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Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment
Waste Management xxx (2015) xxx–xxx
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Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment Jiwan Singh, Byeong-Kyu Lee ⇑ Department of Civil and Environmental Engineering, University of Ulsan, Ulsan 680-749, South Korea
a r t i c l e
i n f o
Article history: Received 11 October 2014 Accepted 26 January 2015 Available online xxxx Keywords: ASR Heavy metals pH effects Kinetic study Leaching test Mobility factor
a b s t r a c t Automobile shredder residue (ASR) is considered as hazardous waste in Japan and European countries due to presence of heavy metals. This study was carried on the extraction characteristics of heavy metals (Mn, Fe, Ni, and Cr) from automobile shredder residue (ASR). The effects of pH, temperature, particle size, and liquid/solid ratio (L/S) on the extraction of heavy metals were investigated. The recovery rate of Mn, Fe, Ni, and Cr increased with increasing extraction temperature and L/S ratio. The lowest pH 2, the highest L/S ratio, and the smallest particle size showed the highest recovery of heavy metals from ASR. The highest recovery rates were in the following order: Mn > Ni > Cr > Fe. Reduction of mobility factor for the heavy metals was observed in all the size fractions after the recovery. The results of the kinetic analysis for various experimental conditions supported that the reaction rate of the recovery process followed a second order reaction model (R2 P 0.95). The high availability of water-soluble fractions of Mn, Fe, Ni, and Cr from the low grade ASR could be potential hazards to the environment. Bioavailability and toxicity risk of heavy metals reduced significantly with pH 2 of distilled water. However, water is a cost-effective extracting agent for the recovery of heavy metals and it could be useful for reducing the toxicity of ASR. Ó 2015 Elsevier Ltd. All rights reserved.
1. Introduction The management and treatment of end-of-life vehicles (ELVs) and the environmental influence of persistent residue disposal have become important issues worldwide. Finding proper treatments for automobile shredder residue (ASR) along with valuable resource recovery is a major problem for the vehicle recycling industries. ASR is the by-product of the vehicle recycling process (Joung et al., 2007) and it contributes up to 20–21% of average ELV weight. According to the report of the European Commission Directive 2000/53/CE, a minimum of 95% (in average weight per vehicle and year) of ELVs must be reused or recovered (including energy recovery) and at least 85% must be reused or recycled by 2015. Therefore, reducing the disposal of ASR in landfills will be less than 10%. Around 87.4% and 82.9% of ASR in Spain were reuse and recovery and reuse and recycling, respectively, and 85.3% and 84.8% in Italy, respectively, in 2011 (Eurostat (2013) ‘‘End of life vehicle statistics’’) (available at http://ec.europa.eu/eurostat/statistics-explained/index.php/End-of-life_vehicle_statistics). In Japan, recycling and material recovery from ASR will be increased up to 70% by 2015 (Sakai et al., 2014). Currently, in Korea, material ⇑ Corresponding author. Tel.: +82 52 259 2864; fax: +82 52 259 2629. E-mail address: [email protected] (B.-K. Lee).
recycling and energy recovery from ASR are about 85% (of which energy recovery rate is within 5%). They are planning to increase up to 95% of ASR for recycling and energy recovery, out of which energy recovery rate is within 10% by 2015 (Sakai et al., 2014). In China, recycling possibility of ASR was about 90% in 2012, and they are try to increase recycling around 95% by 2017 (Sakai et al., 2014). In the United States, around 95% of ELVs enter for recycling route, out of which 80% recovered for the material recycling (Kumar and Sutherland, 2009). Approximately, 0.55 million and 5 million of ELVs are generated per year in Korea and Japan, respectively (Joung et al., 2007). The major parts of ELVs are recycled for reuse with a high rate of 75–80% of the total ELVs generated in Korea (Joung et al., 2007), however, the residual materials (size < 5 mm), termed low grade automobile shredder residues (ASR), are neither reused nor recycled. The ASR is heterogeneous waste material; it is difficult to develop proper technologies for its recycling (Santini et al., 2011). ASR generally categorized into coarse and fine fractions. The composition of coarse fraction is mainly foam, plastics, rubbers and textiles; however, the fine fraction consists of pieces of glass, plastic sand metals along with dust and dirt (Morselli et al., 2010). The composition, density and moisture content of ASR vary depending on the location (Boughton, 2007). The coarse fraction has the lowest ash content and the highest calorific value (15–30 kJ/kg) and highly
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Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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applicable for energy generation. However, the fine fraction (low grade ASR) has higher ash content and a lower calorific value (11–21 MJ/kg) than the coarse fraction. The fine fraction can also generate problems in incineration process, so it is not much considered for energy production (Mancini et al., 2010; Morselli et al., 2010; Vermeulen et al., 2011; Mancini et al., 2014a,b; Cossu et al., 2014). According to the Morselli et al. (2010), about 20–100 mm fractions of ASR are the richest in polymers that would have the highest calorific values among the other fractions of ASR. Fines occupy about 50% of the total ASR, and they are predominantly polluted with hazardous materials, such as heavy metals, mineral oils, and hydrocarbons (Santini et al., 2011). The major sources of heavy metals in ASR are residual metal pieces, solder, plasticizers, and paint (Kurose et al., 2006; Lopes et al., 2009). In the thermal process of ASR, concentration of the heavy metals increased by a factor of up to 20, in the generated ash residues (Mancini et al., 2014a,b). As shown in Fig. 1, the presence of heavy metals in the low grade ASR can pollute ground water and surface water through leaching, so ASR has been classified as a hazardous waste. It cannot be disposed of in uncontrolled landfills without proper treatment. A small fraction of heavy metals present in the water-soluble fraction may readily leach out, and can be bioavailable in the environment (Iwegbue et al., 2007; Liu et al., 2008; Singh and Kalamdhad, 2013a). Thus, study is needed on cost-effective extraction agent(s) for recovery and reduction of toxicity of heavy metals from ASR for its safe disposal in landfill sites. At lower pH, leaching of heavy metals may occur during the landfilling of ASR. The fast recovery rate can be controlled by the rate of diffusion of the ions from the surface of the solid through the boundary layer, while a slow recovery rate can be controlled by chemical reaction (Aydogan et al., 2005). According to Tessier et al. (1979), the chemical speciation of heavy metals in a solid environmental matrix is distributed into five fractions (exchangeable, carbonate, reducible, bound to organic matter, and residual fractions). The exchangeable fraction (F1) is likely to be affected by changes in water ionic composition
as well as sorption/desorption processes. This form can be released by ion-exchange processes. The carbonate or acid-extractable fraction (F2) is susceptible to changes in pH and this can become soluble and be mobilized at lower pH. The F1 and F2 fractions of the total metals are considered to be the most mobile forms. Thus, the mobility, bioavailability, and eco-toxicity of heavy metals depend on their speciation rather than their total content (Singh and Kalamdhad, 2012, 2013a, 2013b; Zhu et al., 2014). The term ‘‘bioavailability’’ of any element may be considered as that part of the total concentration of the element that is readily soluble in water. The bioavailable fractions of the metals may be potential contaminants of the food chain, surface water, and groundwater (Singh and Kalamdhad, 2013b). Few studies are available on the recovery of heavy metals from ASR (Kurose et al., 2006; Gonazalez-Fernandez et al., 2008; Granata et al., 2011), but the recovery and kinetic study of heavy metals (Mn, Fe, Ni, Cr) from the ASR has not been reported before. Thus, the objectives of the present study were to assess the recovery of heavy metals from ASR using water, to examine the governing reaction kinetics during the recovery, and to reduce the bioavailability and toxicity of ASR.
2. Materials and methods 2.1. Physicochemical characterization of ASR The ASR sample (20 kg) was collected from an automobile shredder plant, Nam-gu, Ulsan, Korea. The collected sample was air dried at room temperature with 25 ± 1 °C for two weeks after collection. The ASR sample was sieved using different sizes of sieves and five different grain-size fractions were found, classified as A: 4.75–2.0 mm, B: 2.0–1.19 mm, C: 1.19–0.425 mm, D: 0.425– 0.250 mm, and E: 0.250–0.0 mm. The moisture content of different size fractions was determined after drying in a hot air oven (C-DF forced convectional drying oven, Chang Shin Scientific Co.) at 105 °C for 24 h (Kalamdhad et al., 2009). A digital pH meter (ORION
Fig. 1. Schematic views of ASR processing and heavy metals recovery.
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
J. Singh, B.-K. Lee / Waste Management xxx (2015) xxx–xxx
5 STAR, thermo scientific) was used for measuring the pH of the extraction solution (1:10, w/v sample: distilled water) after extraction by shaking for 2 h at 150 rpm (Kalamdhad et al., 2009). An electronic muffle furnace (C-FMA, Chang Shin Science Co.) was used for the determination of volatile solids and ash content by ignition of the sample at 550 °C (Sprynskyy et al., 2007). An atomic absorption spectrometer (Varian Spectra 55) was used for analysis of Mn, Fe, Ni, and Cr concentrations after the digestion of 1.0 g of ASR sample (dried) with 20 mL of H2SO4 and HClO4 (5:1) mixture in a closed Teflon vessel using a hot plate. During the final stage of digestion, H2O2 was added drop-wise for further oxidation and digestion of ASR materials (Singh and Kalamdhad, 2012, 2013a). All samples were digested triplicates and average values were presented in Table 1. 2.2. Heavy metals recovery study The extraction experiment was designed on the basis of previous studies (Wang et al., 2001; Quina et al., 2009). Extraction of heavy metals was carried out in a series of 250 mL covered conical flasks using a temperature-controlled water bath shaker (HST205SW) at 150 rpm. The extraction of heavy metals from ASR was done at different pHs (2, 4, 6, 8, 10, and 12), L/S ratios (10, 25, 50, and 100 mL/g), temperatures (25, 35, 45, and 55 °C) and particle sizes (4.75–2.0, 2.0–1.19, 1.19–0.425, 0.425–0.250, and 0.25–0.0 mm). For extraction, a batch study was conducted for 120 min and samples were withdrawn at 5, 15, 30, 60, 90, and 120 min for analysis of the heavy metals extracted. The slurry was centrifuged and filtered through Whatman No. 42 filter paper (0.45 lm) and filtrates were used for heavy metal analysis. The pH of deionized water was maintained with 1.0 M HCl and 1.0 M NaOH. Extraction experiments at given conditions were repeated three times and average values were calculated for analysis. Heavy metal recovery was calculated according to Hong et al. (2000): heavy metal recovery (%) = (metals contents in extracted liquid)/ (total contents of metals found after acid digestion). The residues remaining after extractions under different conditions were selected for X-ray diffraction (XRD) and scanning electron microscopy (SEM; Hitachi S-4700) to analyze morphology and compositional change before and after the extraction. The elements such as carbon (C), oxygen (O), nitrogen (N), sulfur (S), sodium (Na), potassium (K), calcium (Ca), and magnesium (Mg) were analyzed by Energy-dispersive X-ray spectroscopy (EDX) coupled with SEM. The XRD patterns were recorded on a Bruker AXN in the 2h range of 10–80° using Cu Ka radiation (k = 1.5418 Å), an accelerating voltage of 40 kV at an applied current of 30 mA to check the crystallinity of the residues. The total organic carbon analyzer (TOC-5000A, Shimadzu) was used for the analysis of dissolved organic carbon (DOC) of extracted liquid samples at different time intervals for different grain fractions. 2.3. Toxicity characteristic leaching procedure (TCLP) test and speciation The residues remained after heavy metal extractions (dried at 105 °C for 24 h) were subjected to the ‘toxicity characteristic Table 1 Total heavy metal contents in different grain fraction of ASR (mean ± SD, n = 3). Grain fraction (mm)
4.75–2.0 (A) 2.0–1.19 (B) 1.19–0.425 (C) 0.425–0.250 (D) <0.250 (E)
Heavy metals (mg/kg dry matter)
3
leaching procedure’ (TCLP) test (US EPA, 1992). According to the TCLP method, 5 g of ASR sample (size less than 9.5 mm) and 100 mL of acetic acid at pH 4.93 ± 0.05 (pH adjusted with 1 N NaOH, solid sample: solution ratio = 1:20) was mixed in a 250mL covered conical flask with stirring at 25 °C for 18 h in a shaker at 30 ± 2 rpm. The suspensions of TCLP test were filtered using 0.45 lm filter paper and stored in a polycarbonate tube for analysis of Mn, Fe, Ni, and Cr. The sequential extraction of sample (residue remained after heavy metal extraction) was carried out according to the procedure proposed by Tessier et al. (1979). The sequential extraction was conducted in polycarbonate centrifuge tubes of 50 mL capacity for the extraction of movable fractions. The following sequential steps were included (a) Exchangeable fraction (F1): a 1 g sample was extracted at 25 °C with 8 mL of 1 M MgCl2 (pH 7) with continuous agitation for 1 h, (b) Carbonate fraction (F2): residue obtained from step (a) above was leached at 25 °C with 8 mL of 1 M of NaOAc (pH 5, adjusted with conc. HOAc) with continuous agitation for 5 h. After each extraction, the supernatant liquid was obtained with a pipette after centrifugation (5000 rpm, 5 min). About 20 mL of deionized water was used for washing the each residue remaining after successive extractions. All extractions were conducted in duplicate and mean values are reported with their standard deviations. An exact quantification of the F1 and F2 fractions for each metal in ASR is necessary for consideration of landfilling disposal of ASR. For all selected metals, the sum of F1 and F2 fractions was used for mobility factor (MF) calculation. The MF was calculated according to Zhu et al. (2014): Mobility factor = [(F1 + F2) fractions/(total contents of metals found after acid digestion)] 100. The MF is generally used to determine element mobility in the environment. Determination coefficients of correlation (R2) were obtained between heavy metal recoveries and different pHs, L/S ratios and temperatures. 2.4. Risk assessment code (RAC) A risk assessment code (RAC) was applied to estimate the environmental risk associated with heavy metals pollution for ASR. This method has been used widely by several researchers for assessing heavy metals pollution in sediments/soils (Liu et al., 2008; Sundaray et al., 2011). RAC measures the bioavailability of heavy metals by applying a scale to the percentage of metals present in the sum of F1 and F2 fractions, since these fractions are weakly bonded, certainly affected by ionic strength, and susceptible to pH changes in soil environment (Yuan et al., 2011). The high mobility of F1 and F2 fractions could equilibrate with the aqueous phase and consequently become more rapidly bioavailable (Sundaray et al., 2011). The heavy metals in the ASR can be classified by using RAC as no risk (RAC < 1), low risk (RAC = 1–10), medium risk (RAC = 11–30), high risk (RAC = 31–50) and very high risk (RAC > 50) (Sundaray et al., 2011). The calculation for RAC is given in Eq. (1).
RAC ð%Þ ¼
F1 þ F2 100 Total concentration of heavy metals
ð1Þ
3. Results and discussion 3.1. Physicochemical properties
Mn
Fe
Ni
Cr
95 ± 10 710 ± 90 700 ± 102 773.5 ± 23 1418 ± 10
5346 ± 604 9270 ± 177 4389 ± 451 8306 ± 410 35091 ± 943
106 ± 9 219 ± 6 118 ± 21 105 ± 23 266 ± 24
15.5 ± 0.3 42.3 ± 2.1 43.2 ± 5.1 38.7 ± 1.3 168.0 ± 5.0
Fig. 2 shows the results from the physicochemical characterizations of the different fractions of the ASR. The moisture content of different size fractions of ASR was found about 1.3–3.2%. The moisture content was measured maximum in fraction E followed by fraction D. Neutral pHs (6.99–7.07) were assessed in all the fractions of
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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ASR. However, the all fractions were mixed with solution of pH 2 and then the pHs of extracted solution (at L/S 50 mL/g and temperature 25 °C, time 2 h) were found 2.1, 2.4, 3.1, 4.4 and 6.4 for fractions A, B, C, D, and E, respectively. The pHs were increased with reduction in ASR grain size fractions. The highest pH was found in the extracted solution of fraction E; this might be due to dissolution of finer particles of ASR in acidic solution. The ash content was found about 80–85% in the fractions B, C and E due to the presence of metals, soil, dust and glass. The similar results were also reported by Kim et al. (2004). The ASR fractions P 4.75 mm have the high content of volatile solid and low concentration of heavy metals. Thus, these size fractions are highly recommended to be utilized for energy production. The high ignition loss was found in fraction D (66.4%) due to the presence of combustible organic materials; which have the lowest weight fraction as compared to fractions A, B, C and E. Fraction D had a high content of fluffy materials (fine textile, paper, sponge, etc.) so that much higher volatile solid was found in the fraction than the other size fractions. The highest volatile content in fraction D resulted in the lowest ash content. Results of metal characterization showed clearly that Fe was the most abundant metal, followed by Mn in all size fractions, consistent with the findings of Granata et al. (2011). The analysis results of total metal content showed that all selected heavy metals were mainly contained in the smallest grain fraction as compared to other larger fractions. 3.2. Instrumental analysis before and after extraction at pH 2 X-ray diffraction is a powerful tool to analyze the crystalline nature of materials. Extraction may lead to change in molecular and crystalline structure of the ASR. Thus, an understanding of the molecular structure and crystalline structure of the ASR and the resulting changes would provide valuable information regarding the recovery of heavy metals. Fig. 3 presents the XRD pattern of fractions D and E of the ASR before and after recovery of heavy metals. The XRD patterns display four peaks at 2h = 21.09°, 27.55°, 29.59° and 42.64° in fraction D before the extraction. The peaks at 2h = 21.09° and 27.55° are likely due to SiO2 and the peak at 29.59° can be assigned to CuO, appearing before and after recovery. The intensities of these two peaks were reduced after recovery.
The peak at 2h = 42.64°was not seen in fractions D and E of ASR after extraction. This peak could correspond to CaO, SiO2, Al2O3, and Fe3O4 (Shibayama et al., 2006). This confirms that these mineral oxides can be extracted by the acidic solution. In fraction E, before recovery, the XRD patterns showed six peaks at 2h = 9.07°, 20.88°, 29.38°, 54.87° and 59.84°.The peaks at 2h = 9.07°, 29.38° and 54.87°may correspond to Al2O3, ZnS/NaCl, CaO and Fe3O4/ CuO, respectively, which did not appear after recovery. The remaining peaks in fraction E after recovery also appeared with very low intensities. The intensities of some peaks in fractions D and E were reduced after the recovery. This may have been due to the suppression of material crystallinity after the recovery of heavy metals (Shibayama et al., 2006). Thus, the results of the present study clearly confirmed that the recovery of heavy metals with distilled water at pH 2 changed the structure of the ASR. The intensity decreases of the diffraction peaks indicated lowering of mesopore uniformity (Anbia and Salehi, 2012). Furthermore, inorganic compounds were completely or partially extracted from the ASR. Fig. 4a and b shows that the solid particles of ASR presented a random surface and many rough holes, respectively, that might be due to attack of acidic water under the optimum temperature (Li et al., 2013). The residue that remained after being treated with acidic water showed clean and clearly scratched particle surfaces. Furthermore, pores were also developed and a surface area increased, compared with an ASR sample before recovery (Souza et al., 2007; Fedje et al., 2010). Table 2 shows the elemental composition of ASR before and after recovery of heavy metals. The C content was increased; however, the O, Na, K, Ca, Mg, S and N contents were declined after metals recovery from the ASR. This might be due to the increased carbonization of ASR and leaching of other elements with hydrochloric acid. Zhang and Itoh (2006) also reported reduction of metallic elements, such as Na, K, Ca and Mg, from the municipal solid waste incinerator fly ash with distilled water. 3.3. Effects of pH Fig. 5 shows the recovery of metals at various pHs while other parameters were kept constant (i.e., temperature 25 °C, L/S ratio 50, and particle size fraction 6 0.25 mm). The pH is one of the most
Fig. 2. Distribution of different particle size fractions of ASR (A: 4.75–2.0 mm, B: 2.0-1.19 mm, C: 1.19–0.425 mm, D: 0.425–0.250 mm, and E: 0.250–0.0 mm).
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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Fig. 3. XRD spectra of ASR (fraction D and E) before and after extraction of heavy metals (pH 2, temperature 25 °C, liquid-to-solid ratio 50 mL/g, stirring speed 150 rpm).
Fig. 4. SEM micrograph of ASR (pH 2, temperature 25 °C, liquid-to-solid ratio 50 mL/g, stirring speed 150 rpm); (a) before extraction-fraction E and (b) residue after extraction of fraction E.
Table 2 Elemental composition of ASR before and after treatment (mean ± SD, n = 3). Elements
ASR (fraction E, %)
ASR (fraction E after recovery, %)
C N O Na Mg S K Ca Total
15.82 ± 3.32 3.78 ± 0.14 78.50 ± 10.9 0.45 ± 0.077 0.32 ± 0.09 0.15 ± 0.03 0.29 ± 0.15 0.68 ± 0.20 100
44.30 ± 12.02 3.24 ± 1.6 51.31 ± 16.23 0.32 ± 0.08 0.27 ± 0.05 0.08 ± 0.013 0.23 ± 0.07 0.24 ± 0.12 100
important factors for heavy metal recovery from ASR. The highest recovery of heavy metals (Mn 9.0%, Fe 0.8%, Ni 6.9%, and Cr 1.3%) was achieved at the lowest pH (pH 2; Fig. 5). In the acidic water, more heavy metals were released versus alkaline pH. Zhang et al. (2008) published similar results in bottom ash generated from the incineration of municipal solid waste. Mn and Fe were recovered at all selected pHs; however, Cr was extracted only at pH 2 and 4. Recovery of Ni was not obtained at pH 8, 10, or 12. The highest recovery of Fe was obtained within 5 min of shaking at pH 2; after that, the recovery declined. At other pHs, the recovery trends were similar to those of Mn, Ni, and Cr. The recovery of Mn, Ni, and Cr increased with increased shaking time and became almost constant after 60 min (Fig. 5). The recovery percentages of all selected heavy metals showed an increasing tendency with decrease in pH level from 12 to 2 due to an ion exchange process in which metal ions are replaced by hydrogen ions (H+) (Chiang et al., 2008; Huang et al., 2011).
The observed recovery results of Cr as a function of pH are shown in Fig. 5. The recovery of Cr at pH 2 was higher than those at other pHs (pH 4–12). The toxicity of Cr depends on its different oxidation states. Cr (VI) compounds are predominant due to the low solubility of Cr (III) compounds in water. Astrup et al. (2006) reported that the predominant chemical species of Cr in solution are Cr (VI) or CrO2 4 at pH above 7. Cr (III) is almost insoluble at pH > 5 (Quina et al., 2009). Thus, ASR may have a potential negative impact on the environment. Table 3 illustrates the maximum recovery of heavy metals from the ASR at different pHs. Final pHs of metal-extracted solutions at pH 4, 6, 8, and 10 were found between 6 and 7; however, with a pH 12 solution, it was observed between 8 and 9 (Fig. 6a). This clearly indicates that pH in the alkaline range is not appropriate for the recovery of heavy metals. 3.4. Effects of liquid/solid (L/S) ratio The effect of L/S ratios on recovery of heavy metals was evaluated at pH 2, extraction temperature 25 °C, and particle size 6 0.25 mm. Recovery of Mn, Fe, Ni, and Cr increased from 2.3%, 0.1%, 1.4%, and 0.7% to 28.8%, 3.1%, 11.1%, and 5.8%, respectively, by increasing from L/S 10–100. This trend is consistent with results of other studies (Zhang and Nicol, 2010; Gharabaghi et al., 2011). Increasing the L/ S ratios greatly increased the recovery of Mn, Fe, Ni, and Cr (Fig. 7). The extraction of these heavy metals is controlled by metal solubility. At higher L/S ratios, the amount of protons available in the extracting medium is sufficient to react with the metal compounds present in ASR material (Sakultung et al., 2007). Table 3 illustrates maximum recovery of heavy metals from the ASR at different extraction conditions. The recovery order of different heavy metals at the highest L/S ratio of 100 was as follows: Mn
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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Fig. 5. Effects of pH on recovery of Mn, Fe, Ni and Cr (50 mL/g liquid-to-solid ratio, stirring rate 150 rpm, grain size fraction 0.25–0.0 mm and temperature 25 °C).
Table 3 Maximum recoveries of heavy metals at variation of pH, temperature, L/S ratio and grain fractions. Variation of parameters
Maximum recovery of heavy metals (%) Mn
Fe
Ni
Cr
pH (L/S-50 mL/g, temperature-25 °C, agitation speed-150 rpm and grain size fraction < 0.25 mm)
2 4 6 8 10 12
8.96 1.17 1.29 1.16 1.36 0.16
0.81 0.08 0.08 0.08 0.15 0.27
6.85 0.66 0.24 0.0 0.0 0.0
1.31 0.13 0.0 0.0 0.0 0.0
Temperature (°C) (L/S-50 mL/g, pH-2, agitation speed-150 rpm and grain size fraction < 0.25 mm)
25 35 45 55
8.91 11.15 11.16 13.28
0.81 0.84 0.92 0.96
6.85 5.52 4.98 4.75
1.31 1.35 1.35 1.46
L/S ratio (mL/g) (pH-2, temperature-25 °C, agitation speed-150 rpm and grain size fraction < 0.25 mm)
10 25 50 100
2.29 5.47 8.96 28.79
0.007 0.22 0.81 3.08
1.51 3.76 6.85 11.13
0.66 1.27 1.31 5.75
Grain fractions (pH-2, L/S-50 mL/g, temperature-25 °C, agitation speed-150 rpm
A B C D E
1.99 4.98 7.90 8.50 8.96
0.52 1.30 1.19 1.66 0.81
2.98 3.34 5.24 5.89 6.85
0.75 1.12 1.17 1.28 1.31
Note – A = ASR size range 4.75–2.0 mm, B = size range 2.0–1.19 mm, C = size range 1.19–0.425 mm, D = size range 0.425–0.25 mm, E = size < 0.25 mm
(28.8%) > Ni (11.1%) > Cr (5.7%) > Fe (3.1%). Recovery of Cr and Fe declined after 5 min. This could be due to the formation of complexes with dissolved organic carbon present in ASR (Singh and Kalamdhad, 2013c). Table 4 shows the reduction of dissolved organic carbon after metals extraction. A stronger pH effect was observed in the highest L/S ratios than the lower one. The final pH of extracting solution of L/S ratio 100 was between 2 and 4; however, in other L/S ratios (10, 25, and 50 mL/g), the final pH observed was near neutral (Fig. 6b). 3.5. Effects of temperature The extraction was carried out in the temperature range of 25–55 °C. Fig. 8 illustrates the effect of reaction temperature on the recovery rate of heavy metals at pH 2, particle size
fraction < 0.250 mm, and L/S ratio 50 mL/g. Temperature is a key parameter that can greatly affect the recovery rate of heavy metals by changing activity levels of components in the ASR sample. Fig. 8 illustrates that the recovery of Mn, Fe, and Cr increased from 9.0%, 0.8%, and 1.3% at 25 °C to 11.8%, 1.0%, and 1.5% at 55 °C, respectively; it confirms that the reactions are endothermic. It could be observed that temperature had a great effect on the Mn, Fe, Ni, and Cr extraction. Sakultung et al. (2007) reported that increasing leaching temperature up to 80 °C led to the rise of the leaching percentages of Ni; however, in the present study recovery of Ni was reduced slightly from 6.8% at 25 °C to 5.4% at 55 °C (Fig. 8). The maximum recovery of heavy metals at different temperatures is illustrated in Table 3. As shown in Fig. 6c, the final pH of extracted solution in the temperature study found to be about 7.
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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Fig. 6. Final pH of solution after extraction at different time intervals with variation in parameters: (a) pH, (b) L/S ratio, (c) temperature and (d) grain fractions.
Fig. 7. Effects of liquid to solid ratio on recovery of Mn, Fe, Ni, and Cr (pH 2, stirring rate 150 rpm, grain fraction < 0.25 mm and temperature 25 °C).
3.6. Effects of particle size The effects of different particle sizes of ASR on recovery of heavy metals were evaluated at pH 2, temperature 25 °C, and L/S 50 mL/g. Fig. 9 shows that the heavy metal recovery increased with increasing extraction time and decreasing particle size. As the particle size was reduced from A to E, the recovery of Mn, Ni, and Cr increased from 2.0% to 9.0%, 3.0 to 6.9%, and 0.7 to 1.3%, respectively. These results were consistent with the finding by Aydogan et al. (2005) from sphalerite in acidic conditions. Inner diffusion was the main
controlling step for the heavy metals extraction process. As surface area of solid particles contacting the fluid increased due to the decreasing particle size, the reactions of metal extraction became faster (Abdel-Aal, 2000). In the smallest size fraction E, Cr recovery increased with increase in extraction time up to 40 min. Recovery of Cr was much higher than in fractions B, C, and D, which was not greatly changed among different grain sizes. Recovery of Fe in fraction E was higher in the initial 5 min, but it then decreased with increased extraction time (Fig. 9). This might be attributable to a Fe complex formed with dissolved organics
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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Table 4 Changes of dissolved organic carbon (DOC) during recovery of Mn, Fe, Ni and Cr in different grain fractions of ASR (50 mL/g liquid-to-solid ratio, pH 2, stirring rate 150 rpm and temperature 25 °C). Contact time in (min)
5 15 30 60 90 120
Concentration of dissolved organic carbon (mg/L) A
B
C
D
E
38.8 ± 0.5 31.6 ± 1.5 31.9 ± 1.5 45.7 ± 0.5 43.2 ± 0.5 41.8 ± 1.0
54.2 ± 0.5 40.7 ± 0.5 42.1 ± 1.5 27.3 ± 1.4 23.5 ± 0.5 23.1 ± 0.5
28.2 ± 0.5 27.4 ± 0.5 27.0 ± 2.0 23.7 ± 0.5 23.8 ± 1.0 22.0 ± 1.0
74.1 ± 1.0 62.2 ± 1.0 59.9 ± 1.0 55.3 ± 0.5 30.5 ± 1.7 30.7 ± 0.2
313.0 ± 5.0 285.8 ± 3.5 136.4 ± 1.5 124.3 ± 2.0 121.7 ± 5.0 121.6 ± 0.0
Note – A = ASR size range 4.75–2.0 mm, B = size range 2.0–1.19 mm, C = size range 1.19–0.425 mm, D = size range 0.425–0.25 mm, E = size < 0.25 mm
Fig. 8. Effects of temperature on recovery of Mn, Fe, Ni, and Cr (50 mL/g liquid-to-solid ratio, pH 2, stirring rate 150 rpm and grain fraction < 0.25 mm).
Fig. 9. Effects of grain fractions on recovery of Mn, Fe, Ni, and Cr (50 mL/g liquid-to-solid ratio, pH 2, stirring rate 150 rpm and temperature 25 °C).
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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J. Singh, B.-K. Lee / Waste Management xxx (2015) xxx–xxx
Fig. 10. Correlation between heavy metal recovery and different parameters (pH, L/S ratio, and temperature).
(DOC) during the extraction process (Singh and Kalamdhad, 2013c). Marschner and Kalbitz (2003) also reported that Fe can form relatively stable complexes with dissolved organic matter in acid soils condition. Furthermore, the final pH of extraction solution was 7. This is because finer particles of ASR were mixed with distilled water of pH 2 (Fig. 6d). The order of recovery of selected heavy metals in different grain fractions was as follows: Mn (9.0%) > Ni (6.9%) > Cr (1.3%). Heavy metals can form metal hydroxides and re-adsorb on the solid ASR residue at higher pHs and thus the recovery of heavy metals decreased (Garau et al., 2007; Singh and Kalamdhad, 2013d) with increasing pH. Dissolved organic matter contains a range of organic substances from defined small molecules to highly polymeric humic substances (Michalzik et al., 2001). Higher concentrations of DOC were observed in the smallest size fraction E of ASR. The DOC was decreased with extraction time for grain fractions B, C, D, and E, but it increased with time in fraction A (Table 4). This can be explained in that in fraction A, the concentration of heavy metals was low compared to with other fractions (B, C, D, and E), and resulting metals did not formed complexes with DOC.
rate of heavy metals strongly depended on L/S ratio and extraction temperature. 3.8. Mechanism for heavy metals extraction The extraction of adsorbed heavy metals from ASR by inorganic acids is based on the exchange of adsorbed metals for hydronium ions, followed by their hydration and solubilization, as described in Eqs. (2) and (3):
ASR Mnþ þ nH3 Oþ ! ASR ðH3 Oþ Þn þ Mnþ
ð2Þ
Mnþ þ mH2 O ! ½MðH2 OÞm nþ
ð3Þ
In the ASR, Ni might be present as NiOOH(s), released in acidic pH water from the ASR, as shown in Eqs. (4) and (5) (Sakultung et al., 2007). The active nickel material is used as nickel oxyhydroxide (NiOOH), which is converted into Ni(OH)2 during battery discharge and reformed into NiOOH during recharging (Coman et al., 2013). 2þ
3.7. Correlation between metal recovery and extraction conditions Recovery of heavy metal did not show great dependency on variable pH values (R2 < 0.60), however, very good correlations were observed between the L/S ratio and heavy metals recovery from ASR (R2 = 0.959 for Mn, R2 = 0.963 for Fe, R2 = 0.985 for Ni and R2 = 0.891 for Cr; Fig. 10). In the correlation analysis between extraction temperature and metal recovery, Fe (0.957), Mn (0.867), Ni (0.880), and Cr (0.830) showed a high value of determination coefficient (R2; Fig. 9). Based on these results, the recovery
NiOOHðsÞ þ 2Hþ ðaqÞ ! Ni ðaqÞ þ 3=2H2 þ O2 2þ
NiðOHÞ2 ðsÞ þ 2Hþ ðaqÞ ! Ni ðaqÞ þ 2H2 O
ð4Þ ð5Þ
Anionic metal contaminants such as hexavalent chromium(HCrO 4 and CrO2 4 ), are adsorbed on positively charged functional groups and, therefore, cannot be removed from ASR by exchange with hydronium ions. Heavy metals retained in ASR can be extracted by inorganic acids as a result of solid phase dissolution and as a result of metal hydration as shown in Eq. (2).
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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J. Singh, B.-K. Lee / Waste Management xxx (2015) xxx–xxx
Fig. 11. TCLP test for heavy metals before and after recovery of heavy metals from the ASR (A = ASR size range 4.75–2.0 mm, B = size range 2.0–1.19 mm, C = size range 1.19– 0.425 mm, D = size range 0.425–0.25 mm, E = size < 0.25 mm and R = ASR after metal recovery).
Fig. 12. Changes of exchangeable and carbonate fractions and mobility factor of heavy metals before and after recovery of heavy metals: (a) Mn, (b) Fe, (c) Ni, (d) Cr, and (e) mobility factor, MF (A = ASR size range 4.75–2.0 mm, B = size range 2.0–1.19 mm, C = size range 1.19–0.425 mm, D = size range 0.425–0.25 mm, E = size < 0.25 mm and R = ASR after metal recovery).
3.9. Toxicity characteristics leaching procedure (TCLP) test Fig. 11 shows that the extraction concentrations of all selected heavy metals (Mn, Fe, Ni, and Cr) were greatly reduced in all grain
fractions (A, B, C, D, and E) of the ASR after metal recovery with distilled water at pH 2. In particular, reduction in metal extraction was mainly observed in grain sizes B and E. The large reduction in size E indicates that heavy metals included in the fine size of ASR are more
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
J. Singh, B.-K. Lee / Waste Management xxx (2015) xxx–xxx Table 5 Determination coefficients (R2) of various kinetic models for recovery of heavy metals. Heavy metals
Extraction temperature (°C)
Determination coefficients (R2) of various kinetic models 1(1a)1/3
1–2a/3 (1a)2/3
1–3(1a)2/3 + 2(1a)
Mn
25 35 45 55
0.8128 0.8516 0.4091 0.4767
0.8244 0.8766 0.4136 0.4554
0.8192 0.8663 0.4117 0.4639
Fe
25 35 45 55
0.7208 0.7863 0.4931 0.5805
0.7765 0.8430 0.5836 0.6755
0.6412 0.6352 0.3735 0.4779
Ni
25 35 45 55
0.3041 0.5091 0.8728 0.8520
0.3047 0.5187 0.8650 0.8447
0.3044 0.5136 0.8495 0.8495
Cr
25 35 45 55
0.3820 0.7323 0.7032 0.6805
0.3783 0.9295 0.9084 0.9288
0.4578 0.1118 0.6254 0.2519
readily extracted, as discussed in the section on particle size effects. If the leachable concentration of heavy metals in the waste is above the regulatory level, the waste is considered ‘hazardous’ (Singh and Kalamdhad, 2013a). The smallest particle size fraction showed more leachability potential than the larger size fractions of the ASR. Chiang et al. (2008) also reported the smallest size fractions of ash generated from incineration of municipal solid waste as having more leachable potential of heavy metals. The concentration of Cr in any fraction of ASR was not beyond the threshold limits specified by the US EPA. The results of the present study suggest that distilled water (pH 2) could be effective for the recovery and reduction of toxicity by heavy metals from ASR. 3.10. Speciation and mobility of heavy metals In a risk assessment, the mobility and bioavailability of heavy metals are affected mainly by changes in pH, different forms of
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metals, and ionic concentrations of leachate during landfilling of ASR. The mobilization of metal depends strongly on their mobility, concentration in the leachate, and solubility in water (Singh and Kalamdhad, 2013e). Fig. 12 plots the quantities of Mn, Fe, Ni, and Cr in exchangeable (F1) and carbonate (F2) fractions in sequential extractions of the different size fractions of ASR and the residues after the recovery of heavy metals under the conditions of pH 2, L/S 50, extraction time 60 min after the recovery. F1 fractions of Mn, Fe, Ni, and Cr were much smaller than F2 fractions for all size fractions (except size fraction A for Cr). F1 fractions of Mn and Fe seemed to increase with decreasing size of ASR; however, F1 fractions of Ni and Cr did not change significantly even with grain size fraction changes. The highest F2 fraction of Mn, Ni, and Cr were observed in size fractions E, B, and C, respectively. The F1 fraction of Cr, Ni, Fe and Mn for all the size fractions of ASR seemed to decrease after the recovery. The F2 fraction of these metals was also reduced after the recovery in the all size fractions (A, B, C, D and E). Fig. 12 shows the mobility factor (MF) of Mn, Fe, Ni, and Cr in different fractions of ASR before and after recovery of heavy metals. The MF was reduced in the range of 50.2–81.3% for Mn, 14.0–65.6% for Fe, 58.0–89.8% for Ni, and 63–73.3% for Cr in the final ASR residue. The total concentration of Fe was found much higher than the Mn and Ni, about 25 and 132 times respectively. However, F1 and F2 fractions of Fe contributed 2.9% of the total Fe, which is less than Mn (25.2%) and Ni (10.9%). The total concentration of Cr was around 209 times less than Fe concentration, but F1 and F2 fractions of Fe contributed similar as Cr. The present study also suggests that toxicity of heavy metals depends on different mobile fractions (F1 and F2) rather than total heavy metals concentration (solid sample digested with strong acids). 3.11. Kinetic of heavy metal recovery A kinetic study of the recovery of heavy metals from the ASR was performed at different temperatures of 25, 35, 45, and 55 °C. For a liquid/solid reaction system, the reaction rate is generally controlled by diffusion through the liquid film, diffusion through the solid or product layer, and the chemical reaction at the surface of the solid particles or a mixed of diffusion and chemical reactions
Fig. 13. Second-order reaction kinetics of Mn, Fe, Ni and Cr at different temperatures.
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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Table 6 Risk assessment code (RAC) for heavy metals before and after recovery of heavy metals from ASR. Heavy metals
Mn Fe Ni Cr
RAC (%) A
AR
B
BR
C
CR
D
DR
E
ER
29.9 2.5 16.4 15.9
5.6 1.6 2.8 5.6
29.0 11.9 13.4 8.1
12.9 4.1 1.6 3.2
32.6 23.8 20.5 13.1
16.2 12.6 2.4 3.5
30.3 12.2 26.2 11.0
14.5 9.1 2.7 4.1
25.2 2.9 10.9 2.9
9.6 2.5 4.5 1.0
Note – A = ASR size range 4.75–2.0 mm, B = size range 2.0–1.19 mm, C = size range 1.19–0.425 mm, D = size range 0.425–0.25 mm, E = size < 0.25 mm, R = recovery of heavy metals
(Gharabaghi et al., 2013). It is important to establish a quantitative measurement of the rate of recovery/extraction kinetics and mechanism. If the ASR particles (size less than 250 mm) were considered to be spherical particles in a liquid–solid reaction system, the kinetics of metal recovery could be described using the shrinking core model (SCM) (Souza et al., 2007; Tian et al., 2010; Li et al., 2013). As shown in Table 5, the SCM model did not, in fact, fit the experimental data very well (R2 < 0.95). Li et al. (2014) also reported similar results from the kinetics of vanadium leaching by sulfuric acid from a spent industrial V2O5/TiO2 catalyst. Initially, the metal concentration increased very rapidly and then slowly reached a saturation value in the bulk liquid solution. This trend suggests that the kinetics model of the recovery process might be represented by a second-order rate law expression (Sakultung et al., 2007). The linear form of a second order kinetic model is expressed according to Ho et al. (2005). Plotting the value of t/Ct versus t will provide a straight line with the slope of 1/Cs and the intercept of 1/KC2: s
t=C t ¼ ðt=KC 2s Þ þ ðt=C s Þ
ð6Þ
where Ct is the concentration of metal ion (mg/L) at given time t (min), Cs is the amount of metal ion released at equilibrium (mg/ L), and K is the second-order rate constant for desorption (L/ mg min). The Cs and K were calculated as Cs = 1/slope and K = slope2/intercept. Fig. 13 shows the series plots of t/Ct as a function of extraction time (5–120 min) at various temperatures (25, 35, 45, and 55 °C) of the Mn, Fe, Ni, and Cr recovery processes using distilled water at pH 2. A coefficient of determination (R2 P 0.95) between the second-order model and the experimental results was found to show very good agreement. This confirms that there were two mechanisms in the recovery process: first, strong dissolution and scrubbing and then a much slower stage occurred related to external diffusion (Sakultung et al., 2007). 3.12. Risk assessment code (RAC) As shown in Table 6, RAC for Mn, Fe, Ni and Cr was reduced after recovery of heavy metals from ASR. The risk of Mn was changed from medium to low level in fraction A and E. However, in fraction B, C and D risk reduced but remained in medium category. Fe changed from medium risk to low risk category in A, B, D and E size fractions but in fraction C it was remained in medium risk category. The toxicity risk of Ni and Cr reduced from medium risk to low risk level in all grain size fractions. It can be attributed as at lower pH (pH 2) of water could be useful for reduction of toxicity risk of heavy metals. 4. Conclusions The important parameters that affect in recovery of heavy metals (Mn, Fe, Ni, and Cr) from the automobile shredder residue (ASR) can be ranked as: pH > L/S ratio > temperature > grain fraction > extracting time. The presence of bioavailable fractions
(water soluble, exchangeable, and carbonate) of heavy metals in the automobile shredder residue (ASR) confirms that ASR is potentially hazardous to the environment without proper disposal. The highest recoveries of heavy metals were found in the lowest size fraction of ASR at pH 2, L/S ratio of 100, and temperature of 25 °C, and thus the suggested conditions can be recommended for working in a real scale plant. The recovery of all selected heavy metals was increased with increase in the L/S ratio from 10 to 100 mL/g. The higher pH (>7) of distilled water was not effective for heavy metals recovery from the ASR. The smallest size fraction is considered the most hazardous but also the most favorable for the highest recovery of heavy metals, compared with the larger fractions. The ASR fractions P 4.75 mm have the high content of volatile solid and low concentration of heavy metals and thus these fractions can be recommended to be utilized for energy production. A very good correlation was observed between the L/S ratio and heavy metals recovery from ASR. Leachable concentration of Cr in all size fractions of ASR was under the threshold limits for leaching of solid wastes. The mobility factors and toxicity risk of the heavy metals were reduced greatly in the final residues of all fractions left after metal recovery. The second-order kinetic model fitted the data very well (R2 > 0.95) for the recovery of all selected heavy metals. Thus, distilled water at pH 2 was found to be a highly effective extracting agent for the recovery and reduction of toxicity of heavy metals in ASR. Acknowledgements This work was supported by the National Research Foundation of Korea (NRF) grant funded by the Ministry of Science, ICT and Future Planning (2013R1A2A2A03013138). References Abdel-Aal, E.A., 2000. Kinetics of sulphuric acid leaching of low grade zinc silicate ore. Hydrometallurgy 55 (3), 247–254. Anbia, M., Salehi, S., 2012. Removal of acid dyes from aqueous media by adsorption onto amino-functionalized nanoporous silica SBA-3. Dyes Pigments 94, 1–9. Astrup, T., Mosb, K.H., Christensen, T.H., 2006. Assessment of long-term leaching from waste incineration air-pollution-control residues. Waste Manage. 26, 803– 814. Aydogan, S., Aras, A., Canbazoglu, M., 2005. Dissolution kinetics of sphalerite in acidic ferric chloride leaching. Chem. Eng. J. 114, 67–72. Boughton, B., 2007. Evaluation of shredder residue as cement manufacturing feedstock. Resour. Conserv. Recycl. 51, 621–642. Chiang, K.Y., Jih, J.C., Chien, M.D., 2008. The acid extraction of metals from municipal solid waste incinerator products. Hydrometallurgy 93, 16–22. Coman, V., Robotin, B., Ilea, P., 2013. Nickel recovery/removal from industrial wastes: a review. Resour. Conserv. Recycl. 73, 229–238. Cossu, R., Fiore, S., Lai, T., Luciano, A., Mancini, G., Ruffino, B., Viotti, P., Zanetti, M.C., 2014. Review of Italian experience on automotive shredder residue characterization and management. Waste Manage. 34 (10), 1752–1762. EU Directive 2000/53/EC of the European parliament and the Council of 18 September 2000 on end-of-life vehicles. Off. J. Eur. Union 2000, L269, pp. 34-42. Fedje, K.K., Ekberg, C., Skarnemark, G., Steenari, Britt-Marie, 2010. Removal of hazardous metals from MSW fly ash-an evaluation of ash leaching methods. J. Hazard. Mater. 173, 310–317. Garau, G., Castaldi, P., Santona, L., Deiana, P., Melis, P., 2007. Influence of red mud, zeolite and lime on heavy metal immobilization, culturable heterotrophic microbial populations and enzyme activities in a contaminated soil. Geoderma 142, 47–57.
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Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment Jiwan Singh, Byeong-Kyu Lee ⇑ Department of Civil and Environmental Engineering, University of Ulsan, Ulsan 680-749, South Korea
a r t i c l e
i n f o
Article history: Received 11 October 2014 Accepted 26 January 2015 Available online xxxx Keywords: ASR Heavy metals pH effects Kinetic study Leaching test Mobility factor
a b s t r a c t Automobile shredder residue (ASR) is considered as hazardous waste in Japan and European countries due to presence of heavy metals. This study was carried on the extraction characteristics of heavy metals (Mn, Fe, Ni, and Cr) from automobile shredder residue (ASR). The effects of pH, temperature, particle size, and liquid/solid ratio (L/S) on the extraction of heavy metals were investigated. The recovery rate of Mn, Fe, Ni, and Cr increased with increasing extraction temperature and L/S ratio. The lowest pH 2, the highest L/S ratio, and the smallest particle size showed the highest recovery of heavy metals from ASR. The highest recovery rates were in the following order: Mn > Ni > Cr > Fe. Reduction of mobility factor for the heavy metals was observed in all the size fractions after the recovery. The results of the kinetic analysis for various experimental conditions supported that the reaction rate of the recovery process followed a second order reaction model (R2 P 0.95). The high availability of water-soluble fractions of Mn, Fe, Ni, and Cr from the low grade ASR could be potential hazards to the environment. Bioavailability and toxicity risk of heavy metals reduced significantly with pH 2 of distilled water. However, water is a cost-effective extracting agent for the recovery of heavy metals and it could be useful for reducing the toxicity of ASR. Ó 2015 Elsevier Ltd. All rights reserved.
1. Introduction The management and treatment of end-of-life vehicles (ELVs) and the environmental influence of persistent residue disposal have become important issues worldwide. Finding proper treatments for automobile shredder residue (ASR) along with valuable resource recovery is a major problem for the vehicle recycling industries. ASR is the by-product of the vehicle recycling process (Joung et al., 2007) and it contributes up to 20–21% of average ELV weight. According to the report of the European Commission Directive 2000/53/CE, a minimum of 95% (in average weight per vehicle and year) of ELVs must be reused or recovered (including energy recovery) and at least 85% must be reused or recycled by 2015. Therefore, reducing the disposal of ASR in landfills will be less than 10%. Around 87.4% and 82.9% of ASR in Spain were reuse and recovery and reuse and recycling, respectively, and 85.3% and 84.8% in Italy, respectively, in 2011 (Eurostat (2013) ‘‘End of life vehicle statistics’’) (available at http://ec.europa.eu/eurostat/statistics-explained/index.php/End-of-life_vehicle_statistics). In Japan, recycling and material recovery from ASR will be increased up to 70% by 2015 (Sakai et al., 2014). Currently, in Korea, material ⇑ Corresponding author. Tel.: +82 52 259 2864; fax: +82 52 259 2629. E-mail address: [email protected] (B.-K. Lee).
recycling and energy recovery from ASR are about 85% (of which energy recovery rate is within 5%). They are planning to increase up to 95% of ASR for recycling and energy recovery, out of which energy recovery rate is within 10% by 2015 (Sakai et al., 2014). In China, recycling possibility of ASR was about 90% in 2012, and they are try to increase recycling around 95% by 2017 (Sakai et al., 2014). In the United States, around 95% of ELVs enter for recycling route, out of which 80% recovered for the material recycling (Kumar and Sutherland, 2009). Approximately, 0.55 million and 5 million of ELVs are generated per year in Korea and Japan, respectively (Joung et al., 2007). The major parts of ELVs are recycled for reuse with a high rate of 75–80% of the total ELVs generated in Korea (Joung et al., 2007), however, the residual materials (size < 5 mm), termed low grade automobile shredder residues (ASR), are neither reused nor recycled. The ASR is heterogeneous waste material; it is difficult to develop proper technologies for its recycling (Santini et al., 2011). ASR generally categorized into coarse and fine fractions. The composition of coarse fraction is mainly foam, plastics, rubbers and textiles; however, the fine fraction consists of pieces of glass, plastic sand metals along with dust and dirt (Morselli et al., 2010). The composition, density and moisture content of ASR vary depending on the location (Boughton, 2007). The coarse fraction has the lowest ash content and the highest calorific value (15–30 kJ/kg) and highly
http://dx.doi.org/10.1016/j.wasman.2015.01.035 0956-053X/Ó 2015 Elsevier Ltd. All rights reserved.
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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applicable for energy generation. However, the fine fraction (low grade ASR) has higher ash content and a lower calorific value (11–21 MJ/kg) than the coarse fraction. The fine fraction can also generate problems in incineration process, so it is not much considered for energy production (Mancini et al., 2010; Morselli et al., 2010; Vermeulen et al., 2011; Mancini et al., 2014a,b; Cossu et al., 2014). According to the Morselli et al. (2010), about 20–100 mm fractions of ASR are the richest in polymers that would have the highest calorific values among the other fractions of ASR. Fines occupy about 50% of the total ASR, and they are predominantly polluted with hazardous materials, such as heavy metals, mineral oils, and hydrocarbons (Santini et al., 2011). The major sources of heavy metals in ASR are residual metal pieces, solder, plasticizers, and paint (Kurose et al., 2006; Lopes et al., 2009). In the thermal process of ASR, concentration of the heavy metals increased by a factor of up to 20, in the generated ash residues (Mancini et al., 2014a,b). As shown in Fig. 1, the presence of heavy metals in the low grade ASR can pollute ground water and surface water through leaching, so ASR has been classified as a hazardous waste. It cannot be disposed of in uncontrolled landfills without proper treatment. A small fraction of heavy metals present in the water-soluble fraction may readily leach out, and can be bioavailable in the environment (Iwegbue et al., 2007; Liu et al., 2008; Singh and Kalamdhad, 2013a). Thus, study is needed on cost-effective extraction agent(s) for recovery and reduction of toxicity of heavy metals from ASR for its safe disposal in landfill sites. At lower pH, leaching of heavy metals may occur during the landfilling of ASR. The fast recovery rate can be controlled by the rate of diffusion of the ions from the surface of the solid through the boundary layer, while a slow recovery rate can be controlled by chemical reaction (Aydogan et al., 2005). According to Tessier et al. (1979), the chemical speciation of heavy metals in a solid environmental matrix is distributed into five fractions (exchangeable, carbonate, reducible, bound to organic matter, and residual fractions). The exchangeable fraction (F1) is likely to be affected by changes in water ionic composition
as well as sorption/desorption processes. This form can be released by ion-exchange processes. The carbonate or acid-extractable fraction (F2) is susceptible to changes in pH and this can become soluble and be mobilized at lower pH. The F1 and F2 fractions of the total metals are considered to be the most mobile forms. Thus, the mobility, bioavailability, and eco-toxicity of heavy metals depend on their speciation rather than their total content (Singh and Kalamdhad, 2012, 2013a, 2013b; Zhu et al., 2014). The term ‘‘bioavailability’’ of any element may be considered as that part of the total concentration of the element that is readily soluble in water. The bioavailable fractions of the metals may be potential contaminants of the food chain, surface water, and groundwater (Singh and Kalamdhad, 2013b). Few studies are available on the recovery of heavy metals from ASR (Kurose et al., 2006; Gonazalez-Fernandez et al., 2008; Granata et al., 2011), but the recovery and kinetic study of heavy metals (Mn, Fe, Ni, Cr) from the ASR has not been reported before. Thus, the objectives of the present study were to assess the recovery of heavy metals from ASR using water, to examine the governing reaction kinetics during the recovery, and to reduce the bioavailability and toxicity of ASR.
2. Materials and methods 2.1. Physicochemical characterization of ASR The ASR sample (20 kg) was collected from an automobile shredder plant, Nam-gu, Ulsan, Korea. The collected sample was air dried at room temperature with 25 ± 1 °C for two weeks after collection. The ASR sample was sieved using different sizes of sieves and five different grain-size fractions were found, classified as A: 4.75–2.0 mm, B: 2.0–1.19 mm, C: 1.19–0.425 mm, D: 0.425– 0.250 mm, and E: 0.250–0.0 mm. The moisture content of different size fractions was determined after drying in a hot air oven (C-DF forced convectional drying oven, Chang Shin Scientific Co.) at 105 °C for 24 h (Kalamdhad et al., 2009). A digital pH meter (ORION
Fig. 1. Schematic views of ASR processing and heavy metals recovery.
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
J. Singh, B.-K. Lee / Waste Management xxx (2015) xxx–xxx
5 STAR, thermo scientific) was used for measuring the pH of the extraction solution (1:10, w/v sample: distilled water) after extraction by shaking for 2 h at 150 rpm (Kalamdhad et al., 2009). An electronic muffle furnace (C-FMA, Chang Shin Science Co.) was used for the determination of volatile solids and ash content by ignition of the sample at 550 °C (Sprynskyy et al., 2007). An atomic absorption spectrometer (Varian Spectra 55) was used for analysis of Mn, Fe, Ni, and Cr concentrations after the digestion of 1.0 g of ASR sample (dried) with 20 mL of H2SO4 and HClO4 (5:1) mixture in a closed Teflon vessel using a hot plate. During the final stage of digestion, H2O2 was added drop-wise for further oxidation and digestion of ASR materials (Singh and Kalamdhad, 2012, 2013a). All samples were digested triplicates and average values were presented in Table 1. 2.2. Heavy metals recovery study The extraction experiment was designed on the basis of previous studies (Wang et al., 2001; Quina et al., 2009). Extraction of heavy metals was carried out in a series of 250 mL covered conical flasks using a temperature-controlled water bath shaker (HST205SW) at 150 rpm. The extraction of heavy metals from ASR was done at different pHs (2, 4, 6, 8, 10, and 12), L/S ratios (10, 25, 50, and 100 mL/g), temperatures (25, 35, 45, and 55 °C) and particle sizes (4.75–2.0, 2.0–1.19, 1.19–0.425, 0.425–0.250, and 0.25–0.0 mm). For extraction, a batch study was conducted for 120 min and samples were withdrawn at 5, 15, 30, 60, 90, and 120 min for analysis of the heavy metals extracted. The slurry was centrifuged and filtered through Whatman No. 42 filter paper (0.45 lm) and filtrates were used for heavy metal analysis. The pH of deionized water was maintained with 1.0 M HCl and 1.0 M NaOH. Extraction experiments at given conditions were repeated three times and average values were calculated for analysis. Heavy metal recovery was calculated according to Hong et al. (2000): heavy metal recovery (%) = (metals contents in extracted liquid)/ (total contents of metals found after acid digestion). The residues remaining after extractions under different conditions were selected for X-ray diffraction (XRD) and scanning electron microscopy (SEM; Hitachi S-4700) to analyze morphology and compositional change before and after the extraction. The elements such as carbon (C), oxygen (O), nitrogen (N), sulfur (S), sodium (Na), potassium (K), calcium (Ca), and magnesium (Mg) were analyzed by Energy-dispersive X-ray spectroscopy (EDX) coupled with SEM. The XRD patterns were recorded on a Bruker AXN in the 2h range of 10–80° using Cu Ka radiation (k = 1.5418 Å), an accelerating voltage of 40 kV at an applied current of 30 mA to check the crystallinity of the residues. The total organic carbon analyzer (TOC-5000A, Shimadzu) was used for the analysis of dissolved organic carbon (DOC) of extracted liquid samples at different time intervals for different grain fractions. 2.3. Toxicity characteristic leaching procedure (TCLP) test and speciation The residues remained after heavy metal extractions (dried at 105 °C for 24 h) were subjected to the ‘toxicity characteristic Table 1 Total heavy metal contents in different grain fraction of ASR (mean ± SD, n = 3). Grain fraction (mm)
4.75–2.0 (A) 2.0–1.19 (B) 1.19–0.425 (C) 0.425–0.250 (D) <0.250 (E)
Heavy metals (mg/kg dry matter)
3
leaching procedure’ (TCLP) test (US EPA, 1992). According to the TCLP method, 5 g of ASR sample (size less than 9.5 mm) and 100 mL of acetic acid at pH 4.93 ± 0.05 (pH adjusted with 1 N NaOH, solid sample: solution ratio = 1:20) was mixed in a 250mL covered conical flask with stirring at 25 °C for 18 h in a shaker at 30 ± 2 rpm. The suspensions of TCLP test were filtered using 0.45 lm filter paper and stored in a polycarbonate tube for analysis of Mn, Fe, Ni, and Cr. The sequential extraction of sample (residue remained after heavy metal extraction) was carried out according to the procedure proposed by Tessier et al. (1979). The sequential extraction was conducted in polycarbonate centrifuge tubes of 50 mL capacity for the extraction of movable fractions. The following sequential steps were included (a) Exchangeable fraction (F1): a 1 g sample was extracted at 25 °C with 8 mL of 1 M MgCl2 (pH 7) with continuous agitation for 1 h, (b) Carbonate fraction (F2): residue obtained from step (a) above was leached at 25 °C with 8 mL of 1 M of NaOAc (pH 5, adjusted with conc. HOAc) with continuous agitation for 5 h. After each extraction, the supernatant liquid was obtained with a pipette after centrifugation (5000 rpm, 5 min). About 20 mL of deionized water was used for washing the each residue remaining after successive extractions. All extractions were conducted in duplicate and mean values are reported with their standard deviations. An exact quantification of the F1 and F2 fractions for each metal in ASR is necessary for consideration of landfilling disposal of ASR. For all selected metals, the sum of F1 and F2 fractions was used for mobility factor (MF) calculation. The MF was calculated according to Zhu et al. (2014): Mobility factor = [(F1 + F2) fractions/(total contents of metals found after acid digestion)] 100. The MF is generally used to determine element mobility in the environment. Determination coefficients of correlation (R2) were obtained between heavy metal recoveries and different pHs, L/S ratios and temperatures. 2.4. Risk assessment code (RAC) A risk assessment code (RAC) was applied to estimate the environmental risk associated with heavy metals pollution for ASR. This method has been used widely by several researchers for assessing heavy metals pollution in sediments/soils (Liu et al., 2008; Sundaray et al., 2011). RAC measures the bioavailability of heavy metals by applying a scale to the percentage of metals present in the sum of F1 and F2 fractions, since these fractions are weakly bonded, certainly affected by ionic strength, and susceptible to pH changes in soil environment (Yuan et al., 2011). The high mobility of F1 and F2 fractions could equilibrate with the aqueous phase and consequently become more rapidly bioavailable (Sundaray et al., 2011). The heavy metals in the ASR can be classified by using RAC as no risk (RAC < 1), low risk (RAC = 1–10), medium risk (RAC = 11–30), high risk (RAC = 31–50) and very high risk (RAC > 50) (Sundaray et al., 2011). The calculation for RAC is given in Eq. (1).
RAC ð%Þ ¼
F1 þ F2 100 Total concentration of heavy metals
ð1Þ
3. Results and discussion 3.1. Physicochemical properties
Mn
Fe
Ni
Cr
95 ± 10 710 ± 90 700 ± 102 773.5 ± 23 1418 ± 10
5346 ± 604 9270 ± 177 4389 ± 451 8306 ± 410 35091 ± 943
106 ± 9 219 ± 6 118 ± 21 105 ± 23 266 ± 24
15.5 ± 0.3 42.3 ± 2.1 43.2 ± 5.1 38.7 ± 1.3 168.0 ± 5.0
Fig. 2 shows the results from the physicochemical characterizations of the different fractions of the ASR. The moisture content of different size fractions of ASR was found about 1.3–3.2%. The moisture content was measured maximum in fraction E followed by fraction D. Neutral pHs (6.99–7.07) were assessed in all the fractions of
Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035
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ASR. However, the all fractions were mixed with solution of pH 2 and then the pHs of extracted solution (at L/S 50 mL/g and temperature 25 °C, time 2 h) were found 2.1, 2.4, 3.1, 4.4 and 6.4 for fractions A, B, C, D, and E, respectively. The pHs were increased with reduction in ASR grain size fractions. The highest pH was found in the extracted solution of fraction E; this might be due to dissolution of finer particles of ASR in acidic solution. The ash content was found about 80–85% in the fractions B, C and E due to the presence of metals, soil, dust and glass. The similar results were also reported by Kim et al. (2004). The ASR fractions P 4.75 mm have the high content of volatile solid and low concentration of heavy metals. Thus, these size fractions are highly recommended to be utilized for energy production. The high ignition loss was found in fraction D (66.4%) due to the presence of combustible organic materials; which have the lowest weight fraction as compared to fractions A, B, C and E. Fraction D had a high content of fluffy materials (fine textile, paper, sponge, etc.) so that much higher volatile solid was found in the fraction than the other size fractions. The highest volatile content in fraction D resulted in the lowest ash content. Results of metal characterization showed clearly that Fe was the most abundant metal, followed by Mn in all size fractions, consistent with the findings of Granata et al. (2011). The analysis results of total metal content showed that all selected heavy metals were mainly contained in the smallest grain fraction as compared to other larger fractions. 3.2. Instrumental analysis before and after extraction at pH 2 X-ray diffraction is a powerful tool to analyze the crystalline nature of materials. Extraction may lead to change in molecular and crystalline structure of the ASR. Thus, an understanding of the molecular structure and crystalline structure of the ASR and the resulting changes would provide valuable information regarding the recovery of heavy metals. Fig. 3 presents the XRD pattern of fractions D and E of the ASR before and after recovery of heavy metals. The XRD patterns display four peaks at 2h = 21.09°, 27.55°, 29.59° and 42.64° in fraction D before the extraction. The peaks at 2h = 21.09° and 27.55° are likely due to SiO2 and the peak at 29.59° can be assigned to CuO, appearing before and after recovery. The intensities of these two peaks were reduced after recovery.
The peak at 2h = 42.64°was not seen in fractions D and E of ASR after extraction. This peak could correspond to CaO, SiO2, Al2O3, and Fe3O4 (Shibayama et al., 2006). This confirms that these mineral oxides can be extracted by the acidic solution. In fraction E, before recovery, the XRD patterns showed six peaks at 2h = 9.07°, 20.88°, 29.38°, 54.87° and 59.84°.The peaks at 2h = 9.07°, 29.38° and 54.87°may correspond to Al2O3, ZnS/NaCl, CaO and Fe3O4/ CuO, respectively, which did not appear after recovery. The remaining peaks in fraction E after recovery also appeared with very low intensities. The intensities of some peaks in fractions D and E were reduced after the recovery. This may have been due to the suppression of material crystallinity after the recovery of heavy metals (Shibayama et al., 2006). Thus, the results of the present study clearly confirmed that the recovery of heavy metals with distilled water at pH 2 changed the structure of the ASR. The intensity decreases of the diffraction peaks indicated lowering of mesopore uniformity (Anbia and Salehi, 2012). Furthermore, inorganic compounds were completely or partially extracted from the ASR. Fig. 4a and b shows that the solid particles of ASR presented a random surface and many rough holes, respectively, that might be due to attack of acidic water under the optimum temperature (Li et al., 2013). The residue that remained after being treated with acidic water showed clean and clearly scratched particle surfaces. Furthermore, pores were also developed and a surface area increased, compared with an ASR sample before recovery (Souza et al., 2007; Fedje et al., 2010). Table 2 shows the elemental composition of ASR before and after recovery of heavy metals. The C content was increased; however, the O, Na, K, Ca, Mg, S and N contents were declined after metals recovery from the ASR. This might be due to the increased carbonization of ASR and leaching of other elements with hydrochloric acid. Zhang and Itoh (2006) also reported reduction of metallic elements, such as Na, K, Ca and Mg, from the municipal solid waste incinerator fly ash with distilled water. 3.3. Effects of pH Fig. 5 shows the recovery of metals at various pHs while other parameters were kept constant (i.e., temperature 25 °C, L/S ratio 50, and particle size fraction 6 0.25 mm). The pH is one of the most
Fig. 2. Distribution of different particle size fractions of ASR (A: 4.75–2.0 mm, B: 2.0-1.19 mm, C: 1.19–0.425 mm, D: 0.425–0.250 mm, and E: 0.250–0.0 mm).
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Fig. 3. XRD spectra of ASR (fraction D and E) before and after extraction of heavy metals (pH 2, temperature 25 °C, liquid-to-solid ratio 50 mL/g, stirring speed 150 rpm).
Fig. 4. SEM micrograph of ASR (pH 2, temperature 25 °C, liquid-to-solid ratio 50 mL/g, stirring speed 150 rpm); (a) before extraction-fraction E and (b) residue after extraction of fraction E.
Table 2 Elemental composition of ASR before and after treatment (mean ± SD, n = 3). Elements
ASR (fraction E, %)
ASR (fraction E after recovery, %)
C N O Na Mg S K Ca Total
15.82 ± 3.32 3.78 ± 0.14 78.50 ± 10.9 0.45 ± 0.077 0.32 ± 0.09 0.15 ± 0.03 0.29 ± 0.15 0.68 ± 0.20 100
44.30 ± 12.02 3.24 ± 1.6 51.31 ± 16.23 0.32 ± 0.08 0.27 ± 0.05 0.08 ± 0.013 0.23 ± 0.07 0.24 ± 0.12 100
important factors for heavy metal recovery from ASR. The highest recovery of heavy metals (Mn 9.0%, Fe 0.8%, Ni 6.9%, and Cr 1.3%) was achieved at the lowest pH (pH 2; Fig. 5). In the acidic water, more heavy metals were released versus alkaline pH. Zhang et al. (2008) published similar results in bottom ash generated from the incineration of municipal solid waste. Mn and Fe were recovered at all selected pHs; however, Cr was extracted only at pH 2 and 4. Recovery of Ni was not obtained at pH 8, 10, or 12. The highest recovery of Fe was obtained within 5 min of shaking at pH 2; after that, the recovery declined. At other pHs, the recovery trends were similar to those of Mn, Ni, and Cr. The recovery of Mn, Ni, and Cr increased with increased shaking time and became almost constant after 60 min (Fig. 5). The recovery percentages of all selected heavy metals showed an increasing tendency with decrease in pH level from 12 to 2 due to an ion exchange process in which metal ions are replaced by hydrogen ions (H+) (Chiang et al., 2008; Huang et al., 2011).
The observed recovery results of Cr as a function of pH are shown in Fig. 5. The recovery of Cr at pH 2 was higher than those at other pHs (pH 4–12). The toxicity of Cr depends on its different oxidation states. Cr (VI) compounds are predominant due to the low solubility of Cr (III) compounds in water. Astrup et al. (2006) reported that the predominant chemical species of Cr in solution are Cr (VI) or CrO2 4 at pH above 7. Cr (III) is almost insoluble at pH > 5 (Quina et al., 2009). Thus, ASR may have a potential negative impact on the environment. Table 3 illustrates the maximum recovery of heavy metals from the ASR at different pHs. Final pHs of metal-extracted solutions at pH 4, 6, 8, and 10 were found between 6 and 7; however, with a pH 12 solution, it was observed between 8 and 9 (Fig. 6a). This clearly indicates that pH in the alkaline range is not appropriate for the recovery of heavy metals. 3.4. Effects of liquid/solid (L/S) ratio The effect of L/S ratios on recovery of heavy metals was evaluated at pH 2, extraction temperature 25 °C, and particle size 6 0.25 mm. Recovery of Mn, Fe, Ni, and Cr increased from 2.3%, 0.1%, 1.4%, and 0.7% to 28.8%, 3.1%, 11.1%, and 5.8%, respectively, by increasing from L/S 10–100. This trend is consistent with results of other studies (Zhang and Nicol, 2010; Gharabaghi et al., 2011). Increasing the L/ S ratios greatly increased the recovery of Mn, Fe, Ni, and Cr (Fig. 7). The extraction of these heavy metals is controlled by metal solubility. At higher L/S ratios, the amount of protons available in the extracting medium is sufficient to react with the metal compounds present in ASR material (Sakultung et al., 2007). Table 3 illustrates maximum recovery of heavy metals from the ASR at different extraction conditions. The recovery order of different heavy metals at the highest L/S ratio of 100 was as follows: Mn
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Fig. 5. Effects of pH on recovery of Mn, Fe, Ni and Cr (50 mL/g liquid-to-solid ratio, stirring rate 150 rpm, grain size fraction 0.25–0.0 mm and temperature 25 °C).
Table 3 Maximum recoveries of heavy metals at variation of pH, temperature, L/S ratio and grain fractions. Variation of parameters
Maximum recovery of heavy metals (%) Mn
Fe
Ni
Cr
pH (L/S-50 mL/g, temperature-25 °C, agitation speed-150 rpm and grain size fraction < 0.25 mm)
2 4 6 8 10 12
8.96 1.17 1.29 1.16 1.36 0.16
0.81 0.08 0.08 0.08 0.15 0.27
6.85 0.66 0.24 0.0 0.0 0.0
1.31 0.13 0.0 0.0 0.0 0.0
Temperature (°C) (L/S-50 mL/g, pH-2, agitation speed-150 rpm and grain size fraction < 0.25 mm)
25 35 45 55
8.91 11.15 11.16 13.28
0.81 0.84 0.92 0.96
6.85 5.52 4.98 4.75
1.31 1.35 1.35 1.46
L/S ratio (mL/g) (pH-2, temperature-25 °C, agitation speed-150 rpm and grain size fraction < 0.25 mm)
10 25 50 100
2.29 5.47 8.96 28.79
0.007 0.22 0.81 3.08
1.51 3.76 6.85 11.13
0.66 1.27 1.31 5.75
Grain fractions (pH-2, L/S-50 mL/g, temperature-25 °C, agitation speed-150 rpm
A B C D E
1.99 4.98 7.90 8.50 8.96
0.52 1.30 1.19 1.66 0.81
2.98 3.34 5.24 5.89 6.85
0.75 1.12 1.17 1.28 1.31
Note – A = ASR size range 4.75–2.0 mm, B = size range 2.0–1.19 mm, C = size range 1.19–0.425 mm, D = size range 0.425–0.25 mm, E = size < 0.25 mm
(28.8%) > Ni (11.1%) > Cr (5.7%) > Fe (3.1%). Recovery of Cr and Fe declined after 5 min. This could be due to the formation of complexes with dissolved organic carbon present in ASR (Singh and Kalamdhad, 2013c). Table 4 shows the reduction of dissolved organic carbon after metals extraction. A stronger pH effect was observed in the highest L/S ratios than the lower one. The final pH of extracting solution of L/S ratio 100 was between 2 and 4; however, in other L/S ratios (10, 25, and 50 mL/g), the final pH observed was near neutral (Fig. 6b). 3.5. Effects of temperature The extraction was carried out in the temperature range of 25–55 °C. Fig. 8 illustrates the effect of reaction temperature on the recovery rate of heavy metals at pH 2, particle size
fraction < 0.250 mm, and L/S ratio 50 mL/g. Temperature is a key parameter that can greatly affect the recovery rate of heavy metals by changing activity levels of components in the ASR sample. Fig. 8 illustrates that the recovery of Mn, Fe, and Cr increased from 9.0%, 0.8%, and 1.3% at 25 °C to 11.8%, 1.0%, and 1.5% at 55 °C, respectively; it confirms that the reactions are endothermic. It could be observed that temperature had a great effect on the Mn, Fe, Ni, and Cr extraction. Sakultung et al. (2007) reported that increasing leaching temperature up to 80 °C led to the rise of the leaching percentages of Ni; however, in the present study recovery of Ni was reduced slightly from 6.8% at 25 °C to 5.4% at 55 °C (Fig. 8). The maximum recovery of heavy metals at different temperatures is illustrated in Table 3. As shown in Fig. 6c, the final pH of extracted solution in the temperature study found to be about 7.
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Fig. 6. Final pH of solution after extraction at different time intervals with variation in parameters: (a) pH, (b) L/S ratio, (c) temperature and (d) grain fractions.
Fig. 7. Effects of liquid to solid ratio on recovery of Mn, Fe, Ni, and Cr (pH 2, stirring rate 150 rpm, grain fraction < 0.25 mm and temperature 25 °C).
3.6. Effects of particle size The effects of different particle sizes of ASR on recovery of heavy metals were evaluated at pH 2, temperature 25 °C, and L/S 50 mL/g. Fig. 9 shows that the heavy metal recovery increased with increasing extraction time and decreasing particle size. As the particle size was reduced from A to E, the recovery of Mn, Ni, and Cr increased from 2.0% to 9.0%, 3.0 to 6.9%, and 0.7 to 1.3%, respectively. These results were consistent with the finding by Aydogan et al. (2005) from sphalerite in acidic conditions. Inner diffusion was the main
controlling step for the heavy metals extraction process. As surface area of solid particles contacting the fluid increased due to the decreasing particle size, the reactions of metal extraction became faster (Abdel-Aal, 2000). In the smallest size fraction E, Cr recovery increased with increase in extraction time up to 40 min. Recovery of Cr was much higher than in fractions B, C, and D, which was not greatly changed among different grain sizes. Recovery of Fe in fraction E was higher in the initial 5 min, but it then decreased with increased extraction time (Fig. 9). This might be attributable to a Fe complex formed with dissolved organics
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Table 4 Changes of dissolved organic carbon (DOC) during recovery of Mn, Fe, Ni and Cr in different grain fractions of ASR (50 mL/g liquid-to-solid ratio, pH 2, stirring rate 150 rpm and temperature 25 °C). Contact time in (min)
5 15 30 60 90 120
Concentration of dissolved organic carbon (mg/L) A
B
C
D
E
38.8 ± 0.5 31.6 ± 1.5 31.9 ± 1.5 45.7 ± 0.5 43.2 ± 0.5 41.8 ± 1.0
54.2 ± 0.5 40.7 ± 0.5 42.1 ± 1.5 27.3 ± 1.4 23.5 ± 0.5 23.1 ± 0.5
28.2 ± 0.5 27.4 ± 0.5 27.0 ± 2.0 23.7 ± 0.5 23.8 ± 1.0 22.0 ± 1.0
74.1 ± 1.0 62.2 ± 1.0 59.9 ± 1.0 55.3 ± 0.5 30.5 ± 1.7 30.7 ± 0.2
313.0 ± 5.0 285.8 ± 3.5 136.4 ± 1.5 124.3 ± 2.0 121.7 ± 5.0 121.6 ± 0.0
Note – A = ASR size range 4.75–2.0 mm, B = size range 2.0–1.19 mm, C = size range 1.19–0.425 mm, D = size range 0.425–0.25 mm, E = size < 0.25 mm
Fig. 8. Effects of temperature on recovery of Mn, Fe, Ni, and Cr (50 mL/g liquid-to-solid ratio, pH 2, stirring rate 150 rpm and grain fraction < 0.25 mm).
Fig. 9. Effects of grain fractions on recovery of Mn, Fe, Ni, and Cr (50 mL/g liquid-to-solid ratio, pH 2, stirring rate 150 rpm and temperature 25 °C).
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Fig. 10. Correlation between heavy metal recovery and different parameters (pH, L/S ratio, and temperature).
(DOC) during the extraction process (Singh and Kalamdhad, 2013c). Marschner and Kalbitz (2003) also reported that Fe can form relatively stable complexes with dissolved organic matter in acid soils condition. Furthermore, the final pH of extraction solution was 7. This is because finer particles of ASR were mixed with distilled water of pH 2 (Fig. 6d). The order of recovery of selected heavy metals in different grain fractions was as follows: Mn (9.0%) > Ni (6.9%) > Cr (1.3%). Heavy metals can form metal hydroxides and re-adsorb on the solid ASR residue at higher pHs and thus the recovery of heavy metals decreased (Garau et al., 2007; Singh and Kalamdhad, 2013d) with increasing pH. Dissolved organic matter contains a range of organic substances from defined small molecules to highly polymeric humic substances (Michalzik et al., 2001). Higher concentrations of DOC were observed in the smallest size fraction E of ASR. The DOC was decreased with extraction time for grain fractions B, C, D, and E, but it increased with time in fraction A (Table 4). This can be explained in that in fraction A, the concentration of heavy metals was low compared to with other fractions (B, C, D, and E), and resulting metals did not formed complexes with DOC.
rate of heavy metals strongly depended on L/S ratio and extraction temperature. 3.8. Mechanism for heavy metals extraction The extraction of adsorbed heavy metals from ASR by inorganic acids is based on the exchange of adsorbed metals for hydronium ions, followed by their hydration and solubilization, as described in Eqs. (2) and (3):
ASR Mnþ þ nH3 Oþ ! ASR ðH3 Oþ Þn þ Mnþ
ð2Þ
Mnþ þ mH2 O ! ½MðH2 OÞm nþ
ð3Þ
In the ASR, Ni might be present as NiOOH(s), released in acidic pH water from the ASR, as shown in Eqs. (4) and (5) (Sakultung et al., 2007). The active nickel material is used as nickel oxyhydroxide (NiOOH), which is converted into Ni(OH)2 during battery discharge and reformed into NiOOH during recharging (Coman et al., 2013). 2þ
3.7. Correlation between metal recovery and extraction conditions Recovery of heavy metal did not show great dependency on variable pH values (R2 < 0.60), however, very good correlations were observed between the L/S ratio and heavy metals recovery from ASR (R2 = 0.959 for Mn, R2 = 0.963 for Fe, R2 = 0.985 for Ni and R2 = 0.891 for Cr; Fig. 10). In the correlation analysis between extraction temperature and metal recovery, Fe (0.957), Mn (0.867), Ni (0.880), and Cr (0.830) showed a high value of determination coefficient (R2; Fig. 9). Based on these results, the recovery
NiOOHðsÞ þ 2Hþ ðaqÞ ! Ni ðaqÞ þ 3=2H2 þ O2 2þ
NiðOHÞ2 ðsÞ þ 2Hþ ðaqÞ ! Ni ðaqÞ þ 2H2 O
ð4Þ ð5Þ
Anionic metal contaminants such as hexavalent chromium(HCrO 4 and CrO2 4 ), are adsorbed on positively charged functional groups and, therefore, cannot be removed from ASR by exchange with hydronium ions. Heavy metals retained in ASR can be extracted by inorganic acids as a result of solid phase dissolution and as a result of metal hydration as shown in Eq. (2).
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Fig. 11. TCLP test for heavy metals before and after recovery of heavy metals from the ASR (A = ASR size range 4.75–2.0 mm, B = size range 2.0–1.19 mm, C = size range 1.19– 0.425 mm, D = size range 0.425–0.25 mm, E = size < 0.25 mm and R = ASR after metal recovery).
Fig. 12. Changes of exchangeable and carbonate fractions and mobility factor of heavy metals before and after recovery of heavy metals: (a) Mn, (b) Fe, (c) Ni, (d) Cr, and (e) mobility factor, MF (A = ASR size range 4.75–2.0 mm, B = size range 2.0–1.19 mm, C = size range 1.19–0.425 mm, D = size range 0.425–0.25 mm, E = size < 0.25 mm and R = ASR after metal recovery).
3.9. Toxicity characteristics leaching procedure (TCLP) test Fig. 11 shows that the extraction concentrations of all selected heavy metals (Mn, Fe, Ni, and Cr) were greatly reduced in all grain
fractions (A, B, C, D, and E) of the ASR after metal recovery with distilled water at pH 2. In particular, reduction in metal extraction was mainly observed in grain sizes B and E. The large reduction in size E indicates that heavy metals included in the fine size of ASR are more
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J. Singh, B.-K. Lee / Waste Management xxx (2015) xxx–xxx Table 5 Determination coefficients (R2) of various kinetic models for recovery of heavy metals. Heavy metals
Extraction temperature (°C)
Determination coefficients (R2) of various kinetic models 1(1a)1/3
1–2a/3 (1a)2/3
1–3(1a)2/3 + 2(1a)
Mn
25 35 45 55
0.8128 0.8516 0.4091 0.4767
0.8244 0.8766 0.4136 0.4554
0.8192 0.8663 0.4117 0.4639
Fe
25 35 45 55
0.7208 0.7863 0.4931 0.5805
0.7765 0.8430 0.5836 0.6755
0.6412 0.6352 0.3735 0.4779
Ni
25 35 45 55
0.3041 0.5091 0.8728 0.8520
0.3047 0.5187 0.8650 0.8447
0.3044 0.5136 0.8495 0.8495
Cr
25 35 45 55
0.3820 0.7323 0.7032 0.6805
0.3783 0.9295 0.9084 0.9288
0.4578 0.1118 0.6254 0.2519
readily extracted, as discussed in the section on particle size effects. If the leachable concentration of heavy metals in the waste is above the regulatory level, the waste is considered ‘hazardous’ (Singh and Kalamdhad, 2013a). The smallest particle size fraction showed more leachability potential than the larger size fractions of the ASR. Chiang et al. (2008) also reported the smallest size fractions of ash generated from incineration of municipal solid waste as having more leachable potential of heavy metals. The concentration of Cr in any fraction of ASR was not beyond the threshold limits specified by the US EPA. The results of the present study suggest that distilled water (pH 2) could be effective for the recovery and reduction of toxicity by heavy metals from ASR. 3.10. Speciation and mobility of heavy metals In a risk assessment, the mobility and bioavailability of heavy metals are affected mainly by changes in pH, different forms of
11
metals, and ionic concentrations of leachate during landfilling of ASR. The mobilization of metal depends strongly on their mobility, concentration in the leachate, and solubility in water (Singh and Kalamdhad, 2013e). Fig. 12 plots the quantities of Mn, Fe, Ni, and Cr in exchangeable (F1) and carbonate (F2) fractions in sequential extractions of the different size fractions of ASR and the residues after the recovery of heavy metals under the conditions of pH 2, L/S 50, extraction time 60 min after the recovery. F1 fractions of Mn, Fe, Ni, and Cr were much smaller than F2 fractions for all size fractions (except size fraction A for Cr). F1 fractions of Mn and Fe seemed to increase with decreasing size of ASR; however, F1 fractions of Ni and Cr did not change significantly even with grain size fraction changes. The highest F2 fraction of Mn, Ni, and Cr were observed in size fractions E, B, and C, respectively. The F1 fraction of Cr, Ni, Fe and Mn for all the size fractions of ASR seemed to decrease after the recovery. The F2 fraction of these metals was also reduced after the recovery in the all size fractions (A, B, C, D and E). Fig. 12 shows the mobility factor (MF) of Mn, Fe, Ni, and Cr in different fractions of ASR before and after recovery of heavy metals. The MF was reduced in the range of 50.2–81.3% for Mn, 14.0–65.6% for Fe, 58.0–89.8% for Ni, and 63–73.3% for Cr in the final ASR residue. The total concentration of Fe was found much higher than the Mn and Ni, about 25 and 132 times respectively. However, F1 and F2 fractions of Fe contributed 2.9% of the total Fe, which is less than Mn (25.2%) and Ni (10.9%). The total concentration of Cr was around 209 times less than Fe concentration, but F1 and F2 fractions of Fe contributed similar as Cr. The present study also suggests that toxicity of heavy metals depends on different mobile fractions (F1 and F2) rather than total heavy metals concentration (solid sample digested with strong acids). 3.11. Kinetic of heavy metal recovery A kinetic study of the recovery of heavy metals from the ASR was performed at different temperatures of 25, 35, 45, and 55 °C. For a liquid/solid reaction system, the reaction rate is generally controlled by diffusion through the liquid film, diffusion through the solid or product layer, and the chemical reaction at the surface of the solid particles or a mixed of diffusion and chemical reactions
Fig. 13. Second-order reaction kinetics of Mn, Fe, Ni and Cr at different temperatures.
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Table 6 Risk assessment code (RAC) for heavy metals before and after recovery of heavy metals from ASR. Heavy metals
Mn Fe Ni Cr
RAC (%) A
AR
B
BR
C
CR
D
DR
E
ER
29.9 2.5 16.4 15.9
5.6 1.6 2.8 5.6
29.0 11.9 13.4 8.1
12.9 4.1 1.6 3.2
32.6 23.8 20.5 13.1
16.2 12.6 2.4 3.5
30.3 12.2 26.2 11.0
14.5 9.1 2.7 4.1
25.2 2.9 10.9 2.9
9.6 2.5 4.5 1.0
Note – A = ASR size range 4.75–2.0 mm, B = size range 2.0–1.19 mm, C = size range 1.19–0.425 mm, D = size range 0.425–0.25 mm, E = size < 0.25 mm, R = recovery of heavy metals
(Gharabaghi et al., 2013). It is important to establish a quantitative measurement of the rate of recovery/extraction kinetics and mechanism. If the ASR particles (size less than 250 mm) were considered to be spherical particles in a liquid–solid reaction system, the kinetics of metal recovery could be described using the shrinking core model (SCM) (Souza et al., 2007; Tian et al., 2010; Li et al., 2013). As shown in Table 5, the SCM model did not, in fact, fit the experimental data very well (R2 < 0.95). Li et al. (2014) also reported similar results from the kinetics of vanadium leaching by sulfuric acid from a spent industrial V2O5/TiO2 catalyst. Initially, the metal concentration increased very rapidly and then slowly reached a saturation value in the bulk liquid solution. This trend suggests that the kinetics model of the recovery process might be represented by a second-order rate law expression (Sakultung et al., 2007). The linear form of a second order kinetic model is expressed according to Ho et al. (2005). Plotting the value of t/Ct versus t will provide a straight line with the slope of 1/Cs and the intercept of 1/KC2: s
t=C t ¼ ðt=KC 2s Þ þ ðt=C s Þ
ð6Þ
where Ct is the concentration of metal ion (mg/L) at given time t (min), Cs is the amount of metal ion released at equilibrium (mg/ L), and K is the second-order rate constant for desorption (L/ mg min). The Cs and K were calculated as Cs = 1/slope and K = slope2/intercept. Fig. 13 shows the series plots of t/Ct as a function of extraction time (5–120 min) at various temperatures (25, 35, 45, and 55 °C) of the Mn, Fe, Ni, and Cr recovery processes using distilled water at pH 2. A coefficient of determination (R2 P 0.95) between the second-order model and the experimental results was found to show very good agreement. This confirms that there were two mechanisms in the recovery process: first, strong dissolution and scrubbing and then a much slower stage occurred related to external diffusion (Sakultung et al., 2007). 3.12. Risk assessment code (RAC) As shown in Table 6, RAC for Mn, Fe, Ni and Cr was reduced after recovery of heavy metals from ASR. The risk of Mn was changed from medium to low level in fraction A and E. However, in fraction B, C and D risk reduced but remained in medium category. Fe changed from medium risk to low risk category in A, B, D and E size fractions but in fraction C it was remained in medium risk category. The toxicity risk of Ni and Cr reduced from medium risk to low risk level in all grain size fractions. It can be attributed as at lower pH (pH 2) of water could be useful for reduction of toxicity risk of heavy metals. 4. Conclusions The important parameters that affect in recovery of heavy metals (Mn, Fe, Ni, and Cr) from the automobile shredder residue (ASR) can be ranked as: pH > L/S ratio > temperature > grain fraction > extracting time. The presence of bioavailable fractions
(water soluble, exchangeable, and carbonate) of heavy metals in the automobile shredder residue (ASR) confirms that ASR is potentially hazardous to the environment without proper disposal. The highest recoveries of heavy metals were found in the lowest size fraction of ASR at pH 2, L/S ratio of 100, and temperature of 25 °C, and thus the suggested conditions can be recommended for working in a real scale plant. The recovery of all selected heavy metals was increased with increase in the L/S ratio from 10 to 100 mL/g. The higher pH (>7) of distilled water was not effective for heavy metals recovery from the ASR. The smallest size fraction is considered the most hazardous but also the most favorable for the highest recovery of heavy metals, compared with the larger fractions. The ASR fractions P 4.75 mm have the high content of volatile solid and low concentration of heavy metals and thus these fractions can be recommended to be utilized for energy production. A very good correlation was observed between the L/S ratio and heavy metals recovery from ASR. Leachable concentration of Cr in all size fractions of ASR was under the threshold limits for leaching of solid wastes. The mobility factors and toxicity risk of the heavy metals were reduced greatly in the final residues of all fractions left after metal recovery. The second-order kinetic model fitted the data very well (R2 > 0.95) for the recovery of all selected heavy metals. Thus, distilled water at pH 2 was found to be a highly effective extracting agent for the recovery and reduction of toxicity of heavy metals in ASR. Acknowledgements This work was supported by the National Research Foundation of Korea (NRF) grant funded by the Ministry of Science, ICT and Future Planning (2013R1A2A2A03013138). References Abdel-Aal, E.A., 2000. Kinetics of sulphuric acid leaching of low grade zinc silicate ore. Hydrometallurgy 55 (3), 247–254. Anbia, M., Salehi, S., 2012. Removal of acid dyes from aqueous media by adsorption onto amino-functionalized nanoporous silica SBA-3. Dyes Pigments 94, 1–9. Astrup, T., Mosb, K.H., Christensen, T.H., 2006. Assessment of long-term leaching from waste incineration air-pollution-control residues. Waste Manage. 26, 803– 814. Aydogan, S., Aras, A., Canbazoglu, M., 2005. Dissolution kinetics of sphalerite in acidic ferric chloride leaching. Chem. Eng. J. 114, 67–72. Boughton, B., 2007. Evaluation of shredder residue as cement manufacturing feedstock. Resour. Conserv. Recycl. 51, 621–642. Chiang, K.Y., Jih, J.C., Chien, M.D., 2008. The acid extraction of metals from municipal solid waste incinerator products. Hydrometallurgy 93, 16–22. Coman, V., Robotin, B., Ilea, P., 2013. Nickel recovery/removal from industrial wastes: a review. Resour. Conserv. Recycl. 73, 229–238. Cossu, R., Fiore, S., Lai, T., Luciano, A., Mancini, G., Ruffino, B., Viotti, P., Zanetti, M.C., 2014. Review of Italian experience on automotive shredder residue characterization and management. Waste Manage. 34 (10), 1752–1762. EU Directive 2000/53/EC of the European parliament and the Council of 18 September 2000 on end-of-life vehicles. Off. J. Eur. Union 2000, L269, pp. 34-42. Fedje, K.K., Ekberg, C., Skarnemark, G., Steenari, Britt-Marie, 2010. Removal of hazardous metals from MSW fly ash-an evaluation of ash leaching methods. J. Hazard. Mater. 173, 310–317. Garau, G., Castaldi, P., Santona, L., Deiana, P., Melis, P., 2007. Influence of red mud, zeolite and lime on heavy metal immobilization, culturable heterotrophic microbial populations and enzyme activities in a contaminated soil. Geoderma 142, 47–57.
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Please cite this article in press as: Singh, J., Lee, B.-K. Pollution control and metal resource recovery for low grade automobile shredder residue: A mechanism, bioavailability and risk assessment. Waste Management (2015), http://dx.doi.org/10.1016/j.wasman.2015.01.035