1 Department of Civil Engineering, National University of Singapore, Blk E1A, #07-03, Engineering Drive 2, Singapore 117576
2 Laboratoire de Microbiologie, Institut de Recherche pour le Développement (IRD-ex ORSTOM), Universités de Provence et de la Méditerranée, ESIL case 925, 163 Avenue de Luminy, 13288 Marseille cedex 09, France
3 Departamento de Biotecnología, Universidad Autónoma Metropolitana-Iztapalapa, Av. San Rafael Atlixco 186, Col. Vicentina, 09340 Iztapalapa, DF, Mexico
Correspondence
Wen-Tso Liu
cveliuwt{at}nus.edu.sg
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ABSTRACT |
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The GenBank/EMBL/DDBJ accession numbers for the 16S rRNA gene sequences reported in this paper are AY297961AY297969, AY297971AY297978, AY297980AY297982, AY297985, AY297987AY297992 and AY661403AY661422.
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INTRODUCTION |
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Past studies have suggested that mesophilic TA degradation follows a two-step process through the syntrophic association between fermentative bacterial groups which convert TA to acetate and hydrogen and methanogens which convert acetate and hydrogen to final gaseous products (CH4 and CO2) (Fajardo et al., 1997; Kleerebezem et al., 1999a
, b
; Wu et al., 2001
; Qiu et al., 2004
). This is because the oxidation of TA to acetate and hydrogen is an endergonic reaction unless coupled to methanogenesis reactions that further convert those intermediates to the final gaseous products (Kleerebezem et al., 1999a
). The observation that degradation of TA could be inhibited in the presence of acetate or benzoate further supports a hypothesis that the degradation pathway is through decarboxylation via benzoate to acetate and hydrogen (Fajardo et al., 1997
; Kleerebezem et al., 1999b
). Using culture-dependent methods and molecular techniques, the major fermentative bacterial populations involved in mesophilic methanogenic TA degradation have been identified as being from the
-Proteobacteria (Wu et al., 2001
) and the subcluster Ih of the group Desulfotomaculum lineage I (Qiu et al., 2004
). In addition, the syntrophic methanogenic counterparts have been identified as Methanosaeta concilii and members of Methanospirillum and Methanobacteriaceae. These results further support the concept that TA degradation is a two-step process under methanogenic conditions.
Unlike mesophilic methanogenic processes, little is known about the feasibility and microbial consortia for the anaerobic thermophilic treatment of PTA wastewater. An initial attempt by Kleerebezem et al. (1999c) failed to establish a thermophilic methanogenic consortium in a UASB (upflow anaerobic sludge bed) reactor fed with TA as the sole carbon source. This failure was probably attributable to a low number of thermophilic TA-degrading organisms in the original seed sludge and inadequate operational conditions to enrich the microbial consortia. S. Thierry, I. Ramirez, C. Allouche, H. Ferrer & H. Macarie (unpublished results) successfully demonstrated that all the chemicals present in PTA wastewaters (i.e. terephthalic, phthalic, benzoic, trimellitic and acetic acids), with the exception of p-toluic acid, were readily degradable in a laboratory-scale thermophilic hybrid anaerobic reactor (i.e. UASB with pack materials) over a period of 800 days. The highest loading rate achieved (16 kg COD m3 day1) was at least comparable to those obtained from full-scale mesophilic systems such as expanded granular sludge bed and internal circulation reactors operated at PTA plants. A good understanding of the thermophilic methanogenic community involved in the TA degradation is necessary in order to facilitate future efforts for full-scale process start-up (good seed sludge selection), to monitor process treatment efficiency and to establish optimal operating conditions. At the moment, the microbial community responsible, which is likely to be different from the mesophilic species (Wu et al., 2001
; Qiu et al., 2004
), remains unknown. Thus, this study focused on the characterization of microbial community structure and dynamics in a thermophilic methanogenic consortium responsible for TA degradation using various molecular techniques. A laboratory-scale hybrid anaerobic reactor degrading TA-containing wastewater was successfully operated for 272 days and characterized.
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METHODS |
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Anaerobic batch experiment and chemical analyses.
The sludge sample taken from the hybrid reactor at day 200 was anaerobically transferred into 120 ml serum bottles containing 50 ml culture medium [final volatile suspended solid (VSS), 200 mg l1]. The medium (pH 7·27·4) contained 5 mM TA, inorganic nutrients (mg l1) (NH4Cl, 170; CaCl2.2H2O, 17; MgCl2.6H2O, 125; FeCl3.6H2O, 4·1; KCl, 90; MnCl2.4H2O, 1·4; CoCl2.6H2O, 2·1; H3BO3, 0·4; CuCl2.2H2O, 0·19; Na2MoO4.2H2O, 0·18; and ZnCl2, 0·15) and micro nutrients (µg l1) (biotin, 4; folic acid, 4; pyridoxine.HCl, 20; riboflavin, 10; thiamine, 10; pantothenic acid, 10; nicotinic acid, 10; vitamin B12, 0·2; 4-aminobenzoic acid, 10; and thioctic acid, 10). All cultivations were carried out at 55 °C and under an atmosphere of 80 % N220 % CO2 (v/v) without shaking. During incubation, a 0·5 ml liquid sample was withdrawn every 23 days and filtered immediately through a 0·45 µm filter prior to chemical analysis.
Soluble volatile fatty acid concentration was determined using a gas chromatograph equipped with a flame-ionization detector (Shimadzu GC-14B), and an HP-FFAP capillary column (Agilent Technologies). The temperatures at the injection port, in the oven and at the detector port were 150, 135 and 150 °C, respectively, with nitrogen as the carrier gas. Aromatic compounds were determined by HPLC using a Shimadzu model FCV-10AL instrument equipped with a Shim_Pack VP-ODS separation column and an SPD-M10A UV-detector. The solvent used was 60 % (v/v) methanol containing 1 % (v/v) acetate at a flow rate of 0·8 ml min1. Biogas composition (methane, hydrogen and carbon dioxide) was analysed with a gas chromatograph GC-17A (Shimadzu) equipped with a thermal conductivity detector, and a 2 m stainless Supelco Porapak Q column (80/100 mesh) with nitrogen as the carrier gas.
Scanning electron microscopy (SEM).
The sludge sample taken at day 200 was fixed with 2·5 % glutaraldehyde overnight, mounted on poly-L-lysine-coated cover slips, and dehydrated in an ethanol series [25 % (v/v), 5 min; 50 %, 5 min; 75 %, 5 min; 95 %, 10 min; 100 %, 10 min; three times], followed by critical-point drying, then sputter coated with gold. The cells were examined using a Philips XL30 FEG scanning electron microscope.
Terminal-RFLP.
Analysis of 16S rRNA gene-based terminal-RFLP (T-RFLP) was performed according to a protocol described previously (Liu et al., 1997). From each sample taken from the sludge bed (at days 103, 172, 200 and 259) and the packing materials (at day 272), total community DNA was extracted using a protocol described previously (Liu et al., 1997
), and used in the subsequent PCR amplification with a domain Bacteria-specific forward primer 47f (5'-Cy5-CYTAACACATGCAAGTCG-3') and a reverse primer 927r (5'-ACCGCTTGTGCGGGCCC-3') (Amann et al., 1995
). The amplified PCR products were purified with the QIAquick PCR purification kit (Qiagen), and digested with three different tetrameric restriction nucleases (MspI, RsaI and HhaI). The digested products were denatured at 95 °C for 2 min, immediately chilled on ice, and analysed using a model CEQ 8000 automated sequencer (Beckman Coulter) at 55 °C and 4·8 kV for 2 h. The lengths of fluorescently labelled fragments were determined by comparison with internal standards using CEQ 8000-genetic analysis system software (Beckman Coulter). Only the T-RFs with abundance greater than 1 % of total intensity were used. For individual samples, the T-RFLP fingerprints were obtained based on the mean fingerprinting profiles of three different analyses. The observed T-RF lengths were later compared with and identified from the predicted T-RF lengths of the dominant clones obtained in the bacterial clone library.
Construction of 16S rRNA gene clone libraries.
Community 16S rRNA genes from domain Bacteria and Archaea were PCR-amplified with bacterial primer set EUB008F (Hicks et al., 1992) and 1512R (Kane et al., 1993
) and archaeal primer set A1F/A1100R (Embley et al., 1992
), respectively. Total community DNA extracted from the samples taken at days 200 and 259 was used as the template in PCR-reaction mixtures, which contained 1x PCR buffer (Invitrogen), 200 µM dNTPs, 2 mM MgCl2, 0·2 µM of each primer, and 2·5 U Taq DNA polymerase (Invitrogen) in a final volume of 100 µl. DNA amplification was performed in a Hybaid PCR Express thermocycler using a thermal program consisting of a hotstart at 94 °C for 1 min, 30 cycles of denaturation (30 s at 94 °C), annealing (30 s at 55 °C) and extension (30 s at 72 °C), and a final extension at 72 °C for 5 min. PCR products after confirmation were used in the construction of the 16S rRNA gene clone libraries as reported previously (Liu et al., 2002
).
Phylogenetic analysis and probe design.
Nearly full-length 16S rRNA gene sequences of representative clones were compared to available rRNA gene sequences in GenBank using the NCBI BLAST program (Altschul et al., 1997) and checked for chimeric artifacts using the CHECK-CHIMERA tool available in the Ribosomal Database Project (RDP) (Maidak et al., 2000
). Sequences of those selected clones and closely related bacterial species were aligned using the CLUSTAL_W program available in the BioEdit software package (Hall, 1999
). A neighbour-joining tree with the JukesCantor method was constructed with bootstrapping (1000 replicates) using the MEGA2 program (Kumar et al., 2001
). For probe design, sequences closely related to the clone sequences were retrieved and imported into a sequence alignment program, ARB (Ludwig et al., 2004
). After sequence alignment, 16S rRNA-targeted oligonucleotide probes targeting specific groups of sequences were designed using the probe design function provided in ARB.
Fluorescence in situ hybridization analysis.
The TA-degrading sludge samples taken were gently washed three times with 0·1 M PBS and fixed with 4 % paraformaldehyde at 4 °C overnight. Fluorescence in situ hybridization (FISH) was conducted according to a protocol previously described (Amann, 1995). In brief, hybridization was carried out at 46 °C for 3 h with hybridization buffer (0·9 M NaCl, 20 mM Tris/HCl pH 7·2, 0·01 % SDS) containing 5 ng µl1 of each fluorescent probe. When necessary, samples were briefly vortexed for 3 min before being immobilized on glass slides, and then were subjected to a freeze-and-thaw protocol (Sekiguchi et al., 1999
) to improve probe hybridization efficiency. Sludge samples were initially hybridized with the probe NON338 labelled with Cy3 to exclude nonspecific-probe-binding (Wallner et al., 1993
), and then were analysed with the domain- and group-specific oligonucleotide probes listed in Table 1
to provide microbial community structure. The hybridization efficiencies of the domain probes were verified by staining with 4',6-diamidino-2-phenylindole (DAPI). For the optimization of newly designed probes, the optimal formamide concentration was determined by simultaneously comparing the hybridization signal of microbial cells in sludge samples under different formamide concentrations (035 %) in the hybridization buffer. An Olympus BX51 epifluorescence microscope equipped with a cooled CCD camera SPOT-RT Slider (Diagnostic Instruments), a 100 W HBO bulb and fluorescence filter sets (U-MWU2, U-MWB2 and U-MF2) was used for the FISH analysis. The image analysis software, MetaMorph (Universal Imagine Corporation), was used to control the camera and to perform image analysis. Semi-quantitative FISH analysis was performed by analysing at least 12 microscopic fields selected randomly from the hybridization of individual probes (Bouchez et al., 2000
; Schmid et al., 2000
). Cells hybridized to a given probe in each field were expressed as a percentage of the total area of archaeal or bacterial domain probes hybridized by the ARC915 or EUB338_I/II/III, respectively, using the functions provided in MetaMorph.
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RESULTS AND DISCUSSION |
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At day 119, the feed was changed from synthetic PTA wastewater to feed with TA as the sole carbon source. Between days 126 and 172 the total and soluble COD removal efficiencies further improved to approximately 92·6 and 97·8 %, respectively. This improvement was attributed to the absence of the refractory p-toluic acid. TA removal was over 99 %, and the in situ apparent specific TA-degrading capacity further increased from 0·7 to 1·6 g (g VSS)1 day1. In addition, no metabolic intermediates of TA degradation were detected.
However, between days 173 and 200, the total COD- and TA-removal efficiency decreased from 90 and 98 % to 55 and 89 %, respectively. Intermediates including acetate, propionate and butyrate were found at concentrations of approximately 4, 2·8 and 6 mM, respectively, and represented approximately 21 % of the carbon coming from the degraded TA. The increases in these intermediates were suspected to be related to organic overload due to low biomass inventory in the reactor, since all other operating conditions (Bv, HRT, temperature, pH, etc.) remained unchanged. However, the decrease in biomass inventory was observed not to be related to sludge washout, and the reactor performances recovered on day 208. Thus, the exact cause of the decrease in removal efficiency remained unclear.
Between days 218 and 219 failure of the temperature controller occurred, which resulted in an increase in operational temperature to 92 °C for a period of approximately 3 h. Another perturbation in the system was encountered between days 225 to 232, when the feed and recycling pump failed, and the reactor operation had to be stopped. However, the reactor showed extremely high resistance to such drastic perturbations, and its performance recovered 1 week after operating conditions were restored. Good removal efficiencies were obtained again from day 241 onwards; 75·8 % for total COD and 95·7 % for TA.
TA degradation was further studied using a batch experiment with a sludge sample taken at day 200. No fermentative intermediates were detected during the incubation, except for trace levels of benzoate (79 µM) observed in the first 4 days. Complete TA degradation with concurrent formation of methane was achieved after 21 days of incubation. The rate of methane production [3·82 mmol (g VSS)1 day1] was three times higher than that of TA degradation [1·35 mmol TA (g VSS)1 day1 or 0·224 g TA (g VSS)1 day1]. The TA-degrading activity obtained under the batch conditions was much lower than the sludge specific removal capacity observed during reactor operation (Table 2
). This decrease was possibly due to the absence of agitation to enhance mass transfer and the use of a low TA concentration (5 mM) in the batch test. Overall performance results suggest that the thermophilic reactor could be equally or more effective in degrading TA than the mesophilic reactor, and would be an attractive biological process for future treatment of PTA wastewater.
SEM-based morphological observations
SEM analysis revealed at least six different dominant morphotypes present in the sludge sample taken at day 200; each morphotype is indicated by an arrow in Fig. 1. These included bamboo-shaped cells (0·60·7x3·315 µm) (arrow 1), rods with flat ends (0·60·7x1·63·3 µm) (arrow 2), fat rods (0·60·8x1·5 µm) (arrow 3), very small rods (0·3x1·62·0 µm) (arrow 4), long and slender rods (0·3x3·34·8 µm) (arrow 5), and thin filaments (0·2x2533 µm) (arrow 6). The fat rods (arrow 3) were the dominant cells found in the thermophilic TA-degrading hybrid reactor, and are presumed to be responsible for TA degradation. The second most dominant cells were bamboo-shaped cells (arrow 1) and rods with flat ends (arrow 2), which closely resembled Methanosaeta-like species, which are methanogens that utilize acetate.
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The MspI-digested T-RFLP profile was further determined for the biofilm sample D272 (i.e. attached growth) obtained from the surface of reactor packing materials. The profile (Fig. 2e), showing 171 bp T-RF as the only major peak (
72 % of total peak area), was very similar to samples D103, D172 and D200 obtained from the sludge bed (i.e. suspended growth). Apparently, the difference in attached growth and suspended growth did not have a significant effect on the TA-degrading community structure when good mixing was provided in the reactor under a steady-state operation. However, when perturbation (e.g. heat shock and pump failure) occurred, the microbial population structure in the sludge bed, as observed in Fig. 2(d)
, could be more susceptible to environmental changes than that on the surface of the packing materials. This was attributed to the physical protection provided to the microbial populations by the formation of a biofilm (Stickler, 1999
). A similar observation was reported in reactors with immobilized biomass under temperature perturbation (van Lier, 1996
).
It was possible that the attached growth could have served as a good seeding source to help the reactor recover rapidly from process upsets, for example, between day 219 and day 258 in this study. The amount of biomass attached on the surface of the packing materials was not determined in this study due to technical difficulty in sampling during reactor operation, but was reported to be 22·28 % of the total sludge biomass from our previous study (S. Thierry, I. Ramirez, C. Allouche, H. Ferrer & H. Macarie, unpublished results). In view of this, reseeding mechanisms could possibly provide full-scale hybrid reactors with better operational stability than systems with suspended sludge biomass. This explanation would be more convincing if the microbial community structures on the surface of the packing material could have been constantly monitored during the reactor operation.
Thermophilic TA-degrading consortium as revealed by 16S rRNA gene clone libraries
To further understand the thermophilic TA-degrading consortia, the phylogenetic composition of the fermentative bacteria and methanogens in the reactor was examined through clone library construction. An archaeal clone library was constructed for the methanogenic counterpart in sample D200. In total, 60 clones were selected and classified into five different sequence types or phylotypes after clone screening and sequencing. Phylogenetic analysis (Fig. 3) showed that four (TTA_A1, 4, 5 and 64; 93 % of total archaeal clones) of those five phylotypes were closely related to Methanothrix thermophila in the acetoclastic Methanosaeta group. The remaining clone (TTA_A24; 7 %) was related to hydrogen-utilizing Methanospirillum species and an environmental clone UASB_TA02 from a mesophilic TA-degrading UASB process (Wu et al., 2001
).
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The phylogenetic tree in Fig. 4 indicated that the 22 phylotypes obtained from sample D200 were affiliated with nine different bacterial divisions. Most phylotypes were from the Gram-positive low-G+C group (LGC) (7 phylotypes, 42·2 % of total bacterial clones) and the candidate division OP5 (5 phylotypes, 30·8 %). In the LGC, the most dominant phylotype (TTA_B12; 31·9 %) and phylotype TTA_B5 (0·7 %) were affiliated with the Desulfotomaculum group, which consists of physiologically diverse isolates and environmental clone sequences from various thermophilic and mesophilic environments (Castro et al., 2000
; Dojka et al., 1998
; Imachi et al., 2002
; Stubner & Meuser, 2000
). The second most dominant phylotype (TTA_B6; 6·1 %) was closely related to environmental clone MUG10, derived from a mesophilic UASB granule (Sekiguchi et al., 1998
). The remaining phylotypes at low abundance were clustered together with an uncultured hot spring clone OPB54 (Hugenholtz et al., 1998
) and environmental clone SHA-118, derived from an anaerobic dechlorinating bioreactor (Schlötelburg, 2001
). The five phylotypes in the candidate division OP5 formed a cluster closely related to environmental clones obtained from an anaerobic dechlorinating bioreactor (Schlötelburg, 2001
) and a contaminated aquifer (Dojka et al., 1998
). The remaining ten phylotypes from sample D200 were found in the CytophagaFlexibacterBacteroides (CFB) group (2 phylotypes, 7·5 % of total clones), the
-Proteobacteria (3 phylotypes, 4·9 %), the
-Proteobacteria (1 phylotype, 0·7 %), Chlorobi (1 phylotype, 1·4 %), the candidate division OP8 (1 phylotype, 4·8 %), Planctomycetes (1 phylotype, 0·7 %), and Thermotogae (1 phylotype, 0·7 %).
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The clone library results suggested that various microbial populations were capable of degrading TA to acetate, CO2 and hydrogen, which were further mineralized by the methanogenic counterpart. The Desulfotomaculum- and OP5-related phylotypes were the main populations in the hybrid reactor under steady-state operation. However, when perturbations occurred, other phylotypes, for example from the LGC, the CFB, the -Proteobacteria and the
-Proteobacteria, could emerge rapidly and become involved in TA degradation.
It was further observed that in anaerobic TA-degrading reactors, the dominant populations in the thermophilic hybrid reactor were different from those obtained from mesophilic reactors, where a yet-to-be-cultured group from the -Proteobacteria was identified as the dominant group (Wu et al., 2001
; Qiu et al., 2004
). Similar observations were reported between mesophilic and thermophilic methanogenic granular reactors degrading a mixture of sucrose, propionate and acetate (Sekiguchi et al., 1998
). Recently, a mesophilic TA-degrading bacterium strain JT within the Desulfotomaculum group was successfully isolated from a UASB granular sludge (Qiu et al., 2004
). This further suggested that the Desulfotomaculum group could possibly contain both mesophilic and thermophilic TA-degrading members.
Phylogenetic identity of T-RFs observed in community T-RFLP profiles
The MspI-digested T-RF sizes of the phylotypes obtained from samples D200 and D259 were initially determined and verified by comparing them with the predicted T-RF sizes from the sequences obtained. The observed T-RF sizes of these phylotypes (Fig. 4) were compared with those observed in the T-RFLP profiles (Fig. 2
). This comparison revealed that the most dominant 171 bp T-RF observed in most of the T-RFLP profiles represented phylotypes related to the Desulfotomaculum group (e.g. TTA_B12). Those T-RFs with sizes of 99, 231 and 262 bp probably represented phylotypes TTA_H16, TTA_H4 and TTA_H122, respectively, in the LGC group. The 469 bp T-RF represented Syntrophus gentianae from the
-Proteobacteria (e.g. TTA_B114). The T-RFs with sizes of 423 and 503 bp represented phylotypes TTA_H6 and TTA_H13 from the CFB group, respectively. It was further observed that some T-RFs (e.g. 138 bp T-RF and 151 bp T-RF) could represent more than one phylotype, sometimes from different bacterial divisions. For example, the 138 bp T-RF observed in samples could represent phylotypes either from the OP5 (e.g. TTA_B4) or related to T. acidaminovorans from the LGC group or possibly both. These results were in synergy with T-RFLP and clone library results.
Thermophilic TA-degrading syntrophic consortium as revealed by FISH
FISH with rRNA-targeted probes specifically for domains Archaea (ARC915) and Bacteria (EUB338_I/II/III) was used to examine the distribution of bacterial and archaeal cells in the sludge bed taken at day 200. Fig. 5(a) indicates that bacterial cells (green) and archaeal cells (red) accounted for 48·1±3·0 and 49·2±2·6 % of the DAPI-stained cells, respectively.
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The 171 bp T-RF phylotypes related to the Desulfotomaculum group were further confirmed using FISH with probe DFMI227a targeting members of the Desulfotomaculum group (Loy et al., 2002). Fig. 5(b)
shows that probe DFMI227a could hybridize to fat rods (green), representing approximately 37·0±6·6 % of total bacterial cells hybridized by probe EUB338_I/II/III. Thus, these fat rods were probably responsible for the thermophilic TA degradation. Furthermore, a new probe TA55_OP5 was designed and optimized for those phylotypes found in the OP5-like division. It hybridized to very small rod-shape cells, accounting for the second largest fraction of the bacterial populations in the sludge (26·9±5·8 %) (data not shown).
FISH results further indicated that members of the CFB group and -Proteobacteria accounted for only 7·7±3·2 and 3·2±0·6 %, respectively, of total bacterial cells in the methanogenic consortium in sample D200, and exhibited morphological traits of either long slender rods or fat rods (data not shown). In addition, a specific probe delta-TA1 targeting the
-proteobacterial TA group found in mesophilic TA-degrading sludge (Wu et al., 2001
) was used in the FISH analysis, but did not hybridize to any cells. This study further suggested that the thin filaments observed under SEM (arrow 6 in Fig. 1
) were members of green non-sulfur bacteria found in thermophilic UASB sludge treating a mixture of sucrose, acetate and propionate (Sekiguchi et al., 1999
). However, this possibility was rejected based on FISH analysis with probe GNSB633 targeting this filamentous group (Sekiguchi et al., 1999
).
Overall, our results based on T-RFLP, clone library and FISH analysis have revealed that the Desulfotomaculum-related phylotypes (171 bp T-RF) were the main populations involved in the initial thermophilic fermentation of TA to give products further mineralized by the methanogenic counterpart. Likewise, other phylotypes representing different T-RFs observed in the T-RFLP profiles of sample D259 were also capable of TA degradation. Further studies on the enrichment and isolation of the thermophilic TA-degrading consortium could support this proposition, and would be useful to better understand the physiological role of the dominant bacterial groups in the thermophilic TA-degrading pathway and for process operation.
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ACKNOWLEDGEMENTS |
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Received 29 March 2004;
revised 2 July 2004;
accepted 9 July 2004.
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