Microbial Treatment of Industrial Wastewater

CHAPTER 1

Microbial Treatment of Industrial Wastewater 1.1 INTRODUCTION While industrial wastewater is a generic term involving a wide array of wastewater discharged out of various industries that could hardly be captured in a book, this book is about wastewater mostly produced by metalliferous industries and mining. The increasing trend of metal consumption has resulted in an increase in industrial activities such as mining, metal processing, smelting, alloy casting, and silver refineries, etc., resulting in the generation of metals containing wastewater with different anions, such as predominantly ions of sulfur, phosphorus, and carbon, etc. The discussions in this book will deal with sulfur, because of the criticality it poses. Sulfur can be present in a long range valence state from 22 (completely reduced sulfide) to 16 (completely oxidized sulfate) in the earth’s crust. Sulfate (SO22 4 ) is the second most abundant anion in the sea, after chloride, and in rivers, after bicarbonate (Middelburg, 2000), and is a common pollutant in most industrial wastewater. The presence of sulfate and metal in natural or wastewater can be due to either natural emission, or anthropogenic activities (Lens et al., 1998). The current global estimate of sulfate flux is 430 Tg/y (teragrams per year), or 430 million tonnes per year, of which the annual anthropogenic sulfate flux range is from 138 to 178 Tg/y. Around 72 Tg/y is considered to be the maximum sulfate flux from nonfertilizers and nonfuel burning industries (Nordstrom, 2011). In the natural sulfur cycle, sulfate can be generated through geochemical or biological pathways. Oxidative weathering of sulfide-bearing rocks leads to the formation of sulfate and dissolved metals. This process can also be accelerated by microbial involvement (Bhattacharya et al., 2006). Sulfide-oxidizing bacteria may lead to the formation of elemental sulfur (S0) from sulfide (S22), which is eventually further oxidized by other microbial species for their metabolic requirements (Bruser et al., 2000). Sulfide may be oxidized aerobically by chemolithotrophic sulfur-oxidizing bacteria (Thiobacillus or Beggiatoa spp.), or anaerobically by phototrophic sulfur bacteria (Chlorobium spp.) to Low Cost Wastewater Bioremediation Technology DOI: http://dx.doi.org/10.1016/B978-0-12-812510-6.00001-2

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elemental sulfur and sulfate (Muyzer and Stams, 2008). Elemental sulfur disproportionation by some microbial species (Desulfovibrio or Desulfobulbus spp.) also leads to the formation of sulphate in water (Muyzer and Stams, 2008; Hardisty et al., 2013). Rain water percolating through sulfate salt reservoirs dissolve sulfate and metal into groundwater. Atmospheric sulfate produced due to sea salt aerosols or by volcanic eruption can be bound with rain water and find its way into surface water as acid rain (Lens et al., 1998).

1.1.1 Sources of Sulfate and Metal-Rich Wastewater Industrial wastewater containing dissolved metals and sulfate is produced from different industries such as electroplating, metal processing, textiles, tanneries, oil refineries, and mining (Barakat, 2011). For example, the wastewater from the electroplating industry contains significant amounts of heavy metals such as Cd, Zn, Cr, Ni, Pb, etc., (Rahman et al., 2015); the tannery industry generates Cr and sulfate-rich wastewater (El-Sherif et al., 2013); the petroleum refinery industry generates wastewater containing Se, V, Zn, Al, Se, Mn, etc., (Diya’Uddeen et al., 2011); the Table 1.1 Lists of the pollutants generated from different industrial activities Name of the industry Pollutants Reference

Electroplating Paper and pulp processing Circuit board printing industry Oil refinery

Paint industry

Textiles Tannery Mining

Cu, Cr, Mg, Fe, SO22 4 SO22 4 , Ca, Fe, Mg, Mn, Cu Cu, Zn, SO22 4 , Ni, Pb, As, Se, V, Cd, Cu, Fe, Pb, Mn, Ni, Zn, Hg, Cr, SO22 4 SO22 4 , Al, Cr, Cu, Fe, Mg, Ni, Pb, Si, Sr, Ti, Zn, Zr SO22 4 , Fe, Cu, Zn, Ni, Cd, Pb, Cr SO22 4 , Cr SO22 4 , Fe, Ni, Cu, Mn, Mg, Pb, Cd, Zn, Co, Al

Yalc¸in et al. (2001), Cavaco et al. (2007), Rahman et al. (2015) Tavares et al. (2002), Arivoli et al. (2015) Chang (1995), Chen and Huang (2014) Ismail and Beddri (2009), Diya’Uddeen et al. (2011) Gondal and Hussain (2007), Malakootian et al. (2009) Malamis et al. (2011), Prabha et al. (2015) Lens et al. (1998), El-Sherif et al. (2013) Johnson and Hallberg (2005), Equeenuddin et al. (2010); Nordstrom et al. (2015)

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wastewater from the mining industry known as acid mine drainage (AMD), in addition to being acidic at times, contains elevated concentrations of SO24 , Fe, Ni, Zn, Co, Mn, and Pb. A general list of the pollutants generated from different industrial activities is presented in Table 1.1. Among the industries mentioned, the major amount of sulfate and metalrich wastewater is generated from the mining industries. Obviously depending on the H1 concentration, such drainage can be acidic, circumneutral, or alkaline. During excavation and mining, the exposure of the sulfide, (oxy) hydroxide, and hydroxysulfate minerals into atmospheric oxygen leads to oxidation, resulting in the release of SO22 4 , Fe21, and H1 (Eq. 1.1). Further oxidation of Fe21 to Fe31 leads to the generation of more acidity (Eq. 1.2 and 1.3). Similarly, the oxidation of trace elements associated with sulfide minerals results in the generation of acid mine drainage (AMD) containing toxic dissolved metals such as As, Cu, Mo, Ni, Cd, Zn, and Pb (Nordstrom et al., 2015). 1 FeS2 1 7=2O2 1 H2 O-Fe21 1 2SO22 4 1 2H

(1.1)

Fe21 1 1=4O2 1 H1 -Fe31 1 1=2H2 O

(1.2)

Fe31 1 H2 O-FeðOHÞ21 1 H1

(1.3)

In addition to the geochemical process, the microbial process also plays an important role in the generation of sulfate and metal-rich wastewater from sites including mines and excavations. Iron oxidizing bacteria such as Acidithiobacillus ferrooxidans, Acidianus brierleyi, Ferroplasma acidarmanus, etc., are reported to oxidize pyrite, arsenopyrite, chalcopyrite, marcasite, and sphalerite, resulting in the dissolution of the metals, acidification, and the production of AMD (Baker and Banfield, 2003; Akcil and Koldas, 2006).

1.1.2 Characteristics of Sulfate and Metal-Rich Wastewater The nature of all sulfate and metal-rich industrial wastewaters is not the same as far as the presence of other pollutants are concerned. The important parameters for the presence of sulfate and dissolved metals are the pH, the abundance of chemical oxygen demand (COD), and the organic source of sulfate and metals in the wastewater. When sulfuric acid is used during the processing steps, most of the wastewater evolved is of low pH (tannery, textile or rubber, and the mineral processing industry). Still, it is fair to state that only in a few cases is the effluent water from these

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sources primarily acidic. More than 70% of mines throughout the world are facing the problem of acidic metal and sulfate-rich wastewater discharges (Lopez et al., 2009). The oxidation of pyritic minerals in active or abandoned coal or metal mines and earth work excavations results in the formation of acidic (pH as low as 2 to 4) metal and sulfate-rich wastewaters, often called acid mine drainage (AMD) or acid rock drainage (ARD). In contrast, the oxidation of common metallic sulfides may not release sufficient acid, and hence can result in sulfate and metal-rich circumneutral water. Even the buffering reaction due to the presence of carbonate rock (calcite or dolomite) or limestone may lead to alkaline water with sulfate and dissolved metals (Banks et al., 1997; Bigham and Nordstrom, 2000). Nevertheless, the presence of sulfate in higher amounts along with heavy metals is always found in low pH waters. Targeted metals are also found in mineral processing wastewater (Usinowicz et al., 2006). Chromium is reported to be present in chromium-tannery effluents (Genschow et al., 1996). These industrial wastewaters can be categorized by the presence of high and low levels of COD. The production processes of food processing, pulp and paper, textile, tannery, petroleum, and rubber processing industries deal with organic substances that discharge high levels of COD with their effluents (Lens et al., 1998; Saritpongteerakaa and Chaiprapat, 2008; Dutta et al., 2010; Li et al., 2012). The COD in such wastewater is generally comprised of low chain volatile fatty acid (VFA), along with some ammonium nitrogen and phosphate. Low chain VFA is an oxidized substance of sulfate that can be reduced by sulfate reducing bacteria (SRB) which naturally treat such water, and can be used for treatment (Liamleam and Annachhatre, 2007; Neculita et al., 2007).

1.1.3 Effect of Sulfate and Metal-Rich Wastewater The impact of the production and discharge of sulfate and metal-rich wastewater can be categorized into three classes of aspects: environmental impacts, impacts on health, and problems faced during industrial operations. Sulfate is a chemically inert, non-volatile, and non-toxic compound; therefore, the damage caused by sulfate emissions is not direct. Sulfate ions individually though are not so toxic, but the dissolved metals which are an obvious companion during sulfate discharge can have a serious impact on the environment, as well as on human health. Different oxidized forms of sulfur (sulfate, sulfite, or thiosulfate) can be present at the same time in the

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wastewater, and each can undergo several reduction paths with different available organic sources (Nikzad, 2007). These organic sources are intermediates of the natural mineralization process. Therefore, discharge of high sulfate concentrations into wastewater can cause various imbalances of the natural sulfur cycle (Lens et al., 1998). Problems caused by sulfates are most frequently related to their ability to form strong acids which change the pH of a water source. Acid, metal, and sulphate-rich wastewater from mines, textiles, rubber, or paper mill industries as a whole can have a severe impact on ecological balance. These waters have a major impact on rivers, lakes, estuaries, and coastal ecosystems. Major effects can be categorized into chemical, physical, biological, and ecological; however, the overall effect is the elimination of species, destroying the primary food chain and thereby reducing the ecological stability of any aquatic ecosystem (Gray, 1997; Jennings et al., 2008). Considering the volume and complexity of wastewater, the severity is practically immeasurable. The migration of sulfate from its source to nearby geological strata is another significant problem. Sulfate from earth works, excavations and mines, or landfill leachates can migrate and contaminate nearby aquifers (MacFarlane et al., 1983). Groundwater contamination from mining or agricultural originated sulfate can lead to an enormous increase in drinking water sulfate levels (Toran, 1987; Liu et al., 2013). Sulfate is harmful to health compared to other pollutants, in different ways. Its concentration at more than 250 mg/L in drinking water gives an unpleasant medicinal taste. Water containing high levels of sulfates, particularly magnesium sulfate (Epsom salts) and sodium sulfate (Glauber’s salt), may have a laxative effect on persons unaccustomed to that water (Nikzad, 2007). People unaccustomed to drinking water with elevated levels of sulfate can experience catharsis, dehydration, diarrhea, and gastrointestinal irritation. Infants are often more sensitive to sulfate than adults. Animals are also sensitive to high levels of sulfate. In young animals, high levels may be associated with severe, chronic diarrhea, and in a few instances, death. Exposure of sulphate-rich wastewaters has been recently reported to have acute genotoxic effect on human leukocytes (Mihaljevic et al., 2011). Mineral processing units and coal washeries are generally located near their respective mines. These processing plants depend on the mines for high amounts of water necessary for their operation. Also, rubber processing and textile industries need high water levels for processing, and they generally recycle parts of the effluent water. The effluents containing high

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sulfates effectively increase the total dissolved solid (TDS) and salinity that often limits the number of cycles for reuse of such water (Bowell, 2004). Scaling problems from sulfate salt produce a major drawback (Bader, 2007). This poses a great challenge where these industries are active in semi-arid to arid environments (Bowell, 2004). Mines are nowadays thought be a source of usable water. Some mining companies are also trying to reuse the wastewater for irrigation purposes, or as drinking water to recoup part of their expenses (Usinowicz et al., 2006). Sulfate, the common pollutant in mine waters, is therefore a target. Regulatory agencies are becoming increasingly concerned over elevated sulfate concentrations in effluent, owing largely to its impact on the salinity of receiving waters.

1.1.4 Treatment Technologies Treatment of sulfate and metal-rich wastewater can be performed using different methods (Fig. 1.1) that are mainly divided into two different groups, such as abiotic treatment and biotic treatment. Abiotic treatments are divided into both active and passive systems (Taylor et al., 2005; USEPA, 2014). 1.1.4.1 Limestone Diversion Well (LDW) The limestone diversion well (LDW) is a well containing crushed limestone aggregate. During treatment, acidic sulfate and metal-rich wastewater are diverted into the well through a pipeline. The hydraulic force of water flow causes grinding of the limestone gravel, and the contact with lime causes

Figure 1.1 Different treatment methods for sulfate and metal-rich wastewater.

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the removal of acidity and precipitation of dissolved metals (Taylor et al., 2005). Armoring of limestone particles over time, however, brings down the efficiency of operation. 1.1.4.2 Adsorption Adsorption techniques use metal absorption sites, e.g., the surface of carbonized peanut shells, leaf litter, etc., for the removal of metals from wastewater. Ferrihydrite adsorption is used for the removal of both Se and As from wastewater. In this method, ferric salt is added into the wastewater to generate ferric hydroxide and ferrihydrite precipitate which adsorbs the Se and As from the wastewater (USEPA, 2014). The process is still limited, due to the early saturation of adsorption sites and operational difficulty for site regeneration, in the face of high loads. 1.1.4.3 Ion Exchange In this process, ion exchange resins are used to remove the metal contaminants from the wastewater. This process is useful for the removal of both hardness and metals (Da˛browski et al., 2004). The presence of strong oxidizing agents can limit the process performance of their ability to degrade the resins. The process is economically expensive (USEPA, 2014). Significant new developments have taken place in the field in recent years (Fu and Wang, 2011). The typical ion exchange process is presented in Fig. 1.2. 1.1.4.4 Evaporation Evaporation techniques, such as solar and mechanical evaporation, are used to treat such industrial wastewater. In this process, the water is evaporated

Figure 1.2 Treatment of sulfate and metal-rich wastewater using the ion exchange method.

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from the wastewater stream, and the contaminants are concentrated as a solid or in a brine stream ( Jakob et al., 1996). This process is not always suitable for wastewater treatment in cold climatic areas, and also where large quantities of wastewater are produced and the space is limited. This treatment method has scope to create groundwater contamination. The requirement for additional storage ponds, regular clean ups, and revegetation make the processing economically ineffective over a long time (USEPA, 2014). 1.1.4.5 Electrodialysis Reversal (EDR) This process applies electrode polarity reversal on an ion exchange membrane during the removal of the contaminants from the wastewater (Chao and Liang, 2008). This process requires the disposal of wastes generated from concentrated waste stream, electrode cleaning, and residuals, etc. This process is used for the removal of As, Ra, SO22 4 , Ca, and nitrates from the wastewater (Chao and Liang, 2008; Montan˜a et al., 2013; USEPA, 2014). Using the EDR system, sulfate was removed at 80% (Chao and Liang, 2008) and 90% (Roquebert et al., 2000; Karimi and Ghassemi, 2016) from industrial wastewater and contaminated water, respectively. The EDR system was reported to successfully treat the metal-rich wastewater generated from the battery manufacturing industry, and removed almost completely Pb, Fe, and Cu from the wastewater (Chen and Jiang, 2011). The treatment of electroplating industry wastewater using the EDR system remove 97% of the Ni from the wastewater (Lu et al., 2014). 1.1.4.6 Electrocoagulation In this technique, graphite or stainless steel is used as a cathode in combination with a metal anode to remove the dissolved metals present in the wastewater. Upon applying a voltage, the anode generates charged ions which subsequently precipitate the dissolved metals present in the wastewater (Al-Shannag et al., 2015). In this process, the anode material depends on the pollutants being removed from the wastewater. Using this process, a lower amount of sludge is generated when compared to the chemical process, and little maintenance is required. Metals such as As, Cu, Zn, Pb, Cr, Ni, etc., are reported to be removed using this technique (Oncel et al., 2013). The electrocoagulation process removed 68.5% (Hossini et al., 2015) and 98.8% ( Jo et al., 2016) of sulfate from synthetic and wet scrubber wastewater, respectively. The efficiency of the treatment process has been reported to be dependent on electrode materials,

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voltage, and reaction time ( Jo et al., 2016). However, the method is still very expensive (USEPA, 2014; Al-Shannag et al., 2015). 1.1.4.7 Reverse Osmosis (RO) This technique is based on the pressure driven separation process of the contaminants from the wastewater using semi-permeable membranes. The successful operation of the process requires high pressure. As scaling or fouling is the major problem in the membrane-based treatment process, there is a requirement for a pretreatment process (Fu and Wang, 2011). Different dissolved metals such as Fe, Cu, Zn, Ni, Mg, etc., and sulfate are removed from wastewater using this technique (USEPA, 2014). The RO process was reported to remove almost all sulfate from the wastewater generated from the steel industry and metal finishing industry (Petrinic et al., 2015; Colla et al., 2016). The process was reported to remove Ni and Cu completely from wastewater (Qdais and Moussa, 2004; Silva et al., 2016). Similarly, all the metals (Cr, Cu, Fe, Ni, Sn, and Na) were reported to be removed completely from metal finishing wastewater (Benito and Ruı´z, 2002). 1.1.4.8 Photoreduction In this process, a photocatalyst (e.g., TiO2) is used to generate electronhole pairs using ultraviolet light. The electrons and holes subsequently induce the redox reaction of contaminants absorbed on the surface of the photocatalyst (Chen and Ray, 2001). After the treatment is over, the photocatalyst is regenerated by desorption (USEPA, 2014). The hollow, spherical-shaped TiO2 particles are reported to remove almost all Cr from water at pH 2.82 (Cai et al., 2017). Similarly, using H2O2 as a photocatalyst, Cr was reported to be removed completely at pH 2 (Chaudhary and Singh, 2014). The authors also reported complete removal of Cu and Zn, and 94.8% Ni removal under alkaline conditions. 1.1.4.9 Anoxic Limestone Drains (ALD) In this treatment process, water to be treated is passed though the limestone drain under anoxic conditions. Near the source of the wastewater, a drainage line with coarse limestone aggregate covered with low permeable synthetic liner is constructed. When the wastewater comes in contact with the limestone, bicarbonate alkalinity is added to the water, and the pH is increased to 6 2 8 (Gazea et al., 1996). The main purpose of

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wastewater treatment using ALD is to remove acidity and increase alkalinity (Taylor et al., 2005; USEPA, 2014). 1.1.4.10 Oxic Limestone Drains (OLD) This treatment system consists of open channels filled with coarse limestone aggregate. The exclusion of oxygen, like ALD, is not required in this system. The use of this system is important when acidic wastewater is transported long distances from the source. The length of the channels ensures a longer contact time between the limestone and wastewater, and increases the treatment efficiency (Johnson and Hallberg, 2005). This treatment system is generally used to remove acidity, Mg, Al, Fe, Cu, Zn, Pb, and Se from wastewater. Because of its low construction and operating costs, it is frequently used in several mining areas for the treatment of AMD (Taylor et al., 2005; USEPA, 2014). 1.1.4.11 Slag Leach Beds (SLB) This system is similar to OLD, where slag is used as a neutralizing material instead of limestone. The slag materials generated from the steel manufacturing process are reported to be extensively used in this SLB-based treatment process, because of their high neutralizing capacity (Simmons et al., 2002). Although the process is cost-effective, there are certain disadvantages like the release of trace metals such as Al, Fe, Mg, Ti, Mn, and silica, if they are present in the slag (Taylor et al., 2005). 1.1.4.12 Alkalinity Producing Cover Alkalinity producing cover is generally used at the mining site to prevent the formation of AMD. In this system, the alkaline materials are mixed with acid generating waste rock. The treatment process becomes ineffective, due to the deployment of carbonates in the limestone cover placed above the acidic materials. The low solubility and slow dissolution rate of carbonates in near neutral water makes the treatment process ineffective. Use of magnesium-based alkaline material instead of limestone often improves the acid neutralization capacity of this process (Taylor et al., 2005). 1.1.4.13 Pyrolusite Limestone Beds Pyrolusite limestone beds are usually used for the treatment of AMD containing high Mn (Milavec, 1999). This system consists of limestone beds that are inoculated with aerobic microorganisms, such as algae.

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The respiration of aerobic microorganisms generates alkalinity which catalyzes the hydrolysis of Mn21 to insoluble MnO2 (Eq. 1.4). 2Mn21 1 2H2 O 1 O2 22MnO2 1 4H1

(1.4)

The acidity generated in the reaction is subsequently neutralized by the limestone. The treatment process is not suitable for the treatment of wastewater containing high concentrations of Fe. To make the treatment process effective, routine maintenance like regular dosing of the organic substrates for microorganisms and removal metal precipitates are important (Taylor et al., 2005). 1.1.4.14 Phytoremediation In these techniques, plants are used to treat the metal contaminated wastewater. Several mechanisms for phytoremediation include sequestration of the contaminants in plant tissue, volatilization of volatile contaminants, degradation of the contaminants, and immobilization of the contaminants in plant roots (USEPA, 2014). This treatment process is beneficial because of its sustainability and eco-friendly nature. It requires no energy and generates much lower amounts of air, water discharge, and secondary wastes (Mendez and Maier, 2008). Phytoremediation is also helpful in land restoration and sequestration of greenhouse gases. However, the process is very slow and its efficiency can decrease over time (Zhang, 2010). 1.1.4.15 Aerobic Wetland The anaerobic wetland in used for the precipitation of certain metals, such as As, Fe, Cr, and Mn, whose solubility is dependent on the pH of the water (Taylor et al., 2005). It consists of a shallow pond having some planted vegetation. The wetlands receive water diverted from other passive water treatment systems, and provide longer residence time for the precipitation to take place completely (Johnson and Hallberg, 2005). The precipitates are retained on the surfaces of the plants or flow downstream from the wetlands. This system is found effective in removing Fe (60 2 95 %) (Taylor et al., 2005). 1.1.4.16 Anaerobic Wetlands Anaerobic wetland is a water retention pond containing organic matter as the substrate and limestone aggregate as the neutralizing agent. As the sulfate and metal-rich wastewater is passed through the organic matter, the oxygen is stripped out of the water, resulting in anaerobic conditions.

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Under anaerobic conditions, the sulfate-reducing bacteria (SRB) generate alkalinity by using the organic matter as a carbon source and sulfate as an electro acceptor ( Johnson and Hallberg, 2005). The growth of SRB also generates hydrogen sulfide (Eq. 1.5). 2 2H2 O 1 SO22 4 1 Corganic matter 2H2 S 1 2HCO3ðbicarbonate alkalinityÞ

(1.5)

The dissolved metal present in the wastewater can be precipitated by the sulfide generated due to SRB growth. Some metals can also be removed due to bicarbonate alkalinity provided by both the SRB growth and limestone aggregates (Taylor et al., 2005). 1.1.4.17 Alkalinity producing system (APS) The alkalinity producing system (APS) is a combination of the ALD and anaerobic compost wetlands. In this system, the organic compost layer (18 inches) is placed on top of the limestone bed layer (18 2 24 inches). The sulfate-rich wastewater falls on the surface of the compost layer, and is subsequently passed through the layer to precipitate out the dissolved metal sulfide and consume the dissolved oxygen through organic matter decomposition (Watzlaf et al., 2000; Jarvis et al., 2002). The contaminants which are removed from the wastewater using this technique include Cu, Zn, Fe, Mn, and Pb (USEPA, 2014). 1.1.4.18 Successive Alkalinity Producing System (SAPS) A successive alkalinity producing system (SAPS) is a pond which is a combination of ALD and organic substrates. The mixture of limestone and organic substrate is overlaid by the wastewater. When the wastewater is passed through the compost, limestone provides the initial alkalinity and the oxidation 2 reduction potential (ORP) of the water is decreased. This creates a favorable condition for the growth of SRB (Bhattacharya et al., 2008). The growth of SRB generates alkalinity and hydrogen sulfide, both resulting in the removal of dissolved metals. Using this method, acidity, Fe, Al, Cu, Mn, and Zn, etc., are removed from the wastewater (Cheong et al., 2010; USEPA, 2014). 1.1.4.19 Gas Redox and Displacement System (GaRDS) This passive treatment technology is used to prevent sulfide oxidation by displacing oxygen from underground workings. To displace the air from mine voids, a mixture of anaerobic gases (CO2 and CH4) is used. Both the gases can be generated using an external bioreactor, and can be

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derived from coal bed methane (Taylor and Waring, 2001). In an anaerobic bioreactor, the biodegradation of organic compost materials generates the CO2 and CH4. When the pressure of the gases is increased, they migrate through the attached pipelines into the underground mine working. This treatment technology can be useful for underground mines where partial or complete flooding of the underground mines is not possible (Taylor et al., 2005). Long-term effectiveness of the technology is doubtful, though. 1.1.4.20 Permeable Reactive Barrier (PRB) Permeable reactive barriers (PRB) are reported to be increasingly used in the treatment of a wide range of polluted ground water (Thiruvenkatachari et al., 2008). The construction of a PRB involves the digging of a trench on the flow path of contaminated groundwater, filling the void with permeable reactive material that could serve as a growth substrate for microorganisms, and finally landscaping the disturbed surface (Johnson and Hallberg, 2005). The reductive microbiological activity within the PRB generates alkalinity and results in the removal of metals as sulfides, carbonates, and hydroxides. The anaerobic conditions within the reactor favors the growth of SRB that are known for sulfate reduction activity (Gibert et al., 2013). 1.1.4.21 Sulfate-Reducing Bacteria Mediated Treatment The SRB mediated passive treatment is routinely followed for the removal of sulfate and metal from wastewater. The SRB uses sulfate as an electron acceptor, and reduces it to sulfide that finally precipitates the dissolved metals as metal sulfide. The substrates provided to grow SRB also serves as the source of electron donor for the biological sulfate reduction process (Pol et al., 1998). The process has certain potential advantages, such as it can be readily controlled, and performance can be predictable. The dissolved metals present in the wastewater could be selectively removed and reused. The concentration of sulfate and dissolved metals can be reduced significantly. The process is cost-effective for the treatment of a high volume of wastewater (Johnson and Hallberg, 2005).

1.1.5 Sulfate-Reducing Bacteria (SRB) Sulfate-reducing bacteria are obligate anaerobic bacteria that use sulfate as an electron acceptor and reduce it to sulfide. The SRB belong to the major bacterial groups such as Gram-negative Eubacteria, Gram-positive Eubacteria, and Archaebacteria. The SRB exhibit diverse morphology such

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as rod, vibrio, coccus, filamentous, etc. The habitat of SRB includes diverse natural environments such as anoxic sediments, hydrothermal vents, hypersaline aqueous environments, hydrocarbon seeps, hydrothermal vents and mud volcanoes, etc. 1.1.5.1 Taxonomy A molecular phylogenetic study based on 16 s rRNA sequences exhibited the phylogenic relationship of SRB. In addition to the genetic diversity, the knowledge of the ecological relationship and metabolic characteristics are also important to ascertain the phylogenetic relationship of the bacterial groups (Barton and Fauque, 2009). The SRB can generally be divided into four groups, such as mesophilic gram-negative sulfate reducers, thermophilic gram-negative sulfate reducers, gram-positive sulfate reducers, and archaeal sulfate reducers (Castro et al., 2000). The SRB genera Desulfovibrio, Desulfobacter, Desulfobacterium, Desulfobulbus, Desulfococcus, Desulfohalobium, Desulfomonas, Desulfomonile, Desulfonema, Desulfosarcina, Desulfoarculus, Desulfobotulus, and Desulfomicrobium come under the group of mesophilic gram-negative sulfate reducers (Barton and Fauque, 2009). The SRB under this group show diverse morphology such as vibrio, oval, rod, spirilloid, spherical, lemon shaped, filamentous, and irregular shaped. This SRB group can use a wide range of electron donors like lactate, acetate, ethanol, fatty acids, methanol, etc. (Widdel and Bak, 1992). The SRB genus Thermodesulfobacterium fall into the group of thermophillic gram-negative sulfate reducers. The SRB under this group show a maximum growth temperature of 85 ˚C, low GC content (34 mole%), and rod-shaped morphology (Barton and Fauque, 2009). The gram-positive sulfate reducers include Desulfotomaculum (Castro et al., 2000). The SRB under the group Desulfotomaculum show rod-shaped morphology, lower GC content (38 mole%), and sporulation capacity (Campbell and Postgate, 1965). The species of Desulfotomaculum lacks desulfoviridin, and instead contains sulfite reductase P582 (Barton and Fauque, 2009). The SRB Archeoglobus fulgidas and Archeoglobus profundus fall in the group of archaeal sulfate reducers. The archaeal SRB show regular to irregular coccoid shaped morphology and are extremely thermophilic, showing a maximum growth temperature of 90 ˚C (Muyzer and Stams, 2008).

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1.1.6 Ecology The SRB is widely distributed in anaerobic aquatic and terrestrial environments containing significant amounts of sulfate. The blackening of the water and sediments due to iron sulfide precipitate is an indication of high SRB activity in that particular area. High sulfate-containing environments such as saltmarsh sediments, marine, estuarine, hypersaline lakes, and ponds are reported as a significant habitat for SRB like Desulfobacterium, Desulfonema, Desulfosarcina, Desulfobacter, and Desulfohalobium (Gibson, 1990). Similarly, activity of SRB (Desulfoarculus, Desulfobotulus, Desulfomicrobium, Desulfomonile, etc.) has also been found in soil and freshwater sediments (Muyzer and Stams, 2008). SRB has also been detected in the environments like sewage plants, spoiled foods, anaerobic digesters and whey digester, etc. (Pfennig et al., 1981). Sulfate-reducing bacteria have been isolated from termite guts, rumen contents, rice fields, oilfields, and feces of man and animals. The anaerobic submarine hydrothermal areas serve as the habitat for Archaeoglobus profundus and Archaeoglobus fulgidas (Barton and Fauque, 2009).

1.2 FIELD APPLICATION OF MICROBIAL SULFATE REDUCTION Constructed or engineered wetlands mimicking natural wetlands usually consist of a series of shallow cells often filled with gravel, soil, and organic matter to support the growth of wetland macrophytes. Microbial sulfate reduction generally takes place in an organic layer along with other reaction processes (Kleinmann, 1990). Primary productivity by inhabitant species and the organic layer acts as nutrient sources for the microbes (Gazea et al., 1996). Mine water (Flege, 2001) and electroplating wastewater (Sochacki et al., 2011), etc., have been tried for sulfate removal in engineered wetland. A maximum of around 52% sulfate was removed from Midwestern wetland in summer, compared to a declining value during winter. Smaller and more controlled systems, however, could remove more than 75% sulfate (Sheridan et al., 2012). Wetlands planted with Typha generally contain higher amounts of SRB, and can produce better sulfate removal (Park et al., 2009; Gruyer et al., 2013). Still, wetlands are more useful for acidity and metal treatment than sulfate removal (Kleinmann, 1990; Russell et al., 2003). Low pH and dissolved organic carbon are the limiting factors for sulfate reduction in this system (Woulds and Ngwenya, 2004; Gruyer et al., 2013).

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The use of permeable and geochemically reactive barriers for microbial sulfate reaction is another viable option. The PRB can be set in situ, and contaminated water can pass across by natural hydraulic gradient. Treatment can be obtained by chemical, physical, and microbial processes which lead to increased pH, and reduced metal and sulfate concentration (Benner, 1999; Gibert et al., 2002). Diverse microbial species including SRB are present in the reactive barrier, and their synergistic effect leads to overall activity (Groudev et al., 2003; Da Silva et al., 2007). Substrate tested for preparing the reactive barrier are bamboo chips, rice husk, pig-farm compost, municipal compost, coconut husk chips, recycled concrete, oyster shells, leaf compost, wood chips, limestone, silica sand, pea gravel, dolomite, fly ash, spent mushroom compost, cow manure, sawdust, peat, cucumber compost, crushed pyrite, etc. (Benner, 1999; Gibert et al., 2002; Vestola, 2009; Kijjanapanich et al., 2012). They can be used individually or as a reactive mixture; however, mixtures containing multiple substrates provide better results (Gibert et al., 2002). Choice of the substrate depends on local availability (Shabalala, 2013). Activity and longevity of PRB generally depends on: (1) selection of appropriate reactive materials; (2) clogging and armoring if metal hydroxide precipitates; (3) barrier thickness; and (4) adequate colonization of microbes, etc. (Benner, 1999; Da Silva et al., 2007). Batch or column laboratory tests aiming PRB application at different substrates showed very high levels of sulfate removal. For example, Waybrant et al. (2002) reported a 14 month average sulfate removal between 500 and 800 mmol/d/m3 in column experiments, Kijjanapanich et al. (2012) reported 30 to 77% of sulfate removal, whereas about 72 to 99% was reported by Shabalala. (2013) in batch tests. Laboratory batch test results do not always reflect original field treatments. In a continuous operation, Ludwig et al. (2002) reported approximately 48 µmol/mg/day sulfate removal in a 21 month evaluation period. Groudev et al. (2003) found maximum sulfate removal of 190 mg/L/h treating copper desposit waste in Elshiza, Central Bulgaria, in an 18 month study. However, due to partial exhaustion of organic substrate, the process slowed down and periodical replacement of fresh organic substrate was necessary. A similar finding was experienced by Benner (1999), when an aquifer contaminated by the Nickel Rim mine site, Ontario, was treated through PRB. Sulfate reduction declined from 58 to 40 mmol/L/annum after 38 months of installation, due to the exhaustion of organic carbon in the reactive matrix. A simpler organic source addition into these systems was recommended by Gibert et al.

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(2002) for process continuation. Other important parameter of PRB treatment is the uncontrolled temperature shift and optimum residence time for microbial activity (Benner, 1999; Gibert et al., 2002). A vertical flow system (VFS) or successive alkalinity producing ponds (SAPS) are semi-biological set ups generally constructed in the flow path of a wastewater creek or stream. The term SAPS was coined by the originators, Kepler and McCleary, in 1994 (Younger et al., 2002; URS report, 2003). Fig. 1.3 shows that, in this system, wastewater is allowed to pass vertically (downward) through a reactor bed generally composed of organic compost at the top followed by a solid alkaline-generating support, generally limestone, and a drainage system at the bottom (Bhattacharya et al., 2007). The organic layer is supposed to provide shelter and nutrients for microbes, including SRB, that would precipitate metal as metal sulfide and generate alkalinity. A head of standing water is maintained above the organic matter to drive the hydraulics of the system, when operational. This long water column over organic layer would help to reduce the dissolved oxygen, thereby creating the anaerobic environment required for SRB growth, colonization, and activity. Fermentative bacteria situated in the upper portion of the organic layer may further remove the dissolved oxygen. The reduced environment is also supposed to convert any Fe13 present to Fe12, thereby preventing the

Figure 1.3 Schematic diagram of a successive alkalinity producing system (SAPS).

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likelihood of “armoring” the next limestone layer. At this layer, existing acidity may be neutralized by additional bicarbonate alkalinity. Finally, the water would be directed through the network of a perforated drainage system to an aerobic settling pond or wetland, to allow metals to form precipitates and polishing (Kepler and McCleary, 1994; Younger et al., 2001; Costello, 2003; Bhattacharya et al., 2007). A SAPS is supposed to be more efficient than anaerobic wetlands, and requires less space to provide the same level of treatment (Kepler and McCleary, 1994; Younger et al., 2001; Costello, 2003). Sucessive alkalinity producing systems are also known as reducing and alkalinity producing systems (RAPS) (INAP, 2003; Watzlaf et al., 2004; Matthies et al., 2010), which more accurately reflects their function (Younger et al., 2001). Most of the study reported the use of spent mushroom compost (SMC) as the organic layer (Kepler and McCleary, 1994; Nairn and Mercer, 2000; Bhattacharya et al., 2007); however, horse manure and straw compost have also been used (Demchak et al., 2001; Matthies et al., 2010). The presence of more than 90% CaCO3 is recommended for limestone used with diameter of 1.3 to 1.9 cm (Kepler and McCleary, 1994; Nairn and Mercer, 2000; Bhattacharya et al., 2008). The presence of enough limestone to generate sufficient alkalinity for a longer treatment period (at least 10 years) is recommended (Bhattacharya et al., 2008). The Pennsylvania Bureau of Abandoned Mine Reclamation has recommended a general design guideline for SAPS: 90 to 180 cm of standing water, 45 to 61 cm of compost, and 45 to 61 cm of limestone with drainage pipes (Doshi, 2006). However, the design has not been streamlined, and varies from site to site. For example, Kepler and McCleary (1994) kept a compost layer of 45 cm and a limestone layer of 45 to 60 cm when they established four SAPS in the Mill creek watershed, Pennsylvania, treating abandoned gas well discharges, abandoned deep mine flows, and untreated surface mine discharges. Demchak et al. (2001) studied diverse designs: Bridge SAPS: 190 cm, compost layer 45 cm, limestone layer 45 cm; Filson 1 SAPS: standing water 125 cm, compost layer 40 cm, limestone layer 65 cm; Sommerville SAPS: standing water100 cm, compost layer 35 cm, limestone layer 40 cm and McKinley SAPS: standing water 40 cm, compost layer 15 cm, limestone layer 40 cm. There are many more examples of different designs of SAPS in the literature (Watzlaf et al., 2000; Bhattacharya et al., 2008). The most important things in the design aspect are, arguably, the thickness, depth, and age of the organic layer, and residence time (Drury, 1999; Demchak et al., 2001). A sufficiently thick

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compost layer may prepare a better reduced zone and provide enough nutrients for SRB (Doshi, 2006; Lee et al., 2010). The oxidation status study by Equilibrator showed that a reduced zone situated near 60 cm of bed depth, and oxidized conditions were found at 30 cm bed depth. Therefore, an organic compost layer of at least 50 cm thick has been recommended (Demchak et al., 2001). An organic budget calculation assuming the rate of sulfate entrance before construction is also an important parameter (Bhattacharya et al., 2008). There are contrasting reports about the performance, especially sulfate reduction, found in the literature. Studying three years after the installation in Simmons run SAPS in Ohio, Riefler et al. (2008) found 61% of the total alkalinity generation was due to microbial sulfate reduction. Sludge formed in the SAPS contained a high ash percentage of sulfur (29%) along with other metals. Lee et al. (2010) found 86% of sulfate removal in the SAPS effluent. Average sulfate removal rate of 8.04 g/m2/d was reported by Matthies et al. (2010) during treatment of discharges from the Durham coalfield, UK, with the highest removal value at 54%. Bhattacharya et al. (2008) found an initial increase in sulfate reduction to 60%, which dropped down sharply within a year in correlation with declining DOC values when treating wastewater coming from the Hanchang coal mine, South Korea. In the later stages, it did not rise more than 20%. However, acidity and metal removal was satisfactory. Nairn and Mercer (2000) reported a high acidity (51 g/m2/day) and metal removal rate against low sulfate reduction at Gowen passive treatment system, Oklahoma. Similar results were found by Demchak et al. (2001) in four different SAPS in Western Pennsylvania. A maximum of 17% sulfate reduction was noticed, against 26 to 91% acidity and 30 to 90% metal removal. In most cases, even the addition of sulfate into the effluent was noticed, albeit there was some definite evidence of microbial sulfate reduction. Overall, SAPS was found to be a reliable technology for treatment of highly acidic, metal, and sulphate-rich mine water, though the technique requires sufficient maintenance. Some important findings in SAPS technology are: leaching of sulfate from the organic layer (Nairn and Mercer, 2000; Bhattacharya et al., 2008); release of bioaccumulated sulfide granules by SRB when oxidizing conditions prevail, which then produce sulfate (Demchak et al., 2001); exhaustion or armoring of limestone results in lower alkalinity generation, thereby ceasing sulfate reduction (Watzlaf et al., 2000; Jage et al., 2001; Matthies et al., 2010); exhaustion of organic layer (URS report, 2003; Bhattacharya et al., 2008); low performance due

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to poor design (Kustula et al., 2005; Ji and Kim, 2008; Matthies et al., 2010); insufficient retention time and unsuitable material in SAPS (Jage et al., 2001; Kustula et al., 2005; Cheong et al., 2007); low microbial sulfate reduction (Ji et al., 2008), etc. Space availability to form a SAPS is a major problem on some sites (Kustula et al., 2005; Ji et al., 2008). Periodic replacement of the organic layer is recommended for process sustenance (Demchak et al., 2001; Matthies et al., 2010); however, disposal of the spent organic layer is very costly, especially if the disposal route is landfill (Matthies et al., 2010). Seasonal variations of performance due to high temperature shifts in these systems is another operational parameter that should be taken into account (Doshi, 2006; Ji and Kim, 2008; Matthies et al., 2010; Lee et al., 2010). In situ biostimulation for sulfate reduction at the origin of pollution is another effective technology. Microbial sulfate reduction can be stimulated by means of organic carbon amendments, adding alkalinity or effective microbial communities depending of the nature of the problem. The advantage of in situ stimulation or amendments is the relatively low labor requirements, and eliminating the need for solids management and disposal (Kempton et al., 2003). In Germany several aquifers contaminated by acidic pit lake impacted water were experimented with in situ manipulation. Enhanced sulfate reduction in impacted lake sediment by organic carbon addition was reported by Kusel and Dorsch. (2000). Schultze et al. (2009) reported methanol injection into the aquifer caused considerable reduction of acidity, iron, and sulfate in the groundwater. Longterm effects were seen in an acidic mining lake in Lusatia, Germany, when amended with straw, carbokalk, and lime to stimulate microbial reduction of both iron and sulfate in the sediment. The amendments induced sulfate reduction which led to an increase of pH in sediment pore-water of the uppermost sediment layers, as well as sufficient sulfate reduction (Herzsprung et al., 2002). Nutrient dosing was sourced by Benthaus and Lucke (2013) to improve the groundwater quality affected by a lignite mine at Pit Lake in the Lusatia and Leipzig district in Germany. Glycerin (2.6 mmol/L), nitrogen, and phosphorus with a C/N and C/P ratio of 25 and 400, respectively, were dosed by injection into the aquifer. Within 400 days of operation, the downstream groundwater showed 40% sulfate reduction and 90% iron reduction. Even after a pause of 100 days of nutrient addition, sulfate reduction immediately started to happen once it was dosed again. Jeschke et al. (2013) also found similar results when they injected nutrient containing glycerol into a

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contaminated aquifer via injection wells. Significant sulfate reduction was noticed in the wells. Groundwater sulfate was found to be reduced stepwise as it progressed through the reactive zone. The residence time and availability of an electron donor influenced the extent of sulfate degradation. Successful sulfate reduction from AMD impacted groundwater near tailing ponds at Burgas Copper Mines, Bulgaria, was reported by Groudev et al. (1998) after nutrient dosing. Acetate-bearing waste product and ammonium phosphate were injected into the subsurface through boreholes. The amendments resulted in an increase of all classes of normal microflora present, along with 100-fold increase of SRB. A maximum of 82% sulfate was removed. The biogenic H2S reduced U61 to U41, and precipitated heavy metals like Cu, Zn, Pb, Mo, and Mn as sulfide. Declining of 45% dissolved sulfate and a significant amount of metal into the wastewater was achieved from Greens Creek mine tailings, Alaska, by amending the tailing with peat and dried spent brewing grain (Lindsay et al., 2009). A tailing discharge from Kidd Creek Metallurgical Site near Timmins, Ontario, was tested for the ability of different organic amendments to induce sulfate reduction (Doshi, 2006). The sulfate reduction rate in woodchip-amended cells was 500 mg/L per year, whereas the pulp-amended cell had a much higher sulfate reduction rate, about 5,000 mg/L per year. Significant alteration was observed at the interface of the amended 2 unamended zones, where sulfate concentrations were 65 to 70% lower, iron concentrations were 80 to 99.5% lower, and zinc concentrations were over 99% lower than the unamended tailings. Alkalinity ranged from 113 to 319 mg/L at the interface, while it was 75 mg/L at the same depth in the unamended cell. The SRB numbers were one to two orders of magnitude higher than the control. Different well-known organic carbon sources were evaluated for sulfate reduction activity by organic carbon amendments in sediment samples from the Canary Creek Marsh in Lewes, Delaware (Dicker and Smith, 1985). Methanol was found to be the most effective stimulant for sulfate reduction activity, along with glucose and cellulose. An increase of 127 to 250% of sulfate reduction against a control rate of 130 to 319 nmol sulphate reduced/gm dry sediment/h was found after amendments. Syntrophy between SRB and other heterotropic microbial communities were suggested as the result after addition of an organic carbon source. The choice of the amendment materials is an important part from the industrial point of view. The availability and costs of acquiring and transporting the material are also of significant importance (Kumar et al., 2011). During nutrient or organic amendments, not only the quantity of organic carbon

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Low Cost Wastewater Bioremediation Technology

or microbial community should be taken into account, but also their biodegradability has to be assured over a reasonable time (Harris et al., 2006). The addition of insufficiently biodegradable material may cause additional pollution problems (Harris et al., 2006). The addition of organic materials into the site before sulfate rich waste starts filling offers the advantage that the material would act as a filter/nutrient medium for the groundwater and can be remediated quickly, due to the presence of available nutrient and SRB communities for microbial sulfate reduction (Kumar et al., 2011). Treatment of wastewater at the origin by compartmentalizing through a bioreactor or making an enclosure (mesocosm or microcosm) using natural conditions is another experimental approach. However, it is difficult to separate an in situ biostimulation and an in-field enclosure treatment. One aspect of the latter is the possibility of testing with original experimental parameters in more controlled steps. Both opting for original treatment, as well as studying the performance with an experimental set up on site, are reported in literature. Three field scale bioreactors (two below the ground and one above the ground) were established to treat waste rock pile emanated wastewater in the abandoned Calliope Mine, Butte, Montana (MWTP Report, 2002). The reactors were composed of organic matter (cow manure and cut straw), crushed limestone, and cobbles. Flow rate was kept at 1 gallon per minute with 4.5 to 5.5 days residence time for the majority of the 2.5 year run. Determination of the amount of sulfate reduction was inconclusive due to leaching of sulfate from substrate; however, around 4 mg/L dissolved sulfide and metal sulfide precipitate in the organic matter was found. After initial establishment, an average SRB population around 8.31 3 102 MPN/cm3 in solid matrix was noticed. Wastewater from a coal fine dump at Landau Colliery near Witbank, South Africa, was treated through an in-field sulfidogenic pilot plant bioreactor (volume 105.5 m3). An integrated treatment approach was applied where wastewater entered into a bioreactor (inoculated with sludge from an anaerobic digester) after a CaCO3 pretreatment step. The temperature was maintained at 17 ˚C. Sucrose, ethanol, propanol, (NH4)2SO4, and H3PO4 were added as a nutrient source, with a COD:N:P ratio of 1000:7:2. pH of the wastewater increased from 7.2 to 7.7 with 91% sulfate removal (2203 mg/L to 198 mg/L) (INAP, 2003). Ethanol was completely oxidized and performance enhancement was observed with increased temperature. A field pilot study for the treatment of wastewater from a reclaimed coal mine in Dunlap, Tennessee, was tested successfully (Jin et al., 2006). Dairy

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wastes (returned milk and spoiled ice cream) were injected gravimetrically into the well and monitored periodically. The resulting pH was stable at around 6.0 to 7.0, from the initial acidic readings. Decreases in sulfate and increases in sulfide concentrations confirmed valid sulfate-reducing activity. In a constructed wetland accepting wastewater from the Ranger Uranium Mine, Australia, Lloyd et al. (2004) tried stimulation of microbial sulfate reduction in two enclosures constructed on site. When sucrose and NH4Cl was added into test enclosures, more than 90% in sulfate reduction was noticed. A field reactor packed with wood chips, straw, seasoned municipal yard waste, and agricultural ground limestone was able to remove 44% sulfate, along with 76 to 99% metal and 84% acidity, from the Tab-Simco Coal Mine wastewater, Illinois, for four years’ running (Behum et al., 2011). Bilek and Wagner (2012) studied the effect of chemolithoautotrophic sulfate reduction in natural conditions to treat AMD influenced ground water for three years. Expanded clay particle was used as a reactor matrix, and H2 was fed as an electron donor. After the initial start up and establishment phase of 1.3 years, sulfate reduction rate of 0.25 to 0.3 mmol/L/ h was found with constant biomass content. Several attempts have been tried to treat mining lake 111 (Lusatian Mining District, Germany, pH 2.4 to 2.6 and SO22 4 around 1200 mg/L) by a test enclosure method (mesocosm) inside the lake. Initial experiments were carried out using Carbokalk and straw as substrate (Koschorreck et al., 2002; Wendt-Potthof et al., 2002; Geller et al., 2009; Friese et al., 2010). Microbial sulfate reduction in these enclosures was a common mechanism; however, precipitation of metal sulfide and neutralization of acidity never gained continuity. The principle reason for these results were the reoxidation of the reduced mineral phases, probably by oxygen supply during seasonal overturn of the water column, and the acidity in the sediment layer that prevented metal sulfide precipitation (Bozau et al., 2007; Geller et al., 2009; Friese et al., 2010). The reason was confirmed by a microcosm experiment using the same substrate and sediment samples in the laboratory. When oxygen entrance was strictly prevented, desired results were obtained (Frommichen et al., 2004). The in lake mesocosms were also tried using some external organic carbon dose like ethanol or whey, however, the decreased pH due to overdose of substrate and reoxidation of sulfides disallowed sufficient net sulfate reduction (Koschorreck et al., 2002; Friese et al., 2010).

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1.3 SULFATE-REDUCING BACTERIA-BASED BIOREACTOR FOR THE TREATMENT OF SULFATE AND METAL-RICH WASTEWATER Different bioreactor configurations have been used to remove sulfate and dissolved metals from wastewater (Barber and Stuckey, 2000; Kaksonen et al., 2006; Jong and Parry, 2006; Kieu et al., 2011). The efficiency of the reactors in removing sulfate and metal depends on their configuration.

1.3.1 Continuous Stirred Tank Reactor The continuous stirred tank reactor (CSTR) configuration is reported to provide efficient mass transfer that may support improved growth of SRB (Kieu et al., 2011). The disadvantage of this bioreactor application is rapid washout of the biomass, resulting incomplete removal of sulfate and metal from wastewater (White and Gadd, 1996). The biomass retention can be improved by providing an internal sedimentation system and cationic flocculants. Similarly, the biomass can be separated from the effluent and recycled back to the reactor. The different techniques for biomass separation may include sedimentation, flocculation, centrifugation, and magnetic separation (Kaksonen and Puhakka, 2007).

1.3.2 Anaerobic Sludge Blanket Reactor The anaerobic sludge blanket (UASB) reactor has been reported to be used for metal recovery in bench, pilot, and full-scale applications (Kaksonen and Puhakka, 2007; Steed et al., 2000). The UASB enables biomass retention based on the settling characteristics of granular sludge. The biomass granulation eliminates the use of packing material and reduces the start up cost (De Smul et al., 1997; 1999). Similarly, the requirement for additional instrumentation to remove the biogas produced extensively inside the reactor may increase the operational cost. The main disadvantage of the UASB reactor is poor granulation and rapid disintegration of the sludge (Kaksonen and Puhakka, 2007).

1.3.3 Membrane Bioreactor Membrane bioreactors (MBR) are of three different kinds, such as the sidestream MBR (SMBR), immersed MBR (IMBR), and extractive MBR (EMBR) (Kaksonen and Puhakka, 2007). The surface area per unit volume is higher in IMBR compared to SMBR. The IMBR is cost-effective, as it

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can operate at a lower transmembrane pressure and lower liquid cross flow velocity (Vallero et al., 2005). The EMBR is used to prevent contact between the SRB and wastewater. In EMBR, the wastewater is selectively passed over one surface of the membrane, while the microbial culture is maintained on the other side. The membrane allows the dissolved sulfide to pass from the biological compartment into the wastewater to precipitate the dissolved metals. The impermeability of the charged molecules through the membrane prevents the SRB from direct exposure to toxic metals (Chuichulcherm et al., 2001). The disadvantages of this bioreactor configuration include membrane fouling, poor mass transfer, and repeated backwashing (Chang et al., 2002; Kaksonen and Puhakka, 2007).

1.3.4 Fluidized Bed Bioreactor The fluidized bed bioreactor (FBR) eliminates channeling and clogging due to the fluidization of biomass carrier materials inside the reactor. The FBR is reported to retain high biomass, and improves the mass transfer and reaction rates (Heijnen et al., 1989). Various carrier materials such as iron chips, pumice, porous glass beads, and silica materials are reported to be used for biomass immobilization inside the reactor (Gundersen and Palmer, 2007; Kaksonen and Puhakka, 2007; Sahinkaya et al., 2007). The packing materials provide a large surface area for biomass retention inside the bioreactor. Treatment of sulfate and metal rich wastewater in FBR is reported to have near complete removal of the pollutants (Kaksonen et al., 2004; Sahinkaya et al., 2007b). The sulfate and metal removal rates per reactor volume and carrier material surface area are reported to be high in FBR, compared to other bioreactors (Kaksonen and Puhakka, 2007).

1.3.5 Anaerobic Hybrid Reactor The anaerobic reactor (AHR) is a combination of the UASB and FBR. The lower and upper part of the reactor behaves like the UASB and FBR, respectively. The AHR was reported to remove sulfate and metal more efficiently than UASB and FBR (Steed et al., 2000). Using the AHR, almost complete sulfate removal at a high rate was achieved from the wastewater (Pender et al., 2004; Sabumon, 2008). Although few studies mentioned sulfate and metal removal from wastewater using AHR (Kaksonen and Puhakka, 2007), their ability for simultaneous removal of acidity, metals, and sulfate has not been studied extensively.

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1.3.6 Packed Bed Bioreactor Efficient biomass retention can be achieved in a packed bed bioreactor (PBR). The bioreactor employs different packing materials such as glass beads, diatomaceous earth pellets, polyurethane beads (McMahon and Daugulis, 2008), sand, straw (Costa et al., 2008), porcelain rings (Mohan et al. 2005), and pumice stone (Alvarez et al., 2006) for the formation of biofilms inside the packed columns. The PBR is reported to have been operated at upflow (Robinson-Lora and Brennan, 2009), downflow (Costa et al., 2009), and horizontal modes (Silva et al., 2002). The operation of a PBR requires much less maintenance, and the packed bed creates anaerobic conditions favorable for SRB growth inside the column (McMahon and Daugulis, 2008; Bernardez et al., 2012). Therefore, the PBR has been extensively used to study sulfate and metal removal in laboratory (Viggi et al., 2010), bench (Jong and Parry, 2003), pilot, and field scale applications (Foucher et al., 2001). Several studies have reported complete removal of sulfate and dissolved metals from wastewater at a high rate using PBR (Drury 1999; Chang, 2000; Jong and Parry, 2003; Costa et al., 2009). 1.3.6.1 Packing Materials Packing materials may be solid inert substances, reactive mixtures, or compost (Viggi et al., 2009). An inert material pack allows the biofilm to grow over its surface, with a constant supply of external electron donors or nutrients into the reactor. On the other hand, a reactive mixture or compost support both the microstructure for biofilm development and nutrients for microbes derived from the matrix. Different inert packing materials have been reported in literature that are used in sulfidogenic packed bed reactors. They are polyurethane foam (Silva et al., 2002; Silva et al., 2006; Wang and Banks, 2007; McMahon and Daugulis, 2008; Camiloti et al., 2013), low-density polyethylene (Silva et al., 2006), propylene cylinder (Jimenez-Rodriguez et al., 2010), propylene pall ring (El Bayoumy et al., 1999), plain glass bead (Amann et al., 1992; Baskaran and Nemati, 2006; Bernardez et al., 2012) and porous glass bead (Kolmert and Johnson, 2001; Alvarez et al., 2006; Kousi et al., 2007; McMahon and Daugulis, 2008; Nancucheo and Johnson, 2012), porous glass pipes (Kousi et al., 2011), k-carrageenan bead and BIO-SEP bead (Selvaraj et al., 1997), rock (Tsukamoto and Weems, 2010), dolomite pebbles (Maree and Strydom, 1987), pumice

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stone (Hilton and Archer, 1988; Alvarez et al., 2006), bone char (McMahon and Daugulis, 2008), vegetal carbon (Silva et al., 2006), alumina based ceramics (Silva et al., 2006), porous ceramics (Glombitza, 2001), ceramic rings (Du Preez et al., 1991), sand particles (Baskaran and Nemati, 2006; Viggi et al., 2009), coarse sand (Jong and Parry, 2003; Costa et al., 2009; Martins et al., 2009), mineral coal (Sarti and Zaiat, 2011), porous volcanic rock (Sekomo et al., 2012), clinoptilolite (naturally occurring porous zeolite) (Bratkova et al., 2013), perlite (Viggi et al., 2009), Flocor (Thabet et al., 2009), 320 Raschig rings (Chou et al., 2008), R-635 diatomaceous earth pellets (McMahon and Daugulis, 2008), etc. “Compost bioreactor” is a generic term to describe a reactor packed with a reactive mixture or compost (Johnson and Hallberg, 2005). The composition of the packing materials is generally prepared by mixing a relatively biodegradable source (generally organic compost or manure) with more recalcitrant materials (generally cellulosic waste) (Johnson and Hallberg, 2005; Neculita et al., 2007). The general thought is that the former one would provide nutrients and electron donors for microbes, whereas the latter one would supply long-term provision for a microbial attachment site (Johnson and Hallberg, 2005). Cellulosic waste also provides the added nutrients; however, it sustains the process for longer period because of its lower degradability (Zagury et al., 2007; Choudhary and Sheoran, 2011). Some compost, for example spent mushroom compost (SMC), was reported to serve both purposes in some reactors (Dvorak et al., 1992). A simple flow through deign of a compost packed bed reactor closely imitates the field permeable reactive barrier (Neculita et al., 2007). There are several studies reporting the composition and selection criteria of a reactive mixture (Gibert et al., 2004; Neculita et al., 2007; Pagnanelli et al., 2000; Kijjanapanich et al., 2012). An efficient reactive mixture should contain an organic carbon source, a nitrogen source, a neutralizing agent, and a source of inoculum (Waybrant et al., 1998; Neculita et al., 2008). It should have both long-term and easily available carbon sources, which should ensure shorter and longer periods of time for degradation with perfect nutritional balance (Prasad et al., 1999; Cocos et al., 2002; Gibert et al., 2004). It should be locally available, with proven effectiveness, and should offer sufficient permeability (Waybrant et al., 1998; Johnson and Hallberg, 2005). For confirming adequate permeability, some “bulking agent” should also be applied

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along with the reactive mixture in some cases (Drury, 1999; Willow and Cohen, 2003; Viggi et al., 2010). More than one organic substance is generally used to prepare the reactive mixture, as it produces better sulfate reduction (Waybrant et al., 1998; Drury, 1999; Neculita et al., 2007; McCauley et al., 2009). Reactive mixtures are generally prepared from locally available agricultural, domestic, or industrial waste products. Major components are plant based materials, manures, composts, sludge, etc. Some substances that are frequently used in sulfidogenic packed bed reactors are: spent mushroom compost (Dvorak et al., 1992; Hammack and Edenborn, 1992; Chang et al., 2000; Neculita et al., 2011; Song et al., 2012), vegetal compost (Gibert et al., 2002), municipal compost (Gibert et al., 2003; Gibert et al., 2004), composted leaf mulch (Waybrant et al., 2002; Neculita et al., 2008; Guo and Blowes, 2009), composted bark (Furuya et al., 2012), cow manure (Cheong et al., 1998; Drury, 1999; Neculita et al., 2011; Song et al., 2012; Furuya et al., 2012), horse manure (Tsukamoto et al., 2004), sheep manure (Gibert et al., 2004), poultry manure (Gibert et al., 2004; Neculita et al., 2008), composted livestock manure (Willow and Cohen, 2003), oyster shells (Banasiak and Indraratna, 2012), sewage sludge (Waybrant et al., 2002), paper sludge (Chang et al., 2000), organic rich soil (Chang et al., 2000), rye grass (Gibert et al., 2002), sawdust (Drury, 1999; Waybrant et al., 2002; Neculita et al., 2008; Neculita et al., 2011; Song et al., 2012),wood chips (Waybrant et al., 2002), rice husk (Furuya et al., 2012), rice straw (Cheong et al., 1998; Neculita et al., 2011; Song et al., 2012), rape straw (Wang et al., 2012), oak chips (Chang et al., 2000; Neculita et al., 2011), pine shavings, beechwood, ground alfalfa (Pereyra et al., 2012), spent oak chips (Chang et al., 2000), maple wood chips (Neculita et al., 2008), post peel, Pinusradiata bark, mussel shells (McCauley et al., 2009), oak leaf and Verdemix-CEREMC compost (Viggi et al., 2009; Viggi et al., 2010). Limestone is also reported to be used in these reactors for neutralizing any acidity (Cheong et al., 1998; Neculita et al., 2008; Guo and Blowes, 2009).

1.4 FACTORS AFFECTING BIOLOGICAL SULFATE REDUCTION Biological sulfate and metal removal is affected by different parameters such as pH, hydrogen sulfide, hydraulic retention time (HRT), temperature, inoculum, and concentration of sulfate and dissolved metals.

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It is reported that at low pH the higher energy consumption for proton pumping across the cell membrane could result in the inhibition of growth of SRB, and alter the microbial community structure (Kaksonen and Puhakka, 2007; Hao et al., 2014; Sa´nchez-Andrea et al., 2014). The pH also regulates the SRB growth, inhibiting metabolic products such as H2S and free NH3. Low pH enables the high dissolution of H2S in the solution, resulting in the inhibition of SRB growth (Kaksonen and Puhakka, 2007). Similarly, pH at alkaline range increases the NH3 concentration, which could inhibit SRB growth by altering the intracellular pH, increasing the cellular maintenance energy, decreasing specific cellular activities, and disrupting cellular homeostasis (Gutierrez et al., 2009). Temperature is also an important factor for biological sulfate reduction. Increase in temperature enables the SRB to outcompete the methanogens (Kallmeyer and Boetius, 2004). The elevated temperature may increase the rate of microbial metabolism and reduce the solubility of H2S that results in the improvement of SRB growth (Moosa et al., 2005). The short HRT may lead to biomass washout from the bioreactor, resulting in a slow rate of sulfate reduction (Sheoran et al., 2010). On the contrary, very high HRT leads to the dominance of methanogens over SRB (Kaksonen and Puhakka, 2007). A high concentration of dissolved metals such as Ni and Zn inhibits the growth of SRB (Poulson et al., 1997). The presence of calcium and magnesium is also reported to improve the dominance of SRB over methanogens (De Smul et al., 1999). It is reported that at high sulfate concentration, SRB gain thermodynamic and kinetic advantages over acetogenic and methanogenic bacteria (Lens et al., 1998; Pol et al., 1998). Very high sulfate concentration is reported to increase redox potential and production of dissolved sulfide, resulting in inhibition of SRB growth (Oyekola et al., 2010).

1.5 SUMMARY In the above, we discussed the evolutionary development of bacterial remediation of industrial wastewater, with particular emphasis on SRB mediated treatments. Many of the developments either remain confined to laboratory-scale treatment, or are conducted in uncontrollable field conditions. The challenges are to retain a healthy and resilient microbial community, reaction rate, scaling-up and control, and economical costs, etc. In the following chapters we see the progress made in making scaledup plant-based treatments, using the microbes in their decomposer role.

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FURTHER READING Banisi, S., Finch, J.A., Laplante, A.R., 1993. Electrical conductivity of dispersions: a review. Miner. Eng. 6 (4), 369385. Cruz Viggi, C., Pagnanelli, F., Cibati, A., Uccelletti, D., Palleschi, C., Toro, L., 2010. Biotreatment and bioassessment of heavy metal removal by sulphate reducing bacteria in fixed bed reactors. Water Res. 44 (1), 151158. Karimi, L., Ghassemi, A., 2015. Effects of operating conditions on ion removal from brackish water using a pilot-scale electrodialysis reversal system. Desalin. Water Treat. 3994, 113. Mcdonald, C.W., Bajwa, R.S., 1977. Removal of toxic metal ions from metal-finishing wastewater by solvent extraction. Sep. Sci. 12 (4), 435445. Silva, A.M., Lima, R.M.F., Lea˜o, V.A., 2012. Mine water treatment with limestone for sulfate removal. J. Hazard. Mater. 221-222, 4555. Velasco, A., Ramı´rez, M., Volke-Sepu´lveda, T., Gonza´lez-Sa´nchez, A., Revah, S., 2008. Evaluation of feed COD/sulfate ratio as a control criterion for the biological hydrogen sulfide production and lead precipitation. J. Hazard Mater. 151, 407413. .