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Weathered Hydrocarbon Biotransformation: Implications for Bioremediation, Analysis, and Risk Assessment

  • S. Cipullo
  • K. J. Brassington
  • S. J. T. Pollard
  • F. CoulonEmail author
Living reference work entry
Part of the Handbook of Hydrocarbon and Lipid Microbiology book series (HHLM)

Abstract

Weathered petroleum hydrocarbons are highly complex and important soil contaminants, which, despite 40 years of petroleum research, are still not sufficiently understood or appropriately characterized for informing contaminated land risk assessments. Improved insights into biotransformation of these contaminants and their residual toxicity are essential for improving risk assessments, bioremediation strategies, and effective regeneration of previously contaminated land. The remediation of land contaminated with weathered hydrocarbons has long been limited by inappropriate analytical methodology, the absence from risk assessment frameworks, reduced stakeholder confidence, lack of ecotoxicological analysis in risk assessments, and a distinct paucity of information regarding weathered hydrocarbon toxicity, distribution, transport, and availability in the environment. Recent research has resulted in the development of a robust analytical method for identification of hydrocarbon residues (weathered hydrocarbons) which are the principal source of the organic carcinogens or suspected carcinogens that drive quantitative risk assessment (e.g., benzo[a]pyrene), development of a tool kit for contaminated sites incorporating ecotoxicological consideration, and an improved understanding of weathered hydrocarbon toxicity and biotransformation chemistry. However, knowledge gaps still remain, and additional implications for bioremediation practitioners have been identified concerning remedial methodology at previously remediated sites.

1 Introduction

Petroleum hydrocarbons are common environmental contaminants. They are components of crude oil and products derived from it and are consequently found on a variety of sites including refineries, chemical materials and by-products storage sites, and manufactured gas production sites. They may also be present as a result of spills and leaks during transportation. They are a highly complex mixture of aliphatic and aromatic hydrocarbons with minor amounts of other heterogenic compounds such as nitrogen, sulfur, and oxygen (Farrell-Jones 2003). Once released to the environment, they are subject to physical, chemical, and biological processes that further change their composition, toxicity, availability, and distribution (partitioning) within the environment. Such degradation processes (weathering processes) include adsorption, volatilization, dissolution, biotransformation, photolysis, oxidation, and hydrolysis (Brassington et al. 2007; Pollard et al. 1994, 2005). The extent of weathering experienced is particularly important when characterizing petroleum contamination prior to remediation (Wang et al. 1998), especially the heavy oils, which have high viscosity (ca. 50–360 mPa s), high boiling point (ca. 300– > 600 °C), and carbon number ranges in excess of C20. These weathering processes shift their chemical composition toward recalcitrant, asphaltenic products of increased hydrophobicity.

Typical concentration of total petroleum hydrocarbons (TPH) of weathered oils is below 5000 mg kg−1, volatile compounds such as benzene, toluene, ethylbenzene, and xylenes (BTEX) are not detected or less than 1 ppm, and the chloride concentration is less than 250 mg kg−1. The aliphatic and aromatic fractions of weathered oils are usually ranging from C12 to C40 (Table 1). These fractions are commonly less bioavailable within the soil due to their unfavorable physicochemical properties (e.g., solubility, volatility, and K ow water/octanol partitioning coefficient) which restricts further microbial attack and degradation (Pollard et al. 1994). However, attempts to improve the bioavailability of the aliphatic and aromatic fractions of weathered hydrocarbons to microorganisms during bioremediation activities may result in increased human exposure. These residual fractions are the principal source of the organic carcinogens or suspected carcinogens that drive quantitative risk assessment (e.g., benzo[a]pyrene, benz[a]anthracene, chrysene) at contaminated sites (Environment Agency 2005).
Table 1

Petroleum hydrocarbon fractions (based on equivalent carbon numbera) for use in UK human health risk assessment. Hydrocarbon fractions usually identified for weathered oils are in bold (Environment Agency 2005)

Aliphatic fraction

Avg EC

Aromatic fraction

Avg EC

EC > 5–6

5.5

EC > 5–7

6.5

EC > 6–8

7.0

EC > 7–8

7.5

EC > 8–10

9.0

EC > 8–10

9.0

EC > 10–12

11

EC > 10–12

11

EC > 12–16

14

EC > 12–16

14

EC > 16–35

25.5

EC > 16–21

18.5

EC > 35–44

39.5

EC > 21–35

28.5

EC > 35–44

39.5

EC > 44–70

57

aThe equivalent carbon (EC) number of a hydrocarbon is related to its boiling point (b.p.) normalized to the boiling point of an n-alkane series or its retention time on a nonpolar b.p. gas chromatographic (GC) column. For hydrocarbons where the boiling points are known, an EC can be calculated. Hexane contains six carbon atoms and has a boiling point of 69 °C and an EC number of 6. Benzene also contains six carbon atoms and has a boiling point of 80 °C. Based on benzene’s b.p. and its retention time on a b.p. GC column, benzene’s EC number is 6.5. This approach has been recognized more appropriate differentiation technique than the actual carbon number of the chemical. For hydrocarbons with higher relative carbon number indices, the disparity (in terms of EC) between aliphatic and aromatic hydrocarbons is substantial (Brassington et al. 2007)

Although these important qualitative and quantitative differences between weathered and non-weathered petroleum hydrocarbons are widely acknowledged (see, for review, Brassington et al. 2007), weathered hydrocarbons are not sufficiently understood or appropriately characterized for assessing risk at contaminated sites.

Measuring the total concentration of petroleum hydrocarbons (TPH) in soil does not give a useful basis for the evaluation of the potential risks to human and the environment. The variety of physical-chemical properties, and thus differences in the migration and fate of individual compounds, and the toxicity of different fractions and compounds in oil products must be taken into account in human health risk assessments .

2 Weathered Petroleum Hydrocarbon Analysis in Soil

While there is a range of methods available for the analysis of weathered hydrocarbons including gravimetric analysis, infrared spectrometry (IR), gas chromatography with flame ionization detector (GC-FID), and gas chromatography coupled to mass spectrometry (GC-MS), method choice may partially be restricted or influenced by the risk assessment being used during the remediation of contaminated land (API 2001; ARCADIS Geraghty and Miller International Inc. 2004). In addition, economics also frequently play an important role in method choice. Many of the risk assessments now used during the remediation of contaminated sites incorporate analytical guidance and reference methods (Table 2). Variations in protocols between frameworks may however give rise to inter-framework variation, affecting the remedial goals set as a result. Indeed, the comparison of reference analytical methods used for petroleum risk assessment protocols highlights the need for practical and simple extraction procedures that allow a better characterization of both aliphatic and aromatic hydrocarbon fractions within oil-contaminated samples, including soil and sediment samples with high moisture levels (Brassington et al. 2007; Environment Agency 2005). In addition to this the Environment Agency of England and Wales notes that currently adopted methods for petroleum hydrocarbon analysis may not be suitable for the heavier compounds (>C16) and questions whether it is practical or relevant for analyzing weathered hydrocarbons (Brassington et al. 2007; Risdon et al. 2008).
Table 2

Summary of different analytical methods developed for risk assessment frameworks (Modified from Brassington et al. (2007 ))

 

Massachusetts Department of Environmental Protection (MaDEP 1994)

Total Petroleum Hydrocarbon Criteria Working Group (TPHCWG 1997a, b, 1998a, b ; 1999 )

Canadian Council of Ministers of the Environment (CCME 2000 )

New Zealand (Ministry for the Environment 1999 )

New South Wales (National Environment Protection Council 1999 )

Risdon et al. (2008)

Extraction technique

Use of two methods. Volatile petroleum hydrocarbon (VPH) method (MaDEP 2004b) and extractable petroleum hydrocarbon (EPH) method developed by MaDEP. VPH method uses purge and trap with methanol. EPH method uses DCM for extraction and solvent exchanges into hexane. Using one of several US EPA solvent extraction methods (MaDEP 2004a)

Vortex or shaker method using n-pentane

Purge and trap for C6 to C10 range using methanol. Soxhlet for the C10 to C50 range

Purge and trap is used for the C6 to C9 range. For the C10 to C36 range, any methods that meet set performance criteria are used

US EPA methods 3540B (US EPA 2005) or 3540C (US EPA 1996a) (Soxhlet extraction), 3550B (US EPA 1996b) (sonication extraction) or sequential bath sonication and agitation described by NEPC (National Environment Protection Council 1999)

Sequential ultrasonic solvent extraction for the nC8 to nC40 using 1:1 acetone/hexane mixture

Evaporation

The EPH method uses those specified by the US EPA. However, after fractionation the use of a gentle stream of air or nitrogen is recommended to bring the sample to the required volume. Evaporation is not applicable to the VPH method

N/A

Uses an evaporation vessel after extraction for the C10 to C50 range. After silica gel cleanup, rotary evaporator is used to reach the required sample volume

Use of any method that meets set of performance criteria

US EPA methods specified for extraction using Kuderna-Danish (K-D) evaporation

Not required but can be used to achieve lower limits

Cleanup/fractionation

Silica gel cleanup for EPH method. Not applicable to VPH method

Extract fractionation using alumina or silica

One of two specified cleanup steps for C10 to C50 range, not fractionated

Cleanup steps and fractionation are optional as this may not be required for each sample/analytical approach

Solvent exchange into hexane followed by K-D evaporation and treated with silica gel as described in US EPA method 1664 (US EPA, 2005; National Environment Protection Council 1999)

Not necessary; however is used to fractionate samples. Uses microscale silica gel column chromatography after a water partition step

Analysis technique

EPH uses GC-FIDa. VPH may use either GC/FIDa or GC/PIDb

GC-FIDa

GC-FIDa

For the C10 to C36 range, GC-FIDa is used, and for the C6 to C9 range, GC-MSc is used

GC-MSc, or GC-FIDa; however the use of GC/MSc to identify unusual mixtures is noted as being necessary when analyzing by GC-FIDa

Combination of GC-FID and GC-MS depending upon level of analysis

aGC-FID refers to gas chromatography with flame ionization detection

bGC-PID refers to gas chromatography with photoionization detection

cGC-MS refers to gas chromatography with mass spectroscopy detection

Various extraction techniques for petroleum hydrocarbons exist within the open literature; however many suffer from inter-method variation and both positive and negative bias (Buddhadasa et al. 2002; Environment Agency 2003; Whittaker et al. 1995). Historically, Soxhlet extraction has been the benchmarked method, due to its exhaustive nature, high recoveries, and ability to be easily standardized (Risdon et al. 2008; Shu et al. 2003). However Soxhlet suffers from long extraction times, a need to concentrate samples, high solvent use, and the degradation of thermally liable compounds (Risdon et al. 2008; Shu et al. 2003). Consequently this has resulted in investigations into and development of the alternative methods (Hawthorne et al. 2000; Hollender et al. 2003; Risdon et al. 2008; Whittaker et al. 1995). Alternative methods to Soxhlet include ultrasonication (Banjoo and Nelson 2005; Risdon et al. 2008; Sanz-Landaluze et al. 2006), pressurized liquid extraction (Hawthorne et al. 2000), supercritical fluid extraction (Hawthorne et al. 2000), subcritical water extraction (Hawthorne et al. 2000), and microwave-assisted extraction (Saifuddin and Chua 2003). Although some of the alternative methods offer improvements over Soxhlet extraction, most of these methods need further refinement and optimization, as there have been conflicting results from different investigations into the same method. For example, Heemken et al. (1997), Sun et al. (1998), Banjoo and Nelson (2005), and Sporring et al. (2005) demonstrated ultrasonic methods that had equivalent or better extraction efficiencies than Soxhlet, whereas investigations by Song et al. (2002) and Hollender et al. (2003) gave worse extraction efficiencies.

Recently, a novel and robust ultrasonic extraction method for contaminated soils with weathered hydrocarbons has been developed and optimized (Risdon et al. 2008). The method covers the determination of TPH between nC8 and nC40 and subranges of hydrocarbons including diesel range organic compounds , kerosene range organic compounds, and mineral oil range organic compounds in soils. Further modifications to the TPH carbon banding may be made as requested for risk assessment including ranges known as Texas Risk banding (TPH C8–C10, C10–C12, C12–C16, C16–C21, and C21–C5) as well as separation of the aliphatic and aromatic fractions as defined in the UK regulatory framework (Environment Agency 2005). The method can be routinely used for measuring hydrocarbons down to 10 mg kg−1 in soil. Detection limits may vary for individual carbon ranges calculated on the percentage of the full range they cover. With an extraction efficiency and recovery between 95% and 99%, this method can be easily positioned as a good alternative to Soxhlet extraction and shows a good potential for implementation as a standard method potentially providing further insight to the contaminated land sector. The method has been accredited ISO17025 for TPH analysis, banding, and class separation.

3 Risk Assessment for Weathered Hydrocarbons

Risk assessment is an established requirement for the management of contaminated land (ARCADIS Geraghty and Miller International Inc. 2004) and now a widely used support tool for environmental management decisions. It is employed as a means of assessing and managing potential impacts to human and ecosystem health (Vegter et al. 2002). Several risk-based frameworks for petroleum hydrocarbons in soil have been published under the auspices of the Total Petroleum Hydrocarbon Criteria Working Group (TPHCWG 1999), the American Society for Testing and Materials (ASTM 1994), the Massachusetts Department of Environmental Protection (MADEP 2002), the Environment Agency of England and Wales (Environment Agency 2005), the American Petroleum Institute (API 2001), and the Canadian Council of Ministers of the Environment (CCME 2000), each reflecting national legislation, a range of expert judgments, and socioeconomic issues. Typically these frameworks adopt a three-tiered approach with increasingly sophisticated levels of data collection and analysis, as an assessor moves through the tiers. However, these frameworks, and the exposure assessment methods embedded within them, do not specifically address weathered hydrocarbons, although some acknowledge that petroleum products released to the environment will have undergone some degree of degradation (API 2001; Environment Agency 2005; MADEP 2002; TPHCWG 1998a).

Assessing the risk of weathered hydrocarbons that they pose at contaminated sites is complicated because the profile of compounds present in weathered oil can be very different from the composition of fresh oil. However, many hydrocarbon compounds would be sufficiently similar in structure to expect that they might have similar toxicities and endpoints. In view of these factors, the UK approach to human health risk assessment of petroleum hydrocarbons favored the adoption of a combined indicator and fraction approach within a tiered risk-based framework (Environment Agency 2005). Specific indicator compounds (genotoxic carcinogens and noncarcinogens) should be assessed because these are often the key risk drivers at petroleum-contaminated sites. Genotoxic carcinogens are assumed not to have a threshold concentration as even very small concentrations (or doses) are assumed to pose some (albeit small) risk of cancer. There are cases in which carcinogenicity can be assumed to occur only after some dose or threshold concentration is reached, depending on the mode of action by which the contaminant is thought to cause cancer. The assessment of fractions should facilitate a more representative picture of risk at sites where the origin of the petroleum hydrocarbons contamination may be unclear. The fractions are typically used to consider threshold health effects (Environment Agency 2005).

Recent work conducted by the research consortium PROMISE on optimizing biopile processes for weathered hydrocarbons within a risk management framework identified suitable potential hydrocarbon fractions and indicator compounds for weathered hydrocarbons (Tables 1 and 3), the adoption of which is essential to the remediation of weathered hydrocarbons in soil (Coulon 2008). Within the aliphatic fraction, only the carbon ranges C12–C16 has been identified as threshold health indicator fractions for weathered hydrocarbons. Additionally the atypically high threshold toxicity, relative to the other hydrocarbon compounds, of 2-methylnaphthalene (oral exposure) and naphthalene (inhalation exposure) indicated that they should be viewed as potential indicator compounds at contaminated sites.
Table 3

List of potential indicator compounds for weathered hydrocarbons

Potential indicator compounds for weathered hydrocarbon-contaminated soils

Naphthalene

Aromatic > EC10–EC12

Acenaphthene

Aromatic > EC12–EC16

l-Methylnaphthalene

Pyrene

Aromatic > EC16–EC21

Phenanthrene

Fluoranthene

Aromatic > EC21–EC35

Benz[a]anthracenea

Benzo[b]fluoranthenea

Benzo[k]fluoranthenea

Benzo[a]pyrenea

Benzo[ghi]perylene

Chrysenea

Dibenz[a,h]anthracenea

Indeno[1,2,3-c,d]pyrene

5-Methylchrysene

aNon-threshold indicator compound, as known to possess some genotoxic carcinogenic potential

As shown in Tables 1 and 3, use of the combined indicator and fraction approach instead of measuring the total concentration of TPH in weathered oils provides a better insight of the carbon range compounds and the key risk drivers within each fraction. Access to this detailed information is important for assessing human and environmental risks and effective remediation at contaminated sites.

4 Use of Fugacity Modeling to Parametrize and Conceptualize Fate and Behavior of Petroleum Hydrocarbons

Understanding the environmental fate of a contaminant is a key requirement when estimating potential risks to human health. To achieve this, meaningful information on a substance’s behavior and distribution, toxicity, concentration, and potential exposure at a site is essential (Allan et al. 2006; Environment Agency 2003).

The overriding dominance of the nonaqueous phase liquid (NAPL) for hydrophobic contaminants is theoretically established but rarely incorporated into the exposure assessment tools used to derive soil screening levels and guideline values (Pollard et al. 2008). This oversight is likely to have a marked influence on soil assessment criteria at hydrocarbon-contaminated sites. Its significance comes into play when one considers the residual risk posed by posttreatment residues. Level I and II fugacity models were developed comprising four phases within the soil matrix, namely, air, water, mineral soil, and NAPL. The implications of the fugacity modeling developed by Pollard et al. 2008 are important for risk analysts and remediation engineers. The fugacity modeling confirmed the propensity for risk critical compounds to be preferentially partitioned to the NAPL and soil phases. However, modeled depletion times for contaminants in the context of authentic soils are immaterial, and thus research efforts should be focused on the likely exposures of humans and other receptors to residual saturation at hydrocarbon-contaminated sites. The results indicate clearly the need for modifications to the exposure assessment models used to generate soil screening guidelines or guideline values, so as to better represent contaminant fate in the multimedia systems.

5 Bioavailability Complexity

Bioavailability has been recognized as a key component in risk assessment, but it has not been widely implemented yet (Harmsen and Naidu 2013). Bioavailability of contaminants can be influenced by a wide range of factors including sorption rate, partitioning, mineral speciation, pH, ageing, soil structure, and soil organic content. The physical-chemical characteristics of soil and contaminants are not the only factors affecting bioavailability in soil. In fact, chemical-biological processes such as microbial degradation also contribute in making contaminants less available for uptake. The relationship between the percentage of bioavailable and non-bioavailable fractions is represented in Fig. 1. While bioavailable fraction is decreasing over time the non-bioavailable fraction is increasing. For example, weathered hydrocarbon residues pose negligible risks to human health, and this should be reflected in posttreatment remedial objectives. Bioavailability is directly related with exposure and risk estimates and inversely related to risk-based cleanup levels. Higher cleanup levels are required when bioavailability is high, mainly due to increased exposure and risk estimates.
Fig. 1

Relationship between the percentage of bioavailable and non-bioavailable fractions, exposure risks, and risk-based cleanup level; (Adapted from Reid et al. 2000)

Given the multiple variables affecting the availability of chemicals in the soil, we should look at bioavailability not as a fixed value (concentration), but as a dynamic process between an organism and the chemical uptake over time (Lanno et al. 2004). However the evaluation of contaminant-soil matrix interactions, that might partially be the cause of the non-accessibility, is still a challenge (Wu et al. 2013, 2014 ). Research should focus on understanding and accurately representing the bioavailable fraction and ensuring that this parameter is correctly quantified in the risk assessment.

6 Bioremediation of Weathered Hydrocarbons

There is a plethora of approaches for the remediation of contaminated land, encompassing physical, chemical, and biological methods which can contain, destroy, or separate the contaminants. Implementation of the EU Landfill Directive (The Council of the European Union 1999) encourages the development and implementation of alternative remediation techniques such as bioremediation in a move away from the mainstay method of excavation and disposal. The organic nature of petroleum hydrocarbons makes these contaminants highly suited to bioremediation techniques as such, and due to the widespread, health, and ecological hazards posed by petroleum hydrocarbons, greater interest has been directed at these contaminants. Bioremediation is a well-established method that works well for remediating petroleum hydrocarbon-contaminated soil (Flathman et al. 1994; Hyman and Dupont 2001). Bioremediation methods are often optimized, using biostimulation and bioaugmentation to enhance biotransformation, reduce cost, and process duration.

Typically, biotransformation is rapid in the initial stages of bioremediation, with rates seen to asymptote toward the end of remediation treatments (Ellis 1994; Fogel 1994; Wood 1997) as the proportion of less bioavailable and recalcitrant compounds increase. Weathered hydrocarbons generally display relatively low bioavailability and are more recalcitrant than their non-weathered counterparts. As such optimization of bioremediation can prove essential to the successful remediation of weathered hydrocarbons (Giles et al. 2001; Guerin 2000).

Recently, Brassington (2008) demonstrated that significant biotransformation of weathered hydrocarbons in soil can be achieved. After 112 days of bioremediation treatments, using biostimulation and bioaugmentation methods (Fig. 2), the residual fraction of weathered petroleum hydrocarbons of two soils (A and B) was degraded up to 85% and 92%, respectively, of its initial content. Soil A was a hydrocarbon-contaminated sandy soil that has previously undergone remediation, and soil B was a clay soil-contaminated with weathered hydrocarbons.
Fig. 2

Change in mean (± se) total petroleum hydrocarbon (TPH) concentration (mg kg−1) over 112 days of treatment for soil a and soil b

Further, on the basis of physicochemical parameters, the study suggested that the success of bioremediation considering both biostimulation and bioaugmentation approaches was largely dependant of the oil contaminant and soil structure characteristics. For the majority of the contaminated soils investigated, mineral nutrients played an essential role without which in some cases bioremediation could not occur. Slow-release fertilizers were shown to be an important alternative to liquid fertilizers, in mitigating issues arising from the addition of liquid fertilizers. Combining bioaugmentation strategies with biostimulation may improve the rate and extent of weathered hydrocarbon degradation, while the potential benefit of bioaugmentation still needs further evidence. Ex situ bioremediation for treatment will allow greater control over soil temperature, water holding capacity, and leaching.

The design of an efficient bioremediation system always requires a careful site assessment. Consideration of the physical, chemical, and biological properties of the contaminated sites is essential in establishing appropriate response and recovery methods. Despite the ability of indigenous microorganisms to degrade petroleum hydrocarbons, there are still situations where the use of a microbial inoculum might enhance petroleum hydrocarbon biodegradation.

A diagnostic strategy toolbox for ecological hazard assessment of weathered hydrocarbons has been developed by the research consortium PROMISE (Brassington 2008; Dawson et al. 2007). The toolbox highlighted the role of a multiple-trophic view when considering both the hazard and remediation of weathered hydrocarbons. The selected bioassay techniques were used in combination with the chemical analysis to allow ecological relevance and a more focused understanding of hydrocarbon transformations (Table 4). The bioassays were selected on their ease of execution and representation of different ecological soil organisms.
Table 4

Ecotoxicological tests used for weathered hydrocarbons (selected species are in bold)

Test

Species considered

Earthworm survival

Eisenia fetida, Lumbricus terrestris, Lumbricus rubellus

Seed germination

Brassica alba mustard white, Triticum aestivum (Consort) wheat, Pisum sativum, pea

Luminescence-based bacterial biosensors

Metabolic: Vibrio fischeri, Escherichia coli HB101, Pseudomonas putida F1 Tn5

Catabolic: Escherichia coli HMS174, Escherichia coli DH5α, Pseudomonas putida TVA8

The research demonstrated that a gross reduction of the hydrocarbons does not represent environmental nor sustainable improvement, as reported by the biological response. In order to integrate and rank the effectiveness of remediation treatments, the biological indicator data were transformed into a soil quality index (SQI) (Dawson et al. 2007). The results highlighted that biological receptor specificity defined the risk-based endpoint and reinforced the concepts of risk reduction within a dynamic and degrading oil matrix. Furthermore, the changes in toxicity and to some extent the bioavailability of fractions within the matrix reflect that some pollutants may be mobilized during remedial activities .

7 Research Needs/Knowledge Gaps

It is clear that successful bioremediation of land contaminated with weathered petroleum hydrocarbons depends upon the successful integration of effective analysis techniques, informed optimization of bioremediation, and an appropriate risk assessment protocol. Remediation of contaminated soils is limited by several factors including a lack of toxicological and environmental fate and behavior data, inadequate and variable chemical analyses, negative stakeholders’ perception of bioremediation techniques, and variable risk assessments that do not always consider weathered hydrocarbons. While the work of researchers has taken important steps toward addressing key knowledge gaps and methodological limitations within analysis and risk protocols, there is still work to be done.

Further investigation is still required to increase our knowledge on weathered hydrocarbon chemical, toxicological, and biological diagnostics as well as environmental fate and behavior. This combined diagnostic approach will significantly help to identify optimal remediation strategies and contribute to change the overconservative nature of the current risk assessments.

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Copyright information

© Springer International Publishing AG 2016

Authors and Affiliations

  • S. Cipullo
    • 1
  • K. J. Brassington
    • 1
  • S. J. T. Pollard
    • 1
  • F. Coulon
    • 1
    Email author
  1. 1.School of Water, Energy and EnvironmentCranfield UniversityCranfieldUK

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