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Biodegradation of pyridine using aerobic granules in the
presence of phenol
Sunil S. Adav
a, Duu-Jong Lee
a,, N.Q. Ren
baDepartment of Chemical Engineering, National Taiwan University, Taipei 10617, Taiwan b
School of Environmental and Municipal Engineering, Harbin Institute of Technology, Harbin 150090, China
a r t i c l e
i n f o
Article history:
Received 21 January 2007 Received in revised form 11 March 2007
Accepted 30 March 2007 Available online 17 May 2007 Keywords: Aerobic granule Pyridine Phenol Inhibition Kinetics
a b s t r a c t
Aerobic granules cultivated with 500 mg/L phenol medium effectively degraded pyridine at a concentration of 250–2500 mg/L; maximum degradation rate was 73.0 mg pyridine g/VSS/h at 250 mg/L pyridine concentration. Phenol concentrations of 500–2000 mg/L limited pyridine degradation in a competitive inhibition pattern, as interpreted using Michaelis– Menten kinetics with corresponding parameters Vmax, Kmand KIof 63.7 mg/L h1, 827.8 and
1388.9 mg/L, respectively. Fluorescent staining and confocal laser scanning microscopy (CLSM) tests suggested that an active biomass accumulated at the granule outer layer. Denaturing gradient gel electrophoresis (DGGE) fingerprint profile demonstrated that dominating microbial strains exist in phenol and pyridine-degrading aerobic granules.
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1.
Introduction
Pyridine and its derivatives are by-products of coal gasifica-tion (Stuermer et al., 1982) and retorting oil shale (Leenheer et al., 1982), and are utilized as a catalyst in the pharmaceu-tical industry. Technologies for removing pyridine from wastewater are biodegradation (Sandhya et al., 2002; Rhee et al., 1996;Lee et al., 1991, 1994, 2001), adsorption (Akita and Takeuchi, 1993; Sabah and Celik, 2002; Yokoi et al., 2002), adsorption and electrosorption (Niu and Conway, 2002), ozonation (Stern et al., 1997), and ion exchange (Akita and Takeuchi, 1993).
Aerobic granules have been employed in treating high-strength wastewaters containing organic compounds (Moy et al., 2002), nitrogen and phosphorus (Yang et al., 2003), and phenol (Jiang et al., 2002, 2004).Tay et al. (2004)showed that their granules degraded phenol at a specific rate 41 g phenol/ g/VSS/d at a phenol concentration of 500 mg/L.Adav et al. (2007a, b) isolated strains in aerobic granules with high
phenol-degrading capability. Other recent works on aerobic granule processes include Nancharaiah et al. (2006a, b), de Kreuk et al. (2005), andSu and Yu (2005, 2006a, b).
This work examines the feasibility of using a cultivated phenol-fed granule to degrade pyridine in water, and dis-cusses the possible inhibition effects of phenol during pyridine degradation. The strains in aerobic granules corresponding to pyridine and phenol degradation were identified.
2.
Materials and methods
2.1. Granule cultivation
Aerobic granules were cultivated in a column-type sequential aerobic sludge reactor. The reactor was seeded with activated sludge and phenol as the sole carbon source
0043-1354/$ - see front matter & 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2007.03.038
Corresponding author. Tel.:+886 2 26325632; fax:+886 2 23623040. E-mail address:djlee@ntu.edu.tw (D.-J. Lee).
using synthetic wastewater with the following composition: 1000 mg/L (NH4)2SO4; 200 mg/L MgCl2; 100 mg/L NaCl; 20 mg/L
FeCl3; 10 mg/L CaCl2; and, phosphate buffer (1350 mg/L
KH2PO4, 1650 mg/L K2HPO4). The synthetic wastewater also
had the following micronutrients (g/L): H3BO3, 0.05; ZnCl2,
0.05; CuCl2, 0.03; MnSO4H2O(NH4)6, 0.05; Mo7O244H2O, 0.05;
AlCl3, 0.05; CoCl26H2O, 0.05; and NiCl, 0.05 (Moy et al., 2002).
The medium was sterilized by autoclaving for 15 min at 121 1C. The phenol solution was filter-sterilized and added to the autoclaved medium. The reactor was operated in 12 h cycles. Air bubbles for aeration were supplied through the base of the reactor. The granules formed and matured at pH 6.8 over 3 weeks.
2.2. Degradation test
Four parallel batch tests were performed. In test batch I, the synthetic wastewater containing pyridine (the sole source of carbon) at 250–3000 mg/L concentrations were added into the batch reactors. The experiment was initiated by inoculating equal volumes of aerobic granules.
In test batch II, the medium with 500 or 1000 mg/L pyridine was mixed with phenol as the carbon source at concentra-tions of 0–2000 mg/L. A constant volume of granules (4 mL) was added to the reactors.
In test batch III, the medium with 1000 mg/L phenol was mixed with pyridine at concentrations of 0–1500 mg/L. Equal volumes of granules were added to each reactor.
To identify the inhibition mechanism, the biomass in all reactors in test batch IV was kept constant with pyridine concentrations of 200–1000 mg/L with or without 250 and 500 mg/L phenol.
Air was supplied to each reactor at a rate of 1 L/min. Liquid samples were collected at 3 h intervals. To estimate the stripping loss of pyridine (and phenol) by airflow during tests, an identical reactor operated with the same synthetic wastewater and without granules was used as a control. Maximum stripping rate was 1.2 mg pyridine/L/h, which was insignificant throughout the test period (27 h). To minimize the stripping effects, condensers were equipped on all reactors to recover the stripped pyridine (and phenol). All experiments were performed in duplicate to assess data reproducibility.
Subsequently, adsorption was studied at various initial concentrations (ranging from 200 to 800 mg/L). The experi-ments were performed in serum bottles containing 50 mL solution and 2.5 g (wet weight) granules in rotary shaker at 100 rpm and samples were analyzed for residual pyridine at 2 h intervals for 18 h. The pyridine and phenol are the substrates for the active biomass hence the active biomass was heat-killed by heating the granules at 75 1C for 2 h. This method did not affect the physical integrity of the granules (Hawari and Mulligan, 2006; Nancharaiah et al. 2006a, b). The difference between uranium (VI) biosorp-tion onto live biomass (51.03 mg U/gm dry wt.) and heat-killed biomass (48.2 mg U/gm dry wt.) was not significant (Nancharaiah et al., 2006a, b). The maximum adsorption rate was 0.11 mg pyridine/L/h, which was negligible throughout the test-period.
2.3. Analytical methods
The dry weight of granules and volatile suspended solids (VSS) in the suspension were measured according to Standard Methods (APHA, 1998). Pyridine concentrations in the reactor were determined via high-performance liquid chromatogra-phy (HPLC) equipped with a C18 column (Varian, Inc., CA, USA), and measured spectrographically at 254 nm. The mobile phase consisted of acetonitrile:water (300:700), 0.11 g heptane sulphonic acid, 0.29 g anhydrous sodium acetate, and 2.5 mL glacial acetic acid. The concentration of phenol was mea-sured spectrographically at 276 nm.
2.4. Granule staining and CLSM imaging
The collected granules were maintained fully hydrated during staining. Fluoresceinisothiocyanate (FITC), an amine reactive dye, was utilized to stain proteins and amino-sugars in cells and the extracellular polymeric substances (EPS). The SYTO 63, which is a cell-wall-permeable nucleic acid stain, was applied to analyze cells. The SYTOX blue, a cell-wall-impermeable stain, was utilized to stain dead cells in the granules. Concanavalin tetramethylrhodamine con-jugate (ConA) was used to bind to a-mannopyranosyl and a-glucopyranosyl sugar residues. Nile red was utilized to stain lipids. Calcofluor white was utilized to stain b-polysacchar-ides. All probes were purchased from Molecular Probes (Carlsbad, CA, USA).
Confocal laser scanning microscopy (CLSM) (Leica TCS SP2 Confocal Spectral Microscope Imaging System, Gmbh, Ger-many) was utilized to visualize cell distributions in aerobic granules. The granules were imaged using a 10 objective and analyzed with Leica confocal software. The fluorescence of SYTO 63 was detected via excitation at 633 nm and emission at 650–700 nm. The fluorescent intensity of SYTOX blue was analyzed via excitation at 458 nm and emission at 460–500 nm. The fluorescence of Nile red, FITC, and Calcofluor white were detected via excitation at 514, 448, and 400 nm and emission at 625–700, 500–550, and 410–480 nm, respec-tively. Staining details are available inChen et al. (2007).
2.5. DNA isolation and DGGE
The DNA extraction method from aerobic granules is avail-able inAdav et al. (2007a). Polymerase chain reaction (PCR) amplification of the 16S ribosomal DNA (16S rDNA) gene was performed using extracted DNA with forward primer P1 and reverse primer P2 as described byMuyzer et al. (1993). The GC-rich sequence of 40 nucleotides (GC clamp) was attached at 50
end of primer P1.
Denaturing gradient gel electrophoresis (DGGE) tests were conducted utilizing the Bio-Rad universal mutation detection system with 10% (w/v) polyacrylamide gels. The range of denaturants (100% denaturant corresponds to 7 M urea and 40% (v/v) deionized formamide) was 35–65%. Electrophoresis was performed at 60 1C for 12 h at 120 V. Gels were stained with ethidium bromide and photographed using a UV trans-illuminator.
3.
Results
3.1. Granule characteristics
The pyridine-degrading granules were compact bioaggregates comprising many bacterial strains (Fig. 1). After degrading 1000 mg/L pyridine, filamentous bacteria were identified at the granule surface (Fig. 1b) together with a large diversity of microbial morphotypes including rods and cocci (Fig. 1c). No detectable changes in surface morphology of granules were observed for fresh phenol-cultivated granules and pyridine-degrading granules.
3.2. Pyridine biodegradation test
The aerobic granules efficiently degraded pyridine over initial concentrations of 200–2500 mg/L (Fig. 2). At initial pyridine concentrations of 250 and 500 mg/L, degradation kinetics followed closely a zero-order kinetics with no time delay. The specific degradation rate of pyridine was 73.0 and 66.8 mg pyridine/g/VSS/h at 250 and 500 mg/L of pyridine, respec-tively. A time lag was noticeable in tests with 2000 and 2500 mg/L pyridine, in which pyridine was completely de-graded at 80 and 110 h, respectively. The corresponding
specific degradation rate declined to 31.0 mg pyridine/g/VSS/h at 2500 mg/L pyridine. At X3000 mg/L pyridine, the rate of pyridine degradation was low, implying a strong inhibitory effect of pyridine on the granules.
3.3. Degradation of pyridine with phenol
Fig. 3 presents pyridine degradation curves with different initial concentrations of phenol. Compared with the control (pure pyridine test), the speed of pyridine degradation was slower, whereas the magnitude of reduction increased as the phenol concentration increased. For example, 12 h was required to completely degrade 500 mg/L pyridine, whereas 18 h was needed to degrade the same amount of pyridine in the presence of 2000 mg/L phenol, and 26 h was required to completely degrade 1000 mg/L pyridine. In the presence of 2000 mg/L phenol, lag time was approximately 10 h; complete degradation was achieved at 442 h.
Fig. 4shows the phenol degradation curves with different initial concentrations of pyridine. Phenol degraded faster than pyridine in pure substrate tests. When both substrates co-existed in the medium, the phenol degradation rate was slightly slower than that of the control; however, the rate of
Fig. 1 – Scanning electron micrographs of pyridine-degrading granules.
0 2 10 12 14 Pyridine concentration (mg/L) 0 Pyridine - 250 mg/L Pyridine - 500 mg/L 0 20 40 60 80 100 Pyridine concentration (mg/L) 0 Pyridine : 3000 mg/L Pyridine : 2500 mg/L Pyridine : 2000 mg/L Pyridine : 1500 mg/L Pyridine : 1000 mg/L 600 500 400 300 200 100 4 6 8 Time (h) 120 4000 3000 2000 1000 Slope = -39.30±0.27 Time (h)
pyridine degradation was largely suppressed. Hence, pyridine has minimal effect on phenol degradation though the opposite is not true.
4.
Discussion
4.1. Co-degradation of phenol and pyridine
Phenol limited the degradation rates for pyridine via granules (Fig. 3), as satisfactory statistics performed by ANOVA (po0.001) (Table 1).
Pyridine reaction rates with and without 250 mg/L phenol were determined at a constant granule mass (1.5970.05 g/VSS/L). The Lineweaver–Burk plot demonstrated a competitive inhibition (Fig. 5). Restated, phenol competed with pyridine for the same enzyme site; however, the formed enzyme–phenol complex could not generate its endproduct in the reaction. The specific enzyme for pyridine degradation was present in the system that has a high affinity for phenol. Thus, phenol granules could be applied for the removal of phenol in the presence of pyridine in industrial wastewater.
The Michaelis–Menten kinetics in the presence of a competitive inhibitor was stated as follows:
V0¼
Vmax½S
aKmþ ½S
, (1)
where a ¼ 1+[I]/KI, Vmaxis the maximum reaction velocity, S
the pyridine concentration, I is the phenol concentration, Km
is the Michaelis–Menten constant, and KI is the inhibitory
constant. Eq. (1) can be rearranged into the Lineweaver–Burk type equation 1 V0 ¼ aKm Vmax½S þ 1 Vmax. (2)
The Vmax, Km, a1Kmand a2Kmdetermined fromFig. 5were
63.7 mg/L/h, and 827.8, 981.8, and 1381 mg/L, respectively. That is, a1¼1.18 and a2¼1.67, and KI was estimated to
be 1389 and 746.3 mg/L. At 250 mg/L phenol concentration,
0 8 10 12 14 16 18 20 Pyridine concentration (mg/L) 0 100 200 300 400 500 600 Pyridine : phenol-500:0mg/L Pyridine : phenol-500:500 mg/L Pyridine : phenol-500:1000 mg/L Pyridine : phenol-500:1500 mg/L Pyridine : phenol-500:2000 mg/L 0 10 20 30 40 Pyridine concentration (mg/L) 0 200 400 600 800 1000 1200 Pyridine : phenol - 1000:0 mg/L Pyridine : phenol - 1000:500 mg/L Pyridine : phenol - 1000:1000 mg/L Pyridine : phenol - 1000:1500 mg/L Pyridine : phenol - 1000:2000 mg/L Time (h) Time (h) 2 4 6
a
b
Fig. 3 – Influence of different initial phenol concentrations on pyridine degradation.
0 10 15 20 25 Phenol concentration (mg/L) 0 200 400 600 800 1000 1200 Phenol : pyrdine (1000:0) mg/L Phenol : pyridine (1000:500) mg/L Phenol : pyridine (1000:1000) mg/L Phenol :pyridine (1000:1500) mg/L 5 Time (h)
Fig. 4 – Influence of different initial pyridine concentrations on phenol degradation.
Table 1 – Comparison of pyridine and phenol degradation rates by aerobic granules
Concentration (pyridine:phenol, mg/L) Rate of degradation (mg pyridine (g1VSS h)1) One-way ANOVA 500:0 (control) 66.970.1 — 500:500 57.570.7** p ¼ 0.002 500:1000 54.671.5** p ¼ 0.006 500:1500 47.671.9*** p ¼ 0.001 500:2000 44.771.6*** po0.001 1000:0 (control) 58.570.1 — 1000:500 51.371.1** p ¼ 0.005 1000:1000 45.272.8*** po0.001 1000:1500 40.371.6*** po0.001 1000:2000 36.674.8*** po0.001 Values represent mean7 SEM; *po0.05; **po0.01; ***po0.001 versus control (Tukey–Kramer multiple comparisons test).
KIwas 29.3% higher than Km, suggesting that the
pyridine-degrading enzyme had greater affinity for pyridine than phenol.
Fig. 5shows a set of double-reciprocal plots, one obtained in the absence of inhibitor and two at different concentrations of a competitive inhibitor. Increasing [I] resulted in a family of lines with a common intercept on the 1/rate of degradation axis but with different slopes. The unchanged value of Vmax
demonstrated the competitive effect.
4.2. Granule structure
Fig. 6 (left panel) presents the CLSM images of stained granules following degradation with 1000 mg/L pyridine. The fluorescent intensity of proteins (green), lipids (yellow), and dead cells (violet) correlated with each other and accumu-lated over the complete granule interior. This observation suggests that the proteins and lipids were generally bound to dead cell membranes. The granule outer layer was composed
y=12.996x+0.0157 R2=0.97 1 L mg-1 -0.001 0.000 0.001 0.002 0.003 0.005 1 L h mg -1 Rate of degradation 0.00 0.02 0.04 0.06 0.08 0.10 0.12 0.14 Pyridine
Pyridine and phenol (250 mg/L) Pyridine and phenol (500 mg/L)
y=15.414x+0.0157 R2=0.98 y=21.675x+0.0157 R2=0.99 -0.02 [Substrate] 0.006 002 0.004
Fig. 5 – Line Lineweaver–Burk plot for bisubstrate reaction.
0 50 100 150 200 250 Fluorescence intensity 0 10 20 30 40 50 60 Col 13 vs Col 14 0 20 40 60 80 100 120 140 160 180 0 10 20 30 40 50 60 0 20 40 60 80 0 20 40 60 80 100 120 140 160 0 20 40 60 80 100 120 140 160 Distance (µm) Col 15 vs Col 16 Col 17 vs Col 18 Col 19 vs Col 20 Col 21 vs Col 22 Col 23 vs Col 24
a
e
b
f
g
c
d
Fig. 6 – CLSM images of pyridine-degrading granule and distribution profiles of extracellular polymeric substances along the line indicated. (Distance 0 indicates surface of granule). Left: (a) lipids (yellow): Nile red; (b) proteins (green): FITC; (c) total cells (red): SYTO 63; (d) a-polysaccharide (light blue): Con A rhodamine; (e) dead cells (violet): Sytox blue; (f) b-polysaccharide (blue): calcofluor white; (g) combined image. Right: fluorescent intensities of six stains.
of live cells (SYTO 63 in red), b-polysaccharide (Calcofluor white in blue), and a-polysaccharide (Con A in light blue). A bioactive layer located on the granule surface, similar to
those observed in phenol-degrading granules (Adav et al., 2007a, b).
Fig. 6 (right panel) presents the intensity distributions of the six stains applied to the granule. Inward from the granule surface is a layer approximately 10 mm thick, primarily composed of lipids and a-polysaccharides. Under this surface layer exists a layer comprising lipids, proteins, a- and b-polysaccharides, and live cells (10–40 mm). Inside this layer, the non-cellular core was principally composed of proteins, lipids, and dead cell nucleic acids.
All granules analyzed in this work have layered structures in their interiors similar to those shown inFig. 6. Hence, only approximately 22% of the granule mass contained active biomass for phenol/pyridine degradation. The non-cellular core comprising mainly proteins provided the granule with its mechanical strength.
4.3. Microbial analysis
Fig. 7represents the DGGE fingerprint profile of PCR amplified sequences for phenol, pyridine, and phenol/pyridine co-degrading aerobic granules. The DNA extraction and DGGE experiments were performed twice. An identical DGGE pattern was acquired for replicate samples. Nine bands were excised, amplified and sequenced to identify microbial species (Table 2).
The following strains were identified in granules fed with 2000 mg/L phenol: Bacillus weihenstephanensis strain RBE1CD (band 1); B. sphaericus strain D45 (band 2); Enterobacter cancerogenus sp. EBD (band 3); B. cereus (band 4); Acinetobacter sp. (band 5); Acinetobacter calcoaceticus strain CBMAI 464 (band 8); and, Pseudomonas sp. Hugh2319 (band 9).
A shift in the microbial community was observed when pyridine was present. When the phenol-fed granules were cultivated in 500 mg/L pyridine solution for 24 h, the E. cancerogenus sp. EBD strain (band 3) was undetectable in Fig. 7 – Ethidium bromide-stained polyacrylamide
denaturing gradient gel with DGGE profiles of 16S rDNA gene fragments after PCR amplification of nucleic acids from (a) phenol granules, (b) pyridine granules, and (c) co-degrading phenol and pyridine granules.
Table 2 – Sequence analyses of bands derived from 16S rDNA of aerobic granules Band
no.
Most closely related bacterial sequence Ref accession no. Similarity (%) GenBank accession no. Phenol test Pyridine test Phenol/ pyridine test 1 Bacillus weihenstephanensis strain RBE1CD EF111134 99 EF190020 J J J
2 Bacillus sphaericus strain D45 DQ923492 98 EF190021 J J J
3 Enterobacter cancerogenus sp. EBD
EF011116 99 EF190022 J – –
4 Bacillus cereus DQ177461 100 EF190025 J J –
5 Acinetobacter sp. DQ837531 100 EF190029 J J J
6 Uncultured bacillus sp. EF072549 98 EF190028 – J J
7 Klebsiella pneumoniae strain IEDC 78
DQ122297 97 EF190024 – – J
8 Acinetobacter calcoaceticus strain CBMAI 464
DQ250143 98 EF190023 J J –
9 Pseudomonas sp. Hugh2319 AB247215 97 EF190026 J J J
the granule; however, an uncultured Bacillus strain (band 6) appeared in the pyridine medium.
When 500 mg/L pyridine combined with 2000 mg/L phenol was utilized to cultivate granules originally grown in 2000 mg/L phenol, B. cereus (band 4) was not present in the granule, whereas both uncultured Bacillus sp. (band 6) and Klebsiella pneumoniae strain IEDC 78 (band 7) appeared in the granules.
Among these strains, B. sphaericus, Pseudomonas sp. Hugh2319, and the Actonetobactor sp. have been shown in the literature to have the capability to degrade aromatic com-pounds (Kaplan and Rosenberg, 1982, Navon-Venezia et al., 1995,Reisfeld et al., 1972). In all three mediums, Acinetobactor sp. was present in a significant quantity, and likely dominated the pyridine–phenol aerobic granule system. The dominating strains in bioactive layers degraded pyridine and phenol with some mass transfer shield provided by the granule surface layer comprised of lipids and a-polysaccharides.
5.
Conclusions
The cultivated phenol-fed aerobic granules could degrade pyridine with zero order kinetics of 73.0 mg pyridine/g/VSS/h at 250 mg/L pyridine concentration. The pyridine degradation tests suggested that granules completely degraded 250–1500 mg/L pyridine at a constant rate with no time lag, and with 12 and 15-h time lag at 2000 and 2500 mg/L pyridine concentration, respectively. Significant effect was noted for pyridine degrada-tion by aerobic granules at concentradegrada-tion 42500 mg/L pyridine. The staining-CLSM test revealed that active microorganisms located at the granule surface while protein and some poly-saccharide comprised the granule core.
The presence of phenol limited pyridine degradation rates by the granules. On the contrary, pyridine had minimal effect on the granules to degrade pyridine. The double reciprocal Lineweaver–Burk plot showed a competitive type of inhibition for phenol on pyridine degradation. The Vmax, Kmand KIin
the Michaelis–Menten kinetics were estimated as 63.7 mg/L/h, 827.8 and 1388.9 mg/L, respectively. The DGGE fingerprint pattern identified nine strains in the aerobic granules. Microbes shift in phenol-cultured granules when pyridine was present. The Acinetobactor sp. was expected to dominate the pyridine–phenol aerobic granule system.
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