- Email: [email protected]
PAH contaminated soils
PII: S0043-1354(00)00475-9
Wat. Res. Vol. 35, No. 10, pp. 2363–2370, 2001 # 2001 Published by Elsevier Science Ltd. Printed in Great Britain 0043-1354/01/$ - see front matter
EFFECTIVENESS OF AN ANAEROBIC GRANULAR ACTIVATED CARBON FLUIDIZED-BED BIOREACTOR TO TREAT SOIL WASH FLUIDS: A PROPOSED STRATEGY FOR REMEDIATING PCP/PAH CONTAMINATED SOILS K.M. KORAN1*, M.T. SUIDAN1, A.P. KHODADOUST1, G.A. SORIAL1, and R.C. BRENNER2 1 Department of Civil and Environmental Engineering, University of Cincinnati,741 Baldwin Hall, Cincinnati, OH 45221-0071, USA and 2 US Environmental Protection Agency National Risk Management Laboratory, 26 W. Martin Luther King Drive, Cincinnati, OH 45268, USA
(First received 1 April 2000; accepted in revised form 1 September 2000) Abstract}An integrated system has been developed to remediate soils contaminated with pentachlorophenol (PCP) and polycyclic aromatic hydrocarbons (PAHs). This system involves the coupling of two treatment technologies, soil-solvent washing and anaerobic biotreatment of the extract. Specifically, this study evaluated the effectiveness of a granular activated carbon (GAC) fluidized-bed reactor to treat a synthetic-waste stream of PCP and four PAHs (naphthalene, acenaphthene, pyrene, and benzo(b)fluoranthene) under anaerobic conditions. This waste stream was intended to simulate the wash fluids from a soil washing process treating soils from a wood-preserving site. The reactor achieved a removal efficiency of greater than 99.8% for PCP with conversion to its dechlorination intermediates averaging 46.5%. Effluent, carbon extraction, and isotherm data also indicate that naphthalene and acenaphthene were removed from the liquid phase with efficiencies of 86 and 93%, respectively. Effluent levels of pyrene and benzo(b)fluoranthene were extremely low due to the high-adsorptive capacity of GAC for these compounds. Experimental evidence does not suggest that the latter two compounds were biochemically transformed within the reactor. # 2001 Published by Elsevier Science Ltd. Key words}anerobic, chlorinated phenol, fluidized bed, expanded bed, PAHs
INTRODUCTION
Contamination of soils and natural waters with wood-treating chemicals is a widespread environmental problem. Pentachlorophenol (PCP), a pesticide and a suspected carcinogen, has been used extensively in the United States as a wood-preserving agent and is one of the primary contaminants found in soils at abandoned and existing wood-preserving sites. In addition to PCP, soils at these sites are often contaminated with polycyclic aromatic hydrocarbons (PAHs), aliphatic and aromatic hydrocarbons, and heavy metals such as copper, chromium, or arsenic. Removal of these contaminants from soils, natural waters, and industrial wastewaters has been mandated by local, state, and federal regulatory agencies. Stringent regulations have led to the development of several remediation technologies for the removal of pesticides from soils. Khodadoust et al. (1999) developed an effective solvent washing procedure for the removal of PCP and PAHs from soils. This procedure, which employs an ethanol–water solution *Author to whom all correspondence should be addressed.
as the solvent, could be used in either in-situ or ex-situ soil washing applications. New technologies that emphasize the detoxification and destruction of these contaminants following their removal from soil are also being developed. Bioremediation, the use of microorganisms or microbial processes to detoxify and degrade environmental contaminants, is one of these growing technologies. Highly halogenated compounds such as PCP are typically less susceptible to aerobic biological treatment than non-halogenated compounds. For highly halogenated contaminants, following anaerobic treatment with aerobic treatment has been shown to be an effective remediation approach. Studies by Wilson et al. (1995) and Khodadoust et al. (1997) have demonstrated that PCP can be degraded anaerobically in a GAC fluidized-bed reactor even at low-empty-bed contact times (EBCTs). Equimolar conversion of PCP to monochlorophenol (MCP) was achieved at an EBCT as low as 1.16 h, and mineralization was shown to occur under these conditions (Wilson et al., 1995). Complete conversion of PCP to carbon dioxide (CO2), methane (CH4) and chloride (Cl) was accomplished using a two-stage
2363
2364
K. M. Koran et al.
anaerobic–aerobic treatment process (Wilson et al., 1996). While PAHs are known to degrade under aerobic conditions, these compounds are typically recalcitrant to strict anaerobic treatment (Hambrick et al., 1980; Madsen et al., 1996). Numerous researchers have shown that some low molecular weight PAHs, such as naphthalene and acenaphthene, can be degraded under denitrifying conditions (Mihelcic and Luthy, 1988; Bregnard et al., 1996) and under sulfate-reducing conditions (Bedessem et al., 1997; Coates et al., 1996; Langenhoff et al., 1996). The mineralization of various aromatic and aliphatic hydrocarbons can also be coupled to the reduction of Fe(III) and Mn(IV) (Baedecker et al., 1993; Essaid et al, 1995; Thierrin et al., 1995). Relatively few studies have confirmed the degradation of naphthalene under methanogenic conditions. Many researchers have reported no success in degrading naphthalene under such conditions (Edwards and Grbic-Galic, 1994; Madsen et al., 1996; Hambrick et al., 1980). Results from an anaerobic digester study (Parker and Monteith, 1995) imply that PAH degradation may be possible under methanogenic conditions. Genthner et al. (1997) observed limited degradation of naphthalene, 1- and 2-methylnaphthalene, biphenyl, 2,6-dimethylnaphthalene, and anthraquinone under methanogenic conditions. None of the 4- or 5-ring PAHs were degraded under methanogenic, sulfidogenic, or nitrate-reducing conditions. The following research was carried out to evaluate the effectiveness of a GAC fluidized-bed reactor to treat a synthetic waste stream of PCP and PAHs under methanogenic conditions. In particular, this work examined the biological degradation of PCP into its various dechlorination intermediates and the removal of naphthalene, acenaphthene, pyrene, and benzo(b)fluoranthene from the system through biological transformation and/or physical adsorption onto the GAC. The removal of PCP, PCP dechlorination intermediates, and the four-feed PAHs from the system by these mechanisms was demonstrated through a mass balance of influent concentrations, effluent concentrations, and solid-phase loadings. Anaerobic adsorption isotherms, were conducted for naphthalene and acenaphthene, and predicted reactor breakthrough curves were generated. A
comparison of measured effluent concentrations and predicted effluent concentrations with time confirmed the primary removal mechanisms for these compounds.
METHODS
Anaerobic GAC fluidized-bed reactor system The reactor system included a water-jacketed main column with a recycle loop, an influent header, an effluent header, a feed system, and an effluent and gas collection system. The main column of the reactor was a Plexiglas tube with an inner diameter of 10.2 cm, a length of 96.5 cm, and a total volume of 11 L including the recycle loop. The reactor temperature was maintained at 358C by circulating water from a constant temperature water bath through the outer jacket of the column. The reactor was initially charged with 1.0 kg of 16 20 US Mesh Filtrasorb 400 GAC (Calgon Corporation, Pittsburgh, PA) and was seeded with a methanogenic culture from a similar anaerobic GAC fluidized-bed reactor treating a synthetic pollutant stream of PCP and ethanol (Khodadoust et al., 1997). Fluidization of the GAC bed, to 30% by volume based on the original carbon volume, was accomplished by recirculation of the effluent at an approximate recycle ratio of 200 : 1. The feed system consisted of a synthetic organic waste stream, a nutrient feed solution (Fox, 1989), and a carbonate buffer solution. The nutrient and buffer solutions were introduced into the suction side of the recycle line using fixed r.p.m. pump drives. Pump drives were controlled by Dayton control timers with an on/off cycle every minute to achieve target flow rates and maintain uniform conditions in the reactor. The synthetic organic waste stream consisted of a mixture of PCP, naphthalene, acenaphthene, pyrene, and benzo(b)fluoranthene in ethanol at influent concentrations of 100, 35, 11, 6 and 0.5 mg/L, respectively (Table 1). Ethanol served as the primary substrate and was fed at a concentration of 690 mg/L. This solution was fed into the recycle line using a high-precision syringe infusion pump with a 10-mL fixed needle syringe via 1/16-in. stainless-steel tubing. System operation Reactor operating parameters are listed in Table 2. The reactor was initially operated at a flow rate of 6 L/d and an EBCT of 9.3 h (Phase I). After several hundred days of stable operation, the organic mass loading rate and the hydraulic loading rate to the reactor were increased in 10% increments every 4 days until the loading rates were doubled (Phase II). Once reactor performance stabilized, the loading rates were again doubled until the desired EBCT of 2.32 h was achieved (Phase III). At day 640, a buffer-feed pump failed and the reactor experienced an operational disturbance due to a sudden
Table 1. Operating parameters for GAC fluidized-bed reactor system (Phases I–VII) Organic loading (g/d)
PCP Naphthalene Acenaphthene Pyrene Benzo(b)fluoranthene Ethanol
Influent conc. (mg/L)
I/V
II/VI
III/VII
IV
Phases I–VII (excl. IV)
0.60 0.21 0.066 0.036 0.003 4.28
1.20 0.42 0.132 0.072 0.006 8.33
2.40 0.84 0.264 0.144 0.012 16.66
0 0 0 0 0 0
100 35 11 6 0.5 690
Anaerobic treatment of soil wash fluids Table 2. Operating parameters for GAC fluidized-bed reactor system Phase I II III IV V VI VII
Days of operation 0–434 435–471 472–639 640–704 705–784 785–949 950–1135
Total flow rate (l/d) 6.0 12.0 24.0 batch/flush 6.0 12.0 24.0
EBCT h 9.30 4.65 2.32 0/2.32 9.30 4.65 2.32
elevation in pH. For a period of 65 days following the pH shock, the reactor was operated as a batch system (Phase IV). This period was devoted to reestablishing an effective biofilm within the reactor. Initially, feed to the reactor was stopped and the reactor was allowed to operate as a batch system in order to reduce the high aqueous phase chemical oxygen demand (COD) already present. The existing biomass, however, was unable to treat this COD, so the reactor was flushed with effluent from a reactor used to initially seed this reactor. This flushing process was intended to dilute the high COD and reintroduce active biomass into the reactor. Once the COD was reduced to levels experienced during normal operating conditions, feed to the reactor was resumed at the initial loading conditions under Phase I. This marks the beginning of Phase V and corresponds to an EBCT of 9.2 h. The loading rates were increased in two stages as before until the desired EBCT of 2.32 h was achieved. Thus, the reactor in this study underwent seven phases of operation, each corresponding to a different EBCT as summarized in Table 2. Analytical methods The reactor was monitored daily for effluent pH, total gas volume production, buffer and nutrient feed rates, and syringe feed rate. Weekly effluent samples were analyzed for COD, gas composition, chlorides, volatile fatty acids (VFAs), alcohols, and concentrations of PCP, PCP dechlorination intermediates, and the four-feed PAHs. The concentrations of PCP and its dechlorination intermediates were analyzed using a gas chromatograph (GC) (HewlettPackard, Palo Alto, CA, Model 5890 Series II) with an electron capture detector (ECD) (Hewlett-Packard, Palo Alto, CA). The concentrations of the four PAHs were measured using the same GC euqipped with a flame ionization detector (FID) (Hewlett-Packard, Palo Alto, CA). Duplicate samples of GAC and the associated biomass, were periodically withdrawn from the reactor and extracted to quantitatively determine masses of adsorbed compounds (method outlined by Fox, 1989). A detailed listing of the analytical methods used in this study is given by Miller et al. (1998). Isotherm setup To determine the adsorptive capacity of GAC for naphthalene and acenaphthene, GAC adsorption isotherm experiments were conducted using the bottle point method. The experiments were performed under anoxic conditions at 358C and in the absence of biological activity according to the methods outlined in Miller et al. (1998). The activated carbon used in this study was Calgon’s Filtrasorb-400, US mesh 16 20, the same carbon used to initially charge the fluidized-bed reactor. The isotherm studies were run at two initial solution concentrations: 15 and 25 mg/L for naphthalene and 1.5 and 2.5 mg/L for acenaphthene. Bottles were tumbled on a rotary shaker for 14 days to establish equilibrium. An equilibrium time of 14 days was determined
2365
to be sufficient by Vidic et al. (1990). Once equilibrium was established, the concentration of adsorbate on the carbon, qe , was calculated from the measured aqueous concentration and plotted against the aqueous concentration, Ce , to obtain an adsorption isotherm curve. The calculated value of qe was obtained from the equation: V ð1Þ m where C0 is the initial liquid phase concentration of solute, V the solution volume, and m the mass of adsorbate. Carbon extractions were performed to determine extraction efficiencies relative to the calculated values. For both naphthalene and acenaphthene, Soxhlet extraction yielded extraction efficiencies of greater than 95%. The Freundlich isotherm equation provided the best fit of the isotherm data (Freundlich, 1906): qe ¼ ðC0 Ce Þ
qe ¼ KCe1=n
ð2Þ
The Freundlich parameters K and 1=n were obtained by nonlinear least-squares regression analysis. From the isotherm data, predicted breakthrough curves for naphthalene and acenaphthene were generated. Equilibrium between the aqueous- and solid-phase concentrations was assumed. Furthermore, it was also assumed that only adsorption took place and no adsorbate–adsorbate interaction occurred. Since the reactor was operated at a recycle ratio of 200 : 1, CSTR hydraulics were assumed. Therefore, a material balance around the expanded-bed bioreactor yields the following first-order differential equation: dc dq þ Mc ð3Þ dt dt where Q is the flow through the reactor; C0 is the influent solute concentration; Ce is the effluent solute concentration; Vt the total volume of reactor; Vc the volume of carbon; Mc the mass of carbon; and q the mass loading on carbon. The above equation was then solved using the fourth-order Runge–Kutta method to obtain the predicted reactor effluent concentration with time for a single-solute system. This predicted breakthrough curve was compared to the observed breakthrough curve for both naphthalene and acenaphthene to analyze the fate of these compounds within the system. QC0 QCe ¼ ðVt Vc Þ
RESULTS AND DISCUSSION
Treatment of chlorophenols by anaerobic GAC bioreactor The molar effluent concentrations of PCP and its chlorinated phenolic intermediates throughout each phase of operation are presented in Fig. 1. The molar influent concentration of PCP was 0.376 mmol/L throughout the study with the exception of Phase IV, during which time feed to the reactor was stopped. The concentration of each of the dechlorination intermediates represents the sum of all measured isomers. The first two phases of operation were characterized by stable operation, with low COD, acid and alcohol concentrations (averaging 59.8, 1.44, and 1.37 mg/L, respectively). Most of the organics were being removed through adsorption onto the GAC, including the four-feed PAHs, which remained at concentrations less than 0.1 mg/L. During this period, the reactor experienced a steady increase in concentration of MCP in the effluent (Fig. 1). The
2366
K. M. Koran et al.
Fig. 1. Effluent concentrations for PCP and dechlorination intermediates as a function of reactor operating time.
Fig. 2. Effluent concentrations for phenol and select MCP and DCP species.
concentrations of all other dechlorination intermediates remained below 3 103 mmol/L, while PCP, tetrachlorophenol (TeCP) and trichlorophenol (TCP) concentrations remained below 1 104 mmol/L. This corresponds to a greater than 99.9% removal of PCP from the influent waste stream. Figure 2 displays the effluent concentrations for phenol and selected MCP and dichlorophenol (DCP) species. Para-chlorophenol (4-CP) and meta-chlorophenol (3-CP) were the primary MCP species present, with 4-CP dominating throughout the first two phases. 3,4-DCP was the primary DCP species present during this period, followed by 3,5-, 2,5- and trace amounts of 2,4-DCP. By the end of Phase I, 3-CP was no longer detected in the effluent. However,
the concentration of 4-CP continued to increase until the middle of Phase III. On day 550, a spike was observed in the effluent COD (810 mg/L) and acetate (187 mg/L) concentrations, indicating stress to the system had occurred. Also at this time, a shift was noted in the predominant MCP species present from 4-CP to 3-CP, the more readily anaerobically biodegradable of the two isomers. 3-CP continued to rise in the effluent as 4-CP declined, and by day 680 the reactor had achieved a near equimolar conversion of PCP to MCP. Following the appearance of 3-CP, a corresponding increase in phenol was noted and that increase continued through the end of phase III.
Anaerobic treatment of soil wash fluids
Following the period of upset in Phase IV, the concentrations of 3-CP, 4-CP, and 3,4-DCP steadily declined while all other chlorinated intermediates remained below 104 mmol/L. By day 880, the molar sum of all dechlorination intermediates accounted for less than 10% of the molar influent concentration of PCP. Following day 880, a shift from 3-CP to 4-CP as the predominant intermediate was observed. PCP removal efficiencies averaged 99.8% for the last three phases of operation, with conversion to measured intermediates averaging 46.5%. MCP and phenol accounted for 94.5% of the intermediates measured in the effluent during this period. Carbon extractions were performed monthly to determine if adsorption could explain the difference between influent PCP and effluent phenolic concentrations. The mass of MCP and intermediates adsorbed to the carbon, however, decreased with reactor operating time and were fairly constant during the last phases of operation (Miller et al., 1998). Thus, the difference between the measured intermediates and the influent PCP is likely due to the complete mineralization of PCP. Throughout this phase, the effluent COD, VFA and alcohol concentrations remained stable. Treatment of PAHs by anaerobic GAC bioreactor Figure 3 shows the effluent concentrations of naphthalene and acenaphthene throughout the course of reactor operation. During the first 400 days of operation, concentrations of all PAHs remained extremely low in the effluent due to the strong affinity of GAC for PAHs. Naphthalene, which was fed at a concentration of 35 mg/L, was present in the effluent at concentrations below 0.2 mg/L. Throughout the entire study, levels of pyrene and benzo(b)fluoranthene remained nearly
2367
constant at concentrations below 0.01 mg/l. Since the effluent concentrations of pyrene and benzo(b)fluoranthene were so low and there was no evidence to suggest that these compounds were undergoing any chemical transformation within the reactor, these compounds will be omitted from the remainder of this discussion. By the end of Phase I, naphthalene was present in the effluent at approximately 0.4 mg/L and these concentrations continued to increase until the end of phase III when the reactor experienced the pH shock. By this time, acenaphthene had similarly begun an increasing trend in the effluent. The concentrations of naphthalene and acenaphthene in the effluent at the time of the upset were 3.8 and 0.1 mg/L, respectively. When the nutrient feed pump failed, the aqueous pH of the system increased to approximately 9.2, causing mass desorption of compounds from the GAC. Naphthalene concentrations in the effluent at this elevated pH were extremely high, creating a toxic environment for the biological community. This resulted in a loss of active biomass and the subsequent failure of the system. During the recovery period following the pH shock, effluent concentrations returned to levels experienced under stable operating conditions prior to the upset. The effluent concentrations of naphthalene and acenaphthene continued to increase during Phases V and VI. By the end of Phase VI, effluent concentrations of naphthalene had leveled off reaching an apparent steady-state condition. Concentrations of naphthalene during this time fluctuated around a mean of approximately 5 mg/L. Similarly during Phase VII, acenaphthene levels appeared to stabilize around an average concentration of 0.8 mg/ L. These concentrations represent removal efficiencies of 86% for naphthalene and 93% for
Fig. 3. Effluent PAH concentrations as a function of reactor operating time.
2368
K. M. Koran et al.
acenaphthene. Carbon extraction data confirm that these compounds were not accumulating on the GAC during these phases (Miller et al., 1998).
Isotherm studies In order to determine whether naphthalene was being degraded in the system, a GAC single-solute adsorption isotherm study was performed (Fig. 4) and, from these data, a predicted breakthrough curve was generated for naphthalene within the GAC fluidized-bed system (Fig. 5). The Freundlich equation offered the best fit of the data with an r2 value of 0.96. The Freundlich regression parameters K and 1=n were 203.1 and 0.21, respectively. This isotherm equation defines the relationship between the aqueous solute concentration, Ce , and the solid phase loading, q. Using the isotherm regression constants, a predicted breakthrough curve was generated by integrating equation (3) using the fourth-order Runge–Kutta method. This breakthrough curve is the predicted aqueous phase concentration for naphthalene assuming no other compounds are competing for adsorption on the GAC and no biological activity is occurring within the system. Since competitive adsorption does occur within the reactor, the solid-phase concentration of any single solute is reduced, and breakthrough may occur earlier than predicted. For the first 520 days, observed effluent naphthalene concentrations followed the predicted breakthrough curve closely. After day 520, effluent concentrations remained much lower than the predicted breakthrough curve. This difference between the observed concentration and the predicted concentration suggests that naphthalene was being removed by some mechanism other than adsorption
alone, such as biological degradation or transformation. A similar isotherm study was performed for acenaphthene (Fig. 6), and a predicted breakthrough curve was generated (Fig. 7). The Freundlich regression parameters K and 1=n were 319.8 and 0.25, respectively, and the r2 value was 0.97. The observed effluent concentrations of acenaphthene followed the predicted breakthrough curve for the first 640 days. After day 640, effluent concentrations of acenaphthene were higher than those predicted by the breakthrough curve. Since the predicted breakthrough curve ignores the effect of competitive adsorption, it represents the latest estimate for breakthrough. Thus, this study suggests that carbon adsorption was a primary means of removal for acenaphthene. These data concur with PAH mass balance data that revealed the mass of acenaphthene removed from the system was nearly equivalent to the mass of that compound adsorbed on the GAC (Miller et al., 1998), as summarized in Table 3. Experimental data do not reveal whether pyrene or benzo(b)fluoranthene were undergoing any chemical transformation within the reactor. However, since after 1200 days of operation there was no evidence of these compounds in the effluent, their removal from the influent wastewater stream could be achieved through periodic carbon replacement. A carbon replacement schedule of 10% every 4 months would be sufficient to remove these contaminants from the influent waste stream. The spent GAC could be thermally regenerated and reintroduced to the reactor.
CONCLUSIONS
The objective of this work was to demonstrate that an anaerobic GAC fluidized-bed bioreactor could
Fig. 4. Anaerobic adsorption isotherm for naphthalene.
Anaerobic treatment of soil wash fluids
2369
Fig. 5. Breakthrough curves for naphthalene for the GAC fluidized-bed reactor system.
Fig. 6. Anaerobic adsorption isotherm for acenaphthene.
effectively treat a synthetic waste stream containing contaminants commonly found in wood-treating wastes. During periods of stable operation following the upset in Phase IV, the reactor achieved greater than 99.8% removal of PCP from the influent waste stream with conversion to measured intermediates averaging 46.5%. Carbon extractions confirm that these compounds were not accumulating on the GAC. Thus, the difference between measured intermediates and influent PCP is likely due to the complete mineralization of PCP. An aerobic fluidized-bed reactor is recommended to remove residual phenol intermediates in the anaerobic reactor effluent.
The reactor removed naphthalene and acenaphthene from the influent waste stream with efficiencies of 86 and 93%, respectively. Carbon extraction, isotherm, and predicted breakthrough data suggest that naphthalene was being removed by some mechanism other than adsorption alone, such as biological degradation. Carbon adsorption was the primary means of removal for acenaphthene. Experimental data do not reveal whether pyrene and benzo(b)fluoranthene were undergoing any chemical transformation within the reactor. However, their removal from the influent wastewater stream could be achieved through periodic carbon replacement.
2370
K. M. Koran et al.
Fig. 7. Breakthrough curves for acenaphthene for the GAC fluidized-bed reactor system. Table 3. Cumulative mass balance for naphthalene and acenaphthene Mass (g)
Naph
Acen
Cumulative influent Cumulative effluent Mass removed Mass adsorbed on GAC Difference (due to biodegradation)
409.3 22.2 387.1 148.0 239.1
128.8 1.6 127.2 114.6 12.6
REFERENCES
Baedecker M. J., Cozzarelli I. M., Eganhouse R. P., Siegel D. I. and Bennett P. C. (1993) Crude oil in a shallow sand and gravel aquifer: III. Biogeochemical reactions and mass balance modeling in anoxic groundwater. Appl. Geochem. 8, 569–586. Bedessem M. E., Swoboda-Colberg N. G. and Colberg P. J. S. (1997) Naphthalene mineralization coupled to sulfate reduction in aquifer-derived enrichments. FEMS Microbiol. Lett. 152(2), 213–218. Bregnard T., Hoehener P., Haener A. and Zeyer J. (1996) Degradation of weathered diesel fuel by microorganisms from a contaminated aquifer in aerobic and anaerobic microcosms. Environmental Toxicology and Chemistry 15(3), 299–307. Coates J. D., Anderson R. T. and Lovely D. R. (1996) Oxidation of polycyclic aromatic hydrocarbons under sulfate-reducing conditions. Appl. Environ. Microbiol. 62(3), 1099–1101. Edwards E. A. and Grbic-Galic D. (1994) Anaerobic degradation of toluene and o-xylene by a methanogenic consortium. Appl. Environ. Microbiol. 60(1), 313–322. Essaid H. I., Bekins B. A., Godsy M. E. and Ean W. (1995) Simulation of aerobic and anaerobic biodegradation processes at a crude oil spill site. Water Resources Research 31, 3309–3327. Fox P. (1989) An innovative reactor design for the treatment of biologically inhibitory wastewater. Ph.D. Dissertation, Department of Civil Engineering, University of Illinois, Urbana-Champaign, IL. Freundlich H. (1906) Ueber die Adsorption in Leosungen. Z. Phys. Chem. 57, 385–470. Genthner B. R. S., Townsend G. T., Lantz S. E. and Mueller J.G. (1997) Persistence of polycyclic aromatic hydrocarbon
components of creosote under anaerobic enrichment conditions. Arch Environ. Contam. Toxicol. 32(1), 99–105. Hambrick G. A., DeLaune R. D. and Patrick W. H. (1980) Effect of estuarine sediment pH and oxidation-reduction potential on microbial hydrocarbon degradation. Appl. Environ. Microbiol. 40(2), 365–369. Khodadoust A. P., Sorial G. A., Wilson G. J., Suidan M. T., Griffith R. A. and Brenner R. C. (1999) An integrated system for remediation of contaminated soils. J. Environ. Eng., ASCE 125(11), 1033–1041. Khodadoust A. P., Wagner J., Suidan M. T. and Brenner R. C. (1997) Anaerobic treatment of PCP in fluidized-bed GAC bioreactors. Water Res. 31(7), 1776–1786. Langenhoff A., Zehnder A. and Schraa G. (1996) Behaviour of toluene, benzene and naphthalene under anaerobic conditions in sediment columns. Biodegradation 7(3), 267–274. Madsen E., Man C. and Bilotta S. (1996) Oxygen limitations and aging as explanations for the field persistence of naphthalene in coal tar-contaminated surface sediment. Environ. Toxicol. Chem. 15(11), 1876–1882. Mihelcic J. R. and Luthy R. G. (1988) Degradation of polycyclic aromatic hydrocarbon compounds under various redox conditions in soil-water systems. Appl. Environ. Microbiol. 54(5), 1182–1187. Miller K. M., Suidan M. T., Sorial G. A., Khodadoust A. P., Acheson C. M. and Brenner R. C. (1998) Anaerobic treatment of soil wash fluids from a wood preserving site. Water Sci. Technol. 38(7), 63–72. Parker W. J. and Monteith H. D. (1995) Fate of polynuclear aromatic hydrocarbons during anaerobic digestion of municipal wastewater sludges. Water Environ. Res. 67(7), 1052–1059. Thierrin J., Davis G. B. and Barber C. (1995) A groundwater tracer test with deuterated compounds for monitoring in-situ biodegradation and retardation of aromatic hydrocarbons. Ground Water 33, 469–476. Vidic R. D., Suidan M. T., Traegner U. K. and Nakhla G. F. (1990) Adsorption isotherms: illusive capacity and role of oxygen. Water Research 24(10), 1187–1195. Wilson G., Khodadoust A., Suidan M., Acheson C., and Brenner R. (1995) Anaerobic/aerobic biodegradation of pentachlorophenol using GAC fluidized bed bioreactors: optimization of the empty bed contact time. Proceedings of the 1995 Water Environment Federation Conference, Water Environment Federation, Los Angeles.
Wat. Res. Vol. 35, No. 10, pp. 2363–2370, 2001 # 2001 Published by Elsevier Science Ltd. Printed in Great Britain 0043-1354/01/$ - see front matter
EFFECTIVENESS OF AN ANAEROBIC GRANULAR ACTIVATED CARBON FLUIDIZED-BED BIOREACTOR TO TREAT SOIL WASH FLUIDS: A PROPOSED STRATEGY FOR REMEDIATING PCP/PAH CONTAMINATED SOILS K.M. KORAN1*, M.T. SUIDAN1, A.P. KHODADOUST1, G.A. SORIAL1, and R.C. BRENNER2 1 Department of Civil and Environmental Engineering, University of Cincinnati,741 Baldwin Hall, Cincinnati, OH 45221-0071, USA and 2 US Environmental Protection Agency National Risk Management Laboratory, 26 W. Martin Luther King Drive, Cincinnati, OH 45268, USA
(First received 1 April 2000; accepted in revised form 1 September 2000) Abstract}An integrated system has been developed to remediate soils contaminated with pentachlorophenol (PCP) and polycyclic aromatic hydrocarbons (PAHs). This system involves the coupling of two treatment technologies, soil-solvent washing and anaerobic biotreatment of the extract. Specifically, this study evaluated the effectiveness of a granular activated carbon (GAC) fluidized-bed reactor to treat a synthetic-waste stream of PCP and four PAHs (naphthalene, acenaphthene, pyrene, and benzo(b)fluoranthene) under anaerobic conditions. This waste stream was intended to simulate the wash fluids from a soil washing process treating soils from a wood-preserving site. The reactor achieved a removal efficiency of greater than 99.8% for PCP with conversion to its dechlorination intermediates averaging 46.5%. Effluent, carbon extraction, and isotherm data also indicate that naphthalene and acenaphthene were removed from the liquid phase with efficiencies of 86 and 93%, respectively. Effluent levels of pyrene and benzo(b)fluoranthene were extremely low due to the high-adsorptive capacity of GAC for these compounds. Experimental evidence does not suggest that the latter two compounds were biochemically transformed within the reactor. # 2001 Published by Elsevier Science Ltd. Key words}anerobic, chlorinated phenol, fluidized bed, expanded bed, PAHs
INTRODUCTION
Contamination of soils and natural waters with wood-treating chemicals is a widespread environmental problem. Pentachlorophenol (PCP), a pesticide and a suspected carcinogen, has been used extensively in the United States as a wood-preserving agent and is one of the primary contaminants found in soils at abandoned and existing wood-preserving sites. In addition to PCP, soils at these sites are often contaminated with polycyclic aromatic hydrocarbons (PAHs), aliphatic and aromatic hydrocarbons, and heavy metals such as copper, chromium, or arsenic. Removal of these contaminants from soils, natural waters, and industrial wastewaters has been mandated by local, state, and federal regulatory agencies. Stringent regulations have led to the development of several remediation technologies for the removal of pesticides from soils. Khodadoust et al. (1999) developed an effective solvent washing procedure for the removal of PCP and PAHs from soils. This procedure, which employs an ethanol–water solution *Author to whom all correspondence should be addressed.
as the solvent, could be used in either in-situ or ex-situ soil washing applications. New technologies that emphasize the detoxification and destruction of these contaminants following their removal from soil are also being developed. Bioremediation, the use of microorganisms or microbial processes to detoxify and degrade environmental contaminants, is one of these growing technologies. Highly halogenated compounds such as PCP are typically less susceptible to aerobic biological treatment than non-halogenated compounds. For highly halogenated contaminants, following anaerobic treatment with aerobic treatment has been shown to be an effective remediation approach. Studies by Wilson et al. (1995) and Khodadoust et al. (1997) have demonstrated that PCP can be degraded anaerobically in a GAC fluidized-bed reactor even at low-empty-bed contact times (EBCTs). Equimolar conversion of PCP to monochlorophenol (MCP) was achieved at an EBCT as low as 1.16 h, and mineralization was shown to occur under these conditions (Wilson et al., 1995). Complete conversion of PCP to carbon dioxide (CO2), methane (CH4) and chloride (Cl) was accomplished using a two-stage
2363
2364
K. M. Koran et al.
anaerobic–aerobic treatment process (Wilson et al., 1996). While PAHs are known to degrade under aerobic conditions, these compounds are typically recalcitrant to strict anaerobic treatment (Hambrick et al., 1980; Madsen et al., 1996). Numerous researchers have shown that some low molecular weight PAHs, such as naphthalene and acenaphthene, can be degraded under denitrifying conditions (Mihelcic and Luthy, 1988; Bregnard et al., 1996) and under sulfate-reducing conditions (Bedessem et al., 1997; Coates et al., 1996; Langenhoff et al., 1996). The mineralization of various aromatic and aliphatic hydrocarbons can also be coupled to the reduction of Fe(III) and Mn(IV) (Baedecker et al., 1993; Essaid et al, 1995; Thierrin et al., 1995). Relatively few studies have confirmed the degradation of naphthalene under methanogenic conditions. Many researchers have reported no success in degrading naphthalene under such conditions (Edwards and Grbic-Galic, 1994; Madsen et al., 1996; Hambrick et al., 1980). Results from an anaerobic digester study (Parker and Monteith, 1995) imply that PAH degradation may be possible under methanogenic conditions. Genthner et al. (1997) observed limited degradation of naphthalene, 1- and 2-methylnaphthalene, biphenyl, 2,6-dimethylnaphthalene, and anthraquinone under methanogenic conditions. None of the 4- or 5-ring PAHs were degraded under methanogenic, sulfidogenic, or nitrate-reducing conditions. The following research was carried out to evaluate the effectiveness of a GAC fluidized-bed reactor to treat a synthetic waste stream of PCP and PAHs under methanogenic conditions. In particular, this work examined the biological degradation of PCP into its various dechlorination intermediates and the removal of naphthalene, acenaphthene, pyrene, and benzo(b)fluoranthene from the system through biological transformation and/or physical adsorption onto the GAC. The removal of PCP, PCP dechlorination intermediates, and the four-feed PAHs from the system by these mechanisms was demonstrated through a mass balance of influent concentrations, effluent concentrations, and solid-phase loadings. Anaerobic adsorption isotherms, were conducted for naphthalene and acenaphthene, and predicted reactor breakthrough curves were generated. A
comparison of measured effluent concentrations and predicted effluent concentrations with time confirmed the primary removal mechanisms for these compounds.
METHODS
Anaerobic GAC fluidized-bed reactor system The reactor system included a water-jacketed main column with a recycle loop, an influent header, an effluent header, a feed system, and an effluent and gas collection system. The main column of the reactor was a Plexiglas tube with an inner diameter of 10.2 cm, a length of 96.5 cm, and a total volume of 11 L including the recycle loop. The reactor temperature was maintained at 358C by circulating water from a constant temperature water bath through the outer jacket of the column. The reactor was initially charged with 1.0 kg of 16 20 US Mesh Filtrasorb 400 GAC (Calgon Corporation, Pittsburgh, PA) and was seeded with a methanogenic culture from a similar anaerobic GAC fluidized-bed reactor treating a synthetic pollutant stream of PCP and ethanol (Khodadoust et al., 1997). Fluidization of the GAC bed, to 30% by volume based on the original carbon volume, was accomplished by recirculation of the effluent at an approximate recycle ratio of 200 : 1. The feed system consisted of a synthetic organic waste stream, a nutrient feed solution (Fox, 1989), and a carbonate buffer solution. The nutrient and buffer solutions were introduced into the suction side of the recycle line using fixed r.p.m. pump drives. Pump drives were controlled by Dayton control timers with an on/off cycle every minute to achieve target flow rates and maintain uniform conditions in the reactor. The synthetic organic waste stream consisted of a mixture of PCP, naphthalene, acenaphthene, pyrene, and benzo(b)fluoranthene in ethanol at influent concentrations of 100, 35, 11, 6 and 0.5 mg/L, respectively (Table 1). Ethanol served as the primary substrate and was fed at a concentration of 690 mg/L. This solution was fed into the recycle line using a high-precision syringe infusion pump with a 10-mL fixed needle syringe via 1/16-in. stainless-steel tubing. System operation Reactor operating parameters are listed in Table 2. The reactor was initially operated at a flow rate of 6 L/d and an EBCT of 9.3 h (Phase I). After several hundred days of stable operation, the organic mass loading rate and the hydraulic loading rate to the reactor were increased in 10% increments every 4 days until the loading rates were doubled (Phase II). Once reactor performance stabilized, the loading rates were again doubled until the desired EBCT of 2.32 h was achieved (Phase III). At day 640, a buffer-feed pump failed and the reactor experienced an operational disturbance due to a sudden
Table 1. Operating parameters for GAC fluidized-bed reactor system (Phases I–VII) Organic loading (g/d)
PCP Naphthalene Acenaphthene Pyrene Benzo(b)fluoranthene Ethanol
Influent conc. (mg/L)
I/V
II/VI
III/VII
IV
Phases I–VII (excl. IV)
0.60 0.21 0.066 0.036 0.003 4.28
1.20 0.42 0.132 0.072 0.006 8.33
2.40 0.84 0.264 0.144 0.012 16.66
0 0 0 0 0 0
100 35 11 6 0.5 690
Anaerobic treatment of soil wash fluids Table 2. Operating parameters for GAC fluidized-bed reactor system Phase I II III IV V VI VII
Days of operation 0–434 435–471 472–639 640–704 705–784 785–949 950–1135
Total flow rate (l/d) 6.0 12.0 24.0 batch/flush 6.0 12.0 24.0
EBCT h 9.30 4.65 2.32 0/2.32 9.30 4.65 2.32
elevation in pH. For a period of 65 days following the pH shock, the reactor was operated as a batch system (Phase IV). This period was devoted to reestablishing an effective biofilm within the reactor. Initially, feed to the reactor was stopped and the reactor was allowed to operate as a batch system in order to reduce the high aqueous phase chemical oxygen demand (COD) already present. The existing biomass, however, was unable to treat this COD, so the reactor was flushed with effluent from a reactor used to initially seed this reactor. This flushing process was intended to dilute the high COD and reintroduce active biomass into the reactor. Once the COD was reduced to levels experienced during normal operating conditions, feed to the reactor was resumed at the initial loading conditions under Phase I. This marks the beginning of Phase V and corresponds to an EBCT of 9.2 h. The loading rates were increased in two stages as before until the desired EBCT of 2.32 h was achieved. Thus, the reactor in this study underwent seven phases of operation, each corresponding to a different EBCT as summarized in Table 2. Analytical methods The reactor was monitored daily for effluent pH, total gas volume production, buffer and nutrient feed rates, and syringe feed rate. Weekly effluent samples were analyzed for COD, gas composition, chlorides, volatile fatty acids (VFAs), alcohols, and concentrations of PCP, PCP dechlorination intermediates, and the four-feed PAHs. The concentrations of PCP and its dechlorination intermediates were analyzed using a gas chromatograph (GC) (HewlettPackard, Palo Alto, CA, Model 5890 Series II) with an electron capture detector (ECD) (Hewlett-Packard, Palo Alto, CA). The concentrations of the four PAHs were measured using the same GC euqipped with a flame ionization detector (FID) (Hewlett-Packard, Palo Alto, CA). Duplicate samples of GAC and the associated biomass, were periodically withdrawn from the reactor and extracted to quantitatively determine masses of adsorbed compounds (method outlined by Fox, 1989). A detailed listing of the analytical methods used in this study is given by Miller et al. (1998). Isotherm setup To determine the adsorptive capacity of GAC for naphthalene and acenaphthene, GAC adsorption isotherm experiments were conducted using the bottle point method. The experiments were performed under anoxic conditions at 358C and in the absence of biological activity according to the methods outlined in Miller et al. (1998). The activated carbon used in this study was Calgon’s Filtrasorb-400, US mesh 16 20, the same carbon used to initially charge the fluidized-bed reactor. The isotherm studies were run at two initial solution concentrations: 15 and 25 mg/L for naphthalene and 1.5 and 2.5 mg/L for acenaphthene. Bottles were tumbled on a rotary shaker for 14 days to establish equilibrium. An equilibrium time of 14 days was determined
2365
to be sufficient by Vidic et al. (1990). Once equilibrium was established, the concentration of adsorbate on the carbon, qe , was calculated from the measured aqueous concentration and plotted against the aqueous concentration, Ce , to obtain an adsorption isotherm curve. The calculated value of qe was obtained from the equation: V ð1Þ m where C0 is the initial liquid phase concentration of solute, V the solution volume, and m the mass of adsorbate. Carbon extractions were performed to determine extraction efficiencies relative to the calculated values. For both naphthalene and acenaphthene, Soxhlet extraction yielded extraction efficiencies of greater than 95%. The Freundlich isotherm equation provided the best fit of the isotherm data (Freundlich, 1906): qe ¼ ðC0 Ce Þ
qe ¼ KCe1=n
ð2Þ
The Freundlich parameters K and 1=n were obtained by nonlinear least-squares regression analysis. From the isotherm data, predicted breakthrough curves for naphthalene and acenaphthene were generated. Equilibrium between the aqueous- and solid-phase concentrations was assumed. Furthermore, it was also assumed that only adsorption took place and no adsorbate–adsorbate interaction occurred. Since the reactor was operated at a recycle ratio of 200 : 1, CSTR hydraulics were assumed. Therefore, a material balance around the expanded-bed bioreactor yields the following first-order differential equation: dc dq þ Mc ð3Þ dt dt where Q is the flow through the reactor; C0 is the influent solute concentration; Ce is the effluent solute concentration; Vt the total volume of reactor; Vc the volume of carbon; Mc the mass of carbon; and q the mass loading on carbon. The above equation was then solved using the fourth-order Runge–Kutta method to obtain the predicted reactor effluent concentration with time for a single-solute system. This predicted breakthrough curve was compared to the observed breakthrough curve for both naphthalene and acenaphthene to analyze the fate of these compounds within the system. QC0 QCe ¼ ðVt Vc Þ
RESULTS AND DISCUSSION
Treatment of chlorophenols by anaerobic GAC bioreactor The molar effluent concentrations of PCP and its chlorinated phenolic intermediates throughout each phase of operation are presented in Fig. 1. The molar influent concentration of PCP was 0.376 mmol/L throughout the study with the exception of Phase IV, during which time feed to the reactor was stopped. The concentration of each of the dechlorination intermediates represents the sum of all measured isomers. The first two phases of operation were characterized by stable operation, with low COD, acid and alcohol concentrations (averaging 59.8, 1.44, and 1.37 mg/L, respectively). Most of the organics were being removed through adsorption onto the GAC, including the four-feed PAHs, which remained at concentrations less than 0.1 mg/L. During this period, the reactor experienced a steady increase in concentration of MCP in the effluent (Fig. 1). The
2366
K. M. Koran et al.
Fig. 1. Effluent concentrations for PCP and dechlorination intermediates as a function of reactor operating time.
Fig. 2. Effluent concentrations for phenol and select MCP and DCP species.
concentrations of all other dechlorination intermediates remained below 3 103 mmol/L, while PCP, tetrachlorophenol (TeCP) and trichlorophenol (TCP) concentrations remained below 1 104 mmol/L. This corresponds to a greater than 99.9% removal of PCP from the influent waste stream. Figure 2 displays the effluent concentrations for phenol and selected MCP and dichlorophenol (DCP) species. Para-chlorophenol (4-CP) and meta-chlorophenol (3-CP) were the primary MCP species present, with 4-CP dominating throughout the first two phases. 3,4-DCP was the primary DCP species present during this period, followed by 3,5-, 2,5- and trace amounts of 2,4-DCP. By the end of Phase I, 3-CP was no longer detected in the effluent. However,
the concentration of 4-CP continued to increase until the middle of Phase III. On day 550, a spike was observed in the effluent COD (810 mg/L) and acetate (187 mg/L) concentrations, indicating stress to the system had occurred. Also at this time, a shift was noted in the predominant MCP species present from 4-CP to 3-CP, the more readily anaerobically biodegradable of the two isomers. 3-CP continued to rise in the effluent as 4-CP declined, and by day 680 the reactor had achieved a near equimolar conversion of PCP to MCP. Following the appearance of 3-CP, a corresponding increase in phenol was noted and that increase continued through the end of phase III.
Anaerobic treatment of soil wash fluids
Following the period of upset in Phase IV, the concentrations of 3-CP, 4-CP, and 3,4-DCP steadily declined while all other chlorinated intermediates remained below 104 mmol/L. By day 880, the molar sum of all dechlorination intermediates accounted for less than 10% of the molar influent concentration of PCP. Following day 880, a shift from 3-CP to 4-CP as the predominant intermediate was observed. PCP removal efficiencies averaged 99.8% for the last three phases of operation, with conversion to measured intermediates averaging 46.5%. MCP and phenol accounted for 94.5% of the intermediates measured in the effluent during this period. Carbon extractions were performed monthly to determine if adsorption could explain the difference between influent PCP and effluent phenolic concentrations. The mass of MCP and intermediates adsorbed to the carbon, however, decreased with reactor operating time and were fairly constant during the last phases of operation (Miller et al., 1998). Thus, the difference between the measured intermediates and the influent PCP is likely due to the complete mineralization of PCP. Throughout this phase, the effluent COD, VFA and alcohol concentrations remained stable. Treatment of PAHs by anaerobic GAC bioreactor Figure 3 shows the effluent concentrations of naphthalene and acenaphthene throughout the course of reactor operation. During the first 400 days of operation, concentrations of all PAHs remained extremely low in the effluent due to the strong affinity of GAC for PAHs. Naphthalene, which was fed at a concentration of 35 mg/L, was present in the effluent at concentrations below 0.2 mg/L. Throughout the entire study, levels of pyrene and benzo(b)fluoranthene remained nearly
2367
constant at concentrations below 0.01 mg/l. Since the effluent concentrations of pyrene and benzo(b)fluoranthene were so low and there was no evidence to suggest that these compounds were undergoing any chemical transformation within the reactor, these compounds will be omitted from the remainder of this discussion. By the end of Phase I, naphthalene was present in the effluent at approximately 0.4 mg/L and these concentrations continued to increase until the end of phase III when the reactor experienced the pH shock. By this time, acenaphthene had similarly begun an increasing trend in the effluent. The concentrations of naphthalene and acenaphthene in the effluent at the time of the upset were 3.8 and 0.1 mg/L, respectively. When the nutrient feed pump failed, the aqueous pH of the system increased to approximately 9.2, causing mass desorption of compounds from the GAC. Naphthalene concentrations in the effluent at this elevated pH were extremely high, creating a toxic environment for the biological community. This resulted in a loss of active biomass and the subsequent failure of the system. During the recovery period following the pH shock, effluent concentrations returned to levels experienced under stable operating conditions prior to the upset. The effluent concentrations of naphthalene and acenaphthene continued to increase during Phases V and VI. By the end of Phase VI, effluent concentrations of naphthalene had leveled off reaching an apparent steady-state condition. Concentrations of naphthalene during this time fluctuated around a mean of approximately 5 mg/L. Similarly during Phase VII, acenaphthene levels appeared to stabilize around an average concentration of 0.8 mg/ L. These concentrations represent removal efficiencies of 86% for naphthalene and 93% for
Fig. 3. Effluent PAH concentrations as a function of reactor operating time.
2368
K. M. Koran et al.
acenaphthene. Carbon extraction data confirm that these compounds were not accumulating on the GAC during these phases (Miller et al., 1998).
Isotherm studies In order to determine whether naphthalene was being degraded in the system, a GAC single-solute adsorption isotherm study was performed (Fig. 4) and, from these data, a predicted breakthrough curve was generated for naphthalene within the GAC fluidized-bed system (Fig. 5). The Freundlich equation offered the best fit of the data with an r2 value of 0.96. The Freundlich regression parameters K and 1=n were 203.1 and 0.21, respectively. This isotherm equation defines the relationship between the aqueous solute concentration, Ce , and the solid phase loading, q. Using the isotherm regression constants, a predicted breakthrough curve was generated by integrating equation (3) using the fourth-order Runge–Kutta method. This breakthrough curve is the predicted aqueous phase concentration for naphthalene assuming no other compounds are competing for adsorption on the GAC and no biological activity is occurring within the system. Since competitive adsorption does occur within the reactor, the solid-phase concentration of any single solute is reduced, and breakthrough may occur earlier than predicted. For the first 520 days, observed effluent naphthalene concentrations followed the predicted breakthrough curve closely. After day 520, effluent concentrations remained much lower than the predicted breakthrough curve. This difference between the observed concentration and the predicted concentration suggests that naphthalene was being removed by some mechanism other than adsorption
alone, such as biological degradation or transformation. A similar isotherm study was performed for acenaphthene (Fig. 6), and a predicted breakthrough curve was generated (Fig. 7). The Freundlich regression parameters K and 1=n were 319.8 and 0.25, respectively, and the r2 value was 0.97. The observed effluent concentrations of acenaphthene followed the predicted breakthrough curve for the first 640 days. After day 640, effluent concentrations of acenaphthene were higher than those predicted by the breakthrough curve. Since the predicted breakthrough curve ignores the effect of competitive adsorption, it represents the latest estimate for breakthrough. Thus, this study suggests that carbon adsorption was a primary means of removal for acenaphthene. These data concur with PAH mass balance data that revealed the mass of acenaphthene removed from the system was nearly equivalent to the mass of that compound adsorbed on the GAC (Miller et al., 1998), as summarized in Table 3. Experimental data do not reveal whether pyrene or benzo(b)fluoranthene were undergoing any chemical transformation within the reactor. However, since after 1200 days of operation there was no evidence of these compounds in the effluent, their removal from the influent wastewater stream could be achieved through periodic carbon replacement. A carbon replacement schedule of 10% every 4 months would be sufficient to remove these contaminants from the influent waste stream. The spent GAC could be thermally regenerated and reintroduced to the reactor.
CONCLUSIONS
The objective of this work was to demonstrate that an anaerobic GAC fluidized-bed bioreactor could
Fig. 4. Anaerobic adsorption isotherm for naphthalene.
Anaerobic treatment of soil wash fluids
2369
Fig. 5. Breakthrough curves for naphthalene for the GAC fluidized-bed reactor system.
Fig. 6. Anaerobic adsorption isotherm for acenaphthene.
effectively treat a synthetic waste stream containing contaminants commonly found in wood-treating wastes. During periods of stable operation following the upset in Phase IV, the reactor achieved greater than 99.8% removal of PCP from the influent waste stream with conversion to measured intermediates averaging 46.5%. Carbon extractions confirm that these compounds were not accumulating on the GAC. Thus, the difference between measured intermediates and influent PCP is likely due to the complete mineralization of PCP. An aerobic fluidized-bed reactor is recommended to remove residual phenol intermediates in the anaerobic reactor effluent.
The reactor removed naphthalene and acenaphthene from the influent waste stream with efficiencies of 86 and 93%, respectively. Carbon extraction, isotherm, and predicted breakthrough data suggest that naphthalene was being removed by some mechanism other than adsorption alone, such as biological degradation. Carbon adsorption was the primary means of removal for acenaphthene. Experimental data do not reveal whether pyrene and benzo(b)fluoranthene were undergoing any chemical transformation within the reactor. However, their removal from the influent wastewater stream could be achieved through periodic carbon replacement.
2370
K. M. Koran et al.
Fig. 7. Breakthrough curves for acenaphthene for the GAC fluidized-bed reactor system. Table 3. Cumulative mass balance for naphthalene and acenaphthene Mass (g)
Naph
Acen
Cumulative influent Cumulative effluent Mass removed Mass adsorbed on GAC Difference (due to biodegradation)
409.3 22.2 387.1 148.0 239.1
128.8 1.6 127.2 114.6 12.6
REFERENCES
Baedecker M. J., Cozzarelli I. M., Eganhouse R. P., Siegel D. I. and Bennett P. C. (1993) Crude oil in a shallow sand and gravel aquifer: III. Biogeochemical reactions and mass balance modeling in anoxic groundwater. Appl. Geochem. 8, 569–586. Bedessem M. E., Swoboda-Colberg N. G. and Colberg P. J. S. (1997) Naphthalene mineralization coupled to sulfate reduction in aquifer-derived enrichments. FEMS Microbiol. Lett. 152(2), 213–218. Bregnard T., Hoehener P., Haener A. and Zeyer J. (1996) Degradation of weathered diesel fuel by microorganisms from a contaminated aquifer in aerobic and anaerobic microcosms. Environmental Toxicology and Chemistry 15(3), 299–307. Coates J. D., Anderson R. T. and Lovely D. R. (1996) Oxidation of polycyclic aromatic hydrocarbons under sulfate-reducing conditions. Appl. Environ. Microbiol. 62(3), 1099–1101. Edwards E. A. and Grbic-Galic D. (1994) Anaerobic degradation of toluene and o-xylene by a methanogenic consortium. Appl. Environ. Microbiol. 60(1), 313–322. Essaid H. I., Bekins B. A., Godsy M. E. and Ean W. (1995) Simulation of aerobic and anaerobic biodegradation processes at a crude oil spill site. Water Resources Research 31, 3309–3327. Fox P. (1989) An innovative reactor design for the treatment of biologically inhibitory wastewater. Ph.D. Dissertation, Department of Civil Engineering, University of Illinois, Urbana-Champaign, IL. Freundlich H. (1906) Ueber die Adsorption in Leosungen. Z. Phys. Chem. 57, 385–470. Genthner B. R. S., Townsend G. T., Lantz S. E. and Mueller J.G. (1997) Persistence of polycyclic aromatic hydrocarbon
components of creosote under anaerobic enrichment conditions. Arch Environ. Contam. Toxicol. 32(1), 99–105. Hambrick G. A., DeLaune R. D. and Patrick W. H. (1980) Effect of estuarine sediment pH and oxidation-reduction potential on microbial hydrocarbon degradation. Appl. Environ. Microbiol. 40(2), 365–369. Khodadoust A. P., Sorial G. A., Wilson G. J., Suidan M. T., Griffith R. A. and Brenner R. C. (1999) An integrated system for remediation of contaminated soils. J. Environ. Eng., ASCE 125(11), 1033–1041. Khodadoust A. P., Wagner J., Suidan M. T. and Brenner R. C. (1997) Anaerobic treatment of PCP in fluidized-bed GAC bioreactors. Water Res. 31(7), 1776–1786. Langenhoff A., Zehnder A. and Schraa G. (1996) Behaviour of toluene, benzene and naphthalene under anaerobic conditions in sediment columns. Biodegradation 7(3), 267–274. Madsen E., Man C. and Bilotta S. (1996) Oxygen limitations and aging as explanations for the field persistence of naphthalene in coal tar-contaminated surface sediment. Environ. Toxicol. Chem. 15(11), 1876–1882. Mihelcic J. R. and Luthy R. G. (1988) Degradation of polycyclic aromatic hydrocarbon compounds under various redox conditions in soil-water systems. Appl. Environ. Microbiol. 54(5), 1182–1187. Miller K. M., Suidan M. T., Sorial G. A., Khodadoust A. P., Acheson C. M. and Brenner R. C. (1998) Anaerobic treatment of soil wash fluids from a wood preserving site. Water Sci. Technol. 38(7), 63–72. Parker W. J. and Monteith H. D. (1995) Fate of polynuclear aromatic hydrocarbons during anaerobic digestion of municipal wastewater sludges. Water Environ. Res. 67(7), 1052–1059. Thierrin J., Davis G. B. and Barber C. (1995) A groundwater tracer test with deuterated compounds for monitoring in-situ biodegradation and retardation of aromatic hydrocarbons. Ground Water 33, 469–476. Vidic R. D., Suidan M. T., Traegner U. K. and Nakhla G. F. (1990) Adsorption isotherms: illusive capacity and role of oxygen. Water Research 24(10), 1187–1195. Wilson G., Khodadoust A., Suidan M., Acheson C., and Brenner R. (1995) Anaerobic/aerobic biodegradation of pentachlorophenol using GAC fluidized bed bioreactors: optimization of the empty bed contact time. Proceedings of the 1995 Water Environment Federation Conference, Water Environment Federation, Los Angeles.