Recent advances in the recovery of metals from waste through biological processes

Journal Pre-proofs Recent advances in the recovery of metals from waste through biological processes Zhengsheng Yu, Huawen Han, Pengya Feng, Shuai Zhao, Tuoyu Zhou, Apurva Kakade, Saurabh Kulshrestha, Sabahat Majeed, Xiangkai Li PII: DOI: Reference:

S0960-8524(19)31646-3 https://doi.org/10.1016/j.biortech.2019.122416 BITE 122416

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Bioresource Technology

Received Date: Revised Date: Accepted Date:

26 September 2019 8 November 2019 10 November 2019

Please cite this article as: Yu, Z., Han, H., Feng, P., Zhao, S., Zhou, T., Kakade, A., Kulshrestha, S., Majeed, S., Li, X., Recent advances in the recovery of metals from waste through biological processes, Bioresource Technology (2019), doi: https://doi.org/10.1016/j.biortech.2019.122416

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Recent advances in the recovery of metals from waste through biological processes Zhengsheng Yua#, Huawen Hana#, Pengya Fenga, Shuai Zhaoa, Tuoyu Zhoua, Apurva Kakadea, Saurabh Kulshresthab, Sabahat Majeedc, Xiangkai Lia* a

Ministry of Education Key Laboratory of Cell Activities and Stress Adaptations, School

of Life Science, Lanzhou University, No. 222 Tianshuinan Road, Lanzhou, Gansu, 730000, People’s Republic of China b

Faculty of Applied Sciences and Biotechnology, Shoolini University of Biotechnology

and Management Sciences, Bajhol, Solan, Himachal Pradesh-173229, India c

Department of biosciences, COMSATS University, Park Road, Tarlai Kalan Islamabad,

Islamabad, 44000, Pakistan #

These authors contributed equally to this work

*Corresponding author: E-mail address: [email protected] Tel/ fax: +86 931 8912562

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Abstract Wastes containing critical metals are generated in various fields, such as energy and computer manufacturing. Metal-bearing wastes are considered as secondary sources of critical metals. The conventional physicochemical methods of metals recovery are energy-intensive and cause further pollution. Low-cost and eco-friendly technologies including biosorbents, bioelectrochemical systems (BESs), bioleaching, and biomineralization,have become alternatives in the recovery of critical metals. However, a relatively low recovery rate and selectivity severely hinder their large-scale applications. Researchers have expanded their focus to exploit novel strain resources and strategies to improve the biorecovery efficiency. The mechanisms and potential applicability of modified biological techniques for improving the recovery of critical metals need more attention. Hence, this review summarize and compare the strategies that have been developed for critical metals recovery, and provides useful insights for energy-efficient recovery of critical metals in future industrial applications. Keywords:

biorecovery,

critical

metals,

improvements

2

biological

techniques,

performance

1. Introduction The rising demand for critical metals has attracted great attention because of their high scarcity, uneven distribution, and low substitutability (Moss et al., 2011). Although the criteria for the criticality of metals varies worldwide and depends on national demands, there is a consensus that critical metals comprise rare earth elements (REEs), platinum group elements (PGMs), and other metals, including nickel (Ni), cobalt (Co), tellurium (Te), indium (In), gallium (Ga), molybdenum (Mo), and tungsten (W) etc. (Hennebel et al., 2015; Won et al., 2014). Among these metals, PGMs are representative of the high-scarcity group, 60% of the world’s cobalt stems from Congo, and 56% of the lithium comes from Chile (Zhang et al., 2017). The increasing demand for computer hard drives, permanent magnets, lamp phosphors, catalysts and rechargeable batteries accelerated the exploitation of PGMs and REEs, resulting in the generation of various metal-bearing solid and liquid wastes. Notably, there has been an explosive increase in waste electric and electronic equipment (WEEE); the generation of WEEE reached 44.7 million metric tons (Mt) in 2016, which equals 6.1 kg per inhabitant (Baldé et al., 2017). These E-wastes are estimated over 52 million metric tons by 2021, which are regarded as better recycling resources for critical metals (Işıldar et al., 2018), including silver (Ag), gold (Au), palladium (Pd), In, and REEs (Dias et al., 2018). The traditional methods (e.g., solvent extraction, ion exchange, coprecipitation, and crystallization) of recovery of critical metals from wastes are cost-inefficient or environment unfriendly (Xie et al., 2014; Yin et al., 2017). Moreover, these methods are unfeasible for the extraction of critical metals at low concentrations (Lo et al., 2014). For example, the industrial practices developed for the separation, purification, and pre-concentration of REEs are carried out in several pretreatment steps using strong acids or bases followed by complex extraction cycles using organic solvents (Quinn et al., 2015). These REE extraction methods are also energy-intensive and produce a larg amount of toxic secondary wastes, including thorium, uranium, hydrogen fluoride, and acidic wastewater, thus causing severe environmental pollution (Dodson et al., 2015). 3

Therefore, the development of alternative technologies that enable the efficient recovery of critical metals from wastes is highly desirable. Biological technologies offer obvious advantages because of their low cost and eco-friendliness. Universally, biosorbents (Abdolali et al., 2017), bioelectrochemical systems (Nancharaiah et al., 2015), bioleaching (Gu et al., 2018), and biomineralization (Lee et al., 2014) have been used to recover metals. Biosorbents showed high multiple-metal binding properties (Abdolali et al., 2015; Abdolali et al., 2014). The application of bioelectrochemical systems to metal removal and recovery was reported recently (Nancharaiah et al., 2015). The use of biomineralization contributed to the development of nanoparticles (Mandal et al., 2006). Although these biotechnologies in recovery of critical metals showed superiority, disadvantages were also existed. A relatively low selectivity always happened in biosorption process, and the application of BESs is still in lab-scale and low recovery rate occurred (Abourached et al., 2014; Qin et al., 2012), the mechanisms of biomineralization processes are not clear. Thus, it is necessary to ameliorate these technologies to obtain high recovery efficiencies. Improvements and modifications of the application of the technologies to the biorecovery of critical metals have attracted interest (Park et al., 2017). For example, a light-emitting diode (LED) powder-adapted Acidithiobacillus ferrooxidans showed a 96% and 60% bioleaching rate for nickel and gallium, respectively, while only 60% and 34% of these metals leached when using a non-adapted A. ferrooxidans (Pourhossein & Mousavi, 2018). Moreover, a combination of microbial fuel cell (MFC) and microbial electrolysis cell (MEC) also improved the recovery rate of critical metals (Shen et al., 2015). Besides, new methods or materials were applied to recover critical metals (Yuan et al., 2018). However, there is a lack of comprehensive and comparative analyses of the various biological recovery methods for guiding the development of suitable techniques for the recovery of targeted critical metals. It is necessary to compare these technologies and explore suitable strategies for recovery of critical metals at specific condition. The objectives of this review were to discuss the enhancement, mechanisms, current

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challenges, and future perspectives in the use of biorecovery technologies, and to explore the large-scale application of these techniques to waste management in the future. 2. Recovery of critical metals through biosorbents 2.1 Biosorbents Biosorption is defined as the removal or binding of substances from solution by bio-derived materials (Fig. 1), such as plant materials and microorganisms (e.g., algae, fungi, and bacteria) (Farooq et al., 2010; Ju et al., 2016). Biosorbents exhibited good performance for metal removal and recovery from industrial effluents from the metallurgical industry and electroplating over the past few decades (Volesky & Holan, 1995). The adsorption process has obvious superiority in the accessibility to renewable materials, simple processing, minimal sludge production, high metal uptake rates (even in trace conditions), and potential for regeneration and reusability (Vijayaraghavan & Yun, 2008). Generally, electrostatic interaction manipulates the mechanism of adsorption of metals onto biosorbents, rather than chemical reaction, ion exchange, and redox reaction (Cui & Zhang, 2008). In practical wastewater treatment, the competitive adsorption between critical metals ions and other ions and organic pollutants resulted in an obvious decline in adsorption efficiency and selectivity for non-living biosorbents, and the growth conditions contributed to a low reuse possibility for living cells as biosorbents (Vijayaraghavan & Yun, 2008). The developments and modifications of biosorbents overcoming these shortcomings which are essential for better application to actual wastewater treatment. 2.2 Improvements on adsorption capacity and selectivity Various biosorbents were developed recently to adsorb critical metals (Table 1). For non-living biosorbents, the introduction of functional groups (e.g., -NH2, -C=O, -COOH, and -OH) contributes to metal sorption (Wang & Chen, 2009). So, improvements on the density of functional groups would increase adsorption capacity. For example, Lysinibacillus sphaericus encapsulated into alginate matrix recovered 5

100% of Au (III) after three hours at an initial concentration of 60 ppm (Páez-Vélez et al., 2019). In some cases, NaOH- or hexadecyl-trimethylammonium bromide (CTAB)-modified bark powder of Mangifera indica improved dysprosium (Dy) (III) removal from aqueous solution (Devi & Mishra, 2019), with a resulting maximum adsorption capacity of 55.04 mg/g and 34.4 mg/g, respectively. NaOH pretreatment provides additional active binding sites (-COOH and -OH) on the bark powder, which yielded better Dy (III) removal than that observed for CTAB modification. A heat-treated E. coli cell suspension (90-100℃) realized 100% W recovery in 1 h, while the biosorption time of un-treated E. coli is more than 7 h. The zeta potential of the E. coli cells which increased from 5 mV (un-treated) to 19-24 mV (heat-treated), contributed to the improved biosorption performance. Further analysis with FTIR analysis, free amino acid, and LC-MS/MS revealed cell membrane of the E. coli isdissolved in heat treatment, resulting in an increase density of amino and carboxyl groups on the surface of E. coli cells. The enhancement of positively charged amino groups in acidic solution after heat treatment alters more tungsten anions adsorbed onto the cell surface (Ogi et al., 2016). The polyethyleneimine (PEI) modified cellulose nanofibril from tunicate showed the highest Pt adsorption capacity with 600 mg/g than PEI modified cellulose nanofibrils and PEI modified cellulose nanocrystals obtained from hardwood pulp, which is due to the highest negative charge and surface area of PEI modified cellulose nano-fibril from tunicate (Hong et al., 2019). Also, this biosorbent exhibit higher selectivity to Pt than other metals. A novel magnetic biosorbent was generated by immobilizing persimmon tannin (PT) onto Fe3O4@SiO2 microspheres (Fe3O4@SiO2@PT) showed high adsorption capacities (917.43 mg/ g) of Au (III) with the participation of phenolic hydroxyl group. The high selectivity of Fe3O4@SiO2@PT towards Au (III) was observed when Au (III), Pd (II), Zn (II), Cu (II) and Fe (III) coexist in solution. The reduction of Au (III) to metallic gold by Fe3O4@SiO2@PT realizes the convenient and efficient recovery of Au (III) from acidic multiply metal ions system (Fan et al., 2019).

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In contrast, extracellular proteins play a crucial role in the biosorption process of living-cell biosorbents. Similarly, an extracellular protein from Tepidimonas fonticaldi sp. AT-A2, displayed an outstanding adsorption capacity for gold (Au) (III) (9.7 mg Au/mg of protein). In the treatment of industrial wastewater obtained from a manufacturing factory of printed circuit boards (PCBs) (Han et al., 2017), the protein-based biosorbent also achieved a high adsorption capacity (1.45 mg Au/mg of protein) and removal efficiency was reached to 71%. Subsequently, the overexpression of the CueR protein on the surface of Saccharomyces cerevisiae cells significantly enhanced the silver (Ag) (I) binding capacity (Tao et al., 2016). The surface-engineered cells can specifically adsorb 2.799  mg/g Ag (I) and exhibit increased adsorption (87.3%) compared with the original cells. Overall, engineered S. cerevisiae-CueR improved the Ag (I) adsorption capacity and selectivity. A similar phage surface display technique also exhibited an affinity for Te of 100-fold and an affinity of Ni of 20-fold compared with the controls via the use of highly selective peptides (Braun et al., 2018). In addition, displaying EC20 on the surface of E.coli increased the density of amine, carboxyl and phosphate groups and the biorecovery of Pt (IV) reached to 112.67 mg/ g, which is 1.6-fold higher than original strain (Tan et al., 2019). In this regard, the expression of specific metal-binding proteins on the surface of cells using the surface display technology may be a promising strategy to improve adsorption ability. It is worth mentioning that the surface display technology offers obvious advantages in the recovery of REEs (Park et al., 2017). Lanthanide-binding tags (LBTs) comprise 15–20 amino acids and can bind terbium (Tb) (III) with high affinity (Sculimbrene & Imperiali, 2006). The expression of the fusion protein OmpA–LBT on the surface of E. coli increased the adsorption capacity to 28.3 ± 1.2 mg/g dry cell weight, which is twice high than that detected in the un-induced control (Park et al., 2017). During the treatment of different leachates (Lower Radical, Round Top Mountain, and Togo leachates), an enhancement of 2–10-folds in the adsorption efficiency of REEs was observed through the expression of LBTs on the cell surface of

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E. coli. The engineered E. coli strain selectively adsorbed REEs from the Lower Radical, Round Top Mountain, and Togo leachates, which only contained 0.9, 5.5, and 0.5% REEs by mass of the total metals, respectively. Little to no adsorption (<10%) to the engineered E. coli strain was observed for Ca, Ba, Zn, Mg, Na, K, Mn, and Rb from three leachates, despite their high abundance. Thus, LBT-displaying E. coli systematically enhanced the affinity and selectivity of REEs. Although the engineered E. coli was selective for REEs over other metals, its performance in batch conditions relied on the growth phase of living cells (Philip et al., 2000). To address its low reuse potential, the engineered E. coli harboring a fused protein (CsgA-LBT) for REE recovery via the secretion of functionalized curli fibers (Tay et al., 2018). This extracellular display of curli-LBT filters showed selective Tb (III)-binding capacity in the presence of 10-fold higher concentrations of competing metals (i.e., Al (III), Fe (III), Ni (II), and Ca (II)). The adsorbed fiber can be regenerated via simple washing with diluted acid, which allows the reuse of the bacteria for further material production. The curli-based filters are also more robust at thetreat conditions that might compromise the viability of cell-based sorbents or the structural integrity of other biopolymeric materials. Overall, pretreatment of biosorbents contributes to the availability of an increased number of metal-binding sites for enhancing adsorption capacity, while filters generated using the surface display technology favor high selectivity and recycling ability. The main advantages of non-living biosorbents are the low cost of raw materials and high adsorption capacity, while the relatively low selectivity is its shortcomings. Expressing metal-binding protein in living cells could improve the selectivity, while there are only a few metal-binding proteinsreported previously. Hence, more attention should be paid to the discovery of metal-binding protein. Most of the biosorbents were applied in lab-scale synthetic wastewater, the breakthrough in performance on industrial application would be the main challenges in future studies. 3. Recovery of critical metals through bioelectrochemical systems

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3.1 Bioelectrochemical systems The bioelectrochemical systems (BESs) usually contains MFC, MEC, microbial desalination cell, and other microbial electrochemical technologies (Logan et al., 2019). Over the past years, BESs have been fabricated to deal with metal removal and recovery (Nancharaiah et al., 2015). In which the MFC generates electricity, where the electrons derived during microorganism metabolism can reduce some metal ions in wastewater on the cathode (Li & Zhou, 2019). This novel MFC-mediated wastewater treatment can recover critical metals and produce additional electricity simultaneously. Unlike that of the MFC, MEC operation requires the use of external voltage, its application shifting from hydrogen production (Cheng & Logan, 2011) to methane production (Park et al., 2018), anaerobic digestion monitoring (Yu et al., 2018), and metal removal (Qin et al., 2012). However, the low and unstable metal-recovery efficiencies of MFCs or MECs hardly meet the requirements of real wastewater treatment systems. 3.2 Improvements in biorecovery rate Many

parameters

(e.g., physicochemical

factors,

pure

culture/microbial

communities, and fabrication) affect the recovery behavior of MFCs and MECs (Ali et al., 2019; Ho et al., 2017a). In the presence of oxygen, MFCs showed an increased recovery rate of tungsten (W) and molybdenum (Mo), by 2.4-fold and 1.3-fold, respectively (Table 2) (Wang et al., 2017). A possible explanation for this observation is that the addition of oxygen accelerates the electron-transfer efficiency and the generation of H2O2, resulting in enhanced reduction of W and Mo. Furthermore, the externally one-time addition of H2O2 also improved W and Mo deposition (by 2.4 and 86.3%, respectively) and increased the coulombic efficiency (CE) to 38.1%. In the MEC system, the combination of light irradiation and in situ H2O2 production significantly increased the W and Mo deposition rate (shifted to 81.5 and 97.2%, respectively) along with 82.5% CE (Wang et al., 2018b). Light irradiation altered the photocatalysis activity of W and Mo via a process in which previous W and Mo deposits catalyzed the subsequent reduction of W (VI) and Mo (VI) by photogenerated electrons. In addition,

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oxygen-exposing biocathode MFCs may exhibit a promising performance for the recovery of cobalt from spent lithium-ions batteries (Huang et al., 2015a). A nearly complete removal of W (97‒98%), Mo (98‒99%), and acetate (95‒96%) was obtained with an influent ratio of W:Mo:acetate of 0.5:1.0:24 mM and a hydraulic residence time of 2 days in a 40 L cylindrical single-chamber MEC (Huang et al., 2019a). More importantly, the resultant wastewater treated with MEC meets national wastewater discharge standards. Microbial communities are the major driving force of metal recovery in the MFC or MEC system (Nancharaiah et al., 2015; Wu et al., 2018). After 25 days (MFC-Cr and MFC-Cu) and 10 days (MEC-Cd) of acclimatized incubation (Huang et al., 2015b), the complete and selective metal reduction rates of Cr (VI), Cu (II), and Cd (II) reached 1.24 mg L−1 h−1, 1.07 mg L−1 h−1, and 0.98 mg L−1 h−1, respectively, which is 2.5, 2.9, and 3.6 times higher than that observed for the non-adaptive controls. 16S rRNA sequencing revealed that microbial communities with a lower diversity were observed in the adaptive groups vs. the control groups, and metal reduction had no significant impact on the microbial communities of adaptive groups at the phylum and genus levels (Huang et al., 2015b). Thus, the acclimatization of microbial communities is beneficial for the improvement of critical metal bioreduction and biorecovery from wastewaters. Another study first used MFC for the biorecovery of indium (In) from synthetic wastewater (Kim et al., 2018), and reported that over 90% of In (III) was removed after 14 days of MFC operation; moreover, the cathode carbon electrode generated amorphous and crystalline indium hydroxides. Moreover, the combination of MFC with MEC yielded great progress in the recovery of critical metals. For example, higher Co (II) removal rates (5.3 ± 0.4 mg L–1 h–1) was achieved at an initial Co (II) concentration of 40 mg/L using MECs with biocathodes and driven by MFCs (Shen et al., 2015). This recovery rate is 3.3 times higher than observed for MECs with abiotic cathodes and driven by MFCs. The MFC–MEC hybrid systems were also applied to the recovery of other metals, such as

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Cr, Pb, and Ni, with high removal rates of 32.7, 32.7, and 8.9 mg L−1 d−1 being achieved, respectively (Li et al., 2015b). Using the MEC-MEC coupled system together with adaptation strategies would be a potential approach to enhance the biorecovery rate of critical metals from wastewaters. The application of MFC and MEC on the recovery of critical metals achieved great progress in recent years. These technologies conduct removal of organic pollutants, recovery of critical metals and energy production simultaneously. However, biorecovery rates of critical metals are unstable and vary in larger range. The initial metal concentration, temperature, and pH value are the main influential factors in the performance of MFC and MEC, which play important role in the metabolic mechanism of microorganisms in MFC and MEC. Besides, the applied voltage used in MEC can affect the biorecovery rate of critical metals. So, the exploration of optimal conditions for practical applications of MEC and MFC on critical metals recovery from wastewater can improve the biorecovery rate and provides a new insight in the actual industrial wastewater treatment. 4 Recovery of critical metals through bioleaching 4.1 Bioleaching Bioleaching has been employed by various researchers to recover metals owing to a higher metal extraction rate from low-grade ores and complex resources (Gu et al., 2018; Xin et al., 2012). However, its application is still in its infancy. A variety of chemolithotrophic bacteria (e.g., Acidithiobacillus thiooxidans and A. ferrooxidans), cyanogenic bacteria (e.g., Chromobacterium violaceum, Pseudomonas sp., and Bacillus megaterium), and fungi (e.g., Aspergillus niger and Penicillium simplicissimum) (Gu et al., 2018) are involved in the mobilization of metals from E-wastes (Table 3). These microorganisms produce inorganic or organic acids or cyanide to perform bioleaching process. The biogenic cyanide from microorganism metabolism was more secure than conventional chemical processes due to extremely lower concentrations (Pollmann et al., 2018). In fact, bioleaching is only one first step in metal recovery, these critical metals

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present dilute leachates can be recovered via classical methods (e.g. precipitation, comentation, solvent extraction, ion exchange) (Li et al., 2013), newer approaches (membrane technologies and dialysis ) and biobased approaches (bio-adsorption, bioflotation, bio-reduction, microbial electrochemistry) (Hsu et al., 2019; Pollmann et al., 2018). Recently, much attention has been attributed to processes that allow the enhancement of the bioleaching rate from various wastes. 4.2 Improvements on the bioleaching rate Regarding the bioleaching process, pH adjustment significantly enhanced the bioleaching rate of Co and Ni from spent electric vehicle Li-ion batteries (Xin et al., 2016). The release efficiencies of Co and Ni increased from 43.5 to 96% and from 38.3 to 97%, respectively, because of the enhancement of the effect on cell growth. Similar results were observed for the recovery of Li (I) and Co (II) from lithium cobalt oxide by sulfur-oxidizing and iron-oxidizing bacteria (Wu et al., 2019b). The recovery of Li (I) and Co (II) by bacterial communities reached 100 and 99.3% at 72 h compared with the chemical leaching method (91.4 and 94.2%, respectively). In addition to pH adjustment, the presence of biochar led to a higher bioleaching efficiency of metals from E-wastes through the regulation of electron transfer from a Fe-mediated bacterial community to metals (Wang et al., 2016). Microorganisms act as the prerequisite of the bioleaching process (Gu et al., 2018; Latorre et al., 2016; Ma et al., 2019). Although cyanogenic bacteria have been widely used for the bioleaching of precious metals, screening of novel isolates may provide increased performance. P. balearica SAE1 isolated from an e-waste recycling facility exhibited a leaching rate of 68.5% for Au and of 33.8% for Ag from a 10 g/L pulp density under optimal conditions (Kumar et al., 2018). The cyanide production determines the recovery efficiency of precious metals. Yuan et al. found that maximum cyanide production by P. fluorescens strain P13 using optimized medium (6 g/ L tryptone and 5g/ L yeast extract) was 1.5 times higher that of LB medium. Furthermore, the relationship between cell growth and cyanide production revealed a novel kinetic

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model for predicting cyanide production (Yuan et al., 2018). Another bacterial strain, Cellulosimicrobium funkei, from cadmium- and arsenic-contaminated soil showed an efficient leaching rate of 70% for gallium recovery from thin-film GaAs solar cell waste (Maneesuwannarat et al., 2016). The high concentration of metals from bioleaching process inhibited bacterial growth and activity (Natarajan & Ting, 2015), while fungus exhibited a high metal-tolerance capacity (Xin et al., 2012; Zotti et al., 2014). In contrast to bacterial bioleaching, fungal bioleaching has several advantages, such as bioleaching at high pH, short lag phase (Moore et al., 2008), and secreting organic acids (Jadhav et al., 2016). Some fungus, such as Aspergillus genus, exhibited a high recovery of 90-95% from Zn–Mn battery and Ni–Cd battery via the formation of citric and oxalic acids (Kim et al., 2016), which was highly comparable or even better than bacterial or acid leaching (Biswal et al., 2018). In addition, a fungus strain Penicillium expansum can bioconcentrate Lanthanum (up to 390 ppm) and Terbium (up to 1520 ppm) from WEEE after three weeks of cultivation (Di Piazza et al., 2017), its relative mechanism affecting this process needs further studies. These evidence indicate that the exploration of novel strain resources is an effective strategy for improving bioleaching rates. During bioleaching, forming critical metals in the leaching solution can cause potential toxicity to microorganisms, thus adaptation strategy can contribute to improving microbial resistance (Ma et al., 2019). Some membrane proteins (transporter) and EPS alleviate the toxicity of metals via adsorbing on the surface of cell or increasing extracellular efflux. The adaptation of Acidithiobacillus ferrooxidans treated with varied concentrations of LED powder (5–25 g/ L) resulted in higher Fe (III) levels, cell amounts, oxidation-reduction potential, and lower pH vs. the non-adapted group (Pourhossein & Mousavi, 2018). The adapted A. ferrooxidans exhibited an increase of 1.6-fold leaching rate of nickel and gallium when compared to that non-adapted A. ferrooxidans. Such adaptation strategy achieved a recovery rate of 100% of indium from discarded liquid crystal displays by A. thiooxidans (Jowkar et al., 2018). Similarly, adapted fungi A. niger improved the organic acid production and leaching efficiency

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(Bahaloo-Horeh et al., 2018), the high metals recovery was achieved by leaching using biogenic acids rather than chemical leaching (Bahaloo-Horeh et al., 2016). A novel strategy of three-step (consortium construction, directed evolution, and chemostat selection) adaptive laboratory evolution was proposed to improve bioleaching efficiency. The results showed that iron recovery rate increased by 26and 55% at pH 1.5 and 0.75 respectively by using developed consortium than original consortium (Liu et al., 2019a). Pathak et al. proposed addition of catalysts can improve the dissolution and bioleaching rate (Pathak et al., 2017). With the addition of citric acid, Fe (II) and elemental sulfur, recovery rate of cobalt and copper by moderate thermophiles (A. caldus S2, Leptospirillum

ferriphilum

DX,

Sulfobacillus

acidophiles

TPY,

Ferroplasma

thermophilum L1) from high pressure acid leaching residue reached to 87.91and 58.52% with pulp density of 50 g/ L (Liu et al., 2019b). Subsequently, aerobic activated sludge was domesticated for leaching indium by incubation with Starky medium, 9 K medium, and a mix of the two media for 7 days, to induce microbial communities via the S-mediated pathway, Fe-mediated pathway, and mixed S- and Fe-mediated pathway (Xie et al., 2019). The bioleaching efficiencies of the S-mediated pathway reached approximately 100%, which was the highest among these three pathways (0% for the Fe-mediated pathway and 78% for the mixed pathway). A reasonable explanation for this observation is that Acidithiobacillus was the dominant genus in the S-mediated pathway communities, based on sequencing data of 16S rRNA gene. The two-step method is a more efficient metal mobilization process when bacteria attains its logarithmic phase in pure culture (Heydarian et al., 2018). This method was proposed for gold recovery from waste PCBs using Chromobacterium violaceum (Li et al., 2015a). Regardless of the high Cu content (268.17 mg/g) and low Au content (0.050 mg/g), pretreatment of waste PCBs with Acidithiobacillus ferrooxidans can remove 80% of Cu and other base metals at optimum conditions. Moreover, pretreatment increased the Au content of waste PCBs by 0.114 mg/g, accompanied by a biorecovery rate of Au that was two times higher than that of the control group. In addition, oxygen

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supplementation and the addition of nutritive salts (NaCl and MgSO4) enhanced leaching efficiency, which is consistent with previous research (Tran et al., 2011). Furthermore, pretreatment with a mixed culture (e.g., A. ferrivorans and A. thiooxidans, or P. fluorescens and P. putida) yielded a 1.6-fold improvement in gold recovery from waste PCBs (Işıldar et al., 2016). A novel step-wise indirect bioleaching method was proposed by addition of biogenic ferric to improve recovery efficiency of copper, nickel, and gallium from waste light-emitting diodes (LED) using adapted A. ferrooxidans. Compared to direct bioleaching, step-wise indirect bioleaching achieved 83, 97, and 84% recovery efficiency for copper, nickel, and gallium respectively at a pulp density of 20 g/ L and the leaching time reduced from 30 days to 15 days (Pourhossein & Mousavi, 2019). Bioleaching has been widely used for commercial metal extraction from mining ores (Panda et al., 2015). Recently, this technology was developed to recover metals from solid wastes. Various mature experience in commercial metal extraction can be used for metal recovery. Nevertheless, the relative low metal concentration and multiple metal toxicity exist in solid wastes hinder the application of bioleaching and decrease recovery efficiency. Different strategies were developed to solve these problems, while these strategies only focus on specific metals. For other critical metals, the feasibility of similar strategies should be investigated for more broad applications. 5 Recovery of critical metals through biomineralization 5.1 Biomineralization Biomineralization of critical metals is performed by microorganisms, usually via detoxification process. As shown in Table 4, the Phomopsis sp. XP-8 fungus selectively recovered around 80% of Au from electronic wastewater without any pretreatment. In simulated electronic water containing different concentrations of Au (III), K (I), Na (I), Ni (II), Ca (II), Co (II), Al (III), and Fe (III), this fungus exhibited a recovery ability of 68.07% (Xu et al., 2019). Moreover, the metal-reducing bacterium Shewanella algae recovered three PGMs (platinum (IV), palladium (II), and rhodium (III)) from the aqua

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regia leachate of spent automotive catalysts simultaneously (Saitoh et al., 2017). Two Fe (III)-reducing bacteria, Acidocella aromatica sp. PFBCT and Acidiphilium cryptum sp. SJH, were applied to recover palladium (Pd) from acidic Pd (II) solutions and spent catalyst leachates via bionanoparticles (Okibe et al., 2017). 5.2 Recent progress in the mechanistic study of biomineralization Despite the wide application of biomineralization to waste treatment and nanoparticle formation, the underlying mechanisms remain elusive. Research has confirmed the crucial role of nitrate reductase in the extracellular bioreduction of silver by P. aeruginosa JP1 (Ali et al., 2017). Purified nitrate reductase treatment accompanied by AgNO3 can transfer electrons from the nitrate molecule to silver, resulting in the formation of nanoparticles. A novel molybdoprotein-mediated mechanism was involved in aerobic Pd (II) reduction by E. coli mutants deficient in different oxidoreductase processes, using formate as the electron donor (Foulkes et al., 2016); this was distinct from the hydrogenase-mediated Pd (II) reduction pathway of anaerobic E. coli cultures described previously. Recently, the mechanism of gold biomineralization associated with extracellular polymeric substances (EPSs) of E. coli was elucidated (Kang et al., 2017), the Au (III) reduction process was mediated by the hemiacetal groups of the reducing saccharides of EPSs via the identification of the relevant functional groups. Such EPSs-mediated U (VI) reduction also exists in Shewanella sp. HRCR-1 (Cao et al., 2011). 5.3 Novel biomineralization isolates and processes Recently, novel microorganisms for the reduction of critical metals were isolated. The isolated Raoultella ornithinolytica and Raoultella planticola strain MoI showed optimal molybdate reduction ability when using glucose as the electron donor (Saeed et al., 2019). Moreover, tellurite reduction was observed intracellularly and extracellularly in the genus Raoultella and Escherichia isolated from wastewater (Nguyen et al., 2019). Tellurite-reducing photosynthetic bacterium, Rhodopseudomonas palustris strain TX618, exhibited a high recovery efficiency with wide variations in pH, temperature, light

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intensity, and initial tellurium concentration (Xie et al., 2018). In addition, the potential functions of biorecovery of critical metals in previously isolated microorganisms were explored. A keratin-degrading Bacillus sp. strain (khayat) was able to reduce molybdate to molybdenum in optimal conditions (pH, 5.8–6.8; temperature, 25–34°C) (Khayat et al., 2016). Moreover, Shinella sp. WSJ-2 was also found in the reduction of Te (IV) (Wu et al., 2019a). Several novel biomineralization processes of critical metals have been reported. The extracellular formation of Pd nanoparticles was observed after supplementation of Pd (II) (12–48 ppm) into the fermentation medium during the growth of Phanerochaete chrysosporium (Tarver et al., 2019). This extracellular biosynthesis of gold nanoparticles exists in A. foetidus-mediated biomineralization (Roy et al., 2016). As reported, a UASB reactor was used to perform tellurium recovery continuously from tellurite-containing wastewater (Mal et al., 2017). The exploration of the mechanisms of biomineralization of different critical metals provides insight on improving the production of metal nanoparticles and preferable applications in industry. The application of biomineralization for biorecovery of critical metals offer an effective way to produce nanoparticles. The metal nanoparticles showed good performance in application, such as Pd nanoparticles in chromate reduction (Ha et al., 2016) and azo dyes bioreduction (Wang et al., 2018a), Au nanoparticles in degradation of toxic dyes (Xu et al., 2019), etc. With exploration of novel biomineralization isolates and processes, more pathways of critical nanoparticles generation are elucidated. However, the detailed mechanisms are not clear. Therefore, mechanisms for better performance of nanoparticle production need be to be explored. 6 Challenges and Future prospects 6.1 Gaps and challenges in knowledge The application of biosorption, BESs, bioleaching, and biomineralization into the recovery of critical metals provides a potential for the recycling of these metals from different kinds of wastes. Although great improvements have been achieved using these

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strategies, gaps and challenges remain regarding their practical applications to the recovery of critical metals. The metal-binding protein exhibits high adsorption ability and selectivity. However, only few metal-binding proteins have been applied to biorecovery. Thus, more attention should be focus on the explorations and mechanisms of novel metal-binding proteins. The MFC and MEC exhibited high efficiencies for simultaneous recovery of critical metals and removal of organic pollutants, which is suitable for wastewater treatment. However, most of the applications of BESs on wastewater treatment are still in lab-scale. The functional microbial communities in biorecovery processes of critical metals are not clear. In addition, more studies are needed to explore the mechanisms underlying the exploitation of the electron transfer rate of BES. The mechanisms of biomineralization should be explored further in future studies. A better understanding of the mechanisms that link microorganisms to the reduction of critical metals would lead to the generation of nanoparticles with high quality in short time. The possibility of applying novel isolates to the biorecovery of critical metals needs to be verified. 6.2 Future prospects The combination of two or three technologies might be more efficient in terms of biorecovery. Combination of bioleaching with a biosorbent strategy yielded a > 86% Cu recovery from waste PCBs using USCT-R010 isolates for bioleaching and the dead biomass of A. oryzae and Baker’s Yeast as biosorbents (Sinha et al., 2018). The hybrid of ammonium thiosulfate (AT) and Lactobacillus acidophilus (LA) achieved 85% gold recovery from waste PCBs, the π-π interaction between AT and LA enhanced amide absorption bond and then lead to improvement of gold recovery (Sheel & Pant, 2018). MFC and bioleaching hybrid technologies also significantly improved the recovery rate of Cu, with 54% leaching and 78% recovery of Cu from the secondary copper tailings (Huang et al., 2019b). A similar concept was used for the recovery of Cu from copper sulfide minerals and enhanced Cu recovery, which was attributed to an MFC-mediated decrease in the pH value derived from the anodic sulfide/sulfur oxidation process

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(Huang et al., 2019d). The successful biorecovery of Cu provides a potential for recycling of critical metals from wastes via a combination of these technologies. Researchers investigated the integration of two stages, i.e., bioleaching and electrochemical extraction, for the recovery of Nd and La from monazite rock ore (Maes et al., 2017). The results showed that Nd was concentrated from 392 mg/L in the lixiviants to 880 mg/L after the electrochemical extraction process through utilization of a CEM under the effect of an electric field. Thus, the combination of these technologies might be a promising way to recover critical metals from wastes in the future. The exponential growth of emerging materials (e.g., nanomaterials) in a wide range of potential applications implies that various types of waste would become an additional secondary source of critical metals (Tan et al., 2017). The interactions between fungi and bacteria occurred in various environments (Deveau et al., 2018), new metal-reducing bacteria could be isolated through the utilization of fungi (Furuno et al., 2012). Furthermore, the co-culture of bacteria and fungi might be a potential way to deal with critical metals for high recovery rate (Deveau et al., 2018), such as REEs (Hopfe et al., 2017). 7 Conclusions The biorecovery of critical metals through biosorption, BESs, bioleaching, and biomineralization are potential approaches for waste management and metals’ recycling. The biosorption technique showed better efficiency and low-cost in recovery of metals from wastewaters as compare to other treatments while bioleaching is appropriate for solid wastes. The combination of these technologies has emerged as a promising application in biorecovery processes. Developments and modifications of biological strategies will provide high-efficient way to recover metals selectively from wastes. A better understanding of the mechanisms underlying these technologies will lead to the generation of more accurate technologies for the biorecovery of metals. Conflict of Interest The authors declare no conflict of interest.

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Acknowledgments The work was supported by National Natural Science Foundation grant (31870082); Gansu province major science and technology projects (No: 17ZD2WA017); Central Universities grant [grant number lzujbky-2018-103].

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29

Figure legends

Fig. 1 Bisorbents derived from various bio-materials and their mechanisms for the recovery of critical metals

30

Table 1 Different adsorbents for recovery of critical metals Critical metals

Biosorbents

Recovery rate %

Ability mg/g

Reus e cycle s

Du rati on

Estimated costs ($/ L)

Adsorptio n model

Mechanism s

Potential application

References

Au (III)

sugarcane bagasse

99

1497.5

NA

4h

0.5

Langmuir

hydroxyl

Lab scale

(Rubcumintara,

isotherm

2015)

model Au (III)

extracellular proteins of

100

9700

NA

1h

0.7

Langmuir

Cysteine and

71% recovery

Tepidimonas fonticaldi

and

histidine

rate from PCB

AT-A2

Freundrich

industrial

models

wastewater

(Han et al., 2017)

Au (III)

Bacillus subtilis

100

119

NA

4h

0.82

ND

Sulfhydryl

Lab scale

(Yu & Fein, 2017)

Au (III)

Tetradesmus obliquus

ND

169

NA

6h

0.5

Langmuir

NA

Lab scale

(Shen & Chirwa,

model Au (III)

Au (III)

Lysinibacillus sphaericus

100

NA

60%

encapsulated into alginate

after 3

matrix

cycles

Fe3O4@SiO2@PT

100

917

NA

3h

0.5

ND

2018) S-layer

Lab scale

protein

24 h

0.8

Langmuir

tannin

model

(Páez-Vélez et al., 2019)

Lab scale,

(Fan et al., 2019)

acidic multiply metal ions system

Ag (I)

Saccharomyces cerevisiae

80

2.799

NA

2h

0.5

Langmuir

CueR protein

with enhanced displaying

and

binding

CueR protein

Freundlich

Lab scale

(Tao et al., 2016)

Lab scale,

(Xu et al., 2017)

model Pd (II)

Providencia vermicola

NA

119

NA

3h

0.64 31

Langmuir

amine,

model

Pd (II)

polyethylenimine- coated

97.4

216.9

polysulfone Escherichia

>5

24 h

0.8

cycles

Langmuir

carboxyl,

potential

hydroxyl, and

multiple

phosphate

industrial

groups

wastewater

NA

Lab scale,

model

(Cho et al., 2016)

potential for

coli biomass composite

Pd(II) recovery

fiber

from acidic solutions

Pd (II)

Aspergillus

65.7-98.8

4.28

NA

24 h

0.7

Langmuir

NA

Lab scale

model Pt (IV)

Aspergillus

37.5-82.7

5.49

NA

24 h

0.7

Langmuir

icz et al., 2019) NA

Lab scale

model Pt (IV)

Providencia vermicola

NA

30.2

NA

3h

0.64

Langmuir

(Godlewska-Żyłkiew

(Godlewska-Żyłkiew icz et al., 2019)

amine groups

model

Lab scale,

(Xu et al., 2017)

potential multiple industrial wastewater

Pt (IV)

surface-displaying EC20 on

NA

239

NA

3h

2.3

Escherichia coli

Langmuir

EC20

Lab scale

(Tan et al., 2019)

NA

Lab scale,

(Hong et al., 2019)

and Redlich-Pet erson models

Pt (IV)

polyethyleneimine

NA

600

NA

24 h

1.1

modified cellulose nano

Langmuir model

fibril from tunicate

simulated spent automobile catalyst leachate

32

Dy (III)

NaOH-treated bark powder

92

55

NA

3h

0.7

of Mangifera indica Eu (III)

Bacillus thuringiensis

98

160

5

50 h

0.64

cycles

Langmuier

Hydroxyl and

model

carboxyl

Langmuir

negative

model

charge density

Lab scale

(Devi & Mishra, 2019)

Lab scale

(Pan et al., 2017)

(Arunraj et al., 2019)

of functional groups Eu (III)

Saccharomyces cerevisiae

97.5

25.9

4

3h

0.69

cycles

Langmuier

amine,

Lab scale in

model

carboxyl,

fluorescent

hydroxyl, and

lamp phosphor

polysaccharid

powder

e Eu (III)

Acutodesmus acuminatus

NA

174.2

NA

9h

2.6

Langmuier

phosphate

model

groups,

Lab scale

(Furuhashi et al., 2019)

carboxyl group Eu (III)

Saccharomyces cerevisiae

NA

19.41

immobilized on the

>4

1h

0.63

cycles

Langmuier

NA

Lab scale

(B et al., 2018)

carboxylic

Lab scale

(Hadjittofi et al.,

model

chitosan matrix Sm (III)

activated biochar from

100

350

NA

24 h

0.5

NA

Opuntia Ficus Indica Sm (III)

Turbinaria conoides

moieties 99.2

151.6

>5

10 h

0.1

cycles

Langmuir

negatively

and

charged

Redlich–Pet

binding sites

erson model

such as

2016) Lab scale

(Vijayaraghavan et al., 2017)

carboxyl REEs

LBT engineered E. coli

42-92

28.3

NA

0.5

0.64

h 33

NA

lanthanide

Lab scale,

binding tag

leachates from

(Park et al., 2017)

metal-mine tailings and rare earth deposits Tb (III)

curli-LBT E. coli

~70

54.3

contin

cont

uous

inuo

0.64

NA

lanthanide

Lab scale,

binding tag

bauxite residue,

us

(Tay et al., 2018)

phosphogypsum , and metallurgical slag

W (VI)

Heat-treated Escherichia

100

297.8

NA

1h

0.2

NA

coli Co (II)

Sargassum glaucescens

free amino

Lab scale

(Ogi et al., 2016)

Lab scale

(Esmaeili & Beni,

acids 91

NA

NA

1.5

0.1

h

Langmuir

NA

and Dubinin–Ra dushkevich models

NA: Not available. The estimated costs were calculated based on medium, instruments and other chemicals used in researches.

34

2015)

Table 2 Applications of MFC and MEC in critical metals recovery Critical metals

Membrane type/ Applied voltage (V)

Concentration mg/L

Recovery rate %

Duration

Estimated costs ($)

Voltage/ Applied voltage (mV)

Power density

Columbic efficiency (%)

References

Ag (I)

CEM

2000

92.5

24 h

265

410

5396mW/m3

19.89

Ag (I)

CEM

2000

96

24 h

265

85

3795mW/m3

12.74

Ag (I)

CEM

1000

99.3

24 h

265

441

8258mW/m3

21.61

Ag (I)

AEM

50

98.2

24 h

288

340

348mW/m3

NA

Ag (I)

AEM

4000

92.3

24 h

288

970

1930mW/m3

NA

Ag (I)

PEM

500

67.8

72 h

56

650

3006mW/m3

8.73

Pt (IV)

CEM

16.88

37.9

24 h

47

710

844.0mW/m2

NA

Mo (VI)

CEM

200

86.4

6h

2

230

280mW/m2

66.4

Co (II)

CEM

30

93

6h

18

340

1500 mW/m3

NA

Co (II)

AEM

40

96.35

18 d

48

863

11340mW/m3

28.74

(Ho et al., 2017a) (Ho et al., 2018) (Ho et al., 2017b) (Lim et al., 2015) (Lim et al., 2015) (Ali et al., 2019) (Liu et al., 2019c) (Wang et al., 2017) (Huang et al., 2015a) (Huang et al., 2019c)

MFC

35

In (III)

PEM

100

90

14 d

72

520

NA

NA

(Kim et al., 2018)

Co (II)

0.2

50

31.8

6h

11

200

NA

NA

(Wang et al., 2015)

Co (II)

0.5

50

72.2

6h

11

500

NA

NA

(Wang et al., 2015)

Co (II)

NA

40

79.5

6h

11

340

490mW/m3

54.2

(Shen et al., 2015)

Mo(VI)

0.5

96

98

30 d

122

210

NA

NA

(Huang et al., 2019a)

Mo(VI)

0.3

200

97.2

4h

48

300

NA

82.5

(Wang et al., 2018b)

MEC

NA: Not available. The estimated costs were calculated based on medium, instruments and other chemicals used in researches.

36

Table 3 Bioleaching of critical metals Critical metals

Wastes sources

microorganisms

Pulp density (g/ L)

Estimated costs（$/ L）

Duration

Au

waste PCBs

Pseudomonas balearica SAE1

10

0.64

7d

68.50

Produces hydrogen cyanide

Au

waste PCBs

Chromobacterium violaceum

200 mesh

0. 4

7d

70.6

Produces hydrogen cyanide

Au

waste PCBs

Aspergillus niger

11

0.5

32 d

56

Produces hydrogen cyanide

Ag

waste PCBs

Pseudomonas balearica SAE2

10

0.64

7d

33.80

Produces hydrogen cyanide

Ag

waste PCBs

Sphingomonas sp. MXB8/C

0.5

0.52

35 d

54

form an Ag mirror

37

Recovery rate %

Mechanisms

Potential application Lab scale, recovering precious gold from e-waste Lab scale, recovering gold from waste PCBs Lab scale, bioleaching of metals from PCBs of cell phones or computers Lab scale, recovering precious silver from e-waste Lab scale, recover metals from WEEE

References

(Kumar et al., 2018)

(Li et al., 2015a)

(Argumedo-Delira et al., 2019)

(Kumar et al., 2018)

(Díaz-Martínez et al., 2019)

Au

waste PCBs

Orthopsilosis MXL20

0.5

0.52

35 d

44.20

form precipitates

REE

FP

Komagataeibacter hansenii

0.85

0.46

14 d

7.90

Dissolve the REE-compounds

REE

FP

Zygosaccharomyces lentus

0.85

0.46

14 d

0.23

Dissolve the REE-compounds

REE

FP

Kombucha culture

0.85

0.32

14 d

7.20

Dissolve the REE-compounds

Sm

EPs

Aspergillus niger

0.0009

0.56

14 d

66.70

Produce Siderophores

La

EPs

Aspergillus niger

0. 362

0.56

14 d

51

Produce Siderophores

Ce

EPs

Aspergillus niger

0.866

0.56

14 d

50.10

Produce Siderophores

38

Lab scale, recover metals from WEEE Lab scale, Leaching REE from FP powder Lab scale, leaching REE from FP powder Lab scale, leaching REE from FP powder Lab scale, extract rare element from phosphorites Lab scale, extract rare element from phosphorites Lab scale, extract rare element from phosphorites

(Díaz-Martínez et al., 2019) (Hopfe et al., 2017)

(Hopfe et al., 2017)

(Hopfe et al., 2017)

(Osman et al., 2019)

(Osman et al., 2019)

(Osman et al., 2019)

Ga

GaAs

Cellulosimicrobium funkei

0.086

0.64

30 d

91

Amino acids of C. funkei are involved in Ga leaching

Ga

LED

Acidithiobacillus ferrooxidans

20

0. 4

30 d

60

NA

Co

arsenide

Acidithiobacillus ferrooxidans and Leptospirillum ferriphilum

20

0.46

8d

92

NA

Co

EV LIBs

Acidithiobacillus thiooxidans and Leptospirillum ferriphilum

10

0.48

11 d

96

Co

EV LIBs

100

0.2

2h

53.2

Co

LiCoO2

Acidithiobacillus ferrooxidans and Acidithiobacillus thiooxidans Leptospirillum ferriphilum and

Contact mechanism between the cathodes material and cells NA

15

0.4

7h

94.2

39

Hypothetical mechanism:

Lab scale, bioleaching of gallium from thin film GaAs solar cells Lab scale, direct metals bioleaching from LED powder Lab scale, new approach to recycle semiconductor wastes Lab scale, recover the valuable metals from EV LIBs Lab scale, recovery of metals from LIB waste Lab scale, commercial

(Maneesuwannarat et al., 2016)

(Pourhossein & Mousavi, 2018)

(Giebner et al., 2019)

(Xin et al., 2016)

(Boxall et al., 2018)

(Wu et al., 2019b)

Sulfobacillus thermosulfidooxidans

Co

high pressure acid leaching residue

Acidithiobacillus caldus S2, Leptospirillum ferriphilum DX, Sulfobacillus acidophiles TPY, Ferroplasma thermophilum L1

50

0.2

15 d

87.9

In

discarded LCDs

Acidithiobacillus thiooxidans

16

0.4

12 d

100

40

electron transfer or extracellular polymeric substances that is beneficial for Co recovery citric acid and sulfate dissolve CoFe3(SO4)2(OH)6, Co (II) dissolved at low pH

NA

application of the bioleaching of lithium-ion batteries. Lab scale, industrial application of the trace heavy metals removal from high-pressure acid leaching residue Lab scale, recover In from discarded LCDs

(Liu et al., 2019b)

(Jowkar et al., 2018)

In

discarded LCDs

Microbial community dominated by Acidithiobacillus

15

0.4

8d

100

secrete enzymes and extracellular polymeric substances

Lab scale, potential biological leaching of In from liquid crystal displays

(Xie et al., 2019)

PCBs: Printed Circuit Boards; FP: fluorescent phosphor; EPs: Egyptian Phosphorites; GaAs: gallium arsenide; LED: Light Emitting Diode; EV LIBs: electric vehicle Li-ion batteries; LCDs: liquid crystal displays; NA: Not available. The estimated costs were calculated based on medium, instruments and other chemicals used in researches.

41

Table 4 Biomineralization of critical metals Critica l metals

Sources

Microorganis ms

Concentratio n mg/ L

Recover y rate %

Estimated Costs ($/ L)

Durat ion

Mechanisms

Potential applicability

References

Au

HAuCl4

300

36

0.64

24 h

HAuCl4

300

19.6

0.64

24 h

Au

HAuCl4

197

NA

0.5

4h

Lab scale, the recovery of Au ions from industrial waste Lab scale, the recovery of Au ions from industrial waste Lab scale.

Au

HAuCl4

100

68

0.4

6h

Ag

AgNO3

Pseudomonas aeruginosa

NA

NA

0.5

NA

sodium lactate as electron donor sodium lactate as electron donor certain protein molecules Adsorb and reduce Au (III) to immobilized AuNPs nitrate reductase mediated mechanism

(Wu & Ng, 2017)

Au

Shewanella xiamenensis BC01 Shewanella oneidensis MR-1 Aspergillus foetidus Phomopsis sp. XP-8

Pt

PtCl4

Escherichia coli

1950

74

0.4

112 day

Pt

K2PtCl4

halophilic cultures

100

98

0.72

3-21 h

42

Bacterial biomineralizatio n Reduction to Pt (0)

Lab scale, complex aqueous solutions, AuNPs can degrade toxic dyes biomineralization and biotransformation processes and biogeochemical cycles for silver and other heavy metals. biomineralize platinum

recovery of platinum from process and waste

(Wu & Ng, 2017) (Roy et al., 2016) (Xu et al., 2019)

(Ali et al., 2017)

(Shar et al., 2019) (Maes et al., 2016b)

Pt

K2PtCl6

halophilic cultures

100

97

0.72

3-21 h

Pt

industrial process streams Pd(II)

halophilic microbial community Escherichia coli

80

99

0.72

24 h

100

100

0.64

0.5 h

platinum(IV ), palladium (II) and rhodium (III) Pd2+

Shewanella algae

0.2wt% PGMs

> 95

NA

2h

Enterococcus faecalis

210

100

0.64

48 h

Acidocella aromatica PFBCT Phanerochaete chrysosporium

100

100

0.5

95 h

48

81.5

0.4

48 h

Pd

Pt, Pd,

Pd

Pd

Pd

spent catalyst leachate K2PdCl6

43

Possible thinner peptidoglycan layer NA

molybdoprotein -mediated mechanism, using formate as the electron donor oxidation of organic acid salts such as lactate and formate sodium formate as electron donor Monosaccharid es or formate as electron donor Amide groups

streams Lab scale

(Maes et al., 2016b)

Lab scale, from process stream

(Maes et al., 2016a)

Lab scale

(Foulkes et al., 2016)

Lab scale, from spent automotive catalysts

(Saitoh et al., 2017)

Lab scale

(Ha et al., 2016)

Lab scale, leaching lixiviant

(Okibe et al., 2017)

Lab scale, catalyze Heck reaction of styrene and iodobenzene

(Tarver et al., 2019)

Eu

Thermus scotoductus SA-01 Phoma glomerata

76

100%

0.5

10 h

Produce Eu2(CO3)3

Lab scale

(Maleke et al., 2019)

10

98.5

2.6

30 d

Lab scale

(Liang et al., 2019)

Lab scale, synthetic wastewater

(Mal et al., 2017)

Lab scale

(Espinosa-Orti z et al., 2017) (Wu et al., 2019a) (Nordmeier et al., 2018)

Te

sodium tellurite

Te

tellurite

anaerobic granular sludge

10

98

110

42 d

Te

tellurite

10

39.8

1.2

8d

Te

tellurite

63.8

96.4

0.64

44 h

NA

Mo

Na2MoO4

Phanerochaete chrysosporium Shinella sp. WSJ-2 Clostridium pasteurianum BC1

intracellular uptake or interaction with surface biomolecules loosely-bound EPS of granular sludge NA

100

88

0.5

4-5 h

NA

Lab scale, simulated lake water Lab scale, degradation of MeO

NA: Not available. The estimated costs were calculated based on medium, instruments and other chemicals used in researches.

44

Highlights 1.

Improved biotechnologies facilitates the recovery of critical metals

2.

Surface-displaying of metal binding proteins enhances selectivity and recovery

3.

Two-steps bioleaching contributes to recovery of critical metals

4.

Combination of biological techniques is a promising strategy for high recovery

45

46