Metals removal and recovery in bioelectrochemical systems: A review

Bioresource Technology 195 (2015) 102–114

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Review

Metals removal and recovery in bioelectrochemical systems: A review Y.V. Nancharaiah a,b,⇑, S. Venkata Mohan c, P.N.L. Lens b,d a

Biofouling and Biofilm Processes Section of Water and Steam Chemistry Division, Bhabha Atomic Research Centre, Kalpakkam 603102, Tamil Nadu, India Environmental Engineering and Water Technology Department, UNESCO-IHE Institute for Water Education, P.O. Box 3015, 2601 DA Delft, The Netherlands c Bioengineering and Environmental Centre (BEEC), CSIR-Indian Institute of Chemical Technology (CSIR-IICT), Hyderabad 500 007, India d Department of Chemistry and Bioengineering, Tampere University of Technology, P.O. Box 541, Tampere, Finland b

h i g h l i g h t s

g r a p h i c a l a b s t r a c t

 Bioelectrochemical systems are

promising for recovering metal from waste streams.  Bioelectrochemical removal of metal ions was reviewed and summarized.  Studies should focus on metal recovery from metallurgical waste streams and leachates.  Long term operation of bioelectrochemical systems for metal recovery is needed.

Electricity supply

e-

Electricity harvesting CO2

Treated water

i n f o

Article history: Received 30 April 2015 Received in revised form 11 June 2015 Accepted 12 June 2015 Available online 17 June 2015 Keywords: Bioelectrochemical treatment (BET) Biorecovery Heavy metals Microbial fuel cells Wastewater treatment

e-

Exhaust

e-

e-

CO2

e-

Mered

H+

PEM

Meox

Biofilm Anode chamber

Cathode chamber

Treated water

Meox=Ag+, Au3+, Co3+, Cr6+, Cu2+, V5+, Se4+, U6+, Co2+, Ni2+, Cd2+

Metal waste water

a b s t r a c t Metal laden wastes and contamination pose a threat to ecosystem well being and human health. Metal containing waste streams are also a valuable resource for recovery of precious and scarce elements. Although biological methods are inexpensive and effective for treating metal wastewaters and in situ bioremediation of metal(loid) contamination, little progress has been made towards metal(loid) recovery. Bioelectrochemical systems are emerging as a new technology platform for removal and recovery of metal ions from metallurgical wastes, process streams and wastewaters. Biodegradation of organic matter by electroactive biofilms at the anode has been successfully coupled to cathodic reduction of metal ions. Until now, leaching of Co(II) from LiCoO2 particles, and removal of metal ions i.e. Co(III/II), Cr(VI), Cu(II), Hg(II), Ag(I), Se(IV), and Cd(II) from aqueous solutions has been demonstrated. This article reviews the state of art research of bioelectrochemical systems for removal and recovery of metal(loid) ions and pertaining removal mechanisms. Ó 2015 Elsevier Ltd. All rights reserved.

1. Introduction Availability and supply of raw materials such as metal(loid)s can greatly influence the economy (in terms of exports and job

⇑ Corresponding author at: Biofouling and Biofilm Processes Section of Water and Steam Chemistry Division, Bhabha Atomic Research Centre, Kalpakkam 603102, Tamil Nadu, India. E-mail addresses: [email protected], [email protected] (Y.V. Nancharaiah). http://dx.doi.org/10.1016/j.biortech.2015.06.058 0960-8524/Ó 2015 Elsevier Ltd. All rights reserved.

Me0=Ag0, Au0, Cu0, Se0, Hg0, Co0, Ni0, Cd0 Mered=Co2+, Cr3+, V4+ ,U4+

Me0 Organic matter Waste water

a r t i c l e

e-

creation) of many countries (Hennebel et al., 2015). Since the reserves of raw materials are finite, unequally distributed in the world and rapidly dwindling due to urbanization, high standard of living and the world population explosion, scarcity of critical raw materials is expected in the coming years. In the case of critical metals, the scarcity is perceived as an increased risk faced by the industry and characterized by the price volatility. To avoid the risk of price volatility and to stockpile raw materials for future generations, there is a need to identify secondary sources and to develop suitable technologies for their recovery. In this context,

Y.V. Nancharaiah et al. / Bioresource Technology 195 (2015) 102–114

metallurgical wastes and process streams could be potential alternative sources for resource recovery, but the economics of metal recovery from such wastes needs re-estimation due to increased prices of certain metals over the years (Hennebel et al., 2015). Metal laden waste streams are generated in various anthropogenic activities such as mining, metallurgical operations, burning fossil fuels, cement production, electroplating, leather tanning, manufacturing plastics, fertilizers, pesticides, anticorrosive agents, Ni–Cd batteries, paints, pigments, dyes and photovoltaic devices (Fu and Wang, 2011). The concentrations of certain metal(loid)s are abundant in metallurgical wastes and process streams. The concentrations of metal ions in wastewaters are usually quite low, often in the range of lg to mg/L (Wang and Ren, 2014). Therefore, the treatment methods should not only be efficient in removing metals from dilute streams but also be able to treat large volumes of waters and concentrate the metal(loid)s in sufficient amounts. Despite the challenges in recovering metals, there is a need to develop cost effective methods for removing metal ions from large volumes of metal laden wastewaters generated in various industries to comply with the discharge limits and to avoid environmental pollution. Metal contamination of natural resources is a health hazard and an environmental concern because metals are not biodegraded unlike organic pollutants and many metals can transfer across trophic levels and accumulate in the biota. Also, the presence of heavy metal ions in wastewaters is a concern. Some metal ions such as As, Pb, Hg, Cd, Cr, Cu, Ni and Zn are commonly encountered in wastewaters (Fu and Wang, 2011). Although many metal ions are essential trace elements in the metabolism of living organisms, they cause acute and chronic toxicity at higher concentrations. Due to potential toxic and carcinogenic effects, as many as 13 metals i.e. Ag, As, Be, Cd, Cr, Cu, Hg, Ni, Pb, Sb, Se, Tl, and Zn are included in the US EPA priority pollutants list (US EPA). Therefore, stringent limits have been adopted for discharge of various metal ions in wastewaters to avoid environmental contamination. Several physical, chemical and biological treatment methods have been developed for removing metal ions from water and wastewaters. Fu and Wang (2011) have provided a comprehensive review of physical and chemical methods for removing heavy metal ions from wastewaters. Microorganisms are well known to interact with a broad range of metals, metalloids and radionuclides, thereby influencing the mobility of metal(loid)s and radionuclides in natural and engineered environments (Francis and Nancharaiah, 2015). The mechanisms of metal removal by microorganisms include, but are not limited to biosorption, bioaccumulation, bioreduction and biomineralization. Microbial transformation of certain metal(loid)s and radionuclides is a useful strategy in in situ bioremediation efforts and for their removal from contaminated waters and wastewaters (Francis and Nancharaiah, 2015). However, more efforts are being made to develop innovative methods to recover metal(loid)s in order to make the treatment cost effective and sustainable (Wang and Ren, 2014). In this context, bioelectrochemical systems (BES) often used in the literature to represent both microbial fuel cells (MFCs) and microbial electrolysis cells (MECs) have emerged as a method of choice because they not only couple the treatment of organic wastewaters with that of metal laden wastewaters, but also offer a possibility to recover the metals. In fact, the evaluation of BES for removing metal ions from wastewaters has prominently begun less than a decade ago (Li et al., 2008). Since then, the utility of BES for removing various metal ions such as Ag(I), Au(III), Co(II), Cd(II), Cr(VI), Cu(II), Hg(II), Pb(II), Se(IV), V(V), U(VI), and Zn(II) has been reported. This review provides a brief introduction to BES and the state of art research performed on metal removal and recovery in both MFCs and MECs.

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2. Bioelectrochemical systems Electricity generation during oxidation of organic matter by microorganisms has been known since 1910, when M.C. Potter made first observations on glucose oxidation coupled to simultaneous production of 0.3–0.5 V electricity using baker’s yeast, Saccharomyces cerevisiae. This research topic has actually gained momentum only after the 1990s and since then seminal work has been carried out to improve the power output as well as to expand MFC applications to address other societal problems. MFCs are electrochemical hybrid systems that integrate microbial and electrochemical processes to release reducing equivalents from organic matter and convert chemical energy to electrical energy through a cascade of redox reactions mediated by microbial metabolism (Venkata Mohan et al., 2014a,b). The MFCs and other BES are being intensively pursued in both basic and applied research as a futuristic and sustainable platform specifically for harnessing energy and generating value added bio-products along with simultaneous contaminant remediation (ElMekawy et al., 2014, 2015). The microbial metabolism is linked via electron donating or accepting conditions through the external electrodes (anode and cathode), which facilitate development of a potential difference leading to bioelectrogenic activity. The research both in basic and applied fronts on MFCs and other BES has been intensified in the last decade due to its inherent ability to produce sustainable energy from renewable organic waste (Logan, 2010; Venkata Mohan et al., 2008). The power production of BES in the order of 10–100 W/m3 is still small compared to energy recovered in other anaerobic treatment technologies (i.e. anaerobic digestion) despite remarkable improvements in the design and electrode materials made during the last decade. Microbially catalyzed electrochemical mechanisms occurring in BES provide an inherent advantage for diverse applications in the arena of energy conservation, value-added product synthesis and waste remediation. The reducing equivalents (e and H+) generated as a result of biodegradation of organic matter can be utilized towards harnessing of power by MFC, waste remediation by bioelectrochemical treatment (BET), microbial electrosynthesis of various value added products and H2 production in MECs (Patil et al., 2015; Schröder et al., 2015; Modestra et al., 2015; Venkata Mohan et al., 2014a,b). All these systems, along with other BES configurations not described here, share similarities on the anode but differ in cathode reactions. The terms such as BES, microbial electrochemical technology (MET) or BET are increasingly and interchangeably referred in the literature to cover all the configurations (Patil et al., 2015; Schröder et al., 2015; Modestra et al., 2015). In MFC, chemical energy available in the organic matter is converted to electricity using the innate biodegradation capability of anaerobic microorganisms. The biodegradable organic matter available in the wastewater is oxidized by the microbial community, growing in a biofilm on the anode, into CO2, electrons and protons (Fig. 1a). CO2 and protons are released into the solution, while the microbial community conducts the electrons to the solid electron acceptor, the anode. These electrons then travel via an external circuit to the cathode, where they are finally accepted by oxygen to form water in combination with the protons available in the solution. By transferring electrons to the anode, bacteria harvest energy through anaerobic respiration. In BES, microbial communities need to use either a solid electron acceptor (bioanode) or an electron donor (biocathode) as part of their metabolism (Fig. 1b). However, it is likely that microorganisms that receive electrons from the solid electron donor (cathode) may not gain energy (Rosenbaum et al., 2011). Bacteria face a serious challenge to either transfer or accept electrons from, respectively, a solid terminal electron acceptor (TEA) or solid electron

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value-added chemicals or reductive precipitation of metal ions (Fig. 1a). Improved understanding of BES and EET led to a shift towards applications in bioelectrochemical degradation of recalcitrant pollutants, sediment bioremediation, biosensors, desalination, production of value added chemicals, removal and recovery of metal ions (Pant et al., 2012; Huang et al., 2011a,b; Velvizhi and Venkata Mohan, 2015; Venkata Mohan et al., 2009, 2010, 2014a,b). MFC and other BES can utilize soluble complex organics in wastewater as substrate and can degrade the pollutants that serve as electron acceptors. The process begins with simple molecule breakdown by the microorganisms where the electrode (bioanode) acts as the electron acceptor. Electrodes offer a potential alternative as an electron acceptor for promoting the anaerobic degradation of organic contaminants. The function of a TEA is crucial in BES, BET or MET operation in addition to the electron donor, especially when waste remediation is considered as focal objective (Velvizhi and Venkata Mohan, 2011). Apart from O2, pollutants such as nitrate, sulfate and dye molecules also function as TEAs in BES. A number of metal ions can actually substitute for oxygen and serve as an effective TEA in MFCs or MECs (Fig. 2). BES have been demonstrated for removal of pollutants such as nitrate, sulfate, phosphate, perchlorate, estrogens, poly aromatic hydrocarbons and metal ions (Krishna et al., 2014; Kumar et al., 2012; Lu et al., 2015; Mohanakrishna et al., 2010; Mathuriya and Yakhmi, 2014; ter Heijne et al., 2006; Thrash et al., 2007;

Fig. 1. Schematic representation of (a) bioelectrochemical systems and (b) extracellular electron transport (EET) mechanisms between electrodes and microorganisms. (i) Direct, (ii) electron shuttle (ES) mediated and (ii) conductive pili mediated EET. PEM = proton exchange membrane, TEAox = terminal electron acceptor oxidised; TEAred = terminal electron acceptor reduced, Meox = metal ion oxidised, Mered = metal ion reduced, ESox = ES oxidised, ESred = ES reduced.

donor, and they are known to employ different extracellular electron transport (EET) mechanisms to achieve this (Fig. 1b). The EET operated by microbial communities in the case of solid TEAs (e.g. iron oxide) and bioanodes is fairly understood as compared to biocathodes. The EET operates in biofilms via a single or a combination of multiple mechanisms, which includes a) direct contact, b) electron shuttle, or c) conductive pili mediated (Fig. 1b). Direct EET requires immediate contact between the cell and the external electron acceptor, thereby c-type cytochromes present in the outer membrane shunt the electrons. Electron shuttle-mediated EET involves the use of soluble molecules (e.g. flavins, phenazines, and quinones) as redox mediators which are produced either by the microbial cells or supplied exogenously to shunt electrons between solid electrodes and microorganisms. The third EET mechanism involves the use of conductive pili, called bacterial nanowires, for long range electron conduction (Pant et al., 2012). The biochemical components involved in EET from the cathode (electron donor) to microorganisms are not very well known. Rosenbaum et al. (2011) reviewed the EET mechanisms of biocathodes and predicted that both c-type cytochromes of outer membranes and electron shuttles (endogenous and exogenous) could participate in the electron transfer from the cathode to the microorganisms. In MEC, an external power source is needed to drive the electron flow and reduction of organic or inorganic compounds/ions with a low redox potential to form H2, H2O2, other

Fig. 2. Microbe-electrode redox tower utilized for removal and recovery of metal ions in microbial bioelectrochemical systems. In microbial fuel cells, oxidation of electron donor at the anode is coupled to the reduction of metal ions with a comparable or higher redox potential at the cathode. The net redox potential of the microbial fuel cell, that is the sum of anodic and cathodic potentials is positive, thus the electron flow from the anode to the cathode continues in a spontaneous manner. Conversely, in microbial electrolysis cells the oxidation of an electron donor at the anode is coupled to a metal ion with a lower redox potential at the cathode, which require an input of power to force the electron flow. The theoretical redox potentials (E0) are under standard conditions sourced from papers on metal removal in bioelectrochemical systems. MFC = microbial fuel cell, MEC = microbial electrolysis cell.

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Velvizhi et al., 2014; Venkata Mohan et al., 2010; Venkata Mohan and Chandrasekhar, 2011). Removal and recovery of nutrients are targeted in BES, BET or MET applications as nutrient removal is an integral part of wastewater treatment (Kelly and He, 2014). Thus, BES, BET or MET offer a flexible platform and allow simultaneous removal of multiple pollutants from wastewater. 3. Metal removal using bioelectrochemical systems In recent years, the applications of BES have been expanded to the removal of metal ions from aqueous solutions with the hope that these systems could be employed for recovering precious and scarce elements from metallurgical wastes, process streams and wastewaters. The general concept is that the biological conversions at the anode are used to deliver electrons to the cathode and drive reductive precipitation of metal(loid)s by using them as TEAs (Fig. 3a). Therefore, BES offer a novel platform for integrated wastewater treatment through oxidation and reduction reactions for recovering energy, water and metal(loid)s. The configurations of BES systems used for removing different metal ions varies between dual and single chambered BES. Often a bioelectrochemical system comprises of an anode, a cathode and an optional separator membrane. The biofilms living on the anode by oxidising organic matter provide the driving force for an electrochemical metal(loid) reduction and recovery process at the cathode. The reduced metal(loid)s are either deposited on the cathode, precipitated in the solution or remain soluble in the solution depending on the speciation of the metal(loid), and the prevailing solution chemistry. Reduction of metal ions becomes spontaneous at the cathode provided the redox potential of a cathodic half-cell reaction is either comparable or higher than the anode potential generated in BES. Consequently, the reductive precipitation of metal(loid) ions with positive redox potentials, e.g. Ag(I), Au(III), Cr(VI), Co(III), Cu(II), Hg(II), Se(IV), and V(V) has been successfully demonstrated in MFCs (Fig. 3a). Here, the removal of metal ions was coupled to organic carbon removal along with electricity production.

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For the reduction of metal(loid)s whose reduction potentials are lower than the anodic potential, the electron flow is thermodynamically not favourable. Under such conditions, an external power supply is required to drive the electron flow from the anode to the cathode in order to achieve metal ion reduction at the cathode. Therefore, reduction of metal(loid) ions such as Ni(II), Pb(II), Cd(II) and Zn(II) at the cathode was demonstrated by operating BES in the MEC mode using an external power supply (Fig. 3b, Table 2). Since the voltage needed for driving these reactions is quite small, reduction of metal ions in MECs can be powered by operating a MFC. Studies on removal of metal ions using MFC or MECs are organized into (i) metal removal in MFCs, (ii) metal removal in MECs, (iii) metal removal using biocathodes and (iv) integrated system operation for metal removal from multi-metal solutions are presented below. Finally, studies on toxicity of metal ions to microbial communities in BES are discussed. 4. Metal removal in microbial fuel cells Removal of metal ions has been studied using dual or single chamber MFCs, in which heavy metals were removed in the anaerobic or anoxic cathode chamber through cathodic metal reduction, while organic substrates were used as the carbon and electron donor in the anodic chamber (ter Heijne et al., 2010; Li et al., 2008, 2009; Huang et al., 2010; Tandukar et al., 2009; Tao et al., 2011a,b; Wang et al., 2008, 2011; Zhang et al., 2012). The catholyte solutions contained only heavy metal ions without organic compounds. Reduction of metal ions such as Ag(I), Au(III), Co(III), Cr(VI), Cu(II), Hg(II), Se(IV), or V(V) at the cathode in a MFC are presented in this section (Fig. 3a, Table 2). 4.1. Chromium(VI) Chromium is a common element used in various industrial processes including metal electroplating, leather tanning, metallurgy, dye manufacturing and corrosion control in cooling towers.

Fig. 3. Removal and recovery of heavy metals in (a) microbial fuel cells, (b) microbial electrolysis cells, (c) microbial fuel cell with bipolar membrane (modified and redrawn after ter Heijne et al. (2010)) and (d) microbial fuel cells and microbial electrochemical cells with biocathodes.

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Hexavalent (Cr(VI)) and trivalent (Cr(III)) forms of chromium are the two most stable forms found in the environment. Cr(VI) is water soluble in the full pH range and extremely toxic to living organisms causing mutagenic and carcinogenic effects. Cr(III) is less water soluble, tends to form chromium hydroxide (Cr(OH)3) precipitates in waters at moderately acidic to alkaline pH and is less toxic to living organisms. Therefore, microbial reduction of Cr(VI) to Cr(III) is a potential detoxification mechanism and a strategy for remediating contaminated waters and wastewaters (Nancharaiah et al., 2010). Microbial reduction of Cr(VI) is promising for bioremediation of contaminated soils and ground waters where the concentrations are relatively low. Biological treatment of Cr(VI) wastewaters is still challenging because toxic Cr(VI) can affect metabolism and viability of microorganisms. Moreover, recent studies showed that Cr(III) exhibits a bioaccumulation potential in both prokaryotic and eukaryotic cells and affects cell viability (Eastmond et al., 2008). Therefore, Cr(VI) reduction does not seem to be a foolproof method for mitigating exposure, toxicity and to avoid re-oxidation. Since the half-cell Cr(VI) reduction reaction has a redox potential of +1.33 V (versus the Standard Hydrogen Electrode (SHE)), much higher than O2, Cr(VI) can act as an efficient TEA in MFCs. This concept was for the first time demonstrated in a dual-chamber MFC by Wang et al. (2008) using acetate and chromate (K2Cr2O7) as electron donor and acceptor, respectively. Cr(VI) removal was determined at different initial catholyte pH (2, 3, 4, 5 and 6) and Cr(VI) concentrations (25, 50, 100 and 200 mg/L Cr(VI)). A higher Cr(VI) removal rate was observed at lower pH because of the pH dependency of the chromate reduction reaction (Table 1). Complete removal of 100 mg/L Cr(VI) was observed in 150 h from a catholyte solution with an initial pH of 2. Cr(VI) removal was also achieved at moderately acidic pH, although the reduction rates were slower. Modification of the cathode with rutile (natural form of TiO2) coating was suggested for improving Cr(VI) reduction through light induced photocatalysis at the cathode of MFCs (Li et al., 2009). The redox potential of Cr(VI)/Cr(III) and V(V)/(IV) couples are 1.33 V and 0.9 V, respectively. But, both Cr(VI) and V(V) were simultaneously reduced at the cathode using acetate as the electron donor in a dual chamber MFC (Zhang et al., 2012). Reduction of Cr(VI) was slightly better as compared to V(V) because of the higher redox potential of Cr(VI)/Cr(III) couple. The

deposits formed on the cathode were mainly composed of Cr(III) with some amount of vanadium. The pH of the catholyte gradually increased to 3.7 from the initial pH 2 during MFC operation and facilitated precipitation of Cr(III) as Cr(OH)3 deposits. While 67.5% of V(V) had been reduced, most of the vanadium (V(IV) and residual V(V)) was present in the solution and requires an additional step such as hydroxide precipitation for removal and recovery. Cr(VI) present in the wastewater was bioelectrochemically reduced to Cr(III) in the presence of an organic electron donor in a dual-chambered MFC with simultaneous generation of electricity (Gangadharan and Nambi, 2015). The Cr(VI) as catholyte (100 mg/L) was completely removed within 48 h (initial pH 2.0) by precipitation on the cathode. In addition to that, 78.4% of total organic carbon removal was achieved in the anode chamber with a maximum power density of 0.767 W/m2 (2.08 mA/m2). 4.2. Cobalt(III) Cobalt is a rare metal with an occurrence of 0.001% in the Earth’s crust. It is mainly produced as a by-product of copper and nickel refining. Cobalt has many diverse industrial and military applications. It is a component of super alloys used in making turbine engines of aircrafts. In many countries, cobalt is classified as a strategic or critical metal. It is one of the 14 critical raw materials identified by the European Union (Hennebel et al., 2015). Lithium ion batteries, which use lithium cobalt oxide (LiCoO2) as the cathode material, are widely used as a power source in mobile phones, laptops, video cameras and other electronic devices. In Li-ion batteries, cobalt constitutes to about 5–10% (w/w), much higher than its availability in ore. Therefore, lithium ion batteries are a potential source for cobalt recovery (Xin et al., 2009). Recovery through a biohydrometallurgy route (e.g. bioleaching) has been evaluated to reduce the cost, chemical usage and waste generation (Xin et al., 2009). Co(III) is an efficient TEA for operation of MFC because the Co(III)/Co(II) redox couple has a high positive potential (+1.61 V vs SHE). Huang et al. (2013) studied cobalt leaching from LiCoO2 particles in MFCs using acetate as the electron donor. Graphite felt was soaked in water containing LiCoO2 particles to deposit the particles on the electrode surface. When LiCoO2 particle loaded graphite felt was used as the cathode, release of Co(II) into the catholyte solution was observed. In fact, the Co(II) leaching rate

Table 1 Half-cell reactions at the anode and cathode of microbial bioelectrochemical systems used for removal and recovery of metal ions. Electron donor/acceptor

Half-cell reaction

Redox potential (E0)

References

Anodic reactions Acetate Glucose

þ  CH3 COO þ 4H2 O ! 2HCO 3 þ 9H þ 8e þ  C6 H12 O6 þ 12H2 O ! 6HCO 3 þ 30H þ 24e

0.289 V 0.41 V

ter Heinje et al. (2010) Catal et al. (2009)

LiCoO2 ðsÞ þ 4Hþ þ e ! Co2þ ðaqÞ þ Li þ ðaqÞ þ 2H2 O

+1.61 V

Huang et al., 2014a,b

3þ þ  Cr2 O2 7 ðaqÞ þ 6e þ 14H ! 2Cr ðaqÞ þ 7H2 O   AuCl4 ðaqÞ þ 3e ! AuðsÞ þ 4Cl

+1.33 V

Wang et al. (2008)

+1.00 V +0.991 V

Choi and Hu (2013) Zhang et al. (2012) Wang et al. (2011)

Cathodic reactions Cobalt(III) Chromium(VI) Gold(III) Vanadium(V)

2þ þ  ðaqÞ þ H2 O VOþ 2 ðaqÞ þ 2H þ e ! VO

Mercury(II)

Hg2þ ðaqÞ þ 2e ! 2Hgþ ðsÞ

+0.91 V

Silver(I)

Agþ ðaqÞ þ e ! Ag0 ðsÞ

+0.799 V

Choi and Cui (2012)

Silver(I)

0  AgðNH3 Þþ 2 ðaqÞ þ e ! Ag ðsÞ þ 2NH3 ðaqÞ

+0.373 V

Wang et al. (2013)

Silver(I)

½AgS2 O3  ðaqÞ þ e ! Ag0 ðsÞ þ S2 O2 3 ðaqÞ

+0.250 V

Tao et al. (2012)

Selenium(IV)

SeðIVÞðaqÞ þ 4e ! Se0 ðaqÞ

+0.41 V

Catal et al. (2009)

Copper(II)

Cu2þ ðaqÞ þ 2e ! Cu0 ðsÞ

+0.286 V

ter Heinje et al. (2010)

Cobalt(II)

Co2þ ðsÞ þ 2e ! Co0 ðsÞ

0.232 V

Huang et al., 2014a,b

Nickel(II)

Ni

0.25 V

Qin et al. (2012)

2þ

0

ðaqÞ þ 2e ! Ni ðsÞ

Cadmium (II)

Cd

ðaqÞ þ 2e ! Cd ðsÞ

0.403 V

Modin et al. (2012)

Zinc(II)

Zn2þ ðaqÞ þ 2e ! Zn0 ðsÞ

0.764 V

Abourached et al. (2014)

2þ



0

Table 2 Bioelectrochemical recovery of metals in microbial fuel cells and microbial electrolysis cells without and with biocathodes. Metals

BES configuration, Electrode materials

Electron donor, concentration

Metal removal efficiency

Maximum power output (W/m2)

References

K2Cr2O7; 25, 50, 100 and 200 mg/L Cr(VI); pH 2–6 K2Cr2O7; 26 mg/L Cr(VI); pH 2

100% from 100 mg/L in 150 h

Wang et al. (2008)

97% in 26 h

0.150 at 200 mg/L Cr(VI) and pH 2 

K2Cr2O7; 100 mg/L Cr(VI); 2

99.8% in 48 h from 100 mg/L

0.767 at 100 mg/L and pH 2

LiCoO2 particles; solid/ liquid ratios of 50–1000 mg/L (w/v); pH 1–3 CuCl2; 1 g/l Cu(II); pH 3

99.1% from solid/liquid ratio of 50 mg/L in 48 h 99.88% in anaerobic cathode in 6 d; 99.95% in aerobic cathode in 7 d; residual Cu below 1.2 mg/L Cu >99% from 10–100 mg/L Cu2+ at pH 2–5 92% from 600 mg/L Cu2+ in 480 h, 48% from 2000 mg/L Cu2+ in 672 h >96% from 200 mg/L Cu(II) in 264 h



Gangadharan and Nambi (2015) Huang et al. (2013) ter Heijne et al. (2010)

CuSO4; 10–200 mg/L Cu(II); pH 2–5 CuSO4; 600 and 2000 mg/L Cu(II); pH 2 CuSO4; 50, 200, 500 or 1000 mg/L Cu(II); pH 4.7 AuCl 4 ; 100–200 mg/L Au(III); pH 2 HgCl2; 25, 50, 100 mg/L Hg(II); pH 2 SeO2 3 ; 50–200 mg/L Se(IV); pH 7 AgNO3; 50, 100, 200 mg/L Ag(I); pH 7 AgNO3; 50, 100 or 200 mg/L Ag(I); pH 2, 4, 6.6; [AgS2O3] Ag(NH3)+2; 1000 mg/L Ag(I); pH 9.2 NaVO3; 500 mg/L V(V); pH 2

0.43 and 0.80 under anaerobic and aerobic cathodes, respectively 0.319 at 200 mg/L Cu2+, pH 3

Li et al. (2009)

Wang et al. (2010)



Tao et al. (2011b)

339 mW/m3 at 6412 mg/L Cu(II), pH 4.7 6.58 for 2000 mg/L Au(III) at pH 2 0.433 at 100 mg/L, pH 2

Tao et al. (2011a) Choi and Hu (2013) Wang et al. (2011)

99% of 75 mg/L in 48 h in acetate fed MFC 99.9% of 50 mg/L in 8 h; 0.05 mg/L remaining 95% in 36 h

2.90 at 25 mg/L Se(IV)

Catal et al. (2009)

4.25 at 1000 mg/L, pH 7 0.109

Choi and Cui, (2012) Tao et al. (2012)

99.9% of 1000 mg/L Ag + in 21 h; <1 mg/L 25% in 72 h

0.3 at 1000 mg/L, pH 9.2

Wang et al. (2013)

0.572 at 500 mg/L, pH 7

Zhang et al. (2009)

99.8% from 200 mg/L; residual 0.22 mg/L Au(III) 98.2–99.5% in 10 h; 0.44–0.69 mg/L

Applied voltage Metal recovery using microbial electrolysis cells with abiotic cathodes Cadmium tMEC, carbon brush anode, carbon Sodium acetate, cloth cathode Cobalt tMEC, graphite felt anode, carbon rod Sodium acetate, cathode Cobalt tMEC, graphite brush anode, graphite Sodium acetate, felt cathode Nickel tMEC, carbon felt anode, stainless Sodium acetate, steel cathode

1 g/l

CdSO4; 50, 100, 200 mg/L Cd(II); pH 6

93.6% from 50 mg/L Cd(II) in 60 h

Cr(VI) reducing MFC

Choi et al. (2014)

1 g/l

Co(II); 50 mg/L Co(II); pH 6

7 mg/L/h Cu(II)

Co(II) leaching MFC

1 g/l

CoCl2; 874 lM Co(II); pH 3.8 to 6.2

92% in 6 h

0.3–0.5 V

Huang et al. (2014a) Jiang et al. (2014)

1 g/l

NiSO4; 50–1000 mg/L Ni(II); pH 5

99% of 50 mg/L Ni(II) in 20 h

0.9 V

Qin et al. (2012)

88% in 6 h

0.2 V

99.2% in 7 h

–

79.3% in 24 h

–

Huang et al. (2014a) Huang et al. (2010) Wu et al. (2015)

19 mg/L/d

–

0.46 mg Cr(VI)/mgVSS/h

0.055 V

Metal recovery using bioelectrochemical systems (microbial fuel cells and microbial electrolysis cells) with biotic cathodes Cobalt tMEC, graphite fiber anode, graphite Sodium acetate, 1 g/l CoCl2; 20 mg/L Co(II); pH 6.2 felt cathode Chromium tMFC, graphite plate anode, graphite Sodium acetate, 1.6 g/l K2Cr2O7; 39.2 mg/L Cr(VI); plate in granular graphite cathode Chromium tMFC, graphite felt for anode and Glucose, 1 g/l K2Cr2O7; 6–20 mg/L Cr(VI); pH 7 cathode Chromium tMFC, graphite brush anode, graphite Sodium acetate, 1 g/l K2Cr2O7; 20 mg/L Cr(VI); pH 7 granules cathode Chromium tMFC, graphite plates for anode and Sodium acetate K2Cr2O7; pH 7 cathode

Huang et al. (2011b) Tandukar et al. (2009) 107

tMFC = two-chamber microbial fuel cell, sMFC = single-)chamber microbial fuel cell, tMEC = two-chamber microbial electrolysis cell.

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Metal recovery using microbial fuel cells with abiotic cathodes Chromium tMFC, graphite plates for anode and Sodium acetate, 2.64 g/l cathode Chromium tMFC, graphite plate anode, rutile CH3COOH, 1.64 g/l coated graphite plate cathode Chromium tMFC, carbon cloth for anode and Sodium acetate, 1 g/l cathode Cobalt tMFC, graphite felt for anode and Sodium acetate, 0.38 COD/l cathode Sodium acetate, 1.64 g/l Copper tMFC, graphite plate for anode, graphite foil pressed on Ti plate for cathode Copper tMFC, graphite felt for anode and Sodium acetate, 0.82 g/l cathode Copper tMFC, graphite felt for anode and Sodium acetate, 1 or 0.5 g/l cathode Copper tMFC, graphite plates for anode and Glucose, 5 g/l cathode Gold tMFC, carbon brush anode, carbon Sodium acetate, 1 g/l cloth cathode Mercury tMFC, graphite felt anode, carbon Sodium acetate, 0.82 g/l paper cathode Selenium sMFC, carbon cloth anode, coated Acetate or glucose, carbon cloth cathode Silver tMFC, carbon brush anode, carbon Sodium acetate, 1 g/l cloth cathode Silver tMFC, graphite plate anode, graphite Sodium acetate, 1.28 g/l felt cathode Silver tMFC, carbon cloth anode, graphite Sodium acetate, 1.6 g/l felt cathode Vanadium tMFC, carbon fiber felt for anode and Glucose, 0.81 g/l; Sulphide, cathode 100 mg/L

Metal salt; metal concentration; pH

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in MFC was 3.4 times higher compared to a chemical leaching process. The leaching was aided by acetate oxidation at the anode and transfer of electrons to the cathode where these electrons were used for the reduction of solid Co(III) to Co(II) and subsequent release of Co(II) ions into the catholyte. In addition, the cobalt leaching capacity of the MFC was increased by almost 3 times by addition of 10 mg/L Cu(II) into the catholyte (Liu et al., 2013). For recovering cobalt from the aqueous solution, subsequent reduction of Co(II) to Co(0) is still needed. Since the redox potential (Co(II)/Co(0)) of the half cell reaction is negative (0.232 V vs SHE) and lower than the bioanode, the cathodic reduction of Co(II) cannot occur spontaneously under MFC conditions. Further reduction of Co(II) to Co(0) can still be achieved in a MEC by providing an external power supply. 4.3. Copper(II) Copper is an essential micronutrient in living organisms. But, it is potentially toxic to living organisms at higher concentrations. Copper is present in wastewaters arising from the mining, smelting, semiconductor, metallurgical, electroplating, wire drawing, and copper polishing industry. Methods such as precipitation, adsorption and reduction are currently employed to remove copper ions from industrial wastewaters. Electrochemical reduction is attractive as it uses electricity to deposit metallic Cu(0) on the cathode. The energy harvested through oxidation of acetate or glucose at the anode of MFCs drives the reductive precipitation of Cu on the cathode. A net theoretical cell voltage of 0.49 or 0.69 V can be achieved using Cu(II) as the TEA, depending on the electron donor acetate or glucose, respectively. Thus, cathodic reduction of Cu(II) to Cu(0) occurs spontaneously in a MFC. Cu(II) removal from aqueous solution was studied in dual-chamber MFC using either acetate or glucose as the electron donor. Copper salts such as CuCl2, CuSO4, and fly ash leachate were used as the catholyte. An acidic pH was chosen for the catholyte to avoid chemical precipitation of Cu(II) ions at moderately acidic and alkaline pH. ter Heijne et al. (2010) used a bipolar membrane between the anode and cathode chambers to avoid H+ ingress to the anode chamber and to keep the pH differences intact. The bipolar membrane produces H+ and OH ions through water splitting that occurs between the cationic and anionic layers using part of the energy generated in the MFC (Fig. 3c). The pH difference between the anode and the cathode compartments was maintained by the migration of H+ to the cathode compartment and OH ions to the anode compartment through the membrane. Removal of almost all the copper (1 g/L) was observed in 6 and 7 days, respectively, under anaerobic and aerobic cathode conditions. Copper removal efficiencies of >99% were achieved from the CuCl2 catholyte (1 g/L Cu(II)) at pH 3 in 6 to 7 days of MFC operation. The remaining copper concentrations were in the range of 0.5–1.2 mg/L, below the US EPA maximum containment limit of 1.3 mg/L for drinking water. Removal of copper from aqueous solution was corroborated through precipitation of Cu(0) on the cathode. X-ray powder diffraction measurements revealed that the deposits formed on the cathode were composed of metallic Cu(0). Formation of neither CuO nor Cu2O deposits on the cathode was observed. Copper recovery from a copper sulphate solution was studied in a dual-chamber MFC at different initial Cu(II) concentrations from 50 to 6412 mg/L (Tao et al., 2011a). The anode chamber was fed with a minimal medium (pH 7) containing glucose, while the cathode chamber was filled with a CuSO4 solution at pH 4.7. The standard potential for glucose at pH 7 is 0.41 V, thus a net theoretical cell voltage of Ecell 0.69 V favours spontaneous reduction of Cu(II) to Cu(0) aided by the glucose oxidation at the anode. Under these operating conditions, almost complete (>99%) removal of Cu(II) ions was achieved from a 196 mg/L Cu(II) solution. The remaining

Cu(II) concentration in the outlet water was <1.3 mg/L. X-ray diffraction analysis of the deposit confirmed the formation of both metallic Cu(0) and Cu2O on the cathode. The sustained removal of Cu(II) was shown by operating a MFC in fed-batch mode for 8 cycles. The same authors successfully recovered Cu in a 16 L pilot scale membrane less MFC (Tao et al., 2011b). Copper was removed and recovered as deposits on the cathode of a MFC by feeding the cathode chamber with a fly ash leachate (Tao et al., 2014). The copper concentration decreased from an initial 52.1 mg/L to 1.5 mg/L and a removal efficiency of 97% of Cu(II) was achieved in 36 h operation. Other metals such as Pb(II) and Zn(II) present in the fly ash leachate were not removed in this MFC. Removing Pb(II) and Zn(II) from the treated fly ash leachate was achieved in an electrochemical reactor (Tao et al., 2014). Integration of BES with conventional electrolysis is still attractive because the overall cost will be lower than using only electrolysis for metal removal. 4.4. Gold(III) Gold, silver, palladium and platinum are used in jewellery, electronics and dental applications. There is a growing interest in recovering these precious metals from scrap and waste streams. Traditionally, chemical precipitation methods are used for recovering precious metals from leachate of scrap and process streams, which are ineffective at low concentrations. Recent studies have shown the potential of biological methods for recovering precious element such as palladium as biomass associated palladium nanoparticles for environmental applications (Suja et al., 2014). Gold removal was studied in a dual-chamber MFC using acetate and Au(III) as electron donor and acceptor, respectively (Choi and Hu, 2013). Reduction of Au(III) is a thermodynamically favoured process at the cathode of a MFC as the standard redox potential of Au(III) or Au(0) is fairly high at 1.002 V. Removal efficiencies of 97.8% and 94.6% were achieved for initial concentrations of 50 and 100 mg/L Au(III), respectively in 12 h. A maximum power output of 0.89 W/m2 was achieved for 100 mg/L Au(III). Removal of Au(III) from the catholyte was associated with deposition of metallic Au(0) on the cathode surface. The study shows potential application of MFCs for recovering precious metals from aqueous solutions. 4.5. Mercury(II) Mercury is the 16th rarest element and available in trace amounts on the earth. It exits as soluble Hg(II) ion bound to chloride, sulfide, or organic acids in waters principally arising from anthropogenic activities such as burning of coal and petroleum, the use of mercurial fungicides, and catalysts. Mercury can also enter air due to the release of Hg(0) vapours from compact fluorescent lamps and partitioning of Hg(0) from water into air due to its high volatility. Mercury is well known to bioaccumulate and biomagnify as methyl mercury in living organisms (Jiang et al., 2012). The toxic effects on ecosystems result from methyl mercury and other chemical forms of mercury. Various chemical and biological methods are available for reduction of soluble Hg(II) to insoluble Hg(0), followed by adsorption and retention of the Hg(0). Reduction of soluble Hg(II) to insoluble elemental Hg(0) is known to occur in a diverse group of microorganisms. Alternatively, elemental selenium nanoparticles have superior affinity to adsorb Hg(0) vapour in a stable HgSe form and can thus be used as adsorbent (Jiang et al., 2012). Simultaneous reduction of Se(IV) and Hg(II) by Shewanella putrefaciens was therefore exploited for capturing Hg(0) onto selenium nanospheres to form stable HgSe (Jiang et al., 2012).

Y.V. Nancharaiah et al. / Bioresource Technology 195 (2015) 102–114

Hg(II) is a potential TEA in a MFC because the Hg(II)/Hg(0) couple has a high theoretical standard redox potential of +0.91 V (vs SHE). Wang et al. (2011) have investigated cathodic reduction of Hg(II) in a dual-chamber MFC using acetate as the electron donor. The cathode chamber was irrigated with an HgCl2 solution and made anaerobic by purging with N2 gas. Rapid removal of Hg(II) was noted for different initial concentrations of 25, 50, and 100 mg/L. A removal efficiency of more than 98% Hg(II) was observed within 10 h. The Hg(II) concentration in the catholyte was decreased to 0.44–0.69 mg/L. Scanning electron microscope-energy dispersive spectroscopy (SEM-EDS) analysis revealed formation of Hg(0) deposits onto the cathode. However, the formation of other Hg compounds such as Hg2Cl2 was also noted at the bottom of the cathode chamber. 4.6. Selenium(IV) Selenium (Se) is a scarce element with uneven geographic distribution in the Earth’s crust. It is required in trace amounts in living organisms for its role in essential metabolic functions and mitigating oxidative stress. Selenium containing wastewaters are generated in coal combustion, oil refining, metal refining, coal and phosphate mining (Nancharaiah and Lens, 2015a). In addition, selenium is used in several industrial processes and products such as electronics, semi-conductors, glass manufacturing, pigments, stainless steel, photoelectric cells, and pesticides, which generate water contaminated with the selenium oxyanions selenate and/or selenite. Stringent limits are in place for regulating discharge because selenium gets enriched in the food chain due to its trophic transfer and bioaccumulation abilities. According to the US EPA, the discharge limit for wastewaters is below 5 lg L1 Se. Biological methods are preferred for treatment of selenium wastewaters over chemical methods because of low cost and wastewater characteristics (dilute and high volume). Microbial reduction of soluble selenium oxyanions into insoluble elemental selenium is a key process exploited in the biological treatment of selenium wastewaters (Nancharaiah and Lens, 2015a). The treatment of selenium bearing wastewaters is challenging because of the stringent discharge limits, occurrence of co-contaminants (e.g. metals, competing electron acceptors), and the fate of bioreduced selenium. There is a need for development of integrated treatment strategies with a potential to recover this scarce element from wastewaters (Nancharaiah and Lens, 2015b). Selenite removal was investigated in a single chamber air cathode MFC using defined substrate (e.g. acetate or glucose) as the electron donor (Catal et al., 2009). Complete reduction of 75 and 200 mg/L of selenite was achieved in 48 and 72 h, respectively, in acetate and glucose fed MFCs. Bright red coloured precipitates were observed on both the electrodes (anode and cathode) and in the solution of the selenite fed MFC. It is desirable to determine the fraction of selenium deposited on the electrodes and suspended in the treated water. Since a fraction of the bioreduced selenium remains suspended in the treated water, most probably in the form of colloidal particles, an additional step is needed to remove the elemental Se from the treated water. The reduction of Se(IV) to Se(0) was caused by anode respiring bacteria. The authors noted that oxygen, and not selenite acted as the TEA at the cathode because no voltage was generated when the MFC was shifted to an anaerobic chamber. It is thus not clear how this approach of selenite reduction in an MFC, but not linked to cathodic reduction, is different from the direct microbial reduction of selenite. Since the redox potential of the Se(IV)/Se(0) couple is +0.41 V (vs SHE) and the net theoretical cell voltage of +0.82 V using glucose as electron donor clearly shows the possibilities for Se(IV) as a TEA at the cathode. Although, oxygen will be preferred over

109

selenite as TEA due to its higher redox potential, cathodic reduction of Se(IV) should be thermodynamically favourable under anaerobic conditions. Additional studies are needed to investigate the reductive precipitation of selenium at the cathode of MFCs using selenate or selenite as TEA. 4.7. Silver(I) Silver is commonly present in effluents which originate from electronic, jewellery and photographic industries. Current methods for removal of Ag(I) ions include precipitation, ion exchange, and electrochemical reduction. Biological methods such as biosorption and bioreduction are less expensive but suffer from low selectivity and recovery. Electrochemical deposition has been the method of choice for recovering Ag(I) ions from aqueous solutions, but requires an input of electrical energy at a rate of 3.81 KWh per kg silver recovered (Choi and Cui, 2012). An MFC was evaluated for removing silver from jewellery wastewater because the energy needed for Ag(I) ion reduction can be harvested from the oxidation of organic matter present in the wastewater. Because of the positive standard redox potential ((Ag(I)/Ag(0), E0 = 0.799 V), Ag(I) is a potential TEA at the cathode of the MFC wherein reduction of Ag(I) ions allows precipitation of metallic Ag(0). Removal of Ag ions from AgNO3, silver thiosulfite ([AgS2O3]) and Ag-ammonia (Ag(NH3)+2) complex was determined in dual-chamber MFCs using an anaerobic cathode. [AgS2O3] and Ag(NH3)+2 complexes were used to simulate chelated silver compounds present in the effluents of the photographic and silver plating industries, respectively. The cathodic reduction of Ag(I) was studied from an AgNO3 solution in a dual-chamber MFC using acetate as the electron donor (Choi and Cui, 2012). The conductivity of the catholyte was improved by adding 0.2 M NaClO4. Silver removal efficiencies of >99% were achieved in 8 h for initial concentrations ranging from 50 to 200 mg/L Ag(I) ion. SEM-EDS analysis showed deposition of metallic Ag(0) crystals onto the cathode. Experiments with simulated photographic wastewater showed the utility of MFCs for silver recovery from the silver thiosulfite ([AgS2O3]) complex (Tao et al., 2012) using acetate as the electron donor. The pH of the catholyte varied from 2.0 to 6.6, but had no influence on Ag(I) ion reduction. Deposition of small amount of acanthite (Ag2S) along with metallic Ag(0) was noted on the cathode during silver sulfite reduction, indicating reduction of thiosulfite in the MFC. The reduction rate of silver was much slower when supplied as silver thiosulfite complex as compared to that achieved using silver nitrate. The potential of the MFC in recovering silver from chelated complexes is very important when these systems are applied for metal recovery from real process streams and wastewaters. 4.8. Vanadium(V) Vanadium is a trace element in living organisms. The vanadium abundance (0.013% w/v) in the Earth’s crust is larger than that of copper and zinc. Vanadium is used in various industrial applications such as vanadium alloys, photographic development, ceramic colouring and also in technologies linked to atomic energy and space operations. Vanadium contamination of water sources can thus arise from both natural and anthropogenic activities. Vanadium containing wastewaters are generated in large quantities in vanadium mining and vanadium pentoxide (V2O5) production. Pentavalent vanadium (V(V)) is considered the most toxic form, while V(IV) is less toxic and insoluble at neutral and alkaline pH. Reduction of V(V) to V(IV) is a remediation method to remove vanadium from contaminated waters and environments. Some microorganisms are able to reduce toxic V(V) to less toxic V(IV) through detoxification and anaerobic respiration using V(V) as the TEA.

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Reduction of V(V) to V(IV) was studied in a two-chamber MFC using sulphide as an electron donor (Zhang et al., 2009). Removal efficiencies of about 82% and 26% were achieved for sulphide and vanadium, respectively. This approach is promising because the reduction products of both pollutants are precipitated and removed from the wastewater. In another study, both V(V) and Cr(VI) were used as the electron acceptors at the cathode of a MFC. Reduction efficiencies of 68 and 75% for Cr(VI) and V(V), respectively, were achieved. Selective recovery of both metals was possible because Cr(III) was mainly deposited on the cathode, while V(IV) remained in the catholyte. The soluble V(IV) could be precipitated out of the treated water by increasing the pH. A recent study showed that V(V) can be reduced to V(IV) by microbial and electrochemical reduction mechanisms in anode and cathode chambers, respectively (Zhang et al., 2012). 5. Metal recovery in microbial electrolysis cells Electron flow from the anode to the cathode is thermodynamically not favourable if metal ions with lower redox potentials than the bioanode are used as the TEA. For example, the standard redox potential of the Ni(II)/Ni(0) couple is 0.25 V and Ni(II) cannot be reduced by accepting electrons from the cathode in a MFC. However, the reduction of such metal ions, i.e. Cd(II), Co(II), Ni(II), Zn(II), and Pb(II) can be made possible by using an external power to drive the electron flow from the high potential anode to the low potential cathode (Fig. 3b, Table 2). Although, cathodic reduction of these metal ions in MECs requires energy input compromising the intended use of BES, the overall voltage needed is still lower than the conventional electrolysis because part of the voltage is supplemented by the anode. In order to further reduce the energy input, MFCs have been evaluated as a power source for driving reduction of metal ions in MECs. 5.1. Cobalt(II) Studies on cobalt recovery using BES have mainly concentrated on leaching of Co(II) from LiCoO2 and subsequent recovery of Co(II) from the aqueous solution. Recycling of cobalt from Li-ion batteries is of primary importance in recycling Li-ion batteries and to prevent environmental pollution. For recovering cobalt from Li-ion batteries, the cathode material needs to be separated from the other materials through a pre-treatment step involving processes such as skinning, removing of crust, crushing and sieving (Xin et al., 2009). Then, the cathode materials are subjected to leaching and recovery processes. MFCs have been investigated for leaching Co(II) from LiCoO2, while Co(II) has been recovered from aqueous solutions in MECs. Since the standard redox potential of the Co(II)/Co(0) couple is negative (0.232 V vs SHE) and comparable to that of the bioanode potential using acetate as the electron donor, an input of energy is needed to drive reduction of cobalt(II) at the cathode. Co(II) ions leached from LiCoO2 particles on the cathode of a MFC were reduced to Co(0) on the cathode of a MEC. Operation of the Co(II) reducing MEC was powered by the Co(II) leaching MFC. Therefore, a sequential MFC-MEC (MFCCo(III)–MECCo(II)) system was proposed for leaching and recovery of cobalt from spent lithium ion batteries (Huang et al., 2014a,b). A cobalt leaching rate of 46 mg l1 h1 was observed in the MFC, while a Co(II) to Co(0) reduction rate of 7 mg l1 h1 was achieved in the MEC. Recently, reduction of Co(II) in a MEC (MECCo(II)) was powered by the Cu(II) reducing MFC (MFCCu(II)) without providing any external power input (Huang et al., 2015). The MFCCu(II)–MECCo(II) system was used for simultaneous removal of Cu(II) and Co(II) from aqueous solutions using acetate as the electron donor.

In another study, Co(II) was recovered as elemental Co(0) deposits on the cathode along with simultaneous H2 production in a MEC with an applied voltage of 0.2–0.5 V (Jiang et al., 2014). A rapid removal of cobalt with an efficiency of 92% from a 847 lM of Co(II) solution was observed within the first 6 h. Cobalt was removed from the aqueous solution via reductive deposition and formation of flakes of metallic Co crystals on the cathode surface. However, adsorption and diffusion of Co(II) from the cathode chamber to the anode chamber accounted for about 27% Co(II) removal. The microbial community of the anode biofilm in the Co(II) reducing MEC was completely different from that present in the start-up period. Co(II) ingress into the anode chamber could have exerted a selection pressure for the development of a cobalt resistance microbial community (Jiang et al., 2014). This again highlights the importance of ingress of metal ions to the anode compartment and there is a need to link metal ion ingress, structure and function of the microbial community to the fate of metal ions. 5.2. Cadmium(II) The standard redox potential of the Cd(II)/Cd(0) couple is 0.40 V. For the reduction of the Cd(II) at the cathode, external voltage was supplied. Choi et al. (2014) used a Cr(VI) reducing MEC to power the operation of a MEC for Cd(II) reductive precipitation. The Cd(II) concentration decreased from the initial 50, 100 and 200 mg/L to 3.2, 6.7 and 20.5 mg/L during 60 h of MEC operation. The authors have successfully demonstrated Cd(II) reduction in a MFCCr(VI)–MECCd(II) system without any extra energy input. 5.3. Copper(II) Copper ion removal, denitrification and neutralization were simultaneously achieved using a bioelectrochemical reactor while treating copper metal pickling wastewater (Watanabe et al., 2001). The reactor was operated by applying external potential and feeding acetate as the carbon source employing carbon based electrodes and denitrifying microorganisms immobilized on the surface of the cathode. The electric current contributed to the removal of the copper ion and generation of hydrogen. The generated hydrogen and the added acetate were effectively utilized for denitrification of high strength nitrate wastewater. Both electric current and an external source of organic matter enhanced the denitrification rate due to simultaneous utilization of hydrogen derived from the electrolysis and the added organic matter. Increment in pH was caused by the generation of hydroxyl ions during denitrification. The inorganic carbon species generated during denitrification with acetate and by the electrochemical oxidation of organic substrate in the anode chamber acted as a buffer to minimize a further increase of pH at higher nitrate removal efficiencies. 5.4. Nickel(II) Aqueous nickel was recovered on the cathode of a dual-chamber MEC at an applied voltage of 0.7 V (Luo et al., 2014). A removal efficiency of 94% was achieved from a 500 mg/L Ni(II) in 40 h. At an applied voltage of 0.5 and 1.1 V, between 51% and 67% Ni(II) was removed from an initial 500 mg/L at pH 5 (Qin et al., 2012). Results showed that the Ni(II) removal efficiency with the MEC was 3 times higher than that with an electrolysis cell and a MFC. Ni(II) removal efficiencies decreased from 99 (±0.6)% to 33 (±4.2)% with the increase in initial Ni(II) concentrations from 50 to 1000 mg/L, while the mass removal of Ni(II) increased consistently with the initial concentrations. During the Ni(II) removal process in the MEC, the nickel was recovered by depositing on the cathode electrode. The MEC performance was affected by

Y.V. Nancharaiah et al. / Bioresource Technology 195 (2015) 102–114

various conditions, including the initial Ni(II) concentration, applied voltages, and initial pH. 5.5. Uranium(VI) The anthropogenic activities such as mining, milling, uranium ore processing, nuclear fuel fabrication, spent fuel processing, and weapon production have led to uranium contamination of water sources (Nancharaiah et al., 2006; Francis and Nancharaiah, 2015). In soils and waters, uranium primarily exists as soluble uranium (VI) species. In anaerobic environments, U(VI) is reduced to the less soluble U(IV). Microbe-uranium interactions revealed that several microorganisms use U(VI) as an alternative electron acceptor in anaerobic respiration, reducing soluble U(VI) to insoluble U(IV) (Francis and Nancharaiah, 2015). Microbial reduction of U(VI) to U(IV) is a potential bioremediation method for in situ immobilization in subsurface environments. This approach is better than the pump and treat remediation strategies, but also has some limitations such as maintenance of suitable conditions for microbial reduction as well as stability of bioreduced U(IV) precipitates. As an alternative method, bioelectrochemical reduction of U(VI) was explored for immobilization and recovery of uranium from the contaminated waters (Gregory and Lovley, 2005). Graphite electrodes poised at 0.5 V were used as the electron donor for the uranium reducing bacterium Geobacter sulfurreducens. U(VI) was effectively recovered from the aqueous solution by the G. sulfurreducens cells through reduction and deposition on the electrode. Under defined conditions, an amount of 80 lM of U(VI) was removed from the aqueous solution and deposited on the cathode. The experiments performed in soil column experiments and ground water augmented with uranium, showed promising results for uranium recovery.

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biofilm was used as the biocathode in a MFC. The authors first developed an electroactive biofilm on the anode in an MFC and subsequently used it as the cathode for Cr(VI) reduction experiments (Wu et al., 2015). The start-up times and Cr(VI) reduction improved in biocathode MFC with a fixed set potential of 0.3 V as compared to a control MFC operated without a set potential (Huang et al., 2011b). A MEC with a biocathode was operated for recovering Co(II) along with the simultaneous formation of methane and acetate (Huang et al., 2014b). Co(II) reduction of 88.1% with simultaneous yields of 0.266 (±0.001) mol Co/mol COD, 0.113 mol CH4/mol COD, and 0.103 (± 0.003) mol acetate/mol COD were reported at an applied voltage of 0.2 V at the biocathode (Huang et al., 2014a,b). The apparent activation energy obtained in MECs was 26.7 kJ/mol compared to 40.5 kJ/mol in the abiotic controls, highlighting the importance of cathodic microbial catalysis on Co(II) reduction. The composition of the bacterial community on the cathodes showed the presence of microorganisms such as Geobacter psychrophilus, Acidovorax ebreus, Diaphorobacter oryzae, Pedobacter duraquae, and Prolixibacter bellariivorans. These studies demonstrated the technical feasibility of removing metal ions from aqueous solutions using biocathodes wherein the metal ion removal is actually driven by a combination of mechanisms such as bioelectrodeposition, bioreduction and biosorption. Adsorption, migration of metal ions from the cathode to the anode chamber and even chemical precipitation may also contribute to the metal removal. Since metal deposits are often enmeshed with the microbial cells and in the extracellular polymeric substances (EPS) matrix of the biofilm, additional separation methods are needed for recovering metals when biocathodes are employed.

7. Integrated system operation for metal removal from multimetal solutions

6. Metal recovery using biocathodes There are limited studies on bioelectrochemical recovery of metals employing biocathodes. In this configuration, microorganisms oxidize the organic matter to release electrons which travel to the cathode via an external resistor, and the protons which migrate to the cathode chamber through a cation exchange membrane. The reduction of metals at the cathode can be enhanced by using metal reducing bacteria, which shunt electrons from the poised cathode to the TEA, i.e. Cr(VI), Co(II) (Fig. 3d, Table 2). At the near neutral pH region, cathodic reduction of Cr(VI) becomes slower because of low H+ availability and formation of a Cr(OH)3 monolayer on the cathode surface. The efficiency of Cr(VI) was improved using Cr(VI)-reducing biofilms at the cathode (biocathodes). The Cr(VI)-reducing bacteria receive electrons from the poised cathode and catalyze the reduction of Cr(VI) to Cr(III) (Xafenias et al., 2013). It is still unclear whether the bacteria gain energy by shunting electrons from the poised cathode to Cr(VI). Improved Cr(VI) removal was observed when the cathode chamber of MFC was inoculated with a denitrifying mixed culture (Tandukar et al., 2009), a chromate contaminated soil (Huang et al., 2010) or Shewanella oneidensis MR-1 (Xafenias et al., 2013). Analysis of the Cr(VI) reducing microbial community from the cathode by 16S rRNA clone library preparation showed the dominance of phylotypes closely related to Trichococcus pasteuri and Pseudomonas aeruginosa (Tandukar et al., 2009). Xafenias et al. (2013) have inoculated the cathode with the S. oneidensis MR-1 strain and fed the MFC with lactate. The use of two electron donors such as a poised cathode and lactate allowed simultaneous use of both bioelectrochemical and microbial bioreduction for removing Cr(VI) from wastewaters. Recently, Wu et al. (2015) reported an efficient Cr(VI) reduction when an anode with a functional electroactive

Studies on recovery of metals from multi-metal solutions (Table 3) are presented in this section. The effect of integration of MEC followed by an electrolysis reactor was tested for recovering Zn, Pb and Cu from fly ash leachate (Tao et al., 2014). Acetic acid was used as the fly ash leaching agent (liquid to solid ratio of 14:1 (w/w), 10 h, and pH 1.0). Heavy metal ions from fly ash leachate with an integrated process showed removal efficiencies of 98.5%, 95.4% and 98.1% for Cu(II), Zn(II), and Pb(II), respectively. In addition to power output, Cu(II) was reduced and recovered mainly as metallic Cu(0) on the cathode. Zn(II) and Pb(II) were recovered in the electrolysis reactor. Simultaneous power generation (3.6 W/m2) and high Cd (90%) and Zn (97%) removal efficiencies were demonstrated in a single chamber air-cathode MFC (Abourached et al., 2014). Biosorption and sulfide precipitation are the major mechanisms reported for the heavy metal removal in the single chamber MFC. Incineration of municipal solid waste (MSW) is a common treatment method for volume reduction through elimination of the organic fraction along with recovering energy. High concentrations of various metals present in the remaining ashes can be leached out using acids. Modin et al. (2012) have demonstrated selective reduction of metals on a titanium cathode from a MSW leachate solution containing Cu(II), Pb(II), Cd(II) and Zn(II) in a dual-chamber BES using acetate as the electron donor in the anode chamber. Cu(II), Pb(II), Cd(II) and Zn(II) were sequentially recovered from the simulated MSW leachate solution applying external voltages of 0, 0.3, 0.5 and 1.7 V, respectively (Modin et al., 2012). Cu(II) was rapidly recovered from the solution with little change in the concentration of other metal ions when the BES was operated in MFC mode without applying any external voltage. Pb(II) was rapidly removed when the cathode potential was controlled

Li et al. (2008)

Zhang et al. (2012)

Modin et al. (2012)

1.6 at 204 mg/L Cr(VI), pH 2.5

0.97

0 V for Cu 0.3 V for Pb 0.5 V for Cd 1.7 V for Zn 1.0 V

at 0.51 V. But, some amount of Cd(II) was also removed during this stage. Cd(II) and Zn(II) were rapidly removed when the cathode potentials were controlled at 0.66V and 1.0V, respectively. Cu was recovered on the cathode with high efficiency (77.2%), while other metals were recovered with much lower efficiencies (1.2–5.3%) as a result of simultaneous H2 generation or reduction of dissolved O2. Cu and Zn were also deposited on the cathode surface as pure metals without detectable contamination of the other metals. However, when Pb was recovered, small amounts of Cu could also be detected in the deposits and when Cd was recovered, small amounts of Cu and Pb were also detected. Therefore, the BES technology offers a possibility for recovering metals from metallurgical waste streams and leachates with limited energy input as compared to conventional electrolysis.

Luo et al. (2014)

References

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Maximum power output (W/m2)/Applied voltage

112

Cu(II) was first recovered, followed by Ni(II) and then Fe(II); Some amount of Fe was also recovered during Ni(II) recovery stage. tMFC = two-chamber microbial fuel cell, tMEC = two-chamber microbial electrolysis cell.

CuSO4, FeSO4, NiSO4; 9 mM Fe(II) + 5 mM Cu(II) + 5 mM Ni(II), pH 2.8 Sodium acetate, 1 g/l tMEC, graphite brush anode, carbon cloth with platinum layer as cathode Copper, Iron, Nickel

Sodium acetate, 1.64 g/l

tMFC, carbon fiber felt for anode and cathode tMEC, carbon felt anode, titanium wire cathode Chromium, Vanadium, Cadmium, Copper, Lead, Zinc

Glucose, 0.81 g/l

67.9% for V(V) and 75.4% for Cr(VI) in 240 h Cu(II) was first recovered, followed by Pb(II), Cd(II) and finally Zn(II).

99.5% in 25 h from 204 mg/L;

K2Cr2O7; Cr(VI) in real electroplating wastewater; 204 mg/L; pH 2.5 K2Cr2O7, NaVO3; 250 mg/L Cr(VI) + 250 mg/L V(V); pH 2 CuCl2, Pb(NO3)2, CdCl2, ZnCl2; 0.8 g/l Cu(II), 0.4 g/l Pb(II), 0.8 g/l Cd(II), 0.3 g/l Zn(II) tMFC, carbon felt anode, graphite paper cathode Chromium

Sodium acetate, 1 g/l

Metal removal efficiency Metal salt, concentration, catholyte pH Electron donor, concentration BES configuration, electrode materials Metal(loid)

Table 3 Bioelectrochemical recovery of metals from multi-metal or real wastewaters using microbial fuel cells or microbial electrolysis cells.

8. Toxicity of metal ions in bioelectrochemical systems The concentration of heavy metals encountered in domestic wastewaters is often in the mg/L range, unlikely that they exert any notable inhibition on electroactive biofilms. But the concentration of metal ions can be much higher when BES are used for metallurgical wastes, process streams or industrial wastewaters. For removing and recovering metal ions, the anode and cathode chambers are separately fed with organic and metal containing aqueous solutions using dual-chamber MFCs or MECs. This approach possibly avoids the toxicity of metal ions on the bioanode microbial community, essential to the functioning of the BES. Also, there is a need to operate bioelectrochemical metal recovery systems for a long term to address ingress of metal ions into the anode chamber and their effect on bioanode performance. Besides, there is little understanding on the influence of metal ions on the structure and function of microbial communities along with the fate of metal ions in the anode chamber. Often both organic substances and metals are co-contaminants in waste streams. This is beneficial for BES, as supply of organic rich wastewaters is required for recovering metal ions from metallurgical wastes and process streams. The effect of a Cu(II) shock loading on the performance of an MFC was assessed by adding 25 or 125 mg/L Cu(II) in the influent for a few days (Feng et al., 2013). The carbon and nitrogen removal efficiencies were unaffected by the addition of 25 mg/L Cu(II) to the influent. Whereas supplementing the MFC with 125 mg/L Cu(II) in the influent for a week inhibited carbon and nitrogen removal efficiencies as well as voltage generation. A Cu(II) shock loading shifted the microbial community and decreased diversity and evenness due to selection of Cu(II) resistant strains. Nevertheless, nutrient removal and the voltage generation performance of the MFC were recovered within 30 days after stopping exposure to Cu(II). The function of the electroactive microbial community was unaffected by 25 mg/L of Cu(II), probably because of biofilm resilience to toxic pollutants. Conversely, the added Cu(II) was removed by biosorption and precipitation reactions. The maximum tolerable concentrations of Cd(II) and Zn(II) for an electroactive microbial community of a single chamber air cathode MFC were found to be fairly high at 200 and 400 lM, respectively (Abourached et al., 2014). This was because >90% of the supplied 200 lM Cd(II) and 400 lM Zn(II) were removed in the MFC through biosorption and sulfide precipitation. In spite of microbial reduction of selenite to elemental Se(0) in a single chamber MFC, selenite concentrations exceeding 25 mg/L in the feed affected the power output (Catal et al., 2009). The understanding of the toxic effect of metal(loid) ions on the microorganisms in MFCs or MECs is at present limited and additional systematic studies are needed to understand toxicity along with the fate of metal ions.

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