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Inhibition of anaerobic digestion process - A review


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Bioresource Technology 99 (2008) 4044–4064

Review

Inhibition of anaerobic digestion process: A review
Ye Chen, Jay J. Cheng *, Kurt S. Creamer
Department of Biological and Agricultural Engineering, North Carolina State University, Raleigh, NC 27695-7625, USA Received 5 April 2005; received in revised form 25 January 2007; accepted 25 January 2007 Available online 30 March 2007

Abstract Anaerobic digestion is an attractive waste treatment practice in which both pollution control and energy recovery can be achieved. Many agricultural and industrial wastes are ideal candidates for anaerobic digestion because they contain high levels of easily biodegradable materials. Problems such as low methane yield and process instability are often encountered in anaerobic digestion, preventing this technique from being widely applied. A wide variety of inhibitory substances are the primary cause of anaerobic digester upset or failure since they are present in substantial concentrations in wastes. Considerable research e?orts have been made to identify the mechanism and the controlling factors of inhibition. This review provides a detailed summary of the research conducted on the inhibition of anaerobic processes. The inhibitors commonly present in anaerobic digesters include ammonia, sul?de, light metal ions, heavy metals, and organics. Due to the di?erence in anaerobic inocula, waste composition, and experimental methods and conditions, literature results on inhibition caused by speci?c toxicants vary widely. Co-digestion with other waste, adaptation of microorganisms to inhibitory substances, and incorporation of methods to remove or counteract toxicants before anaerobic digestion can signi?cantly improve the waste treatment e?ciency. ? 2007 Elsevier Ltd. All rights reserved.
Keywords: Anaerobic digestion; Agricultural wastes; Industrial wastes; Inhibition; Municipal wastes

1. Introduction Anaerobic digestion involves the degradation and stabilization of organic materials under anaerobic conditions by microbial organisms and leads to the formation of biogas (a mixture of carbon dioxide and methane, a renewable energy source) and microbial biomass (Kelleher et al., 2000). Anaerobic treatment provides a method of reducing pollution from agricultural and industrial operations while at the same time o?setting the operations’ usage of fossil fuels. As one of the most e?cient waste and wastewater treatment technologies, anaerobic digestion has been widely used for the treatment of municipal sludge and limited application in the treatment of organic industrial wastes including fruit and vegetable processing wastes, packinghouse wastes, and agricultural wastes (Parkin and Miller, 1983). Anaerobic digestion o?ers numerous signi?*

Corresponding author. Tel.: +1 919 515 6733; fax: +1 919 515 7760. E-mail address: jay_cheng@ncsu.edu (J.J. Cheng).

cant advantages, such as low sludge production, low energy requirement, and possible energy recovery (Ghosh and Pohland, 1974; van Staikenburg, 1997). Compared to mesophilic digestion, thermophilic anaerobic digestion has additional bene?ts including a high degree of waste stabilization, more thorough destruction of viral and bacterial pathogens, and improved post-treatment sludge dewatering (Lo et al., 1985). In spite of these bene?ts, however, poor operational stability still prevents anaerobic digestion from being widely commercialized (Dupla et al., 2004). In anaerobic digestion, the acid forming and the methane forming microorganisms di?er widely in terms of physiology, nutritional needs, growth kinetics, and sensitivity to environmental conditions (Pohland and Ghosh, 1971). Failure to maintain the balance between these two groups of microorganisms is the primary cause of reactor instability (Demirel and Yenigun, 2002). Inhibitory substances are ¨ often found to be the leading cause of anaerobic reactor upset and failure since they are present in substantial concentrations in wastewaters and sludges. A wide variety of

0960-8524/$ - see front matter ? 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.biortech.2007.01.057

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Nomenclature CSTR continuously stirred tank reactor FA free ammonia HRT hydraulic retention time IC50, IC90, IC100 the toxicant concentration that causes 50%, 90%, and 100% reduction in cumulative methane production, respectively, over a ?xed period of exposure time LCFAs MPB SRB TAN VFAs UASB long chain fatty acids methane producing bacteria sulfate reducing bacteria total ammonia nitrogen volatile fatty acids up?ow anaerobic sludge blanket reactor

substances have been reported to be inhibitory to the anaerobic digestion processes. A material may be judged inhibitory when it causes an adverse shift in the microbial population or inhibition of bacterial growth. Inhibition is usually indicated by a decrease of the steady-state rate of methane gas production and accumulation of organic acids (Kroeker et al., 1979). The aim of this review is to present a detailed comparative summary of the previous and current research on the inhibition of anaerobic processes by various inorganic and organic substances, focusing on: (1) mechanisms of inhibition, (2) factors a?ecting inhibition, and (3) common operating problems encountered in waste treatment processes. 2. Inhibitors Literature on anaerobic digestion shows considerable variation in the inhibition/toxicity levels reported for most substances. The major reason for these variations is the complexity of the anaerobic digestion process where mechanisms such as antagonism, synergism, acclimation, and complexing could signi?cantly a?ect the phenomenon of inhibition. 2.1. Ammonia Ammonia is produced by the biological degradation of the nitrogenous matter, mostly in the form of proteins and urea (Kayhanian, 1999). The quantity of ammonia that will be generated from an anaerobic biodegradation of organic substrate can be estimated using the following stoichiometric relationship (Tchobanoglous et al., 1993): Ca Hb Oc Nd ? 4a ? b ? 2c ? 3d H2 O 4 4a ? b ? 2c ? 3d 4a ? b ? 2c ? 3d CH4 ? CO2 ! 8 8 ? dNH3

cipal forms of inorganic ammonia nitrogen in aqueous solution. FA has been suggested to be the main cause of inhibition since it is freely membrane-permeable (Kroeker et al., 1979; de Baere et al., 1984). The hydrophobic ammonia molecule may di?use passively into the cell, causing proton imbalance, and/or potassium de?ciency (Sprott and Patel, 1986; Gallert et al., 1998). Among the four types of anaerobic microorganisms, the methanogens are the least tolerant and the most likely to cease growth due to ammonia inhibition (Kayhanian, 1994). As ammonia concentrations were increased in the range of 4051–5734 mg NH3–N L?1, acidogenic populations in the granular sludge were hardly a?ected while the methanogenic population lost 56.5% of its activity (Koster and Lettinga, 1988). There is con?icting information in the literature about the sensitivity of aceticlastic and hydrogenotrophic methanogens. Some research based on the comparison of methane production and growth rate indicated that the inhibitory e?ect was in general stronger for the aceticlastic than for the hydrogenotrophic methanogens (Koster and Lettinga, 1984; Zeeman et al., 1985; Sprott and Patel, 1986; Bhattacharya and Parkin, 1989; Robbins et al., 1989; Angelidaki and Ahring, 1993; Borja et al., 1996a), while others observed the relatively high resistance of acetate consuming methanogens to high total ammonia nitrogen (TAN) levels as compared to hydrogen utilizing methanogens (Zeeman et al., 1985; Wiegant and Zeeman, 1986). Among the methanogenic strains commonly isolated from sludge digesters, i.e. Methanospirillum hungatei, Methanosarcina barkeri, Methanobacterium thermoautotrophicum, and Methanobacterium formicicum, Methanospirillum hungatei was the most sensitive, being inhibited at 4.2 g/L; the other three strains tested were resistant to ammonia levels higher than 10 g/L (Jarrell et al., 1987). 2.1.1. Factors controlling ammonia inhibition 2.1.1.1. Concentration. It is generally believed that ammonia concentrations below 200 mg/L are bene?cial to anaerobic process since nitrogen is an essential nutrient for anaerobic microorganisms (Liu and Sung, 2002). A wide range of inhibiting ammonia concentrations has been reported in the literature, with the inhibitory TAN concentration that caused a 50% reduction in methane production ranging from 1.7 to 14 g/L (Kroeker et al., 1979; van Vel-

?1?

Several mechanisms for ammonia inhibition have been proposed, such as a change in the intracellular pH, increase of maintenance energy requirement, and inhibition of a speci?c enzyme reaction (Whittmann et al., 1995). Ammonium ion ?NH? ? and free ammonia (FA) (NH3) are the two prin4

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sen, 1979; Braun et al., 1981; Parkin and Miller, 1983; de Baere et al., 1984; Zeeman et al., 1985; Hashimoto, 1986; Jarrell et al., 1987; Koster and Lettinga, 1988; Bhattacharya and Parkin, 1989; Hendriksen and Ahring, 1991; Angelidaki and Ahring, 1993; Angelidaki and Ahring, 1994; Soubes et al., 1994; Kayhanian, 1994; Borja et al., 1996b; Boardman and McVeigh, 1997; Gallert and Winter, 1997; Guerrero et al., 1997; Krylova et al., 1997; PoggiVaraldo et al., 1997; Chamy et al., 1998; Gallert et al., 1998; Hansen et al., 1998; Bujoczek et al., 2000; Sung and Liu, 2003). The signi?cant di?erence in inhibiting ammonia concentration can be attributed to the di?erences in substrates and inocula, environmental conditions (temperature, pH), and acclimation periods (van Velsen et al., 1979; de Baere et al., 1984; Hashimoto, 1986; Angelidaki and Ahring, 1994). 2.1.1.2. pH. During treatment of waste containing high concentrations of TAN, pH a?ects the growth of microorganisms as well as the composition of TAN (Kroeker et al., 1979; Hashimoto, 1983, 1984; Hansen et al., 1999). Since the FA form of ammonia has been suggested to be the actual toxic agent, an increase in pH would result in increased toxicity (Borja et al., 1996b) because of the shift to a higher FA to ionized ?NH? ? ammonia ratio at higher 4 pH. Process instability due to ammonia often results in volatile fatty acids (VFAs) accumulation, which again leads to a decrease in pH and thereby declining concentration of FA. The interaction between FA, VFAs and pH may lead to an ‘‘inhibited steady state’’, a condition where the process is running stably but with a lower methane yield (Angelidaki and Ahring, 1993; Angelidaki et al., 1993). Control of pH within the growth optimum of microorganisms may reduce ammonia toxicity (Bhattacharya and Parkin, 1989). Acidi?cation of crab wastewater has been reported to enhance UASB reactor performance, as indicated by the lower e?uent COD concentration (Boardman and McVeigh, 1997). Reducing pH from 7.5 to 7.0 during thermophilic anaerobic digestion of cow manure also increased the methane production by four times (Zeeman et al., 1985). During anaerobic digestion of liquid piggery manure (pH 8), VFAs accumulated to 316 mg/L. Adjustment of pH to 7.4 led to reutilization of VFAs and lowered VFAs concentrations to 20 mg/L. The better performance at pH 7.4 has been attributed to the relief of ammonia-induced inhibition at low pH (Braun et al., 1981). It should also be noted that both methanogenic and acidogenic microorganisms have their optimal pH. Failing to maintain pH within an appropriate range could cause reactor failure although ammonia is at a safe level (Kroeker et al., 1979). 2.1.1.3. Temperature. Both microbial growth rates and FA concentration are a?ected by temperature change. An increased process temperature in general has a positive e?ect on the metabolic rate of the microorganisms but also results in a higher concentration of FA. Several authors

have found that anaerobic fermentation of wastes with a high concentration of ammonia was more easily inhibited and less stable at thermophilic temperatures than at mesophilic temperatures (Braun et al., 1981; Parkin and Miller, 1983). Thermophilic digestion at 50 °C of cow manure with TAN above 3 g/L was found to be very di?cult (Hashimoto, 1983). A decrease in operating temperature from 60 °C to 37 °C in anaerobic digesters with a high ammonia concentration provided relief from inhibition caused by FA, as indicated by an increase in biogas yield (Angelidaki and Ahring, 1994; Hansen et al., 1999). Contrary to these ?ndings, Gallert and Winter (1997) studied the anaerobic digestion of organic wastes and reported that methane production was inhibited 50% by 0.22 g/L FA at 37 °C and by 0.69 g/L FA at 55 °C, indicating that thermophilic ?ora tolerated at least twice as much FA as compared to mesophilic ?ora. 2.1.1.4. Presence of other ions. Certain ions such as Na+, Ca2+, and Mg2+ were found to be antagonistic to ammonia inhibition, a phenomenon in which the toxicity of one ion is decreased by the presence of other ion(s) (McCarty and McKinney, 1961; Braun et al., 1981; Hendriksen and Ahring, 1991). Ammonia and sodium showed mutual antagonism, a situation where each ion can antagonize the toxicity produced by another ion. While 0.15 M ammonia reduced the methane production from acetic acid by 20%, addition of 0.002–0.05 M Na+ produced 5% more methane compared to that from the control (a sample without addition of inhibitor). Combination of Na+ and K+ or Na+ and Mg2+ resulted in around 10% increase in methane yield compared to that produced by Na+ alone (Kugelman and McCarty, 1964). The addition of 10% (w/v) phosphorite ore was also reported to stimulate biogas generation from poultry manure when NH? Cl was as high as 30 g/L 4 (Krylova et al., 1997). This stimulation e?ect of phosphorite can be partially attributed to the immobilization of the biomass on mineral particles, which prevented biomass washout from the reactor. Alleviation of ammonia inhibition was also thought to be partially due to the antagonistic e?ect provided by minerals in the phosphorite ore (K+, Ca2+, Mg2+). However, inhibition caused by more than 50 g/L of NH? Cl was irreversible and could not be elimi4 nated by addition of phosphorite (Krylova et al., 1997). 2.1.1.5. Acclimation. Acclimation is another factor that can in?uence the degree of ammonia inhibition. One of the ?rst reports dealing with adaptation of methanogens to ammonia by exposing them to slowly increasing concentrations was the sludge digestion study of Melbinger and Donnellon (1971). At present, adaptation of methanogens to a wide variety of potentially inhibitory substances has been reported (Parkin and Miller, 1983; Speece, 1983; Speece and Parkin, 1983). The adaptation may be the result of internal changes in the predominant species of methanogens, or of a shift in the methanogenic population (Zeeman et al., 1985).

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Once adapted, the microorganisms can retain viability at concentrations far exceeding the initial inhibitory concentrations (Kroeker et al., 1979; Parkin and Miller, 1983; Bhattacharya and Parkin, 1989; Angelidaki and Ahring, 1993). Koster and Lettinga (1988) reported that while unacclimated methanogens failed to produce methane at 1.9–2 g N/L, they produced methane at 11 g N/L after adaptation. Hashimoto (1986) observed that ammonia inhibition began at about 2.5 g/L and 4 g/L for unacclimated and acclimated thermophilic methanogens, respectively. Successful operation of anaerobic ?lters has been achieved at 6 g/L and 7.8 g/L after adaptation (Parkin et al., 1983; de Baere et al., 1984). Parkin and Miller (1983) reported that levels as high as 8–9 g/L of TAN could be tolerated with no signi?cant decrease in methane production after acclimation. The experiments clearly demonstrated the possibility of obtaining stable digestion of manure with ammonia concentrations exceeding 5 g N/L after an initial adaptation period. However, the methane yield was lower than that for reactors with a lower ammonia load (Koster and Lettinga, 1988; Borja et al., 1996a). 2.1.2. Methods to counteract ammonia inhibition To remove ammonia from the substrate, two physical– chemical methods can be utilized: air stripping and chemical precipitation. Both have been proven to be technically feasible at high ammonia concentrations and in a complex wastewater matrix (Kabdasli et al., 2000). A common approach to ammonia inhibition relies on dilution of the manure to a total solid level of 0.5–3.0%. However, the resulting increase in waste volume that must be processed makes this method economically unattractive (Callaghan et al., 1999). Various types of inhibition can be counteracted by increasing the biomass retention in the reactor. It was found that the methane yield in a CSTR could be increased by switching o? the stirrer half an hour before and after substrate addition. This operation increased biomass retention due to improved sedimentation resulting in an e?uent with a reduced concentration of biomass solids. This method, where particles within the reactor were allowed to settle, was promising since it was easy and economical to achieve (Hansen et al., 1998). Immobilizing the microorganisms with di?erent types of inert material (clay, activated carbon, zeolite) has been demonstrated to reduce inhibition of the biogas process and make the process more stable (Angelidaki et al., 1990; Nakhla et al., 1990; Borja et al., 1993; Hanaki et al., 1994; Hansen et al., 1998). Addition of ionic exchangers or adsorbants which can remove inhibitors mitigates the ammonia inhibition (Borja et al., 1996a). Natural zeolite and glauconite show high selectivity for ammonium ion and can be used as an ionic exchanger for ammonia (Borja et al., 1996a; Hansen et al., 1998). When treating swine manure, addition of activated carbon at concentrations equal to 2.5% (w/w) or higher or FeCl2 removed most of the sul?de in solution. Although activated carbon did not adsorb ammonia, it reduced inhibition of

ammonia by removing sul?de, which otherwise would act synergistically with ammonia (Hansen et al., 1999). Addition of antagonistic cations such as Mg2+ or Ca2+ stabilizes anaerobic degradation (McCarty and McKinney, 1961). The positive e?ect of zeolite on the anaerobic process could partially be attributed to the presence of cations such as Ca2+ and Na+ that have been shown to counteract the inhibitory e?ect of ammonia (Borja et al., 1996a). 2.2. Sul?de Sulfate is a common constituent of many industrial wastewaters (O’Flaherty et al., 1998a). In anaerobic reactors, sulfate is reduced to sul?de by the sulfate reducing bacteria (SRB) (Koster et al., 1986; Hilton and Oleszkiewicz, 1988). Sulfate reduction is performed by two major groups of SRB including incomplete oxidizers, which reduce compounds such as lactate to acetate and CO2, and complete oxidizers, which completely convert acetate to CO2 and HCO? . Two stages of inhibition exist 3 as a result of sulfate reduction. Primary inhibition is due to competition for common organic and inorganic substrates from SRB, which suppresses methane production (Harada et al., 1994). Secondary inhibition results from the toxicity of sul?de to various bacteria groups (Anderson et al., 1982; Oude Elferink et al., 1994; Colleran et al., 1995; Colleran et al., 1998). 2.2.1. Competition of SRB and other anaerobes SRB are very diverse in terms of their metabolic pathways (Oude Elferink et al., 1994). Compounds which can be completely or partially degraded by SRB include branched-chain and long chain fatty acids, ethanol and other alcohols, organic acids, and aromatic compounds (Oude Elferink et al., 1994). Laanbroek et al. (1984) ranked the a?nity of SRB for reduced substrates in the order of H2 > propionate > other organic electron donors. Because of the variety in substrate utilization exhibited by SRB, they compete with several di?erent types of microorganisms involved in anaerobic digestion. SRB may compete with methanogens, acetogens, or fermentative microorganisms for available acetate, H2, propionate, and butyrate in anaerobic systems (McCartney and Oleszkiewicz, 1993; Colleran et al., 1995). The outcome of the competition between SRB and other anaerobic microorganisms determines the concentration of sul?de in the reactor system. Sul?de is toxic to methanogens as well as to the SRB themselves (Winfrey and Zeikus, 1977; Karhadkar et al., 1987; McCartney and Oleszkiewicz, 1991; Reis et al., 1992; Okabe et al., 1995). Thus the concentration of sul?de and the susceptibility of anaerobes to sul?de feed back into the competition between SRB and other anaerobes. 2.2.1.1. Competition between SRB and hydrolytic and acidogenic bacteria. SRB do not degrade natural biopolymers such as starch, glycogen, protein, or lipids and thus

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depend on the activity of other organisms for providing them with degradation products (Hansen, 1993). Consequently, competition does not occur in the hydrolysis stage. Although a few strains of SRB have been shown to utilize sugars and amino acids as substrate (Klemps et al., 1985; Min and Zinder, 1990), vigourous growth of SRB on typical acedogenic substrates is not common (Hansen, 1993). It is generally agreed that SRB cannot e?ectively compete against the fast-growing fermentative microorganisms involved in monomer degradation (Postgate, 1984). O’Flaherty et al. (1999) conducted tests to detect SRB in an anaerobic digester fed with glucose and lactose. No change of their degradation rates was detected upon addition of sulfate, indicating that SRB species did not play any substantial role in the degradation of glucose and lactose. 2.2.1.2. Competition between SRB and acetogens. From a purely thermodynamic and kinetic standpoint, SRB should out-compete other anaerobes for substrate (Oude Elferink et al., 1994; Colleran et al., 1995; O’Flaherty et al., 1998a). In practice, however, factors such as COD=SO2? ratio, the 4 relative population of SRB and other anaerobes, and the sensitivity of SRB and other anaerobes to sul?de toxicity in?uence the competition. As a result, the literature on anaerobic digestion of sulfate-containing wastewaters is highly complex and often contradictory. Propionate is a key intermediate in anaerobic digestion and a substrate for all SRB. Degradation of propionate by SRB involves an incomplete conversion to acetate (O’Flaherty et al., 1998a). SRB show a higher a?nity for propionate and faster growth rates than the propionateutilizing syntrophic species (Parkin et al., 1990; Uberoi and Bhattacharya, 1995; Omil et al., 1996a). The Ks and lmax values for SRB were 0.15 d?1 and 23 mg/L in full scale anaerobic digester while Ks and lmax values for syntrophic bacteria were 0.05 d?1 and 34 mg/L, respectively (O’Flaherty et al., 1997, 1998b). As a result, sul?dogenic oxidation of propionate should be favored over the syntrophic route (Colleran et al., 1995). Several studies using various anaerobic systems and sludges have con?rmed the importance of SRB in the degradation of propionate, indicating that su?dogenic oxidation is the key degradation pathway of this substrate (Mulder, 1984; Ukei et al., 1988; Qatibi et al., 1990; Hepner et al., 1992; Colleran et al., 1994, 1998; O’Flaherty et al., 1997, 1998a). Butyrate and ethanol are also important fermentation intermediates in anaerobic digestion. Butyrate was utilized exclusively by SRB in a UASB reactor fed mixed VFAs and sucrose ?COD=SO2? ? 0:5? (Visser et al., 1993). In 4 another hybrid reactor, both sul?dogenic and methanogenic anaerobes were present at a COD=SO2? ratio of 3 4 and 5.6, indicating that competition occurred among SRB and other anaerobes for butyrate and ethanol (Colleran et al., 1998; O’Flaherty et al., 1998a). The e?ective competition of non-SRB was attributed to the lower a?nity of SRB for butyrate and ethanol (Laanbroek et al., 1984; Overmeire et al., 1994).

2.2.1.3. Competition between SRB and hydrogenotrophic methanogens. From thermodynamic and substrate a?nity considerations, H2-oxidizing SRB should e?ectively outcompete hydrogenotrophic methanogens under the conditions prevailing in anaerobic digesters (Zinder, 1993). This view was supported by data reported previously indicating that in reactors treating sulfate-containing wastewaters, H2 oxidation is almost exclusively catalyzed by SRB (Rinzema and Lettinga, 1988; Visser et al., 1993; Alphenaar et al., 1993; Harada et al., 1994; Uberoi and Bhattacharya, 1995; Omil et al., 1996a; Colleran et al., 1998; O’Flaherty et al., 1999). Methanogenesis appeared to occur simultaneously with sulfate-reduction, but methanogens could not compete for H2 with the SRB (Oremland and Taylor, 1978). The predominance of SRB in H2 utilization has been related to the more favorable kinetic parameters for SRB. Hydrogenotrophic SRB have a lower hydrogen threshold concentration than hydrogenotrophic methanogens (Oude Elferink et al., 1994; Colleran et al., 1995). Temperature has been reported to impact the outcome of the competition between SRB and hydrogenotrophic methane producing bacteria (MPB). SRB were dominant at mesophilic condition (37 °C), while MPB outcompeted SRB at thermophilic conditions (55 °C). An explanation for this di?erence was not o?ered (Colleran and Pender, 2002). 2.2.2. Competition between SRB and aceticlastic methanogens Literature data on the outcome of competition between SRB and MPB for acetate are contradictory, with some authors reporting successful competition of SRB (Rinzema and Lettinga, 1988; Alphenaar et al., 1993; Stucki et al., 1993; Gupta et al., 1994), whereas others reported dominance of MPB (Rinzema et al., 1988; Isa et al., 1986a,b; Visser et al., 1993; Omil et al., 1996a; Oude Elferink et al., 1994; Colleran et al., 1998; O’Flaherty et al., 1998a; De Smul et al., 1999; Colleran and Pender, 2002). Various mechanisms have been proposed to explain the observed discrepancies. Choi and Rim (1991) attributed the outcome of the competition to the COD=SO2? ratio. Ace4 ticlastic MPB predominated when the COD=SO2? was 4 above 2.7; SRB predominated when this ratio was below 1.7. Active competition occurred between these ratios. O’Flaherty et al. (1998b) correlated the performance of MPB and SRB to the di?erent growth properties at di?erent pH values. Oude Elferink et al. (1994) observed that the initial population of SRB played a role in the competition between SRB and MPB. They calculated that by starting with a ratio of aceticlastic MPB/SRB of 104:1 and with a biomass retention time in the reactor of 0.02 d?1, it would take one year before the number of SRB equaled that of the MPB in the reactor. Isa et al. (1986a,b) attributed the successful competition of MPB to their superior attachment capabilities. In ?xed?lm reactors, better attachment of microorganisms can e?ectively prevent biomass washout (Omil et al., 1996a).

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Colleran and Pender (2002) concluded that aceticlastic methanogens predominated because SRB have a lower a?nity for acetate than for other substrates. Under sulfate-limiting conditions, acetate was believed to be the least favored substrate for sulfate reduction (Uberoi and Bhattacharya, 1995). However, the dominance of SRB in acetate degradation was attributed to the kinetic advantages of SRB over MPB (Rinzema and Lettinga, 1988; Gupta et al., 1994; Harada et al., 1994). Alphenaar et al. (1993) attributed the higher extent of organic removal by SRB to the long HRT used in the UASB/CSTR reactor, which led to the washout of the dispersed growing MPB. 2.2.3. Sul?de inhibition towards di?erent trophic groups There is considerable confusion in the literature with respect to the nature of sul?de toxicity and the e?ect of different sul?des on microorganisms. Tursman and Cork (1988) reported that H2S was the toxic form of sul?de since it can di?use into the cell membrane. Once inside the cytoplasm, H2S may be inhibitory by denaturing native proteins through the formation of sul?de and disul?de cross-links between polypeptide chains (Conn et al., 1987), interfering with the various coenzyme sul?de linkages, and interfering with the assimilatory metabolism of sulfur (Vogels et al., 1988). This theory was supported by the studies of Speece (1983). By contrast, McCartney and Oleszkiewicz (1991) observed that sul?de toxicity increased with increasing pH. Other studies on sul?de inhibition indicated that more than one inhibition threshold might be present under di?erent conditions. Koster et al. (1986) observed a high correlation between the unionized sul?de concentration and the maximum speci?c aceticlastic methanogenic activity in the pH range of 6.4–7.2. At pH 7.8– 8.0, total sul?de concentration dictated the degree of inhibition. O’Flaherty et al. (1998b) observed that sul?de inhibition for all of the groups of bacteria was related to the unionized sul?de concentration in the pH range of 6.8– 7.2 and total sul?de concentrations above pH 7.2. Hilton and Oleszkiewicz (1990) showed that inhibition of SRB and MPB was correlated with the total sul?de and unionized sul?de concentration, respectively. There is also considerable discrepancy in the literature with respect to the levels of sul?de which can cause inhibition to various trophic groups and which steps of the anaerobic transformation are most adversely a?ected by sul?de. Much of the data reported in the literature was obtained by adding sul?de to a system rather than by feeding sulfate. Thus, the interaction between SRB and nonSRB was not considered (Parkin et al., 1990). In addition, information on pH was rarely included, making it impossible to draw reliable conclusions on the inhibition concentrations. The inhibitory sul?de levels reported in the literature were in the range of 100–800 mg/L dissolved sul?de or approximately 50–400 mg/L undissociated H2S (Parkin et al., 1990). Fermentative microorganisms which are responsible for the breakdown of monomers into smaller products were less a?ected by sul?de toxicity than SRB

or MPB (McCartney and Oleszkiewicz, 1991; Maillacheruvu et al., 1993). Acetogens were found to be less susceptible to sul?de inhibition than MPB; toxicity thresholds for acetogens were comparable with those of the SRB (O’Flaherty et al., 1998b). Sulfur is a required nutrient for methanogens (O’Flaherty et al., 1999). It has been shown that the sulfur content of methanogens was higher than in other groups of microorganisms generally found in anaerobic systems (Speece, 1983). The optimal level of sulfur reported in the literature varies from 1 to 25 mg S/L (Scherer and Sahm, 1981). The levels reported in the literature for inhibition of MPB also vary, with IC50 values of 50–125 mg H2S/L at pH 7–8 for suspended sludge and 250 mg H2S/L and 90 mg H2S/L at pH 6.4–7.2 and pH 7.8–8.0, respectively (Parkin et al., 1983; Koster et al., 1986; Oleskiewicz et al., 1989; McCartney and Oleszkiewicz, 1993; Maillacheruvu et al., 1993; O’Flaherty et al., 1998a). 2.2.4. Sulfate/sul?de toxicity control Several processes can be applied to promote the removal of dissolved sulfate. One method to prevent sul?de toxicity is to dilute the wastewater stream, although in general this approach is considered undesirable because of the increase in the total volume of wastewater that must be treated. An alternative way to reduce the sul?de concentration in an anaerobic treatment system is by incorporating a sul?de removal step in the overall process. Sul?de removal techniques include physico-chemical techniques (stripping), chemical reactions (coagulation, oxidation, precipitation), or biological conversions (partial oxidation to elemental sulfur) (Oude Elferink et al., 1994; Song et al., 2001). Adaptation of the MPB to free H2S, particularly in reactors with ?xed biomass, could increase the tolerance of MPB to sul?de. Isa et al. (1986a) reported that acclimated aceticlastic and hydrogenotrophic MPB were only slightly inhibited at more than 1000 mg/L free H2S. 2.3. Light metals ions (Na, K, Mg, Ca, and Al) Salt toxicity has been studied in the biological ?eld for several decades. High salt levels cause bacterial cells to dehydrate due to osmotic pressure (de Baere et al., 1984; Yerkes et al., 1997). Although the cations of salts in solution must always be associated with the anions, the toxicity of salts was found to be predominantly determined by the cation (McCarty and McKinney, 1961). The light metal ions including sodium, potassium, calcium, and magnesium are present in the in?uent of anaerobic digesters. They may be released by the breakdown of organic matter (such as biomass), or added as pH adjustment chemicals (Grady et al., 1999). They are required for microbial growth and, consequently, a?ect speci?c growth rate like any other nutrient. While moderate concentrations stimulate microbial growth, excessive amounts slow down the growth, and even higher concentrations can cause severe inhibition or toxicity (Soto et al., 1993a).

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2.3.1. Aluminum Information in the literature about the e?ect of aluminum on anaerobic digestion is minimal. The mechanism of aluminum inhibition was reported to be due to its competition with iron and manganese or to its adhesion to the microbial cell membrane or wall, which may a?ect microbial growth (Cabirol et al., 2003). Both acetogenic and methanogenic microorganisms were inhibited by the addition of Al(OH)3. After exposed to 1000 mg/L Al(OH)3 for 59 days, the speci?c activity of methanogenic and acetogenic microorganisms decreased by 50% and 72%, respectively (Cabirol et al., 2003). Jackson-Moss and Duncan (1991) reported that 2,500 mg/L Al3+ could be tolerated by anaerobes after acclimation. 2.3.2. Calcium Calcium is known to be essential for the growth of certain strains of methanogens (Murray and Zinder, 1985). It is also important in the formation of microbial aggregates (Thiele et al., 1990; Huang and Pinder, 1995). Excessive amounts of calcium lead to precipitation of carbonate and phosphate, which may result in (i) scaling of reactors and pipes, (ii) scaling of biomass and reduced speci?c methanogenic activity, (iii) loss of bu?er capacity and essential nutrients for anaerobic degradation (Keenan et al., 1993; El-Mamouni et al., 1995; van Langerak et al., 1998). Very little is known about the toxicity of Ca2+ in the anaerobic system. Jackson-Moss et al. (1989) observed that Ca2+ concentrations of up to 7000 mg/L had no inhibitory e?ect on anaerobic digestion. A large proportion of the Ca2+ passed through the digester and was present in the e?uent. Kugelman and McCarty (1964) reported a much lower toxicity threshold. They showed that the optimum Ca2+ concentration for methanation of acetic acid was 200 mg/L. Ca2+ was moderately inhibitory at a concentration of 2500–4000 mg/L, but was strongly inhibitory at a concentration of 8000 mg/L. Addition of calcium can have a positive impact on reactors in which retention of biomass is desired. Addition of Ca2+ increased the accumulation of bio?lm when Ca2+ concentration in the feed was below 120 mg/L. For Ca2+ concentrations higher than 120 mg/ L, an accumulation of minerals and a decrease in water content in the bio?lm caused an inhibition of cellular metabolism (Huang and Pinder, 1995). Similarly in UASB reactors, low Ca2+ concentrations from 100 to 200 mg/L were reported to be bene?cial for sludge granulation (Cail and Barford, 1985; Mahoney et al., 1987; Yu et al., 2001), whereas high Ca2+ concentrations (greater than 300 mg/L) were reported to be detrimental (Hulsho? Pol et al., 1983; Thiele et al., 1990; Yu et al., 2001). Calcium carbonate precipitation could also impact biomass activity. Calcium carbonate precipitation is dependent on the Ca2+ concentration and on the COD removal e?ciency (van Langerak et al., 1998). The impact of precipitate on the biomass activity is complex. Highly scaled biomass is less active because of mass transfer limi-

tations. However, active biomass could be formed in thin bio?lms on the surface of the precipitates. The overall activity of the biomass would be the average of the two e?ects (van Langerak et al., 1998). 2.3.3. Magnesium The optimal Mg2+ concentration was reported to be 720 mg/L for the anaerobic bacterium Methanosarcina thermophila TM1 and a Methanosarcinae-dominated UASB reactor (Ahring et al., 1991; Schmidt and Ahring, 1993). Cultures could be adapted to 300 mM Mg2+ without a change in growth rate, but growth ceased at 400 mg/L Mg2+ (Schmidt and Ahring, 1993). Magnesium ions at high concentrations have been shown to stimulate the production of single cells (Harris, 1987; Xun et al., 1988; Schmidt and Ahring, 1993). The high sensitivity of single cells to lysis is an important factor in the loss of aceticlastic activity in anaerobic reactors. 2.3.4. Potassium Maintenance of high levels of potassium in culture media or in a digester is undesirable since pure culture studies have shown that high levels of extracellular potassium (1.0 M) lead to a passive in?ux of potassium ions that neutralize the membrane potential (Jarrell et al., 1984). In addition, potassium is one of the best extractants for metals bound to the exchangeable sites in sludge. Ilangovan and Noyola (1993) observed the increase of micronutrients (Cu2+, Zn2+, Ni2+, Mo2+, Co2+) in a UASB reactor treating molasses stillage containing a high concentration of potassium. The removal of the essential micronutrients from active sludge was believed to be responsible for the low activity of anaerobic methanogenic population. The toxic e?ect of potassium is rarely referenced in the literature. Low concentrations of potassium (less than 400 mg/L) were observed to cause an enhancement in performance in both the thermophilic and mesophilic ranges while at higher concentrations there was an inhibitory e?ect that was more pronounced in the thermophilic temperature range. Slug feed studies, in which the concentration of the cation was suddenly increased in actively fermenting cultures, were conducted to determine the toxicity of individual cations (Kugelman and McCarty, 1964). It was observed that 0.15 M K+ caused 50% inhibition of acetate-utilizing methanogens. A series of studies have shown that K+ inhibits the thermophilic digestion of simulated co?ee wastes (Fernandez and Forster, 1993, 1994; Shi and Forster, 1994). Information about the sensitivity of di?erent groups of microorganisms to potassium is con?icting. The results of batch tests using acetate as the carbon source showed that the gas production from both the control and samples with elevated K+ were identical, indicating that the inhibition could be at the acidogenic stage (Fernandez and Forster, 1994). Mouneimne et al. (2003) investigated the biotoxicity of potassium using acetate and glucose as substrates and anaerobic sludge as inoculum. The IC50 for acetate-utilizing microorganisms was

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found to be 0.74 mol/L. However, the degradation rates of glucose were virtually una?ected by potassium, indicating that the acetate-utilizing microorganisms exhibited a greater sensitivity to the toxic e?ects of cations than the acid-forming ones. Sodium, magnesium, and ammonium were observed to mitigate potassium toxicity, with sodium producing the best results. Combinations of cations produce antagonism superior to that of single cations. The best results were obtained for combinations of sodium and calcium, and combinations of sodium, calcium and ammonia (Kugelman and McCarty, 1964). 2.3.5. Sodium Wastewaters with high concentrations of sodium are produced in the food processing industry (Soto et al., 1991). Mesophilic and thermophilic anaerobic ?lters treating e?uents from a mussel cooking factory were compared. The mesophilic reactor exhibited better performance than the thermophilic reactor, which was attributed to the more rapid adaptation of mesophilic sludges to the high salinity of the wastewater (Soto et al., 1991; Soto et al., 1992). In comparing VFA-degrading bacteria, sodium was more toxic to propionic acid-utilizing microorganisms than to acetic acid-utilizing ones (Soto et al., 1993b). This result was in agreement with the ?ndings of Liu and Boone (1991), who found the NaCl toxicity decreased in the order of lignocellulose-degrading > acetate-utilizing > propionate-utilizing > H2/CO2-utilizing organisms. 2.3.5.1. Concentration. At low concentrations, sodium is essential for methanogens, probably because of its role in the formation of adenosine triphosphate or in the oxidation of NADH (Dimroth and Thomer, 1989). McCarty (1964) reported sodium concentrations in the range of 100–200 mg/L to be bene?cial for the growth of mesophilic anaerobes. According to Kugelman and Chin (1971), the optimal sodium concentration for mesophilic aceticlastic methanogens in waste treatment processes was 230 mg Na+/L. The optimal growth conditions for mesophilic hydrogenotrophic methanogens reportedly occurred at 350 mg Na+/L (Patel and Roth, 1977). At high concentrations, sodium could readily a?ect the activity of microorganisms and interfere with their metabolism (Kugelman and McCarty, 1964; Rinzema et al., 1988; Gourdon ? et al., 1989; Balsleve-Olsen et al., 1990; Mendez et al., 1995). The level of inhibition depends on the concentration of sodium ions. An early study reported sodium concentrations ranging from 3500 to 5500 mg/L to be moderately and 8000 mg/L to be strongly inhibitory to methanogens at mesophilic temperatures (McCarty, 1964). The IC50 for sodium inhibition has been reported to be 5.6–53 g/L, depending on the adaptation period, antagonistic/synergistic e?ects, substrate, and reactor con?guration (Patel and Roth, 1977; Rinzema et al., 1988; Liu and Boone, 1991; Soto et al., 1993b; Feijoo et al., 1995; Omil et al.,

? 1995a,b; Aspe et al., 1997; Kim et al., 2000; Vallero et al., 2002; Chen et al., 2003; Vallero et al., 2003a,b). 2.3.5.2. Acclimation. Acclimation of methanogens to high concentrations of sodium over prolonged periods of time could increase the tolerance and shorten the lag phase before methane production begins (de Baere et al., 1984; Feijoo et al., 1995; Omil et al., 1995a,b, 1996b; Chen et al., 2003). The tolerance is related to the Na+ concentration the methanogens acclimated to and the time of exposure. The IC100 of methanogens increased from 12.7 to 22.8 g/L when the methanogens were acclimated to 4.1 and 12.0 g/L of Na+, respectively (Chen et al., 2003). Inoculum showed increased tolerance for sodium as the time of ? acclimation increases. Mendez et al. (1995) reported that IC90 of inocula was 12.0 g/L when sludge was taken from anaerobic reactor after one day of acclimation and greater than 17.0 g/L when the acclimation period was 719 days. Anaerobic ?lter sludge that treating high salinity wastewater for 2 years also exhibited better performance than sludge that had been sampled from a central activity digester employed for wastewater treatment for 1 year (Feijoo et al., 1995). Application of experimental results obtained from a batch reactor to a continuous reactor often overestimated the sensitivity of the microorganisms. This may be attributable to the sudden change in sodium levels that microorganisms were exposed to in batch testing, giving them minimum time to adapt. In continuous testing, sodium concentration was often increased gradually, allowing su?cient time for the microorganisms to adapt (Feijoo et al., 1995). Contrary to the ?ndings in the previous section, Rinzema et al. (1988) found no adaptation of Methanothrix sp. to high sodium concentrations after 12 weeks. Similarly, when treating methanol in a sulfate-reducing reactor, stepwise increases in NaCl could not increase the tolerance of SRB to sodium, indicating that the adaptation of thermophilic, sul?dogenic methanol-degrading bacteria to a high NaCl environment was unlikely to occur (Vallero et al., 2002, 2003a,b). 2.3.5.3. Antagonistic/synergistic e?ects. Microorganisms accumulate cations and/or low-molecular-weight organic compounds, known as compatible solutes, when the extracellular solute concentration exceeds that of the cell cytoplasm (Lai and Gunsalus, 1992). The role of compatible solutes in osmoregulation was recognized and the antagonistic e?ect of cations and betaine towards sodium has been investigated. Kugelman and McCarty (1964) showed that a combination of potassium and calcium signi?cantly increased the antagonism over that achieved by potassium alone. An anaerobic toxicity assay of sludge from musselprocessing wastewater also con?rmed that when using the e?uent of the anaerobic ?lter as assay medium, the tolerance to sodium was highly increased compared to distilled water (Soto et al., 1993b). This e?ect was attributed to the antagonism exerted by the presence of sea salts, probably

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K+, Mg2+, and Ca2+. Information in the literature about the e?ect of Mg2+ is con?icting. Ahring et al. (1991) reported an antagonistic e?ect of Mg2+ for Na+. The inhibition by Na+ was directly related to the Mg2+ concentration. When the Mg2+ was 0.05 mM or less, 0.35 M Na+ completely inhibited growth. More Na+ was required for inhibition at higher Mg2+ concentrations. However, beyond 0.01 M/L, Mg2+ was reported to start showing a synergistic e?ect towards Na+ (Kugelman and McCarty, 1964). Ca2+ and NH? also showed synergistic e?ects 4 towards Na+ (Kugelman and McCarty, 1964). The antagonistic e?ect of the compatible solute betaine ((CH3)3N+CH2COO?) towards sodium toxicity was investigated by Yerkes et al. (1997). Bacteria subjected to salinity stress have been shown to accumulate betaine in proportion to the salinity of the medium (Poukomailian and Booth, 1992). Concentrations of betaine as low as 1 mM have been shown to be e?ective in reducing the toxicity of high concentrations of sodium by reducing acclimation or lag time, increasing substrate uptake rate, and increasing gas production (Yerkes et al., 1997). 2.4. Heavy metals Heavy metals can be present in signi?cant concentrations in municipal sewage and sludge. The heavy metals identi?ed to be of particular concern include chromium, iron, cobalt, copper, zinc, cadmium, and nickel (Jin et al., 1998). A distinguishing feature of heavy metals is that, unlike many other toxic substances, they are not biodegradable and can accumulate to potentially toxic concentrations (Sterritt and Lester, 1980). In one extensive study of anaerobic digester performance, it was found that heavy metal toxicity is one of the major causes of digester upset or failure (Swanwick et al., 1969). The toxic e?ect of heavy metals is attributed to disruption of enzyme function and structure by binding of the metals with thiol and other groups on protein molecules or by replacing naturally occurring metals in enzyme prosthetic groups (Vallee and Ulner, 1972). 2.4.1. Factors controlling heavy metal inhibition Many heavy metals are part of the essential enzymes that drive numerous anaerobic reactions. Analysis of ten methanogenic strains showed the following order of heavy metal composition in the cell: Fe ) Zn P Ni > Co = Mo > Cu (Takashima and Speece, 1989). Whether heavy metals would be stimulatory or inhibitory to anaerobic microorganisms is determined by the total metal concentration, chemical forms of the metals, and process-related factors such as pH and redox potential (Mosey et al., 1971; Lin and Chen, 1999; Zayed and Winter, 2000). It is generally believed that acidogens are more resistant to heavy metal toxicity than methanogens (Zayed and Winter, 2000). However, Hickey et al. (1989) have speculated that some trophic group(s) or organisms within the anaerobic consortia in digesters might be more severely inhibited by

a pulsed addition of heavy metals than the methanogenic populations. 2.4.1.1. Chemical forms of heavy metal. Because of the complexity of the anaerobic system, heavy metals may be involved in many physico-chemical processes including (1) precipitation as sul?de (except Cr), carbonate and hydroxides (Lawrence and McCarty, 1965; Mosey et al., 1971), (2) sorption to the solid fraction, either biomass or inert particulate matter (Shen et al., 1993; Shin et al., 1997), and (3) formation of complexes in solution with intermediates and product compounds produced during digestion (Hayes and Theis, 1978; Hickey et al., 1989; Callander and Barford, 1983a,b). Among these metal forms, only metals in soluble, free form are toxic to the microorganisms (Lawrence and McCarty, 1965; Mosey and Hughes, 1975; Oleszkiewicz and Sharma, 1990). Several studies have con?rmed that the heavy metal toxicity correlated better to the metal’s free ionic concentration (determined through a combination of dialysis and ion exchange) than to its total concentration (Bhattacharya and Sa?erman, 1989; Bhattacharya et al., 1995a,b). In previous reports, the various physico-chemical forms of a particular heavy metal were rarely distinguished due to the complex interactions between the heavy metals and anaerobic sludge and/or lack of analytical techniques for separating metal species (Gould and Genetelli, 1978; Hayes and Theis, 1978; Oleszkiewicz and Sharma, 1990; Zayed and Winter, 2000). This is one factor that explains the wide variation in reported toxic concentrations of heavy metals. 2.4.1.2. Concentrations. In addition to physico-chemical form, di?erences in substrate, bacteria genre, and environmental factors also explain the wide variation (from several to several hundreds of mg/L) in both the reported dosages of heavy metals and their relative toxicity (Lawrence and McCarty, 1965; Hickey et al., 1989; Bhattacharya et al., 1995a; Jin et al., 1998; Lin and Chen, 1999; Zayed and Winter, 2000). Moreover, the operating solids level signi?cantly impacts the heavy metal toxicity in anaerobic digesters by providing protection from metal inhibition. It has been suggested that inhibition due to heavy metals would be more comparable if metal dosage was expressed as milligram of metal per gram of volatile solids (Hickey et al., 1989). Unfortunately, most of the literature only reported the inhibition concentration values in mg/L, which makes the comparison of inhibition concentrations more di?cult. Heavy metal concentrations that caused 50% inhibition of methanogenesis during whey methanation indicated that toxicity decreased in the order of Cu > Zn > Ni. Similar results were obtained by Lin (1992, 1993) and Lin and Chen (1999). This is, however, not surprising since Zn and Ni are components of several enzymes in anaerobic microorganisms (Nies, 1999). The relative sensitivity of acidogenesis and methanogenesis to heavy metals is Cu > Zn > Cr > Cd > Ni > Pb and Cd > Cu > Cr > Zn > Pb > Ni, respectively (Lin, 1992, 1993). The relative

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toxicity of four metals to the anaerobic digestion of sewage sludge was reported to be Cr > Ni > Cu > Zn (Wong and Cheung, 1995). 2.4.1.3. Antagonistic and synergistic e?ects. Industrial wastewaters or sludges generally contain many kinds of heavy metals which cause synergistic or antagonistic e?ects on anaerobic digestion. The level of inhibition is determined by the species and the ratio of the individual components. Although toxicity of most mixed heavy metals such as Cr–Cd, Cr–Pb, Cr–Cd–Pb, and Zn–Cu–Ni was synergistic (Lin, 1992), some of the metal mixtures showed antagonistic inhibition (Lin, 1993). In a variety of aerobic, facultative and anaerobic studies reviewed by Babich and Stotzky (1983), Ni was shown to act synergistically in Ni–Cu, Ni–Mo–Co, and Ni–Hg systems; antagonistically in Ni–Cd, Ni–Zn systems. Ahring and Westermann (1985) found that Ni decreased the toxicity of Cd and Cu. 2.4.2. Detoxi?cation of heavy metals The most important methods for mitigating heavy metal toxicity are precipitation, sorption and chelation by organic and inorganic ligands (Oleszkiewicz and Sharma, 1990). Sul?de has been the main agent used to precipitate heavy metals. Reactor recovery from 20 mg/L of copper exposure was observed when sul?de was added after copper exposure. The addition of sul?de before copper exposure can signi?cantly shorten the time required for recovery (Jin et al., 1998; Zayed and Winter, 2000). However, caution must be exercised since excess sul?de can also be an important inhibitor to methanogens (Anderson et al., 1982). Excessive quantities of sul?de might be minimized by adding ferrous sulfate, which has the highest solubility of all toxic heavy metals. Heavy metals would combine with the sul?de in FeS, releasing Fe2+, which is relatively non-toxic up to several hundred mg/L. Presence of a solid phase in the reactor provides protection for anaerobic microorganisms from heavy metal inhibition (Jarrell et al., 1987). The protection e?ect is proportional to the surface area or the amount of solids. The mechanism is believed to be chemisorption (Gould and Genetelli, 1984; Alibhai et al., 1985). The a?nity of sludge for heavy metals has been proposed as (in decreasing order): Cu > Cd > Zn > Ni (Gould and Genetelli, 1984). Similarly, sorption of heavy metals to activated carbon, kaolin, bentonite, diatomite and waste materials such as compost and cellulose pulp waste can also mitigate inhibition (Ulmanu et al., 2003). Chelation by organic ligands has been well documented for several metals. Babich and Stotzky (1983) have shown a decrease in nickel toxicity by EDTA, PDA, NTA, aspartate, and citrate, in that order. Exposure of microorganisms to heavy metals is known to activate a wide variety of intracellular detoxi?cation strategies (Gadd and Gri?ths, 1978). The intracellular defense systems include biologically mediated precipitation or chelation of metal ions at the cell surface (Wood and Wang, 1983), biomethylation (Summers, 1986), exocytosis

(expulsion of metals after their chemical inactivation inside the cell) (Silver and Phung, 1996) and plasmid-mediated resistance (Wood and Wang, 1983). 2.5. Organics A wide range of organic compounds can inhibit anaerobic processes. Organic chemicals which are poorly soluble in water or adsorbed to the surfaces of sludge solids may accumulate to high levels in anaerobic digesters. The accumulation of apolar pollutants in bacterial membranes causes the membrane to swell and leak, disrupting ion gradients and eventually causing cell lysis (Heipieper et al., 1994; Sikkema et al., 1994). Organic compounds which have been reported to be toxic to the anaerobic processes include alkyl benzenes (Yang and Speece, 1986; Renard et al., 1993), halogenated benzenes (van Beelen and van Vlaardingen, 1994), nitrobenzenes (Bhattacharya et al., 1996), phenol and alkyl phenols (Sierra-Alvarez and Lettinga, 1991a; Soto et al., 1991; Fang et al., 1995), halogenated phenols (Shin and Kwon, 1998), nitrophenols (Borja et al., 1997; Uberoi and Bhattacharya, 1997a; McCue et al., 2003), alkanes (Mormile and Su?ita, 1996), halogenated aliphatics (Stuckey et al., 1980; Boucquey et al., 1995), alcohols (Dimirer and Speece, 1998), halogenated alcohols (Blum and Speece, 1991), aldehydes (Gonzales-Gil et al., 2002), ethers (Playne and Smith, 1983; Hayward and Lau, 1989), ketones (Playne and Smith, 1983; Hayward and Lau, 1989), acrylates, carboxylic acids, amines, nitriles, amides (Blum and Speece, 1991; Stergar et al., 2003), and pyridine and its derivatives (Liu et al., 1998). Moreover, some LCFAs (Koster and Cramer, 1987), surfactants, and detergents were also reported to adversely impact anaerobic digestion (Madsen and Rasmussen, 1996; Gavala and Ahring, 2002). The inhibition concentration ranges vary widely for speci?c toxicants. The parameters that a?ect the toxicity of organic compounds include toxicant concentration, biomass concentration, toxicant exposure time, cell age, feeding pattern, acclimation, and temperature (Yang and Speece, 1986). At lower concentrations, biodegradation of some toxicants can prevent inhibition; higher concentrations of toxicants generally lead to signi?cant inhibition of anaerobic processes (Anthony and Breimhurst, 1981; O’Connor and Young, 1989). With higher biomass concentration, reactors exhibit greater process stability in the presence of toxic shocks (Uberoi and Bhattacharya, 1997a). At equal solids concentrations, younger cultures were proved to be more robust and resistant to toxicants than older cultures (Yang and Speece, 1986). Inhibition has been usually quanti?ed by determining the IC50 concentration. Because of a lack of consistency in exposure times between studies, which varied from 30 min (Dutka et al., 1983) to 285 h (Johnson and Young, 1983), the inhibition concentrations can vary signi?cantly. As with other inhibitory substances, microbial acclimation is an important parameter in assessing the inhibitory

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e?ects of organic substances. Four interrelated mechanisms by which adaptation can occur have been suggested: (1) enrichment of organisms which can degrade the toxic compounds, (2) induction of speci?c enzymes for the degradation, (3) genetic engineering and (4) exhaustion of preferential substrates before switching to the xenobiotic substrate, i.e. a diauxic pattern (Spain et al., 1980; Spain and van Veld, 1983; van der Meer, 1994). Acclimation of anaerobic microorganisms both increases their tolerance to the toxicants shock and enhances toxicant biodegradability (Stuckey et al., 1980; Wu et al., 1993). 2.5.1. Chlorophenols Chlorophenols include monochlorophenols (CPs), dichlorophenols (DCPs), trichlorophenols (TCP), tetrachlorophenols (TeCPs), and pentachlorophenol (PCP). Chlorophenols are toxic to many organisms by disrupting the proton gradient across membranes and interfering with energy transduction of cells. The e?ect of aromatic compounds on membrane processes has been reviewed by Sikkema et al. (1995). The relative toxicity of chlorophenols has been investigated by many researchers and the results are somewhat contradictory. Among di?erent isomer series, PCP was the most toxic to acidogens and methanogens. Approximately 0.5–10 mg/L PCP caused inhibition to acidogenic and methanogenic populations (Bauer and Capone, 1985; Godsy et al., 1986; Blum and Speece, 1991; Patel et al., 1991; Uberoi and Bhattacharya, 1997b; Piringer and Bhattacharya, 1999). Sierra-Alverez and Lettinga (1991b) have described the relationship between the increase in the number of chloro-substituents in the aromatic benzene rings and their toxicity to methanogenic processes. This theory was supported by the ?nding of Jin and Bhattacharya (1996) that TCPs were more toxic than DCPs and CPs. However, other studies found no correlation between the toxicity of DCPs and TCPs and the number of chloro-substituents (Blum and Speece, 1991). Within individual isomer series, toxicity due to DCPs and TCPs to both propionate and acetate degradation was dependent on the substitution position of chlorine atoms on the benzene ring. Di?erent orders of relative toxicity have been reported (Jin and Bhattacharya, 1996; Uberoi and Bhattacharya, 1997b). The chlorine position on CPs did not signi?cantly e?ect toxicity to either propionate or acetate degradation (Kim et al., 1994; Uberoi and Bhattacharya, 1997b). Davies-Venn et al. (1992), however, found that the toxicity of CPs to aceticlastic methanogenesis increased as the substituted chlorine group changed from the ortho to the meta to the para position on the benzene ring. In relation to physico-chemical properties, previous studies with aromatic compounds indicated that those structural characteristics that decrease polarity increase toxicity (Kamlet et al., 1986; Patel et al., 1991). Compounds of greater hydrophobicity accumulate more e?ciently in membranes, causing a greater disturbance to the membrane structure (Heipieper et al., 1994; Sikkema

et al., 1994). A high correlation of the methanogenic toxicity to the log P (logarithm of the octano-water partition coe?cient) was obtained for chloro-substituted benzenes and phenols (Sierra-Alverez and Lettinga, 1991b; EnnikMaarsen et al., 1998). The relative hydrophobicity also provides an explanation for the relative toxicity of nitrophenols and hydroxyphenols (Wang et al., 1991). Literature data on the tolerance of di?erent important subpopulations in methanogenic sludges to the chlorophenols are also con?icting. This variation could be attributed to variations in microbial speci?c growth rates and/or the physiological state, which may a?ect the tolerance to an inhibitory compound as well as the accumulation of organic compounds in the cell (Ennik-Maarsen et al., 1998). Colleran et al. (1992) reported that aceticlastic methanogens, butyrate oxidizers and ethanol oxidizers, were similarly sensitive to halogenated aromatics, while hydrogenotrophs were less sensitive (Kim et al., 1996). In other studies, propionate degraders were found to be the most sensitive to chlorophenols (Johnson and Young, 1983; Wu et al., 1993; Jin and Bhattacharya, 1996), whereas Kim et al. (1994) concluded that ethanol degraders were not as sensitive as aceticlastic methanogens towards chnorophenols and chloroanilines. Degradation of initially inhibitory compounds was observed for 2,4-DCP, 2,3,6TCP, 2,3,5-TCP (Uberoi and Bhattacharya, 1997b), 4CP, and 2,4,6-TCP (Fantroussi et al., 1998) after acclimation. Acclimation was related to the concentration of substrates and their chemical structures (Link?eld et al., 1989). 2.5.2. Halogenated aliphatics Most of the halogenated aliphatics are strong inhibitors of methanogenesis. In general, the brominated compounds were more inhibitory to methanogens than their chlorinated analogs (Belay and Daniels, 1987). Renard et al. (1993) measured the toxicity of a mixture of polychlorinated organic compounds including C4Cl6, C2Cl6, C2Cl4, etc. and found that 50% inhibition of methanogenesis started at 3.3 mg/L. Complete inhibition of methanogenesis occurred at 100 mg/L. Methanogenic toxicity data for other chlorinated aliphatic hydrocarbons are lacking (Stuckey et al., 1980; Anthony and Breimhurst, 1981). However, the methanogenic toxicity of chloroform, the most widely used chloroaliphatics, has been studied extensively. Among six chlorinated hydrocarbon solvents including carbon tetrachloride and 1,1,1-trichloroethane, chloroform was found to be the most toxic to the anaerobic digestion of sewage sludge (Swanwick and Foulkes, 1971). A concentration of 0.01 mg/L or more of chloroform in the sewage was likely to have an adverse e?ect on sludge digestion (Stickley, 1970). The IC50 concentration for this compound has been reported to range from 0.15 mg/L in unacclimated methanogenic consortia to 50 mg/L in acclimated consortia (Anthony and Breimhurst, 1981; Parkin and Speece, 1983; Salenieks and Henry, 1986; Yang and Speece, 1986; Hickey et al., 1987; In et al., 1992). During anaerobic degradation of chloroform, reactive and toxic intermediates

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were formed, which partly contributed to its strong toxicity (van Beelen and van Vlaardingen, 1994). Unlike chloroaromatics, no relationship between the number of chloro-substituents and toxicity could be identi?ed for chloroaliphatics. Polarity is an important factor, but it is insu?cient for the prediction of the toxicity of the di?erent compounds (Sanz et al., 1997). It was reported that tri- and tetrachloride derivatives of methane and ethane were more toxic than dichlorinated compounds. Perchlorinated derivatives of ethane and ethene were scarcely inhibitory at concentrations near their maximum water solubility (Sanz et al., 1994). Compared to their saturated counterparts, unsaturated chloroaliphatics were less toxic (Chou et al., 1978; Sanz et al., 1997). Acclimation of methanogenic consortium to polychlorinated aliphatic compounds is possible. Yang and Speece (1986) found that an anaerobic culture was able to acclimate to the presence of chloroform while fermenting acetate to methane. Inhibition of unacclimated cultures by chloroform was noted at 0.5 mg/L, but with acclimation 15 mg/L could be tolerated. Filho et al. (1992) reported that fermentative microorganisms acclimate more quickly than hydrogenotrophic ones, followed by H2 producing acetogens and aceticlastic methanogens. 2.5.3. N-substituted aromatics N-substituted aromatics are reactive toxicants including nitrobenzenes, nitrophenols, aminophenols, aromatic amines, etc. (Blum and Speece, 1991). Reactive toxicity is caused by speci?c chemical interactions with enzymes or interference with metabolic pathways (Balderston and Payne, 1976). Nitroaromatics are very toxic compounds to methanogens, with IC50 values generally ranging from 0.014 to 0.12 mM (Johnson and Young, 1983; Donlon et al., 1995; Bhattacharya et al., 1996). Aromatic amines, in contrast, are less inhibitory; the IC50 values were between 3.2 and 67 mM, perhaps due to their lower hydrophobicity (Razo-Flores et al., 1997). Nitroanilines were found to be the most toxic among N-substituted aromatics, partly because of their high chemical reactivity (Razo-Flores et al., 1997). Increasing the number of nitro groups beyond one had limited e?ect in increasing the toxicity of nitrobenzenes. The addition of an extra amino group to aminophenol resulted in more toxic compounds, while the addition of an amino group to aniline resulted in less toxic phenylenediamines (Donlon et al., 1995). However, the combination of nitro and amino groups, e.g. nitroanilines, was found to be the most toxic substitute pattern (Donlon et al., 1995). Toxicity due to the mononitrophenols also depends on the substitution position. Among the nitrophenols studied, the toxicity increased in the order of 4-nitrophenol > 2-nitrophenol > 3-nitrophenol for an acetate enrichment culture (Haghighi-Podeh et al., 1995; Haghighi-Podeh and Bhattacharya, 1996). Similar results were observed for acetate and propionate enrichment culture (Uberoi

and Bhattacharya, 1997a). Another study found that para-nitrophenol was more toxic than meta-nitrophenol and the ortho-nitrophenol was the least toxic to methanogens (Tseng and Yang, 1994). Acclimation of anaerobes to N-substituted aromatics decreases their toxicity and therefore enhances biodegradation. 4-Nitrophenol and 2,4-dinitrophenol can be degraded at concentrations up to 200 mg/L after su?cient acclimation. The lag time before biodegradation started increased with increasing toxicant concentration (O’Connor and Young, 1989).

2.5.4. LCFAs Treatment of fatty materials by anaerobic digestion is often hampered because of the inhibitory e?ect of LCFAs. LCFAs have been reported to be inhibitory at low concentrations for gram-positive but not gram-negative microorganisms (Kabara et al., 1977). Methanogens can be inhibited by LCFAs due to their cell wall, which resembles that of gram-positive bacteria (Zeikus, 1977). LCFAs show acute toxicity towards anaerobic consortium by adsorption onto the cell wall/membrane, interference with the transport or protective function (Rinzema et al., 1994). In addition, sorption of a light layer of LCFAs to biomass leads to the ?otation of sludge and consequent sludge washout (Rinzema et al., 1989). In UASB reactors, granular sludge ?otation sometimes occurred at concentrations far below the toxicity limit (Hwu et al., 1998). Oleic acid was almost as inhibitory as lauric acid, which exhibited I50 of 4.3 mM. Cyprylic acid was only slightly inhibitory (Koster and Cramer, 1987). Koster and Cramer (1987) also observed enhanced toxicity of capric acid and myristic acid when lauric acid was present. LCFA toxicity varied with the type of anaerobic sludges and was more correlated to the sludges’ physical characteristics (speci?c surface area and size distribution) than to their biological characteristics. Suspended and ?occulent sludges, which have a higher speci?c surface area, su?ered much greater inhibition than did granular sludge (Hwu et al., 1996). Thermophiles have been reported to be more sensitive to LCFAs than mesophiles, possibly due to the di?erent composition of cell membranes (Hwu and Lettinga, 1997). Biodegradation of LCFAs has been reported in both mesophilic and thermophilic environments (Hanaki et al., 1981; Angelidaki and Ahring, 1992). It has been suggested that LCFAs exerted a bactericidal e?ect and no adaptation of methanogens occured (Hanaki et al., 1981; Koster and Cramer, 1987; Angelidaki and Ahring, 1992). However, recent studies based on the degradation of oleic acid in an anaerobic ?xed-bed reactor showed that acclimation improved the resistance of the bio?lm to the presence of oleate and improved the biodegradation capacity compared to the bio?lm formed in the absence of lipids (Alves et al., 2001a,b). Addition of calcium has been shown to reduce LCFA inhibition, probably because of the formation of insoluble salts (Hanaki et al., 1981; Angelidaki

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and Ahring, 1990). However, calcium addition cannot solve the problem of sludge ?otation (Alves et al., 2001a,b). 2.5.5. Lignins and lignin related compounds Lignin derivatives with aldehyde groups or apolar substituents are highly toxic to methanogens. The aromatic carboxylic acids, however, were only mildly toxic. Op den Camp et al. (1988) tested the toxicity of several lignin model phenolic acids to the anaerobic degradation of cellulose and observed that the acids only caused inhibition of methane production at very high concentrations. Benjamin et al. (1984) evaluated the methanogenic toxicity of various lignin derived monomers present in kraft condensates and found that eugenol, with an apolar side chain, was more toxic than its counterpart guaiacol which lacks the side chain. 3. Engineering signi?cance 3.1. Agricultural wastes

bark are produced from a variety of crops, which could be used for energy generation (Kalra and Panwar, 1986). Crop residues typically contain a high lignocellulosic content. Problems such as low gas yield during anaerobic digestion of these materials are usually associated with a high C/N ratio or high lignin content. In addition, the inhibition caused by pesticide and herbicide residues would a?ect digestion process kinetics (Khalil et al., 1991; Chakraborty et al., 2002). Certain plants generate resin extracts which protect them from biological damage. These extracts may be inhibitory to the digestion process (Speece, 1987). Pretreatments such as acid or base hydrolysis are often employed before anaerobic digestion to increase biogas yield. However, byproducts formed in the pretreatment (fufural, hydroxymethyl fufural, formic acid, and levulinic acid) are potential inhibitors of anaerobic digestion. Microorganisms may eventually adapt and/or degrade these byproducts, but process kinetics could be a?ected (Speece, 1987). 3.2. Municipal wastes

Animal waste includes voided waste from livestock and poultry, wastewater, feedlot runo?, silage juices, bedding, and feed. These wastes are a substantial contributor to non-point source pollution and can a?ect wetland habitats and contaminate drinking water sources. Animal waste often has very high total ammonia nitrogen concentrations due to the presence of ammonia as well as protein and urea that readily release ammonia upon anaerobic treatment (Zeeman et al., 1985; Krylova et al., 1997; Hansen et al., 1998). Consequently, the principal instability associated with the anaerobic digestion of animal waste is ammonia inhibition (Zeeman et al., 1985; Hashimoto, 1986; Kayhanian, 1994). Sudden increases in ammonia concentration in the feedstock are unusual (Hobson, 1991). However, feed slurry that has been stored for some time in the animal house often contains high concentration of ammonia released from decomposition of organic nitrogen. Shock loading of this feed slurry can cause inhibition of anaerobic digesters (Hobson, 1991). In addition to ammonia, swine manure also contains a high sulfate concentration derived from a protein-rich diet. The inhibition caused by ammonia and by sul?de in?uences each other (Hansen et al., 1999). Feed additives (antibiotics, chemotherapeutics) for improving food utilization and disinfectants for preventing infectious diseases have been widely used in intensive animal production (Hilpert et al., 1984). In most cases, these compounds are in very low concentrations (less than 30 ppm) in the waste and are generally not inhibitory (Hobson, 1991). However, some synthetic chemotherapeutics such as Olaquindox may be strongly inhibitory even at 1 mg/L (Hilpert et al., 1984). This concentration may be reached in practice and special treatments such as predilution may be needed before anaerobic digestion (Varel and Hashimoto, 1981; Hilpert et al., 1984; Poels et al., 1984). Crop residues represent another fraction of agricultural waste. Substantial quantities of unused stalks, straws, and

More than 181.4 million metric tons of municipal solid wastes are produced in the United States annually (US EPA, 2000) with up to 60% organic material (de Laclos et al., 1997). Due to the increase of source and on-site recycling and improved refuse-handling equipment, substantial fractions of the discarded paper fraction, metals, and glass are being recycled, resulting in the production of more organic-rich and less biotoxic biowaste (Ghosh et al., 2000). Because of its value as a potential renewable energy source and high biodegradability, there is a growing interest in anaerobic digestion of the biowaste. The principal instability associated with the anaerobic digestion of biowaste is ammonia inhibition due to the degradation of protein-containing materials (Kayhanian, 1994; Gallert and Winter, 1997; Gallert et al., 1998). The composition of the organic matter depends greatly on the source of the organic fraction. It was found that the digestion process deteriorated with increasing levels of ammonia-N, with total process cessation at a COD/N ratio of 50 (Poggi-Varaldo et al., 1997, 1998). Two practical methods, dilution of digester contents with water and adjustment of feedstock C/N, have been tested successfully to mitigate ammonia inhibition (Kayhanian, 1999). In general, for solid wastes with a C/N ratio above 20, the ammonia inhibition e?ect can be compensated by dilution with water which lowers the concentration of potential inhibitors. Sludge production is an integral part of the domestic sewage treatment process. Due to their resistance to biodegradation, heavy metals present in the raw sewage often accumulate in the sludge to potentially toxic concentrations. Petrasek and Kugelman (1983) found that most metals were concentrated 10–30-fold in the sludges produced in conventional wastewater treatment plants. Anaerobic digestion is the most typical method for stabilization of sludge prior to ultimate disposal in many countries

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(Hobson et al., 1981) and is usually the most sensitive process to heavy metal toxicity (Lester et al., 1983). The e?ects on digestion of chemical treatment of sludge were investigated by Gossett et al. (1978). Alum, ferric chloride and organic ?occulent were reported to reduce: (1) gas production, (2) the methane content in the biogas, (3) COD removal, and (4) volatile suspended solids removal. A similar e?ect was observed for sewage emulsi?ers such as alkyl benzene sulphonate (ABS). 3.3. Industrial wastes 3.3.1. Food industry wastes The food industries that could bene?t from anaerobic treatment include fruit and vegetable canning, edible oil re?ning, dairy production, seafood processing, meat processing, starch and sugar production, brewing, and fermentation. Wastes from food processing are high in organic matter and are therefore ideal for anaerobic digestion. However, the application of this technique may be hindered by the presence of various inhibitors. Seafood processing wastewaters contain high concentrations of di?erent cations and anions, mainly Na+, Cl?, and SO2? 4 (Feijoo et al., 1995). The sodium concentration in these wastewaters may reach that of sea water (approximately 12 g/L) (Rinzema et al., 1988). Hypersaline wastewaters are also generated in vegetable, vegetable oil, and dairy processing industries. Dairy wastewaters are rich in fats, proteins and carbohydrates (Rico et al., 1991). Wastewaters from the dairy industry are usually generated in an intermittent way, and the ?ow and characteristics of wastewaters change from one factory to another depending on the type of systems and the methods of operation (Rico et al., 1991). It has been reported that the intermediates of fat degradation, glycerol and LCFAs, were below inhibitory concentrations. However, ammonia produced from the degradation of milk proteins was 62.2 mg/L, close to the inhibition level of the anaerobic digestion process (Vidal et al., 2000). Meat processing wastes are substantially di?erent from other food industry wastes. They are very strong wastes containing grease, blood, faeces, and recalcitrant organic matter such as straw and hair. During anaerobic digestion, protein and lipids degradation leads to the accumulation of ammonia and LCFAs, which are important inhibitors of the anaerobic microorganisms (Salminen and Rintala, 1999). The wastes also frequently include high concentrations of biocides and disinfectants such as hypochlorite (Tritt, 1992). The di?cult nature of these wastes could be overcome by co-digestion, which could be advantageous due to an improved C/N ratio and dilution of the inhibitory compounds (Tritt, 1992). 3.3.2. Paper and pulp industry wastes The pulp and paper industry has several high strength waste streams that are of concern from an environmental standpoint. Since the pulp produced corresponds to only

40–45% of the original weight of the wood, the e?uents exhibit high COD concentrations (Ali and Sreekrishnan, 2001), which together with e?uent’s warm temperature (the waste is typically around 35 °C), makes anaerobic digestion a favorable waste treatment technique. Satisfactory digestion of paper mill e?uent has been reported (Rintala et al., 1991). The most common inhibitors to the anaerobic digestion process include sul?de, tannins, resin acids, LCFAs, and halogenated compounds (Ali and Sreekrishnan, 2001). Sulfate is primarily produced in pulping operations using the sul?te process (Thompson et al., 2001). Sul?de removal can be achieved by sulfur bacteria which convert the sul?de ions to elemental sulfur (Buisman et al., 1991). Chen and Horan (1998) reported the use of a two-stage anaerobic-aerobic approach to remove COD and sulfate from newsprint mill wastewater. Tannins were found to contribute up to 50% of the COD of the debarking process wastewater. They are known to exhibit methanogenic toxicity proportional to the extent of polymerization (Field et al., 1988). Naturally occurred resin acids and LCFAs in the wood and bark could be transferred to process waters during pulping operations. They have been shown to inhibit methanogens, especially the aceticlastic methanogens (Hanaki et al., 1981; Koster and Cramer, 1987). Halogenated compounds are produced in the bleaching process. Their toxicity to anaerobes has been well documented. Most of the organic inhibitors are biodegradable to a certain extent (Ali and Sreekrishnan, 2001). However, knowledge of the possible contaminants present in the wastewater, their origins, and their degree of toxicity is essential to successful anaerobic treatment. 3.3.3. Textile industrial wastes The main sources of wastewater generated by the textile industry originate from the washing (or scouring) and bleaching of natural ?bers and from the dyeing and ?nishing steps. Given the great variety of ?bers, dyes, process aids, and ?nishing products in use, these processes generate wastewaters of great chemical complexity (Vandevivere et al., 1998). Several laboratory-scale investigations have illustrated the potential of sequential anaerobic/aerobic biotreatment steps for textile wastewaters. Other studies have shown, however, that methanogenesis, and hence COD removal, is easily inhibited by textile e?uents (Athanasopoulos, 1992). The components of textile wastewater that could be potential inhibitors are dye, dyeing auxiliaries (polyacrylates, phosphonates), surfactants (alkyl phenol ethoxylates), adsorbable organic halogens (chloroform), and heavy metals (Feitkenhauer, 2004; Lee and Pavlostathis, 2004). 3.3.4. Petrochemical re?neries wastes Anaerobic digestion could also be of use in petrochemical re?neries. It has been found that after prolonged acclimation, aldehydes, acids, alcohols, and esters could be used for methane production (Chou et al., 1978). The presence of hydroxyl groups and an increasing carbon chain reduced

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Y. Chen et al. / Bioresource Technology 99 (2008) 4044–4064 Angelidaki, I., Ahring, B.K., 1994. Anaerobic thermophilic digestion of manure at di?erent ammonia loads: e?ect of temperature. Water Res. 28, 727–731. Angelidaki, I., Petersen, S.P., Ahring, B.K., 1990. E?ects of lipids on thermophilic anaerobic digestion and reduction of lipid inhibition upon addition of bentonite. Appl. Microbiol. Biotechnol. 33, 469–472. Angelidaki, I., Ellegaard, L., Ahring, B.K., 1993. A mathematical model for dynamic simulation of anaerobic digestion of complex substrates: focusing on ammonia inhibition. Biotechnol. Bioeng. 42, 159–166. Anthony, R.M., Breimhurst, L.H., 1981. Determining maximum in?uent concentrations of priority pollutants for treatment plants. J. Water Pollut. Control Fed. 53, 1457–1468. ? ? Aspe, E., Mart?, M., Roeckel, M., 1997. Anaerobic treatment of ?shery wastewater using a marine sediment inoculum. Water Res. 31 (9), 2147–2160. Athanasopoulos, N., 1992. Cotton yarn and fabric ?nishing wastewater treatment using an anaerobic expanded bed reactor. Bioresour. Technol. 39, 291–293. Babich, H., Stotzky, G., 1983. Toxicity of nickel to microbes, environmental aspects. Adv. Appl. Microbiol. 29, 195–295. Balderston, W.L., Payne, W.J., 1976. Inhibition of methanogenesis in salt marsh sediments and whole-cell suspensions of methanogenic bacteria by nitrogen oxides. Appl. Environ. Microbiol. 32, 264–269. Balsleve-Olsen, P., Lynggaard-Jensen, A., Nickelsen, C., 1990. Pilot-scale experiments on anaerobic treatment of wastewater from a ?sh processing plant. Water Sci. Technol. 22, 463–474. Bauer, J.E., Capone, D.G., 1985. E?ects of four aromatic organic pollutants on microbial glucose metabolism and thymidine incorporation in marine sediments. Appl. Environ. Microbiol. 49, 828–835. Belay, N., Daniels, L., 1987. Production of ethane, ethylene, and acetylene from halogenated hydrocarbons by methanogenic bacteria. Appl. Environ. Microbiol. 53 (7), 1604–1610. Benjamin, M.M., Woods, S.L., Ferguson, J.F., 1984. Anaerobic toxicity and biodegradability of pulp mill waste constituents. Water Res. 18, 601–607. Bhattacharya, S.K., Parkin, G.F., 1989. The e?ect of ammonia on methane fermentation process. J. Water Pollut. Control Fed. 61 (1), 55–59. Bhattacharya, S.K., Sa?erman, A.G., 1989. Determination of bioavailable nickel concentrations in inhibited methanogenic systems. Environ. Technol. Lett. 10 (8), 725–730. Bhattacharya, S.K., Madura, R., Uberoi, V., Haghighi-Podeh, M.R., 1995a. Toxic e?ects of cadmium on methanogenic systems. Water Res. 29 (10), 2339–2345. Bhattacharya, S.K., Uberoi, V., Madura, R.L., Haghighi-Podeh, M.R., 1995b. E?ect of cobalt on methanogenesis. Environ. Technol. 16 (3), 271–278. Bhattacharya, S., Qu, M., Madura, R.L., 1996. E?ects of nitrobenzene and zinc on acetate utilizing methanogens. Water Res. 30 (12), 3099– 3105. Blum, D.J.W., Speece, R.E., 1991. A database of chemical toxicity to environmental bacteria and its use in interspecies comparisons and correlations. J. Water Pollut. Control Fed. 63, 198–207. Boardman, G.D., McVeigh, P.J., 1997. Use of UASB technology to treat crab processing wastewaters. J. Environ. Eng. 123, 776–785. ? Borja, R., Sanchez, E., Weiland, P., Travieso, L., 1993. E?ect of ionic exchanger addition on the anaerobic digestion of cow manure. Environ. Technol. 14, 891–896. ? Borja, R., Sanchez, E., Duran, M.M., 1996a. E?ect of the clay mineral zeolite on ammonia inhibition of anaerobic thermophilic reactors treating cattle manure. J. Environ. Sci. Health A31 (2), 479–500. ? Borja, R., Sanchez, E., Weiland, P., 1996b. In?uence of ammonia concentration on thermophilic anaerobic digestion of cattle manure in up?ow anaerobic sludge blanket (UASB) reactors. Process Biochem. 31 (5), 477–483. Borja, R., Alba, J., Banks, C.J., 1997. Impact of the main phenolic compounds of olive mill wastewater (OMW) on the kinetics of aceticlastic methanogenesis. Process Biochem. 32 (2), 121–133.

the toxicity of compounds to the digester micro?ora. Acclimation to aromatic ring and double-bond compounds was also possible. Chou et al. (1978) concluded that digestion of petrochemical wastes would not only result in a saving of energy over aerobic processes but would also produce methane on a scale for use as a fuel. 4. Conclusion Anaerobic digestion is an e?cient waste treatment technology that harnesses natural anaerobic decomposition to reduce waste volume and generate biogas at the same time. It has been widely applied to the treatment of waste from agricultural and industrial operations. Depending on the origin, the waste stream may contain inhibitory or even toxic substances such as ammonia, sul?de, heavy metals, and organics. Accumulation of these substances may cause reactor upset, as indicated by reduced biogas production and/or biogas methane content, and possible reactor failure. Due to the di?erence in anaerobic microorganisms, waste composition, and experimental methods and conditions, results from previous investigations on inhibition of anaerobic processes vary substantially. Obtaining information on waste components is necessary for successful application of anaerobic digestion. It has been suggested that co-digestion with other waste, adaptation of microorganisms to inhibitory substances, and incorporation of methods to remove or counteract toxicants before anaerobic digestion can signi?cantly improve the waste treatment e?ciency. References
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