J . Chrm. Trch. Biotrchriol. 1990, 47, 219-233
Landfill Co-disposal of Phenol-bearing Wastewaters: Organic Load Considerations Irene A. Watson-Craik & Eric Senior Department of Bioscience and Biotechnology, Applied Microbiology Division, University of Strathclyde, 204 George Street, Glasgow G1 lXW, UK (Received 21 November 1988; revised version received 13 March 1989; accepted 10 April 1989)
ABSTRACT A multi-stage model, operated with single elution, was used to investigate the effrcts of organic loadings on the attenuation of a model phenolic wastewater in domestic refuse. Although 100 % dissimilation of influent phenol (25 mmol dm - 3, was recorded at a dilution rate of 0.007 h - partial inhibition of both phenol degradation and species competing with methanogens for a common electron donor(s) was apparent at concentrations 3 4 mmol dm-3. On extended perfusion with 8 mmol phenol dm- 3, the progressive inhibition of phenol dissimilation was not obviated by nutrient supplementation. Simultaneous degradation of the catabolic intermediate, hexanoic acid, and elevated methane release rates suggested that the transformation of phenol to hexanoate was rate limiting.
',
Key words: codisposal, phenol, refuse, organic load, landfill, met hanogenesis.
1 INTRODUCTION
In the UK, the official endorsement of hazardous waste codisposal with refuse in landfill sites' and the relative cheapness of the method' ensure the continuation of the practice, which accounts for the disposal of more than 78 % of all hazardous It is unfortunate, however, that guidelines on permissible loading rates have been issued in the absence of definitive studies on the performance of landfill sites as bioreactors. The landfill bioreactor contains a wide range of natural and xenobiotic molecule^,^ some or all of which are susceptible to microbial challenge. The 219 J . Chrm. Tech. Bioteclrriol.0268-2575/90/%03.50 0 1990 Society of Chemical Industry. Printed in Great
Britain
220
1. A . Warsou-Craik.E . Scwior
ecosystem is dominated by irregularly distributed solid surface components and liquid (leachate) of changing composition. Consequently, physicochemical parameters such as pH, E,,, a, and temperature vary both temporally and spatially. The complexity and dynamics of the environment thus militate against in-situ research, and necessitate the development of heterogeneous laboratory models.’ Studies have been made with conventional small-scale (0.11 m3)6and pilot-scale (8 m3)7systems but, although the latter have been employed to examine leachate reactions, their high capital cost, considerable operator input, protracted operational period’ and difficulties in maintaining anaerobiosis7 have limited their value. Previous co-disposal studies have focused on the effects of increased loadings of heavy metals on refuse metabolism.’.’ Work on phenol codisposal, with specific reference to hydraulic loading rates, has also been reported’ although organic loads have not been considered. The aim of the present study, using a small-scale multi-stage system, was to examine the impacts of phenol organic loads on refuse metabolism. Phenol was selected as being the simplest aromatic core molecule characteristic of the high-volume wastewaters produced in coke, fibreboard, phenolic resin and nylon manufacture.
’
2 EXPERIMENTAL MATERIALS AND METHODS 2.1 Refuse
Refuse, from a depth of 2 2 m and emplaced 2 1 month, was sampled from Kilgarth Landfill (Glasgow). Prior to use, it was handsorted to remove visible stones, glass, metals and wood, then homogenised for 30 s in a domestic blender (Krups). 2.2 Molecule
Stock solutions of phenol (100 mmol dm-3 in glass-distilled water) were used to modify influent phenol concentrations.
2.3 Mineral salts solution The mineral salts solution used was the same as that described by Coutts et al.” 2.4 Multi-stage refuse column
Eight glass columns (length 90 cm, internal diameter 5 5 cm), each filled with 1 kg uniformly compacted refuse and insulated with domestic quality (1 5 mm) pipelagging, were serially linked with oxygen-impermeable butyl rubber tubing (Esco Rubber, Feltham) (Fig. 1). Sampling ports ( l d ) , fitted with three-way sampling valves, were sited at discrete intervals down the array. The model was maintained at ambient temperature for 67 days before influent substrate was introduced to the top of column (a) by use of a Watson-Marlow 202U Flow Inducer. After percolating under gravity through columns (a) and (b) the resulting leachate was pumped to the top of the next column. This serial perfusion was continued through the array. A mean empty-bed dilution rate ( D ) of 0026 h-’ was imposed on column (a)
Oryuiiic loud qf1i.m 011 co-disposal
22 I
7 a.
Y
L Fig. 1. Multi-stage refuse column array, with sampling ports 1-6 and temperature probes i and ii.
(sampling port 1 ) and the resultant dilution rates at ports 2-5 were 0.013, 0.007, 0.004 and 0003 h-', respectively. Any transient over- or under-pressuresin the column headspaces were modulated by the insertion into columns (a), (c), (e) and (g) of syringes with sliding barrels.
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I. A . Watson-Craik, E . Settior
2.5 Sample preservation and storage Leachate samples for sulphate, phosphate or residual phenol analysis were stored at - 10°C in polypropylene bottles for < 7 days. Samples (0.9 cm3)for volatile fatty acid (C, to C,) analysis were acidified with formic acid (0.1 cm3) (BDH, Aristar) before storage in glass vials at - 10°C for 6 30 days.
2.6 Analytical methods 2.6.1 Phenols Liquid samples (0.001 pm3) were injected into a Perkin Elmer Sigma 115 gas chromatograph (GC), equipped with a flame ionisation detector, in which the temperatures of the injector, oven and detector were maintained at 200"C, 150°C and 210"C, respectively. The flow rate of the oxygen-free nitrogen (OFN) (British Oxygen) carrier gas was set at 40cm3 min-'. The glass column (length 2 m, i.d. 2 mm) was packed with 5 % polymeta phenyl ether with 6 rings on Chromosorb WAW (80-100 mesh) (Phase Separations). Leachate phenol and m-lpcresol concentrations were determined, after standard curve construction, by mean peak area comparison with phenol and pcresol standards (2 mmol dm-3). 2.6.2 Volatile fatty acids Leachate volatile fatty acid (VFA) concentrations were determined by use of the same GC.The flow rate of the carrier gas (OFN) was set at 30 cm3 min-' and the oven temperature, initially held at 100°C for 2 min, was programmed to increase to 150°C at a ramp rate of 30°C min-I. Acidified samples (0.001 pm3) were injected on to the glass column (length 2 m, i.d. 2 mm) which was packed with 5 % neopentyl glycol sebacate+ 1 % H3P0, on Anakrom polyester (80-100 mesh) (Chromatographic Services).Acidified solutions (10mmol dm-3) of fatty acid standards were used, and leachate VFA concentrations calculated by peak area comparison, after computer-generated baseline correction. 2.6.3 Sulphate After sparging leachate samples (10 cm3)for 10 min with OFN to remove sulphides, total sulphate concentrations were estimated by use of the standard BaCI, turbidometric method." 2.6.4 Phosphate Total acid-hydrolysable phosphate was determined by the standard ascorbic acid method.'
2.6.5 pH pH values of the liquid samples were recorded by use of a Pye Unicam PU9148 pH meter. 2.6.6 Refuse dry weight Triplicate fresh refuse samples (lOOg),dried to constant weight at 60°C were used to estimate refuse moisture content.
Orgunic loud ejjicts on co-disposul
223
2.6.7 Solubilised methane Triplicate universal bottles ($ ounce) were closed with suba-seals (Gallenkamp)and evacuated, for 30s, with a vacuum-pressure pump (Millipore XX60 22050). A saturated methane solution was prepared by sparging cold glassdistilled water (10cm3)with methane (CP grade, British Oxygen)for 30 min,I4 then tightly sealing the vessel. The solubilisation temperature was noted and the soluble methane concentration then determined from a standard curve." Samples ( 1 cm3) of the standard solution were injected into pre-evacuated universal bottles which were then equilibrated to atmospheric pressure with hypodermic needles. After sealing the bottles by needle withdrawal, triplicate gas phase samples (0.05pm3) were taken and injected into the GC under the operating conditions described for VFA analysis although, in this case, the oven temperature was held at 80°C. Leachate samples ( 1 cm3) were similarly processed and leachate methane concentrations calculated by comparison with the methane standard. The rates of methane release at each discrete dilution rate were calculated from the equation: r=
22.4jc(1 + l/a,) W
where r is the rate (cm3g refuse-' d - ') of methane release, f is the liquid flow rate (dm3 d - ' ) through the refuse, c is the leachatedissolved methane concentration (pmol ~ m - ~a,) is, the solubility coeflicient of methane" at ambient temperature t"C, and w is the cumulative refuse weight (fresh) (g).
3 RESULTS AND DISCUSSION 3.1 Introduction To facilitate realistic enrichments of microorganisms the columns were allowed to self-generate redox and pH gradients at temperatures which, as a result of column insulation, were 0.5-14T above ambient (Probes i and ii (Fig. 1)). Since previous studies16had shown that a pre-perfusion incubation period effected a 50% increase in phenol catabolism compared with controls, this strategy was again adopted. The parameters selected for monitoring included those known to act as indicators of refuse age and degree of stabilisation.' These comprised: refuse compaction/ density; leachate pH, sulphate and VFA concentrations; and methane release. Despite initial processing the refuse moisture capacity (64%)approximated to insitu values. Similarly, the applied refuse density (513 kg m-3) was comparable to site conditions." After 548 days perfusion the final density was 696 kg m - 3 due to 26 % (v/v) settlement.
3.2 Influent phenol: 2-6 mmol dm-3 The discrete phenol concentrations recorded in the leachate samples may have originated from three possible sources: the influent substrate; as a component of the refuse used; or as a product of refuse degradation. Phenol concentrations of 0.33 and 016 mmol dm-3 of leachate were recorded at sampling ports 1 and 5 after 4 and
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1 . A . Wurson-Craik, E . Senior
96 h perfusion, respectively. These concentrations were greater than might have
been expected if the model operated with plug-flow. It was possible that either the column array was operating on a partial plug-flow basis, due possibly to a limited degree of channelling, or phenol was displaced from the refuse. The equivalent pcresol concentrations of 0.38 and 0.10 mmol dm-3 were attributed to a comparable displacement of phenolics from the refuse. After 24 days perfusion the residual phenol concentration (port 5 ) progressively decreased and was no longer detected after 45 days (Fig. 2). A similar trend was observed with leachate sampled at ports 4 and 3, whilst over the subsequent 34 days the average residual phenol concentrations of leachate samples 2 and I approximated to 14% and 32% of the influent. On day 70 the influent phenol concentration was increased to 4 mmol dm-3. Within three days, elevated phenol concentrations were recorded at sampling ports 1 and 2, and within 10 days at ports 1 to 5 (Fig. 2). After day 88, however, residual concentrations progressively declined until by day 101 no phenol was detected in the final efiluent. On day 119 residual phenol (0-94mmol dm-3) was detected; traces remained detectable until day 136, and were coincident with low laboratory temperatures ( < 6°C). Subsequently, the temperature was increased (17-19°C) and phenol concentrations fell until by day 143 no residual phenol was detected in leachate sampled at ports 4 and 5.
Fig. 2. Changes in leachate residual phenol concentrations (mmol dm-’) at sampling ports I (0 ), 2 (a), 3 4 (m)and 5 (A) of the multi-stage refusecolumn array, perfused by 2 (a),4 (b),5 (c)and 6 (d)mmol phenol dm-3.
(a),
Biodegradation was adversely affected by reducing the temperature to < IO'C whilst a previous study had shown that a temperature increase from 21 1.5'C to 30' C had no significant effect." The bactericidal activity of phenol is promoted by temperature increase." Thus it is possible that at temperatures > 20°Cincreased metabolism is negated by the increased bactericidal activity. Clearly, landfill temperature is an operational parameter that may require optimisation in relation to phenol codisposal. When the influent concentration was increased to 5 mmol dm-3 (day 161) and 6 mmol dm-3 (day 244) no residual phenol was detected in leachates sampled at ports 4 and 5. Thus at an empty-bed dilution rate 60404 h - ' attenuation of 6 mmol phenol dmP3is possible. There are few data on phenol removal rates in other continuous anaerobic reactors. Removal (1OOXJ of influent 2.1 and 4.3 mmol phenol d m - 3 was recorded2' in an activated carbon-packed upflow system with recirculation, at a dilution rate of 0.01 1 h-I; even at an influent concentration of 10-6mmol dm-3, removal was 98.1 "/,. Increased inhibitory effects were recorded, however, in reactors supplemented with 1.6 and 3.2 mmol phenol dm-3 in which there was no provision of biomass supports.2' Although 100% removal was noted at dilution rates from 0.002 to 0403 h increasing emuent concentrations were recorded after periods from 17 (3.2 mmol dm-j) to 53 (1.6 mmol dm-3) days. The provision of refuse surfaces for biofilm attachment may thus be an integral factor in the efficacy of phenolic wastewater treatment by co-disposal. 3.3 p H and volatile fatty acids Leachate pH and VFA analyses showed that the increase in influent phenol from 2 to 4 mmol dm - 3 resulted in a pH decrease, within 10 days, from 6.55 to 5.70 at sampling port 1, although this was not coincident with elevated VFA concentrations. Only acetate, on days 80 and 89, at concentrations of 0.81 and 0.03 mmol dm-3, respectively, was recorded. By day 101, however, leachate pH (port 1) had increased to 6.4. However, when the phenol concentration was increased to 5, and, later, to 6 mmol dm ', the pH profiles did not show a similar dip and recovery but remained relatively constant; phenol degradation also continued smoothly. This indicated that no shock-loading had occurred under these conditions. 3.4 Methane From the concentrations of phenol removed from the influent, theoretical rates of methane release were calculated and compared with the actual rates recorded (Fig. 3). Phenol was assumed to be mineralised to CO, and CH, according to the equation:22 CnH,O,+(n-~/4-b/2)HZO
+ (n/2-~/8+b/4)CO,
+(n/2+~/8-h/4)CH,
(2) Such stoichiometric metabolism under defined non-SO,-supplemented conditions has been r e p ~ r t e d . ~In ~ ,the ' ~ present study, however, where leachate SO, concentrations between days 1, and 70 ranged from 0.02 to 056 mmol dm-3,
226
I . A . Watson-Craik, E . Senior
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(0) and theoretical ( 0 )phenolderived rates of methane release (cm3 g refuse-'d-')atsamplingports 1-5ofthemulti-stagemodel.perfusedwith2(a).4(b).5(c)or 6(d)mmol phenol dm-3. Sampling ports 1-5 correspond to dilution rates 0.026,0~013.0~007.0004 and 0403 h-', respectively. Fig. 3. Changes in recorded
electron flow towards sulphate reduction with a concomitant decrease in methanogenesis must also be considered. From Fig. 3a it can be seen that, in the presence of an influent phenol concentration of 2 mmol dm-3, the actual rates of methane release were higher than predicted. The discrepancy was assumed to reflect methanogenesis from refuse components. With time the discrepency decreased, possibly due to metabolism of
Organic loud yff~crs oti co-disposul
227
the pool of methane precursors. At the higher dilution rates, however, the observed rates were lower than were theoretically expected. This was possibly due to spatial separation of phenol degradation from methanogenesis, although previous studies with single-stage columns16 maintained at 30°C had shown that under these conditions the two metabolic processes were coincident. When the influent phenol concentration was increased to 4 mmol dmimmediate increases in the rates of methane release were apparent at all dilution rates, particularly at 0.026 h - (Fig. 3b). These increases were significantly greater than could be accounted for by phenol mineralisation and possibly resulted from a diversion of electron flow from sulphate-reduction to methanogenesis. A comparable inhibition of group(s) competing with methanogens for common electron donor@)and/or acceptor(s) was also observed in a phenol-supplemented (2 mmol dm- ’) veratric acid degrading association (K. K. Abdul-Halim, personal communication). At dilution rates > 0.013 h-’ reduced theoretical rates of phenol-derived methane release were apparent and were possibly indicative of partial inhibition of a metabolic group. At this stage it can only be speculated as to the identity of the group. Work with batch methanogenic fermentations by Fedorak & Hrudey (1984)*’ also demonstrated this inhibitory effect. They showed that methanogens were inhibited only at phenol concentrations 2 12.75mmol dm-3, whilst phenolcatabolising species were increasingly inhibited by concentrations > 5.3 mmol dm-’. Unfortunately, they were unable to determine if non-phenoldegrading species were adversely affected by the concentrations used. The effects of phenol concentration on rates of methane release were also examined with influent concentrations of 5 and 6 mmol dm-3 (Fig. 3c and 3d). With 5 mmol dm-’ phenol and dilution rates >0*007 h-’, methane release rates were significantly reduced (Fig. 3c). Although possibly due to partial inhibition of methanogenic activity, this was not confirmed by the presence of elevated concentrations of methanogenic precursors. However, at dilution rates < 0404 h - and in the presence of phenol concentrations < 0.71 mmol dm methanogenic activity was sustained. With an influent concentration of 6 mmol phenol dm-3, there was an initial dramatic increase in the methane release rate at the highest dilution (Fig. 3d), although by day 285 partial methanogenic inhibition at dilution rates 20.013h-’ was again apparent. At the lowest dilution rate, however, the observed methane release rate increased relative to the theoretical rate and reflected increased total phenol removal.
’,
’,
’,
3.5 Influent phenol: 8 mmol dm-3 On day 285 the influent phenol concentration was increased to 8 mmol dm-’. For the next 38 days no residual phenol was detected at sampling port 5 (Fig. 4) and it thus appeared that, with appropriate dilution rates, influent phenol concentrations < 8 mmol dm- could be effectively attenuated by single-elution through refuse. By day 330, however, residual phenol was recorded at ports 4 (0.56 mmol dm-’) and 5 (0.06mmol dm-7. These residual concentrations then progressively increased to 6.52, 5.01 and 3.25 mmol dm-3 at ports 3, 4 and 5, respectively (day
I . A . WofsowCrctik. E . Senior
280
(0 ] , 2 (@ 1, (0). 4) . ( and 5 (A)of the multi-stage column system, perfused with 8 mmol phenol dm-3. either
Fig. 4. Changes in leachate residual phenol concentrations (mmol dm -’) at sampling ports I
3
unsupplemented (a) or supplemented with N + P (b) or mineral salts (c).
379). This progressive loss of phenol attenuation may be attributed to toxic/ inhibitory effects of the elevated influent phenol concentration or nutrient/element limitat ion. Although kinetic studies with short-term batch cultures have shown that biofilm development conferred partial protection from substrate inhibition,26 data are lacking on continuously perfused models. It is possible, however, that microbial attachment to the heterogeneous refuse surfaces may have afforded initial protection. Reduced phenol catabolism was not reflected by low leachate pH values or high VFA concentrations. This suggested that either the phenol concentration used inhibited the acidogenic species also, or the acidogenic substrate pool was depleted. The rates of methane release, however, were not unduly affected. To establish if nutrient/element limitation was operative, the influent was supplemented, on day 383, with nitrogen (NH,CI, 0 9 g dm-3) and phosphorus (K2HP04,1.5 g; NaH2P0,, 0.85 g dm-3). These elements were chosen since they are often considered limiting in refuse.I6 Prior to supplementation, however, acidhydrolysable phosphate concentrations of 0.63, 0 8 0 and 0-77 mg dm-3 were recorded in leachate samples at ports 1,3 and 5 , respectively. Supplementation with N + P or, subsequently, with complete mineral salts (day 421) did not promote phenol catabolism (Fig. 4). This progressive loss of phenol attenuation was also reflected by reduced methane release rates (Fig. 5a, b and c).
229
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Fig. 5. Changes in recorded (0) and theoretical ( 0 )phenol-derived rates of methane release (cm3 g refuse-' d - ' ) at sampling ports 1-5 of the multi-stage column system, perfused with 8mmol phenol dm-3 in the absence (a) or presence (bk(d) of nutrient supplementation: (b) N and P; (c) mineral salts; and (d) hexanoate (2 mmol dm-'). In (d) the theoretical rates of release have been corrected, to include theoretical hexanoatederived release rates.
230
I . A . Watson-Craik, E . Senior
Fig.6. Changes in leachate residual phenol concentrations (mmol dm-3)at sampling ports 1 (0). 2 (a), 3 (O), 4 ( W ) and 5 (A) of the multi-stage system, sequentially perfused with 8 mmol phenol dm either unsupplemented (a) or supplemented with 2 mmol hexanoate dm-’ (b), and 2 mmol phenol dm-3 (c).
’,
Since hexanoic acid has been suggested as an intermediate in the anaerobic degradation of phenol in the presence of nitrate reduction27 and methanogenesis28 it was questioned whether 8 mmol phenol dm-3 would inhibit its dissimilation. The influent was, therefore, supplemented, on day 453,with 2 mmol hexanoate dm-3 in PO4 buffer (4 mmol dm-3). Within 8 days, the hexanoate was completely dissimilated ( D =0.007 h-’), with concomitantly increased CH, release (Fig. 5d), although residual phenol concentrations simultaneously increased (Fig. 6). Since anaerobic hexanoate degradation was not constrained by nutrient/element limitation or potentially toxic/inhibitory effects of phenol (8 mmol dm- 3), it appeared that the limited attenuation of phenol at dilution rates 3 0.007 h - was due to inhibition of phenol transformation to hexanoate. A similar differential inhibitory effect was reported by Fedorak and H r ~ d e ywho , ~ ~showed that in batch culture acetogenesis and methanogenesis were inhibited by phenol concentrations > 21.5 mmol dm-3 in contrast with the phenoldegrading acid-formers who were inhibited by concentrations 2 8.5 mmol dm-3.
3.6 Influent phenol: 2 mmol dm-3 To investigate if the observed inhibitory effects were bactericidal or bacteriostatic, the influent phenol concentration was reduced (day 509) to 2 mmol dm-3. After a
Orgunic load
Landfill Co-disposal of Phenol-bearing Wastewaters: Organic Load Considerations Irene A. Watson-Craik & Eric Senior Department of Bioscience and Biotechnology, Applied Microbiology Division, University of Strathclyde, 204 George Street, Glasgow G1 lXW, UK (Received 21 November 1988; revised version received 13 March 1989; accepted 10 April 1989)
ABSTRACT A multi-stage model, operated with single elution, was used to investigate the effrcts of organic loadings on the attenuation of a model phenolic wastewater in domestic refuse. Although 100 % dissimilation of influent phenol (25 mmol dm - 3, was recorded at a dilution rate of 0.007 h - partial inhibition of both phenol degradation and species competing with methanogens for a common electron donor(s) was apparent at concentrations 3 4 mmol dm-3. On extended perfusion with 8 mmol phenol dm- 3, the progressive inhibition of phenol dissimilation was not obviated by nutrient supplementation. Simultaneous degradation of the catabolic intermediate, hexanoic acid, and elevated methane release rates suggested that the transformation of phenol to hexanoate was rate limiting.
',
Key words: codisposal, phenol, refuse, organic load, landfill, met hanogenesis.
1 INTRODUCTION
In the UK, the official endorsement of hazardous waste codisposal with refuse in landfill sites' and the relative cheapness of the method' ensure the continuation of the practice, which accounts for the disposal of more than 78 % of all hazardous It is unfortunate, however, that guidelines on permissible loading rates have been issued in the absence of definitive studies on the performance of landfill sites as bioreactors. The landfill bioreactor contains a wide range of natural and xenobiotic molecule^,^ some or all of which are susceptible to microbial challenge. The 219 J . Chrm. Tech. Bioteclrriol.0268-2575/90/%03.50 0 1990 Society of Chemical Industry. Printed in Great
Britain
220
1. A . Warsou-Craik.E . Scwior
ecosystem is dominated by irregularly distributed solid surface components and liquid (leachate) of changing composition. Consequently, physicochemical parameters such as pH, E,,, a, and temperature vary both temporally and spatially. The complexity and dynamics of the environment thus militate against in-situ research, and necessitate the development of heterogeneous laboratory models.’ Studies have been made with conventional small-scale (0.11 m3)6and pilot-scale (8 m3)7systems but, although the latter have been employed to examine leachate reactions, their high capital cost, considerable operator input, protracted operational period’ and difficulties in maintaining anaerobiosis7 have limited their value. Previous co-disposal studies have focused on the effects of increased loadings of heavy metals on refuse metabolism.’.’ Work on phenol codisposal, with specific reference to hydraulic loading rates, has also been reported’ although organic loads have not been considered. The aim of the present study, using a small-scale multi-stage system, was to examine the impacts of phenol organic loads on refuse metabolism. Phenol was selected as being the simplest aromatic core molecule characteristic of the high-volume wastewaters produced in coke, fibreboard, phenolic resin and nylon manufacture.
’
2 EXPERIMENTAL MATERIALS AND METHODS 2.1 Refuse
Refuse, from a depth of 2 2 m and emplaced 2 1 month, was sampled from Kilgarth Landfill (Glasgow). Prior to use, it was handsorted to remove visible stones, glass, metals and wood, then homogenised for 30 s in a domestic blender (Krups). 2.2 Molecule
Stock solutions of phenol (100 mmol dm-3 in glass-distilled water) were used to modify influent phenol concentrations.
2.3 Mineral salts solution The mineral salts solution used was the same as that described by Coutts et al.” 2.4 Multi-stage refuse column
Eight glass columns (length 90 cm, internal diameter 5 5 cm), each filled with 1 kg uniformly compacted refuse and insulated with domestic quality (1 5 mm) pipelagging, were serially linked with oxygen-impermeable butyl rubber tubing (Esco Rubber, Feltham) (Fig. 1). Sampling ports ( l d ) , fitted with three-way sampling valves, were sited at discrete intervals down the array. The model was maintained at ambient temperature for 67 days before influent substrate was introduced to the top of column (a) by use of a Watson-Marlow 202U Flow Inducer. After percolating under gravity through columns (a) and (b) the resulting leachate was pumped to the top of the next column. This serial perfusion was continued through the array. A mean empty-bed dilution rate ( D ) of 0026 h-’ was imposed on column (a)
Oryuiiic loud qf1i.m 011 co-disposal
22 I
7 a.
Y
L Fig. 1. Multi-stage refuse column array, with sampling ports 1-6 and temperature probes i and ii.
(sampling port 1 ) and the resultant dilution rates at ports 2-5 were 0.013, 0.007, 0.004 and 0003 h-', respectively. Any transient over- or under-pressuresin the column headspaces were modulated by the insertion into columns (a), (c), (e) and (g) of syringes with sliding barrels.
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I. A . Watson-Craik, E . Settior
2.5 Sample preservation and storage Leachate samples for sulphate, phosphate or residual phenol analysis were stored at - 10°C in polypropylene bottles for < 7 days. Samples (0.9 cm3)for volatile fatty acid (C, to C,) analysis were acidified with formic acid (0.1 cm3) (BDH, Aristar) before storage in glass vials at - 10°C for 6 30 days.
2.6 Analytical methods 2.6.1 Phenols Liquid samples (0.001 pm3) were injected into a Perkin Elmer Sigma 115 gas chromatograph (GC), equipped with a flame ionisation detector, in which the temperatures of the injector, oven and detector were maintained at 200"C, 150°C and 210"C, respectively. The flow rate of the oxygen-free nitrogen (OFN) (British Oxygen) carrier gas was set at 40cm3 min-'. The glass column (length 2 m, i.d. 2 mm) was packed with 5 % polymeta phenyl ether with 6 rings on Chromosorb WAW (80-100 mesh) (Phase Separations). Leachate phenol and m-lpcresol concentrations were determined, after standard curve construction, by mean peak area comparison with phenol and pcresol standards (2 mmol dm-3). 2.6.2 Volatile fatty acids Leachate volatile fatty acid (VFA) concentrations were determined by use of the same GC.The flow rate of the carrier gas (OFN) was set at 30 cm3 min-' and the oven temperature, initially held at 100°C for 2 min, was programmed to increase to 150°C at a ramp rate of 30°C min-I. Acidified samples (0.001 pm3) were injected on to the glass column (length 2 m, i.d. 2 mm) which was packed with 5 % neopentyl glycol sebacate+ 1 % H3P0, on Anakrom polyester (80-100 mesh) (Chromatographic Services).Acidified solutions (10mmol dm-3) of fatty acid standards were used, and leachate VFA concentrations calculated by peak area comparison, after computer-generated baseline correction. 2.6.3 Sulphate After sparging leachate samples (10 cm3)for 10 min with OFN to remove sulphides, total sulphate concentrations were estimated by use of the standard BaCI, turbidometric method." 2.6.4 Phosphate Total acid-hydrolysable phosphate was determined by the standard ascorbic acid method.'
2.6.5 pH pH values of the liquid samples were recorded by use of a Pye Unicam PU9148 pH meter. 2.6.6 Refuse dry weight Triplicate fresh refuse samples (lOOg),dried to constant weight at 60°C were used to estimate refuse moisture content.
Orgunic loud ejjicts on co-disposul
223
2.6.7 Solubilised methane Triplicate universal bottles ($ ounce) were closed with suba-seals (Gallenkamp)and evacuated, for 30s, with a vacuum-pressure pump (Millipore XX60 22050). A saturated methane solution was prepared by sparging cold glassdistilled water (10cm3)with methane (CP grade, British Oxygen)for 30 min,I4 then tightly sealing the vessel. The solubilisation temperature was noted and the soluble methane concentration then determined from a standard curve." Samples ( 1 cm3) of the standard solution were injected into pre-evacuated universal bottles which were then equilibrated to atmospheric pressure with hypodermic needles. After sealing the bottles by needle withdrawal, triplicate gas phase samples (0.05pm3) were taken and injected into the GC under the operating conditions described for VFA analysis although, in this case, the oven temperature was held at 80°C. Leachate samples ( 1 cm3) were similarly processed and leachate methane concentrations calculated by comparison with the methane standard. The rates of methane release at each discrete dilution rate were calculated from the equation: r=
22.4jc(1 + l/a,) W
where r is the rate (cm3g refuse-' d - ') of methane release, f is the liquid flow rate (dm3 d - ' ) through the refuse, c is the leachatedissolved methane concentration (pmol ~ m - ~a,) is, the solubility coeflicient of methane" at ambient temperature t"C, and w is the cumulative refuse weight (fresh) (g).
3 RESULTS AND DISCUSSION 3.1 Introduction To facilitate realistic enrichments of microorganisms the columns were allowed to self-generate redox and pH gradients at temperatures which, as a result of column insulation, were 0.5-14T above ambient (Probes i and ii (Fig. 1)). Since previous studies16had shown that a pre-perfusion incubation period effected a 50% increase in phenol catabolism compared with controls, this strategy was again adopted. The parameters selected for monitoring included those known to act as indicators of refuse age and degree of stabilisation.' These comprised: refuse compaction/ density; leachate pH, sulphate and VFA concentrations; and methane release. Despite initial processing the refuse moisture capacity (64%)approximated to insitu values. Similarly, the applied refuse density (513 kg m-3) was comparable to site conditions." After 548 days perfusion the final density was 696 kg m - 3 due to 26 % (v/v) settlement.
3.2 Influent phenol: 2-6 mmol dm-3 The discrete phenol concentrations recorded in the leachate samples may have originated from three possible sources: the influent substrate; as a component of the refuse used; or as a product of refuse degradation. Phenol concentrations of 0.33 and 016 mmol dm-3 of leachate were recorded at sampling ports 1 and 5 after 4 and
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1 . A . Wurson-Craik, E . Senior
96 h perfusion, respectively. These concentrations were greater than might have
been expected if the model operated with plug-flow. It was possible that either the column array was operating on a partial plug-flow basis, due possibly to a limited degree of channelling, or phenol was displaced from the refuse. The equivalent pcresol concentrations of 0.38 and 0.10 mmol dm-3 were attributed to a comparable displacement of phenolics from the refuse. After 24 days perfusion the residual phenol concentration (port 5 ) progressively decreased and was no longer detected after 45 days (Fig. 2). A similar trend was observed with leachate sampled at ports 4 and 3, whilst over the subsequent 34 days the average residual phenol concentrations of leachate samples 2 and I approximated to 14% and 32% of the influent. On day 70 the influent phenol concentration was increased to 4 mmol dm-3. Within three days, elevated phenol concentrations were recorded at sampling ports 1 and 2, and within 10 days at ports 1 to 5 (Fig. 2). After day 88, however, residual concentrations progressively declined until by day 101 no phenol was detected in the final efiluent. On day 119 residual phenol (0-94mmol dm-3) was detected; traces remained detectable until day 136, and were coincident with low laboratory temperatures ( < 6°C). Subsequently, the temperature was increased (17-19°C) and phenol concentrations fell until by day 143 no residual phenol was detected in leachate sampled at ports 4 and 5.
Fig. 2. Changes in leachate residual phenol concentrations (mmol dm-’) at sampling ports I (0 ), 2 (a), 3 4 (m)and 5 (A) of the multi-stage refusecolumn array, perfused by 2 (a),4 (b),5 (c)and 6 (d)mmol phenol dm-3.
(a),
Biodegradation was adversely affected by reducing the temperature to < IO'C whilst a previous study had shown that a temperature increase from 21 1.5'C to 30' C had no significant effect." The bactericidal activity of phenol is promoted by temperature increase." Thus it is possible that at temperatures > 20°Cincreased metabolism is negated by the increased bactericidal activity. Clearly, landfill temperature is an operational parameter that may require optimisation in relation to phenol codisposal. When the influent concentration was increased to 5 mmol dm-3 (day 161) and 6 mmol dm-3 (day 244) no residual phenol was detected in leachates sampled at ports 4 and 5. Thus at an empty-bed dilution rate 60404 h - ' attenuation of 6 mmol phenol dmP3is possible. There are few data on phenol removal rates in other continuous anaerobic reactors. Removal (1OOXJ of influent 2.1 and 4.3 mmol phenol d m - 3 was recorded2' in an activated carbon-packed upflow system with recirculation, at a dilution rate of 0.01 1 h-I; even at an influent concentration of 10-6mmol dm-3, removal was 98.1 "/,. Increased inhibitory effects were recorded, however, in reactors supplemented with 1.6 and 3.2 mmol phenol dm-3 in which there was no provision of biomass supports.2' Although 100% removal was noted at dilution rates from 0.002 to 0403 h increasing emuent concentrations were recorded after periods from 17 (3.2 mmol dm-j) to 53 (1.6 mmol dm-3) days. The provision of refuse surfaces for biofilm attachment may thus be an integral factor in the efficacy of phenolic wastewater treatment by co-disposal. 3.3 p H and volatile fatty acids Leachate pH and VFA analyses showed that the increase in influent phenol from 2 to 4 mmol dm - 3 resulted in a pH decrease, within 10 days, from 6.55 to 5.70 at sampling port 1, although this was not coincident with elevated VFA concentrations. Only acetate, on days 80 and 89, at concentrations of 0.81 and 0.03 mmol dm-3, respectively, was recorded. By day 101, however, leachate pH (port 1) had increased to 6.4. However, when the phenol concentration was increased to 5, and, later, to 6 mmol dm ', the pH profiles did not show a similar dip and recovery but remained relatively constant; phenol degradation also continued smoothly. This indicated that no shock-loading had occurred under these conditions. 3.4 Methane From the concentrations of phenol removed from the influent, theoretical rates of methane release were calculated and compared with the actual rates recorded (Fig. 3). Phenol was assumed to be mineralised to CO, and CH, according to the equation:22 CnH,O,+(n-~/4-b/2)HZO
+ (n/2-~/8+b/4)CO,
+(n/2+~/8-h/4)CH,
(2) Such stoichiometric metabolism under defined non-SO,-supplemented conditions has been r e p ~ r t e d . ~In ~ ,the ' ~ present study, however, where leachate SO, concentrations between days 1, and 70 ranged from 0.02 to 056 mmol dm-3,
226
I . A . Watson-Craik, E . Senior
"4
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(0) and theoretical ( 0 )phenolderived rates of methane release (cm3 g refuse-'d-')atsamplingports 1-5ofthemulti-stagemodel.perfusedwith2(a).4(b).5(c)or 6(d)mmol phenol dm-3. Sampling ports 1-5 correspond to dilution rates 0.026,0~013.0~007.0004 and 0403 h-', respectively. Fig. 3. Changes in recorded
electron flow towards sulphate reduction with a concomitant decrease in methanogenesis must also be considered. From Fig. 3a it can be seen that, in the presence of an influent phenol concentration of 2 mmol dm-3, the actual rates of methane release were higher than predicted. The discrepancy was assumed to reflect methanogenesis from refuse components. With time the discrepency decreased, possibly due to metabolism of
Organic loud yff~crs oti co-disposul
227
the pool of methane precursors. At the higher dilution rates, however, the observed rates were lower than were theoretically expected. This was possibly due to spatial separation of phenol degradation from methanogenesis, although previous studies with single-stage columns16 maintained at 30°C had shown that under these conditions the two metabolic processes were coincident. When the influent phenol concentration was increased to 4 mmol dmimmediate increases in the rates of methane release were apparent at all dilution rates, particularly at 0.026 h - (Fig. 3b). These increases were significantly greater than could be accounted for by phenol mineralisation and possibly resulted from a diversion of electron flow from sulphate-reduction to methanogenesis. A comparable inhibition of group(s) competing with methanogens for common electron donor@)and/or acceptor(s) was also observed in a phenol-supplemented (2 mmol dm- ’) veratric acid degrading association (K. K. Abdul-Halim, personal communication). At dilution rates > 0.013 h-’ reduced theoretical rates of phenol-derived methane release were apparent and were possibly indicative of partial inhibition of a metabolic group. At this stage it can only be speculated as to the identity of the group. Work with batch methanogenic fermentations by Fedorak & Hrudey (1984)*’ also demonstrated this inhibitory effect. They showed that methanogens were inhibited only at phenol concentrations 2 12.75mmol dm-3, whilst phenolcatabolising species were increasingly inhibited by concentrations > 5.3 mmol dm-’. Unfortunately, they were unable to determine if non-phenoldegrading species were adversely affected by the concentrations used. The effects of phenol concentration on rates of methane release were also examined with influent concentrations of 5 and 6 mmol dm-3 (Fig. 3c and 3d). With 5 mmol dm-’ phenol and dilution rates >0*007 h-’, methane release rates were significantly reduced (Fig. 3c). Although possibly due to partial inhibition of methanogenic activity, this was not confirmed by the presence of elevated concentrations of methanogenic precursors. However, at dilution rates < 0404 h - and in the presence of phenol concentrations < 0.71 mmol dm methanogenic activity was sustained. With an influent concentration of 6 mmol phenol dm-3, there was an initial dramatic increase in the methane release rate at the highest dilution (Fig. 3d), although by day 285 partial methanogenic inhibition at dilution rates 20.013h-’ was again apparent. At the lowest dilution rate, however, the observed methane release rate increased relative to the theoretical rate and reflected increased total phenol removal.
’,
’,
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3.5 Influent phenol: 8 mmol dm-3 On day 285 the influent phenol concentration was increased to 8 mmol dm-’. For the next 38 days no residual phenol was detected at sampling port 5 (Fig. 4) and it thus appeared that, with appropriate dilution rates, influent phenol concentrations < 8 mmol dm- could be effectively attenuated by single-elution through refuse. By day 330, however, residual phenol was recorded at ports 4 (0.56 mmol dm-’) and 5 (0.06mmol dm-7. These residual concentrations then progressively increased to 6.52, 5.01 and 3.25 mmol dm-3 at ports 3, 4 and 5, respectively (day
I . A . WofsowCrctik. E . Senior
280
(0 ] , 2 (@ 1, (0). 4) . ( and 5 (A)of the multi-stage column system, perfused with 8 mmol phenol dm-3. either
Fig. 4. Changes in leachate residual phenol concentrations (mmol dm -’) at sampling ports I
3
unsupplemented (a) or supplemented with N + P (b) or mineral salts (c).
379). This progressive loss of phenol attenuation may be attributed to toxic/ inhibitory effects of the elevated influent phenol concentration or nutrient/element limitat ion. Although kinetic studies with short-term batch cultures have shown that biofilm development conferred partial protection from substrate inhibition,26 data are lacking on continuously perfused models. It is possible, however, that microbial attachment to the heterogeneous refuse surfaces may have afforded initial protection. Reduced phenol catabolism was not reflected by low leachate pH values or high VFA concentrations. This suggested that either the phenol concentration used inhibited the acidogenic species also, or the acidogenic substrate pool was depleted. The rates of methane release, however, were not unduly affected. To establish if nutrient/element limitation was operative, the influent was supplemented, on day 383, with nitrogen (NH,CI, 0 9 g dm-3) and phosphorus (K2HP04,1.5 g; NaH2P0,, 0.85 g dm-3). These elements were chosen since they are often considered limiting in refuse.I6 Prior to supplementation, however, acidhydrolysable phosphate concentrations of 0.63, 0 8 0 and 0-77 mg dm-3 were recorded in leachate samples at ports 1,3 and 5 , respectively. Supplementation with N + P or, subsequently, with complete mineral salts (day 421) did not promote phenol catabolism (Fig. 4). This progressive loss of phenol attenuation was also reflected by reduced methane release rates (Fig. 5a, b and c).
229
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Fig. 5. Changes in recorded (0) and theoretical ( 0 )phenol-derived rates of methane release (cm3 g refuse-' d - ' ) at sampling ports 1-5 of the multi-stage column system, perfused with 8mmol phenol dm-3 in the absence (a) or presence (bk(d) of nutrient supplementation: (b) N and P; (c) mineral salts; and (d) hexanoate (2 mmol dm-'). In (d) the theoretical rates of release have been corrected, to include theoretical hexanoatederived release rates.
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I . A . Watson-Craik, E . Senior
Fig.6. Changes in leachate residual phenol concentrations (mmol dm-3)at sampling ports 1 (0). 2 (a), 3 (O), 4 ( W ) and 5 (A) of the multi-stage system, sequentially perfused with 8 mmol phenol dm either unsupplemented (a) or supplemented with 2 mmol hexanoate dm-’ (b), and 2 mmol phenol dm-3 (c).
’,
Since hexanoic acid has been suggested as an intermediate in the anaerobic degradation of phenol in the presence of nitrate reduction27 and methanogenesis28 it was questioned whether 8 mmol phenol dm-3 would inhibit its dissimilation. The influent was, therefore, supplemented, on day 453,with 2 mmol hexanoate dm-3 in PO4 buffer (4 mmol dm-3). Within 8 days, the hexanoate was completely dissimilated ( D =0.007 h-’), with concomitantly increased CH, release (Fig. 5d), although residual phenol concentrations simultaneously increased (Fig. 6). Since anaerobic hexanoate degradation was not constrained by nutrient/element limitation or potentially toxic/inhibitory effects of phenol (8 mmol dm- 3), it appeared that the limited attenuation of phenol at dilution rates 3 0.007 h - was due to inhibition of phenol transformation to hexanoate. A similar differential inhibitory effect was reported by Fedorak and H r ~ d e ywho , ~ ~showed that in batch culture acetogenesis and methanogenesis were inhibited by phenol concentrations > 21.5 mmol dm-3 in contrast with the phenoldegrading acid-formers who were inhibited by concentrations 2 8.5 mmol dm-3.
3.6 Influent phenol: 2 mmol dm-3 To investigate if the observed inhibitory effects were bactericidal or bacteriostatic, the influent phenol concentration was reduced (day 509) to 2 mmol dm-3. After a
Orgunic load
