bioavailability of hydrophobic organic contaminants in soils fundamental concepts and techniques for...

Bioavailability of hydrophobic organic contaminants insoils: fundamental concepts and techniques for analysis

K. T. SEMPLEa, A. W. J. MORRISS

a & G. I . PATONb

aDepartment of Environmental Science, Institute of Environmental and Natural Sciences, Lancaster University, Lancaster LA1 4YQ,

and bDepartment of Plant and Soil Science, University of Aberdeen, Aberdeen AB24 3UU, UK

Summary

Soils represent a major sink for organic xenobiotic contaminants in the environment. The degree to

which organic chemicals are retained within the soil is controlled by soil properties, such as organic

matter, and the physico-chemical properties of the contaminant. Chemicals which display hydrophobic

and lipophilic characteristics, as well as a recalcitrant chemical structure, will be retained within the soil,

and depending on the ‘strength’ of the association may persist for long periods of time. This review

describes the behaviour of hydrophobic organic contaminants in soils, focusing on the mechanisms

controlling interactions between soil and contaminants. The bioavailability of contaminants in soil is

also discussed, particularly in relation to contact time with the soil. It considers the degradation of

organic contaminants in soil and the mechanisms microbes use to access contaminants. Finally, the

review discusses the ‘pros’ and ‘cons’ of chemical and biological techniques available for assessing

bioavailability of hydrophobic organic chemicals in soils, highlighting the need to quantify bioavailability

by chemical techniques. It concludes by highlighting the need for understanding the interactions between

the soil, contaminants and biota which is crucial to understanding the bioavailability of contaminants in

soils.

Introduction

The term ‘hydrophobic organic contaminant’ (HOC) is a

generic term covering a wide range of organic xenobiotic

chemicals that have found their way into the environment,

that are characteristically barely soluble in water, as well as

being fairly resistant to biological, chemical and photolytic

breakdown. They include the simple aromatic compounds,

namely benzene, toluene, ethylbenzene and xylenes (BTEX),

polycyclic aromatic hydrocarbons (PAHs), including naphtha-

lene, phenanthrene and benzo[a]pyrene, and polychlorinated

biphenyls (PCBs). People are concerned about the occurrence

and concentration of HOCs in the environment because the

compounds are potentially toxic, carcinogenic, have mutagenic

activities, and are persistent. Their potential impact on ecolo-

gical receptors is also a cause for concern. Concern was so

great that in the late 1970s the United States Environmental

Protection Agency (USEPA) listed 16 PAHs as ‘Priority

Pollutants’, and this list was subsequently adopted by the

European Union. A similar scenario has been applied to

many other contaminants.

The soil plays an important role in the fate of HOCs in the

environment. Contaminants enter the soil mainly by deliberate

application, by spillage and leakage, and by atmospheric

deposition. As a result, the soil is a sink for them. While

HOCs may be lost from the soil, significant concentrations

may be retained within soils (Figure 1). Consequently, the

fate and behaviour of organic contaminants in soils has been

the subject of intensive research, with particular interest dir-

ected at the bioavailability of contaminant in the soil.

In this review we discuss the interactions between contamin-

ants and the soil, evaluate the biological, chemical and phys-

ical factors that determine bioavailability, and compare

chemical and biological approaches for quantifying contamin-

ant behaviour in the soil.

The fate of hydrophobic organic contaminants in the

soil environment

After its arrival in the soil, an organic contaminant may be lost

by biodegradation, leaching or volatilization, or it may

accumulate within the soil biota or be retained or sequestered

within the soil’s mineral and organic matter fractions (Figure 1).

The fate and behaviour of HOCs in the soil is controlled byCorrespondence: K. T. Semple. E-mail: [email protected]

Received 19 June 2002; revised version accepted 19 November 2002

European Journal of Soil Science, December 2003, 54, 809–818 doi: 10.1046/j.1365-2389.2003.00564.x

# 2003 Blackwell Publishing Ltd 809

several factors including soil type (mineral and organic matter

content) and physico-chemical properties (e.g. aqueous solubil-

ity, polarity, hydrophobicity, lipophilicity and molecular

structure) of the contaminant(s) (Reid et al., 2000a).

Weakly polar hydrophobic and lipophilic compounds that

are sparingly soluble in water have a low vapour pressure

and a recalcitrant molecular structure and are retained strongly

within the soil (Reid et al., 2000a).

Contaminants can be removed from the soil at varying rates

and to varying extents. Figure 2 describes the theoretical loss

curves for three classes of contaminants: A represents a

water-soluble, highly mobile, easily degradable contaminant;

B represents the biphasic behaviour of most contaminants in

soil, where losses are clearly occurring – however, as the time

of contact increases between soil and contaminant, the rate

and extent of loss diminish; and C represents the slow loss of a

highly intractable chemical (Jones et al., 1996). If HOCs enter-

ing the soil are not completely removed by leaching, volatil-

ization or degradation, then their interaction with the

components of the soil must be considered, as denoted by the

arrow under curve B (Figure 2).

Normally, as the time of contact between contaminant and

soil increases there is a decrease in chemical and biological

availability, a process termed ‘ageing’ (Hatzinger & Alexander,

1995). Figure 3 demonstrates the influence of contact time on

the extractability and bioavailability of HOCs in soil. Over time,

the readily available fraction (easily extractable or bioavailable

fraction) diminishes in a biphasic manner, i.e. some is degraded

or lost from the soil and some is transformed into the

recalcitrant fraction. There is an increase in the recalcitrant

fraction, which can be accessed only by specific and sometimes

aggressive extractions, followed by a slower increase in a frac-

tion deemed to be non-extractable (Macleod & Semple, 2000).

The mechanisms of ageing have been much investigated

(Reid et al., 2000a). As a result, it is now known that interac-

tions between soil and HOCs are influenced by the soil organic

matter, both its amount (Hatzinger & Alexander, 1995) and its

nature (Piatt & Brusseau, 1998); inorganic constituents (Ball &

Roberts, 1991a,b; Mader et al., 1997) with particular reference

to pore size and structure (Nam & Alexander, 1998); microbial

activity (Guthrie & Pfaender, 1998), and contaminant concen-

tration (Divincenzo & Sparks, 1997).

The main mechanisms involved in the ageing are sorption and

diffusion (collectively termed sequestration), which are inter-

actions between the contaminant and the solid fractions within

soil, namely the mineral and organic matter fractions (Xing &

Pignatello, 1997; Schlebaum et al., 1998). These, together with

the contaminants’ physico-chemical characteristics, largely deter-

mine the rate and extent of ageing in soils (Alexander, 2000).

Degradation

Volatilization

Leaching

SequestrationBioaccumulation

Figure 1 Putative fate and behaviour of a model hydrophobic organic

contaminant (phenanthrene) in soil.

A

Time

Con

tam

inan

t con

cent

ratio

n

B

?????

C

Figure 2 Theoretical loss curves for three

classes of contaminants: A, a water-soluble,

highly mobile, easily degradable contaminant;

B, the biphasic behaviour of most contaminants

in soils, where losses are occurring, but diminish

with time; C, a highly intractable chemical. The

arrow under curve B represents the need to

characterize the processes retarding the loss of

the contaminant from the soil.

810 K. T. Semple et al.

# 2003 Blackwell Publishing Ltd, European Journal of Soil Science, 54, 809–818

Organic contaminants generally exhibit two kinetic (biphasic)

stages within the soil (Figure 3). Initially, a portion of the

contaminant can be sorbed quickly (in minutes to a few

hours), whereas the remaining fraction is sorbed more slowly

over weeks or months (Xing & Pignatello, 1997). The initial

rapid sorption is generally by hydrogen bonding and van der

Waals forces, mechanisms that are expected to occur instant-

aneously upon contact of the HOC with the soil surface (Dec &

Bollag, 1997; Gevao et al., 2000). Covalent bonds are most

likely to be associated with contaminants that are similar to

organic matter, i.e. they have phenolic structure. The resulting

bonds lead to stable almost irreversible incorporation into the

soil (Dec & Bollag, 1997; Gevao et al., 2000). However, as

HOCs’ sorption is generally governed by a partitioning between

the solution and organic matter phases, the specific interactions

between sorbate and sorbent causing chemisorption are unlikely

to affect these contaminants (Brusseau et al., 1991).

Two concepts have been proposed to describe the sequestra-

tion of HOCs in soils: (i) diffusion through organic matter and

(ii) sorption-retarded pore diffusion. Both have been described

in detail elsewhere (Pignatello & Xing, 1996; Cornelissen et al.,

1998) and are summarized below.

1 Diffusion into organic matter. The soil organic matter is

envisaged as comprising rubbery and glassy phases, both of

which contain dissolution sites. The glassy phase is also

thought to contain more rigid cavities (holes) where

contaminants can interact with the organic matter (Xing &

Pignatello, 1997). A contaminant thus diffuses into this complex

structure and is sequestered in the organic matter as it does so.

2 Sorption-retarded diffusion. A contaminant diffusing through

the pore water in the soil is retarded by sorption to surfaces

within the nano- and micropores in soils containing little

organic matter. The rates of diffusion are controlled by

the radii of soil particles, the tortuosity of pores, and steric

hindrance within pore spaces (Pignatello & Xing, 1996).

Many soils contain an abundance of pores with diameters of

20 nm or less (Alexander, 1995). Such pores are too small to

allow the smallest bacteria (1�m), higher organisms (protozoa

10�m) or root hairs (7�m) to penetrate them. A contaminant

residing in such fine pores is therefore protected from attack

by biota in the soil; it is not bioavailable.

Bioavailability

The term ‘bioavailability’ refers to the fraction of a chemical in

a soil that can be taken up or transformed by living organisms.

Two important factors determine the amount of a chemical

that is bioavailable: (i) the rate of transfer of the compound

from the soil to the living cell (mass transfer) and (ii) the rate

of uptake and metabolism (the intrinsic activity of the cell).

The bioavailability of a chemical is determined by the rate of

mass transfer relative to the intrinsic activity of the soil biota

(Bosma et al., 1997). Bioavailability has also been defined as

the degree to which a compound is free to move into or on to

an organism, and as such the term is best used in the context of

a specific organism(s) because it is known that bioavailability

differs between organisms and even species (Reid et al., 2000a).

For example, Kelsey et al. (1997) and White et al. (1997) found

that the extent of bioaccumulation by the earthworm (Eisenia

fetida) and mineralization by a bacterium (Pseudomonas sp.) of

phenanthrene in soil were different, though bioavailability to

both decreased with increased contact time. Further, Guerin &

Boyd (1992) reported differences in bioavailability between

species of bacteria in that Pseudomonas putida ATCC 17484

mineralized 32% of the sorbed naphthalene in 225 hours

whereas a Gram-negative isolate (NP-Alk) mineralized only

Degradable, removable fraction

Readily availablefraction

Recalcitrant fraction

Time

Con

tam

inan

t con

cent

ratio

n

Non-extractable fractionFigure 3 The influence of contact time on the

extractability and bioavailability of a con-

taminant.

Organic contaminants in soil 811


18%. Assessment of the bioavailability of contaminants in soil

is essential to understanding the risk posed by the contaminant

and the means required for successful remediation.

Temporal implications in contaminant fate and

bioavailability

It is well established that sequestration of organic contamin-

ants in soil reduces the bioavailability of organic chemicals

and results in a non-degraded residue in the soil. Contamin-

ants that have aged in soil are not available for degradation

even though freshly added compounds are still degradable

(Alexander, 1995). Sorption is the major factor preventing

complete bioremediation of hydrocarbon contamination in

soil (Bosma et al., 1997). Slow sorption results in a fraction

of the HOC becoming resistant to desorption and in increased

persistence within the soil matrix (Hatzinger & Alexander,

1995). The following hypotheses have been suggested as

explanations for ageing.

1 The aged fraction results from the slow diffusion of the

HOCs within the solid organic matter fraction of soil,

possibly the lipid fraction (Alexander, 2000). This concept is

supported by Nam et al. (1998), who suggested that the

organic matter sequesters the HOC, as described previously.

2 The contaminant slowly diffuses through the soil and

becomes sorbed and entrapped within nano- and micropores

within the soil (Hatzinger & Alexander, 1995), again as

described previously.

Of course, contaminants may become sequestered by a com-

bination of both the above mechanisms (Figure 4).

Evidence for the sequestration of contaminants includes

(i) laboratory and field investigations, which demonstrate a

decreasing availability to organisms (Chung & Alexander,

1998); (ii) investigations into the extractability of aged HOCs

and the kinetics of sorption and desorption (Hatzinger &

Alexander, 1995); (iii) temporal changes in the rate and extent

of contaminant mineralization (Hatzinger & Alexander, 1995;

Reid et al., 2000b), and (iv) the assessment of toxicity. This last

is very important for decisions regarding risk and environmen-

tal regulations; however, the evidence is based on only a few

studies by Salanitro et al. (1997) and Saterbak et al. (1999,

2000). Simplistically, ageing may be associated with

the continuous diffusion of HOCs into small pores where the

organic molecules are retained by sorption. This explains the

Diffusion into rubberyorganic matter

Diffusion in pores

Surfacesorption

Surfacesorption

Organic matter

Mineralfraction

Organic matter

Water-solublefraction

Diffusion into glassyorganic matter

Figure 4 A summary of the physical behaviour of a contaminant within the soil. For explanation see text.



decreases in solvent extractability and bioavailability of

HOCs. It also means that toxic organic chemicals that have

been in contact with the soil matrix for a long time are unlikely

to be available to humans, animals or plants (Alexander,

1995). However, we still do not know how long this fraction

will remain in this state or whether the contaminant(s) will

remobilize and so become extractable and bioavailable.

Degradation: evolution of catabolism

The ability of the soil’s microbial community to degrade

HOCs is fundamental to soil health and fertility. One of the

principal mechanisms that accounts for the removal of HOCs

from soils is the catabolic activity of the microbes (Pritchard &

Bourquin, 1984). Soil microflora have a diverse capacity for

attacking HOCs. This catabolic ability is due primarily to the

co-evolution of soil microflora and naturally occurring organic

compounds, which contain chemical structures analogous to

those of HOCs (Dagely, 1975). The rate of microbial decom-

position of HOCs in soils is a function of several factors, either

singly or in combination (Macleod et al., 2001):

1 the availability of the contaminants to the microorganisms

that have the catabolic ability to degrade them;

2 the numbers of degrading microorganisms present in the soil;

3 the activity of degrading microorganisms, and

4 the molecular structure of the contaminant.

However, the processes that control the evolution of catabolic

activity in soils are not well understood. The catabolic activity

can develop by adaptation, by the following processes:

1 the induction or depression of specific enzymes;

2 the development of new metabolic capabilities through

genetic changes, such as plasmid transfer or mutation, and

3 selective enrichment of organisms able to transform the

target contaminant(s) (Spain & van Veld, 1983; Pritchard &

Bourquin, 1984).

Adaptation is thought to be controlled by the concentration of

the HOC interacting with the microflora, as well as the length

of time the chemical is in contact with the soil (Bosma et al.,

1997; Alexander, 2000; Macleod et al., 2001). For example,

Macleod & Semple (2002) investigated the development of

pyrene catabolic activity in two soils (pasture and woodland)

with disparate amounts of organic matter amended with

100mg pyrene kg�1. Pyrene mineralization was observed in

the pasture soil after 8weeks’ incubation, whereas it took

76weeks in the woodland soil. Degradative investigations on

the woodland soil showed that pyrene was bioavailable but

that the microbial community in the woodland soil could not

mineralize the pyrene. The observers thought the disparity in

catabolic activity was due to the slower transfer of pyrene from

the soil to the microorganisms in the woodland soil caused by

its larger organic matter content. The frequency of HOC

additions to the soil is also thought to be important in

determining the rate and extent of the catabolic activity

(Carmichael et al., 1997; Thompson et al., 1999).

Degradation: accessing soil-associated contaminants

Microbial catabolism is the principal mechanism for the

removal of contaminants, such as PAHs, from the soil. For

sparingly soluble contaminants, biodegradation is generally

slower than for more soluble contaminants, as the chemicals

will partition more readily with the solid phases of the soil

(Bosma et al., 1997). Microorganisms can utilize contaminants

in the liquid phase by direct contact of cells with the organic

contaminant, or with submicrometric particles dispersed in the

aqueous phase (Nakahara et al., 1977). Microbial interaction

with HOCs involves two processes (Bosma et al., 1997):

1 a physical or chemical component involving the movement

of the chemical in the physical environment, in relation to the

degrading microorganisms, and

2 a biological component involving the metabolism of the

chemical.

The relative importance of these mechanisms depends on

how strongly the contaminant is sequestered as well as the rate

of degradation. The rate at which a sequestered HOC becomes

available is influenced by the ability of the microorganisms to

reduce the concentration in the aqueous phase and the ten-

dency of the organisms to adhere to the sorbent (Calvillo &

Alexander, 1996). This is shown in Figure 5, where the possible

mechanisms of microbial attack are described, i.e. direct

contact or in the aqueous phase. Increased contact time

reduces the magnitude of the rapidly desorbing phase and

extent of biodegradation (Hatzinger & Alexander, 1995;

Pignatello & Xing, 1996; Cornelissen et al., 1998). Guerin &

Boyd (1992) and Calvillo & Alexander (1996) have shown that

the presence of degrading microorganisms alters the

desorption rates of contaminants from sorbed surfaces. The

change arises because the microorganisms utilize the contamin-

ants, which are readily available through the aqueous phase,

leading to the desorption of more contaminant from the solid

phase to the aqueous phase (Figure 5) (Bosma et al., 1997).

Techniques for appraising availability of HOCs in soil:

chemical assays

Mass transfer of a contaminant governs microbial bioavail-

ability (Bosma et al., 1997; Carmichael et al., 1997) and, in

particular, the size of the rapidly and slowly desorbable

fractions (Cornelissen et al., 1998). Uptake of contaminants is

far greater from fluid than from sorbed states (Ogram et al.,

1985), and, because HOCs are sparingly soluble in water and

strongly sorbed, at any one time only a small proportion of a

given contaminant will be present in the soil solution. Thus,

water is a poor choice of solvent for assessing bioavailability,

as a large labile pool of HOC will be present in the solid phase

of the soil (Reid et al., 2000a). The ideal extractant may there-

fore be one that can access the entire labile fraction of the

contaminant in the soil, perhaps mimicking the interactions

between microbes and the contaminant.



Conventional extraction involves organic solvents to remove

as much as possible of the contaminant. This does not give a

good representation of the fraction of HOCs available to the

soil biota. An alternative approach, pioneered by Hatzinger &

Alexander (1995), is to use mild organic solvents and extrac-

tion conditions. The hypotheses on which this approach is

based are (i) HOCs in soil are comprised of a ‘readily extract-

able fraction’ as well as a more strongly sequestered hydro-

carbon fraction and (ii) this readily extractable fraction is

more representative of bioavailability than the total contamin-

ant. Hatzinger & Alexander (1995) tested the extractability of

phenanthrene aged in a sterile soil. The amount of phenan-

threne mineralized on addition of a degrading inoculum

decreased with length of ageing; the amount of phenanthrene

extractable by butanol also decreased. Kelsey et al. (1997)

tested nine different combinations of mild solvent for

non-exhaustive extraction and concluded that butanol

(without and with shaking, respectively) was the most

appropriate solvent for predicting bioavailability to earth-

worms and a bacterial inoculum. However, non-exhaustive

solvent extraction has not been shown to be a reliable predic-

tor of bioavailable fractions, because of the complexity of

interactions between the physico-chemical properties of the

contaminant, the soil and the biota in question. To date, the

selection of mild organic solvent and extraction conditions has

been developed empirically, from observed correlations with

experimentally measured and defined bioavailability.

One of the most significant advances relates to the applica-

tion of solid-phase microextraction (SPME), avoiding the use

of solvents entirely. If a solid phase adsorbent is placed into

contact with a soil–water slurry, HOCs can diffuse out of the

soil and on to the adsorbent. This can be used as a rapid

and straightforward alternative to conventional extraction

(Eriksson et al., 1998) and may be used to assess bioavailability.

Using a suitable adsorbent, such as Tenax, will ensure that

the aqueous concentration of HOCs in a soil–water slurry will

be effectively zero (Yeom et al., 1996). The adsorbent maximizes

the diffusion gradient by acting as an infinite sink for HOCs:

exchangeable soil-sorbed HOCs thus transfer into the adsorbent.

Cornelissen et al. (1998) used Tenax beads for SPME to

measure the rapidly desorbing fractions of PAHs from

environmental matrices and concluded that the technique

mimicked the bioremediation capacity of an active population

of PAH-degrading microorganisms. The rapidly depleted

HOCs were renewed by desorption from the matrix. They

reported that bioremediation was slightly under-predicted by

desorption, and they suggested that the total amount bio-

degraded was equal to the readily exchangeable (rapidly

desorbing) fraction plus a small fraction of the strongly sorbed

contaminant. They demonstrated the merit of this approach

by the strong correlation obtained across a range of authenti-

cally contaminated samples. Ramos et al. (1998) reported that

freely dissolved concentrations of HOCs, and hence bioaccu-

mulation, could be measured by SPME. White et al. (1999)

used the same technique to demonstrate that changes in bio-

availability of phenanthrene (to earthworms and to two differ-

ent bacterial species) were strongly correlated with changes in

the proportion desorbed.

Attempts have been made to improve the validity of

non-exhaustive extraction by methods that more closely

mimic the transfer that occurs during biodegradation. In the

last few years, much research has been devoted to developing

methods for measuring the bioavailability of HOCs which per-

haps mimic microbial interactions with the HOCs (Figure 6).

Techniques include persulphate oxidation (Cuypers et al., 2000)

and cyclodextrin extraction (Reid et al., 2000b; Cuypers et al.,

2001). All of these techniques have shown strong correlations

between the microbially degradable fraction and the fraction

available to the chemical extractants. They have focused on

PAHs. The techniques vary in their complexity. The advantage

Degradation bydirect contact

DesorbedHOC

Degradation in theaqueous phase

HOC in soil

A B

Figure 5 Microbial attack on hydrophobic organic contaminants in soils: A represents direct contact and B represents degradation in the aqueous

phase. The direction and size of the arrows represent the interaction (sorption–desorption) between the contaminant and the soil and the pore water.



that the cyclodextrin extraction has over the solid-phase extrac-

tions and persulphate oxidation is its simplicity: it requires only

a simple shaking, and it is highly reproducible. However, little is

known of the applicability of these techniques to other organic

contaminants and soil biota, and this is a matter for future

research.

Techniques for appraising the availability of HOCs in

soil: biological assays

We clearly need biological methods to complement the trad-

itional chemical analysis approach so as to verify the concept

of bioavailability. Traditional methods of monitoring soil

microbes are based on measuring changes in the microbial

biomass, the numbers of organisms, enzyme activity and the

response of key soil processes (Macleod et al., 2001). More

recently, community-based methods such as phospholipid

fatty acid analysis and Biolog have been applied, further

widening the scope of tools available for study of the microbial

ecology of the soil. In general, these techniques have proved

insensitive to HOCs, and the response of the assays often

reflects effects such as incubation and soil physical characteris-

tics rather than the contaminant (Bundy et al., 2001). As a

consequence methods that focus on specific catabolic processes

are likely to yield the most relevant information.

Respirometry

Microbial respiration can be used to quantify the impact of a

contaminant on microorganisms in the soil or to measure the

catabolism of a contaminant to CO2. The latter, often termed

mineralization, is routinely used to assess the microbially available

fraction of contaminants in soil (Hatzinger & Alexander, 1995;

White et al., 1997; Macleod & Semple, 2000; Reid et al., 2000b).

The use of 14C-labelled substrates is necessary to trace the fate of

added organic compounds in the soil and allows complete mass

balances to be calculated. This, combined with the sensitivity

of radiometric methods, has led to the widespread use of14C-labelled compounds in studies investigating the fate of xeno-

biotics. By measuring the biodegradation of a 14C-labelled HOC

to 14CO2, the catabolic potential of the soil microbial community

can be determined (Reid et al., 2001). Further, the impact on

microbial catabolic activity depends on a contaminant’s bioavail-

ability in the soil (Reid et al., 2000a). The use of 14C-labelled

substrates allows the fate of an organic contaminant and its

bioavailability to be followed successfully, even where complete

destruction of the original carbon skeleton occurs (Hatzinger &

Alexander, 1995; Reid et al., 2000b).

Several devices have been developed to follow the release of14CO2 from 14C-labelled substrates in static and flow-through

systems. Static systems are widely used and have the advantages

of simplicity, low cost and few uncertainties concerning flow

rates, leakage and sorption of the 14C-labelled materials. Simple

HOC in soil

DesorbedHOC

Interacting cell

A

B

Chemical mimic

Figure 6 Theoretical mechanisms for the biodegradation of phenanthrene (A) and a putative chemical mimic, allowing the determination of the

extent of biodegradation (B).



respirometric systems are effective in quantifying the rates and

extents of HOC mineralization in soils (Reid et al., 2001).

Molecular probes

The impacts of HOCs on microbial communities have been

studied at the genetic level by applying techniques capable of

assessing the frequency and distribution of specific degradation

genes. Probes have been developed for genes from pathways for

aliphatic, monoaromatic and PAH degradation (Stapleton et al.,

1998). These probes can be applied to DNA, mRNA or rRNA.

An initial stage of preliminary polymerase chain reaction (PCR)

is usually done to enhance the sensitivity of detection (Barkay

et al., 1995). Although the probes can be applied to environ-

mental isolates, a more valuable approach is to extract nucleic

acids directly from the soil, thereby avoiding the bias associated

with culturing (Sayler, 1991; Atlas et al., 1992). It is widely

acknowledged that mRNA transcripts are short-lived in soil,

and so measurement of in situ mRNA can provide an estimate

of highly specific degradative activity (Wilson et al., 1999).

These techniques have been widely applied, which demonstrates

the merits of this approach. Stapleton & Sayler (1998) used six

gene probes taken from aerobic hydrocarbon degradation

pathways and two associated with methanogens to assess an

aquifer contaminated with jet fuel. Almost all of the samples

gave a positive response for all genes, and, correspondingly,14C-labelled hydrocarbons added to the samples were rapidly

degraded.

Existing gene probes, however, cannot completely charac-

terize the ecological effects of the HOCs, and Ahn et al. (1999)

and Lloyd-Jones et al. (1999) have reported that not all PAH

degraders isolated from contaminated soils could be hybridized

with existing probes. Stapleton et al. (1998) reported that DNA

from a very acid environment (pH< 2) did not hybridize at all

with standard genes, although biochemical techniques verified

that hydrocarbons were being degraded. To recognize the value

of these technologies fully, new gene probes are required to

assess the degradation of PAHs (Ahn et al., 1999).

Microbial monitoring at the species level

One method that is well suited for interrogating the microbial

response at the genetic or species level is the use of reporter

genes to produce biosensors because these permit rapid, cheap

and sensitive biomonitoring, and they can be selected for their

environmental relevance (Paton, 2001). A biosensor can be

defined as a receptor (biological unit, e.g. enzyme, whole cell,

tissue) linked to a transducer mechanism (de la Guardia,

1995). The most suitable reporters for microbial systems are

reporter genes designed to enable rapid quantification of the

target product. The most widely used is the lacZ gene, which

encodes �-galactosidase, antibiotic resistance, catechol-2,3-

oxygenase and �-glucuronidase (Atlas et al., 1992). Biolumin-

escence, based on the lux genes, is a particularly useful

reporter mechanism because it is very sensitive and is rapid

enough to allow real-time monitoring, as well as permitting

non-destructive, in situ measurement (Stewart, 1990). Biolumin-

escence-based biosensors can be considered to be as least as

widely useful and applicable as lacZ systems (Atlas et al., 1992).

Microbes can be marked with the lux genes fused to specific

HOC degradation genes, allowing in situ monitoring of gene

expression (Barkay et al., 1995). For example, Sanseverino et al.

(1993) used Pseudomonas fluorescens HK44 (fused with the nap

genes) to characterize seven contaminated soils (by manufactured

gas plant and by creosote). A later study by Burlage et al. (1994)

showed that this sensor could detect middle-range refined oil

contamination in soils. More recently, Ripp et al. (1999) used the

sensor for in situ, on-line monitoring of naphthalene bioremedia-

tion. Escherichia coli DH5� (pGEc74, pJAMA7) (induced by

octane) responded to heating-oil that was contaminating ground-

water with concentrations too small to permit characterization by

chemical analysis (Sticher et al., 1997). Willardson et al. (1998)

showed that a similar catabolic biosensor designed to luminesce

in the presence of toluene responded to BTEX contaminants in

well water and soil.

Recently, Bundy et al. (2001) compared three catabolic-

based luminescence biosensors and observed that the induc-

tion of the biosensors varied greatly according to the nature of

the HOC and the extraction technique. They concluded that in

combination the biosensors offered useful information in the

prediction of bioremediation over time. Another reporter gene

system, gfp, which encodes for green fluorescent protein

(GFP), has also been used but to a lesser extent. This protein

also permits highly sensitive detection, to the single cell level,

and can be used for constitutive and specific marking. The

combination of gfp- and lux-based systems has great potential,

as the two systems report on complementary aspects of micro-

bial performance. For example, Unge et al. (1999) used a dual

gfp- and lux-marked bacterium simultaneously to monitor

both the cell numbers (by GFP) and the metabolic activity

(by bioluminescence) of a bacterial population.

Conclusions

The total burden of organic xenobiotic compounds present

cannot be fully quantified in contaminated soil because we

cannot extract all of the chemicals concerned. To predict the

behaviour of contaminants in the soil, the mechanisms of

interaction between the soil, the contaminant and the biota

must be understood. Further, it is far from certain that the

total concentrations of contaminant, which are adopted by

regulatory organizations, are what are really needed. In the

last 3 years, studies have shown that bioavailability can be

quantified chemically. However, there is no all-encompassing

extractant to describe bioavailability, as it differs between the

types of biota. The class of contaminant and the chemical

techniques used are likely to be important when determining



bioavailability. With so much uncertainty, it is clear that more

research is required to determine bioavailability and its

quantification in an environment as complex as soil.

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