Published online 27 October 2006
Published in J Environ Qual 35:2021-2025 (2006)
DOI: 10.2134/jeq2006.0080
© 2006 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
677 S. Segoe Rd., Madison, WI 53711 USA
TECHNICAL REPORTS
Bioremediation and Biodegradation
Persistence of Fermentative Process to Phenolic Toxicity in Groundwater
Youxian Wua,*,
David N. Lernera,
Steven A. Banwarta,
Steven F. Thorntona and
Roger W. Pickupb
a Groundwater Protection and Restoration Group, Department of Civil and Structural Engineering, University of Sheffield, Mappin St., Sheffield S1 3JD, UK
b Centre for Ecology and Hydrology, Lancaster Environment Centre, Library Avenue, Lancaster LA1 4AP. Y. Wu, current address, Department of Civil and Environmental Engineering, San Diego State University, San Diego, CA 92182, USA
* Corresponding author (ywu{at}mail.sdsu.edu)
Received for publication February 22, 2006.
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ABSTRACT
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The fermentation process is an important component in the biodegradation of organic compounds in natural and contaminated systems. Comparing with terminal electron-accepting processes (TEAPs), however, research on fermentation processes has to some extent been ignored in the past decades, particularly on the persistence of fermentation process in the presence of toxic organic pollutants. Both field and laboratory studies, presented here, showed that microbial processes in a groundwater-based system exhibited a differential inhibitory response to toxicity of phenolic compounds from coal tar distillation, thus resulting in the accumulation of volatile fatty acids (VFAs) and hydrogen. This indicated that fermentation processes could be more resistant to phenol toxicity than the subsequent TEAPs such as methanogenesis and sulfate reduction, thus providing us with more options for enhancing bioremediation processes.
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INTRODUCTION
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PHENOL AND ITS DERIVATIVES ARE WIDELY USED in the manufacture of a variety of chemicals, such as antioxidants, biocides, and disinfectants. They are also the major constituents of wastes produced by many industrial processes such as coal carbonization, petroleum, and coal tar distillation. Phenols often enter the environment as a result of uncontrolled discharges through spills from these industrial sources or long-term low level seepage (Williams et al., 2001). The high aqueous solubility and weak adsorption of phenols to most soils can result in the leaching and low retention of phenols in soils and a more rapid entry into groundwater (Rao and Asolekar, 2001; Williams et al., 2001).
At concentrations above some threshold levels, phenols have been considered to be toxic to microorganisms and refractory to biodegradation (Kirk-Othmer, 1978). Studies have shown that sufficiently high concentrations of phenols are toxic to anaerobic microorganisms, but many phenolic compounds can be degraded by microorganisms in engineered systems (Di Gioia et al., 2002) although the inhibitory threshold is different depending on the composition of impacted microbial communities (Thompson et al., 2005). Sulfate reduction and methanogenesis are suppressed by phenols in chemostats (Daneel et al., 1997), in upflow anaerobic sludge blankets (UASBs) (Fang and Chan, 1997), and in the subsurface (Nielsen et al., 1995b; Spence et al., 2001). Methanogenesis from the degradation of benzoate and acetate is depressed by phenol at 10 mM in UASB biogranules (Wang et al., 1991; Fang and Chan, 1997). Indigenous microorganisms in groundwater appear to be more susceptible to phenol toxicity than microorganisms in engineered systems (Williams et al., 2001; Thompson et al., 2005), but the relative extent of such phenol toxicity to fermentation and different terminal electron-accepting processes (TEAPs) is not clear. Furthermore, cresol or phenol mixtures are more toxic than phenol alone, particularly in groundwater (Dyreborg and Arvin, 1995; Nielsen et al., 1995a).
Fermentation is the initial indispensable but limiting process followed by the TEAPs for anaerobic biodegradation of most organics in groundwater. Stable associations of syntrophic fermentative microorganisms with a capacity to carry out TEAPs consume fermentation products and thus complete the process of organic biodegradation. Recent studies suggested that fermentation processes could play a more significant role in completing the biodegradation of some chlorinated organic compounds (Becker et al., 2005). Such fermentation activities of indigenous microorganisms, however, could be affected by the presence of toxic compounds. A better understanding of the effects of chemical toxicity on these indigenous processes could help improve bioremediation technology by making prediction of bioremediation or natural attenuation more accurate.
Previous studies have found ineffective natural attenuation of degradable organic compounds in a phenol-contaminated aquifer in the UK (Lerner et al., 2000; Williams et al., 2001), which was likely to be associated with the toxicity of phenols at high concentrations under anaerobic conditions (Pickup et al., 2001). The present work is assessing the impact of phenol toxicity to anaerobic processes of indigenous microorganisms with emphasis on fermentation and methanogenesis in the contaminated aquifer.
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MATERIALS AND METHODS
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Site Description and Sampling
This study was based on a phenolic contaminated site, which overlies the Permo-Triassic Sherwood Sandstone in central England (Williams et al., 2001). The pollution is linked to a chemical plant, which originally operated in the 1950s as a coal-tar distillation plant and now uses feedstocks sourced from other chemical plants (Williams et al., 2001). Groundwater in the vicinity of the coal-tar distillation plant has been contaminated with BTEX and a range of coal-tar compounds, including phenol, cresols, and xylenols. Bulk groundwater velocity is 4 to 11 m yr–1. The site and contamination have previously been described in detail (Lerner et al., 2000). Groundwater samples were taken from multilevel samplers (MLS) in a closed system using peristaltic pumps downstream of a wellhead sampling manifold (Lerner et al., 2000).
Microcosm Experiments
Microcosms were prepared in 1-liter glass bottles with 1-inch Teflon gas-tight caps and equipped with both liquid and gas sampling ports to permit periodic sampling. A series of microcosms labeled P0 (control), P1, P10, and P50 were established by diluting varying amounts of phenolic-contaminated groundwater containing phenol, m/p-cresol and o-cresol (90, 45, and 35 mM, respectively) with synthetic groundwater (Table 1), giving final concentrations of 0, 1, 10, and 50 mM total phenols, respectively. In addition, the microcosms also comprised approximately 600 g of sterile sediment taken from an uncontaminated borehole close to the plume and 700 mL of synthetic groundwater supplemented with 3% of yeast extract for enhancing microbial growth and activity. The microcosms were sterilized at 121°C (1.05 kg cm–2) for 3 periods of 20 min and then flushed with pure N2 gas through a sterile 0.2-µm filter followed by addition of 0.2 mM of Na2S. The microcosms were anaerobically incubated with a positive pressure at 20°C in the dark until strictly anaerobic conditions were achieved constantly for 2 wk. Microcosms were then inoculated with 100 mL of groundwater from the contaminated site, which had been incubated in laboratory in absence of any additional supplements and was active with respect to methanogenesis.
Chemical Analysis
Liquid samples (3 to 5 mL) were removed from the microcosms at approximately 1-wk intervals for chemical analysis after filtration through 0.20-µm membrane filters (Whatman Nylon). Average values from duplicate or triplicate sample analyses were reported. Phenolic compounds were analyzed by reverse-phase gradient-elution HPLC (high performance liquid chromatography). The apparatus consisted of two Isco Model 2350 pumps equipped with an Isco Chemresearch System controller, two Pye-Unicam 4020 Model UV variable wavelength detectors set at wavelengths of 280 nm and 266 nm, respectively, and a Jones-Genesis C18 chromatography column (250 mm x 4 um) and C18 guard column. The eluent was a 40:60 mixture of acetonitrile and 1% acetic acid at a flow rate of 1.0 mL min–1.
Ions, including sulfate and nitrate, were analyzed by Dionex 2000 IC (ion chromatograph). Anions were separated using an AS14 column (Dionex) with AG14 guard column (Dionex). Cations were analyzed on a CS12 column with a CG12 guard column (Dionex). The instrument had both anion and cation micro-membrane suppressers and a conductivity detector. The eluent for anion analysis was 3.5 mM Na2CO3 + 1.0 mM NaHCO3 at a flow rate of 1.2 mL min–1. The cation eluent was 18 mM methanesulphonic acid at a flow rate of 1 mL min–1. Some elements such as dissolved Fe and Mn were analyzed by ICP–OES (Inductively Coupled Plasma-Optical Emission Spectroscopy) with a Spectra Analytical Instruments Spectroflame M120E.
The concentrations of CH4 and CO2 were determined from gaseous subsamples taken from the headspace of microcosms using a Varian 3400 GC equipped with a methaniser and flame ionization detector (FID). CH4 and CO2 were separated at 140°C on 80/100 Carbonsphere in a stainless-steel column (1.83 m x 3.2 mm id) using N2 (40 mL min–1) as carrier. The temperatures of the injector and detector were 340 and 250°C, respectively. Gas phase hydrogen concentration was measured using a Trace Analytical RG3 Reduced gas analyzer. Hydrogen was analyzed at 165°C on 60/80 Spherocarb in a stainless-steel column (0.92 m x 3.2 mm id) using nitrogen (30 mL min–1) as carrier. The reduced gas detector was operated at a temperature of 250°C.
The concentrations of VFAs from acetic to heptoic acid and benzoate in water were measured by a Varian 3400 GC-FID. The GC was equipped with a 25 m x 0.22 mm ID BP21 (Phenomenex) polyethylene-glycol (TPA treated) capillary column and used N2 as the carrier gas at 10 mL min–1. The injector and detector temperatures were 200 and 250°C, respectively. The fluid sample was filtered through a 0.20-µm membrane filter and acidified to below pH 3 with concentrated phosphoric acid before analysis. The initial temperature of the column was 70°C for 2 min followed by a ramp of 7°C min–1 to 140°C for 2 min and a second ramp of 5°C min–1 to the final temperature of 180°C for 3 min. VFA standards (Supelo, Bellefonte, PA) and reagent grade sodium benzoate were used for calibration of the FID.
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RESULTS AND DISCUSSION
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Field investigation showed the changes in activities of TEAPs with concentrations of total phenols in the contaminated plume (Fig. 1). Oxygen, nitrate, and sulfate were present at background levels of 0.3, 1.6, and 0.65 mM while Fe(II) and Mn(II) were not found outside the phenol plume (Fig. 1A). Additional sulfate was supplied by an independent pollution source so that total sulfur concentrations were higher in the plume than in background groundwater. Oxygen and nitrate were not found inside the plume where phenols were present (Fig. 1A1), indicating that anaerobic processes were dominant there. Soluble Fe(II) and Mn(II) increased with increasing phenol concentrations where phenols were below 2 mM, and then remained stable when phenols were in the range from 2 to 45 mM (Fig. 1A2). This suggests that the reduction of iron/manganese oxides could be enhanced by phenols at the concentrations below 2 mM but remain evident at the high concentrations of phenols (2 to 45 mM). Accordingly, sulfate decreased when phenols were below 1 to 2 mM but remained at the background concentration of 0.65 mM or higher at higher phenol concentrations (Fig. 1A3). This indicated that sulfate-reducing processes could be active when phenols were below 1 to 2 mM and weak or stopped at higher concentrations of phenols.
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Fig. 1. Changes of (A) inorganic electron acceptors and (B) fermentation products (acetate, hydrogen, and methane) with different phenol concentrations.
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Acetate accumulation was also found to change with phenol concentrations (Fig. 1B1). Acetate increased with increasing phenol concentrations when phenols were below 10 mM but decreased with increasing phenol concentrations that were higher than 10 mM. Methane accumulation had a similar trend, increasing with phenol concentrations until phenols were at 10 mM and then decreasing with phenol concentrations (Fig. 1B3). It seemed that acetogenic and methanogenic processes remained active until phenols were at approximately 10 mM, indicating that this was the threshold for inhibition. Hydrogen increased with increasing phenol concentrations when phenols were below 10 mM but remained at a relatively low level (Fig. 1B2), suggesting that hydrogen-consuming processes were inactive at the high concentrations of phenols, although hydrogen production appeared to remain active.
Laboratory microcosm experiments were conducted to observe the microbial activities under different phenol concentration regimes and strict anaerobic conditions. The results showed effects of different phenol concentrations on the accumulation of acetate, hydrogen, methane, and sulfate as a function of incubation time (Fig. 2). The accumulation of methane occurred in the treatment with 0 (control) and 1 mM of phenol, while no pronounced accumulation of methane was found in the presence of 10 or 50 mM phenols. It indicated that methanogenic processes were active in the presence of phenols below 1 mM but completely inhibited by phenols higher than 10 mM. The range between 1 and 10mM was not assessed so it is possible that the inhibition threshold lies between these values. Sulfate-reducing processes showed a similar trend as a function of phenol concentrations.
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Fig. 2. Accumulation of (A) acetate, (B) hydrogen, (C) methane, and (D) sulfate in groundwater of microcosms containing different concentrations of total phenols (P0: no phenols; P1: 1 mM phenols; P10: 10 mM phenols; P50: 50 mM phenols).
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The accumulation of acetate and hydrogen had a similar trend for the four treatments with different phenol concentrations of 0, 1, 10, and 50 mM, respectively (Fig. 2). In the treatment with 0 and 1 mM of total phenols, acetate and hydrogen increased sharply and then decreased rapidly to a low level after 100 d of incubation; whereas with 10 mM of phenols, acetate increased gradually with time and remained at a high level after 100 d but hydrogen rose sharply up to a high level. Little acetate and relatively low concentrations of hydrogen were produced in the treatment with 50 mM of total phenols. This suggested that the accumulation of acetate and hydrogen was inhibited by phenols higher than 10 mM.
It is known that acetate and hydrogen are the main products of fermentation processes in complex biodegradation systems and they are also the reactants/reductants for subsequent anaerobic respiration and methanogenic processes (Schink, 1997; Karlsson et al., 2000). The accumulation of hydrogen and acetate has been used as the indicator of relative activities of fermentative microorganisms against those of anaerobic respiratory microorganisms for TEAPs (Lovley and Chapelle, 1996; Jakobsen et al., 1998). Fermentation processes have been found to dominate TEAPs in aquitards where the supply of electron acceptors is limited, thus causing accumulation of high concentrations of acetate and hydrogen (McMahon and Chapelle, 1991; Routh et al., 2001). In our field investigation, therefore, fermentation processes were observed to be relatively more active than anaerobic TEAPs with increasing phenol concentrations up to the threshold of 10 mM. This is probably because anaerobic TEAPs are more susceptible to phenol toxicity at high concentrations and exhibit more inhibition than fermentation. Thus, the difference in subsequent activity between fermentation and the depressed or inhibited anaerobic TEAPs leads to the accumulation of hydrogen and acetate. This conclusion is supported by the laboratory study although the effects of the subsurface and hydrogeologic complexity on microbial activities (Barbaro et al., 1994; Bekins et al., 1999) made the comparison of field investigation and laboratory study difficult.
The greater resilience of fermentation to chemical toxicity may be of great significance for biodegradation of organic contaminants by indigenous microbial communities in groundwater, although the mechanism of resilience is unknown. However, the resilience may be due to the largely unknown diversity of microorganisms that carry out fermentative processes being large with a subset being particularly robust when confronted by normally inhibitory concentrations of toxic hydrocarbons (Head et al., 1998). Fermentation is the first step for anaerobic biodegradation of many organic pollutants and their products, hydrogen and low molecular weight organic acids, are sequentially utilized in the TEAPs to produce CO2 and/or CH4. When TEAPs are suppressed by the toxicity of pollutants such as phenols, a shift of the relative rates between fermentation and subsequent TEAPs leads to acetate and hydrogen accumulation. The accumulation of fermentation products can improve the tolerance of methanogenic cultures to higher concentrations of phenolic compounds (Kennes et al., 1997), and when phenols are subsequently diluted, the surviving microorganisms recover their activity. Therefore, even though fermentation is slow in heavily contaminated groundwater, it does ensure that biodegradation continues and that the potential for natural attenuation remains.
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ACKNOWLEDGMENTS
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The work was sponsored by UK Environment Agency and EPSRC. The authors wish to thank Professor Bob Watkinson for his helpful discussions during the study.
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ABSTRACT
INTRODUCTION
