Hydrometallurgical recovery of heavy metals from low grade automobile shredder residue (ASR): An application of advanced Fenton process (AFP)

Journal of Environmental Management 161 (2015) 1e10

Contents lists available at ScienceDirect

Journal of Environmental Management journal homepage: www.elsevier.com/locate/jenvman

Research article

Hydrometallurgical recovery of heavy metals from low grade automobile shredder residue (ASR): An application of advanced Fenton process (AFP) Jiwan Singh a, b, Byeong-Kyu Lee a, * a b

Department of Civil and Environmental Engineering, University of Ulsan, Daehkro 93, Namgu, Ulsan 680-749, South Korea Department of Environmental Engineering, Kwangwoon University, Seoul 139-701, South Korea

a r t i c l e i n f o

a b s t r a c t

Article history: Received 14 March 2015 Received in revised form 18 June 2015 Accepted 19 June 2015 Available online xxx

To investigate the leaching and recovery of heavy metals from low-grade automobile shredder residue (ASR), the effects of nitric acid (HNO3) and hydrogen peroxide (H2O2) concentrations, liquid/solid (L/S) ratio, leaching temperature and ASR particle size fractions on the heavy metal leaching rate were determined. The heavy metals were recovered by fractional precipitation and advanced Fenton process (AFP) at different pHs. The toxicity characteristic leaching procedure (TCLP) test was also performed in the residue remaining after heavy metal leaching to evaluate the potential toxicity of ASR. The heavy metal leaching efficiency was increased with increasing HNO3 and H2O2 concentrations, L/S ratio and temperature. The heavy metal leaching efficiencies were maximized in the lowest ASR size fraction at 303 K and L/S ratio of 100 mL/g. The kinetic study showed that the metal leaching was best represented by a second-order reaction model, with a value of R2 > 0.99 for all selected heavy metals. The determined activation energy (kJ/mol) was 21.61, 17.10, 12.15, 34.50, 13.07 and 11.45 for Zn, Fe, Ni, Pb, Cd and Cr, respectively. In the final residue, the concentrations of Cd, Cr and Pb were under their threshold limits in all ASR size fractions. Hydrometallurgical metal recovery was greatly increased by AFP up to 99.96% for Zn, 99.97% for Fe, 95.62% for Ni, 99.62% for Pb, 94.11% for Cd and 96.79% for Cr. AFP is highly recommended for the recovery of leached metals from solution even at low concentrations. © 2015 Elsevier Ltd. All rights reserved.

Keywords: ASR Heavy metals Leaching Kinetics Analysis of variances (ANOVA) AFP

1. Introduction Hydrometallurgical recovery of toxic heavy metals from the hazardous waste of automobile shredder residue (ASR) is a possible alternative to reduce the environmental impacts of low-grade ASR, which is disposed of into landfill (Tuncuk et al., 2012; Kubonova et al., 2013). ASR has been considered hazardous waste due to presence of heterogeneous material and alterable conformation such as toxic heavy metals and organic pollutants (Granata et al., 2011). ASR contributes about 20e25 % of the total volume of end of life vehicle (Kurose et al., 2006; Joung et al., 2007). Despite containing heavy metals, most low-grade ASR is directly disposed of into landfills (Donaja et al., 2010). The leaching of heavy metals from landfill ASR is a key environmental problem. The heavy metals released from ASR may raise public concern by polluting soil,

* Corresponding author. E-mail address: [email protected] (B.-K. Lee). http://dx.doi.org/10.1016/j.jenvman.2015.06.034 0301-4797/© 2015 Elsevier Ltd. All rights reserved.

ground water and surface water. Generally, strong oxidizing conditions are needed for leaching of heavy metals from ASR. The combination of nitric acid (HNO3) and hydrogen peroxide (H2O2) can act as a strong oxidizing agent. Therefore, this combination can be highly proficient for the leaching of heavy metals from ASR (Bas et al., 2014). Nitric acid is generally used with hydrochloric acid for the determination of heavy metals in ASR (Granata et al., 2011). Nitric acid alone is a highly proficient extracting agent for the leaching of heavy metals such as Cu, Ni, Pb and Ag (Le et al., 2011; Bas et al., 2014). Therefore, the treatment of low-grade ASR with its relatively high heavy metal content has not been considered effectively. Separation of heavy metals from low-grade ASR by physical separation methods has not been considered a viable option (Kurose et al., 2006). Therefore, an appropriate management tactic needs to be found for the handling of ASR before landfill disposal. Heavy metal leaching using acids has been considered a viable option because of its ease of operation, rapid effect and high recovery efficiency (Deng et al., 2009). Therefore, hydrometallurgical recoveries of heavy metals from ASR

2

J. Singh, B.-K. Lee / Journal of Environmental Management 161 (2015) 1e10

can be useful for conserving metals resource and preventing heavy metal pollution associated with ASR disposal into landfill, as shown in Fig. S1. The application of zero-valent iron (ZVI) has become the most auspicious novel technology for pollution remediation due to its low toxicity, economy, ease of handling, and low Fe concentration in effluent, with no further treatment being needed. The nano/ micro ZVI has been applied for the degradation of chelated heavy metal in wastewater (Fu et al., 2012). However, it has also been applied for the degradation of organic pollutants such as tetrachloroethene (chlorinated organic compounds) (Amir and Lee, 2011), 2, 4, 6-trinitrotoluene (Zhang et al., 2010) and nitrate (Kassaee et al., 2011), as well as for the removal of dyes (He et al., 2012). The present study investigated the application of the advanced Fenton process (AFP) for increasing the heavy metals recovery through precipitation study. The study objectives are: i) leaching of heavy metals (Zn2þ, Fe3þ/Fe2þ, Pb2þ/Pb4þ, Cd2þ, Ni2þ and Cr3þ/Cr6þ)) from low-grade ASR with HNO3, ii) enhancing metals recovery by AFP from the leaching solution of ASR through the precipitation study, and iii) reducing the toxicity risk of metals in ASR under landfilling condition. 2. Materials and methods In order to remove heavy metals (Zn, Fe, Ni, Pb, Cd and Cr) from ASR, a sequence of leaching and recovery procedures was used, as shown in Fig. S2. 2.1. Collection and preparation of ASR About 20 kg of ASR sample (a representative sample) was taken from the automobile shredder plants of Steel SCRAP Ulsan factory, Nam-gu, Ulsan, South Korea. The following various ASR size fractions were obtained by passing through different sieves: the ranges of 4750-2000 mm, 2000e1190 mm, 1190e425 mm, 425e250 mm and <250 mm, and these particle sizes were denoted as A, B, C, D and E, respectively. These fractions were dried in a hot air oven at 105  C for 24 h before being leached. 2.2. Leaching of heavy metals A temperature-controlled water bath shaker (HST-205SW) was used for the heavy metal leaching. A series of 250 mL covered conical flasks were located in a water bath shaker for stirring at 150 rpm. Heavy metal leaching from ASR was performed at different HNO3 concentrations (0.5, 1.0, 2.0 and 3.0 M), L/S ratios (10, 25, 50 and 100 mL/g), temperatures (303, 313, 323 and 333 K) and particle sizes (4750e2000 mm, 2000e1190 mm, 1190e425 mm, 425e250 mm and <250 mm). The effect of different H2O2 concentrations on heavy metal leaching was investigated at 0.5 M HNO3 concentration. During 120 min of leaching, samples were collected at 5, 15, 30, 60, 90 and 120 min. For heavy metals analysis, mixtures of ASR with solution were centrifuged and then filtered by filter paper (0.45 mm). Experiments were conducted three times and the averaged results are presented. Heavy metal leaching efficiency was determined by using the following formula (Hong et al., 2000):

Leaching efficiency ð%Þ ¼ ðCL =CT Þ  100

(1)

where CL is the concentration of metals in liquid (mg/kg) and CT is the total concentration of metals obtained after strong acid digestion (mg/kg). Residue samples were washed and dried after leaching. Residues were used for the toxicity characteristic leaching procedure

(TCLP) test according to the standard method (United States Environmental Protection Agency, US EPA 1311, 1992).

2.3. Recovery of heavy metals through fractional precipitation method The precipitation study was carried out according to the jar test method (Sthiannopkao and Sreesai, 2009; Fu et al., 2012) in a 500 mL beaker with 200 mL leaching liquid. A temperature controlled magnetic stirrer (GLHF-G, Global lab) was used for the precipitation chemical reaction, with an agitation rate of 80 rpm to increase the precipitation. The precipitation by chemical reaction was controlled at 35  C (Chen et al., 2014). Heavy metals were precipitated at different pHs (pH adjusted using 1.0 N NaOH and 1.0 N HCl) of leaching liquid. The required pH for each metal was maintained at the beginning of the stirring. The pH was maintained in the range of pH 2e3 for Fe (Chen et al., 2014), 9e10 for Pb (Zhang et al., 2008), 10e11 for Zn (Yan et al., 2008), 11e12 for Ni (Fu et al., 2012), 8e9 for Cr and 12e14 for Cd (Quina et al., 2009). For measuring pH during metal precipitation, a digital pH meter (ORION 5 STAR, thermo scientific) was used. The residue filtrate remaining after each step was used for the next precipitation step. The precipitation study was also conducted to enhance the fractional precipitation of heavy metals by adding nano ZVI (nZVI) under Fenton's type reaction system (Fu et al., 2012). After each successive step, nZVI (concentration of 0.125 g/L) was added separately with 0.5% (V/V) H2O2 (conc. 30%) under acidic condition (pH 1.5e3.0) (Fu et al., 2012). The mixture of ZVI and H2O2 was stirred for 20 min for reaction and then the required pH was maintained during the repeated Jar tests method. Each fractional precipitation experiment was repeated at least three times. After each precipitation step, the solution was centrifuged at 4000 rpm for 10 min and then filtered with a 0.45 mm filter. A 10 mL sample was collected from each step for analyzing the residual concentration of heavy metals remaining after each precipitation. Heavy metals recovery was calculated by using the following Eq. (2) (Fu et al., 2012):

Metal recovery ð%Þ ¼ ½ðA  BÞ=A  100

(2)

where A is the metals in the extracted liquid (mg/L) and B is the metals in the solution remaining after precipitation (mg/L).

2.4. Synthesis of nano zero valent iron (nZVI) The nZVI was synthesized according to the previous research (Fang et al., 2011; Patela et al., 2013). Two grams of FeCl3 salt was dissolved in 100 mL of ethyl alcohol. An appropriate amount of polyvinylpyrrolidone (PVP) (nanoparticle/PVP, w/w ¼ 1/1) was added into the FeCl3 solution with agitation to entirely mix the solution. The solution of reducing agent was prepared by dissolving 2 g of NaBH4 in 100 mL of distilled water. The borohydride solution was added drop-wise to the ferric solution while being stirred using a temperature control magnetic stirrer. Black agglomerated particles appeared immediately after the addition of NaBH4, indicating the presence of nZVI. After complete addition of the NaBH4 solution, stirring was stopped after the mixture had changed completely black. The mixture was centrifuged (Centrifuge MF 600, Hanil Science industrial Co. Ltd., South Korea) and the separated nanoparticles (NPs) were washed with ethyl alcohol three times to remove any excess borohydride solution. Finally, all the NPs were dried in vacuum overnight at 50  C before use in the precipitation study.

J. Singh, B.-K. Lee / Journal of Environmental Management 161 (2015) 1e10

2.5. Characterization of residues remaining after leaching and nZVI particles The morphology of ASR before and after leaching, as well as of the synthesized nZVI particles, was viewed with a field emission scanning electron microscope (FE-SEM JEOL JSM-6500F, USA). The crystal structure of the residues remaining after leaching was investigated by X-ray powder diffraction (Bruker AXN, Germany) in the 2q range of 10e80 using Cu-Ka radiation (l ¼ 1.5418 Å), with an accelerating voltage of 40 kV at an applied current of 30 mA. A Thermo Fisher K-alpha model was used for X-ray photoelectron spectroscopy (XPS) measurements to determine the chemical forms of the metals present in the original ASR. The morphology of the synthesized nZVI particles was viewed with FE-SEM and the composition of nZVI was analyzed with energy dispersive X-ray spectrometry (EDX). The size and morphology of the synthesized nZVI were characterized by transmission electron microscopy (TEM; H-8100, Hitachi). Before TEM analysis, nZVI particles were dispersed by an ultrasonicator for 1 h, and then the mixture of NPs and ethanol was deposited as several droplets on a carbon-coated Cu-grid. The samples were placed in a vacuum chamber after complete ethanol evaporation. 2.6. Heavy metal analysis A flame atomic absorption spectrometer (FAAS) (Varian Spectra 55) was used to analyze the concentrations of Zn, Fe, Ni, Pb, Cd and Cr in the leachate and the solution remaining after precipitation. One gram of ASR was digested (triplicate) for measuring the total metal concentrations. A mixture of 20 mL of H2SO4 and HClO4 (5:1) was used for the digestion ASR samples in a closed Teflon vessel (Singh and Klamdhad, 2012, 2013). For FAAS calibration, standard solutions of all selected metals were purchased from Cica-reagent, Kanto Chemical Co. Inc., Japan. Statistical analysis of the leaching data was performed by analysis of variance (ANOVA) to analyze any significant difference between the different leaching conditions. The Statistical Package for the Social Sciences (SPSS) Package (Version 21) was used for the statistical analysis. The likelihood ratio (F) and probability (p) were used for statistical significance for the tests. 3. Results and discussion 3.1. Heavy metal leaching 3.1.1. Characterization of ASR fraction E before and after heavy metal leaching X-ray photoelectron spectroscopy (XPS) was used to analyze the availability of metals and their compounds in ASR. The ASR composition was characterized by XPS. Characteristic photoelectron spectra and peaks were detected for Zn, Cu, Mn, Fe, Ni, Pb, Cd, Cr, Co and Ag, as well as oxygen and carbon (Fig. S3a). The order of relative elemental atomic concentrations was as follows: C1s (61.41%) > O1s (25.14%) > Si2p (8.72%) > Cr2p (1.29%) > Co2p (0.68%) > Zn2p3 (0.67%) > Fe2p (0.65%) > Mn2p3 (0.61%) > Cu2p3 (0.57%) > Ag3d (0.16%) and Pb4f (0.11%) (Table 1). The high levels of C and O are indicative of hydrocarbons, phenolic compound, polyesters, hydroxyl groups and epoxies; which are present in polypropylene, polyurethane, polyvinylchloride, acrylonitrile butadiene styrene, poly-methyl-methacrylate and polyethylene terephthalate (Vermeulen et al., 2011). As shown in Figs. S3b and c, carbon was found in CeC or CeH, carbonate, and metals carbonate form, whereas oxygen occurred as SiO2, metal CO3 and metal oxides. Figs. S3 d, e, f, g, h and i show that heavy metals were mainly found in ASR in the form of elements, metal CO3, metal oxides and

3

metal SiO4. The morphology of the E fraction of ASR before and after heavy metal leaching was examined using SEM (Fig. 1). Fig. 1a and b illustrate the images of ASR before metal leaching. After the leaching, the particles in the residues were characterized by a random surface, many rough holes and a high degree of organic degradation (Li et al., 2013). The surface of ASR appeared damaged after the leaching, possibly due to attack of the acidic solution. Fig. 1d, e and f clearly show that the ASR particle size was reduced during the leaching with 1.0, 2.0 and 3.0 M HNO3; however, Fig. 1c shows a slight change in the structure of ASR by leaching with 0.5 M HNO3. Fig. 2 presents the XRD pattern of fraction E before and after heavy metal recovery. Before the recovery, the XRD patterns showed seven peaks at 2q ¼ 20.88, 26.68, 27.08, 29.56, 42.26, 55.08 and 60.12 . The peaks at 2q ¼ 29.56, 42.26 and 54.88 may correspond to ZnS/NaCl/ZnCl2, CaO and Fe3O4/CuO, respectively (Shibayama et al., 2006; Ferella et al., 2014). The peak at 2q ¼ 54.88 did not appear in the residue remaining after leaching, which confirmed that oxides of Fe and Cu were extracted by acid solution. The peaks at 2q ¼ 28.18, 36.52 and 50.32 may correspond to ZnCl2, AlCl3, and Al2Fe2Si, respectively; however, the peaks at 2q ¼ 59.96 and 75.95 correspond to SiO2 (Wang et al., 2001). The peaks at 2q ¼ 36.52 and 50.32 appeared in all acid-treated residues but not in the original ASR. The peak at 2q ¼ 59.96 was reduced in acid-treated ASR as compared to the original ASR. The peak at 2q ¼ 28.18 appeared with high intensity in ASR treated with 0.5 M HNO3, but the peak intensity was decreased in ASR treated with 1.0e3.0 M HNO3. This was explained by the ability of the higher acid concentration to extract more metal chlorides compared to the 0.5 M HNO3 concentration (Wang et al., 2001). The intensities of some peaks were reduced after the metal leaching due to the suppression of material crystallinity (Shibayama et al., 2006). These study results clearly confirmed that heavy metal leaching with acids changes the composition and structure of ASR.

3.1.2. Effect of nitric acid concentration on metal leaching Fig. 3 presents the effect of HNO3 concentration (0.5 Me3.0 M) on the heavy metal leaching time at constant experimental conditions (L/S-50 mL/g, agitation speed-150 rpm, particle size< 0.25 mm and temperature 303 K). Most of the metal leaching efficiencies increased very quickly in the initial 30 min of leaching and gradually reached constant values at the end of the leaching time. The metal leaching efficiencies were increased from 89.6 to 96.5% for Zn, from 27.8 to 32.6 % for Fe, from 25.9 to 40.4 % for Ni, from 81.9 to 96.8 % for Pb, from 31.6 to 50.0 % for Cd and from 20.0 to 25.2 % for Cr, with increasing HNO3 concentration from 0.5 to 3.0 M. The leaching efficiency of all selected metals became constant faster with 3.0 M HNO3 than with 0.5 M HNO3. The economically optimal HNO3 concentration for metal leaching was kept at 0.5 M Deng et al. (2009) and Gheju et al. (2011) also reported an increase in metal extraction with increasing HNO3 concentration from bio solids. The changes of heavy metals with various HNO3 concentrations were significant (F ¼ 37.0, p < 0.001 for Zn, F ¼ 23.0, p < 0.001 for Fe, F ¼ 28.8, p < 0.001 for Ni, F ¼ 12.4, p < 0.001 for Pb, F ¼ 32.6, p < 0.001 for Cd and F ¼ 25.6, p < 0.001 for Cr). The increased metal leaching from ASR was attributed to increased dissolution of fine particles of ASR and increased conversion of metals to their ionic forms with increasing HNO3 acidity. The leaching of heavy metals from ASR with HNO3 is based on the exchange of H3Oþ with metals present in ASR, as shown in Eqs. (3) and (4):

4

J. Singh, B.-K. Lee / Journal of Environmental Management 161 (2015) 1e10

Table 1 Binding energy (BE), full width half maximum (FWHM), atomic percentage and chemical state from XPS spectra. Elements

Peak BE (eV)

FWHM (eV)

Atomic content (%)

Chemical state

O1s C1s Ag3d Zn2p3 Fe2p Cr2p Pb4f Mn2p3 Co2p Cu2p3 Si2p

532.27 285.21 368.49 1022.50 726.85 579.96 139.87 641.31 787.73 942.88 102.48

3.25 2.98 1.47 1.93 5.38 16.29 2.75 1.42 1.79 1.71 1.98

25.14 61.41 0.16 0.67 0.65 1.29 0.11 0.61 0.68 0.57 8.72

SiO2, Metal CO3, Metal oxide CeC or CeH, carbonate, metals carbonate Metal CO3, Metal oxide, elemental form, Metal SiO4

Fig. 1. Scanning electron microscopy image of fraction E of ASR: before treatment (a and b), and after treatment with 0.5 M (c), 1.0 M (d), 2.0 M (e), and 3.0 M HNO3 (f).

  ASR  Mnþ þ nH3 Oþ /ASR  H3 Oþ þ Mnþ

(3)

 nþ Mnþ þ mH2 O/ MðH2 OÞm

(4)

n

3.1.3. Effect of hydrogen peroxide concentration on metal leaching The effect of H2O2 concentration (0.5, 1.0 and 2.0%) on metal

leaching was studied with 0.5 M HNO3, while maintaining all other experimental conditions constant. Mixing HNO3 with H2O2 gives a good oxidizing agent. Therefore, the metal leaching was increased effectively by increasing the H2O2 concentration. The metal leaching efficiencies were increased from 89.6 to 98.2 % for Zn, from 27.8 to 30.1 % for Fe, from 25.9 to 50.0 % for Ni, from 81.9 to 98.5 % for Pb, from 31.6 to 39.6 % for Cd and from 20.0 to 26.1% for Cr with the increase in H2O2 concentration from 0.0 to 2.0 % after 2 h of leaching (Fig. S4). A leaching equilibrium was achieved very quickly

J. Singh, B.-K. Lee / Journal of Environmental Management 161 (2015) 1e10

5

ASR with HNO3. The changes of heavy metals with various H2O2 concentrations were significant (F ¼ 21.6, p < 0.001 for Zn, F ¼ 14.3, p < 0.001 for Fe, F ¼ 36.6, p < 0.001 for Ni, F ¼ 18.6, p < 0.001 for Pb, F ¼ 71.4, p < 0.001 for Cd and F ¼ 23.0, p < 0.001 for Cr).

Fig. 2. XRD spectra of ASR (fraction E) before and after treatment with acid.

(30 min) in all H2O2 (0.5e2.0%) treatments as compared to that (60 min) in 0.5 M HNO3 concentration without H2O2 addition. All metal leaching efficiencies were increased greatly with increasing H2O2 concentration. A leaching equilibrium for Zn and Pb was obtained after 15 min of leaching with 2.0% H2O2 treatment. The increased leaching efficiency of metals with increasing H2O2 concentration was attributed to the action of H2O2 as a powerful oxidant that can be converted into hydroxyl radical (OH) (Badmus et al., 2007). Hydroxyl radicals are involved in the dissolution of

3.1.4. Effect of liquid/solid (L/S) ratio on metal leaching The effect of L/S ratio (10, 25, 50 and 100 mL/g) was studied under the following conditions: 0.5 M HNO3, temperature 303 K and ASR particle size  250 mm. Fig. 4 shows that the metal leaching efficiency greatly increased with increasing L/S ratio. The metal leaching efficiencies increased from 53.8 to 98.6 % for Zn, from 23.2 to 44.0 % for Fe, from 19.9 to 35.9% for Ni, from 54.8 to 99.2 % for Pb, from 8.6 to 50.6 % for Cd and from 13.5 to 38.7 % for Cr with increasing L/S ratio from 10 to 100 mL/g after 2 h of leaching. The leaching equilibrium for all selected metals was achieved within 15 min of leaching at the L/S ratio of 100 mL/g Aydogan et al. (2005) has also reported that the Zn recovery increased with an increasing in the L/S ratio during Zn extraction from sphalerite. The metal leaching efficiency is strongly expected to increase with increasing amount of solution; however, the process economy and separation of metals from the liquid will be affected. Considering the process economy and performance, the present study results suggested 50 mL/g L/S as the optimum for metal leaching. The metal leaching efficiency improved with decreasing ASR quantity, because the increased volume of leaching solution promotes the mass transfer process of ASR and acid solution during the leaching (Aydogan et al., 2005). A higher volume of leaching solution leads to increased leaching but lower metal contents in the leaching liquid, which may affect the separation efficiency of metals from

Fig. 3. Effect of HNO3 concentration on metal leaching (temperature 303 K, particle size < 250 mm, and L/S ratio 50 mL/g).

6

J. Singh, B.-K. Lee / Journal of Environmental Management 161 (2015) 1e10

Fig. 4. Effect of liquid/solid (L/S) ratio on heavy metal leaching (temperature 303 K, particle size < 250 mm, and 0.5 M HNO3).

the ASR leachate. The changes of heavy metals at all selected L/S ratios were significant (F ¼ 15.7, p < 0.001 for Zn, F ¼ 122.2, p < 0.001 for Fe, F ¼ 17.8, p < 0.001 for Ni, F ¼ 30.8, p < 0.001 for Pb, F ¼ 59.1, p < 0.001 for Cd and F ¼ 243.5, p < 0.001 for Cr). 3.1.5. Effect of temperature on metal leaching The effect of leaching temperature (303, 313, 323 and 333 K) was studied on metal leaching at different conditions: HNO3 concentration 0.5 M, L/S ratio 50 mL/g and ASR particle size  0.25 mm. Fig. 5 shows the effect of temperature on the metal leaching efficiency from the low-grade ASR. As predicted, metal leaching increased from 89.6 to 97.7 % for Zn, from 27.8 to 48.7 % for Fe, from 25.9 to 39.4 % for Ni, from 81.9 to 97.9 % for Pb, from 31.6 to 49.8 % for Cd, and from 20.0 to 31.0 % % for Cr as the temperature was increased from 303 to 333 K. Temperature had a strong effect on the metal leaching from ASR. The reaction rate increased greatly at high leaching temperature, which was attributed to the increased kinetic energy of the water molecules present in the leaching solution (Sakultung et al., 2007). The changes of heavy metals with increasing leaching temperature were significant (F ¼ 17.1, P < 0.001 for Zn, F ¼ 37.0, P < 0.001 for Fe, F ¼ 44.3, P < 0.001 for Ni, F ¼ 5.3, P < 0.001 for Pb, F ¼ 74.3, P < 0.05 for Cd and F ¼ 13.1, P < 0.001 for Cr). 3.1.6. Effect of different particle sizes on metal leaching Metal leaching was carried out with different particle sizes at different experimental conditions: HNO3 concentration 0.5 M, L/S 50 mL/g, and temperature 303 K. Fig. S5 illustrates the effect of different particle sizes of ASR on the metal leaching. The leaching

efficiency of all selected metals was increased from 8.9 to 89.6 % for Zn, from 10.7 to 27.8 % for Fe, from 3.8 to 25.9 % for Ni, from 14.4 to 81.9 % for Pb, from 3.3 to 31.6 % for Cd and from 2.6 to 20.0 % for Cr with decreasing ASR particles size from fraction E (<0.25 mm) to fraction A (4.75e2.0 mm). The leaching efficiency of Zn and Cd was increased about 10 times; however, Ni, Pb and Cr increased around 6 times in fraction E compared to fraction A of ASR. The initial extraction rate of Fe was faster in fraction E in comparison to the fraction D; however, extraction became almost constant in fractions D and E after 30 min of leaching. The metal leaching was maximized in the smallest size fraction (E), which was attributed to the more expansive exchange surfaces available between the acid solution and the fine particles (Gharabaghi et al., 2013). A leaching equilibrium for most of the metals was achieved within 15 min of leaching, particularly in fraction A of ASR; it can be attributed as HNO3 was highly efficient for washing of metals from larger ASR particles. The changes of heavy metals with all ASR size fractions were significant (F ¼ 25.4, p < 0.001 for Zn, F ¼ 38.0, p < 0.05 for Fe, F ¼ 37.6, p < 0.001 for Ni, F ¼ 21.7, p < 0.001 for Pb, F ¼ 100.7, p < 0.05 for Cd and F ¼ 80, p < 0.05 for Cr). 3.1.7. Kinetics of metal leaching The kinetics of the metal leaching from ASR was investigated at different temperatures (303, 313, 323 and 333 K). Fig. 6 shows the plots of t/Ct as a function of leaching time for different metals during 120 min of leaching. The metal leaching rate in the bulk liquid solution firstly increased very quickly, and then gradually extended to a saturation value. This trend suggested a second-order reaction model to characterize the kinetics model for the leaching

J. Singh, B.-K. Lee / Journal of Environmental Management 161 (2015) 1e10

7

Fig. 5. Effect of temperature on heavy metal leaching (particle size < 250 mm, L/S ratio 50 mL/g, and 0.5 M HNO3).

process, which is demonstrated in Eqs. (5) and (6) (Sakultung et al., 2007). The slope of 1/Cs and the intercept of 1/kC2s can be obtained by plotting t/Ct versus t as a straight line.

dC=dt ¼ kðCs  Ct Þ2

2

t=Ct ¼ t=kðCs Þ þ t=Cs

(5)

(6)

where Ct is the concentration of metal leached (mg/L) at given time t (min), Cs the concentration of metal leached at equilibrium (mg/L) and k the rate constant for leaching (L/mg min). The experimental results for all the metals were best fitted with the second-order reaction kinetic model (R2  0.999). Therefore, the following two mechanisms contributed during the metal leaching process: i) fast leaching by increased metal scrubbing accompanied by digestion of ASR in acidic solution, and ii) slow leaching by external diffusion for the remaining metals in ASR (Sakultung et al., 2007). The values of Cs and k for different metals are given in Table 2 at different leaching temperatures. The Cs value for all selected metals was increased by increasing the leaching temperature: from 86.96 to 94.34 mg/L for Zn, from 196.08 to 357.14 mg/L for Fe, from 1.45 to 2.14 mg/L for Ni, from 25.84 to 30.40 mg/L for Pb, from 0.42 to 0.64 mg/L for Cd and from 0.7 to 0.95 mg/L for Cr as the temperature was increased from 303 to 333 K. The rate constant (k) also increased by increasing the leaching temperature from 303 to 333 K (Table 2).

3.1.8. Determination of activation energy The activation energy (Ea) was determined according to Arrhenius equation, given in Eq. (7) (Szymczycha-Madeja, 2011). The plot of ln k against 1/T should be a straight line with a slope of Ea/R and the intercept of ln A.

lnk ¼ lnA  ðEa=RTÞ

(7)

where k is the rate constant (L/g. min), A is denoted as the preexponential factor, Ea represents the activation energy of metal leaching (kJ/mol), R is the gas constant (8.314 J/mol. K), and T is the leaching temperature (K). Ea was calculated by the plot of ln (k) against 1/T (Fig. S6). Ea values (kJ/mole) were determined as around 21.61 for Zn, 17.09 for Fe, 12.15 for Ni, 34.50 for Pb, 13.07 for Cd and 11.45 for Cr. The Ea values of Fe, Ni, Cd and Cr were <20 kJ/mol, which indicating the diffusion process was involved in the leaching process. However, the Ea values of Zn and Pb were between 20 and 40 kJ/mol (Table 2), which suggests that the leaching rates of both metals are controlled by both diffusion and chemical processes (Sakultung et al., 2007). 3.2. Toxicity characteristics leaching procedure (TCLP) test Different ASR particle size fractions before and after treatment (details given in Section 3.1.6) were subjected to TCLP test. Fig. S7 shows that the leaching concentrations of all selected metals were greatly decreased in all ASR grain size fractions after treatment, even with 0.5 M HNO3. The threshold limits in the waste for Cd, Cr and Pb are 20, 100 and 100 mg/kg, respectively (US EPA

8

J. Singh, B.-K. Lee / Journal of Environmental Management 161 (2015) 1e10

Fig. 6. Second-order leaching kinetics of heavy metals at different temperatures as a function of leaching time.

Table 2 Linearization of the kinetic model of the second-order leaching of heavy metals at various leaching temperatures and activation energies. Heavy metals

Temperature (K)

Cs (mg/L)

k (L/mg min)

Activation energy (kJ/mole)

Zn

303 313 323 333 303 313 323 333 303 313 323 333 303 313 323 333 303 313 323 333 303 313 323 333

86.96 91.74 92.59 94.34 196.08 263.18 294.12 357.14 1.45 1.72 1.90 2.14 25.84 26.95 28.74 30.40 0.42 0.52 0.62 0.64 0.71 0.75 0.87 0.95

0.0121 0.0186 0.0224 0.0268 0.0009 0.0010 0.0013 0.0016 0.1260 0.1379 0.1710 0.1901 0.0173 0.0212 0.0402 0.0553 0.3134 0.3639 0.4109 0.6983 0.2169 0.2456 0.2719 0.3311

21.61

Fe

Ni

Pb

Cd

Cr

Method 1311, 1992). However, in the present study, the leached concentrations (mg/kg) of these metals were in the range of 0.98e1.1 for Cd, 0.25e2.50 for Cr and 22.5e38.2 for Pb in all ASR size fractions, i.e., within the threshold limits. These results confirmed that even 0.5 M HNO3 could be highly effective for reducing the toxicity of metals in different ASR size fractions. Therefore, the treated ASR was not considered hazardous waste

17.10

12.15

34.50

13.07

11.45

(Singh and Kalamdhad, 2013). 3.3. Recovery of heavy metals 3.3.1. Characterization of nano zero valent iron (nZVI) The morphology and structure of the ZVI NPs are shown in Fig. S8. The nZVI were mostly spherical in shape and formed

J. Singh, B.-K. Lee / Journal of Environmental Management 161 (2015) 1e10

9

precipitated as hydrated oxides/oxy-hydroxide, which has comparatively steady and adsorptive surface characteristics. The Fe2þ/Fe3þ hydroxide precipitate has high capability to adsorb metals ions in single and multi-adsorbate systems (Benjamin, 1983; Feng et al., 2000). All selected metals could be absorbed as cations (Zn2þ, Cd2þ, Ni2þ, Pb2þ Cr3þ) on the surface of Fe2þ/Fe3þ hydroxides in alkaline pH (8e12). According to Fu et al. (2012), Eqs. (8)e(10) show that the surface of nZVI is generating Fe2þ and H2 gas under acidic condition; Fe2þ reacts quickly with H2O2 to generate OH0 in AFP, and is then converted into Fe3þ in the solution. The nZVI surface can then reduce the Fe3þ down to Fe2þ.

chain-like structure clumping, which was attributed to magnetic interactions between the ZVI particles (Fang et al., 2011). These chain-like iron clumps were also found by other researchers (Shih et al., 2011; Li et al., 2013). The SEM and TEM images of nZVI show that the diameters of the produced nZVI particles were less than 100 nm (Fig. S8a and b). Fig. S8c shows the XRD characteristic peaks at 2q of 35.56 and 44.94 , which specify the existence of iron oxides (FeO and Fe3O4) and ZVI NPs, respectively. The wide-ranging iron peak reveals the amorphous structure of the produced ZVI particles (Chen et al., 2012). The XRD analysis shows that the iron oxides could be precipitated on the ZVI surface during reduction reaction with sodium borohydride. 3.3.2. Precipitation study The leaching efficiencies and heavy metal contents in the original ASR and in the filtrate are given in Table 3. The fractional precipitation study was carried out with the metal leached solution extracted from the fraction E of ASR under the following conditions: 1.0 M HNO3, L/S 50 mL/g, temperature 333 K, shaking speed 150 rpm and leaching time 2 h. As shown in Table 4, the recoveries of metals from the filtrate were increased effectively by AFP from 98.24 to 99.97 % for Fe, from 92.41 to 96.79 % for Cr, from 98.64 to 99.62 % for Pb, from 98.57 to 99.96 % for Zn, from 84.04 to 95.62 % for Ni and from 90.33 to 94.11 % for Cd. The highest metal recovery occurred in steps 1e3 of AFP, after which the recovery was slightly enhanced further in the following steps. The recovery of some heavy metals (Fe, Zn and Pb) was >99% in AFP, which is much higher than in the case without AFP. This is because of the high concentrations of these metals (Fe: 205 mg/L, Zn: 91 mg/L and Pb: 27.5 mg/L) in the solution as compared to other metals. A clear solution was observed after step 3 of AFP and centrifugation was not required after this step. Recovery of Ni was increased about 4 times in step 1 of AFP as compared to that without AFP. AFP was highly effective even at very low concentration of metals (Cr, Cd and Ni) present in the solution. The possible mechanism for the increased recovery of heavy metals in AFP is as follows: at alkaline pH, Fe2þ/Fe3þ were

nZVI þ 2Hþ /Fe2þ þ H2 [

(8)

Fe2þ þ H2 O2 /Fe3þ þ OH þ OH

(9)

2Fe3þ þ nZVI/3Fe2þ

(10)

4. Conclusions The heavy metal leaching efficiency increased greatly with increasing HNO3 concentration, L/S ratio and temperature. The combined effect of HNO3 and H2O2 was also highly supportive for metal recovery. Activation energy study revealed that the leaching rates of all selected metals were regulated by both diffusion and chemical processes. The leaching data were best fitted by the second-order kinetic model (R2 > 0.99) for the leaching of all six selected heavy metals. The TCLP test revealed that all metal contents were under their threshold limits in all ASR size fractions. The application of AFP greatly increased the metal recovery from the ASR leachate even at very low metal concentrations in the solution. The metal recovery was maximized within step 3 of AFP. Therefore, the present study results suggest that further precipitation steps are not required for metal recovery by AFP. The recovery of heavy metals from ASR is important for conserving metals and reducing the potential toxicity of ASR prior to landfill disposal.

Table 3 Heavy metals concentration in ASR, leaching filtrate and leaching efficiencies. Metals

Fe

Cr

Pb

Zn

Ni

Cd

Metal contents in ASR (mg/kg) Metal contents in filtrate (mg/L) Metal leaching efficiencies (%)

35091 ± 943 205.0 ± 2.32 29.22 ± 0.33

168.0 ± 5.0 0.69 ± 0.03 20.36 ± 1.02

1550 ± 30 27.52 ± 0.06 88.78 ± 0.21

4823 ± 70.6 91.13 ± 0.53 94.46 ± 0.54

266 ± 24 1.58 ± 0.01 29.67 ± 0.21

63.0 ± 9.0 0.45 ± 0.01 35.70 ± 0.80

Table 4 Heavy metals recovery from the leaching filtrate. Steps precipitation pH

Recovery efficiency (%)

Recovery efficiency (%) with advanced Fenton process

Fe Cr Pb Zn Ni Cd Fe Cr Pb Zn Ni Cd

1

2

3

4

5

2.5e3

8e9

9e10

10e11

11e12.5

94.71 65.85 41.32 79.53 10.32 75.55 99.01 88.91 62.25 91.39 40.15 83.11

± ± ± ± ± ± ± ± ± ± ± ±

0.10 2.92 0.55 0.06 0.0 2.22 0.001 1.46 1.27 0.04 0.32 2.22

95.63 90.90 98.40 96.94 75.49 88.44 99.96 94.31 98.84 99.85 92.21 90.44

± ± ± ± ± ± ± ± ± ± ± ±

0.01 0.05 0.07 0.15 0.32 0.45 0.003 0.73 0.07 0.01 0.95 1.11

97.76 91.17 98.49 96.97 81.22 88.44 99.96 96.06 98.84 99.94 95.25 91.66

± ± ± ± ± ± ± ± ± ± ± ±

0.24 0.22 0.02 0.04 0.10 0.56 0.001 0.15 0.07 0.005 0.32 1.22

97.98 91.61 98.53 98.03 82.65 89.55 99.97 96.28 98.98 99.94 95.31 92.11

± ± ± ± ± ± ± ± ± ± ± ±

0.01 0.37 0.02 0.05 0.63 0.67 0.00 0.07 0.11 0.001 0.25 0.11

98.0 92.34 98.60 98.26 83.44 89.77 99.97 96.50 99.25 99.95 95.50 93.09

6

± ± ± ± ± ± ± ± ± ± ± ±

0.001 0.7 0.09 0.20 0.10 0.45 0.00 0.29 0.02 0.002 0.06 0.20

12e14 98.24 92.41 98.64 98.57 84.04 90.33 99.97 96.79 99.62 99.96 95.62 94.11

± ± ± ± ± ± ± ± ± ± ± ±

0.24 0.15 0.10 0.06 0.32 0.33 0.01 0.15 0.02 0.004 0.06 0.78

10

J. Singh, B.-K. Lee / Journal of Environmental Management 161 (2015) 1e10

Acknowledgment This research work was funded by a National Research Foundation of Korea (NRF) grant funded by the Ministry of Science, ICT and Future Planning (2013R1A2A2A03013138). This research work was also partially supported by the Research Grant of Kwangwoon University in 2015. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.jenvman.2015.06.034. References Amir, A., Lee, W., 2011. Enhanced reductive dechlorination of tetrachloroethene by nano-sized zero valent iron with vitamin B12. Chem. Eng. J. 170, 492e497. Aydogan, S., Aras, A., Canbazoglu, M., 2005. Dissolution kinetics of sphalerite in acidic ferric chloride leaching. Chem. Eng. J. 114, 67e72. Badmus, M.A.O., Audu, T.O.K., Anyata, B.U., 2007. Removal of heavy metal from industrial wastewater using hydrogen peroxide. Afr. J. Biotechnol. 6 (3), 238e242. Bas, A.D., Deveci, H., Yazici, E.Y., 2014. Treatment of manufacturing scrap TV boards by nitric acid leaching. Sep. Purif. Technol. 130, 151e159. Benjamin, M.M., 1983. Adsorption and surface precipitation of metals on amorphous iron oxyhydroxide. Environ. Sci. Technol. 17 (11), 686e692. Chen, J., Qiu, X., Fang, Z., Yang, M., Pokeung, T., Gu, F., Cheng, W., Lan, B., 2012. Removal mechanism of antibiotic metronidazole from aquatic solutions by using nanoscale zero-valent iron particles. Chem. Eng. J. 181e182, 113e119. Chen, T., Yan, B., Lei, C., Xiao, X., 2014. Pollution control and metal resource recovery for acid mine drainage. Hydrometallurgy 147e148, 112e119. Deng, J., Feng, X., Qiu, X., 2009. Extraction of heavy metal from sewage sludge using ultrasound-assisted nitric acid. Chem. Eng. J. 152, 177e182. Donaja, P., Yanga, W., Błasiaka, W., Forsgren, C., 2010. Recycling of automobile shredder residue with a microwave pyrolysis combined with high temperature steam gasification. J. Hazard. Mater. 182, 80e89. Fang, Z.Q., Qiu, X.H., Chen, J.H., Qiu, X.Q., 2011. Debromination of polybrominated diphenyl ethers by Ni/Fe bimetallic nanoparticles: influencing factors, kinetics, and mechanism. J. Hazard. Mater. 185, 958e969. Feng, D., Aldrich, C., Tan, H., 2000. Treatment of acid mine water by use of heavy metal precipitation and ion exchange. Min. Eng. 13, 623e642.  , F., 2014. Extraction of Ferella, F., Michelis, I.D., Scocchera, A., Pelino, M., Veglio metals from automotive shredder residue: preliminary results of different leaching systems. Chin. J. Chem. Eng. 23, 417e424. Fu, F.L., Xie, L.P., Tang, B., Wang, Q., Jiang, S.X., 2012. Application of a novel strategydadvanced fenton-chemical precipitation to the treatment of strong stability chelated heavy metal containing wastewater. Chem. Eng. J. 189e190,283e287. Gharabaghi, M., Irannajad, M., Azadmehr, A.R., 2013. Leaching kinetics of nickel extraction from hazardous waste by sulphuric acid and optimization dissolution conditions. Chem. Eng. Res. Des. 91, 325e331. Gheju, M., Pode, R., Manea, F., 2011. Comparative heavy metal chemical extraction from anaerobically digested biosolids. Hydrometallurgy 108, 115e121. Granata, G., Moscardini, E., Furlani, G., Pagnanelli, F., Toro, L., 2011. Automobile shredded residue valorization by hydrometallurgical metal recovery. J. Hazard. Mater. 185 (1), 44e48. He, Y., Gao, J.-F., Feng, F.Q., Liu, C., Peng, Y.-Z., Wang, S.-Y., 2012. The comparative study on the rapid decolorization of azo, anthraquinone and triphenylmethane dyes by zero-valent iron. Chem. Eng. J. 179, 8e18. Hong, K.J., Tokunaga, S., Kajiuchi, T., 2000. Extraction of heavy metals from MSW

incinerator fly ashes by chelating agents. J. Hazard. Mater. B75, 57e73. Joung, H.T., Seo, Y.C., Kim, K.H., Hong, J.H., Yoo, T.W., 2007. Distribution and characteristics of pyrolysis products from automobile shredder residue using an experimental semi-batch reactor. Korean J. Chem. Eng. 24, 996e1002. Kassaee, M.Z., Motamedi, E., Mikhak, A., Rahnemaie, R., 2011. Nitrate removal from water using iron nanoparticles produced by arc discharge vs. reduction. Chem. Eng. J. 166, 490e495. Kubonova, L., Langova, S., Nowak, B., Winter, F., 2013. Thermal and hydrometallurgical recovery methods of heavy metals from municipal solid waste fly ash. Waste Manag 33, 2322e2327. Kurose, K., Okuda, T., Nishijima, W., Okada, M., 2006. Heavy metals removal from automobile shredder residues (ASR). J. Hazard. Mater. B137, 1618e1623. Le, H.L., Jeong, J., Lee, J.C., Pandey, B.D., Yoo, J.M., Huyunh, T.H., 2011. Hydrometallurgical process for copper recovery from waste printed circuit boards (PCBs). Min. Process. Extr. Metall. Rev. Int. J. 32, 90e104. Li, X., Zhang, M., Liu, Y., Li, X., Liu, Y., Hua, R., He, C., 2013. Removal of U (VI) in aqueous solution by nanoscale zero-valent iron (nZVI). Water Qual. Expo. Health 5, 31e40. Petala, E., Dimos, K., Douvalis, A., Bakas, T., Tucek, J., Zboril, R., 2013. Nanoscale zerovalent iron supported on mesoporous silica: characterization and reactivity for Cr (VI) removal from aqueous solution. J. Hazard. Mater. 261, 295e306. Quina, M.J., Bordado, J.C., Quinta-Ferreira, R.M., 2009. The influence of pH on the leaching behaviour of inorganic components from municipal solid waste incineration residues. Waste Manag 29, 2483e2493. Sakultung, S., Pruksathorn, K., Hunsom, M., 2007. Simultaneous recovery of valuable metals from spent mobile phone battery by an acid leaching process. Korean J. Chem. Eng. 24, 272e277. Shibayama, A., Otomo, T., Takasaki, Y., Cao, Y., Murakami, T., Watanab, K., Inoue, H., 2006. Separation and recovery of valuable metals from automobile shredder residue. Int. J. Soc. Mater. Eng. Resour. (ASR) Fly. Ash Wet Process. 13 (2), 54e59. Shih, Y.H., Chungyu, H., Yuhfan, S., 2011. Reduction of hexachlorobenzene by nanoscale zero-valent iron: kinetics, pH effect, and degradation mechanism. Sep. Purif. Technol. 76, 268e274. Singh, J., Kalamdhad, A.S., 2012. Concentration and speciation of heavy metals during water hyacinth composting. Bioresour. Technol. 124, 169e179. Singh, J., Kalamdhad, A.S., 2013. Assessment of bioavailability and leachability of heavy metals during rotary drum composting of green waste (Water hyacinth). Ecol. Eng. 52, 59e69. Sthiannopkao, S., Sreesai, S., 2009. Utilization of pulp and paper industrial wastes to remove heavy metals from metal finishing wastewater. J. Environ. Manag 90, 3283e3289. Szymczycha-Madeja, A., 2011. Kinetics of Mo, Ni, V and Al leaching from spent hydrodesulphurization catalyst in a solution containing oxalic acid and hydrogen peroxide. J. Hazard Mater. 186, 2157e2161. Tuncuk, A., Stazi, V., Akcil, A., Yazici, E.Y., Deveci, H., 2012. Aqueous metal recovery techniques from e-scrap: hydrometallurgy in recycling. Min. Eng. 25, 28e37. US Environmental Protection Agency Method 1311, July 1992. Toxicity characteristic leaching procedure (TCLP), p. 35. Vermeulen, I., Van Caneghem, J., Block, C., Baeyens, J., Vandecasteele, C., 2011. Automotive shredder residue (ASR): reviewing its production from end-of-life vehicles (ELVs) and its recycling, energy or chemicals' valorisation. J. Hazard. Mater. 190, 8e27. Wang, K.S., Chiang, K.Y., Lin, K.L., Sun, C.J., 2001. Effects of a water-extraction process on heavy metal behavior in municipal solid waste incinerator fly ash. Hydrometallurgy 62, 73e81. Yan, Z., Jianguo, J., Maozhe, C., 2008. MINTEQ modeling for evaluating the leaching behavior of heavy metals in MSWI fly ash. J. Environ. Sci. 20, 1398e1402. Zhang, H., He, P.-J., Shao, L.-M., Li, X.-J., 2008. Leaching behavior of heavy metals from municipal solid waste incineration bottom ash and its geochemical modelling. J. Mater. Cycles Waste Manag 10, 7e13. Zhang, X., Lin, Y., Shan, X., Chen, Z., 2010. Degradation of 2,4,6-trinitrotoluene (TNT) from explosive wastewater using nanoscale zero-valent iron. Chem. Eng. J. 158, 566e570.