Industrial development has led to generation of large volumes of wastewater containing heavy metals, which need to be removed before the wastewater is released into the environment. Chemical and electrochemical methods are traditionally applied to treat this type of wastewater. These conventional methods have several shortcomings, such as secondary pollution and cost. Bioprocesses are gradually gaining popularity because of their high selectivities, low costs, and reduced environmental pollution. Removal of heavy metals by sulfate-reducing bacteria (SRB) is an economical and effective alternative to conventional methods. The limitations of and advances in SRB activity have not been comprehensively reviewed. In this paper, recent advances from laboratory studies in heavy metal removal by SRB were reported. Firstly, the mechanism of heavy metal removal by SRB is introduced. Then, the factors affecting microbial activity and metal removal efficiency are elucidated and discussed in detail. In addition, recent advances in selection of an electron donor, enhancement of SRB activity, and improvement of SRB tolerance to heavy metals are reviewed. Furthermore, key points for future studies of the SRB process are proposed.

  • Recent researches on the mechanism of sulfate-reducing bacteria (SRB) heavy metal removal are introduced.

  • Elements affecting SRB activity are reviewed.

  • The suitabilities of SRB electron donors are analyzed and some cheap electron donors are listed.

  • Strategies for enhancing SRB activity are outlined.

  • Ways to improve SRB tolerance to heavy metal toxicity are discussed.

Graphical Abstract

Graphical Abstract
Graphical Abstract

Wastewater containing heavy metals is produced by electronics and metallurgy, automobile, steel forging, mining and electroplating industries. Discharge of contaminated wastewater has received attention worldwide because it can lead to environmental pollution. After entering microbial cells, heavy metals supersede essential cations in cellular structures and combine with functional parts in microorganisms, like thiols, to replace enzyme prosthetic groups' metal cofactors, inactivating enzymes, and change microbial characteristics such as cell assembly and morphology (Chakravarty & Banerjee 2008; Guo et al. 2017a). Metal toxicity is one of the great challenges for biological treatment of waste and wastewater.

In addition to their concentrations, the bioavailabilities and reactivities of heavy metals are dependent on their forms, which are influenced by several factors, for instance, oxidation–reduction potential (ORP), pH, temperature, the ionic strength of the medium, and the existence of chelating agents (e.g., ethylene diamine tetraacetic acid, humic acids and amino acids). Heavy metal bioavailability can be increased through a synergistic effect between heavy metals and acidity, which greatly enhances the toxicity. It has been proved that complexed metal ions are less toxic to microorganisms than free metal ions, and many methods for reducing metal toxicity by transformation of a free ion state have been reported (Hughes & Poole 1991; Van Nostrand et al. 2005; Qian et al. 2015; Wakeel et al. 2020). For instance, some organic acids (such as citrate) can change metal free ion states through chelating with them, and some inorganic anions (such as OH or Cl) can form complexes with metals. In addition, some cations (such as iron, magnesium, and calcium) can compete with cationic metals for anionic sites on cell surfaces, and then prevent cells from coming into contact with heavy metals. In addition, metals in different oxidation states, such as As5+/As3+ and Se6+/Se4+, show different levels of toxicity towards microorganisms. In addition to heavy metal toxicity, hydrogen sulfide (H2S) release by sulfate reduction in wastewater is highly corrosive and corrodes pipes. Therefore, it is of utmost importance to remove heavy metals and sulfate from wastewater before discharging it to the environment.

Various conventional physicochemical methods have been used to remove heavy metals and sulfate from wastewater, including reverse osmosis, soil washing, coagulation–flocculation, ion exchange, adsorption, membrane technology, electrolysis, thermal extraction, cementation, direct solvent extraction, and precipitation of intermediates (Kikot et al. 2009, 2010; Iakovleva et al. 2015). These methods perform well for heavy metal removal but generate high volumes of sludge. Biological techniques are economical and efficient alternatives to chemical methods because of their advantages over traditional processes, including low cost, high selectivity and better removal efficiency when treating sewage with a low metal concentration (Kikot et al. 2009, 2010). Using biogenic sulfide mediated by sulfate-reducing bacteria (SRB) to precipitate heavy metals is an attractive way to treat the heavy metals and sulfate-containing wastewater (Blazquez et al. 2016). In recent decades, the SRB process has been widely used for treatment of industrial wastewater from pulp, textile, mining, tanning and paper industries, which produce wastewater with high concentrations of heavy metals and sulfate (Liamleam & Annachhatre 2007b; Rampley et al. 2020). The SRB are a group of anaerobic prokaryotes that can tolerate low oxygen levels and exist in a wide range of lakes, marshes, underground pipelines, paddy fields, industrial wastewater, petroleum deposits and other anoxic habitats. They grow over a wide pH range (5.0–9.0). At low metal concentrations, the SRB process has a higher efficiency for heavy metal precipitation than other processes, and it is easier to separate the metals from the wastewater because the solubility is several orders of magnitude lower than that of hydroxides (Kikot et al. 2009). Metal sulfides are dense and have dewatering and good settling properties. Effective precipitation caused by SRB occurs over a broad pH range and valuable metals can be recycled from the precipitated sulfides (Gallegos-Garcia et al. 2009). In summary, the SRB process can remove organic compounds and heavy metals simultaneously from wastewater, and reduce the sulfate concentration to a level much lower than that obtained with any chemical method. However, the SRB system is usually operated in sub-optimal conditions, which limits the removal efficiency. This may be caused by several factors, such as the acidity, ORP, concentrations of sulfides, temperature, ratio of chemical oxygen demand (COD)/SO42−, the solid retention time (SRT) and the hydraulic retention time (HRT). The performance is also affected by the toxicities of heavy metals and excess sulfide, a lack of sufficient sulfide to precipitate heavy metals, growth conditions, SRB activity, the availabilities and quantities of carbon sources, the diversity and abundance of microbial species, and the synergistic effect between microorganisms. This paper reviews advances in the laboratory studies on understanding of the mechanism of heavy metal removal by SRB, factors affecting the microbial activity and removal efficiency, selection of the electron donor, enhancement of SRB activity, improvement of SRB tolerance to heavy metals and SRB biomass disposal methods. And then, the limitations existing in the SRB process are pointed out.

SRB with different respiration types (litho-autotrophic, autotrophic, and heterotrophic) have been observed under anaerobic conditions (Hussain et al. 2016). Autotrophic SRB can utilize CO2 as a carbon source and H2 as an electron donor to obtain energy for growth, whereas heterotrophic SRB can utilize a broad range of substrates (e.g., alcohols, hydrogen, organic acids and fatty acids (FAs)). Macromolecular substrates should be degraded by fermentative bacteria to produce low-molecular-weight organic matter firstly, which then serves as the SRB carbon source. Some low-molecular-weight organic matter (such as lactic acid) can only be partially oxidized to acetic acid (Sahinkaya et al. 2007). When lactic acid is used as carbon source of SRB, the reaction is as follows (Equations (1)–(3)) (Tasaki et al. 1993):
formula
(1)
formula
(2)
formula
(3)
The bicarbonates are produced during the degradation (Equations (1) and (3)) and they can neutralize the pH of the solution (Wang et al. 2013a). And the pathway of sulfate reduction is shown as follows (Equations (4)–(6)). Sulfate is a thermodynamically stable oxidation form of sulfur, which cannot be directly reduced to sulfide by SRB (Keller & Wall 2011):
formula
(4)
formula
(5)
formula
(6)
formula
(7)

APS: adenosine 5′-phosphosulfate; AMP: adenosine monophosphate; M: metal ion; MS: metal sulfide precipitation.

In this process, the formation of sulfide (such as HS) can reduce the pH of the solution. The acidity can be neutralized by bicarbonates generated by electron donor oxidation (Equations (1) and (3)) (Wang et al. 2013a). Thus, most reactors are operated efficiently without pH control (Sanchez-Andrea et al. 2014). Bio-produced H2S can react with dissolved heavy metals (e.g., Cd2+, Cu2+, Ni2+, Pb2+, U6+, and Sb5+) to form insoluble metal sulfide precipitates (Equation (7)). And the precipitates can be removed efficiently from the wastewater due to their favorable densities, low solubility, and good dewatering and settling properties (Thi Quynh Hoa et al. 2015).

In addition to the formation of metal precipitates, the separation of precipitates is also crucial for the SRB process. With the extension of the operation time of a biological system, the surface of sludge will be covered by many sulfides, which will affect the process of action. Sulfides exposed to water/air are oxidized directly, which is beneficial to metal leaching and recovery (Kousi et al. 2018). Valuable metals can be recovered in the form of precipitates through pH control because metal sulfides have different solubilities at different pH values. The separation of metal precipitates from wastewater can maintain high production rate of sulfides under steady-state conditions. As mentioned in Figure 2, with the fluctuation of pH, heavy metals can be graded precipitated (Johnson & Hallberg 2005; Cao et al. 2009; Yang et al. 2020a). The sustainability of the metal removal biological process depends on the fate of sludge to a great extent. The sludge rich in metal sulfides in a sulfate reduction bioreactor was characterized in detail, including chemical composition, granulometry and mineralogical properties (Kousi et al. 2018). This is very conducive to our further study on the SRB process.

Besides the metal precipitation by biologically produced sulfides, there are other heavy metal removal effects in the SRB system, such as the precipitation of carbonates and hydroxides related to the SRB cell surfaces and the auxiliary function of secreted extracellular polymeric substances (EPS), that aid heavy metal removal (Goncalves et al. 2007; Xie et al. 2009; Kiran et al. 2017; Paul et al. 2020). In a previous study, the high removal efficiency observed for Mn was explained by (oxy)hydroxide but not metal sulfide precipitation because of the low correlation of Mn removal with sulfate reduction and the low solubility product constant (ksp) of metal sulfide precipitates (Miao et al. 2018). In addition to these biochemical reactions, organic functional groups, which exist in the SRB process, play an important role in fixing heavy metals such as mineral nucleation and metal sorption at sites on cell surfaces and extracellular structures or polymers (Kiran et al. 2017). EPS produced by SRB contain nonpolar amino acids (mass fractions of 12.0% alanine, 11.0% glycine, and 7.0% Phe) and acidic amino acids (mass fractions of 14% glutamic acid (Glu) and 13% aspartic acid (Asp)), which can promote the precipitation of metal sulfides due to the large negative charges in Glu and Asp (Qi et al. 2019). As shown in Figure 1, large negative charges in Glu and Asp contribute a lot to interactions between EPS and heavy metal cations (Qi et al. 2019). An unparalleled biosynthesis mediation function of EPS is endowed by their unique acidic amino acids and characteristic nonpolar structure. It has been observed that the main components of EPS which react with heavy metals are tryptophan-like substances. In addition, different components of EPS own different capacities to adsorb heavy metals; for example, the binding sites of proteins and polysaccharides can adsorb Zn2+, whereas only polysaccharides can adsorb Cu2+ (Wang et al. 2014). Furthermore, the toxicity of metal sulfides toward SRB can be decreased by EPS firmly adhering to the surfaces of metal sulfides, which is important for their application in bio-treatment of wastewater containing heavy metals (Bao et al. 2010). Besides EPS, active heavy metals can react with some organic ligands so as to alter metal existing forms. Furthermore, microbial enzymes can change active states of some heavy metal ions into inactive ones, which can reduce metal toxicity (Sahinkaya et al. 2011; Sheng et al. 2011).

Figure 1

Mechanisms of heavy metal removal by SRB (HMs: heavy metals).

Figure 1

Mechanisms of heavy metal removal by SRB (HMs: heavy metals).

Close modal

pH value

SRB can be divided into two categories: neutrophilic and eosinophilic. The optimum pH range of neutrophilic SRB is 7.0–7.8 (Kikot et al. 2010). The maximum rate of sulfate reduction is achieved in the range of pH 7.0–7.5 (Sharma et al. 2014). Inhibition is observed when the pH is lower than 5 or higher than 9, and there is no activity when the pH is below 2 (Janyasuthiwong et al. 2016). Eosinophilic SRB can withstand and even flourish at pH values between 2 and 4 (Zhao et al. 2017). Figure 2 shows the interactions between different pH values and SRB activity. The presence of H2S and acetate decreases the pH, whereas production of HCO3 increases the pH. Fluctuations in the pH affect SRB activity seriously, which in turn greatly affects the consumption of sulfate and the production of acetate and H2S. Therefore, pH control greatly affects heavy metal removal in the SRB process. For the treatment of acid mine wastewater by an anaerobic packed-bed reactor, the pH of influent water is maintained at 7 by adding alkaline substances. Before the enrichment of SRB, the precipitation of metal is caused by the added alkaline substances. Then, the metal precipitation is mainly caused by their combination with sulfides produced by SRB, because Fe, Cu, Ni, and Zn cannot be precipitated well when pH is 7, but can be precipitated completely and efficiently only when pH is higher than 9.5 (Sharma et al. 2014). Cu is removed first, followed by Zn, Ni, and Fe, with Mn last on the basis of the solubilities (ksp) of the metal sulfides (Dev et al. 2017). It has been proved that the removal of Mn is dependent on the pH value (higher than 8) of the solution but not the precipitation of sulfides (Song et al. 2012). Oxidized Mn can be reduced to a less soluble form by Fe, so the removal efficiency of Mn is also dependent on the Fe concentration of the influent (Song et al. 2012).

Figure 2

Influence of pH fluctuations on SRB process.

Figure 2

Influence of pH fluctuations on SRB process.

Close modal

The mechanisms of the pH influence are summarized in Table 1. H2S, mentioned in Table 1, is toxic to SRB and does not easily form precipitates with heavy metals, and is present at acidic pH values. A low pH value can intensify the toxicity of H2S and organic acids, because organic acids and H2S are in their un-dissociated forms and not charged at low pH, which means they can go into the cell membrane easily and then make the proton motive force uncouple and the cytoplasm acidify. Moreover, an influx of more un-dissociated acid will be caused by dissociation of un-dissociated acids in the circumneutral internal cell cytoplasm so as to affect SRB growth (Gyure et al. 1990; Kimura et al. 2006; Koschorreck 2008; Yang et al. 2020b). The pH also greatly influences the size, separation, precipitation, solubility and quality of the metal sulfide, and even the charge and binding character of the metal (Cao et al. 2009; Sanchez-Andrea et al. 2014; Kefeni et al. 2017). As the pH increases, bio-sorption and bio-precipitation play important roles in the immobilized SRB system. When the pH is beyond the optimum range, the influences of bio-sorption and bio-precipitation decrease (Park et al. 2008; Pinto et al. 2011; Zhang et al. 2018b). Furthermore, many microbial metabolic mechanisms can be affected by the variation of pH, for instance, dissociation and homeostasis of electron donors (Janyasuthiwong et al. 2016). pH can determine how the treatment proceeds and affect the energy performance of electron donors, so control of it is essential for utilization of electron donors. It is reported that nonionic substrates, such as glycerol, hydrogen, sugars and alcohols, are more suitable for fermentation at low pH values than at other pH values (Nancucheo & Johnson 2014).

Table 1

Effects of pH on SRB process

Affected aspectsInfluences in detailReferences
Competition between SRB and MB At alkaline pH values, acetotrophic SRB outcompete with acetotrophic methanogenic bacteria (MB).
At acidic or neutral pH values, MB are dominant. 
Das et al. (2015)  
pH = 6.75 and 7.75, SRB outcompete methane-producing archaea. Dai et al. (2017)  
Sulfide speciation At acidic pH values: H2S; at basic and neutral pH values: HS and S2−Janyasuthiwong et al. (2016)  
Recovery of metals Graded precipitation of metal sulfides. Johnson & Hallberg (2005), Cao et al. (2009), Yang et al. (2020a)  
Existing form of metal Forming different states of metal (metal complexes or non-complexed metals). Kaksonen & Puhakka (2007), Zhang et al. (2018b)  
Microbial metabolic mechanisms Dissociation of electron donors and homeostasis. Janyasuthiwong et al. (2016), Nancucheo & Johnson (2014)  
Production of more available short chain fatty acids for SRB by fermentation at alkaline pH values (at 10.0). Chen et al. (2013)  
Affected aspectsInfluences in detailReferences
Competition between SRB and MB At alkaline pH values, acetotrophic SRB outcompete with acetotrophic methanogenic bacteria (MB).
At acidic or neutral pH values, MB are dominant. 
Das et al. (2015)  
pH = 6.75 and 7.75, SRB outcompete methane-producing archaea. Dai et al. (2017)  
Sulfide speciation At acidic pH values: H2S; at basic and neutral pH values: HS and S2−Janyasuthiwong et al. (2016)  
Recovery of metals Graded precipitation of metal sulfides. Johnson & Hallberg (2005), Cao et al. (2009), Yang et al. (2020a)  
Existing form of metal Forming different states of metal (metal complexes or non-complexed metals). Kaksonen & Puhakka (2007), Zhang et al. (2018b)  
Microbial metabolic mechanisms Dissociation of electron donors and homeostasis. Janyasuthiwong et al. (2016), Nancucheo & Johnson (2014)  
Production of more available short chain fatty acids for SRB by fermentation at alkaline pH values (at 10.0). Chen et al. (2013)  

From the perspective of thermodynamics, sulfate reduction becomes energetically more favorable at low pH values (as shown in Equation (5)), because the existence of hydrogen ions facilitates the reaction. At pH 3, the change in Gibbs free energy of the reaction is −198 kJ/mol but only −152 kJ/mol at pH 7, which shows that the energy gain is higher at low pH than high pH (Meier et al. 2012). Larger increase in cell numbers and a higher sulfate reduction rate are caused by higher acidity (Meier et al. 2012). Although low pH values have several advantages as mentioned above, the growth of SRB cells is inhibited significantly. As shown in Figure 3, most energy is used to support various pH homeostasis strategies due to the large pH gradient between extracellular and intracellular environments, including passive mechanisms, such as changes in lipid composition and expression of positively charged surface proteins, and active mechanisms, such as proton pumps and amino acid decarboxylases. Little energy obtained from sulfate reduction is used for cell growth (Meier et al. 2012). Low pH is the most crucial factor which affects SRB activity seriously (Kikot et al. 2010; Sheoran et al. 2010). Most SRB bioreactor applications use water recycling or ‘off-line’ bioreactor units and separation of precipitates to prevent SRB from coming into direct contact with acidic water. And many efforts have been made to isolate and enrich acidophilic SRB that can operate at lower pH values (Sanchez-Andrea et al. 2013; Sanchez-Andrea et al. 2015; Antsiferov et al. 2017; Willis et al. 2019).

Figure 3

Inhibition effect of low pH value on SRB growth.

Figure 3

Inhibition effect of low pH value on SRB growth.

Close modal

Sulfide concentrations

Anaerobic SRB can reduce sulfate and metabolize sulfur-containing amino acids to produce sulfides, which exert various functions in the SRB process. The ability of SRB to generate sulfides takes an important role in determining the metal removal efficiency. An SRB system, which can produce sulfides abundantly, has sustainable highly efficient metal precipitation even after dilution several times (Thi Quynh Hoa et al. 2015). Moreover, sulfide has a strong reducing ability due to its low valence state of sulfur, so it can reduce some oxidized metals and alleviate their toxicity. For example, the toxicity of Cr6+ can be decreased because Cr6+ is reduced by sulfide but not removed by precipitation (Sheng et al. 2011). However, because sulfide is a toxic substance, it is best to contain the sulfide concentration as much as possible. Studies have used the metal to sulfide ratio (M/S2−) as an index to measure the ability of SRB to remove heavy metals from wastewater and the quantity of sulfides (Villa-Gomez et al. 2015). With a low M/S2− ratio (<1), reduction of sulfate produces abundant sulfides which can efficiently precipitate heavy metals, but also decrease the growth rate of SRB because a strongly toxic environment for microorganisms is created. By contrast, with a M/S2− ratio of >1, the toxicity of residual metals decreases the efficiencies of metal removal and sulfate reduction (Villa-Gomez et al. 2015; Kiran et al. 2017). SRB growth is inhibited initially and then increases when the heavy metals and sulfate are removed through precipitating (Kaksonen et al. 2004). The toxicity of excessive sulfide to SRB can be reduced by biochar (Oliveira et al. 2020). In summary, both an excess and lack of sulfide can be detrimental to heavy metal removal. The effects of sulfide on SRB system are as follows. Sulfides can be present in different forms, such as H2S, HSand S2−, which present different level of toxicity to SRB. The main toxic form of sulfide is un-dissociated H2S; high concentrations of H2S can inhibit the SRB metabolic processes significantly but not completely stop them (Kushkevych et al. 2019b). Firstly, H2S is a neutral molecule, which can permeate through the cell membrane into the cytoplasm without specific receptors, and then act as a de-coupler, affect sulfur assimilation, adjust the intracellular pH, damage DNA and denature primary proteins (e.g., certain enzymes for maintaining well-balanced metabolism) (Figure 4) (Reis et al. 1992; Pol et al. 1998; Kaksonen & Puhakka 2007; Dai et al. 2017; Kushkevych et al. 2019a). H2S can inhibit cytochrome oxidase to transport electron donors from respiratory substrates to molecular oxygen so as to block microbial metabolism (Kushkevych et al. 2019a). Secondly, H2S causes the accumulation of acetic acid, which can lead to a high COD-containing effluent (Oyekola et al. 2009), because acetoclastic bacteria are more susceptible to H2S toxicity than are other microorganisms. Studies showed that the toxicity of H2S is dependent on the bacterial species because the hydrogen-utilizing SRB are more tolerant to high H2S concentrations than are the acetoclastic SRB group (Kaksonen et al. 2004; Icgen & Harrison 2006; Oyekola et al. 2009). A high COD/SO42− ratio is good for the growth of acetogenic bacteria instead of acetoclastic strains, which have have high resistance to H2S (Wang et al. 2008). The lack of acetoclastic SRB produces an effluent with a low pH and a high COD because of the accumulation of acetic acid. Thirdly, H2S can precipitate some valuable metals that are needed by SRB, such as iron. Because of concern about the tolerance of SRB to H2S toxicity, researchers have demonstrated that more complex electron donor molecules result in higher tolerance of SRB (Cao et al. 2012).

Figure 4

Cytotoxic mechanism of H2S on SRB at low pH.

Figure 4

Cytotoxic mechanism of H2S on SRB at low pH.

Close modal

Temperature

Most SRB are mesophilic, and some are psychrophilic and thermophilic. In an immobilized mesophilic SRB system, the optimum temperature for effective elimination of heavy metals is 37 °C (Kim et al. 2016). Temperature higher than 35 °C can inhibit the SRB activity because of bacterial inactivation and protein denaturation (Hao et al. 2014). The optimal growth rate of psychrophilic SRB is 7–18 °C, although the living environment temperature is −1.8 °C. However, the maximum rate of sulfate reduction is obtained when the temperature is 2–9 °C higher than their optimal growth temperature (Knoblauch & Jorgensen 1999). The best growth rate of thermophilic SRB is obtained at temperature 65–70 °C (Krukenberg et al. 2016) and moderately thermophilic SRB are obtained at temperature 50 °C (Cha et al. 2013). Reports have confirmed that the microbial growth rate is closely related to the temperature and drops sharply outside the optimum temperature range (Okabe et al. 1992; Hao et al. 2016; Zhang et al. 2018b; Jung et al. 2019). A low temperature was detrimental to the efficiency of passive biochemical reactors and resulted in low removal efficiencies for heavy metals and sulfate, and decreased alkalinity and SRB activity. Furthermore, the removal of heavy metals and sulfate decreased from 70–90% in summer (14–18 °C) to 0–39% in winter (approximately 5 °C) (Nielsen et al. 2018). Different operating processes, which are usually classified as suspended and immobilized biomass systems, can be adapted to different temperature ranges. When the temperature is lower than 15 °C, the removal efficiency of heavy metals in the immobilized system is significantly higher than that in the suspended system (Hao et al. 2016).

The mechanisms of the influence of temperature on SRB activity are discussed below. The rates of de-protonation and protonation of functional groups, together with the activity of SRB, increase at higher temperatures (Bajpai et al. 2004). High temperatures can reduce the solubility of toxic H2S in wastewater (Kaksonen & Puhakka 2007). Furthermore, studies have shown that the temperature can greatly affect competition between SRB and methanogenic bacteria (MB), and increases in the temperature can help SRB to outcompete MB in wastewater treatment (Kaksonen & Puhakka 2007). Reduction of sulfate is quicker than methane production when a high temperature (55–65 °C) is applied to a mesophilic system which contains SRB and MB (Omil et al. 1997). When considering the microorganisms, the cellular lipid content is decreased under a sub- or supra-optimal growth temperature. Changes in the temperature will change the average structural compositions of membrane lipids such as acyl and alkyl chains but not change the proportions of different lipid classes (Vincon-Laugier et al. 2017). Temperature can change lipid ordering, the rotational and lateral diffusion of proteins and the resistance of the membrane to shear forces so as to influence the membrane fluidity (Sperotto et al. 1989). Microorganisms adjust their FA composition and the nature of the head group of their phospholipids so as to control the phase transition temperature of membrane lipids and keep adequate membrane properties in response to changes in temperature (Russell 1984; Ernst et al. 2016). Adjusting the proportion of branched (classically iso/anteiso) FAs and the chain length and degree of unsaturation (addition/removal of double bonds or rings) of the acyl chains are the main adaptive mechanisms (Koga 2012). Changing the carbon chain lengths of phospholipid FAs is a classic way for microorganisms to adapt to temperature changes. Higher temperatures will lead to longer carbon chain lengths and higher melting points, which will decrease the membrane fluidity, and vice versa (Denich et al. 2003; Mykytczuk et al. 2007). For both mesophilic and thermophilic organisms, unsuitable temperatures will lead to changes in enzymatic production/activity and gene regulation, and formation of different lipid and protein assemblages (Aerts et al. 1985).

Ratio of COD to sulfate

As well as the nature of the organic substrate, the removal efficiencies of heavy metals by SRB may be associated with the amount of the substrate. Studies have shown that changes in the organic substrate to sulfate ratio (COD/SO42−) greatly affect the removal efficiency of heavy metals by mixed SRB cultures (both COD and sulfate are in mg/L). Theoretically, under standard conditions, reduction of 1 g of SO42− needs 0.67 g of COD. The COD/SO42− ratio decides the removal efficiencies of heavy metals and sulfate through affecting the competitiveness between SRB and other microorganisms, because some carbon sources are common for both their growth. When the ratio of COD/SO42− is under the theoretical value of 0.67, all the electrons will flow to SO42−. SRB competition for the common electron donors with other microorganisms will be intensified when the ratio is >0.67 (Dar et al. 2008). SRB can outcompete MB at the ratio of <1.7 because more electrons are used for sulfate reduction (Das et al. 2015). But at ratio of >2.7, MB outcompetes SRB, whereas when the ratio is between 1.7 and 2.7, active competition exists between SRB and MB (Mulopo & Schaefer 2013). Research has shown that less acetate is found in effluent with a ratio of 8 than with other ratios, which could be explained by the use of some electrons from acetate for methane production (Das et al. 2015). High COD/SO42− ratios are conducive to biomass production, and several authors have observed acetate accumulation in SRB systems with COD/SO42− ratios of >1.5, even with organic substrates other than lactate (Martins et al. 2009b; Oyekola et al. 2009). High COD/SO42− ratios could favor the growth of the acetogenic bacteria instead of acetoclastic strains and improve the resistance of the culture to H2S because acetoclastic strains are more susceptible to high H2S concentrations. High ratios also increase the chemical sulfate reduction and metal precipitation, probably as metal sulfides (Icgen & Harrison 2006). Studies have been conducted to determine how different COD/SO42− ratios affect the performance of reactors. Until reaching a maximum, the higher COD/SO42− ratio leads to a higher sulfate removal efficiency, and then the sulfate removal efficiency declines with further increase in the ratio whether MB exists together or not (Choi & Rim 1991; O'Reilly & Colleran 2006; Ren et al. 2007; Pina-Salazar et al. 2011).

HRT/SRT

The HRT is recognized as a parameter that can affect microbial activity and the reactive mixture composition, which might result in physical and biological inhibition of SRB, and even failure of the bioreactor. The optimum HRT varies with different treatment processes and pH environments. A study conducted in a laboratory-scale packed-bed bioreactor showed that the most suitable HRT under neutral conditions was 6 h, whereas under acidic conditions, the suitable HRT was 20 h (Aoyagi et al. 2017). This paper summarizes the different effects of different HRT values on SRB in different reactors, as shown in Table 2. In a biochemical passive reactor (a common method to treat acid mine wastewater by using biological, chemical and physical processes in natural environment), HRT greatly affects the dynamic of a microbial community (Skousen et al. 2017). A low HRT is not conducive to the growth and activity of microorganisms, because a high and constant input of acidic wastewater can wash out biomass in bioreactors (Mulopo & Schaefer 2013). Meanwhile, operation of fluidized-bed reactors is limited by the decreased pH caused by a lack of alkalinity, incomplete oxidation of carbon sources and accumulation of acetate (Kaksonen et al. 2004). Moreover, input of dissolved oxygen by rapid flow makes the ORP higher than −100 mV, which decreases the relative sulfate reduction and abundance of SRB, destroys the SRB living anaerobic environment and decreases the heavy metal removal efficiency (<60%) (Zheng et al. 2014; Gomez et al. 2015; Vasquez et al. 2018). The growth of microorganisms is favored and degradation of organic nitrogen and organic carbon is enhanced when the HRT is kept long (4 days) (Vasquez et al. 2018). So a sufficient quantity of available low-molecular-weight organic matter is provided for SRB growth.

Table 2

Effects of HRT on SRB process

Type of reactorType of sewageValue and influence of HRTReferences
Anaerobic membrane bioreactor Municipal wastewater HRT = 2.2 h, SRT = 60 d, SRB biomass accounts for less than 2% of total biomass. Mei et al. (2018)  
Anaerobic sequential batch reactors Sulfate-rich wastewater HRT = 12 h: sufate reducing rate = 75%, SRB outcompetes MB;
HRT = 36 h: MB outcompetes SRB. 
Moon et al. (2015)  
Laboratory-scale methane fermentation reactor Synthetic peptone wastewater HRT = 1 d, MB stops methane production, SRB grows and functions. Kwon & Nakasaki (2015)  
Up-flow anaerobic sludge bed Sulfate-rich wastewater HRT = 6 h, sufate reducing rate = 30%. Jing et al. (2013)  
Up-flow anaerobic sludge bed source-diverted blackwater HRT = 2.2 h, COD removal rate > 80%, SRB outcompetes MB. Gao et al. (2020)  
Up-flow anaerobic sludge bed Acid mine drainage HRT: (24 h → 16 h), sulfate removal rate increased to 92.1%;
HRT = 24 h, sulfate removal rate < 80%. 
Cunha et al. (2019)  
Bioreactor Acid mine drainage HRT = 4 d, the relative abundance of SRB increased, and the residual sulfide increased;
HRT = 1 d, it affects the anaerobic environment and is beneficial to the existence of acidophilic fossil nutrient microorganisms. 
Vasquez et al. (2018)  
Bioreactor Acid mine drainage HRT = 4 d, it increases residual sulfide, worsens effluent quality and affects SRB activity;
HRT = 1 d, the biomass was washed away, and the increase of dissolved oxygen resulted in the increase of ORP and the decrease of metal removal efficiency. 
Vasquez et al. (2016)  
Type of reactorType of sewageValue and influence of HRTReferences
Anaerobic membrane bioreactor Municipal wastewater HRT = 2.2 h, SRT = 60 d, SRB biomass accounts for less than 2% of total biomass. Mei et al. (2018)  
Anaerobic sequential batch reactors Sulfate-rich wastewater HRT = 12 h: sufate reducing rate = 75%, SRB outcompetes MB;
HRT = 36 h: MB outcompetes SRB. 
Moon et al. (2015)  
Laboratory-scale methane fermentation reactor Synthetic peptone wastewater HRT = 1 d, MB stops methane production, SRB grows and functions. Kwon & Nakasaki (2015)  
Up-flow anaerobic sludge bed Sulfate-rich wastewater HRT = 6 h, sufate reducing rate = 30%. Jing et al. (2013)  
Up-flow anaerobic sludge bed source-diverted blackwater HRT = 2.2 h, COD removal rate > 80%, SRB outcompetes MB. Gao et al. (2020)  
Up-flow anaerobic sludge bed Acid mine drainage HRT: (24 h → 16 h), sulfate removal rate increased to 92.1%;
HRT = 24 h, sulfate removal rate < 80%. 
Cunha et al. (2019)  
Bioreactor Acid mine drainage HRT = 4 d, the relative abundance of SRB increased, and the residual sulfide increased;
HRT = 1 d, it affects the anaerobic environment and is beneficial to the existence of acidophilic fossil nutrient microorganisms. 
Vasquez et al. (2018)  
Bioreactor Acid mine drainage HRT = 4 d, it increases residual sulfide, worsens effluent quality and affects SRB activity;
HRT = 1 d, the biomass was washed away, and the increase of dissolved oxygen resulted in the increase of ORP and the decrease of metal removal efficiency. 
Vasquez et al. (2016)  

Competition between MB and SRB is affected by HRT control. A high HRT intensifies the competition between them, because enhanced production of H2, a common substrate for them, is observed when the HRT is controlled long (Jing et al. 2013). A low HRT can increase the concentration of dissolved oxygen in wastewater and stimulate bacteria to produce high concentration of volatile fatty acids and ammonium ions, which is conducive to the formation of SRB (Kwon & Nakasaki 2015). In addition, in an inverse fluidized-bed reactor and up-flow anaerobic sludge reactor, it was reported that low HRT value is beneficial to improve sulfate reduction rate, which is also the precondition for economic feasibility and industrial application of the sulfur reduction process (Reyes-Alvarado et al. 2018; Gao et al. 2020). In order to investigate the effect of HRT on sequential batch and continuous anaerobic digestion reactors, it was found that operation of a sequential batch reactor improved the activity of MB and SRB and the biomass retention rate (Jung et al. 2017).

SRT is another important parameter in the SRB process. When SRT was shortened from 40 days to 5 days, sulfide production decreased by 33.7%, indicating that 40 days was more favorable for SRB enrichment compared with 5 days (Huang et al. 2020). In a sequencing batch anaerobic–aerobic membrane bioreactor, the reduction rate of SRB and the oxidation rate of COD decreased with the decrease of SRT (from 60 d to 30 d) (Yurtsever et al. 2017). In the sequencing batch reactor, the reduction of SRT (from 1.2D to 0.68d) inhibited the activity of SRB (Saha & Sinha 2018). In an anaerobic membrane bioreactor, 64.4% COD degradation rate was gained in an ultra-high SRT (140 days), 45.9% of which was removed by SRB (Seco et al. 2018). Therefore, a long SRT is necessary for SRB enrichment.

ORP

Redox potential is used to reflect the macro redox properties of all substances in aqueous solution, which is an important factor affecting the activity of SRB. Research shows that a reducing environment with ORP ≤ −100 mV is beneficial to maintain the activity of SRB, but an oxidizing environment with ORP > −100 mV can inhibit SRB activity (Wang et al. 2005). ORP is affected by the existence of redox substances in water, such as nitrate, nitrite, and zero valent iron. In many literature reports, zero valent iron (ZVI) can promote the activity of SRB partly due to the reduction of ORP in solution caused by ZVI, which can create a more favorable environment for SRB sulfate reduction (Kumar et al. 2014, 2015; Zhang et al. 2016a). The existence of nitrate can keep the ORP value at a high level so as to create an oxidizing environment, which is not conducive to the sulfate reduction by SRB. When treating wastewater containing U(VI) and nitrate, the nitrite formed will reoxidize and reactivate the precipitated uranium (Wu et al. 2006). It was shown that before the complete removal of nitrate, the generation of oxidation intermediates increased the ORP of the solution and hindered the bio-reduction of oxidized metal U(VI) (Yi et al. 2007). SRB are microaerophilic microorganisms whose respiratory rate increases with the decrease of oxygen concentration or stops after repeated oxygen addition (Dannenberg et al. 1992). Some practical operations will also regulate oxygen dosing by controlling ORP parameters to reduce the sulfide toxicity in the anaerobic treatment of sulfate-rich synthetic wastewater on-line (Khanal & Huang 2003).

The cost and availability of the carbon sources (electron donors) is the greatest challenge among all the limitations in application of the SRB process (Bertolino et al. 2014). The following two points are necessary to be taken into consideration when choosing appropriate carbon sources: (1) the availability of carbon sources for SRB to completely reduce sulfate and other pollutants in wastewater while increasing the removal efficiency of sulfate and heavy metals, (2) the cost of carbon sources consumed for reduction of sulfate to sulfides (Vanhouten et al. 1994). Sulfate-rich sewage usually has a high COD or concentration of organic compounds, which can serve as electron donors for SRB (Cao et al. 2012). However, most types of industrial wastewater (except for several kinds of wastewater rich in organic matrix, such as food processing wastewater) cannot successfully support the growth of SRB because of their low dissolved organic carbon content, especially the acid mine drainage (Li et al. 2016). Therefore, it is necessary to add essential electron donors during the treatment process to keep the basic ratio of COD/SO42− of 0.67, so as to ensure the complete reduction of sulfate (Liamleam & Annachhatre 2007a). The availability of added carbon sources determines the performance of the microbial process; some cannot be completely degraded by SRB, which leads to the high COD of the effluent. Therefore, selection of a suitable organic substrate is necessary for the SRB process and greatly affects application of the process on an industrial scale (Barbosa et al. 2014).

Although high sulfate removal performance has been observed using low-molecular-weight compounds as direct carbon sources, the compounds (e.g., lactate and ethanol) are too expensive. Instead, various agriculture residues or plant materials, such as rice/maize straw, compost, leaf mulch, organic-rich soil, oak chips, oak leaves, reed canary grass and its hydrolysate, animal manure, silage, peanut shells, algae, and wood waste, have been applied as carbon sources for SRB in laboratory- and field-scale acid mine drainage treatment (Choudhary & Sheoran 2012). In sulfate-reduction bioreactors, organic substrates mentioned above can also serve as a porous material and filtration media to support SRB growth and metal removal (Al-Abed et al. 2017). If the buffering capacities of those organic substrates are not enough, adding lime or paper mill waste (Neculita et al. 2007), mussel shells (McCauley et al. 2009) or steel slag (Batty & Younger 2004) is a typical solution. Because of the low cost and accessibility of organic substrates, many studies have investigated optimization of their use. Various inexpensive substrates have been evaluated for removal of sulfate and heavy metals (Table 3).

Table 3

Low cost substrates for SRB reported in the literatures

Carbon sourceWastewater typeBioreactor typeTemperaturepH(t0)TimepH(final)Sulfate removal: rate, percentageHeavy metals removalReferences
Chicken manure Acidic coal refuse Culture medium 30 ± 2 °C 5.5 20 d 6.6 –,71.2% Fe (99.7%), Mn (91.1%), Ni (40.8%), Zn (95.6%), Cu (93.2%), Cd (68.3%) Zhang & Wang (2014)  
Straw with ethanol Acidic saline drainage Culture medium 18–30 °C 4.5 50 d 5.1–5.9 0.1054–0.1110 mmol/(dm3·d),– Co (95%), Pb (96%), La (98.6%) Santini et al. (2010)  
Algae Acid rock drainage Permeable reactive barriers 24 ± 1 °C 4.0 123 d 6.5 –,80% Cu (>99.5%) Ayala-Parra et al. (2016)  
Microalgae – Immobilized SRB beads 30 5.5 45 d – –,72.4–74.4% Cu (91.7–98.2%) Li et al. (2018b)  
Phalaris arundinacea plant material hydrolyzate Mine wastewater Fluidized-bed bioreactor 35 35 d – 2.2–3.3 g/(L·d),– Fe (99%, 0.84 g Fe/(L·d)), Zn (99%, 15 mg Zn/L) Lakaniemi et al. (2010)  
Sweetmeat waste fractions Acid mine drainage Culture medium 37 7.3 ± 0.2 56 d – –,70% – Das et al. (2013)  
Tannery effluent – Up-flow anaerobic sludge blanket (UASB) 27 7.5 20 d 7.4–7.9 0.6 g/(L·d) 80% – Boshoff et al. (2004)  
Tannery effluent – Stirred tank reactor 27 7.5 33 d 7.6–8.0 025 g/(L·d), 78% – Boshoff et al. (2004)  
Tannery effluent – Trench reactor 27 7.5 30 d 7.9–8.2 0.4–0.5 g/(L·d), 72.12% (s.d. ± 11.72) – Boshoff et al. (2004)  
Chitinous materials Mining influenced water Sulfate-reducing bioreactors – 2.48 ± 0.06 400 d – 2.03–3.97 mmol/(dm3·d),– Cd (97.38%), Zn (99.45%) Al-Abed et al. (2017)  
Crab-shell chitin complex Mine-impacted water Laboratory bottles 20 ± 1 °C 2.95 50 d 6.5–6.7 17.8 mgSO42−/(L·d),– Al (100%), Mn (>73%), Fe (96%) Robinson-Lora & Brennan (2010)  
Crab shell and spent mushroom compost Mine-impacted water Continuous-flow columns – 2.5 ± 0.1 181 d 6.0 – Ni (10–31%), Zn (25–66%), Co (41–66%) Grembi et al. (2016)  
SC-20 grade crab-shell chitin Acid mine drainage Laboratory bottles – 3.21 10 d 6.79 –,37% Al (>99%),Fe (>99%),Mn (81%) Daubert & Brennan (2007)  
Herbaceous organic substrates Acid mine drainage In situ continuous-flow reactors 5–28 °C 4.5 460 d 6.4 –,26–35% Fe (36–62%), Al (78–83%), Mn (2–6%), Ni (64–81%), Zn (88–95%), Cu (72–85%), Cd (90–92%) Lefticariu et al. (2015)  
Crude glycerol – fluidized bed reactors 25–28 °C 6.5 610 d 7.0 –,89% – Bertolino et al. (2014)  
Municipal sewage and green waste Acidic and metalliferous drainage  25 ± 1 °C 2.4 145 d 5.5 –,– Al (99%), Ca (87%), Mg (87%), Fe (99%) McCullough & Lund (2011)  
Mixed organic substrates compared to mushroom compost Mine drainage Passive bioreactors 20 ± 1 °C 21 d –,83–88% Al (47–100%), Fe (21–100%), Mn (9–90%) Neculita et al. (2011)  
Goat/cow/buffalo manure Acid mine drainage Sulfate-reducing bioreactor 27 2.70 40 d 6.25–7.50 –,48.95–55% Fe (51.49–99.32%), Cu (84.95–99.97%),Zn (35.11–99.78%),Ni (17.87–99.14%),Co (63.55–99.02%),Mn (12.68–73.86%) Choudhary & Sheoran (2012)  
A chitinous product obtained from crushed crab shells Mining-influenced water Anaerobic experimental columns 27 6.6 500 d 5–6 mol/(m3·d), 80–98% Mn (73%), Zn (99.48%), Cu (98.99%), Fe (70.59%) Pinto et al. (2018)  
Mixed ligneous substrate (wood chips, hay, SC-20 and manure) Mining-influenced water Anaerobic experimental columns 27 6.6 500 d 1–2 mol/(m3·d), 80–98% Zn (99.48%), Cu (98.99%), Fe (70.59%) Pinto et al. (2018)  
Grass cellulose Mine water A hybrid reactor system 25 °C 6.6–6.9 (controlled) 49 d 6.6–6.9 650 mg/(L·d),– – Mulopo et al. (2011)  
Waste sludge Acid mine drainage Batch bioassays 8 °C 7.5 65 d  10.70 mg/(L·d),68% – Sahinkaya (2009)  
Silage Synthetic waste water Batch and semi-batch reactors 30 °C; 20 °C; 9 °C 4.2 84 d 5.5 34 mg/(L·d); 22 mg/(L·d); 6 mg/(L·d),– – Wakeman et al. (2010)  
A waste from the wine industry with calcite tailing added Sulfate-rich effluents  21 °C 7.3 65 d –,95% – Martins et al. (2009a)  
Molasses Rayon industry wastewater UASB  1.1–1.5 250 d 7.0–7.5 7.22 ± 1.91 g SO4/(L·d), 74 ± 22% Zn (95%) Liamleam & Annachhatre (2007b)  
Carbon sourceWastewater typeBioreactor typeTemperaturepH(t0)TimepH(final)Sulfate removal: rate, percentageHeavy metals removalReferences
Chicken manure Acidic coal refuse Culture medium 30 ± 2 °C 5.5 20 d 6.6 –,71.2% Fe (99.7%), Mn (91.1%), Ni (40.8%), Zn (95.6%), Cu (93.2%), Cd (68.3%) Zhang & Wang (2014)  
Straw with ethanol Acidic saline drainage Culture medium 18–30 °C 4.5 50 d 5.1–5.9 0.1054–0.1110 mmol/(dm3·d),– Co (95%), Pb (96%), La (98.6%) Santini et al. (2010)  
Algae Acid rock drainage Permeable reactive barriers 24 ± 1 °C 4.0 123 d 6.5 –,80% Cu (>99.5%) Ayala-Parra et al. (2016)  
Microalgae – Immobilized SRB beads 30 5.5 45 d – –,72.4–74.4% Cu (91.7–98.2%) Li et al. (2018b)  
Phalaris arundinacea plant material hydrolyzate Mine wastewater Fluidized-bed bioreactor 35 35 d – 2.2–3.3 g/(L·d),– Fe (99%, 0.84 g Fe/(L·d)), Zn (99%, 15 mg Zn/L) Lakaniemi et al. (2010)  
Sweetmeat waste fractions Acid mine drainage Culture medium 37 7.3 ± 0.2 56 d – –,70% – Das et al. (2013)  
Tannery effluent – Up-flow anaerobic sludge blanket (UASB) 27 7.5 20 d 7.4–7.9 0.6 g/(L·d) 80% – Boshoff et al. (2004)  
Tannery effluent – Stirred tank reactor 27 7.5 33 d 7.6–8.0 025 g/(L·d), 78% – Boshoff et al. (2004)  
Tannery effluent – Trench reactor 27 7.5 30 d 7.9–8.2 0.4–0.5 g/(L·d), 72.12% (s.d. ± 11.72) – Boshoff et al. (2004)  
Chitinous materials Mining influenced water Sulfate-reducing bioreactors – 2.48 ± 0.06 400 d – 2.03–3.97 mmol/(dm3·d),– Cd (97.38%), Zn (99.45%) Al-Abed et al. (2017)  
Crab-shell chitin complex Mine-impacted water Laboratory bottles 20 ± 1 °C 2.95 50 d 6.5–6.7 17.8 mgSO42−/(L·d),– Al (100%), Mn (>73%), Fe (96%) Robinson-Lora & Brennan (2010)  
Crab shell and spent mushroom compost Mine-impacted water Continuous-flow columns – 2.5 ± 0.1 181 d 6.0 – Ni (10–31%), Zn (25–66%), Co (41–66%) Grembi et al. (2016)  
SC-20 grade crab-shell chitin Acid mine drainage Laboratory bottles – 3.21 10 d 6.79 –,37% Al (>99%),Fe (>99%),Mn (81%) Daubert & Brennan (2007)  
Herbaceous organic substrates Acid mine drainage In situ continuous-flow reactors 5–28 °C 4.5 460 d 6.4 –,26–35% Fe (36–62%), Al (78–83%), Mn (2–6%), Ni (64–81%), Zn (88–95%), Cu (72–85%), Cd (90–92%) Lefticariu et al. (2015)  
Crude glycerol – fluidized bed reactors 25–28 °C 6.5 610 d 7.0 –,89% – Bertolino et al. (2014)  
Municipal sewage and green waste Acidic and metalliferous drainage  25 ± 1 °C 2.4 145 d 5.5 –,– Al (99%), Ca (87%), Mg (87%), Fe (99%) McCullough & Lund (2011)  
Mixed organic substrates compared to mushroom compost Mine drainage Passive bioreactors 20 ± 1 °C 21 d –,83–88% Al (47–100%), Fe (21–100%), Mn (9–90%) Neculita et al. (2011)  
Goat/cow/buffalo manure Acid mine drainage Sulfate-reducing bioreactor 27 2.70 40 d 6.25–7.50 –,48.95–55% Fe (51.49–99.32%), Cu (84.95–99.97%),Zn (35.11–99.78%),Ni (17.87–99.14%),Co (63.55–99.02%),Mn (12.68–73.86%) Choudhary & Sheoran (2012)  
A chitinous product obtained from crushed crab shells Mining-influenced water Anaerobic experimental columns 27 6.6 500 d 5–6 mol/(m3·d), 80–98% Mn (73%), Zn (99.48%), Cu (98.99%), Fe (70.59%) Pinto et al. (2018)  
Mixed ligneous substrate (wood chips, hay, SC-20 and manure) Mining-influenced water Anaerobic experimental columns 27 6.6 500 d 1–2 mol/(m3·d), 80–98% Zn (99.48%), Cu (98.99%), Fe (70.59%) Pinto et al. (2018)  
Grass cellulose Mine water A hybrid reactor system 25 °C 6.6–6.9 (controlled) 49 d 6.6–6.9 650 mg/(L·d),– – Mulopo et al. (2011)  
Waste sludge Acid mine drainage Batch bioassays 8 °C 7.5 65 d  10.70 mg/(L·d),68% – Sahinkaya (2009)  
Silage Synthetic waste water Batch and semi-batch reactors 30 °C; 20 °C; 9 °C 4.2 84 d 5.5 34 mg/(L·d); 22 mg/(L·d); 6 mg/(L·d),– – Wakeman et al. (2010)  
A waste from the wine industry with calcite tailing added Sulfate-rich effluents  21 °C 7.3 65 d –,95% – Martins et al. (2009a)  
Molasses Rayon industry wastewater UASB  1.1–1.5 250 d 7.0–7.5 7.22 ± 1.91 g SO4/(L·d), 74 ± 22% Zn (95%) Liamleam & Annachhatre (2007b)  

The chemical and physical compositions of carbon substrates can determine the rate of degradation by SRB. There is a correlation between the chemical composition and capacity to drive sulfate reduction. Firstly, from the aspect of chemical composition, substrates containing high concentrations of crude fat, carbohydrate and protein, like municipal waste and especially the food industry wastewater, can make a high sulfate-reducing efficiency (e.g., wine wastewater). Various animal manures are sometimes used for SRB because of their contents of proteins and lipids (Halofsky & McCormick 2005; Zhang et al. 2014b). By contrast, higher fiber and lignin contents in a carbon source decrease its capacity to drive sulfate reduction (Coetser et al. 2006). Degradation of the enzymatic cellulose can be obstructed and the hydrolysis process can be hindered by lignin. So the substrates which contain lignin are not suitable for the SRB process, such as solid silage. For instance, different grass types have high protein contents and low lignin concentrations and are suitable as carbon sources for SRB. Secondly, the matrix carbon-to-nitrogen ratio is also an important factor in selection of SRB electron donors. Some farm wastes, such as cattle manure, which is an abundant biodegradable waste, have low C/N ratios that limit the effectiveness of the biological treatment process. An additional advantage of chitin is its carbon/nitrogen ratio of 6.9, which makes it an effective substrate because low nitrogen availability limits SRB growth (Neculita et al. 2007; Robinson-Lora & Brennan 2009). And chitin is the only substrate able to partially remove Mn efficiently (>73%), maybe resulting from the formation of rhodochrosite (Robinson-Lora & Brennan 2010; Rodrigues et al. 2019). Thirdly, some substrates have some special properties, such as the ability to neutralize pH. Crab-shell chitin complex has a strong ability to quickly remove acidity and generate alkalinity steadily and increase pH from 3.0 to close to neutral in 3 days (Robinson-Lora & Brennan 2010). And some substrates can adjust the pH of solution during their fermentation, such as the crab-shell chitin complex (Robinson-Lora & Brennan 2010). Fourthly, from the aspect of physical composition, the presence of organic functional groups is important in metal removal processes with SRB, such as sulfate, hydroxyl, amino, and carboxyl groups, which enhance the removal efficiency. Algal biomass has various active sites in its extracellular polymeric substances, and utilization of algae as carbon resource generates a lower COD level in effluent than using ethanol; even the masses of ethanol and algae are identical (Xiao & Zheng 2016; Flores-Chaparro et al. 2017; Henriques et al. 2017; Li et al. 2018b). In the SRB biological treatment system, there are competitive and synergistic effects between SRB and other bacteria for electronic donors. SRB compete with Acetobacter for butyrate and propionate, and with methanogenic archaea for acetate and hydrogen (Stams et al. 2005). And SRB can have a faster growth rate than syntrophic consortia on butyrate and propionate (Stams et al. 2005). An increasing ratio of COD/SO42− makes SRB lose the ability to drive more available electrons toward sulfate reduction (Das et al. 2015). Methanogens and acetogens gain better dominance over SRB at a higher COD/SO42− ratio (Das et al. 2015). Acid environment is more favorable for MB to compete with SRB (Das et al. 2015). SRB are mainly thermophilic bacteria, and increasing temperature may help SRB to outcompete MB (Ayangbenro et al. 2018). Lactic acid and ethanol are the most studied pure electron donors, but they cost a lot in practical application, and induce an accumulation of acetic acid, which improves the competition between SRB and other bacteria (Liu et al. 2018). As an intermediate of degradation by organic acids, acetate can reduce the pH, and create an acidic environment by its accumulation, which is harmful for SRB. However, acetic acid can be degraded by other co-existing microorganisms, such as chitin-degrading bacteria (Logan et al. 2005; Rodrigues et al. 2019). Therefore, the cooperation of different microorganisms plays an important role in the SRB process.

Chitin-degrading bacteria have synergistic interactions with SRB. These co-existing bacteria, which are fermentative or cellulose degraders, include Bacillus sp., Bacteroides sp., Proteiniclasticum sp., Sedimentibacter sp., Porphyromonadaceae bacteria, Petrimonas sulfuriphila, Parabacteroides chartae, and Macellibacteroides fermentans (Logan et al. 2005). Co-existing Bacteroides sp. can utilize a wide range of organic carbon and energy sources, including pectin, hemicellulose, cellulose, and starch. Bacteroides sp. plays an important role in fermentation of silage, which contains lignocelluloses that are not easy to be fertilized, and cooperate with SRB for sulfate reduction (Hiibel et al. 2008; Wakeman et al. 2010). Among these co-existing species, Citrobacter sp. is interesting and can reduce sulfate and remove copper (Qiu et al. 2009). They have also been used for bioremediation of wastewater contaminated with textile dyes. Kaksonen et al. (Kaksonen et al. 2004) reported on the function of Citrobacter sp. in the treatment of acidic metal- and sulfate-containing wastewater in a sulfidogenic filter bio-reactor. Citrobacter sp. was found to live with SRB on a steel surface, and cooperated with traditional SRB to form a biofilm and produce hydrogen. The Citrobacter sp. has a strong ability for sulfate reduction and they are tolerant to oxygen. They can reduce 10 mM sulfate completely to sulfides within 7 days, and then recover their sulfate-reducing ability after 7 days of aerobic growth (Qiu et al. 2009). Because of their strong ability to reduce sulfate and tolerate oxygen, Citrobacter sp. could be a better inoculum to start a bioreactor than usual substances. Some studies have listed a large number of bacteria co-existing with SRB in heavy metal- and sulfate-containing wastewater (Li et al. 2018b; Qiu et al. 2009; Stams et al. 2009; Zhang & Wang 2016; Zhang et al. 2016a).

Of course, there will be some in situ advantages of some waste substrates, but they always have the disadvantage of poor degradation. The use of mixed media is more conducive to the cultivation of a rich microbial population. The use of mixed matrices is particularly important when substrates containing cellulose or lignin are used as the carbon source. A mixture of different substrates that provide mutual assistance to each other is beneficial for abundant cultivation of various microorganisms (Pinto et al. 2018). And the function of synergism when using a mixture of materials usually yields better efficiencies in the SRB process rather than using a single substrate (Zhang et al. 2016b). Energy sources and organic carbon can be utilized by co-existing Bacteroides sp., including starch, cellulose, hemicellulose and pectin. So Bacteroides sp. helps a lot in fermentation of lignocelluloses containing substrates (Hiibel et al. 2008; Wakeman et al. 2010). It has been confirmed that Citrobacter sp. has a strong ability to tolerate oxygen and reduce sulfate, and it can be a better inoculum to start a bioreactor (Qiu et al. 2009). And a high rate of pH neutralization was found in a previous study when using a mixture substrate of complex manure and cellulosic material (straw) (Zagury et al. 2006).

There are several methods to enhance SRB activity. SRB activity is confirmed by the observation of the formation of black precipitates in the solution and the rotten-egg smell produced by the hydrogen sulfide gas (Cao et al. 2009). From a microbial perspective, vitality of SRB can be increased by increasing salinity or adding inorganic cations. Non-toxic inorganic cations can reduce the toxicities of heavy metals toward SRB by combining with anionic sites on cell surfaces instead of heavy metals, such as calcium, magnesium and iron. It has been demonstrated that the Mg2+ ion concentration is a favorable and important factor for SRB activity and growth. The activity of SRB will be not inhibited but stimulated by a high concentration of Mg2+ through improving the diffusion of SO42− into SRB and the mass transfer (Figure 5) (Cao et al. 2009). An energy load can be imposed on bacteria by the ionic strength in solution due to the osmotic gradient between the exterior and interior of SRB (Blight & Ralph 2004). Fe2+ ion acts as a trace element for SRB and takes part in the synthesis of some enzymes that are active in contaminant removal (Xu et al. 2011; Lee & Oh 2016). In addition, Fe2+ can precipitate with sulfides and then reduce their toxicity to SRB (Zhang et al. 2011, 2016a; Liu et al. 2015c). And the insoluble sulfides formed with Fe2+ have a better ability to adsorb and remove heavy metals. Research has shown that the addition of 200 mg/L Fe2+ can enhance the removal efficiency of As from 78% to 98.2% and slightly increases the removal rate of Sb from 98.8% to 99.4% (Liu et al. 2018). It has also been demonstrated that adding NaCl can enhance the growth rate and activity of SRB more than by adding MgCl2·6H2O (pH of 7.6 for NaCl compared with 7.0 for MgCl2·6H2O). Consequently, there is better affinity between SRB and NaCl than SRB and MgCl2·6H2O (Liu et al. 2018).

Figure 5

Mechanisms of enhancement on metal removal efficiency in SRB process by adding cations (HM: heavy metal).

Figure 5

Mechanisms of enhancement on metal removal efficiency in SRB process by adding cations (HM: heavy metal).

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The ZVI can efficiently improve the activity of SRB and the removal rate of heavy metals in the SRB system. ZVI and microorganisms act cooperatively in heavy metal removal (Figure 6). ZVI acts as an electron donor to decrease competition between MB and SRB for electrons in sulfate-containing wastewater, and SRB adhered to ZVI can gain electrons from the H2 produced during ZVI corrosion. This reduces the COD for reduction of sulfate (Dinh et al. 2004; Kong et al. 2014; Liu et al. 2015c; Ou et al. 2016). Furthermore, the decreased ORP caused by ZVI (about 100 mV) creates a more reductive environment for SRB (Zhang et al. 2011, 2014c; Kumar et al. 2015). Research has shown that the reducing capacities of SRB such as Desulfobacterium sp. and Desulfovibrio strains are driven by H2 although both H2 and organic compounds are electron donors for SRB (Carepo et al. 2002; Seth & Edyvean 2006). Furthermore, pH can be neutralized through the consumption of H+ and generation of alkaline by-products (Zhang et al. 2011; Liu et al. 2015c). Also, as a strong reductant, ZVI can alleviate the toxicity of contaminants and remove them through directly reacting with inorganic or organic compounds, such as by reacting with Cr6+ (Gheju 2011; Zhang et al. 2011, 2016a; Liu et al. 2015c). Research has shown that with ZVI, SRB can adapt to wider pH and temperature ranges with a greater Cu and sulfate loading rate (Xin et al. 2008; Bai et al. 2013). In a coupled system, enrichment of SRB has been observed (Zhang et al. 2011, 2016a; Liu et al. 2015c). In addition, the precipitates of heavy metals formed synergistically by SRB and ZVI are more stable and don't leach out with pH or redox changes (Gu et al. 1999; Kumar et al. 2015). So ZVI is always used as a promoter to stimulate the performance of SRB, resulting in better sulfate and heavy metal removal (Guo et al. 2017a). However, the cost of iron is high at approximately $70 for 100 g in South Africa, which should be considered in the financial viability assessment (Mulopo & Schaefer 2013).

Figure 6

Mechanism of enhancement by ZVI on SRB activity.

Figure 6

Mechanism of enhancement by ZVI on SRB activity.

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Some other materials can enhance the microorganism activity by enhancing electron transfer in the SRB system. Graphene oxide (GO) introduced to a SRB culture as a carrier can enhance the activity of SRB. Compared with conventional carbon materials, GO has a higher electro-catalytic activity (20 m2/(V·s)), higher thermal conductivity (5,000 W/(m·K)) and larger surface area (2,630 m2/g) (Brownson et al. 2011). Research has shown that the electron transfer can be highly improved by the presence of GO (Yan et al. 2018). In biocathodes, the maximum current of the cathodic reduction peak achieved with GO is four times that without GO (12.0 vs. 3.0 mA), the population of SRB in effluent with GO was 277 times that without GO, and the sulfate reduction rate in microbial electrolysis cells with GO was 70% higher than that without GO (Hu et al. 2019). These results indicate that GO can greatly enhance the growth of SRB by transferring electron donors to SRB (Hu et al. 2019). Furthermore, GO can act as an electron acceptor for SRB, and using GO as a sole electron donor for SRB can generate a higher sulfate removal efficiency. Reduced graphene oxide (rGO) has better ability of electron transfer and structural characteristics than GO (Wu et al. 2019). And rGO can induce formation of pili in a sulfate-reducing biofilm (Adhikari et al. 2016). Self-secreted mediators and direct electron transfer (e.g., outer membrane c-type cytochromes) are involved at the cell/GO interface during the reduction of GO using microorganisms (Wang et al. 2011). Research has shown that when the rGO embeds in the biofilm, electron transfer in the biofilm can be accelerated (Bo et al. 2014), because rGO has a positive surface, which is beneficial to electron inflow and electrostatic interactions with bacterial membranes (negatively charged) (Sun et al. 2017). The electron inflow can also restore toxic heavy metals so as to alleviate their toxicity toward microbes. And microbial reduction of GO to rGO by SRB is cost-effective, simple, biocompatible and environmentally friendly (Cheng et al. 2018; Zhang et al. 2018a). The properties and purification methods of microbially reduced graphene should be explored in future research.

Some additives can enhance the accumulation of volatile fatty acids for a better growth of SRB. In a previous study, adding a cysteine dose of 50 mg/L in the acidification stage could increase the accumulation of propionate, butyrate and acetate by 49.3% (Liu & Chen 2018). The acidification of l-glucose increased significantly, resulting from the addition of cysteine, and the activities of homoacetogens improved by 34.8% and hydrogenotrophic methanogens by 54% through acceleration of the electron transfer process (Liu & Chen 2018). The solubilization of carbohydrate and sludge protein can be improved by the humic acid due to its protection of enzyme activity and high hydrophobicity (Liu et al. 2015a; Luo et al. 2016).

Toxicity of heavy metals to SRB is one of the biggest challenges in the SRB process, especially in suspended SRB systems. Several measures have been taken to increase the tolerance of SRB to heavy metals, including immobilization of SRB and masking metal toxicity by additives. Citric acid, a multi-dentate chelating agent and naturally occurring carboxylic acid which can form bidentate, tridentate, and polymeric stable metal–citric acid complexes and binuclear with heavy metals, and the form depends on the metal/citrate acid ratio, pH of the solution and type of metal (Dodge & Francis 2002; Gamez et al. 2009; Kou et al. 2020). Changes in the speciation of heavy metals can reduce their toxicity to SRB in batch reactors (Paulo et al. 2015; Qian et al. 2015; Chen et al. 2020) (Kieu et al. 2011). Research has shown that Cu2+ adsorption is dominated by formation of complexes between Cu2+ and citrate3− anions (Li et al. 2015). And both Cd and U form tridentate and binuclear complexes with citric acid (Qian et al. 2013). In the presence of iron and uranium, a ternary iron–uranium–citric acid complex forms (Dodge & Francis 2002). When removing Ni2+ from citric acid-containing wastewater using SRB, the Ni–citrate complex is predominant rather than Ni–lactate complex. Results showed that citric acid can decrease the amount of free Ni2+ and enhance the quantity of SRB, which suggests that it can effectively alleviate the toxicity of Ni for SRB (Basnakova & Macaskie 2001; Van Nostrand et al. 2005; Qian et al. 2015). Citric acid is also used as a chelator to extract heavy metals from sewage (Al-Qahtani 2017; Zhang et al. 2020). Citric acid is also used as an assistant in heavy metal plant adsorption (Wang et al. 2019b; Mallhi et al. 2020). Citric acid treatment with 63% uranium solubilization was found to be the most efficient method in studies on the optimization of uranium solubilization (Lozano et al. 2011). Because ammonium citrate is non-toxic and biodegradable, ammonium citrate is usually used to form stable complexes with heavy metals, such as Cd, Zn and Cu (Pedersen et al. 2005). Citric acid is sometimes used to reduce the toxicity of heavy metals in soil (Shi et al. 2020). And citric acid can be fermented to acetate, CO2, and hydrogen under anaerobic conditions so as to feed SRB to reduce sulfate (Gamez et al. 2009; Stams et al. 2009; Xia et al. 2019). Moreover, the degradation rate of citric acid remains at a very high level, which is difficult to be affected by the types of sludge and sulfate concentration (Gamez et al. 2009).

Immobilization treatment can protect SRB from heavy metal toxicity significantly. Suspended SRB systems are more vulnerable to environmental factors and poor cell retention in continuous bioreactors. SRB lose their activity due to washout of cells in continuous HRT bioreactors. Addition of rich carbon sources to the suspended system is needed to feed the SRB, which leads to a high COD in the effluent. Many methods have been investigated to solve these problems. SRB immobilization can enhance the sulfate reduction efficiency by 33–40% (Li et al. 2018b; Zhang et al. 2018b; Dong et al. 2019). There are two ways to immobilize SRB: the first with only the SRB fixed, without an internal carbon source in the beads, sometimes through the way of biofilm formation; and the second with the carbon source and SRB fixed simultaneously in the beads, usually by gel entrapment (Figure 7). Immobilization with only SRB can produce an effluent with a high COD, which requires further treatments. Immobilization with an internal carbon source in the beads can prevent the effluent from having a high COD (Min et al. 2008; Li et al. 2017; Zhang et al. 2018c). Moreover, low-molecular-weight organic matter degradation from organic substrates can occur near SRB, which creates a high COD/SO42− ratio for SRB growth (Li et al. 2018b).

Figure 7

Mechanism of enhancement in SRB activity by immobilization: (a) with inner substrates fixed; (b) without inner substrates fixed.

Figure 7

Mechanism of enhancement in SRB activity by immobilization: (a) with inner substrates fixed; (b) without inner substrates fixed.

Close modal

The mechanism of enhancement involves synergistic effects such as provision of a large surface area for SRB growth, adsorption of sulfate and heavy metals onto the carriers, prevention of heavy metal toxicity toward SRB, and creation of neutral sites in the inner part of the cell, which prevents death or inhibition of cells. A hydrophobic surface (Hadjiev et al. 2007), iron reduction state (Li et al. 2019) and proper porosity (Wang & Wang 1989) are good properties of SRB carrier materials. Biofilms can reductively precipitate heavy metals, such as U. Ensuring there is sufficient biofilm surface coverage rather than increasing the biofilm thickness and biomass is a better strategy for improving heavy metal removal (Cologgi et al. 2014). Various physical carriers show a protective effect on SRB activity in different ways. The total surface area provided by the immobilization carrier matrix can decide the bioreactor performance (Baskaran & Nemati 2006; Zhang et al. 2019). In one study, sand exhibited superior performance (the highest sulfate-reducing rate was 1.7 g/(L·h)) in the shortest residence time (0.5 h) because it had a higher total surface area (145.1 m2) than other materials (biomass support particles 3.2 m2 and glass beads, 0.6 m2). The reduction rate with sand was eight- and 40-fold faster than those obtained with biomass support particles or glass beads, respectively (Baskaran & Nemati 2006). Researchers have shown that supports with higher contact angles and hydrophobicity have larger numbers of adhered microorganisms because of the lower interface energy between the support and SRB (Hadjiev et al. 2007; Sohn et al. 2010). In terms of the pore structure of the adhesive material, studies have indicated that macroporous supports with pore diameters one to five times those of the microorganisms provide the best biomass performance (Silverstein 2014). Nanoscale materials, such as lava, ceramics, natural activated carbons and other synthetic compounds, usually have enhanced adhesive abilities for SRB because of their high surface roughness. By contrast, µm-scale materials, such as poly (high internal phase emulsion) beads, have strong abilities to produce a dense bacterial colony and stimulate the generation of SRB (Depardieu et al. 2016; Li et al. 2018a). Moreover, in an immobilized system, the biomass residence time is uncoupled from the HRT. Immobilization could shorten the HRT and maintain a longer microbial retention time, resulting in a high concentration of biomass and increased reaction rate (Baskaran & Nemati 2006; Silva et al. 2006; Kiran et al. 2018). In addition, energy dispersive X-ray spectroscopy analysis has indicated that the outer precipitate of metal sulfides on the cells can prevent metal ions from entering the granules and directly contacting SRB in immobilized systems (Min et al. 2008). Therefore, it is an important strategy to provide an internal carbon source, and before the reactors start, immobilized beads should be immersed in sulfate-containing water to create a sulfide-rich layer around the beads, which can prevent heavy metals from entering the beads (Min et al. 2008; Zhang et al. 2018c). And immobilized SRB beads can be reused for several times to enhance the heavy metal removal efficiency.

Many carriers have been used to promote SRB activity, some of which act as an electron shuttle to enhance the ability of electron transportation in the SRB system, such as GO and copper–iron bimetallic particles (Figure 8). Bagasse (Grubb et al. 2018; Dong et al. 2020) and corncob (Di et al. 2019; Dong et al. 2019) were used to prepare immobilized SRB sludge particles in the treatment of acid mine wastewater with low pH. For copper–iron bimetallic particles, compared with particles containing only iron, the presence of copper can promote the electron transfer and eliminate iron passivation. Some other materials such as recycled aggregate bio-carriers, which have negatively charged surfaces, can adsorb heavy metals. In this paper, several carriers were reviewed according to their different functions and increased efficiencies of heavy metal and sulfate removal (Table 4). The mechanisms are shown in Figure 8.

Table 4

Materials reported for SRB immobilization

Carrier materialFeaturesComparison of removal efficiency before and after fixing
References
Free SRB systemImmobilized SRB system
GO 
  • large surface area enabled growth of microbes which is better than activated carbon and carbon filter (Liu et al. 2015b).a

  • the redox-mediating and electron transfer capacity of grapheme remarkably promote bioelectricity generation, biotransformation and biodegradation of pollutants (Zhang et al. 2014a; Colunga et al. 2015; Yoshida et al. 2016; Toral-Sanchez et al. 2017).a

  • rGO particles were formed naturally without artificial effort (Zhang et al. 2014a, 2014b).a

  • rGO prevented microbial activity from inhibition caused by toxic and unfavorable condition.a

  • excess rGO addition has an inhibitory effect on the growth of SRB (Guo et al. 2017b).b

  • efficiency of sulfate removal by pure rGO adsorption could only reach 5%.b

 
  • SO42− (70%)

  • Pb (51%), Ni (46%), Cu (42.1%), Tl (35.2%), Fe (24.1%), Cd (13.5%), Cr (9.2%)

 
  • SO42− (about 100%)

  • Pb(II) (97.1%), Tl(I) (91%), Cu(II) (89.2%), Fe(III) (77.0%), Cd(II) (51.5%) and Cr(III) (12.4%)

  • SRB growth rate (0.27 h−1), doubling time (2.5 h) (around 3 times faster than free system)

 
Zhang et al. (2016b), Yan et al. (2018)  
Copper-iron bimetallic particles 
  • enhanced Cu2+ and Zn2+ removal and SRB resistance (from 100 to 200 mg/L for Cu2+ and 300 to 400 mg/L for Zn2+) to metals.a

  • Cu(0) could serve as an electron shuttle to alleviate the electronic transfer inhibition caused by the insoluble film formed or iron sulfide deposited on the iron surface (Hu et al. 2010).a

 
  • SO42− (45–70%)

  • after 24 h: Cu2+ (20.7%), Zn2+ (78.4%)

  • after 48 h: Cu2+ (29.8%), Zn2+ (90.9%)

 
  • SO42− (96%)

  • after 24 h: Cu2+ (98.2%), Zn2+ (99.5%)

  • after 48 h: Cu2+ (98.17%), Zn2+ (99.67%)

 
Hu et al. (2010), He et al. (2011), Zhou et al. (2013)  
Polyvinyl alcohol-sodium alginate 
  • polyvinyl alcohol, sodium alginate and microalgae had good adsorption capacity for sulfate and heavy metals (Xiao & Zheng 2016).a

  • sodium alginate, a linear polysaccharide composed of b-D -mannuronic (M) and a-L-guluronic acid (G) residues, could strongly bind metal ions and is currently used for heavy metal adsorption (Wang et al. 2013b), had easy availability, low-cost and non-biodegradability.a

  • microalgae serve as an inner substrates, have abundant polysaccharides and proteins in their extracellular polymeric substances, and contain functional groups such as amino, hydroxyl, carboxyl, and sulfate groups which could adsorb heavy metals strongly (Xiao & Zheng 2016; Henriques et al. 2017).a

  • carbon sources were in the vicinity of the SRB, maintaining a high COD/SO42−.a

  • the immobilized SRB beads can be reused several times for achieving the best metal removal results.a

 
  • SO42− (34.7–39.4%)

  • Zn (49.1–76.8%)

 
  • SO42− (61–88%)

  • Cu(II) (91.7–98.2%); Fe, Cd, Zn (over 99%); Mn (42.1–99.3%); Ni(II) (75%); Pb(II) (92%)

 
Zhang & Wang (2016), Zhang et al. (2016b, 2018b), Li et al. (2017), Kiran et al. (2018), Li et al. (2018b)  
Recycled aggregate bio-carrier 
  • the recycled aggregate (RA) is a renewable minerals resource that is available in large quantities and can be used as a good base for the development of adsorbent materials (Coleman et al. 2005).a

  • slightly negative charged RA contains about 95% (CaO, SiO2, Al2O3 and Fe2O3), which are major ingredients of adsorbents for heavy metal removal from wastewater.a

  • heavy metal removal by RA was relatively lower than that of other types of sorbents and other biomasses (Babel & Kurniawan 2003).b

 
– 
  • Zn2+ (100%), Ni2+ (98.0%), Cr6+ (87.8%) (nine times greater than what was obtained with the RA carrier)

 
Kim et al. (2016)  
Carrier materialFeaturesComparison of removal efficiency before and after fixing
References
Free SRB systemImmobilized SRB system
GO 
  • large surface area enabled growth of microbes which is better than activated carbon and carbon filter (Liu et al. 2015b).a

  • the redox-mediating and electron transfer capacity of grapheme remarkably promote bioelectricity generation, biotransformation and biodegradation of pollutants (Zhang et al. 2014a; Colunga et al. 2015; Yoshida et al. 2016; Toral-Sanchez et al. 2017).a

  • rGO particles were formed naturally without artificial effort (Zhang et al. 2014a, 2014b).a

  • rGO prevented microbial activity from inhibition caused by toxic and unfavorable condition.a

  • excess rGO addition has an inhibitory effect on the growth of SRB (Guo et al. 2017b).b

  • efficiency of sulfate removal by pure rGO adsorption could only reach 5%.b

 
  • SO42− (70%)

  • Pb (51%), Ni (46%), Cu (42.1%), Tl (35.2%), Fe (24.1%), Cd (13.5%), Cr (9.2%)

 
  • SO42− (about 100%)

  • Pb(II) (97.1%), Tl(I) (91%), Cu(II) (89.2%), Fe(III) (77.0%), Cd(II) (51.5%) and Cr(III) (12.4%)

  • SRB growth rate (0.27 h−1), doubling time (2.5 h) (around 3 times faster than free system)

 
Zhang et al. (2016b), Yan et al. (2018)  
Copper-iron bimetallic particles 
  • enhanced Cu2+ and Zn2+ removal and SRB resistance (from 100 to 200 mg/L for Cu2+ and 300 to 400 mg/L for Zn2+) to metals.a

  • Cu(0) could serve as an electron shuttle to alleviate the electronic transfer inhibition caused by the insoluble film formed or iron sulfide deposited on the iron surface (Hu et al. 2010).a

 
  • SO42− (45–70%)

  • after 24 h: Cu2+ (20.7%), Zn2+ (78.4%)

  • after 48 h: Cu2+ (29.8%), Zn2+ (90.9%)

 
  • SO42− (96%)

  • after 24 h: Cu2+ (98.2%), Zn2+ (99.5%)

  • after 48 h: Cu2+ (98.17%), Zn2+ (99.67%)

 
Hu et al. (2010), He et al. (2011), Zhou et al. (2013)  
Polyvinyl alcohol-sodium alginate 
  • polyvinyl alcohol, sodium alginate and microalgae had good adsorption capacity for sulfate and heavy metals (Xiao & Zheng 2016).a

  • sodium alginate, a linear polysaccharide composed of b-D -mannuronic (M) and a-L-guluronic acid (G) residues, could strongly bind metal ions and is currently used for heavy metal adsorption (Wang et al. 2013b), had easy availability, low-cost and non-biodegradability.a

  • microalgae serve as an inner substrates, have abundant polysaccharides and proteins in their extracellular polymeric substances, and contain functional groups such as amino, hydroxyl, carboxyl, and sulfate groups which could adsorb heavy metals strongly (Xiao & Zheng 2016; Henriques et al. 2017).a

  • carbon sources were in the vicinity of the SRB, maintaining a high COD/SO42−.a

  • the immobilized SRB beads can be reused several times for achieving the best metal removal results.a

 
  • SO42− (34.7–39.4%)

  • Zn (49.1–76.8%)

 
  • SO42− (61–88%)

  • Cu(II) (91.7–98.2%); Fe, Cd, Zn (over 99%); Mn (42.1–99.3%); Ni(II) (75%); Pb(II) (92%)

 
Zhang & Wang (2016), Zhang et al. (2016b, 2018b), Li et al. (2017), Kiran et al. (2018), Li et al. (2018b)  
Recycled aggregate bio-carrier 
  • the recycled aggregate (RA) is a renewable minerals resource that is available in large quantities and can be used as a good base for the development of adsorbent materials (Coleman et al. 2005).a

  • slightly negative charged RA contains about 95% (CaO, SiO2, Al2O3 and Fe2O3), which are major ingredients of adsorbents for heavy metal removal from wastewater.a

  • heavy metal removal by RA was relatively lower than that of other types of sorbents and other biomasses (Babel & Kurniawan 2003).b

 
– 
  • Zn2+ (100%), Ni2+ (98.0%), Cr6+ (87.8%) (nine times greater than what was obtained with the RA carrier)

 
Kim et al. (2016)  

aAdvantages.

bDisadvantages.

Figure 8

Mechanisms of carriers supporting SRB process: (a) reduced graphene oxide; (b) copper–iron bimetallic particles; (c) polyvinyl alcohol; (d) recycled aggregate bio-carrier.

Figure 8

Mechanisms of carriers supporting SRB process: (a) reduced graphene oxide; (b) copper–iron bimetallic particles; (c) polyvinyl alcohol; (d) recycled aggregate bio-carrier.

Close modal

The sludge or sewage rich in SRB produced in the subsequent SRB process is a major environmental hazard and will continue to react and corrode the pipeline. So, SRB biomass disposal is very important. This paper lists some methods for disposal of SRB and compares their characteristics. The bactericidal effect of ultraviolet light-emitting diode on SRB and other bacteria on the sea floor has been studied, and it was concluded that the ultraviolet light-emitting diode with a peak wavelength of 280 nm is the most practical in this field (Qiao et al. 2018). However, the field applicability of ultraviolet sterilization and high concentration bactericide is poor, and the number of bacteria is obviously restored (Ye 2019). In contrast, negative pressure chlorine dioxide shows superiority; for example, the sterilization rate is close to 100% (Ye 2019). The presence of chlorine dioxide inhibited the reduction ability of SRB and promoted the oxidation of sulfide (Okoro 2015). Due to the fact that SRB cells are stuck in the polymer matrix in practical application, biofilm is extremely difficult to remove. Therefore, many researchers have tried to find some enzymes that can disperse SRB biofilm. A supernatant of Pseudomonas aeruginosa dispersed SRB biofilm via rhamnolipids (Wood et al. 2018). Dvua0066 (a hypothetical phospholipase) can reduce the formation of SRB biofilm by 5.6 times (Wood et al. 2018). Glycoside hydrolase DisH (Zhu et al. 2018), mannose, 2-deoxy-d-glucose and N-acetylgalactosaminidase (Poosarla et al. 2017) play inhibitory and dispersal roles on the biofilm of Desulfovibrio vulgaris.

A method of biological competition exclusion can create an environment which is conducive to the survival of other bacteria so as to outcompete SRB, thus limiting the growth of SRB. Injection of nitrate or nitrite can promote SRB control by stimulating the growth of nitrite-reducing or organic nitrate-reducing bacteria, which can outcompete SRB for electron donors (bio-competitive exclusion), and/or by lithotrophic nitrate- or nitrite-reducing sulfide-oxidizing bacteria, which can remove H2S directly (Hubert et al. 2009; Rajeev et al. 2015; Dolfing & Hubert 2017). Biological competition is the main factor when nitrite is added (Hubert et al. 2009). The addition of nitrate promoted the growth of Rhizobiaceae and Xanthophytaceae, while the addition of chlorate had a general inhibition on the microbial community, at the same time, significantly reduce in the sulfate-reducing species, showing a specific toxicity. Chlorate stimulates Pseudoalternates and Pseudomonaceae. The addition of perchlorate promotes the production of Thiomonaceae and Thiobaceae, which contain the types of elemental sulfur reduction and sulfide oxidation respectively. Therefore, in addition to the biological competition, the concentration of sulfide in wastewater can be controlled through the sulfur oxidation–reduction cycle (Engelbrektson et al. 2014). In a recent study, the inhibition effect of adding perchlorate was not as good as that of adding nitrate due to the fact that the perchlorate-reducing bacteria enriched in the study cannot utilize the substrate such as alkylbenzene, which can be easily used by nitrate-reducing bacteria (Okpala & Voordouw 2018). The concentration of sulfide decreases significantly in a system of chlorine dioxide with the presence of nitrate reducing bacteria. Therefore, it is cost-effective to explore the synergistic effect of low concentration chlorine dioxide and metabolic inhibitors (such as nitrite) (Okoro 2015).

When considering both the entire process and each stage, various factors can influence the heavy metal and sulfate removal efficiency of SRB. The two steps in the SRB process are sulfate reduction and precipitation of metals with sulfide. Sulfate reduction is the determining step for removal of sulfate and heavy metals (Hsu et al. 2010). Therefore, strategies focus on promotion of the efficiency of sulfate reduction. Whatever method is used to enhance the removal efficiency, the main focus is increasing SRB activity, which can be achieved by increasing the availability of carbon sources, mediating the environment for SRB survival, enhancing SRB activity using additives, and preventing direct contact between SRB and heavy metals. Studies have shown that the cost and availability of the electron source are challenges that limit implementation of the biological sulfate-reduction process. The most commonly used sources are lactate and ethanol. The efficiency of sulfate removal with ethanol is worse than that using lactate because degradation of lactate results in production of more alkaline species than with ethanol, which is beneficial to neutralize the acidity and support SRB growth. The utilization of both lactate and ethanol leads to production of acetate, which greatly reduces the pH and SRB activity, resulting in a high COD in the effluent. Furthermore, the acetoclastic SRB group is susceptible to high H2S concentrations, and the lack of a mechanism for acetyl-CoA oxidation means that acetate is not degraded. Instead of using expensive small molecule matrices in large-scale applications, a more feasible strategy is to identify inexpensive and available carbon sources, which can satisfy the relatively high efficiency and cost requirements of the SRB process. There is a close correlation between the chemical compositions and capacities of carbon sources and sulfate reduction. High concentrations of protein, carbohydrates, and crude fats can increase the capacity of a carbon source to drive sulfate reduction. By contrast, higher lignin and fiber contents in a carbon source reduce the capacity to drive sulfate reduction. When choosing a suitable substrate to feed SRB, the composition should be determined first. Utilization of cost-effective substrates produces an effluent with a high COD. Considering the availability and cost of carbon sources, it is worthwhile to explore waste carbon as a source.

Synergistic functions between microorganisms have been investigated thoroughly. Many studies have indicated that the use of mixed SRB cultures presents advantages over the use of pure cultures because of co-degradation of substrates by other microorganisms and production of low-molecular-weight acids for SRB growth. In addition, some microorganisms can adsorb heavy metals or promote generation of products such as FeS, which also shows excellent adsorption of heavy metals. Inorganic ions added to the SRB system can compete with heavy metals for anion binding sites to reduce heavy metal toxicity towards SRB. They can also enhance electron transfer, such as the participation of ZVI and GO as electron shuttles. Among the strategies available for enhancing the efficiency of the SRB process, immobilization seems to the most feasible because of its simplicity and greatly increased removal rate of heavy metals. Immobilization creates a perfect barrier to prevent direct contact between SRB and heavy metals, which alleviates the toxicity issue. Compared with unfixed carbon source systems, immobilization of the carbon source shows better performance because it avoids production of an effluent with a high COD. In addition, carbon sources surrounding SRB increase the COD/SO42− ratio, which is conducive to SRB growth. Most immobilized materials can adsorb heavy metals, and some immobilized materials have negatively charged surfaces that can adsorb heavy metal cations very well. Whatever materials are used in the SRB process, a large increase in the efficiency must be obtained. GO, which acts as an immobilization carrier and is also an excellent electron acceptor for SRB growth, can be reduced to rGO, which has the ability to enhance SRB activity through electron transportation. It has been explored for use in the SRB process for heavy metal and sulfate removal.

Although the SRB process is a promising microbial method for heavy metal removal, it also has disadvantages, and many aspects still need to be explored.

The treatment of mercury-containing wastewater by SRB will produce toxic methylmercury. At present, there are few methods (Wang et al. 2019a) to overcome this problem, which needs further study.

Acetoclastic SRB are very susceptible to the toxicity of H2S. The accumulating and inhibiting effect of acetic acid on SRB is an urgent problem to be solved in SRB methods.

Strengthening electron transfer is an effective way to improve SRB activity, and more substances that could improve electron transfer conditions should be identified. Synergistic bioprecipitation and adsorption would be beneficial for heavy metal removal. To date, many materials have been discovered or fabricated for heavy metal adsorption. In future studies, these materials should be investigated in combination with SRB because of their excellent qualities.

Metal recovery from different types of solid waste is becoming a priority in the new production cycle of reintroducing resources and treatment schemes that tend toward zero emissions. Therefore, the potential value of metal-rich sludge from metal-containing wastewater treatment is worth exploring.

This work was financially supported by the National Science Foundation of China (51778454 and 51425802).

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