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Bioleaching of electronic waste using acidophilic sulfur oxidising bacteria
Journal of Cleaner Production 65 (2014) 465e472
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Journal of Cleaner Production journal homepage: www.elsevier.com/locate/jclepro
Bioleaching of electronic waste using acidophilic sulfur oxidising bacteria Y. Hong*, M. Valix The University of Sydney, Chemical and Biomolecular Engineering, Darlington, Sydney, NSW 2006, Australia
a r t i c l e i n f o
a b s t r a c t
Article history: Received 1 May 2013 Received in revised form 27 July 2013 Accepted 31 August 2013 Available online 12 September 2013
Bioleaching of copper-rich electronic waste material was carried out using mesophillic Acidithiobacillus thiooxidans. The leaching behaviour of copper was investigated using three methods: abiotic chemical leaching using inorganic sulphuric acid, indirect leaching using bacterially generated sulphuric acid and by direct leaching using the acidophilic bacteria. The yield of bacterially generated sulphuric acid used for both indirect and direct leaching was 14.9 g/dm3 grown in a medium containing 25 g/dm3 of elemental sulphur and basalt salts for 14 days at 30 C. This acid was diluted to achieve various pHs for the leaching tests. The variables tested were solution pH, temperature, time, pulp density and copper concentration in the waste. The results indicated that copper dissolution is influenced by passivation and galvanic coupling, both of which reduced the Cu yield and resulted in slower leaching. Increasing the acid concentration, copper concentration in the waste, higher temperature and prolonged leaching favoured higher yields and higher copper selectivity. e-waste toxicity had little effect on direct bioleaching at the pulp density used (10 g/dm3). However the growth medium containing partially oxidised sulphide compounds promoted copper surface passivation resulting in lower Cu recovery (60%) relative to abiotic leaching (98%). Ó 2013 Elsevier Ltd. All rights reserved.
Keywords: Copper e-waste Acidithiobacillus thiooxidans Passivation Galvanic coupling
1. Introduction The continual and unprecedented consumer demand for the latest gadgets and devices has led to a technological boom that has brought huge technical benefits to society; creating jobs, wealth and generally increasing living standards across the globe. However, in the process large amounts of electric and electronic wastes (e-wastes) have been generated. The problem with e-waste is its growing volume, its toxicity and its content of valuable resources (e.g., gold, copper) which are lost when e-waste is disposed (Zheng et al., 2013). The challenge in managing e-waste will be in developing sustainable recycling technologies that are able to address the volume and complexity of this waste using cost effective and ecologically sensitive methods. Many articles are widely available in literature giving reviews of the metallurgical recovery of metals from electronic wastes, discussing in detail the pyrometallurgical, hydrometallurgical and bioprocessing of electronic wastes (Cui and Zhang, 2008; Lee and Pandey, 2012; Tuncuk et al., 2012) and hybrid technologies
* Corresponding author. Tel.: þ 61 2 9351 4995; fax: þ61 2 9351 2854. E-mail addresses: [email protected] (Y. Hong), marjorie.valix@ sydney.edu.au (M. Valix). 0959-6526/$ e see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.jclepro.2013.08.043
combining both hydrometallurgical and bioprocessing methods (Pant et al., 2012). It is apparent that despite significant work that has been achieved in this area, these processes continue to be challenged by economic, technical and ecological issues. Although the technologies for recycling of relatively pure forms of metals including steel, aluminum, copper, zinc, lead and nickel are well established, recycling metals from e-wastes is greatly more challenging. There is often reluctance by smelters and scrap metal buyers to recycle e-waste. This is attributed to complexity, composition and the presence of toxic materials in the waste. Only a limited number of thermal processes are available for formal processing of the metallic components of e-waste worldwide. This include Aurubis smelter in Germany (Kahhat and Williams, 2009), Noranda copper smelting in Quebec, Canada and Ronnskar smelter in Sweden (Cui and Zhang, 2008). These pyrometallurgical processes are challenged in upgrading the final metal products because of the nature of the metallic feedstock, which in e-wastes consist of both pure metals and alloys. Metals that are in their pure forms are easily processed by melting in smelters (Reck and Graedel, 2012). However the stability of the alloys and their separation behavior during melting makes the recovery of the pure metals by thermochemical methods more energy intensive and relatively difficult or impossible (Nakajima et al., 2010). Upgrading of the final metals often necessitates the subsequent use of hydrometallurgical
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processing (Cui and Zhang, 2008). In addition smelting of electronic waste can generate dioxin from the plastic components of e-wastes (Kahhat and Williams, 2009). Heavy investments by thermal processes are therefore required to manage gaseous emissions resulting from e-waste re-processing. Unfortunately an alternative route for processing the metallic components of e-waste is through informal or un-regulated processes. Often conducted in open burning sites (Chi et al., 2011; Labunsk et al., 2013; Tsydenova and Bengtsson, 2011) these emit large quantities of toxic heavy metals (Hg, Zn) and polychlorinated dibenzodioxins and dibenzofurans resulting in both unprecedented and long term ecological and health problems (Cui and Zhang, 2008; Kahhat and Williams, 2009; Tsydenova and Bengtsson, 2011). Interest, on the other hand in using hydrometallurgical methods to process e-waste has increased because of its simplicity, the exactness of the technology and the requirement of lower capital investment and operating costs in comparison to pyrometallurgical methods (Oishi et al., 2007). Hydrometallurgical methods involve the dissolution of the metallic fractions of e-wastes in either acidic or alkali solutions, the nature of which has been driven by the recovery of precious metals (Cui and Zhang, 2008). Leaching reagents that have been used include cyanide, thiourea, thiosulfate and halide solutions (Cui and Zhang, 2008; Tuncuk et al., 2012; Zhang et al., 2012). Leaching is followed by conventional metal recovery methods including precipitation, solvent extraction, adsorption and ion-exchange and electrowinning (Coman et al., 2013; Cui and Zhang, 2008; Rimaszeki et al., 2012; Robotin et al., 2012). Although the technical feasibility of using hydrometallurgical routes have been proven, the economics of processing and the environmental impact, in particular of using toxic reagents such as cyanide and thiourea, continue to be an issue (Cui and Zhang, 2008; Tuncuk et al., 2012). In the light of these concerns, the search for simple, cost effective and environmentally friendly method for reclaiming value added materials and energy from e-waste is an urgent goal. Bioleaching is a promising technology that uses naturally occurring biological micro-organisms and their metabolic products in extracting valuable metals from the waste such spent catalyst and electronic waste (Brandl et al., 2008; Ivanus, 2010; Mishra et al., 2008; Valix et al., 2001). The wide range of waste treated by bioleaching is a clear indication of the ease of implementation of the process (Mishra et al., 2009). The increase in interest in the use of biohydrometallurgical route for re-processing wastes is driven by the fact that this method is environmentally sound with a huge potential to lower operational cost and energy requirements. Chemolithoautotrophs bacteria (e.g., Acidithiobacillus ferrooxidans and Acidithiobacillus thiooxidans), which uses CO2 as carbon source and inorganic compounds (Fe2þ, reduced S) as an energy source, has been the most widely considered group of microorganism in terms of bioleaching applications due to their ability to facilitate metal dissolution through a series of biooxidation and bioleaching reactions (Brandl et al., 2001). Other organisms including thermophiles Sulfobacillus thermosulfidooxidans and Bacillus stearothermophilus and Metallosphera sedula and heterotrophic fungi including Aspergillus niger and Penicillium Simplicissimum and Cyanobacterium violaceum have also been used to effectively dissolve various metallic fractions from e-wastes (Brandl et al., 2001, 2008; Ilyas et al., 2010). The numerous studies that have been devoted in the use of microorganisms in leaching e-wastes (Brandl et al., 2001; Ilyas et al., 2010; Pradhan et al., 2010) primarily feature metal recoveries and the efficacy of the various organisms in the mobilisation of the metals, however very few studies examined the mechanisms that influences the metal solubilisation. Biological leaching is thought to occur by abiotic or chemical dissolution
Table 1 Element composition of the Cu rich waste. Metals
Al
Cu
Fe
Mg
Pb
Sn
Zn
(wt%)
0.25
86.63
0.063
0.028
0.026
0.029
0.167
where protons carry out metal solubilisation by biooxidation and leaching reaction. However secondary reactions including metal adsorption, precipitation and passivation can hamper metal solubilisation (Valix et al., 2001) resulting in both poor metal yields and slow leaching rates. The complex composition of the wastes and the components of growth media for organisms can make the leaching of e-wastes prone to these reactions (Sasaki, 2011). The objective of this paper was to establish the factors that influence copper mobilisation bioleached indirectly and directly from ewaste. 2. Material and methods 2.1. e-waste material Ground electronic wastes (copper-rich) were obtained from Total Union PCB Recycle in Hong Kong. Metallic components of the PCB were separated by crushing, milling and magnetic separation. The waste from the recycling company was used as received with minor size separation. After undergoing sieving, waste fraction with a particle size range of 40e104 mm was collected and used throughout all experiments. The elemental composition of this waste is summarised in Table 1. A photograph of the waste is shown in Fig. 1. As shown the waste consists primarily of metallic components consisting of copper with minor quantities of plastic. 2.2. Bioleaching A pure culture microorganism, Acidithiobacillus thiooxidans (ATCC8085) was used in this study. The acidophilic bacteria was grown in 25 g of elemental S per litre of basalt salts at 30 C for 14 days, after which the acid was harvested and used for in-direct bioleaching tests. This yielded 14.9 g/dm3 of sulphuric acid. For direct leaching, sterilised waste was added directly to the growing bacteria. The in-direct bioleaching and abiotic leaching with sulphuric acid tests were carried out in a series of temperature controlled 50 ml Variomag batch reactors at temperature from 30 to 90 C with a pulp density of 10e100 g/dm3 for periods of 0.5e 24 h. Metal recoveries were estimated from the ratio of dissolved metals to the original metal content of the waste. Because of the heterogeneity of the waste, the original metal content the waste was determined by summing the dissolved metal and metal present from the leaching residue. The metal content in the residues
Fig. 1. Photograph of the Cu-rich waste.
Y. Hong, M. Valix / Journal of Cleaner Production 65 (2014) 465e472
were determined by aqua regia digestion (HCl and HNO3, ratio 3:1) of the waste (Meng and Zhang, 2012); the dissolved metals were analysed by Varian Vista AX CCD Inductively Coupled Plasma Atomic Emission Spectrometer (ICP-AES) using standard procedures. The pH and oxygen reduction potential of the solutions were monitored during leaching with a dedicated pH-mV-temp meter (TPS WP-80D). The chemical structure and morphology of the copper surface was examined after leaching using XPS (Thermo ESCALAB250i) and SEM-EDS Phillips XL 30CP. 3. Results and discussion Bioleaching of metals from the non-sulphide wastes such as ewaste may involve indirect leaching mechanism by the biogenic H2SO4 reagent. Where the role of the organism, A. thiooxidans, in this process is in catalyzing the oxidation of elemental sulfur by dissolved oxygen to sulfuric acid (Sand et al., 2001)
þ
S þ 1:5O2 þ H2 O/2H þ
SO2 4
(1)
To decouple the microbiological activity from the process of leaching, copper from e-wastes was leached by three methods: 1) chemical or abiotic, 2) in-direct or spent acid leaching and by direct leaching. The chemical method involved using sulphuric acid reagent, spent acid leaching involved using biogenic sulphuric acid that has been harvested from the fermentation of elemental S nutrient and direct leaching involved the addition of waste in a growing culture. 3.1. Abiotic chemical leaching Chemical or abiotic leaching was carried out using sulphuric acid reagent to examine the factors influencing Cu yield and selectivity. Variables that were examined were acid pH, temperature, period of leaching and pulp density. The effect of acid concentration on the rate of copper dissolution is shown in Fig. 2. As shown increasing the acid concentration increased the rate of copper dissolution. This is consistent with all bioleaching efforts of metallic wastes using acidophiles (Brandl et al., 2008; Brandl et al., 2001; Ivanus and Ivanus, 2009). The rate of leaching was characterised by slow Cu dissolution, followed by rapid then again a slower Cu dissolution. At pH 0.5, almost full Cu recovery was achieved. However above this pH, the Cu dissolution was characterised by the slower Cu dissolution and reduced Cu yields. The nature of the pH effect was examined in this study to establish if other factors other than the acid concentration affected copper dissolution.
Fig. 2. Recovery of Cu from Cu-rich waste by biogenic sulphuric acid at various pH.
467
Copper dissolution will proceed by two stage redox reactions (see Equations (2) and (3)). The relative extent of each reaction is determined by pH, the nature of anions that are present, temperature, and dissolved oxygen concentration. Here reduced copper metal is first oxidised rapidly to cuprous ion (Cuþ) with an E ¼ 0.52V followed by its subsequent slower oxidation to cupric ion (Cu2þ) with E ¼ 0.16V (Mattsson and Bockris, 1959; Andersen et al., 1975).
Cu/Cuþ þ e
(2)
Cuþ /Cu2þ þ e
(3)
The corresponding cathodic reaction to support the oxidation of copper with E ¼ 1.23 V is:
O2 þ 4Hþ þ 4e 42H2 O
(4)
The resulting reaction is (Nandeesh and Sheshadri, 1988):
2Cuþ þ O2 þ 4Hþ /4Cu2þ þ 2H2 O Copper ions then react with
Cu2þ þ SO2 4 /CuSO4
SO2 4
(5) ion to form a soluble salt:
(6)
The electrochemical reactions 2 to 5 should proceed spontaneously because the cell potential (Ecathode Eanode) is positive. However as shown in Fig. 2, this process only occurs at the very high acid concentrations. At the lower acid concentration, the reaction ceases leading to less than full Cu recovery inferring possible surface passivation. This is confirmed by the formation cuprous and cupric ions on the copper surface established by XPS analysis of the copper particles after leaching (see Fig. 3). In the present work, the Cu(2p)3/2 spectrum was fitted to four components. The component at binding energy 932.5 eV represents the Cu(I) state. The three peaks located at 935, 939.5 and 943.5 eV are due to Cu(II) species (Cano et al., 2001). Further analysis of the surface with SEM-EDS spot analysis in Fig. 4 shows the copper surface is covered by two specific copper species. It can be inferred from the atomic ratios of the elemental analysis in Table 2 that the copper surface is primarily covered with Cu2O (spectrum 1) with deposits of CuSO4 precipitate (spectrum 2 and 3) (Table 3).
Fig. 3. Curve-fitting results for the Cu2þ/Cuþ XPS spectrum of Cu-rich waste-after leaching.
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Fig. 4. SEM micrograph showing the surface morphology of the copper particle leached in H2SO4, pH 1.5 for 6 h at 90 C and EDS analysis of spectrums 1, 2 and 3.
Examination of the electrochemical reactions also sheds some light into the nature of this passivation. The first stage oxidation of copper has been suggested to occur as follows (Schweickert et al., 1976):
Cu/Cuþ ads
þe
/Cuþ sol
þe
(7)
Because the rate determining step in the dissolution of copper is the oxidation of cuprous to cupric ions (Mattsson and Bockris, 1959), this infers that cuprous ions, at least at the beginning of the leaching process, are present in large concentrations in the vicinity of the dissolving copper particles. Cuprous ions are considered thermodynamically unstable (Chen and Dutrizac, 1989) and their fate, that is they can either be displaced or disproportionate, determines the recovery of copper. The work by Liu et al. (2010) suggests that one fate would be the conversion of Cuþ to Cu2O. This begins with the electrochemical reaction of molecular dissolved oxygen with E ¼ 0.4 V:
This leads to the precipitation (see Equation (9)) and dehydration of the precipitate to cuprous oxide (see Equation (10)).
Cuþ þ OH /CuOH
(9)
2CuOH/Cu2 O þ H2 O
(10)
An alternative process for the oxidation of the cuprous ion has been postulated to occur through the peroxide route. According to (Andersen et al., 1975) the oxidation reaction in Equation (5) can also proceed as follows:
2Cuþ þ O2 þ 2Hþ /2Cu2þ þ H2 O2
(11)
(8)
The presence of H2O2 as suggested by the work of (Stewart and Gewirth, 2007) can also lead to the formation of CueOH. This in turn could also induce the formation of Cu2O as suggested by Equation (10). These results suggest passivation of the copper surface during bioleaching occur principally from the formation of
Table 2 EDS-elemental analysis of copper particle leached in H2SO4, pH 1.5 for 6 h at 90 C (see Fig. 4).
Table 3 EDS-elemental analysis of copper particle leached in biogenic sulphuric acid, pH 1.0 for 6 h at 90 C (see Fig. 12).
O2 þ 2H2 O þ 4e /4OH
Elements Spectrum 1 (atomic %) Spectrum 2 (atomic %) Spectrum 3 (atomic %)
Elements
Spectrum 1 (atomic %)
Spectrum 4 (atomic %)
O Al S Cu
O Al S Cu
4.51 0.68 41.55 53.26
30.35 Neg 0.23 68.22
26.93 0.33 0.38 72.16
66.39 0.29 7.89 25.43
48.32 0.58 7.91 43.19
Y. Hong, M. Valix / Journal of Cleaner Production 65 (2014) 465e472
oxidised cuprous ions and support the SEM-EDS analysis in Fig. 4. Cupric ions (Cu2þ) on the copper surface are present as deposits of CuSO4 precipitate. These may foul the surface but would not have similar passivating effects as Cu2O. These results suggest that competitive reactions of the Cuþ disproportionation to Cu2þ and reaction with sulphate or hydroxyl ions in the vicinity of the copper ions dictate if leaching or passivation will proceed. This is consistent with the observed results in Fig. 2 that infer the higher acidity at the lower pH (pH < 1.0) is able to avoid passivation by neutralising the OHions leading to improved leaching. Whilst those conducted at the higher pH (>1.0) are more susceptible to passivation effects. The implication of this is the need to maintain low pH in bioleaching efforts. Another feature of e-waste that could influence metal recovery is the galvanic coupling of Cu with the other metallic components in the waste. This process, which is also referred to as cementation involves the electrochemical interaction of metals where metals with more electropositive electrode reduction potential have a tendency to be reduced, whilst those with lower standard potential are oxidised and are preferentially dissolved (Aktas, 2012; Demirkiran et al., 2007). The ranking of electrode potential at 25 C among the metals in the waste studied is: Cuþ/Cu (0.52 V) > Cu2þ/Cu (0.34 V) > Zn2þ/Zn (0.76 V) > Al3þ/Al (1.67 V) > Mg2þ/Mg (2.35 V). The presence therefore of metals with lower electrode potentials has the potential to limit the selective leaching of copper from the wastes. Recent studies of metal cementation reveal that in addition to standard potential various factors including solution temperature, pH, concentration, presence of dissolved oxygen, leaching reagents and time can also influence the selective removal or preferential dissolution of metals (Keles, 2009; Sulka and Jaskula, 2003, 2004). The effects of pH, temperature and time on the galvanic coupling of metals were considered in this study. The effect of acid pH on the relative metal mobilisation in Fig. 5 showed that at the higher pH (pH 1.4 and 1.7) the selective dissolution of the metals followed the ORP ranking. Metals with the least ORP, for example Mg, were preferentially dissolved in comparison to Cu. However the lower pH (<1.0) appears to overcome this effect leading to the selective dissolution of Cu. Previous studies of cementation have also demonstrated that lowering pH can retard the displacement of the metal of higher reduction potential and in effect promoted their dissolution (Demirkiran et al. 2007, Sulka and Jaskula 2003, Lamya and Lorenzen 2005).
Fig. 5. Metal leaching as a function of sulphuric acid pH (10 g/dm3, 90 , 24 h).
469
The effect of increasing temperature in overcoming the galvanic interaction effect on Cu dissolution as shown in Fig. 6, is inconsistent with the observed effect of temperature in cementation studies which promoted the displacement of copper in the presence of metals with lower standard potentials (Demirkiran et al., 2007; Lamya and Lorenzen, 2005). In previous studies however only two metals were coupled, whereas in e-waste several metals are involved. Similarly the effect of increasing the period of leaching in overcoming galvanic interaction, as shown in Fig. 7, is again inconsistent with observed effect of time in cementation of copper. It would be anticipated that longer periods of leaching would leach the less electropositive metals first eventually allowing the dissolution of the remaining electronegative metals. This appears true for Al and Zn, however despite only dissolving about 40% of Mg, about 99% of Cu was dissolved in 24 h. These results show the galvanic effects in leaching e-wastes used in this study do not follow conventional electrochemical
Fig. 6. Metal leaching in sulphuric acid as a function of temperature (10 g/dm3, pH 1.0, 24 h).
Fig. 7. Metal leaching in sulphuric acid as a function of time (10g/dm3, pH 1.0, 90 C).
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rules. Fig. 8 show that the percentage of copper compared to the remaining metal constituents of the waste has a significant impact on effect of galvanic interactions during leaching. At 89e 96% Cu, the effect of galvanic interaction does not appear to have a significant impact on Cu dissolution. However at 12.9% Cu, the effect of galvanic interaction on copper recovery in 24 h of leaching is significant in reducing copper recovery. The effect of copper concentration on selective leaching of copper in a mixed metal environment is consistent with observations made by Lamya and Lorenzen (2005). These results demonstrate the potential challenge in selective leaching of metallic wastes with variable compositions of metal by hydrometallurgical and biohydro-metallurgical routes. The effect of increasing pulp density on Cu recovery is shown in Fig. 9. This effect has often been attributed to the toxic effect of the waste on the bioleaching organisms (Ilyas et al., 2010; Pradhan et al., 2010). These abiotic tests suggest that other factors must contribute to the reduced Cu recovery as pulp density is increased. This also infers the potential challenge in implementing the use of very low pulp density in commercial application of this process.
3.2. Indirect and direct leaching To validate the effects observed under chemical leaching, direct leaching of waste with a growing organism was conducted and compared to biogenic acid or indirect leaching and abiotic leaching with sulphuric acid. Direct leaching was performed with 10 g/dm3 of waste with Cu-rich waste being added to the microbe culture once the pH reached 1.0. For indirect leaching, the biogenic acid was harvested and used in leaching the waste. Fig. 10 shows there is very little difference in the Cu mobilisation in both direct and indirect leaching at 30 C. It would appear that at this concentration (10 g/dm3), the toxic effects of the waste have little effect on the organism metabolism and would be consistent with the reported e-waste threshold for acidophilic bacteria (Brandl et al., 2001). The effects of abiotic and indirect leaching at the elevated temperature are compared in Fig. 11. It is apparent that at the early phase, there is little difference in the Cu mobilisation. However prolonged leaching resulted in lower Cu dissolution when biogenic acid was used in indirect leaching. Examination of the copper surface after leaching for 6 h (see Fig. 12) and the corresponding elemental analysis on the surface (see Table 2) reveal the formation of CuS (spectrum 1) and Cu2O (spectrum 4) on the surface. Spectrums 2, 3, and 5 were of the plastic components. We proposed that the reduction in copper mobilisation with time in biogenic acid results from the passivation of the Cu surface with polythionates and So of incomplete oxidation products of sulphide compounds that are present in the biogenic acid (Sasaki, 2011) and from cuprous oxidation.
Fig. 8. Effect of copper content of e-waste on Cu leaching in sulphuric acid as a function of pH (10g/dm3, 24 h, 90 C).
Fig. 10. Direct and abiotic leaching at 30 C (pH 1.0, 10 g/dm3).
Fig. 9. Metal leaching in sulphuric acid as a function of pulp density (pH 1.0, 90 C, 24 h).
Fig. 11. Indirect and abiotic leaching at 90 C (pH 1.0, 10 g/dm3).
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Fig. 12. SEM micrograph showing the surface morphology of the copper particle leached in biogenic acid, pH 1.0 for 6 h at 90 C and EDS analysis of spectrums 1 and 4.
4. Conclusions
Acknowledgement
Almost complete mobilisation of Cu can be achieved using Acidithiobacillus thiooxidans providing evidence for the efficacy of acidophiles in the recovery of metals from e-wastes. However this was achieved within a window of conditions reflecting the various challenges in bioprocessing of e-wastes. Chemical leaching of copper rich waste has been demonstrated to be affected by passivation and galvanic coupling. Passivation of copper with chemical leaching appears to result from the oxidation of cuprous ion promoted by oxygen and peroxide dissociation. Passivation by copper oxidation and galvanic coupling effects can be overcome in wastes containing a high percentage of Cu (>90%) by using low pH (<1.0), high temperature (90 C), low pulp density (10 g/dm3) and long periods of leaching. The effect of galvanic coupling however is prevalent and is more difficult to control as the percentage of copper decreases. Leaching using biogenic acid or of direct leaching with the micro-organisms has added complexities. It is apparent that at 10 g/dm3 of waste do not pose any toxic effects on the A. thiooxidans and is the optimal pulp density to achieve maximum copper dissolution. However the presence of incompletely oxidised sulphide compounds and the sulphates leads to the formation of CuS and copper sulphate precipitates. Both of which appear to contribute to copper passivation.
This research was supported under Australian Research Council Discovery Project Scheme (DP1096342).
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Pant, D., Joshi, D., Upreti, M.K., Kotnala, R.K., 2012. Chemical and biological extraction of metals present in e waste: a hybrid technology. Waste Manag. 32, 979e990. Pradhan, D., Kim, D.J., Ahn, J.G., Chaudhury, G.R., Lee, S.W., 2010. Kinetics and statistical behavior of metals dissolution from spent petroleum catalyst using acidophilic iron oxidizing bacteria. J. Industrial Eng. Chem. 16, 866e871. Reck, B.K., Graedel, T.E., 2012. Challenges in metal recycling. Science 337, 690e695. Rimaszeki, G., Kulcsar, T., Kekesi, T., 2012. Application of HCl solutions for recovering the high purity metal from tin scrap by electrorefining. Hydrometallurgy 125, 55e63. Robotin, B., Coman, V., Ilea, P., 2012. Nickel recovery from electronic waste III iron nickel separation. Stud. Univ. Babes-bolyai Chem. 57, 81e90. Sand, W., Gehrke, T., Jozsa, P.G., Schippers, A., 2001. (Bio) chemistry of bacterial leaching e direct vs. indirect bioleaching. Hydrometallurgy 59, 159e175. Sasaki, K., 2011. Characterization of passivation layers in bioleaching of Sulfides. Bunseki Kagaku 60, 911e919. Schweickert, H., Elmiligy, A.A., Melendez, A., Lorenz, W.J., 1976. Mathematical treatment of galvanostatic double pulse method applied to consecutive chargetransfer reactions. J. Electroanal. Chem. 68, 19e30. Stewart, K.L., Gewirth, A.A., 2007. Mechanism of electrochemical reduction of hydrogen peroxide on copper in acidic sulfate solutions. Langmuir 23, 9911e 9918. Sulka, G.D., Jaskula, M., 2003. Study of the kinetics of silver ions cementation onto copper from sulphuric acid solution. Hydrometallurgy 70, 185e196. Sulka, G.D., Jaskula, M., 2004. Study of the mechanism of silver ions cementation onto copper from acidic sulphate solutions and the morphology of the silver deposit. Hydrometallurgy 72, 93e110. Tsydenova, O., Bengtsson, M., 2011. Chemical hazards associated with treatment of waste electrical and electronic equipment. Waste Manag. 31, 45e58. Tuncuk, A., Stazi, V., Akcil, A., Yazici, E.Y., Deveci, H., 2012. Aqueous metal recovery techniques from e-scrap: hydrometallurgy in recycling. Minerals Eng. 25, 28e37. Valix, M., Usai, F., Malik, R., 2001. The electro-sorption properties of nickel on laterite gangue leached with an organic chelating acid. Minerals Eng. 14, 205e215. Zhang, Y.H., Liu, S.L., Xie, H.H., Zeng, X.L., Li, J.H., 2012. Current status on leaching precious metals from waste printed circuit boards. In: Jinhui, L., Hualong, H. (Eds.), Seventh International Conference on Waste Management and Technology. Elsevier Science Bv, Amsterdam, pp. 560e568. Zheng, G.N., Xu, X.J., Li, B., Wu, K.S., Yekeen, T.A., Huo, X., 2013. Association between lung function in school children and exposure to three transition metals from an e-waste recycling area. J. Exposure Sci. Environ. Epidem. 23, 67e72.
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Journal of Cleaner Production journal homepage: www.elsevier.com/locate/jclepro
Bioleaching of electronic waste using acidophilic sulfur oxidising bacteria Y. Hong*, M. Valix The University of Sydney, Chemical and Biomolecular Engineering, Darlington, Sydney, NSW 2006, Australia
a r t i c l e i n f o
a b s t r a c t
Article history: Received 1 May 2013 Received in revised form 27 July 2013 Accepted 31 August 2013 Available online 12 September 2013
Bioleaching of copper-rich electronic waste material was carried out using mesophillic Acidithiobacillus thiooxidans. The leaching behaviour of copper was investigated using three methods: abiotic chemical leaching using inorganic sulphuric acid, indirect leaching using bacterially generated sulphuric acid and by direct leaching using the acidophilic bacteria. The yield of bacterially generated sulphuric acid used for both indirect and direct leaching was 14.9 g/dm3 grown in a medium containing 25 g/dm3 of elemental sulphur and basalt salts for 14 days at 30 C. This acid was diluted to achieve various pHs for the leaching tests. The variables tested were solution pH, temperature, time, pulp density and copper concentration in the waste. The results indicated that copper dissolution is influenced by passivation and galvanic coupling, both of which reduced the Cu yield and resulted in slower leaching. Increasing the acid concentration, copper concentration in the waste, higher temperature and prolonged leaching favoured higher yields and higher copper selectivity. e-waste toxicity had little effect on direct bioleaching at the pulp density used (10 g/dm3). However the growth medium containing partially oxidised sulphide compounds promoted copper surface passivation resulting in lower Cu recovery (60%) relative to abiotic leaching (98%). Ó 2013 Elsevier Ltd. All rights reserved.
Keywords: Copper e-waste Acidithiobacillus thiooxidans Passivation Galvanic coupling
1. Introduction The continual and unprecedented consumer demand for the latest gadgets and devices has led to a technological boom that has brought huge technical benefits to society; creating jobs, wealth and generally increasing living standards across the globe. However, in the process large amounts of electric and electronic wastes (e-wastes) have been generated. The problem with e-waste is its growing volume, its toxicity and its content of valuable resources (e.g., gold, copper) which are lost when e-waste is disposed (Zheng et al., 2013). The challenge in managing e-waste will be in developing sustainable recycling technologies that are able to address the volume and complexity of this waste using cost effective and ecologically sensitive methods. Many articles are widely available in literature giving reviews of the metallurgical recovery of metals from electronic wastes, discussing in detail the pyrometallurgical, hydrometallurgical and bioprocessing of electronic wastes (Cui and Zhang, 2008; Lee and Pandey, 2012; Tuncuk et al., 2012) and hybrid technologies
* Corresponding author. Tel.: þ 61 2 9351 4995; fax: þ61 2 9351 2854. E-mail addresses: [email protected] (Y. Hong), marjorie.valix@ sydney.edu.au (M. Valix). 0959-6526/$ e see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.jclepro.2013.08.043
combining both hydrometallurgical and bioprocessing methods (Pant et al., 2012). It is apparent that despite significant work that has been achieved in this area, these processes continue to be challenged by economic, technical and ecological issues. Although the technologies for recycling of relatively pure forms of metals including steel, aluminum, copper, zinc, lead and nickel are well established, recycling metals from e-wastes is greatly more challenging. There is often reluctance by smelters and scrap metal buyers to recycle e-waste. This is attributed to complexity, composition and the presence of toxic materials in the waste. Only a limited number of thermal processes are available for formal processing of the metallic components of e-waste worldwide. This include Aurubis smelter in Germany (Kahhat and Williams, 2009), Noranda copper smelting in Quebec, Canada and Ronnskar smelter in Sweden (Cui and Zhang, 2008). These pyrometallurgical processes are challenged in upgrading the final metal products because of the nature of the metallic feedstock, which in e-wastes consist of both pure metals and alloys. Metals that are in their pure forms are easily processed by melting in smelters (Reck and Graedel, 2012). However the stability of the alloys and their separation behavior during melting makes the recovery of the pure metals by thermochemical methods more energy intensive and relatively difficult or impossible (Nakajima et al., 2010). Upgrading of the final metals often necessitates the subsequent use of hydrometallurgical
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processing (Cui and Zhang, 2008). In addition smelting of electronic waste can generate dioxin from the plastic components of e-wastes (Kahhat and Williams, 2009). Heavy investments by thermal processes are therefore required to manage gaseous emissions resulting from e-waste re-processing. Unfortunately an alternative route for processing the metallic components of e-waste is through informal or un-regulated processes. Often conducted in open burning sites (Chi et al., 2011; Labunsk et al., 2013; Tsydenova and Bengtsson, 2011) these emit large quantities of toxic heavy metals (Hg, Zn) and polychlorinated dibenzodioxins and dibenzofurans resulting in both unprecedented and long term ecological and health problems (Cui and Zhang, 2008; Kahhat and Williams, 2009; Tsydenova and Bengtsson, 2011). Interest, on the other hand in using hydrometallurgical methods to process e-waste has increased because of its simplicity, the exactness of the technology and the requirement of lower capital investment and operating costs in comparison to pyrometallurgical methods (Oishi et al., 2007). Hydrometallurgical methods involve the dissolution of the metallic fractions of e-wastes in either acidic or alkali solutions, the nature of which has been driven by the recovery of precious metals (Cui and Zhang, 2008). Leaching reagents that have been used include cyanide, thiourea, thiosulfate and halide solutions (Cui and Zhang, 2008; Tuncuk et al., 2012; Zhang et al., 2012). Leaching is followed by conventional metal recovery methods including precipitation, solvent extraction, adsorption and ion-exchange and electrowinning (Coman et al., 2013; Cui and Zhang, 2008; Rimaszeki et al., 2012; Robotin et al., 2012). Although the technical feasibility of using hydrometallurgical routes have been proven, the economics of processing and the environmental impact, in particular of using toxic reagents such as cyanide and thiourea, continue to be an issue (Cui and Zhang, 2008; Tuncuk et al., 2012). In the light of these concerns, the search for simple, cost effective and environmentally friendly method for reclaiming value added materials and energy from e-waste is an urgent goal. Bioleaching is a promising technology that uses naturally occurring biological micro-organisms and their metabolic products in extracting valuable metals from the waste such spent catalyst and electronic waste (Brandl et al., 2008; Ivanus, 2010; Mishra et al., 2008; Valix et al., 2001). The wide range of waste treated by bioleaching is a clear indication of the ease of implementation of the process (Mishra et al., 2009). The increase in interest in the use of biohydrometallurgical route for re-processing wastes is driven by the fact that this method is environmentally sound with a huge potential to lower operational cost and energy requirements. Chemolithoautotrophs bacteria (e.g., Acidithiobacillus ferrooxidans and Acidithiobacillus thiooxidans), which uses CO2 as carbon source and inorganic compounds (Fe2þ, reduced S) as an energy source, has been the most widely considered group of microorganism in terms of bioleaching applications due to their ability to facilitate metal dissolution through a series of biooxidation and bioleaching reactions (Brandl et al., 2001). Other organisms including thermophiles Sulfobacillus thermosulfidooxidans and Bacillus stearothermophilus and Metallosphera sedula and heterotrophic fungi including Aspergillus niger and Penicillium Simplicissimum and Cyanobacterium violaceum have also been used to effectively dissolve various metallic fractions from e-wastes (Brandl et al., 2001, 2008; Ilyas et al., 2010). The numerous studies that have been devoted in the use of microorganisms in leaching e-wastes (Brandl et al., 2001; Ilyas et al., 2010; Pradhan et al., 2010) primarily feature metal recoveries and the efficacy of the various organisms in the mobilisation of the metals, however very few studies examined the mechanisms that influences the metal solubilisation. Biological leaching is thought to occur by abiotic or chemical dissolution
Table 1 Element composition of the Cu rich waste. Metals
Al
Cu
Fe
Mg
Pb
Sn
Zn
(wt%)
0.25
86.63
0.063
0.028
0.026
0.029
0.167
where protons carry out metal solubilisation by biooxidation and leaching reaction. However secondary reactions including metal adsorption, precipitation and passivation can hamper metal solubilisation (Valix et al., 2001) resulting in both poor metal yields and slow leaching rates. The complex composition of the wastes and the components of growth media for organisms can make the leaching of e-wastes prone to these reactions (Sasaki, 2011). The objective of this paper was to establish the factors that influence copper mobilisation bioleached indirectly and directly from ewaste. 2. Material and methods 2.1. e-waste material Ground electronic wastes (copper-rich) were obtained from Total Union PCB Recycle in Hong Kong. Metallic components of the PCB were separated by crushing, milling and magnetic separation. The waste from the recycling company was used as received with minor size separation. After undergoing sieving, waste fraction with a particle size range of 40e104 mm was collected and used throughout all experiments. The elemental composition of this waste is summarised in Table 1. A photograph of the waste is shown in Fig. 1. As shown the waste consists primarily of metallic components consisting of copper with minor quantities of plastic. 2.2. Bioleaching A pure culture microorganism, Acidithiobacillus thiooxidans (ATCC8085) was used in this study. The acidophilic bacteria was grown in 25 g of elemental S per litre of basalt salts at 30 C for 14 days, after which the acid was harvested and used for in-direct bioleaching tests. This yielded 14.9 g/dm3 of sulphuric acid. For direct leaching, sterilised waste was added directly to the growing bacteria. The in-direct bioleaching and abiotic leaching with sulphuric acid tests were carried out in a series of temperature controlled 50 ml Variomag batch reactors at temperature from 30 to 90 C with a pulp density of 10e100 g/dm3 for periods of 0.5e 24 h. Metal recoveries were estimated from the ratio of dissolved metals to the original metal content of the waste. Because of the heterogeneity of the waste, the original metal content the waste was determined by summing the dissolved metal and metal present from the leaching residue. The metal content in the residues
Fig. 1. Photograph of the Cu-rich waste.
Y. Hong, M. Valix / Journal of Cleaner Production 65 (2014) 465e472
were determined by aqua regia digestion (HCl and HNO3, ratio 3:1) of the waste (Meng and Zhang, 2012); the dissolved metals were analysed by Varian Vista AX CCD Inductively Coupled Plasma Atomic Emission Spectrometer (ICP-AES) using standard procedures. The pH and oxygen reduction potential of the solutions were monitored during leaching with a dedicated pH-mV-temp meter (TPS WP-80D). The chemical structure and morphology of the copper surface was examined after leaching using XPS (Thermo ESCALAB250i) and SEM-EDS Phillips XL 30CP. 3. Results and discussion Bioleaching of metals from the non-sulphide wastes such as ewaste may involve indirect leaching mechanism by the biogenic H2SO4 reagent. Where the role of the organism, A. thiooxidans, in this process is in catalyzing the oxidation of elemental sulfur by dissolved oxygen to sulfuric acid (Sand et al., 2001)
þ
S þ 1:5O2 þ H2 O/2H þ
SO2 4
(1)
To decouple the microbiological activity from the process of leaching, copper from e-wastes was leached by three methods: 1) chemical or abiotic, 2) in-direct or spent acid leaching and by direct leaching. The chemical method involved using sulphuric acid reagent, spent acid leaching involved using biogenic sulphuric acid that has been harvested from the fermentation of elemental S nutrient and direct leaching involved the addition of waste in a growing culture. 3.1. Abiotic chemical leaching Chemical or abiotic leaching was carried out using sulphuric acid reagent to examine the factors influencing Cu yield and selectivity. Variables that were examined were acid pH, temperature, period of leaching and pulp density. The effect of acid concentration on the rate of copper dissolution is shown in Fig. 2. As shown increasing the acid concentration increased the rate of copper dissolution. This is consistent with all bioleaching efforts of metallic wastes using acidophiles (Brandl et al., 2008; Brandl et al., 2001; Ivanus and Ivanus, 2009). The rate of leaching was characterised by slow Cu dissolution, followed by rapid then again a slower Cu dissolution. At pH 0.5, almost full Cu recovery was achieved. However above this pH, the Cu dissolution was characterised by the slower Cu dissolution and reduced Cu yields. The nature of the pH effect was examined in this study to establish if other factors other than the acid concentration affected copper dissolution.
Fig. 2. Recovery of Cu from Cu-rich waste by biogenic sulphuric acid at various pH.
467
Copper dissolution will proceed by two stage redox reactions (see Equations (2) and (3)). The relative extent of each reaction is determined by pH, the nature of anions that are present, temperature, and dissolved oxygen concentration. Here reduced copper metal is first oxidised rapidly to cuprous ion (Cuþ) with an E ¼ 0.52V followed by its subsequent slower oxidation to cupric ion (Cu2þ) with E ¼ 0.16V (Mattsson and Bockris, 1959; Andersen et al., 1975).
Cu/Cuþ þ e
(2)
Cuþ /Cu2þ þ e
(3)
The corresponding cathodic reaction to support the oxidation of copper with E ¼ 1.23 V is:
O2 þ 4Hþ þ 4e 42H2 O
(4)
The resulting reaction is (Nandeesh and Sheshadri, 1988):
2Cuþ þ O2 þ 4Hþ /4Cu2þ þ 2H2 O Copper ions then react with
Cu2þ þ SO2 4 /CuSO4
SO2 4
(5) ion to form a soluble salt:
(6)
The electrochemical reactions 2 to 5 should proceed spontaneously because the cell potential (Ecathode Eanode) is positive. However as shown in Fig. 2, this process only occurs at the very high acid concentrations. At the lower acid concentration, the reaction ceases leading to less than full Cu recovery inferring possible surface passivation. This is confirmed by the formation cuprous and cupric ions on the copper surface established by XPS analysis of the copper particles after leaching (see Fig. 3). In the present work, the Cu(2p)3/2 spectrum was fitted to four components. The component at binding energy 932.5 eV represents the Cu(I) state. The three peaks located at 935, 939.5 and 943.5 eV are due to Cu(II) species (Cano et al., 2001). Further analysis of the surface with SEM-EDS spot analysis in Fig. 4 shows the copper surface is covered by two specific copper species. It can be inferred from the atomic ratios of the elemental analysis in Table 2 that the copper surface is primarily covered with Cu2O (spectrum 1) with deposits of CuSO4 precipitate (spectrum 2 and 3) (Table 3).
Fig. 3. Curve-fitting results for the Cu2þ/Cuþ XPS spectrum of Cu-rich waste-after leaching.
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Y. Hong, M. Valix / Journal of Cleaner Production 65 (2014) 465e472
Fig. 4. SEM micrograph showing the surface morphology of the copper particle leached in H2SO4, pH 1.5 for 6 h at 90 C and EDS analysis of spectrums 1, 2 and 3.
Examination of the electrochemical reactions also sheds some light into the nature of this passivation. The first stage oxidation of copper has been suggested to occur as follows (Schweickert et al., 1976):
Cu/Cuþ ads
þe
/Cuþ sol
þe
(7)
Because the rate determining step in the dissolution of copper is the oxidation of cuprous to cupric ions (Mattsson and Bockris, 1959), this infers that cuprous ions, at least at the beginning of the leaching process, are present in large concentrations in the vicinity of the dissolving copper particles. Cuprous ions are considered thermodynamically unstable (Chen and Dutrizac, 1989) and their fate, that is they can either be displaced or disproportionate, determines the recovery of copper. The work by Liu et al. (2010) suggests that one fate would be the conversion of Cuþ to Cu2O. This begins with the electrochemical reaction of molecular dissolved oxygen with E ¼ 0.4 V:
This leads to the precipitation (see Equation (9)) and dehydration of the precipitate to cuprous oxide (see Equation (10)).
Cuþ þ OH /CuOH
(9)
2CuOH/Cu2 O þ H2 O
(10)
An alternative process for the oxidation of the cuprous ion has been postulated to occur through the peroxide route. According to (Andersen et al., 1975) the oxidation reaction in Equation (5) can also proceed as follows:
2Cuþ þ O2 þ 2Hþ /2Cu2þ þ H2 O2
(11)
(8)
The presence of H2O2 as suggested by the work of (Stewart and Gewirth, 2007) can also lead to the formation of CueOH. This in turn could also induce the formation of Cu2O as suggested by Equation (10). These results suggest passivation of the copper surface during bioleaching occur principally from the formation of
Table 2 EDS-elemental analysis of copper particle leached in H2SO4, pH 1.5 for 6 h at 90 C (see Fig. 4).
Table 3 EDS-elemental analysis of copper particle leached in biogenic sulphuric acid, pH 1.0 for 6 h at 90 C (see Fig. 12).
O2 þ 2H2 O þ 4e /4OH
Elements Spectrum 1 (atomic %) Spectrum 2 (atomic %) Spectrum 3 (atomic %)
Elements
Spectrum 1 (atomic %)
Spectrum 4 (atomic %)
O Al S Cu
O Al S Cu
4.51 0.68 41.55 53.26
30.35 Neg 0.23 68.22
26.93 0.33 0.38 72.16
66.39 0.29 7.89 25.43
48.32 0.58 7.91 43.19
Y. Hong, M. Valix / Journal of Cleaner Production 65 (2014) 465e472
oxidised cuprous ions and support the SEM-EDS analysis in Fig. 4. Cupric ions (Cu2þ) on the copper surface are present as deposits of CuSO4 precipitate. These may foul the surface but would not have similar passivating effects as Cu2O. These results suggest that competitive reactions of the Cuþ disproportionation to Cu2þ and reaction with sulphate or hydroxyl ions in the vicinity of the copper ions dictate if leaching or passivation will proceed. This is consistent with the observed results in Fig. 2 that infer the higher acidity at the lower pH (pH < 1.0) is able to avoid passivation by neutralising the OHions leading to improved leaching. Whilst those conducted at the higher pH (>1.0) are more susceptible to passivation effects. The implication of this is the need to maintain low pH in bioleaching efforts. Another feature of e-waste that could influence metal recovery is the galvanic coupling of Cu with the other metallic components in the waste. This process, which is also referred to as cementation involves the electrochemical interaction of metals where metals with more electropositive electrode reduction potential have a tendency to be reduced, whilst those with lower standard potential are oxidised and are preferentially dissolved (Aktas, 2012; Demirkiran et al., 2007). The ranking of electrode potential at 25 C among the metals in the waste studied is: Cuþ/Cu (0.52 V) > Cu2þ/Cu (0.34 V) > Zn2þ/Zn (0.76 V) > Al3þ/Al (1.67 V) > Mg2þ/Mg (2.35 V). The presence therefore of metals with lower electrode potentials has the potential to limit the selective leaching of copper from the wastes. Recent studies of metal cementation reveal that in addition to standard potential various factors including solution temperature, pH, concentration, presence of dissolved oxygen, leaching reagents and time can also influence the selective removal or preferential dissolution of metals (Keles, 2009; Sulka and Jaskula, 2003, 2004). The effects of pH, temperature and time on the galvanic coupling of metals were considered in this study. The effect of acid pH on the relative metal mobilisation in Fig. 5 showed that at the higher pH (pH 1.4 and 1.7) the selective dissolution of the metals followed the ORP ranking. Metals with the least ORP, for example Mg, were preferentially dissolved in comparison to Cu. However the lower pH (<1.0) appears to overcome this effect leading to the selective dissolution of Cu. Previous studies of cementation have also demonstrated that lowering pH can retard the displacement of the metal of higher reduction potential and in effect promoted their dissolution (Demirkiran et al. 2007, Sulka and Jaskula 2003, Lamya and Lorenzen 2005).
Fig. 5. Metal leaching as a function of sulphuric acid pH (10 g/dm3, 90 , 24 h).
469
The effect of increasing temperature in overcoming the galvanic interaction effect on Cu dissolution as shown in Fig. 6, is inconsistent with the observed effect of temperature in cementation studies which promoted the displacement of copper in the presence of metals with lower standard potentials (Demirkiran et al., 2007; Lamya and Lorenzen, 2005). In previous studies however only two metals were coupled, whereas in e-waste several metals are involved. Similarly the effect of increasing the period of leaching in overcoming galvanic interaction, as shown in Fig. 7, is again inconsistent with observed effect of time in cementation of copper. It would be anticipated that longer periods of leaching would leach the less electropositive metals first eventually allowing the dissolution of the remaining electronegative metals. This appears true for Al and Zn, however despite only dissolving about 40% of Mg, about 99% of Cu was dissolved in 24 h. These results show the galvanic effects in leaching e-wastes used in this study do not follow conventional electrochemical
Fig. 6. Metal leaching in sulphuric acid as a function of temperature (10 g/dm3, pH 1.0, 24 h).
Fig. 7. Metal leaching in sulphuric acid as a function of time (10g/dm3, pH 1.0, 90 C).
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rules. Fig. 8 show that the percentage of copper compared to the remaining metal constituents of the waste has a significant impact on effect of galvanic interactions during leaching. At 89e 96% Cu, the effect of galvanic interaction does not appear to have a significant impact on Cu dissolution. However at 12.9% Cu, the effect of galvanic interaction on copper recovery in 24 h of leaching is significant in reducing copper recovery. The effect of copper concentration on selective leaching of copper in a mixed metal environment is consistent with observations made by Lamya and Lorenzen (2005). These results demonstrate the potential challenge in selective leaching of metallic wastes with variable compositions of metal by hydrometallurgical and biohydro-metallurgical routes. The effect of increasing pulp density on Cu recovery is shown in Fig. 9. This effect has often been attributed to the toxic effect of the waste on the bioleaching organisms (Ilyas et al., 2010; Pradhan et al., 2010). These abiotic tests suggest that other factors must contribute to the reduced Cu recovery as pulp density is increased. This also infers the potential challenge in implementing the use of very low pulp density in commercial application of this process.
3.2. Indirect and direct leaching To validate the effects observed under chemical leaching, direct leaching of waste with a growing organism was conducted and compared to biogenic acid or indirect leaching and abiotic leaching with sulphuric acid. Direct leaching was performed with 10 g/dm3 of waste with Cu-rich waste being added to the microbe culture once the pH reached 1.0. For indirect leaching, the biogenic acid was harvested and used in leaching the waste. Fig. 10 shows there is very little difference in the Cu mobilisation in both direct and indirect leaching at 30 C. It would appear that at this concentration (10 g/dm3), the toxic effects of the waste have little effect on the organism metabolism and would be consistent with the reported e-waste threshold for acidophilic bacteria (Brandl et al., 2001). The effects of abiotic and indirect leaching at the elevated temperature are compared in Fig. 11. It is apparent that at the early phase, there is little difference in the Cu mobilisation. However prolonged leaching resulted in lower Cu dissolution when biogenic acid was used in indirect leaching. Examination of the copper surface after leaching for 6 h (see Fig. 12) and the corresponding elemental analysis on the surface (see Table 2) reveal the formation of CuS (spectrum 1) and Cu2O (spectrum 4) on the surface. Spectrums 2, 3, and 5 were of the plastic components. We proposed that the reduction in copper mobilisation with time in biogenic acid results from the passivation of the Cu surface with polythionates and So of incomplete oxidation products of sulphide compounds that are present in the biogenic acid (Sasaki, 2011) and from cuprous oxidation.
Fig. 8. Effect of copper content of e-waste on Cu leaching in sulphuric acid as a function of pH (10g/dm3, 24 h, 90 C).
Fig. 10. Direct and abiotic leaching at 30 C (pH 1.0, 10 g/dm3).
Fig. 9. Metal leaching in sulphuric acid as a function of pulp density (pH 1.0, 90 C, 24 h).
Fig. 11. Indirect and abiotic leaching at 90 C (pH 1.0, 10 g/dm3).
Y. Hong, M. Valix / Journal of Cleaner Production 65 (2014) 465e472
471
Fig. 12. SEM micrograph showing the surface morphology of the copper particle leached in biogenic acid, pH 1.0 for 6 h at 90 C and EDS analysis of spectrums 1 and 4.
4. Conclusions
Acknowledgement
Almost complete mobilisation of Cu can be achieved using Acidithiobacillus thiooxidans providing evidence for the efficacy of acidophiles in the recovery of metals from e-wastes. However this was achieved within a window of conditions reflecting the various challenges in bioprocessing of e-wastes. Chemical leaching of copper rich waste has been demonstrated to be affected by passivation and galvanic coupling. Passivation of copper with chemical leaching appears to result from the oxidation of cuprous ion promoted by oxygen and peroxide dissociation. Passivation by copper oxidation and galvanic coupling effects can be overcome in wastes containing a high percentage of Cu (>90%) by using low pH (<1.0), high temperature (90 C), low pulp density (10 g/dm3) and long periods of leaching. The effect of galvanic coupling however is prevalent and is more difficult to control as the percentage of copper decreases. Leaching using biogenic acid or of direct leaching with the micro-organisms has added complexities. It is apparent that at 10 g/dm3 of waste do not pose any toxic effects on the A. thiooxidans and is the optimal pulp density to achieve maximum copper dissolution. However the presence of incompletely oxidised sulphide compounds and the sulphates leads to the formation of CuS and copper sulphate precipitates. Both of which appear to contribute to copper passivation.
This research was supported under Australian Research Council Discovery Project Scheme (DP1096342).
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