Journal of Applied Microbiology ISSN 1364-5072
REVIEW ARTICLE
Microbe-aliphatic hydrocarbon interactions in soil:
implications for biodegradation and bioremediation
J.L. Stroud1, G.I. Paton2,3 and K.T. Semple1
1 Department of Environmental Science, Faculty of Science and Technology, Lancaster University, Lancaster, UK
2 School of Biological Sciences, University of Aberdeen, Aberdeen, UK
3 Remedios Ltd, Aberdeen Science and Technology Park, Aberdeen, UK
Keywords
bioaccessibility, bioavailability, biodegradation,
contaminated land, organic contaminants.
Correspondence
K.T. Semple, Department of Environmental
Science, Faculty of Science and Technology,
Lancaster University, Lancaster, LA1 4YQ, UK.
E-mail: k.semple@lancaster.ac.uk
2006 ⁄ 1369: received 29 September 2006,
revised 14 March 2007 and accepted 20
March 2007
doi:10.1111/j.1365-2672.2007.03401.x
Summary
Aliphatic hydrocarbons make up a substantial portion of organic contamination in the terrestrial environment. However, most studies have focussed on
the fate and behaviour of aromatic contaminants in soil. Despite structural differences between aromatic and aliphatic hydrocarbons, both classes of contaminants are subject to physicochemical processes, which can affect the degree of
loss, sequestration and interaction with soil microflora. Given the nature of
hydrocarbon contamination of soils and the importance of bioremediation
strategies, understanding the fate and behaviour of aliphatic hydrocarbons is
imperative, particularly microbe–contaminant interactions. Biodegradation by
microbes is the key removal process of hydrocarbons in soils, which is controlled by hydrocarbon physicochemistry, environmental conditions, bioavailability and the presence of catabolically active microbes. Therefore, the aims of
this review are (i) to consider the physicochemical properties of aliphatic
hydrocarbons and highlight mechanisms controlling their fate and behaviour in
soil; (ii) to discuss the bioavailability and bioaccessibility of aliphatic hydrocarbons in soil, with particular attention being paid to biodegradation, and (iii) to
briefly consider bioremediation techniques that may be applied to remove aliphatic hydrocarbons from soil.
Introduction
Hydrocarbons in the environment
Hydrocarbon pollution is ubiquitous in the environment
and accounts for over 15% of all pollution incidents in
England and Wales (Environment Agency 2006). Oil contamination is reported to be a common pollution incident with almost nine incidents per day in 2005
(Environment Agency 2006). Details of actual contamination of the terrestrial environment by oil are difficult to
quantify because of the unintentional nature of contamination (largely through accidental spillage). However, it is
estimated that over one million tonnes of oil are spilled
into UK terrestrial ecosystems every year (Ripley et al.
2002). This is a significant problem receiving considerable
political and scientific interest, and the Environment
Agency of England and Wales had recently published a
report investigating the risk from petroleum hydrocarbons in soils (Askari and Pollard 2005).
Contaminated land is costly to clean up; for example, in
the USA, the cost is expected to exceed US $1 trillion
(Maier et al. 2000), where 90% of the sites undergoing
remediation are linked to petroleum hydrocarbons (Cole
1994). Bioremediation is a cost-effective strategy, and it is
reported that approximately 25% of all petroleum-contaminated land is being cleaned up using natural attenuation
strategies (Holden et al. 2002). Thus, in order to predict
how successful bioremediation will be, research into the
fate and behaviour of hydrocarbons in soil is important.
To date, research has centered on aromatic contaminants, particularly polycyclic aromatic hydrocarbons
(PAHs), as these contaminants are the current risk drivers
for land remediation (Askari and Pollard 2005). However,
ª 2007 The Authors
Journal compilation ª 2007 The Society for Applied Microbiology, Journal of Applied Microbiology 102 (2007) 1239–1253
1239
J.L. Stroud et al.
aliphatic hydrocarbons are significant contaminants; oil
pollution in the UK is dominated by diesel, with petrol
and diesel containing up to 90% of aliphatic hydrocarbons by volume dominated by C14-C20 alkanes. However,
despite aromatic contaminants receiving the greatest
research attention, this class of chemical only equates to
less than 5% by volume (Block et al. 1991).
their fate and behaviour in soil; (ii) to discuss the bioavailability and bioaccessibility of aliphatic hydrocarbons
in soil, with particular attention being paid to biodegradation, and (iii) to briefly consider bioremediation techniques that may be applied to remove aliphatic
hydrocarbons from soil.
Aliphatic hydrocarbons in soil
Aims
Given the nature of hydrocarbon contamination of soils
and the importance of bioremediation strategies, understanding the fate and behaviour of aliphatic hydrocarbons
is imperative, particularly microbe–contaminant interactions. Biodegradation by microbes is the key removal process of hydrocarbons in soils, which is controlled by
hydrocarbon physicochemistry, environmental conditions,
bioavailability and the presence of catabolically active
microbes. Therefore, the aims of this review are (i) to
consider the physicochemical properties of aliphatic
hydrocarbons and highlight the mechanisms controlling
Physicochemical properties of aliphatic hydrocarbons
Aliphatic hydrocarbons are defined as open-chain
methane derivatives, which are both non-aromatic and
non-cyclic organic compounds, containing carbon and
hydrogen. Aliphatic hydrocarbons can be subdivided into
three structurally different groups: (i) alkanes – saturated
hydrocarbons with single C-C bonds; (ii) alkenes – unsaturated hydrocarbons containing double C=C bonding;
and (iii) alkynes – unsaturated hydrocarbons containing a
triple C ” C bond. Table 1 shows the members of the aliphatic groups and their properties.
Table 1 Physicochemical properties of selected hydrocarbons (Verschueren 1983; Howard and Meyln 1997)
Hydrocarbon group
Name
Formula
Molecular
weight
(g mol)1)
Aliphatic
Alkane
Tetradecane
C14H30
198Æ38
5Æ5
253
0Æ000 282
7Æ2
MODEL
Alkane
Hexadecane
C16H34
226Æ44
18
287
0Æ0009
9Æ1
Alkene
Hexadecene
C16H32
224Æ43
3–5
274
N⁄A
N⁄A
Alkyne
Hexadecyne
C16H30
222Æ42
15
148
N⁄A
N⁄A
PAH
Naphthalene
C10H8
128Æ18
79–83
217Æ9
30
3Æ36
MODEL
PAH
Phenanthrene
C14H10
178Æ22
97–101
340
1Æ1
4Æ16
PAH
Pyrene
C16H10
202Æ6
156
404
0Æ135
5Æ19
PAH
Benzo[a]Pyrene
C20H12
252Æ31
175–179
495
0Æ0038
6Æ06
Aromatic
Structure
Melting
point
(C)
Boiling
point
(C)
Solubility
(mg l)1)
log
Kow
N ⁄ A, Data not available.
1240
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Bioavailability of hydrocarbons in soil
Mid-length (C14-C20) alkanes are non-polar, virtually
water insoluble hydrocarbons with increasing melting and
boiling points as carbon number increases within the
molecule. Typically, these alkanes have low aqueous solubilities; for example, hexadecane has a water solubility of
0Æ9 lg l)1 and is in a liquid state at room temperature.
Collectively, these physicochemical properties mean that
mid-length aliphatic contaminants are not readily volatilized or leached from soil. Despite this, hydrocarbons may
be subjected to a number of physical, chemical and biological processes in soil, with microbial degradation representing the major loss process. Further, it is well
established that interactions between hydrophobic organic
contaminants (HOCs), of which most studies have focussed on aromatic hydrocarbons, may also be important
in controlling the fate and behaviour in soil, involving
interactions with the mineral and organic fractions, which
may result in a reduction in biological and chemical
availability and allowing these contaminants to persist in
soil. A comparison between the properties of aliphatic
and aromatic hydrocarbons is shown on Table 1. Hexadecane is a very hydrophobic hydrocarbon, being three
orders of magnitude more insoluble and having a higher
octanol-water partition coefficient than phenanthrene.
Further, hydrophobicity has been determined to be a crucial property in controlling the organic contaminant
behaviour in soil, affecting sequestration and both chemical and biological availability (Reid et al. 2000a). Due to
the limited research on aliphatic sequestration and bioavailability, their behaviour is discussed in relation to
PAHs in this review, where appropriate.
Bioavailability of hydrocarbons in soil
as key components with sorption described as reversible
(Pan et al. 2006), and (ii) hard carbon (glassy) defined as
rigid, condensed structures with humin (Pan et al. 2006),
kerogen (Cornelissen et al. 2005) and pyrogenic carbons
(Cornelissen et al. 2005) as commonly identified components. Sorption of hydrocarbons within the glassy region
is characterized by irreversible sequestration (Xing and
Pignatello 1996).
The extent to which a chemical partitions into the
organic matter is described by Koc, but may also be described by Kow (Table 1); aliphatic hydrocarbons can
strongly partition into organic matter and diffuse into the
three-dimensional structure of the organic matter. Hydrocarbons may be sequestered within the soil through sorption to organic matter and mineral fractions and ⁄ or
diffuse into the three-dimensional structure of the soil
(Fig. 1); the degree to which these physical interactions
occur increases with time, and has been termed ‘ageing’
(Hatzinger and Alexander 2005). It has been well documented that as soil–contaminant contact time increases,
there are commensurate decreases in mild chemical
extraction and biodegradation (Semple et al. 2003). Interactions between soil components and chemical availability
and biological interactions have mainly focussed on aromatic hydrocarbons, and have only been reported for aliphatic hydrocarbons in a few studies. For example,
Noordman et al. (2002) reported that mass transport
limitations were important after investigating the biodegradation of hexadecane in a variety of matrix sizes. The
smaller the pore size, the lower the extent of mineralization; the smallest pore size of 6 nm (Silica 60) led to the
lowest extent of degradation.
Interactions between soil and aliphatic hydrocarbons
Before considering the interactions between soil microorganisms and aliphatic hydrocarbons, putative interactions between soil and these organic contaminants must
be briefly considered, as it is well known that soil–contaminant interactions can affect microbial degradation.
Soil is composed of inorganic and organic components
separated by pores containing air or water. The interactions between hydrocarbons and mineral surfaces (clay,
silt and sand) are only significant when the organic matter content is <0Æ1% (Schwarzenbach and Westall 1981).
Thus, organic matter is very important in the fate and
behaviour of organic contaminants, including aliphatic
hydrocarbons, in soil. Such mechanisms of interaction
between organic matter and organic contaminants are not
the major focus of this review and have been reviewed
elsewhere (Pignatello and Xing 1996; Cornelissen et al.
1998). In summary, organic matter can be divided into
two distinct phases: (i) soft carbon (rubbery) is defined as
expanded, flexible structures with humic and fulvic acids
Interactions between microflora and
hydrocarbons in soil
Bioavailability and bioaccessibility – definition and
measurement
Understanding the term ‘bioavailability’ is complicated
because of a number of interpretations in published literature. In an attempt to clarify this matter, Semple
et al. (2004) proposed two linked definitions, bioavailability and bioaccessibility. A bioavailable compound is
defined as ‘a compound which is freely available to cross
an organism’s membrane from the medium the organism inhabits at a given point in time’. A bioaccessible
compound is described as ‘a compound which is available to cross an organisms’ membrane from the environment it inhabits, if the organism has access to it;
however, it may either be physically removed from the
organism, or only bioavailable after a period of time’
(Semple et al. 2004). In considering the methods used
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J.L. Stroud et al.
J.L. Stroud et al.
NAPL
fraction
Surface
sorption
Water soluble
fraction
Diffusion into glassy
organic matter
Mineral
fraction
Surface
sorption
Microbes
Diffusion
into pores
Diffusion into rubbery
organic matter
Figure 1 Possible interactions between soil matrices and aliphatic hydrocarbons.
to determine microbial degradation as well as the use of
chemical extractions to predict microbial degradation in
published literature, it is reasonable to presume that the
latter definition (bioaccessibility) is more likely to be
measured both biologically and chemically. For the purposes of this review, bioaccessibility will relate to microbial degradation.
Measuring bioaccessibility is difficult, as this will be
controlled by the nature of interaction between the
organism and the chemical, and the physical location of
the chemical in relation to the organism and likelihood of
desorption from soil particles with time (Semple et al.
2003, 2004). Further, the methods used to chemically or
biologically determine microbe–contaminant interactions
will also constrain the measurements. By way of explanation, Reid et al. (2000a) and Stokes et al. (2005)
comprehensively reviewed the methods for assessing bioavailability with biological and chemical techniques. The
reviews highlighted that biodegradation assays rely on soil
slurries and solubilization target contaminants, which
gives an estimation of bioaccessibility rather than bioavailability. For example, 14C-hydrocarbon respirometry
(Reid et al. 2001) involved mixing soil with liquid media,
forming a slurry. Recently, it has been shown that this
puts considerable bias towards a more planktonic aqueous phase biodegradation, reducing sorbed phase biodegradation (Woo et al. 2004).
1242
The chemical determination of bioaccessibility has
attracted considerable attention, with comparisons
between microbial degradation assays and various chemical extractions being the most widely investigated (Semple et al. 2003; Stokes et al. 2005). In comparison to
microbial degradation, mild organic solvent extractions
have been reported extensively, with some success (Kelsey
et al. 1997; Liste and Alexander 2002), whereas other
authors have reported no correlation (Macleod and Semple 2000; Macleod and Semple 2003). More promisingly,
Tenax, XAD-4 and cyclodextrin extractions have been
more successful as they rely on desorbing fractions of
HOCs, with good correlations with microbial degradation
mainly for aromatic hydrocarbons (Cornelissen et al.
1997; Reid et al. 2000b; Dew et al. 2005; Stokes et al.
2005; Doick et al. 2006) However, only a few studies have
investigated this for aliphatic hydrocarbons, showing that
the desorbed fraction is less than that of the microbially
degraded fraction, indicating that contaminants can be
degraded without prior desorption to the aqueous phase
(Huesemann et al. 2003, 2004). Thus, can bioaccessibility
be adequately measured for aliphatic hydrocarbons using
these extraction techniques described for aromatic hydrocarbons?
To some extent, the use of biomarkers in the field of
environmental forensics has enabled a thorough technique
to connect genuine environmental behaviour (sequestra-
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Bioavailability of hydrocarbons in soil
tion, degradation and ageing) to bioavailability. For
example, Wang et al. (2006) reviewed this area extensively
and related the nature of refined and crude oils to certain
biomarker chemicals. For example, terpanes and steranes
(multi-ringed cycloalkanes found in crude oils) of varying
sizes are widely used because of their recalcitrant nature,
and their concentration can be used to identify both the
source of the hydrocarbon pollution and the relative time
since release. The application of this approach is widely
acknowledged for crude oils. Such petroleum biomarkers
are widely used in the investigation of hydrocarbon spills
(Page et al. 1988; Barakat et al. 1999; Zakaria et al. 2001).
However, most of these examples are concerned with
marine incidents with none reported in the terrestrial
environment. These data enabled the users to establish
the actual origin of the material and the likely discharge
volume that had occurred. Refined petroleum fuels are
obtained from crude oil through refining processes, and
thus the chemical composition of this feedstock, the refinery techniques and conditions and the intended products determine the relationship between the crude and
the refined material. In the case of lubrication oils,
Speight (1999) points out that biomarkers are located in
the high carbon number end and these products have
more markers than most crude oils and petroleum products. In oils and petroleum products, aromatic steranes
are additional useful compounds for biomarking on
account of their resistance to biodegradation.
Biodegradation of hydrocarbons in soil
The initial step in the aerobic degradation of saturated, aliphatic (n-alkanes) compounds involves the oxygenase
enzyme ‘attacking’ the terminal methyl group where a primary alcohol is formed (Sepic et al. 1995; Koma et al.
2001; van Hamme et al. 2003). The alcohol is further oxidized to the corresponding aldehyde and fatty acid. The
fatty acid is oxidized by cytoplasmic b-oxidation enzymes
to tricarboxylic acid (TCA) (van Hamme et al. 2003).
Studies on the oxidation of alkenes have been limited
to the degradation of terminal alkenes (1-alkenes) with
little attention to the internal alkene oxidation (Morgan
and Watkinson 1994). The double bond position is a factor affecting alkene degradability as 1-alkenes are more
easily degradable than alkenes with an internal double
bond. Based upon the initial attack of aerobic oxidation,
which can occur at either the double bond or the saturated chain end (i.e. methyl group), four main products
can be recognized. The oxidation of the methyl group can
form either (i) alkenol or fatty acids or (ii) primary or
secondary alcohols or methyl ketones, whereas the oxidation of the double bond produces either (i) an epoxide or
(ii) a diol (Britton 1984; Morgan and Watkinson 1994).
Bioavailability of hydrocarbons in soil
Although a number of studies (e.g., Chayabutra and Ju
2000; Grishchenkov et al. 2000; van Hamme et al. 2003)
have illustrated that n-alkanes and branched alkanes can
be degraded under anaerobic conditions, the anaerobic
pathway of aliphatic hydrocarbon metabolism has not
been described in great detail. A number of studies have
suggested that the degradation of n-hexane can be used as
a model for the degradation pathway of n-alkanes by a
range of anaerobic micro-organisms (Heider et al. 1999;
Widdel and Rabus 2001; Wilkes et al. 2003). The initial
step of this pathway is by the activation of n-alkane at C2
with fumarate to form (1-methylalkyl) succinate, which is
further activated by HSCoA (coenzyme A, CoA) to yield
(1-methylalkyl) succinyl-CoA (Heider et al. 1999; Wilkes
et al. 2003). The latter compound goes though carbon
skeleton rearrangement to form (2-methylalkyl) malonylCoA, which is then decarboxylated to 4-methylalkanoylCoA and then oxidized by b-oxidation to produce CO2
as the end product (Wilkes et al. 2003).
The biodegradable fraction of organic contaminants in
soil is defined throughout the literature as the fraction
that may be easily desorbed to or is desorbed from soil
and present in the aqueous phase (Alexander 2000; Reid
et al. 2000a; Semple et al. 2003). Whilst older microbiological studies show that microbes are only able to utilize
dissolved HOCs (Wodzinski and Coyle 1974; Ogram
et al. 1985), those studies had limitations as it is now
known that common bacterial culturing techniques may
have led to the isolation of degraders that cannot utilize
hydrocarbons sorbed to soil. Micro-organisms capable of
sorbed-hydrocarbon biodegradation require a sorbed
form of growth substrate (Tang et al. 1998). Thus, more
recent studies comparing the biodegradation of organic
contaminants, which are present as dissolved or sorbed
substrates, have shown that sorption is not limiting to the
biodegradation of either aromatic (Laor et al. 1996) or
aliphatic hydrocarbons (Holden et al. 2002). Indeed, Pignatello et al. (1983) hypothesized that the biodegradation
of HOCs occurs at the soil surface rather in the aqueous
phase. Additionally, much our understanding of this
phenomenon comes from the measured response of the
relatively water-soluble herbicides such as 2,4-dichlorophenolxyacetic acid (2, 4-D) (Ogram et al. 1985) or benzylamine (Miller and Alexander 1991), which have water
solubilities of 900 mg l)1 and 1 kg l)1, respectively (Howard and Meyln 1997). However, a study by Park et al.
(2001) showed that bacteria are able to utilize both
sorbed and nonsorbed 2,4-D through rate calculations
based on the liquid-phase concentration of 2,4-D. This is,
therefore, in (Askari and Pollard 2005) direct conflict to
the original study by Ogram et al. (1985), who used
mathematical models based on laboratory data to show
that sorbed 2,4-D was unavailable and not degraded by
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J.L. Stroud et al.
micro-organisms, and only the solution phase 2,4-D was
degraded. Additionally, the chemicals 2,4-D and benzylamine are up to 1Æ11 · 109 more water-soluble than
hexadecane, and up to 9Æ09 · 105 times more watersoluble than phenanthrene (Table 1).
Hydrocarbons are considered to be poorly water-soluble
organic contaminants; yet this is relative; for example,
naphthalene and phenanthrene are 33 000 and over 1200
times more water-soluble than hexadecane. This raises
several issues when comparing the bioaccessibility and biodegradation of aromatic and aliphatic hydrocarbons in soil.
i Is the emphasis on the aqueous phase and passive
uptake actually pertinent to the very low water soluble
hydrocarbons?
ii Does the readily desorbed fraction adequately describe
the size of a bioaccessible fraction?
iii Are there different modes of biodegradation?
These questions are addressed in Fig. 2, in which (a)
illustrates the simple mass transfer and mass transport of
a relatively water-soluble chemical, benzylamine, which is
not limited by low availability; (b) illustrates the mass
transport limitations of a poorly water-soluble hydrocarbon (naphthalene); (c) shows that the biodegradation rate
of phenanthrene is controlled by the rate of dissolution,
and (d and e) illustrate the specialized mechanisms that
bacteria have to degrade HOCs, such as hexadecane,
enabling them to overcome the low contaminant bioaccessibility.
The degree of biodegradation of aliphatic hydrocarbons
is typically lower than their aromatic counterparts. For
example, Huesemann et al. (2004) carried out a biodegradation study using hydrocarbon contaminated soils
inoculated with degraders adapted to use Minas crude oil
in slurry reactors. The results showed that, in the Belhaven soil that had an initial 1Æ6% hydrocarbon contamination, less than 20% of the initial concentration of
hexadecane remained after 27 d of remediation. The PAH
phenanthrene in the same time period had decreased to
less than 5%, showing the higher residual fractions of an
aliphatic hydrocarbon. Further, a study by Chaineau et al.
(1995) investigated the biodegradation of drill cuttings,
showing that the concentration of saturated hydrocarbons decreased from 1350 mg TPH kg)1 to about
400 mg TPH kg)1. The aromatic fraction decreased from
an initial 600 mg TPH kg)1 to about 200 mg TPH kg)1.
Thus, whilst the percentage of biodegradation was similar,
double the concentration of saturated hydrocarbons resided in soil after biodegradation. A key study illustrating
the degradability of hydrocarbons was carried out by
Loser et al. (1999) in which a microporous, sandy soil
was spiked with 0Æ3% hexadecane or phenanthrene, and
biodegradation was measured to study a pilot scale percolator. The biodegradation experiment was carried out in
1244
J.L. Stroud et al.
inoculated shake flasks. The degree of biodegradation of
hexadecane was 80%, highly consistent with the degree of
biodegradation of the aliphatic fraction found in remediation studies, equating to a residual fraction of
600 mg kg)1. However, the degree of phenanthrene biodegradation was significantly higher (96%) with a residual
concentration of only 100 mg kg)1. A range of phenanthrene concentrations was tested, which showed different
percentages of biodegradation; nonetheless, a residual
concentration
was
always
detected
at
about
100 mg TPH kg)1. Thus, the aliphatic hydrocarbon had a
significantly higher concentration remaining in the soil
after biodegradation as compared with phenanthrene,
under the same experimental conditions. These results
indicate that aliphatic hydrocarbons may be constrained
by factors affecting bioaccessibility to a greater extent
than PAHs. Although this aliphatic component may be
degradable, degradation in soils does not necessarily
occur, and can therefore be considered as an important
contaminant fraction.
The physicochemical properties of the aliphatic hydrocarbon hexadecane show that it is very hydrophobic, and
so it extensively partitions into the solid phase. For example, a study by Watts and Stanton (1999) investigated the
mineralization of hexadecane sorbed to silica sand in slurries, and showed that 0Æ36% of hexadecane was present in
the aqueous phase and that the sorbed phase was not
readily desorbed to the aqueous phase. For example, Huesemann et al. (2003, 2004) reported minimal desorption
to the aqueous phase by aliphatic hydrocarbons with an
XAD-2 extraction, suggesting that there was a minimal
accessible fraction available for biodegradation. However,
this does not adequately describe the biodegradable fraction, as aliphatic hydrocarbons are widely reported to be
extensively biodegraded in studies in both the laboratory
and environment (Huesemann 1995; Gogoi et al. 2003;
Huesemann et al. 2003, 2004). Biodegradation of aliphatic
hydrocarbons has been shown to significantly exceed desorption to the aqueous phase (Thomas et al. 1986; Huesemann et al. 2003, 2004).
Liquid culture experiments as well as NAPL-soil investigations have widely reported that the dissolution of
HOCs to the aqueous phase cannot account for the biodegradation rate and postulate the presence of specialized
uptake mechanisms (Thomas et al. 1986; Goswami and
Singh 1991; Huesemann 1995; Huesemann et al. 2003;
Rapp and Gabriel-Jurgens 2003; Huesemann et al. 2004).
There are a wide range of mechanisms that lead to the
biodegradation of hydrocarbons, and are well described
by Johnsen and Karlson (2004). Two of the most important mechanisms, solubilization and direct contact, would
result in the degradation of the poorly water-soluble
hydrocarbons.
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Bioavailability of hydrocarbons in soil
Bioavailability of hydrocarbons in soil
(a) e.g. Benzylamine (1 kg l –1) at a concentration where water solubility is not exceeded
Mass transfer
Solubilised substrate (water
solubility not exceeded)
Simple diffusion into the
cell
Mass transport
–1
(b) e.g. Naphthalene (30 m gl ) at a concentration where water solubility is not exceeded
Mass transfer
Solubilised substrate
(water solubility not
exceeded)
Simple diffusion into the
cell
Minimal aqueous phase
presence, therefore mass
transport limitations
Mass transport
–1
(c) e.g. Phenanthrene (1·1 mg l ) at a typical concentration (lower solubility)
Mass transfer
Simple diffusion into the cell
Minimal aqueous phase
presence, therefore mass
transport limitations
Rate of uptake dependent on
dissolution to aqueous phase
(d) e.g. Hexadecane (0·0009 mg l –1) at a typical concentration (lower solubility)
Mass transfer
Bacterium produced surfactant
– enhances solubility
(e)
Simple diffusion into the cell
Minimal aqueous phase presence,
therefore mass transport limitations
Rate of uptake dependent on
dissolution to aqueous phase
Microbes have adaptations to
overcome limitations
e.g. Hexadecane (0·0009 mg l –1) at a typical concentration (lower solubility)
Hydrophobic
exterior/
biosurfactant
production to
allow direct
contact
Mass transfer
Biosurfactants also mediate
the rate of uptake and
movement into cell
Biosurfactant mediated
uptake into cell to minimise
mass transfer limitations
Direct contact, therefore
minimal mass transport
limitations
Figure 2 Range of possible bacterial uptake mechanisms for hydrocarbons in soil.
Solubilization. It involves the production of biosurfactants by microbes, which increase the concentration of
hydrocarbons in the aqueous phase (illustrated in
Fig. 2c). The solubilization of hydrocarbons by biosurfactants is widely reported (Goswami and Singh 1991),
where higher concentrations of hydrocarbons were found
in the aqueous phase than was expected. Bouchez-Naitali
et al. (1999) also noted the importance of solubilization
while examining the biodegradation of hexadecane by a
variety of strains of bacteria. Bai et al. (1997) showed that
the solubility of hexadecane in a 500 mg l)1 rhamnolipid
solution was 19 mg l)1, showing that the increase of
hydrocarbon concentrations in the aqueous phase was
significant (illustrated in Fig. 2d). Whyte et al. (1999b)
reported that invagination of hydrocarbons occurred,
where inclusions of hydrocarbons in cells formed, followed by the uptake. Noordman et al. (2002) reported
that the role of rhamnolipids was to mediate the mass
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J.L. Stroud et al.
transfer of hexadecane into cells (illustrated in Fig. 2e),
causing biodegradation.
Direct contact. In direct contact, the bacterial cells
adhere to the surface of the hydrocarbon. This is key to
biodegradation as shown by Holden et al. (2002), where
direct contact was crucial to the degradation of hexadecane by the bacterium under investigation (illustrated in
Fig. 2e). Direct contact can be facilitated by biosurfactants
and bioemulsifiers produced by the cells that enhance
adhesion between the cell wall and the accessible hydrocarbon (Bouchez-Naitali et al. 1999). For example, the
Gram-negative bacterium Acinetobacter spp. (Margesin
et al. 2003) is widely reported to produce biosurfactants ⁄ bioemulsifiers; thus, it has a hydrophobic exterior
to allow cellular contact with the hydrocarbon. Additionally, some bacteria naturally have hydrophobic cell
surfaces enabling cellular adhesion to hydrocarbons
(Bouchez-Naitali et al. 1999; Whyte et al. 1999b).
Bacteria capable of utilizing sorbed contaminants are
widely reported for a range of hydrocarbons, such as
naphthalene (Guerin and Boyd 1992) and phenanthrene
(Laor et al. 1996; Tang et al. 1998). Deziel et al. (1996)
showed in a microbiological study based on biodegradation of PAHs that over 40% microbes produced biosurfactants. Thus, whilst the literature has a strong bias
towards biodegradation requiring abiotic desorption of
HOCs to the aqueous phase prior to biodegradation, a
wide range of sorbed hydrocarbons are available to the
degraders (Table 2). However, the role of biosurfactants
is not always clear; for example, biosurfactants are also
produced by bacteria when only a solubilized contaminant is available (Bouchez-Naitali et al. 1999). Furthermore, the biodegradation extent by bacteria that do not
J.L. Stroud et al.
produce biosurfactants is not enhanced by the presence of
biosurfactants, indicating that the role of biosurfactants is
species specific (Noordman and Janssen 2002). Additionally, the ability to produce surface active agents does not
necessarily correspond to mineralization of PAHs (Willumsen and Karlson 1997).
Clearly, more research is required to investigate the
role of biosurfactants on hydrocarbon biodegradation in
soils. This is particularly important as biodegradation is
understood to be the principal mechanism for the
removal of aliphatic hydrocarbons from soil and essential
for the successful bioremediation of contaminated sites.
Bioremediation of hydrocarbons in soil
Soils, considered suitable for bioremediation, are typically
contaminated with hydrocarbons at levels of 0Æ2–55% by
volume of oil concentration (Huesemann et al. 2002).
Targets for the remediation of hydrocarbon-contaminated
land are currently based upon the total petroleum hydrocarbon (TPH) content (Gogoi et al. 2003), which measures hydrocarbons within the range of C6-C40, although
in the UK this would be determined by a risk-based
approach based on future intended land use (Askari and
Pollard 2005).
Bioremediation relies on the breakdown of target pollutant compounds in the soil through microbial degradation, by using either in situ (including strategies involving
soil amendment, bioventing ⁄ biosparging, natural attenuation and phyto- ⁄ rhizo-remediation) or ex situ (biopiling,
composting, bioreactors and land farming) techniques.
The remedy may be achieved by an intense engineered
biological solution or by means of a less intensive attenu-
Table 2 Studies investigating the uptake and biodegradation of hydrocarbons in soil
Chemical
Biodegradation and
bioavailability mechanism
PAHs (e.g. phenanthrene,
naphthalene)
Soil-sorbed availability
Biosurfactant production
Hydrophobic cell walls ⁄ Direct contact
Aliphatics (e.g. dodecane,
hexadecane)
Soil-sorbed availability
Hydrophobic cell walls ⁄ Direct contact
Biosurfactant production
HOCs (e.g. biphenyl,
styrene)
1246
Hydrocarbon inclusions in cells
Hydrophobic cell walls ⁄ Direct contact
Soil-sorbed availability
Reference
Tang et al. 1998
Deziel et al. 1996; Dean et al. 2001; Wick et al. 2001; Wick et al. 2002;
Bogan et al. 2003
Guerin and Boyd 1992; Ortega-Calvo and Alexander 1994; Guerin and
Boyd 1997; Bastiaens et al. 2000; Holden et al. 2002; Park et al. 2002;
Bogan et al. 2003; Johnsen and Karlson 2004
Thomas et al. 1986; Huesemann 1995; Huesemann et al. 2003; 2004
Bouchez-Naitali et al. 1999; Bouchez-Naitali et al. 2001; Holden et al.
2002; Bogan et al. 2003
Goswami and Singh 1991; Zhang and Miller 1995; Herman et al. 1997;
Bouchez-Naitali et al. 1999; Barathi and Vasudevan 2001; Noordman
and Janssen 2002; Noordman et al. 2002; Bogan et al. 2003; Gogoi
et al. 2003; Rapp and Gabriel-Jurgens 2003
Scott and Finnerty 1976a,b; Whyte et al. 1999b
Calvillo and Alexander 1996; Feng et al. 2000
Fu et al. 1994
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Bioavailability of hydrocarbons in soil
ation stage. The selection of a method is often driven by
economics. Currently, for example, (monitored) natural
attenuation accounts for over one-quarter of all petroleum contaminated land bioremediation strategies in the
USA (Holden et al. 2002). In some cases, the physicochemical and biological constraints of a remediation project negate the selection of an attenuation approach and a
bioengineered solution is required. Such techniques may
be selected when there is low microbial catabolic activity,
a lack of available nutrients, poor pollutant bioavailability, pH extremes or soil physical factors. A slow degradation rate is unsuitable from a regulatory perspective, and
enhanced bioremediation techniques have been developed
to speed up degradation. To ensure optimal biological
performance, it is common to manipulate the degrading
microbial populations and the environment. Several general approaches can be applied: (i) environmental modification, (ii) bioaugmentation and (iii) enhancement of
bioavailability.
Environmental modification involves the manipulation
of the physicochemical nature of the contaminated soil by
altering pH, O2, H2O, and ⁄ or nutrient levels. Nutrient
deficiencies can occur due to the enrichment of carbon
caused by the pollution event (Smith et al. 1998). The
optimization of the C : N : P ratio is thought to be one
of the most important amendments enhancing the biodegradation rate and extent (Smith et al. 1998). However,
whilst these nutrients are routinely added, biostimulation
has been found to be inconsistent. Inorganic nutrients
(fertilizers) are the most widely used amendment, and
success has been reported (Lin and Mendelssohn 1998;
Whyte et al. 2001; Trindade et al. 2005; Ayotamuno et al.
2006; Perfumo et al. 2007). For example, Ayotamuno
et al. (2006) found that the addition of NPK fertilizer to
a polluted agricultural soil in Nigeria had significantly
enhanced the biodegradation rate of the crude oil, initially present at 84 mg TPH kg)1 and reduced by 50–95%
in the test cells. However, the traditional addition of fertilisers or urea has been shown to have no affect (Nyman
1999; Ruberto et al. 2003; Bento et al. 2005; Sarkar et al.
2005; Fernandez-Alvarez et al. 2006; Sabate et al. 2006).
This traditional application has been found to have a negative impact on bacteria (Sarkar et al. 2005) and fungi
(Chaillan et al. 2006), and thus may significantly inhibit
biodegradation. Sarkar et al. (2005) reported that fertilizer
addition caused either NH3 overdosing and ⁄ or fertilizerinduced acid toxicity, which significantly affected the
microbial community. Chaillan et al. (2006) found that
the addition of urea had a fungicidal effect and caused a
toxic concentration of ammonia gas, significantly limiting
the degradation extent of soil contaminated with weathered oils and drill cuttings. Thus, alternative methods of
nutrient amendment to these traditional applications were
Bioavailability of hydrocarbons in soil
receiving increased research attention. A novel technique
involves the application of biosolids, which are cheap and
widely available from sewage treatment works, and have a
slow release of highly available nutrients, avoiding the
problems associated with traditional applications (Sarkar
et al. 2005). A study by Sarkar et al. (2005) compared the
traditional fertilizer addition and biosolids application
and monitored natural attenuation on Tarpley clay soil
contaminated with 3500 mg TPH kg)1 diesel. The results
showed that, after 8 weeks of incubation, the biosolid
amended conditions showed a superior TPH reduction of
96%. Further, the traditional fertilizer application exerted
toxic effects on the soil microflora, while the biosolid
amendment showed no such impacts.
Bioaugmentation involves the addition of an enriched
degrading microbial inoculum to remove target contaminant molecules. However, the success of applying commercially available microbes is limited (Jones 1998;
Whyte et al. 1999a). Bioaugmentation research is developing into two distinct approaches: (i) environmental specificity and (ii) microbe specificity. The main limitation
in commercial preparations and laboratory based inocula
is the inability to rapidly adapt to the local field conditions (Vogel 1996). Environmentally-specific bioaugmentation can be defined as the use of naturally adapted
indigenous microbes that have been cultured and
enriched in the laboratory and then applied back to the
local contaminated soil. This approach has shown considerable success. For example, Bento et al. (2005) showed
up to four-fold increase in the microbial activity of bioaugmented diesel-contaminated soils. This corresponded
with differences in the degradation extent; for example,
the light hydrocarbon fraction (C12-C23) degradation
showed 48Æ7 ± 0Æ33% in the naturally attenuated treatment as compared with 75Æ2 ± 0Æ17% in the bioaugmented treatment. In the heavy hydrocarbon fraction
(C23-C40), 45Æ7 ± 0Æ41% degradation was found in the
biostimulated condition as compared with 72Æ7 ± 0Æ37%
degradation in the bioaugmented condition using Long
Beach soil, initially contaminated with 2800 mg TPH kg)1
TPH (C12-C23) and 9450 mg TPH kg)1 TPH (C23-C40).
In terms of microbe specificity, a new approach is the
molecular engineering of microbes to produce highly
effective degrading microbes as reviewed by Ang et al.
(2005). This report identifies that research has focussed
on enhancing the presence of enzymes that cause PAH
biodegradation, such as laccases and cytochrome P450 oxidases, through either rational design or directed evolution to produce highly effective degrading organisms
(Ang et al. 2005). This research is in its infancy, with the
focus on the complex procedure of genetically modifying
organisms and currently little practical application to biodegradation. Nonetheless, whilst there are obvious advan-
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J.L. Stroud et al.
tages of producing efficient degrading organisms, the field
application of genetically modified organisms is improbable given the current environmental regulations and
increasing unpopularity with the general public.
Enhancing the bioavailability of organic contaminants
may also be used in the bioremediation of contaminated
land. For example, heat has been used to enhance the
removal of aliphatic hydrocarbons from soil; Perfumo
et al. (2007) reported that indigenous thermophilic bacteria are common hydrocarbon degraders and that bioremediation at 60C showed significant hexadecane intrinsic
degradation. Soil was amended with 2% (w ⁄ v) hexadecane;
after 40 d, 70% degradation was measured in microcosms
maintained at 60C. However, hexadecane degradation
was only approximately 38% in soil microcosms maintained at room temperature. Furthermore, with the boiling
point of hexadecane being 287C, losses are unlikely to
have been due to volatilization. This enhanced degradation
is hypothesized to be caused by the mobilization and
solubilization of hexadecane at a higher temperature as
well as the stimulation of thermophilic degrading microorganisms.
Chemical surfactants are able to emulsify or pseudosolubilize poorly water-soluble compounds and, as such,
enhance bioavailability. Attributes such as charge, hydrophilic balance and micellar concentrations define the performance and applicability of the surfactant (van Hamme
et al. 2003), and the enhancement of bioavailability can
be easily measured (Efroymson and Alexander, 1991).
Although widely used in industrial processing applications, surfactants are more difficult to apply to soils
where there may be significant forms of carbon present
other than just the hydrocarbons of interest. For example,
Mulligan et al. (2001), on reviewing the most significant
findings in the literature, reported that degradation of
poorly soluble hydrocarbons could be inhibited because
of toxicity associated with the surfactants, preferential
metabolism of the surfactant, cellular membrane interference and ⁄ or reduced hydrocarbon bioavailability. Rouse
et al. (1994) also observed the potential application of
these chemicals in wastewater and slurry applications, but
were critical about their applicability to oil contaminated
soils. Indeed, it was further stressed that many surfactants
are designed to operate independent of microbial degradation, and thus such combined degradative capacity may
be an exception rather than a rule.
Of wider application is biosurfactant enhanced remediation. Many biological molecules are amphiphilic and
hence could partition at interphases. Microbial compounds that exhibit particularly high surface activity and
emulsifying activities are classified as biosurfactants,
which include lipopeptides, glycolipids, neutral lipids and
fatty acids. Cameotra and Bollag (2003) considered the
1248
J.L. Stroud et al.
way in which communities exhibit biosurfactant production, putatively enhancing degradation through increased
bioavailability. It was concluded that these structurally
diverse compounds produced by hydrocarbon degraders
exhibit surface activities that are difficult to produce synthetically. These compounds are also biodegradable and
have low microbial toxicity, and their concentration and
particular traits are likely to be defined by environmental
parameters.
Commercial microbial remediation of hydrocarbon and
crude oil-contaminated soils is an emerging technology
involving the application of biosurfactants (Bartha 1986;
Harvey et al. 1990; Banat 1995; Ghosh et al. 1995).
Hydrocarbon degradation by indigenous microbial populations is the main mechanism employed (Atlas and Bartha, 1992) because the performance of enhanced
degradation through the addition of prepared microbial
inocula (bioaugmentation) has been ambiguous (Atlas,
1991). However, the amendment of biosurfactant to stimulate indigenous populations has been shown to
degrade hydrocarbons at rates higher than those achievable through nutrient addition alone. For example, rhamnolipids from Pseudomonas aeruginosa were shown to
remove substantial quantities of oil from contaminated
Alaskan gravel from the Exxon Valdez oil spill (Harvey
et al. 1990). Further, Bragg et al. (1994) reported the
effectiveness of in situ bioremediation on the Exxon Valdez oil spill using biosurfactant methods in a large-scale
experiment. More recently, Van Dyke et al. (1993) demonstrated more than a doubling of hydrocarbon recovery
(from 25 to 70% and 40 to 80%) from contaminated soil
using rhamnolipids from P. aeruginosa. Glycolipid biosurfactants have also been shown to enhance the hydrocarbon removal (from 80 to 90–95%) from soil;
furthermore, the biosurfactant was reported to increase
hydrocarbon mineralization by two-fold and shorten the
adaptation time of microbial populations to fewer hours
(Mulligan and Gibbs, 1993).
Conclusion
This review discussed the presence, consequence and
removal of aliphatic hydrocarbons in soils. Aliphatic
hydrocarbons (C14-C20) present in fuels are ubiquitous
soil contaminants and are carcinogenic at high concentrations and represent a considerable hazard to biological
receptors because of the formation of toxic and carcinogenic metabolites as a result of biodegradation. Aliphatic
hydrocarbon biodegradation is constitutive to the indigenous microbial community, which have specialized
uptake degradation mechanisms that enable sorbed-phase
degradation. However, significant concentrations of aliphatics persist in soils because of their interaction with
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Bioavailability of hydrocarbons in soil
soil components resulting in bioavailability limitations.
Remediation techniques attempt to enhance the removal
of aliphatic hydrocarbons from soils, through modifications to the microbial community, enhanced biostimulation and increasing bioavailability using cyclodextrin
amendments. The success of these techniques is important given the nature, consequences and bioavailability
limitations associated with aliphatic hydrocarbon contamination in soils.
Acknowledgements
The authors would like to thank the Natural Environment
Research Council, UK and Remedios, for financially supporting this work.
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