Heavy hydrocarbon fate and transport in the environment

David M. Brown1*, Matthijs Bonte1, Richard Gill1, James Dawick2 & Peter J. Boogaard31 Shell Global Solutions International BV, Lange Kleiweg 40, 2288 GK Rijswijk, Netherlands2 Shell Health – Global Risk Sciences, Shell International B.V., Manchester, M22 0RR, UK3 Shell Health – Global Risk Sciences, Shell International B.V., The Hague, 2596 HP, Netherlands

D.M.B., 0000-0001-8248-056X; J.D., 0000-0002-7450-6114; P.J.B., 0000-0002-6964-6681*Correspondence: david.bro@shell.com

Abstract: Heavy hydrocarbons are a heterogeneous mixture of compounds consisting mainly of alkylated cyclics, resins andasphaltenes and, depending on the source, can form a significant proportion of crude oil. Their prevalence is expected toincrease in the future as heavy oil reserves are increasingly exploited for growing worldwide energy demands. Despite theirgrowing use, heavy hydrocarbons are generally overlooked when assessing the risk of hydrocarbons to human health, ecologyand water reserves. Although their human and environmental health risks are considered low, heavy hydrocarbons are known topersist in the environment. This review considers the fate, transport and toxicity of heavy hydrocarbons. It provides adescription of the possible mechanisms involved in heavy hydrocarbon attenuation and offers some interpretation of data thatprovides insight into their persistence in the environment.

Supplementary material: The effect of hydrocarbon concentration and soil organic content on the partitioning of heavyhydrocarbon (EC >44) into soils is available at https://doi.org/10.6084/m9.figshare.c.3791473.

Received 15 December 2016; revised 14 April 2017; accepted 24 April 2017

Petroleum hydrocarbon compounds can enter groundwater owing tospills or leaks associated with exploration, production anddistribution of oil and gas. Releases can also enter the subsurfaceas a result of natural oil and gas seeps. These releases may result incontaminated sites that pose an unacceptable risk to human health,ecology and/or water resources, and have prompted research on theenvironmental fate and transport and consequent risk of a range ofpetroleum hydrocarbons found in crude oil or refined products.These include relatively soluble and volatile compounds, such asbenzene, toluene, ethyl-benzene and xylene (BTEX) and morerecalcitrant compounds such as polyaromatic hydrocarbons(PAHs). An alternative approach is to assess risks for a number ofdefined hydrocarbon fractions (e.g. C8–C10, C16–C35 up to C44)using representative toxicity and physico-chemical parameters foreach fraction (CONCAWE 1996; TPHCWG 1997; Verbruggen2004). Whereas there is a reasonable understanding of the fate andtoxicity of these selected components up to carbon length of C44, thebehaviour of heavier hydrocarbons has received far less attention.

In general, heavy (i.e. high molecular weight) hydrocarbons areconsidered less mobile and less (eco)toxic, which implies thatassociated health and environmental risks are also lower comparedwith lighter hydrocarbons. The persistence of heavier hydrocarbons,however, is also much greater compared with lighter fractions. Thisraises the question of how to deal with cases of heavy hydrocarbonsoil and groundwater contamination, knowing that the risks tohuman health, ecology and water resources are small, yet on theother hand biodegradation and natural attenuation processes aresuch that the contamination may persist for a significant length oftime (years). This is especially relevant to spills of fresh heavy crudeoils and weathered light crude oils. It is also an issue for theredevelopment of legacy brownfield sites (with ubiquitous residualheavy hydrocarbon contamination) where problem-owners andregulators are faced with decisions on how best to manage theselow-toxicity, yet persistent, impacts.

To address these issues and contribute towards developing anappropriate risk management strategy, this paper presents a

literature review of the fate of heavy petroleum hydrocarbons. Thereview describes the following: (1) heavy hydrocarbons and theiroccurrence in crude oil; (2) the physical and chemical properties ofheavy petroleum hydrocarbons; (3) major attenuation mechanismsand reasons for persistence in the environment; (4) their fate,transport and associated risks to man and the environment.

Defining heavy hydrocarbons

Crude oil is a complex mixture of hydrocarbon- and non-hydrocarbon-containing constituents. Hydrocarbons, moleculescontaining only hydrogen and carbon, consist of a large array oflinear, branched, cyclic and aromatic components, whereas non-hydrocarbon fractions, termed resins (carbazoles, thiophenes andoxygenated hydrocarbons) and asphaltenes, contain additionalelements, such as nitrogen, sulphur and oxygen. Hydrocarbonsexist at a range of molecular weights, from very light (C1,16 g mol−1) to very heavy (∼C80, up to 1000 g mol−1).Asphaltenes are generally considered to be one the largestcomponents in crude oil with molecular weight up to1000 g mol−1 (Mullins 2007, 2008).

Crude oil is refined in part by distillation to separate products byboiling point. Much of the lighter, more valuable hydrocarbon isused as a feedstock for gasoline and jet fuel, whereas the higherboiling point hydrocarbons may be used for diesel, heavy-duty fuelsand lubricating oils. For the purposes of this review heavyhydrocarbons are defined as the fraction that cannot be distilledunder atmospheric and vacuum distillation (McMillen et al. 2001).This is equivalent to starting boiling temperature of c. 550°C atatmospheric pressure. Very little heavy hydrocarbon finds its wayinto fuel and lubricant products and most will end up as bitumen,petroleum pitch or asphalt at the end of the petroleummanufacturingand refining processes. As boiling points are dependent onmolecular structure, chemical composition and molecular weight,hydrocarbons with a range of carbon numbers will constitute thestarting point that defines the heavy hydrocarbon fraction. The

© 2017 Shell Global Solutions International BV. This is an Open Access article distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0/). Published by The Geological Society of London. Publishing disclaimer: www.geolsoc.org.uk/pub_ethics

Thematic set:Organic contaminants in groundwater Quarterly Journal of Engineering Geology and Hydrogeology

Published online June 15, 2017 https://doi.org/10.1144/qjegh2016-142 | Vol. 50 | 2017 | pp. 333–346

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proportion of heavy hydrocarbon in crude oils can vary consider-ably. A survey of 41 crudes by McMillen et al. (2001) found thatheavy hydrocarbons ranged between 0.6 and 57% and had a meanvalue of 23%. This broadly agrees with work by Coleman et al.(1978), who showed heavy hydrocarbon to be 0 – 70% in anassessment of over 800 US crudes.

Characterization of heavy hydrocarbons inenvironmental samples

The principal problem in characterizing hydrocarbon contaminationin soils and waters is that crude oil consists of a complex mixture ofcomponents, rather than a single chemical. The Total PetroleumHydrocarbon Criteria Working Group (TPHCWG) grouped hydro-carbons, up to C35, into a number of fractions based on carbonnumber ranges and structure (aliphatic or aromatic). This wasjustified by differences in solubility and other fate and transportcharacteristics (TPHCWG 1997). This approach, using gaschromatography with flame ionization detection (GC-FID), wasextended by the American Petroleum Institute in 2001 (Douglaset al. 2001) to include hydrocarbons up to equivalent carbon (EC)number 44. No standard analytical method exists for the detectionand measurement of heavy hydrocarbons beyond EC44. Douglaset al. (2001) proposed the use of vacuum distillation to quantifyhydrocarbons beyond the operating limit of GC. Although thismethod is reliable and provides an accurate measurement of heavyhydrocarbon content, it can only be undertaken using specialistequipment. Other ways to estimate the total mass of heavyhydrocarbon concentration in environmental samples includegravimetric hydrocarbon measurement after solvent extraction(e.g. ASTM 2016a) or by simulated distillation by GC (e.g.ASTM 2016b). It should be noted that gravimetric approaches willgenerate a total hydrocarbon measurement, and estimation of theheavy fraction can be achieved only by determining the proportionof the lighter hydrocarbon (≤C44) fraction using GC or similarmethods. Whichever approach is used, the choice of solvent forextraction of hydrocarbons from the environmental matrix requiresfurther investigation, as they are known to affect the miscibility ofasphaltenes (Miadonye & Evans 2010) and potentially have animpact on the effectiveness of the extraction step.

Although these techniques provide insight into the fraction ofheavy hydrocarbons, they do not provide information on thecompositional makeup of the heavy hydrocarbons. The mostcommonly used analytical technique to characterize complexhydrocarbon mixture analysis, combined gas chromatography–mass spectrometry (GC–MS) (Hsu &Drinkwater 2001), often lackssufficient chromatographic resolution and mass resolving power tocharacterize the heavy hydrocarbon fraction. Other methods arebetter suited for this analysis, such as fractionation coupled withfield ionization mass spectrometry (FIMS) (Boduszynski 1985;Malhotra et al. 1993), thin layer chromatography with flameionization detection (TLC-FID) (Ugochukwu et al. 2013), orFourier transform ion cyclotron resonance (FT-ICR) (Marshall &Rodgers 2008; Hegazi et al. 2012).

FIMS uses nominal mass resolution to its advantage by providingone mass (m/z) value for all of the isomers present for a particularcompound. Analysis by this method shows heavy hydrocarbons tobe a mixture of alkylated cyclic, aromatic and naphthoaromatichydrocarbons that have up to seven ring structures, but typicallythree to four (Fig. 1). They are extensively alkylated, with chainlengths of between C15 and C40 (Sinninghe Damsté et al. 1991; Luet al. 2011), and constitute 25 – 40% of the heavy hydrocarbonfraction. Very high molecular weight (>500 g mol−1) resins andasphaltenes will form a substantial component of this fraction.Based on studies by Tissot &Welte (1984) and Boduszynski (1985)resins and asphaltenes constitute between 40 and 80% of the heavy

hydrocarbon fraction. Smaller fractions of n- and iso-alkanes willalso be present (up to 5%). Carbon chain lengths can reach beyondC80 but a range between C40 and C60 is more typical (Tissot &Welte1984). TLC-FID is a semi-quantitative method used to determinethe relative percentage distribution of the saturates, aromatics, resinsand asphaltenes (SARA) fractions in oil. Although this techniquedoes not have high reproducibility, it can be used to rapidly assessgross changes in heavy hydrocarbon composition (Ugochukwuet al. 2013). The relatively recent development of Fourier transformion cyclotron resonance (FT-ICR) coupled with electrosprayionization (ESI) or atmospheric pressure chemical ionization(APCI) mass spectroscopy has further opened up the field ofhydrocarbon characterization. These techniques provide the highestresolving power, mass accuracy and dynamic range amongst alltypes of mass spectrometers and have revolutionized our under-standing of the composition and characterization of crude oils,refined petroleum products and their degradation metabolites(Marshall & Rodgers 2008; Hegazi et al. 2012; Lemkau et al.2014). One area in which FT-ICRMSmay prove to be an extremelyuseful technique is the analysis of hydrocarbon biodegradationintermediates. Recent work by Islam et al. (2016) applied thetechnique to characterize dissolved organic carbon (DOC) down-gradient from a crude spill site. This work was able to clearlyidentify products of hydrocarbon biodegradation and differentiatethem from natural (background) DOC. Others have also used FT-ICR to follow the biodegradation processes of oil released intomarine environments (Chen et al. 2016) to show the identificationand fate of oxygenated compounds in marine sediments.

Fate and behaviour of heavy hydrocarbons

The fate of crude oil in the environment is an important aspect of riskassessment because it determines the exposureof ahumanorecologicalreceptor to the mixture (McMillen et al. 2000). To achieve this,information on hydrocarbon behaviour and distribution, concentrationand potential exposure is essential (Coulon et al. 2010). A conceptualdiagram of heavy hydrocarbon fate and transport is presented inFigure 2 to highlight the major processes involved in the fate of heavyhydrocarbons when released into the environment. It is assumed thatheavy hydrocarbons are present together with lighter hydrocarbonfractions that are found in most crude oils. The conceptual diagram issplit into three stages that describe transport and transformationprocesses: (1) initial surface release of crude oil and infiltration into theunsaturated zone; (2) partitioning of a non-aqueous-phase liquid(NAPL) body near the water table; (3) attenuation of oil resulting inprogressive increase in proportion of heavy hydrocarbons and masssubsequent removal of heavy hydrocarbons by biodegradation. Thesestages can be used to infer at what point heavy hydrocarbons impart thegreatest potential exposure to human health and the environment.

Initial surface release and infiltration into theunsaturated zone

If crude oil is released in sufficient quantities, there is the potentialto infiltrate into the subsurface and form an NAPL body at the watertable. The rate and extent of infiltration will depend on the soilproperties (porosity, pore size, moisture levels) as well as thephysical properties (viscosity, density, surface tension) and thevolume of the oil spilled. Generally, the greater the soil permeabilityand the lower the product viscosity, the greater the rate and extent ofsoil penetration (API 1972; Brost & DeVaull 2000; Keller &Simmons 2005). Spilled oil will penetrate the soil to a point wherethe NAPL head is insufficient to overcome pore entry pressures, andfurther migration ceases. As such, the volume of NAPL migratingdecreases as immobile residual NAPL is left behind owing toretentive capillary forces in the soil column. The result is a partially

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saturated mass zone in the soil that is subsequently subjected to thevarious attenuation and transport processes typical of hydrocarboncontaminants (McCarty & Ellis 2002; Bear & Cheng 2010). Anumber of studies have shown that downward migration towards thewater table is not an issue when hydrocarbons are present at lowconcentrations in the soil. Raymond et al. (1976) assessed migrationin a range of soils (silty loam, sandy loam and black clay loam) withseveral oils dosed at 2% (w/w). In that study 99% of the oil mixedinto the top 15 cm of soil remained within 20 cm of the surface after1 year. Brost & DeVaull (2000) determined the concentration limitin soil at which NAPL would be immobile (not migrate). It wasfound that a range of heavy petroleum products had soil residualsaturation points of 0.5–15% (w/w). The variability was in partattributed to differences in soil types, with much lower valuesobserved in coarse soils.

Partitioning of NAPL

If residual soil saturation points are overcome, then NAPL willdescend through the soil column until it reaches the water table.Dissolution of hydrocarbons from the NAPL into the groundwatermay then generate contaminant plumes that are transported tosensitive receptors. Mass transfer of hydrocarbons from NAPL intothe aqueous phase is determined by Raoult’s law. Level I fugacitymodels, developed by Pollard et al. (2008), illustrate the basicequilibrium partitioning behaviour in air, water, NAPL and soilsystems. Calculations with heavy hydrocarbons show an extremepreference for partitioning into NAPL and soils (Table 1) comparedwith water and air. This analysis indicates that mass transfer of heavyhydrocarbons into groundwater will be extremely limited.Subsequent transport of dissolved hydrocarbons relative to ground-water can be estimated using retardation factors (Nham et al. 2015).Based on the physical and chemical properties proposed byMcMillen et al. (2001), the estimated retardation factor for heavyhydrocarbons is very high (Table 1), suggesting that potentialtransport after dissolution in groundwater will be tremendouslylimited. As a consequence, heavy hydrocarbons are unlikely to betransported any meaningful distance via groundwater flow. Karsticgeological environments may be an exception to this position, wheresuspended solids and/or NAPL may potentially be transported oversignificant distances. Although long-distance transportation ispossible, recent research (Schwarz et al. 2011) suggests the likelihoodof karstic transport of heavy hydrocarbons causing an unacceptableimpact on groundwater is low. Based on these considerations, heavy

hydrocarbons are unlikely to be of interest in hydrogeological riskassessments. Conversely, consideration of heavy hydrocarbons issubstantially more important during the risk management of soils orfree-floating NAPLs to control points of exposure.

Attenuation mechanisms of heavy hydrocarbons

Based on level II fugacity modelling proposed by Coulon et al.(2010), the ultimate fate of heavy hydrocarbons will be dependenton the effectiveness of biodegradation (data not shown), althoughother attenuation mechanisms such as photo-oxidation maycontribute to mass removal mechanisms. Further details of theseprocesses are described in the next section.

Photo-oxidation

Photo-oxidation has the potential to transform the composition ofoil spills at the soil surface. The process requires oil spills to beexposed to sunlight and involves the absorption of photons bymolecules within the crude oil mixture, which then initiatesoxidation reactions by two main mechanisms: singlet oxygen andradial chain oxidation (Shankar et al. 2015). The result is an increaseof oxygenated hydrocarbons that may be more susceptible tobiodegradation (Maki et al. 2001; Lee 2003). It has been suggestedthat photo-oxidation could be an effective attenuation mechanismfor heavy hydrocarbons. Garrett et al. (1998) proposed that photo-oxidation preferentially transformed and oxygenated high molecu-lar weight aromatic hydrocarbons. In agreement with this,Prince et al. (2003) and co-workers used PAHs to demonstrateincreasing levels of photo-oxidation with increasing aromaticity andalkylation. This pattern of photo-oxidative loss is very differentfrom that seen during biodegradation (as described in the nextsection). Less in known about the photo-oxidative propensity ofasphaltenes or resins, which can form a significant component ofheavy hydrocarbons. Current knowledge suggests that these speciesare not subject to significant photo-oxidative destruction(Prince et al. 2003).

Most of the understanding relating to photo-oxidation has beengenerated using marine-based studies. Less is known about thesignificance of photo-oxidation of hydrocarbons in soils. As theprocess requires direct sunlight it will be effective only at the soilsurface. Braddock et al. (2003) attributed poor removal rates ofcrude oil in soil horizons (>8 cm from the surface) to a lack ofphoto-oxidation caused by limited sunlight penetration over a

Fig. 1. Chemical structures of heavyhydrocarbon classes. Figure showsexample structures for the major classes ofheavy hydrocarbon. Alkylations portrayedin the figure are illustrative only.

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25 year period at an Alaskan test site. Whether this limitation can beovercome through simple soil turning (tilling) has not yet beendetermined; however, the potential benefits of photo-oxidationwarrant further investigation of this mechanism as a form ofattenuation.

Biodegradation

Biodegradation is a major attenuation process for hydrocarbonsreleased into the environment. The underlying mechanisms thatultimately drive the mineralization of hydrocarbons have been

Fig. 2. A conceptual model of heavy hydrocarbon fate and transport in the environment. Heavy hydrocarbons are typically released together with lighterhydrocarbon fractions that are found in most crude oils. This modifies the physical and chemical properties that significantly affect the transport on soils andin the subsurface. Releases of viscous refined bitumen or extremely heavy crude oils (APIo < 12) are unlikely to travel any significant distance in the soil orsubsurface.

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the subject of numerous studies (Atlas 1981; Leahy & Colwell1990; Cerniglia 1992; Boonchan et al. 2000; Widdel & Rabus2001; Seo et al. 2009; Abbasian et al. 2015). Much of thisknowledge has been established through the study of lighterhydrocarbon components, whereas knowledge of heavy hydro-carbons biodegradation is still in its infancy. This disparity ishighlighted in the number of identified enzymes known to beinvolved in heavy hydrocarbon biodegradation. To date only oneenzyme, a novel monooxygenase from Acinetobacter sp. DSM17874 (Throne-Holst et al. 2007), has been identified in thebiodegradation of hydrocarbons, which contrasts completely withlighter hydrocarbons for which multiple enzymes in multiplepathways are known. Heavy hydrocarbon biodegradation hasmainly been demonstrated through laboratory-scale microcosmstudies. For example, long-chain n-alkanes have been shown tocompletely degrade under aerobic conditions (Sakai et al. 1994;Wang et al. 2002). Similarly, asphaltenes and resins aredegradable to varying degrees under aerobic and anaerobicconditions (Kim et al. 2005; Strapoc et al. 2008; Uribe-Alvarezet al. 2011; Singh et al. 2012; Tabatabaee et al. 2012; Furmannet al. 2013; Meslé et al. 2013). Thus from a biochemicalviewpoint, heavy hydrocarbons are expected to be susceptible tobiodegradation. This is supported by Gibbs free energycalculations, which show that potential energy availablethrough biodegradation is only marginally affected by hydrocar-bon size under most redox conditions (Fig. 3).

Observations of biodegradation in the field

Efforts to understand heavy hydrocarbon biodegradation throughlaboratory-based studies, performed under ideal conditions, do notaccurately reflect observations and experiences in the field. Manystudies have shown the rate and extent of hydrocarbon biodegrad-ation to be limited under field conditions. Hydrocarbons present insurface soils and subject to predominantly aerobic conditions areknown to degrade extensively (Kincannon 1972; Francke & Clark1974; Raymond et al. 1976) but complete removal is oftenunachievable. In these circumstances removal of lighter hydro-carbons is predominant, leaving behind a residue of heavyhydrocarbons. Huesemann (1995) found that about 30% heavysaturates and at least 75% of heavy aromatics remained from a rangeof crude oils and petroleum products after 1 year of landfarming. Ina longer study, Loehr et al. (1992) demonstrated that a recalcitrantfraction (20 – 24%) of an oily sludge remained after 2 years of landtreatment. Others have shown similar trends. Salanitro et al. (1997)tested the limits and extent of hydrocarbon biodegradation soilscontaining crude oils with heavy, medium and light API gravity (14,30 and 55 respectively). In agreement with other studies, arecalcitrant fraction was observed after 11 months of treatment.Heavy oil contamination, containing the greatest proportion ofheavy hydrocarbons, was biodegraded the least, with only 6 – 44%removed during the experiment. The large range in removal rateswas attributed to differences in non-hydrocarbon organic carbon

Table 1. The physical properties for heavy hydrocarbons compared with lighter saturated and aromatic hydrocarbons as described by TPHCWG (1997) andMcMillen et al. (2001)*

Water solubility(mg l−1)

Vapour pressure(atm)

Log Koc (cm3 g−1) H (cm3 cm−3) Retardation

factor, Rf†Fugacity levelI partitioning (%)‡

EC > 12-16 aliphatic 7.6 × 10−4 4.8 × 10−5 6.7 520 1.3 × 108 Air 1.8 × 10−2

Water 1.9 × 10−5

NAPL 70.0Soil 29.7

EC > 12-16 aromatic 5.8 4.8 × 10−5 3.7 5.3 × 10−2 1.3 × 105 Air 4.7 × 10−4

Water 4.9 × 10−2

NAPL 24.2Soil 75.8

EC > 16-35 aliphatic 2.5 × 10−6 9.8 × 10−7 8.8 4.9 × 103 1.6 × 1010 Air 5.4 × 10−4

Water 6.1 × 10−8

NAPL 88.0Soil 12.0

EC > 16-21 aromatic 0.65 1.1 × 10−6 4.2 1.3 × 10−2 4.0 × 105 Air 3.4 × 10−4

Water 1.5 × 10−2

NAPL 28.6Soil 71.4

EC > 21-35 aromatic 6.6 × 10−3 4.4 × 10−10 5.1 6.7 × 10−4 3.2 × 106 Air 1.9 × 10−6

Water 1.6 × 10−3

NAPL 38.9Soil 61.1

EC > 44 heavy hydrocarbon 1.0 × 10−4 4.1 × 10−12 8.7 4.1 × 10−8 1.3 × 1010 Air 1.2 × 10−14

Water 1.7 × 10−7

NAPL 74.4Soil 25.6

Benzo[a]pyrene 3.8 × 10−3 2.1 × 10−10 5.9 5.7 × 10−4 2.0 × 107 Air 1.4 × 10−7

Water 1.4 × 10−4

NAPL 4.6Soil 95.4

*The 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 non-polar b.p.gas chromatographic (GC) column (TPHCWG 1997).†Aquifer properties: effective porosity 0.25, fOC 0.025 to 0.0025 and bulk density 2500 kg m−3. Rf = 1 + (Kd × bulk density)/porosity.Koc is the soil organic carbon–water partitioningcoefficient. H is Henry’s Law Constant.‡Fugacity model level I generated as described by Pollard et al. (2008) with foc 0.025 and air/water/NAPL/soil ratio of 0.23:0.13:0.003:0.64 based on McKone (1996). Data for benzo[a]pyrene are shown for comparison only.

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content of the soils tested, with soils containing the greatest amountof organic carbon demonstrating the least amount of oil biodegrad-ation. McMillen et al. (2001) took the correlation between crude oilAPI gravity and extent of biodegradation a step further. Their workdemonstrated an inverse correlation between API gravity and extentof biodegradation, showing that heavy hydrocarbons are a majordeterminant of the final extent of oil removal from soils.

The potential for heavy hydrocarbon biodegradation in thesubsurface, where anaerobic conditions are prevalent, is much morelimited than in surficial soils (Grishchenkov et al. 2000). A numberof well-studied sites have demonstrated limited removal ofhydrocarbons many years after the original spill. A study byReddy et al. (2002) of subsurface soils revealed very little evidenceof biodegradation from aNo. 2 fuel oil spilled 30 years previously inWest Falmouth (MA), USA. Their work showed that much of the oilremained within the first 20 cm of the soil surface and that onlypartial degradation of the alkane fraction had occurred. A similarlack of biodegradation was observation by Wang et al. (1998) on a25-year-old crude spill in Alberta, Canada. Again, oil foundjust beneath the soil surface, at 10 – 40 cm, was relatively freshand showed little sign of biodegradation. The biodegradationof hydrocarbons in NAPL, where water, electron acceptorsand nutrients are limiting, supports the idea of slow hydrocarbonremoval rates in sub-optimal environments. Lundegard & Johnson(2006) found that hydrocarbons were removed at a rate of0.1 – 1.0 kg m−2 a−1 at a 3000 acre oilfield site in California, USA.This is equivalent to an average zero-order attenuation rate constant of1.5 × 10−3 a−1 and a half-life of over 400 years. Others have shownsimilar findings.McCoyet al. (2015) usedCO2 traps to estimateNAPLbiodegradation at a disused refinery. Based on their measurements,average rates of biodegradationwere found tobe 6.3 × 10−3 a−1 and thecalculated half-life was 108 years. These observations highlight thelong-term persistence of oil in the subsurface. As heavy hydrocarbonsare likely to be the most refractory to biodegradation, their persistencein the subsurface environment will be greater.

Factors causing heavy hydrocarbon persistence

Structure

The relative susceptibility of different classes of heavy hydrocarbonis not fully understood. Because thousands, or possibly millions, of

chemical compounds constitute the compositional fraction of heavyhydrocarbons (CONCAWE 2012), the extent of biodegradation isnot easily predicted and broad statements regarding biodegradationare more appropriate in describing the susceptibility of thesehydrocarbons to biodegradation. Based on information from lighterfractions they are expected to degrade in the following order (Leahy& Colwell 1990): n-alkanes > (branched) iso-alkanes > naphthenicand polyaromatic hydrocarbons > heteroatom-containing species.Asphaltenes, resins and highly condensed alkylated polyaromatics(with more than three rings) are more resistant to biodegradationthan other components owing to their condensed and fused aromaticring structures.

One aspect that may influence recalcitrance is physical size. It hasbeen proposed that heavy hydrocarbons are too large to bebiodegraded by intracellular mechanisms considered typical formicroorganisms (Lehmann & Kleber 2015). Overcoming thisobstacle, however, may be achieved through the production ofextracellular enzymes that ‘break down’ or convert large hydro-carbons into more metabolically manageable products. Forexample, white-rot fungi produce ligninolytic enzymes, such asperoxidases and laccases, that have activity towards hydrocarbons(Hammel 1995; Cerniglia & Sutherland 2010). Extracellularesterase and lipase are other classes of enzyme consideredcandidates for hydrocarbon biodegradation. Recent studiessuggest that they can hydrolyse both C–C and C–O bonds atmuch higher levels than previously thought (Alcaide et al. 2013).Work using Phanerochaete chrysosporium by Wang et al. (2009)showed that the same extracellular oxidases responsible for PAHbiodegradation were also excreted during the biodegradation of soilorganic matter. This is significant because it suggests that the sameenzymatic mechanism can be used to degrade heavy hydrocarbons.Further support for this mechanism can be found in studies on coalbiodegradation that have shown coals such as lignite can beliquefied through the action of extracellular enzymes produced bymicrobes (Fakoussa &Hofrichter 1999). Fungi are generally aerobicorganisms and require some oxygen to function (McGinnis &Tyring 1996). Therefore, their contribution to overall degradationrates under strongly anaerobic conditions may be questioned.Studies have shown, however, that several species (Aspergillus sp.,Trichocladium canadense and Fusarium oxysporum) are able todegrade PAHs under low-oxygen conditions (<1% O2) (Silva et al.

Fig. 3. Gibbs free energy calculations ofsaturated and aromatic hydrocarbonsunder various redox conditions. Gibbsfree energy values are from a range ofpublished data (McFarland & Sims 1991;Fuchs 1998; Spormann & Widdel 2000;Widdel & Rabus 2001; Weelink et al.2010; Mbadinga et al. 2011; Bose et al.2013; Acuna-Askar et al. 2015) plusadditional calculations-basedthermodynamic properties reported byAlberty et al. (1990).

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2009), suggesting that fungal catalysed degradation of hydro-carbons in oxygen-poor environments may be understated. Bacteriaare also sources of extracellular enzymes that may be active againstheavy hydrocarbons. For example, recent work by Zeng et al.(2016) demonstrated the oxidation of PAHs by a laccase derivedfrom Bacillus subtilis, and Shekoohiyan et al. (2016) showed thatPseudomonas spp. and Bacillus spp. produced peroxidases that areactive against hydrocarbons in a setup to treat waste.

Bioavailability

Microbial degradation of hydrocarbons requires physical contactbetween the microorganism and the hydrocarbon. Thereforebioavailability, or the amount of hydrocarbon that is accessible tothe microorganism, is a key factor in the overall effectiveness of thebiodegradation of hydrocarbons. Matrices, such as NAPL and soilorganic and mineral matter, that interfere with the contact betweenhydrocarbon and microorganism represents a major obstacle forbiodegradation (Brassington et al. 2007). Owing to their lowerwater solubility and vapour pressure and higher hydrophobicity,heavy hydrocarbons show a greater tendency to remain in NAPL orpartition into soil compared with lighter fractions. As such,bioavailability is relatively more important as a determinant ofbiodegradation of heavy hydrocarbon than that of lighter fractions.Partitioning of heavy hydrocarbons between NAPL and soil isdetermined by the concentration of oil present, the organic carboncontent of the soil and the hydrophobicity (Kow) of the hydrocarbon,and can be estimated using level I fugacity modelling (Table 1). Inmost circumstances, biodegradation will occur at the interface ofboth NAPL and soil matrices. In cases where hydrocarbon mass isdecreasing (biodegradation) or soil organic content is high,partitioning into soils becomes more predominant (supplementarymaterial). This conceptualization is important when considering theprocesses that influence bioavailability and biodegradationdescribed in the subsequent sections of this review.

Many hydrophobic soil contaminants show biphasic biodegrad-ation behaviour, whereby, in the initial phase, the rate of removal ishigh and is primarily limited by microbial degradation kinetics. Inthe second phase, the rate of removal is low and is generallyrestricted by mass transfer limitations (Barnier et al. 2014). In anexperiment using several PAHs in contaminated sediments, Kanet al. (1994) demonstrated extremely slow desorption of con-taminants compared with the initial sorption process. In this workonly 30 – 50% of the adsorbed pollutant could be desorbed. It isgenerally accepted that sequestration of hydrocarbons is caused bydiffusion into the soil particles and entry into soil micro- andnanopores (Reid et al. 2000). Subsequent desorption from the soilmatrix then becomes a rate-limiting step for biodegradation (Sempleet al. 2003; Trindade et al. 2005; Dungait et al. 2012). The mainsorbent for hydrophobic molecules is the organic matter of soils,and this fraction is assumed to be where hydrophobic moleculesbecome entrapped (Alexander 2000), although partitioning into themineral soil matrix is also possible (Reid et al. 2000). The age of thecontamination is also known to affect bioavailability, with sourcesof hydrocarbon becoming less bioavailable in soils with time(Barnier et al. 2014; Wu et al. 2014).

The interactions between bioavailability and biodegradation inoil-contaminated soils have been well studied. Efforts by Oleszczuk(2007, 2009) demonstrated that changes in soil organic matter(SOM) composition can result in significant changes in thebioavailability of adsorbed hydrocarbons. It was proposed that aweakening of the sorption between organic matter and hydrocarbon,through mineralization, results in increased bioavailability. Thisfinding was further developed by others. Wu et al. (2013b,c) addedvarious composts to contaminated soils to increase organic carbonmatter. They found that after an initial period of high sorption to the

SOM, bioavailability and biodegradation of PAHs increasedsignificantly. These increases were partly attributed to the presenceof humic acids that were probably generated during the compostingprocess. Humic acids are considered important because they act as acarrier to provide mass transport (Wu et al. 2013b) and increasehydrocarbon bioavailability (Smith et al. 2011). Modelling usingmolecular dynamic (MD) simulation has contributed further tounderstanding. Wu et al. (2015) assessed the intricate interactionsbetween heavy hydrocarbons and soil at the molecular scale. Theydemonstrated that although soil organic matter (humic acid)increased the overall hydrocarbon adsorption capacity of the soil,it also increased mobilization and bioavailability by driving oil fromthe NAPL and mineral soil phases. It is this mechanism, termedcompetitive sorption, that probably describes observations thatincreased organic matter enhances the biodegradation of heavyhydrocarbons in soils (Liu et al. 2017). For example, Tang et al.(2012) found that the addition of humic acid, dosed at 0.02 g kg−1,resulted in an increase of aromatic and polar (asphaltene and resin)biodegradation after 210 days of incubation and Gao et al. (2015)demonstrated that the presence of low molecular weight organicacids increased the bioavailability of PAHs in soils. Interestingly,low molecular weight organic acids, derived from plants, are alsoknown to increase removal of hydrocarbon in soils (Martin et al.2014), suggesting that phytoremediation may facilitate the remedi-ation of heavy hydrocarbons.

Several studies have shown a correlation between extent ofhydrocarbon biodegradation and laboratory-based methods tomeasure bioavailability in soils (Rhodes et al. 2008; Dandie et al.2010; Adetutu et al. 2013). Most make use of the cyclodextrinextraction to determine bioavailability of hydrocarbons, althoughother methods, involving mild organic solvents or solid-phaseadsorbents such as XAD resin or Tenax are available. Further workwith high molecular weight PAHs has suggested a direct correlationbetween the rapidly desorbing fractions and the extent to whichremoval by biodegradation or chemical treatment is feasible(Richardson & Aitken 2011; Lemaire et al. 2013; Barnier et al.2014). This correlation, however, remains to be validated, as doother forms of the technique, for heavy hydrocarbons. For example,Dandie et al. (2010) showed that a number of bioavailabilitytechniques overestimated the biodegradative propensity of a C37–40

hydrocarbon fraction by up to a factor of three, and Wu et al.(2013b) highlighted the potential inefficiencies of cyclodextrin as amethod for predicting the bioavailability of PAHs with a highmolecular size or polar substituents. Although this predictive toolshows promising signs, more research is required to validate its usefor heavy hydrocarbons under a range of conditions. Indeed, Umehet al. (2017) suggested that the most appropriate bioavailabilityapproach is still to be determined. One possible explanation forthese observations can be found in work by Huesemann et al.(2004), who demonstrated that in some cases hydrocarbons thatwere bioavailable in soils did not undergo complete biodegradation.It was suggested that this phenomenon was due to microbial factorssuch as lack of cometabolic substrates or insufficient numbers ofhydrocarbon-degrading populations.

Nutrients and water availability

Availability of inorganic nutrients, particularly nitrogen andphosphorus, is often an important control on hydrocarbondegradation in soils and in the subsurface. These nutrients aregenerally needed to support microbial biomass production thatoccurs concurrently with the biodegradation of hydrocarbons. Therequirement for nutrients has been shown in both surface soils andthe subsurface. Coulon et al. (2012) and Brown et al. (2017) bothdemonstrated that the addition of inorganic nutrients stimulatedbiodegradation rates of hydrocarbons in aerobic landfarmed soils.

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Similarly, Essaid et al. (2011) identified nutrient limitation to be oneof the main factors controlling biodegradation in subsurfacecontamination at Bemidji, USA, and Leewis et al. (2013) showeda similar impact in the colder climate of Alaska, USA. Manyhydrocarbon-contaminated soils become hydrophobic, whichreduces water retention in soils (Adams et al. 2008). As water isnecessary for microbial activity, a lack of it will have a negativeimpact on the biodegradation of hydrocarbons. The higherpropensity of heavy hydrocarbons to bind to soils suggests thatthey are major determinants of the process that drives soils tobecome more hydrophobic.

Toxicological impact of heavy hydrocarbons

Humans can be exposed to hydrocarbons by inhalation, by oraluptake and via dermal contact. From a mobility perspective, heavyhydrocarbon will not move significantly from the area of release viagroundwater. In addition, these materials are not volatile so exposureby inhalation would not be expected. Therefore, exposure to heavyhydrocarbons would be limited to direct contact by oral or dermalroutes. Basic physical and chemical properties suggest that heavyhydrocarbons are not bioavailable and therefore are unlikely to exerttoxic effects.As a consequence, there is a lackof toxicological studieswith heavy hydrocarbons. It is generally accepted that hydrocarbonswith carbon numbers ≥C35 are not bioavailable because they cannotbe absorbed via biological membranes (CONCAWE 1993;EFSA Panel on Contaminants in the Food Chain (CONTAM)2012). As such, both oral and dermal exposure is not likely to be anissue with heavy hydrocarbons. A survey of dermal uptakecoefficients for petroleum hydrocarbons was performed recently toassess the toxicity of petroleum hydrocarbons (Jakasa et al. 2015).This work demonstrated that models used to predict uptake ofhydrocarbons tended to over-predict uptake of smaller molecules butwere good at predicting the penetration of heavier products. Whenthese models were applied to heavy hydrocarbons, they predicted nodermal penetration whatsoever. This suggests that acute or chronictoxicity of heavy hydrocarbon through dermal exposure is expectedto be negligible. In a 90 day oral repeat dose study on a heavy gas-to-liquid oil, which is predominantly composed of linear, branched andcyclic alkanes C40–70, there was, as expected, a total lack of toxicity(Boogaard et al. 2016).

Ecotoxicological impact of heavy hydrocarbons

The ecotoxicity of petroleum hydrocarbons in soil has been studiedusing a range of species, including bacteria, earthworms and plants(Table 2). Factors that are known to have an impact on toxicity effectsinclude the type and quantity of oil, the length of weathering (extentof biodegradation) of the oil and the type of soil (Chaîneau et al.1997, 2003; Dorn et al. 1998). Most studies have been undertakenusing whole product mixtures or crude oils in which the proportionof heavy hydrocarbon varies. These studies have shown thatproducts containing higher amounts of heavy hydrocarbon aremuch less toxic than those containing lighter fractions (Table 2). Forexample, Dorn et al. (1998) demonstrated that the no observed effectconcentration (NOEC) in seed germination tests was up to 19.5%(195 000 mg kg−1) for heavy crude oil (API 17). This contrastedsignificantly with light crude (API 55), which showed an NOEC of2.4% in the same test. Chaîneau et al. (1997) compared LD50

(median lethal dose) values for light and heavy fuel oil fractionsusing Zea mays germination tests. The work clearly showed theheavy fuel product to be less toxic than the corresponding lightfraction. Direct toxicity of hydrocarbons to organisms is stronglydependent on bioavailability (Peterson 1994). It is manifested in theorganism once a critical body burden is attained, thus the organism–

water partitioning (with octanol–water partitioning used as a

surrogate) of the constituents is an important aspect of the process(van Wezel et al. 1995). Single hydrocarbons have differentorganism–water partition coefficient values that varywithmolecularweight, class and molecular structure. Previous work has suggestedthat the bioavailability and toxicity of compounds such as heavyhydrocarbons that have extremely high log Kow (>6.0) values arelimited (Verbruggen 2004; Redman et al. 2012). Understanding ofhydrocarbon ecotoxicity has been developedmainly through studieson aquatic systems. For example, aromatic extracts of vacuumdistillates, predominantly composed of alkylated polyaromatichydrocarbons (C20–50) show no acute or chronic toxicity in testson aquatic organisms (API 2009). Similar observations werereported for lubricating oil base stocks, which are predominantlycomposed of saturated aliphatic hydrocarbon constituents (C15–50)(API 2009). These observations are consistent with modelledtoxicity predictions for heavy hydrocarbons using the PETROTOXTarget Lipid Model (Redman et al. 2012), which shows thathydrocarbon constituents with octanol–water partition coefficient(Kow) values of >6.0 are predicted to be non-toxic to aquaticorganisms. Recent work extending the PETROTOX to soil andsediment organisms using equilibrium partitioning (EqP) theoryalso demonstrated that for hydrocarbons with logKow values >6.0 anincreasing incidence of no observed toxicity consistent with thedataset for aquatic organisms was observed (Redman et al. 2014).

Recently, there has been interest in metabolic intermediateproducts of hydrocarbon biodegradation (Bekins et al. 2016).These components, typically more polar and soluble than theirparent hydrocarbon, are likely to have higher mobility in soil andgroundwater (Boll et al. 2015). The human and ecological risk ofthese metabolic intermediates is unclear at this point. Bekins et al.(2016) suggested that metabolites in groundwater can pose a risk toaquatic and mammalian species, whereas others such as Zemo et al.(2016) indicated they had low toxicity to humans and were notexpected to pose significant risk to human health. Studies ingroundwater lag behind the work in marine environments, whichhave been stimulated by the Deepwater Horizon oil spill in 2010.Reddy et al. (2012) was the first to show that polar compoundsaccounted for a significant fraction of the oil released into the Gulf ofMexico. These compounds, generated through biodegradation andphoto-oxidation, have been suggested to bemore persistent and toxicthan their parent compound (Aeppli et al. 2012; Chen et al. 2016).Naphthenic acids, have been identified as one potential source ofthese metabolites with several studies indicating that di, tri and tetra-cyclics become increasingly predominant during biodegradation(Angolini et al. 2015; Kim et al. 2005). Whether the biodegradationof heavy hydrocarbons in terrestrial environments would yield moreor less toxic compounds is not known, but would depend on theirresulting structure and properties that govern hydrophobicity (i.e. logKow). Heavy hydrocarbons contain numerous compound classeswith a wide range of carbon number distributions. Therefore,addition of oxygen during biodegradation alters each compound’shydrophobicity and toxicity in a potentially different way. The vastmajority of heavy hydrocarbon constituents have log Kow valuesconsiderably greater than 6.0, which is the approximate cut-off fortoxicity of hydrocarbons to aquatic, sediment and terrestrial dwellingorganisms (Redman et al. 2012, 2014) Thus, significant changes inoverall polarity, through increased heteroatom (N, S, O) content or areduction in molecular mass, are necessary before heavy hydrocar-bon metabolic intermediates can be expected to influence toxicity.

Issues related to heavy hydrocarbon contamination

Surface releases of crude oil can negatively affect local flora andfauna and cause severe loss of vegetation of the area affected by thespill. The natural rehabilitation of contaminated soils may take timeto accomplish and it may take months or years for plant growth to

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return. The process can be further delayed by the residual, non-toxiceffects of oil that remains in the soil. For example, water repellencyreduces the rate of wetting and retention of water in soils which inturn will create a less than ideal environment plant growth (Roy etal. 2003). Water repellency reduces the rate of wetting and retentionof water in soils, which in turn will create a less than idealenvironment for plant growth (Roy et al. 2003). This process hasbeen observed in soils of all textures as well as organic-rich soils,but there seems to be a greater probability of it developing in sandysoils (Doerr et al. 2000). Oil contamination is one of many ways inwhich soil can become hydrophobic and repel water. Waterrepellency may also develop after episodes of fire, through thebreakdown of plant matter or the action of microbes (Doerr et al.2000). Its prevalence in agricultural soils has prompted thedevelopment of strategies to overcome this phenomenon usingwetting agents (Hallett 2008). Despite success in agriculture, thisapproach is not likely to be useful for oil-contaminated soils, wherethe risk of remobilizing subsurface oil would be a source of concern.Alternative approaches to treat water repellency in hydrocarbon-contaminated soils have been explored with some success recently.Adams et al. (2016) found that washing with alkali (0.1M NaOH)significantly reduced water repellency in soils. In other work, Cann(2000) showed that addition of clay had a similar impact in reducingwater repellency. Because of the simplicity of these methods theylend themselves to implementation under field conditions and it isworth monitoring their progress in the future.

The formation of ‘tar crusts’ on the surface of soils is anotherphysical impact that can hinder the rehabilitation of an area (Fig. 4).These crusts are formed by either extensive weathering (Volk 1980)or burning of oil through artisanal refining. In some cases, tar crustscan exist for decades (UNEP 2011). They form a physical barrierthat can smother the soil, reduce the ingress of oxygen and limit theinitial rooting of plants (Coulon et al. 2012), thus make re-vegetation very difficult. For example, Murygina et al. (2016)described the occurrence of a large asphalt crust in Siberia, Russia.It could withstand a weight up to 60 kg without being destroyed andhad zero plant growth 20 years after the spill. There are other similarexamples across most oil-producing regions in the world.Eliminating the physical barrier caused by these crusts, byhomogenization or removal, is likely to be the simplest means ofstimulating natural rehabilitation of the affected area. It should benoted, however, that these crusts are usually associated with lighterhydrocarbons, which have a different risk profile and may requireadditional consideration during treatment.

Management of sites contaminated by heavyhydrocarbon

Risk assessment is an established requirement for effectivemanagement of contaminated land, and is now a widely usedsupport tool for environmental management decisions. Heavyhydrocarbons are unlikely to be major risk drivers duringcontaminated land management but need to be consideredbecause they will generally be a component of the larger mixtureof hydrocarbons. Fugacity modelling (Pollard et al. 2008; Coulonet al. 2010) demonstrates that there are limited exposure pathwayswith heavy hydrocarbons (Table 1). Based on these assumptions themost relevant exposure scenario is via contact with hydrocarbon-affected soils. Remedial actions for soil clean-up may be triggeredby soil hydrocarbon concentrations greater than calculated orarbitrary screening values or through non-risk-based measures suchas stakeholder concerns over soil aesthetics or impacts to soiltexture.

A shift to sustainable remediation practices recently (Bardos et al.2011) has pushed soil treatment options towards active biologicaltreatments such as landfarming or biopiling. These approaches use

Tab

le2.

Selected

ecotoxicitycharacteristicsof

hydrocarbons

insoils

Product

Com

positio

nEnd-point

Value

Com

ment

Reference

Crude

oil,Villeperdue

field

Fresh

oil:48%

saturates;26%

arom

atics;26%

resins

andasphaltenes

Inhibitio

nof

seed

germ

ination

(%)

100%

Oildosedto

1.5%

(w/w)in

soil;

5plant

speciestested

(Chaîneauetal.

2003)

Worm(Eiseniafetidia)

mortality

(%)

100%

Oildosedto

1.5%

(w/w)

Rem

ediatedoil(480days):6%

saturates;15%

arom

atics;79%

resins

andasphaltenes

Inhibitio

nof

seed

germ

ination

(%)

0.3%

Zea

mays;10%

Lactuca

sativa

Oilconcentration0.4%

(w/w)

Worm(Eiseniafetidia)

mortality

(%)

0%Oilconcentration0.4%

(w/w)

Fueloil,lig

htfractio

nUnknown(fresh)

Seedgerm

inationLC50

0.3%

(w/w)fueloilZea

mays

(Chaîneauetal.

1997)

Fueloil,heavyfractio

nUnknown(fresh)

Seedgerm

inationLC50

6%(w

/w)fueloilZea

mays

Heavy

crudeoil

Fresh

oilAPI17:20%

saturates;29%

arom

atics;44%

resins

and

asphaltenes

Worm(Eiseniafetidia)

mortality

LC50

6.3%

(w/w)oilin

soil

(Dornetal.1

998)

Microtox©

mortalityLC50

>10%

(w/w)oilin

soil

Seedgerm

inationNOEC

4.8–19.5%

(w/w)oilin

soil

5plantspeciestested

Light

crudeoil

Fresh

oilAPI55:87%

saturates;6%

arom

atics;1%

resins

and

asphaltenes

Worm(Eiseniafetidia)

mortality

LC50

0.73%

(w/w)oilin

soil

Microtox©

mortalityLC50

2.6%

(w/w)oilin

soil

Seedgerm

inationNOEC

<2.4–2.4%

(w/w)oilin

soil

5plantspeciestested

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biodegradation as the principal form of contaminant destruction andthus require conditions that are favourable to microbial growth. Onemajor limitation of this remediation option is that it is less effectivewith soils containing high levels of contamination (>50 000 mghydrocarbon kg−1 soil) (Kuppusamy et al. 2017). In thesecircumstances innovative alternatives have been proposed. Forexample, Wu et al. (2013a) carried out a laboratory trial of acombined solvent extraction and bioremediation process to reduceoil contamination in soil from Shengli Oilfield, China. They foundthat this process reduced contamination from 140 000 to<4000 mg kg−1 but required the use of bioaugmentation (Bacillussubtilis FQ06) and biosurfactant (rhamnolipid) for full effective-ness. Whether this process competes on a cost-effective basis atlarger scale remains to be seen, but it does highlight efforts todevelop strategies to tackle highly contaminated soils. For in-situtreatment of heavy hydrocarbons, smouldering technologies, such

as Self-sustaining Treatment for Active Remediation (STAR), haveproven successful recently and have effectively treated subsurfaceheavy coal tar and crude oil contamination (Pironi et al. 2011;Scholes et al. 2015).

A number of tools are now available to assist with determiningrisk-based remediation end-points and timeframes. Level II fugacitymodelling has been used successfully to predict the removal ofhydrocarbons from constructed biopiles (Coulon et al. 2010). In thiswork partitioning propensity and fate of hydrocarbon fractions weremodelled to produce time explicit data that adequately predictedhydrocarbon fate during the remediation. There is growingawareness that determination of remediation end-points shouldaccount for the bioavailability of hydrocarbons in soils (Harmsen &Naidu 2013; Kördel et al. 2013; Umeh et al. 2017). From a riskassessment and bioremediation perspective it is the bioavailablehydrocarbon fraction that is important rather than total concentration

Fig. 4. Tar crusts formed after extensive weathering of crude oil (a–c) or the undistillable waste products after artisanal refinement (d). Scale bar = 0.1m.

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(Guo et al. 2016). Wu et al. (2013c) demonstrated the applicabilityof various machine learning models to predict changes in PAHbioavailability, based on soil physicochemical properties, duringcomposting of contaminated soils (Wu et al. 2013c). These modelswere found to be useful for determining remediation end-pointsbased on bioavailability and could be applied to other remedialsituations with relative ease. In similar work Harmsen & Naidu(2013) used Tenax extraction to predict PAH biodegradation basedon various levels of bioavailability. Further developments inmodelling are considering the interactions between factors thatinfluence remediation and their relative contribution to predictingremediation success (Wu et al. 2014). The bioavailability ofhydrocarbons has substantial ramifications for the remediation ofcontaminated soils. If it can be demonstrated that greater levels ofcontamination can be left in soil without additional risk, lower costsand smaller remediation volumes could be achieved (Wu et al.2013b). There is evidence, however, that bioavailability is not fixedin time. Changes in environmental conditions can affect thebioavailability of hydrocarbons, as demonstrated in work by Wuet al. (2013b). This indicates that a better understanding of theimpacts of bioavailability of hydrocarbons is required to fullycomprehend the risks associated with residual contamination. Mostof the current tools used for assessing risk posed by hydrocarbonsthat determine the end-point of remedial activities are based on totalextractable concentrations and do not account for bioavailability.Clear framework guidance documents to support wide regulatoryadoption of the bioavailability concept have recently been published(reviewed by Umeh et al. 2017). There has been some limitedrecognition by European-based environmental regulators (inGermany, the Netherlands and the UK) that bioavailability can bean aspect of contaminated land risk management, but generallyregulatory and public acceptance is still limited (Harmsen & Naidu2013).

Conclusions and future perspectives

In most circumstances, the unintentional release of heavy hydro-carbons will result in soil contamination and, depending on thevolume released, the formation of free phase NAPL. Given the lowvolatility and solubility, the only attenuation mechanism ofsignificance is biodegradation, which, from thermodynamic con-siderations, can be expected to occur under different redoxconditions. Field evidence shows that biodegradation is dependenton the prevailing environmental conditions and that heavyhydrocarbon contamination can persist over long time periods inthe subsurface (years to decades). Recent developments inmodelling have demonstrated that humic acids can influence thepartitioning of heavy hydrocarbons in NAPL and soil. Experimentalobservations support this and have demonstrated that humic acidsand soil organic matter can improve the bioavailability andbiodegradation of heavy hydrocarbons. Although the low bioavail-ability of heavy hydrocarbons limits their toxicity, sites contami-nated with heavy hydrocarbons may still require action to overcomethe physical effects they may have on soils, including increasedhydrophobicity, negatively affected soil texture, or the formation ofoil crusts limiting oxygen diffusion into the soil.

Overall, it is concluded that further research is needed tounderstand the fate and transport of heavy hydrocarbons. Thisincludes developing a better understanding of heavy hydrocarbonbiodegradation, which is required to determine their ultimate fateand risks. This will assist with the development of more effectiveremediation strategies, provide a scientific basis for establishingremediation end-points and help shed light on whether heavyhydrocarbons are a significant source of metabolic intermediatesthat change the risk profile of contaminated soils or become mobilein groundwater. Recent analytical and modelling techniques can

contribute to this goal. Improved analytical capabilities willcontribute to characterizing the composition of heavy hydrocarbonfractions and degradation intermediates. More detailed under-standing of heavy hydrocarbon composition is needed to developmore realistic physical and chemical properties for this highlyheterogeneous fraction. These data can be used to developmodelling methods that encompass the overall biodegradationroute to establish science-based remediation end-points includingbioavailability for heavy hydrocarbons. The combination of higherresolution analytical data with process-based models can also helpto identify bottlenecks in biodegradation and potentially open upnew sustainable remediation options for soils that contain a highproportion of heavy hydrocarbon. This includes understanding anddeveloping strategies to overcome some of the gross contaminationissues discussed in this review (soil hydrophobicity, oil crusts, soiltexture) that may affect the rehabilitation of the contaminated soil.

Acknowledgements The authors acknowledge and thank several internaland external reviewers, including Matthew Lahvis, who provided insightfulcomments and guidance throughout the writing of this document.

Funding This work was funded by Shell Global Solutions International BV.The views expressed in this paper are those of the authors and may not representthe policy or position of Shell Global Solutions International BV.

Scientific editing by Stephen Buss; Jonathan W.N. Smith

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