Microbiology (1998), 144, 599408
Printed in Great Britain
Bioremediation:towards a credible technology
ARTICLE
Ian M. Head
Tel: +44 191 222 7024. Fax: +44 191 222 5431. e-mail: [email protected]
Fossil Fuels and Environmental Geochemistry (Postgraduate Institute), Drummond Building,
University of Newcastle, Newcastle upon Tyne NE1 7RU, UK
Keywords : bioremediation, bioavailability
Context
Bioremediation is the technological process whereby
biological systems are harnessed to effect the clean-up of
environmental pollutants. Currently, microbial systems
are most widely employed in bioremediation programmes, generally in the treatment of soils and waters
contaminated with organic pollutants. Micro-organisms
have a huge metabolic repertoire that enables them to
degrade a panoply of organic pollutants and in many
cases the complex biochemistry and molecular biology
of the catabolic pathways involved have been unravelled
(e.g. Gibson, 1984; Frantz et al., 1987; Evans & Fuchs,
1988; Burlage et al., 1989; Abramowicz, 1990; Assinder
8c Williams, 1990; Chaudhry & Chapalamadugu, 1991 ;
Cerniglia, 1992 ; Knackmuss, 1996). Despite valuable
basic knowledge on the mechanisms of pollutant biodegradation, bioremediation has yet to be accepted as a
routine treatment technology and the environmental
industry is wary of applying bioremediation for the
treatment of contaminated sites.
Bioremediation has the potential to treat contaminants
on site with relatively little disturbance to the contaminated matrix. Furthermore, micro-organisms offer
the possibility that organic pollutants can be completely
mineralized to inorganic materials (e.g. CO,, H,O, C1-,
NO,), making bioremediation an attractive treatment
strategy. In contrast, removal of contaminated material
to landfill sites, or extraction of contaminants using
physical processes such as soil washing, do not destroy
the contaminants present, but simply concentrate the
contaminated material in a different location. Why then,
if bioremediation offers benefits over other technologies,
has it not been more widely adopted to treat environmental contamination ? One reason is that physical
treatments are rapid and their outcome is generally
predictable in the short term; they are also relatively
inexpensive. Bioremediation too, can be relatively
cheap, but methods to confirm efficacy on a field scale
have either been unavailable or simply not applied
routinely. The unpredictability of bioremediation,
meanwhile, stems from a lack of understanding of the
behaviour of microbial populations in natural environ0002-1241 Q 1998SGM
ments and how physical, biological and chemical factors
interact to control their activity against environmental
pollutants. With this realization, has come a different
philosophy towards bioremediation ; the focus of bioremediation research has shifted from isolation and
construction of ' superbugs ' to determining the factors
that limit pollutant transformations and mineralization
in natural environments. It is now clear that access of
micro-organisms to pollutants in situ is a critical factor
in determining the success of bioremediation. Consequently, methods are being developed to enable predictions regarding the feasibility of bioremediation
based on pollutant bioavailability and biodegradation.
With knowledge of the causes of unsuccessful bioremediation, methods can be formulated to overcome
these limitations. Assessing the feasibility of bioremediation within a predictive framework and confirming its
efficacy in the field is the focus of this review.
Assessing bioremediation potential:a
problem of bioavailability
It is relatively simple to demonstrate biodegradation of
specific compounds in a contaminated environmental
sample. Spiking a sample of contaminated soil with a
pollutant chemical of interest and monitoring its loss
and the appearance of degradation end products in
comparison with sterilized control samples often demonstrates rapid biodegradation. Conversion of 14Clabelled compounds to 14C0, is undoubtedly the best
evidence that a microbial community has the ability to
mineralize an organic compound, but is insufficient to
demonstrate convincingly that bioremediation will be
successful. For example, added naphthalene and phenanthrene were rapidly degraded in a polycyclic aromatic
hydrocarbon (PAH)-contaminated soil, but levels of
indigenous naphthalene and phenanthrene were not
reduced appreciably (Erickson et al., 1993). Comparable
observations have been made in other soils and with
different chemicals (Steinberg et al., 1987; Doelman et
al., 1990; Weissenfels et al., 1992; Beurskens et al.,
1993). Reduced bioavailability results from the interaction of pollutants with both organic and inorganic
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I. M. HEAD
components of the soil matrix. Access of micro-organisms or their enzymes to the pollutant molecules is
constrained and with increasing contact time (‘ ageing’)
the proportion of the pollutant that becomes biologically
unavailable increases. A number of mechanisms are
involved: diffusion limitation due to sequestration of the
pollutant in micropores ( < 1 pm diameter ;Steinberg et
al., 1987; Wu & Gschwend, 1986) or humic and mineral
flocs (Kan et al., 1994);binding to soil minerals by ionic
or electrostatic interactions (e.g. Knaebel et al., 1994);
oxidative covalent coupling of the pollutant with soil
organic matter via enzymic or chemical catalysis (e.g.
Bhandari et al., 1996); and partition/dissolution of the
pollutants into the soil organic matter (e.g. Kan et al.,
1994). There is consensus that the last of these is
quantitatively the most important in immobilizing and
reducing the bioavailability of organic pollutants, particularly in organic-rich soils and sediments (Pignatello
& Xing, 1996). However, as the soil organic matter
content decreases (less than 0-4% organic carbon ;
Carmichael et al., 1997), catalytic effects of soil minerals
may result in a greater proportion of pollutant immobilization via oxidative covalent coupling with soil organic
matter (Bhandari et al., 1996). Clearly, several mechanisms operate in consort to influence bioavailability, and
different mechanisms will predominate in any given
situation (Pignatello & Xing, 1996).
The phenomenon of pollutant ‘ageing ’ in soils has been
investigated by examining adsorption and desorption
isotherms which, under ideal, equilibrium conditions,
should be identical. Experimental measurements typically indicate that desorption occurs at a much lower
rate than would be predicted from a simple equilibrium
model (e,g. Kan et al., 1994; Pignatello & Xing, 1996).
Furthermore, a discontinuity in the desorption of
pollutants from a soil or sediment is often observed.
Following initial rapid desorption of a fraction of the
bound contaminant, the remaining material desorbs at a
considerably reduced rate. Kan et al. (1994) demonstrated that equilibrium for adsorption of naphthalene
to soil was achieved within 24 h, but with increasing
adsorption time up to 30 d, the amount of ‘desorptionresistant ’ material increased from 15 to 23 YO.It has been
suggested that this is due to slow diffusion of pollutant
molecules from soil micropores (Steinberg et al., 1987;
Wu & Gschwend, 1986) or through soil organic matter
that is in a glassy liquid state (Carroll et al., 1994;
Pignatello & Xing, 1996). It has however been noted
that if diffusion kinetics are the cause, then adsorption
should be affected to the same extent as desorption (Kan
et al., 1994). This is not the case and it has been recently
demonstrated by the use of laboratory-constructed
homogeneous mineral/organic matter matrices lacking
tortuous, diffusion-limiting micropores that similar
‘two-phase’ desorption behaviour occurs (Hunter et al.,
1996).Whilst this is difficult to explain in the framework
of equilibrium kinetics, it is feasible that the activation
energy for desorption may be greater than the activation
energy for adsorption (Cornelissen et al., 1997 ; Pignatello & Xing, 1996).
600
At present, we do not fully understand the mechanisms
involved in reducing bioavailability, but it is certainly a
key factor in determining the feasibility of bioremediation. Bosma et al. (1997)recently commented that ‘ ... a
critical assessment of bioremediation data reveals that
the intrinsic microbial activities limit bioremediation in
only a few cases. In most cases, mass transfer limitation
prevented the full exploitation of the microbial degradative potential ’. While reduced bioavailability may
lead to failure of bioremediation, it can also be
responsible for reduced toxicity of pollutant residues.
Can bioavailability be measured and its e m c t on
bioremediationpredicted ?
Bioavailability has been estimated using microbial
bioassays (Heitzer et al., 1992), differential solvent
extraction techniques (Kelsey et al., 1997), analysis of
desorption thermodynamics and kinetics (Cornelissen et
al., 1997), and a combination of desorption, transport
and biodegradation kinetics (Bosma et al., 1997).
A microbial bioassay for assessment of bioavailability has been developed for naphthalene and its
degradation intermediate salicylate (Heitzer et al., 1992).
This system is based on a recombinant Pseudomonas
fluorescens strain (KH44)containing a plasmid-encoded
nahG-ZuxCDABE transcription fusion. In response to
induction by naphthalene or salicylate, nahG and the
lux operon are co-expressed, resulting in bioluminescence. When a whole-cell bioassay was tested in
aqueous extracts of soil artificially contaminated with
naphthalene, the reduction in aqueous phase concentration (approx. 96% by HPLC) due to sorption to the
soil matrix corresponded well with the reduction in
bioluminescence determined by the bioassay. However,
in soil slurries the bioassay was compromised due to
luminescence quenching (Heitzer et al., 1992). The
system was further developed to produce a flow-through
biosensor that enabled on-line monitoring of naphthalene and salicylate in process waters (Heitzer et al.,
1994). Although the biosensor showed a reproducible
response to pulsed concentrations of naphthalene and
salicylate, the response time was slow (8-24 min) and
dependent on the concentration of the analyte. Further
limitations of the bioassay included a long signal decay
period (approx. 30 min) and sensitivity to the presence
of toxic substances. Rather than bioavailability, it was
essentially aqueous phase pollutant concentration that
the biosensor responded to and significant refinement
would be required to compete with more traditional
measurement techniques.
Bioassay.
Differential extraction. A simple approach to determining
the bioavailability of aged residues of atrazine and
phenanthrene has been adopted by Kelsey et al. (1997)
and is based on the differential extraction of pollutants
from soil using a range of solvents. Up to 11 different
extraction regimes were compared for their ability to
reflect the degree of pollutant uptake by earthworms,
and the degree of bacterial mineralization in sterilized
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soils containing aged pollutants and inoculated with
specific pollutant-degrading pure cultures. Recovery of
atrazine with 1 :1 methanol :water for example, was
similar to the degree of bacterial mineralization of the
herbicide, whereas n-butanol extraction efficiency mirrored the level of phenanthrene mineralization. This
preliminary study was limited in the scope of pollutants
and soil tested and it is likely that empirical determination of the appropriate extractant for any particular
soil/pollutant system will be required. Nonetheless,
because the methodology is straightforward it could
prove an attractive means to assess bioavailability.
Mode//ingapproaches. The above methods may be of use
in the empirical determination of bioavailability, but do
not by themselves allow the likely success of bioremediation to be determined. Potentially more valuable
in this respect is the development of mathematical
models that can explain biodegradation behaviour in
relation to both biological and abiotic factors such as
pollutant sorption and diffusion.
Rigorous kinetic and thermodynamic analyses have
recently been applied to the problem of desorption of
pollutants from soil (Cornelissen et al., 1997) and the
effect of desorption on biodegradation rates (Bosma et
al., 1997). The desorption of chlorobenzenes, polychlorinated biphenyls (PCBs) and PAHs was studied in
laboratory-contaminated and field-aged sediments (Cornelissen et al., 1997).The pollutants were found to occur
in three ' compartments '. The pollutants desorbed from
one of the compartments rapidly, from a second slowly
and from a third very slowly. From 8 to 54% of the
pollutant was present in the slowly desorbing fraction,
with the exact value depending on the particular
chemical and sediment. Up to 47%, though more
typically 0.5-16 Yo, of some contaminants was found to
be present in the very slowly desorbing compartment
(Cornelissen et al., 1997). The latter can equate to
significant quantities of pollutant in a contaminated soil
and its occurrence in this form has considerable implications for the time that may be required to effect
satisfactory bioremediation. The very slow desorption
phase could only be measured in reasonable time scales
if desorption was carried out at elevated temperatures,
and values obtained for the fraction of very slowly
desorbed material at 20 "C and 60 "C were generally in
good agreement. This approach offers the possibility
that not only the amount of pollutant in differentially
desorbing pools can be determined, but also the time
taken for complete desorption to occur under field
conditions can be estimated. This may be invaluable in
determining the length of time required for bioremediation if it is limited by mass transfer of pollutants from
a sorbed phase to solution phase, as is generally the case
with aged pollutant residues (Bosma et al., 1997;
Carmichael et al., 1997).
The biodegradation of a contaminant in situ is a function
of both the catabolic activity of the micro-organisms
present and transport of the contaminant to microbial
cells with the ability to degrade the contaminant. It is the
latter that is probably the most important factor in
determining the success of a bioremediation programme.
By considering both of these factors, a parameter termed
the bioavailability number (Bn)has been derived (Bosma
et al., 1997). The Bn is equal to a first-order exchange
constant k (kcontrols the rate of desorption/diffusion of
a sorbed contaminant), divided by the product of the
maximal transformation rate (qmax) and the reciprocal
of the half-saturation constant for contaminant transformation (K;') (equation 1).This is essentially the ratio
of the exchange constant and the first-order rate
constant applicable to Michaelis-Menten kinetics
(equation 2) at substrate concentrations much less than
K , (equation 3).
Bn
= k/qmaXK;'
(1)
4 = qmax [Sl/(Km+ [Sl)
(2)
When [S] is much lower than K , this approximates to
(3) with qmaX/Kmas a first-order rate constant.
4 = (qmax/Km) [SI
(3)
The Bn is hence a measure of the importance of mass
transfer relative to the intrinsic catabolic activity of the
microbial cell or population. Thus, for values of Bn < 1
desorption and diffusion (i.e. mass transfer) are more
important than biodegradation in determining the rate
of transformation of a pollutant. At values of Bn > 1,
the reverse is true and the rate of transformation is not
limited by desorption and diffusion. In other words, if
mass transfer (k) is large relative to biodegradation
(qmax/Km) bioavailability is not likely to limit biodegradation. The value of k can be calculated for
different modes of mass transfer (e.g. linear diffusion vs
radial diffusion vs dissolution) using experimentally
determined values for distribution coefficients and
dissolution rate constants (Rijnaarts et al., 1990; Bosma
et al., 1997). Values of qmax and K , can be determined
by measurement of initial biotransformation rates at
different added substrate concentrations in soil slurries.
The parameters can then be determined graphically (e.g.
using a Lineweaver-Burk-type plot).
This approach was used recently to investigate the
importance of mass transport and hence bioavailability
on a-hexachlorocyclohexane (a-HCH)-degradation in
soil slurries (Bosma et al., 1997). The exchange constant
was calculated from a-HCH desorption data using a
first-order model (Rijnaarts et al., 1990) and values for
Km and qmaxwere determined using Lineweaver-Burk
plots (Bachmann et al., 1988). The Bn for this system
was calculated at 00164-030 under different mixing
regimens that affected mass transfer. Both values were
however less than one and hence biodegradation was
likely to be sorption limited. When the biodegradation
process was modelled using an expression that takes
account of both mass transfer and biodegradation, the
fit with experimental data was extremely good (Bosma
et al., 1997).This relatively simple method may therefore
prove useful in predicting the applicability of bioremediation if it can be shown to be valid over a wide
range of conditions.
The same authors also developed an expression relating
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I. M. HEAD
the exchange coefficient and the quantity of pollutant
transformed to satisfy cell maintenance requirements to
the threshold concentration below which no further
biodegradation would be observed. High and low values
for substrate fluxes required to satisfy maintenance
energy were estimated at
pg s-l for copiotrophs
and
pg s-l for oligotrophic bacteria, using published data on the maintenance energy requirements for
a range of bacteria (e.g. Chesboro et al., 1979; Bouillot
et al., 1990; Tros et al., 1996a, b). Exchange coefficients
representing a range of mass transfer conditions from
the diffusion of hydrophilic compounds in water to the
diffusion of hydrophobic compounds in soils were also
calculated. With this information, threshold pollutant
concentrations expected under different mass transfer
conditions and with bacterial populations with various
maintenance energy requirements were determined
(Bosma et al., 1997). The range from low maintenance
energy systems with high mass transfer to high maintenance energy systems with low mass transfer exhibited
threshold concentrations that varied from
pg 1-' to
greater than lo' pg l-l, respectively. It was apparent
from this analysis that under certain conditions residual
concentrations of pollutant, even in microbiologically
active, environments, can be very high and this is
consistent with the levels of residual pollutant often
measured in the field.
The fundamental importance of bioavailability in determining the success of bioremediation is becoming more
widely appreciated (e.g. Tabak et al., 1994; Bosma et al.,
1997).The development of methods that can potentially
predict the extent of bioavailability based on relatively
simple measurements and calculations is likely to have a
significant impact on decisions to bioremediate. Certainly, more research is required to determine if predictions made from kinetic and thermodynamic data are
widely applicable and to extend their utility to complex
mixtures. As we increase our understanding of the
mechanisms involved, it is likely that predictions will
become more reliable and the efficacy, and hence the
reputation, of bioremediation will undoubtedly be
improved. Continued isolation of particular organisms
with the ability to catabolize specific pollutant chemicals
may yet reveal much regarding the biochemistry and
genetics of biological transformations of novel xenobiotics and may even prove useful in 'end-of-pipe'
processes for the treatment of well-defined wastestreams. However, improvement of the efficacy of
bioremediation of contaminated matrices in the environment will require a redistribution of research effort,
away from such studies and towards a greater understanding of the limitations on microbial transformations
in situ.
Can bioavailability limitation be overcome ?
The prospects for predictable evaluation of bioremediation efficacy are improving, but it is not clear that this
knowledge will allow development of more effective
means to treat organic pollutants biologically in contaminated soils and sediments.
602
A number of strategies for improving mass transfer have
been suggested. Diffusion and desorption of pollutants
are temperature-dependent (Cornelissen et al., 1997)
and this is the basis of physico-chemical treatments such
as thermal desorption and steam-stripping. In the
context of bioremediation it has been suggested that the
use of composting systems, where elevated temperatures
are maintained, should improve mass transfer and hence
bioremediation rates (Pignatello & Xing, 1996); rapid
degradation of chlorophenols in contaminated soil by
composting has been demonstrated accordingly (Laine
& Jarrgensen, 1996; Jaspers et al., 1997). Physical
grinding or mixing to disaggregate soils has been
suggested as a means of alleviating mass transfer
limitation, and there is some evidence that this is
effective (e.g. Rijnaarts et al., 1990). The cost involved in
large-scale application of physical disruption to soils to
the degree required to improve mass transfer substantially may, however, be prohibitive. The most widely
investigated method for improving bioavailability of
sorbed pollutants is treatment with surfactants. Surfactant-enhanced biodegradation of sorbed and poorly
soluble chemicals is probably due to decreasing the
distribution coefficient (K,) for the pollutant in a soilwater system, i.e. increasing the equilibrium concentration in the aqueous phase relative to the sorbed phase.
This would require the surfactant to be provided at a
concentration greater than the critical micelle concentration (CMC; Alexander, 1994; Deitsch & Smith,
1995), unfortunately requiring the use of substantial
amounts of surfactant that can be both costly and toxic
to micro-organisms. There is however evidence that
mineralization rates can be increased even at surfactant
concentrations well below the CMC (Aronstein et al.,
1991). Recent evidence suggests that although high
surfactant concentrations do lower K,, concentrations
below the CMC increase the mass transfer exchange
coefficient (equivalent to k in equation 1; Deitsch &
Smith, 1996). In essence, this means that in abiotic
systems the equilibrium concentration in the aqueous
phase does not change, but the time for equilibrium to
be reached is reduced. Consequently, enhanced rates of
mineralization can be observed at surfactant concentrations below the CMC despite no increase in the
aqueous equilibrium concentration being noted in
abiotic control systems.
In a number of cases however, surfactant addition has
had no effect or been detrimental (see Liu et al., 1995 for
a summary) due to toxicity of the surfactant at high
concentrations (e.g. Aronstein et al., 1991) or retardation of surface attachment of bacterial cells to the
hydrophobic contaminant (e.g. Efroymson & Alexander, 1991). Furthermore, the identity of the surfactant
and its concentration are critical. One study compared
the effects of seven surfactants (anionic and non-ionic)
on desorption of phenanthrene, of which five had no
effect on desorption of the PAH from sterile soil
(Aronstein et al., 1991). The effect of the remaining two
non-ionic surfactants on the mineralization of phenanthrene and biphenyl was quite different in mineral and
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organic-rich soils. It is interesting to note that one of the
surfactants, Alfonic 810-60, did not increase the apparent aqueous concentration of phenanthrene at equilibrium in sterile slurries of the mineral soil. Despite this,
mineralization was markedly enhanced, supporting the
view that it is the mass transfer coefficient and not K ,
that is affected by surfactant addition.
On the basis of pollutant- and soil-specific effects
(Aronstein et al., 1991; Providenti et al., 1995), it seems
likely that a universally applicable surfactant treatment
to promote the biodegradation of poorly bioavailable
contaminants will be elusive. Consequently, surfactantenhanced bioremediation is likely to rely on ad hoc
rather than proprietary solutions. A need for case by
case evaluation has considerable cost implications for
bioremediation and the formulation of generic treatment
practices would reduce the unit cost. The development
of reliable, widely applicable practices will be essential if
bioremediation is to compete succesfully with engineering solutions. In order to achieve this, it is important
to understand processes rather than rely on empirical
observations, and the causes of reduced bioavailability
are now being rigorously investigated.
Field demonstration of the efficacy of
bioremediation
One of the greatest challenges faced by advocates of
bioremediation is proof that a chosen treatment is
effective under field conditions. Concentration of contaminant compounds and residue toxicity are the most
common criteria specified by legislative bodies to define
an end point for successful bioremediation. Chemical
analyses (e.g. GC-MS) and toxicological assays (e.g.
Microtox) can be applied to accurately identify and
quantify organic contaminants or assess residual toxicity
following treatment. Nevertheless, a major factor
hindering the acquisition of statistically valid proof of
bioremediation efficacy in the field is the heterogeneous
distribution of pollutants. For example, simple gravimetric analysis of oil residues from the Exxon Valdez oil
spill was inadequate for the determination of hydrocarbon bioremediation. The variability in oil loading on
beach sediments, determined from replicate samples,
was so great (SEM as high as 224 YO; Bragg et al., 1992)
that valid conclusions could not be reached. Even after
sieving a contaminated soil from a derelict coal extraction plant, the variation in GC-MS analyses of
certain PAHs in replicate samples can be greater than
100% (Mundell, 1994). To overcome this problem a
number of strategies have been devised that either alone
or in combination may provide the body of evidence
required to demonstrate that bioremediation has been
effective.
Poorly biodegradable components of complex
mixtures
The overall biodegradation of contaminants that comprise a complex mixture of compounds with different
susceptibilities to biodegradation can be assessed by
*.O
r
3 7.5
x
C
0
r
3n 7.0
U
E
= 6.5
17a(H)21P(H)-hopane
Fig, 1. Changes in the ratio of total GC-detectable
hydrocarbons (TGCDHC) t o hopane during the bioremediation
of the Exxon Valdez oil spill. The solid black line represents
data for fertilized beach sediments and the dashed line data for
untreated beach material. The grey areas represent the 95 %
confidence intervals. Inset, structure of 17a(H)Zlj?(H)-hopane, a
poorly biodegradable component of crude oil that has been
used t o index the degradation of more labile components of
crude oil. Redrawn from Bragg e t a / . (1992).
measuring the ratio of the degradable components to a
poorly degradable component in the mixture. Crude oil
is an excellent example of such a contaminant mixture
and this approach has been used successfully to confirm
biodegradation of weathered oil spilled from the Exxon
Valdez (Bragg et al., 1994). Crude oil biodegradation
studies have commonly used the ratio of the linear
alkane n-heptadecane (n-C,,) to the more resistant
branched alkane pristane (or alternatively n-octadecane: phytane ratios) to evaluate the extent of biodegradation (e.g. Fayad et al., 1992; Pritchard & Costa,
1991). A decrease in the ratio of the n-alkane to the
branched alkane is taken as evidence of crude oil
biodegradation. However studies of beach sediments
from Prince William Sound, Alaska, demonstrated that
even the branched alkanes were readily degraded,
resulting in anomalously high n-C,, :pristane ratios in
highly degraded oils (Bragg et al., 1992). For this reason,
17a(H)21j?(H)-hopane,a minor component of the North
Slope crude spilled at Prince William Sound, was used to
index the biodegradation of the more labile components
of the oil (Bragg et al., 1994). 17a(H)21p(H)-hopaneis a
pentacyclic triterpane (Fig. 1) and was chosen as a
conserved internal marker because is known to be
poorly biodegradable. Hopanoids are membrane lipids
found in many bacterial taxa (Rohmer et al., 1984) but
they exhibit stereochemical differences to those occurring in crude oil and can be distinguished readily.
Comparisons of fertilizer-treated beach plots with untreated plots using the ratio of either the total GCdetectable hydrocarbons, total PAH or total resolvable
hydrocarbons to the hopane concentration provided
convincing evidence that bioremediation had been
effective (Fig. 1; Bragg et al., 1994).
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I . M. HEAD
A,,,,
9
Succinyl product
COOH
Succinyl product
6OOH
CH3
Fumaryl product
Xylene
I
cn3
coon
................................................................................ ................................................................
Fig. 2. Succinyl and fumaryl products from the anaerobic
*
catabolism of substituted benzenes. These compounds have
been used as indicators of in situ biodegradation of BTEX in
ground water.
Whilst assessment of crude oil biodegradation has been
the main application of this approach to date, it can be
applied to other complex mixtures. PCBs, for example,
are normally found as a mixture of congeners with
similar physico-chemical properties but different biological fates and are therefore amenable to this approach. 3,4-3',4'-tetrachlorobiphenyl is more readily
biodegradable than 2,3,6-3',4'-pentachlorobiphenyl and
the ratio of the two congeners has been used to monitor
bioremediation of PCBs in river sediments (Harkness et
al., 1993). There is potential for 2,3,6-3',4'-pentachlorobiphenyl to be transformed by 3-, 3'- and 4'-dehalogenation reactions, compromising its use as a conserved
tracer. This is because anaerobic dechlorination has
been shown to result in preferential dehalogenation at
meta- and para-positions in PCB molecules (Bedard &
May, 1996). However, independent studies have demonstrated that the meta and para chlorines of the 2,3,63',4'-substituted congener are not susceptible to reductive dehalogenation in anoxic sediments (Bedard et
al., 1996).
Analysis of degradation products
Analysis of metabolic products or intermediates of
contaminant metabolism in situ is not universally
applicable as many intermediates may be short-lived or
only accumulate to very low levels. Some features of
metabolic products applicable in the diagnosis of
successful bioremediation are : an unequivocal biochemical relationship with the parent compound ; no
604
exogenous sources of contamination ; and biological
and chemical stability under in situ conditions (Beller et
al., 1995). Degradation products of trichloroethylene
(TCE) metabolism meet some of these criteria in that
biological degradation produces cis-1,2-dichloroethylene whereas abiotic TCE transformation results in the
formation of 1,l-dichloroethylene (Kastner, 1991).
Anaerobic biodegradation of the alkyl substituted components of BTEX (benzene, toluene, ethylbenzene,
xylenes) is also known to produce characteristic byproducts under denitrifying and sulphate reducing
conditions (Beller et al., 1992; Evans et al., 1992) and
this has been exploited recently under field conditions
(Beller et al., 1995). Benzylsuccinate, benzylfumarate
and corresponding homologues accumulate during anaerobic degradation of BTEX in laboratory incubations
(Fig. 2). These compounds could be detected in samples
from an anaerobic aquifer contaminated with gasoline,
implying that anaerobic transformations of the BTEX
had occurred in the aquifer (Beller et al., 1995).
Furthermore, in an aquifer spiked with BTEX-contaminated water (containing bromide as a conservative
tracer) it was possible to demonstrate that the accumulation of the metabolic products followed the
disappearance of the corresponding contaminant hydrocarbons. When BTEX became undetectable, the concentration of the metabolic products began to decrease.
This approach is not fail-safe in determining biodegradation at field sites, as work on pentachlorophenol
degradation has illustrated. Both photolytic (Steiert &
Crawford, 1986) and biodegradative reactions (Apajalahti & Salkinoja-Salonen, 1987) produce the same
intermediates. It may thus be impossible to distinguish
these reaction mechanisms under field conditions, and
unequivocally associate disappearance of pentachlorophenol with a biological process, simply by detection of
putative metabolic products.
Geochemical indicators of bioremediation
Except in fermentation and disproportion reactions, the
oxidation of organic compounds is coupled to the
reduction of exogenous electron acceptors. In soil and
sediment environments these are usually O,, NO,, Fe3+,
Mn4+, SO:- or CO,. Depletion of these species or
accumulation of their reduced products in contaminated
relative to uncontaminated material may be indicative
of organic pollutant metabolism (Borden et al., 1995).
This can be achieved most convincingly if the concentrations are compared with levels of a conservative
tracer (e.g. chloride or bromide) that has similar
physico-chemical properties to the analyte of interest,
but is not susceptible to biological reduction. Indirect
evidence of this nature is far from definitive unless
reduction in contaminant concentrations can be related
to changes in oxidant concentrations and a mass balance
constructed. In the field, these objectives are often
difficult to meet and geochemical analysis of redoxsensitive species at best provides supplementary evidence
for biodegradation.
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Bioremediation
Stable isotopes
A rapid method that is used to monitor the progress of
bioremediation is measurement of CO, production.
While this gives an indication of increased rates of
breakdown of organic matter it does not necessarily
follow that the CO, originates from the target pollutant.
The advantage of using stable isotope analysis is that it
is possible to determine the source of CO, evolved
during a remediation process. This is feasible because
there is very little isotopic fractionation associated with
the aerobic mineralization of organic matter by microorganisms (Chapelle et al., 1988; Jackson et al., 1996).
Whilst potentially a very useful technique to rapidly
assess the biodegradation of contaminants, it is only
applicable if the 613C of the endogenous organic matter
and the contaminant are measurably different.
A good example of the potential of this approach has
been published recently (Jackson et al., 1996). The 613C
ratio of crude oils is typically -29% to -32%,. In
contrast, organic carbon in soils and sediments dominated by C-4 plants exhibits a 613C signature of between
-14.4ym and -17.7%. Thus it has been possible to
clearly distinguish the rate of CO, production from
hydrocarbon degradation and from mineralization of
endogenous organic matter in oil-contaminated salt
marsh sediments dominated by Spartina (Jackson et al.,
1996). In addition, kinetic constants calculated using the
613C data for CO, were statistically no different from
those calculated from data on alkane degradation. That
hydrocarbon degradation had occurred was further
indicated by reductions in the ratio of labile hydrocarbons to hopane.
The carbon isotope ratio of CO, can be readily and
inexpensively determined. This method of assessing the
biodegradability of organic pollutants under field conditions therefore has considerable promise for adoption
in bioremediation studies (e.g. Aggarwal & Hinchee,
1991; Landmeyer et al., 1996; Aggarwal et al., 1997). Its
application, however, will be limited to situations where
the 613C of the contaminant and endogenous organic
matter are known and distinct.
A mle for molecular biology in assessing
bioremediation?
The ability to isolate and enumerate bacteria from
contaminated sites capable of degrading a particular
pollutant is one line of evidence often used to support
the feasibility of bioremediation. This is particularly
true if an increase in the population of degradative
bacteria occurs following implementation of a treatment
to stimulate biodegradation. Culture-based methods
underestimate both qualitative and quantitative
measures of microbial populations by orders of magnitude. This has led to the development and application
of nucleic-acid-based techniques to study the ecology
and diversity of micro-organisms in nature. Since a more
complete understanding of microbial ecology will undoubtedly be required to gain maximum benefits from
bioremediation, it is not surprising that molecular
biological methods are now being employed to study
bioremediation. This can be viewed as analogous to the
isolation of bacteria with appropriate catabolic properties from a polluted site, as a means of implying that
competent degradative populations are present at the
site, and thus that bioremediation potential exists.
Molecular biological methods have already been used
successfully to evaluate the bioremediation potential of
a contaminated site (Fleming et al., 1993). The expression of a catabolic gene (nahA) involved in the
initial oxidation of naphthalene to 1,2-dihydroxynaphthalene was investigated at a manufactured gas
plant site contaminated with PAHs. RNA was extracted
from the contaminated soil and mRNA for the nahA
gene quantified using an RNase protection assay (Belin,
1996). The extracted RNA was hybridized in solution
with a radioactively labelled antisense nahA RNA probe
transcribed from a cloned nahA gene. Treatment of the
hybridized RNA preparation with RNase preferentially
degrades single-stranded RNA leaving pro be
RNA:mRNA hybrids intact. The amount of
radiolabelled probe RNA: mRNA hybrid is quantified
in relation to known amounts of target RNA analysed
using the assay. Comparable samples of the contaminated soil were used to determine the frequency of nahA
genes in a cultured fraction of the bacterial population
by a colony hybridization protocol. Residual naphthalene concentrations and the rate of radiolabelled
naphthalene catabolism were also measured. Comparison of the data revealed a good correlation between
nahA gene expression, as determined by quantification
of nahA mRNA, and the other three parameters. The
ability to monitor expression of specific catabolic genes
may be of great importance in determining the feasibility
of bioremediation in situ, particularly since samples can
be taken from the site and processed with minimal
disturbance due to storage and transport or alteration of
conditions in the sample, as occurs with culture based
techniques.
More recently, a similar approach has been adopted to
examine the expression of lignin peroxidases (Lamar et
al., 1995)and manganese-dependent peroxidases (Bogan
et al., 1996) in soil inoculated with Phanerochaete
chrysosporium. These enzymes are believed to be
important in the degradation of organic pollutants by
white-rot fungi. In these studies, reverse transcriptase
competitive PCR was used to quantify different levels of
the gene transcripts and transcription of the fungal genes
was shown to be related to the degradation of contaminant PAHs (Bogan et al., 1996). This was in turn
related to the activity of manganese peroxidase in
extracts of the soil.
Although these examples demonstrate the feasibility of
using molecular methods to monitor bioremediation,
routine practical application is perhaps a little way off.
A number of factors will determine how widely adopted
these approaches will be for routine evaluation and
monitoring of bioremediation programmes (Brockman,
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605
I . M. HEAD
1995).To date, molecular biological studies of catabolic
genes in contaminated samples have focused on well
characterized metabolic traits. For example, the occurrence of meta-cleavage dioxygenase genes involved
in the catabolism of substituted benzenes, naphthalene
and biphenyl have been examined in a number of
samples. These have naturally targeted genes for which
most sequence data are available, for instance xyZE-like
(e.g. Joshi & Walia, 1996) and 6phC genes (Erb &
Wagner-Dobler, 1993). However, meta-cleavage genes
(approx. 30 sequences in the public databases) exhibit
considerable diversity (e.g. Kukor & Olsen, 1996; Daly
et al., 1997) and no single set of PCR primers or gene
probe will allow detection of all variants of known
meta-cleavage genes, let alone those that remain to be
discovered. This is a particular problem with analysis of
anaerobic pathways in pollutant degradation, for which
there is growing evidence of importance, but which
remain largely uncharacterized at the molecular level.
Notwithstanding methodological limitations, regulators
and practitioners of bioremediation must be convinced
of the benefits of molecular biological techniques. For
these approaches to be widely embraced, they must
provide information of practical benefit in assessing the
success or progress of bioremediation that cannot be
accessed by other means. Molecular biological analyses
may also be more costly than conventional chemical
analyses. Nonetheless, if a sufficiently large market for
molecular-biological environmental diagnostics exists,
it is not unreasonable to assume that expertise in medical
diagnostics could be adapted to produce cheap, reliable,
quantitative molecular assays for biodegradative microorganisms or catabolic genes. A further hurdle that must
be overcome is the acceptance of molecular data, which
is often not quantitative and unfamiliar to those not
directly involved in research using these methods. It is
therefore important that researchers communicate the
benefits and limitations of the techniques available and
do not attempt to apply them where tangible benefits
over other approaches cannot reasonably be claimed.
Currently, molecular biological techniques are excellent
research tools that increase our understanding of the
distribution and expression of important catabolic genes
and microbial population dynamics during biotreatment
(e.g. Massol-Deya et al., 1997). They are not yet
sufficiently robust for routine monitoring applications,
and as ‘stand-alone’ assays their use will always be
limited. Ultimately. if the techniques can be developed
to provide rapid reliable information on the occurrence,
quantity and activity of important biodegradative microbial populations that is complementary to chemical
and physical data, it will only be a matter of time before
they become more widely used.
Concluding remarks
Degradation of organic pollutants by micro-organisms
has been studied for many decades. The fruit of this
labour is a remarkable understanding of the biochemical
606
pathways and molecular genetics involved in the catabolism of a relatively small number of intensively
studied pollutants by a relatively small group of microorganisms. It may have been assumed from this knowledge that the catabolic diversity of micro-organisms
could be harnessed to solve a wide range of pollution
problems. On this basis, initial predictions for the
benefits of bioremediation were overstated with the
consequence that it was never likely to live up to
expectations.
It is now clear that it is not our knowledge of pollutant
catabolism that limits the success of bioremediation, but
rather a restricted understanding of the interplay between the biotic and abiotic factors that determine the
outcome of any particular remedial strategy. Treating
bioremediation as a natural bioengineering process that
takes account of these interactions, coupled with rigorous quantitative assessment of the outcome of remedial treatments is required to ensure bioremediation
is successful and to raise the credibility of the technology.
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