrogenated to allow quaternary ammonium salt - CEC = 95 meq-100 g-1, was obtained from Southern Clay Products, Texas, USA. Concentrated acids (HCl and H2SO4) and KOH were sourced from Aldrich Chemicals, South Africa.

Preparation of adsorbent

The EVA/C20A was fabricated and characterised for thermal and morphological properties as discussed in our work previously published (Dlamini et al., 2011a; b). The adsorbent was derived through acid and heat treatments. To prepare the acid-treated adsorbent, a composite strip was oxidised with concentrated H2S04 at a ratio of 1:2 (weight) and heated at 130°C in an oven for 24 h. The dry curing was meant to catalyse the oxidation process. The acidified composite was allowed to cool and thereafter the free acid was removed by rinsing the composite material with deionised water until the filtrate reached a pH of between 6.5 and 7.5, after which it was then dried in an oven at 100°C for 24 h. Heat-treated composite adsorbents were derived from the ΕVA/C20A composite strip at different temperatures: 400°C, 500°C, 600°C, 700°C, 800°C, 900°C, and 1 000°C, fora period of 15 min in air andN2 atmospheres.

Characterisation of adsorbent

Thermogravimetric analysis (TGA)

Thermogravimetric analyses were performed to approximate the EVA residue after heating. The analyses were done in a Perkin Elmer TGA 4000 Analyzer equipped with PyrisTM Software. The sample mass was 8-10 mg and the temperature ranged from 100 °C to 600°C at a heating rate of 15°C-min-1. The analyses were performed under nitrogen atmosphere at a flow rate of 15 mf-min-1.

Contact angle measurements

The hydrophilic/hydrophobic nature of the composite strips, before and after acid treatment, was examined by using contact angle (CA) measurements using the sessile-drop method. The measurements were done on a Dataphysics Optical Contact Angle SCA20 at a dosing volume of 12 μ£. The values are given from an average of 6 measurements made at different locations on the specimen surface.

Zetasizer analysis

Zeta potential and particle size were investigated using a Zetasizer Nano ZS from Malvern Instruments. The dispersant was water with a pH of 5.5.

ATR-FT-IR analysis

The functional properties were measured directly by Bruker Tensor 27 FT-IR spectrometer and analysed with OPUS software. With this ATR-FT-IR model (attenuated total reflectance (ATR) Fourier transform infrared (FT-IR) model) the samples were analysed as they were. A small piece of sample was sliced from the polymeric composite material and was analysed. Prior to analysis, the samples were dried in an oven overnight.

Batch adsorption studies

The batch experiment technique was adopted for bench-scale adsorption experiments to study the heavy-metal adsorption capabilities of the composites using Pb2+ as an analyte. The experiments were carried out in 25 ml stopper reagent bottles. A stock solution of Pb2+ solution (200 mg-f-1) was prepared by dissolving 0.319 g of lead nitrate salt in deionised water. Batch adsorption experiments were conducted to establish the optimum pH, contact time and adsorbent dose, and the effect of temperature and initial concentration on Pb2+ adsorption. The adsorbent weight was 20 mg and the volume of the analyte solutions was 20 mf, unless otherwise specified. Initial pH of the solution was adjusted using 0.1M KOH or HCl. After adsorption the solutions were analysed for the remaining Pb2+ concentration using atomic adsorption spectroscopy (AAS).

Adsorption isotherms and kinetic models

Methods and supporting information used for the analysis of adsorption isotherms and kinetics are described in Appendix 1 (Langmuir, Freundlich, pseudo-first-order, pseudo-second- order, Natarajan and Khalaf, and Elovich models).


Results and discussion


The thermal degradation mechanism in char-forming polymers like EVA may be described as a generalised chemical bond scission process consisting of primary and secondary decomposition events (Bahramian et al., 2008). Generally, the degradation of EVA occurs via a 2-step mechanism with the loss of acetic acid during the first step (300°C to 400°C) and random chain scission of the remaining material in the second step, to form unsaturated vapour species (= 430°C), such as butene and ethylene (Hull et al., 2003). The thermographs of the composite and neat EVA are shown in Fig. 1.



The catalytic effect of layered silicates on crosslinking/ charring reactions derives mainly from the acid sites formed on silicates due to the degradation of the organic treatment of the clay (Kiliaris and Papaspyrides, 2010). The degradation of C20A is discussed in our previous work published on EVA/ C20A composites (Dlamini et al., 2011a; b).

The percentage weight loss in each TGA curve clearly suggests that the clay residue after thermal degradation corresponds to the 5% clay dosage. Of most interest to this study was the residue at 500°C marked with the horizontal arrows in each TGA curve. It can be seen that about 35.5% (weight) EVA residue remains at 500°C. The EVA residue is essential in the sense that it suggests that we still have a composite.

FT-IR analysis

The FT-IR results of the EVA/C20A composite and the func-tionalised EVA/C20A composite are shown in Fig. 2. The peaks located at 2 917 cm-1 and 2 849 cm-1 were attributed to v(CH) and v(CH2) groups, respectively, (Dlamini et al., 2011a) and the broad peak located at 3 351 cm-1 was assigned to the v(OH) group.



In the carbonyl region (1 600-1 to 1 800 cm-1) there is a peak at 1 704 cm-1 attributed to v(C=O). Noteworthy, the v(C=O) peak in the un-functionalised composite is shifted about 35 cm-1 higher indicating that the carbonyl bond is shorter in C=O, most likely because it belongs to the COOC component whereas in the oxidised composite it belongs to the COOH component, which may have resulted from the elimination of the acetate group in EVA. The band at 1 439 cm-1 resulted from the C-O stretching of the carboxylate anion in carboxylic groups and the band at 1 654 cm-1 was attributed to C=O stretching vibration modes in carbonyl groups (Han et al., 2011; Wang et al., 2011; Wei et al., 2011). The spectra show a broad and strong peak at 1 020 cm-1 which was confidently assigned to Si-O adsorption.

The FT-IR results of the heat-treated composites are shown in Fig. 3. The heat-treated composites were obtained by heating the acid-functionalised composites at different temperatures ranging from 400°C to 1 000°C.



The results indicate that the heating at higher temperatures resulted in the degradadation of the functional groups introduced during the heat treatment.

Contact angle

Contact angles provide useful information regarding surface hydrophilicity. The rate at which water wets the composite determines how easily the water can penetrate the composite assemblage. Mixing EVA with the clay increased the contact angle from 95.60° to 99.73° and the polar component consequently decreased from 19.18 mJ-nf2 to 10.31 mJ-nf2. This is understandable because both materials are hydrophobic. After oxidation, the contact angle of the composite substantially decreased from 99.73° to 30.40° with a polar component of 44.86 mJ-nf2. The reductions in the polar components suggest a decrease in the hydrophobic character (Bessadok et al., 2007).

Zetasizer analysis

The zeta potential results are summarised in Table 1. For the heat-treated composites, the composites derived at 500°C were used.



Visual comparison of the composite produced under the different atmospheres showed that the composite derived in air was more like ash. The argument that the heat-treated composite was composed of EVA residue was justified by heating pure EVA as a reference experiment. Zeta potential results show that all the materials are negatively charged. The nucleophilic nature of the adsorbents suggests that the composites will interact chemically with Pb2+.

Heavy-metal adsorption

Effect of heating temperature

The purpose of acid/heat treatment of the composites was to improve the surface area and provide physicochemical alterations in the structure of the clay. The effect of heating temperature (in air) on the modified composite and its adsorption capacity for Pb2+ was investigated. The uptake of Pb2+ was calculated by using the following equation:


Co and Ct (mg-f-1) are the highest initial Pb2+ concentration and remaining concentration at time (t), respectively.

The temperatures selected were 400°C, 500°C, 600°C, 700°C,

800°C, 900°C and 1 000°C. The results, given in Fig. 4, show a sharp increase from 400°C to 500°C and thereafter a substantial decrease from 500°C to 900°C in per cent Pb2+ uptake.



Koyuncu (2008) activated bentonite at 600°C. In the present work, 500°C was found to be the optimum temperature for heat treatment of carbonised EVA/C20A composites for the adsorption of Pb2+. This can be attributed to the destruction of the adsorption active sites (-COOH) as shown in FT-IR at temperatures higher than 500°C leaving only silica and alumina behind.

Effect of heating atmosphere

The atmospheres used were air and N The results are shown in Fig. 5. The effect of heating atmosphere on the adsorption properties of the adsorbents was tested on EVA (Fig. 5A), C20A (Fig. 5B), and on the EVA/C20A composite (Fig. 5C). Under N2 atmosphere, EVA formed a gel which solidified after cooling.



Apparently, there is little change in the adsorption properties of the adsorbent thermally treated in either air or N2. The EVA residue had the lowest adsorption capacity, probably because of the absence of silicates. The results are consistent with the data obtained from Zetasizer analyses. Based o these results, the composite derived under air conditions was used forthwith for all adsorptions with heat-treated composites.

Effect of pH

Optimisation of pH for sorption medium plays an important role in sorption studies (Hosseini-Bandegharaei et al., 2011). This is because the hydronium ions are strong competing adsorbate ions and partly due to the fact that the pH of a solution influences the chemical speciation of the metal ions in solution. The amount of Pb2+ adsorbed to the heat-treated (in air) composite, expressed in terms of adsorption efficiency as a function of pH, is shown in Fig. 6.



The adsorption increased with increase in the pH of the solution until pH 5.5. The low uptakes at low pH were attributed to a competitive adsorption as a result of high concentrations of hydronium ions (H3O+) which compete for adsorption sites with Pb2+. The observed adsorption increased with increasing pH and this was attributed to an increase in the concentration of Pb(OH)+ resulting in the decline in the hydronium ion concentration. For pH values >7.0, a new increase in Pb2+ uptake is observed as a result of the chemical precipitation of the metal in the form of hydroxide. Therefore, the Pb2+ adsorption phenomenon in aqueous solution at 100 mg-l-1 occurs at pH below 6.0.

Adsorption time profile: Acid vs. heat treatment

The adsorbents prepared by acid treatment were compared to the adsorbents prepared via heat treatment in terms of adsorption capacity (qt). The sorption capacity is one of the most important parameters of the sorbent characteristics in the sense that it determines how much of the heavy-metal pollutant can be removed from the aqueous solution by a unit mass of the sorbent (Bystrzejewski et al., 2011). The adsorption capacity (qt) was calculated using the following equation:


Co and Ct represent the initial concentration and the remaining concentration at different time intervals ν is the volume of the solution ws is the weight of the composite

Figure 7 depicts the adsorption capacity of untreated EVA/ C20A composite. The plots given in Fig. 8 show the adsorption kinetics of Pb2+ adsorbed onto the acid-treated (A) and heat-treated in air (B) composite. It can be seen that the adsorption of Pb2+ on the composite increased steadily until equilibrium was attained, i.e, after about 6 h for acid-treated and 4 h for heat-treated composites, respectively. The maximum adsorption capacities were 153 mg-g-1 for heat-treated and 49 mg-g-1 for acid-treated composites, respectively, from an initial concentration of 200 mg-f-1. At an initial Pb2+ concentration of 100 mg-f-1, the heat-treated (in air) composite had an adsorption capacity of 80 mg-f-1. This is a significant increase from the 6 mg-g-1 recorded with untreated EVA/C20A (Fig. 7).





The steep increase in the adsorption capacity during the early stages of the adsorption may be attributed to the abundance of adsorption sites available for binding. The increase in adsorption capacity with an increase in initial concentration can be attributed to the higher concentration of Pb2+ ions vying for adsorption sites, resulting in a higher driving mass force. This effect can be attributed to adsorbate-adsorbent ratios at low initial concentrations which imply that metal ion adsorption involves higher energy sites (Naiya et al., 2009).

Effect of adsorbent dose and temperature

The effect of adsorbent quantity and temperature on the adsorption of Pb2+ on the composites heat-treated in air was investigated. Fig. 9 shows how the adsorption capacity changed with an increase in composite dose.

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