- Email: [email protected]
Anaerobic biological treatment of phenolic wastewater at 15–18 °C
ARTICLE IN PRESS
Water Research 39 (2005) 1614–1620 www.elsevier.com/locate/watres
Anaerobic biological treatment of phenolic wastewater at 15–18 1C Gavin Collins, Clare Foy, Sharon McHugh1, The´re`se Mahony, Vincent O’Flaherty Microbial Ecology Laboratory, Department of Microbiology and Environmental Change Institute (ECI), National University of Ireland, Galway (NUI, Galway), University Road, Galway, Ireland. Received 24 November 2004; received in revised form 20 January 2005; accepted 25 January 2005 Available online 17 March 2005
Abstract Low-temperature, or psychrophilic (o20 1C) anaerobic digestion has been proven feasible for the mineralisation of simple wastewaters. In this study, hybrid expanded granular sludge bed-anaerobic filter (EGSB-AF) bioreactors were used to evaluate the feasibility of psychrophilic digestion for the treatment of phenol-containing wastewater. Efficient chemical oxygen demand and phenol removal were observed at organic and phenol loading rates of 5 kg COD m 3 d 1 and 0.4–1.2 kg phenol m 3 d 1 (400–1200 mg phenol [l wastewater] 1), respectively. There was no long-term accumulation of volatile fatty acids in the reactor systems. Methanogenic activity was developed under psychrophilic conditions but anaerobic methane-producing populations remained mesophilic throughout the trial of 415 days. r 2005 Elsevier Ltd. All rights reserved. Keywords: Psychrophilic anaerobic digestion; Phenol; EGSB; Specific methanogenic activity.
1. Introduction Phenolic compounds are present in the liquid effluent of coal gasification plants, coking plants, petroleum refineries, pharmaceutical, fertiliser and dye manufacturing plants, as well as degreasing and paint stripping operations (Khan et al., 1981) and fibreboard manufacturing (Eroglu et al., 1994). Although not found to be bioaccumulative (Loehr and Krishnamoorthy, 1988), humans exposed to phenol in well water at concentraCorresponding author. Tel.: +353 91 493734; fax: +353 91 525700. E-mail address: vincent.ofl[email protected] (V. O’Flaherty). 1 Present address: Environmental Microbiology Research Unit, Department of Microbiology, NUI, Galway, University Road, Galway, Ireland.
tions of 1300 mg l 1 exhibited a statistically significant increase in diarrhoea, mouth sores, dark urine and burning of the mouth (US EPA, 1980). Anaerobic digestion of phenolic wastewaters enjoys increasing application, and upflow anaerobic sludge blanket (UASB) and expanded granular sludge bed (EGSB) reactor configurations have been successfully demonstrated for the remediation of this category of effluent (Chang et al., 1995; Li et al., 1996). Continued advances in this field, however, are dependent on concerted efforts to improve and amend existing reactor configurations to design next generation technologies. In this context, the application of EGSBbased state-of-the-art reactor designs currently plays a significant role in the search for broader and newer applications of anaerobic digestion, and the successful widespread adoption of anaerobic reactors for the treatment of the vast and important range of phenolic
0043-1354/$ - see front matter r 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2005.01.017
ARTICLE IN PRESS G. Collins et al. / Water Research 39 (2005) 1614–1620
discharges is strongly dependant on the bioengineering ingenuity required to provide increasingly efficient and versatile technology. Further, and in parallel to this, there is growing interest in maximising methane extraction for energy recovery from anaerobic treatment processes. As the popularity of anaerobic digestion increases, so too does the urgent demand for greater economic efficiency from these processes. This requirement has prompted recent study into the feasibility of low-temperature reactor operations (Rebac et al., 1995; Rebac et al., 1999; Collins et al., 2003). One of the most important parameters for anaerobic treatment of wastewater is operating temperature. The mesophilic range is traditionally used since it is generally thought that maintaining a high (thermophilic) temperature is uneconomical for many wastestreams, whereas degradation within the psychrophilic range is too slow. However, efficient degradation, which was comparable to that from mesophilic trials, has been demonstrated for a range of wastewater types (Rebac et al., 1995; Lettinga et al., 1999; Rebac et al., 1999; Nozhevnikova et al., 2000; Collins et al., 2003; McHugh et al., 2004). Furthermore, the majority of industrial effluents released for treatment in temperate climates are below 20 1C. It follows, therefore, that psychrophilic anaerobic treatment, which returns satisfactory methane production, presents these countries with an attractive and economically sound option for sustainable remediation regimes. The aim of this study was to evaluate the feasibility of anaerobic biological treatment of phenolic wastewaters under ambient conditions (15–18 1C).
1615
Table 1 Particulars of wastewater components Phase/day
I II III IV V VIa
0–69 70–138 139–246 247–354 355–395 396–415
Phenol (mg l 1)
COD (g l 1) Phenol
VFA
Total
R1 R2
R1 R2
R1
R2
R1 R2
0 0 0 0 0 0
0 0 400 800 1200 1200
0 0 0 0 0 0
0 10 10 10 10 0 5 5 5 5 0.96 5 4.04 5 5 1.92 5 3.08 5 5 2.88 5 2.12 5 5 2.88 5 2.12 5 5
a Phase VI was characterised by the reduction in reactor operating temperature from 18 to 15 1C
A mesophilic, anaerobic granular sludge was obtained from a full-scale (1500 m3) internal circulation (IC) reactor, at Carbery Milk Products, Ballineen, Co. Cork, Ireland. The small (Ø, 0.4–0.8 mm), black and grey granules were of a regular form and settled well in liquid phase.
fatty acid (VFA)-based, industrial wastewater (pH 7.570.2) consisting of ethanol, butyrate, propionate and acetate, in the COD ratio of 1:1:1:1, to a total of 10 g COD l 1. The influent was buffered with NaHCO3 and fortified, as described by Shelton and Tiedje (1984), with macro- (10 ml l 1) and micro- (1 ml l 1) nutrients. The trial period was divided into six operational phases as detailed in Table 1. A loading rate of 10 kg COD m 3 d 1 was applied to each reactor, with a hydraulic retention time (HRT) of 24 h, and effluent was recycled at a rate of 5 m h 1. The loading rate was reduced to 5 kg COD m 3 d 1, on day 70, when the COD concentration of the influent was decreased to 5 g COD l 1, in order to advance the start-up period of the reactors. The liquid upflow velocity applied through the recirculation facility was increased to 7.5 m h 1 and to 10 m h 1, on days 34 and 98, respectively. The operational temperature was maintained at 18 1C until day 395 and at 15 1C for the remainder of the 415-day trial period. R2 influent was supplemented with phenol (Sigma), to a final concentration of 400 mg l 1, on day 138 (Table 1). Phenol was not added to R1, thereby maintaining this reactor as an experimental control. Phenol concentration in R2 influent was increased to 800 mg l 1 and 1200 mg l 1, on days 246 and 354, respectively.
2.2. Reactor set-up and operation
2.3. Biogas and effluent analysis
Two 3.5 l glass laboratory-scale expanded granular sludge bed (EGSB)-based reactors (R1 and R2), which were of the same design as that described by Colleran and Pender (2002), with the addition of an upper fixedfilm section, which was randomly packed with polyethylene rings of 1 cm diameter each, were used here. These hybrid expanded granular sludge bed-anaerobic filter (EGSB-AF) reactors were each inoculated with 70 g volatile suspended solids (VSS) of the seed sludge. The reactors were used for the stabilisation of a volatile
Samples of reactor effluent were taken for VFA and COD analysis and biogas was sampled for CH4 determination (American Public Health Association (APHA), 1992). In addition, samples of reactor liquor were retrieved from a sampling port located mid-way up the reactor column for COD analysis. Effluent phenol concentrations were ascertained using a colorimetric spectrophotometer (Odyssey DR/2500, Hach) and the 4aminoantipyrine technique (APHA, 1992). Briefly, this method involved the reaction of phenols with
2. Materials and methods 2.1. Biomass
ARTICLE IN PRESS G. Collins et al. / Water Research 39 (2005) 1614–1620
4-aminoantipyrine in the presence of potassium ferricyanide, to form a coloured antipyrine dye. The dye was extracted from the aqueous dilutions of effluent samples with chloroform and the colour was measured at 460 nm. 2.4. Batch specific methanogenic activity, biodegradability and toxicity assays
COD Removal Efficiency (%)
Maximum specific methanogenic activity (SMA) profiles of the seed sludge and the biomass from both reactors at the conclusion of the trial (day 415) were determined for the substrates acetate, propionate, butyrate, ethanol, phenol and H2/CO2 using the pressure transducer technique (Colleran and Pistilli, 1994; Coates et al., 1996). In addition, substrate utilisation rate (SUR)-based biodegradability tests were used to assess the biodegradability of the test chemical (phenol) using the seed sludge and biomass from both the control and test reactors. Biomass samples were incubated (at 15 1C or 37 1C) with oxygen-free phenolic solutions and anaerobic test buffer, which was fortified with microand macro-nutrients (Shelton and Tiedje, 1984). Substrate (phenol) depletion was determined by collecting homogenous samples from vials, as a function of time. Phenol concentrations were determined as described above. Biodegradability tests were carried out at an initial concentration of 200 mg phenol l 1 and SUR values were expressed as the rate of phenol depletion (mg phenol [g VSS] 1 d 1). SUR-based tests were carried out in 60 ml sealed hypovials to a total volume of 30 ml. Phenol toxicity was assessed using the SMA-based toxicity assay described by Colleran and Pistilli (1994). Phenol-induced inhibition of methanogenic (acetoclastic and hydrogenotrophic) and syntrophic populations was
100 90 80 70 60 50 40 30 20 10 0
determined for seed sludge and reactor biomass samples. Toxicity was defined in terms of the IC50 value, i.e. the concentration of phenol that resulted in 50% inhibition of the control vial SMA, which was calculated from the linear regression of SMA as a function of phenol concentration. All activity, biodegradability and toxicity test vials contained 2–5 g VSS l 1.
3. Results 3.1. Reactor performance A lengthy start-up period, of over 135 days, was recorded for both reactors, after which COD removal efficiencies of over 80% (at the applied loading rate of 5 kg COD m 3 d 1) were recorded (Fig. 1). The dominant VFA present in the reactor effluents during phases I and II were acetate and propionate (Fig. 2). Upon supplementation of R2 influent with phenol (Table 1), reduced COD removal (Fig. 1) and biogas methane composition (data not shown) were observed. However, by the end of phase III, decreased VFA concentrations in R2 effluent coincided with increased phenol removal. After day 230, comparable performance data were recorded for both reactors (Fig. 1). There was a decline in R2 COD removal upon the application of the 0.8 kg phenol m 3 d 1 loading rate on day 247, which was accompanied by an accumulation of phenol in the reactor effluent (Fig. 1). However, by day 260, there was an evident improvement in the phenol removal rate and COD removal equal to R1 was recorded by day 340 (Fig. 1). Sporadic accumulation of acetate was observed in R1 effluent (Fig. 2) Nevertheless, R1 acetate levels did not exceed 500 mg l 1 at any stage and propionate
400 350 300 250 200 150 B
C
D
A 0
50
100 150 200 250 300 350 400 Time (Days)
100 50
Effluent Phenol (mg/l)
1616
0
Fig. 1. COD removal efficiency of R1 (’) and R2 (J), and R2 effluent phenol concentration (m). Applied phenol loading rate: (A) 0.4 kg phenol m 3 d 1; (B) 0.8 kg phenol m 3 d 1; (C) 1.2 kg phenol m 3 d 1. (D) Operating temperature of R1 and R2 reduced to 15 1C.
ARTICLE IN PRESS G. Collins et al. / Water Research 39 (2005) 1614–1620
II
I
III
IV
V
VI
VFA Concentration (mg/l)
2200 2000 1800 1600 1400 1200 1000 800 600 400 200 0 (A) 0
Table 2 Specific methanogenic activity (ml CH4 [g VSS] seed sludge and R1 and R2 biomass
Ethanol Acetate Propionate Butyrate H2/CO2 Phenol
50
100
150
200
250
300
350
1617
1
d 1) of the
Seed sludge
R1
R2
15 1C
37 1C
15 1C
37 1C
15 1C
37 1C
312 140 39a 158 40.3 0.6
553 285 110 516 52 0
627 598 202 967.5 610 2.8
889 824 705 1215 843 3.6
371.4 321.4b 146c 496 347 87
678.2 398.9 209 651 510 120
400
VFA Concentration (mg/l)
a
(B)
2200 2000 1800 1600 1400 1200 1000 800 600 400 200 0
Lag 120 h. Lag 73 h. c Lag 382 h. b
0
50
100
150
200 250 300 Time (Days)
350
400
Fig. 2. (A) R1 and (B) R2 effluent VFA concentrations; acetate (’), ethanol (m), propionate (J) and butyrate (K). Timeline illustrates the six operational phases studied.
concentrations remained steadily low throughout. In contrast, an accumulation of acetate and propionate, to concentrations of 1000 and 1400 mg l 1, respectively, was observed in R2 effluent during the initial days of phase IV. However, an abatement of VFA accumulation, in parallel with increased phenol degradation, was observed by day 270, signalling a stabilisation of R2 performance. R2 COD removal efficiency decreased again after the addition of phenol to a concentration of 1200 mg l 1 (Fig. 1). Propionate, at maximum levels of 1200 mg l 1, constituted the predominant VFA in R2 effluent (Fig. 2). The recovery period recorded after this increase in the phenol loading to 1.2 kg phenol m 3 d 1 was shorter than those observed after day 138 (0.4 kg phenol m 3 d 1) and 247 (0.8 kg m 3 d 1). Improved R2 COD removal was evident within 40 days of instigating this increment and the performance of both digesters was comparable by day 390 (Fig. 1). Modest increases in VFA concentrations were observed in R1 effluent after the temperature decrement to 15 1C and were accompanied by a temporary accumulation of butyrate (Fig. 2). On the other hand, substantially greater concentrations of acetate and propionate
(200 and 600 mg l 1, respectively) were detected in R2 effluent on the days after the temperature change. However, a rapid recovery period of about 10 days was recorded whereby R2 performance data were comparable to those from the control reactor by day 405. Data from COD and phenol assays using the mixed reactor liquor sampled from the port mid-way up the reactor columns indicated that the anaerobic biofilm, which developed on the polyethylene rings, contributed 473% to overall COD removal in both reactors and 371.5% to total phenol removal by R2. 3.2. Metabolic characteristics of biomass as determined by specific methanogenic activity, biodegradability and toxicity assays The seed granules exhibited high SMA values when tested within the mesophilic range (37 1C) (Table 2). However, relatively low hydrogenotrophic activity was recorded at 37 1C and low SMA values against phenol suggested poor degradation of the test compound (Table 2). Lower SMA values were recorded, against all substrates when the assays were carried out within the psychrophilic range (15 1C), thereby confirming the mesophilic status of the reactor inoculum (Table 2). R1 SMA values at the conclusion of the trial were substantially higher than the respective values collated for the seed sludge at both 37 1C and 15 1C. High SMA values for H2/CO2 utilisation were recorded from R1, thereby suggesting satisfactory development of the hydrogenotrophic methanogenic community (Table 2). In fact, the R1 SMA values at 15 1C were higher than those recorded for the seed sludge at 37 1C, thus pointing towards substantial development of the methanogenic activity of this biomass against all substrates tested (except phenol). However, the optimum temperature of all of the trophic groups evaluated here remained within
ARTICLE IN PRESS G. Collins et al. / Water Research 39 (2005) 1614–1620
1618
Table 3 Substrate utilisation rates (SURs) for seed sludge and reactor biomass samples using phenol Biomass
SUR (mg phenol [g VSS]
1
d 1)
15 1C
37 1C
Seed Sludge 13.97 R1 sludge (Day 415) 15.4 R2 sludge (Day 415) 88.96
19.9 24.9 165.1
Table 4 IC50 values for the methanogenic and syntrophic (propionatedegrading) microorganisms in seed sludge, R1 and R2 biomass. Values are mg l 1 Seed sludge
R1
R2
Acetoclastic
15 1C 37 1C
305 341
426.9 436
1181 1410
Hydrogenotrophic
15 1C 37 1C
331 761
840 942.4
1147 1681
Propionate degraders
15 1C 37 1C
121.6 199
147 217
231 349.3
otrophs (Table 4). The low SMA of the seed and R1 sludges against phenol as substrate (Table 2), and the poor biodegradability of phenol in batch assays were confirmed by low toxicity thresholds observed within both the psychrophilic and mesophilic ranges (Table 4). However, higher IC50 values were obtained for R2 granules at the conclusion of the trial (Table 4). This was despite the observation that R2 SMA values were generally lower than those determined for R1 (Table 2). The reduced sensitivity (higher IC50) to the presence of phenol exhibited by all three populations tested was in agreement with the increased SMA against phenol (Tables 2 and 4). The toxicity profile obtained for R2 sludge suggested that phenol generally inhibited the trophic groups in the same order as in both the seed and R1 sludges, i.e. propionate-degraders IC50 valueoacetoclastic methanogensohydrogenotrophic methanogens. Finally, all IC50 values calculated for the reactor biomass were higher when tested within the mesophilic range than at 15 1C, indicating greater sensitivity to the test compound at lower temperatures and thus broadly supporting the conclusion that the reactor biomass remained substantially mesophilic, despite some evidence of psychrotolerance.
4. Discussion the mesophilic range (Table 2), suggesting psychrotolerant, rather than psychrophilic populations within the biomass. A similar SMA profile, in terms of temperature optima, was obtained from R2 biomass at day 415. As with R1, the activity values recorded for R2 sludge at 15 1C were greater than those determined for the seed within the mesophilic range (for all substrates except butyrate). The development of phenol-degrading activity was also evident, reflecting the efficient phenol mineralisation of the R2 reactor. All SMA values recorded for this biomass were lower than corresponding R1 values, indicating inhibition of the development of methanogenic activity in R2. Low-substrate utilisation rates (SURs) were calculated for both the seed and R1 biomass (Table 3). Although poor substrate depletion was observed from the unacclimated seed and R1 sludges, the mesophilic optimum of the biomass was evident from these tests (Table 3). As expected, higher substrate depletion rates were recorded for R2 biomass, at both 15 1C and 37 1C (Table 3). Although a mesophilic optimum was also obvious for this sludge, only 6.1% of the initial substrate introduced to the vials incubated at 15 1C was present at the conclusion of the tests. IC50 values indicated that propionate-utilising organisms were most sensitive to phenol, while the inhibition order among methanogens was acetoclasts4hydrogen-
It can be said from the results of this study that efficient phenol degradation was achieved in a psychrophilic anaerobic reactor operated at a constant temperature of 18 1C. In addition, the successful operation of the phenol-mineralising reactor at 15 1C (in phase VI) serves to confirm the feasibility of this treatment approach and further underlines the potential for costeffectiveness and rationalisation of many economically non-sustainable remediation practices. The principle VFA detected in the effluent of the test reactor immediately after the drop in operating temperature to 15 1C was propionate. Previous research has indicated that propionate-degradation may be the rate-limiting step in the psychrophilic anaerobic digestion process (Rebac et al., 1999; Collins et al., 2003). However, stable propionate degradation was observed within 10 days of the temperature reduction applied to both reactors. Furthermore, the development of propionate-degrading activity was also observed (Table 2). Reactor performance was found to improve with the number of phenol amendments, thus showing that microbial adaptation can occur throughout the trial. The length of the acclimation phase decreased after each phenol amendment, from initially being 100 days (during the 107 days of phase II) to 97 days (during phase III) and then 20 days (during the 40 days of phase IV). This is in agreement with the findings of Guieysse et al. (2001) and confirms that the use of a mesophilic
ARTICLE IN PRESS G. Collins et al. / Water Research 39 (2005) 1614–1620
granular sludge as an inoculum for psychrophilic acclimation of suitable degraders and digestion of phenolic wastewater is a feasible treatment option. Interestingly, the contribution to overall COD and phenol removal of the upper fixed-film section was low. This may be due to (1) low levels of biomass washout from the sludge bed and consequent lack of seed biomass for biofilm development and/or (2) the low COD concentration of reactor liquor entering this section from the sludge bed. Nevertheless, the benefits of employing this reactor design has been recently demonstrated in a 2,4,6-trichlorophenol-degrading reactor, where partially dechlorinated phenols from the sludge bed section were degraded by fixed-film consortia in the upper section of that reactor (Collins et al., 2005). The SMA profiles collated for both reactor sludges in this study indicated the development of methanogenic activity, even under low-temperature conditions. In fact, R1 SMA values at 15 1C were higher than those recorded for the seed sludge at 37 1C. However, the optimum temperature of all of the trophic groups evaluated here remained within the mesophilic range (Table 2), suggesting psychrotolerant, rather than psychrophilic populations within the biomass and thus supporting the conclusions of Rebac et al. (1999), Collins et al. (2003) and McHugh et al. (2004). Furthermore, reduced methanogenic development was observed in R2 sludge, which pointed towards inhibition of some or all of the trophic groups involved in the processes of this reactor. In addition, increased activity against H2/CO2 was observed at the conclusion of the trial. This phenomenon has also been observed during previous low-temperature treatment trials (McHugh et al., 2004; Collins et al., 2005). Notwithstanding the presence of hydrogenic substrates (ethanol, propionate, butyrate and phenol) in the influent wastewater in this study, the reason for high hydrogenotrophic methanogenic activity in low-temperature reactors is largely unknown. A possible explanation is based on thermodynamic considerations, as lowering the operating temperature may favour hydrogenotrophic methanogenesis (Conrad and Wetter, 1990), as the H2 threshold concentrations for hydrogenotrophic methanogens are significantly lowered under psychrophilic conditions. Furthermore, due to the increased solubility of gases at low temperatures (Perry and Green, 1984; Lettinga et al., 2001), it is possible that elevated dissolved H2 levels were present within the mixed liquor of R1 and R2, encouraging the growth of hydrogenotrophic methanogens. It is clear that many questions remain to be answered in this field. One such question is whether the apparent shift towards hydrogenotrophic methanogenesis in psychrophilic anaerobic reactors is primarily a temperature-related issue or is simply a response to environmental stress. In addition, a greater insight to the degradation of propionate in psychrophilic environ-
1619
ments—and the development of reactor designs to overcome propionate accumulation—is desirable.
5. Conclusions The following conclusions can now be drawn: (i) psychrophilic anaerobic digestion of phenolic wastewaters is feasible at 15–18 1C and at applied loading rates of up to 1.2 kg phenol m 3 d 1; (ii) further research is required to establish the upper limits of phenol loading for low-temperature anaerobic treatment; (iii) the use of an unacclimated mesophilic sludge is a satisfactorily efficient means of cultivating a phenoldegrading consortium for psychrophilic reactor operation—this is of particular importance if phenol-acclimated, psychrophilically cultivated biomass is not readily available.
Acknowledgements Financial support of the Higher Education Authority (HEA) Programme for Research in Third Level Institutes (PRTLI)-Cycle II, through the Environmental Change Institute (ECI), NUI, Galway, and an Enterprise Ireland Research Scholarship to G.C. are gratefully acknowledged. Paula McDermott is thanked for her assistance with data collation.
References American Public Health Association, 1992. Standard methods for the examination of water and wastewater, 18th ed. Washington, DC. Chang, B., Chen, K., Yuan, S., 1995. Dechlorination of 2,4,6TCP by an anaerobic mixed culture. Chemoshpere 31, 3803–3811. Coates, J.D., Coughlan, M.F., Colleran, E., 1996. Simple method for the measurement of the hydrogenotrophic methanogenic activity of anaerobic sludges. J. Microbiol. Methods 26, 237–246. Colleran, E., Pender, S., 2002. Mesophilic and thermophilic amaerobic digestion of sulphate-containing wastewaters. Water Sci. Technol. 45, 231–235. Colleran, E., Pistilli, A., 1994. Activity test system for determining the toxicity of xenobiotic chemicals to the methanogenic process. Ann. Microbiol. Enzymol. 44, 1–20. Collins, G., Woods, A., McHugh, S., Carton, M.W., O’Flaherty, V., 2003. Microbial community structure and methanogenic activity during start-up of psychrophilic anaerobic digesters treating synthetic industrial wastewaters. FEMS Microbiol. Ecol. 46, 159–170. Collins, G., Foy, C., McHugh, S., O’Flaherty, V., 2005. Anaerobic treatment of 2,4,6-trichlorophenol in an expanded granular sludge bed-anaerobic filter (EGSB-AF)
ARTICLE IN PRESS 1620
G. Collins et al. / Water Research 39 (2005) 1614–1620
bioreactor at 15 1C. FEMS Microbiology Ecology in press (Online 21st November, 2004). Conrad, R., Wetter, B., 1990. Influence of temperature on energetics of hydrogen metabolism in homoacetogenic, methanogenic and other anaerobic bacteria. Arch. Microbiol. 155, 94–98. Eroglu, V., Ozturk, I., Ubay, G., Demir, I., Korkurt, E.N., 1994. Feasibility of anaerobic pre-treatment for the effluents from hardboard and laminated board industry. Water Sci Technol. 29, 391–397. Guieysse, B., Wickstro¨m, P., Forsman, M., Mattiasson, B., 2001. Biomonitoring of continuous microbial community adaptation towards more efficient phenol degradation in fed-batch reactor. Appl. Microbiol. Biotechnol. 56, 780–787. Khan, K.A., Suidan, M.T., Cross, W.H., 1981. Anaerobic activated carbon filter for the treatment of phenol-bearing wastewater. J. Water Poll. Control Fed. 53, 1519–1532. Lettinga, G., Rebac, S., Parshina, A., Nozhevnikova, J.B., van Lier, J.B., Stams, A.J.M., 1999. High-rate anaerobic treatment of wastewater at low temperatures. Appl. Environ. Microbiol. 65, 1696–1702. Lettinga, G., Rebac, S., Zeeman, G., 2001. Challenge of psychrophilic anaerobic wastewater treatment. Trends Biotechnol. 19, 363–370. Li, Y.Y., Fang, H.H.P., Chui, H.K., Chen, T., 1996. UASB treatment of wastewater with concentrated benzoate. J. Environ. Eng. ASCE. 12, 401–411.
Loehr, R.C., Krishnamoorthy, R., 1988. Terrestrial bioaccumulation potential of phenolic compounds. Haz. Waste Haz. Mater. 5, 109–119. McHugh, S., Carton, M.W., Collins, ., O’Flaherty, V., 2004. Reactor performance and microbial community dynamics during anaerobic biological treatment of wastewaters at 16–37 1C. FEMS Microbiol. Ecol. 48, 369–378. Nozhevnikova, A.N., Rebac, S., Kotsyrurbenko, O.R., Parshina, S.N., Holliger, C., Lettinga, G., 2000. Anaerobic production and degradation of volatile fatty acids in low temperature environments. Water Sci. Technol. 41 (12), 39–46. Perry, R.H., Green, D.W. (Eds.), 1984. Perry’s Chemical Engineers’ Handbook (Section 14), 6th ed. McGraw-Hill, New York. Rebac, S., Ruskova, J., Gerbens, S., van Lier, J.B., Stams, A.J.M., Lettinga, G., 1995. High-rate anaerobic treatment of wastewater under psychrophilic conditions. J. Ferm. Bioeng. 80, 499–506. Rebac, S., Gerbens, S., Lens, P.N., van Lier, J.B., Stams, A.J.M., Keesman, K.J., Lettinga, G., 1999. Kinetics of fatty acid degradation by psychrophilically grown anaerobic granular sludge. Bioresource Technol. 69, 241–248. Shelton, D.R., Tiedje, J.M., 1984. General method for determining anaerobic biodegradation potential. Appl. Environ. Microbiol. 47, 850–857. US EPA, 1980. Treatability Manual, US Environmental Protection Agency, Washington, DC.
Water Research 39 (2005) 1614–1620 www.elsevier.com/locate/watres
Anaerobic biological treatment of phenolic wastewater at 15–18 1C Gavin Collins, Clare Foy, Sharon McHugh1, The´re`se Mahony, Vincent O’Flaherty Microbial Ecology Laboratory, Department of Microbiology and Environmental Change Institute (ECI), National University of Ireland, Galway (NUI, Galway), University Road, Galway, Ireland. Received 24 November 2004; received in revised form 20 January 2005; accepted 25 January 2005 Available online 17 March 2005
Abstract Low-temperature, or psychrophilic (o20 1C) anaerobic digestion has been proven feasible for the mineralisation of simple wastewaters. In this study, hybrid expanded granular sludge bed-anaerobic filter (EGSB-AF) bioreactors were used to evaluate the feasibility of psychrophilic digestion for the treatment of phenol-containing wastewater. Efficient chemical oxygen demand and phenol removal were observed at organic and phenol loading rates of 5 kg COD m 3 d 1 and 0.4–1.2 kg phenol m 3 d 1 (400–1200 mg phenol [l wastewater] 1), respectively. There was no long-term accumulation of volatile fatty acids in the reactor systems. Methanogenic activity was developed under psychrophilic conditions but anaerobic methane-producing populations remained mesophilic throughout the trial of 415 days. r 2005 Elsevier Ltd. All rights reserved. Keywords: Psychrophilic anaerobic digestion; Phenol; EGSB; Specific methanogenic activity.
1. Introduction Phenolic compounds are present in the liquid effluent of coal gasification plants, coking plants, petroleum refineries, pharmaceutical, fertiliser and dye manufacturing plants, as well as degreasing and paint stripping operations (Khan et al., 1981) and fibreboard manufacturing (Eroglu et al., 1994). Although not found to be bioaccumulative (Loehr and Krishnamoorthy, 1988), humans exposed to phenol in well water at concentraCorresponding author. Tel.: +353 91 493734; fax: +353 91 525700. E-mail address: vincent.ofl[email protected] (V. O’Flaherty). 1 Present address: Environmental Microbiology Research Unit, Department of Microbiology, NUI, Galway, University Road, Galway, Ireland.
tions of 1300 mg l 1 exhibited a statistically significant increase in diarrhoea, mouth sores, dark urine and burning of the mouth (US EPA, 1980). Anaerobic digestion of phenolic wastewaters enjoys increasing application, and upflow anaerobic sludge blanket (UASB) and expanded granular sludge bed (EGSB) reactor configurations have been successfully demonstrated for the remediation of this category of effluent (Chang et al., 1995; Li et al., 1996). Continued advances in this field, however, are dependent on concerted efforts to improve and amend existing reactor configurations to design next generation technologies. In this context, the application of EGSBbased state-of-the-art reactor designs currently plays a significant role in the search for broader and newer applications of anaerobic digestion, and the successful widespread adoption of anaerobic reactors for the treatment of the vast and important range of phenolic
0043-1354/$ - see front matter r 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2005.01.017
ARTICLE IN PRESS G. Collins et al. / Water Research 39 (2005) 1614–1620
discharges is strongly dependant on the bioengineering ingenuity required to provide increasingly efficient and versatile technology. Further, and in parallel to this, there is growing interest in maximising methane extraction for energy recovery from anaerobic treatment processes. As the popularity of anaerobic digestion increases, so too does the urgent demand for greater economic efficiency from these processes. This requirement has prompted recent study into the feasibility of low-temperature reactor operations (Rebac et al., 1995; Rebac et al., 1999; Collins et al., 2003). One of the most important parameters for anaerobic treatment of wastewater is operating temperature. The mesophilic range is traditionally used since it is generally thought that maintaining a high (thermophilic) temperature is uneconomical for many wastestreams, whereas degradation within the psychrophilic range is too slow. However, efficient degradation, which was comparable to that from mesophilic trials, has been demonstrated for a range of wastewater types (Rebac et al., 1995; Lettinga et al., 1999; Rebac et al., 1999; Nozhevnikova et al., 2000; Collins et al., 2003; McHugh et al., 2004). Furthermore, the majority of industrial effluents released for treatment in temperate climates are below 20 1C. It follows, therefore, that psychrophilic anaerobic treatment, which returns satisfactory methane production, presents these countries with an attractive and economically sound option for sustainable remediation regimes. The aim of this study was to evaluate the feasibility of anaerobic biological treatment of phenolic wastewaters under ambient conditions (15–18 1C).
1615
Table 1 Particulars of wastewater components Phase/day
I II III IV V VIa
0–69 70–138 139–246 247–354 355–395 396–415
Phenol (mg l 1)
COD (g l 1) Phenol
VFA
Total
R1 R2
R1 R2
R1
R2
R1 R2
0 0 0 0 0 0
0 0 400 800 1200 1200
0 0 0 0 0 0
0 10 10 10 10 0 5 5 5 5 0.96 5 4.04 5 5 1.92 5 3.08 5 5 2.88 5 2.12 5 5 2.88 5 2.12 5 5
a Phase VI was characterised by the reduction in reactor operating temperature from 18 to 15 1C
A mesophilic, anaerobic granular sludge was obtained from a full-scale (1500 m3) internal circulation (IC) reactor, at Carbery Milk Products, Ballineen, Co. Cork, Ireland. The small (Ø, 0.4–0.8 mm), black and grey granules were of a regular form and settled well in liquid phase.
fatty acid (VFA)-based, industrial wastewater (pH 7.570.2) consisting of ethanol, butyrate, propionate and acetate, in the COD ratio of 1:1:1:1, to a total of 10 g COD l 1. The influent was buffered with NaHCO3 and fortified, as described by Shelton and Tiedje (1984), with macro- (10 ml l 1) and micro- (1 ml l 1) nutrients. The trial period was divided into six operational phases as detailed in Table 1. A loading rate of 10 kg COD m 3 d 1 was applied to each reactor, with a hydraulic retention time (HRT) of 24 h, and effluent was recycled at a rate of 5 m h 1. The loading rate was reduced to 5 kg COD m 3 d 1, on day 70, when the COD concentration of the influent was decreased to 5 g COD l 1, in order to advance the start-up period of the reactors. The liquid upflow velocity applied through the recirculation facility was increased to 7.5 m h 1 and to 10 m h 1, on days 34 and 98, respectively. The operational temperature was maintained at 18 1C until day 395 and at 15 1C for the remainder of the 415-day trial period. R2 influent was supplemented with phenol (Sigma), to a final concentration of 400 mg l 1, on day 138 (Table 1). Phenol was not added to R1, thereby maintaining this reactor as an experimental control. Phenol concentration in R2 influent was increased to 800 mg l 1 and 1200 mg l 1, on days 246 and 354, respectively.
2.2. Reactor set-up and operation
2.3. Biogas and effluent analysis
Two 3.5 l glass laboratory-scale expanded granular sludge bed (EGSB)-based reactors (R1 and R2), which were of the same design as that described by Colleran and Pender (2002), with the addition of an upper fixedfilm section, which was randomly packed with polyethylene rings of 1 cm diameter each, were used here. These hybrid expanded granular sludge bed-anaerobic filter (EGSB-AF) reactors were each inoculated with 70 g volatile suspended solids (VSS) of the seed sludge. The reactors were used for the stabilisation of a volatile
Samples of reactor effluent were taken for VFA and COD analysis and biogas was sampled for CH4 determination (American Public Health Association (APHA), 1992). In addition, samples of reactor liquor were retrieved from a sampling port located mid-way up the reactor column for COD analysis. Effluent phenol concentrations were ascertained using a colorimetric spectrophotometer (Odyssey DR/2500, Hach) and the 4aminoantipyrine technique (APHA, 1992). Briefly, this method involved the reaction of phenols with
2. Materials and methods 2.1. Biomass
ARTICLE IN PRESS G. Collins et al. / Water Research 39 (2005) 1614–1620
4-aminoantipyrine in the presence of potassium ferricyanide, to form a coloured antipyrine dye. The dye was extracted from the aqueous dilutions of effluent samples with chloroform and the colour was measured at 460 nm. 2.4. Batch specific methanogenic activity, biodegradability and toxicity assays
COD Removal Efficiency (%)
Maximum specific methanogenic activity (SMA) profiles of the seed sludge and the biomass from both reactors at the conclusion of the trial (day 415) were determined for the substrates acetate, propionate, butyrate, ethanol, phenol and H2/CO2 using the pressure transducer technique (Colleran and Pistilli, 1994; Coates et al., 1996). In addition, substrate utilisation rate (SUR)-based biodegradability tests were used to assess the biodegradability of the test chemical (phenol) using the seed sludge and biomass from both the control and test reactors. Biomass samples were incubated (at 15 1C or 37 1C) with oxygen-free phenolic solutions and anaerobic test buffer, which was fortified with microand macro-nutrients (Shelton and Tiedje, 1984). Substrate (phenol) depletion was determined by collecting homogenous samples from vials, as a function of time. Phenol concentrations were determined as described above. Biodegradability tests were carried out at an initial concentration of 200 mg phenol l 1 and SUR values were expressed as the rate of phenol depletion (mg phenol [g VSS] 1 d 1). SUR-based tests were carried out in 60 ml sealed hypovials to a total volume of 30 ml. Phenol toxicity was assessed using the SMA-based toxicity assay described by Colleran and Pistilli (1994). Phenol-induced inhibition of methanogenic (acetoclastic and hydrogenotrophic) and syntrophic populations was
100 90 80 70 60 50 40 30 20 10 0
determined for seed sludge and reactor biomass samples. Toxicity was defined in terms of the IC50 value, i.e. the concentration of phenol that resulted in 50% inhibition of the control vial SMA, which was calculated from the linear regression of SMA as a function of phenol concentration. All activity, biodegradability and toxicity test vials contained 2–5 g VSS l 1.
3. Results 3.1. Reactor performance A lengthy start-up period, of over 135 days, was recorded for both reactors, after which COD removal efficiencies of over 80% (at the applied loading rate of 5 kg COD m 3 d 1) were recorded (Fig. 1). The dominant VFA present in the reactor effluents during phases I and II were acetate and propionate (Fig. 2). Upon supplementation of R2 influent with phenol (Table 1), reduced COD removal (Fig. 1) and biogas methane composition (data not shown) were observed. However, by the end of phase III, decreased VFA concentrations in R2 effluent coincided with increased phenol removal. After day 230, comparable performance data were recorded for both reactors (Fig. 1). There was a decline in R2 COD removal upon the application of the 0.8 kg phenol m 3 d 1 loading rate on day 247, which was accompanied by an accumulation of phenol in the reactor effluent (Fig. 1). However, by day 260, there was an evident improvement in the phenol removal rate and COD removal equal to R1 was recorded by day 340 (Fig. 1). Sporadic accumulation of acetate was observed in R1 effluent (Fig. 2) Nevertheless, R1 acetate levels did not exceed 500 mg l 1 at any stage and propionate
400 350 300 250 200 150 B
C
D
A 0
50
100 150 200 250 300 350 400 Time (Days)
100 50
Effluent Phenol (mg/l)
1616
0
Fig. 1. COD removal efficiency of R1 (’) and R2 (J), and R2 effluent phenol concentration (m). Applied phenol loading rate: (A) 0.4 kg phenol m 3 d 1; (B) 0.8 kg phenol m 3 d 1; (C) 1.2 kg phenol m 3 d 1. (D) Operating temperature of R1 and R2 reduced to 15 1C.
ARTICLE IN PRESS G. Collins et al. / Water Research 39 (2005) 1614–1620
II
I
III
IV
V
VI
VFA Concentration (mg/l)
2200 2000 1800 1600 1400 1200 1000 800 600 400 200 0 (A) 0
Table 2 Specific methanogenic activity (ml CH4 [g VSS] seed sludge and R1 and R2 biomass
Ethanol Acetate Propionate Butyrate H2/CO2 Phenol
50
100
150
200
250
300
350
1617
1
d 1) of the
Seed sludge
R1
R2
15 1C
37 1C
15 1C
37 1C
15 1C
37 1C
312 140 39a 158 40.3 0.6
553 285 110 516 52 0
627 598 202 967.5 610 2.8
889 824 705 1215 843 3.6
371.4 321.4b 146c 496 347 87
678.2 398.9 209 651 510 120
400
VFA Concentration (mg/l)
a
(B)
2200 2000 1800 1600 1400 1200 1000 800 600 400 200 0
Lag 120 h. Lag 73 h. c Lag 382 h. b
0
50
100
150
200 250 300 Time (Days)
350
400
Fig. 2. (A) R1 and (B) R2 effluent VFA concentrations; acetate (’), ethanol (m), propionate (J) and butyrate (K). Timeline illustrates the six operational phases studied.
concentrations remained steadily low throughout. In contrast, an accumulation of acetate and propionate, to concentrations of 1000 and 1400 mg l 1, respectively, was observed in R2 effluent during the initial days of phase IV. However, an abatement of VFA accumulation, in parallel with increased phenol degradation, was observed by day 270, signalling a stabilisation of R2 performance. R2 COD removal efficiency decreased again after the addition of phenol to a concentration of 1200 mg l 1 (Fig. 1). Propionate, at maximum levels of 1200 mg l 1, constituted the predominant VFA in R2 effluent (Fig. 2). The recovery period recorded after this increase in the phenol loading to 1.2 kg phenol m 3 d 1 was shorter than those observed after day 138 (0.4 kg phenol m 3 d 1) and 247 (0.8 kg m 3 d 1). Improved R2 COD removal was evident within 40 days of instigating this increment and the performance of both digesters was comparable by day 390 (Fig. 1). Modest increases in VFA concentrations were observed in R1 effluent after the temperature decrement to 15 1C and were accompanied by a temporary accumulation of butyrate (Fig. 2). On the other hand, substantially greater concentrations of acetate and propionate
(200 and 600 mg l 1, respectively) were detected in R2 effluent on the days after the temperature change. However, a rapid recovery period of about 10 days was recorded whereby R2 performance data were comparable to those from the control reactor by day 405. Data from COD and phenol assays using the mixed reactor liquor sampled from the port mid-way up the reactor columns indicated that the anaerobic biofilm, which developed on the polyethylene rings, contributed 473% to overall COD removal in both reactors and 371.5% to total phenol removal by R2. 3.2. Metabolic characteristics of biomass as determined by specific methanogenic activity, biodegradability and toxicity assays The seed granules exhibited high SMA values when tested within the mesophilic range (37 1C) (Table 2). However, relatively low hydrogenotrophic activity was recorded at 37 1C and low SMA values against phenol suggested poor degradation of the test compound (Table 2). Lower SMA values were recorded, against all substrates when the assays were carried out within the psychrophilic range (15 1C), thereby confirming the mesophilic status of the reactor inoculum (Table 2). R1 SMA values at the conclusion of the trial were substantially higher than the respective values collated for the seed sludge at both 37 1C and 15 1C. High SMA values for H2/CO2 utilisation were recorded from R1, thereby suggesting satisfactory development of the hydrogenotrophic methanogenic community (Table 2). In fact, the R1 SMA values at 15 1C were higher than those recorded for the seed sludge at 37 1C, thus pointing towards substantial development of the methanogenic activity of this biomass against all substrates tested (except phenol). However, the optimum temperature of all of the trophic groups evaluated here remained within
ARTICLE IN PRESS G. Collins et al. / Water Research 39 (2005) 1614–1620
1618
Table 3 Substrate utilisation rates (SURs) for seed sludge and reactor biomass samples using phenol Biomass
SUR (mg phenol [g VSS]
1
d 1)
15 1C
37 1C
Seed Sludge 13.97 R1 sludge (Day 415) 15.4 R2 sludge (Day 415) 88.96
19.9 24.9 165.1
Table 4 IC50 values for the methanogenic and syntrophic (propionatedegrading) microorganisms in seed sludge, R1 and R2 biomass. Values are mg l 1 Seed sludge
R1
R2
Acetoclastic
15 1C 37 1C
305 341
426.9 436
1181 1410
Hydrogenotrophic
15 1C 37 1C
331 761
840 942.4
1147 1681
Propionate degraders
15 1C 37 1C
121.6 199
147 217
231 349.3
otrophs (Table 4). The low SMA of the seed and R1 sludges against phenol as substrate (Table 2), and the poor biodegradability of phenol in batch assays were confirmed by low toxicity thresholds observed within both the psychrophilic and mesophilic ranges (Table 4). However, higher IC50 values were obtained for R2 granules at the conclusion of the trial (Table 4). This was despite the observation that R2 SMA values were generally lower than those determined for R1 (Table 2). The reduced sensitivity (higher IC50) to the presence of phenol exhibited by all three populations tested was in agreement with the increased SMA against phenol (Tables 2 and 4). The toxicity profile obtained for R2 sludge suggested that phenol generally inhibited the trophic groups in the same order as in both the seed and R1 sludges, i.e. propionate-degraders IC50 valueoacetoclastic methanogensohydrogenotrophic methanogens. Finally, all IC50 values calculated for the reactor biomass were higher when tested within the mesophilic range than at 15 1C, indicating greater sensitivity to the test compound at lower temperatures and thus broadly supporting the conclusion that the reactor biomass remained substantially mesophilic, despite some evidence of psychrotolerance.
4. Discussion the mesophilic range (Table 2), suggesting psychrotolerant, rather than psychrophilic populations within the biomass. A similar SMA profile, in terms of temperature optima, was obtained from R2 biomass at day 415. As with R1, the activity values recorded for R2 sludge at 15 1C were greater than those determined for the seed within the mesophilic range (for all substrates except butyrate). The development of phenol-degrading activity was also evident, reflecting the efficient phenol mineralisation of the R2 reactor. All SMA values recorded for this biomass were lower than corresponding R1 values, indicating inhibition of the development of methanogenic activity in R2. Low-substrate utilisation rates (SURs) were calculated for both the seed and R1 biomass (Table 3). Although poor substrate depletion was observed from the unacclimated seed and R1 sludges, the mesophilic optimum of the biomass was evident from these tests (Table 3). As expected, higher substrate depletion rates were recorded for R2 biomass, at both 15 1C and 37 1C (Table 3). Although a mesophilic optimum was also obvious for this sludge, only 6.1% of the initial substrate introduced to the vials incubated at 15 1C was present at the conclusion of the tests. IC50 values indicated that propionate-utilising organisms were most sensitive to phenol, while the inhibition order among methanogens was acetoclasts4hydrogen-
It can be said from the results of this study that efficient phenol degradation was achieved in a psychrophilic anaerobic reactor operated at a constant temperature of 18 1C. In addition, the successful operation of the phenol-mineralising reactor at 15 1C (in phase VI) serves to confirm the feasibility of this treatment approach and further underlines the potential for costeffectiveness and rationalisation of many economically non-sustainable remediation practices. The principle VFA detected in the effluent of the test reactor immediately after the drop in operating temperature to 15 1C was propionate. Previous research has indicated that propionate-degradation may be the rate-limiting step in the psychrophilic anaerobic digestion process (Rebac et al., 1999; Collins et al., 2003). However, stable propionate degradation was observed within 10 days of the temperature reduction applied to both reactors. Furthermore, the development of propionate-degrading activity was also observed (Table 2). Reactor performance was found to improve with the number of phenol amendments, thus showing that microbial adaptation can occur throughout the trial. The length of the acclimation phase decreased after each phenol amendment, from initially being 100 days (during the 107 days of phase II) to 97 days (during phase III) and then 20 days (during the 40 days of phase IV). This is in agreement with the findings of Guieysse et al. (2001) and confirms that the use of a mesophilic
ARTICLE IN PRESS G. Collins et al. / Water Research 39 (2005) 1614–1620
granular sludge as an inoculum for psychrophilic acclimation of suitable degraders and digestion of phenolic wastewater is a feasible treatment option. Interestingly, the contribution to overall COD and phenol removal of the upper fixed-film section was low. This may be due to (1) low levels of biomass washout from the sludge bed and consequent lack of seed biomass for biofilm development and/or (2) the low COD concentration of reactor liquor entering this section from the sludge bed. Nevertheless, the benefits of employing this reactor design has been recently demonstrated in a 2,4,6-trichlorophenol-degrading reactor, where partially dechlorinated phenols from the sludge bed section were degraded by fixed-film consortia in the upper section of that reactor (Collins et al., 2005). The SMA profiles collated for both reactor sludges in this study indicated the development of methanogenic activity, even under low-temperature conditions. In fact, R1 SMA values at 15 1C were higher than those recorded for the seed sludge at 37 1C. However, the optimum temperature of all of the trophic groups evaluated here remained within the mesophilic range (Table 2), suggesting psychrotolerant, rather than psychrophilic populations within the biomass and thus supporting the conclusions of Rebac et al. (1999), Collins et al. (2003) and McHugh et al. (2004). Furthermore, reduced methanogenic development was observed in R2 sludge, which pointed towards inhibition of some or all of the trophic groups involved in the processes of this reactor. In addition, increased activity against H2/CO2 was observed at the conclusion of the trial. This phenomenon has also been observed during previous low-temperature treatment trials (McHugh et al., 2004; Collins et al., 2005). Notwithstanding the presence of hydrogenic substrates (ethanol, propionate, butyrate and phenol) in the influent wastewater in this study, the reason for high hydrogenotrophic methanogenic activity in low-temperature reactors is largely unknown. A possible explanation is based on thermodynamic considerations, as lowering the operating temperature may favour hydrogenotrophic methanogenesis (Conrad and Wetter, 1990), as the H2 threshold concentrations for hydrogenotrophic methanogens are significantly lowered under psychrophilic conditions. Furthermore, due to the increased solubility of gases at low temperatures (Perry and Green, 1984; Lettinga et al., 2001), it is possible that elevated dissolved H2 levels were present within the mixed liquor of R1 and R2, encouraging the growth of hydrogenotrophic methanogens. It is clear that many questions remain to be answered in this field. One such question is whether the apparent shift towards hydrogenotrophic methanogenesis in psychrophilic anaerobic reactors is primarily a temperature-related issue or is simply a response to environmental stress. In addition, a greater insight to the degradation of propionate in psychrophilic environ-
1619
ments—and the development of reactor designs to overcome propionate accumulation—is desirable.
5. Conclusions The following conclusions can now be drawn: (i) psychrophilic anaerobic digestion of phenolic wastewaters is feasible at 15–18 1C and at applied loading rates of up to 1.2 kg phenol m 3 d 1; (ii) further research is required to establish the upper limits of phenol loading for low-temperature anaerobic treatment; (iii) the use of an unacclimated mesophilic sludge is a satisfactorily efficient means of cultivating a phenoldegrading consortium for psychrophilic reactor operation—this is of particular importance if phenol-acclimated, psychrophilically cultivated biomass is not readily available.
Acknowledgements Financial support of the Higher Education Authority (HEA) Programme for Research in Third Level Institutes (PRTLI)-Cycle II, through the Environmental Change Institute (ECI), NUI, Galway, and an Enterprise Ireland Research Scholarship to G.C. are gratefully acknowledged. Paula McDermott is thanked for her assistance with data collation.
References American Public Health Association, 1992. Standard methods for the examination of water and wastewater, 18th ed. Washington, DC. Chang, B., Chen, K., Yuan, S., 1995. Dechlorination of 2,4,6TCP by an anaerobic mixed culture. Chemoshpere 31, 3803–3811. Coates, J.D., Coughlan, M.F., Colleran, E., 1996. Simple method for the measurement of the hydrogenotrophic methanogenic activity of anaerobic sludges. J. Microbiol. Methods 26, 237–246. Colleran, E., Pender, S., 2002. Mesophilic and thermophilic amaerobic digestion of sulphate-containing wastewaters. Water Sci. Technol. 45, 231–235. Colleran, E., Pistilli, A., 1994. Activity test system for determining the toxicity of xenobiotic chemicals to the methanogenic process. Ann. Microbiol. Enzymol. 44, 1–20. Collins, G., Woods, A., McHugh, S., Carton, M.W., O’Flaherty, V., 2003. Microbial community structure and methanogenic activity during start-up of psychrophilic anaerobic digesters treating synthetic industrial wastewaters. FEMS Microbiol. Ecol. 46, 159–170. Collins, G., Foy, C., McHugh, S., O’Flaherty, V., 2005. Anaerobic treatment of 2,4,6-trichlorophenol in an expanded granular sludge bed-anaerobic filter (EGSB-AF)
ARTICLE IN PRESS 1620
G. Collins et al. / Water Research 39 (2005) 1614–1620
bioreactor at 15 1C. FEMS Microbiology Ecology in press (Online 21st November, 2004). Conrad, R., Wetter, B., 1990. Influence of temperature on energetics of hydrogen metabolism in homoacetogenic, methanogenic and other anaerobic bacteria. Arch. Microbiol. 155, 94–98. Eroglu, V., Ozturk, I., Ubay, G., Demir, I., Korkurt, E.N., 1994. Feasibility of anaerobic pre-treatment for the effluents from hardboard and laminated board industry. Water Sci Technol. 29, 391–397. Guieysse, B., Wickstro¨m, P., Forsman, M., Mattiasson, B., 2001. Biomonitoring of continuous microbial community adaptation towards more efficient phenol degradation in fed-batch reactor. Appl. Microbiol. Biotechnol. 56, 780–787. Khan, K.A., Suidan, M.T., Cross, W.H., 1981. Anaerobic activated carbon filter for the treatment of phenol-bearing wastewater. J. Water Poll. Control Fed. 53, 1519–1532. Lettinga, G., Rebac, S., Parshina, A., Nozhevnikova, J.B., van Lier, J.B., Stams, A.J.M., 1999. High-rate anaerobic treatment of wastewater at low temperatures. Appl. Environ. Microbiol. 65, 1696–1702. Lettinga, G., Rebac, S., Zeeman, G., 2001. Challenge of psychrophilic anaerobic wastewater treatment. Trends Biotechnol. 19, 363–370. Li, Y.Y., Fang, H.H.P., Chui, H.K., Chen, T., 1996. UASB treatment of wastewater with concentrated benzoate. J. Environ. Eng. ASCE. 12, 401–411.
Loehr, R.C., Krishnamoorthy, R., 1988. Terrestrial bioaccumulation potential of phenolic compounds. Haz. Waste Haz. Mater. 5, 109–119. McHugh, S., Carton, M.W., Collins, ., O’Flaherty, V., 2004. Reactor performance and microbial community dynamics during anaerobic biological treatment of wastewaters at 16–37 1C. FEMS Microbiol. Ecol. 48, 369–378. Nozhevnikova, A.N., Rebac, S., Kotsyrurbenko, O.R., Parshina, S.N., Holliger, C., Lettinga, G., 2000. Anaerobic production and degradation of volatile fatty acids in low temperature environments. Water Sci. Technol. 41 (12), 39–46. Perry, R.H., Green, D.W. (Eds.), 1984. Perry’s Chemical Engineers’ Handbook (Section 14), 6th ed. McGraw-Hill, New York. Rebac, S., Ruskova, J., Gerbens, S., van Lier, J.B., Stams, A.J.M., Lettinga, G., 1995. High-rate anaerobic treatment of wastewater under psychrophilic conditions. J. Ferm. Bioeng. 80, 499–506. Rebac, S., Gerbens, S., Lens, P.N., van Lier, J.B., Stams, A.J.M., Keesman, K.J., Lettinga, G., 1999. Kinetics of fatty acid degradation by psychrophilically grown anaerobic granular sludge. Bioresource Technol. 69, 241–248. Shelton, D.R., Tiedje, J.M., 1984. General method for determining anaerobic biodegradation potential. Appl. Environ. Microbiol. 47, 850–857. US EPA, 1980. Treatability Manual, US Environmental Protection Agency, Washington, DC.