Identification of important microbial populations in the mesophilic and thermophilic phenol-degrading methanogenic consortia

ARTICLE IN PRESS WAT E R R E S E A R C H

42 (2008) 1963 – 1976

Available at www.sciencedirect.com

journal homepage: www.elsevier.com/locate/watres

Identification of important microbial populations in the mesophilic and thermophilic phenol-degrading methanogenic consortia Chia-Lung Chena,b, Jer-Horng Wuc, Wen-Tso Liub, a

Department of Civil Engineering, National University of Singapore, Singapore 117576, Singapore Division of Environmental Science and Engineering, 9 Engineering Drive 1, Block EA-07-23, National University of Singapore, Singapore 117576, Singapore c Sustainable Environment Research Center, National Cheng Kung University, Tainan 701, Taiwan b

art i cle info

ab st rac t

Article history:

Active mesophilic and thermophilic phenol-degrading methanogenic consortia were

Received 16 May 2007

obtained after an 18-month acclimation and enriching process in the serum bottles, and

Received in revised form

characterized using the rRNA-based molecular approach. As revealed by cloning,

22 November 2007

fluorescence in situ hybridization (FISH) and terminal restriction fragment length

Accepted 24 November 2007

polymorphism (T-RFLP), these two enrichments differed greatly in the community

Available online 5 December 2007

structures. The results for the first time suggest that group TA in the Deltaproteobacteria

Keywords: Anaerobic degradation Microbial consortia Phenol Phylogeny Cloning

(88.0% of EUBmix FISH-detectable bacterial cell area) and Pelotomaculum spp. in the Desulfotomaculum family (81.2%) were the predominant fermentative bacteria under mesophilic (37 1C) and thermophilic (55 1C) conditions, respectively. These populations closely associated with mesophilic and thermophilic members of Methanosaetaceae, Methanobacteriaceae and Methanomicrobiales to mineralize phenol as the sole carbon substrate to carbon dioxide and methane. Moreover, these two enrichments could mineralize terephthalate and benzoate. During benzoate degradation in the mesophilic

FISH

enrichment, a shift in the predominant bacterial population from Deltaproteobacteria group TA to Syntrophus spp. was observed, suggesting Syntrophus-related spp. could have a higher substrate affinity for benzoate. FISH further revealed that member of the Deltaproteobacteria group TA represented more than 68.3% of EUBmix FISH-detectable bacterial cell area in a full-scale mesophilic bioreactor treating phenol-containing wastewaters. & 2007 Elsevier Ltd. All rights reserved.

1.

Introduction

Phenol-containing wastewater is generated during coal gasification, petroleum refining and petrochemical manufacturing. This waste stream is usually treated with aerobic or anaerobic biological processes, and based on the energy required and sludge disposal cost; anaerobic methanogenic processes are considered as a more attractive solution (Fang et al., 1996; Corresponding author. Tel.: +65 65161315; fax: +65 67744202.

E-mail address: [email protected] (W.-T. Liu). 0043-1354/$ - see front matter & 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2007.11.037

Young and Rivera, 1985). So far, a number of phenol-degrading methanogenic bioreactors operated under ambient and mesophilic conditions have been described (Kleerebezem and Macarie, 2003; Macarie, 2000; Veeresh et al., 2005). Observations from operation of these processes indicate that a long start-up period is often required, and this further suggests that the degradation pathways and the microbial populations involved in this anaerobic process are poorly understood.

ARTICLE IN PRESS 1964

WAT E R R E S E A R C H

42 (2008) 1963– 1976

Under methanogenic conditions, phenol degradation requires the coupling of fermentative steps (by fermentative bacteria) with methanogensis (by methanogens) in close physical proximity to overcome thermodynamics limit. Therefore, it is suggested that phenol is initially transformed to benzoate via reductive carboxylation (Kobayashi et al., 1989; Gallert et al., 1991) and then through the benzoyl-CoA pathway producing acetate and hydrogen (Harwood et al., 1999). It is also reported that phenol can be degraded via n-caproate producing acetate and hydrogen (Evans, 1977). Fang and co-workers (Fang et al., 2004, 2006; Zhang et al., 2005) have studied the microbial populations involved in the phenol degradation under methanogenic conditions at 26 and 55 1C. At 26 1C, they reported that phenol is first converted to benzoate possibly by Desulfotomaculum spp. and Clostridium spp., and then to acetate and H2/CO2 by Syntrophus spp. These intermediate by-products are then converted to methane and CO2 by members of Methanosaetaceae, Methanomicrobiales and Methanobacteriaceae (Fang et al., 2004; Zhang et al., 2005). At 55 1C, a unique community structure was observed and discriminated from that associated at 26 1C (Fang et al., 2006). In the study, however, the key populations involved in the degradation of phenol at 55 1C were not clearly identified. With respect to the two temperatures, most anaerobic bioreactors treating phenol-containing wastewaters were operated under mesophilic (3037 1C) conditions (Fang et al., 1996; Chang et al., 1995; Veeresh et al., 2005). To our knowledge, current understanding as to the diversity of phenol-degrading anaerobes in the temperature range remained unclear. To better understand the effect of the temperature on the diversity of phenol-degrading methanogenic consortia, it is crucial to extend the community analysis to the phenol-degrading consortia acclimated under 37 and 55 1C conditions. In this study, two microbial communities that degrade phenol efficiently under mesophilic or thermophilic methanogenic conditions were obtained and their microbial community structures characterized using full-cycle 16S rRNA gene-based approach. In addition, the abilities of these two enrichments to degrade terephthalate and benzoate were determined. The abundance of the predominant microbial population found in the mesophilic enrichment was examined in a full-scale bioreactor treating actual phenol-containing wastewater.

2.

Materials and methods

2.1.

Cultivation of phenol-degrading consortia

The seeding sludge used for the cultivation of phenoldegrading consortia was taken from an anaerobic digester at a local wastewater treatment plant receiving both industrial (66.6% of total influent volume) and domestic wastewaters. The anaerobic digester was operated at ambient temperature (28–30 1C), and the concentrations of suspended solid and volatile suspended solid (VSS) were approximately 30 and 14 g l1, respectively. Five milliliters of seeding sludge were inoculated into 120-ml serum bottles containing 50 ml of anaerobic culture medium pre-sparged with 80% N2–20% CO2

(v/v) (Angelidaki et al., 1990; Chen et al., 2004). Concentrated phenol solution was added as the sole carbon source to a final concentration of 100 mg l1, and incubated at 37 and 55 1C without shaking. Once the phenol was completely degraded, the concentration used in subsequent culture transfers (10% inoculum, v/v) was increased stepwise to 250, 500 and finally 750 mg l1. Phenol at a concentration of 750 mg l1 was used in subsequent culture transfers. The cultures (5 ml) were anaerobically transferred using a syringe and needle. This enrichment process was conducted for up to 18 months. Afterward, the effective phenol-degrading mesophilic enrichment (MP) (after 10 successive transfers) and thermophilic enrichment (TP) (after 7 successive transfers) were used for batch degradation assays and microbial community analyses.

2.2.

Biological granular-activated carbon sample

A biological granular-activated carbon (GAC) sample was obtained from a full-scale anaerobic fluidized bed reactor (275 m3) treating resin manufacturing industrial wastewater under mesophilic conditions for more than 6 years in Taiwan. The size of biological GAC particles was between 2 and 4 mm. The influent contained 2500–3000 mg l1 chemical oxygen demand (COD) (approximately 60% was contributed by phenol). The reactor had a COD removal efficiency of about 85% under a loading rate of 3–5 kg COD m3 day1.

2.3.

Batch experiment of substrate utilization

Microbial biomass from enrichments MP and TP were transferred into 120-ml serum bottles containing 50 ml of freshly prepared media containing 500 mg l1 (5.3 mM) of phenol to obtain the degradation profile. During incubation, phenol depletion and methane production were measured every 2–3 days. Phenol depletion was monitored till the concentration of phenol was below to the detection limit and the VSS concentrations were determined at the end of the experiment. Another batch assay was conducted to test the degradability of benzoate (8 mM) and terephthalate (5 mM) as the sole carbon source by the enrichments MP and TP. In this assay, phenol (5 mM) was also included as the positive control to ensure the phenol-degrading activity of incoculum. Concentration of phenol, benzoate or terephthalate was determined at the start and the end of 28-day incubation. Three portions of 0.5 ml liquid samples were withdrawn separately with a syringe and needle, and filtered immediately through a 0.45 mm filter prior to chemical analysis. Soluble short-chain volatile fatty acids (i.e. acetate, propionate and butyrate) were determined using a gas chromatograph equipped with a flame-ionization detector (Shimadzu GC-14B), an HP-FFAP capillary column (Agilent Technologies, USA) and with nitrogen as the carrier gas. Aromatic compounds (phenol, benzoate and terephthalate) were determined using high-pressure liquid chromatography (model FCV-10AL, Shimadzu, Japan) equipped with an Eclipse XDBC18 column (Agilent, USA) and an SPD-M10A UV-detector. The solvent used was 50% (v/v) of acetonitrile as mobile phase at a flow rate of 0.8 ml min1. Bio-gases (i.e., methane, hydrogen and carbon dioxide) were analyzed with a gas chromatograph GC-17A (Shimadzu, Japan) equipped with a

ARTICLE IN PRESS WAT E R R E S E A R C H

1965

42 (2008) 1963 – 1976

thermal conductivity detector, and a 2-m stainless Supelco Porapak Q column (80/100 mesh) with nitrogen as the carrier.

program (Kumar et al., 2004) was used to construct the neighbor-joining tree with the Juke-Cantor correction and bootstrapping for 1000 replicates to estimate the confidence of the tree topology.

2.4. DNA extraction, 16S rRNA gene clone library and phylogeny analysis Genomic DNA from enrichments was extracted using chemical lysis and phenol–chlorform–isoamyl alcohol (25:24:1, v:v:v) purification protocol as described previously (Liu et al., 1997). Community 16S rRNA genes from domains Bacteria and Archaea were amplified with bacterial primer 8F and prokaryotic primer 1490R (Weisburg et al., 1991) and archaeal primer A1F (Embley et al., 1992) and 1490R, respectively (Table 1). ‘‘Reconditioning PCR’’ protocol was used to minimize the production of PCR artifacts (Thompson et al., 2002). TOPO TA cloning kit (Invitrogen, CA) was used for clone library construction according to manufacturer’s instruction. The PCR product of 16S rRNA gene inserts from each clone was screened using restriction fragment length polymorphism (RFLP) for archaeal clone library and using denaturing gradient gel electrophoresis (DGGE) for bacterial clone library, respectively. For RFLP screening, three tetramer restriction enzymes, namely, MspI, RsaI and HhaI (New England Biolabs Inc., MA) were used. For DGGE screening, the PCR products amplified by primers 968F-GC and 1392R (Table 1) were analyzed on a DGGE gel with the DeCode system (BioRad) as described previously (Liu et al., 2002). Nearly full-length 16S rRNA gene sequences (41350 bp) of representative clones were obtained by sequencing with three primers (i.e., 8F, 1088R and 1490R for Bacteria; A1F, Arc915R and 1490R for Archaea) (Table 1) on CEQ 8000 (Beckman Coulter) and compared to available rRNA gene sequences in GenBank using the NCBI BLAST program (Altschul et al., 1997). Chimeric artifacts were determined using Bellerophon (Huber et al., 2004) and Pintail (Ashelford et al., 2005). MEGA3

2.5. Probe design and fluorescence in situ hybridization (FISH) Semi-quantitative FISH measurements were performed on paraformaldehyde-fixed enrichment samples and biological GAC samples from the full-scale anaerobic fluidized bed reactor according to protocols described previously (Amann et al., 1995). For the biological GAC sample, the paraformaldehyde-fixed particles were embedded with Jung Tissue Freezing MediumTM (Leica, Germany), and cut into 20 mmthick sections with a cryotome (model CM3050S, Leica). The oligonucleotide probes used included EUBmix (i.e., EUB338, EUB338-II, EUB338-III), targeting most of Bacteria (Amann et al., 1995; Daims et al., 1999); ARCH915, targeting most of Archaea (Amann et al., 1995); Ih820, targeting members of Desulfotomaculum subcluster Ih (Imachi et al., 2006) (Table 1). In addition, an oligonucleotide probe Delta-TA664 (664–681 in Escherichia coli positions) (Table 1) targeting members of Deltaproteobacteria group TA was re-designed using the probe design function in ARB (Ludwig et al., 2004), and the specificity confirmed using the PROBE-MATCH program in the Ribosomal Database Project (Cole et al., 2005). The stringency of FISH hybridization was optimized using different formamide concentrations in the hybridization buffer (25% [v/v] for Delta-TA664, and 35% for EUBmix, ARCH915 and Ih820). For probe Delta-TA664, the optimal hybridization condition was determined by comparing the hybridization signal of Deltaproteobacteria group TA-containing enrichment under different formamide concentrations with Desulfonema limicola (ATCC 33961) as a negative control.

Table 1 – 16S rRNA genes oligonucleotide used in this study Oligonucleotide

Specificity

Sequence (50 –30 )a

Analysis

Reference

Primer 8F 1490R 968F-GCb 1392R 47F 927R 1088R A1F Arc915R

Eubacteria Prokaryote Eubacteria Prokaryote Eubacteria Eubacteria Eubacteria Archaea Archaea

AGAGTTTGATYMTGGCTC GGYTACCTTGTTACGACTT AACGCGAAGAACCTTAC ACGGGCGGTGTGTAC CYTAACACATGCAAGTCG ACCGCTTGTGCGGGCCC GCTCGTTGCGGGACTTAACC TCYGKTTGATCCYGSCRGAG GTGCTCCCCCGCCAATTCCT

Phylogeny Phylogeny DGGE DGGE T-RFLP T-RFLP Phylogeny Phylogeny Phylogeny

Weisburg et al. (1991) Weisburg et al. (1991) Heuer et al. (1997) Ferris et al. (1996) Chen et al. (2004) Amann et al. (1995) Amann et al. (1995) Embley et al. (1992) Stahl and Amann (1991)

Eubacteria Archaea Deltaproteobacteria group TA Desulfotomaculum subcluster Ih

GCWGCCWCCCGTAGGWGT GTGCTCCCCCGCCAATTCCT CGGGAATTCCGTTTCCCT CCTCCTACACCTAGCACC

FISH FISH FISH FISH

Daims et al. (1999) Stahl and Amann (1991) This study Imachi et al. (2006)

Probe EUBmix ARC915 Delta-TA664 Ih820 a b

Y: C and T; K: G and T; S: G and C; R: A and G; M: A and C; W: A and T. For primer 968F-GC, a 40-bp GC clamp (50 -CGCCCGGGGCGCGCCCCGGGCGGGGCGGGGGCACGGGGG G-30 ) was included at the 50 end.

ARTICLE IN PRESS 1966

WAT E R R E S E A R C H

42 (2008) 1963– 1976

When dual-staining FISH was performed, probes labeled with cyanine 3 (Cy3) or cyanine 5 (Cy5) were used. FISH-stained images were captured by a confocal laser scanning microscope (CLSM) (model LSM Pascal, Carl Zeiss) equipped with two helium/neon lasers (543 and 633 nm) under X100 objective lens. At least 10 microscopic fields were randomly acquired to obtain statistically valid determinations. For biological GAC samples, biofilm sections from at least five pellets were examined. The quantitative analyses were based on counting stained cell area using the functions provided in the image analysis software, Metamorph (Universal Image). The numbers obtained for each were averaged and standard derivations of the mean were calculated.

Through the successive culture transfer, the cultures were able to consistently mineralize phenol to final products (i.e. methane and carbon dioxide). The effective phenol-degrading mesophilic enrichment (MP) and thermophilic enrichment (TP) after 10 and 7 successive transfers, respectively, were used for anaerobic degradation and microbial analyses. As shown in Fig. 1, complete phenol degradation at 500 mg l1 (5.3 mM) was observed together with methane production for both MP and TP enrichments within 20 days. The specific phenol degradation rate for the MP enrichment was approximately 1.9  103 mmol mg VSS1 day1 (Fig. 1a).

2.6. Terminal restriction fragment length polymorphism (T-RFLP)

Phenol 12.0 Substrate and Products (mM)

16S rRNA gene T-RFLP was performed as described previously (Liu et al., 1997). Briefly, total community DNA extracted from phenol-, terephthalate- and benzoate-degrading cultures was used in the PCR amplification using domain Bacteria-specific primer set 47F fluorescently labeled with Cy5 at the 50 end and 927R (Table 1) (Chen et al., 2004). The tetramer restriction enzyme MspI was used in the microbial fingerprinting analysis. The MspI-digested terminal restriction fragment (T-RF) lengths were separated and detected using automatic DNA sequencer (CEQ 8000, Beckman coulter). Fragment sizes were determined based on internal standards (DNA size standard kit-600, Beckman Coulter) using the software provided with the CEQ 8000-genetic analysis system (Beckman coulter). Only those T-RFs with abundance greater than 1% of total intensity were used. T-RFLP fingerprints were obtained based on the average of two duplicates, and reproduced by software MicrocalTM Origins Version 6.0 (Microcal Software Inc., USA). The phylogeny affiliation of those major peaks observed in the T-RFLP analyses were identified based on the in-silico terminal restriction fragment analysis done on the sequences of those 16S RNA gene clones obtained.

14.0

Methane Hydrogen

10.0

Acetate 8.0 6.0 4.0 2.0 0.0 0

Results

3.1.

Enrichment of phenol-degrading consortia

Initial phenol degradation and methane production in mesophilic and thermophilic cultures were observed 3 weeks and 7 months, respectively, after incubation. The long acclimation period required for thermophilic culture was likely attributed to the low abundance of those slow-growing thermophilic fermentative populations present in the seeding sludge. The cultures (10%, v/v) were successively transferred repetitively into fresh medium containing phenol with the concentration increased gradually from 100 to 750 mg l1.

Substrate and Products (mM)

3.

20

25

10 15 Time (days)

20

25

Phenol Methane

Nucleotide sequence accession numbers

The GenBank/EMBL/DDBJ accession numbers for the 16S rRNA gene sequences reported in this paper are EF198024– EF198052.

10 15 Time (days)

14.0 12.0

2.7.

5

Hydrogen

10.0

Acetate 8.0 6.0 4.0 2.0 0.0 0

5

Fig. 1 – Degradation of phenol in (a) mesophilic and (b) thermophilic enrichments.

ARTICLE IN PRESS WAT E R R E S E A R C H

42 (2008) 1963 – 1976

Fig. 1b indicates that the TP enrichment exhibited a slow phenol degradation rate (5.6  104 mmol mg VSS1 day1) during the first 8 days of incubation. However, the maximal specific phenol degradation rate (5.2  103 mmol mg VSS1 day1) of the TP enrichment is observed from day 10 to 14, which is 2.74-fold faster than that of the MP enrichment. Correspondingly, the maximal specific methane production rate (1.0  102 mmol mg VSS1 day1) of the TP enrichment is higher than that 7.5  103 mmol mg VSS1 day1 of the MP enrichment. During the incubation, no fermentative intermediates except acetate, which was momentarily detected at a maximum concentration of 0.8 mM on day 8 in the MP enrichment and 0.2 mM on day 10 in the TP enrichment, respectively. These observations validated the activity of phenol-degrading microbial communities and suggested the difference of phenol degradation kinetics under mesophilic and thermophilic conditions.

3.2. Microbial compositions as revealed by 16S rRNA gene clone library Microbial compositions of the MP and TP enrichments were characterized by cloning and sequencing of 16S rRNA gene sequences that were PCR amplified individually with domain Archaea- and Bacteria-specific primer sets. For the members in the domain Archaea, 61 and 68 clones were randomly selected from MP and TP, respectively. Of the total clones screened by

1967

RFLP, four unique sequence types or operational taxonomy units (OTUs) were obtained from each library, and separately classified into three different phylogenetic groups (Fig. 2). OTUs MP101 and MP104 (88.5% of total archaeal clones) from the MP enrichment, and OTUs TP103, TP114 and TP123 (92.6%) from the TP enrichment were closely related to mesophilic Methanosaeta concilii and thermophilic Methanothrix thermophila (i.e., Methanosaeta thermoacetophila) in the acetotrophic methanogens Methanosaetaceae, respectively. OTU MP123 and OTU TP120 were closely related to mesophilic Methanobacterium beijingense and thermophilic Methanothermobacter thermoautotrophicus, respectively, from Methanobacteriaceae. The remaining mesophilic OTU MP126 was clustered with environmental clones and a novel acidiphilic methanogen (Candidatus Methanoregula boonei) isolated from acidic peat bog (Brauer et al., 2006) in Methanomicrobiales. This cluster was branched from its closest known genera of Methanoculleus and Methanospirillum. For the members in the domain Bacteria, 107 and 114 clones from MP and TP samples were constructed for clone libraries, respectively, and after screened by DGGE, they were classified to 10 and 11 OTUs, respectively. These 21 OTUs were phylogenetically affiliated with seven different taxonomic clades (Fig. 3a). For MP, six of the ten OTUs (76.6% of total clones) obtained were affiliated with the Deltaproteobacteria, two (11.1%) with green nonsulfur bacteria (Chloroflexi), one

Fig. 2 – Phylogenetic tree of 16S rRNA gene sequences constructed for archaeal clones obtained from mesophilic and thermophilic phenol-degrading enrichments. The relevant sequences were aligned using ClustalW program provided in MEGA3 package. The tree was constructed using the neighbor-joining algorithm with Jukes-Cantor model (bootstrapping number ¼ 1000). The 16S rRNA gene sequence of Methanopyrus kandleri (M59932) was used as the outgroup. Only bootstrap values greater than 50% of the replicates are indicated at branch points. The scale bar represents the estimated number of nucleotide changes per sequence position.

ARTICLE IN PRESS 1968

WAT E R R E S E A R C H

42 (2008) 1963– 1976

Fig. 3 – Phylogenetic trees of bacterial clones obtained from mesophilic and thermophilic phenol-degrading enrichments for (a) domain Bacteria, (b) Deltaproteobacteria group TA and (c) Desulfotomaculum subcluster Ih. The trees were rooted with the 16S rRNA gene sequences of Methanococcus vanniellii (M36507), Desulfotomaculum thermobenzoicum (L15628) and Desulfotomaculum guttoideum (Y11568), respectively. Only bootstrap values greater than 50% of the replicates are indicated at branch points. The scale bar represents the estimated number of nucleotide changes per sequence position (see Chouari et al., 2005 in Fig. 3(b)).

ARTICLE IN PRESS WAT E R R E S E A R C H

42 (2008) 1963 – 1976

1969

Fig. 3 – (Continued)

(3.7%) with the Spirochaetes and one (4.7%) with Verrucomicrobia. In Deltaproteobacteria, OTU MP18 (57.0%) and three other OTUs (MP1, MP27 and MP29; 9.4%) present with high abundance in the clone library were affiliated with a recently recognized lineage, group TA (Wu et al., 2001) (Fig. 3b). OTU MP17 (9.3%) was closely related to Desulfovibrio species and OTU MP91 (0.9%) to Syntrophus spp. For TP, seven of the 11 OTUs were affiliated with the Grampositive low G+C group (LGC) (Firmicutes) (78.9% of total clones screened by DGGE), one with candidate division OP8 (5.3%), one with Verrucomicrobia (3.5%) and two with Thermotogae (6.1%) (Fig. 3a). In the LGC group, OTU TP18 (40.4%) and OTU TP15 (14.9%) predominant in the clone library were closely

related to environmental clones of the Desulfotomaculum subcluster Ih (Fig. 3c) that were obtained from a thermophilic terephthalate-degrading sludge (Chen et al., 2004); OTUs TP16 and TP41 (14.0%) to a syntrophic fatty acids-degrading bacterial strain TOL; OTUs TP25 and TP34 (7.0%) to a Clostridium sp.; and OTU TP45 to clone TTA_H39 obtained from a thermophilic terephthalate-degrading consortium. In the candidate division OP8, OTU TP11 (5.3%) was closely related to clone TTA_B3 retrieved from a thermophilic terephthalate-degrading community (Chen et al., 2004). In the Thermotogae, two similar OTUs TP29 and TP35 were placed together with the environmental clone obtained from another thermophilic phenol-degrading sludge (Fang et al., 2006).

ARTICLE IN PRESS 1970

WAT E R R E S E A R C H

42 (2008) 1963– 1976

Fig. 3 – (Continued)

ARTICLE IN PRESS WAT E R R E S E A R C H

42 (2008) 1963 – 1976

1971

Fig. 4 – FISH analyses of mesophilic (a and c) and thermophilic (b and d) phenol-degrading enrichments. In panels (a) and (b), samples were hybridized with Cy3-labeled EUBmix probe specific for the domain Bacteria (green), and the Cy5-labeled ARC915 probe specific for the domain Archaea (red). (c) Samples were hybridized with Cy3-labeled EUBmix probe (green) and Cy5-labeled Delta-TA664 probe specific for Deltaproteobacteria group TA (red). (d) Samples were hybridized with Cy3-labeled EUBmix probe (green) and Cy5-labeled Ih820 probe specific for Desulfotomaculum subcluster Ih (red). Arrow indicates a sporeforming cell. Bar ¼ 10 lm.

3.3. FISH

Predominant microbial populations as revealed by

FISH results (Fig. 4a) indicated that oval- (green) and filamentshaped (red) cells were the most predominant morphotypes for the bacterial and archaeal populations, respectively, found in MP. In TP, the most predominant archaeal morphotype (in red) was filament-shaped cells, and the predominant bacterial morphotypes (in green) were fat rods and thin rods. The filamentous archaea observed in both mesophilic and thermophilic enrichments morphologically resembled Methanosaeta-like spp. Sludge samples taken from MP and TP were examined using FISH with probes DETLA-TA664 and Ih820 targeting the Deltaproteobacteria group TA and Desulfotomaculum subcluster Ih, respectively. In MP, probe DELTA-TA664 could bind on to the oval-shaped cells (yellow), representing approximately

88.076.6% of total bacterial cells hybridized by EUB338mix (Fig. 4c). In TP, probe Ih820 could hybridize to fat rods (yellow), representing 81.275.8% of EUB338mix-hybridized bacterial cells (Fig. 4d). These Ih820-targeted fat rods were observed to have a spore-like cellular trait (arrow in Fig. 4d), suggesting that they were likely the spore-forming organisms.

3.4. Population changes of phenol-degrading MP and TP enrichments associated with terephthalte and benzoate degradation We further evaluated the abilities of MP and TP to degrade benzoate and terephthalate as the sole carbon source (Table 2). In the batch tests, the degradability was inferred by the consumption of substrate using the highly enriched phenol-degrading consortia after 28 days of incubation.

ARTICLE IN PRESS 1972

WAT E R R E S E A R C H

42 (2008) 1963– 1976

Table 2 – Degradability of benzoate and terephthalate of phenol-degrading enrichments within 28 days Phenol-degrading enrichment used

Substrate

Initial substrate concentration (mM)

% Degraded in 28 days

Methane production (mM)

MP

Phenol Terephthalate Benzoate

5.0270.14 4.9370.37 8.2170.21

100 67.572.02 29.2370.28

12.1770.34 8.2870.36 6.0370.21

TP

Phenol Terephthalate Benzoate

4.9670.38 4.8570.13 8.4570.35

100 85.672.24 100

12.0571.20 10.1470.35 20.7672.74

Fig. 5 – T-RFLP fingerprints of 16S rRNA genes PCR-amplified from (a) mesophilic phenol-degrading batch culture, (b) mesophilic terephthalate-degrading batch culture, (c) mesophilic benzoate-degrading batch culture, (d) thermophilic phenol-degrading batch culture, (e) thermophilic terephthalate-degrading batch culture, and (f) thermophilic benzoatedegrading batch culture.

Table 2 shows that phenol as a positive control was completely degraded by the enrichments. Both MP and TP enrichments could degrade most of the terephthalate as the sole substrate (67.572.02% and 85.672.24%, respectively). Moreover, TP enrichment could completely degrade benzoate without a lag phase. In contrast, MP enrichment could only degrade 29.2370.28% of benzoate added. T-RFLP analysis revealed that 119-bp T-RF (79.2% of total fragment intensity) was the predominant T-RF observed in the phenol-enriching MP culture (Fig. 5a). Using terephthalate or benzoate as the sole carbon source, 119-bp peak (64.4%) or 469-bp peak (88.1%) was separately observed as the predominant T-RFs in the T-RFLP pattern obtained from each

cultures (Fig. 5b and c). In TP enrichment, 171-bp T-RF (50.878.2%) was the predominant T-RF in the phenolenriching culture associated with phenol, terephthalate and benzoate degradation (Fig. 5d–f). In addition, three to 11 detectable T-RFs were observed in MP and TP cultures. Cloning analysis (20 clones randomly selected from individual experiments) suggested that 119-bp, 171-bp and 469-bp T-RFs were represented by the OTUs or bacterial populations associated with Deltaproteobacteria group TA- (MP18), Pelotomaculum- (TP18), and Syntrophus aciditrophicus-(MP91) related populations, respectively. In addition, 189-bp T-RF and 252-bp T-RF were probably represented by the OTU TP16 in LGC and OTU TP29 in Thermotogae, respectively.

ARTICLE IN PRESS WAT E R R E S E A R C H

42 (2008) 1963 – 1976

1973

3.5. Localization and abundance of Deltaproteobacteria group TA in a mesophilic phenol-degrading full-scale reactor FISH results in Fig. 6 showed that the thickness of biofilms on GAC pellets was approximately 100–300 mm. The nonlayered structure was observed with the distribution of bacterial and archaeal (i.e., methanogens) populations (Fig. 6a). Both bacterial cells (green) and archaeal cells (red) were randomly distributed with approximately equal abundance in the thinsectioned biofilms. Yellow signals showed where bacterial and archaeal cells were closely associated. FISH with probes EUBmix and DETLA-TA664 further suggested that the members of the group TA (yellow) accounted for a majority (68.377.6%) of the total bacterial cells (Fig. 6b). Close-up FISH image revealed that the DETLA-TA664-postive cells had an oval-shaped morphotype similar to those observed in the MP enrichment.

4.

Discussion

Effective phenol-degrading mesophilic and thermophilic methanogenic consortia were successively obtained in this study. Due to temperature difference, microbial community structures in these two enrichments differed greatly, and displayed differences in physiological acclimation and degradation kinetics. Under mesophilic and thermophilic conditions, microbial consortia differing in community structures responsible for anaerobic degradation of substrates have also been reported in sucrose/propionate/acetate (Sekiguchi et al., 1998) and terephthalate (Chen et al., 2004; Wu et al., 2001) systems. In this study, archaeal OTUs found in the MP and TP were closely related to acetotrophic methanogens (Methanosaetaceae) and hydrogenotrophic methanogens (Methanobacteriaceae and Methanomicrobiales) (Fig. 2). Those OTUs related to acetotrophic methanogens were closely affiliated with M. concilii and M. thermoacetophila in MP and TP, respectively. It is reported that Methanosaeta spp. were the key population in facilitating sludge granulation (Fang et al., 1996; Zhang et al., 2005; Chang et al., 1995). The observation of predominant Methanosaeta-related populations in the phenol degradation suggests that these methanogens are also important in suspended-growth systems. The detection of Methanobacterium beijingense- and Methanothermobacter thermoautotrophicus-related species in the MP and TP enrichments support that these methanogens can play important roles in hydrogen consumption and in maintaining hydrogen partial pressure at a level low enough for phenol degradation. For the bacterial populations, the MP enrichment is predominated by members of the deltaproteobacterial group TA, which mostly consists of environmental clone sequences from various anoxic/anaerobic environments (Fig. 3a, b). Most members of the group TA exhibited degradation activities for different organics such as terephthalate (Wu et al., 2001), isophthalate (Qiu et al., 2004), trichlorobenzene (von Wintzingerode et al., 1999), 1,2-dichloropropane (Schlotelburg et al., 2000) and phenol (Fang et al., 2004, 2006; this study). Four OTUs (MP1, MP27, MP18 and MP29) found here were closely affiliated with the mesophilic isophthalate-degrading clone UI (Qiu et al., 2004) and terephthalate-degrading clone

Fig. 6 – FISH images of thin-sectioned biofilms obtained from a mesophilic full-scale phenol-degrading granules sample using a confocal laser scanning microscopy. (a) The sample was hybridized with Cy3-labeled domain Bacteriaspecific EUBmix probe (green) and Cy5-labeled domain Archaea-specific ARC915 probe (red). (b) The sample was hybridized with Cy3-labeled EUBmix probe (green) and Cy5labeled Detla-TA664 probe specific for Deltaproteobacteria group TA (red). Bar ¼ 100 lm. Inset shows a close-up view of cells targeted by Delta-TA664 (Bar ¼ 10 lm).

TA11 (Wu et al., 2001) (Fig. 3b) and differed from those reported by Fang et al. (2004, 2006) in ambient and thermophilic phenol-degrading communities. The results in this study suggested that Deltaproteobacteria group TA was also one of the most predominant bacterial populations in a mesophilic full-scale bioreactor treating phenol-containing wastewater. Like previous reports (Chang et al., 1995; Fang et al., 1996), no layered microstructure was observed with the

ARTICLE IN PRESS 1974

WAT E R R E S E A R C H

42 (2008) 1963– 1976

phenol-degrading granule, which is different from the layered structure observed with granules degrading easily degraded substrates (i.e., glucose, hydrolyzed proteins, sucrose, brewery wastes, and a mixture of sucrose/acetate/propionate) at high concentration (Fang et al., 1994; Guiot et al., 1992; MacLeod et al., 1990; Sekiguchi et al., 1999). This discrepancy might be due to effects of degradation kinetics and GAC absorption/desorption. In the TP consortium, the predominant bacterial group was related to Pelotomaculum spp. from the Desulfotomaculum subcluster Ih. This cluster contains a large number of closely related sequences from methanogenic environments (Imachi et al., 2006; Stubner and Meuser 2000) and a few bacterial isolates that can degrade phthalate isomers in co-culture with methanogen (i.e., Pelotomaculum terephthalicum and Pelotomaculum isophthalicum) (Qiu et al., 2004, 2006), or to transform phenol to benzoate anaerobically at 30–37 1C (i.e., Cryptanaerobacter phenolicus) (Juteau et al., 2005). The most predominant clone OTUs TP15 and TP18 found are closely affiliated with the predominant bacterial population observed in a thermophilic terephthalate-degrading consortium (Chen et al., 2004) (Fig. 3c), but different from those clones and isolates found in mesophilic phthalate isomers-degrading consortia (Qiu et al., 2004, 2006) and various methanogenic environments (Imachi et al., 2006). These observations suggested that Pelotomaculum-related populations can play an important role in the degradation of aromatic compounds such as phenol, terephthalate and benzoate under methanogenic conditions. Both MP and TP enrichments could further degrade terephthalate and benzoate as sole carbon sources without a lag phase. During terephthalate degradation, no noticeable change in the microbial community structures of MP and TP enrichments was observed. It is reported that a long start-up period from a few months to more than 1 year is required for mesophilic terephthalate-degrading methanogenic reactors (Kleerebezem, 1999; Razo-Flores et al., 2006). In this study, the MP enrichment can degrade 67.5% of terephthalate within 28 days. Based on this, we think that it is possible to use the seeding sludge from a phenol-degrading bioreactor for startup of another bioreactor treating terephthalate-containing wastewaters, or to use phenol as a co-substrate to enrich terephthalate-degrading microbial consortia, shortening the start-up time. During benzoate degradation with the MP enrichment, Syntrophus-like microorganisms were suggested to be the predominant populations instead of the deltaproteobacterial group TA by T-RFLP and cloning analyses. This also suggests that Syntrophus-like cells could have a higher substrate affinity to benzoate or a higher growth rate than the deltaproteobacterial group TA. It is reported that phenol is degraded via benzoate to hydrogen and acetate in a phenol-degrading consortium operated under ambient temperature (Fang et al., 2004). It is possible that this pathway is also used by the MP and TP phenol-degrading enrichments in this study as both can degrade benzoate. Fang and co-workers (Zhang et al., 2005; Fang et al., 2004) suggested that more than one bacterial population are involved in the transformation of phenol to benzoate (Desulfotomaculum and Clostridium), and benzoate to acetate and hydrogen (Syntrophus) in their granulated system.

In contrast, we could only observe one major bacterial population in the MP and TP enrichments using cloning, FISH and T-RFLP analyses. This difference in community structure remains to be further studied.

5.

Conclusions

Using polyphasic molecular biological tools, Deltaproteobacteria group TA- and Pelotomaculum-related bacterial populations as well as the acetotrophic and hydrogenotrophic methanogens were suggested to play an important role in the degradation of phenol at 37 and 55 1C conditions. More importantly, it was found that phenol- and terephthalatedegrading consortia at 37 and 55 1C shared the community structures. These findings greatly improve our understanding on the diversity of microbial populations involved in the degradation of aromatic compounds under methanogenic conditions, and can provide crucial microbial information to start up a full-scale phenol- or terephthalate-degrading reactor in the future. R E F E R E N C E S

Altschul, S.F., Madden, T.L., Schaffer, A.A., Zhang, J., Zhang, Z., Miller, W., Lipman, D.J., 1997. Gapped BLAST and PSI-BLAST: a new generation of protein database search programs. Nucleic Acids Res. 25 (17), 3389–3402. Amann, R.I., Ludwig, W., Schleifer, K.H., 1995. Phylogenetic identification and in situ detection of individual microbial cells without cultivation. Microbiol Rev. 59 (1), 143–169. Angelidaki, I., Petersen, S.P., Ahring, B.K., 1990. Effects of lipids on thermophilic anaerobic digestion and reduction of lipid inhibition upon addition of bentonite. Appl. Microbiol. Biotechnol. 33 (4), 469–472. Ashelford, K.E., Chuzhanova, N.A., Fry, J.C., Jones, A.J., Weightman, A.J., 2005. At least 1 in 20 16S rRNA sequence records currently held in public repositories is estimated to contain substantial anomalies. Appl. Environ. Microbiol. 71 (12), 7724–7736. Brauer, S.L., Cadillo-Quiroz, H., Yashiro, E., Yavitt, J.B., Zinder, S.H., 2006. Isolation of a novel acidiphilic methanogen from an acidic peat bog. Nature 442 (7099), 192–194. Chang, Y.J., Nishio, N., Nagai, S., 1995. Characteristics of granular methanogenic sludge grown on phenol synthetic medium and methanogenic fermentation of phenolic wastewater in a UASB reactor. J. Fermentation Bioeng. 79 (4), 348–353. Chen, C.L., Macarie, H., Ramirez, I., Olmos, A., Ong, S.L., Monroy, O., Liu, W.T., 2004. Microbial community structure in a thermophilic anaerobic hybrid reactor degrading terephthalate. Microbiology 150 (Pt 10), 3429–3440. Chouari, R., Le Paslier, D., Daegelen, P., Ginestet, P., Weissenbach, J., Sghir, A., 2005. Novel predominant archaeal and bacterial groups revealed by molecular analysis of an anaerobic sludge digester. Environ. Microbiol. 7 (8), 1104–1115. Cole, J.R., Chai, B., Farris, R.J., Wang, Q., Kulam, S.A., McGarrell, D.M., Garrity, G.M., Tiedje, J.M., 2005. The Ribosomal Database Project (RDP-II): sequences and tools for high-throughput rRNA analysis. Nucleic Acids Res. 33 (Database issue), D294–D296. Daims, H., Bruhl, A., Amann, R., Schleifer, K.H., Wagner, M., 1999. The domain-specific probe EUB338 is insufficient for the detection of all Bacteria: development and evaluation of a

ARTICLE IN PRESS WAT E R R E S E A R C H

42 (2008) 1963 – 1976

more comprehensive probe set. Syst. Appl. Microbiol. 22 (3), 434–444. Embley, T.M., Finlay, B.J., Thomas, R.H., Dyal, P.L., 1992. The use of rRNA sequences and fluorescent probes to investigate the phylogenetic positions of the anaerobic ciliate Metopus palaeformis and its archaeobacterial endosymbiont. J. Gen. Microbiol. 138 (7), 1479–1487. Evans, W.C., 1977. Biochemistry of the bacterial catabolism of aromatic compounds in anaerobic environments. Nature 270 (5632), 17–22. Fang, H.H.P., Chui, H.K., Li, Y.Y., Chen, T., 1994. Performance and granule characteristics of UASB process treating wastewater with hydrolyzed proteins. Water Sci. Technol. 30 (8), 55–63. Fang, H.H.P., Chen, T., Li, Y.Y., Chui, H.K., 1996. Degradation of phenol in wastewater in an upflow anaerobic sludge blanket reactor. Water Res. 30 (6), 1353–1360. Fang, H.H.P., Liu, Y., Ke, S.Z., Zhang, T., 2004. Anaerobic degradation of phenol in wastewater at ambient temperature. Water Sci. Technol. 49 (1), 95–102. Fang, H.H., Liang, D.W., Zhang, T., Liu, Y., 2006. Anaerobic treatment of phenol in wastewater under thermophilic condition. Water Res. 40 (3), 427–434. Ferris, M.J., Muyzer, G., Ward, D.M., 1996. Denaturing gradient gel electrophoresis profiles of 16S rRNA-defined populations inhabiting a hot spring microbial mat community. Appl. Environ. Microbiol. 62 (2), 340–346. Gallert, C., Knoll, G., Winter, J., 1991. Anaerobic carboxylation of phenol to benzoate: use of deuterated phenols revealed carboxylation exclusively in the C4-position. Appl. Microbiol. Biotechnol. 36, 124–129. Guiot, S.R., Pauss, A., Costerton, J.W., 1992. A structured model of the anaerobic granule consortium. Water Sci. Technol. 25, 1–10. Harwood, C.S., Burchhardt, G., Herrmann, H., Fuchs, G., 1999. Anaerobic metabolism of aromatic compounds via the benzoyl-CoA pathway. FEMS Microbiol. Rev. 22, 439–458. Heuer, H., Krsek, M., Baker, P., Smalla, K., Wellington, E.M., 1997. Analysis of actinomycete communities by specific amplification of genes encoding 16S rRNA and gel-electrophoretic separation in denaturing gradients. Appl. Environ. Microbiol. 63 (8), 3233–3241. Huber, T., Faulkner, G., Hugenholtz, P., 2004. Bellerophon: a program to detect chimeric sequences in multiple sequence alignments. Bioinformatics 20 (14), 2317–2319. Imachi, H., Sekiguchi, Y., Kamagata, Y., Loy, A., Qiu, Y.L., Hugenholtz, P., Kimura, N., Wagner, M., Ohashi, A., Harada, H., 2006. Non-sulfate-reducing, syntrophic bacteria affiliated with desulfotomaculum cluster I are widely distributed in methanogenic environments. Appl. Environ. Microbiol. 72 (3), 2080–2091. Juteau, P., Cote, V., Duckett, M.F., Beaudet, R., Lepine, F., Villemur, R., Bisaillon, J.G., 2005. Cryptanaerobacter phenolicus gen. nov., sp. nov., an anaerobe that transforms phenol into benzoate via 4-hydroxybenzoate. Int. J. Syst. Evol. Microbiol. 55 (Pt 1), 245–250. Kleerebezem, R., 1999. Anaerobic treatment of phthalates: microbiological and technological aspects. Ph.D. Thesis, Wageningen University, Wageningen, The Netherlands. Kleerebezem, R., Macarie, H., 2003. Treating industrial wastewater: anaerobic digestion comes of age. Chemical Engineering 110 (4), 56–64. Kobayashi, T., Hashinaga, T., Mikami, E., Suzuki, T., 1989. Methanogenic degradation of phenol and benzoate in acclimated sludge. Water Sci. Technol. 21, 55–65. Kumar, S., Tamura, K., Nei, M., 2004. MEGA3: Integrated software for molecular evolutionary genetics analysis and sequence alignment. Brief Bioinform. 5 (2), 150–163.

1975

Liu, W.T., Marsh, T.L., Cheng, H., Forney, L.J., 1997. Characterization of microbial diversity by determining terminal restriction fragment length polymorphisms of genes encoding 16S rRNA. Appl. Environ. Microbiol. 63 (11), 4516–4522. Liu, W.-T., Huang, C.L., Hu, J.Y., Song, L.F., Ong, S.L., Ng, W.J., 2002. Denaturing gradient gel electrophoresis polymorphism for rapid 16S rDNA clone screening and microbial diversity estimation. J. Biosci. Bioeng. 93 (101–03). Ludwig, W., Strunk, O., Westram, R., Richter, L., Meier, H., Yadhukumar, Buchner, A., Lai, T., Steppi, S., Jobb, G., Forster, W., Brettske, I., Gerber, S., Ginhart, A.W., Gross, O., Grumann, S., Hermann, S., Jost, R., Konig, A., Liss, T., Lussmann, R., May, M., Nonhoff, B., Reichel, B., Strehlow, R., Stamatakis, A., Stuckmann, N., Vilbig, A., Lenke, M., Ludwig, T., Bode, A., Schleifer, K.H., 2004. ARB: a software environment for sequence data. Nucleic Acids Res. 32 (4), 1363–1371. Macarie, H., 2000. Overview of the application of anaerobic treatment to chemical and petrochemical wastewaters. Water Sci. Technol. 42 (5–6), 201–214. MacLeod, F.A., Guiot, S.R., Costerton, J.W., 1990. Layered structure of bacterial aggregates produced in an upflow anaerobic sludge bed and filter reactor. Appl. Environ. Microbiol. 56 (6), 1598–1607. Qiu, Y.L., Sekiguchi, Y., Imachi, H., Kamagata, Y., Tseng, I.C., Cheng, S.S., Ohashi, A., Harada, H., 2004. Identification and isolation of anaerobic, syntrophic phthalate isomer-degrading microbes from methanogenic sludges treating wastewater from terephthalate manufacturing. Appl. Environ. Microbiol. 70 (3), 1617–1626. Qiu, Y.L., Sekiguchi, Y., Hanada, S., Imachi, H., Tseng, I.C., Cheng, S.S., Ohashi, A., Harada, H., Kamagata, Y., 2006. Pelotomaculum terephthalicum sp. nov. and Pelotomaculum isophthalicum sp. nov.: two anaerobic bacteria that degrade phthalate isomers in syntrophic association with hydrogenotrophic methanogens. Arch. Microbiol. 185 (3), 172–182. Razo-Flores, E., Macarie, H., Morier, F., 2006. In: Cervantes, F.J., Pavlostathis, S.G., van Haandel, A.C. (Eds.), Advanced Biological Treatment Processes for Industrial Wastewaters. IWA Publishing, London, UK, pp. 267–297. Schlotelburg, C., von Wintzingerode, F., Hauck, R., Hegemann, W., Gobel, U.B., 2000. Bacteria of an anaerobic 1,2-dichloropropane-dechlorinating mixed culture are phylogenetically related to those of other anaerobic dechlorinating consortia. Int. J. Syst. Evol. Microbiol. 50 (Pt 4), 1505–1511. Sekiguchi, Y., Kamagata, Y., Syutsubo, K., Ohashi, A., Harada, H., Nakamura, K., 1998. Phylogenetic diversity of mesophilic and thermophilic granular sludges determined by 16S rRNA gene analysis. Microbiology 144 (Pt 9), 2655–2665. Sekiguchi, Y., Kamagata, Y., Nakamura, K., Ohashi, A., Harada, H., 1999. Fluorescence in situ hybridization using 16S rRNAtargeted oligonucleotides reveals localization of methanogens and selected uncultured bacteria in mesophilic and thermophilic sludge granules. Appl. Environ. Microbiol. 65 (3), 1280–1288. Stahl, D.A., Amann, R., 1991. In: Stackebrandt, E., Goodfellow, M. (Eds.), Nucleic Acid Techniques in Bacterial Systematics. Wiley, New York, pp. 205–248. Stubner, S., Meuser, K., 2000. Detection of Desulfotomaculum in an Italian rice paddy soil by 16S ribosomal nucleic acid analyses. FEMS Microbiol. Ecol. 34 (1), 73–80. Thompson, J.R., Marcelino, L.A., Polz, M.F., 2002. Heteroduplexes in mixed-template amplifications: formation, consequence and elimination by ‘reconditioning PCR’. Nucleic Acids Res. 30 (9), 2083–2088. Veeresh, G.S., Kumar, P., Mehrotra, I., 2005. Treatment of phenol and cresols in upflow anaerobic sludge blanket (UASB) process: a review. Water Res. 39 (1), 154–170.

ARTICLE IN PRESS 1976

WAT E R R E S E A R C H

42 (2008) 1963– 1976

von Wintzingerode, F., Selent, B., Hegemann, W., Gobel, U.B., 1999. Phylogenetic analysis of an anaerobic, trichlorobenzene-transforming microbial consortium. Appl. Environ. Microbiol. 65 (1), 283–286. Weisburg, W.G., Barns, S.M., Pelletier, D.A., Lane, D.J., 1991. 16S ribosomal DNA amplification for phylogenetic study. J. Bacteriol. 173 (2), 697–703. Wu, J.H., Liu, W.T., Tseng, I.C., Cheng, S.S., 2001. Characterization of microbial consortia in a terephthalate-degrading

anaerobic granular sludge system. Microbiology 147 (Pt 2), 373–382. Young, L.Y., Rivera, M.D., 1985. Methanogenic degradation of four phenolic componds. Water Res. 19 (10), 1325–1332. Zhang, T., Ke, S.Z., Liu, Y., Fang, H.P., 2005. Microbial characteristics of a methanogenic phenol-degrading sludge. Water Sci. Technol. 52 (1–2), 73–78.