Journal of Applied Microbiology ISSN 1364-5072
ORIGINAL ARTICLE
Polyhydroxyalkanoate (PHA) accumulation potential and PHA-accumulating microbial communities in various activated sludge processes of municipal wastewater treatment plants K. Sakai1, S. Miyake1, K. Iwama1, D. Inoue2, S. Soda1 and M. Ike1 1 Division of Sustainable Energy and Environmental Engineering, Osaka University, Suita, Osaka, Japan 2 Department of Health Science, Kitasato University, Sagamihara-Minami, Kanagawa, Japan
Keywords activated sludge, microbial community analysis, phaC gene, polyhydroxyalkanoateaccumulating organism, polyhydroxyalkanoate accumulation potential. Correspondence Michihiko Ike, Division of Sustainable Energy and Environmental Engineering, Osaka University, 2-1 Yamadaoka, Suita, Osaka 565-0871, Japan. E-mail:
[email protected] 2014/1364: received 4 July 2014, revised 26 October 2014 and accepted 26 October 2014 doi:10.1111/jam.12683
Abstract Aims: To clarify the polyhydroxyalkanoate (PHA) accumulation potential and the PHA-accumulating microbial community structure in activated sludge in municipal wastewater treatment plants (WWTPs) and to identify their influential factors. Methods and Results: Nine activated sludge samples were collected from municipal WWTPs employing various biological treatment processes. In acetate-fed 24-h batch experiments under aerobic and nitrogen- and phosphorus-limited conditions, polyhydroxybutyrate (PHB) content of activated sludge increased from 0–13 wt% to 79–24 wt%, with PHB yields of 022–050 C-mol 3-hydroxybutyrate (C-mol acetate)−1. Microbial community analyses found that activated sludge samples that accumulated >20 wt% of PHB after 24-h PHA accumulation experiments had >50 9 108 copies g1mixed liquor-suspended solid of phaC genes. Conclusions: Results indicated that (i) activated sludge in municipal WWTPs can accumulate up to approx. 20 wt% of PHA without enrichment processes, (ii) PHA accumulation potential of activated sludge varied depending on the operational conditions (treatment processes) of WWTPs, and (iii) phaC gene number can provide a simple indication of PHA accumulation potential. Significance and Impact of the Study: This is the first study to compare the PHA accumulation potential and PHA-accumulating microbial communities in activated sludge of various treatment processes. Our findings may be useful for enhancing the resource recovery potential of wastewater treatment systems.
Introduction Polyhydroxyalkanoates (PHAs) are polyesters of hydroxyl fatty acids that are stored by a variety of bacteria as intracellular carbon and energy reserve materials. To date, it has been reported that more than 300 different species of bacteria have an ability to synthesize PHAs (Dias et al. 2006). It is also well known that activated sludge, which consists of micro-organisms that vary phylogenetically and functionally and has been used for municipal and industrial wastewater treatment, is capable of producing PHAs (Satoh et al. 1999). PHAs are a
group of bioplastics that exhibit biodegradable, biocompatible and thermoplastic properties and are recognized as having a high resource value as an environmentally compatible material (Steinb€ uchel 1992; Sudesh et al. 2000). A recent study reported that methane production by anaerobic digestion using excess sludge as the substrate could be improved using sludge in which PHA was accumulated (Huda et al. 2013). Therefore, PHA accumulation by activated sludge can improve the value of activated sludge as a bioresource and may be very useful in upgrading the resource recovery potential of wastewater treatment systems.
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Recent studies have demonstrated the construction of enrichment cultures that can accumulate up to 90 wt% (per cell dry weight) PHA using activated sludge as the inoculum (Johnson et al. 2009; Jiang et al. 2011a). Furthermore, Plasticicumulans acidivorans TUD-YJ37T gen. nov., sp. nov., a novel bacterium with a very high PHA accumulation ability, has been isolated from an acetatefed enrichment culture that was inoculated with activated sludge (Jiang et al. 2011b). Together, these findings suggest the presence of bacteria with a high PHA accumulation potential in activated sludge. Research on the characteristics that provide activated sludge with a high PHA accumulation potential and on the identification of PHA-accumulating microbial populations can provide information on ways to increase the PHA content in activated sludge. This information may then be applied to research and development efforts to convert activated sludge into a valuable resource. However, while many studies have succeeded in the enrichment of highly effective PHA-accumulating populations from activated sludge in municipal wastewater treatment plants (WWTPs) under the specific conditions of laboratory-scale or bench-scale reactors (e.g. Serafim et al. 2004; Johnson et al. 2009; Albuquerque et al. 2010; Jiang et al. 2011a; Moralejo-Garate et al. 2011), there is little knowledge available on the inherent PHA accumulation potential of activated sludge without experiencing such selective enrichment process (Takabatake et al. 2002; Oshiki et al. 2013). Furthermore, only a few studies have investigated the abundance and taxonomic compositions of PHA-accumulating bacteria in activated sludge using culture-dependent (Vishnuvardhan Reddy et al. 2008) and culture-independent molecular techniques (Oshiki et al. 2008, 2013). To obtain generalizable knowledge on PHA accumulation by activated sludge, it is of great importance to compare PHA accumulation potential and PHAaccumulating microbial populations or their structure in different activated sludges and to explore possible links between them. Very recently, Oshiki et al. (2013) investigated the PHA accumulation potentials and analysed the PHA-accumulating microbial communities in activated sludge samples; however, they focused on activated sludge of the fully aerobic process. Similar studies have not been conducted with activated sludges acclimated with different wastewater treatment processes in municipal WWTPs, in spite that the treatment process is an important factor affecting the activated sludge microbial community (Victorio et al. 1996; Matsuda et al. 2010) and thus may influence on the PHA accumulation potential of activated sludge. In this study, we aimed to clarify the PHA accumulation potential and the PHA-accumulating microbial community structure in activated sludge in municipal 256
WWTPs and to identify their influential factors. Activated sludge samples were collected from municipal WWTPs employing various biological treatment processes. PHA accumulation potential of the samples was evaluated by PHA accumulation experiments under nitrogen limited and aerobic conditions and using acetate as the substrate. In addition, terminal restriction fragment length polymorphism (T-RFLP) analysis and most probable numberpolymerase chain reaction (MPN-PCR) targeting the phaC genes, which encode PHA synthase, were carried out to analyse the PHA-accumulating communities in the activated sludge samples, in addition to those targeting eubacterial 16S rRNA genes for the analysis of total microbial community. Materials and methods Activated sludge samples Nine activated sludge samples were collected from the outlet of final biological treatment tanks in six full-scale municipal WWTPs in Osaka, Japan (Table 1). The WWTPs applied one or more of the following processes for secondary biological treatment: activated sludge (AS), anaerobic–oxic (AO), anaerobic–anoxic–oxic (A2O), step-feed multistage denitrification–nitrification and membrane bioreactor (MBR) processes. All of the AS processes investigated in this study were operated as reduced aeration–high aeration sequential processes in which the aeration is partially reduced in the former part of the aeration tank(s) for enhanced biological phosphate removal (EBPR). Because multistage denitrification–nitrification process corresponds to a biological nutrient removal process, step-feed multistage denitrification– nitrification process is referred to as Step-feed BNR in this study. All samples were transported to the laboratory on ice and stored at 4°C until further processing. PHA accumulation experiments PHA accumulation potential of activated sludge samples was evaluated by batch experiments under nitrogen limited and aerobic conditions, using acetate as the substrate. Activated sludge samples were harvested by centrifugation (8500 g, 4°C, 10 min) and inoculated to a mixed liquor-suspended solid (MLSS) concentration of 1000 mg l−1 in a 500-ml Erlenmeyer flask containing 200 ml of nitrogen-free basal salt medium (MgSO47H2O 6840 mg l−1, KCl 2684 mg l−1, allylthiourea 5 mg l−1 and trace elements solution 15 ml l−1 (Vishniac and Santer 1957)) to which sodium acetate was added at 1000 mg-C l−1 as the sole carbon source (system A). As the presence of phosphate and the fluctuation of pH have
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Table 1 Operational conditions of WWTPs and activated sludge samples analysed in this study
Plant
Process
Sample name
A B C D E
AO AS* AO AS* A2O MBR AS* A2O Step-feed BNR
A-AO B-AS C-AO D-AS D-A2O E-MBR F-AS F-A2O F-Step-feed BNR
F
SRT (d)
HRT (h)
F/M ratio§(g-BOD (g-MLSS d)–1)
Temp. (°C)
pH
MLSS (mg l–1)
63† 60† 77† 50† 67† 148†, 28‡ 32† 13† 22†
74† 74† 75† 84† 70† 78†, 69‡ 61† 130† 130†
023 020 022 021 023 003 082 011 008
259 256 256 244 247 300 265 257 256
63 72 64 69 71 71 62 62 62
1640 1720 1380 880 1640 8860 430 1270 1900
Sampling date May. 29, 2012 Jun. 11, 2012 Jun. 11, 2012 Jun. 26, 2012 Aug. 21, 2012 Jul. 10, 2012
AO, anaerobic–oxic; A2O, anaerobic–anoxic–oxic; AS, activated sludge; BOD, biochemical oxygen demand; F/M ratio, food to micro-organisms ratio; HRT, hydraulic retention time; MBR, membrane bioreactor; MLSS, mixed liquor-suspended solids; SRT, sludge retention time; Step-feed BNR, step-feed multistage denitrification–nitrification. *All of the AS processes investigated in this study were operated by a reduced aeration–high aeration sequential process for enhanced biological phosphate removal. †Monthly average value. ‡Value obtained on a day close to the sampling. §F/M ratio was expressed as BOD-MLSS loading.
been reported to affect PHA accumulation (Sudesh et al. 2000; Serafim et al. 2004), the following two test systems were also prepared: test system B, in which KH2PO4 was added at 3856 mg-P l−1, and test system C, in which HEPES and KH2PO4 were added at 20 mmol l−1 and 3856 mg-P l−1, respectively. Test systems A, B and C were prepared in duplicate for each sludge sample. In addition, a control test system that did not contain sludge was prepared. All test systems were incubated at 28°C on a rotary shaker at 120 rev min1 for 24 h. The pH of media and the concentrations of MLSS, acetate and PHA were monitored periodically. Analytical methods The MLSS concentrations of sludge samples were measured according to standard methods (American Public Health Association 1998). Acetate concentrations in the supernatant of sludge samples were determined by high-performance liquid chromatography (HPLC). Prior to HPLC analysis, the supernatants were filtered through a cellulose acetate filter (pore size 045 lm; Advantec, Tokyo, Japan). HPLC analysis was conducted according to the methods developed by Mishima et al. (2008) using a Shimadzu LC-10Avp system (Shimadzu, Kyoto, Japan) with an RID10A refractive index detector and an Aminex HPX-87H column (300 9 78 mm; Bio-Rad, Hercules, CA). Polyhydroxybutyrate (PHB), which is a polymer of 3-hydroxybutyrate (3HB), in sludge samples was determined as the representative of PHA because it has been reported that the majority of PHA was composed of PHB
when acetate was used as the sole substrate (Lemos et al. 2006). Prior to the analysis of PHB, sludge samples were pretreated according to the methods developed by Michinaka et al. (2007) with the following minor modifications: acidified methanol (20% (v/v) with sulphuric acid) was used for methanolysis, the reaction time for methanolysis was shortened to 7 h, and aqueous ammonia (28% (v/v)) was used for neutralization. Benzoic acid was used as the internal standard. Methyl-3-hydroxybutyrate, which is a derivative of PHB through the aforementioned methanolytic decomposition, was analysed by gas chromatography–mass spectrometry (GC-MS) using Varian 450-GC equipped with Varian 220-MS (Varian, Palo Alto, CA). Chromatographic separation was performed with an InertCap WAX-HT capillary column (30 m 9 025 mm ID, 025 lm df; GL Sciences, Tokyo, Japan). Helium (99998%) was used as the carrier gas at a flow rate of 10 ml min−1. The injector temperature was set at 230°C. Sample was injected with a split ratio of 1 : 20. The column temperature was held at 80°C for 4 min, increased to 200°C at a rate of 12°C min−1, then held at 200°C for 6 min. GC/MS analysis was operated in scan mode (m/z 40–300). The compounds were confirmed by comparing their retention times and mass spectra with the methyl ester derivatives of authentic standards and quantified by the internal standard method. Microbial community analysis Bacterial DNA was extracted with ISOIL for Beads Beating (Nippon Gene, Tokyo, Japan), then purified with
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MagExtractor-PCR&Gel Clean up (Toyobo, Tokyo, Japan) according to the manufacturers’ instructions. T-RFLP analysis of eubacterial 16S rRNA genes was performed according to the methods of Matsuda et al. (2010) with minor modifications. Briefly, the conserved region of eubacterial 16S rRNA genes was PCR-amplified using the 27F (50 -AGA GTT TGA TCC TGG CTC AG-30 ), the 50 -end of which was labelled with phosphoramidite fluorochrome 5-carboxyfluorescein (6-FAM), and 1392R (50 -ACG GGC GGT GTG TAC A-30 ) primer set (Amann et al. 1995). The PCR amplification was conducted with the following thermal profile: initial denaturation at 95°C for 10 min; 18–20 cycles of denaturation at 95°C for 1 min, annealing at 57°C for 1 min and extension at 72°C for 3 min; and final extension at 72°C for 10 min. The cycle number for each sample was determined within the early exponential amplification phase. The PCR products (around 1300 bp) were purified with NucleoSpin Extract II (Macherey-Nagel, D€ uren, Germany) according to the manufacturer’s instructions and digested with HhaI at 37°C for 5 h. Fluorescently labelled terminal restriction fragments (T-RFs) were separated and detected using an ABI Prism 310 genetic analyzer (Applied Biosystems, Foster City, CA), and their size and abundance were determined with GeneScan ver. 3.7 (Applied Biosystems). For T-RFLP analysis of phaC genes encoding the PHA synthase, those of Classes I and II were PCR-amplified using the CF1 (50 -ATC AAC AAR TWC TAC RTC YTS GAC CT-30 ), the 50 -end of which was labelled with 6-FAM, and CR4 (50 -AGG TAG TTG TYG ACS MMR TAG KTC CA-30 ) primer set (Sheu et al. 2000; Michinaka et al. 2007). The PCR amplification was conducted with the following thermal profile: initial denaturation at 94°C for 5 min; 37–39 cycles of denaturation at 94°C for 30 s, annealing at 57°C for 45 s and extension at 72°C for 1 min; and final extension at 72°C for 10 min. The cycle number for each sample was determined within the early exponential amplification phase. The PCR products (around 500 bp) were purified as described above and digested separately with MboI and AccII at 37°C for 5 h. Fluorescently labelled T-RFs were analysed in the same manner as above. For in silico analysis of T-RFLP, the phaC gene sequences of taxonomically identified bacteria were obtained from the DNA sequence database of the National Center for Biotechnology Information on June 30, 2011, and a database for the phaC genes was constructed using a TRiFLe package (Junier et al. 2008). The copy numbers of eubacterial 16S rRNA genes were determined by MPN-PCR as described previously (Sang et al. 2008), while those of phaC genes were determined by MPN-PCR using the CF1 and CR4 primer set with 258
the aforementioned thermal profile, excepting that the thermal cycling number was fixed to 35 cycles. Numerical analysis Microbial community diversity was evaluated using the Shanon–Weaver index (H0 ) and Shannon effective num0 0 ber of species (eH ). H0 and eH values were calculated from the T-RFLP profile by the following equation: H0 ¼
X
ðPi ln PiÞ
0
eH ¼ expðH 0 Þ where Pi is the relative abundance of the targeted T-RF (Blackwood et al. 2007). To compare the bacterial community composition of different samples, cluster analysis and principal component analysis (PCA) were conducted based on the Pi matrix obtained from T-RFLP analysis using PAST ver. 1.3.4 (http://folk.uio.no/ohammer/past/). In the cluster analysis, a dendrogram was constructed from Dice’s coefficient (SD) using the unweighted pair group method with arithmetic averages (UPGMA). To identify the factors affecting PHA accumulation potential and microbial communities, single linear regression analysis was conducted between the indicators associated with the PHA accumulation potential (acetate consumption and PHB yield) and the Pi values or operating conditions of the WWTPs. Regression analysis was conducted using PAST ver. 1.3.4. Correlation coefficient (r) and significance probability (P) were calculated, and correlations were considered statistically significant at P < 005. Results PHA accumulation potential PHA accumulation experiments were conducted with three different test systems to evaluate the PHA accumulation potential of activated sludge samples. Time courses of acetate consumption, pH increase and PHB accumulation in the three test systems using sample B-AS are shown in Fig. 1. In the control system without sludge inoculation, the acetate concentration did not decline, and the pH remained constant throughout the experimental period (data not shown). In test systems A-C with sludge inoculation, the acetate concentration decreased, and the pH and PHB content increased over time. The consumption of acetate and the increase in pH and PHB content were slowest in test system C, in which HEPES and phosphate salt were added, and fastest in test system
Journal of Applied Microbiology 118, 255--266 © 2014 The Society for Applied Microbiology
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MagExtractor-PCR&Gel Clean up (Toyobo, Tokyo, Japan) according to the manufacturers’ instructions. T-RFLP analysis of eubacterial 16S rRNA genes was performed according to the methods of Matsuda et al. (2010) with minor modifications. Briefly, the conserved region of eubacterial 16S rRNA genes was PCR-amplified using the 27F (50 -AGA GTT TGA TCC TGG CTC AG-30 ), the 50 -end of which was labelled with phosphoramidite fluorochrome 5-carboxyfluorescein (6-FAM), and 1392R (50 -ACG GGC GGT GTG TAC A-30 ) primer set (Amann et al. 1995). The PCR amplification was conducted with the following thermal profile: initial denaturation at 95°C for 10 min; 18–20 cycles of denaturation at 95°C for 1 min, annealing at 57°C for 1 min and extension at 72°C for 3 min; and final extension at 72°C for 10 min. The cycle number for each sample was determined within the early exponential amplification phase. The PCR products (around 1300 bp) were purified with NucleoSpin Extract II (Macherey-Nagel, D€ uren, Germany) according to the manufacturer’s instructions and digested with HhaI at 37°C for 5 h. Fluorescently labelled terminal restriction fragments (T-RFs) were separated and detected using an ABI Prism 310 genetic analyzer (Applied Biosystems, Foster City, CA), and their size and abundance were determined with GeneScan ver. 3.7 (Applied Biosystems). For T-RFLP analysis of phaC genes encoding the PHA synthase, those of Classes I and II were PCR-amplified using the CF1 (50 -ATC AAC AAR TWC TAC RTC YTS GAC CT-30 ), the 50 -end of which was labelled with 6-FAM, and CR4 (50 -AGG TAG TTG TYG ACS MMR TAG KTC CA-30 ) primer set (Sheu et al. 2000; Michinaka et al. 2007). The PCR amplification was conducted with the following thermal profile: initial denaturation at 94°C for 5 min; 37–39 cycles of denaturation at 94°C for 30 s, annealing at 57°C for 45 s and extension at 72°C for 1 min; and final extension at 72°C for 10 min. The cycle number for each sample was determined within the early exponential amplification phase. The PCR products (around 500 bp) were purified as described above and digested separately with MboI and AccII at 37°C for 5 h. Fluorescently labelled T-RFs were analysed in the same manner as above. For in silico analysis of T-RFLP, the phaC gene sequences of taxonomically identified bacteria were obtained from the DNA sequence database of the National Center for Biotechnology Information on June 30, 2011, and a database for the phaC genes was constructed using a TRiFLe package (Junier et al. 2008). The copy numbers of eubacterial 16S rRNA genes were determined by MPN-PCR as described previously (Sang et al. 2008), while those of phaC genes were determined by MPN-PCR using the CF1 and CR4 primer set with 258
the aforementioned thermal profile, excepting that the thermal cycling number was fixed to 35 cycles. Numerical analysis Microbial community diversity was evaluated using the Shanon–Weaver index (H0 ) and Shannon effective num0 0 ber of species (eH ). H0 and eH values were calculated from the T-RFLP profile by the following equation: H0 ¼
X
ðPi ln PiÞ
0
eH ¼ expðH 0 Þ where Pi is the relative abundance of the targeted T-RF (Blackwood et al. 2007). To compare the bacterial community composition of different samples, cluster analysis and principal component analysis (PCA) were conducted based on the Pi matrix obtained from T-RFLP analysis using PAST ver. 1.3.4 (http://folk.uio.no/ohammer/past/). In the cluster analysis, a dendrogram was constructed from Dice’s coefficient (SD) using the unweighted pair group method with arithmetic averages (UPGMA). To identify the factors affecting PHA accumulation potential and microbial communities, single linear regression analysis was conducted between the indicators associated with the PHA accumulation potential (acetate consumption and PHB yield) and the Pi values or operating conditions of the WWTPs. Regression analysis was conducted using PAST ver. 1.3.4. Correlation coefficient (r) and significance probability (P) were calculated, and correlations were considered statistically significant at P < 005. Results PHA accumulation potential PHA accumulation experiments were conducted with three different test systems to evaluate the PHA accumulation potential of activated sludge samples. Time courses of acetate consumption, pH increase and PHB accumulation in the three test systems using sample B-AS are shown in Fig. 1. In the control system without sludge inoculation, the acetate concentration did not decline, and the pH remained constant throughout the experimental period (data not shown). In test systems A-C with sludge inoculation, the acetate concentration decreased, and the pH and PHB content increased over time. The consumption of acetate and the increase in pH and PHB content were slowest in test system C, in which HEPES and phosphate salt were added, and fastest in test system
Journal of Applied Microbiology 118, 255--266 © 2014 The Society for Applied Microbiology
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activated sludge samples. Figure 3 shows the T-RFLP profiles of eubacterial 16S rRNA genes. Nineteen to thirty-four T-RFs were detected in the samples, and H0 0 and eH values were calculated to be 27–34 and 15–30, respectively (Table 2). T-RFs of 59–60 and 203–204 bp were predominantly detected in the samples taken from the treatment processes that include a final settling tank (all samples except for sample E-MBR), which is consistent with the results of a previous study (Matsuda et al. 2010). In sample E-MBR, the T-RFs of 84–85 and 405 bp were predominantly detected, and the whole community structure was quite different from that of the other samples. In PCA of the T-RFLP profiles, the first, second and third PCs (PC1, PC2 and PC3, respectively) accounted for 413, 223, and 116%, respectively, of the total variation. Although no clear grouping was found in a scatter diagram with PC1 and PC2 values, another scatter diagram with PC1 and PC3 values showed that the nine samples fell into three distinct groups (Fig. 4): group A consisted mainly of AS and AO, group B consisted of
0
100
200
300
400
500
600
700
800
A2O and Step-feed BNR, and group C consisted of MBR. The copy numbers of phaC genes in the investigated MPN-copies samples were 29 9 108–12 9 109 −1 (g-MLSS) and 20 wt% in the samples with more than 50 9 108 copies (g-MLSS)−1 of phaC genes. 262
500 bp
5B : A11_Sjk CAS AccII 37c.fsa/
11B : A5_Mhm CAS Mbol 37c.fsa/
0·4
400
A-AO
A-AO 8B : A3_Sjk CAS Mbol 37c.fsa/
(a)
300
–0·5 0 PC1 (56·6%)
0· 5
Figure 6 Dendrogram showing clustering of the T-RFLP profiles of phaC genes using Dice’s coefficient as a similarity index (a) and ordination produced from principal component analysis based on the same T-RFLP profiles (b). Symbol: filled square, A-AO; filled diamond, B-AS; open square, C-AO; grey diamond, D-AS; filled triangle, D-A2O; filled circle, E-MBR; open diamond, F-AS; open triangle, F-A2O; cross, F-Step-feed BNR.
Discussion Very low PHB contents of activated sludge samples investigated in this study prior to the PHA accumulation experiment (0–13 wt%; Fig. 2) suggested that PHB was not significantly accumulated in activated sludge of municipal WWTPs. By contrast, PHB contents in the activated sludge samples after a 24-h PHA accumulation experiment (79–24 wt%) were similar to those observed in a previous study (60–295 wt%) in which PHA accumulation by activated sludge samples taken from AS, AO and A2O was examined under continuous acetate amendment (Takabatake et al. 2002). These findings indicate that activated sludge of municipal WWTPs has the potential to accumulate PHA up to approx. 20 wt% by single batch cultivation on acetate without enrichment of PHA-accumulating organisms.
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30
25
PHB (wt%)
20
15
10
5
0 0·0E+00
5·0E+08
1·0E+09
1·5E+09
Number of phaC gene (MPN-copies (g-MLSS)–1) Figure 7 Relationships between the copy number of phaC genes and PHB content after 24-h PHA accumulation experiment. Dashed line represents 50 9 108 MPN-copies (g-MLSS)1. Symbol: filled square, A-AO; filled diamond, B-AS; open square, C-AO; grey diamond, DAS; filled triangle, D-A2O; filled circle, E-MBR; open diamond, F-AS; open triangle, F-A2O; cross, F-Step-feed BNR.
Comparison of the PHA accumulation potentials of activated sludge samples collected from various processes revealed that after 24-h PHA accumulation experiments, PHB contents increased by more than 20 wt% and reached 20–24 wt% in AS, AO and MBR sludge samples, while increases were lower in A2O and Step-feed BNR samples. As mentioned above, all of the AS processes investigated were operated as a reduced aeration–high aeration sequential process for EBPR. Therefore, polyphosphate-accumulating organisms (PAOs), which play an important role in EBPR, and glycogen-accumulating organisms (GAOs), which compete with PAOs, might predominate the AO and AS processes investigated although their presence could not be confirmed in this study and needs further study to be clarified. PAOs take up organic substrates and accumulate them as PHA by consuming intracellular polyphosphate as the energy source in the anaerobic tank. In the subsequent aerobic tank, they grow by consuming intracellular PHA and simultaneously accumulate polyphosphate (Smolders et al. 1995). GAOs take up organic substrates and accumulate them as PHA by consuming intracellularly accumulated glycogen as the energy source in the anaerobic tank, while they grow by consuming PHA and accumulate glycogen in the aerobic tank (Mino et al. 1995). However, Bengtsson (2009) reported that a sequencing batch reactor (SBR) sludge in which GAOs accounted for 54–70% of total populations could accumulate more
PHA under aerobic conditions than under anaerobic conditions when the PHA accumulation experiment was conducted in the absence of nitrogen and phosphate salt, indicating that it is likely that GAOs are also capable of accumulating PHA under aerobic and nitrogen- and phosphate-free conditions, which are coincident with the experimental conditions of test system A that was applied to evaluate the PHA accumulation potential of activated sludge in this study. Also, PHA accumulation by PAOs under the aerobic condition is not necessarily ruled out although there has been no available knowledge that directly confirmed it. Together, it could be hypothesized that high PHB contents observed in AO and AS samples in this study were partially attributable to the predominance of GAOs and PAOs, and further study is needed to clarify this hypothesis. By contrast, A2O and Step-feed BNR processes commonly have anoxic tanks and are operated for effective nitrogen removal. Considering these features, it appears that the operational conditions specific to the nitrogen removal processes may be unsuitable for enrichment of PHA-accumulating organisms. Beun et al. (2000) found using SBR incorporating an anoxic phase that activated sludge in which PHB had accumulated required a chemical oxygen demand to nitrogen (COD/N) ratio for denitrification 17 times higher than activated sludge with no accumulation of PHB. The low COD/N ratio of municipal sewage has been generally known (Beun et al. 2000; Peng et al. 2007; Pramanik et al. 2012). Thus, it was suggested that PHA accumulation may be an obstacle for the survival of micro-organisms in the nitrogen removal processes; PHA accumulation potentials were relatively low in A2O and Step-feed BNR samples. T-RFLP analysis targeting the phaC genes was previously established as a useful method to investigate the diversity of PHA-accumulating organisms (Michinaka et al. 2007). Application of the T-RFLP analysis in this study revealed the presence of PHA-accumulating organisms belonging to the order Burkholderiales, family Alcaligeneceae and genus Sinorhizobium. Alcaligenes, which belong to the family Alcaligeneceae, was reported to have a high PHA accumulation potential (Vishnuvardhan Reddy et al. 2008). However, this study could not find specific PHA-accumulating organisms that might be related to the PHA accumulation potentials observed in the PHA accumulation experiments (Figs 5 and 6). Therefore, the PHA accumulation potential of activated sludge in municipal WWTPs, which consists of various PHA-accumulating organisms, could not be easily predicted by specific PHA-accumulating organisms. By contrast, the quantification of phaC gene copy numbers revealed that PHB contents after PHA accumulation experiments reached more than 20 wt% in activated sludge samples with more
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than 50 9 108 copies (g-MLSS)−1 of phaC genes (Fig. 7). This indicates the presence of a threshold phaC gene number (PHA-accumulating organism number) to achieve a high PHB content and that phaC gene copy number could be used as a simple tool to provide a general indication of the PHA accumulation potential of activated sludge. The low PHA accumulation potentials of activated sludge in nitrogen removal processes could indicate that the specific operational conditions for those processes had negative effects on the enrichment of PHA-accumulating organisms. However, PHA accumulation potentials varied largely for activated sludge samples with similar phaC gene numbers (Fig. 7). Given the diversity of PHA-accumulating communities, it is likely that difference in the phaC gene expression, rather than the abundance of PHA-accumulating organisms (or phaC genes), significantly affected the accumulation of PHB observed in this study, which is an area that requires further research. PHB yields determined in this study (Fig. 2) were equivalent to or slightly lower than PHA yields indicated in a previous study using activated sludge samples collected from WWTPs operated in fully aerobic mode (029–064 C-mol 3-hydroxyalkanoate (C-mol acetate)−1) (Oshiki et al. 2013). However, considering that PHA amounts were determined as the sum of PHB and polyhydroxyvalerate in that study, PHB yields in our study may be equivalent to those reported in Oshiki et al. (2013). In the MBR sample (E-MBR), acetate consumption was relatively low and PHB yield was quite high (050 C-mol 3HB (C-mol acetate)−1) compared with the other samples investigated. The MBR process is operated with a high sludge concentration; consequently, the F/M ratio in MBR is quite low compared with the other processes (Fenu et al. 2010; Sipma et al. 2010). The F/M ratios of the processes that include a final settling tank were reported to be 02–04 g-BOD (g-MLSS day)−1 (Bitton 2005), while those of MBR were 0017–0092 g-BOD (g-MLSS day)−1 (Delrue et al. 2011). The MBR process investigated in this study was also operated with a very low F/M ratio compared with the other processes (Table 1). Rapid uptake of limited substrates and efficient accumulation of them as PHA are critical for survival under those conditions. Microbial community analyses indicated that MBR sludge had low phaC gene diversity and was dominated by a few PHA-accumulating organisms (Table 2). Furthermore, the relative abundance of dominant T-RFs in sample E-MBR had a significant positive correlation with the PHB yield. It is therefore possible that the specific conditions of the MBR process enabled selective enrichment of PHA-accumulating organisms capable of synthesizing PHB very efficiently; this would lead to a very high PHB yield. It should be noted, however, that only one MBR sample was 264
investigated in this study and further research is needed to verify the hypothesis. HRT and SRT are important operational parameters for biological wastewater treatment processes. In this study, a negative correlation was found between HRT and the acetate consumption for 24 h. This would be reasonable, given that all of the wastewater treatment processes investigated have achieved sufficient removal of organic compounds with the respective HRT settings. By contrast, regarding SRT, a previous study by Chua et al. (2003) investigated the PHB accumulation abilities of activated sludge acclimated from the same seed sludge under different SRTs (3 and 10 days) in 24-h batch experiments under the aerobic condition. They found that a higher PHA content after 24 h was achieved with the sludge acclimated under lower SRT condition. However, in this study, SRT did not show a significant correlation with either the acetate consumption or the PHB yield. This might be caused by the employment of activated sludge samples acclimated under various conditions regarding not only the treatment process but also operational parameters. Thus, further study with activated sludge samples acclimated from the same seed sludge under different SRT conditions is needed to clarify the effect of SRT on the PHA accumulation potential of activated sludge. In conclusion, this study revealed that activated sludge in municipal WWTPs has the potential to accumulate up to approx. 20 wt% of PHA in acetate-fed 24-h batch experiments under aerobic and nitrogen- and phosphatelimited conditions and that the phaC gene number can be used to predict the PHA accumulation potential. Results also indicate that the PHA accumulation potential of activated sludge depends largely on the operational conditions (treatment processes) of WWTPs: PHA accumulation potential was high in EBPR and low in nitrogen removal processes, and MBR would be particularly effective in selectively enriching PHA-accumulating organisms that synthesize PHA from limited substrates very efficiently. Acknowledgements We appreciate the anonymous WWTP officials for allowing us to collect the samples used in this study and for providing helpful information. This study was partly supported by Grants-in-Aid for Young Scientists (A) (No. 24681010) from the Japan Society for the Promotion of Science (JSPS). Conflict of Interest No conflict of interest declared.
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Supporting Information Additional Supporting Information may be found in the online version of this article: Table S1 Correlations between acetate consumption or PHB yield and various parameters
Journal of Applied Microbiology 118, 255--266 © 2014 The Society for Applied Microbiology