microbiological analysis of multi-level borehole samples from a contaminated groundwater system

Microbiological analysis of multi-level borehole

samples from a contaminated groundwater system

R.W. Pickup a,*, G. Rhodes b, M.L. Alamillo a, H.E.H. Mallinson a,S.F. Thornton c, D.N. Lerner c

aNERC-CEH Institute of Freshwater Ecology, Windermere, The Ferry House, Far Sawrey, Ambleside,

Cumbria LA22 0LP, UKbFreshwater Biological Association, Windermere Laboratory, The Ferry House, Far Sawrey, Ambleside,

Cumbria LA22 0LP, UKcGroundwater Protection and Restoration Group, Civil and Structural Engineering, University of Sheffield,

Mappin Street, Sheffield S1 3JD, UK

Received 9 March 2000; received in revised form 7 September 2000; accepted 5 October 2000

Abstract

A range of bacteriological, geochemical process-related and molecular techniques have been used

to assess the microbial biodegradative potential in groundwater contaminated with phenol and other

tar acids. The contaminant plume has travelled 500 m from the pollutant source over several decades.

Samples were obtained from the plume using a multi-level sampler (MLS) positioned in two

boreholes (boreholes 59 and 60) which vertically transected two areas of the plume. Activity of the

microbial community, as represented by phenol degradation potential and ability to utilise a range of

substrates, was found to be influenced by the plume. Phenol degradation potential appeared to be

influenced more by the concentration of the contaminants than the total bacterial cell numbers.

However, in the areas of highest phenol concentration, the depression of cell numbers clearly had an

effect. The types of bacteria present were assessed by culture and DNA amplification by polymerase

chain reaction (PCR). Bacterial groups or processes associated with major geochemical processes,

such as methanogenesis, sulphate reduction and denitrification, that have the potential to drive

contaminant degradation, were detected at various borehole levels. A comparative molecular analysis

of the microbial community between samples obtained from the MLS revealed the microbial

community was diverse. The examination of microbial activity complemented those results obtained

through chemical analysis, and when combined with hydrological data, showed that MLS samples

provided a realistic profile of plume effects and could be related to the potential for natural

attenuation of the site. D 2001 Elsevier Science B.V. All rights reserved.

Keywords: Groundwater; Phenol; Microbial ecology; Natural attenuation

0169-7722/01/$ - see front matter D 2001 Elsevier Science B.V. All rights reserved.

PII: S0169-7722 (01 )00169 -3

* Corresponding author. Tel.: +44-1539442468; fax: +44-1539446914.

E-mail address: [email protected] (R.W. Pickup).

www.elsevier.com/locate/jconhyd

Journal of Contaminant Hydrology 53 (2001) 269–284

1. Introduction

Natural attenuation is rapidly gaining favour as a feasible option in the remediation of

polluted aquifers (Stapleton and Sayler, 1998) and has been reported in a number of

environments. However, some are better suited to bioremediation than others due to

limitations imposed by hydrology and the nature of the pollutants (Lui and Suflita, 1993;

Bouwer et al., 1994; Stapleton and Sayler, 1998). The natural attenuation process relies on

the collective abilities of microorganisms to degrade pollutants under prevailing environ-

mental conditions (Zarda et al., 1998). To demonstrate natural attenuation, the degradation

of pollutants must be associated with degradative activity/capability of the microbial

community. The limitations of classical microbiological methods coupled with the

inherent non-representative nature of sampling strategies hinder any full characterisation

of a microbial community (Pickup, 1995) and their specific role in the attenuation process.

Conversely, the impact of high pollutant loads on bacterial communities is poorly

understood. However, by applying a multidisciplinary approach, comprising bacteriolog-

ical, molecular analyses and biogeochemical process-related studies coupled with detailed

hydrological data, the potential for attenuation can be assessed. Using this approach, this

study has focused on a tar acid-polluted aquifer (Lerner et al., 2000). The sampling site,

which overlies a major aquifer of Triassic Sandstone in the Midlands of England, was

chosen because a significant amount of site characterisation was already available. This

revealed the presence of a major groundwater plume comprising a complex mixture of

phenolic compounds derived from the distillation of acidified coal tars (Lerner et al.,

2000).

Prior to detailed investigation, the site history and historical groundwater data were

reviewed. Existing boreholes were also sampled and groundwater subjected to chemical

and microbial analysis. The data have been used to construct a conceptual model of the

plume and to identify factors, which may be tested to assess the degree of natural

attenuation (Williams et al., 2001). Since the primary attenuation process affecting organic

compounds is biodegradation, the focus of the study was to identify microbiologically

active zones within the groundwater for comparison with trends in the distribution of

organic and inorganic species to determine those conditions favourable or unfavourable for

natural attenuation (Williams et al., 2001). To further this aim, this manuscript presents a

detailed microbial analysis of two areas within the plume sampled using vertical multilevel

sampling (MLS) devices constructed within two boreholes (boreholes 59 and 60; Thornton

et al., 2001).

2. Materials and methods

2.1. Sampling and sample handling

Groundwater samples (not aquifer matrix material) were collected from two multi-level

samplers (MLS; boreholes 59 and 60) located at the contaminated site in the West

Midlands (UK) in May 1998 according to Thornton et al. (2001). Samples were transferred

to sterile 2.5-l bottles until all the headspace was removed and then sealed. Samples for

R.W. Pickup et al. / Journal of Contaminant Hydrology 53 (2001) 269–284270

isolation of culturable bacteria were processed within 72 h of sampling whilst samples for

activity analysis remained sealed until required.

2.2. Isolation of culturable bacteria

Culturable aerobic heterotrophs were isolated on a medium comprising tryptone soy

broth (TSB; Oxoid, Basingstoke, UK; 50% of manufacturer’s recommended strength)

supplemented with Bacteriological Agar No. 1 (Oxoid, UK) to give a final concentration

of 1.3% w/v. Bacteria were incubated at 20 �C for 1 week prior to enumeration.

2.3. Diversity/activity assessment of MLS samples

2.3.1. Assessment of bacterial enzyme activity

A comparison of the potential metabolic activity of the microbial community was

carried out according to Morgan and Pickup (1992) using the range of carbon sources on

the API 20NE test strip (BioMerieux, France). Samples (0.5 ml) from each of the

boreholes were inoculated into API 20NE strips and the ability to utilise 20 substrates

was tested. Positive reactions were scored and grouped in accordance with the manu-

facturer’s instructions. The strips were examined daily and recorded for a maximum of 10

days.

2.3.2. Potential degradative activity assessment using 14C-substrates

Phenol degradation potential of samples was determined aerobically and anaerobically

using 14C-phenol (representing the major pollutant at the site) as a substrate by measuring

the evolution of 14CO2 in a Packard Liquid Scintillation counter (1900TR). Twenty

millilitres of sample was placed in a 40-ml Wheaton vial and 100 ml of phenol-UL-14C(0.05 mCi; Sigma, UK, specific activity 41.22 mCi mmol � 1) was added. Anaerobic

samples were degassed with oxygen-free N2 (5 min) prior to adding phenol-UL-14C.

Controls comprised 18 ml sample and 2 ml 2% (v/v) formaldehyde with phenol-UL-14C

added 4 h after adding formaldehyde and a filter sterile water control comprising 18 and 2

ml formaldehyde (blow over control). All samples were incubated at 20 �C for 5 days. The

reaction was terminated by adding 200 ml of 0.1 M HCl. The headspace of each sample

was removed from Wheaton vials by displacement with air into a scintillation vial

containing 2 ml Carbo-sorb (Packard Bioscience) and 15 ml Permafluor E + (Packard

Bioscience). Degradation potential was calculated as the difference between the exper-

imental sample and its corresponding control after the blow over control value had been

subtracted from each. Each activity was expressed in disintegrations per minute (dpm).

2.4. Methanogenesis, sulphate reduction and denitrification potentials within the

contaminant plume

Methanogenesis, sulphate reduction and denitrification potentials were measured in

samples taken from the contaminant plume. Methanogenesis was measured in 40 ml

samples placed in 125-ml sealed jars pre-flushed with N2 and degassed by bubbling

N2 through the liquid for 5 min. Degassing for longer than 5 min changed the pH con-

R.W. Pickup et al. / Journal of Contaminant Hydrology 53 (2001) 269–284 271

siderably. Headspace sub-samples were taken every 3 days and analysed for methane

concentration using a Perkin-Elmer Gas chromatograph as described by Hall et al.

(1996). Denitrification activity was assessed as previously described (Miskin et al., 1998)

and using the API 20NE test strips (BioMerieux). Sulphate-reducing bacteria were

enumerated using the multiple tube MPN method as described by Miskin et al. (1998).

2.5. Total bacterial cell counts

Acridine orange direct count (AODC) of bacterial cells were performed as described by

Miskin et al. (1998). The sample was filtered through a 0.22-mm pore size black

polycarbonate membrane (Sigma, UK) and then stored in the dark before counting within

20 min. AODC stained cells were visualised with a Zeiss axioplan fluorescence micro-

scope (Carl Zeiss, Germany) using appropriate optical filters. Triplicate filters were wet

mounted in low fluorescence immersion oil (Leica, Germany). Counts were made in

accordance with the recommendations of Fry (1990).

2.6. DNA extraction from MLS samples

Samples (10 and 100 ml) were filtered through sterile 0.2-mm pore size filter (Supor

200, Gelman, UK), the filter then placed in 1–10 ml PBS and cell resuspended by gentle

agitation. Cells were recovered by centrifugation at 13,000� g on a MSE Microcentaur

microcentrifuge (Fisher-Scientific, UK). DNA was extracted from these cells using the

Masterpure DNA purification kit (Cambio, UK) as described by the manufacturer.

2.7. Polymerase chain reaction (PCR) amplification

PCR was carried out using a Hybaid Omnigene thermal cycler using pre-aliquoted PCR

Master mixes (1.5 mM MgCl2; Advanced Biotechnologies, UK). For the amplification of

bacterial 16S rDNA primers pA and pHVwere used (Edwards et al., 1989). Primers and the

conditions ofMorgan et al. (1993) were used for amplification of xylE.Methanogen-specific

PCR amplification used primers 1100F and 1400R (Kudo et al., 1997). Enterobacterial

repetitive intergenic consensus (ERIC)-PCR was carried out according to Di Giovanni et al.

(1999) and thermal gradient gel electrophoresis (TGGE) with primers F984GC and R1378

(Heuer et al., 1999). Primers were added to reaction tubes at a concentration of 20 pmol. For

16S rDNA andmethanogen amplifications, the following protocol was used: 1 cycle of 95 �Cfor 3min; 55 �C for 1min and 72 �C for 2min followed by 29 cycles of 95 �C for 1min; 55 �Cfor 1min and 72 �C for 2min finishing with 1 cycle differing only in the 72 �C extension step

increased to 10 min. Conditions for ERIC-PCR and analysis and generation of binary ma-

trices and dendrograms from banding patterns were as described byDi Giovanni et al. (1999)

using average linking distance and Euclidian distance package within Systat version 8.0.

2.8. Thermal gradient gel electrophoresis (TGGE) analysis

TGGE analysis of amplified products was carried out using the TGGE apparatus

(BioRad, UK). Aliquots of (5–20 ml) of PCR product were electrophoresed in a 6%


acrylamide, 8 M Urea, 20% formaldehyde gel (pre-equilibrated to the temperature

gradient of 45 �C for 30 min) in TAE buffer (Sambrook et al., 1989) at a constant

voltage of 100 V for 16 h with a thermal gradient of 47–60 �C. The gel was then stained

in 200 ml 1�TAE buffer containing 4 ml silver green in a sealed light-proof container

for 30 min with mild agitation. The gel was then scanned using Imagequant (Amersham-

Pharmacia Biotech).

3. Results

3.1. Bacterial content of samples from MLS

Total bacterial cell counts were determined in acridine orange stained samples. General

non-specific staining of bacteria using acridine orange enabled the total number of bacteria

to be estimated. For borehole 59 (less contaminated than BH60; Fig. 1a), bacterial

numbers ranged from 2� 105 to 7� 106 cells ml � 1 with no obvious trend with respect to

sample depth or phenol concentration. Samples from borehole 60 (Fig. 1b) contained

between 7� 104 and 6� 106 cells ml � 1 apart from samples from 23, 26 and 33 m where

no count was possible and coincided with the maximal concentrations of phenol. At these

depths, the contaminant levels prevented an accurate counting with acridine orange due to

the formation of an opaque precipitate. However, further microscopical investigation

showed that bacteria were present but in numbers too low to count statistically (Fry, 1990).

However, these areas were flanked by regions containing bacteria in concentrations in

excess of 105 cells ml� 1 (Fig. 1b).

3.2. Culturability of bacteria within the contaminant plume

In this particular survey, culturability was not examined, however, subsequent sampling

showed that the numbers of culturable bacteria on high nutrient media were approximately

two orders of magnitude lower than the direct count and severely depressed in the centre of

the plume where direct counts were below 105 cells ml� 1 (data not presented). In

borehole 60, where the direct count was not possible, no culturable bacteria were detected

(data not presented).

3.3. Bacterial activity assessment of MLS samples

3.3.1. Assessment of bacterial enzyme activity

API 20NE strips containing a range of test substrates were used to assess the

metabolic activity of the whole bacterial community in each sample. Using the API

index where growths on individual substrates are scored from a minimum value of zero

(no growth) to a maximum 49 (growth on all substrates being positive), borehole 59

showed a depression of activity between 10 and 30 m depth (Fig. 2a). For borehole 60,

no activity was detected at 23 and 26 m, this region was flanked at 21 and 33 m with

depressed activity compared to higher activity detected outside this region (Fig. 2b). In

each case, the depression of activity was associated with increasing phenol concentration.


This indicates that the diversity of organisms that are active within the region of high

phenol concentration has changed and reduced to a population of active bacteria that

grow on a limited substrate range.

3.3.2. Phenol degradation potential

Biodegradation of phenol-UL-14C to 14CO2 under aerobic and anaerobic conditions was

used to show the phenol degradation potential of microbial populations at different depths

within boreholes 59 and 60 (Fig. 3a,b). In general for both boreholes, the aerobic activity was

greater than that obtained under anaerobic conditions. Both boreholes showed a depression

of activity between 10 and 30 m (borehole 59) and 17 and 42 m (borehole 60). Regions of

higher activity flanked each depression. Similar profiles were obtained with the API data

(Fig. 3a,b) and both could be related directly to the increase in phenol concentration. It should

be noted that the differences in degradation potential observed may be due either to the

Fig. 1. Acridine orange direct count of bacteria from MLS samples with depth below surface (m) for boreholes 59

(a) and 60 (b).


depressed bacterial degradative activity caused by the concentration of phenol (plume effect)

or a competition effect between the phenol present in the sample and the added radioactive

substrate (experimental effect). In this case of the latter, high indigenous concentrations of

phenol would result in lower phenol-UL-14C assimilation. This would result in lower 14CO2

production compared to less polluted samples and would suggest an apparent depression of

activity where, in reality, the samples may have the same activity. However, given the

depression of substrate utilization (Section 3.2), cell numbers and culturability, the depres-

sion of degradation potential is considered to be a true reflection of the effect of the phenol

concentration in the plume rather than an artefact of the experimental system.

3.4. Denitrification, sulphate reduction and methanogenic potential

The potential activity of several geochemical cycles was assessed in a number of

ways along the vertical transects of each borehole (Table 1). The presence of sulphate-

Fig. 2. Metabolic activity of total bacteria community from MLS samples with depth below the surface (m) for

boreholes 59 (a) and 60 (b).


reducing bacteria (SRB) was determined by a culture-dependent MPN method. For

borehole 59, SRB were detected throughout the MLS apart from the 5 m sample. They

peaked in numbers (60 cells per ml) at 17 m depth, below which they were detectable

but in very low numbers (2–15 cells per ml; Table 1). In borehole 60, SRB were not

detected at 23 and 26 m but, in the flanking regions, their numbers increased the further

they were from this region of inactivity to maximum of 60 cells per ml (Table 1).

Methanogenesis was problematic to measure with methane being released from the

control samples even after formaldehyde or oxygen treatment. Therefore, no evidence of

Fig. 3. Potential of MLS samples to degrade 14C-phenol with depth below the surface (m) for boreholes 59 (a) and

60 (b).


the active process could be provided unlike the samples in the general site survey which

showed regions where methanogenesis was active (Williams et al., 2001). However,

PCR using methanogen-specific primers indicated that methanogens were present at all

depths in both boreholes apart from the sample obtained from 33 m in borehole 60

(Table 1). Bacteria capable of reducing nitrate to nitrite and nitrite to nitrogen gas were

detected in both boreholes but not at the depths 17/23 m (borehole 59) and 23/26/33 m

(borehole 60; Table 1). This indicated that denitrification as a process was potentially

active at these depths.

3.5. Molecular analysis of bacterial diversity at a community level

3.5.1. Detection of xyIE gene

DNA samples were subjected to PCR using primers specific for genes homologous to

xylE which encodes aromatic ring cleavage enzyme catechol 2,3 dioxygenase (Morgan et

al., 1993). Amplified products generated using xylE-specific primers (Morgan et al., 1993)

were detected in samples borehole 59 at depths of 10, 17 and 30 m and borehole 60 at 30

m (data not shown). No other sample depths were examined.

3.5.2. TGGE analysis

The resolution of 16S rDNA amplifications by TGGE analysis is shown in Fig. 4.

For borehole 59, there appears to be a high level of diversity apart from a common

Table 1

Detection of bacteria associated with reductive processes in MLS

Borehole/

depth (m)

Phenol

(mg ml� 1)

Bacterial group or activity

Sulphate reducing

bacteriaaMethanogensb Denitrifying

activityc16S rRNA

extractiond

59/5 < 1 � + + +

59/8 < 1 + + + +

59/10 < 1 + + + +

59/17 50 + + � +

59/23 95 + + � +

59/30 56 + + + +

60/5 < 1 + + + +

60/8 < 1 + � + +

60/17 < 1 + + + +

60/23 1007 � + � +

60/26 1004 � + � +

60/33 2665 + � � +

60/42 73 + + + +

60/44 52 + + + +

a Determined using MPN count method (Miskin et al., 1998).b Determined using PCR (Kudo et al., 1997).c Determined using API data (Biomerieux, manufacturer’s instruction).d Determined using PCR (Edwards et al., 1989).


dominant band identified in samples from depths of 5, 8 and 23 m and a different but

common band at depths 5 and 17 m. The co-migrating bands were more intense and are

likely to have originated from common species that may have been numerically

dominant at that depth. At 30 m, a band not detected at other depths dominated the

sample. For borehole 60, there appears to be more commonality between the predom-

inant bacterial communities at different depths. However, some diversity is apparent and

the sample obtained from 33 m appears to be dominated by species common to both

borehole depths 5 and 17 m. Some bands may be inferred as common to boreholes 59

and 60, although sequencing would be required to confirm whether they are identical.

Although the resolution of bands in borehole 60 is poor, there seems to be a greater

number of bands present that do not co-migrate with others within borehole 60 or with

those from borehole 59.

3.5.3. ERIC-PCR

The presence of enterobacterial repetitive intergenic consensus (ERIC) sequences

generated by ERIC-PCR was used to generate a high resolution fingerprint of the

bacterial populations from different depths in the plume (Fig. 5a,b). For boreholes 59

Fig. 4. TGGE analysis of PCR-amplified 16S rDNA gene fragments showing diversity of bacteria in groundwater

samples. (1–6) Borehole 59: 5, 8, 10, 17, 23, and 30 m depth, respectively; (7–14), borehole 60: 5, 8, 17, 23, 26,

42, and 44 m depth, respectively. (15–16) Positive controls (known standards), Dbmv and BG8, respectively.

Arrows indicate some common bands.


and 60, the complexity of the banding patterns produced was high with up to 25

distinct amplification products detected in some samples. The patterns (denoted by

arrows, Fig. 5a) showed that although some common bands were observed there was a

difference in the population fingerprint between different depths and between boreholes

(Fig. 5a). The increase in levels of phenol in borehole 59 did not markedly affect the

numbers of bands within each fingerprint observed. In borehole 60, the banding

patterns of regions experiencing the highest phenol insult were clearly affected (below

23 m). However, the dendrogram (Fig. 5b) generated by cluster analysis using a binary

matrix from the observed banding patterns showed that there was no clear relationship

Fig. 5. (a) ERIC-PCR profiles of MLS samples from boreholes 59 and 60. (1) l DNA size marker; (2–7)

borehole 59: 5, 8, 10, 17, 23, and 30 m depth, respectively; (8–15) borehole 60: 5, 8, 17, 23, 26, 42, and 44 m

depth, respectively. Arrows indicate some common bands. (b) Dendrogram from cluster analysis of ERIC-PCR

profiles.


between the amplified products from between levels, between boreholes or with phenol

concentration.

4. Discussion

Application of a wide range of molecular and bacteriological analyses on the samples

from the MLS have permitted the potential for microbially mediated natural attenuation to

be assessed within a plume containing high levels of phenol and associated aromatic

contamination (Williams et al., 2001). The strategy adopted was to relate the biodiversity

and microbial activity to the prevailing chemical environment. It is well established that

the representative nature of a sample to its environment decreases as the degree of

destruction of the sample increases (Pickup, 1996). Alfreider et al. (1997) emphasised that

groundwater per se did not reflect the bacterial densities and activity of the solid aquifer

matrix. They showed that the proportion of respiring bacteria in sandy sediment was 25%

compared to 5% for pumped groundwater (Alfreider et al., 1997). When sand enclosed in a

fine mesh was introduced for two months into a number of boreholes at the site, the

activity measured was similar to that generated by the pumped samples as was the

depression of activity which was related to the TOC concentration (data not presented).

Despite this, some caution should be exercised when extrapolating results from pumped

water samples to core samples unless comparative experiments have been carried out

(Pickup, 1995). In this case, the effects of environmental variables such as plume

contaminants on activity can be related as the experimental protocols permit in an analysis

of ‘potential’ rather than the actual in situ activities (Pickup, 1995; Hall et al., 1996;

Alfreider et al., 1997; Stapleton and Sayler, 1998).

Fig. 5 (continued ).


This study was able to show that in regions where bacteria were present at approx-

imately 104 cells ml� 1, both the anaerobic and aerobic degradation potential and bacteria

associated with anaerobic processes such as sulphate-reduction, methanogenesis and

denitrification, known to drive degradative processes under anaerobic conditions (Ander-

son and Lovely, 1998), were present. Chemical analysis showed that the phenol concen-

trations in borehole 60 were greater than borehole 59 (Thornton et al., 2001). The chemical

profile of the samples from the MLS showed that the phenol (and total organic carbon)

reached a peak between 17 and 23 m depth for borehole 59, and 23 and 33 m depth for

borehole 60. The rise in phenol concentration is concomitant with a decrease in

degradative activity, catabolic activity and a depression in total bacterial numbers and

culturability. The differences in phenol insult can be seen from a rapid fall in activity in the

central depths of borehole 60 which was less pronounced in borehole 59. In the case of

borehole 60, bacterial numbers were depressed to > 102 cells per ml� 1 where peak phenol

concentrations were encountered, which in turn returned very little or no bacterial activity

as shown by the degradation of phenol-UL-14C and the response of the bacterial

population to a range of substrates (API20NE). In addition, we have previously shown

that 98% of aerobic bacteria isolated from deep subsurface sediments are facultative

anaerobes (Miskin et al., 1998) so that by adopting techniques that were not rigorously

anaerobic, the obligate anaerobes and oxygen sensitive organisms (e.g. methanogens) will

be the most affected by the ingress of oxygen (Hall et al., 1996), although others may be

affected by sampling per se (Pickup, 1995). Therefore, although methanogens probably

dominated the anoxic zones, the culture and process measuring limitations imposed

through non-rigorous anaerobic sampling were offset by detection using culture-inde-

pendent techniques such as PCR.

Molecular analyses permit an assessment of both culturable and nonculturable bacteria

(Head et al., 1998). Studies within this site suggest that some common organisms exist

although none appear to be ubiquitous and that there is considerable diversity between

depths and between sites. The diverse organisms, however, are linked probably by

common phenotypes such as degradative and biogeochemical potential rather than

showing genotypic conservation for these processes (Ridgway et al., 1990; Daly et al.,

1997; Bergwall and Bengtsson, 1999; Wilson et al., 1999). The use of ERIC-PCR,

although limited in its powers of discrimination at species-level, showed the differences

between samples inferring some diversity at a cellular level (Versalovic et al., 1991;

Hulton et al., 1991; Bruijn, 1992; Di Giovanni et al., 1999). Potentially higher resolution at

a species level from amplification of 16S rDNA was achieved through TGGE analysis

(Muyzer and Smalla, 1998). This showed that no particular species dominated the entire

vertical transect through the plume in boreholes 59 and 60. However, at a few sample

depths common and potentially dominant species were apparent as shown by band

intensity. In general, the bacterial populations between the boreholes and at depths within

each borehole were different. Only the sequencing of TGGE bands and the subsequent

development of fluorescent probes from these sequences for fluorescent in situ hybrid-

ization (FISH) could assess whether these bands are of any ecological significance (Porter

et al., 1996). The number of bands observed in ERIC-PCR and TGGE analyses and the

differences between samples through the variety of banding patterns observed is proof of

diversity. Moreover, both techniques showed responses to changes in levels of phenol


although no consistent pattern emerged that was common to both. This was probably due

to the limitations inherent in both techniques.

Although our data show natural attenuation to be slow or non-existent (Lerner et al.,

2000; Williams et al., 2001), there is a potential outside the central core of pollutants for

degradation to occur. However, combining chemical and molecular data extended our

powers of assessment of degradative potential within this plume over a conventional

chemical/hydrological analysis. For instance, the demonstration of the presence of

methanogens by PCR and its linkage with chemical data, despite our inability to recreate

the process experimentally, suggested that in some regions methanogenesis was actually

occurring. This contrasts with the findings of Shi et al. (1999) who were unable to detect

any Archaea. Stapleton and Sayler (1998) used a similar combined approach to assess

BTEX degradation in aquifer material and concluded that favourable conditions existed for

natural attenuation on the basis of microbial activity. Stapleton and Sayler (1998), like this

study and Williams et al. (2001), found that bacteria with multi-aromatic degradative

capability were ubiquitous and molecular techniques showed the presence of key genes

necessary for this process (e.g. xylE) at a number of depths within the boreholes. However,

without the chemical data this study would have come to a similar but incorrect conclusion

that natural attenuation was active. The chemical data suggest otherwise that although the

potential is present, attenuation is not active apart from the plume fringes (Thornton et al.,

2001). Many of the aromatic compound degrading genotypes studied in groundwater

(Stapleton and Sayler, 1998; van der Meer et al., 1998; Hosein et al., 1997) are linked to

oxidative processes requiring molecular oxygen (e.g. dioxygenase genes; Daly et al.,

1997) and, therefore, are only active in oxygenated areas such as the plume fringe

(Davison and Lerner, 1999). An anaerobic core dominates the plume so the presence of

these genotypes will have no significant effect on the degradative process. There is ample

evidence of degradation being coupled to oxidative processes such as denitrification

(Edwards et al., 1992; Edwards and Grbic-Galic, 1994; Anderson and Lovely, 1998;

Rooney-Varga et al., 1999). Conditions for coupling to oxidative processes are present and

above 25 m are active given, for example, the sulphide:sulphate ratios within the plume

(Thornton et al., 2001). The conclusions drawn from this study are therefore that if the

conditions were favourable then the biological components are in place for attenuation to

occur apart from the region of very high total organic carbon (Thornton et al., 2001).

However, the ratio of plume components (e.g. phenol:cresol) suggested that natural

attenuation is not a feasible process to remove the plume contaminants in the timescale

required (Williams et al., 2001). In the absence of favourable conditions, and given the

potential for degradation, then intervention may be an option to initiate the biodegradative

process (Head, 1998). This study would direct such an approach to suitable areas in the

plume (Power et al., 1998).

Acknowledgements

Support from EPSRC and environment Agency for RWP, HEHM and MLA is

gratefully acknowledged. We would also wish to thank the EU Leonardo Da Vinci

programme (MLA).


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