Keywords

8.1 Introduction

Generation of million tons of waste in metalliferous/coal mines containing sulfidic minerals is a major environmental and economic concern worldwide (Nordstrom 2000; Baker and Banfield 2003). These sulfidic minerals upon oxidation generate highly acidic, metal- and sulfate-rich mine waters known as acid mine drainage (AMD) (Nordstrom 2000). AMD is considered to be a multifactor pollutant, detrimental for the receiving ecosystems including aquatic and terrestrial (Gray 1997). Numerous studies have reported the lethal impact of AMD on the aquatic life-forms (Soucek et al. 2000; Schmidt et al. 2002; Gerhardt et al. 2004; Martin and Goldblatt 2007; Van Damme et al. 2008; Gray and Delaney 2008; Jennings et al. 2008; Hogsden and Harding 2011). Owing to its extreme and toxic nature, AMD degrades the quality of river, lakes, and groundwater and also affects the terrestrial ecosystem (Nieto et al. 2007; Olıas et al. 2004; Alhamed and Wohnlich 2014; Gupta and Nikhil 2016; Roychowdhury et al. 2017; Grande et al. 2018). Both active and abandoned mines generate millions of cubic tons of AMD worldwide. In the reports from the US Forest Service 2005, 55700 mines are abandoned in the USA. Almost 15,000–23,000 km AMD-impacted streams are present across the USA (Roychowdhury et al. 2019 and references therein). Near about 600 coal mines in the Zipaquirá Mining District of Colombia generate 70,400 m3 of contaminated mine water each month (Fenalcarbón 2006). Pérez-Ostalé et al. (2013) determined the mining-affected area in the Iberian Pyrite Belt (TPB) region and found that more than 4800 ha areas of 88 sulfidic mines were occupied by mine waste, open pit, tailing dams, etc. Recently, Grande et al. (2018) mapped the impact of AMD on the river network of the IPB. They also demonstrated that 23 water reservoirs near the IPB are also affected by AMD and have excess mining leachates than the established range by Directive 98/83/CE (regulation for the quality of the drinking water in Europe). In the preliminary evaluation of AMD in the Minas Gerias State of Brazil, it was identified that four mining sites can generate about billions of m3 of water at pH 2–3 (de Mello et al. 2006). China has more than 63,433 metal and nonmetal mines, but the quantity of nonmetal mines surpasses 56,000 including 15,000 coal mines (Liu et al. 2018b and references therein). Shu (2003) reported the generation of 25,000 tons of waste rocks and 30,000 tons of tailings annually from the Lechang Pb/Zn mine, China. Saha and Sinha (2018) reported that more than billion tons of waste rocks and tailings are present in the mining area of Canada and have the potential to generate AMD. In India, problem of AMD is also associated with metal and coal mining areas (Tiwary 2001; Pandey et al. 2007; Equeenuddin et al. 2010; Sahoo et al. 2012). Problem of mine drainage is also present in South African mining areas (Manders et al. 2009; Ochieng et al. 2010; McCarthy 2011). Groundwater near the mining area of Johannesburg was contaminated with AMD (Naicker et al. 2003). Chon and Hwang (2000) reported that 41 coal mines in Korea discharge about 141,000 m3/day AMD waste.

In order to reduce the AMD impact on the ecosystem, different prevention measures have been recommended. However, owing to the various environmental factors associated with the mine environment, capital cost, and sustainability factor, application of these preventive methods to combat AMDs is restricted to only few mines/affected areas (Sahoo et al. 2013). Several remediation technologies including chemical and biological options are used to treat AMD and AMD-impacted sites (Johnson and Hallberg 2005a; Neculita et al. 2007; Skousen et al. 2017). In comparison to chemical treatments, biological treatments are preferred as they are eco-friendly, cost-effective, and sustainable in nature (Kieu et al. 2011; Hao et al. 2014). Sulfidogenic activity of the sulfate-reducing bacteria (SRB) is the prime principle of the biological treatments. In order to enhance the growth and activities of SRB, several organic carbon sources were tested, which in turn increase the alkalinity and metal precipitation process during the biological treatment of AMD. Nowadays, acidophilic/neutrophilic SRB consortium/mixture is also applied to enhance the treatment of AMD in sulfidogenic bioreactors (Nancucheo et al. 2017).

The present review chapter provides a broad account of the occurrence of AMD, its ecological impact on ecosystem, and available prevention measures and highlights the remediation options. Bioremediation-based options utilizing the SRB are described in detail including their taxonomic identity, metabolic capabilities, and processes.

8.2 Formation of Acid Mine Drainage

Exposure of sulfidic minerals/ores to air and water during mining activity generates highly acidic, metal- and sulfate-rich acid mine drainage (AMD) in the mining area (Nordstrom 2000). The process of AMD generation starts with the oxidation of sulfidic ores (pyrite) in presence of water and air resulting in the formation of Fe2+, SO42-, and H+ (Eq. 8.1).

$$ 2{\mathrm{Fe}\mathrm{S}}_2+7{\mathrm{O}}_2+2{\mathrm{H}}_2\mathrm{O}\to 2{\mathrm{Fe}}^{2+}+4{{\mathrm{SO}}_4}^{2-}+4{\mathrm{H}}^{+} $$
(8.1)
$$ 4{\mathrm{Fe}}^{2+}+{\mathrm{O}}_2+4{\mathrm{H}}^{+}\to 4{\mathrm{Fe}}^{3+}+2{\mathrm{H}}_2\mathrm{O} $$
(8.2)

The ferrous ions further oxidize to ferric ions (Fe3+), which act as an oxidizing agent for sulfidic ores (Eqs. 8.2 and 8.3) and also form ferric hydroxide as iron precipitates (Eq. 8.4).

$$ 2{\mathrm{Fe}\mathrm{S}}_2+14{\mathrm{Fe}}^{3+}+8{\mathrm{H}}_2\mathrm{O}-\to 15{\mathrm{Fe}}^{2+}+2{{\mathrm{SO}}_4}^{2-}+16{\mathrm{H}}^{+} $$
(8.3)
$$ 4{\mathrm{Fe}}^{3+}+12{\mathrm{H}}_2\mathrm{O}\to 4\mathrm{Fe}{\left(\mathrm{OH}\right)}_3+12{\mathrm{H}}^{+} $$
(8.4)

The generation of H+ ion increases acidity by reducing the pH of mine drainage which leads to the dissolution of other toxic heavy metals such as cadmium, copper, zinc, cobalt, nickel, lead, chromium, etc. from the host rocks or ores. These toxic metals are highly soluble in acidic environment, hence making the mine drainage hazardous to the environment (Nordstrom and Alpers 1999; Nordstrom 2000). At low pH (3–4), the oxidation of sulfidic ores is very slow (Diaby et al. 2007), but the presence of highly acidophilic chemolithothrophic iron and sulfur oxidizers in mine environment makes this process feasible, as these organisms are capable of oxidizing Fe/S for their energy-generation process (Baker and Banfield 2003). It was observed that the rate of Fe2+ oxidation is enhanced to several folds by these iron- and sulfur-oxidizing bacteria/archaea (Diaby et al. 2007 and references therein). These microorganisms use the sulfur and iron as electron donors for their energy generation (Baker and Banfield 2003; Mendez-Garcia et al. 2015) which leads to further decrease in pH of the mine drainage. Acidithiobacillus, Ferrithrix, Leptospirillum, Ferrovum, Picrophilus, Sulfobacillus, Ferroplasma, etc. are the typical iron and sulfur oxidizers involved in acid mine generation (Mendez-Garcia et al. 2015; Chen et al. 2016; Teng et al. 2017).

8.3 Ecological Impact of Acid Mine Drainage

Acid mine drainage is a multifactor system which causes ecological instability due to its hazardous nature. The major factors such as acidity and metal and sulfate concentrations vary and depend upon the prevailing geochemistry of the environment. These factors make the life of both terrestrial and aquatic systems unsustainable (Fig. 8.1). It affects the food chain, increases loss of species, changes the community structure of ecosystem, increases acidity/metal precipitation/soluble metal concentration, and reduces pH (Gray 1997). These deleterious effects of AMD on ecosystem are more pronounced than single-factor pollutants and can be divided into four different categories (chemical, physical, biological, and ecological) (Gray 1997).

Fig. 8.1
figure 1

Overview of AMD generation and its impact on environment, prevention options, and remediation strategies

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High concentration of heavy metals in AMD acts as a metabolic poison for the life in both terrestrial and aquatic systems. These metals when mixed with river streams, ponds, lakes, etc. produce hydroxide precipitate, which in turn reduces the oxygen content for life and also covers the body surface of several organisms and bed of these water bodies, hence making the life of both pelagic and benthic population inhabitable (Soucek et al. 2000; Schmidt et al. 2002; Gerhardt et al. 2004; Jennings et al. 2008; Qiu et al. 2009; Casiot et al. 2009; Hogsden and Harding 2011). Brown and Sadler (1989) reported that the loss of sodium ions from blood and oxygen in the tissue due to the accumulation of precipitates of heavy metals in gills (affect the breathing) was the main cause of fish death in AMD-contaminated water bodies. It has been found that streams that have normal pH levels but high iron concentrations have a small fish population (Koryak et al. 1972). Slaninova et al. (2014) also reported the loss of fish due to the inflow of acidic water which contained higher amount of Al and Fe into the rearing pond. The presence of heavy metals in high amount reduces the reproduction capabilities of organisms (Walton et al. 1982; Marsden and DeWreede 2000). When AMD runoff passes through the field, it reduces pH (increase in acidity) and increases heavy metal and sulfate contents of the soil, which change the nature of the soil (Gray 1997). Heavy metals have negative influences on soil fertility by diminishing the population of soil microbiota and other decomposers of soil organic matter, hence acting as a serious threat to plant productivity (Leita et al. 1995; Giller et al. 1998). Heavy metals as micronutrient support the metabolic precursor for many enzymatic reactions in plant metabolism, but when it exceeds the permissible limit, it starts hampering the growth, reproduction, flowering, etc. (Chibuike and Obiora 2014; Shahid et al. 2015). Plants have several mechanisms to overcome the stress of heavy metals, and one of them is phytoaccumulation, which leads to the bioaccumulation of these metals in the food chain/web and hence alters the ecosystem (Chibuike and Obiora 2014). Near to the mining regions, farmers use AMD-infested water for irrigation purpose, which severely affects the fertility of field (due to the acidity of AMD and accumulation of excessive heavy metal contents) and crop productivity (Garrido et al. 2009; Sun et al. 2015; Liao et al. 2016). Humans also have an indirect effect of AMD on their life. AMD-affected water bodies have low pH and high toxic metal concentration, which are used by the local population for their day-to-day life (He et al. 2018). These toxic metals cause serious diseases related to liver, kidney, brain, reproduction, etc. in humans (Jaishankar et al. 2014). It is of utmost importance to develop proper remediation technologies that can reduce its hazardous impacts on the ecosystem and continue with mining in more sustainable ways.

8.4 Acid Mine Drainage (AMD) Prevention and Control Options

Control and prevention options are considered to be of great importance to eliminate the sulfur oxidation event in waste rock and tailings, thus reducing AMD generation. There are several ways through which this can be achieved, but this is not feasible in many mining sites due to short life span or cost-intensiveness (Sahoo et al. 2013 and references therein). Five different prevention options are available: (i) physical barriers, (ii) bacterial inhibition (bactericides), (iii) chemical passivation, (iv), electrochemical, and (v) desulfurization (Kuyucak 2002) (Fig. 8.1). Physical barriers include dry and wet covers which are used to prevent the oxidation of pyritic minerals. Soil covers, plastic liners, and organic wastes have been used as dry covers to limit the oxygen transfer and oxidation of sulfidic minerals in mining region (Backes et al. 1987; Yanful and Payant 1992; Yanful and Orlandea 2000; Timms and Bennett 2000; Kuyucak 2002; Pandey et al. 2011). Water covers have also been used to prevent the oxidation of sulfidic minerals due to two reasons (i) low solubility of oxygen and (ii) diffusion rate in water as compared to air (Kuyucak 2002; Demers et al. 2008). Prevention of water migration from mining wastes via constructing interceptor structures, slurry walls, and diversion ditches also play an important role in limiting AMD generation (Kuyucak 2002). Vegetative covers are also reported as an option for reducing the acid generation, as well as phytostabilization potential of plants for accumulation of heavy metals (Reid and Naeth 2005; Tamas and Kovacs 2005; Valente et al. 2012; Wang et al. 2017). Bactericidal effect on mine tailings also provides a limited time span prevention option by killing the iron- and sulfur-oxidizing bacteria responsible for AMD generation (Kuyucak 2002; Sahoo et al. 2013 and references therein). Anion surfactants such as sodium dodecyl sulfate (SDS), sodium lauryl sulfate (SLS), and organic acids were found to be effective in reducing the growth and activity of iron- and sulfur-oxidizing bacteria (Kuyucak 2002; Lottermoser 2007). Benzoic acids, fatty acids, amines, etc. were tested at lab scales and were found to be effective in limiting the pyrite oxidation (Kleinmann et al. 1981; Onysko et al. 1984). Chemical passivation is another technique for AMD prevention which deals with the use of chemically inert organic and inorganic materials for coating the surface of sulfidic minerals to limit its oxidation (Zhang and Evangelou 1998). Various organic (humic acid, lipids, fatty acids, polyethylene polyamine, alkoxysilanes, oxalic acid, catechol, 1, 3-benzenediamidoethanthiol, etc.) and inorganic (phosphate, silica, permanganate, alkaline materials, etc.) coating materials are used nowadays to prevent AMD generation from sulfidic minerals (Elsetinow et al. 2003; Zhang et al. 2003a, b; Khummalai and Boonamnuayvitaya 2005; Hao et al. 2009; Ačai et al. 2009; Liu et al. 2013; Shu et al. 2013; Park et al. 2018). Huang and Evangelou (1994) introduced a new microencapsulation technique that uses soluble phosphate salt, hydrogen peroxide, and sodium acetate to form a stable ferric phosphate coating on iron sulfide minerals, which inhibits its oxidation by limiting the oxidant. Electrochemical is a process in which sulfidic minerals act as an electrode as these minerals are known to behave like electronic conductors (Shelp et al. 1995). Several laboratory and field tests were conducted by ENPAR Technology Inc. 2000 and Shelp et al. (1995) where sulfide tailings were converted into electrodes of an electrochemical cell which removed the dissolved oxygen and lowered the redox potential, forming a thermodynamically stable environment and eventually inhibiting the pyrite oxidation. ENPAR Technology Inc. installed a pilot-scale electrochemical system on the pyrrhotite tailings in the Sudbury basin (Ontario, Canada) and Golden Sunlight Mine tailings of Montana, USA, in 2000 and 2002, respectively. Desulfurization is a method based on froth flotation principle which is widely used these days to separate sulfide-rich minerals from the waste, resulting in low-sulfur containing tailings which are non-acid generating (Benzaazoua et al. 2000; Hesketh et al. 2010; Nagase et al. 2011).

8.5 Acid Mine Drainage Treatment Options

In order to minimize the impact of AMD on the receiving environments such as rivers, soil, and groundwater, two major types of treatment options are available: (i) abiotic treatment (application of neutralizing agents to neutralize the AMD with precipitation of metals) and (ii) biotic treatment (application of biostimulation and bioaugmentation agents). Both abiotic and biotic remediation strategies are divided into two parts: (i) active and (ii) passive (Fig. 8.1).

8.5.1 Abiotic Remediation Strategies

8.5.1.1 Active Technologies

A wide range of neutralization materials is available that can be used for the treatment of AMD (Johnson and Hallberg 2005a; Skousen et al. 2017). These neutralization agents (limestone, quicklime, hydrated lime, magnesite, dolomite, ammonia, sodium hydroxide, cement kiln dust, barium hydroxide, barium carbonate, fly ash, lime kiln dust, etc.) increase the pH of the AMD which lead to the precipitation of certain heavy metals. This pH-dependent metal precipitation is applicable for most of the heavy metals (Cu, Ni, Zn, Fe, Ag, Se, Pb, Cd, Th, Be, and Sb) but few remain unaffected [Hg. Cr (VI), Mo, and As (III)] or uncertain (Co and Bi) [Taylor et al. 2005]. This neutralization technology is adopted by several mines worldwide to reduce the acidity and recover the metals from the AMD (Skousen et al. 2017 and references therein). The choice of the neutralization agent depends upon the mine type or the metal contents of the AMD (Trumm 2010). Each neutralization agent has specific chemical properties like solubility and pH (Taylor et al. 2005). Calcium- and magnesium-based carbonate materials can be used as neutralization agents for this application but few carbonate minerals such as siderite, rhodochrosite, and ankerite may increase the acidity of Fe and Mn upon treatment of AMD (Taylor et al. 2005). There are other important active treatment options available including reverse osmosis, aeration, ion exchange, electrodialysis, and natural zeolites (Feng et al. 2000; Potgieter-Vermaak et al. 2006; Zhong et al. 2007; Motsi et al. 2009; Buzzi et al. 2013; Rodríguez-Galán et al. 2019 and references therein). Active treatments are useful in minimizing the detrimental effect of AMD, but it has few disadvantages like (i) it requires continuous input of chemicals, (ii) it is expensive, (iii) some of the chemicals are toxic and require safe handling, and (iv) it is labor- intensive.

8.5.1.2 Passive Technology

Abiotic passive treatment of AMD includes anoxic limestone drains (ALD), open limestone channels (OLC), limestone leach bed (LLB), limestone sand treatment (LST), steel slag leach bed (SSLB), limestone diversion well (LDW), and low-pH Fe oxidation channels (Skousen et al. 2017). These strategies are economically more cheap and eco-friendly than active treatment. Anoxic limestone drain (ALD) is the approach where limestone drain is covered to avoid the passage of air into the drain system, and to keep the dissolved iron in the reduced form (Skousen et al. 2017 and references therein). This process is applicable if the dissolved oxygen concentration is below 1mg/L (Taylor et al. 2005). In this system limestone dissolves and increases the pH by adding bicarbonate alkalinity (Taylor et al. 2005). Long residence time is encouraged for higher efficiency as well as low DO is required; otherwise armoring of the limestone takes place with iron minerals, reducing the rate of limestone dissolution. Open limestone channel (OLC) is the drainage system equipped with large coarse limestone over which AMD flows (Skousen et al. 2017). There are several studies in which the efficiency of OLCs has been assessed for the treatment of AMD (Ziemkiewicz et al. 1997; Ziemkiewicz and Brant 1996; Cravotta 2007; Cravotta III 2008a, b; Cravotta and Ward 2008). These studies reported the higher neutralization efficiency of the incoming AMD. It requires a large area for effective operation and gradient of the slope of the drainage such that it neither passes the AMD very quickly nor accumulates the metal precipitates within the void spaces (Skousen et al. 2017). Limestone diversion well (LDW) is consisted of well with crushed limestone. The AMD stream is diverted into these wells through pipelines to increase the alkalinity of the AMD and reduce the acidity with metal precipitation. Armoring of limestone is less due to the adequate mixing of water. Reduction of acidity with increase in pH was observed in a few studies using LDW for treatment of AMD (Arnold 1991; Faulkner and Skousen 1994; Ziemkiewicz and Brant 1996). Limestone leach bed (LLB) is a small basin carrying small-sized limestones to treat AMD or mine discharge. It can be constructed near an AMD seep or underground mine drainage (Skousen et al. 2017 and references therein). A minimum of 30 min time is sufficient to reduce 50% acidic load of moderately acidic AMD (Black et al. 1999). It was suggested that LLB was useful at the upstream of OLC, as it shortened the time and improved the efficiency (Ziemkiewicz et al. 2003). A manually flushed upflow LLB was constructed at Strattanville, PA, to treat water with 400– 650 mg L−1 acidity and pH 4.5 at a flow rate of 380–570 L min−1 (Schueck et al. 2004). Limestone sand treatment (LST) can be used for treatment of AMD stream by the addition of sand-size limestone to the stream (Skousen et al. 2017). It causes precipitation of heavy metals due to neutralization of AMD, and armoring of limestone is less due to agitation caused by the flowing streams. Steel slag leach beds (SSLB) are used to treat AMD due to the alkalinity content (45–78% CaCO3) of basic steel slag. Slags generated from stainless steel are toxic than basic steel slags. So it can be used as a potent neutralization agent for AMD treatment. AMD and metal-free water with acidity were treated by slag (Ziemkiewicz and Skousen 1998; Simmons et al. 2002a, b; Skousen et al. 2017 and references therein). Low-pH Fe oxidation channels can be used to partially treat high Fe discharges (Skousen et al. 2017). In this technique, a shallow channel is built lined with limestone or sandstone to increase the Fe oxidation.

8.5.2 Biological Remediation Technologies

The capacity of iron- and sulfate-reducing microorganisms plays substantial roles in generating alkalinity, heavy metal removal, sulfate reduction, and metal recovery, thereby attenuating the AMD-generation process (Johnson and Hallberg 2005a; Roychowdhury et al. 2015). Macrophytes are also known to play an important role in metal removal in various wetland systems (Roychowdhury et al. 2015). The excessive presence of Fe3+ and SO42- in the AMD system tends to be of major importance as they can generate alkalinity with the help of their respective reducing microbial populations. AMD system is very low in organic carbon content (less than 10 mg/L), as a result of which these indigenous microorganisms cannot promote iron- and sulfate-reducing events naturally (Johnson 2012). In order to promote the growth and activities of these microorganisms, supplementation of organic carbon substrates is required. These organic carbons act as electron donors while the available Fe3+ and SO42- act as electron acceptors for these microbial populations. Sulfate-reducing members on undergoing dissimilatory sulfate reduction process generate H2S which increases the pH of the system. On the contrary, iron reduction by iron-reducing populations generates Fe2+ in the system which on reaction with H2S forms iron sulfide, thereby removing the iron and sulfate contents from the AMD system with raise in pH (Fig. 8.2). Other metals present in the AMD such as Cu, Ni, Zn, and Cd also get precipitated in the form of their metal sulfides upon reaction with H2S. Biologically mediated remediation technologies are presented in this section in details.

Fig. 8.2
figure 2

Mechanism of sulfate- and iron-reducing bacteria in bioremediation of AMD. SRB and IRB indicate sulfate- and iron-reducing bacteria. Biochemical reactions operated by SRB/IRB during sulfate/iron reduction processes while using organic carbons (lactate, formate, and acetate) and H2. Metal precipitation reaction is represented in red (M: Metal)

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8.5.2.1 Passive Treatment Options

Different passive biological treatment options are available which can be used for the development of environment-friendly and sustainable technology for remediation of AMD (Johnson and Hallberg 2005a). Wetland systems (aerobic and anaerobic) have been employed to eliminate metals and sulfate and reduce the acidity of AMD by chemical, physical, microbial, and plant-mediated processes (Pat-Espadas et al. 2018). Aerobic wetlands depend upon the oxidation of metals and precipitation of metals in the form of metal hydroxides (Roychowdhury et al. 2015). Wetland plants remove the metals from the AMD via rhizofiltration (adsorption/adsorption/precipitation of metals in rhizosphere) and phytoextraction (uptake of metals and store them in plant bodies) (Roychowdhury et al. 2015 and references therein). Anaerobic wetland system utilizes the potential of SRB for removal of heavy metal via metal precipitation through sulfate reduction (Pat-Espadas et al. 2018). There are reports where constructed wetland systems decreased the heavy metal contents up to a great extent and reduced the acidity of AMD (Sheoran and Sheoran 2006; Nyquist and Greger 2009; Sheoran 2017; Pat-Espadas et al. 2018). Acidity of AMD, metal contents, and seasonal variations are considered to be important factors of wetland system to enhance the performance of AMD treatment. Permeable reactive barrier (PRB) is the treatment option for AMD where a reactive barrier is installed in the path of contaminated water flow (Benner et al. 1999). Reactive mixture includes limestone and organic matter which promote the activities of SRB, leading to the decrease in metal and sulfate contents and acidity (Benner et al. 1997; Waybrant et al. 1998)). The potential of different organic matters for reactive mixture had been tested for treatment of AMD, and many other studies have been conducted to assess the potential of PRB in remediation (Blowes et al. 2000; Gibert et al. 2004; Golab et al. 2006; Pagnanelli et al. 2009; Gibert et al. 2011; Shabalala et al. 2014 and references therein). Infiltration bed technology passes the surface water contaminated with AMD from the channel comprising a bed of organic substrates that promote and sustain the SRB activities (Vestola 2004). This study reported the removal of Cu, Zn, Fe, and Mn from the AMD-contaminated water up to 99%. Anoxic pond is the open pit which receives AMD and treatment of AMD is performed via addition of organic substrates and SRB (Riekkola-Vanhanen and Mustikkamaki 1997). It was reported that increase in sulfate-reduction activity was observed during treatment of AMD in anoxic pond upon supplementation with liquid manure and press juice from silage. Increase in SRB activities resulted in the decrement of heavy metal and sulfate concentration with slow rise in pH (Riekkola-Vanhanen and Mustikkamaki 1997). Injection of substrates into subsurface is the method to treat the groundwater contaminated by AMD by enhancing the growth and activity of SRB through injecting the substrates into the subsurface by boreholes (Canty 1999). Canty (1999) reported the increase in pH and removal of Al, Cd, and Zn from mine wastewater flowing through the mine shaft containing organic substrates. Unwanted discharge of substrates, leakage, and spill outs are the major concerns of this method (Canty 1999).

8.5.2.2 Active Treatment Options

Active treatment options of biologically mediated remediation technology depend upon the sulfidogenic activity of the sulfate-reducing microorganism. The sulfidogenic activity of the microorganisms radically improves the process of remediation to mitigate the impact of AMD on ecosystem. There are several advantages of active biological treatment systems such as recovery of selective heavy metals, lowering of significant amount of sulfate and heavy metals from the AMD, along with predictable and readily controllable performance of the treatment system (Jonhson and Hallberg 2005a, b; Neculita et al. 2007; Kefeni et al. 2017; Nancucheo et al. 2017; Kiran et al. 2017). Due to these properties, this technology is preferred over the passive treatment options. There are several reports in the literature where sulfidogenic bioreactors were used to treat AMD or wastewater containing heavy metals through sulfate reduction process (Hao et al. 2014; Kefeni et al. 2017; Nancucheo et al. 2017 and references therein). In the current scenario, in order to remediate AMD/acidic mine water, researchers have been employing neutrophilic sulfate-reducing populations in offline bioreactor system, i.e., avoiding the direct exposure of SRB to the acidic water which are lethal to them (Nancucheo et al. 2017). In contrast, few studies have been conducted to treat AMD with the help of acidophilic SRB in sulfidogenic bioreactors (Kimura et al. 2006; Nancucheo and Johnson 2012, Hedrich and Johnson 2014; Santos and Johnson 2017; Johnson and Santos 2018; González et al. 2019).

Several studies have been conducted using the different kinds of sulfidogenic bioreactors/microcosms/mesocosms for the treatment of acid mine drainage which include either of the strategies: (i) biostimulation, (ii) bioaugmentation, and (iii) biostimulation and bioaugmentation (Jong and Parry 2003; Kaksonen et al. 2004; Church et al. 2007; Neculita et al. 2007; Hiibel et al. 2008; Becerra et al. 2009; Bijmans et al. 2009; Hiibel et al. 2011; Neculita et al. 2011; Burns et al. 2012; Nancucheo and Johnson 2012; Xingyu et al. 2013; Zhang and Wang 2016; Nancucheo and Johnson 2014; Lefticariu et al. 2015; Zhang et al. 2016; Kefeni et al. 2017; Vasquez et al. 2018; Johnson and Santos 2018; González et al. 2019). Biostimulation and bioaugmentation are complementing each other as these strategies provide substantial remediation efficiencies than any other systems (Tyagi et al. 2011). Biostimulation denotes the addition of organic carbon as electron donor or nutrient in nutrient-deficient systems to enrich the indigenous population for the remediation purpose. Bioaugmentation describes the addition of extraneous culture of single isolates capable of performing desired functions. Biostimulation and bioaugmentation approaches denote the supplementation of nutrient and biological culture/strain in the treatment systems.

8.5.2.2.1 Biostimulation-Based Bioremediation of AMD or Mine Discharge

Different biostimulation agents were used to promote the growth and activities of SRB for the treatment of AMD. Table 8.1 tabulates the amendment use for the biostimulation-based bioremediation of AMD. A detailed list of biostimulation studies performed in last 15 years was prepared and tabulated with their major findings and type of amendments used for achieving the bioremediation goals (Table 8.1). In this section, only those studies are described where microbiological investigation was performed through mesocosms, microcosms, and bioreactors-based treatment of AMD to understand the role of indigenous microbiota in such process (Table 8.2). Kaksonen et al. (2004) reported the use of lactate and ethanol for sulfate reduction to treat AMD and found that microbial diversity of ethanol-fed bioreactors was more diverse than lactate-fed ones and were mostly dominated by members of Proteobacteria (especially Deltaproteobacteria). Kaksonen et al. (2006) reported the use of ethanol in sulfate reduction and higher HRT promoted sulfate removal up to 81%. Members such as Desulfovibrio, Desulfotomaculum, Desulfobulbus, Desulforhabdus, and Desulfobacca were found to be involved in sulfate reduction process in this study. Zagury et al. (2006) attempted to remediate the AMD through addition of six different organic carbon substrates to enhance the sulfate reduction and metal precipitation in batch reactors. The results suggested enhanced metal removal efficiencies (94–99% for Zn, Ni, Cd, Mn, and Fe). This study also suggested that mixture of organic materials was found to better in terms of sulfate reduction followed by ethanol, maple woods, and single substrates. Increase in several orders of SRB cell numbers from initial day to the end of experiment suggested the role of organic carbons in enhancing the activities of SRB. Church et al. (2007) reported the sulfate reduction at low pH and confirmed the involvement of SRB members such as Desulfosporosinus, Bacillus, and Clostridium in sulfate reduction process. Hiibel et al. (2008) compared the microbial community structure of two field-scale pilot sulfate-reducing bioreactors different in configurations used for treating AMD. The results showed that around 60% sulfate were removed from the AMD in one of the reactors, which was almost 40% greater than other bioreactors. The increase in abundance of Desulfovibrio taxa was quantified by qPCR, and clone library suggested that the presence of SRB (Desulfosporosinus, Desulfobulbus, and Desulfovibrio) in the bioreactors confirmed the enhanced reduction of sulfate. The effect of three different organic substrates (ethanol, hay and wood chips, and corn stover and wood chips) on the performance and microbial communities present in pilot-scale biochemical reactors treating mine drainage were observed (Hiibel et al. 2011). Results revealed the differences in microbial diversity across the bioreactors fed with different substrates. Desulfovibrio and Desulfomicrobium were the dominant species detected in the bioreactors. Behum et al. (2011) reported successful treatment of AMD through anaerobic sulfate-reducing bioreactor with increase in pH up to 6.8, 99% removal of Fe and Al, and 42% reduction in sulfate content. Desulfobacca was found to be the dominant SRB in this treatment. Burns et al. 2012 demonstrated the performance of pilot-scale bioreactor and microbial community dynamics treating AMD. The results indicated the successful increase in pH from 3.09 to 6.56 with lowering of total iron and sulfate levels upto 95.9% and 67.4%, respectively. The microbial diversity was assessed through 16S rRNA gene and dsrA gene-based clone library, and the results suggested the involvement of various SRB members. Nancucheo and Johnson (2012) treated synthetic AMD using glycerol and confirmed 98% reduction in sulfate with the help of sulfate-reducing bacteria (Desulfosporosinus). Lefticariu et al. (2015) suggested that organic carbon substrates are promising agents for the enhancement of sulfate reduction. In this study different herbaceous and woody organic substrate along with limestone amended bioreactors were used to treat AMD and the results showed the increase in herbaceous content resulted in more sulfate and metal removal. Microbial diversity analysis also confirmed the increase in cellulolytic and fermentative bacteria along with sulfate reducers. In another study, where different organic reactive mixtures were tested on the treatment of AMD along with role of different hydraulic retention times (HRT) on efficiency, reactive mixture and microbial diversity was also assessed (Vasquez et al. 2016a, b; Vasquez et al. 2018). The results from these studies suggested that mixtures were found to increase the pH and alkalinity, reduced the sulfate (>70%), and removed Fe2+ and Zn2+ (>99%). Different retention times also enhanced the pH and reduced the sulfate content, but the best was observed with 4 days HRT. The microbial diversity assessment of these bioreactors during the treatment suggested the enhancement in abundance of SRB belonging to genera Desulfovibrio, Desulfomicrobium, Desulfobulbus, Syntrophobacter, Desulfococcus, Desulfomona, Desulfobacter, and other fermentative groups. Influence of inorganic ligand availability and organic carbon on the performance of bioreactors and microbial community composition on the treatment of Zn-laden mine water was investigated. The results revealed the removal of zinc in the form of zinc sulfide after few days of treatment and no difference was observed in the microbial community composition of the bioreactors amended with alfalfa and wood chips. Rice bran was used as a carbon source to treat AMD through sulfate-reducing bioreactors (Sato et al. 2018). The results revealed the reduction of sulfate with successful enrichment of SRB member (Desulforhabdus), and decrement of metal concentration was observed. Giachini et al. (2018) used the poultry-derived biochar to treat AMD and was successful in the removal of sulfate and metals. Grubb et al. (2018) reported the successful removal of metals (>90%) from AMD with active involvement of SRB members.

Table 8.1 Selected AMD remediation studies conducted in last 15 years through biostimulation-based approach in bioreactors, mesocosms, and microcosms
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Table 8.2 Bioaugmentation- and biostimulation-based approaches used for bioremediation of AMD
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8.5.2.2.2 Bioaugmentation or Biostimulation and Bioaugmentation-Based Bioremediation of AMD or Mine Discharge

In contrast to biostimulation, there are several studies where the potential of acidophilic or neutrophilic sulfate-reducing bacteria was used to treat AMD. These sulfate-reducing microbial consortia were developed either from municipal waste/activated anaerobic sludge (Bai et al. 2013; Zhang et al. 2016; Zhang and Wang 2016) or mine environment (Elliott et al. 1998; Jong and Parry 2003; Nancucheo and Johnson 2012; Jing et al. 2018; Gonzalez et al. 2019). Low-pH sulfidogenic bioreactor was operated by acidophilic SRB as these members were capable to perform sulfidogenesis at low pH (Kimura et al. 2006; Nancucheo and Johnson 2012; Johnson and Santos 2018). A detailed list of bioaugmentation- and biostimulation-based treatment studies were tabulated and represented in Table 8.2. The role of SRB-containing consortium in remediation of acidic metal and sulfate-rich water has been successfully demonstrated (Nancucheo et al. 2017). Thiopaq®, by Paques, the Netherlands, and BioSulphide®, developed by BioteQ Environmental Technologies Inc., Canada, are the two patented bioreactors using SRB which are being implemented in various sites including Budelco zinc refinery, the Netherlands; Gold Mine Pueblo Viejo, Dominican Republic; Kennecott Utah Copper Mine, Utah; and Copper Queen Mine, Bisbee, Arizona, respectively, for AMD remediation and metal recovery (Nancucheo et al. 2017).

8.6 Sulfate-Reducing Bacteria and Their Application in Remediation of Acid Mine Drainage

8.6.1 Taxonomy

Sulfate-reducing prokaryotes are anaerobic and utilize sulfate (most oxidized form of sulfur) as their terminal electron acceptor for oxidation of organic compounds (fatty acids, alcohol, hydrocarbon, etc.) and hydrogen (Plugge et al. 2011; Muyzer and Stams 2008). These bacteria are known to play an important role in global carbon and sulfur cycles (Crowe et al. 2014). Sulfate-reducing prokaryotes are mainly present in Deltaproteobacteria (subdivision of Proteobacteria), while few other members are present in domain of Nitrospira, Thermodesulfobacteria, Clostridia, and Euryarchaeota/Crenarchaeota (Stahl et al. 2002; Rabus et al. 2006; Rabus et al. 2013) (Fig. 8.3). Based on rRNA sequence analysis, these SRB can be classified into four major phylogenetic lineages: Gram-negative mesophilic SRB (e.g., Desulfovibrio, Desulfomonile, Desulfobulbus), Gram-positive spore-forming SRB (e.g., Desulfotomaculum), thermophilic bacterial SRB (e.g., Thermodesulfobacterium and Thermodesulfovibrio), and thermophilic archaeal SRB (e.g., Achaeoglobus) (Castro et al. 2000). Both 16S rRNA and dsrAB (dissimilatory sulfite reductase) genes were used as molecular markers for identification of SRB in any environment. Fingerprinting of dsrAB gene through t-RFLP, DGGE, and gel-retardation analyses were used to determine the SRB populations of any ecosystems (Wagner et al. 2005; Geets et al. 2006). It is important to understand the nutritional requirement of diverse SRB to utilize their potential for bioremediation processes.

Fig. 8.3
figure 3

Neighbor-joining phylogenetic tree based on 16S rRNA gene sequence of type strain of sulfate-reducing taxa at 1000 bootstrap. Scale bar refers to a phylogenetic distance of 0.05 nucleotide substitutions per site. All the sequences of the type strains were recovered from Ribosomal Database Project except for Caldivirga and Archaeoglobus whose 16S rRNA gene sequences were retrieved from NCBI while n represents the number of type stains used for construction of phylogenetic tr

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8.6.2 Metabolic Versatility

Metabolic versatility of SRB is well established in the 1970s (Plugge et al. 2011 and references therein). In the last few decades, several biochemical studies have been conducted to understand their mode of metabolism [autotrophic (grow autotrophically with hydrogen and fixing CO2), litho-autotrophic (reduced mineral and CO2), or heterotrophic (utilize organic carbons)], suitable carbon substrates (electron donor), or terminal electron acceptors for efficient growth and energy generation process (Rabus et al. 2006; Fichtel et al. 2012; Hussain and Qazi 2016; Agostino and Rosenbaum 2018; Qian et al. 2018 and references therein). In addition to sulfate, various sulfur species (sulfite, thiosulfate, and tetrathionate) have been used by SRB as terminal electron acceptors (Rabus et al. 2006; Muyzer and Stams 2008). Various carbon compounds can be used by heterotrophic SRB including sugars, amino acids, alcohols monocarboxylic acids, dicarboxylic acids, and aromatic compounds (Fauque 1995; Fauque and Ollivier 2004; Rabus et al. 2006; Zagury et al. 2007; Muyzer and Stams 2008). Sulfate reducers are known to grow syntrophically with methanogenic populations in both sulfate-rich and sulfate-depleted marine environments for mineralization of organic carbon (Odom and singleton 1993; Raskin et al. 1996; Finke et al. 2007; Xiao et al. 2017). Sulfate-reducing populations coexist with methanogens by competing for common substrates, i.e., hydrogen for their growth and metabolism (Stams and Plugge 2009; Paulo et al. 2015; Ozuolmez et al. 2015). Thus, some SRB showed dual life styles: sulfidogenic and syntrophic metabolism (Plugge et al. 2011). These dual metabolic potentials improve the chance of survival of SRB under sulfate depletion condition.

8.6.3 Dissimilatory Sulfate Reduction Pathway

Sulfate reduction process occurs in three steps: (i) activation of sulfate with two ATP molecules by ATP sulfurylase to generate adenosine phosphosulfate (APS) and pyrophosphate, (ii) reduction of APS into sulfite by APS reductase with release of AMP and consumption of two electrons, and (iii) reduction of sulfite to sulfide. Two mechanisms were proposed for the last step of sulfate reduction: (i) direct pathway, in which sulfite is directly converted to sulfide with the help of bisulfite reductase (Peck and LeGall 1982), and (ii) trithionate pathway, in which sulfite is first converted to trithionate which is further converted into thiosulfate, followed by reduction of thiosulfate to sulfide with the help of sulfite reductase, trithionate reductase, and thiosulfate reductase, respectively (Kobayashi et al. 1969, 1974). Chambers and Trudinger (1975) reported that neither thiosulfate nor trithionate was a normal intermediate in the reduction pathway. Santos et al. (2015) reported that sulfite was first converted to DrsC trisulfide by dsrAB/dsrC genes which was further reduced to DsrCr and sulfide by dsrMKJOP complex. They further reported that DsrC was a co-substrate for sulfite reduction by dsrAB and absence of DsrC or high bisulfite concentration lead to the production of thiosulfate and trithionate as in vitro products of dsrAB. In order to understand the rate of sulfate reduction and sulfate transportation with the formation of different intermediary products intracellularly, isotopic fraction of sulfur was examined to understand the sulfur transfer pathway in dissimilatory sulfate reduction (Sim et al. 2017).

8.6.4 Substrate-Level Phosphorylation

Sulfate-reducing microbes are known to utilize diverse organic carbon for their metabolism. Based on their mode of oxidation, they are categorized into (i) complete oxidizers – which completely oxidize organic carbon to CO2, and (ii) incomplete oxidizers, which form acetate as an end product on oxidation of organic carbon (Castro et al. 2002). Out of the known SRB species, 70% of the species fall into incomplete oxidation category (Qian et al. 2018).

Lactate is considered to be the most common organic carbon substrate/electron donor for the SRB (Plugge et al. 2011). It is oxidized to pyruvate by a D-or L-lactate dehydrogenase which is ultimately oxidized to acetate. Lactate oxidation yields two molecules of ATP through substrate-level phosphorylation, which are used for activation of sulfate during first step of sulfate reduction pathway resulting in no net gain of ATP. Two hypothetical models are proposed for electron transfer from lactate oxidation to sulfate reduction process: (i) H2 cycling pathway in which H2 is produced from cytoplasmic and periplasmic hydrogenases which ultimately transfers electron to terminal reductase via cytochrome and transmembrane complexes (Heidelberg et al. 2004) and (ii) second pathway that transfers electrons from lactate oxidation directly to the membrane-bound menaquinone pool which further transfers the electrons to terminal reductase via transmembrane-bound complexes (Keller and Wall 2011). Citric acid cycle (CAC) and oxidative carbon monoxide dehydrogenase (OCMD) pathways are operated in SRB to completely oxidize the acetate (Goevert and Conrad 2008 and references therein). But most of the organisms utilize OCMD as it operates at more negative redox potential [greater than that of the APS/HSO3- couple (-60 mV) and HSO3- /HS- couple (-116 mV)] than CAC pathway (Thauer et al. 2007). Propionate is another electron donor and carbon source for several incomplete and complete SRB (Widdel and Pfennig 1982; Suzuki et al. 2007; El Houari et al. 2017). Methylmalonyl-CoA pathway was considered to be operated in Desulfobulbus propionicus for its incomplete oxidation (Widdel and Pfenning 1982). This pathway was found to be more thermodynamically/energetically favorable than oxidation of propionyl-CoA via acrylyl-CoA, because redox potential of the acrylyl-CoA/propionyl-CoA couple is more positive (+69 mV) than fumarate/succinate couple (þ33 mV) (Barton and Hamilton 2007). The completely oxidizing sulfate reducers use the classical b-oxidation pathway to degrade fatty acids (Janssen and Schink 1995a, b). The electron donor or carbon source plays an important role in the metabolic pathways of SRB. It was reported that several other important organic carbons were utilized by SRB for sulfate reduction. Complex polysaccharides, amino acid/protein, or alcohol enhance the sulfate-reduction ability of SRB (Neculita et al. 2007; Schmidtova and Baldwin 2011; Ayala-Parra et al. 2016; Hussain et al. 2016).

8.6.5 Electron Transport Phosphorylation

Electron transport phosphorylation occurs in SRB where H2 acts as an electron donor for dissimilatory sulfate reduction. In this pathway, H2 is oxidized by periplasmic hydrogenase (Hase1) of SRB to generate electron and H+ ion. The generated electron transfers to cytoplasmic heme type cytochrome C3, which transfers it to cytoplasmic dissimilatory sulfite reductase via transmembrane complex (da Silva et al. 2013). The H+ ion generated in this mechanism is pumped across the cytoplasmic membrane with proton motive pump. This results in a H+ gradient between the cytoplasm and periplasm resulting in the generation of ATP through phosphorylation of ADP with the help of membrane-bound ATP synthase (Fitz and Cypionka 1991). Odom and Peck (1981) reported the production of H2 during the growth of Desulfovibrio species on lactate/sulfate media. But genetic evidence confirmed that H2 cycling in Desullfovibrio alaskensis G20 mutant of type 1 cytochrome c3, Desulfovibrio vulgaris Hildenborough, mutants of Tp1-c3, Desulfovibrio gigas mutants of cytoplasmic hydrogenase or periplasmic hydrogenase, was not required for sulfate reduction (Heidelberg et al. 2004; Li et al. 2009; Keller and wall 2011; Sim et al. 2013; Morais-Silva et al. 2013; Keller et al. 2014; Price et al. 2014). Another potential mechanism for electron-transport phosphorylation is formate cycling. Heidelberg et al. (2004) reported that in Desulfovibrio vulgaris Hildenborough, formate is oxidized by formate dehydrogenases present in the periplasm. A putative cytoplasmic formate: hydrogen lyase was reported from the genome of D. alaskensis G20 which could convert cytoplasmic formic acid to H2 and CO2 or vice versa (Pereira et al. 2011). Electrons generated from the oxidation of formate can be transferred to cytoplasmic membrane for sulfate reduction via periplasmic electron carriers and the produced H+ ion creates a gradient across the membrane for ATP generation. Knockout of Desulfovibrio vulgaris Hildenborough for formate dehydrogenase showed reduced growth in lactate/sulfate media indicated the role of formate cycling in energy production (da Silva et al. 2013). In a cycA mutant of Desulfovibrio alaskensis G20, accumulation of formate was observed on lactate/sulfate media which suggested that formate formed during oxidation of lactate oxidation was immediately reoxidized. (Li et al. 2009). There are several transmembrane redox complexes such as Qrc complex, Qmo complex, DsrMKJOP complex, Hmc (high-molecular weight cytochrome) and Tmc (with a periplasmic type II cytochrome c3 subunit), Hdr/flox, nfn, and Rnf identified in SRB which are known to be involved in electron transfer to cytoplasm (Pereira et al. 1998, 2006; Li et al. 2009; Venceslau et al. 2010; Grein et al. 2010; Pereira et al. 2011; Ramos et al. 2012; Quintas et al. 2013; Keller et al. 2014; Meyer et al. 2014).

8.6.6 Extracellular Electron Transfer Between ANME and SRB

Anaerobic methane oxidation by archaeal members in association with sulfate-reducing bacteria offers extracellular electron transfer to SRB for dissimilatory sulfate reduction process (Boetius et al. 2000). Various studies have been conducted to understand the genetic potential of ANME archaea and syntrophic association with SRB members (Skennerton et al. 2017 and references therein). The two possible mechanisms of methane oxidation coupled with sulfate reduction were proposed (Milucka et al. 2012; McGlynn et al. 2015; Wegener et al. 2015; Scheller et al. 2016; Skennerton et al. 2017). Wegener et al. (2015) suggested that under thermophilic anaerobic methane oxidation (TAOM) conditions, both ANME and the HotSeep-1 bacteria overexpressed the extracellular cytochrome production genes and formed cell-to-cell connections by expressing the genes for pili production which is responsible for interspecies electron transfer between syntrophic consortia. In contrast to the previous mechanism, Milucka et al. (2012) confirmed that ANME-2 oxidized methane with a concomitant reduction of sulphate to zero-valent sulfur for extracellular electron transfer. Hence, it can be concluded that AOM is not an obligate syntrophic process. Produced S0 is exported outside the cell where it reacts with sulfide to form disulfide. Disulfide is disproportionated to sulfate and sulfide by sulfate-reducing Deltaproteobacterial members.

8.6.7 SRB in Acidic Condition and Factors Affecting SRB Activities

It was believed that SRB thrive in environmental conditions above pH 5.0 (Hao et al. 1996). In the last decades, several studies confirmed the possibility of sulfate reduction at low pH condition (pH < 5.0) (Sánchez-Andrea et al. 2014). The presence of SRB in acidic environment has been reported in several AMD sites and was suggested that the existence of micro-niches of higher pH in such environments supported the growth and metabolism of SRB (Sánchez-Andrea et al. 2014 and references therein). It is reported that low pH destabilizes the macromolecules (proteins get denatured), changes the conformation of surface protein due to protonation as well as increases energy cost for pH homeostasis (Golyshina and Timmis 2005; Krulwich et al. 2011). Acidophilic organisms have different strategies to maintain the circum-neutral pH in the cytosol by pumping out the H+ ion, changing the permeability of membrane, increasing the positive redox potential of cytoplasm through potassium transporter, and decarboxylation of amino acids (Chen et al. 2015 and references therein). At low pH, H2S and organic acids are also toxic to SRB (Koschorreck 2008). Sulfate reduction process generates H2S as the main product, which remains freely present in undissociated form at pH < 5.0 (Moosa and Harrison 2006). Furthermore, it has a negative impact on cells due its toxic nature as well as high membrane permeability. There are several reports where inhibition of sulfate reduction with increase in sulfiide concentration was observed (Reis et al. 1992; Icgen and Harrison 2006; Koschorreck 2008; Cassidy et al. 2015 and references therein). Similarly, low molecular weight organic compounds such as acetate and lactate are mostly in undissociated state at low pH, which may pass the cell membrane and get dissociated inside the cytoplasm releasing proton which in turn lowers the pH (Sanchez-Andrea et al. 2014 and references therein). The nonionic substrates such as glycerol, hydrogen, and ethanol were more preferred at low pH (Nagpal et al. 2000; Kaksonen et al. 2004; Bijmans et al. 2009). Benner et al. 1999 reported that anaerobic bacteria change their mode of carbohydrate metabolism to avoid the accumulation of organic acid inside the cytoplasm at low pH condition. There are reports where SRB activity decreases with increasing undissociated concentration of lactate or acetate (Reis et al. 1990; Jong and Parry 2006). Other than the abovementioned factors, several other factors are known to be inhibitory for sulfate-reducing taxa. Heavy metals are considered to be toxic to several SRB above certain threshold (Utgikar et al. 2002; Sani et al. 2001). Effect of copper toxicity on sulfate removal efficiency was observed and found that 1.57mM copper concentration reduced the efficiency of SRB for sulfate removal to 50% (Song et al. 1998). Sani et al. 2001 showed the toxic effects of copper, zinc, and lead on Desulfovibrio desulfurican G20. It was found that at copper concentration of 13.3uM, specific growth rate of D. desulfurican G20 got reduced to 50%. In another study 10uM concentration of lead was shown as minimum inhibitory concentration for Desulfovibrio desulfurican G20 at pH 6.0 (Sani et al. 2003). These studies suggest that heavy metals increased the lag phase and lower the specific growth rates of the tested species. Toxic effects of several heavy metals [Cu(II), Zn(II)], Mn(II), Ni(II), and Cr(III)] were evaluated through batch studies on two SRB cultures, Desulfovibrio vulgaris and Desulfovibrio sp (Cabrera et al. 2006 and references therein). The results indicated Cu > Ni > Mn > Cr > Zn showed maximum inhibitory effect on both the SRB cultures. Higher concentration of molybdate may act as an inhibitor for sulfate reduction as it is a structural analog of sulfate (Biswas et al. 2009; de Jesus et al. 2015). Hussain and Qazi (2016) reported that higher concentration (10 and 15 ppm) of Cu, Cr, and Ni inhibited bioprecipitation and sulphate reduction by Desulfotomaculum reducens-HA1 and Desulfotomaculum hydrothermale-HA2. It was reported that sulfate reduction could not be observed at oxic condition or in the presence of nitrate and nitrite (Rubio-Rincón et al. 2017). It was suggested that nitrite formation by nitrate inhibited the SRB activities, because nitrite suppressed the sulfate reduction (Barton and Hamilton 2007; Mohanakrishnan et al. 2008). SRB are known to reduce oxygen as a protective mechanism against its harmful effects (Lamrabet et al. 2011 and reference therein).

8.6.8 Resource Recovery from AMD Treatment Using SRB

Sulfate-reducing bacteria are known to have diverse metabolic properties and have been used in various treatment processes. The ability of SRB to produce biohydrogen, aliphatic hydrocarbon, polyhydroxyalkanoates (PHA), magnetite, and metal sulfides is well documented (Gramp et al. 2007; Gramp et al. 2010; Yu et al. 2011; Peltier et al. 2011; Friedman and Rude 2012; Martins and Pereira 2013; Hai et al. 2004; Hedrich and Johnson 2014; Martins et al. 2015). Sulfate reducers are capable of oxidizing various organic carbons and produce sulfide as the main product of sulfate reduction. Interaction of sulfide with the metals such as Fe, Cu, Ni, Zn, and Cd lead to the formation of metal sulfides. Sulfidogenic treatment of AMD leads to generation of metal sulfide, and also the generated sulfide could be used for the production of elemental sulfur via oxidation (aeration) (Nancucheo et al. 2017). There are two strategies generally used for metal sulfide production during treatment of AMD: (i) direct interaction of sulfide and metals (in-line sulfidogenic reactors) and (ii) sulfide produced during the sulfidogenesis is separated and used as a metal precipitating agent (offline sulfidogenic reactors). Thioteq (Paques) and BioSulfide (BioteQ) are the two offline sulfidogenic systems used for metal recovery from the AMD/wastewater treatment. Biogenic Ni and Zn sulfide precipitation were observed during the growth of SRB in Ni- and Zn-containing medium (Gramp et al. 2007). The Ni-sulfidic phase (Heazelwoodite) was more crystalline than its abiotic control setup. Gramp et al. (2009) reported the formation of iron sulfides (mackinawite and greigite) by the sulfate-reducing bacteria, and with the increase in incubation temperature of the bacteria, crystallinity of the iron sulfides got enhanced. Castillo et al. (2012) observed the zinc precipitation in the two polymorphs of ZnS (sphalerite and wurztite) while investigating the tolerance of SRB against zinc-rich sulfate medium for the application of SRB in AMD remediation. Hedrich and Johnson (2014) used the modular bioreactor for the recovery of metal sulfide (ZnS) during the treatment of metal-containing wastewaster. Zhou et al. (2014) reported the iron sulfide precipitation during the extended incubation of Desulfovibrio vulgaris, and higher sulfide accumulation led to the transformation of mackinawte to gregite. Murray et al. (2017) reported the use of waste H2S gas from AMD remediation to synthesize zinc sulfide quantum dots, which was found to be different in its physical and optical properties from its other chemical counterparts. Picard et al. (2018) reported that Desulfovibrio hydrothermalis AM13 influenced the nucleation and growth of iron sulfidic minerals. Overall, metal sulfides are known to have diverse applications including cancer therapy, antimicrobial agents, textile, optoelectronics, hydro-cracking in fuel refineries, and hydro-processing (Qian et al. 2018 and references therein).

8.7 Conclusion

The toxic and lethal effects of AMD or acidic mine water on environmental health are of major concern. This problem has gained attention from the researchers to find out a promising technology for its treatment or measures to prevent its generation. There are several prevention measures through which its generation can be checked, but several factors limit these options. There are remediation technologies being used including both passive and active treatment strategies. But the choice of remediation technologies depends upon various factors such as (i) mine type, (ii) type of waste generated, (iii) geochemistry of AMD, (iv) cost of the treatment system, and (v) sustainability and long-term treatment options. Due to high capital investment, requirement of large area, continuous input of neutralization agents, sludge accumulation and its disposal, as well as sustainability of chemical treatment options, biological treatments are found to be more attractive nowadays. Biological treatment provides cost-effective, sustainable, and eco-friendly options to treat AMD and recover metals from the AMD environment. Advancement in the molecular approaches allowed the researchers to gain a deeper insight of microbial community and composition, their functions, and interaction in such processes. Sulfidogenic activity and metabolic versatility of SRB play an important role in attenuation of AMD and metal precipitation. In future, research is needed to develop this biological technology for its application in more promising ways to treat AMD or AMD-impacted areas.