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Denitrification and phenol degradation in a contaminated aquifer
Journal of Contaminant Hydrology 53 (2001) 305 – 318 www.elsevier.com/locate/jconhyd
Denitrification and phenol degradation in a contaminated aquifer M.J. Spence a,*, S.H. Bottrell a, J.J.W. Higgo b, I. Harrison c, A.E. Fallick d b
a School of Earth Sciences, University of Leeds, Leeds LS2 9JT, UK Fluid Processes Group, British Geological Survey, Keyworth, Nottingham NG12 5GG, UK c Analytical and Regional Geochemistry Group, British Geological Survey, Keyworth, Nottingham NG12 5GG, UK d Scottish University Reactor Research Centre, East Kilbride, Glasgow G75 0QF, UK
Received 1 March 2000; received in revised form 11 October 2000; accepted 14 December 2000
Abstract A natural groundwater system modified by pollutant phenols and agricultural nitrate has been modelled in the laboratory by a series of sacrificial microcosm experiments. Samples of aquifer sediment and groundwater from the margin of the phenol plume were used to inoculate anaerobic microcosms enriched in nitrate and pollutant phenols. Rapid degradation of phenol and p-cresol was observed over a 35-day period leading to the generation of inorganic carbon and a number of transient intermediates. O-cresol proved to be recalcitrant on the experimental time-scale. A mass balance calculation shows that, during degradation, carbon was conserved in the aqueous phase. Groundwater – sediment interactions were monitored using carbon stable isotope data. A mass balance for solution TIC indicates that p-cresol degradation stimulated the dissolution of sedimentary carbonate phases due to the formation of carbonic acid. Compound-specific carbon isotope analysis (GC-IRMS) was used to search for 13C enrichment in residual p-cresol. A slight enrichment trend (e = 2.5x) was tentatively identified. The potential of this fractionation effect for obtaining in situ degradation rates is discussed. Results from the microcosm experiments help to explain the observed distribution of nitrate and phenols within the polluted aquifer. D 2001 Elsevier Science B.V. All rights reserved. Keywords: Bacterial nitrate reduction; Phenols; Stable isotopes; Bioremediation
*
Corresponding author.
0169-7722/01/$ - see front matter D 2001 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 9 - 7 7 2 2 ( 0 1 ) 0 0 1 7 1 - 1
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1. Introduction Toxic and highly soluble phenolic compounds have been introduced into many aquifers around the UK as a result of past industrial activity and careless waste disposal practices. At many sites, phenol plumes pose a significant threat to future water supplies, and hence, it is important to understand the fate of these compounds in the subsurface environment. The capacity of the subsurface environment for rapid organic carbon oxidation is generally limited by the availability of reactive chemical oxidants (electron acceptors). Oxygen and nitrate are the first electron acceptors to be consumed, followed by less energy-efficient species including manganese/ferric iron oxides and sulphate. The exhaustion of available oxygen and nitrate is generally associated with a significant fall in degradation rate: engineered remediation strategies overcome this problem by bringing the pollution to the surface where abundant oxygen is available. Alternatively, electron acceptors can be injected directly into the aquifer. In some cases, however, groundwater affected by phenol pollution already has an unnaturally high nitrate concentration due to agrochemical pollution, raising the possibility of beneficial pollutant – pollutant interactions. At Four Ashes in the West Midlands, a large plume of phenols is surrounded by nitrate-enriched groundwater in the Triassic Sherwood Sandstone aquifer. The phenol pollution extends in the direction of regional groundwater flow for a distance of 500 m and reflects 50 years of contaminant migration at an average velocity of 10 m/year. A large number of conventional boreholes have been installed to delineate the plume and sample the pollutant source area (Williams et al., 2001, this issue). Additional highresolution groundwater samples have been obtained from two multilevel sampling boreholes (MLS boreholes 59 and 60) that penetrate the tail and leading edge of the plume. These are 130 and 350 m from the source term and are 30- and 44-m deep, respectively. Total coal tar concentrations in excess of 6 g/l render the plume centre highly toxic and the dispersion of contaminants across the plume margin appears to be minimal. The groundwater surrounding the plume contains between 80 and 100 mg/l nitrate, a concentration well in excess of the recommended UK limit for drinking water quality (50 mg/l). The major organic pollutants, in the order of decreasing concentration, are phenol, o/m/p cresols and xylenols (Thornton et al., 2001, this issue). The inorganic carbon present in association with the pollution today could be derived from degradation of a mere 7% of the contaminant mass. Given that the plume is approximately 50 years old, it therefore appears that degradation processes are extremely slow (Williams et al., 2001, this issue). Much of this inorganic carbon may, in fact, be the product of aerobic processes operating early during plume development, implying that anaerobic degradation may be negligible. The high phenol concentration inhibits biodegradation via sulphate reduction (Spence et al., 2001, this issue; Kuhlmann and Schottler, 1996) and more reactive electron acceptors, such as oxygen and nitrate, are absent from the plume core. Outside the plume, organic carbon concentrations are low (total organic carbon (TOC) = 5 mg/l) and electron acceptors are relatively abundant. Two high-resolution
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porewater profiles through the strata above the plume show the presence of sulphate ( 30 mg/l), dissolved oxygen (DO) ( 8 mg/l), and in particular, nitrate ( 100 mg/ l). The high nitrate concentration persists to the plume margin, falling to zero (along with DO) over a depth interval of only 2 m. In the middle of this interval, dispersive processes mix organic pollutants with dissolved nitrate and provide an environment potentially favourable for the growth of denitrifying bacteria (Flyvbjerg et al., 1993). Denitrifying bacteria are ubiquitous to most near-surface environments. They characteristically obtain energy from the stepwise reduction of the nitrate ion but can also utilise atmospheric oxygen if available (Kornaros and Lyberatos, 1998). Most subsurface environments contain a diverse, though not necessarily active, bacterial population existing under nutrient limited conditions (Bengtsson and Bergwall, 1995). When organic contaminants are introduced into a nitrate-rich groundwater, the activity of in situ nitrate reducers may therefore be stimulated to remediate the pollution hazard. In this paper, we present the results of a microcosm investigation into denitrification in the mixing zone between a phenol plume and groundwater rich in agriculturally derived nitrate. The degradation of contaminants in this zone is of special relevance to the outcome of a planned pump and treat remediation scheme that will remove the most concentrated part of the plume and result in increased mixing of these groundwater types. The study therefore provides an insight into how in situ bacterial denitrification could help to reduce the residual phenol loading.
2. Experimental background The processes of denitrification and pollutant oxidation were characterised by microcosm experiments that simulated the nitrate-enriched margin of the phenol plume. Groundwater containing trace phenols was collected from the lower plume margin (borehole 12) at a depth of 60 –70 m. Aquifer sediment (also for use in the microcosms) was equilibrated in this borehole for 6 months prior to use by means of a nylon mesh sack. Both the groundwater and the sediment sock were collected anaerobically from borehole 12 and transported back to the laboratory in sealed glass bottles prior to use. The use of both groundwater and sediment from the plume margin ensured that conditions in the microcosms were a good approximation to those in the aquifer. Table 1 shows XRF analyses of six rock core samples recovered from the aquifer adjacent to borehole 59. The sediment is from the same formation as that used in the microcosm experiments and varies in composition from a medium grained, pale yellow/green carbonate-cemented sandstone (A) to a dark red, clay-rich siltstone (F). The sediment used in the microcosms was derived from a poorly cemented dark red siltstone that contained 0.77 wt.% CaCO3 (determined by CO2 yield on a cryogenic gas purification line). It was therefore similar in composition to aquifer core samples D, E and F. During the experiment, the nitrate-enriched environment of the plume margin was recreated by spiking the bottle microcosms with potassium nitrate solution. Additional pollutant also had to be added to the microcosms (the trace pollutants in the marginal groundwater were at the detection limit of the analysis technique). The pollutant chosen for the trial
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Table 1 XRF and carbonate analyses of six aquifer core samples recovered from the same rock formation as the sediment used in the microcosm experiment Component
XRF analyses (wt.%) A
B
C
D
E
F
SiO2 CaO Al2O3 K2O Fe2O3 MgO Na2O TiO2 P2O5 MnO
84.62 6.51 4.70 2.59 0.57 0.33 0.28 0.14 0.08 0.04
88.47 3.10 5.30 2.87 0.43 0.36 0.36 0.17 0.08 0.03
87.43 1.88 6.44 3.28 0.83 0.41 0.35 0.26 0.11 0.02
89.98 1.01 6.37 3.27 0.78 0.41 0.45 0.24 0.10 0.02
89.31 0.60 7.10 3.56 0.93 0.48 0.44 0.28 0.13 0.02
88.29 0.70 7.05 3.55 1.25 0.46 0.41 0.28 0.12 0.02
Total
99.86
101.17
101.01
102.63
102.85
102.13
5.39
2.93
1.96
0.91
0.22
0.39
Wt.% CaCO3 (by CO2 yield)
was p-cresol. This compound is known to degrade under nitrate reducing conditions with a relatively short lag time (Bossert and Young, 1986) and is a major constituent of the plume phenolics ( 10 wt.%). The experiment was not intended to last for a period much in excess of 40 days, and so, recalcitrant phenolics such as o-cresol and 2,6xylenol were not added.
3. Microcosm design and construction Five sets of sacrificial bottle microcosms were produced with 10 bottles per set, two sets of which (D and E) were sterile controls containing 2000 mg/l sodium azide. The other three sets (A, B, C) were active nonsterile microcosms, one of which (set A) was used for subsequent total inorganic carbon (TIC) 13C/12C isotope analysis. Sets A, B, D, E were spiked with 1000 mg/l nitrate on day 12, whereas set C was spiked with 40 mg/l of nitrate on day 0 and 1000 mg/l on day 12. The initial spiking of set C with 40 mg/l nitrate demonstrated that degradation would occur at low nitrate concentration. Once this had been proved, it was spiked with additional nitrate to bring about more rapid degradation. The microcosms contained about 20% by volume equilibrated aquifer sediment. During incubation at 15 C, microcosm bottles were stored in gas-tight, nitrogen-filled steel vessels mounted on a rolling table. This prevented air incursion in the event of imperfect sealing, and the rolling action served to keep the microcosms homogenised. As an additional measure, bottle lids were sealed by dipping them in molten paraffin wax. Gas samples were withdrawn by syringe and analysed immediately. Liquid samples were stored at 3 C prior to analysis to slow the degradation process.
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4. Analysis methods Dissolved organic compounds were initially analysed by gas chromatography (GC) after extraction into di-isopropyl ether. It was subsequently found that in the presence of dissolved nitrite, an acidification step in the extraction procedure led to p-cresol degradation through the formation of nitrophenols. As a result, samples were subsequently re-analysed by direct injection into an HPLC system. HPLC analyses were performed using a C18 column with UV spectrophotometric detection at 254 nm. Compounds were identified on the basis of their retention times relative to known standards. Dissolved bicarbonate was determined by titration to pH 4 with 0.001 M HCl (all microcosm samples had a pH close to 7.00 before titration so that 71.5% of the TIC was present as titratable HCO 3 ). A correction factor of 1.4 was therefore used to convert this to TIC. Nitrate, nitrite and sulphate were determined by ion chromatography on samples that had been held in storage at 3 C to prevent further reaction (although it is likely that this did not halt degradation completely, see later discussion). Prior to ion chromatography, potentially column-damaging organics were removed with C18 solid phase extraction cartridges. Initially, the microcosm headspace was flushed with pure nitrogen. Before sacrificial sampling of each microcosm, changes in the headspace gas composition were monitored by packed column gas chromatography. Gas samples were withdrawn through PTFEcoated silicone septa. Oxygen and carbon dioxide concentrations were monitored to check for signs of air incursion and phenol degradation, respectively. During degradation, the carbon isotope ratio of TIC present in microcosm set A was expected to change due to the addition of carbon derived from p-cresol breakdown. Samples of TIC for d13C analysis were fixed as strontium carbonate by the addition of ammoniacal strontium chloride solution. The insoluble precipitate was then collected and dried prior to conversion to CO2 with 100% phosphoric acid in a vacuum line (McCrea, 1950). Purified gas samples were analysed using a VG Sira-10 gas source mass spectrometer and the data corrected using standard procedures (Craig, 1957). The precision, obtained from repeat analyses of a carbonate standard, was F 0.1x. Samples of microcosm sediment were analysed to determine the carbon isotope composition of any carbonate phases present. Sediment samples were tested both before and after exposure to the microcosm environment. Sediment carbonates were converted directly to CO2 by the addition of 100% phosphoric acid. Samples of CO2 were then purified and the 13C/12C ratio determined as described above. p-Cresol 13C/12C ratios were determined by a VG isochrom GC-IRMS. Unacidified microcosm samples containing p-cresol were extracted into a small volume of di-isopropyl ether (IPE) containing 200 mg/l bromobenzene (internal standard); 0.1– 3 ml of IPE was then injected into a BP 20 GC column with a split outlet. One-half of the split went to a flame ionization detector (FID) and the other into a copper oxide furnace at 800 C. Here, organic compounds were oxidised to CO2 and water. The stream of CO2 was purified by freezing out water at 100 C prior to continuous determination of the 13C/12C ratio. The precision, obtained from repeat analyses of an internal standard (bromobenzene), was F 0.2x.
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5. Results and discussion Fig. 1 shows the concentration of p-cresol in active microcosm sets A and B vs. a sterile control over the 35-day period as measured by GC and HPLC. Analyses obtained by GC were more precise until the addition of nitrate on day 12 gave problems with the extraction procedure. Hence, GC data are presented from day 1 to day 11 and HPLC data from day 15 to day 35. Prior to nitrate addition on day 12, C/Co (Measured/Initial concentration) values for pcresol in active microcosm sets A and B and the control sets D and E were close to 1.00, showing that no degradation had taken place. Degradation proceeded rapidly in the nitrate-bearing set C with the 40 mg/l of nitrate initially present being entirely converted to nitrite between days 4 and 11. This indicates a lag time of less than 3 days for nitrate reduction. After nitrate addition to sets A and B, a steady increase in alkalinity and headspace CO2 was observed that coincided with the appearance of three new organic compounds in solution (Figs. 2A and B). HPLC analysis tentatively identified these as 4-hydroxybenzyl alcohol (4HBalc), 4-hydroxybenzaldehyde (4HBald) and 4-hydroxybenzoic acid
Fig. 1. Variation in C/Co ( p-cresol) over the 35-day period for p-cresol degradation in the active microcosm sets A and B vs. the azide sterilised controls (set E). The addition of 1000 mg/l of nitrate on day 12 stimulated rapid pcresol degradation in the active sets.
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(4HBacid). The identity of these compounds in selected microcosms was confirmed by GC-MS. Previous mechanistic studies (Bossert and Young, 1986) show that they form through progressive oxidation of the methyl group on the p-cresol molecule. After pcresol exhaustion in set A, concentrations of 4HBalc and 4HBald began to diminish as they were converted to 4HBacid. If the experiment had been continued, then the concentration of all the intermediates would be expected to fall to zero after removal of the precursor p-cresol. The groundwater used for the microcosms from borehole 12 contained low concentrations of phenol, o-cresol and m + p cresol (5, 7, and 5 mg/l, respectively). The p-cresol concentration was increased by spiking with 200 mg/l, with the result that phenol and ocresol were insignificant in terms of the overall carbon budget. Phenol degradation coincided with the main phase of p-cresol breakdown whereas o-cresol was recalcitrant over the 35-day period. This recalcitrance is in accordance with previous studies conducted under comparable conditions (Bakker, 1977). Throughout the 35-day period, the sulphate concentration remained constant at 61 mg/l, indicating that sulphate had not been used as an electron acceptor. Similarly, microcosms sampled on day 19 after the onset of degradation contained no dissolved iron, indicating
Fig. 2. Concentration profiles of key species in microcosm sets A and B over the 35-day period. Analyses of dissolved phenolics and TIC are presented as mM C equivalents. Nitrate concentration is plotted in mM and the percentage of headspace CO2 is shown on the secondary axis. Nitrate reduction is associated with the appearance of p-cresol degradation products in solution, an increase in TIC and a rise in headspace CO2.
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that iron (III) oxide was not being used as an oxidant. An electron balance was calculated for each microcosm to determine whether denitrification could account for the observed pcresol removal and production of partially oxidised intermediates. Electron acceptor/donor reactions were used as shown in Table 2. The electron-accepting capacity of denitrification was expressed in terms of nitrate to nitrite conversion (deduced from the nitrite concentration). Note that nitrate remained at high concentration throughout the duration of the experiment, thereby inhibiting the secondary reduction of the nitrite ion (Kornaros et al., 1996). Most of the added nitrate was reduced to nitrite, leaving only 10% unaccounted
Table 2 Table of reactions used to calculate an electron balance for p-cresol degradation Electron-donating reactions Stepwise oxidation of methyl group: HOC6 H4 CH3 þ H2 O Z HOC6 H4 CH2 OH þ 2Hþ þ 2e
ðIÞ
HOC6 H4 CH2 OH þ H2 O ZHOC6 H4 COH þ 4Hþ þ 4e
ðIIÞ
HOC6 H4 COH þ H2 O ZHOC6 H4 COOH þ 2Hþ þ 2e
ðIIIÞ
Terminal oxidation step (mechanism unknown) HOC6 H4 COOH þ 11H2 O Z 7CO2 þ 28Hþ þ 28e
ðIVÞ
Summary equation HOC6 H4 CH3 þ 13H2 O Z 7CO2 þ 34e þ 34Hþ
ðVÞ
Electron-accepting reactions Nitrate reduction þ NO 3 þ 2e þ 2H Z NO2 þ H2 O
ðVIÞ
Denitrification þ NO 3 þ 5e þ 6H Z 1=2N2 þ 3H2 O
ðVIIÞ
Summary equation for p-cresol breakdown by nitrate reduction HOC6 H4 CH3 þ 17NO 3 Z17NO2 þ 4H2 O þ 7CO2
ðVIIIÞ
Summary equation for p-cresol breakdown by denitrification 102NO 3 þ 15HOC6 H4 CH3 Z102HCO3 þ 3CO2 þ 60H2 O þ 51N2
ðIXÞ
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for. This may have undergone complete reduction, or been converted to other partially reduced species such as N2O. For the electron balances presented here, we have assumed reduction to N2. Fig. 3 shows electron balances for sets A and B; these indicate that denitrification could account for 74% to 81% of the observed p-cresol degradation. The remaining 19– 26% is likely to be the result of aerobic sample degradation between sampling and HPLC analysis; chilling samples to 3 C slowed the rate of degradation rather than halting it. Carbon isotope ratios for microcosm TIC are plotted in Fig. 4, along with the initial groundwater and sediment carbonate ratios. The initial ratio in set A (d13C (TIC) = 17.3x) was consistent with a mixture of heavy carbonate (d13C = 1.1x) and organically derived carbon (d13C = 25x). Nitrate addition on day 15 promoted a shift to more negative values (d13C(TIC) = 19.6x by day 35) that coincided with the main phase of p-cresol degradation (d13C p-cresol = 27x). The change in d13C (TIC) was used to calculate the inorganic carbon contributions from p-cresol oxidation and carbonate dissolution. From days 25 to 35, the percentage of additional TIC derived from p-cresol fell from 98% to 82%, reflecting the dissolution of carbonate due to rising PCO2. This 18% contribution to the TIC from carbonate dissolution would remove 0.063 wt.% CaCO3 from the sediment (10% of that initially present ( 0.77 wt.%)). On a field scale, this reduction would not be noticeable from whole-rock analyses.
Fig. 3. Electron balances for active sets A and B over the 35-day period presented as measured nitrate loss vs. the predicted nitrate loss as calculated from the products of degradation. The gradient of the lines is 0.77, indicating that nitrate reduction could account for 77% of the organic matter oxidised.
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Fig. 4. A reduction in d13C(solution TIC) is coincident with the main phase of p-cresol degradation between days 15 and 35. This is consistent with the addition of 12C-enriched TIC derived from the oxidation of p-cresol (d13C p-cresol = 27x ).
A microcosm carbon balance was obtained by recalculating inorganic and organic carbon components as mM C and summing their concentrations to obtain total dissolved carbon. Carbon balances for active sets A and B are presented in Fig. 5a and b as stacked area plots showing the relative contributions of organic and inorganic carbon sources during the experiment. The carbon mass balance was corrected for the proportion of inorganically derived TIC found above (18% at end of experiment). Any carbon removed into the biomass would be reported as a deficit in the dissolved carbon budget. No deficit is apparent, and hence, the carbon available as p-cresol at the start of the experiment was primarily consumed in respiration processes. As p-cresol degradation proceeded in the nitrate reducing experiments, a small but potentially significant enrichment in 13C was seen in the residual p-cresol. Fig. 6 shows a best-fit curve through the data points assuming a Rayleigh fractionation model with an isotope enrichment factor (e) of 2.5x. This assumes a constant selectivity for 12C bearing p-cresol during the degradation process leading to a build-up of 13C in the residual compound. The curve was plotted from the following equation (Fritz and Fontes, 1980): dt ¼ do þ e Ln f ; where: do = initial d13C of p-cresol (x); dt = final d13C of p-cresol (x); e = enrichment factor ( 2.5x) f = C/Co = fraction of original p-cresol remaining.
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Fig. 5. Analyses of dissolved phenolics and bicarbonate presented as mM C equivalents and plotted in a stacked form over the duration of the experiment. The upper boundary of this plot represents the sum total of all the carbon compounds analysed. The dissolved nitrate concentration is also presented. Note that the number of moles of carbon in solution remains approximately constant during degradation.
Fig. 6. Experimental data are plotted that show a slight increase in d13C (residual p-cresol) with decreasing f ( = C/ Co). A best fit Rayleigh fractionation curve for this trend has been plotted that corresponds to an isotope enrichment factor (e) of 2.5x .
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Potentially, the change in residual p-cresol d13C during reaction can be used to monitor the extent to which degradation has proceeded in the field. However, the small enrichment factor (e = 2.5x) will only produce a significant change in d13C (residual p-cresol) when C/Co values become very low. In Fig. 6, this is illustrated by the extrapolation of the enrichment trend. For e = – 2.5x, kinetic fractionation will shift the isotopic composition of the parent p-cresol by > 4x only when C/Co < 0.2. This limits the use of this technique to situations where either the initial d13C variability in the compound is small or the extent of degradation is large. In the core of the plume, high C/Co values for the major pollutants are inferred from the fact that the present-day TIC could be derived from less than 7% of the present-day TOC (Thornton et al., 2001, this issue). At the plume margin, however, dispersion into surrounding groundwater reduces the initial pollutant concentration and the high nitrate concentration facilitates degradation. Hence, C/Co values may fall sufficiently to produce recognisable isotope enrichment trends. As a result, the plume margin is the only place where isotope fractionation is likely to be observed. In the present-day plume, the rate of dispersive mixing is slow with the result that denitrification is organic carbon limited just outside the plume and nitrate limited
Fig. 7. Depth section through the plume margin at MLS borehole 59 showing concentration variation in TOC, nitrate and nitrite. An analogue for the nutrient-enriched microcosm environment exists at a depth of 10 m where dissolved nitrate co-exists with pollutant phenols derived from coal tar.
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immediately inside it. This lack of electron acceptor –electron donor overlap is illustrated by high-resolution nitrate and TOC profiles for the plume margin at borehole 59 (Fig. 7, Thornton et al., 2001, this issue). The transition from high NO3 low TOC, to high TOC low NO 3 occurs over a depth interval of less than 2 m. Only one sample, taken from a depth of 10 m, straddles the transition and this contains 27 mg/l of nitrate, 1 mg/l of nitrite and 118 mg/l of TOC. This is the only depth where nitrate and TOC are present in sufficient concentration to produce appreciable degradation, and the presence of nitrite suggests that denitrification may follow a similar mechanism to that noted in the microcosms.
6. Conclusions A microcosm investigation into the potential for denitrification in the margin of the phenol plume has revealed that nitrate-reducing bacteria can rapidly degrade a relatively high concentration (200 mg/l) of p-cresol with the simultaneous degradation of trace phenol. After nitrate addition, p-cresol was degraded by stepwise oxidation of the methyl group prior to the formation of inorganic carbon. In the plume core, low dissolved nitrate concentration limits the activity of nitrate reducers, as demonstrated by the rapid onset of nitrate reduction in the microcosms after nitrate addition. At the plume margin, however, dispersive mixing of phenols and groundwater nitrate may stimulate the in situ microbial community to degrade at least part of the pollutant mixture. Recalcitrant pollutants such as o-cresol may also degrade, but most workers report lag times longer than the duration of this experiment. On a longer time-scale, degradation by alternative pathways, such as bacterial sulphate reduction and methanogenesis, may also be important for the reduction of contaminant load. These processes are harder to simulate in the laboratory, however, since they may have very long lag times. Experimental conditions can be modified to increase rates of reaction, but then, the microcosm may not be a good proxy for the real system. In a separate study (Spence et al., 2001, this issue), in situ rates of bacterial sulphate reduction have been estimated from stable sulphur isotope analyses of dissolved sulphate within the plume. A stable carbon isotope investigation of TIC composition has revealed that p-cresol degradation via the reduction of nitrate to nitrite was associated with the dissolution of carbonates in the aquifer sediment. This was attributed to the production of carbonic acid (Eq. (VIII) above). In a nitrate-limited system, however, the complete reduction of nitrate may occur according to Eq. (IX). In this case only 3% of the inorganic carbon produced would be in the form of carbonic acid and carbonate precipitation could therefore occur due to the production of bicarbonate ions. The carbon isotope fractionation during p-cresol degradation by denitrifying bacteria has been shown to be small (e = 2.5x). As such, it would be difficult to use this phenomenon to track degradation in the aquifer. The best location for applying this technique would be at the extreme outer margin of the plume where dispersion has increased the nitrate concentration relative to TOC. In summary, field data from the site have been shown to broadly agree with results presented from the microcosm study. Pollutant degradation by nitrate reduction will
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remain confined to a narrow interval surrounding the plume unless the plume is manipulated to artificially increase the degree of dispersive mixing. Following such remediation, a potential exists for significant in situ biodegradation through denitrification.
Acknowledgements This work was supported by NERC grant GT22/97/6/ENVD as part of the Environmental Diagnostics Initiative, and was made possible by the co-operation of the site owner and EnvirosAspinwall. We also thank R. Leader for assistance in field sampling and microcosm construction.
References Bakker, G., 1977. Anaerobic degradation of aromatic compounds in the presence of nitrate. FEMS Microbiology Letters 1, 103 – 108. Bengtsson, G., Bergwall, C., 1995. Heterotrophic denitrification potential as an adaptive response in groundwater bacteria. FEMS Microbiology Ecology 16 (4), 307 – 317. Bossert, I.D., Young, L.Y., 1986. Anaerobic oxidation of para-cresol by a denitrifying bacterium. Applied and Environmental Microbiology 52 (5), 1117 – 1122. Craig, H., 1957. Isotopic standards for carbon and oxygen and correction factors for mass spectrometric analysis of carbon dioxide. Geochimica et Cosmochimica Acta 12, 133 – 149. Flyvbjerg, J., Arvin, E., Jensen, B.K., Olsen, S.K., 1993. Microbial degradation of phenols and aromatic-hydrocarbons in creosote-contaminated groundwater under nitrate-reducing conditions. Journal of Contaminant Hydrology 12 (1 – 2), 133 – 150. Fritz, P., Fontes, J.C., 1980. Introduction. In: Fritz, P., Fontes, J.Ch. (Eds.), Handbook of Environmental Isotope Geochemistry, vol. 1. Elsevier, Amsterdam, pp. 1 – 19. Kornaros, M., Lyberatos, G., 1998. Kinetic modelling of Pseudomonas denitrificans growth and denitrification under aerobic, anoxic and transient operating conditions. Water Research 32 (6), 1912 – 1922. Kornaros, M., Zafiri, C., Lyberatos, G., 1996. Kinetics of denitrification by Pseudomonas denitrificans under growth conditions limited by carbon and/or nitrate or nitrite. Water Environment Research 68 (5), 934 – 945. Kuhlmann, B., Schottler, U., 1996. Influence of different redox conditions on the biodegradation of the pesticide metabolites phenol and chlorophenols. International Journal of Environmental Analytical Chemistry 65 (1 – 4), 289 – 295. McCrea, J.M., 1950. On the isotope chemistry of carbonates and a palaeotemperature scale. Journal of Chemical Physics 18, 849 – 857. Spence, M.J., Bottrell, S.H., Thornton, S.F., Lerner, D.N., 2001. Isotopic modelling of the significance of bacterial sulphate reduction for phenol attenuation in a contaminated aquifer. Journal of Contaminant Hydrology 53, 285 – 304. Thornton, S.F., et al., 2001. Processes controlling the distribution and natural attenuation of phenolic compounds in a deep sandstone aquifer. Journal of Contaminant Hydrology 53, 233 – 267. Williams, G.M., et al., 2001. Biogeochemical characterisation of a coal tar distillate plume. Journal of Contaminant Hydrology 53, 175 – 198.
Denitrification and phenol degradation in a contaminated aquifer M.J. Spence a,*, S.H. Bottrell a, J.J.W. Higgo b, I. Harrison c, A.E. Fallick d b
a School of Earth Sciences, University of Leeds, Leeds LS2 9JT, UK Fluid Processes Group, British Geological Survey, Keyworth, Nottingham NG12 5GG, UK c Analytical and Regional Geochemistry Group, British Geological Survey, Keyworth, Nottingham NG12 5GG, UK d Scottish University Reactor Research Centre, East Kilbride, Glasgow G75 0QF, UK
Received 1 March 2000; received in revised form 11 October 2000; accepted 14 December 2000
Abstract A natural groundwater system modified by pollutant phenols and agricultural nitrate has been modelled in the laboratory by a series of sacrificial microcosm experiments. Samples of aquifer sediment and groundwater from the margin of the phenol plume were used to inoculate anaerobic microcosms enriched in nitrate and pollutant phenols. Rapid degradation of phenol and p-cresol was observed over a 35-day period leading to the generation of inorganic carbon and a number of transient intermediates. O-cresol proved to be recalcitrant on the experimental time-scale. A mass balance calculation shows that, during degradation, carbon was conserved in the aqueous phase. Groundwater – sediment interactions were monitored using carbon stable isotope data. A mass balance for solution TIC indicates that p-cresol degradation stimulated the dissolution of sedimentary carbonate phases due to the formation of carbonic acid. Compound-specific carbon isotope analysis (GC-IRMS) was used to search for 13C enrichment in residual p-cresol. A slight enrichment trend (e = 2.5x) was tentatively identified. The potential of this fractionation effect for obtaining in situ degradation rates is discussed. Results from the microcosm experiments help to explain the observed distribution of nitrate and phenols within the polluted aquifer. D 2001 Elsevier Science B.V. All rights reserved. Keywords: Bacterial nitrate reduction; Phenols; Stable isotopes; Bioremediation
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Corresponding author.
0169-7722/01/$ - see front matter D 2001 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 9 - 7 7 2 2 ( 0 1 ) 0 0 1 7 1 - 1
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1. Introduction Toxic and highly soluble phenolic compounds have been introduced into many aquifers around the UK as a result of past industrial activity and careless waste disposal practices. At many sites, phenol plumes pose a significant threat to future water supplies, and hence, it is important to understand the fate of these compounds in the subsurface environment. The capacity of the subsurface environment for rapid organic carbon oxidation is generally limited by the availability of reactive chemical oxidants (electron acceptors). Oxygen and nitrate are the first electron acceptors to be consumed, followed by less energy-efficient species including manganese/ferric iron oxides and sulphate. The exhaustion of available oxygen and nitrate is generally associated with a significant fall in degradation rate: engineered remediation strategies overcome this problem by bringing the pollution to the surface where abundant oxygen is available. Alternatively, electron acceptors can be injected directly into the aquifer. In some cases, however, groundwater affected by phenol pollution already has an unnaturally high nitrate concentration due to agrochemical pollution, raising the possibility of beneficial pollutant – pollutant interactions. At Four Ashes in the West Midlands, a large plume of phenols is surrounded by nitrate-enriched groundwater in the Triassic Sherwood Sandstone aquifer. The phenol pollution extends in the direction of regional groundwater flow for a distance of 500 m and reflects 50 years of contaminant migration at an average velocity of 10 m/year. A large number of conventional boreholes have been installed to delineate the plume and sample the pollutant source area (Williams et al., 2001, this issue). Additional highresolution groundwater samples have been obtained from two multilevel sampling boreholes (MLS boreholes 59 and 60) that penetrate the tail and leading edge of the plume. These are 130 and 350 m from the source term and are 30- and 44-m deep, respectively. Total coal tar concentrations in excess of 6 g/l render the plume centre highly toxic and the dispersion of contaminants across the plume margin appears to be minimal. The groundwater surrounding the plume contains between 80 and 100 mg/l nitrate, a concentration well in excess of the recommended UK limit for drinking water quality (50 mg/l). The major organic pollutants, in the order of decreasing concentration, are phenol, o/m/p cresols and xylenols (Thornton et al., 2001, this issue). The inorganic carbon present in association with the pollution today could be derived from degradation of a mere 7% of the contaminant mass. Given that the plume is approximately 50 years old, it therefore appears that degradation processes are extremely slow (Williams et al., 2001, this issue). Much of this inorganic carbon may, in fact, be the product of aerobic processes operating early during plume development, implying that anaerobic degradation may be negligible. The high phenol concentration inhibits biodegradation via sulphate reduction (Spence et al., 2001, this issue; Kuhlmann and Schottler, 1996) and more reactive electron acceptors, such as oxygen and nitrate, are absent from the plume core. Outside the plume, organic carbon concentrations are low (total organic carbon (TOC) = 5 mg/l) and electron acceptors are relatively abundant. Two high-resolution
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porewater profiles through the strata above the plume show the presence of sulphate ( 30 mg/l), dissolved oxygen (DO) ( 8 mg/l), and in particular, nitrate ( 100 mg/ l). The high nitrate concentration persists to the plume margin, falling to zero (along with DO) over a depth interval of only 2 m. In the middle of this interval, dispersive processes mix organic pollutants with dissolved nitrate and provide an environment potentially favourable for the growth of denitrifying bacteria (Flyvbjerg et al., 1993). Denitrifying bacteria are ubiquitous to most near-surface environments. They characteristically obtain energy from the stepwise reduction of the nitrate ion but can also utilise atmospheric oxygen if available (Kornaros and Lyberatos, 1998). Most subsurface environments contain a diverse, though not necessarily active, bacterial population existing under nutrient limited conditions (Bengtsson and Bergwall, 1995). When organic contaminants are introduced into a nitrate-rich groundwater, the activity of in situ nitrate reducers may therefore be stimulated to remediate the pollution hazard. In this paper, we present the results of a microcosm investigation into denitrification in the mixing zone between a phenol plume and groundwater rich in agriculturally derived nitrate. The degradation of contaminants in this zone is of special relevance to the outcome of a planned pump and treat remediation scheme that will remove the most concentrated part of the plume and result in increased mixing of these groundwater types. The study therefore provides an insight into how in situ bacterial denitrification could help to reduce the residual phenol loading.
2. Experimental background The processes of denitrification and pollutant oxidation were characterised by microcosm experiments that simulated the nitrate-enriched margin of the phenol plume. Groundwater containing trace phenols was collected from the lower plume margin (borehole 12) at a depth of 60 –70 m. Aquifer sediment (also for use in the microcosms) was equilibrated in this borehole for 6 months prior to use by means of a nylon mesh sack. Both the groundwater and the sediment sock were collected anaerobically from borehole 12 and transported back to the laboratory in sealed glass bottles prior to use. The use of both groundwater and sediment from the plume margin ensured that conditions in the microcosms were a good approximation to those in the aquifer. Table 1 shows XRF analyses of six rock core samples recovered from the aquifer adjacent to borehole 59. The sediment is from the same formation as that used in the microcosm experiments and varies in composition from a medium grained, pale yellow/green carbonate-cemented sandstone (A) to a dark red, clay-rich siltstone (F). The sediment used in the microcosms was derived from a poorly cemented dark red siltstone that contained 0.77 wt.% CaCO3 (determined by CO2 yield on a cryogenic gas purification line). It was therefore similar in composition to aquifer core samples D, E and F. During the experiment, the nitrate-enriched environment of the plume margin was recreated by spiking the bottle microcosms with potassium nitrate solution. Additional pollutant also had to be added to the microcosms (the trace pollutants in the marginal groundwater were at the detection limit of the analysis technique). The pollutant chosen for the trial
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Table 1 XRF and carbonate analyses of six aquifer core samples recovered from the same rock formation as the sediment used in the microcosm experiment Component
XRF analyses (wt.%) A
B
C
D
E
F
SiO2 CaO Al2O3 K2O Fe2O3 MgO Na2O TiO2 P2O5 MnO
84.62 6.51 4.70 2.59 0.57 0.33 0.28 0.14 0.08 0.04
88.47 3.10 5.30 2.87 0.43 0.36 0.36 0.17 0.08 0.03
87.43 1.88 6.44 3.28 0.83 0.41 0.35 0.26 0.11 0.02
89.98 1.01 6.37 3.27 0.78 0.41 0.45 0.24 0.10 0.02
89.31 0.60 7.10 3.56 0.93 0.48 0.44 0.28 0.13 0.02
88.29 0.70 7.05 3.55 1.25 0.46 0.41 0.28 0.12 0.02
Total
99.86
101.17
101.01
102.63
102.85
102.13
5.39
2.93
1.96
0.91
0.22
0.39
Wt.% CaCO3 (by CO2 yield)
was p-cresol. This compound is known to degrade under nitrate reducing conditions with a relatively short lag time (Bossert and Young, 1986) and is a major constituent of the plume phenolics ( 10 wt.%). The experiment was not intended to last for a period much in excess of 40 days, and so, recalcitrant phenolics such as o-cresol and 2,6xylenol were not added.
3. Microcosm design and construction Five sets of sacrificial bottle microcosms were produced with 10 bottles per set, two sets of which (D and E) were sterile controls containing 2000 mg/l sodium azide. The other three sets (A, B, C) were active nonsterile microcosms, one of which (set A) was used for subsequent total inorganic carbon (TIC) 13C/12C isotope analysis. Sets A, B, D, E were spiked with 1000 mg/l nitrate on day 12, whereas set C was spiked with 40 mg/l of nitrate on day 0 and 1000 mg/l on day 12. The initial spiking of set C with 40 mg/l nitrate demonstrated that degradation would occur at low nitrate concentration. Once this had been proved, it was spiked with additional nitrate to bring about more rapid degradation. The microcosms contained about 20% by volume equilibrated aquifer sediment. During incubation at 15 C, microcosm bottles were stored in gas-tight, nitrogen-filled steel vessels mounted on a rolling table. This prevented air incursion in the event of imperfect sealing, and the rolling action served to keep the microcosms homogenised. As an additional measure, bottle lids were sealed by dipping them in molten paraffin wax. Gas samples were withdrawn by syringe and analysed immediately. Liquid samples were stored at 3 C prior to analysis to slow the degradation process.
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4. Analysis methods Dissolved organic compounds were initially analysed by gas chromatography (GC) after extraction into di-isopropyl ether. It was subsequently found that in the presence of dissolved nitrite, an acidification step in the extraction procedure led to p-cresol degradation through the formation of nitrophenols. As a result, samples were subsequently re-analysed by direct injection into an HPLC system. HPLC analyses were performed using a C18 column with UV spectrophotometric detection at 254 nm. Compounds were identified on the basis of their retention times relative to known standards. Dissolved bicarbonate was determined by titration to pH 4 with 0.001 M HCl (all microcosm samples had a pH close to 7.00 before titration so that 71.5% of the TIC was present as titratable HCO 3 ). A correction factor of 1.4 was therefore used to convert this to TIC. Nitrate, nitrite and sulphate were determined by ion chromatography on samples that had been held in storage at 3 C to prevent further reaction (although it is likely that this did not halt degradation completely, see later discussion). Prior to ion chromatography, potentially column-damaging organics were removed with C18 solid phase extraction cartridges. Initially, the microcosm headspace was flushed with pure nitrogen. Before sacrificial sampling of each microcosm, changes in the headspace gas composition were monitored by packed column gas chromatography. Gas samples were withdrawn through PTFEcoated silicone septa. Oxygen and carbon dioxide concentrations were monitored to check for signs of air incursion and phenol degradation, respectively. During degradation, the carbon isotope ratio of TIC present in microcosm set A was expected to change due to the addition of carbon derived from p-cresol breakdown. Samples of TIC for d13C analysis were fixed as strontium carbonate by the addition of ammoniacal strontium chloride solution. The insoluble precipitate was then collected and dried prior to conversion to CO2 with 100% phosphoric acid in a vacuum line (McCrea, 1950). Purified gas samples were analysed using a VG Sira-10 gas source mass spectrometer and the data corrected using standard procedures (Craig, 1957). The precision, obtained from repeat analyses of a carbonate standard, was F 0.1x. Samples of microcosm sediment were analysed to determine the carbon isotope composition of any carbonate phases present. Sediment samples were tested both before and after exposure to the microcosm environment. Sediment carbonates were converted directly to CO2 by the addition of 100% phosphoric acid. Samples of CO2 were then purified and the 13C/12C ratio determined as described above. p-Cresol 13C/12C ratios were determined by a VG isochrom GC-IRMS. Unacidified microcosm samples containing p-cresol were extracted into a small volume of di-isopropyl ether (IPE) containing 200 mg/l bromobenzene (internal standard); 0.1– 3 ml of IPE was then injected into a BP 20 GC column with a split outlet. One-half of the split went to a flame ionization detector (FID) and the other into a copper oxide furnace at 800 C. Here, organic compounds were oxidised to CO2 and water. The stream of CO2 was purified by freezing out water at 100 C prior to continuous determination of the 13C/12C ratio. The precision, obtained from repeat analyses of an internal standard (bromobenzene), was F 0.2x.
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5. Results and discussion Fig. 1 shows the concentration of p-cresol in active microcosm sets A and B vs. a sterile control over the 35-day period as measured by GC and HPLC. Analyses obtained by GC were more precise until the addition of nitrate on day 12 gave problems with the extraction procedure. Hence, GC data are presented from day 1 to day 11 and HPLC data from day 15 to day 35. Prior to nitrate addition on day 12, C/Co (Measured/Initial concentration) values for pcresol in active microcosm sets A and B and the control sets D and E were close to 1.00, showing that no degradation had taken place. Degradation proceeded rapidly in the nitrate-bearing set C with the 40 mg/l of nitrate initially present being entirely converted to nitrite between days 4 and 11. This indicates a lag time of less than 3 days for nitrate reduction. After nitrate addition to sets A and B, a steady increase in alkalinity and headspace CO2 was observed that coincided with the appearance of three new organic compounds in solution (Figs. 2A and B). HPLC analysis tentatively identified these as 4-hydroxybenzyl alcohol (4HBalc), 4-hydroxybenzaldehyde (4HBald) and 4-hydroxybenzoic acid
Fig. 1. Variation in C/Co ( p-cresol) over the 35-day period for p-cresol degradation in the active microcosm sets A and B vs. the azide sterilised controls (set E). The addition of 1000 mg/l of nitrate on day 12 stimulated rapid pcresol degradation in the active sets.
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(4HBacid). The identity of these compounds in selected microcosms was confirmed by GC-MS. Previous mechanistic studies (Bossert and Young, 1986) show that they form through progressive oxidation of the methyl group on the p-cresol molecule. After pcresol exhaustion in set A, concentrations of 4HBalc and 4HBald began to diminish as they were converted to 4HBacid. If the experiment had been continued, then the concentration of all the intermediates would be expected to fall to zero after removal of the precursor p-cresol. The groundwater used for the microcosms from borehole 12 contained low concentrations of phenol, o-cresol and m + p cresol (5, 7, and 5 mg/l, respectively). The p-cresol concentration was increased by spiking with 200 mg/l, with the result that phenol and ocresol were insignificant in terms of the overall carbon budget. Phenol degradation coincided with the main phase of p-cresol breakdown whereas o-cresol was recalcitrant over the 35-day period. This recalcitrance is in accordance with previous studies conducted under comparable conditions (Bakker, 1977). Throughout the 35-day period, the sulphate concentration remained constant at 61 mg/l, indicating that sulphate had not been used as an electron acceptor. Similarly, microcosms sampled on day 19 after the onset of degradation contained no dissolved iron, indicating
Fig. 2. Concentration profiles of key species in microcosm sets A and B over the 35-day period. Analyses of dissolved phenolics and TIC are presented as mM C equivalents. Nitrate concentration is plotted in mM and the percentage of headspace CO2 is shown on the secondary axis. Nitrate reduction is associated with the appearance of p-cresol degradation products in solution, an increase in TIC and a rise in headspace CO2.
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that iron (III) oxide was not being used as an oxidant. An electron balance was calculated for each microcosm to determine whether denitrification could account for the observed pcresol removal and production of partially oxidised intermediates. Electron acceptor/donor reactions were used as shown in Table 2. The electron-accepting capacity of denitrification was expressed in terms of nitrate to nitrite conversion (deduced from the nitrite concentration). Note that nitrate remained at high concentration throughout the duration of the experiment, thereby inhibiting the secondary reduction of the nitrite ion (Kornaros et al., 1996). Most of the added nitrate was reduced to nitrite, leaving only 10% unaccounted
Table 2 Table of reactions used to calculate an electron balance for p-cresol degradation Electron-donating reactions Stepwise oxidation of methyl group: HOC6 H4 CH3 þ H2 O Z HOC6 H4 CH2 OH þ 2Hþ þ 2e
ðIÞ
HOC6 H4 CH2 OH þ H2 O ZHOC6 H4 COH þ 4Hþ þ 4e
ðIIÞ
HOC6 H4 COH þ H2 O ZHOC6 H4 COOH þ 2Hþ þ 2e
ðIIIÞ
Terminal oxidation step (mechanism unknown) HOC6 H4 COOH þ 11H2 O Z 7CO2 þ 28Hþ þ 28e
ðIVÞ
Summary equation HOC6 H4 CH3 þ 13H2 O Z 7CO2 þ 34e þ 34Hþ
ðVÞ
Electron-accepting reactions Nitrate reduction þ NO 3 þ 2e þ 2H Z NO2 þ H2 O
ðVIÞ
Denitrification þ NO 3 þ 5e þ 6H Z 1=2N2 þ 3H2 O
ðVIIÞ
Summary equation for p-cresol breakdown by nitrate reduction HOC6 H4 CH3 þ 17NO 3 Z17NO2 þ 4H2 O þ 7CO2
ðVIIIÞ
Summary equation for p-cresol breakdown by denitrification 102NO 3 þ 15HOC6 H4 CH3 Z102HCO3 þ 3CO2 þ 60H2 O þ 51N2
ðIXÞ
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for. This may have undergone complete reduction, or been converted to other partially reduced species such as N2O. For the electron balances presented here, we have assumed reduction to N2. Fig. 3 shows electron balances for sets A and B; these indicate that denitrification could account for 74% to 81% of the observed p-cresol degradation. The remaining 19– 26% is likely to be the result of aerobic sample degradation between sampling and HPLC analysis; chilling samples to 3 C slowed the rate of degradation rather than halting it. Carbon isotope ratios for microcosm TIC are plotted in Fig. 4, along with the initial groundwater and sediment carbonate ratios. The initial ratio in set A (d13C (TIC) = 17.3x) was consistent with a mixture of heavy carbonate (d13C = 1.1x) and organically derived carbon (d13C = 25x). Nitrate addition on day 15 promoted a shift to more negative values (d13C(TIC) = 19.6x by day 35) that coincided with the main phase of p-cresol degradation (d13C p-cresol = 27x). The change in d13C (TIC) was used to calculate the inorganic carbon contributions from p-cresol oxidation and carbonate dissolution. From days 25 to 35, the percentage of additional TIC derived from p-cresol fell from 98% to 82%, reflecting the dissolution of carbonate due to rising PCO2. This 18% contribution to the TIC from carbonate dissolution would remove 0.063 wt.% CaCO3 from the sediment (10% of that initially present ( 0.77 wt.%)). On a field scale, this reduction would not be noticeable from whole-rock analyses.
Fig. 3. Electron balances for active sets A and B over the 35-day period presented as measured nitrate loss vs. the predicted nitrate loss as calculated from the products of degradation. The gradient of the lines is 0.77, indicating that nitrate reduction could account for 77% of the organic matter oxidised.
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Fig. 4. A reduction in d13C(solution TIC) is coincident with the main phase of p-cresol degradation between days 15 and 35. This is consistent with the addition of 12C-enriched TIC derived from the oxidation of p-cresol (d13C p-cresol = 27x ).
A microcosm carbon balance was obtained by recalculating inorganic and organic carbon components as mM C and summing their concentrations to obtain total dissolved carbon. Carbon balances for active sets A and B are presented in Fig. 5a and b as stacked area plots showing the relative contributions of organic and inorganic carbon sources during the experiment. The carbon mass balance was corrected for the proportion of inorganically derived TIC found above (18% at end of experiment). Any carbon removed into the biomass would be reported as a deficit in the dissolved carbon budget. No deficit is apparent, and hence, the carbon available as p-cresol at the start of the experiment was primarily consumed in respiration processes. As p-cresol degradation proceeded in the nitrate reducing experiments, a small but potentially significant enrichment in 13C was seen in the residual p-cresol. Fig. 6 shows a best-fit curve through the data points assuming a Rayleigh fractionation model with an isotope enrichment factor (e) of 2.5x. This assumes a constant selectivity for 12C bearing p-cresol during the degradation process leading to a build-up of 13C in the residual compound. The curve was plotted from the following equation (Fritz and Fontes, 1980): dt ¼ do þ e Ln f ; where: do = initial d13C of p-cresol (x); dt = final d13C of p-cresol (x); e = enrichment factor ( 2.5x) f = C/Co = fraction of original p-cresol remaining.
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Fig. 5. Analyses of dissolved phenolics and bicarbonate presented as mM C equivalents and plotted in a stacked form over the duration of the experiment. The upper boundary of this plot represents the sum total of all the carbon compounds analysed. The dissolved nitrate concentration is also presented. Note that the number of moles of carbon in solution remains approximately constant during degradation.
Fig. 6. Experimental data are plotted that show a slight increase in d13C (residual p-cresol) with decreasing f ( = C/ Co). A best fit Rayleigh fractionation curve for this trend has been plotted that corresponds to an isotope enrichment factor (e) of 2.5x .
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Potentially, the change in residual p-cresol d13C during reaction can be used to monitor the extent to which degradation has proceeded in the field. However, the small enrichment factor (e = 2.5x) will only produce a significant change in d13C (residual p-cresol) when C/Co values become very low. In Fig. 6, this is illustrated by the extrapolation of the enrichment trend. For e = – 2.5x, kinetic fractionation will shift the isotopic composition of the parent p-cresol by > 4x only when C/Co < 0.2. This limits the use of this technique to situations where either the initial d13C variability in the compound is small or the extent of degradation is large. In the core of the plume, high C/Co values for the major pollutants are inferred from the fact that the present-day TIC could be derived from less than 7% of the present-day TOC (Thornton et al., 2001, this issue). At the plume margin, however, dispersion into surrounding groundwater reduces the initial pollutant concentration and the high nitrate concentration facilitates degradation. Hence, C/Co values may fall sufficiently to produce recognisable isotope enrichment trends. As a result, the plume margin is the only place where isotope fractionation is likely to be observed. In the present-day plume, the rate of dispersive mixing is slow with the result that denitrification is organic carbon limited just outside the plume and nitrate limited
Fig. 7. Depth section through the plume margin at MLS borehole 59 showing concentration variation in TOC, nitrate and nitrite. An analogue for the nutrient-enriched microcosm environment exists at a depth of 10 m where dissolved nitrate co-exists with pollutant phenols derived from coal tar.
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immediately inside it. This lack of electron acceptor –electron donor overlap is illustrated by high-resolution nitrate and TOC profiles for the plume margin at borehole 59 (Fig. 7, Thornton et al., 2001, this issue). The transition from high NO3 low TOC, to high TOC low NO 3 occurs over a depth interval of less than 2 m. Only one sample, taken from a depth of 10 m, straddles the transition and this contains 27 mg/l of nitrate, 1 mg/l of nitrite and 118 mg/l of TOC. This is the only depth where nitrate and TOC are present in sufficient concentration to produce appreciable degradation, and the presence of nitrite suggests that denitrification may follow a similar mechanism to that noted in the microcosms.
6. Conclusions A microcosm investigation into the potential for denitrification in the margin of the phenol plume has revealed that nitrate-reducing bacteria can rapidly degrade a relatively high concentration (200 mg/l) of p-cresol with the simultaneous degradation of trace phenol. After nitrate addition, p-cresol was degraded by stepwise oxidation of the methyl group prior to the formation of inorganic carbon. In the plume core, low dissolved nitrate concentration limits the activity of nitrate reducers, as demonstrated by the rapid onset of nitrate reduction in the microcosms after nitrate addition. At the plume margin, however, dispersive mixing of phenols and groundwater nitrate may stimulate the in situ microbial community to degrade at least part of the pollutant mixture. Recalcitrant pollutants such as o-cresol may also degrade, but most workers report lag times longer than the duration of this experiment. On a longer time-scale, degradation by alternative pathways, such as bacterial sulphate reduction and methanogenesis, may also be important for the reduction of contaminant load. These processes are harder to simulate in the laboratory, however, since they may have very long lag times. Experimental conditions can be modified to increase rates of reaction, but then, the microcosm may not be a good proxy for the real system. In a separate study (Spence et al., 2001, this issue), in situ rates of bacterial sulphate reduction have been estimated from stable sulphur isotope analyses of dissolved sulphate within the plume. A stable carbon isotope investigation of TIC composition has revealed that p-cresol degradation via the reduction of nitrate to nitrite was associated with the dissolution of carbonates in the aquifer sediment. This was attributed to the production of carbonic acid (Eq. (VIII) above). In a nitrate-limited system, however, the complete reduction of nitrate may occur according to Eq. (IX). In this case only 3% of the inorganic carbon produced would be in the form of carbonic acid and carbonate precipitation could therefore occur due to the production of bicarbonate ions. The carbon isotope fractionation during p-cresol degradation by denitrifying bacteria has been shown to be small (e = 2.5x). As such, it would be difficult to use this phenomenon to track degradation in the aquifer. The best location for applying this technique would be at the extreme outer margin of the plume where dispersion has increased the nitrate concentration relative to TOC. In summary, field data from the site have been shown to broadly agree with results presented from the microcosm study. Pollutant degradation by nitrate reduction will
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remain confined to a narrow interval surrounding the plume unless the plume is manipulated to artificially increase the degree of dispersive mixing. Following such remediation, a potential exists for significant in situ biodegradation through denitrification.
Acknowledgements This work was supported by NERC grant GT22/97/6/ENVD as part of the Environmental Diagnostics Initiative, and was made possible by the co-operation of the site owner and EnvirosAspinwall. We also thank R. Leader for assistance in field sampling and microcosm construction.
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