Bioremediation of acid-rock drainage by sulphate-reducing prokaryotes: A review A.S. Sheoran a, * , V. Sheoran b , R.P. Choudhary a a Department of Mining Engineering, Faculty of Engineering, Jai Narain Vyas University, Jodhpur 342 011, India b Department of Zoology, Faculty of Science, Jai Narain Vyas University, Jodhpur 342 011, India article info Article history: Received 26 November 2009 Accepted 4 July 2010 Available online 24 July 2010 Keywords: Mining Sulphide ores Acid-rock drainage Reclamation Bacteria abstract Acid-rock drainage (ARD) is a widespread environmental problem that causes adverse effects to the qual- ity of ground water and surface water through acidification, high concentration of the iron, sulphate, and elevated levels of soluble toxic metals. Active treatment technologies are often expensive and require regular attention resulting in increased overall costs due to operation and maintenance expenses. One of the effective treatment methods is to use sulphate-reducing prokaryotes (SRP) in bioreactors. They offer advantages such as high metal removal at low pH, stable sludge, very low operation costs, and min- imal energy consumption. Sulphide precipitation is the desired mechanism of contaminant removal; however, many mechanisms including adsorption and precipitation of metal carbonates and hydroxides also occur in passive bioreactors. Several factors influencing the performance of the bioreactors are reviewed. The fundamental biochemical and microbiological reactions that occurs in the bioreactors has been dealt in detail. The present review presents performance of bioreactors, chemical characterisa- tion of organic substrates for successful treatment of ARD. Moreover, design parameters, longevity and future scope of the study on bioreactors is also discussed in this review. Ó 2010 Elsevier Ltd. All rights reserved. Contents 1. Introduction ........................................................................................................ 1074 2. Treatment of acid-rock drainage ........................................................................................ 1075 3. Sulphate-reducing prokaryotes (SRP) .................................................................................... 1075 4. Basic components of sulphate-reducing bioreactors (SRBR) .................................................................. 1076 5. Geochemistry of metal precipitation in SRBR.............................................................................. 1076 5.1. Other contaminant removal processes in SRBR ....................................................................... 1077 5.1.1. Adsorption............................................................................................. 1078 5.1.2. Biosorption ............................................................................................ 1078 5.1.3. Precipitation ........................................................................................... 1078 5.1.4. Co-precipitation ........................................................................................ 1078 5.1.5. Sedimentation and filtration .............................................................................. 1078 5.2. Behaviour of various metals in bioreactor ........................................................................... 1078 5.2.1. Iron .................................................................................................. 1078 5.2.2. Zinc and Lead .......................................................................................... 1079 5.2.3. Aluminum ............................................................................................. 1079 5.2.4. Manganese ............................................................................................ 1079 5.2.5. Arsenic................................................................................................ 1079 5.2.6. Uranium .............................................................................................. 1079 6. Organic substrate .................................................................................................... 1079 0892-6875/$ - see front matter Ó 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.mineng.2010.07.001 Abbreviations: AMD, acid mine drainage; ARD, acid-rock drainage; ASF, acid soluble fraction; DOC, dissolved organic carbon; EC, electric conductivity; EIF, ethanol- insoluble fraction; ESF, easily soluble fraction; HRT, hydraulic retention time; Mg/L as CaCO 3 , milligrams per litre equivalent to calcium carbonate; OC, oak chips; ORS, organic rich soil; pH, potential hydrogen; SMC, spent mushroom compost; SRB, sulphate reducing bacteria ; SRBR, sulphate reducing bioreactor; SRP, sulphate reducing prokaryotes ; SWP, sludge from a wastepaper recycling plant; TKN, nitrogen content (total Kjeldahl nitrogen); TOC, total organic carbon; USEPA, United States Environmental Protection Agency; VFW, vertical flow wetlands; WIF, water-insoluble fraction; WSF, water soluble fraction. * Corresponding author. Tel.: +91 291 2611512, +91 9414411623. E-mail addresses: [email protected], [email protected](A.S. Sheoran). Minerals Engineering 23 (2010) 1073–1100 Contents lists available at ScienceDirect Minerals Engineering journal homepage: www.elsevier.com/locate/mineng
Bioremediation of acid-rock drainage by sulphate-reducing prokaryotes: A review
A.S. Sheoran a,*, V. Sheoran b, R.P. Choudhary a
a Department of Mining Engineering, Faculty of Engineering, Jai Narain Vyas University, Jodhpur 342 011, Indiab Department of Zoology, Faculty of Science, Jai Narain Vyas University, Jodhpur 342 011, India
a r t i c l e i n f o a b s t r a c t
Article history:Received 26 November 2009Accepted 4 July 2010Available online 24 July 2010
Acid-rock drainage (ARD) is a widespread environmental problem that causes adverse effects to the qual-ity of ground water and surface water through acidification, high concentration of the iron, sulphate, andelevated levels of soluble toxic metals. Active treatment technologies are often expensive and requireregular attention resulting in increased overall costs due to operation and maintenance expenses. Oneof the effective treatment methods is to use sulphate-reducing prokaryotes (SRP) in bioreactors. Theyoffer advantages such as high metal removal at low pH, stable sludge, very low operation costs, and min-imal energy consumption. Sulphide precipitation is the desired mechanism of contaminant removal;however, many mechanisms including adsorption and precipitation of metal carbonates and hydroxidesalso occur in passive bioreactors. Several factors influencing the performance of the bioreactors arereviewed. The fundamental biochemical and microbiological reactions that occurs in the bioreactorshas been dealt in detail. The present review presents performance of bioreactors, chemical characterisa-tion of organic substrates for successful treatment of ARD. Moreover, design parameters, longevity andfuture scope of the study on bioreactors is also discussed in this review.
Mining activities generate waste products such as mine over-burden and mine tailings (waste soil). The management of thesewaste materials is an important issue for mining industry world-wide. This becomes challenging especially, if the waste materialcontains sulphide ores. Dissolution of sulphide ores exposed tooxygen, water and microorganisms results in the production ofenvironmentally detrimental acid mine drainage (AMD), morecommonly acid-rock drainage (ARD) that causes adverse effectsto the quality of ground water and surface water through acidifi-cation, high concentrations of iron, sulphate, and elevated levelsof soluble toxic metals (Evangelou and Zhang, 1995; Elliot et al.,1998; Mohan and Chander, 2001; Ritcey, 2005; Sheoran, 2006a,b;Nehdi and Tariq, 2007; Natarajan, 2008; Riefler et al., 2008). About19,300 km of rivers and streams and more than 72,875 hectares oflakes and reservoirs in the continental US have been damaged byARD (Kleinmann, 1989; Wilkin and McNeil, 2003; Hallberg andJohnson, 2005).
Many metals of commercial value occur as metal sulphides,found in association with the most abundant sulphide mineral pyr-ite (FeS2). Additionally variable amounts of pyrites are found in thecoal deposits. Mining of these coals and metals exposes the pyriteto oxygen and water, which coupled with microbial activity, leadsto formation of water that are highly enriched with sulphate, alu-minium and range of heavy metals, the most significant of which isiron (Johnson, 2003; Ackil and Koldas, 2006; Koski et al., 2008).Both operational and abandoned mine works contribute to ARD.Under acidic conditions resulting from ARD, the oxidation of pyriteproceeds by the following reaction (Stumm and Morgan, 1981):
FeS2 þ 14Fe3þ þ 8H2O! 15Fe2þ þ 2SO2�4 þ 16Hþ ð1Þ
This reaction demonstrates the polluting capability of the oxi-dation of pyrite that every mole of pyrite can be converted to16 mol of hydrogen and 2 mol of sulphate (Zhang et al., 2005). Thisreaction serves as a template for the similar oxidation reactions ofmost metal sulphides which also contributes acidity, sulphate andtoxic metal ions to the aquatic environment resulting in a ‘classical’ARD discharge (Cohen, 2006).
Remote abandoned metal mines and coal mines around theworld also discharge acidic, metal laden waters, which can persistfor a very long time largely because of its spontaneous nature ofreaction and the financial constraints. The combination of acidity,heavy metals, and sediment loading associated with ARD fromthe abandoned mine can have severe detrimental impacts onreceiving ecosystems. Conventional treatment methods for suchsites with no production prove to be too expensive to be econom-ically attractive (Amos and Younger, 2003; Kalin and Chaves, 2003;Ntengwe, 2005; Younger et al., 2005). The increasing awarenessand concern about the environment has motivated in the recentyears extensive research into developing new efficient technolo-gies for treatment of the ARD (Gibert et al., 2005a). Nearly US $1million is spent each day on ARD prevention and abatement bymining companies throughout the US (Kleinmann and Hedin,1993). A study by the US Environmental Protection Agency esti-mated the cost of remediation by conventional treatment technol-ogies to be more than $ 5 billion in the state of Pennsylvania alone(Hedin et al., 1994).
Nordstrom and Alpers (1999) estimated that without preventa-tive measures the Richmond Mine, at California’s Iron Mountain,
would generate ARD, with pH < 1 and containing several g/L of dis-solved metals, for 3000 years. Kalin (2001) estimated, based onoxidation rates derived from tailings pore water samples, thatthe site of a small zinc/copper mine in northwest Ontario, Canada,would generate ARD for 1000–35,000 years. These may be extremeexamples but it is not uncommon for base metal mines and theirwaste products to generate acid for more than 100 years. SinceARD is self-renewing, an ideal solution to it would also be self-perpetuating.
Today there are legislation and regulations concerning miningand mined land reclamation in all most all countries to prevent/minimise damage to mined lands. To conform to these regulations,the mining industry has developed and continues to develop pollu-tion control measures and treatment strategies until some selfregenerative treatment system with minimum maintenance andeconomically viable replaces conventional treatment system(Cohen, 2006). A growing body of evidence now suggests a causalrelationship between the environmental and economic perfor-mance of the companies. Since the reduction of the pollutionenhances profit by increasing efficiency, reducing compliance costsand minimizing future liabilities; economics, should all be ofinterest to the mining sector (Kalin et al., 2006).
2. Treatment of acid-rock drainage
Numerous techniques are available for ARD treatment. Manyare established methods, while others are still in experimentalstage. Often, only a combination of various treatment processescan provide the effluent quality desired. There are two distinctstrategies for treating ARD. The conventional solution is to collectand chemically treat acidified effluents in a centralized treatmentplant (Yu et al., 2000; Kaksonen and Puhakka, 2007; Gallegos-Garciaet al., 2009). Alternatively, effluents can be routed through naturalor constructed wetlands within which microbial communities per-form the same function (Kalin et al., 2006). It is economical, non-polluting, and is not a source of secondary wastes. Moreover, awell-engineered passive treatment system is a closed ecologicalsystem and hence is self-renewing.
Passive treatment options are often more appealing than activetreatments because they are relatively less expensive to install, andrequire little maintenance. Since many of the sites are remote andwithout access to power, active treatment is not a viable option.Biotic passive treatment systems rely on microbial processes toremediate acidity and dissolved metals, and include aerobic andanaerobic wetlands and bioreactors (Sheoran, 2004, 2005; Cohen,2006; Bartzas et al., 2006; Sheoran and Sheoran, 2006; Komnitsaset al., 2007). Sulphate-reducing bioreactors (SRBRs) rely on themicrobially mediated reduction of sulphate to sulphide. This pro-cess generates alkalinity and the biogenic sulphide can precipitatedissolved metals as highly insoluble solids (Cabrera et al., 2006;Qiu et al., 2009).
Sulphate-reducing bioreactors rely primarily on the metabo-lism of sulphate-reducing prokaryotes (SRP) for the process ofmetal precipitation; however these systems are populated byan entire community of microorganisms that coexist and interact.Much work has been done in describing the microbial communi-ties responsible for the generation of ARD (Benner et al., 2001;Johnson and Hallberg, 2005a). The role of SRP in the remediationprocess has also been well elucidated (Dvorak et al., 1992;Christensen et al., 1996; Chang et al., 2000; Natarajan, 2009).With a necessity of improvement in the biological remediationtechniques compost bioreactors using anaerobic sulphate-reduc-ing prokaryotes (SRP) has been receiving increased attention withwetlands. Although many of the initial systems were designed onempirical relationships, recent technology of bioreactors is based
on a more thorough understanding of chemical processes, and onboth the successes and failures of the past. Sulphate-reducingprokaryotes have the advantage of growing in the miningenvironments.
Recent developments and improvements have resulted in con-struction of bioreactors that have smaller footprints, and treatthe metals and acidity more efficiently. Interest has also risenrecently in treating ARD with the help of natural absorbents assubstrates in the bioreactors, providing a medium to cultureSulphate-reducing bacteria. In these systems, ARD (mine waste)passes through a reactor with bio-degradable solid waste such asmanure, compost or woodchips. Compost bioreactors – biotechno-logical treatment system offers a less expensive alternative to theconventional precipitation technologies.
As the traditional conventional (active) systems are becomingcostly in time or in applicable in abandoned and remote regions,research needs to be focused on passive biological systems. Thecompost based bioreactors certainly have advantages. It is timeto replace the environment with an alternative, economicallyattractive process which is efficient, long lasting, with no sludgegeneration, a low maintenance system – the compost bioreactorsto treat metal laden acidic drainages. The overall benefits of thetechnology will out weigh the previous treatment technologiesapplied by developing countries. The objective of this paper is toreview all information available on SRP and their application forthe treatment of ARD in bioreactors.
3. Sulphate-reducing prokaryotes (SRP)
Recognition of the biological nature of sulphate reduction innatural environments, and identification of the bacterial species in-volved deals to the later part of nineteenth century (Moosa et al.,2002). For many years, sulphate-reducing bioreactors (SRBRs) havebeen treated as black boxes without any thorough understandingof the members or roles of the microorganism involved. Since theoperation of SRBR is highly dependent on microbial activity, abetter understanding of the role of the microbial community inthese systems will help improve their design and performance(Kaksonen et al., 2004a; Hallberg and Johnson, 2005). From thestandpoint of this study, a complex community is responsible formine wastewater remediation (Logan et al., 2005).
Sulphate-reducing prokaryotes are characterised by anaerobicrespiration using sulphate as a terminal electron acceptor and anorganic source as an electron donor. They are classified into the fol-lowing four taxonomic groups (Odum and Singleton, 1993; Moosaet al., 2005):
� The d-Proteobacteria subdivision contains Gram-negative meso-philic SRP, including the genera Desulfovibrio, Desulfomicrobium,Desulfobulbus, Desulfobacter, Desulfobacterium, Desulfococcus,Desulfosarcina, Desulfomonile, Desulfonema, Desulfobotulus, andDesulfoarculus. These bacteria have optimal growth tempera-tures ranging from 20 to 40 �C. This group is diverse, with a vari-ety of shapes and physiological traits represented.� The Gram-positive spore-forming SRP are mainly represented
by the genus Desulfotomaculum, and form heat-resistant endo-spores. Most species require a similar temperature range toGroup 1, though some withstand higher temperatures.� The bacterial thermophilic SRP group contains the genera Ther-
modesulfobacterium and Thermodesulfovibrio These bacteriahave optimal growth at 65–70 �C, and inhabit high-temperatureenvironments.� Archaeal thermophilic SRP thrive at temperatures above 80 �C,
and have been found only in marine regions. All SRP in thisgroup belong to the genus Archaeoglobus (Castro et al., 2000).
Bhattacharya et al. (2008) reported that various studies on thespecies of algal and fungi also effectively enrich the carbon sourcesthat help to maintain the SRP populations in predominantly anaer-obic environment. Fungi show capacity to absorb significantamount of metals in their cell wall, or by extracellular polysaccha-ride slime (Das et al., 2009).
Sulphate-reducing prokaryotes inhabit a variety of sulphate-rich, reducing environments. High numbers of SRP have beenfound in lacustrine and wetland sediments, cattle rumens, and geo-thermal vents. They can also thrive in human-impacted environ-ments such as rice paddies, paper mills, and streams impacted bysewage or ARD (Postgate, 1965). To stimulate bioremediation, anSRP source such as cattle rumens or organic matter from one ofthe environments listed above is generally added to passive treat-ment systems. Suitable substrate temperature, pH and ARD chem-istry is needed to propagate them to ensure treatment success(Tsukamoto et al., 2004).
4. Basic components of sulphate-reducing bioreactors (SRBR)
A SRBR is a shallow basin filled with substrate. Water is intro-duced at the top/bottom and flows through the substrate and dis-
Fig. 1. Schematic diagram of a su
Fig. 2. Sulph
charged at the other end through the discharge pipe fitted at thebottom/top, which controls the depth of water. A liner is also in-cluded beneath the substrate. The SRBR, a new approach fordecreasing environmental pollution, consists of a properly de-signed shallow basin with a sub-surface flow of ARD. Sulphatereduction has been shown to effectively treat ARD containing dis-solved heavy metals, including aluminum, in a variety of situa-tions. The chemical reactions are facilitated by Desulfovibrio,Desulfotomaculum and Desulfobacterium bacteria in a SRBR asshown in Fig. 1.
5. Geochemistry of metal precipitation in SRBR
Many studies have demonstrated that the primary removalmechanisms for the metals in sub-surface wetlands are sulphate-reducing prokaryotes (SRP). These microbes facilitate the conver-sion of sulphate to sulphide. The sulphides react with metals toprecipitate them as metal sulphides, many of which are stable inthe anaerobic conditions of the treatment system. It is necessaryto have an understanding of dissimilatory sulphate reduction,which is a part of the natural sulphur cycle (Fig. 2). Green plantsand some microorganisms can consume the sulphur in its oxidized
state (sulphate), and this sulphur is utilized for the formation ofamino acids, proteins, nucleic acids, and various sulphur-contain-ing co-enzymes. These amino acids are stored in the form of micro-bial protein in the plants (Postgate, 1984; Baskaran, 2005; Tanget al., 2009). When these plants are consumed by animals the re-duced form of sulphur (sulphide) is virtually returned back to thedead organism-called assimilatory sulphate reduction. This studydeals with the type of sulphate reduction carried out by a uniqueanaerobic micro organism called SRBs. SRBs are either heterotro-phic or autotrophic anaerobes, capable of reducing sulphate to sul-phide by a dissimulator, bioenergetic metabolism when providedwith a suitable organic carbon source (Postgate, 1984).
Microbes facilitate the transfer of electrons from electron rich(reduced) substrates (i.e., oxidants such as oxygen or sulphate) togenerate energy for metabolic activity. Substrate (electron donor)oxidation is coupled to sulphate (terminal electron acceptor)reduction. The resulting energy is used by SRP for growth anddevelopment. The reaction is generally expressed as (Widdel,1988; McMohan and Daugulis, 2008):
2CH2Oþ SO2�4 ! S2� þ 2CO2 þ 2H2O ð2Þ
S2� þ 2CO2 þ 2H2O! 2HCO�3 þH2S ð3Þ
where CH2O represents a simple organic carbon source. The dis-solved inorganic carbon neutralizes the pH and favours the precip-itation of metal carbonate minerals. The soluble sulphides (H2S,HS–, and S2–) react with metals to form metal sulphide precipitates;
H2SþM2þ !MS # þ2Hþ ð4Þ
where M is a cationic metal such as Cd, Fe, Ni, Cu, or Zn symbolises adivalent metal with a low metal sulphide solubility product. Alter-natively, divalent metal may precipitate as metal hydroxide withneutral pH. The two protons released by Reactions (4) and (5) areneutralised by the alkalinity generated by Reactions (2) and (3)(Drury, 1999).
M2þ þ 2H2O!MðOHÞ2ðsÞ þ 2Hþ ð5Þ
In acid mine water lakes, bacterial sulphate reduction occurs of-ten in combination with microbial and chemical Fe(III) reduction.
With overall reaction summarized in following equation:
30ðCH2OÞ þ 12FeOOHðsÞ þ 14SO2�4 þ 28Hþ
! 30CO2 þ 12FeSþ 2S0 þ 50H2O ð8Þ
The produced H2S may precipitate as sedimentary iron sulphidemineral and subsequently converted to pyrite. This reduction ofiron (III) and sulphates by carbohydrates leads to generation ofalkalinity. Thereby potentially resulting in an effective neutralisa-tion (Schindler et al., 1986). Thomas et al. (2000) reported the sus-pected geochemical behaviour of aluminum in sulphate-reducingbioreactors. It is suspected that insoluble aluminium sulphateforms in the reducing environments found in the sulphate-reduc-ing bioreactors, perhaps in accordance with the following reactionwhich is one of many possible:
The key conditions for SRP health are a pH of 5.0 (maintained bythe SRPs themselves through the bicarbonate reaction), the pres-ence of a source of sulphate (typically from the ARD), and organicmatter (CH2O) from the substrate (Gusek, 2005). Sulphate-
reducing prokaryotes use a number of volatile fatty acids andhydrogen (H2) as electron donors (Castro et al., 2000). SRP activityconsumes approximately one or two moles of protons (H+) permole of sulphate (SO2�
4 ) reduced and produces approximatelytwo equivalents of alkalinity per mole of sulphate reduced. Exactnumber of protons consumed and alkalinity produced vary withthe electron donor, i.e. the carbon source [acetate (CH3COO�), lac-tate (CH3CHOHCOO�) and propionate (CH3CH2COO�)].
CH3COO� þ SO2�4 þHþ ! H2Sþ 2HCO�3 ð10Þ
CH3CHOHCOO� þ 3=2SO2�4 þHþ ! 3=2H2Sþ 3HCO�3 ð11Þ
4CH3CH2COO� þ 7SO2�4 þ 6Hþ ! 7H2Sþ 12HCO�3 ð12Þ
4H2 þ SO2�4 þ 2Hþ ! H2Sþ 4H2O ð13Þ
Reactions (10)–(13) neutralize two to three equivalents of acid-ity per mole of sulphate reduced at pH ranging from 6.3 to 7.0. Forlower pH values, carbon dioxide will be produced instead of bicar-bonate. For such a pH Reaction (10) would be rewritten as:
CH3COO� þ SO2�4 þ 3Hþ ! H2Sþ 2CO2 þ 2H2O ð14Þ
From Eq. (14) it is clear that the SRP activity neutralizes two tothree acidity equivalents per mole of sulphate reduced for the pHrange of 7.0 down to the lowest pH at which SRP will remain active(Postgate, 1979).
Thus in SRBR cells, the geochemical conditions are predomi-nantly reducing (that is, oxidation reduction potentials less than�100 mV) and neutral to slightly alkaline pH. In the presence ofdissolved sulphide ions (H2S), metals can be precipitated/formedas the sulphides (Wildeman et al., 1993). Apart from sulphatereduction metals are also form hydrous sulphates (Thomas et al.,2002), carbonates (Wildeman et al., 1993), oxides, and nativeforms in the SRBR.
5.1. Other contaminant removal processes in SRBR
The main mechanisms of metal removal in bioreactors are pre-cipitation in the form of sulphides (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 biologicallymediated processes, ARD quality is improved by filtration of thesuspended and colloid materials (Wildeman and Updegraff,1997). The pH is important because it influences both the solubilityof hydroxides and carbonates and the kinetics of hydrolysis andprecipitation processes.
Gibert et al. (2002) reported that the low degradability of someof the current carbon sources (typically complex plant derivedmaterials) used in such treatments may limit the activity of SRBs.In these non sulphate-reducing conditions field and laboratoryexperiences has shown that mechanism other than sulphide pre-cipitation should be considered in the metal removal, i.e. metal(oxy) hydroxides precipitation, co-precipitation with there precip-itates, and sorption of heavy metals onto a broad range of low costand waste organic materials has successfully been applied to thetreatment of industrial effluents and natural water.
Hard and Higgins (2003) also concluded that the metals of con-cern in mine drainage can be precipitated as (oxy) hydroxides and/or sulphides. The most important metal removal processes in-volved are redox reactions, and these are complemented by otherssuch as precipitation; the sorption of metal by algae, bacteria, plantdebris, organic substrates; and ferric hydroxides; and plant uptake.Any metals in mine drainage that do not get fixed to organic ormineral matter probably will not exist as free ions but rather asions associated with suspended colloids.
5.1.1. AdsorptionAdsorption is essentially a surface phenomenon in which accu-
mulation of metal ions takes place at the surface rather than inbulk of a solid or liquid. The metal removal mechanisms changeduring the life of a passive bioreactor. Upon start-up of a passivebioreactor, the adsorption of dissolved metals onto organic sitesin the substrate material will be an important process. Gibertet al. (2005a) found that the strength of sorption to humic materi-als varied with metal species as follows:
Fe > Cu� Zn�Mn
The adsorption process also varies with pH. In bioreactors, me-tal removal due to adsorption takes place until sorption sites aresaturated. Thus, the system will have a fixed life cycle, perhapsas short as a month (Willow and Cohen, 2003). Utgikar et al.(2001) confirmed competition among Fe, Cu, Zn, and Mn by labora-tory tests for organic adsorption sites, and field tests in wetlands.Whereas at pH > 5.0 for Cu(II) or 6.6 for Zn(II), the desorption orprecipitation contribution increased significantly compared toadsorption. Functional groups capable of metal sorption such ascarboxylic and phenolic groups are deprotonated at high pH andpresumably available for binding dissolved metals. Therefore, atthe slightly acidic to neutral pH of on-site sulphate-reducing biore-actors, adsorption of dissolved metals on the substrate material isan important metal removal mechanism. Over time, however, theadsorption sites become saturated. This saturation may take from3 to 8 weeks (Waybrant et al., 1998; Willow and Cohen, 2003) to4 to 8 months (Zaluski et al., 2003).
Katsumata et al. (2003) proposed a simple and cheap adsorptionprocess for removal of heavy metals in wastewater by using eco-nomic adsorbent such as silica gel, illonite, tobermorite and mag-netite etc. The removal efficiency by adsorption onto silica gel,illonite, tobermorite and magnetite was high value for all metals.Gibert et al. (2004, 2005b) presented some laboratory data high-lighting the Zn and Cu adsorption on vegetal compost, and to de-velop a general and simple model for prediction of theirdistribution in organic based passive treatment for ARD.
Cohen (2006) reported that adsorption is a process in which acation, like Fe2+, is bound to the solid phase that contains a residualnegative charge on its surface (usually in the form of a hydroxideion). Complexation can be the result of humic materials terminat-ing in phenolic and carboxylic groups that dissociate under partic-ular pH conditions.
5.1.2. BiosorptionBiosorption is another important mechanism in bioreactors for
the removal of metal ions, in which adsorption and absorption takeplace simultaneously. Factors such as availability of nutrients dur-ing growth, age and physiological state of bacterial cells, environ-mental conditions (pH, ionic strength, and temperature),presence of competitive ions, and concentration of the biomasscan influence biosorption (Chen et al., 2000; Utgikar et al., 2001;Kosolapov et al., 2004; Santos et al., 2004; Martins et al., 2010).
Chen et al. (2000) reported that at pH 7.0, biosorption capacitywas constant regardless of the experimental conditions (e.g., stir-ring and biomass type). The differing results may be due to a moreor less active microbial population in the biomass used as well asto a possible competition between metals. Also it is a physico-chemical process that occurs both on living and dead cells irrespec-tive of biochemical activity. In the pH range 4–7, SRP retain metalsvia biosorption due to the neutral and/or deprotonated state ofbinding ligands on cell walls. The biosorption by SRP is metabo-lism-independent (sorption onto the cell wall) or metabolism-related (transport, internal compartmentalization, and extracellu-lar precipitation by metabolites).
In the study of Chen et al. (2000), biosorption on Desulfovibriodesulfuricans was strongly pH-dependent. For Cu (II) and Zn(II),biosorption increased within a pH range of 4.0–6.6. At pH below3.0, metal biosorption was insignificant due to the strong affinityof protons onto metal binding sites on biomass cell walls. Becauseof several factors of influence, the experimental results are not al-ways in agreement. At pH 3.0, a biomass content >6 g/L increasedthe efficiency of metal removal, favouring sedimentation of theiron precipitates and rates of filtration (Santos et al., 2004).
5.1.3. PrecipitationOnce sulphate-reducing conditions are established, sulphide
precipitation becomes the predominant mechanism of metal re-moval from ARD (Machemer et al., 1993; Béchard et al., 1994; Song,2003). Sulphide precipitation is the desired mechanism of contam-inant removal because metal sulphides are highly insoluble andless bio-available compared with other metal species (Wildemanand Updegraff, 1997). Kuyucak (2002) observed improvement inthe water quality due to precipitation of the metals as sulphideswith the H2S generated in organic substrate and neutralisation ofacidity due to bicarbonate released during sulphate reduction inSRBR process. This process is particularly effective for removingheavy metals such as cadmium, copper, lead, mercury, zinc andiron to low concentration.
Zaluski et al. (2003) concluded after monitoring study on theperformance of a passive field-bioreactor over 32 months that onlyZn, Cu, and Cd were removed as sulphides at thresholds indepen-dent of the initial concentrations in ARD. Iron, Mn, Al, and Znseemed to be removed following precipitation or co-precipitationas hydroxides. Johnson and Hallberg (2005b) observed that sul-phate reduction in passive bioreactors is confirmed by lower con-centrations of sulphates in the effluent than in the influent waters,and the presence of free sulphides (depending on metal concentra-tions and water pH) and lower redox potentials in the effluentwaters.
5.1.4. Co-precipitationMetals can also be removed by co precipitation with (or adsorp-
tion onto) Fe and Mn oxides and bacterially produced metal sulp-hides (Jong and Parry, 2003).
5.1.5. Sedimentation and filtrationSedimentation and filtration are required in order to remove the
solid phase precipitates in the wet sludge prior to discharge intoreceiving waters. Solid phase analysis is also an important stepfor elucidating the metal removal processes.
5.2. Behaviour of various metals in bioreactor
5.2.1. IronThe term ‘‘ochre” is commonly used as a collective term for the
red, orange and yellow iron salts produced by hydrolysis of ferriciron. The precise composition of ochre varies with pH and theavailability of dissolved anions such as sulphate. Under circum-neutral conditions (pH 6–8), a mixture of amorphous iron hydrox-ide and goethite (a-FeOOH) precipitates. At more elevated pH (>8)ferrihydrite (Fe(OH)3) is more commonly precipitated. Which alsoyield three protons (H+) for every mole of ferric iron which hydro-lyses. At lower pH conditions, substitutes of SO2�
4 for OH� occur,resulting in the formation of ‘‘oxyhydroxysulphate” minerals suchas schwartmanite.
The rate limiting aspect of the oxidative iron removal process inpassive mine water systems varies primarily with pH, and alsowith iron and dissolved oxygen concentrations (e.g., Watzlafet al., 2000a). At pH values less than 3, microbial processes oxidizeFe2+ to Fe3+, but the slow kinetics of hydrolysis limit the formation
Table 1Solubilities of metal sulphides and hydroxides (Ksp, 25 �C) (Machemer et al., 1993).
of a solid. At pH values between 3 and 6, iron removal is limited bythe oxidation step. At pH values greater than 6, iron removal ap-pears to be limited by the oxidation step at concentrations greaterthan 10–20 mg/L, but becomes limited by the solid precipitationstep at concentrations less than 10 mg/L. Under reducing condi-tions, iron can be removed from mine waters in passive treatmentsystems by the formation of iron sulphides and iron carbonates.
The formation and stability of both of these solids requires cir-cum-neutral pH conditions, as the production of the necessary HS�
and HCO�3 anions relies on microbial alkalinity-generatingprocesses.
5.2.2. Zinc and LeadGibert et al. (2005b) estimated that 60% of the influent Zn was
removed by co-precipitation with Fe and Al (oxy) hydroxides and40% by sorption onto compost before saturation of compost sorp-tion sites occurred. Once sorption sites of compost were saturated,co-precipitation appeared to be the only mechanism responsiblefor the Zn removal.
Various low cost adsorbent such as onion skin, tea leaves andpeat moss are known to adsorb Pb(II) ions from solutions in theirnative state and with suitable chemical treatment, the adsorptioncapacity can be significantly enhanced (Raji and Anirudhan, 1997).
5.2.3. AluminumAluminum solubility is governed largely by water pH and is
unaffected by the oxidation and reduction processes that affectother metals (Stumm and Morgan, 1996). Dissolved Al concentra-tions remain high at pH 3.5 or less, but decrease to less than1 mg/L at a pH between 5 and 8 (Hedin et al., 1994). The Al com-pounds in the bioreactors are generally hydroxides and sulphates(Thomas et al., 2006). In reality, Al(OH)3 is Al(H2O)3(OH)3; this neu-tral substance has a very low solubility. Hard and Higgins (2003)reported that the aluminum that cannot be removed by precipita-tion are taken out by biosorption, either through accumulation ofthe ions in the bacterial cell or by adsorption on the cell surface.At the same time, the reduction of sulphate uses up protons, whichresults in neutralisation of the ARD.
5.2.4. ManganeseMetal manganese is more challenging. Their removal as sulp-
hides is less effective in passive bioreactors (Dvorak et al., 1992;Wildeman and Updegraff, 1997; Cheong et al., 1998; Chang et al.,2000; Jong and Parry, 2003; Zaluski et al., 2003). Cheong et al.(1998) reported that in the case of manganese, it is related to therelatively high solubility of MnS, which forms only when the Mnconcentrations are very high compared with other metals (Table1). Furthermore Mn is generally weakly sorbed (Willow and Cohen,2003).
Gordon (1985) reported successful removal of manganese usinga packed column reactor filled with 2–5 cm chert stones from theDuck River below Normandy Dam, Tennessee. A ‘‘black slime’’developed on the stones, but the exact nature of the removalmechanisms is yet undetermined. Possibilities range from adsorp-tion, chemical oxidation and even bacterial mediation to achievethe near 100% removal efficiency. It was later discovered that the‘‘black slime’’ could be successfully transferred to other medium,such as glass marbles. The efficiency of the system was solely afunction of substrate surface area and hydraulic loading rates(Gordon and Burr, 1989).
Yoo et al. (2004a,b) thoroughly investigated the optimum condi-tion for Mn removal using a SRP bioreactor. Results showed that an
excess of H2S was required to remove Mn2+. Johnson and Hallberg(2005b) reported a novel enhanced bioremediation system thatconsists of a passively aerated sub-surface gravel bed. The provi-sion of air and the use of catalytic substrates helped to overcomethe slow kinetics of manganese oxidation.
According to Johnson and Hallberg (2005b) a pH > 8 is requiredto abiotically oxidize Mn(II) to insoluble Mn(IV) and to form insol-uble hydroxides and carbonates. Similarly, Zagury et al. (2006) re-ported a rapid removal of Mn as MnCO3 (initial concentration of14 mg/L, pH around 8) during batch experiments with poultrymanure.
5.2.5. ArsenicIn the case of arsenic, the exact process responsible for its initial
removal is not clear but adsorption or concomitant co-precipita-tion with other metal sulphides or with ferrihydrite has been sug-gested (Jong and Parry, 2003; Zaluski et al., 2003; Cheng et al.,2009; Corkhill and Vaughan, 2009). Tsukamoto and Miller (1999)reported efficient removal of arsenic along with other divalentmetals (Fe, Ni) for a period of almost 2 years in a pilot-scale biore-actor. Formation of insoluble arsenic sulphide may occur laterwhen reducing conditions are established. Proportions of As(III)and As(V) species in the ARD were suggested as a critical factoraffecting the rate of arsenic reduction in different environments(Jong and Parry, 2003; Corkhill and Vaughan, 2009). Cohen(2006) reported that oxyanions such as chromate and arsenatecan be removed using passive bioreactor technology. Arsenic is re-moved as an arsenic sulphide compound and chromate is reducedto Cr(III) and precipitated as a hydroxide.
5.2.6. UraniumTrace levels of soluble U found in waste streams are typically
found in the form of the mobile uranyl tricarbonate or uranylsulphate complex. When these waste streams encounter thereducing conditions of a SRBR, the U changes oxidation state andcan form a variety of reduced U species in the tetravalent formsuch as the insoluble UO2. The presence of organic matter in theSRBR provides the reducing environment for this precipitation tooccur (Machemer et al., 1993).
6. Organic substrate
The substrate in bioreactor usually serves two purposes: to pro-vide a carbon source and to maintain flow through the system.Quite a few organic materials can be used as the organic carbonsource for SRP bioreactors as indicated in literature. This includesthe list of references regarding substrate mixture components usedin SRBR treatment systems and their effectiveness. More than 90publications that dealt with the use of organic substrates as mix-tures for SRP-mediated treatment of ARD have been identified.These publications identified 36 organic substrates that included7 direct and 29 indirect substrates (Zaluski et al., 2006).
The direct/simple organic substrates are those that do not re-quire decomposition by other microorganisms to provide SRPnutrition. Such substrates include: alcohols (e.g., methanol andethanol); organic acids (e.g., acetate, lactate, formate, and pyru-vate) and sugars (e.g., sucrose).
Sulphate-reducing prokaryotes use the easily degradable frac-tion of organic matter such as low molecular weight compoundswith simple structures (e.g., methanol, ethanol, and lactate) (Dvo-rak et al., 1992; Nagpal et al., 2000a; Tsukamoto et al., 2004), poly-lactic acid (Edenborn, 2004), simple carbohydrate monomers (e.g.,glucose or sucrose) (Mizuno et al., 1998), and whey (Christensenet al., 1996). In terms of energy and biomass produced, lactate isa superior electron donor compared to others such as ethanol,acetic acid, propionate, and acetate (Nagpal et al., 2000b). In termsof moles of bicarbonate produced per mole of substrate consumed,the lactate-utilizing processes are superior to ethanol-utilizingprocesses (3 vs. 2, respectively) since they are better at neutraliz-ing the acidity in the treated effluent (Kaksonen et al., 2004a).The main drawback is that only certain species of SRPs (Desulfoto-maculum) 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 acetatealone (Desulfotomaculum acetoxidans) (Nagpal et al., 2000b).
Volatile fatty acids (acetate, propionate, butyrate) and shortchain fatty acids (lactate, pyruvate, and malate) are among themain source of SRP. Carbon and long chain fatty acid and certainaromatic compounds are occasional substrates (Postgate, 1984;Beaulieu et al., 2000). Fermentative products such as methanol,ethanol, and acetate are additional sources and polymers such ascellulose is not degraded by known SRP (Béchard et al., 1994;Chang et al., 2000). The polysaccharide must be degraded byhydrolytic fermentative anaerobes into sulphidogen-supportingfatty acids and (Johnson and Hallberg, 2005a); once these sub-stances are depleted, hydrolytic fermentation is the rate limitingbecause the products of fermentation are completely used up(Chang et al., 2000). Similarly, proteins, carbohydrates and lipidsor even simple sugars are generally not accessible to SRP (Tuttleet al., 1969a,b) and depends on other heterotrophic bacteria to sup-ply them with degradation and fermentation end products (Prasadet al., 1999).
Drury (1999) used lactate and methanol as carbon source inSRBR and observed that lactate broken down only to acetate andminimal methanol utilization. According to Drury (1999) lactateis a known energy source for a variety of SRP and was a logicalchoice as a positive control. Methanol as a substrate does notfreeze at a remote high elevation site during winter months thushas been reported to be utilized by several species of Desulfovibrio,Desulfotomaculum and Desulfobacterium although its metabolism isless common among SRP.
Tsukamoto and Miller (1999) reported that methanol andlactate contribute 6 and 12 electrons respectively per moleculeoxidized, assuming complete oxidation to carbon dioxide.
Electron accounting in this manner will allow a determinationof the number of the moles of carbon source needed to reduceone mole of sulphate, and the efficiency of the utilization of thereducing equivalents.
Edenborn (2004) investigated the ability of poly lactic acid(PLA) to serve as a long term source of lactic acid for bacterial sul-phate reduction activity in zinc smelter tailings. As PLA becomes
more common and disposable waste products generated by thisindustry become more abundant, the applied use of these materi-als to facilitate environmental remediation efforts is feasible. SolidPLA polymers mixed in water hydrolyzed abiotically to release lac-tic acid into solution over an extended period of time. The additionof both PLA and gypsum was required for indigenous bacteria tolower redox potential, raise pH and stimulate sulphate reductionactivity in highly oxidized smelter tailings after 1 year of treat-ment. Bio-available cadmium, copper, lead, nickel and zinc wereall lowered significantly in PLA (gypsum treated soil), but PLAamendments alone increased the bioavailability of lead, nickeland zinc. Similar PLA amendments may be useful in constructedwetlands and reactive barrier walls for the passive treatment ofmine drainage, where enhanced rates of bacterial sulphate reduc-tion are desirable.
Column and batch studies showed that cellulose hydrolysis wasa rate-limiting factor in hydrogen sulphide production by SRPwhen more liable carbon sources were absent (Logan et al.,2005). Thus, if substrates are comprised largely of cellulose, whichmay be the case after long periods of treatment, cellulolytic bacte-ria exert a significant impact on SRP activity. Since cellulolysis ismost effective at a pH of 6.0 or above, additional buffering maybe required to utilize this substrate (Logan et al., 2005).
6.2. Indirect/complex organic substrate
Indirect/complex organic substrates are those requiring decom-position by other microorganisms to provide SRPs nutrition. Thesesubstrates require complex microbial communities to degrade theorganic matter and support SRP growth. Quite a variety of suchsubstrates has been reported and they can be classified as: com-posts (e.g., spent mushroom compost, leaves); wood/paper wastes(e.g., sawdust, leaf mulch, wood chips); food production byprod-ucts (e.g., molasses, cheese whey, potato processing waste); Agri-cultural products (e.g., hay, straw); manure (e.g., cow, horse,dried poultry waste); and sewage waste (e.g., digested sludge, sew-age sludge).
Release of organic acid from substrates for a long time is veryimportant for sustaining SRP. In other words, proper selectionand quantities of substrates are related to sustainability of sul-phate-reducing prokaryotes (Lim, 1997). Less expensive organiccarbon sources such as waste material from the agricultural andfood processing industries have been assessed for their potentialto sustain sulphate reduction. The alternative organic carbonsources may be selected between two groups of materials—cellu-losic wastes and organic wastes (Kuyucak and St-Germain, 1994).Generally, cellulosic wastes include sawdust, hay, alfalfa, and woodchips, 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.
There is a general consensus that these substrates alone do notsignificantly promote the activity of SRP (Christensen et al., 1996;Gibert et al., 2003). Higher sulphate-reduction rates have beenobtained with reactive mixtures containing more than one organiccarbon source (Waybrant et al., 1998, 2002; Cocos et al., 2002;Zagury et al., 2006). Generally, these mixtures contain relativelybio-degradable sources (poultry manure, cow manure, or sludge)and more recalcitrant ones (sawdust, hay, alfalfa, or wood chips).
Hedin et al. (1988), Wildeman et al. (1994), Gusek et al. (1998)reported that many systems have been constructed using compostor other organic wastes to generate an anaerobic environment andprovide a source of organic carbon. Complex organics present inthe substrate are microbiologically degraded to simpler organics,which are utilized by the SRP. Although wetland plants are some-times present, many systems have been built without them.
Cheong et al. (1998) recommended a mixture of rice stalk, cowmanure and limestone. Hard et al. (1997) observed that the carbonsource can be any type of carbon material (e.g., sawdust, biosolids,manure), the decaying roots/detritus of the wetland plants, anadded soluble carbon-based liquid material (e.g., methanol), or alayer of carbonaceous material such as municipal compost or bios-olids. The organic layers in an anaerobic biofilter not only serve ascarbon sources, they also physically retain metal sulphides (Hardand Higgins, 2003).
Waybrant et al. (1998) concluded that after the acclimation per-iod (20–65 days), sulphate reduction rate was higher in the reac-tive mixture that contained a variety of organic carbon sources.Further it was observed that the cellulose entailed a slightly lowersulphate reduction rate compared to other substrates tested (sew-age sludge, leaf mulch, wood chips, sheep manure, and sawdust)alone or in mixture. Cellulosic material alone could sustain satis-factory bacterial activity in a 125-day column test.
Chang et al. (2000) observed performance at later stages ofexperiments (after 20 weeks) in bioreactors using several sourcesof waste materials containing cellulose. The cellulose was the maincomponent used during 35 weeks of operation. Spent mushroomcompost (SMC) proved to be the best substrate due to nutrient con-tent, slow degradation and its bulk form but showed no sulphatereduction for ten weeks when no inoculum was used. The effi-ciency of cellulosic substrates for the biological treatment of ARDhas been confirmed by several studies (Tuttle et al., 1969a,1969b; Waybrant et al., 1998; Chang et al., 2000; Tassé andGermain, 2002; Johnson and Hallberg, 2005b), while other studiessuggested that cellulosic wastes alone entailed carbon-limitingconditions (Béchard et al., 1994) or did not sustain SRP growth(Kuyucak and St-Germain, 1994). With sawdust as the sole nutri-ent source, a mixed bacterial culture containing cellulose-degrad-ing bacteria and SRP was capable of reducing sulphate at pH 3.0,whereas pure cultures of SRP did not reduce sulphate below pH5.5 (Tuttle et al., 1969b). These results stress the importance of awell established microflora in the presence of mixtures of cellu-losic and other complex natural organic carbon sources.
Cohen and Staub (1992) used composted cow manure and hayin a 4–1 by volume ratio. The hay served as a bulking agent to helpmaintain hydraulic conductivity. The manure pH was 8–9 and ithad high buffering capacity. The hay has been demonstrated to en-hance SRP activity by 250–700%. The composted manure can neu-tralize the pH of the mine drainage. The organic matter in themanure decomposes to forms available to SRP and low redoxpotentials are generated making the conditions ideal for SRP. Thus,the SRP can become established and can then modify their ownmicroenvironments. Eger and Wagner (2003) used severalsubstrates in the study; the results from a 45-day old municipalsolid waste compost were impressive and prove to be one of themost reactive substrates for over a period of 10 years.
Tassé and Germain (2002) reported sulphate reduction withconifer barks, especially when mixed with leaf bearing tree barks.Indeed, a mixture of these conifer substrates and poultry manureand/or leaf compost proved to be effective. A novel mixture of var-ious substrates has been proposed by Cocos et al. (2002) in a41 day study using a multiple factor design and the modelling ofthe sulphate reduction rate. Seventeen mixtures of maple woodchips, leaf compost and poultry manure (and inorganic porous sup-ports) was assessed in batch experiments. The best mixture interms of sulphate reduction was composed of 3% wood chips,30% leaf compost, 20% poultry manure (maximum proportion usedin study), and 5% silica sand (wt.%). The components at constantproportions were 37% bacterial source (Creek sediment), 2%calcium carbonates and 3% urea. In this short-term study it wasreported that there was limited degradability of lignin-cellulosicsubstrates in a 41-d batch test. A higher proportion of poultry
manure was essential for promoting higher sulphate-reductionrates.
Amos and Younger (2003) reported that the selection of thereactive media to be used is of paramount importance, with partic-ular reference to permeability and reactivity. A number of reactivemedia mixture containing varying proportions of cattle slurryscreenings, green waste compost, calcite limestone clips and peagravel were prepared and their respective permeability and reac-tivity were investigated. Media mixture containing 50% 10 mmgrade mixture limestone chips showed better alkalinity additionand metals removal than a blank containing 50% pea gravel. Amedia mixture containing 50% limestone chips and 50% greenwaste compost showed a 24 h period to achieve maximum addi-tion of alkalinity and maximum removal of acidity and metals.Mixtures containing 25% green waste compost and 25% slurryscreenings achieved maximum removal in 4 h.
Tsukamoto et al. (2004) selected the matrix consisting sand andhorse manure, which was approx. 5–10 years old and had under-gone appreciable decomposition. Gibert et al. (2004) assessed thedegradability of compost, Sheep and poultry manures and oak leaf,concluding that the sheep manure was the best substrate for SRP,followed by poultry manure and oak leaf. It was reported that low-er the lignin content of substrate, the greater the degradability andthat municipal compost was too low in organic carbon to supportSRP growth. Unfortunately, the efficacy of the mixtures of variousnatural organic substrates was not assessed. Further, no metalswere used in the study and thus the sulphate reduction processwas not true ARD treatment, not having taken into account the po-tential metal toxicity to SRP.
Zagury et al. (2006) evaluated the performance of various com-positions of substrates in sulphate reduction and metal removalARD in a 70 days batch experiment. Maple wood chips, sphagnumpeat moss, leaf compost, conifer compost, poultry manure andconifer sawdust were investigated in terms of their carbon (TOC,TLC, DOC), and nitrogen (TKN) content, as well as their easily avail-able substances content (EAS). It was reported that Sphagnum peatmoss is not recommended for large scale operations because of itslow density. Conifer sawdust and composted spruce chips havehigher complex organic C and lower nitrogen and showed no mea-surable sulphate reduction. Zagury et al. (2006) also reported thatC/N ratio alone is not a good indicator of the potential carbondegradability by the sulphate-reducing micro flora.
Neculita et al. (2007) obtained higher sulphate-reduction rateswith reactive mixtures containing more than one organic carbonsource. Generally, these mixtures contain relatively bio-degradablesources (poultry manure, cow manure or sludge) and more recalci-trant ones (sawdust, hay, alfalfa or wood chips). In preparation ofthe substrate mixture to be used for biological passive treatmentof ARD, it is essential to assess its potential effectiveness by know-ing its chemical make up, notably in terms of organic carboncontent.
In fact, the comparison of different studies dealing with thesame substrate or different organic substrates is very difficult be-cause of different durations for each study. For example, studieshave been performed over 14 day (Jong and Parry, 2003), 70 day(Tassé and Germain, 2002; Zagury et al., 2006), 23 months (Drury,1999), or 32 months (Zaluski et al., 2003). In very short-termexperiments the aging of the material and the clogging of the ma-trix are not addressed (Jong and Parry, 2003; Zagury et al., 2006).Higher proportions of coniferous bark and/or sawdust have beenassociated with sluggish sulphate-reduction rates in short-timeexperiments (Tassé and Germain, 2002), whereas a mixture con-taining a high content of sawdust (40% sawdust, 10% wood chips,10% alfalfa hay, 10% cow manure, 29% limestone, and 1% cementkiln dust) gave the best efficiency in a long-term field study(Reisman et al., 2003).
Fig. 3. Anaerobic pathway of microbial processes of sulphate-reduction in an organic carbon substrate (modified after Logan et al., 2005).
Furthermore, contradictory conclusions emerge from studiesperformed with the same organic carbon source but using differentproportions in the reactive mixture the results proved to beencouraging. In an attempt to find the best reactive mixture foruse in permeable reactive walls, Waybrant et al. (1998) concludedthat sheep manure (100%) did not produce the reducing conditionsnecessary for bacterial activity and excluded this organic sourcefrom their batch assays. In contrast, Gibert et al. (2004) clearlyindicated sheep manure (15% of reactive mixture) as the most suc-cessful organic material for creating reducing conditions and sus-taining active sulphidogenesis (sulphate removal >99%) in abatch experiment. Similarly, in the experimental study of Amosand Younger (2003), cattle manure (100%) was rejected at an earlystage due to low permeability, whereas a mixture of cow manure(80%) and cut straw (20%) was successfully used over 32 monthsby Zaluski et al. (2003).
McCauley et al. (2008) concluded after carrying out laboratorystudies that sulphate-reducing bioreactors containing musselshells and forestry waste products offer a promising technologyfor mitigating acidity and sequestering metals associated withARD at Stockton Mine in New Zealand. Mussel shells tend to dis-solve more readily than limestone in the presence of ARD as indi-cated by alkalinity generation in the SRBRs evaluated in theirstudy. Possible contributing factors include grain size, shape, reac-tive surface area, unique mineralogy (aragonite and calcite) ormineralogical dynamics and consequent structural change whendissolved. Therefore, SRBRs containing mussel shells and a diver-sity of carbon sources exhibit more efficient ARD treatment thansystems utilizing limestone as the sole alkalinity source. Addition-ally, labile carbon attached to the mussel shells and nitrogen with-in the mussel shell matrix may potentially benefit the consortiumof microorganisms which develop as systems reach stable treat-ment conditions (after 5 weeks in this study). Forestry waste prod-
ucts including Pinus radiata bark, post peel and composted woodprovide sustainable short and long-term carbon sources for micro-organisms. Substrate mixtures used in their study was consideredlow risk for plugging but is of potential concern if used on a long-term basis in a SRBR.
6.3. Microbial processes of sulphate-reduction in an organic carbonsubstrate
Sulphate-reducing prokaryotes are obligate anaerobes that useonly a limited range of organic substrates which are simple organiccompounds including low molecular weight fatty acids, alcohols;easily degradable like methanol, ethanol, lactate, poly acetic acid,simple carbohydrate monomers and whey (Dvorak et al., 1992;Nagpal et al., 2000a; Edenborn, 2004; Liamleam and Annachhatre,2007; Neculita et al., 2007; Neculita et al., 2008).
The pathway of anaerobic degradation of organic matterinvolves the hydrolysis of the large molecular weight compoundssuch as proteins, nucleic acids, carbohydrates and lipids to lowermolecular weight products such as organic acids and alcohols(Gibson, 1990; Seyler et al., 2003; Coetser et al., 2006;) as givenin Eqs. (19)–(27) and Fig. 3.
Reactions of hydrolysis:
ðC6H10O5Þnþ nH2O! nC6H12O6 ð19Þ
These may be fermented into volatile fatty acids and gasses
Among the methanogens two main groups are distinguished;one group (acetoclastic ethanogens) uses acetate as a substrate
CH3COOH! CO2ðgÞ þ CH4ðgÞ ð22Þ
The group (hydrogenotrophic methanogens) uses H2(g) as anelectron donor for the reduction of CO2.
Methanogenesis:
CO2 þ 4H2 ! CH4 þ 2H2O ð23Þ
SRBs further uses mainly H2(g) or lactate as substrates thus pre-cipitating metals as metal precipitates
4H2ðgÞ þ SO2�4 þ 2Hþ ! H2SðgÞ þ 4H2O ð24Þ
Terminal oxidative processes are than able to degrade thesesubstrates further
Sulphate reduction: (Kosolapov et al., 2004):
2Hþ þ SO2�4 þ 2\CH2O"! H2Sþ 2H2CO3 ð25Þ
Metal reduction (iron):
Fe3þ þ e� ! Fe2þ ð26Þ
Metal reaction can be also expressed (Lintern, 1994) as
M2þ þH2S!MS # þ2Hþ " ð27Þ
Eq. (27) is possible because H2S is a very strong reducing agent.Sulphate reducers rely on microorganisms SRP that can hydro-
lyze solid phase organic material to produce soluble organic com-pounds that are further degraded to appropriate substrates forsoluble reducers (Figueroa et al., 2004). The sustainability of SRBRsis dependent on the composition of the substrate, i.e. availability ofcarbon sources. The major pools of carbon that make up organicmatter include lipids, protein, simple sugars, organic acids, cellu-lose/hemicelluloses and lignin. Simple sugars, lipids, protein, or-ganic acids and cellulose/hemicelluloses are acid soluble, thesecomponents are readily available to the SRBs. These componentsmust be hydrolyzed before they are available to the microbial com-munity. Hydrolysis is a rate-limiting step in the anaerobic environ-ment supported by solid phase organic matter.
Sulphate reduction seems to be controlled by cellulose degrada-tion and therefore, future research for exploring means by which toenhance cellulose hydrolysis is needed. More work must be con-ducted to understand and differentiate the fundamental biochem-ical and microbiological reactions that occur in anaerobicbioreactors with complex organic substrates. This might be a keystep for the successful implementation of SRP-based ARD remedi-ation systems. Today a routine and rigorous method of analysis oforganic waste materials is still warranted to predict organic sub-strate biodegradability (Gibert et al., 2004; Zagury et al., 2006).
6.4. Selection of suitable substrate
Careful selection of suitable carbon source is of paramountimportance to ensure performance and longevity in ARD treatment(Tuttle et al. 1969b; Dvorak et al., 1992; Hedin et al., 1994). Acid-rock drainage generally contains relatively low concentrations ofdissolved organic carbon (<10 mg/L) (Kolmert and Johnson, 2001).Therefore, the most critical limiting factor for the microbial activityis the availability of carbon from an additional organic source(Gibert et al., 2004; Zagury et al., 2006; Neculita et al., 2007). Releaseof organic acid from substrates for a long time is very importantfor sustaining SRP. In other words, proper selection and quantitiesof substrates are related to sustainability of the SRBR.
Various waste materials could support bacterial sulphate reduc-tion in compost bioreactor to treat ARD. There is growing evidencethat sulphate reduction efficacy can be augmented by the use ofnatural organic substrate mixtures versus single substrate(Waybrant et al., 1998; Cocos et al., 2002; Amos and Younger,2003). The challenge for having an efficient on-site bioreactor isto select a suitable organic substrate to make the process efficientand economically feasible. Selection of the organic carbon source isusually made on the basis of availability and costs of the addedelectron donor per unit of reduced sulphate. The remainingcontaminants in the treated water must be present in low concen-trations or easy to remove (Hulshoff Pol et al., 2001).
Gavaskar et al. (1998) give a key element to the design of SRBRfor successful treatment of contaminated ground water is theselection of the reactive media. Identified essential propertiesrequired for the reactive media are: reactivity, permeability orhydraulic conductivity, environmental compatibility and stability.First two characteristics are of principle interest. They will havethe greatest effect on the success or failure of the treatmentsystem. The later three characteristics should not; however bediscarded (Amos and Younger, 2003).
The use of direct substrates promises to allow more stringentcontrol of biofueling but requires more complicated reactor designand may not be suitable for remote mine sites. The use of somedirect substrates, such as ethanol, at remote mine sites is alsocomplicated by public safety concerns. Indirect substrates aremore feasible than direct substrates for low maintenance systemsat remote mine sites requiring long-term operation.
The choice of an effective substrate mixture is dependent on thecomposition of the ARD and the types of substrates available at lowcost. Overall, substrate mixture containing both easily bio-degrad-able materials and slow bio-degradable (recalcitrant) materials arethe most effective for supporting sustained SRP activity. The easilybio-degradable substrate ensures a quick start of a bioreactor.More recalcitrant materials provide the best long-term bioreactorperformance. The substrate mixture should also provide adequatesurface area for bio-film development, buffering and adsorptioncapacity, and adequate hydraulic conductivity. The suitability ofa substrate mixture for treating a particular composition of ARDis best determined empirically using laboratory-scale tests. Over-all, the through study indicates that a wide range of organic sub-strate materials can be used to effectively treat ARD using SRPtechnology.
7. Factors affecting the performance of SRP
Various parameters such as pH, temperature, sulphide and me-tal concentrations in the ARD undergoing active biological treat-ment will affect the growth and activity of SRP. Apart from thesefactors it is a well known fact that SRP are strict anaerobes (Thaueret al., 2007). The effects of these parameters on SRP activity are dis-cussed in the following sections.
7.1. Effect of pH
Sulphate-reducing prokaryote strains identified are sensitive toacidic waters (Hard et al., 1997). Pure culture of SRP was isolatedfrom a pond with a pH of 3.38 (Tuttle et al., 1969a) and from anabandoned mine, Kam Kotia, at slightly oxidising and acidic condi-tions (Fortin et al., 1996). But when attempted to grow these SRP inlaboratory, there was no sufficient growth below pH 5.5 in both thecases. Based on the results it was concluded that SRP isolated fromoxidising conditions make the environment favourable by reducingthe sulphate to sulphide which will provide an alkaline environ-ment suitable for their activity (Fortin et al., 1996). Johnson et al.
(1993) reported that a species belonging to the genera Desulfoto-maculum can grow in an environment with a pH of 2.9. Later theytermed it as acid tolerant bacteria rather than acidophilic bacteria.
Kolmert and Johnson (2001) reported that a mixed acidophilicSRP culture was able to grow in a medium with a pH of 3.0 andthe result supports the view by Postgate (1984) that mixed SRPcultures are more tolerant to extreme conditions than pure cul-tures. Elliott et al. (1998) investigated the effect of acidic condi-tions on SRP species at pH values of 4.5, 4.0, 3.5, 3.25 and 3.0 ina porous up-flow bioreactor. SRP removed 38.3% of sulphate atpH of 3.25 and 14.4% at pH of 3.0. The reactor was operated at aflow rate of 0.6 ml/min with feed sulphate concentration of 1.0 g/L.
In order for SRP to thrive, they require a pH in the range of 5–8(Willow and Cohen, 2003). Outside this range, the rate of microbialsulphate reduction generally declines and the metal removalcapacity is reduced. Low pH (<5) normally inhibits sulphate reduc-tion and increases the solubility of metal sulphides (Dvorak et al.,1992). In any case, the presence of SRP has been detected in naturalwaters with pH < 3 (Gyure et al., 1990; Kolmert and Johnson, 2001;Koschorreck et al., 2003, Koschorreck and Tittel, 2007). Isolatedstrains were mostly acidophilic SRP, which are more efficient thanthe neutrophilic ones for remediating acidic waste waters (Kolmertand Johnson, 2001).
Elliot et al. (1998) observed that SRP were capable of sulphatereduction in a column bioreactor operated under acidic conditionswith lactate as organic carbon source. At pH 3.25, 38.3% of influentsulphate was removed and pH of the medium rose to 5.82, whereasat pH 3.0 sulphate removal fell to 14.4% and sulphide productiondropped below detection. Nevertheless, viable SRBs were recov-ered from the bioreactor after operation at pH 3.0 for 3 weeks.However, the existence of truly acidophilic SRP is currently notclear. Reduction of sulphate to sulphide has been demonstratedas occurring in extremely acidic environments, but attempts at iso-lating pure culture of acidophilic (acid-tolerant) SRP failed (Gyureet al., 1990; Johnson, 1998). Nevertheless, a pH of 5.5 or higher ispreferred for efficient treatment of ARD in an on-site passive biore-actor (URS Report, 2003).
Willow and Cohen (2003) have shown that pH is critical to reac-tor efficiency. The metal loading rate capacity of a wet-substratebioreactor can be enhanced and hydraulic detention times re-duced, by modifying the pH of the influent to near neutral. Theneutral pH permits enhanced activity of SRP, the production of sul-phide, and the removal of metals as metal sulphide precipitates.Further they reported that the rapid influx of acidic, aerobic watersappears to drive the pH of the treatment system down and redoxup, thus inhibiting microbial sulphate-reducing processes. The me-tal removal efficiency and loading capacity of the treatment systemthen becomes a function of not only size and hydraulic conductiv-ity, but also the acidity and oxygen content of the influent water.
Jong and Parry (2005) found that SRP in laboratory scale bio-reactors sustained sulphate-reduction rates of 553–1052 mmol/m3/day when the pH was lowered from 6.0 to 4.0. However,when the pH was lowered to 3.5, this rate dropped to3.35 mmol/m3/day. Similarly, SRP in ethanol-fed columns sur-vived at the lowest pH tested (2.5), but were less effective ingenerating alkalinity below pH 3.0 (Tsukamoto et al., 2004).Acid-tolerant strains of SRP have been characterised and isolated,and their introduction to ARD treatment systems may improveperformance (Johnson and Hallberg, 2005a). However, a higherpH may be required for effective metal precipitation and organiccarbon degradation.
Cohen (2006) observed from experiments that pH is more crit-ical to reactor efficiency than dissolved oxygen. The metal loadingrate capacity of a wet-substrate bioreactor can be enhanced, andhydraulic detention times reduced to as low as 16 h by modifyingthe pH of the influent to near neutral . The neutral pH permits en-
hanced activity of SRBs, the production of sulphide, and the re-moval of metals as metal sulphide precipitates. The dissolvedoxygen was completely removed in the first few centimetres uponentering the reactors, while the pH required a greater proportion ofthe substrate in order to reach suitable levels for SRP activity. Asthe pH of the influent increased, the rate of sulphate reduction in-creased, increasing the metal removal capacity of the system.
7.2. Redox potential (Eh)
Garcia et al. (2001) reported that negative redox potential pro-vides suitable environment to grow bacteria appropriately. Thebacteria generate an appreciable negative potential, even startingfrom the potential of an oxidizing medium. Therefore, the chemi-cal and microbiological data of the water indicated that the SRPwere active. Decreasing redox potential, formation of black precip-itate in the medium and the presence of hydrogen sulphide can bedetected in the aqueous phase with the classical strong smell(Christensen et al., 1996). For optimal performance, SRP need ananaerobic medium and an anoxic and reduced microenvironmentwith a redox potential (Eh) lower than �100 mV (Postgate, 1984).
Sulphate reduction was often observed in passive field-bioreac-tors at positive Eh values (Reisman et al., 2003; Zaluski et al., 2003;Neculita et al., 2007). Eh measurements of aqueous samples col-lected at the outlets of the bioreactors might not reflect the realvalues present in pockets of organic matter where SRP live (Zaluskiet al., 2003). Their survival in these adverse conditions may also beexplained by the formation of favourable anoxic microenviron-ments in the reactive mixtures (Lyew and Sheppard, 1999). Batchand column laboratory bioreactors successfully treated ARD at Ehvalues of –100 to –200 mV or lower during 23 days (Cocos et al.,2002), 30 days (Beaulieu et al., 2000), or 150 days (Gibert et al.,2004) retention period. In passive field-bioreactors, Eh values aslow as –200 mV were maintained for periods ranging from2 month to more than 2 years (Cheong et al., 1998; Reisingeret al., 2000).
7.3. Effect of temperature on sulphate-reducing prokaryotes
The SRP can be classified into mesophiles (growth temperature<40 �C), moderate thermophiles (growth temperature: 40–60 �C)and extreme thermophiles (growth temperature >60 �C) based ontheir optimum growth temperature. Van Houten et al. (1997) re-ported that sulphate reduction increases when the reaction tem-perature is increased from 20 to 32 �C by employing a mesophilicSRP culture. Moosa et al. (2002) employed a mixed culture consist-ing of acid producers, methane producers and sulphate reducersand conducted batch experiments. It was reported that sulphatereduction rate increased with increasing the reaction temperaturefrom 20 to 35 �C. Further increase of temperature to 40 �C led toinactivity of bacteria. Stetter et al. (1993) isolated a number of ther-mophilic SRP strains from the Thistle reservoir. Some sulphatereducers such as Desulfotomaculum species are endospore formers,and are considered to survive in extreme environments (Widdel,1988).
In passive bioreactors, the operating temperature affects bacte-rial growth, kinetics of organic substrate decomposition, as well ashydrogen sulphide solubility. Generally, SRP can tolerate tempera-tures from below �5 to 75 �C (Postgate, 1984). Temperature signif-icantly affects the rate of microbiological processes. At lowtemperature level (<10 �C) the reaction rate decrease more than50% of that which could be obtain at 20 �C. The average sulphatereduction or (H2S generation) rate has found to be 0.3 mol/m3 ofthe nutrient per day at temperature levels higher than 10 �C inthe test conducted where agriculture waste was used as nutrient(Kuyucak and St-Germain, 1994; Neculita et al., 2007).
The methanogens, which are found when bioreactors are sup-plied with complex organic carbon sources, are mainly mesophilicmicroorganisms. Therefore, they are more sensitive to low temper-atures than SRP (Kuyucak and St-Germain, 1994). Low tempera-tures particularly affect the ability of SRP to acclimate, but onceacclimated at higher temperature, SRP are not affected by thanlow temperature. Column experiments showed that the efficiencyof ARD treatment was not significantly reduced at temperatures aslow as 6 �C (Tsukamoto et al., 2004). Cold-adapted species are ableto function at temperatures as low as 4 �C, and increased popula-tions may offset lower activity (Higgins et al., 2003).
Zaluski et al. (2003) operated passive on-site bioreactors suc-cessfully for 32 months at temperatures between 2 �C and 16 �Cor over 2 years at near-freezing temperatures (1–8 �C). In field-bio-reactors started during the winter, a 4-months lag phase was ob-served for SRP to be established. However, winter freezing of awell established SRP population had little or no effect on theiractivity. Drury (2000) reported that temperature significantly af-fects sulphate reduction. The required hydraulic retention timefor 50% sulphate reduction varies from 8 days at 17 �C to 41 daysat 1 �C.
7.4. Effect of solid support (solid surface for the SRP to grow)
Sulphate-reducing prokaryotes require a solid support (sandand/or gravel), onto which they can establish microenvironmentsfor their survival in the presence of extreme conditions such aslow pH or high oxygen concentrations (Lyew and Sheppard,1997). Higher sulphate-reduction rates are achieved if SRP haveaccess to a porous surface, compared to suspended bacteria(Glombitza, 2001). A medium with large pore spaces, low surfacearea, and a large void volume is generally preferred because it min-imizes the plugging of the bioreactor. In terms of efficiency, bettertreatment occurs with greater surface area. Surface area and poresize need to be balanced in field-bioreactors (Tsukamoto et al.,2004; Neculita et al., 2008).
7.5. Hydraulic retention time (HRT)
The efficiency of bioreactors is also affected by the hydraulicretention time (HRT) and has been widely reported (Dvoraket al., 1992; Béchard et al., 1994; Rockhold et al., 2002; Kaksonenet al., 2004b). The variability of hydraulic properties of porousmedia used in reactive mixtures may result in HRTs specific to eachbioreactor. It is usually accepted that precipitation of metalsulphides occurs in at least 3–5 days of HRT (URS Report, 2003;Kuyucak, 2002; Kuyucak et al., 2006). A shorter HRT may not allowadequate time for SRP activity to neutralize acidity and precipitatemetals or may result in biomass being washed out of the bioreac-tor. A longer HRT may imply depletion of either the available or-ganic matter source or the sulphate source for SRP (Dvorak et al.,1992). In a semi-continuous anaerobic laboratory bioreactor, moresulphates were reduced to sulphides with a 3-d HRT compared to a1-d HRT, regardless of the organic carbon/sulphate ratio (Gibertet al., 2004) emphasized the importance of residence time by Col-umn experiments, considering the residence time as a key factor inthe performance of continuous systems. With a residence time of0.73 days, sheep manure did not promote sulphidogenesis. How-ever, extending residence time to 2.4 and 9.0 days resulted in anincrease in the sulphate removal to 18% and 27%, respectively(Neculita et al., 2007).
7.6. Hydraulic conductivity (permeability)
The hydraulic conductivity of the substrate material is also animportant variable because this will affect the HRT (Bolis et al.,
1992; Benner et al., 2001, 2002). Several studies were conductedto evaluate the effect of microbial growth and biomass accumula-tion on porosity and permeability of saturated porous media.During treatment, bacteria induce changes in the mixture proper-ties due to accumulation of biomass and generation of metabolicbyproducts. The characteristics of the accumulated biomass aredependent on the type of bacteria, the substrate and loadingrate, and the flow rate (Rockhold et al., 2002). Bacterial activitymight cause a decreased surface tension, decreased porosity andpermeability, and pore clogging. A sand-packed column reactorusing sewage bacteria and methanol as a substrate showed adecrease in the saturated hydraulic conductivity (Ks) of threeorders of magnitude following biomass accumulation (Taylor andJaffe, 1990). A relatively small variation in hydraulic conductivitycould entail important differences in residence times, and mightresult in decreased efficiency (Benner et al., 2002). Efficient com-post substrate passive bioreactors have a hydraulic conductivityabout 1 � 10–4 cm/s (URS Report, 2003). Recently, sawdust has beenincreasingly used in reactive mixtures, due to a significantly higherconductivity of around 10�2 to 10�3 cm/s. When sawdust is used,however, there is an increased vulnerability for mixture compac-tion. Pre-soaking the substrate before ARD treatment can helpprovide a more stable hydraulic conductivity and a more consis-tent flow rate through the system (Bolis et al., 1992). Compostedsubstrates should be at least 0.6 m in thickness but should notexceed 0.9–1.2 m; if not, the substrate tends to compact withdepth and the permeability becomes too low for effective treat-ment (URS Report, 2003; Neculita et al., 2007).
Waybrant et al. (1998) reported that in bio-reactive walls, aportion of the existing aquifer was excavated and the originalmaterial was replaced with an organic substrate. The porosity ofthe wall is an important factor; the wall should be sufficient per-meable (e.g., 10�3 cm/s to allow water flow through).
7.7. Type of reactors employed for anaerobic sulphate reduction
A variety of reactor configurations such as up-flow anaerobicsludge bed reactors (Colleran et al., 1994; Sanchez et al., 1997),stirred tank reactors (Herrera et al., 1997; Moosa et al., 2002;2005) packed bed reactors (Chang et al., 2000; Jong and Parry,2003) and membrane (Chuichulcherm et al., 2001) reactors havebeen used to study anaerobic reduction of sulphate and to treatARD (Fig. 4). Biological sulphate reduction can be achieved withfreely suspended bacterial cells or immobilized cells. Applicationof freely suspended cells in continuous bioreactors dictates a highresidence time to prevent cell washout. In other words a continu-ous reactor with freely suspended cells has to be operated at lowflow rate and high residence time. In an immobilized cell bioreac-tor the biomass residence time becomes uncoupled from thehydraulic residence time; therefore it is possible to operate thereactor at high flow rate without cell washout. The bio-film formedin the immobilised cell bioreactors also offers more resistance toextreme conditions such as low pH, high metal concentrations etc.
Chen et al. (1994) studied the kinetics and stoichiometry of sul-phide formation in a packed-bed bioreactor using sea sand as car-rier matrix. Lactate was used as a carbon source and the SRPspecies Desulfovibrio desulfuricans was used as an inoculum. Atthe volumetric loading rate of 0.138 g/L/h the maximum volumet-ric reduction rate achieved was 0.015 g/L/h. Waybrant et al. (2002)investigated the effect of packing reactive mixtures which werebasically waste products. Two up-flow packed-bed bioreactor con-taining two different reactive mixtures were used: first one con-taining leaf mulch, sawdust, sewage sludge, and wood chips andthe second containing leaf mulch and sawdust. The maximum vol-umetric reduction rates achieved in the first and second columnswere 0.003 and 0.005 g/L/h, respectively.
Fig. 4. Configurations of various reactors. (after Kaksonen and Puhakka, 2007).
Elliott et al. (1998) conducted experiments in a packed-bed bio-reactor to investigate the effect of pH on the anaerobic sulphatereduction. The column was packed with sand and the pore volumewas 783 ml. In this study Postgate (1979) Medium B without ironsulphate was pumped through the column at a rate of 0.6 ml/min.The bioreactor was operated at a given pH until a steady state wasreached. After attaining the steady state, the pH of the feed waslowered step by step. Initially the pH of the feed was adjusted to4.5 and then it was decreased to 4.0, 3.5, 3.25 and 3.0 under con-tinuous flow conditions. The bioreactor removed 45.1%, 44.6%,35.5%, 38.3% and 14.4% of initial sulphate at pH 4.5, 4.0, 3.5, 3.25and 3.0, respectively.
Chang et al. (2000) demonstrated that solid waste materialsincluding oak chips (OC), spent oak from shiitake mushroom farms(SOS), spent mushroom compost (SMC), sludge from a wastepaperrecycling plant (SWP) and organic-rich soil (ORS) can be used aselectron donors and immobilisation matrices to treat ARD. The bio-reactors were inoculated with an anaerobic digester fluid. The feedsulphate concentration was 2.58 g/L and total dissolved metal con-
centrations were 500 mg/L iron, 100 mg/L zinc, 50 mg/L manga-nese and 50 mg/L copper. Temperature was maintained at 25 �Cand the pH of the medium was adjusted to 6.8. At a volumetricloading rate of 0.005 g/L/h, the highest volumetric reduction rateof 0.005 g/L/h was achieved in the bioreactor packed with sludgefrom wastepaper recycling plant.
Kolmert and Johnson (2001) investigated the tolerance of mixedSRP culture to acidic environment in an up-flow packed-bed biore-actor, using porous glass beads as a carrier matrix. The average vol-umetric reduction rates of 0.010–0.013 g/L/day were achieved inbioreactors containing mixed culture of acidophilic and neutro-philic SRP with a feed pH of 4.0. Kolmert and Johnson (2001)reported that sulphate reduction occurred at a pH of 3.0 but witha lower rate. Jong and Parry (2003) used an up-flow packed-bedbioreactor with sand as carrier matrix for anaerobic reduction ofsulphate with mixed culture of SRP. Feed contained 2.5 g/Lsulphate and 10 mg/L of each Al, As, Cu, Zn, Ni and Fe metals.The highest volumetric reduction rate of 0.019 g/L/h was observedat a volumetric loading rate of 0.155 g/L/h at 25 �C.
It has always been challenging to use CO2 and H2 as the carbonand energy source for sulphate reduction in a packed-bed bioreac-tor. Foucher et al. (2001) successfully proved that CO2 and H2 canbe used to treat Chessy mine drainage in an upflow packed-bedbioreactor with a special packing to provide good mass transfer be-tween hydrogen and liquid. The pH of the feed was 2.55 and thesulphate concentration was 5.8 g/L and metals like Fe2+, Fe3+, Zn,Cu, Al, Mn, Co, Ni and Pb were present in concentrations of 1470,70, 320, 160, 210, 5.5, 0.06, 0.4 and 0.5 mg/L, respectively.Although the feed sulphate concentration was 5.8 g/L, a part ofthe effluent stream was recycled and the concentration of sulphatein the inlet stream was reduced to 0.6–0.8 g/L. The maximum flowrate employed was 900 ml/h (residence time of 0.9 days), and thecorresponding volumetric reduction rate achieved was 0.2 g/L/h.
Lin and Lee (2001) studied anaerobic sulphate reduction in afixed bed bio-film column bioreactor. The Plastic Ballast rings werechosen as the supporting media for bio-film formation. The feedsulphate concentration was 0.9 g/L. The reactor volume was42.65 L, which yields a hydraulic residence time of 2.5 days. Thereactor temperature was controlled at 35 �C. The conversionachieved was 98%.
7.8. Flow direction
Vertical flow bioreactors have been used in numerous labora-tory and field studies (Dvorak et al., 1992; Cheong et al., 1998; Elli-ott et al., 1998; Drury, 1999; Tsukamoto and Miller, 1999; Changet al., 2000; Willow and Cohen, 2003; Tsukamoto et al., 2004; John-son and Hallberg, 2005b; Kuyucak et al., 2006). In downward flowmode bioreactors, the influent is fed through the top, while in theupward flow mode it is fed through the reactor bottom (URS Re-port, 2003). Recently, flow in a horizontal plane was reported ina field study (Zaluski et al., 2003). A three-step system separatingSRP activity from metal precipitation units and from a pH controlsystem was also proposed at the laboratory scale (Prasad et al.,1999). The flow pattern can affect both the transport of metalsand their interaction with the substrate (Song, 2003). Bioreactorswith vertical flow may show preferential channels of influentARD percolating through the reactive mixture. The upward flowbioreactors tend to last longer because upward flow limits compac-tion and preferential flow paths (URS Report, 2003). However, re-lease of metals by treated effluent is a potential problem. Ahorizontally oriented bioreactor using a mixture of cow manureand cut straw did not show preferential flow patterns during a32-months field operation period (Zaluski et al., 2003). This config-uration seems more promising, whereas the three-step process re-quires higher maintenance costs.
7.9. Sulphate concentration (SO2�4 )
Under optimum field conditions, sulphate reduction occurs atrates of about 0.3 mol/m3/d (URS Report, 2003). Comparison of lab-oratory and field passive bioreactors in terms of the sulphate re-moval rate is, however, not feasible because of several factors ofinfluence specific to each study. First, the HRTs in batch experi-ments are far longer than in columns and field-bioreactors. Second,there is the variability of the initial concentrations, ranging from229 mg/L (Zaluski et al., 2003) to 4800 mg/L (Waybrant et al.,1998), which influences SRP growth and sulphate reduction kinet-ics (Moosa et al., 2002). Third, some studies provide the calculatedpercentages of sulphate removal without presenting the initialand final concentrations (Cheong et al., 1998; Elliott et al., 1998;Johnson and Hallberg, 2005b). For example, a sulphate removalefficiency of 42% from 900 mg/L SO2�
4 translates to a residual sul-phate concentration about 500 mg/L (Tsukamoto et al., 2004),whereas an efficiency higher than 82% from 2315 mg/L SO2�
4 yields
about 400 mg/L of sulphate in the treated effluent (Jong and Parry,2003).
Typically, mine drainage contains >500 mg/L of sulphates andthese can be removed from it by reduction to sulphides, by biolog-ical uptake, and/or by the formation of organic esters on plantdecomposition in a vegetated wetland cell (biosorption). Sulphatereduction is carried out by anaerobic, sulphate-reducing bacteria.Moreover, some researcher’s report sulphate removal taking intoaccount initial sulphate concentration and equivalents of organiccarbon consumed (Tsukamoto and Miller, 1999; Tsukamoto et al.,2004). Therefore, great care should be taken when comparing theefficiency of passive systems in terms of sulphate removal.
7.10. Effect of sulphide on SRP
The state of sulphide solely depends on the pH of the environ-ment. The sulphide can be present in both HS� and S2� forms ata pH of 7.0 most of the sulphide concentration is in the hydrogensulphide form (Perry and Green, 1984). The relation between theconcentrations of undissociated hydrogen sulphide in the liquidand gas phase is based on the Henry’s law as shown by Reactions(28)–(30). The value of absorption coefficient, a, at 30 �C is equalto 1.99 (Lens et al., 1998).
Information available on sulphide toxicity and the mechanism oftoxicity is vague. It has been reported that sulphide is absorbed intothe cell and destroys the proteins thereby making the cell inactive(Postgate, 1984). If this is the case, bacteria should not be able to re-sume its activity once all the sulphide is removed. By contrast, it wasreported that the sulphide inhibition is reversible in SRP inoculatedbioreactors (Reis et al., 1992). Another theory states that the precip-itation of trace element metals, as metal sulphides, which are essen-tial for the growth of SRP, is the cause for the decreased activity(Bharathi et al., 1990). In addition to the uncertainty with respectto inhibitory mechanisms of sulphide, contradictory reports existwith respect to inhibitory effects of various forms of sulphide. Someresearchers report the sulphide inhibition based on the total sul-phide (Hilton and Oleszkiewicz, 1988), and some based on theundissociated H2S (McCartney and Oleskiewicz, 1991; McCartneyand Oleskiewicz, 1993). Speece (1983) stated that only the undisso-ciated H2S is capable of entering into the cell membrane. Later it wasshown that the bacteria has two threshold inhibition levels, one forthe undissociated H2S and the other for the total sulphide and thislevel depends on the environmental pH. At a pH less than 7.2, undis-sociated H2S is dominant and it will reach the threshold limit. At apH above 7.2 the total sulphide is responsible for the inhibitory ef-fect (O’Flaherty and Colleran, 1998). It is not easy to compare theinhibitory/toxic values reported in the literature, as the inhibitionhas been assessed based on growth, substrate degradation, sulphatereduction or cellular yield. The SRP are less sensitive to total sul-phide when the pH is increased from 6.8 to 8.0 and more sensitiveto the undissociated sulphide concentration. In addition, as the pHincreases less concentration of undissociated H2S is needed to inhi-bit the growth by 50% (O’Flaherty and Colleran, 1998).
The non-competitive inhibition model has been used by differ-ent workers to describe the effects of total sulphide and undissoci-ated H2S (Maillacheruvu and Parkin 1996; Kaksonen et al., 2004a).A previous report indicated that there is a difference in activity offree cells and bio-films employed for sulphate reduction (Kaksonenet al., 2004b). It is also said that an extracellular polymeric sub-stance that binds the bio-films protects the cells from the toxic ef-fects to some extent (Teitzel and Parsek, 2003).
Whether passive SRBR treatment is sufficient to treat a particu-lar ARD stream also depends on its chemistry. Up to a point, higherconcentrations of metal lead to higher metal precipitation rates.Under these conditions, potential metal precipitation can be calcu-lated from sulphate-reduction rates and reaction kinetics. How-ever, batch studies showed that high metal concentrations couldslow bacterial population growth, decrease sulphate-reducingcapacity, and ultimately cause death (Cabrera et al., 2006).
In general, ARD incorporates heavy metals such as iron, zinc,copper, manganese and lead at high concentrations which maybe toxic or inhibitory to the activity of SRP. Steps are taken to pre-vent the toxic effects of metals on SRP by finding new strains ofmetal tolerating bacteria and by employing bioreactors with spe-cial designs. Utgikar et al. (2002) reported that the inhibition ki-netic constants (based on rate of sulphate reduction) for Cu andZn as 17.9 ± 2.5 and 25.2 ± 1.0 m M�1, respectively and the toxic ki-netic constants (based on microbial growth) for Cu and Zn as 10.6and 2.9 m M�1. Another study reported that a strain of SRP (UFZ B407) was able to tolerate the presence of aluminium, lead, ura-nium, iron, chromium, copper, silver, nickel, manganese and cobaltat maximum concentrations of 50 mM, 10 mM, 0.01 mM, 50 mM,30 mM, 10 mM, 100 mM, 10 mM, 50 mM and 10 mM, respectively(Hard et al., 1997). In addition, strains of metal tolerating sulphate-reducing bacteria SRP were isolated and when tested, displayedgood activity in the stream containing 100 ppm of Cu and30 ppm of Fe (Garcia et al., 2001).
Contrary to common belief that only soluble metallic ions canbe toxic or inhibitory, Utgikar et al. (2001) demonstrated thatinsoluble metal compounds could also affect the activity of SRP.It was found that the insoluble metal sulphide formed is not toxicto the SRP by itself but it blocks the access to substrate and thenutrients that are essential for bacteria by forming a precipitatecoating on the SRP. To reduce the inhibitory effects of the insolublemetals and to increase the pH of ARD, a part of the treated ARD can
Fig. 5. Breakdown of o
be recycled and mixed with the influent ARD (Glombitza, 2001).The sulphide present in the treated ARD will react with the presentmetals and precipitate. In another study an extractive silicon mem-brane module was accommodated with the original bioreactor set-up (Chuichulcherm et al., 2001). The treated ARD from thebioreactor was passed to the outer part of the membrane (whichwas selectively permeable for H2S) while the untreated ARD waspumped through the inner part of the membrane. The H2S perme-ates from the outer side through the membrane and reacts withmetal ions present in the inner part to form metal sulphides. Themetal sulphides were removed as fine suspensions and the ARDwithout metal ions was fed to the bioreactor.
7.12. Chemical characteristics of substrate
In the past, selection of the organic substrates placed in SRBRhas been based on descriptive characteristics (i.e., chicken vs.cow manure, or leaf vs. municipal compost) (Gilbert et al., 1999;Pinto et al. 2001). These systems exhibited high levels of sulphateand metal reduction in the initial operating period but lacked long-term sustainability. Examples of substrates used include: chitin,leaf compost, various types and ages of manure, walnut hulls, mu-nicipal compost, different types of wood shavings, sewage sludge,hay, walnut shells, and pecan shells (Benner et al., 1997; Wildemanand Updegraff, 1997; Waybrant et al., 1998; Gilbert et al., 1999).Identification of SRP rates associated with specific fractions of or-ganic matter will allow for specification of organic substrate com-binations that provide rapid start-up and long sustainability.
It is essential to assess the chemical characteristics of organicmatter and the relationship between different fractions and effecton SRP rates. Important characteristics of organic matter may in-clude moisture content, organic fraction, nutrient content andthe composition of the organic fraction. The nutrient content isconstituted by protein, lipid and carbohydrates. Soluble proteins,simple sugars, organic acids and lipids can be separated and cate-gorized by using various solvents. These soluble components are
readily available to the microbial community. Ethanol and watercan be used for extractions and the extracts are designated etha-nol-soluble fraction (ESF) or water-soluble fraction (WSF), respec-tively. The residues are termed the ethanol-insoluble fraction (EIF)and water-insoluble fraction (WIF), respectively. The EIF and WIFrepresent fractions that must be hydrolyzed before they are avail-able to the microbial community. Hydrolysis is typically the rate-limiting step in anaerobic environments supported by solid phaseorganic matter. These insoluble fractions are further divided intothe acid-soluble (e.g., protein and cellulose/hemicellulose) andacid-insoluble fraction (e.g., lignin). Fig. 5 illustrates key microbialprocesses in an anaerobic environment (e.g., PTS) and shows thedifferent organic matter fractions associated with these processes.
Figueroa et al. (2004) on the basis of the chemical characterisa-tion and their solubility in anaerobic biozones categorized sub-strates into the following fractions (Fig. 6).
Another important characteristic of substrates is the metalsorption capacity. Batch experiments can be used to estimate theequilibrium capacity of substrates. Column studies can be usedto estimate kinetic effects and the sorption zone. The sorptioncapacity of the substrate is important during the start-up periodof the microbial populations and may be the primary mechanismfor metal removal for some metals (e.g., manganese).
Chang et al. (2000) considered the importance of the nitrogenand phosphorous contents of in the selection of the substrate. Sug-gested the enough requirement of nitrogen and phosphorous forthe growth of sulphidogens. A C/N ratio around 10 is generally con-sidered suitable for biological degradation of complex substrates(Reinertsen et al., 1984; Béchard et al., 1994). Higher ratio indicatesexcessive carbon or nitrogen deficiency, whereas lower ratio maysuggest a lack of carbon (Prasad et al., 1999). Poultry manure hasthe lowest C/N ratio (3.3) and the highest DOC, TKN and EAS, sug-gesting that the nitrogen was easily accessible (63% soluble sugars,hemicellulose, amino acids and proteins).
Gibert et al. (2004) correlated the chemical composition (lignincontent) of four different natural organic substrates (compost,sheep and poultry manures, oak leaf) and their capacity to sustainbacterial activity in an attempt to predict biodegradation fromchemical characterisation. The results showed that the lower thecontent of lignin in the organic substrate, the higher its biodegrad-ability and capacity for developing bacterial activity.
Zagury et al. (2006) reported that the least effective reactorswere those containing only leaf compost or poultry manure, thesubstrates with the least amount of TOC. However poultry manure
Fig. 6. Chemical composition of subst
has the highest DOC, EAS content. The wood materials (maplewood chips, conifer compost and conifer sawdust) had lower nitro-gen content (C/N ratio higher than 460) and less degradable carbon(lignocellulose and lignin). The C/N ratio of organic substrate wasdecrease from 6.5 to 3.2 after urea addition.
Coetser et al. (2006) performed a study on chemical character-isation of organic substrate with following conclusions: The mostreadily bio-degradable carbon source should be high in proteincontent and low in lignin contents. The higher the carbohydratecontent and crude fat content of a carbon source, the higher thecapacity to drive sulphate reduction. The higher the crude fibrecontent of carbon source, the lower the capacity to drive sulphatereduction. Authors also reported that chemical characterisationcan be used to assist in predicting sulphate reduction capacity ofa carbon source and the selection of organic electron donors for po-tential use in ARD treatment.
Although it is generally assumed that the chemical compositionof an organic substrate controls the patterns of its degradabilitybut no minimum data set to predict it has yet been established. Ani-mal nutritionists have developed chemical and enzymatic proce-dures to estimate feedstuff digestibility for ruminant diets (Rahnet al., 1999) and engineers have attempted to quantify the degrada-bility of a raw organic waste in methane fermentation research(Chandler et al., 1980), but unfortunately little standard practicehas been developed in environmental engineering (Prasad et al.,1999). Chemical approaches to predict the nutritive value of an or-ganic material are classically based on the quantification of poorlydigestible fractions (mostly structural organics, such as lignin andcellulose, which are resistant to microbial decomposition) and rap-idly digested fractions (Chandler et al., 1980; Prasad et al., 1999;Rahn et al., 1999).
Although chemical characterisation of the substances on anindividual basis provided insight on their organic carbon composi-tion and solubility. It did not give a good idea of ability to promotesulphate reduction and metal removal. It would be interesting toassess the influence of the chemical oxygen demand/SO2�
4 ratioand the quantity the lignin content of the organic materials degra-dability and sulphate removal efficiency, respectively (Gibert et al.,2004).
Chandler et al. (1980) proposed a model equation to estimatethe substrate bio-degradable fraction (B) based on the lignin con-tent as represented in the following equation:
where the bio-degradable fraction (B) is expressed on a volatile so-lid (VS) basis and X is the lignin content of the VS, expressed as a percent of the dry weight.
8. Selection and design considerations of SRBR
8.1. Selection of sulphate-reducing bioreactors
Brodie (1993) sorted out the empirical relationships in a mile-stone design flow chart that provided the foundation for a morecomprehensive design flow chart subsequently developed by He-din and Nairn at the former US Bureau of Mines as shown inFig. 7. This figure, in one form or another, continues to guide engi-neers and practitioners in the passive treatment cell design pro-cess. It has been modified by Gusek (2000) to include the passivetreatment of heavy metal-bearing ARD based on observations since1988. The sulphate-reducing bioreactor as shown reflects wherethis particular technology fits in the design philosophy. Althoughthe technology is well suited for ARD with net acidity and/or heavymetals, it can also be effectively applied to net alkaline watersources as indicated by the arrow drawn from the settling pondon the left hand side of the flow chart.
Fig. 7. Selection of passive treatment method
There is no ‘‘cookbook” design manual for passive treatmentsystems although the design flow chart above is a safe startingpoint. A phased approach design project is recommended; it typi-cally begins in the laboratory with static tests, graduating to finaltesting phases (bench and pilot) performed at the site on the actualARD. Bench scale testing will determine if the treatment technol-ogy is a viable solution for the ARD and will narrow initial designvariables for the field pilot. A proper bench scale test will certainlyreduce the duration of the more costly field pilot test. Field pilottest duration can range from days, to months, to years, dependingon the nature of the technology.
As shown in Fig. 7 sulphate-reducing bioreactors can be appliedin a number of different ARD situations. While most passive treat-ment systems (both aerobic zone and anaerobic zone types) offersimplicity of design and operation and economic advantages overactive/chemical treatment, sulphate-reducing bioreactors haveadvantages worth considering (Gusek, 2000):
� No aluminium plugging.� Can easily handle low flow net acidic water or high flow
net alkaline water.� Uses waste organic materials.
� Resilient to loading and climate variations.� Consumes sulphate; capable of treating selenium and uranium.� Generates more net alkalinity in effluent.� Burial to minimize vandalism.� Opportunities for community involvement in organic
procurement.� Might be able to construct them in abandoned underground
mines.
8.2. Design considerations
Once technology and application methods have been selected,site and source specific information can be combined with generaldesign guidelines. No specific design is appropriate for every site;technical consulting, bench test, and often pilot scale tests are nec-essary to correct size and configure bioreactor treatment system.However, general guidelines have been developed that may guidethe early stages of the design process. The final design and con-struction decisions will be based on the flow rate to be treated,the loading rates of metals, and the space available for thebioreactor.
8.2.1. Flow rateLarge flow variation can overwhelm anaerobic bioreactors if
they are sized for lower flow rates. Conversely, if SRP bioreactorsare sized for maximum probable flow and the flow decrease signif-icantly, some of anaerobic substrate can become oxidized and re-lease metals and had previously been precipitated, if the cellsurface is not designed to be permanently submerged. Release ofmetals is a potential problem especially in up-flow bioreactors be-cause the treated water will flush metals released during oxidationof the upper substrate as it leaves the bioreactors. If the flow and/or metal loading to the bioreactors increased significantly abovethe design flow, especially as in the spike increase event, the bac-terial community in the bioreactor may suffer severe die-off andsignificantly reduce system effectiveness and require additionalsystem maintenance (e.g., the microorganism consortium allowto re-establish, the addition of the new inoculums) (Gusek et al.,1998).
8.2.2. pHThe SRP are not only obligate anaerobes, but also require a nar-
row pH range near neutral for optimal sulphate reduction. It is rea-sonable to assume that the pH is the predominant controllingfactor. Optimization and enhancement of the reactor may byaccomplish through the modification of the mine drainage tonear-neutral pH. Perhaps the use of anoxic limestone drains beforethe SRP bioreactor would raise the pH to levels more acceptable tothe bacteria. In the improvement of water quality of the treatedwater, pH also played a very important role, revealed efficient re-moval of metals and acidity. For SRP bioreactors, a pH of approxi-mately 5.5 or higher is preferred, but lower values have beensuccessfully treated when the hydraulic loading factor of the reac-tor was proportionately decreased. That is, the more the acidicARD, larger the surface area of the SRP bioreactor is required.
8.2.3. Sulphate concentration desiredSulphate loading is a very conservative factor in treating metals
because it assumes that the bacteria reduce all sulphates. Typicallysulphate concentrations do not drop below a few hundred milli-grams per litre. The calculation is useful, however, because in orderto function properly, the system requires sulphate to be present atlevels in excess of metals. Moosa et al. (2002, 2005) studied thekinetics of anaerobic sulphate reduction and it was observed thatthe kinetics of reaction was shown to be dependent on the initialconcentration of sulphate in the feed. An increase in initial concen-
tration of sulphate from 1.0 to 5.0 g/L enhanced the maximum vol-umetric reduction rate from 0.007 to 0.075 g/L/h. Similar resultswere also observed by Baskaran (2005) and concluded at highervolumetric loading rates of sulphate (shorter residence times),while maintaining a high conversion and volumetric reduction rate.
8.2.4. Metal loading rateThe flux of heavy metals, including iron, copper, zinc, and cad-
mium, into the system must be less than the rate of sulphatereduction. Typically, flux values of 0.15–0.30 mol/m3/day heavymetal are used. Metal concentrations must be converted frommass/volume (mg/L) to mol/L. Then: mol/day = Q � c, whereQ = discharge in volume/time (l/day) and c = concentration, mass/volume (mg/L).
Once metal loading rates are known, the total volume of sub-strate required must be determined using estimated SRP activity.Waters with lower metals concentrations could either be treatedwith a smaller system or with higher flow rates. Alternativelyand perhaps even better is to assume an empty bed hydraulic res-idence time of 20–40 h. Forty hours would be used if the feedwaters are below pH 5 and 20 h if the feed waters were near-neutral pH (Cohen, 2006).
8.2.5. Retention timeZinc and some other metals sulphides can require up to 3 days
for effective precipitation. Therefore, the bioreactor should be sizedto allow minimum of 3 days retention time if zinc is to be treated.The porosity should be sufficient high for optimum residence time(Filipek et al., 2003). Present bench scale study also consistencewith the finding of Cohen and Staub (1992), who found that reten-tion period of about 4 days were sufficient to ensure metal removalefficiency more than 98% in anaerobic reactors used to treat minewater discharges.
8.2.6. Available areaArea available at the location for treating heavy metals and sul-
phate-laden ARD is the important inter related parameter. Thisbasically depends on the metals loading rate, sulphate loading rateand flow rate of the bioreactor. Higher amount of contaminateneeds more surface area when the flow rate remain a constraintdue to removal efficiency.
8.2.7. Sulphate-reduction ratesThe molar sulphate removal rate is typically estimated in moles
per day per cubic meter of organic medium or substrate (moles/day/m3). Cell design protocol typically attempts to match the vol-umetric removal rates for metals and sulphate at 0.3 mol/day/m3.This benchmark value has been established over dozens of SRBRapplications ranging from bench- to pilot- to full-scale systems(Gusek, 2005). Hedin et al. (1988) have suggested that a good esti-mate for rates of sulphate reduction was 300 nmole/S producedper cm3 of substrate per day.
Reynolds et al. (1991) found that water extracts of hay could in-crease sulphate reduction between 2.5 and 7 times. SRP cannot uti-lize complex organics. They require simple, volatile, organic acidssuch as acetate and lactate. Indeed, additions of these nutrients en-hanced SRP activity. There probably is a consortium of heterotro-phic bacteria in the substrate that decomposes complex organicsinto forms available to the SRP. Reynolds et al. (1991) also demon-strated that SRP activity is approximately 600 nmole of sulphideproduced per cm3 of substrate per day. Calculations based on themetal loading rates at the Eagle Mine and the 11.28 m3 of substrateof the bioreactor, that the system should theoretically reach itslimit of metal removal at a flow of between 200 and 400 ml/min.Maximum removal rates of 97–100% have occurred for all metalsexcept for manganese at the 200 ml/min and 400 ml/min flow rate.
Reynolds et al. (1991) also reported typical levels of sulphatereduction to be 600–1200 nmole/cm3/day.
Machemer et al. (1993) concluded from the experimental re-sults, a universal sorption capacity cannot be assigned to passivemine drainage systems as a whole, but adsorptive capacity lastsfor 30–60 days; reactor life-span is 4–6 years. Thus, most atten-tion should be focused on optimizing conditions for sulphatereduction.
Eger and Lapakko (1988) and Wildeman et al. (1993) describeda design process for solid substrate bioreactors and SSF wetlands.The acidity load to the system is calculated by summing the freehydrogen ions in the influent plus hydrogen ions that would be re-leased through metal precipitation. Since one mole of sulphatereduction typically produces two mole of alkalinity, the system issized by dividing the acidity load by one half of the anticipated sul-phate reduction rate in moles/day/m3. A major problem is forecast-ing what the sulphate reduction rate will be, since it depends onthe biodegradability of the organic substrate and the temperature.Additionally, the hydraulic retention time affects the amount ofsulphate reduction, because more electrons from solid substratedegradation will be transferred to the water if the retention timeis long (Drury, 2000).
Wildeman et al. (1994) reported typical sulphate-reductionrates range from 200 to 600 mmole/m3/day but an average rateof 300 mmole/m3/day has been recommended for the design ofsulphate reduction systems. The above equation means that a dailysupply of 150–400 (average 300) mmole/m3 of organic matterneeds to be available (molar ratio being 2:3). Carbon productionin natural wetlands has been reported to be about 1 kg/m2/year(Whittakar and Likens, 1972). A 1 m thick substrate can provide23 mmole of carbon/m3/day. Eger and Wagner (2003) reported for1 m thick substrate, this would provide 23 mmole carbon/m3/day.Since two moles of carbon are required to reduce each mole ofsulphate, the reaction rate based on the annual input of carbonwould be only about 12 mmole/m3/day or a factor of 25 less thanthe typical design rate of 300 mmole/m3/day.
Eger and Lapakko (1988) also calculated the rate of sulphatereduction from the difference in input and output concentrations.Sulphate-reduction rates measured during the first 2 years of thestudy were generally within the range of rates measured in otherinvestigations. However, rates decreased over time. When the col-umn studies began, sulphate-reduction rates were higher than thevalues measured at the end of the field experiment, but rates de-creased over time. When the MSW compost columns were shutdown (‘‘resting period”), sulphate-reduction rates increased tem-porarily but then decreased within a month. Reduction rates inthe MSW compost was from around 290 mmole/m3/day in thebegning to about 20 mmole/m3/day at the end of 1999.
Glombitza (2001) after conducting column study with acetate-amended compost and a residence time of 0.73 days, also reportedthat although longer residence times resulting higher sulphateremovals but the rate of removal was slow.
Drury (2000) concluded that the sulphate reduction rate was:
SRR ¼ f0:5ð1000� 10�pHÞ þ 1:5Xðmmole=l of Al3þ þ Fe3þÞ
þXðmmole metal sulphide formed; neglecting Fe3þÞgQ
ð32Þ
where SRR is the required sulphate reduction in mmole/timeand Q is the flow rate through the reactor in litres/time.
Eq. (32) assumes that the Al is all 3+, one mole of Fe3+ reductionproduces one equivalent of acidity, and two equivalents of alkalin-ity are formed per mole of sulphate reduced. The 1:1 relationshipbetween sulphide production and metal sulphide precipitation isnot meant to replace traditional equilibrium chemistry as a
description of metal sulphide precipitation, and is only relevantto the required sulphate reduction rate calculation.
8.2.8. Metal and acidity removal rateThe flux of heavy metals, including iron, copper, zinc, and cad-
mium, into the system must be less than the rate of sulphatereduction. Design criteria typically recommended for SRP are basedon metal molar volumetric and acidity areal loadings. Rose andDietz (2002) found acidity removal rates between 25 and 50 g Ca-CO3/m2/day during evaluation of 12 VFW systems with net-acideffluent. Their study also reported net-alkaline effluent when acid-ity loading was <40 g CaCO3/m2/day.
Thomas et al. (2002) found average acidity removal of 87.8 g Ca-CO3/m2/day; however, average acidity feed rate for the experi-ments was 57.8 g CaCO3/m2/day. These results indicate higheralkalinity generation rates occur when influent acidity loading ishigh, which corresponds with findings by Rose and Dietz (2002)who showed positive correlations between retention time, pH, Feconcentrations and alkalinity generation. A comprehensive evalua-tion by Ziemkiewicz et al. (2003) showed average acidity removalof 62.3 g CaCO3/m2/day. Watzlaf et al. (2000b) recommendedapplying areal removal rates of 25–30 g acidity as CaCO3/m2 sur-face area/day. Rose (2004) re-evaluated their earlier study (Roseand Dietz, 2002) and proposed a non-Mn acidity design criteriaof about 35 g CaCO3/m2/day for VFWs. The removal rate doubledwhen the system were incorporates with fine limestone in thecompost mixture.
Skousen and Ziemkiewicz (2005) evaluated 16 VFWs and foundacidity removal rates >200 g CaCO3/m2/day for two systems, fivebetween 39 and 87 g/m2/day, eight between 2 and 17 g CaCO3/m2/day and one system that did not reduce acidity. Gusek (2005)recommended that SRBR cell design criteria included satisfying avolumetric metal loading factor of 0.3 mol of metal loading perday per cubic meter of organic media, as well as an acidity loadingfactor. Wildeman et al. (1994) recommended design criteria of0.3 mol of metal removal/m3 of substrate/day for SRBRs with amixture of organic materials and crushed limestone. They alsoindicated that removal efficiencies reduced about 25% in cold cli-mates. Clarke et al. (2005) proposed a non-Mn acidity design crite-ria of about 35 g CaCO3/m2/day. Cohen (2006) proposed typicalflux values of 0.15–0.30 mol/m3/day for heavy metal for the designof bioreactors.
McCauley et al. (2008) proposed a design criterion for metal andacidity removal for future bioreactor systems based on treatmentperformance of the laboratory scale SRBR. These values range be-tween 0.807 and 1.25 mol of metals removed/m3 substrate/day oracidity removal of 66.7–126 g CaCO3/m2/day. Therefore, conserva-tive design criteria based on the performances of laboratory SRBRare 0.80 mol of metals/m3 substrate/day (32.3 g metals/m3 sub-strate/day) or acidity removal of 66 g CaCO3/m2/day. Therefore, de-sign criteria used for future pilot-scale research was basedprimarily on the maximum removal capabilities of bioreactorsmeasured during stable operation. Overall, the treatment perfor-mance of the SRBRs in this study exceeded recommended designcriteria used for similar systems including SRBRs and VFWs. How-ever, most other systems incorporated limestone as the primaryalkalinity material, not mussel shells. The findings reported>0.8 mol of metal removed/m3 substrate/day. Further it waspointed out that contaminant loading rates should be re-evaluatedor reduced to avoid potentially altering the biogeochemical systembalance if effluent pH falls below 6.7. System hydraulics is also animportant design consideration. Maximizing bioreactor substratedepth and minimizing surface area footprint should be consideredto reduce treatment footprint and reduce discrepancies betweenactual hydraulic residence time and theoretical hydraulic residence
time (tracer studies are currently being conducted to assess actualresidence time and system hydraulics).
Cohen (2006) explained with an example, at the Eagle Mine, themass loadings for the major metals were 1.49 mol metals/day for100 ml/min, 2.97 mol metals/day for 200 ml/min. Once metal load-ing rates are known, the total volume of substrate required mustbe determined using estimated SRP activity. At the Eagle Mine,the substrate volume was 2.46 � 106 cm3. The estimated sulphideproduction rate was from 300 to 1200 nmole S�/cm3/day. There-fore, moles of S- produced/day = V (volume of substrate) � SRPactivity rate mass/volume/day) and for this system the sulphideproduced was 0.74–2.95 mol S�/day. Water with lower metal con-centrations could either be treated with a smaller system or withhigher flow rates. Alternatively and perhaps even better is to as-sume an empty bed hydraulic residence time of 20–40 h. Fortyhours would be used if the feed waters were below pH 5 and20 h if the feed waters were near-neutral pH.
8.2.9. Final sizing of organic portionAcid-rock drainage is a multi-component pollutant and the
characteristics of each component will vary with each site. There-fore, individual information of each pollutant removal can be ex-plained comprehensively for the polluting extent of ARD, thougha more conservative estimate may be preferable (Gusek, 2005).Model should be developed that integrate of these characteristics(i.e., SRR, MRR and acidity removal rate).
The final sizing will of the sulphate-reducing bioreactors will becalculated after calculating the sizing based on metal, acidity andsulphate loading rate. The size of the organic portion of con-structed SRP bioreactors can be calculated based on: flow rate ofthe ARD, sulphate loading rate, heavy metal and acidity loading.
Volumes of the substrate are calculated according the sulphatereduction rate, metal removal rate and acidity removal rate as(m3):
Sizing calculations must be conducted for each criterion todetermine which one is the limiting factor and that will be the finalsize of the substrate.
8.2.10. The final design and construction decisionsThe final design and construction decisions will be based on the
flow rate to be treated, the loading rates of metals, sulphate andthe space available for the bioreactor. The decision whether ornot to use plants should be based solely on the aesthetic and ero-sion considerations, not on plants as a major contributor to metalremoval or system longevity. Once operating, one can expect aneffective life of 4–6 years from a single load of substrate, basedon experience gained by bench scale bioreactors performance(Gusek, 2000).
On start-up of the SRP bioreactor, the soluble organic com-pounds are especially prevalent, usually giving the effluent abrownish colour. Also, the bacterial population needs time (i.e.,15–30 days) to become established in the bioreactor. Ideally onlya portion of the total effluent load is allowed to enter the bioreactorduring this time period. Due to the reactions in the SRP bioreactors,soluble organic and nitrogen compounds, as well as sulphides, areproduced. These compounds must be removed from the effluentbefore release off-site. Typically this is accomplished by placing asmall aerobic polishing cell in line after the SRP bioreactor, which
can also function to remove any manganese present. Between 3and 4 years is the maximum reported period of an SRP bioreactorwithout a major design related retrofit (Filipek et al., 2003).
9. Performance
Passive treatment technologies, including SRP, have been usedto treat ARD from coal mines for over 20 years (Ziemkiewiczet al., 1997). Several authors have evaluated the performance ofmultiple passive treatment systems, to develop general perfor-mance expectations.
Benner et al. (1997) conducted a field test on Bio-reactive wallsin 1995, where mine drainage flowing within an aquifer is inter-cepted and treated using a 15 m wide, 36 m deep and 4 m thickwall. The design rate of flux through the wall was 288 m3/years.Results showed that down streams sulphate concentration hasbeen reduced by 50% and iron concentration was reduced by95%; pH raised by 5.8–7.0 coinciding with the increase in the alka-linity from 0 mg/L to 50 mg/L as CaCO3.
Cheong (1998) evaluated the performance of a pilot reactoroperated at the Dalsung mine for treatment of ARD. The reactor,containing a mixture of rice stalks, cow manure, and limestone,showed removal of 98% Cu, 100% Zn, 99% Fe, 100% Cd, 97% Al,61% Mn and 100% Pb when the effluents from the reactor had apH of 6 and an Eh of about �300 mV. However, as time passed,the Eh rose and the amount of metals removed decreased, exceptAl. This indicated that maintaining reducing conditions was veryimportant for continued metal removal. During the operating per-iod, there were some problems such as volume change in the sub-strate within the reactor and scaling on pipes. These problemsappeared to reduce the flow of mine drainage in pipes and thereactor over time.
Elliott et al. (1998) developed an upflow porous medium biore-actor was inoculated with Sulphate-reducing prokaryotes (SRP)and operated under acidic conditions. The reactor was operatedunder continuous flow and was shown to be capable of sulphatereduction at pH 4.5, 4.0, 3.5 and 3.25 in a medium containing16.1 mM sodium lactate.
Tsukamoto and Miller (1999) reported that pilot scale anaerobicbioreactor was working at Leviathan Mine, Alpine county; CA. thissystem initially utilized a mixture of horse manure and sand as asubstrate. The microbiological community of SRPs and carbonsource were contained in manure. Although the bioreactor workedefficiently for 1 year reducing sulphate concentrations along withthe removal of arsenic and divalent metals, including iron andnickel, the system began to loose its ability to remove sulphateand iron. By the end of second year, the system removed less than10% of the influent sulphate and iron, and use of the bioreactor wasdiscontinued. This depleted substrate was used for the substraterevitalizing. Methanol was delivered to the depleted pilot-scalebioreactor that was no longer effectively treating ARD. Methanoladdition (200–800 mg/L) resulted in reactivation of the bioreactorwith sulphate removal of up to 69%, iron removal of up to 93%, anda pH increase from 3.6 in the influent to 6.5 in effluent.
Garcia et al. (2001) investigated the reduction and removal ofsulphate. The biological reduction of sulphates was effective aswell and in agreement with the preceding results the bacteria wereactive, from the point of view of their metabolism, both at pH 7 andat pH 5. The sulphate concentration decreased by about 85% in27 days at pH 7 whereas at pH 5 a similar result was obtained in9 days (from 9000 ppm to 1350 ppm). As it was predictable pH in-creased to alkaline values (pH 8.2) and Eh decreased to negativevalues (�250 mV) during the tests. In all the tests the SRP werecapable of inducing their own environmental conditions (pH, Eh,etc.).
Eger and Wagner (2003) also reported an encouraging perfor-mance of bioreactor treatment system. The pH in the outflow fromthe overall system increased from about 5.0 to over 6.0, sulphatedecreased by at least 50%, and over 99% of the input copper andnickel were removed. Although the system effectively removedcopper and nickel for the entire experiment, sulphate removal de-creased and outflow pH fluctuated, reaching a minimum of around4 during the column study. By reducing the flow rate and periodi-cally stopping the columns for a period of 1–2 months, treatmentwas re-established and outflow pH increased. At the end of theexperiment, the outflow pH was above 7.5 but the inflow ratehad been reduced to about 10% of the original flow.
Kalin and Chaves (2003) observed the performance of a micro-bial treatment system, partially implemented, receiving aban-doned mine portal effluent in the gold mining district of NovaLima, Minas Gerais, Brazil. The system consisted of four oxida-tion–precipitation–settling ponds and three microbial treatmentponds operating in series. The mine portal effluent flow volumeaveraged 0.6 L/s, increased during the rainy season to 1.0 L/s. Theportal effluent pH ranged from 1.4 to 3.4, measured between1995 and 2001. The particle-settling rates specific to the portaleffluent were 25–3.9 g/m2/day, which decreased from the oxida-tion–precipitation ponds. Most of the pyrite oxidation and subse-quent iron hydrolysis to iron hydroxide occurred in theunderground workings. The resulting weathering products wereflushed out during high flows. For these effluents, the function ofthe oxidation ponds was mainly iron hydroxide particle removal.Oxidation–settling ponds were necessary to prevent or reduce ironcoating of the limestone and organic carbon. From oxidation–settling pond, the effluent flowed into a series of microbial treat-ment ponds. These ponds were ‘‘seeded’’ with 10 tones of rawpotatoes and 4.5 tones of whole sugar cane. These ponds weredesigned to have a floating cattail cover to reduce turbulence andstabilize anaerobic sediments, and to provide a continuous organiccarbon source to the sediments. The pH of discharge from themicrobial treatment ponds ranged from 4.5 to 7.2 between June1999 and June 2001. As yet incomplete treatment system however,removed on average 37–87% of the monthly Ni load (0.6–1.7 kg/month), 77–98% of Al load (19–49 kg/month), 74–82% of Zn load(1.7–0.6 kg/month), and 78–95% of Fe load (71–236 kg/month).
Ziemkiewicz and his co-workers surveyed the passive treat-ment plants and found that SAPS had an average acid removal rateof 62.3 g/m2/day and cost $253/ton/year, while anaerobic wetlandsremoved 24.5 g/m2/day and cost $527/ton/year on average. Thoughlarge variation in performance and cost made generalization diffi-cult, this study showed that SRP treatment had potential to exceedaccepted design factors of 20 g acidity/m2/day for SAPS and3.5 g acidity/m2/day for anaerobic wetlands (Ziemkiewicz et al.,2003).
NRMRL (2004) begin long term evolution of a compost free bio-reactor developed by researchers at the University of Nevada, Reno.The reactors relied on sulphate-reducing microbial organisms suchas Desulfovibrio sp. to neutralize acidity and to precipitate metalsulphide from the ARD at a flow rate ranging from 8 to 30 gpm. Un-like compost bioreactors, this technology used a liquid carbonsource and a rock matrix (rather than a conventional compost orwood chips matrix) that is consumed by bacteria and collapses overtime. A benefit of this technology includes better control of biolog-ical activity and improved hydraulic conductivity and precipitateflushing. The bioreactor treatment began with the introduction ofARD to a pre-treatment pond where sodium hydroxide is addedto increase pH from 3.1 to 4.0. Alcohols also added to serve as a car-bon source for microbes. Acid-rock drainage from pre-treatmentpond was passes to an upstream, 3810 m3 bioreactor for additionalmetal removal. Both bioreactors contained 6–24 inch river rockaggregate that serve as a substrate for SRP growth. Each bioreactor
had three effluent distribution lines and three effluent collectionlines located at different elevations to allow variable flow opera-tions. Precipitates from the second bioreactor settled in 5000 m3,continuous flow pond. From this pond the effluents were allowedto flow to a rock lined aeration channel that promotes degassingof residual hydrogen sulphide prior to discharge. Precipitate slurrywas flushed periodically from bioreactors to prevent plugging of theriver rock matrix and was allowed to settle in a 5488 m3 flushingpond. Solids generated by the technology were non hazardousand may be used (pending additional studies) as soil amendmentsduring future reclamation of the site. Preliminary results indicatedthat this bioreactor system is achieving a metal removal efficiencyof 91–99%. Further it was found that each technology promotesARD neutralisation and metal precipitation was also meeting thesite discharge standards. The field studies suggested that activebiphasic lime treatment may be more effective in applicationsinvolving a high rate of flow and a short treatment season, whilethe semi-passive alkaline lagoon favoured a low flow rate and ex-tended treatment season. The passive compost free bioreactor how-ever is not constrained by seasonal conditions and can be sealed totreat the low to moderate flow common at ARD sites. In additionboth these biphasic and alkaline treatment lagoon technologiesgenerated larger quantities of sludge than the bioreactor.
Luptakova and Kusnierova (2005) conducted a study to deter-mine the bacterial produced hydrogen sulphide could be used forthe elimination of soluble heavy metals from ARD in the form ofsparingly soluble sulphides. They investigated the kinetics of thecopper precipitation in the form of sulphides from the model solu-tion containing Cu2+ by Sulphate-reducing prokaryotes on theground of two different approaches. In the first approach one reac-tor was used, which produced the hydrogen sulphide as a by-prod-uct of SRPs and thus the copper precipitation by the bacterialproduced hydrogen sulphide as a copper sulphide. The secondapproach allowed the successive running of aforementionedprocesses and used two interconnected reactors. The hydrogensulphide bacterial production was realised in the first reactor andthe copper precipitation by the bacterial produced hydrogensulphide was realised consequently in the second reactor. Underthese conditions this method involves three stages such as: thehydrogen sulphide production by sulphate-reducing bacteria, thecopper precipitation by the bacterial produced hydrogen sulphideand the copper sulphides filtration from the liquid phase. Theadvantage of the second approach is the fast running of the Cu2+
elimination, as well as the possibility of the selective metal precip-itation in the form of sulphides.
A case study of yellow creek phase 2B bioreactors at BlacklickCreek Watershed (1090 km2, contains 300 surface coal mines and170 coal refuse dumps, which contribute an average of136,000 kg of acid per day to the streams) in Western Pennsylvania(Black Creek Watershed Association, 2006). One passive treatmentsystem included in this project is the phase 2B bioreactor. Duringthe time of treatment, ARD entering the Yellow Creek TreatmentSystem Phase 2B (YCTS) had an average pH of 2.8, 574.7 mg/L acid-ity, 45 mg/L iron, 33 mg/L aluminum, 2.6 mg/L manganese, and791.7 mg/L sulphate. The YCTS sulphate-reducing bioreactor pondhas a 0.13 ha bottom area and is 0.76 m deep. The substrate is com-prised of 50% woodchips, 30% limestone, 10% cow manure, and 10%hay. The bioreactor was designed for an average flow rate of 38 L/min. the construction cost of the system, including engineering de-sign, was $158,000 (Gusek, 2005). During 4 years of operationsYCTS continue to reduce sulphate, generate alkalinity and loweriron and aluminum.
Zagury et al. (2006) tested to evaluate the performance of singlesubstrate, ethanol, a mixture of leaf compost (30% w/w), poultrymanure (18% w/w) and maple wood chip (2% w/w), and the samemixture supplied with formaldehyde. The lowest metal and
sulphate removal efficiency was found in the reactor containingpoultry manure as a single carbon source despite its high DOCand EAS content. The mixture of the organic materials was mosteffective in promoting sulphate reduction, followed by ethanoland maple wood chips; and single natural organic substrates gen-erally showed low reactivity. Manure was partly mixed in bottomsand bed and partly suspended in the water phase of cylinders.
Viggi et al. (2010) investigated batch-optimised mixture (w/w %:6% leaves, 9% compost, 3% Fe(0), 30% silica sand, 30% perlite, 22%limestone) in a continuous fixed bed column reactor for the treat-ment of synthetic acid-rock drainage (ARD). A column reactor wasinoculated with sulphate-reducing prokaryotes and fed with asolution containing sulphate and heavy metals (As(V), Cd, Cr(VI),Cu and Zn). At steady state, sulphate abatement was 50–100%,while metals were totally removed. A degradation rate constant(k) of 0.015–0.001 h�1 for sulphate removal was determined fromcolumn data by assuming a first order degradation rate. Reductionof ARD toxicity was assessed by using the nematode Caenorhabdi-tis elegans as a test organism. A lethality assay was performed withthe toxicants before and after the treatment, showing that only 5%of the animals were alive after 48 h in presence of the contami-nants, while the percentage increased to 73% when the nematodeswere exposed to the solution eluted from the column.
10. Longevity predictions
The lives of sulphate-reducing bioreactor system for the treat-ment of ARD depend/rely primarily on SRP, which reduce sulphateto sulphide and generate alkalinity. The bacteria will remain activeas long as the system remains anaerobic and contains an adequatesupply of sulphate and small chain organics. Although estimates oflifetime made on the total carbon in the system suggested that life-time should exceed 20 years, data from field and laboratory studiesshowed that the rate of sulphate reduction decreases with time,and after several years’ rates decreased substantially. To maintainacceptable treatment, either additional organic material must beadded or the total metal and acid load to the system would needto be reduced substantially (Eger et al., 2000). Numerous studieshave documented the ability of sulphate reduction to treat bothcoal and metal mine drainage, but many of the early studies wereconducted for only one to 2 years (Hedin et al., 1994; Cohen andStaub, 1992; Dvorak et al., 1992; Wildeman et al., 1994). Initiallythe results were impressive; pH increased from less than 4 to over7, and typical trace metal removal exceeded 90%.
The bioreactors were originally considered passive because theywere believed to be able to function on the order of 20 or moreyears before requiring major maintenance (i.e., replacement ofthe treatment media). In the past 10 years, a number of pilot SRPbioreactors have been used successfully over short periods to re-duce heavy metals such as copper and zinc from sulphate-ladenARD. However, none have remained operational for more than afew years without significant overhaul or modification. Between3 and 4 years is the maximum reported period of an SRP bioreactorwithout a major design related retrofit (Gusek, 2005). Treatmentlifetime depends on the types of processes that provide the major-ity of the metal removal. Anaerobic systems (vertical flow systems)rely primarily on SRP, which reduce sulphate to sulphide and gen-erate alkalinity. Metals can be precipitated as sulphides, and acid-ity is neutralized. These systems can remove over 90% of the metalsand increase pH from around 4 to over 8. The bacteria will remainactive as long as the system remains anaerobic and contains anadequate supply of sulphate and small chain organics. Estimatesof lifetime made on the total carbon in the system. Data from fieldand laboratory studies show that the rate of sulphate reaction de-creases with time, and that after several years’ rates decreased sub-
stantially. To maintain acceptable treatment, either additionalorganic material must be added or the total metal and acid loadto the system would need to be reduced substantially (Eger andWagner, 2003).
For each mole of the sulphate reduction that is reduced, twomoles of carbon are required (Eq. (36)). Sulphate reduction isdependent on a continued supply of sulphate and organic com-pounds produced by the decomposition of the organic matter inthe substrate (Jorgenson, 1983).
Life of substrate ¼ TOW� OC� g12� SRRdesired
ð36Þ
where TOW is the total organic waste, OC the percent of organiccarbon, g the efficiency of organic matter (assumed 90%), and SRRis the sulphate reduction rate desired.
Lifetime estimates based on the total amount of carbon in thesesystems suggested that the substrate should last for several dec-ades. Data collected from a long-term field and laboratory studyindicate that these predictions may substantially overestimatethe lifetime of the substrate (Hedin et al., 1988; Cohen and Staub,1992; Dvorak et al., 1992; Wildeman et al., 1993;Hedin et al., 1994;Wildeman et al., 1994).
11. Conclusions
From the above review of literature it is concluded that the bio-remediation of dissolved metals in ARD using SRP in anaerobic bio-reactors is rapidly becoming a ‘‘best available technology”. Thecombination of this anaerobic bioreactor technology with ad-vanced, semi-passive engineered wetland technology presents apowerful new way to treat mine drainage. Sulphate-reducing bio-reactor is reasonable alternative technology for ARD treatmentpossible to be applied even on remote sites, without power, andwith extreme winter conditions also. The simple and continuousflow sulphate-reducing bioreactor can be efficient for the neutral-isation of the ARD and for an efficient removal of their metals(Costa et al., 2008; Dvorak et al., 1992). The SRP mediated anaero-bic bioremediation process is not limited to acidic mine waters, itcan be applied to various industrial waste waters of different pHthat are high in sulphates and metals and/or organic contaminants.The process can be designed in different ways to accommodate dif-ferent locations and natures of the contaminated waters (Higginset al., 2003).
The efficiencies of passive bioreactors depend on the activity ofSRP, which is mainly controlled by the composition of the reactivemixture. The most important component is the organic carbonsource. Many studies have attempted to predict the biodegradabil-ity of complex organic substrates by using chemical extractions;however, they have not been successful. Higher sulphate-reductionrates have been reported with reactive mixtures containing morethan one organic carbon source. Various organic waste materialcan be used in the bioremediation of the ARD. Bioreactors are rec-ommended to be allowed to ‘‘mature” before fed with ARD, espe-cially when recalcitrant materials are included in the substrate toprovide long-term provision of organic carbon. The capacity ofthe organic substrates to promote sulphate removal by SRP activi-ties are directly correlated with their chemical characterisation.Most of the metal sulphides that were formed due to the SRP activ-ity precipitated within the organic matter. That seems to be truefor the rest of the metals that must have formed hydroxides andcarbonate compounds. Even though formation of metal sulphidesis the preferred metal removal process, many metal removal mech-anisms including adsorption and precipitation of metal carbonatesand hydroxides occur in passive bioreactors. Furthermore, thesemechanisms change during the life-span of a passive bioreactor.
The lifetime of SRP systems is a function of the input waterquality and the type of removal processes. For anaerobic systems,removal will occur as long as a carbon source and sulphate arepresent in an anaerobic environment. Since problematic ARD gen-erally has elevated sulphate concentrations, the limiting reactant isthe amount of available carbon. For organic substrate systems, theestimated lifetime based on total carbon generally exceeds 6–10 years, but effective treatment is likely to last less than 5 yearsunless the substrate is replaced or supplemental carbon is addedto the system. To maximize bioreactors efficiency, the bioreactorsneed to be covered with a plastic liner that would minimize oxygenintrusion either directly from the atmosphere or through atmo-spheric precipitation.
The extreme rainfall quantities and intensities typical at minesite need to be considered when scaling the reactors up to pilotor industrial scale. Since land disturbances are typical at activemine sites, extreme care and planning is essential to ensure thatchemistry and flow of ARD seeps is not exacerbated. Treatmenteffectiveness of SRP will be reduced if metal and acidity loading ex-ceeds system limitations (or design criteria). Implementation of anSRP to treat ARD should incorporate contingency overflow diver-sion to prevent system overloading and damage in the event ofunexpected site disturbances.
Maximizing bioreactor substrate depth and minimizing surfacearea footprint should be considered to reduce treatment footprintand reduce discrepancies between actual hydraulic residence timeand theoretical hydraulic residence time. The technology should beused at actual mine site using the design criteria and the recom-mended methodology developed, which may help to reveal the ac-tual difficulties for construction, operation and maintenance ofsulphate-reducing bioreactors.
12. Future scope for research
Different aspects need to be further investigated for better de-sign and operation of on-site passive treatment systems. Thedepletion rate of organic matter is a key problem. An improvedmethodical analysis of natural organic substrates is warranted toassess their ability to promote sulphate reduction and metal re-moval. Anaerobic degradation of complex organic carbon com-pounds to simpler molecules by other bacterial activity producedmay limit the rate at which substrates become available to SRP.More work must be conducted to understand and differentiatethe fundamental biochemical and microbiological reactions thatoccur in anaerobic bioreactors with complex natural organic sub-strates. Sulphate reduction seems to be controlled by cellulosedegradation and therefore, future research for exploring meansby which to enhance cellulose hydrolysis is needed. More workmust be conducted to understand and differentiate the fundamen-tal biochemical and microbiological reactions that occur in anaer-obic bioreactors with complex organic substrates. This might be akey step for the successful implementation of SRP-based ARDremediation systems. Today a routine and rigorous method ofanalysis of organic waste materials is still warranted to predict or-ganic substrate biodegradability (Gibert et al., 2004; Zagury et al.,2006).
Limited work has been carried out on the direct assessment ofthe ecotoxicological potential of biologically treated ARD waters.Correlation of metal speciation in the treated effluent and in thereactive mixture with toxic effects of treated waters could help im-prove our understanding of bioreactor systems. Sulphate-reducingprokaryotes can also be implicated in the bioremoval of heavy met-als as illustrated already. Hydrogen sulphide produced during thegrowth of SRP aids in the precipitation of the metal ions as theirmetallic sulphides. Future work is required to determine whether
recovery can be efficient enough to justify recycling the metal val-ues from the laden sludge (Cohen, 2006; Natarajan, 2009).
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