Biotechnology Advances 29 (2011) 111–123

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Research review paper

Bioremediation of wastewaters with recalcitrant organic compounds and metals byaerobic granules

A.M. Maszenan a,b, Yu Liu a,b,⁎, Wun Jern Ng a,b

a Division of Environmental and Water Resources Engineering, School of Civil and Environmental Engineering, Nanyang Technological University, 50 Nanyang Avenue,Singapore 639798, Singaporeb Nanyang Environment and Water Research Institute, Nanyang Technological University, 50 Nanyang Avenue, Singapore 639798, Singapore

⁎ Corresponding author. Division of Environmental aing, School of Civil and Environmental Engineering, Nan50 Nanyang Avenue, Singapore 639798, Singapore. Tel.:

E-mail address: cyliu@ntu.edu.sg (Y. Liu).

0734-9750/$ – see front matter © 2010 Elsevier Inc. Aldoi:10.1016/j.biotechadv.2010.09.004

a b s t r a c t

a r t i c l e i n f o

Article history:Received 5 July 2010Received in revised form 17 August 2010Accepted 28 September 2010Available online 16 October 2010

Keywords:Aerobic granulesBioremediationRecalcitrant organic compoundsWastewater treatment

Compared to activated sludge flocs, aerobic granules have a regular shape, and a compact and dense structurewhich enhances settleability, higher biomass retention, multi-microbial functions, higher tolerance totoxicity, greater tolerance to shock loading, and relatively low excess sludge production. The potential forimproved process efficiency and cost-effectiveness can be attractive when it is applied to both municipal andindustrial wastewaters. This review discusses potential applications of aerobic granulation technology inwastewater treatment while drawing attention to relevant findings such as diffusion gradients existing inaerobic granules which help the biomass cope with inhibitory compounds and the ability of granules tocontinue degradation of inhibitory compounds at extreme acid and alkaline pHs.

nd Water Resources Engineer-yang Technological University,+65 6790 5254.

l rights reserved.

© 2010 Elsevier Inc. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1112. Degradation of phenol . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1123. Degradation of p-nitrophenol . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1134. Degradation of chlorinated phenols . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1145. Degradation of pentachlorophenol . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1146. Degradation of pyridine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1157. Degradation of phthalic acids and esters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1158. Degradation of tert-butyl alcohol . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1169. Degradation of methyl tert-butyl ether . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 116

10. Degradation of chloroanilines . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11711. Degradation of metal-chelating agents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11712. Bioaccumulation of pigments/dyes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11813. Bioaccumulation of heavy metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11814. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 119References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121

1. Introduction

Aerobic granulation is a highly complex process involvingbiological, chemical and physical factors (Liu and Tay, 2002, 2004).

Aerobic granules are formed by self-aggregation of microorganisms inthe absence of a support carrier. The granules are dense microbialaggregates assembled by consortia of microbes wherein the variousspecies perform possibly different and specific roles in biodegradationof organics during wastewater treatment (Beun et al., 1999;Morgenroth et al., 1997). Close association among the microbialentities and degradation of organic pollutants via multiple steps resultin the layered structures seen in aerobic granules (Tay et al., 2002; Liet al., 2008). This layered structure also create concentration gradients

112 A.M. Maszenan et al. / Biotechnology Advances 29 (2011) 111–123

of target compounds inside the granule and these protect micro-organisms from the impact of direct acute toxicity associated with thecompounds. There is strong evidence aerobic granulation is poten-tially capable of handling a wide spectrum of wastewaters includinghigh-strength recalcitrant industrial and low-strength municipalwastewaters (Maszenan and Liu, 2009).

Man-made organic compounds or xenobiotics are characterized bytheir complex ring structures or substituents that make them resistant tobiodegradation and so persistent in the environment. Generally, suchorganics canbe classified into4main types: (i) the saturatedhydrocarboncompounds which has no double or triple bonds. The degradability ofsuch compounds decreases as their molecular weight and degree ofbranching increase, (ii) the aromatic hydrocarbon compounds whichcontain conjugated double bonds. The aromatics with one or two ringsdegrade easily, while higher molecular weight aromatics are lessbiodegradable. An example of a relatively degradable aromatic is benzenewhich is the simplest of the aromatics, (iii) the asphalteneswhich includethe phenols, fatty acids, ketones, esters and phorphyrins, and (iv) theresins which include the pyridines, quinolines, carbozoles, sulfoxides andamides (Colwell and Walker, 1977; Leahy and Colwell, 1990). Thebiodegradation of these organics in terms of increasing difficulty is: n-alkanesbbranched alkanesb low molecular weight aromaticbcyclicalkanesbaromatic alcohols and esterbnitrobenzenebchlorinated ben-zene (Perry, 1984).

Industrial wastewater can contain a variety of aromatic pollutantssuch as benzoate, phthalate, phenol and its derivatives, and varioushalogenated compounds. Some of these are toxic and have been listedas priority pollutants by the US Environmental Protection Agency (USEPA). These compounds can have been extraneously introduced intothe environment in large quantities due to their widespread use asherbicides, insecticides, fungicides, solvents, plasticizers, cleaningagents, propellants, gasoline additives and degreasers (Bhatt et al.,2007). Some of these compounds can be degraded by acclimatedmicroorganisms in wastewater treatment systems. However, otherscan remain undegraded and therefore require application of systemswhich support novel consortia of organisms to be effective.

The performance of a biological wastewater treatment system isdependant on appropriate active biomass density and its ability todegrade the organics present in the wastewater, the biodegradationkinetics, reactor configuration, and the availability of nutrients such asoxygen. The morphology of the biomass can impact on processefficiency and the latter can be enhanced by selecting an appropriatemorphology such as granular sludge. Aerobic granular sludge canhave a much higher conversion capacity and rate due to the higherbiomass concentration. Additionally granular sludge has the advan-tage of enhanced liquid–solids separation in the mixed liquor due tosuperior settling properties. Granulation can, however, be compro-mised with loss of auto aggregating capability arising from lowexopolysaccharides (EPS) protein. This can be associated with highorganic loadings (Adav et al., 2010a). Other factors which can impactgranulation include filamentous overgrowth, cells lysis in granules,and disintegration of aerobic granule due to anaerobic core formation.There are differences between the EPS in sludge flocs and aerobicgranules and one of these is the existence of a reversible, pHdependent sol gel transition which only exists in EPS extracted fromaerobic granules and not present in sludge floc EPS (Seviour et al.,2009). In granular EPS, EPS gel, exopolysaccharides or glycoside, arethe important gelling agents for the granules (Seviour et al., 2009). Ina study on starved aerobic granules, where the carbon substrate hasbeen exhausted, α and β amylase were detected in the inner core ofthe aerobic granule, which suggested endogenous respiration oc-curred at the core of the granules. Amylase activities promotepolysaccharide hydrolysis and the formation of cavities in the innercore. However, this internal polysaccharide hydrolysis would lead toeventual loss of granules integrity (Lee et al., 2009). The studiessuggested EPS polysaccharide hydrolysis adversely impacted aerobic

granules stability and eventually led to disintegration. This reviewdiscuss the application of aerobic granules for treating industrialwastewater, the mechanism involve in granule formation withmicrobial population selection and EPS types.

2. Degradation of phenol

Phenol can be inhibitory, even at low concentrations, and mayconsequently upset the conventional activated sludge process.Industrial wastewaters such as those from oil refineries, andpharmaceutical and pesticide plants can be major sources of phenolicwastewater. Aerobic granules used to treat such wastewaters havebeen found to be less susceptible to phenol inhibition due to thecompact and dense granule structure, which then served as a phenoldiffusion barrier. For instance, aerobic granules were able to removealmost all incoming phenol within 1 h at an influent phenolconcentration of 500 mg L−1 (Tay et al., 2004a). The phenol-degrading aerobic granules were developed from activated sludgeacclimated to phenol (Tay et al., 2004a). The phenol loading wasincreased stepwise until 2.5 kg phenol m−3 day−1. It was postulatedthat by increasing phenol concentration, microorganisms whichbetter resist phenol toxicity will be selected and aggregated to formthe granules. The compact structure of the granules retained thephenol degraders and protected the microbial population in thegranules' inner core from the inhibitory effect of phenol (Tay et al.,2004a; Jiang et al., 2002, 2006; Liu et al., 2009; Ho et al., 2010).

While the granular structure protected the microorganisms livingwithin the granules, the existence of diffusion gradients can poseproblems with regards to nutrient and oxygen transport within thegranules. Thus, application of granules does require granule size to becontrolled (Tay et al., 2004a). Evidence has shown that the appliedphenol loading had significant effect on granule structure, microbialactivity, and their metabolism. Good settling granules with compactstructure, high specific oxygen uptake rate, and catechol dioxygenaseactivities have been developed at phenol loadings to 2.0 kgphenol m−3 day−1 beyond which aerobic granules with loosestructure and poor settleability resulted (Jiang et al., 2004b).Compared to the activated sludge process, aerobic granules have ashorter lag time for phenol biodegradation and could cope withhigher phenol concentrations before inhibition set in. A desirablefeature for industrial application is capacity to cope with extreme pHs,either acidic or basic while remaining capable to degrade therecalcitrant compound. Recent study by Ho and co-workers, havereported that aerobic granules can degrade phenol at extreme pHs, pH3 and 12 (Ho et al., 2010).

Phenol degradation by aerobic granules has been reported to besubject to the meta cleavage pathway as 2,3 dioxygenase activity ismuch higher than 1,2 dioxygenase activity which would haveinvolved the ortho cleavage pathway (Jiang et al., 2004b, 2006; Tayet al., 2005c). Anabolic and catabolic regulation favour phenoldegradation via the meta cleavage pathway so as to produce energyfor the synthesis of proteins and enzymes and this is triggered byhigher phenol concentrations (Jiang et al., 2004b, 2006). Higherphenol loadings drive themicrobial population towards themetabolicpathway which can cope with energy demand related to non-growthrelated activities which are needed to counter the inhibition ofcellular activity (Jiang et al., 2004b, 2006). Some of this non-growthrelated energy is also used to maintain cell membrane integrity andactive transport of substrate into the cell. This energy requirement isexpected to be relatively high due to the higher phenol concentration.The latter may also result in greater EPS synthesis which would alsorequire energy (Jiang et al., 2004b, 2006).

Enzymatic activity assay of phenol degradation has shown therewas no predominant phenol degradation pathway among the strainsat high phenol concentrations (Jiang et al., 2006). As mentionedearlier, the meta pathway was strongly induced and favoured under

Fig. 1. Phenol-degrading aerobic granule with fluffy appearance and highly dominatedby filamentous bacteria (a) and its close up view of the filamentous bacteria residing inaerobic granule (b).

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high phenol concentrations as Type I strain with high C23O activitywill have advantage for survival (Jiang et al., 2006). However, in anaerobic granule challenged with high phenol concentration, microbialstrains without this advantage could still grow as shown by thepresence of other strains belonging to Types I, II and III catabolic traits(Jiang et al., 2006). To investigate the functional role of microbialpopulations in aerobic granules, 2 out of 8 dominant strains have beenstudied; one had demonstrated highest phenol degradation capabil-ity, while the other possessed strong aggregation capability tomaintain the granular structure. It is important to note that bothhigh and low Ks phenol degraders have been isolated from aerobicgranules exposed to high phenol concentration (Tay et al., 2004b).High Ks phenol degraders will reside near the granule surface while,due to diffusion limitation, low Ks phenol degraders will be localisedin a granule's interior.

In addition to phenol degrading capability, a microbial populationwith good flocculation capability is also crucial for providingstructural stability of phenol degrading aerobic granules (Tay et al.,2004b) and the speed with which granules can be formed. Forinstance it was possible to acclimate acetate-fed aerobic granules forphenol biodegradation within a short start-up period. The acetate-fedgranules were developed in sequencing batch reactors with acetate asthe sole carbon source at a loading rate of 3.8 kg m−3 day−1, phenolwas then introduced at 0.6, 1.2 and 2.4 kg m−3 day−1. The granulesadapted to the phenol and stabilized within a week. Granules exposedto 0.6 and 1.2 kg m−3 day−1 phenol did not suffer from inhibitionalthough there was a lag period before phenol biodegradationoccurred. At the high phenol loading rate of 2.4 kg m−3 day−1,there was a sharp build up of phenol initially, as the aerobic granuleswere unable to biodegrade phenol. However, the build up subse-quently declined when the granules had adapted to the high phenolconcentration by increasing surface hydrophobicity and/or increasingextracellular polymeric substances (EPS) production (Tay et al.,2005b,c). The microbial population then also developed the necessaryenzymes for phenol degradation. The compact structure of theacetate-fed granules had provided protection for the microbialpopulation (Tay et al., 2005b,c). This would suggest aerobic granulesgrown on a more benign substrate can be used as seed to generatespecific functional granules for toxic substrate degradation. Thiswould further suggest it may be possible to seed a new reactor withgranules drawn from a reactor treating a different wastewater.

When Adav et al. (2007a) explored aerobic granule degradation ofphenol up to 1000 mg/L, it was revealed by fluorescent staining andconfocal laser scanning microscopy that an active biomass accumu-lated at the granule outer layer, and a strain with high ability todegrade and tolerate phenol toxicity could be identified, Candidatropicalis, by 18S rRNA sequencing. This removed phenol at amaximum rate of 390 mg-phenol/g VSS h at pH 6 and 30 °C. Inhibitionwas observed only at phenol concentrations higher than 1000 mg/L.Adav et al. (2007a) reported the Candida strain was primarilydistributed throughout the surface layer of the granules. Thus themass transfer barrier provided by the granule matrix was not aconstraint to the reaction rates of the strain by virtue of its location onthe granule.

The presence of alternative carbon substrates such as phenol andacetate can modify the catabolic potential of a microbial populationwithin the aerobic granule (Tay et al., 2005b). A study by denaturinggradient gel electrophoresis (DGGE) revealed that the microbialpopulation in acetate-fed aerobic granules was more diversecompared to acetate-fed aerobic granules subsequently adapted tophenol, indicating community restructuring have taken place as thebiomass acclimated to become stable phenol-degrading aerobicgranules (Tay et al., 2005c).

Overgrowth of filamentous bacteria had been observed in phenol-degrading aerobic granules. The slow-growing filamentous bacteriatypically have lower Ks compared to floc forming bacteria. Filamen-

tous bacteria have been found to be dominant on the granule surface(Fig 1a, b). Adav et al. (2007b) had attributed the filamentous growthto nutrient deficiency since a low-nutrient environment would favourfilamentous growth due to their low Ks. It is also important to notefilamentous bacteria may able to tolerate phenol toxicity because theyare protected by a sheath structure formed from EPS (Tay et al.,2005c). The sheath–EPS relationship was reported by Jiang and co-workers. There has, however, been no agreement on the relevance ofthe sheath as Adav and co-workers have seen no correlation betweenEPS and filamentous bacteria proliferation in phenol challengedaerobic granules (Adav et al., 2007b). Recently, the role of EPS inabsorbing inhibitory compounds has been proposed as anothermechanism to overcome an inhibitory effect (Adav et al., 2010b).

In summary, phenol-degrading granules have exhibited excellentability to degrade phenol, and the compact structure of the granulesprovide spatial niches for coexistence of competitive superior andinferior strains with similar functions. This suggested it would bepossible to develop aerobic granules targeting specific recalcitrantpollutants by bioaugmenting with a group of strains with similarfunctions. The consortia so developed can then deal with differentenvironmental conditions by exhibiting different responses. Aerobicgranulation may therefore be applicable to other aromatic com-pounds, such as nitrophenol, nitrobenzene and nitrotoluene whichare of industrial importance as the main raw materials for theproduction of dyes, pharmaceuticals, pesticides and explosives.

3. Degradation of p-nitrophenol

p-Nitrophenol (PNP) is a widely used nitroaromatic compoundthat poses a significant risk to the environment and public health. PNP

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is used in themanufacture of the drug acetaminophen, pesticides suchas methyl and ethyl parathion, and as a conditioner for leathertreatment (Spain and Gibson, 1991). Due to its persistence, PNP iscommonly found in industrial wastewater and has contaminated bothsubsurface water and groundwater (Labana et al., 2005). The risksassociated with PNP include its toxicity and mutagenic effects. USEPAhas listed PNP as a priority pollutant and recommended a concentra-tion limit of 10 ng L−1 in natural waters (EPA, 1976) and a limit onindustrial effluent discharge of not more than 162 μg/L (EPA, 1988).PNP can be degraded by microbial consortia under aerobic conditions(Crawford, 1995), but acute toxicity of PNP can seriously hindermicrobial activity (Yi et al., 2006).

Yi et al. (2006) successfully developed PNP-degrading aerobicgranules at a PNP loading rate of 0.6 kg m−3 day−1 with glucose asco-substrate at 500 mg/L to promote active biomass growth duringthe initial acclimation stage. The glucose did not interfere with PNPdegradation nor was glucose preferentially degraded over PNP. Theseindicated the co-substrate strategy is useful for developing PNPdegrading aerobic granules. The compact structure of the PNPgranules would shield the microbial population from PNP toxicity.Specific degradation rates of PNP by granules tended to increase withthe PNP concentration until 40.1 mg PNP L−1, peaked at 19.3 mg ofPNP g−1 VSS h−1, and declined with further increases in the PNPconcentration (Yi et al., 2006).

The PNP-degrading granules contained diverse microbial mor-photypes, and PNP-degrading bacteria accounted for 49% of the totalculturable heterotrophic bacteria. Denaturing gradient gel electro-phoresis analysis of 16S rRNA gene fragments showed a gradualtemporal shift in microbial community succession as the granulesdeveloped from the seed activated sludge (Yi et al., 2006). The specificoxygen utilization rate at 100 mg L−1 PNP was found to increase withthe evolution of smaller granules to large granules, suggesting that thegranulation process enhanced metabolic efficiency toward biodegra-dation of PNP (Yi et al., 2006).

4. Degradation of chlorinated phenols

Chlorinated phenolic compounds are widely used as industrialchemicals for wood preservation, manufacturing of pulp and paper aswell as in agricultural activities via application of pesticides. Due totheir toxicity, even at low concentrations, chlorinated phenoliccompounds have been classified as priority pollutants. The persisten-cy of chlorinated phenolic compounds and their toxicity is mainly dueto the strong C–Cl bonds, the degree of chlorination and its “lipidloving” properties (Loehr and Krishnamoorthy, 1988; Mohn andKennedy, 1992). Chemically, the accumulation of chlorinated pheno-lic compounds in a biological entity is correlated to the octanol waterpartition coefficient. The coefficient increases with increase in thenumber of chlorine atoms attached to the phenol's benzene ring(Annachhatre and Gheewala, 1996), whereas the aerobic biodegra-dation rate of chlorinated phenolic compounds decreases withincreasing number of substituted chlorine atoms in the compound.The substituted chlorine atom inhibits monooxygenase and dioxy-genase enzymatic action which cleave the benzene ring under aerobiccondition. On the other hand, under anaerobic condition, thebiodegradation rate increases with increased number of chlorineatoms due to reductive dechlorination activity (Wang et al., 1998;Tartakovsky et al., 2001). This is supported by experimentalobservation of the rate of biodegradation of monochlorophenoloccurring at much slower rate compared to the initial dechlorinationstep. Thus, it has been suggested to include anaerobic and aerobicbiodegradation steps to enhance the overall removal efficiency ofchlorinated phenolic compounds (Armenante et al., 1999).

4-mono-chlorophenol (4CP), a by-product of wastewater chlori-nation process, can be derived after breakdown of pesticides andchlorinated aromatic compounds (Pritchard et al., 1987). Compared to

other highly chlorinated compounds, the presence of only onechlorine atom in 4CP makes this compound much easier to degradeunder aerobic conditions. Carucci et al. (2009) successfully developedaerobic granules for the biodegradation of 4CP in the presence ofacetate as co-substrate. It was found that at high 4CP concentration,even unacclimated granules can overcome the toxic effects due toboth the high biomass concentration and diffusion gradient whichexist in the granules.

Wang et al. (2007) cultivated aerobic granules for 2,4-dichlorophenol(DCP) bioremediation by controlling settling time and gradually replacingglucosewith 2,4-DCP. 95%CODand 94% 2,4-DCP removalswere achieved,and the biodegradationwas found to obey theHaldanemodelwith a peakrate of 39.6 mg 2,4-DCP g−1 VSS h−1 at a 2,4 DCP concentration of105 mg L−1. In this study, the removal characteristic of 2,4-DCP wassimilar to glucose, suggesting that the benign substrate was notpreferentially degraded over 2,4-DCP when both were supplied together(Wang et al., 2007). It appears that the key strategy in cultivating aerobicgranules for 2,4-DCP is to introduce glucose as co-substrate by whichmicrobial growth for granule formation would be enhanced. In fact, theaddition of glucose not only increased the utilization rate of 2,4-DCP, butalso accelerated NADH production which is needed for 2,4-DCPdegradation (Bali and Sengul, 2002; Wang et al., 2007). Wang et al.(2007) also investigated the release of chloride from 2,4-DCP biodegra-dation, and found that the key step in the degradation of chlorinatedcompounds was the cleavage of carbon–chlorine bond, and this can beachieved by either dehalogenation after ring cleavage or dehalogenationbefore ring cleavage (Snyder et al., 2006). In the 2,4-DCP degradation, theinitial product, 3,5-dichlorocatechol,wasproducedby enzymatic action of2,4-dichlorophenol hydroxylase. The 3,5-dichlorocathechol was furthertransformed to 2,4-dichloro cis, cismuconate by the 3,5-dichlorocate-1,2-oxygenase activity and then finally mineralised to CO2. It was concludedby Wang et al. (2007) that the inducement and availability of thesespecific enzymeswithin an aerobic granular sludge consortia is importantfor reductive dechlorination after ring cleavage.

Carucci et al. (2008) investigated the possibility of using acetate-fedaerobic granules for the degradation of 4-mono-chlorophenol (4-CP)and 2,4,6-tri-chlorophenol (TCP). The concentrations in the feed werecontrolled at the levels between 0 and 50mg L−1 for 4-CP, and between0 and 15 mg L−1 TCP. A possible biodegradation pathway of chlorinatedphenol involves meta cleavage of 4-chlorophenol which would yield achlorinated aliphatic compound, 5-chloro-2 hydroxymuconic acidsemi-aldehyde as end product. This can then be further degraded withthe release of chloride ions (Westmeier and Rehm, 1987; Hollender etal., 1997). Carucci et al. (2008) observed the formation of 5-chloro-2hydroxymuconic acid semi-aldehyde followed by together withchloride release which confirmed complete mineralization of 4-CP byaerobic granules. For TCP degradation, anaerobic feeding and control ofdissolved oxygen concentration in the bulk liquid were employed inorder to create an anaerobic core in the aerobic granules due to diffusionlimitation. This meant TCP reductive dechlorination could be achievedwhen thebulk liquidwasunder aerated conditions. The studybyCarucciet al. (2008) showed that aerobic granules grown on acetate could notonly degrade 4CP but also showed good resistance to high 4CPconcentrations in the influent even when unacclimated. The presenceof TCP did not irreversibly inhibit microbial activity, and complete TCPdegradation was achieved after acclimation. Based on smaller footprint,simple and flexible system, and short start-up time as decision makingparameters, aerobic granules operated under granular sequencingbatchreactors (GSBR) is the preferred treatment technology for treating 4-CPcontaminated wastewater (Carucci et al., 2010).

5. Degradation of pentachlorophenol

Pentachlorophenol (PCP) is a highly recalcitrant compoundwith toxicand carcinogenic properties, and has been used extensively in herbicides,fungicides, and inpreservatives forwoodandwoodproducts. It is listedon

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the Priority List of Pollutants by US EPA (1988). Under aerobic condition,PCP can be hydroxylated to tetrachloro-p-hydroxyquinone, which isconverted further to trichlorohydroxyquinone and dichlorohydroxyqui-none.However, the substituted chlorine inPCP inhibits the activity of bothmonoxygenase and dioxygenase, which are required for breaking up thebenzene ring structure. Consequently, PCP resists aerobic degradation(Chen et al., 2007, 2009). On the other hand, under anaerobic condition,PCP can be reductively dechlorinated to tetra (4C), tri (3C), di (2C) andmono (1C) chlorophenol in stages, but these intermediates are as toxic asor evenmore toxic than PCP andoften aremuchmore difficult to degrade.However, these intermediates can be more easily biodegraded underaerobic conditions. It has been shown that the reductive and oxidativesteps could be coupled and integrated into aerobic granules for PCPdegradation (Chen and Zhan, 2003). Chen and Zhan (2003) developed acoupled reactor system which consists of an upflow sludge bed typereactor and oxygen contactor with partial effluent recycle between thetwo reactors. This coupled system has the advantage of easier andindependent control of dissolved oxygen (DO) and hydrodynamic shearforce by adjusting the aeration and liquid recirculation rates. The aerobicand anaerobic coupled granules were successfully cultivated for biore-mediation of the PCP containingwastewater. These aerobic granuleswerefirst cultivated under low DO condition of 0.5–0.6 mg L−1 (Lan et al.,2005; Chen et al., 2009), followed by PCP acclimation at low DOconcentration as well, in which the PCP concentration was graduallyincreased up to 80 mg L−1. The synergistic involvement of microaer-ophilic and anaerobes for reductive dechlorination was proposed forgranular sludge (Chen et al., 2009).

The microbial populations in the microaerophilic and methano-trophic granular sludge were investigated for Eubacteria and Archaeaby denaturing gradient gel electrophoresis (DGGE). DGGE analysis onseed sludge, and granular sludge with or without exposure to PCPrevealed that Eubacteria was abundant with about 20 bands in seedsludge, 16 bands in granular sludge without PCP, and 11 bands ingranular sludge with PCP, indicating that Eubacteria diversitydecreased as the seed sludge developed into micro-aerophilicgranules and subsequently into PCP degrading granules. In contrast,the Archaea community diversity increased from 12 bands in the seedsludge, to 14 and 18 bands in granular sludges with and without PCPrespectively. Eubacteria bands which are related to Sphingomonas,Desulfobulbus sp, Actinobacterium and α-Proteobacteriawere detectedin granular sludge with PCP. Eubacteria band associated withε-Proteobacteria and Archaea bands were found in the seed sludgeand PCP granular sludge. An interesting observation is the presence ofε-Proteobacteria, which is a symbiont, and a Desulfobulbus sp in thecoupled microaerophilic–methanogenic granules (Chen et al., 2009).Chen et al. (2009) proposed a three layered structure with each layerbeing represented by a specific bacterial group. Near the surface of thegranule are aerobic bacteria, followed by the middle layer consistingin syntrophic bacteria such acidogenic bacteria, hydrogen and aceticacid producing bacteria, and sulfate reducing bacteria, with themethanogens at the core of the coupled granule.

6. Degradation of pyridine

Pyridine is used widely in the manufacture of agrochemicals andpharmaceuticals. Pyridine and its byproducts are derived from coalgasification (Stuermer et al., 1982) and they have also been widelyused as catalyst by the pharmaceutical industry (Leenheer et al.,1982). Pyridine is harmful if inhaled, swallowed or absorbed via theskin. Symptoms of acute intoxication by pyridine include dizziness,headache, nausea, anorexia, abdominal pain and pulmonary conges-tion (Gilchrist, 1997; Shimizu, et al., 2007). Current evidence alsosuggests Pyridine is a possible carcinogen. Pyridine in drinking waterhas been linked to reduction in sperm motility and also increase inestrous cycles in rats. Pyridine can be removed from wastewater bybiodegradation, adsorption, electrosorption, ozonation, and ion

exchange. In fact, pyridine is readily degraded by bacteria to ammoniaand carbon monoxide. For example, Adav et al. (2007c) found thataerobic granules cultivatedwith phenol at 500 mg L−1 as co-substratecan biodegrade pyridine at a concentration of 250–2500 mg L−1. Atthe lower pyridine concentration of 250 mg L−1, the specificdegradation rate of pyridine was 73.0 mg g−1 VSS h−1, while at thehigher concentration of 2500 mg L−1 the specific degradation ratehad declined to 31.0 mg g−1 VSS h−1. It was also observed thatphenol degradation was faster than pyridine when these werepresented as individual substrates (Adav et al., 2007c). As co-substrates, the phenol degradation rate was slower in the presenceof pyridine, thus concluding that pyridine had minimum impact ofphenol degradation, while pyridine degradation was inhibited athigher phenol concentration of above 500 mg L−1 (Adav et al.,2007c). Thus, it can be concluded that pyridine had some effect onphenol degradation, but phenol would severely affect pyridinebiodegradation. Population study by polymerase chain reaction(PCR) DGGE showed a shift in dominant strain in the originalphenol-degrading granules when pyridine was introduced (Adav etal., 2007c). For example, as the pyridine concentration increased,Bacillus cereus strain was replaced by uncultured Bacillus andKlebsiella pneumoniae strain IEDC (Adav et al., 2007b). This wassupported by the observation that Bacillus sphaericus, Psuedomonasand Acinetobacter sp can degrade aromatic compounds (Reisfeld et al.,1972; Kaplan and Rosenberg, 1982; Navon-Venezia et al., 1995).Acinetobacter sp was the dominant species, indicating that they wouldbe the main bacterial species capable of pyridine and phenolbiodegradation in aerobic granules (Adav et al., 2007d). Acinetobactercan produce copious amount of exopolymers which might haveprovided protection from the toxic effect of both pyridine and phenol(Kaplan and Rosenberg, 1982; Bach et al., 2003). Confocal microscopystudies on aerobic granules capable of degrading pyridine and phenol,further revealed that active microorganisms were localized at theperiphery of granules, while proteins and polysaccharides whichconsisted of lipids and polysaccharide were found in the inner core ofthe granules (Adav et al., 2007c).

7. Degradation of phthalic acids and esters

Phthalic acid (PA) (IUPACname:benzene-1,2 dicarboxylic acid) is anaromatic dicarbocylic acid. It has two isomers — isophthalic acid andterephthalic acid, and is used in the production of dyes, perfumes,saccharin, and phthalates. It is an eye, skin and respiratory irritant.Phthalic aciddegradinggranules havebeendeveloped in theSBRsystemusing acclimated activated sludge as microbial inoculum within sevendays of operation (Zenget al., 2007). Itwas shown that the PAdegradingaerobic granules after 8 weeks storage at 4 °C still remained intact withno disintegration, and the stored granules could recover bioactivityquickly (Zeng et al., 2007). The ability of aerobic granules to cope withinhibitory compounds can be attributed to genes which are frequentlycarried by the plasmid. These genes are easily shuttled to and fromorganisms.Oneof thedrawbackof granules storage is the loss of plasmidduring storage due to condition change, i.e. the gene regulating thebiodegradation of recalcitrant compound could be lost as the recalci-trant compound is not present in the storage medium and the bacteriawill need to expense energy to conserve the gene and synthesize theproteins and enzymes responsible for the recalcitrant chemicaldegradation when exposed to it. The excellent recovery of bioactivitycan, however, be attributed to the granule's compact and densestructure enablingmicrobial interaction andmobile exchange of geneticmaterial bymicroorganisms residing within the granules via horizontalgene transfer (Zeng et al., 2007).

Phthalate esters (PAEs) are compounds used widely as plasticizersin plastic resins manufacturing, such as polyvinyl resins, and cellulosicand polyurethane polymers used for the manufacturing of buildingmaterials, home furniture, transportation equipment, clothing, and

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packaging material for food and pharmaceutical industries. The PAEsare mobile as they are not covalently bound to plastics resins in orderto maintain plasticity. This leaching of PAEs into the environment is aserious concern as some of the PAEs and their intermediates aresuspected carcinogens and mutagens. Even at its lowest concentra-tion, these PAEs may disrupt the reproductive systems of human andwild life animals. Thus, some of the PAEs, such as dimethyl phthalate(DMP), diethyl phthalate (DEP) and dibuthyl phthalate (DBP) havebeen listed as top-priority pollutants by US Environmental ProtectionAgency (1988).

In the course of investigating possible phthalic acid bioremedia-tion using aerobic granules, a strain of PA degrading bacterium withPAE degrading capability, strain PA-02, was isolated from PAdegrading aerobic granules. Based on 16S ribosomal RNA sequencing,this isolate belongs to the genus Sphingomonas with 100% identitymatch with Sphingomonas strain D84532. This strain is versatile anddoes not need acclimation for degradation of PAEs. The strain candegrade dimethyl phthalate (DMP), dibuthyl phthalate (DBP) anddiethylhexyl phthalate (Zeng et al., 2008). The availability of strainPA-02 which has versatile capability to degrade PA and DMP at thehigh concentration of 1000 mg L−1 might be useful for bioaugmenta-tion purpose in order to deal with high concentration of PA and itsintermediate products in industrial wastewater treatment.

8. Degradation of tert-butyl alcohol

Tert-butyl alcohol (TBA) is directly added to fuels, and oftencombined with methyl tert-butyl ether (MTBE) as an octane indexenhancer to reduce vehicle emissions (Piveteau et al., 2001) and iswidely used as a solvent for the manufacturing of plastics, resinpolymers, perfumes, paint remover, insecticides and pharmaceuticalproducts. TBA is also a known potent toxin and carcinogen (Cirvello etal., 1995; Bradley et al., 2002; Schmidt et al., 2004). TBA isbiodegradable even though the tertiary butyl structure of TBAmolecules consists of 3 methyl groups attached to a carbon — makingit more resistant to biological degradation compared to less complexgasoline fractions such as benzene and toluene (Fortin et al., 2001).Therefore, TBA has been found to accumulate as the rate limitingintermediates in MTBE biodegradation (Hatzinger et al., 2001;Francois et al., 2003; Somsamak et al., 2005). TBA mineralizationcan only occur at nearly complete removal of MTBE (Salanitro et al.,1994; Steffan et al., 1997; Fayolle et al., 2001).

Zhuang et al. (2005) developed TBA-degrading aerobic granulesby step-wise increment in TBA loading with 180 days of start-upoperation. Several Gram-negative bacteria which can produce extra-cellular polysaccharides and cell surface carbohydrate were isolated,and the latter might have provided protection against environmen-tal stresses such as toxicity (Tay et al., 2005a; Zhuang et al., 2005).The study of Zhuang and co-workers also emphasized thatcombination of hydraulic and microbial selection pressures neededregulation for granule formation. Zhuang et al. (2005) adopted amoderate selection strategy with a cycle time of 24 hour coupledwith low TBA concentration of 100 mg L−1 to promote the growth ofbiomass and thus avoided cell washout. This strategy allowed themicrobial biomass to synthesize enzymes or to shuttle plasmidwhich activate the gene involved in the synthesis of enzymes orprotein for TBA degradation by exposing the microbial biomass tolow concentration of TBA. The compact structure of TBA degradingaerobic granules would also protect the microbial population fromdirect TBA toxicity.

Microbial analysis by DGGE showed that the TBA degradingaerobic granules had a stable microbial population with very lowphylogenetic diversity. The low microbial population diversity iscaused by the strong selection pressure due to the toxicity of TBA andthus only a microbial population that can degrade TBA will survive(Tay et al., 2005a). After sequencing the DGGE bands, it was found

that these bands included Proteobacteria and Cytophaga Flavobacter-ium (CFB) groups.Methylobacterium organophilum, Sphingomonas andRhizobium/Agrobacterium within the α-Proteobacteria and Rubrivivaxgelatinosus with the β-Proteobacteria were detected in the TBAdegrading aerobic granule bacterial populations. This group was alsoreported to be dominant in a MTBE degrading biofilm reactor (Zein etal., 2004). However, it should be noted that the absence of members ofthe high G+C group of bacteria which have been previouslydocumented to be TBA degraders was observed in the TBA-degradingaerobic granules (Mo et al., 1997; Deeb et al., 2000; Steffan et al.,1997; Francois et al., 2003; Schmidt et al., 2004). It would beinteresting to develop aerobic granules which degrade bothMTBE andTBA and to further explore the synergistic effect of MTBE degradationover TBA degradation as the microbial population involved are muchmore related to each other.

9. Degradation of methyl tert-butyl ether

Methyl tertiary-butyl ether (MTBE) is added to gasoline as anoctane enhancer and to reduce emissions of carbon monoxide andsmog associated air pollutants. The wide use of MTBE in gasolinerepresents one of the largest chemical use in industrialized countries.MTBE is introduced in reformulated and reoxygenated gasoline inmany countries worldwide (Depasquier et al., 2002). A serious issuearising from the wide use of MTBE is its leakage from undergroundstorage tanks which can contaminate groundwater. It is because of itsmassive production coupled with its mobility, persistence, toxicityand high solubility in water which has made it an important pollutantin drinking and ground water (Squillace et al., 1997). The USEnvironmental Protection Agency has classified MTBE as a possiblehuman carcinogen and has issued a drinking water advisory onpermissible concentration range of 20–40 ppb (Fortin andDeschusses, 1999). In recent years, the US EPA has taken steps toreduce and eliminate this oxidant in fuels. However, the riskassociated with groundwater contamination from leakage fromservice stations and fuel storage facilities is real and has attractedattention. In other parts of the world, a recent study estimated thatMTBE production rate will increase in many developing economies inAsia and China in the next decade (Dong et al., 2007). In order toreduce MTBE contamination, China EPA has move toward stricterregulation of MTBE discharge. Thus, there will be continued interest todevelop novel and cost effective processes for the remediation ofMTBE by biological processes.

Biological treatment of MTBE contaminated water and wastewatermay be the most economical, energy efficient and environmentallyfriendly method available (Zein et al., 2004). This is based on recentstudies which have shown the ability of several bacterial cultures andmicrobial consortia to aerobically biodegradeMTBE as sole carbon andenergy source or by cometabolism in the presence of other organicsubstrates (Fortin et al., 2001; Piveteau et al., 2001; Francois et al.,2002; Lin et al., 2007). Earlier studies employing activated sludge tobioremediate MTBE had been unsuccessful due to inhibition ofmicrobial growth and biodegradation activity (Deeb et al., 2000).Based on numerous studies on MTBE bioremediation, it wasconcluded that a high biomass concentration is essential for MTBEmineralization. To overcome toxicity of the recalcitrant xenobiotic,the strategy of initially introducing a benign substrate has beensuccessfully used in the biofilm reactor to facilitate biofilm adhesionand formation (Bhatti et al., 2002). Zhang et al. (2008a) developedaerobic granules using MTBE as target contaminant with or withoutethanol as a co-substrate. An initial MTBE concentration of 50 mg L−1

was adopted with step wise increment up to 400 mg L−1, whileethanol was slowly reduced from 500 to 100 mg L−1 (Zhang et al.,2008a). As mentioned earlier, TBA is the main intermediate of MTBEbiodegradation. However, it was found that the TBA concentrationwas low and less than 25 μg L−1 in the aerobic granular sludge

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biomass, indicating a complete mineralization of MTBE (Zhang et al.,2008a,b).

The study of MTBE-degrading aerobic granular sludge by DGGEshowed that the dominant bacterial population was closely related toα-Protoebacteria belonging to the Rhizobium/Agrobacterium group. Thisis a species belonging to the genus Methylobacterium and Hyphomicro-bium vulgare (Zhang et al., 2008a). The fast degradation of tert-butylalcohol (TBA) in aerobic granules has been attributed to H. vulgarewhich can degrade alcohol (Nikitin et al., 1998). H. vulgare have beenreported to degrade MTBE and ethanol in a porous pot reactor (Prudenet al., 2001). In addition to strains related to Flavobacteria–Cytophagagroup (e.g. Flavobacterium) — Biziona and Coccinimonas and aSphingomonas sp. were also detected (Zhang et al., 2008a) which is inconcurrence with the Flavobacteria–Cytophaga group being MTBEdegraders (Sedran et al., 2002). It is postulated that the MTBEdegradationwould be carried out by various bacterial species at variousstages within the MTBE biodegradation pathway or all strains involvedinMTBE degradation are in synergistic cooperation (Zhang et al., 2008a,b). As seen again with other recalcitrant and toxic chemicals, thecompact and dense structure of the aerobic granule had minimized theinhibitory effects of MTBE (Zhang et al., 2008b).

10. Degradation of chloroanilines

Chloroanilines (CIA) are widely used as intermediates for thesynthesis of organic chemicals and industrial polymers such aspolyurethanes, rubber additives, dyes, pharmaceuticals, pesticidesand herbicides. CIA widespread applications coupled with its toxicityand inherent recalcitrant nature meant they are also considered animportant environmental pollutants which have attracted strictlegislative control by environmental protection agencies worldwide.Biological treatment to remove CIA are feasible and have attractedinvestigations employing various reactor configurations. However,reactor systems such as fluidized bed bioreactors and membranebioreactor have been sensitive to fluctuations in influent substrateconcentrations and organic loading rates (Latorre et al., 1984; Boon etal., 2002). These have failed to effectively degrade xenobioticcompounds such as CIA due to its toxicity and consequent killing ofthe substrate utilizing bacteria present in the consortia (Boon et al.,2002; 2003).

CIA biodegradation by microorganisms can be completed by amodified ortho- or meta cleavage pathway after oxidative deaminationof CIA to the chlorocatechol, which is followed by dechlorination andfinally mineralization via the tricarboxylic acid cycle (Latorre et al., 1984;Schlomann, 1994; Moiseeva et al., 2002). Recently, Zhu et al. (2007)successfully cultivated CIA degrading aerobic granules by adopting astepwise strategy to increase CIA loading rates in a sequential airliftbioreactor (SABR). It was postulated that the introduced CIA exertedselection pressure on the microbial consortia with capability tobiodegrade CIA at CIA loadings up to 800 mg L−3 day−1. However, theSABRwas unstablewhen the CIA loading rate reached 1.0 kg m−3 day−1,and the granules were then dominated by filamentous microorganismsand started to disintegrate (Zhu et al., 2007). Again, it appeared thecompact and dense structure of aerobic granules protected the microbialconsortia against the toxicity of chloroanilines. The DGGE results revealedthat β and γ-Proteobacteria and Flavobacteria were dominant classes,while the genus Pseudomonas and Flavobacterium were detected in theexcised gel. In fact, both Pseudomonas sp and Flavobacterium had beenpreviously reported to degrade CIA and pentachlorophenol in chemostatculture (Saber and Crawford, 1985; Latorre et al., 1984).

11. Degradation of metal-chelating agents

Synthetic chelating agents are widely used in many industrialapplications for their metal binding and masking of metal ions.Aminopolycarboxylic acids (APCAs) are one such important group of

chelating agents, as it can formstable andwater-soluble complexeswithmany metal ions and radionuclides (Venugopalan et al., 2005;Nancharaiah et al., 2006a). Nitrilotriacetate (NTA) and ethylene diaminetetra acetate (EDTA) are two of the most widely used APCAs, and havebeen extensivelyused for chemical decontaminationof nuclear reactors,chemical cleaning of steam generators and nuclear waste processing(Bolton et al., 1996;Witschel and Egli, 2001;White and Knowles, 2003;Venugopalan et al., 2005; Nancharaiah et al., 2006a). These chelatingagents are also used in detergent, food processing, pharmaceutical,cosmetic, metal-finishing, photographic, textile, and paper manufac-turing processes (Venugopalan et al., 2005; Nancharaiah et al., 2006a).The disposal of heavy metals with radionuclides together with thesesynthetic chelating agents poses a serious environmental challengebecause chelating agents may promote undesirable movement of toxicmetals and radionuclides away from their primary waste disposal siteinto the environment (Bolton et al., 1996; Thomas et al., 1998).Furthermore, the metal complexes are much more difficult tobiodegrade and very few bacterial strains can biodegrade these. Thus,the bioremediation of such complex wastes is much more challengingthan dealing with the individual pollutants, and the ability to degradeone component may be completely inhibited by another toxiccomponent, and this is particularly true for a mixture of organicpollutants such as chelating agents and metal ions and it is much moresevere in the presence of radionuclides (Venugopalan et al., 2005;Nancharaiah et al., 2006a). Some studies on microbial degradation ofnitrilotriacetic acid and other chelating agents have revealed thatmetalcomplexation has a significant effect on biodegradation. Thus, there isurgency to develop effective biological treatment processes for degrad-ing synthetic chelating agents.

Venugopalan et al. (2005) had successfully cultivated aerobicgranules in a laboratory scale SBR seeded with activated sludge. Thereactor was fed with synthetic wastewater containing NTA andacetate as co-substrate. It was observed that the growth rate of theseed sludge was much lower in the presence of NTA (Nancharaiah etal., 2008). Similarly the growth rate of acetate-fed culture was alsoreduced in the presence of NTA (Nancharaiah et al., 2008). Aerobicbiodegradation of NTA by granules is possible in the presence ofacetate as co-substrate because, NTA exerted selection pressure on themicrobial consortia to favour slow growing bacterial strains over fastgrowers (Nancharaiah et al., 2008), and select microorganisms thatsynthesize inducible enzymes for NTA biodegradation. After twoweeks following seeding, aerobic granules appeared.

Confocal microscopy analysis revealed that cell clusters packed asrod or cocci bacterial cells were in the aerobic granules. Thesegranules can biodegrade NTA concentration up to 600 mg L−1, andthe granules could be used repeatedly for degradation of NTA, and forup to four consecutive cycles of NTA degradation, without losing thedegradative capability (Venugopalan et al., 2005; Nancharaiah et al.,2008). The biodegradation rate was faster with subsequent cycle ofoperation, thus indicating good adaptation of bacterial cells tobiodegrade NTA (Nancharaiah et al., 2008). The specific degradationrate of free NTA and ferric NTA are 0.7 mM and 0.37 mM respectively(Nancharaiah et al., 2006a). It was found that the degradability couldbe rejuvenated by acetate treatment, with the granules retaining itsstructural integrity and settling ability even after repeated cycles ofdegradation (Venugopalan et al., 2005; Nancharaiah et al., 2006a). Inaddition, the aerobic granules could degrade free NTA and Fe boundNTA complex. However, it was found that biodegradation of Fe boundNTA is unstable when compared to free NTA (Venugopalan et al.,2005), and this can be attributed to the long lag time required toinduce specific metal bound NTA degrading enzymes and theadditional time involved in the transport of the complex across thebacterial cell wall and membrane (Venugopalan et al., 2005;Nancharaiah et al., 2006a).

Complete degradation of the chelant and its complex can also beaccomplished in distilled water suggesting the treatment process may

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not require addition of essential nutrients necessary for supportingmicrobial growth (Venugopalan et al., 2005; Nancharaiah et al.,2006a). Venugopalan and co-workers observed that during thebiodegradation of ferric-NTA, a gradual increase in a dark orange-red precipitate (of iron released from the degraded Fe-NTA complex)in the glass cylinders. They also postulated that complete degradationof chelating agents and their metal complexes may be mediated byone or many different types of bacterial strains present in the granularsludge biomass, as granular sludge can facilitate a large number ofmicro-organisms to be retained in the bioreactor (Venugopalan et al.,2005). This also explained the rapid rate of pollutant transformationwhich meant large volumes of waste can be treated using compactbioreactors. Moreover, the large size and relatively high specificgravity (about 1.5) of the granules caused them to settle rapidly,leading to easy separation of treated effluent from the biomass(Venugopalan et al., 2005). The concentration of NTA and ferric-NTAdid affect microbial aggregation (Nancharaiah et al., 2006a).

Efficient degradation rates of NTA and ferric-NTA by aerobicgranules observed in the study pointed to the possibility of developingan efficient technology for rapid biodegradation of dilute industrialwastes using compact bioreactors. However, there is a need forfurther exploration of the capability to handle other metal-NTAcomplexes (e.g., Co-NTA) and radioactive isotopes. The ability ofmicrobial granules to remove heavy metals by biosorption has beendemonstrated by other researchers. Nevertheless, it is desirable toelucidate strain diversity in the granules using molecular microbiol-ogy methods such as 16S rRNA sequence analysis. This will help openup the possibility of applying aerobic granules for bioaugmentation,whereby special natural or engineered catabolic strains could beintegrated into pre-cultured granules to enhance their substrateconversion capabilities.

12. Bioaccumulation of pigments/dyes

Most industrial application such textile manufacturing, andpaper and pulp processing involves the use of dyes and pigmentswhich contribute to environmental pollution (Gupta et al., 2003;Wang et al., 2006). Furthermore, many dyes and pigments areknown carcinogens and mutagens which can be trapped in the foodchain and eventually pose human health risks (McKay et al., 1985;Gregory et al., 1991; Rahman et al., 2005). Technologies involvingchemical coagulation/precipitation and adsorption have been usedfor their removal (Malik and Saha, 2003; Malik and Sanyal, 2004;Rahman et al., 2005; Wang et al., 2006). These techniques do not,however, mineralize the dyes and pigments and are typicallyexpensive to use compared to biological treatment. A wide varietyof biological treatments such fungal cell mass, activated sludge,biosolids, and microalgae have been explored for pigments/dyeremoval (Kumari and Abraham, 2007; Maurya et al., 2006; Patel andSuresh, 2007; Otero et al., 2003; Wang et al., 2006; Sirianuntapiboonand Srisornsak, 2007; Vasanth Kumar et al., 2006; Daneshvar et al.,2007). However, a drawback of the biosorbents widely used arethese are in the form of suspended growth biomass, which sufferedfrom poor settling, and hence washout of active biomass. Aerobicgranulation can be an attractive alternative.

Recently, Sun and co-workers had studied the feasibility oftreating malachite green wastewater with aerobic granules (Sun etal., 2008a). The study by Sun et al., revealed that Malachite greenremoval was rapid in the initial stage, but decreased as time isextended. Sun et al., had argued that in the initial phase, the aerobicgranules surface was available and this explained the fast rate ofbiosorption. After the initial phase, fewer sorption sites remainedavailable on the aerobic granule surface. They found the biosorptionequilibrium comparable to the adsorption of dyes by activated sludge,saw dust, and activated powdered carbon (Kargi and Ozmichi, 2004;Garg et al., 2003; Aksu, 2001; Ozacar and Sengil, 2005; Kadirvelu et al.,

2005). TheMalachite Green biosorption is a single curve, and this maybe because of mono layer coverage of Malachite green on the aerobicgranule surface (Namasivayam et al., 1998). The study also found thatthe initial concentration was the driving force which overcame masstransfer resistance of the dye with respect to the aqueous and thesolid phase. A higher initial concentration of malachite greenenhanced the biosorption process in aerobic granules (Sun et al.,2008a). In the same study, it was found that Malachite Greenbiosorption increased with increasing pH. Sun et al., postulated thatlow biosorption of Malachite green by aerobic granules in acidiccondition could be attributed to the development of cations on thebiosorbent, which prevented biosorption of malachite green onto thesurface of aerobic granules. The presence of excessive H ions wouldalso compete with the positively charged malachite green fornegatively charged biosorption sites (Vasanth Kumar et al., 2006;Sun et al., 2008a). The formation of dimers when pH increased alsoenhanced malachite green biosorption onto aerobic granules surfaceby electrostatic interactions betweenmalachite green and the granulesurface (Sun et al., 2008a). The maximum biosorption capacity ofmalachite green onto aerobic granules was 56.8 mg g−1 (Sun et al.,2008a).

13. Bioaccumulation of heavy metals

Heavymetals are present in high concentration in a wide variety ofindustrial effluents such as textile, leather, tannery, electroplating,galvanizing, pigment and dyes, metallurgical, and paint industries(Zhang et al., 2005; Ahluwalia and Goyal, 2007). Stringent limits onheavy metal discharge concentrations have been imposed byenvironmental regulatory agencies worldwide due to their toxicityand non-biodegradability. The heavy metals can be trapped andbiomagnified along the food chain via consumption of affected plantsand animals. Conventionally, these metals are removed with physicaland chemical methods such as chemical precipitation, chemicaloxidation and reduction, ion exchange, filtration, electrochemicaltreatment, reverse osmosis, membrane technologies, and evaporation(Zhang et al., 2005; Ahluwalia and Goyal, 2007). However, thesemethods may result in the generation of toxic chemical sludges.Recent research efforts have been made on the development ofefficient and low cost metal organic adsorbents such as alginate(Gokhale et al., 2009; Maiti et al., 2009), carragenan (Baysal et al.,2009), algae biomass (Ajjabi, and Chouba, 2009; Ebrahimi et al.,2009), bacterial spp (Freitas et al., 2008; Ziagova et al., 2008; Mertogluet al., 2008; Qiu et al., 2009), and activated sludge biomass (Renningeret al., 2001; Boswell et al., 2001). A drawback of these biosorbents isthat they are in the form of suspended flocs and these may have poorliquid–solids separation properties. Aerobic granules have superiorsettling capability. Liu et al., developed aerobic granules usingsynthetic wastewater for bioaccumulation of zinc (Liu et al., 2002).The biosorption of zinc as Zn(II) was governed by the concentration ofZn(II). This phenomenon had also been seen with other biosorbentsuch as marine algae (Liu et al., 2002; Taniguchi et al., 2000). Themaximum biosorption capacity of Zn(II) by aerobic granules was270 mg g−1 (Liu et al., 2002). In another study, Liu investigated thebiosorption of Ni by aerobic granules, and in this study, the Ni2+

biosorption were investigatedwith release of Ca2+, Mg2+ and K+. Thestudy found that the released Ca2+ (0.52 meq g−1) was much higherthan Mg2+ (0.045 meq g−1) and K+ (0.061 meq g−1) for a Ni2+

uptake of 0.76 meq g−1 dry weight. They concluded that Ni biosorp-tion onto aerobic granules involved an ion exchange mechanism andnot via chemical precipitation of Ni2+. The study also found that Ni2+

was uniformly distributed from the surface to the inner core of aerobicgranules (Liu and Xu, 2007). They also investigated the role ofextracellular polymeric substances (EPS), and found that only a smallfraction, 14.2%, of the total Ni2+ biosorption was due to EPS (Liu andXu, 2007). Biosorption of Ni2+ was endothermic and so is

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temperature dependent (Liu and Xu, 2007). In another studyconducted by Zhang et al., cerium (Ce) adsorbed by aerobic granuleswas investigated. Ce ions adsorbed onto the granules increasedgradually with respect to time, and this was observed at various initialCe ion concentrations (Zhang et al., 2005). Ce biosorption wasattributed to functional groups or biopolymers, and can be describedas a physicochemical process which obeyed first order reversiblekinetics as seen with cadmium (Cd) biosorption (Liu et al., 2003;Zhang et al., 2005). An interesting observation was the rapid decreaseof residual Ce ions in solution i.e. less than 10 mg L−1, which would beinsufficient to drive the uptake of Ce into the inner portions of theaerobic granules (Zhang et al., 2005). Yao et al. (2009) found that Cr3+

biosorption capacity increased with increasing pH at pH range of 2–6.At initial Cr3+ concentration of 50 mg L−1 and pH 5, the maximumbiosorption capacity of aerobic granules is 37.8 mg g−1. Yao et al., hadargued the predominance of negatively charged functional groupssuch as carboxyl, amine, phosphate groups had drawn Cr3+ ions. AtpH greater than 5, precipitation of Cr(OH)3 occurred and competedwith functional groups on biosorption (Yao et al., 2009). As in allprevious heavy metal biosorption studies before Cr3+, biosorptioncapacity increased with increase in initial Cr3+ concentration.Scanning electron microscopy, X-ray energy dispersion and Fouriertransform infra red studies showed that biosorption of Cr3+ involvedchemical precipitation, metal complex formation, and ion exchangephenomenon (Yao et al., 2009). Later, Liu et al., extended the heavymetal study to multi-elements, and introduced cadmium and copperin addition to Zinc. They found that maximum biosorption capacitiesof Cd2+, Zn2+, and Cu2+ by aerobic granules were 625, 204, and52.9 mg g−1 respectively, and these values were much higher whencompared to other biosorbents. In addition, aerobic granules havesuperior settling velocity 5–10× of microbial flocs which madeaerobic granules an excellent heavy metal biosorbent (Liu et al.,2003, 2004). Liu et al., also investigated the biosorption of multi-elements consisting of Ni2+, Cd2+ and Cu2+. They successfullydemonstrated aerobic granules adsorbed Ni2+, Cd2+ and Cu2+ usingX-ray diffraction analysis, and contribution of Cd2+ and Cu2+ removalby chemical precipitation in the form of CdCO3 and Cu2(OH)3Clprecipitates had been minor (Xu and Liu, 2008). No Ni chemicalprecipitates were detected in Ni2+ removal by aerobic granules (Xuand Liu, 2008). An ion exchange mechanism was proposed and thisinvolved the release of light metals such as Ca2+, Mg2+ and K+ andthe uptake of heavy metals Ni2+, Cd2+ and Cu2+ (Xu and Liu, 2008).The ion exchange mechanisms contributed to 71–82% of the solubleheavy metal removal, while extracellular polymeric substancesaccounted for 12–19% of the heavy metal removal (Xu and Liu,2008). In a two element study conducted by Sun et al., involving Zn2+

and Co2+, maximum adsorption of 55.25 mg g−1 for Co2+ was notedat pH 7 and 62.50 mg g−1 for Zn2+ at pH 5. In binarymetal adsorption,competitive adsorption occurred (Sun et al., 2008b). Fourier trans-form infra-red spectroscopy and X-ray photonelectron spectroscopyspectral analysis showed that functional groups such as alcohol andcarboxylate functional groups were involved. Sun et al., also foundthat small amounts of loosely bound extra polymeric substances(LB-EPS) and large amount of tightly bound extra polymericsubstances (TB-EPS) were present (Sun et al., 2009). Their binarymetal study noted competition for active sites and this reduced theadsorption capacity for both EPSs (Sun et al., 2009). The study notedthat functional groups such as alcohol, carboxyl, and amino groups onthe EPS were involved in the biosorption of Zn(II) and Co(II) (Sun etal., 2009). The LB-EPS did not only act as chelating sorbents but alsofacilitated flocculationwhich further enhanced biosorption (Sun et al.,2009). Nancharaiah et al., investigated biosorption of the radionuclideuranium, and found that the maximum uptake capacity was218 mg g−1 (Nancharaiah et al., 2006b). They also found that cationssuch as Na+1, K+1, Mg2+ and Ca2+ were released during U(VI)biosorption which suggested an ion exchange mechanism was

involved, the amount of metals released was in the order K+,Ca2+bMg2+bNa+. Uranium biosorption was pH dependent, and theeffective pH range was 3 to 5 (Nancharaiah et al., 2006b). Influence ofpH was attributed to functional groups on bacterial cell surface andEPS (Sar et al., 2004; Nancharaiah et al., 2006b). Aerobic granularsludge can be used to remove uranium at very low concentrations andis potentially cost effective (Nancharaiah et al., 2006b).

14. Conclusions

Table 1 shows the aerobic granules so far developed which aresuitable for treatment of wastewaters containing potentiallyinhibitory organic compounds, toxic heavy metals and radioactivematerial. In process start-up and initiation of granulation, gradualstepwise introduction of target chemicals and with these replacinga non-inhibitory carbon source has been a successful procedure. Insome cases, in order to overcome slow formation of granules,especially for those which are exposed to highly inhibitorycompounds, and to maintain these once formed, non-inhibitorysubstrates, such as glucose and acetate can be used as co-substrate.However, this strategy may not work for some recalcitrantchemicals due to competition from microbes which can rapidlymetabolize the easily biodegradable substrates in preference overthe recalcitrant/inhibitory compounds. The slow growing special-ist strain shall then be overwhelmed in terms of numbers by thefast grower in the consortia that utilize the easily biodegradablesubstrates, but not the recalcitrant/inhibitory compounds. Shouldthis be the case, it may be necessary to develop the aerobicgranules by directly exposing seed sludge to the target inhibitorycompound without a co-substrate.

The compact structure of the aerobic granule protects themicrobes residing inside from the inhibitory effects of the targetcompounds. Aerobic granules developed following exposure toinhibitory organics will generally have the degrading capability asthe microbial population with degrading capability will have beenpositively selected. More than one trait may be involved in thedegradation. Depending on the micro-niche within the granules,factors such as substrate utilization, growth rate, and availability ofsubstrates such as oxygen will also select the microbial population.Many industrial wastewaters exist at very extreme pHs, the ability ofmicroorganisms in the aerobic granules to tolerate extreme pH whilepreserving the degradative capability is important. Therefore forsuccessful treatment of such industrial wastewaters, it may benecessary to develop processes that can tolerate such pH extremes.Utility of bioaugmentation with a stable strain to an existing culturehas also been demonstrated for treating inhibitory compounds. Thereare also instances when it would be useful to couple aerobic andanaerobic conditions and thesemay be generated within the granules.The presence of bioreductive processes within the granules can beused to potentially dehalogenate more effectively. In conclusion,aerobic granulation is possible even when treating potentiallyinhibitory substances and may well provide the basis for more stablebiotreatment processes. This and the higher cell densities offer apotentially more economic treatment approach because of the lowerplant footprint. The reported work has, however, been essentiallydrawn from laboratory studies. To progress application of aerobicgranulation in treatment of potentially inhibitory industrial waste-waters will now require pilot studies and scale up. One importantchallenge is tolerance to high organic loading which may impede bio-degradative capability and also aerobic granule integrity. It isimportant to establish physico-chemical interaction with biologicaldegradative capability. Laboratory scale study of operational para-meters, reactor configuration, and analytical tools to establish aerobicgranule performance must be integrated and optimized. Due toaerobic granules much better performance compared to flocs in termsof settling and tolerance to high organic loading, research focus has

Table 1Aerobic granulation process for the treatment of recalcitrant chemicals for industrial wastewater treatment.

Recalcitrant chemicals Granulationprocess

Inoculums used Co-metabolismwith co-substrate

Removalefficiency

Organisms involved in degradation Reference

1. Chloroanilines (CIA) Sequencing airliftbioreactor (SABR).

Activated sludge 1. Glucose/sodiumacetate as co substrate

1.99% Acidovorax, Nitrosomonas communis,Psuedomonas, Flavobacterium columnare

1. Zhu et al.,2007.

2. Tertiary-butylalcohol (TBA)

Sequencing batchreactor (SBR)

Activated sludge frommunicipal wastewatertreatment plant

1. No co-substrate

1. 97.8–91.3%

Sphingomonas, Denitromonas aromaticus,Roseateles depolymerans, Bradyrhizobiumsp, Methylobacterium organophilum,methylobacterium extorquens

1. Zhuang et al.,2005;

3. Phthalic acid (PA) 1. Column typeSBR

Acclimated activatedsludge

1. No co-substrate

1. 100% Sphingomonas sp 1. Zeng et al.,2007, 2008.

4. Para-nitrophenol(PNP)

1. Column typeSBR

Activated sludge frommunicipal wastewatertreatment plant

1. Glucose as cosubstrate

1. 99.9% NA 1. Yi et al., 2006

5. Methyl tert-butylether (MTBE)

1. Column typeSBR

Activated sludge formrefinery wastewatertreatment plant

1. Ethanol as cosubstrate

1. 99.5% 1. Sphingomonas sp 1. Zhang et al.,2008a

2. Column typeSBR

2. No co-substrate

2. 95.5–99.8%

Flavobacterium sp, Bizionia sp, Moraxellaosloensis, Methylobacterium sp,Hyphomicrobium vulgare,Pseudoalteromonas sp, Arthrobacter sp,Coccinimonas marina

2. Zhang et al.,2008b

6. 1.Chlorinated phenols:4-chlorophenol (4-CP)and 2,4,6-trichlorophenol(2,4, 6-TCP);

1. Granulatedsequencing batchreactor (GSBR)

1. Activated sludgefrom municipalwastewater treatmentplant.

1. Sodium acetate 1. 100%(4-CP),100% (TCP)

1. NA2. NA 1. Carucci et al.,2008, 2009,2010.

2. 2,4-Dichlorophenol(2,4-DCP).

2. SBR 2. Activated sludgeform secondaryclarifier of a municipalwastewater treatmentplant

2. Glucose 2. 94% 2,4-DCP

2. Wang et al.,2007

7. Phenol and p-cresol Sequencing batchreactor

Aerobic digestedsludge from secondaryclarifier of paper mill

No co-substrate Phenol97%,

NA Usmani et al.,2008

8. Penta-chlorophenol 1. Coupledbioreactor ofupflow sludge bedreactor andcolumn reactor.

Anaerobic granularsludge from municipalupflow anaerobicsludge blanket reactor

1.Glucose 1.85% 1.Sphingomonas sp, Desulfolbus sp,Actinobacterium, ε-Proteobacteria,α-Proteobacteria, Methanogenic archae

1. Chen et al.,2007, 2009

Operated in SBRmode

2. Lan et al.,2005.

9. Pyridine SBR Activated sludge withphenol as sole carbonsource

phenol 99.00% Bacillus weihenstephanensis, Bacillussphaericus, enterobacter cancerogenus,Bacillus cereus, Acinetobacter sp, Klebsiellapneumonia, strain IEDC 78, Acinetobactercalcoaceticus strain CBMAI 464,Pseudomonas sp Hugh2319

Adav et al.,2007c.

10. Pigments and dyes SBR Activated sludge frommunicipal wastewatertreatment plant.

Glucose+sodiumacetate

97% NA Sun et al.,2008a

11. Metal chelating agents(nitrilotriacetic) acid(NTA)

SBR 1. Activated sludgefrom municipalwastewater treatmentplant.

Sodium acetate 1.99% NA 1. Venugopalanet al., 20052. Nancharaiahet al., 2006a,b,2008.3. Nancharaiahet al., 2008

11. Heavy metals Activated sludge frommunicipal wastewatertreatment plant.

1. Cd2+, Cu2+, Zn2+ 1. SBR 1. no co-substrate 1. 173, 60,165 mgg−1

1. NA 1. Liu et al.,2003, 2004.

2. Ce2+ 2. SBR 2. Glucose andacetate

2. 357 mgg−1

2. NA 2. Zhang et al.,2005.

3. Ni2+ 3. SBR 3. no co-substrate 3. NA 3. NA 3. Liu and Xu,2007

4. Cd2+, Cu2+, Ni2+ 4. SBR 4. no co-substrate 4. NA 4. NA 4. Xu and Liu,2008; Sun et al.,2008b,c

5. Zn2+, Co2+ 5. SBR 5. Glucose andsodium acetate

5. 63, 55 mgg−1

5. NA

6. Cr3+ 6. SBR 6. no co-substrate 6. 38 mgg−1

6. NA 6. Yao et al.,2009

12. Radionuclei SBR Activated sludge frommunicipal wastewatertreatment plant.

No co substrate 218 mgg−1

NA Nancharaiahet al., 2006a,b

120 A.M. Maszenan et al. / Biotechnology Advances 29 (2011) 111–123

Table 1 (continued)

Recalcitrant chemicals Granulationprocess

Inoculums used Co-metabolismwith co-substrate

Removalefficiency

Organisms involved in degradation Reference

13. Phenol 1. SBR 1. Activated sludgefrom municipalwastewater treatmentplant.

1. No co-susbtrate 1. 99–100% 1. 6 β-Proteobacteria,3 Actinobacteria andone γ-Proteobacteria.

1. Jiang et al.,2002, 2004a,b

2. SBR 2. Activated sludgefrom municipalwastewater treatmentplant.

2. No co-substrate 2. 99–100% 2. Sphaerotilus natans 2. Tay et al.,2005b,c.

3. SBR 3. Activated sludgefrom municipalwastewater

3. No co substrate 3. 99–100% 3. Candida tropicalis,Acinetobacter

3. Adav et al.,2007a,b,c,d.

121A.M. Maszenan et al. / Biotechnology Advances 29 (2011) 111–123

nowmoved on to the aerobic granular membrane bioreactor since theaerobic granule is considered will minimize biofouling of themembranes.

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