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Passive Treatment of Acid Mine Drainage in Bioreactors using Sulfate-Reducing Bacteria

Critical Review and Research Needs

^a Dep. of Civil, Geological, and Mining Engineering, École Polytechnique de Montréal, Montreal, QC, Canada H3C 3A7
^b Dep. of Applied Sciences, Univ. du Québec en Abitibi-Témiscamingue, Rouyn-Noranda, QC, Canada J9X 5E4

^* Corresponding author (gerald.zagury{at}polymtl.ca)

Received for publication February 15, 2006.

	ABSTRACT

TOP ABSTRACT INTRODUCTION Passive Bioreactors: Principle,... Factors of Influence on... Performance of Passive... Conclusions and Research Needs REFERENCES

 
Acid mine drainage (AMD), characterized by low pH and high concentrations of sulfate and heavy metals, is an important and widespread environmental problem related to the mining industry. Sulfate-reducing passive bioreactors have received much attention lately as promising biotechnologies for AMD treatment. They offer advantages such as high metal removal at low pH, stable sludge, very low operation costs, and minimal energy consumption. Sulfide precipitation is the desired mechanism of contaminant removal; however, many mechanisms including adsorption and precipitation of metal carbonates and hydroxides occur in passive bioreactors. The efficiency of sulfate-reducing passive bioreactors is sometimes limited because they rely on the activity of an anaerobic microflora [including sulfate-reducing bacteria (SRB)] which is controlled primarily by the reactive mixture composition. The most important mixture component is the organic carbon source. The performance of field bioreactors can also be limited by AMD load and metal toxicity. Several studies conducted to find the best mixture of natural organic substrates for SRB are reviewed. Moreover, critical parameters for design and long-term operation are discussed. Additional work needs to be done to properly assess the long-term efficiency of reactive mixtures and the metal removal mechanisms. Furthermore, metal speciation and ecotoxicological assessment of treated effluent from on-site passive bioreactors have yet to be performed.

Abbreviations: AMD, acid mine drainage • SRB, sulfate-reducing bacteria • DOM, dissolved organic matter • COD, chemical oxygen demand • HRT, hydraulic retention time • PRB, permeable reactive barrier • HRC, hydrogen release compounds

	INTRODUCTION

TOP ABSTRACT INTRODUCTION Passive Bioreactors: Principle,... Factors of Influence on... Performance of Passive... Conclusions and Research Needs REFERENCES

 
MINE WASTES REPRESENT a source of potential environmental risk, particularly when the wastes contain sulfide minerals that can oxidize and generate acid mine drainage (AMD) (Aubertin and Bussière, 2001; Blowes et al., 2003; Tabak and Govind, 2003; Willow and Cohen, 2003). The main characteristics of AMD are low pH and high concentrations of dissolved heavy metals and sulfates (Tsukamoto et al., 2004). Acid mine drainage is considered the most important and widespread mining industry related pollution problem around the world (Tsukamoto and Miller, 1999). For example, massive sulfide tailings from ore having a pyrite content higher than 95% generated one of the worst AMD contaminations, with hydrogen ion concentrations (H⁺) as high as 10^3.6 mol L^–1, total dissolved metal concentrations as high as 200 000 mg L^–1, and sulfate concentrations as high as 760 000 mg L^–1 (Nordstrom et al., 2000).

Numerous approaches have been used to prevent AMD generation or to treat, control, and mitigate its effects. Several technologies, for example, have been developed to stop weathering processes by controlling the waste deposits and reducing the transfer of oxygen and water to the waste [Mine Environment Neutral Drainage (MEND) Report, 2001]. However, these prevention techniques do not represent the scope of the present review.

Acid Mine Drainage Formation
The processes by which AMD is generated are currently quite well understood. Acid mine drainage is generated through a combination of chemical and biological processes by which pyrite is converted to sulfates and iron oxyhydroxides. However, the detailed mechanisms still need to be clarified (Usher et al., 2004). The AMD is generated when sulfides, in particular pyrite and pyrrhotite, are exposed to water and oxygen. Acid mine drainage generation is further amplified when the reactions are catalyzed by aerobic bacteria such as Acidithiobacillus (formerly Thiobacillus) ferrooxidans (Zagury et al., 1997; Brown et al., 2002). Factors such as bacterial activity, pH, sulfide mineral surface area, crystallography, type of sulfide minerals, temperature, and oxygen concentration control the rates of AMD generation (Berghorn and Hunzeker, 2001).

Several reactions are involved in the weathering of pyrite (Stumm and Morgan, 1981). The process is initiated at neutral pH by the release of ferrous iron (Fe²⁺) into solution by pyrite oxidation, according to the following reaction:

[1]

The next step occurs at lower pH values (<4):

[2]

Iron-oxidizing bacteria, many of which tend to be most active at pH 2.0 to 4.0, can increase the rate of Fe²⁺ oxidation by factors greater than 10⁶ (Brown et al., 2002).

Ferric iron is not soluble in water if the pH is higher than 2.3 to 3.5, depending on total iron concentration. Therefore, it precipitates as oxyhydroxide releasing H⁺ and lowering the pH:

[3]

As the pH decreases, the cycle reinitiates because ferric iron remains in solution and is reduced by pyrite, which generates additional ferrous iron and acidity (a self-perpetuating process), until either ferric iron or pyrite is depleted:

[4]

The rate of pyrite oxidation by Fe³⁺ is much higher than oxidation by O₂. Furthermore, this oxidation of 1 mole of pyrite releases 16 moles of H⁺ (reaction 4) compared to 2 moles of H⁺ in reaction 1. For these reasons, the oxidation of Fe²⁺ to Fe³⁺ (reaction 2) is often referred to as the "rate determining step" in the acid-generating process.

The overall reaction is given by adding reactions [1] through [3]:

[5]

These reactions progressively increase water acidity, resulting in mobilization of metals from mine wastes.

To avoid significant environmental impacts, waters contaminated by AMD must be collected and treated to remove metals and to increase the pH before being discharged into the environment. Traditionally, AMD treatment was performed with alkali to neutralize the acidity, increase water pH, and precipitate metals as hydroxides and carbonates (Ritcey, 1989; Santos et al., 2004). Unfortunately, dumping of limestone in streams succeeds only as long as the water is anoxic. When the neutralized water is exposed to the atmosphere, ferrous iron oxidizes, hydrolyses, precipitates, coats the limestone, and slows its rate of dissolution. The effect may therefore be limited (Gazea et al., 1996).

Other technologies to treat AMD, such as ion exchange, reverse osmosis, electrodialysis, and electrolytic recovery are also available but are expensive and not commonly used (Prasad et al., 1999). The solubility product of most metal hydroxides is higher than that of metal sulfides (Gazea et al., 1996). Therefore, stabilization of metals is preferred in the form of sulfides. In the past 20 yr, research has focused on passive biological methods for AMD treatment because of their numerous advantages. They produce a high degree of metal removal at low pH (pH 3 to 6), denser, less voluminous, and more stable sludge compared to sludge obtained during AMD chemical treatment. Moreover, they allow lower operation costs, and minimal energy consumption (Gazea et al., 1996; Willow and Cohen, 2003). Nevertheless, treatment performance and long-term efficiency still need to be improved (Beaulieu et al., 1999, 2000; Tsukamoto et al., 2004; Johnson and Hallberg, 2005a; Kalin et al., 2006).

Classification of Passive Treatment Systems
Passive treatment technologies have received much attention lately and the literature offers extensive studies related to these systems (Gazea et al., 1996; Ziemkiewicz et al., 2003). Several classifications have been proposed on the basis of different criteria such as: (1) aerobic or anaerobic processes, (2) complexity and requirements for maintenance, and (3) dominant chemical or biological processes occurring during treatment.

One classification separates aerobic passive treatment systems such as aerobic wetlands, open limestone channels, diversion wells, oxic limestone drains, and pyrolusite treatment beds from anaerobic systems, namely compost and/or anaerobic wetlands, anoxic limestone drains, and vertical flow reactors (Berghorn and Hunzeker, 2001). In another classification, "passive" systems such as aerobic wetlands and compost reactors/wetlands are separated from "active" systems such as sulfidogenic bioreactors and accelerated iron oxidation with immobilized biomass (Johnson and Hallberg, 2002). In the authors' opinion, the most appropriate classification to date (Fig. 1) separates the chemical passive treatment from the biological passive treatment, which involves sulfate-reducing bacteria (SRB). Permeable reactive barriers (PRBs) for groundwater treatment can be classified either as biological or as chemical passive treatment systems (Brown et al., 2002).

View larger version (6K): [in this window] [in a new window]   Fig. 1. Classification of AMD passive treatment systems.

New Trends in US and Canadian Mining Legislation
North American mining legislation requires effluent toxicity assessment. Therefore, the performance of passive bioreactors will also need to be evaluated in terms of ecotoxicity reduction. In Canada, the Metal Mining Effluent Regulations (MMER) (Environment Canada, 2002) under the authority of the Fisheries Act requires that all mines with an effluent flow higher than 50 m³ d^–1 conduct an Environmental Effects Monitoring (EEM) program. The regulation prescribes limits for the discharge of deleterious substances, and a requirement for effluent to be nonacutely lethal to rainbow trout and Daphnia magna. Since 2003, the metal mines are also required to conduct, two to four times per year, a series of four freshwater sublethal tests to determine if the effluent has the potential to affect fish, invertebrates, algae, and plants. In marine ecosystems, three sublethal toxicity tests (fish, invertebrate, and algae) are also required. The sublethal toxicity tests to be used are the following: invertebrate Ceriodaphnia dubia reproduction test, algal growth inhibition test using Pseudokirchneriella subcapitata (formerly Selenastrum capricornutum), fathead minnow Pimephales promelas or rainbow trout Oncorhynchus mykiss embryo development inhibition test, and macrophyte Lemna minor growth inhibition test.

Similarly, in the US, all point source discharges from mining operations are to be authorized under a National Pollutant Discharge Elimination System (NPDES) permit, as described in the Clean Water Act [Clean Water Act (CWA), 1977]. In 2002, the USEPA set new guidelines establishing test procedures for the analysis of pollutants to add a series of standardized acute and short-term chronic whole effluent toxicity (WET) tests to the list of approved methods under the CWA (USEPA, 2002). The methods measure the toxicity of effluents to freshwater, marine, and estuarine organisms. The ratified WET methods are the following: a survival and reproduction test of the invertebrate Ceriodaphnia dubia, a growth inhibition test of algae Selenastrum capricornutum, a larval survival and growth inhibition test of sheepshead minnow Cyprinodon variegatus and inland silverside Menidia beryllina, and a survival, growth inhibition, and fecundity test of crustacean Mysidopsis bahia.

Objective of the Critical Review
Biological passive treatment of AMD has been the focus of numerous studies, but in fact much remains to be learned regarding fundamental interactions within these complex biological reactors. To the authors' knowledge, there has been so far no integrated review dealing with critical factors for the design and long-term operation of passive on-site bioreactors for AMD treatment. However, several papers are available on physicochemical and biological processes occurring within wetlands (Gazea et al., 1996; Wildeman and Updegraff, 1997; Sheoran and Sheoran, 2006). These papers address potential problems related to constructed wetlands—these systems allow little or no system control, are subject to seasonal and other variations, do not allow accurate assessment of their sizing and performance because of insufficient data available, may be ineffective when used in isolation, and may be subject to catastrophic system failure (Johnson and Hallberg, 2002). Comprehensive evaluations on the performance of passive systems excluding on-site bioreactors are also available (Ziemkiewicz et al., 2003). Moreover, other papers have discussed the advantages which sulfur cycle bacteria offer for sulfate-rich wastewater treatment in wetlands and in active bioreactors (Hulshoff Pol et al., 2001; Lens et al., 1998, 2002). Recently, a review paper assessed the applicability, suitability, efficiency, and cost-effectiveness of various AMD treatment schemes based on available monitoring data from the UK. However, in this work special emphasis was given to the use of wetlands as a passive biotechnology (Brown et al., 2002). Finally, a book describing passive mine water remediation with a special emphasis on field studies in the UK is also available (Younger et al., 2002).

The present article intends to be an integrated critical review of current knowledge about passive treatment of AMD in on-site bioreactors. Critical parameters for design and long-term operation such as the composition of reactive mixtures are reviewed. Special attention is accorded to natural organic carbon sources which offer long-term availability, and improved efficiency of the passive bioreactors; an assessment of their potential biodegradability is also provided. Considering the new trends in North American legislation, non-previously treated aspects in available review papers such as organic carbon sources, sulfides, and dissolved metal toxicity to SRB and effluent ecotoxicity are also discussed. Finally, some unexplored research needs and perspectives are suggested.

	Passive Bioreactors: Principle, Characteristics, and Mechanisms

TOP ABSTRACT INTRODUCTION Passive Bioreactors: Principle,... Factors of Influence on... Performance of Passive... Conclusions and Research Needs REFERENCES

 
Sulfate Reduction Principles
Sulfate-reducing bacteria are either heterotrophic or authotrophic anaerobes, capable of reducing sulfate to sulfide by a dissimulator, bioenergetic metabolism when provided with a suitable organic carbon source (Postgate, 1984). Substrate (electron donor) oxidation is coupled to sulfate (terminal electron acceptor) reduction. The resulting energy is used by SRB for growth and development. The reaction is generally expressed as (Widdel, 1988):

[6]

where CH₂O represents a simple organic carbon source. The dissolved inorganic carbon neutralizes the pH and favors the precipitation of metal carbonate minerals. The soluble sulfides (H₂S, HS^–, and S^2–) react with metals to form metal sulfide precipitates:

[7]

where M is a cationic metal such as Cd, Fe, Ni, Cu, or Zn.

Further information on the principles of sulfate reduction can be found in Widdel (1988) and in Hao et al. (1996).

Organic Carbon Sources
Acid mine drainage generally contains relatively low concentrations of dissolved organic carbon (<10 mg L^–1) (Kolmert and Johnson, 2001). Therefore, the most critical limiting factor for the microbial activity is the availability of carbon from an additional organic source (Gibert et al., 2004; Zagury et al., 2006). The challenge for having an efficient on-site bioreactor is to select a suitable organic substrate to make the process efficient and economically feasible. Selection of the organic carbon source is usually made on the basis of availability and costs of the added electron donor per unit of reduced sulfate. The remaining contaminants in the treated water must be present in low concentrations or easy to remove (Hulshoff Pol et al., 2001).

Simple Organic Carbon Sources
Sulfate-reducing bacteria use the easily degradable fraction of organic matter such as low molecular weight compounds with simple structures (e.g., methanol, ethanol, lactate) (Dvorak et al., 1992; Nagpal et al., 2000a; Tsukamoto et al., 2004), polylactic acid (Edenborn, 2004), simple carbohydrate monomers (e.g., glucose or sucrose) (Mizuno et al., 1998), and whey (Christensen et al., 1996). In terms of energy and biomass produced, lactate is a superior electron donor compared to others such as ethanol, acetic acid, propionate, and acetate (Nagpal et al., 2000b). In terms of moles of bicarbonate produced per mole of substrate consumed, the lactate-utilizing processes are superior to ethanol-utilizing processes (3 vs. 2, respectively) since they are better at neutralizing the acidity in the treated effluent (Kaksonen et al., 2004a). The main drawback is that only certain species of SRB (Desulfotomaculum) are capable of oxidizing lactate and ethanol to CO₂, whereas others (Desulfovibrio) can partially oxidize the C2-C4 organic carbon molecules to acetate, and very few can use acetate alone (Desulfotomaculum acetoxidans) (Nagpal et al., 2000b).

Complex Organic Carbon Sources
Alternatively, less expensive organic carbon sources such as waste material from the agricultural and food processing industry have been assessed for their potential to sustain sulfate reduction. The alternative organic carbon sources may be selected between two groups of materials—cellulosic wastes and organic wastes (Kuyucak and St-Germain, 1994). Generally, cellulosic wastes include sawdust, hay, alfalfa, and wood chips, whereas organic wastes include cattle manure, cow manure, horse manure, poultry manure, sheep manure, rabbit manure, granular or sewage sludge, peat, pulp mill, molasses, and compost (see Table 1).

Furthermore, contradictory conclusions emerge from studies performed with the same organic carbon source but using different proportions in the reactive mixture. In an attempt to find the best reactive mixture for use in permeable reactive walls, Waybrant et al. (1998) concluded that sheep manure (100%) did not produce the reducing conditions necessary for bacterial activity and excluded this organic source from their batch assays. In contrast, Gibert et al. (2004) clearly indicated sheep manure (15% of reactive mixture) as the most successful organic material for creating reducing conditions and sustaining active sulfidogenesis (sulfate removal > 99%) in a batch experiment. Similarly, in the experimental study of Amos and Younger (2003), cattle manure (100%) was rejected at an early stage due to low permeability, whereas a mixture of cow manure (80%) and cut straw (20%) was successfully used over 32 mo by Zaluski et al. (2003).

Gibert et al. (2004) report that the findings of Cocos et al. (2002) (who found limited degradability of lignin-cellulosic substrates in a 41-d batch test) and those of Waybrant et al. (1998) (who concluded that a cellulosic material alone could sustain satisfactory bacterial activity in a 125-d column test) are contradictory. In fact, the conclusion of Waybrant et al. (1998) was that after the acclimation period (20 to 65 d), sulfate reduction rates were higher in the reactive mixtures that contained a variety of organic carbon sources. In the short-term study of Cocos et al. (2002), a higher proportion of poultry manure was essential for promoting higher sulfate reduction rates. Further, the observation made by Waybrant et al. (1998) was that the cellulose entailed a slightly lower sulfate reduction rate compared to other substrates tested (sewage sludge, leaf mulch, wood chips, sheep manure, and sawdust) alone or in mixture. This is in agreement with the results of Chang et al. (2000) who observed similar performance at later stages of experiments (after 20 wk) in bioreactors using several sources of waste materials containing cellulose. In this later study, the cellulose was the main component used during 35 wk of operation.

The efficiency of cellulosic substrates for the biological treatment of AMD has been confirmed by several studies (Tuttle et al., 1969a, 1969b; Waybrant et al., 1998; Chang et al., 2000; Tassé and Germain, 2002; Johnson and Hallberg, 2005b), while other studies suggested that cellulosic wastes alone entailed carbon-limiting conditions (Béchard et al., 1994) or did not sustain SRB growth (Kuyucak and St-Germain, 1994). With sawdust as the sole nutrient source, a mixed bacterial culture containing cellulose-degrading bacteria and SRB was capable of reducing sulfate at pH 3.0, whereas pure cultures of SRB did not reduce sulfate below pH 5.5 (Tuttle et al., 1969b). These results stress the importance of a well established microflora in the presence of mixtures of cellulosic and other complex natural organic carbon sources.

Microflora
Sulfate-reducing bacteria cannot directly oxidize complex organic carbon compounds such as carbohydrates, proteins, lipids, cellulose, and hemicellulose polymers (Postgate, 1984). When such organic carbon sources are provided, synergism between three groups of microorganisms (acidogens, methanogens, and sulfate reducers) is essential to provide short-chain organic carbon compounds for SRB (Tuttle et al., 1969a, 1969b; Kuyucak and St-Germain, 1994). Hydrogen produced by acidogenic bacteria during anaerobic digestion can be utilized by all three groups of bacteria and competition may occur (Raskin et al., 1996; Mizuno et al., 1998). The study of changes in Gibbs free energies for hydrogen-consuming reactions under standard and steady-state conditions showed that SRB have a thermodynamic advantage over methane-producing bacteria and homoacetogenic bacteria (Mizuno et al., 1998). Therefore, SRB would outcompete methane-producing bacteria if sulfate is provided as a final electron acceptor.

A recent study indicated that compost-based sulfate-reducing bioreactors are dominated by non-sulfate-reducing bacteria (Hallberg and Johnson, 2005a). Three to four metabolic groups involved in AMD treatment processes were found, using cellulosic or organic substrates as a solid support and cellulolytic microorganisms, fermenters and respirers, methanogens, and SRB (Béchard et al., 1994; Logan et al., 2005) as organic carbon sources for microbial growth. Recently, cellulolytic and fermenter communities were found limited at no time by organic carbon availability during column experiments conducted for 100 d. However, SRB showed great limitation through all but the early establishment phase (Logan et al., 2005). In a passive field bioreactor containing a mixture of 95% softwood and 5% hay, a 10-mo shut-down period before its monitoring appeared a long enough lag time to allow key microbial populations to increase in numbers and activities in the absence of AMD throughput (Johnson and Hallberg, 2005b). Availability of nutrient sources for cellulolytic and fermenter microbes is vital to the long-term sustainability of passive treatment systems (Lynd et al., 2002). Anaerobic degradation of complex organic carbon compounds to simpler molecules by cellulolytic and fermenter microbes may limit the rate at which substrates become available to SRB (Logan et al., 2005). Sulfate reduction seems controlled by cellulose degradation and therefore, future research for exploring means by which to enhance cellulose hydrolysis is needed. More work must be conducted to understand and differentiate the fundamental biochemical and microbiological reactions that occur in anaerobic bioreactors with complex organic substrates. This might be a key step for the successful implementation of SRB-based AMD remediation systems.

Tests for Assessing the Biodegradability of Complex Organic Substrates
Organic carbon available for bacteria is contained in the dissolved organic matter (DOM), which consists of a rapidly degradable fraction (e.g., simple organic compounds), the polysaccharides fraction that is degraded more slowly, and the recalcitrant fraction that remains in solution up to 180 d (Marschner and Kalbitz, 2003). For the quantification of DOM biodegradability, the duration of incubation and the measure of biodegradation are crucial parameters (Marschner and Kalbitz, 2003). Chemical composition of an organic substrate controls its biodegradability pattern (Gibert et al., 2004; Zagury et al., 2006). Results of biodegradation studies strongly depend on the experimental methodology; for example, the duration of incubation and the method of quantification [monitoring of dissolved organic carbon (DOC) concentration vs. CO₂ efflux from the sample during incubation] (Marschner and Kalbitz, 2003). Some researchers attempted to predict the degradability of complex organic substrates by assessing their chemical composition (protein and carbonate, cellulose, hemicellulose, and lignin content, solvent and water-extractable organic matter, and easily extractable fractions) (Prasad et al., 1999; Chang et al., 2000; Gibert et al., 2004; Zagury et al., 2006). The conclusion is that none of these chemical extractions alone is sufficient to accurately predict the degradability of organic materials. Therefore, a standardized method is still warranted to predict the ability of organic substrates to promote sulfate reduction and metal removal. Several models were developed to describe the influence of organic materials degradation on sulfate reduction rates in passive bioreactors and permeable reactive barriers (Drury, 2000; Benner et al., 2002; Mayer et al., 2002; Amos et al., 2004). They all potentially oversimplify sulfate reduction because they do not consider bacterial growth and decay (Hemsi et al., 2005). On the contrary, the biochemical model developed by Hemsi et al. (2005) coupled sulfate reduction kinetics to organic materials decomposition and biochemical processes. The simulated results showed the importance of kinetics used to describe the decomposition of solid organic materials.

Configurations of Organic Substrates and Depletion of Organic Carbon
In many passive bioreactors, the organic matter matrix also serves as a support for microbial attachment and metal precipitation (Tsukamoto et al., 2004). The most effective design is typically when the substrate is sandwiched between pipes set in inert gravel in the top and bottom of the basin (URS Report, 2003). The lifetime of such bioreactors is limited by the amount of reducing equivalents readily available to SRB affecting the extent of microbial activity and treatment efficiency (Gibert et al., 2004; Tsukamoto et al., 2004). Other configurations include bioreactors filled with a combination of organic matter, crushed limestone, and cobbles placed in two to four discrete chambers (Zaluski et al., 2003) and site-specific passive systems that incorporate anaerobic and aerobic cells and limestone and rock filters (Johnson and Hallberg, 2005b; Kuyucak et al., 2006). Evaluation of substrate longevity based on fixed amounts of organic carbon and limestone and the relative consumption rates observed is, however, useless due to high variability in kinetics of sulfate reduction during the treatment (Reisman et al., 2003). A proposed solution to extend the long-term performance of a bioreactor was the addition of an alternative organic source to a depleted matrix. Simple organic compounds such as methanol (Tsukamoto et al., 2004), sucrose (Béchard et al., 1994; Lloyd et al., 2004), lactate (Tsukamoto and Miller, 1999; Tsukamoto et al., 2004), and acetate (Gibert et al., 2004) were successfully tested. Another solution was to bioactivate the bacterial consortia with an easily available organic source (e.g., lactate) and then to replace it with a less expensive source such as ethanol (Kaksonen et al., 2003) or a mixture of wood chips, leaf compost, and poultry manure (Beaulieu et al., 2000). In the former study, Kaksonen et al. (2003) reported partial degradation of ethanol to acetate that increased residual organic carbon in the effluent. In the latter study, a longer lag period was necessary for SRB to get acclimated to the more complex organic carbon sources. New formulations of suitable organic carbon sources, such as patented mixtures of organic materials (methonak, molasses, methanol, and wood chips) commercialized by ARCADIS treatment systems or hydrogen release compounds (HRC) by REGENESIS technologies are also available. Commercialized reactive mixtures can be used for in situ treatment of AMD-contaminated waters in pit lakes, smelter ponds, and flooded and underground workings (e.g., Gilt Edge Mine, Anchor Hill Pit Lake, Hollister Mine, and Sweetwater Mine). Hydrogen release compounds are an electron donor organic material designed to produce controlled release of lactic acid when hydrated. Hydrogen release compounds can be directly injected in the contamination source area or used in PRB applications.

Metal Removal Mechanisms
According to the literature, the main mechanisms of metal removal in bioreactors are precipitation in the form of sulfides (Pb²⁺, Co²⁺, Cd²⁺, Cu²⁺, Ni²⁺, Fe²⁺, Zn²⁺), hydroxides (Fe³⁺, Cr³⁺, and Al³⁺), and carbonates (Fe²⁺, Mn²⁺). Sorption mechanisms such as adsorption, surface precipitation, and polymerization on inorganic support, solid organic matter, bacteria, and metal precipitates also occur. Besides biologically mediated processes, AMD quality is improved by filtration of the suspended and colloid materials (Wildeman and Updegraff, 1997).

The metal removal mechanisms change during the life of a passive bioreactor. Upon startup of a passive bioreactor, the adsorption of dissolved metals onto organic sites in the substrate material will be an important process (Machemer and Wildeman, 1992; Gibert et al., 2005a). In the pH range 4 to 7, SRB retain metals via biosorption due to the neutral and/or deprotonated state of binding ligands on cell walls. The biosorption by SRB is metabolism-independent (sorption onto the cell wall) or metabolism-related (transport, internal compartmentalization, and extracellular precipitation by metabolites) (Chen et al., 2000). Factors such as availability of nutrients during growth, age and physiological state of bacterial cells, environmental conditions (pH, ionic strength, and temperature), presence of competitive ions, and concentration of the biomass can influence biosorption (El Bayoumy et al., 1997; Chen et al., 2000; Utgikar et al., 2000; Santos et al., 2004). Because of several factors of influence, the experimental results are not always in agreement. At pH 3.0, a biomass content >6 g L^–1 increased the efficiency of metal removal, favoring sedimentation of the iron precipitate and rates of filtration (Santos et al., 2004). In another study conducted at pH 7.0, biosorption capacity was constant regardless of the experimental conditions (e.g., stirring and biomass type) (El Bayoumy et al., 1997). These differing results may be due to a more or less active microbial population in the biomass used as well as to a possible competition between metals. Competition among Fe, Cu, Zn, and Mn for organic adsorption sites was confirmed by laboratory tests, and field tests in wetlands (Machemer and Wildeman, 1992; Utgikar et al., 2000). In the study of Chen et al. (2000), biosorption on Desulfovibrio desulfuricans was strongly pH-dependent. For Cu (II) and Zn(II), biosorption increased within a pH range of 4.0 to 6.6. At pH below 3.0, metal biosorption was insignificant due to the strong affinity of protons onto metal binding sites on biomass cell walls, whereas at pH > 5.0 for Cu(II) or 6.6 for Zn(II), the desorption or precipitation contribution increased significantly compared to adsorption. Functional groups capable of metal sorption such as carboxylic and phenolic groups are deprotonated at high pH and presumably available for binding dissolved metals (Dudal and Gérard, 2004). Therefore, at the slightly acidic to neutral pH of on-site sulfate-reducing bioreactors, adsorption of dissolved metals on the substrate material is an important metal removal mechanism. Over time, however, the adsorption sites become saturated. This saturation may take from 3 to 8 wk (Waybrant et al., 1998; Willow and Cohen, 2003) to 4 to 8 mo (Zaluski et al., 2003).

Once sulfate-reducing conditions are established, sulfide precipitation becomes the predominant mechanism of metal removal from AMD (Machemer and Wildeman, 1992; Béchard et al., 1994; Song et al., 2001). Sulfide precipitation is the desired mechanism of contaminant removal because metal sulfides are highly insoluble and less bioavailable compared with other metal species (Wildeman and Updegraff, 1997). Sulfate reduction in passive bioreactors is confirmed by lower concentrations of sulfates in the effluent than in the influent waters, and the presence of free sulfides (depending on metal concentrations and water pH) and lower redox potentials in the effluent waters (Johnson and Hallberg, 2005b).

Metals can also be removed by coprecipitation with (or adsorption onto) Fe and Mn oxides and bacterially produced metal sulfides (Jong and Parry, 2004a; Watson et al., 1995).

Proper studies of metal removal mechanisms should be based on data obtained from the effluent water chemistry and from the solid phase analysis of the bioreactor mixture. Geochemical modeling of water chemistry data in bioreactors can be performed with thermodynamic chemical equilibrium models such as WATEQ4F (Amos and Younger, 2003) and VMINTEQ (Waybrant et al., 1998, 2002; Zagury et al., 2006). These models give useful insights into the chemical equilibrium reactions potentially controlling the concentrations of dissolved metals. Published results generally suggest that early decreases in metal concentration can be attributed to adsorption or precipitation of (oxy)hydroxides and carbonates. However, thermodynamic modeling results should be interpreted with caution since these models do not take into account the bacterial activity that entails the precipitation of metal sulfides (Zagury et al., 2006). Solid phase analysis is also an important step for elucidating the metal removal processes (Machemer et al., 1993). Chemical analyses such as extraction procedures complemented with determination of acid volatile sulfides and simultaneously extracted metals are efficient tools to assess metal fractionation (Jong and Parry, 2004b, 2005). Moreover, mineralogical analyses can help to identify the chemical form of metals in the solid phase. In some studies, the number of techniques for collecting mineralogical data has been limited by the poor crystallinity of the precipitates and the relatively low concentrations of metal sulfides (Song, 2003; Gibert et al., 2005b). Among these methods, scanning electron microscopy has been the most successful technique, whereas X-ray diffraction or Mossbauer analyses have been less effective in detecting amorphous metal sulfides (Machemer et al., 1993). Additional work is needed to accurately assess the various metal removal mechanisms occurring in passive bioreactors.

	Factors of Influence on Sulfate-reducing Bacteria-based Reactors Efficiency

TOP ABSTRACT INTRODUCTION Passive Bioreactors: Principle,... Factors of Influence on... Performance of Passive... Conclusions and Research Needs REFERENCES

 
In passive bioreactors the activity of SRB is the rate-limiting step and bacteria have specific requirements that must be fulfilled. The design of a passive bioreactor requires a near-neutral pH, a source of sulfate, a source of organic carbon, a reducing environment, a solid support for microbial attachment and development, and a way to physically retain metal sulfide precipitates (Dvorak et al., 1992; Lyew and Sheppard, 1997). At the start-up phase, the critical parameters are pH and reducing conditions, flow rate, AMD composition, nutrients, and temperature (Kuyucak and St-Germain, 1994). During treatment, the most limiting factors for the growth of SRB are sulfate concentrations and the type of organic carbon source (Garcia et al., 2001).

Effect of pH, Eh, and Temperature
In order for SRB to thrive, they require a pH in the range of 5 to 8 (Willow and Cohen, 2003). Outside this range, the rate of microbial sulfate reduction generally declines and the metal removal capacity is reduced. Low pH (<5) normally inhibits sulfate reduction and increases the solubility of metal sulfides (Dvorak et al., 1992). In any case, the presence of SRB has been detected in natural waters with pH <3 (Gyure et al., 1990; Kolmert and Johnson, 2001; Koschorreck et al., 2003). Isolated strains were mostly acidophilic SRB, which are more efficient than the neutrophilic ones for remediating acidic waste waters (Kolmert and Johnson, 2001). The SRB were capable of sulfate reduction in a column bioreactor operated under acidic conditions with lactate as organic carbon source (Elliott et al., 1998). At pH 3.25, 38.3% of influent sulfate was removed and pH of the medium rose to 5.82, whereas at pH 3.0 sulfate removal fell to 14.4% and sulfide production dropped below detection. Nevertheless, viable SRB were recovered from the bioreactor after operation at pH 3.0 for 3 wk (Elliott et al., 1998). Efficient AMD treatment was also achieved at a pH as low as 2.5 (Tsukamoto et al., 2004). However, the existence of truly acidophilic SRB is currently not clear. Reduction of sulfate to sulfide has been demonstrated as occurring in extremely acidic environments, but attempts at isolating pure culture of acidophilic (acid-tolerant) SRB failed (Gyure et al., 1990; Johnson, 1998). Nevertheless, a pH of 5.5 or higher is preferred for efficient treatment of AMD in an on-site passive bioreactor (URS Report, 2003).

For optimal performance, SRB need an anaerobic medium and an anoxic and reduced microenvironment with a redox potential (Eh) lower than –100mV (Postgate, 1984). However, sulfate reduction was often observed in passive field bioreactors at positive Eh values (Reisman et al., 2003; Zaluski et al., 2003). Eh measurements of aqueous samples collected at the outlets of the bioreactors might not reflect the real values present in pockets of organic matter where SRB live (Zaluski et al., 2003). Their survival in these adverse conditions may also be explained by the formation of favorable anoxic microenvironments in the reactive mixtures (Lyew and Sheppard, 1999). Batch and column laboratory bioreactors successfully treated AMD at Eh values of –100 to –200 mV or lower during 23 d (Cocos et al., 2002), 30 d (Beaulieu et al., 2000), or 150 d (Gibert et al., 2004). In passive field bioreactors, Eh values as low as –200 mV were maintained for periods ranging from 2 mo to more than 2 yr (Cheong et al., 1998; Reisinger et al., 2000).

In passive bioreactors, the operating temperature affects bacterial growth, kinetics of organic substrate decomposition, as well as hydrogen sulfide solubility. Generally, SRB can tolerate temperatures from below –5 to 75°C (Postgate, 1984). Column experiments showed that the efficiency of AMD treatment was not significantly reduced at temperatures as low as 6°C (Tsukamoto et al., 2004). Passive on-site bioreactors successfully operated for 32 mo at temperatures between 2 and 16°C (Zaluski et al., 2003) or over 2 yr at near-freezing temperatures (1 to 8°C) (Reisinger et al., 2000; Kuyucak et al., 2006). Low temperatures particularly affect the ability of SRB to acclimate, but once acclimated at higher temperature, SRB are not that affected by low temperature (Kuyucak and St-Germain, 1994; Tsukamoto et al., 2004). In field bioreactors started during the winter, a 4-mo lag phase was observed for SRB to be established. However, winter freezing of a well established SRB population had little or no effect on their activity (Zaluski et al., 2003; Kuyucak et al., 2006). The methanogens, which are found when bioreactors are supplied with complex organic carbon sources, are mainly mesophilic microorganisms. Therefore, they are more sensitive to low temperatures than SRB (Kuyucak and St-Germain, 1994).

Effect of Solid Support, Hydraulic Retention Time (HRT), and Hydraulic Conductivity
Sulfate-reducing bacteria require a solid support (sand and/or gravel), onto which they can establish microenvironments for their survival in the presence of extreme conditions such as low pH or high oxygen concentrations (Lyew and Sheppard, 1997). Higher sulfate reduction rates are achieved if SRB have access to a porous surface, compared to suspended bacteria (Glombitza, 2001). A medium with large pore spaces, low surface area, and a large void volume is generally preferred because it minimizes the plugging of the bioreactor (Tsukamoto et al., 2004). In terms of efficiency, better treatment occurs with greater surface area. Surface area and pore size need to be balanced in field bioreactors (Tsukamoto et al., 2004).

Effects of hydraulic retention time (HRT) on efficiency of bioreactors have been widely studied (Dvorak et al., 1992; Al-Ani, 1994; Béchard et al., 1994; Rockhold et al., 2002; Kaksonen et al., 2004b). The variability of hydraulic properties of porous media used in reactive mixtures may result in HRTs specific to each bioreactor. It is usually accepted that precipitation of metal sulfides occurs in at least 3 to 5 d (URS Report, 2003; Kuyucak et al., 2006). A shorter HRT may not allow adequate time for SRB activity to neutralize acidity and precipitate metals or may result in biomass being washed out of the bioreactor. A longer HRT may imply depletion of either the available organic matter source or the sulfate source for SRB (Dvorak et al., 1992). In a semicontinuous anaerobic laboratory bioreactor, more sulfates were reduced to sulfides with a 3-d HRT compared to a 1-d HRT, regardless of the organic carbon/sulfate ratio (Al-Ani, 1994). During treatment, bacteria induce changes in the mixture properties due to accumulation of biomass and generation of metabolic byproducts. The characteristics of the accumulated biomass are dependent on the type of bacteria, the substrate and loading rate, and the flow rate (Rockhold et al., 2002). Bacterial activity might cause a decreased surface tension, decreased porosity and permeability, and pore clogging. The hydraulic conductivity of the substrate material is also an important variable because this will affect the HRT (Bolis et al., 1992; Benner et al., 2001, 2002). Several studies were conducted to evaluate the effect of microbial growth and biomass accumulation on porosity and permeability of saturated porous media. A sand-packed column reactor using sewage bacteria and methanol as a substrate showed a decrease in the saturated hydraulic conductivity (K_s) of 3 orders of magnitude following biomass accumulation (Taylor and Jaffe, 1990). A relatively small variation in hydraulic conductivity could entail important differences in residence times, and might result in decreased efficiency (Benner et al., 2002). Efficient compost substrate passive bioreactors have a hydraulic conductivity about 1 x 10^–4 cm s^–1 (URS Report, 2003). Recently, sawdust has been increasingly used in reactive mixtures, due to a significantly higher conductivity of around 10^–2 to 10^–3 cm s^–1. When sawdust is used, however, there is an increased vulnerability for mixture compaction. Presoaking the substrate before AMD treatment can help provide a more stable hydraulic conductivity and a more consistent flow rate through the system (Bolis et al., 1992). Composted substrates should be at least 0.6 m in thickness but should not exceed 0.9 to 1.2 m; if not, the substrate tends to compact with depth and the permeability becomes too low for effective treatment (URS Report, 2003).

Effect of Chemical Oxygen Demand (COD)/Sulfate (SO₄^2–) Ratio and Nutrients
Several studies have been conducted to find the best chemical oxygen demand (COD)/SO₄^2– ratios for AMD treatment under sulfate-reducing conditions but the results are not consistent. With sludge as the organic carbon source (Al-Ani, 1994) the best performance was found for a COD/SO₄^2– ratio of 5.0, whereas other studies using natural or synthetic substrates found that SRB were predominant for a ratio below 1.7 (Choi and Rim, 1991; Prasad et al., 1999). When complex organic carbon sources were used as substrate, higher optimal ratios were likely obtained because not all the carbon present was used by SRB (Prasad et al., 1999). The assessment of the optimal COD/SO₄^2– ratio for efficient operation of passive bioreactors under sulfate-reducing conditions deserves further investigation.

A C/N ratio around 10 is generally considered suitable for biological degradation of complex organic substrates (Reinertsen et al., 1984; Béchard et al., 1994). Higher ratios indicate an excessive carbon content or nitrogen deficiency, whereas lower ratios may suggest a lack of carbon (Prasad et al., 1999). Nevertheless, Zagury et al. (2006) found that the C/N ratio of reactive mixtures was not a good indicator of the sulfate-reducing activity when they tested six natural organic materials. Moreover, when lactate is used as a substrate, the reported optimal values fluctuate from 15.7 to <45 or between 45 and 120 (Gerhardt et al., 1981; Okabe et al., 1992).

Inhibitory/Toxic Effects of Metals, Sulfides (H₂S, HS^–, and S^2–), Oxygen, and Organic Carbon
Several studies have been conducted to assess metal toxicity to SRB (Gray and O'Neill, 1997; Poulson et al., 1997; Sani et al., 2001a, 2001b, 2003; Utgikar et al., 2001, 2002, 2003, 2004). Some studies were performed with artificially contaminated sulfate-rich waters similar to AMD (Elliott et al., 1998; El Bayoumy et al., 1999; Gibert et al., 2004) or with mine waters pretreated with Na₂S to remove heavy metals (Glombitza, 2001). Results have clearly shown that the effect of heavy metals on SRB can be stimulatory at lower concentrations and inhibitory or even lethal at higher concentrations (Poulson et al., 1997; Utgikar et al., 2002; Sani et al., 2003). Metals may inactivate the enzymes, denature the proteins, and compete with essential cations (Utgikar et al., 2002). In the laboratory, the use of an artificial AMD without metals will simplify the biological system by avoiding reactor clogging. However, an artificial AMD might affect the optimum operating conditions. Evaluation of potential toxic effects of dissolved metals is essential to successfully operate AMD biological treatment systems (Utgikar et al., 2001). The following factors have been identified as important in quantifying the toxicity of heavy metals to SRB—initial metal concentrations, metal sorption and precipitation, and metal complexation with organic ligands (Poulson et al., 1997).

Toxic effects of metals have been reported for concentrations varying from a few mg L^–1 to more than 100 mg L^–1 (Utgikar et al., 2002). Toxic levels are difficult to assess as several studies use "ill-defined" components such as strong chelators, buffers, or reductants in the media that can affect metal activity (Poulson et al., 1997; Sani et al., 2001a, 2001b, 2003). Without chelators in the bacterial growth media, a threshold Zn concentration of 13 mg L^–1 was found toxic to Desulfovibrio desulfuricans (Poulson et al., 1997). In the presence of these compounds, however, toxic concentrations of Zn ranging from 13 to 40 mg L^–1 were reported (Hao et al., 1994; Poulson et al., 1997; Utgikar et al., 2001). To more accurately assess the toxicity of metals to SRB by eliminating abiotic metal precipitation and minimizing formation of metal complexes in solution, a new metal toxicity medium (MTM) was developed and tested (Sani et al., 2001a). Metals can also cause synergetic or cumulative toxic effects as in the case of Ni and Zn (Poulson et al., 1997) or Cu and Zn (Utgikar et al., 2004). During laboratory studies, toxic effects of binary mixtures of Cu and Zn were substantially higher than expected on the basis of additive individual metal toxicity (Utgikar et al., 2004). Therefore, successful operation of a passive bioreactor for treatment of a heavily contaminated AMD might require metal concentrations below inhibitory/toxic levels to maintain a maximum rate of sulphidogenesis.

Exposure to O₂ can inhibit SRB metabolism, although the inhibition is reversible (Nagpal et al., 2000b). The presence of enzymes responsible for O₂ tolerance detected in some SRB can explain their tolerance to low levels of oxygen. In any case, pH was found more critical to bioreactor efficiency than dissolved oxygen (Willow and Cohen, 2003).

Inhibition potential differs among sulfur compounds with toxicity increasing in the following order: sulfate<thiosulfate<sulfite<total sulfide<H₂S (Al-Ani, 1994). Hydrogen sulfide was reported to have a direct and reversible toxic effect on SRB (Reis et al., 1992). Toxic effects were reported for H₂S concentrations varying from 477 to 617 mg L^–1 (Okabe et al., 1992; Reis et al., 1992; Al-Ani, 1994; Kolmert et al., 1997). These H₂S concentrations are, however, much higher than those observed in passive field bioreactors. The inhibition is caused by undissociated H₂S that easily permeates the cell membrane, and by the removal of nutrients as metal sulfides (Reis et al., 1992; Nagpal et al., 2000b; Hulshoff Pol et al., 2001). Furthermore, sulfide toxicity is metal concentration-related because most metals react with sulfide to give insoluble metal sulfides (Hulshoff Pol et al., 2001). According to Utgikar et al. (2002), metal sulfides can inhibit SRB activity by deposition on bacterial cells. Sulfide toxicity is also related to pH, temperature, and organic carbon source. Low pH and low temperature conditions entail higher toxicity because they favor nonionized hydrogen sulfide formation (Hulshoff Pol et al., 2001). Undissociated acetic acid also entailed inhibitory effects to SRB, when hydrogen sulfide was continuously removed from the system. At lower pH (5.8), acetic acid exerted inhibitory effects, whereas at higher pH (6.6), hydrogen sulfide prevailed as the inhibitor of concern (Reis et al., 1992).

Chemical characterization of the organic matter source can provide insights into potential inhibitory effects on SRB (Tassé and Germain, 2002). Phenolic compounds and plant-derived tannins can inhibit the activity of various enzymes, whereas compounds such as terpenoids, amino acids, or hydrophilic base-extracted fractions may exert toxic effects (Chang et al., 2000; Tassé and Germain, 2002; Marschner and Kalbitz, 2003). The exact mechanisms responsible for the observed effects have not been made clear yet.

Ecotoxicity Assessment of Treated Effluent
To our knowledge, Song et al. (2001) is the only study that performed whole effluent toxicity assays on undiluted wetland effluent. In this study, a laboratory-scale wetland was used to treat slightly alkaline (8.0 to 8.5) synthetic lead mine drainage and synthetic lead smelter wastewater. A significant toxicity decrease in wetland effluent for all organisms studied and 100% survival of fathead minnows and Daphnia magna was observed. However, lethality of Ceriodaphnia dubia was 100% in an undiluted effluent. Dilution of effluent to half strength increased survival to 75 to 100%. Wetlands thus offer encouraging promise for decreasing the toxicity of lead-contaminated wastewater.

Configuration of Passive Bioreactors
Vertical flow bioreactors have been used in numerous laboratory and field studies (Dvorak et al., 1992; Cheong et al., 1998; Elliott et al., 1998; Drury, 1999; Tsukamoto and Miller, 1999; Chang et al., 2000; Willow and Cohen, 2003; Tsukamoto et al., 2004; Johnson and Hallberg, 2005b; Kuyucak et al., 2006). In downward flow mode bioreactors, the influent is fed through the top, while in the upward flow mode it is fed through the reactor bottom (URS Report, 2003). Recently, flow in a horizontal plane was reported in a field study (Zaluski et al., 2003). A three-step system separating SRB activity from metal precipitation units and from a pH control system was also proposed at the laboratory scale (Prasad et al., 1999). The flow pattern can affect both the transport of metals and their interaction with the substrate (Song, 2003). Bioreactors with vertical flow may show preferential channels of influent AMD percolating through the reactive mixture. The upward flow bioreactors tend to last longer because upward flow limits compaction and preferential flow paths (URS Report, 2003). However, release of metals by treated effluent is a potential problem. A horizontally oriented bioreactor using a mixture of cow manure and cut straw did not show preferential flow patterns during a 32-mo field operation period (Zaluski et al., 2003). This configuration seems more promising, whereas the three-step process requires higher maintenance costs.

	Performance of Passive Bioreactors

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Neutralization and Alkalinity Generation
Mine water acidity arises from low pH and from dissolved metals (Fe, Mn, and Al) which undergo hydrolysis reactions producing H⁺ (Gazea et al., 1996). The AMD pH must generally be corrected before or during biological treatment. Limestone was successfully used for increasing influent pH either in anoxic drains before alimentation into the bioreactor (Johnson and Hallberg, 2005b) or integrated into the composition of reactive mixture (Reisinger et al., 2000; Cocos et al., 2002; Reisman et al., 2003; Zagury et al., 2006; Kuyucak et al., 2006). During biotreatment, acidity removal and effluent alkalinity are due to bicarbonate produced by bacterial sulfate reduction (Dvorak et al., 1992). Furthermore, generation of ammonia can contribute to an increase in alkalinity (Béchard et al., 1994).

Sulfate Removal
Acid mine drainage frequently contains sulfate levels from 100 to 5000 mg L^–1 (Kolmert and Johnson, 2001). While several mechanisms are responsible for metal removal, sulfate reduction determines sulfate removal and is the best indicator of SRB activity (Johnson and Hallberg, 2005b). Until recently, sulfate reduction was considered as a single mechanism of sulfate removal (Lyew and Sheppard, 1997). However, at early times of treatment, loss of sulfate due to adsorption onto or coprecipitation with ferric (oxy)hydroxides may occur (Christensen et al., 1996; Waybrant et al., 1998; Zagury et al., 2006). A distinction must be made between the factors affecting the amount of sulfate removed and the rates of its removal by sulfate reduction. While the first is related to the available surface area and HRT, the second is dependent on the initial concentration of sulfate in AMD (Lyew and Sheppard, 1999; Chang et al., 2000).

Under optimum field conditions, sulfate reduction occurs at rates of about 0.3 mol m^–3 d^–1 (URS Report, 2003). Results reported in the literature indicate higher sulfate removal in batch experiments compared to column and field bioreactors (Table 1). Comparison of laboratory and field passive bioreactors in terms of the sulfate removal rate is, however, not feasible because of several factors of influence specific to each study. First, the HRTs in batch experiments are far longer than in columns and field-bioreactors. Second, there is the variability of the initial concentrations, ranging from 229 mg L^–1 (Zaluski et al., 2003) to 4800 mg L^–1 (Waybrant et al., 1998), which influences SRB growth and sulfate reduction kinetics (Moosa et al., 2002). Third, some studies provide the calculated percentages of sulfate removal without presenting the initial and final concentrations (Cheong et al., 1998; Elliott et al., 1998; Johnson and Hallberg, 2005b). For example, a sulfate removal efficiency of 42% from 900 mg L^–1 SO₄^2– translates to a residual sulfate concentration about 500 mg L^–1 (Tsukamoto et al., 2004), whereas an efficiency higher than 82% from 2315 mg L^–1 SO₄^2– yields about 400 mg L^–1 of sulfate in the treated effluent (Jong and Parry, 2003).

Moreover, some researchers report sulfate removal taking into account initial sulfate concentration and equivalents of organic carbon consumed (Tsukamoto and Miller, 1999; Tsukamoto et al., 2004). Therefore, great care should be taken when comparing the efficiency of passive systems in terms of sulfate removal.

Metal Removal
As discussed earlier, heavy metals can be removed from AMD via various mechanisms. Increasing the pH of acidic water effectively removes some metals due to precipitation under the form of hydroxides. Bacterial H₂S produced during sulfate reduction results in precipitation of other metals such as Cu²⁺, Zn²⁺, Cd²⁺, Pb²⁺, Ag²⁺, and Fe²⁺ as sulfides. The pH is important because it influences both the solubility of hydroxides and carbonates and the kinetics of hydrolysis and precipitation processes.

Very high initial concentrations of metals in AMD fed to bioreactors (Table 2) may lead to a higher metal load of treated effluent (Zaluski et al., 2003) than in a less contaminated water (Johnson and Hallberg, 2005b). Generally, the proportion of metals removed via sorption has not been clearly quantified. In a monitoring study on the performance of a passive field bioreactor over 32 mo, only Zn, Cu, and Cd were removed as sulfides at thresholds independent of the initial concentrations in AMD. Iron, Mn, Al, and Zn seemed to be removed following precipitation or co-precipitation as hydroxides (Zaluski et al., 2003).

In the case of arsenic, the exact process responsible for its initial removal is not clear but adsorption or concomitant coprecipitation with other metal sulfides or with ferrihydrite has been suggested (Jong and Parry, 2003; Zaluski et al., 2003). Formation of insoluble arsenic sulfide may occur later when reducing conditions are established. Bioreactors that efficiently removed arsenic along with other divalent metals (Fe, Ni) for a period of almost 2 yr were reported in a pilot-scale bioreactor (Tsukamoto and Miller, 1999). Proportions of As(III) and As(V) species in the AMD were suggested as a critical factor affecting the rate of arsenic reduction in different environments (Jong and Parry, 2003).

Long-Term Performance of Pilot-Scale and Full-Scale Systems
Theoretically, AMD-contaminated water can be successfully treated using simple anaerobic passive bioreactors for many years. However, between 3 and 4 yr is the maximum reported operating period (URS Report, 2003). In practice, several problems related to flow patterns, plugging, compacting, overloading, and exhausting of carbon available to SRB occurred during the treatment. These problems affected both the longevity and the performance of bioreactors and ultimately resulted in their failure.

Treatment in a passive bioreactor filled with a mixture of rice stalk, cow manure, and limestone succeeded for only 56 d in removing metals and reducing acidity (Cheong et al., 1998). After 118 d of operation, the substrate thickness had compacted by 15 cm and the pipes for the conveyance of AMD were clogged by brownish iron hydroxide. Another pilot-scale anaerobic bioreactor utilizing a mixture of horse manure and sand as substrate removed <10% of the influent sulfate and iron by the end of the second year. Its use was discontinued because of the exhaustion of available organic carbon (Tsukamoto and Miller, 1999). In a full-scale study, two bioreactors using a mixture of sawdust and hay gave significantly greater iron concentrations in the effluent than in the source water, indicating net mobilization by reductive dissolution of colloidal and/or solid-phase ferric iron compounds (Johnson and Hallberg, 2005b). This phenomenon was attributed to the high proportion of recalcitrant organic materials used in the mixture. Clogging of bioreactors due to formation of sulfide precipitates and microbial biomass is an important operational problem minimized by the use of a physical matrix with large pore spaces, a good hydraulic conductivity, and the ability to flush out the precipitates (Lyew and Sheppard, 1997; Tsukamoto et al., 2004). Overloading is another common reason for passive treatment failure (Reisman et al., 2003).

	Conclusions and Research Needs

TOP ABSTRACT INTRODUCTION Passive Bioreactors: Principle,... Factors of Influence on... Performance of Passive... Conclusions and Research Needs REFERENCES

 
What We Know
Passive on-site bioreactors offer promising perspectives for treatment of AMD-contaminated waters due to the relative simplicity of the system and low operating costs. These technologies are well suited for remediation of abandoned, remote mine sites or for sites located in cold areas. The efficiencies of passive bioreactors depend on the activity of SRB, which is mainly controlled by the composition of the reactive mixture. The most important component is the organic carbon source. Many studies have attempted to predict the biodegradability of complex organic substrates by using chemical extractions; however, they have not been successful. Higher sulfate reduction rates have been reported with reactive mixtures containing more than one organic carbon source. Even though formation of metal sulfides is the preferred metal removal process, many metal removal mechanisms including adsorption and precipitation of metal carbonates and hydroxides occur in passive bioreactors. Furthermore, these mechanisms change during the life span of a passive bioreactor. High concentrations of dissolved metals and sulfides can be toxic to SRB. However, passive bioreactors can still operate at temperatures as low as 2 to 6°C and pH 3.0. Furthermore, SRB can tolerate low levels of oxygen. In pilot and field studies, several problems related to flow patterns, plugging, compacting, overloading, and exhausting of carbon available to SRB have been reported. In addition to reactive mixture composition and presence of SRB, reactor configuration, Eh, hydraulic retention time, and COD/sulfate ratios are critical factors for a long-term operation of passive sulfate-reducing bioreactors.

Research Needs
Different aspects need to be further investigated for better design and operation of on-site passive treatment systems. The depletion rate of organic matter is a key problem. An improved methodical analysis of natural organic substrates is warranted to assess their ability to promote sulfate reduction and metal removal. Anaerobic degradation of complex organic carbon compounds to simpler molecules by other microflora may limit the rate at which substrates become available to SRB. More work must be conducted to understand and differentiate the fundamental biochemical and microbiological reactions that occur in anaerobic bioreactors with complex natural organic substrates. Bioreactors are recommended to be allowed to "mature" before fed with AMD, especially when recalcitrant materials are included in the substrate to provide long-term provision of organic carbon. After maturation, however, the amount of colloids and DOM in pore water and within the effluent should be assessed. Metals bound to DOM and colloids are highly mobile and can flow out of the treatment system. Additional work to accurately assess the various metal removal mechanisms occurring in passive bioreactors is strongly recommended. Geochemical modeling, solid phase speciation analysis, and mineralogical and microbial characterization should be performed to assess the various metal removal mechanisms. The inclusion of natural organic matter in geochemical equilibrium models and metal speciation analysis should be further studied. This knowledge should lead to a more efficient long-term operation of passive bioreactors. Limited work has been done on the direct assessment of the ecotoxicological potential of biologically treated AMD waters. Characteristics of natural organic materials, high concentrations of dissolved organic carbon, low redox potential, high concentrations of dissolved sulfides, and enhanced metal availability are among the parameters that might influence effluent toxicity. Correlation of metal speciation in the treated effluent and in the reactive mixture with toxic effects of treated waters could help improve our understanding of passive bioreactor systems.

	REFERENCES

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