Reductive biodegradation of 1,2-dichloroethane by methanogenic granular sludge in lab-scale UASB reactors

Advances in Environmental Research 6 Ž2001. 17᎐27

Reductive biodegradation of 1,2-dichloroethane by methanogenic granular sludge in lab-scale UASB reactors Stefaan De Wildemana , Hendrik Nollet a , Herman Van Langenhove b, Willy Verstraete a,U a

Laboratory for Microbial Ecology and Technology (LabMET 1), Faculty of Agricultural and Applied Biological Sciences, Ghent Uni¨ ersity, Coupure Links 653, B-9000 Ghent, Belgium b Laboratory of Organic Chemistry, Faculty of Agricultural and Applied Biological Sciences, Ghent Uni¨ ersity, Coupure Links 653, B-9000 Ghent, Belgium

Abstract Dechlorination of 1,2-dichloroethane Ž1,2-DCA. dosed to a model wastewater in lab-scale upflow anaerobic sludge blanket ŽUASB. reactors was examined. Anaerobic granular sludge was used as a biocatalyst. Ethanol served as the main methanogenic substrate. For 3 months, two types of UASB reactors were studied, the first type consisting of a sludge blanket and the second type containing an additional layer of activated carbon. When subjected to 1,2-DCA at an average volumetric loading rate of 87.6 mg ly1 dayy1, the latter type obtained an average removal efficiency of 82%. Increasing the volumetric loading rate of ethanol from 5 to 15 g COD ly1 dayy1 resulted in higher 1,2-DCA conversion rates. No chlorinated intermediates or residues were found. 1,2-DCA was converted mainly to ethene Ž65᎐80%. and ethane Ž- 1%.. Both autoclaved sludge and cell extracts were not able to degrade 1,2-DCA, which indicates the need for metabolic activity. The reactor effluents were less toxic relative to the influent when analyzed by Nitrox tests, indicating that such UASB treatments can protect a subsequent aerobic nitrifying system. The 1,2-DCA removal rates achieved, and the safe nature of the endproducts, warrant the combination of granular sludge and UASB technology for practical decontamination of waters containing such types of organochlorines. 䊚 2001 Elsevier Science Ltd. All rights reserved. Keywords: Methanogenic activity; Reductive dechlorination; Upflow anaerobic sludge blanket ŽUASB. reactor

1. Introduction The anthropogenic chemical 1,2-DCA is reported to be one of the predominant organohalogen pollutants in

U

Corresponding author. Tel.: q32-9-264-5976; fax: q32-9264-6248. E-mail address: [email protected] ŽW. Verstraete .. 1 http:rrwww.welcome.torlabMET.

groundwater and industrial effluents. Concentrations range from ␮g to g ly1 , caused by the use of chlorinated ethane in the production of chlorinated solvents Že.g. trichloroethene and tetrachloroethene . and as an intermediate in the synthesis of vinylchloride and fine chemicals. Since 1,2-DCA has good solubility in water Ž8.7 g ly1 ., a low sorption coefficient Žlog K oc s 1.28. and a low Henry coefficient Ž1.1= 10y3 atm m3 moly1 . ŽDewulf et al., 1995., it remains in the water phase under average environmental conditions Žgroundwater:

1093-0191r01r$ - see front matter 䊚 2001 Elsevier Science Ltd. All rights reserved. PII: S 1 0 9 3 - 0 1 9 1 Ž 0 1 . 0 0 0 6 7 - 8

18

S. De Wildeman et al.D,(,R,216,456,57, r Biophysical Chemistry 6 (2001) 17᎐27

15⬚C, pH 7.. It is a known carcinogen ŽVogel et al., 1987. because of the conversion into chloroacetaldehyde, which is considered to have mutagenic properties ŽMcCann et al., 1975.. The combination of its persistence and toxicity makes 1,2-DCA a target molecule to degrade chemically or biologically, rather than to replace it from water into another phase Že.g. sorption on activated carbon, or gas stripping.. Aerobic biodegradation of 1,2-DCA has been carried out successfully, involving specific strains, such as Pseudomonas sp. strain DE2 ŽStucki et al., 1983., Pseudomonas sp. strain DCA1 ŽHage Jacobus and Hartmans, 1999. and Xanthobacter autotrophicus GJ10 ŽDossantos and Livingston, 1995; Stucki and Thuer, 1995.. Several anaerobic 1,2-DCA-degrading microorganisms, dechlorinating via metabolic or cometabolic reactions, have been isolated in order to study their degradation characteristics ŽFathepure and Boyd, 1988; Holliger et al., 1990.. Dealing with wastewaters contaminated with organohalogens, anaerobic degradation has several interesting advantages compared with aerobic treatments. Firstly, the stripping effect of C 1 and C 2 organohalogens is minimal under anaerobic conditions. Secondly, as multiple-halogenated organic compounds are persistent in aerobic conditions, anaerobic organisms are able to transform a broader spectrum of organohalogen substrates ŽVogel et al., 1987.. In this work, we examined the ability of UASB technology to dehalogenate strongly reduced compounds, such as 1,2-DCA, at high volumetric loading rates. Therefore, the dechlorination capacity of granular methanogenic biomass in the presence of high amounts of 1,2-DCA Ž20᎐100 mg ly1 . and EtOH as the main carbon source was studied. This report demonstrates that the concept of UASB technology under optimal dechlorination conditions can be applied to treat 1,2-DCA-contaminated waters.

2.2. Analytical methods Concentrations of 1,2-DCA were determined by analyzing 400-␮l headspace samples on a Chrompack CP9001 gas chromatograph equipped with a flame ionization detector ŽFID., set at 250⬚C. The GC contained a 50-m long, 0.53-mm I.D. capillary column coated with CP-Sil 5CB Ž95% dimethyl 5% phenylpolysiloxane, film thickness 1.25 ␮m.. The GC was operated in the split injection mode. The carrier gas utilized was helium at a flow rate of 40 ml miny1. Peak areas were calculated using the Cecil software tool and by comparing them to standard curves and a standard mixture of 29 organohalogens ŽSigma, EPA 601r602 Purgeables Kit, Cat. No. 5-06915.. The oven temperature was kept constant at 100⬚C. The samples were prepared in 10-ml crimp neck vials ŽChrompack, Cat. No. 738200. by adding 2 ml of water sample. The vials were closed with gas-tight silicone liners and magnetic caps. They were kept at 50⬚C for 2 min to equilibrate with stirring at 600 rev. miny1 before the autosampler injection Žinjector temperature at 220⬚C.. Volatile fatty acids ŽVFA. were determined with a capillary FID gas chromatograph, Carlo Erba Fractovap 4160. A separation of the VFA was obtained, going from acetic acid to capronic acid. The column used was a FFAP Ž25 m.. The operational temperature was controlled at 135⬚C for the isothermal oven, and 175⬚C for the detector and injector. Nitrogen was used as the carrier gas at a flow rate of 10 ml miny1. Detection of ethene and ethane was performed on a Varian 3700 gas chromatograph with a packed column Ž10 m. under isothermal conditions Ž50⬚C., connected to a flame ionization detector ŽFID.. All compounds had detection limits below 0.1 ␮mol ly1. Gas volumes were measured and accumulated in separate tedlar sampling bags Ž1 l, Alltech, Laarne, Belgium..

2.3. Granular sludge and lab-scale UASB reactors 2. Materials and methods 2.1. Chemicals The compounds used in this study were 1,2-DCA Ž99%, Fluka, Switzerland. and EtOH Ž99%, Fluka.. Standard gases included a pure mixture of CO 2 and CH 4 Ž40:60 ratio, Messer, Germany. and a mixture of acetylene, carbon dioxide, carbon monoxide, ethane, ethene and methane, all at 1 vol.%, in nitrogen ŽScott Specialty Gases, Supelco Park, Bellefonte, PA.. Chloroethane was obtained as a standard solution Ž2000 ␮g mly1 . in methanol from Supelco Inc. ŽMisa group 16, Mix 6, 48799-U, Supelco Park, Bellefonte, PA.. Cyanocobalamin was purchased from Fluka.

Dechlorinating granular sludge was grown in two UASB reactors, which had originally been inoculated with granular methanogenic sludge from a full-scale UASB reactor treating potato processing water ŽPrimeur, Waregem, Belgium.. This methanogenic sludge had a dry matter content of 142 g ly1 , a volatile suspended solids ŽVSS. value of 114 g ly1 and a mineral ash residue of 28.5 g ly1. Reactor A contained a blanket of 200 g of sludge. Reactor B contained an underlying layer of 125 g of granular activated carbon ŽGAC. in addition to the blanket of 200 g of sludge. Both reactors Žvolume, 2.3 l. were fed with EtOH as the only carbon and energy source. Nitrogen and phosphorous sources were supplied by adding ureum wŽNH 2 . 2 CO, 134 mg ly1 x and potassium dihydrogen

S. De Wildeman et al.D,(,R,216,456,57, r Biophysical Chemistry 6 (2001) 17᎐27

phosphate ŽKH 2 PO4 , 55 mg ly1 . in the molar ratio of CODrNrPs 100:1.25:0.25. Trace elements were added by using a stock solution containing Žly1 of distilled water.: 112 mg Ni 2q Ž500 mg NiSO4 ⭈ 6H 2 O.; 140 mg Mn 2q Ž500 mg MnCl 2 ⭈ 4H 2 O.; 101 mg Fe 2q Ž500 mg FeSO4 ⭈ 7H 2 O.; 23 mg Zn2q Ž100 mg ZnSO4 ⭈ 7H 2 O.; 18 mg B 3q Ž100 mg H 3 BO 3 .; 20 mg Mo 6q Ž50 mg NaMoO4 ⭈ 2H 2 O.; 12 mg Co 2q Ž50 mg CoCl 2 ⭈ 6H 2 O.; and 1.3 mg Cu2q Ž5 mg CuSO4 ⭈ 5H 2 O.. This stock solution was dosed to the synthetic influent at 1 ml ly1. Sodium bicarbonate ŽNaHCO3 , 2.05 g ly1 . was used as buffer. The operating temperature of both reactors was 32⬚C, and the hydraulic retention time ŽHRT. was 20 h. The loading rate of COD before adding 1,2-DCA was increased over 24 days to 4.0 g COD ly1 dayy1. From that moment on, 1,2-DCA was added in increasing amounts, the COD loading rate was varied and the HRT was changed. Reactors A and B were kept within the pH ranges of 6.9᎐7.1 and 7.2᎐7.3, respectively. Gas production was measured by connecting the headspace of the reactors with gas columns, in which the acidic solution ŽpH 2. was replaced by the gas produced. Details of the reactor set-ups are outlined in previous work ŽThaveesri et al., 1995..

2.4. Transformations of 1,2-DCA in batch experiments Tests were carried out in penicillin bottles Ž120 ml. at 32⬚C to investigate the conversion products of 1,2DCA. 1,2-DCA was fed to 40 ml of EtOH-adapted methanogenic granular sludge at concentrations of 20 mg ly1. In tests where cyanocobalamin was added, it was dosed to the aqueous supernatant. The experiments consisted of: a control Žwithout granular sludge.; E5rEc5 Žgranular sludge fed with 5 g EtOH᎐COD ly1 withoutrwith addition of cyanocobalamin at a molar ratio of 1,2-DCArcyanocobalamin s 2.; AE Žautoclaved granular sludge.; and CErCcE Žgranular sludge without COD addition, withoutrwith addition of cyanocobalamin at a molar ratio of 1,2-DCAr cyanocobalamin s 2.. The bottles were purged with oxygen-free nitrogen gas. 1,2-DCA was prepared in a saturated stock solution of 1600 mg ly1 , which was fed Ž1 ml. to the bottles. Supernatant Ž39 ml, withrwithout CO D source . with trace elem ents and a nitrogenrphosphor source was prepared anaerobically and added together with the stock solution. The pH was stabilized at 7.5 by adding 2.05 g ly1 NaHCO3 to the mineral medium. The headspace of 40 ml was connected to tedlar gas-sampling bags. Concentrations of 1,2-DCA and chlorinated intermediates, gas volumes and compositions were measured on days 1, 2, 3, 7 and 8.

19

2.5. Nitrox bio-assay To evaluate the residual toxicity of the UASB effluent towards activated sludge, the oxygen uptake rate ŽOUR. of Nitrosomonas sp. was taken as variable ŽSurmacz Gorska et al., 1995.. For this measurement, 180 ml of cultured nitrifying sludge was mixed with 20 ml of UASB effluent. This ratio was found to be optimal in terms of nitrification activity of the sludge and dilution effect of the water sample. Before this, within 4 min, the nitrifying sludge was aerated and brought to 20⬚C. The NHq 4 N concentration was set to 50 mg ly1 wby dosing with FeŽNH 4 . 2 ŽSO4 . 2 x and the pH was adjusted to 7.5. After this pretreatment, the aeration was stopped and the sludge was mixed with the water sample. Oxygen measurements were taken with an interval of 5 s by means of an oxygen electrode. Total inhibition of nitrification and OUR was achieved by adding 1 mg of allylthioureum ŽATU., a selective inhibitor of the Nitrosomonas group at a concentration of 5 mg ly1. Before and after the addition of ATU, the respective slopes ŽOUR, mg O 2 ly1 miny1 . of the respirogram were calculated with linear regression. The difference between both OUR values was defined as the ⌬ value ŽOUR, mg O 2 ly1 miny1 ., giving a measure for the activity of the Nitrosomonas group. The use of this ⌬ value in Eq. Ž1. thus expresses the procentual inhibition of the UASB effluent related to the control Žtap-water and influent without 1,2-DCA.. % inhibition s

⌬ control y ⌬ UASByeffluent )100 ⌬ control

Ž1.

2.6. Adsorption isotherms The adsorption of 1,2-DCA on granular activated carbon ŽGAC. and on granular methanogenic sludge ŽGMS. was calculated. The GAC was powdered and dried at 105⬚C; the GMS was decanted Ž12 wt.% dry matter.. Aqueous solutions of 1,2-DCA Ž50, 100, 150, 200 and 250 mg ly1 ; 200 ml each. were brought into contact with the GAC Ž0.25 g. and the GMS Ž4 g. in 250-ml flasks. For each of the five concentrations of 1,2-DCA, four samples Ž20 in total. were set up to collect the adsorption data. The flasks were closed Žgas-tight. and shaken for 18 h at 28⬚C. After the equilibration time of 18 h, samples were filtered through 8-␮m pore-size filters ŽWhatman, Inc.; used in each filtration step of the adsorption tests. and 1,2-DCA concentrations in each filtrate were measured with GCrFID. The adsorption of EtOH on GAC and on GMS was measured similarly to the 1,2-DCA adsorption tests, except for the amounts that were used and the EtOH

20

S. De Wildeman et al.D,(,R,216,456,57, r Biophysical Chemistry 6 (2001) 17᎐27

measurements in the filtrate. Aqueous solutions of EtOH Ž2, 4, 6, 8, 10 g ly1 ; 200 ml each. were brought into contact with the conditioned GAC Ž80 g. and the decanted GMS Ž40 g. in 250-ml flasks. For each of the five concentrations of EtOH, three samples Ž15 in total. were set up to collect the adsorption data. To avoid EtOH losses due to biological degradation, GMS samples were filtered after 1 h of shaking and EtOH concentrations were evaluated by COD measurements. GAC samples were shaken for 18 h, centrifuged Ž10 000 = g; 10 min. and the supernatants were filtered. EtOH concentrations in each filtrate were evaluated by means of COD measurements.

sources and trace elements were dosed in specific ratios, as described above. Two ranges of 1,2-DCA concentrations were set up Ž0᎐100 at 10-mg intervals and 0᎐500 at 50-mg intervals of 1,2-DCA ly1 .. The flasks were flushed with nitrogen, closed gas-tight and incubated at 32⬚C. Gas pressure on the headspace was measured using a tensiometer with a needle entering the rubber stopper at the top of the flask. This was carried out every hour until gas production decreased Žafter 10 h.. After 10 h, the composition of the gas ŽCH 4rCO 2 ratio in the headspace. was measured.

2.7. Toxicity tests

In order to investigate the dechlorination capacity of bacterial cell extracts, the following extraction procedure was followed. Sludge Ž2 g tubey1 . was brought into five polypropylene round bottom tubes Ž14 ml, Falcon.. TrisrHCl buffer Ž6 ml, 10 mM, pH 9. and

Granular methanogenic sludge Ž40 ml. was mixed with 40 ml of distilled water containing 100 ␮l of EtOH in a penicillin flask Ž120 ml.. Nitrogen and phosphorous

2.8. Cell extraction

Fig. 1. 1,2-DCA dechlorination performance and reactor behavior of UASB reactor A fed with EtOH at several loading rates of 1,2-DCA and COD. Ža. Elimination efficiency of 1,2-DCA Ž⽧. and volumetric loading rate of 1,2-DCA Žfull line without markers.. Žb. COD removal efficiency Ž⽧. and volumetric loading rate of COD Žfull line without markers.. Žc. Biogas production Ž⽧. and hydraulic retention time Žfull line without markers..

S. De Wildeman et al.D,(,R,216,456,57, r Biophysical Chemistry 6 (2001) 17᎐27

glass beads Ž3 g, 0.10᎐0.11 mm. were added. The mixture was beaten five times for 45 s with a rest time of 30 s Ž2000 rev. miny1 . using a bead beater ŽB. Braun Biotech International, Melsungen, Germany.. The content of the tubes was centrifuged at 7000 = g for 10 min. After filtration over a 0.22-␮m pore filter ŽWhatman., part of the supernatant Ž5 ml. was added to a new tube, together with 5 ml of a standard solution containing 100 mg ly1 1,2-DCA and 400 mg ly1 sodiumthioglycollate ŽC 2 H 3 NaO 2 S.. Tubes were flushed with nitrogen, closed gas-tight, and stored under anaerobic conditions in complete darkeness. This procedure was repeated each time after sampling the tubes.

21

3. Results

reactors A and B in relation to the corresponding loading rates of 1,2-DCA and COD, the COD removal and the biogas production. The hydraulic retention time varied between 10 and 20 h. During the start-up period, the granular methanogenic sludge used in both reactors was brought into an active methanogenic state by adding a mineral medium and EtOH as the easily degradable carbon source. After 4 weeks without any 1,2-DCA and with a COD load of 5 g ly1 dayy1, a COD removal of at least 92% was achieved and the residual volatile fatty acid concentration was below 320 mg ly1 Ždata not shown.. After this start-up period, from the first day of dosing on, the model compound 1,2-DCA was degraded by the methanogenic sludge, unadapted to 1,2-DCA. The degradation occurred without any lag phase.

3.1. Degradation of 1,2-DCA in UASB reactors

3.1.1. Reactor A

Figs. 1 and 2 show the dechlorination efficiency of

Within the first 30 days, the loading rate of 1,2-DCA was increased to 39 mg ly1 dayy1. The elimination of

Fig. 2. 1,2-DCA dechlorination performance and reactor behavior of UASB reactor B Žadditional activated carbon. fed with EtOH at several loading rates of 1,2-DCA and COD. Legends for Ža., Žb. and Žc. are analogous to those of Fig. 1.

22

S. De Wildeman et al.D,(,R,216,456,57, r Biophysical Chemistry 6 (2001) 17᎐27

1,2-DCA decreased to 70%. This decrease was reversed by increasing the COD loading rate to 10 g ly1 dayy1. The production of biogas was linked to the 1,2-DCA and COD removal efficiencies. The COD removal efficiency decreased between days 25 and 28 as a consequence of the sudden increase of COD and higher 1,2-DCA load. The fact that the latter influenced the COD removal efficiency was confirmed by the evolution pattern of reactor A between days 33 and 36. In that period, the 1,2-DCA loading rate was decreased to 12 mg ly1 dayy1, bringing the removal efficiency of 1,2-DCA to the level of 80᎐90% and increasing the COD removal efficiency up to 90%. This was accompanied by an increase in the CH 4rCO 2 ratio of the biogas produced from 3.5 to 4.2. Moreover, as the COD removal regained strength, the biogas production increased. The reactor pattern was stabilized at this level for 9 days. From day 48 on, the 1,2-DCA loading rate was increased again, resulting in a COD removal efficiency drop. Although the 1,2-DCA loading rate reached 58 mg ly1 dayy1 on days 61 through 68, only a small drop in removal efficiency occurred, which was followed by a strong recovery of the 1,2-DCA removal efficiency up to 85%. The CH 4rCO 2 ratio increased accordingly from 3.7 to 4.2. In these stressed conditions, the COD removal efficiency stabilized at approximately 60%. An accumulative effect was reached on days 68 through 75 when the 1,2-DCA loading rate was increased further to 85 mg ly1 dayy1 by decreasing the COD loading rate to 4 g ly1 dayy1. Removal efficiencies of 1,2-DCA and COD dropped to 60 and 50%, respectively, within almost 3 days. Biogas production dropped to only 40% of the expected value Ž0.2 instead of 0.5 l gy1 CODremoved .. Increasing the COD loading rate on day 75 and decreasing that of 1,2-DCA on day 81 resulted in the same reaction patterns as before. In the period of 90 days, reactor A performed an average elimination rate of 39.2 ␮mol of 1,2-DCA g of volatile suspended solidsy1 dayy1 Ž38.8 mg ly1 dayy1 . ŽFig. 1.. Applying an average volumetric loading rate of 49.4 mg ly1 dayy1 of 1,2-DCA, 78.6% was eliminated.

3.1.2. Reactor B Analogous tests were carried out on the reactor which contained an additional layer of activated carbon. Similar reaction patterns were observed. Due to the buffering characteristics of the carbon layer, higher 1,2-DCA loading rates of almost 200 mg ly1 dayy1 caused only small decreases in methanogenic activity. At a weight-based CODr1,2-DCA ratio of 100, the average 1,2-DCA elimination efficiency and degradation rate were 81.9% and 72.5 ␮mol of 1,2-DCA g of volatile suspended solidsy1 dayy1 Ž71.8 mg ly1 dayy1 ., respectively ŽFig. 2.. Only a small amount Ž- 5%. of granular sludge was washed out after 90 days. Moreover, the system appeared to be stable against shock

loading rates of 1,2-DCA and COD, and the granular structure of the biomass was conserved.

3.2. Adsorption tests Adsorption tests were performed to investigate the 1,2-DCA removal effect of the granular activated carbon ŽGAC.. Analogous tests were carried out with the granular anaerobic sludge. The adsorption data were fitted to the Freundlich and Langmuir isotherm models, resulting in the following model parameter values. For the Freundlich isotherms: ny1 F,GAC s 0.33, K GAC s 18.19 and R 2 s 0.83; ny1 F,sludge s 0.52, K sludge s 0.021 and R 2 s 0.79. For the Langmuir isotherms: aGAC s 0.42, X m,GAC s 109.89 and R 2 s 0.86; and asludge s 0.0032, X m,sludge s 11.83 and R 2 s 0.14. For the adsorption of 1,2-DCA on GAC, the Langmuir isotherm equation gave a slightly better fit to the data, which was reflected by the higher correlation coefficient. The adsorption of 1,2-DCA on sludge gave no significant determination coefficient in the Langmuir model. However, the Freundlich equation indicates the minor adsorption capacity of the sludge compared to the GAC Ž1000-fold.. Analogous tests were carried out with EtOH and resulted in the following model parameter values. For the Freundlich isotherms: ny1 F,GAC s 0.83, K GAC s 0.035 and R 2 s 0.99; ny1 F,sludge s 1.93, K sludge s 2.65 E y 7 and R 2 s 0.96. For the Langmuir isotherms: aGAC s 2.19, X m,GAC s 64.10 and R 2 s 0.81; and asludge s 1.04, X m,sludge s 2.61 and R 2 s 0.95.

3.3. Electron donor For all dechlorination tests, EtOH was used as the easily degradable carbon source. The volumetric loading rates varied between 1.5 Žbatch tests. and 15 Žreactor B. g COD ly1 dayy1. Higher loading rates of the carbon source enhanced the degradation rate of 1,2DCA. However, this tendency was limited, as higher COD loading rates gave rise to high amounts of volatile organic acids, decreasing the pH and the methanogenic activity. In addition, the latter seemed to be inhibited by the 1,2-DCA itself, although methanogenic bacteria are assumed to play a key role in the dechlorination of 1,2-DCA. Periods with high degradation rates of 1,2DCA were accompanied by an increase in the CH 4rCO 2 ratio in the biogas.

3.4. Transformation of 1,2-DCA to ethene and ethane The dechlorination of 1,2-DCA was tested extensively by conducting batch experiments ŽFig. 3.. In living sludge ŽE5., the main transformation of 1,2-DCA was carried out by dichloroelimination reactions, re-

S. De Wildeman et al.D,(,R,216,456,57, r Biophysical Chemistry 6 (2001) 17᎐27

23

Fig. 3. Dechlorination of 1,2-DCA in penicillin bottles. All tests were carried out in the presence of 40 ppm of 1,2-DCA. AE: autoclaved sludge adapted to EtOH; control: 40 ppm 1,2-DCA solution in demineralised water with a nitrogen headspace; CcE: sludge adapted to EtOH without COD pulse q addition of cyanocobalamin in molar ratio 1,2-DCArcobalamin of 2; CE: idem without addition of cobalamin; Ec5: pulse feed of EtOH Ž5 g CODrl. q addition of cyanocobalamin in molar ratio 1,2DCArcobalamin of 2; E5: idem without addition of cyanocobalamin. Statistical 95% intervals were taken at days 1, 7 and 8 for tests AE, CE and E5.

sulting in the production of ethene Ž65᎐80% recovery.. Another type of reductive dechlorination mechanism, reductive hydrogenolysis, produced only small amounts of ethane Ž- 1%.. In the latter conversion process, ethylchloride was the only chlorinated intermediate detected. It was also present at low concentrations Ž- 50 ␮g ly1 . during the operation of the UASB reactors. The addition of cobalamin ŽEc5. in a molar ratio of 1,2-DCArcobalamin of 2 had no positive effect on the conversion rate of 1,2-DCA, nor did it change the degradation pathway. Autoclaved granular sludge ŽAE. almost overlapped with the control, the latter dropping in concentration due to spontaneous degradation of 1,2-DCA in non-strictly anaerobic and darkened conditions Žreactor conditions.. This drop could be due to absorption and diffusion phenomena of 1,2-DCA caused by rubber stoppers. Recently, in our laboratory, it was confirmed that the use of VITON stoppers enhanced the recovery of chlorinated compounds and their endproducts. Besides this, a conversion of 1,2DCA into carbon dioxide cannot be excluded ŽBouwer and McCarty, 1983.. In tests without the addition of electron donor ŽCE and CcE., 1,2-DCA changed very little. These tests confirmed the need for metabolic activity to enhance the reductive dechlorination reactions of 1,2-DCA. Only a slight removal of 1,2-DCA occurred when the chlorinated compound was stored in a sterile medium in the dark with a nitrogen headspace Ž5 vol.% of the total volume.; i.e. less than 10% within 10 days.

3.5. Biological dechlorination To elucidate the dechlorination capacity of free enzymatic centra of the unadapted methanogenic granu-

lar sludge, cells of the biomass were extracted and filtered over a sterile membrane Ž0.22 ␮m.. The filtered extract of the biomass was mixed with the synthetic influent water of the UASB reactors. For 10 days, the dechlorination of the 1,2-DCA was measured Ždata not shown.. The starting concentration of 1,2DCA was set at 50 mg ly1. Relative to the control tests, only low degradation kinetics of 1,2-DCA were observed Ž- 1 mg ly1 dayy1 .. Ethene and ethane were detected in the gas phase, indicating that the reductive dechlorination mechanisms were carried out at very low catalytic turnover. Chlorinated intermediates or endproducts could not be detected. The same procedure was established in the test with autoclaved sludge. Degradation curves were not significantly different from those of controls. No intermediates were found.

3.6. Toxicity tests towards anaerobic digestion Concentrations of 1,2-DCA tested on the unadapted granular methanogenic sludge ranged from 0 to 500 mg ly1, with EtOH as the sole carbon source Ž2 g ly1 COD. ŽFig. 4.. It was found that concentrations lower than 90 mg ly1 gave no significant toxicity within 12 h towards the amount of biogas production of the biomass. Moreover, the composition of the biogas only gave a significant decrease in the CH 4rCO 2 ratio with concentrations higher than 250 mg ly1. At lower concentrations Ž- 250 mg ly1 ., the biogas produced had a composition of 76᎐78 vol.% of CH 4 and 22᎐24 vol.% of CO 2 , while at higher concentrations, the production of CH 4 decreased almost linearly to 68 vol.% at 500 mg ly1 Ždata not shown.. The amount of 1,2-DCA in the water thus inhibits the overall biogas production

24

S. De Wildeman et al.D,(,R,216,456,57, r Biophysical Chemistry 6 (2001) 17᎐27

Fig. 4. Influence of increasing 1,2-DCA concentrations on the methanogenic activity of granular anaerobic sludge.

and simultaneously tends to decrease the CH 4 formation.

3.7. Nitrification bio-assay The Nitrox results ŽFig. 5. indicated that 1,2-DCA has an inhibitory effect on the nitrification process. At a 1,2-DCA concentration of 100 mg ly1 in the influent, inhibition of the nitrification process reached a value of almost 40% ŽFig. 5.. Up to 20 mg ly1 , a small toxic

effect towards Nitrosomonas species was observed. Moreover, the inhibition effect stabilized at 40% for concentrations of 1,2-DCA up to 500 mg ly1 Ždata not shown.. Equal concentrations of 1,2-DCA within different media Žtap water with 1,2-DCA, influent and effluent of the UASB reactors. did not change the inhibition characteristics. After dechlorination of 1,2DCA in the model influent in the UASB reactors, almost no inhibition could be detected. As the concentrations of 1,2-DCA in the effluents of both UASB

Fig. 5. Inhibition curves Ž%. of the nitrification process of nitrifying sludge in the presence of 1,2-DCA. To account for the possible variability of nitrifying sludge, two totally independent measurements were made: measurement A Žtime x . Ž⽧. and measurement B Žtime x q 12 months. ŽB.. The measurements were carried out with different sludgerwater sample ratios, ranging from 1 to 9 Ž100r100 ml to 180r20 ml., the latter ratio being optimal in terms of nitrification capability of the nitrifying sludge and the dilution effect of the watersample.

S. De Wildeman et al.D,(,R,216,456,57, r Biophysical Chemistry 6 (2001) 17᎐27

reactors never exceeded 40 mg ly1 and had an average value that was lower than 20 mg ly1 , the average inhibition effect was found to be almost 5%. The inhibition similarity between the reference samples and the effluents indicates that no toxic compounds are produced during the conversion process of 1,2-DCA.

4. Discussion This work reveals that living methanogenic granular sludge grown in anaerobic wastewater systems ŽUASB reactor. is able to degrade 1,2-DCA without any adaptation. During the period of the tests Ž90 days., there was no indication of gradual acceleration of dechlorinating capacity ŽFig. 1.. An enhancement of the degradation rate was achieved by increasing the COD loading rates, although this procedure was limited by the production of volatile fatty acids, causing a pH drop. As the pH should not be lower than 6.8 to guarantee good metabolic activity of the methanogenic biomass ŽKato et al., 1997., a maximum loading rate of COD was found to be 15 g ly1 dayy1. Although UASB reactors are known to have the capacity to deal with loading rates up to 30 g ly1 dayy1 ŽKato et al., 1997., the presence of 1,2-DCA Ž20᎐100 mg ly1 . seemed to interfere in the acetotrophic methanogenesis. Adsorption of 1,2-DCA seemed to play a minor role in the removal processes. The fact that no breakthrough of 1,2-DCA was detected in reactor B indicates that the removal of 1,2-DCA was not only due to the adsorption on GAC. Moreover, the degradation rate in reactor A Žwithout GAC. exceeded half of that of reactor B. This indicates that the addition of activated carbon primarily had a stabilizing effect, buffering the anaerobic biomass against variable inputs of 1,2-DCA. The advantages of combining GAC with microbial processes were investigated in several previous experiments ŽBouwer and McCarty, 1982; Suidan et al., 1996; Khodadoust et al., 1997.. In tests with nitrifying sludge, the nitrifying bacteria appeared to be sensitive to 1,2-DCA ŽFig. 5.. Probably, only a part of the involved bacteria andror enzymes for nitrification are blocked by 1,2-DCA. Speece Ž1996. reported an IC 50 value of 29 mg 1,2-DCA ly1 for Nitrosomonas. In general, the influence of different effluents on the nitrification process was studied earlier ŽGruttner et al., 1994.. It was found that heavy metals, such as copper, zinc and nickel, but also cyanide and organic chlorinated compounds, have an inhibiting effect on nitrification processes. The results of the inhibition tests with 1,2-DCA presented in this paper clearly indicate that the anaerobic treatment of 1,2-DCA contaminated wastewaters in UASB reactors is a protec-

25

tive option before subjecting the water to an aerobic treatment plant. Different mechanisms can play a role in the degradation of 1,2-DCA by granular methanogenic sludge. However, as ethene and small amounts of ethane are the only endproducts of the degradation of 1,2-DCA, the reductive dechlorination reactions established in the literature Žvan Eekert et al., 1999. are envisioned as the main conversion mechanisms. The small concentrations of ethylchloride in the effluent and the high ethene recovery of 1,2-DCA in the gas produced indicate that dichloroelimination of 1,2-DCA was a preferred reaction mechanism in relation to a successive reductive hydrogenolysis. As was confirmed earlier ŽHanselmann, 1991., dichloroelimination reactions are favored in relation to reductive hydrogenolysis reactions on the basis of their enthalpic reaction values. In this work, EtOH was used as the only additional carbon source. EtOH is known as an electron donor, generating higher amounts of hydrogen than are produced by propionate or butyrate ŽFennell et al., 1997.. As at least some specific dechlorinating bacteria are able to outcompete methanogens at very low hydrogen concentration ŽSmatlak et al., 1997., the latter organisms are envisioned to carry out the conversions of 1,2-DCA investigated in the experiments. However, Holliger et al. Ž1990. found that ethane was the main conversion product of several methanogenic species. In contrast, dominant ethene production was only seen by specific dechlorinating strains, such as Dehalococcoides ethenogenes 195 ŽMaymo-Gatell et al., 1999.. Although this work focussed on reactor technological aspects, further microbiological research to elucidate the role and long-term stability of different dechlorinating strains in the granular sludge is warranted. This work confirms the hypothesis that higher hydrogen concentrations stimulate the 1,2-DCA degradation mechanism ŽBallapragada et al., 1997.. This can be deduced from the reactor patterns when the COD volumetric loading rates were increased. In cases where the higher COD loading rates resulted in higher metabolic activity, producing relatively more biogas, degrading more 1,2-DCA and increasing the CH 4rCO 2 ratio, the relation between the catalytic degradation, hydrogen production and methane formation is rather evident. Corrinoid centra are known to play a role in methane formation and to be a catalyst in reductive dechlorination reactions Ž Ferry, 1992 . . Some methanogenic strains are able to carry out both reactions by the same catalyzing centra ŽHolliger et al., 1992.. In view of the dominance of methanogens in the reactors, the dechlorination phenomena observed are probably based on the involvement of corrinoids. Extracts of the anaerobic methanogenic cells and autoclaved biomass were almost unable to degrade the

26

S. De Wildeman et al.D,(,R,216,456,57, r Biophysical Chemistry 6 (2001) 17᎐27

1,2-DCA. This suggests the fact that the degradation is driven by electron flows derived from anaerobic digestion processes normally giving rise to the formation of methane. Both dichloroelimination and reductive hydrogenolysis reactions can be carried out by electrons from the strong dechlorinative cobalt centra of the corrinoids Žvan Eekert, 1999.. In our experiments, free cobalamin complexes and autoclaved sludge did not degrade 1,2-DCA. Redox measurements of the water phase at the top of the reactors ranged from y200 to y300 mV, where optimal conditions for dechlorination chemistry with corrinoid mediators need stronger reductive conditions ŽNeumann et al., 1995; Holliger et al., 1998.. Based on the degradation rates obtained in reactors A and B in this work, it appears feasible to develop UASB reactors treating high volumetric flows of water containing high concentrations of 1,2-DCA Žup to 50 mg ly1 ..

5. Conclusions The reactor technology studied was able to detoxify 1,2-DCA-contaminated waters. Concentrations of 1,2DCA between 20 and 100 mg ly1 were dechlorinated to ethene. The best configuration degraded the toxic compound for 82% at 72 mg ly1 dayy1, which is approximately 20-fold faster than other anaerobic bioreactors reported in the literature ŽWild et al., 1995.. The rate of dechlorination of 1,2-DCA was closely related to methane production. As the endproducts of the degradation processes consist of non-toxic molecules, the use of UASB reactors, inoculated with methanogenic granular sludge and operated under-well defined conditions, could be an attractive purification tool to remediate organochlorine polluted waters.

Acknowledgements This manuscript was supported by the Flemish Institute for the Improvement of Scientific and Technological Research in Industry ŽIWT.. References Ballapragada, B.S., Stensel, H.D., Puhakka, J.A., Ferguson, J.F., 1997. Effect of hydrogen on reductive dechlorination of chlorinated ethenes. Environ. Sci. Technol. 31 Ž6., 1728᎐1734. Bouwer, E.J., McCarty, P.L., 1982. Removal of trace chlorinated organic compounds by activated carbon and fixed-film bacteria. Environ. Sci. Technol. 16 Ž12., 836᎐843. Bouwer, E.J., McCarty, P.L., 1983. Transformations of 1carbon and 2-carbon halogenated aliphatic organic com-

pounds under methanogenic conditions. Appl. Environ. Microbiol. 45 Ž4., 1286᎐1294. Dewulf, J., Drijvers, D., Van Langenhove, H., 1995. Measurement of Henry’s law constant as function of temperature and salinity for the low-temperature range. Atmos. Environ. 29, 323᎐331. Dossantos, L.M.F., Livingston, A.G., 1995. Novel membrane bioreactor for detoxification of VOC wastewaters ᎏ biodegradation of 1,2-dichloroethane. Water Res. 29 Ž1., 179᎐194. Fathepure, B.Z., Boyd, S.A., 1988. Dependence of tetrachloroethylene dechlorination on methanogenic substrate consumption by Methanosarcina sp. strain DCM. Appl. Environ. Microbiol. 54 Ž12., 2976᎐2980. Fennell, D.E., Gossett, J.M., Zinder, S.H., 1997. Comparison of butyric acid, ethanol, lactic acid, and propionic acid as hydrogen donors for the reductive dechlorination of tetrachloroethene. Environ. Sci. Technol. 31 Ž3., 918᎐926. Ferry, J.G., 1992. Biochemistry of methanogenesis. Crit. Rev. Biochem. Mol. Biol., 27Ž6., 473-503. Gruttner, H., Winthernielsen, M., Jorgensen, L., Bogebjerg, P., Sinkjaer, O., 1994. Inhibition of the nitrification process in municipal wastewater treatment plants by industrial discharges. Water Sci. Technol. 29 Ž9., 69᎐77. Hage Jacobus, C., Hartmans, S., 1999. Monooxygenase-mediated 1,2-dichloroethane degradation by Pseudomonas sp. strain DCA1. Appl. Environ. Microbiol. 65 Ž6., 2466᎐2470. Hanselmann, K.W., 1991. Microbial energetics applied to waste repositories. Experimentia 47 Ž7., 645᎐687. Holliger, C., Schraa, G., Stams, A.J., Zehnder, A.J., 1990. Reductive dechlorination of 1,2-dichloroethane and chloroethane by cell suspensions of methanogenic bacteria. Biodegradation 1 Ž4., 253᎐261. Holliger, C., Schraa, G., Stupperich, E., Stams, A., Zehnder, A., 1992. Evidence for the involvement of corrinoids and Factor F430 in the reductive dechlorination of 1,2-dichloroethane by Methanosarcina barkeri. J. Bacteriol. 174 Ž13., 4427᎐4434. Holliger, C., Wohlfarth, G., Diekert, G., 1998. Reductive dechlorination in the energy metabolism of anaerobic bacteria. FEMS Microbiol. Rev. 22 Ž5., 383᎐398. Kato, M., Field, J., Lettinga, G., 1997. The anaerobic treatment of low strength wastewaters in UASB and EGSB reactors. Water Sci. Technol. 36 Ž6r7., 375᎐382. Khodadoust, A.P., Wagner, J.A., Suidan, M.T., Brenner, R.C., 1997. Anaerobic treatment of PCP in fluidized-bed GAC bioreactors. Water Res. 31 Ž7., 1776᎐1786. Maymo-Gatell, X., Anguish, T., Zinder, S.H., 1999. Reductive dechlorination of chlorinated ethenes and 1,2-dichloroethane by Dehalococcoides ethenogenes 195. Appl. Environ. Microbiol. 65 Ž7., 3108᎐3113. McCann, J., Simmon, V., Streitwieser, D., Ames, B.N., 1975. Mutagenicity of chloroacetaldehyde, a possible metabolic product of 1,2-dichloroethane Žethylene dichloride., chloroethanol Žethylene chlorohydrin., vinyl chloride, and cyclophosphamide. PNAS USA 72 Ž8., 3190᎐3193. Neumann, A., Wohlfarth, G., Diekert, G., 1995. Properties of tetrachloroethene and trichloroethene dehalogenase of Dehalospirillum multi¨ orans. Arch. Microbiol. 163, 276᎐281.

S. De Wildeman et al.D,(,R,216,456,57, r Biophysical Chemistry 6 (2001) 17᎐27 Smatlak, C.R., Gossett, J.M., Zinder, S.H., 1997. Comparative kinetics of hydrogen utilization for reductive dechlorination of tetrachloroethene and methanogenesis in an anaerobic enrichment culture. Environ. Sci. Technol. 30 Ž9., 2850᎐2858. Speece, R.E., 1996. Anaerobic Biotechnology for Industrial Wastewater. Archae Press, Nashville, Tenessee ŽUSA. Ž393 pp.. Stucki, G., Krebser, U., Leisinger, T., 1983. Bacterial growth on 1,2-dichloroethane. Experimentia 39 Ž11., 1271᎐1273. Stucki, G., Thuer, M., 1995. Experiences of a large-scale application of 1,2-dichloroethane-degrading microorganisms for groundwater treatment. Environ. Sci. Technol. 29 Ž9., 2339᎐2345. Suidan, M.T., Flora, J.R.V., Boyer, T.K., Wuellner, A.M., Narayanan, B., 1996. Anaerobic dechlorination using a fluidized-bed reactor. Water Res. 30 Ž1., 160᎐170. Surmacz Gorska, J., Gernaey, K., Demuynck, C., Vanrolleghem, P., Verstraete, W., 1995. Nitrification process control in activated sludge using oxygen uptake rate measurements. Environ. Technol. 16 Ž6., 569᎐577. Thaveesri, J., Daffonchio, D., Liessens, B., Verstraete, W., Ahring, B.K., 1995. Different types of sludge granules in UASB reactors treating acidified wastewaters. Anaerobic Proc. Bioenergy Environ. 68 Ž4., 329᎐337. van Eekert, M.H.A., 1999. Transformation of Chlorinated Compounds by Methanogenic Granular Sludge. Wageningen University, The Netherlands ŽPhD dissertation, 149 pp.. van Eekert, M.H.A., Stams, A.J.M., Field, J.A., 1999. Gratuitous dechlorination of chloroethanes by methanogenic granular sludge. Appl. Microbiol. Biotechnol. 51 Ž1., 46᎐52.

27

Vogel, T.M., Criddle, C.S., McCarty, P.L., 1987. Transformations of halogenated aliphatic compounds. Environ. Sci. Technol. 21 Ž8., 722᎐736. Wild, A.P., Winkelbauer, W., Leisinger, T., 1995. Anaerobic dechlorination of trichloroethene, tetrachloroethene and 1,2-dichloroethane by an acetogenic mixed culture in a fixed-bed reactor. Biodegradation 6, 309᎐318. Ir. Stefaan De Wildeman, a chemical bio-engineer, is preparing a Ph.D. study in the domain of biocatalytic degradation processes of lower organohalogens in UASB reactors. He has experience in the field of Žheterogeneous. catalysis. Ir. Hendrik Nollet, environmental bio-engineer, focuses on the biodegradation of PCBs in water ŽPh.D. study.. His work is based on experience with biotechnological processes. Prof. Dr. ir. Herman Van Langenhove is head of the Laboratory of Environmental Organic Chemistry. His laboratory has experience in the field of biotreatment of gas and monitoring tools for natural water and gas disposals. Prof. Dr. ir. Willy Verstraete, coordinator of LabMET ŽLaboratory for Microbial Ecology and Technology. has experience in the biotechnological treatment of water and sludge.