Pandey2007.pdf

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International Biodeterioration & Biodegradation 59 (2007) 73–84 www.elsevier.com/locate/ibiod

Review

Bacterial decolorization and degradation of azo dyes Anjali Pandey, Poonam Singh, Leela Iyengar Department of Chemistry, Biotechnology Laboratory, I.I.T., Kanpur 208016, India Received 11 May 2006; received in revised form 25 August 2006; accepted 26 August 2006 Available online 27 October 2006

Abstract Azo compounds constitute the largest and the most diverse group of synthetic dyes and are widely used in a number of industries such as textile, food, cosmetics and paper printing. They are generally recalcitrant to biodegradation due to their xenobiotic nature. However microorganisms, being highly versatile, have developed enzyme systems for the decolorization and mineralization of azo dyes under certain environmental conditions. Several genera of Basidomycetes have been shown to mineralize azo dyes. Reductive cleavage of azo bond, leading to the formation of aromatic amines, is the initial reaction during the bacterial metabolism of azo dyes. Anaerobic/anoxic azo dye decolorization by several mixed and pure bacterial cultures have been reported. Under these conditions, this reaction is nonspecific with respect to organisms as well as dyes. Various mechanisms, which include enzymatic as well as low molecular weight redox mediators, have been proposed for this non-specific reductive cleavage. Only few aerobic bacterial strains that can utilize azo dyes as growth substrates have been isolated. These organisms generally have a narrow substrate range. Degradation of aromatic amines depends on their chemical structure and the conditions. It is now known that simple aromatic amines can be mineralized under methanogenic conditions. Sulfonated aromatic amines, on the other hand, are resistant and require specialized aerobic microbial consortia for their mineralization. This review is focused on the bacterial decolorization of azo dyes and mineralization of aromatic amines, as well as the application of these processes for the treatment of azo-dye-containing wastewaters. r 2006 Elsevier Ltd. All rights reserved. Keywords: Azo dyes; Decolorization; Biodegradation; Aromatic amines; Anaerobic/aerobic treatment

Contents 1. 2.

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4.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Decolorization of azo dyes by bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1. Azo dye decolorization under anaerobic conditions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2. Azo dye decolorization under anoxic conditions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3. Azo dye decolorization under aerobic conditions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4. Mechanism of azo dye reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.1. Direct enzymatic azo dye reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.2. Mediated biological azo dye reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.3. Azo dye decolorization by biogenic inorganic compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Degradation of aromatic amines . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Degradation of aromatic amines under anaerobic conditions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2. Aerobic biodegradation of aromatic amines. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3. Anaerobic–aerobic biological treatment for azo dye decolorization and degradation . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Corresponding author. Tel.: +91 512 2597160; fax: +91 512 2597437.

E-mail address: [email protected] (A. Pandey). 0964-8305/$ - see front matter r 2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.ibiod.2006.08.006

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1. Introduction Azo dyes, which are aromatic compounds with one or more –NQN– groups, constitute the largest class of synthetic dyes used in commercial applications (Zollinger, 1991). In 1994 estimates, the world production of dyes was around 1 million tons, of which more than 50% were azo dyes (Ollgaard et al., 1999; Stolz, 2001). These dyes are widely used in a number of industries, such as textile dyeing, food, cosmetics, paper printing, with the textile industry as the largest consumer. All dyes do not bind to the fabric; depending on the class of the dye, its loss in wastewaters could vary from 2% for basic dyes to as high as 50% for reactive dyes, leading to severe contamination of surface and ground waters in the vicinity of dyeing industries (Ganesh et al., 1994; O’Neill et al., 1999). Many dyes are visible in water at concentrations as low as 1 mgl1. Textile processing wastewaters with dye contents in the range of 10–200 mgl1 are highly colored. Some of the dyes and their degradation products are carcinogenic in nature (Levine, 1991). A review of the mutagenicity of effluents showed that textile and other dye-related industries produce consistently more potent wastewaters when compared to other industrial discharges (Houk, 1992). Recent studies by Rajaguru et al. (2002) and Umbuzeiro et al. (2005) have shown that azo dyes contribute to mutagenic activity of ground and surface waters polluted by textile effluents. Further, their discharge into surface water leads to aesthetic problems and obstructs light penetration and oxygen transfer into water bodies, hence affecting aquatic life. Thus, the removal of color from textile effluents has been a major concern. There are many reports on the use of physicochemical methods for color removal from dyes containing effluents (Churchley, 1994; Vandevivere et al., 1998; Swaminathan et al., 2003; Behnajady et al., 2004; Wang et al., 2004; Golab et al., 2005; Lopez-Grimau and Gutierrez, 2005). Extensively used coagulation/ flocculation techniques produce large amounts of sludge, which requires safe disposal. Adsorption and, to a certain extent, membrane filtration techniques lead to secondary waste streams which need further treatment. These constraints have led to the consideration of advanced oxidation processes (AOP) and biological methods as attractive options for the treatment of dye-containing wastewaters. AOP are defined as those processes that use strong oxidizing agents (H2O2, Fenton’s reagent) or heterogenous photocatalysts such as TiO2, ZnO2, Mn and Fe in the presence or absence of an irradiation source. These involve mainly the generation of (OH) radical for the destruction of refractory and hazardous pollutants (Vandevivere et al., 1998; Alaton et al., 2002; Al-Kdasi et al., 2004). These methods do not produce solid waste. However, both AOP and membrane filtration methods are energy and cost intensive. Biological methods are generally considered environmentally friendly, as they can lead to complete mineralization of organic pollutants at low cost. Azo compounds are xenobiotic in

nature; only one natural azo compound (4–40 dihydroxy azo benzene) has been reported so far (Gill and Strauch, 1984). Thus they can be expected to be recalcitrant to biodegradation. It is generally observed that dyes resist biodegradation in conventional activated sludge treatment units (Stolz, 2001). It is now known that several microorganisms, including fungi, bacteria, yeasts and algae, can decolorize and even completely mineralize many azo dyes under certain environmental conditions. Many reviews are available on the physicochemical and microbiological methods for decolorization of azo dyes (Banat et al., 1996; Delee et al., 1998; Vandevivere et al., 1998; O’Neill et al., 1999; McMullan et al., 2001; Stolz, 2001; Rai et al., 2005;Van der Zee and Villaverde, 2005). This review focuses on the pathways and mechanisms by which aerobic and anaerobic bacteria decolorize azo dyes and degrade the aromatic amines generated by this reaction. We also briefly discuss the setting up of potential dual aerobic and anaerobic processes to remediate azo-dye-containing wastewaters. 2. Decolorization of azo dyes by bacteria Reductive cleavage of the –NQN– bond is the initial step of the bacterial degradation of azo dyes. Decolorization of azo dyes occurs under anaerobic (methanogenic), anoxic and aerobic conditions by different trophic groups of bacteria. Decolorization of azo dyes under these different conditions is briefly discussed in subsequent sections. 2.1. Azo dye decolorization under anaerobic conditions Methanogenesis from complex organic compounds requires the coordinated participation of many different trophic groups of bacteria, including acidogenic, acetogenic and methanogenic bacteria (Kasper and Wuhrmann, 1978). Dye decolorization under these conditions requires an organic carbon/energy source. Simple substrates like glucose, starch, acetate, ethanol and more complex ones, such as whey and tapioca, have been used for dye decolorization under methanogenic conditions (Chinwetkitvanich et al., 2000; Willetts et al., 2000; Talarposhti et al., 2001; Yoo et al., 2001; Isik and Sponza, 2005a; Van der Zee and Villaverde, 2005). Extensive studies have been carried out to determine the role of the diverse groups of bacteria in the decolorization of azo dyes. Carliell et al. (1996) and Razo-Flores et al. (1997a) have associated the decolorization with methanogens, whereas studies by other investigators showed that acidogenic as well as methanogenic bacteria contribute to dye decolorization (Chinwetkitvanich et al., 2000; Bras et al., 2001). Use of molecular methods to characterize the microbial populations in anaerobic-baffled reactors treating industrial dye waste showed that members of the g-proteobacteria, together with sulfate reducing bacteria (SRB), were prominent members of mixed bacterial

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populations. Along with these, a methanogenic population dominated by Methanosaeta species and Methanomethylovorans hollandica contributed to the treatment of industrial wastewater (Plumb et al., 2001). Yoo et al. (2001) showed that decolorization of Orange 96 was not significantly affected by 2-bromoethanesulfonic acid (BES), an inhibitor specific to methanogens. This suggests that methanogens took no part in the decolorization, and it contradicts the findings of many other investigators. On the other hand, in the presence of acetate and sulfate, molybdate, an inhibitor specific to SRB, caused a significant decrease in the decolorization rate (Yoo et al., 2001). Reduction under anaerobic conditions appears to be nonspecific, as most of a varied group of azo compounds are decolorized, although the rate of decolorization is dependent on the added organic carbon source, as well as the dye structure (Bromley-challenor et al., 2000; Stolz, 2001). Furthermore, there is no correlation between decolorization rate and molecular weight, indicating that decolorization is not a specific process and cell permeability is not important for decolorization. Thus, anaerobic azo dye decolorization is a fortuitous process, where dye might act as an acceptor of electrons supplied by carriers of the electron transport chain. Alternatively, decolorization might be attributed to non-specific extracellular reactions occurring between reduced compounds generated by the anaerobic biomass. (Van der Zee et al., 2001a). First-order kinetics with respect to dye concentration has been generally reported for the course of dye decolorization, although zero order was also observed in few cases (Van der Zee et al., 2001a; Isik and Sponza 2005 b). With a few specific dyes, such as acid orange 7 (AO7), autocatalysis by quinone-like compounds, formed during azo dye reduction, contributes to a significant extent to the overall reduction process (Van der Zee et al., 2000; Mendez-Paz et al., 2005). 2.2. Azo dye decolorization under anoxic conditions Anoxic decolorization of various azo dyes by mixed aerobic and facultative anaerobic microbial consortia has been reported (Nigam et al., 1996; Kapdan et al., 2000; Padmavathy et al., 2003; Khehra et al., 2005; Moosvi et al., 2005). Although many of these cultures were able to grow aerobically, decolorization was achieved only under anaerobic conditions. Pure bacterial strains, such as Pseudomonas luteola, Aeromonas hydrophila, Bacillus subtilis, Pseudomonas sp. and Proteus mirabilis, decolorized azo dyes under anoxic conditions (Chang et al., 2001; Chen et al., 1999, 2003; Yu et al., 2001) Azo dye decolorization by mixed, as well as pure, cultures generally required complex organic sources, such as yeast extract, peptone, or a combination of complex organic source and carbohydrate (Chen et al., 2003; Khehra et al., 2005). Glucose is the preferred substrate in anaerobic dye decolorization under methanogenic conditions, but its suitability for anoxic dye

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decolorization by facultative anaerobes and fermenting bacteria seems to vary, depending on the bacterial culture. Decolorization of Mordant Yellow 3 by Sphingomonas xenophaga Strain BN6 was greatly enhanced by glucose, whereas a significant decrease in azo dye decolorization in its presence was reported for P leuteola, Aeromonas sp. and few other mixed cultures (Haug et al., 1991; Kapdan et al., 2000; Chang et al., 2001; Chen et al., 2003). The negative effect of glucose on anoxic decolorization has been attributed either to a decrease in pH due to acid formation, or to catabolic repression (Chen et al., 2003). HPLC and mass spectrometery data from culture filtrates after the decolorization of reactive red 22 by P. leuteola, showed the presence of two aromatic amines, as well as a partially reduced product (Chang et al., 2001). This is in accordance with the two-step reduction mechanism of the azo bond proposed by Gingell and Walker (1971). 2.3. Azo dye decolorization under aerobic conditions Several bacterial strains that can aerobically decolorize azo dyes have been isolated during the past few years. Many of these strains require organic carbon sources, as they cannot utilize dye as the growth substrate (Stolz, 2001). P. aeruginosa decolorized a commercial tannery and textile dye, Navitan Fast blue S5R, in the presence of glucose under aerobic conditions. This organism was also able to decolorize various other azo dyes (Nachiyar and Rajkumar, 2003). There are only very few bacteria that are able to grow on azo compounds as the sole carbon source. These bacteria cleave –NQN– bonds reductively and utilize amines as the source of carbon and energy for their growth. Such organisms are specific towards their substrate. Examples of bacterial strains with this trait are Xenophilus azovorans KF 46 (previously Pseudomonas sp. KF46) and Pigmentiphaga kullae K24 (previously Pseudomonas sp. K24), which can grow aerobically on carboxy-orange I and carboxy-orange II, respectively (Zimmermann et al., 1982; Kulla et al., 1983). These organisms, however, could not grow on structurally analogous sulfonated dyes, acid orange 20 (Orange I) and AO7. Long adaptation of 4-aminobenzenesulfonate (4-ABS) degrading Hydrogenophaga intermedia strain S1 for growth on 4-carboxy-40 -sulfoazobenzene (CSB) as the sole organic carbon source led to the isolation of strain S5, which reduced CSB and utilized the two amine metabolites (Blumel et al., 1998). Coughlin et al. (1999) have reported the isolation of a Sphingomonas sp, strain 1CX, an obligate aerobe, which can grow on an azo dye, AO7, as sole carbon, energy and nitrogen source. This strain degraded only one of the component amines (1-amino 2-naphthol) formed during AO7 decolorization. 4-aminobenzesulfonate (4-ABS) degradation, however, required the additional presence of an unidentified strain, SAD4i (Coughlin et al., 2003). Sphingomonas ICX could also decolorize several azo dyes consisting of either 1-amino-2-naphthol or 2-amino-1naphthol coupled via the azo bond to a phenyl or naphthyl

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moiety (Coughlin et al., 1999). Similar azo dyes, such as AO6 or AO20, which lack these structures, were not decolorized. Three bacterial strains that could utilize azo dye (AO 7 or acid red 88) as sole carbon source were isolated from soil and sewage samples and were identified as Bacillus sp. OY1-2, Xanthomonas sp. NR25-2 and Pseudomonas sp. PR41-1. (Sugiura et al., 1999). Recently, four bacterial species have been isolated using methyl red as the sole carbon source. Two of these strains have been identified as Vibrio logei and P. nitroreducens. Amine products were not detected in the culture medium, indicating their degradation (Adedayo et al., 2004). The structures of few azo dyes that are mineralized under aerobic conditions are presented in Fig. 1. 2.4. Mechanism of azo dye reduction The first step in the bacterial degradation of azo dyes, in either anaerobic or aerobic conditions, is the reduction of the –NQN– bond. This reduction may involve different mechanisms, such as enzymes, low molecular weight redox

mediators, chemical reduction by biogenic reductants like sulfide, or a combination of these (Fig. 2). Additionally, the location of the reactions can be either intracellular or extracellular. 2.4.1. Direct enzymatic azo dye reduction This mechanism involves enzyme-mediated transfer of reducing equivalents, generated from the oxidation of the substrate/coenzyme to azo dyes. These enzymes can be specific, catalyzing only azo dye reduction, or nonspecific. The latter enzymes catalyze the reduction of a wide range of substrates. Due to their nonspecific nature, these enzymes gratuitously reduce azo dyes. 2.4.1.1. Under anaerobic conditions. The presence of azoreductases in anaerobic bacteria that decolorized sulfonated azo dyes during growth on solid or complex media was first reported by Rafii et al. (1990). These strains belonged mainly to the genera Clostridium and Eubacterium. Azoreductases from these strains were oxygen– sensitive and were produced constitutively and released

Fig. 1. Structures of a few azo dyes degraded under aerobic conditions.

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Fig. 2. Schematic representation of different mechanisms of anaerobic azo dye reduction. RM ¼ Redox mediator; ED ¼ electron donor; b ¼ bacteria (enzymes).

extracellularly. Later investigations made with C. perfringens showed that azo dye reduction is catalyzed by an enzyme presumed to be flavin adenine dinucleotide dehydrogenase, which can also reduce nitro aromatic compounds (Rafii and Cerniglia, 1995). Immunoelectron microscope analyses showed that the enzyme was secreted as it was synthesized (Rafii and Cerniglia, 1995). The gene for this enzyme from C. perfringens has been cloned and expressed in Escherichia coli (Rafii and Coleman, 1999). In spite of such extensive studies on oxygen-sensitive azo reductases from anaerobic bacteria by Rafii and coworkers, the source of NADH necessary for the extracellular enzyme activity is still not clear. Another mechanism of dye decolorization could involve cytosolic flavin-dependent reductases, which transfer electrons via soluble flavins to azo dyes. However, recent studies of Russ et al. (2000) with a recombinant Sphingomonas strain-BN6 have shown that the reduction of sulfonated azo dyes by cytosolic flavin-dependent azo reductases is mainly observed in vitro and is of little importance in vivo. A further indication that flavindependent azoreductases are almost completely laboratory artifacts was shown by the addition of flavins to resting cells of strain BN6, which resulted in no significant increase in azo dye reduction. These findings led Russ et al. (2000) to suggest that, in living cells with intact cell membranes, other enzyme systems and/or other redox mediators are responsible for the reduction of azo dyes. In bacteria that possess electron transport systems in their membranes, as in the case of aerobic or facultatively anaerobic bacteria, such as Sphingomonas strain BN6, the transfer of electrons from the respiratory chain to appropriate redox mediators could take place directly. If intracellular reductases are involved in the process, it may be assumed that mediators different from flavin cofactors, with a higher ability to pass through the membranes, must be involved.

2.4.1.2. Under aerobic conditions. The existence of azo reductases in obligate aerobic bacteria was first proven when two azo reductases were isolated and characterized from carboxy-orange-degrading strains Pseudomonas K22 (reclassified as Pigmentiphaga kullae K24) and Pseudomonas KF46 (reclassified as Xenophilus azovorans KF46F) (Zimmermann et al., 1982, 1984). These intracellular azoreductases showed high specificity to dye structures and reductively cleaved their carboxylated, as well as their sulfonated, structural analogs. Purified azo reductases from these two organisms were distinctly different, although both were small polypeptides without any bound flavin and required NADPH for activity. A survey of various orange dyes as substrates for carboxy orange II azo reductase showed that a hydroxy group in the ortho position to the azo bond was required. In contrast, the enzyme from P. kullae KF24 required a hydroxy group at the para position to the azo group. Although AO7 and AO20 could be decolorized by the respective azoreductases and even serve as inducers, the organisms could not utilize sulfonated dyes as the carbon source. The azoreductase from Bacillus sp. strain OYl-2 decolorized AR88, AO7 and a series of proprietary reactive dyes (Suzuki et al., 2001). This enzyme was also found to be a small monomeric protein with a molecular mass of approximately 20 KDa. Sequences of genes encoding these three specific azoreductases are now known (Suzuki et al., 2001; Blumel et al., 2002; Blumel and Stolz, 2003). Amino acid sequence alignments did not show any noticeable homology between this azoreductase and two other well-characterized azoreductases from X. azovorans KF 46 and P. kullae (Blumel and Stolz, 2003). Nonspecific enzymes catalyzing azo bond reduction have been isolated from aerobically grown cultures of Shigella dysenteriae (Ghosh et al., 1992), E. coli (Nakanishi et al., 2001), Bacillus sp. (Maier et al., 2004), Staphylococcus aureus (Chen et al., 2005) and P. aeruginosa (Nachiyar and

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Rajkumar, 2005). Where characterized, these enzymes have been shown to be flavoproteins. It has been reported that intracellular sulfonated azo dye reduction requires not only the presence of azo reductases but also a specific transport system (s), which allows the uptake of the dye into the cells (Russ et al., 2000). Currently there appears to be no available information about the transport systems for these dyes, although there are few reports on systems which are involved in the transport into bacterial cells of other sulfonated substrates, such as p-toluene sulfonate, taurine and alkane sulfonates (Locher et al., 1993; Eichhorn et al., 2000). The ability of X. azovorans KF46F and Sphingomonas sp. strain ICX to take up AO7 and reduce the dye in vivo shows the presence of transport, as well as azoreductase enzymes, in these organisms. Construction of recombinant organisms with sulfonated dye decolorizing ability, therefore, may require the transfer of the aerobic azoreductase gene into bacterial strains that are able to grow on sulfonated aromatics. These organisms generally exhibit narrow substrate specificity. Studies on 4-ABS degrading strains have also shown that they are highly specific, as they can utilize only 4-ABS and not other benzenesulfonates. (Feigel and Knackmuss, 1993; Singh et al., 2004). The 2-ABS degrading Alcaligenes sp. strain O-1 can utilize two other aromatic sulfonates, benzene and toluene sulfonate, for growth. However, cell extracts of this strain can desulfonate at least six substrates (Thurnheer et al., 1986). This suggests the presence of highly specific transport systems for the uptake of aromatic sulfonates in these cultures. Thus the derived recombinants may still have restricted substrate specificity. 2.4.2. Mediated biological azo dye reduction As highly polar sulfonated, as well as high molecular weight, polymeric azo dyes are unlikely to pass through the

cell membrane (Levine, 1991), it was suggested that reduction of these dyes could also occur through mechanisms that are not dependent on their transport into the cell. There are now many reports on the role of redox mediators in azo bond reduction by bacteria under anaerobic conditions (Keck et al., 1997; Van der Zee et al., 2001b; Dos Santos et al., 2003). Riboflavin in catalytic amounts significantly enhanced the reduction of mordant yellow 10 by anaerobic granular sludge (Field and Brady, 2003). 1-amino 2-napthol, one of the constituent amines of the azo dye, AO7, increased its decolorization rate, possibly by mediating the transfer of reducing equivalents (MendezPaz et al., 2005). The addition of synthetic electron carriers such as anthraquinone-2,6-disulphonate could also greatly enhance the decolorization of many azo dyes (Van der Zee et al., 2001b). Keck et al. (1997) reported the first example of the anaerobic cleavage of azo dyes by redox mediators formed during the aerobic degradation of a xenobiotic compound. Cell suspensions of Sphingomonas sp. strain BN6 grown aerobically in the presence of 2-naphthyl sulfonate (NS), exhibited a 10–20 fold increase in decolorization rate of an azo dye, amaranth, under anaerobic conditions, over those grown in its absence. Even the addition of culture filtrates from these cells could enhance anaerobic decolorization by cell suspensions grown in the absence of NS. Based on these observations, a mechanism was proposed for the mediated reduction of azo dyes by S. xenophaga (Fig. 3). Other bacterial cultures generating redox intermediates during the aerobic degradation of aromatic compounds can also lead to the enhancement of dye decolorization in anaerobic conditions (Keck et al., 1997). Recently, Chang et al. (2004) also showed that the addition of culture supernatants containing metabolites of a dye-decolorizing strain, E. coli strain NO3, enhanced azo dye decolorization rates.

Fig. 3. Proposed mechanism for the redox mediator (RM)—dependent reduction of azo dyes by strain BN6 (adapted from Keck et al., 1997).

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Van der Zee et al. (2003) have reported that activated carbon, which is known to have quinone groups at its surface, enhanced dye decolorization. This is probably one of the first examples of biocatalysis mediated by activated carbon. 2.4.3. Azo dye decolorization by biogenic inorganic compounds Azo dye decolorization can occur from purely chemical reactions with inorganic compounds such as sulfide and ferrous ion that are formed as end products of metabolic reactions under anaerobic conditions. It has been shown that H2S generation by SRB resulted in the extracellular decolorization of azo dyes (Yoo et al., 2000; Diniz et al., 2002). Sulfate-influenced dye reduction correlated with biogenic sulfide formation under methanogenic conditions. In the absence of sulfur compounds, dye decolorization readily occurred in the presence of granular sludge, demonstrating the importance of enzymatic mechanisms. An analysis of decolorization kinetics in batch reactor and in laboratory scale anaerobic sludge bed reactors indicated that the relative importance of chemical dye reduction mechanisms in high rate anaerobic bioreactors is small, due to the high biomass in the reactors (Van der Zee et al., 2003). 3. Degradation of aromatic amines Aromatic compounds possess a large negative resonance energy, resulting in thermodynamic stability. Microorganisms, particularly bacteria, have evolved enzyme systems that degrade the benzene structure under aerobic and anoxic conditions (Gibson and Subramanian, 1984; Schink et al., 2000). Common to both oxygen-dependent and anoxic metabolism of aromatic compounds is a separation into peripheral and central pathways (Heider and Fuchs, 1997). Peripheral pathways convert the large variety of compounds into a few central intermediates. In aerobic metabolism, the initial reactions involve the replacement of other functional groups of the aromatic ring with hydroxyl groups, followed by cleavage by incorporating two oxygen atoms. These reactions are catalysed by hydroxylases and oxygenases. Under anoxic conditions, dearomatization is achieved by ring reduction and also includes other unique reactions such as carboxylation, reductive dehydroxylation and addition reactions, which are absent in the aerobic metabolism (Heider and Fuchs, 1997). 3.1. Degradation of aromatic amines under anaerobic conditions Decolorization of azo dyes in anaerobic environments leads to the formation of aromatic amines, many of which were assumed to resist further degradation under these conditions (Stolz, 2001). Nevertheless, mineralization of few simple aromatic amines has been reported under

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methanogenic conditions. They include the three isomers of aminobenzoate, 2- and 4-aminophenols, 2, 4-dihydroxyaniline and 5-aminosalicylic acid (5-ASA). (Connor and Young, 1993; Razo-Flores et al., 1997b; Kalyuzhnyl et al., 2000; Yemashova et al., 2004). Complete degradation of azo disalicylate and partial mineralization of the azodyes, Mordent Orange 1, AO6, AO7 and AO52, under methanogenic conditions, has been reported (Donlon et al., 1997; Razo-Flores et al., 1997b; Savelieva et al., 2004; Yemashova et al., 2004). Many reports have shown that sulfonated aromatic amines (SAA) are nonbiodegradable under methanogenic conditions (Tan et al., 2005). 3.2. Aerobic biodegradation of aromatic amines Aromatic amines formed from azo dye reduction, have been reported to be more easily degraded under aerobic conditions (Brown and Laboureur, 1983; Haug et al., 1991; Ekici et al., 2001). The aerobic biodegradation of aniline and p- amino benzoate appears to be ubiquitous (Gheewala and Annachhatre, 1997; Blumel et al.,1998). Bacteria capable of biodegrading 5-ASA are less numerous (Stolz et al., 1992). Tan et al. (1999) have reported that 4-aminophenol (4-AP) and 5-ASA tend to autoxidise in the presence of oxygen. However, the autoxidation rate for 4-AP was orders of magnitude greater than that for 5-ASA. Hence biodegradation of 5-ASA was possible, whereas 4-AP removal was mainly due to autoxidation under aerobic conditions (Tan et al., 1999). A group of aromatic amines difficult to degrade even under aerobic conditions are represented by aryl sulfonates, aminobenzene (ABS) and aminonaphthyl sulfonates (ANS), which are the constituents of many azo dyes. Among the three isomers of ABS, 4-ABS appears to be more susceptible to biodegradation than 2- and 3-ABS. A co-culture consisting of H. palleroni (Strain S1) and Agrobacterium radiobacter (Strain S2), as well as a few pure bacterial strains, degrade 4-ABS (Feigel and Knackmuss, 1993; Perei et al., 2000; Singh et al., 2006). The uniqueness of these strains is that they can degrade only 4-ABS and not other aromatic sulfonates. 2-ABS-degrading cultures appear to be rare (Thurnheer et al., 1986; Tan et al., 2005).There is only one report of the isolation of a 2-ABS-degrading pure culture, which was characterized to be of the genus Alcaligenes (strain O-1). This strain can utilize two other sulfonated aromatics, benzene and toluene sulfonates, as growth substrates. (Thurnheer et al., 1990). 2-ABS degradation is a plasmid-associated trait in this organism (Jahnke et al., 1990). Not many reports are available on 3-ABS degradation (Kolbener et al., 1994; Nachiyar and Rajkumar, 2003). Recent studies by Nachiyar and Rajkumar (2004) seem to suggest aniline as an intermediate during the aerobic degradation of 3-ABS in P. aeruginosa. This appears to be an unusual desulfonation reaction, as aerobic desulfonation reactions generally involve hydroxylation catalysed by either mono or dioxygenases and catechol sulfonates have been identified

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as intermediates during 2- as well as 4-ABS (Feigel and Knackmuss, 1993; Junker et al. 1994). Few bacterial cultures utilizing naphthyl amines as the sole organic carbon source have been reported. Sphingomonas sp. strain ICX decolorized AO7 and degraded 1-amino-2-naphthol (Coughlin et al., 1999) P. aeruginosa degraded 1,4-diaminonaphthalene (Nachiyar and Rajkumar, 2004). Sulfonated naphthylamines are among the most common products of bacterial decolorization of azo dyes. Pure cultures have been isolated belonging to the genera Pseudomonas, Moraxella and Arthrobacter, which can either degrade 2-aminonaphthyl sulfonate (2ANS) or utilize it as sulfur source (Nortemann et al., 1986; Wittch et al., 1988; Ohe et al., 1990; Rozgaj and Glancer, 1992). Extensive studies have been carried on the degradation of 6-aminonaphthyl-2-sulfonate (6A2NS). A co-culture consisting of 11 bacterial strains was able to utilize 6AN2S. None of the individual strains could degrade 6AN2S, indicating that apparently each strain participates in the degradation process (Rozgaj and Glancer, 1992). A mixed bacterial culture utilized 6A2NS as the sole source of carbon, energy, nitrogen and sulfur. From this culture, strain BN6 was isolated. This strain could convert 6A2NS to pyruvate and 5-ASA. 5-ASA was excreted in almost stiochiometric amounts and was utilized by other members of the culture (Nortemann et al., 1986). After some years of cultivation of S. xenophaga strain BN6 and some accompanying bacteria, strain BN12, which could completely degrade 6A2NS, was isolated. This strain was identified as Pseudoaminobacter salicylatoxidans (Kampfer et al., 1999). To date, this is the only organism that is known to completely degrade 6A2NS. Based on their studies with strain BN6, Nortemann et al. (1994) suggested the following requirements for the productive breakdown of substituted naphthalene sulfonates (i) an uptake system and naphthyl sulfonate dioxygenase with low substrate specificity (ii) counteraction of rapid autoxidation of dihydroxy naphthalenes and other metabolites. Studies on the degradation of naphthalenesulfonates by S. xenophaga are reviewed by Stolz (1999). Available literature on degradation of sulfonated aromatic amines (SAA) show that cultures generally have a narrow substrate range. Mixtures of aromatic sulfonates can only be degraded by mixed bacterial consortia, which harbor diverse activities for congeneric substrates and metabolites (Hopper, 1991). Few aromatic amines, such as 5-ASA, phenylenediamines, aminophenols and aminonaphthols, tend to autoxidize in the presence of oxygen. (Jensen et al., 1993; Kudlich et al., 1999). Oxygen reacts with the aromatic products, via free radical reactions, resulting in the formation of undesirable colored oligomers and polymers that may be toxic and mutagenic (Field et al., 1995). Though the autoxidation process eliminates the aromatic amines, the products formed are more recalcitrant to biological degradation. Thus, for biological degradation of

these compounds, the rate of degradation must be higher than the rate of autoxidation. 3.3. Anaerobic–aerobic biological treatment for azo dye decolorization and degradation A wide range of structurally diverse dyes is used in the textile industry in the same unit. The emerging effluent thus contains many of these dyes. Treatment of such an effluent requires the application of non-specific processes (Correia et al., 1994). Owing to this constraint, treatment of textile and other dye-containing wastewaters is commonly carried out using chemical coagulation/flocculation, followed by activated sludge (Pophali et al., 2003; Georgiou et al., 2005). Bacterial decolorization of azo dyes under methanogenic conditions is non-specific. Although azo dye decolorization under anoxic conditions is also non-specific, limitations of this method seem to be the requirement for yeast extract or peptone, thus making the process economically inviable for industrial-scale application unless alternate cheaper sources are identified (Nigam et al., 1996; Chen et al., 2003; Moosvi et al., 2005). Except for a few, the aromatic amines formed from decolorization of azo dyes are recalcitrant to biodegradation under anaerobic conditions. These anaerobically decolorized effluents can still be hazardous, as many aromatic amines are toxic. Thus their removal, which requires aerobic conditions, is essential. Decolorization and degradation of azo dyes in biological processes based on bacterial activities thus requires anaerobic–aerobic conditions. These treatments can be carried out either sequentially or simultaneously. For the sequential treatment, anaerobic and aerobic environments can be provided either in a single reactor for different periods or in two separate reactors. The feasibility of this strategy was first demonstrated for the sulfonated azo dye, Mordant Yellow 3 (Haug et al., 1991). In a simultaneous treatment system, decolorization takes place in anaerobic zones of the biofilm or in immobilized microbes entrapped in a gel matrix (Field et al., 1995; Kudlich et al., 1996). Different reactor configurations used for anaerobic/ aerobic steps, and their efficiencies, have been excellently reviewed recently by Van der Zee and Villaverde (2005). These include anaerobic high-rate reactors, such as upflow anaerobic sludge blanket, fixed film, rotating biological contactors and anaerobic baffled reactors for anaerobic processes and activated sludge and rotating biological contactors for aerobic treatment (Isik and Sponza, 2004; Ong et al., 2005; Van der Zee and Villaverde, 2005). An auxiliary substrate is generally needed for the decolorization. Theoretically, the required amount of electrondonating substrate is low, 32 mg COD per mmol monoazo dye. However, the actual requirement is much higher due to competition from other reactions for reducing equivalents. In most of the studies, primary substrate COD was usually in great excess. The major part of the substrate may be consumed in the anaerobic step. A few amines may also

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be mineralized. As a result, a small fraction of the auxiliary substrate or its metabolites and undegraded aromatic amines serve as the carbon source for the organisms in the aerobic reactor. Color removal levels ranging from 70–95% have been reported in anaerobic/aerobic reactors (Van der Zee and Villaverde, 2005). However, the fate of the aromatic amines has been specifically addressed by few investigators. Some of these studies show partial or complete removal of many aromatic amines in the aerobic stage. Furthermore, decrease in toxicity (shown by suppression of bacterial luminescence) between the effluents of the anaerobic stage and anaerobic–aerobic stage seems to provide indirect evidence for the removal of aromatic amines. As reviewed by Pinheiro et al. (2004), various substituted aminobenzene and aminonaphthalene and aminobenzidine compounds have been found to be aerobically biodegradable. However, they generally require the enrichment of specialized cultures. Biodegradability of SAA has been demonstrated only for a few simple ABS and ANS compounds. Most of these studies are still at the laboratory scale and with synthetic wastewaters. Reports on pilot-scale and fullscale implementation of anaerobic–aerobic biological treatment are still scarce. Carliell et al. (1996) investigated laboratory and full-scale trials on the treatment of exhausted reactive dye bath effluent and reported that cotreatment of the effluent in an existing sewage sludge digester exhibited promising results. Delee et al. (1998) have reviewed the reports on full-scale and pilot-scale plants and some of their limitations. A two-stage fixed bed pilot plant for on-site anaerobic decolorization of textile wastewater was reported by Georgiou et al. (2005). Anaerobic (with facultative anaerobic bacterial culture)– aerobic sequential system was used for color and COD removal from real textile wastewater at pilot scale (Kapdan and Alparslan, 2005). These recent studies with real wastewaters have shown that the addition of an anaerobic unit prior to activated sludge significantly improves the effluent quality with respect to color. There appears to be only one report on the successful full-scale implementation of anaerobic technology in combination with aerobic membrane technology for the treatment of more than 1000 L3 wastewaters per day from textile processing industry. (Stolz, 2001, VA Tech WABAG, 2003). 4. Conclusion The presence of dyes imparts an intense color to effluents, which leads to environmental as well as aesthetic problems. The treatment of azo-dye-containing wastewaters still presents a technical challenge. As regulations are becoming even more stringent, there is urgent need for technically feasible and cost-effective methods. The available literature seems to indicate that anaerobic–aerobic biological methods may be appropriate for the treatment of dye-containing wastewaters. However, there is a still a need to assess the extent of mineralization of aromatic amines,

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as many amines can undergo autoxidation, leading to the formation of soluble recalcitrant polymers, which may be toxic. Degradation of many amines, including SAA, requires the presence of specialized cultures. SAA degraders have a very narrow substrate range. Hence there is a requirement for developing microbial consortia that harbor genes for the rapid degradation of mixtures of aromatic amines. Such cultures may have to be used for the bioaugmentation of aerobic treatment units. Molecular biology techniques may also be used to improve the strains so that rapid mineralization of aromatic amines can be achieved. Their use, however, requires caution. It may also be necessary to combine AOP with biological processes to achieve the required degree of treatment of dye-containing wastewaters so that regulatory standards can be met.

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