Typical oil discharge scenarios include process losses or accidental releases from exploration, production, processing or transport related activities. However, the risk of accidental oil discharges will always be present in an environment with oil-related activities. In addition, the diminishing Arctic ice zones will likely increase ship activity through the Northwest and Northeast passages, activities which may contribute to an increased risk of oil discharges in Arctic marine environments. Biodegradation of oil compounds is regarded as the most complete process, as it is able to completely remove oil compounds, through the mineralization to carbon dioxide and water. In particular, two large oil spills in marine water have drawn public attention to oil biodegradation issues in general, the Exxon Valdez grounding in Prince William Sound in 1989 and the recent Deepwater Horizon blowout in 2010, both recently reviewed by Atlas and Hazen (2011). Media attention on the deep water plume from the Deepwater Horizonblowout has resulted in increased focus on the potential effects of oil spills in cold water environments. 

Physico-chemical properties of the spilled oil will affect biodegradation. Crude oils may be separated into 4 main types; paraffinic, asphalthenic, naphthenic and wax-rich (Moldestad et al. 2003). For instance, paraffinic oils will have a distinct chromatographic n-alkane pattern and a high content of light compounds like BTEX and naphthalenes. These oils therefore contain a high degree of easily biodegradable compounds. However, naphthenic oils are often biodegraded in-reservoir (Head et al. 2003), and in these oils most of the easily biodegradable compounds may have been consumed before oil production.  Asphalthenic oils are rich in poorly biodegradable asphaltenes and resins, while wax-rich oils will often have a high pour point and may be solidified in cold water, a physical factor that will affect their availability for biodegradation.  Marine fuel oils may also be spilled.

When oil is discharged to the marine environment a number of weathering processes occur:

Volatile compounds with low boiling points (e.g. saturates up to nC₁₁, mono- and some diaromatic hydrocarbons) are rapidly evaporated after surface spills. These are compounds that are normally rapidly biodegradable in the water column, but evaporation is normally more rapid than biodegradation after a surface spill. Evaporation is slower in cold than in temperate seawater (Brandvik et al. 2005), and this may result in temporarily higher concentrations of volatile toxic compounds (e.g. BTEX) in the seawater. At high concentrations these compounds may prolong microbial lag-phases and delay the onset of biodegradation (Atlas and Bartha 1972; Hokstad et al. 1999), although it is not known if this will have an impact on biodegradation under field conditions. In subsurface releases these volatiles are rapidly dissolved from dispersed oil and we suspect may be biodegraded rather than evaporated. Evaporation also results in increased viscosity of the residual oil (Faksness 2008), which will negatively affect the ability of oil to disperse, thereby slowing biodegradation.

Components dissolved from the oil phase are available for biodegrading microbes in the water column. In cold seawater the dissolution of oil compounds is decreased compared to temperate water (Faksness, 2008). The typical oil compounds in a water-soluble fraction (WSF) from fresh oils include phenols, naphthalenes and 2-3 ring PAHs. In addition, the WSF contains considerable amounts of highly polar compounds with nitrogen, sulphur, and oxygen atoms in their structures (so-called NSO compounds), often present as a chromatographic "hump", termed the "unresolved complex mixture" (UCM). In a study of WSF from an in-reservoir biodegraded oil (Troll) approximately 70 % of the WSF was separated by preparative high-pressure liquid chromatography, into a polar fraction (Melbye et al. 2009).  The non-polar compounds of the WSF are often considered to be rapidly biodegraded in the marine environment (Brakstad and Faksness 2000), and biodegradation of these compounds may result in a significant increase in the UCM  concentration relative to other crude oil components (e.g. Meredith et al. 2000), highlighting their persistence (Han et al. 2008).

Photooxidation is an important process in degrading and transforming crude oil compounds after release to the environment. The polar region exhibits vast seasonal differences in light conditions, and as a result photooxidation varies significantly between the polar summer and winter. UV-irradiation of crude oils has shown that aliphatic compounds are mainly resistant to photodegradation, while aromatic compounds appear particularly sensitive to this process (Maki et al. 2001). In contrast to biodegradation, increased size and alkyl substitution result in increased sensitivity of aromatic hydrocarbons to photochemical oxidation. As photooxidation leads to the inclusion of oxygen atoms in the structures of these compounds, photooxidized products appear mainly in the polar resin fraction of the oil (Maki et al. 2001; Garrett et al. 1998; Prince et al. 2003). Additionally, the average molecular weight of oil compounds is reduced and the oxygen content increased.  Studies have also shown that the photooxidized compounds subsequently exhibit increased susceptibility to biodegradation (Dutta and Harayama 2000; Maki et al. 2001; Ni'matuzahroh et al. 1999). Consistent with the physico-chemical properties of the photooxidized compounds, both the dissolved organic carbon concentration and acute toxicity of the water-soluble fraction of oil increased during the irradiation period (Maki et al. 2001). Thus, photooxidation results in a greater proportion of oxidized compounds that exhibit increased water-solubility and subsequently more significant impacts on toxicity and biodegradation. However, investigation of the relationship between photooxidation, biodegradation and toxicity would be of interest as part of the fate-determination of different oil compounds during the Arctic summer.

In shallow seawater and at higher levels of suspended sediments (e.g. after a storm), sediment particles may adhere to the oil and sink to the subtidal seabed sediments. Oil spills may also drift to shore and be mixed into the intertidal sediments. Oils mixed into the sediments will be subject to microbial processes in the sediment (Refer to Section 3). If seawater replenishment is poor, aerobic processes may consume most of the oxygen, resulting in anoxic conditions. Anaerobic biodegradation of hydrocarbons, by several alternative mechanisms, will occur in the absence of oxygen, as reviewed by Heider (2007).  Genes associated with anaerobic hydrocarbon degradation (e.g. benzyl- and alkylsuccinate synthase genes) have been detected in hydrocarbon-contaminated sediments (Callaghan et al. 2010).  

Water uptake into the spilled oil may cause the formation of viscous and often stable water-in-oil emulsions. Emulsions have been shown to be poorly biodegradable (Brakstad et al. 2011; Cook et al., 2011). Water taken up in the emulsions may contain oil-degrading microbes, but if water is trapped in the emulsions this will not promote biodegradation on the bulk oil as the emulsions may be depleted in essential nutrients. 

With sufficient energy from wave action the oil may break up into droplets in the water column. If oils are easily dispersed, small droplets are generated. The rising or settling rate of the droplets is related to the size and specific gravity of particles. As an example droplets 100 µm in diameter and a lower specific gravity that the surrounding seawater have been observed to rise with a velocity of approximately 1.5 m/h. Thus, larger droplets will resurface rapidly and thin oil films (sheens) may be formed. Oil dispersion is important for biodegradation. Dispersible oils will generate relatively small droplets, resulting in large surface areas for bacterial attachment. For instance, fresh Louisiana Sweet Crude oil can be made to generate dispersions with a median droplet size of 50-150 µm under continuous breaking wave conditions in an oil-on-seawater flume experiment (Brakstad et al. 2011). Several biodegradation studies of dispersed oil in cold seawater (5-8°C) have shown bacterial colonization of oil-droplets and biodegradation of dispersible oils (e.g. Lindstrom and Braddock 2002; MacNaughton et al. 2003; Venosa and Holder 2007; Prince et al. 2012). This colonization may tend to generate flocs of oil and biomass (MacNaughton et al. 2003; Bælum et al. 2012). However, for waxy oils with high pour points, evaporation, dilution and dispersion may be reduced in cold seawater, since precipitated wax may form a matrix which limits internal mixing and acts as a diffusion barrier between the oil and the water (Faksness 2008).

As described above, surface and resurfaced oil may generate thin films on the sea surface. In a series of studies with thin oil films immobilized on hydrophobic adsorbents, n-alkanes in these films were rapidly biodegraded in temperate and cold seawater (0-13°C), while aromatic compounds were subject to mixed dissolution and biodegradation (e.g. Brakstad and Bonaunet 2006; Brakstad et al. 2004). In experiments over a period of 112 days with different oil thicknesses of a wax-rich oil it was apparent that a thickness limit for measurable biodegradation (nC₁₇/Pristane and nC₁₈/Phytane) was between 0.1 and 1.0 mm in cold (6-10°C) seawater (Brandvik et al. 2006). 

In the aftermath of the Deepwater Horizon incident, a large body of new information has been collected and integrated with our already existing understanding of the microbial response to oil spilled in the marine environment (Hazen et al. 2010; Mason et al. 2012; Valentine et al. 2012). In general, in situ sampling and analysis revealed unexpectedly rapid disappearance of released oil in the Gulf of Mexico environment, which is characterized by a temperate climate (Hazen et al. 2010). This rapid disappearance was affected by the prevalence of water-soluble constituents in the crude oil (Reddy et al. 2012), injection of subsea dispersant into the erupting oil flow (Kujawinski et al. 2011), and presence of indigenous oil-degrading microorganisms in this area that is well known for natural seeps of crude oil from reservoirs (Lu et al. 2012). Such indigenous oil-degrading microorganisms are the topic of this section.

Following the Deepwater Horizon incident, extensive analysis of microbial responses was done both in situ and in laboratory microcosms. These analyses support, in general, a paradigm of successive blooms of taxonomically distinct indigenous microbial populations as the oil weathers and labile components are sequentially degraded leaving less-readily degraded components to feed subsequent blooms (Hazen et al. 2010; Valentine et al. 2010; Kostka et al. 2011; Baelum et al. 2012; Beazley et al. 2012; Lu et al. 2012; Mason et al. 2012; Valentine et al. 2012).

Conditions are very different in high latitude marine environments.  As described in previous sections, the Arctic and Antarctic marine environments are characterized by seasonal extremes of photoperiod, spatial variability in salinity and temperature, as well as generally colder surface temperatures compared to the temperate latitudes. These differences may result in different expectations about the rate of oil degradation, as described in previous sections.  They also result in different expectations about the indigenous populations of oil degrading microorganisms.

As mentioned in previous sections, microbial responses to oil in marine environments generally are dominated by bacteria rather than archaea (Roling et al. 2004). Although fungi are known to degrade petroleum compounds in some marine settings (Zinjarde and Pant 2002), few surveys of fungal abundance in high latitude marine environments have been done (Butinar et al. 2011) and thus far none have addressed oil degradation by fungi in high latitude environments. For these reasons, this section focuses on the bacterial component of the marine microbiological community. 

Among the bacterial taxa catalogued in high latitude marine environments, many appear to be specific to that environment (Ghiglione et al. 2012; Sul et al. 2013).  This apparent specificity may be due to truly unique populations, or it may be a function of the limit of detection.  Community members that thrive in the high latitude marine environment grow to relatively high cell densities and are therefore more easily detected.  Various investigations have found that microbial species richness curves are not saturated with typical levels of effort.  This finding has led to the hypothesis that there is an under sampled “rare biosphere” of organisms with low population density (Sogin et al. 2006) that, despite low population levels, can respond to changes in environment and energy source.  This phenomenon may be typified by the explosion of Oceanospiralles and Colwellia populations in the presence of different partitions of spilled oil during the Deepwater Horizon incident (Hazen et al. 2010; Bælum et al. 2012).

Marine ice represents an extreme biosphere with below-zero-centigrade temperatures and high salt concentrations. It has been demonstrated by field studies that bacterial populations in Arctic marine ice are affected by oil pollution, stimulating species of a few genera like Colwellia, Marinomonas and Glaciecola (Brakstad et al. 2008). 

Not all of the microorganisms found in the Arctic oceans are adapted to that environment.  The various currents carry viable microorganisms from diverse locations to the Arctic (Rosnes et al. 1991; Hubert et al. 2009; Hubert et al. 2010); thus, there is an expectation of cosmopolitanism among the free-living microorganisms.  This is not to say that the population structure is homogenous as if the Arctic were a giant mixing bowl.  In fact, there is documented variability in population structures, with different communities associated with water masses of different origins (Galand et al. 2010; Sul et al. 2013).   The presence of non-adapted microorganisms such as thermophiles does, however, indicate that microbial populations adapted to the consumption of natural or human-induced oil releases might be transported to and be present in areas that are not commonly exposed to oil.

Crude mineral oil is degradable by indigenous microorganism populations in the Arctic marine environment, even at near-freezing temperatures (Brakstad and Bonaunet 2006), although at slower rates compared to higher temperatures (Margesin et al. 2003; Michaud et al. 2004).  Nevertheless, over a time course on the order of weeks substantial biodegradation can be observed in nutrient-enriched cold Arctic seawater (Brakstad and Bonaunet 2006). Community analysis of oil-degrading Arctic microbial consortia indicated that several taxa of bacteria are involved in biodegradation in this environment, including genera related to Pseudoalteromonas, Pseudomonas, Shewanella, Marinobacter, Psychrobacter, and Agreia (Deppe et al. 2005).  Of interest, these are different organisms from those directly associated with degradation in the Deepwater Horizon spill in the Gulf of Mexico, specifically bacteria of the orders Oceanospiralles (Hazen et al. 2010; Kostka et al. 2011) and Alteromonadales (Bælum et al. 2012), among others.

Linear alkanes often are characterized as an easily accessible carbon source, either through degradation or direct incorporation into microbial biomass, in the marine environment (Harayama et al. 1999).  The metabolic pathways for linear, branched, and cyclic alkanes have been studied and described since the 1960s (Jobson et al. 1972, Coates et al. 1997, Feng et al. 2007, Rojo 2009, Gray et al. 2011).  Preferential degradation of short-chain alkanes (represented by C₁₅) over long-chain alkanes (represented by C₂₆) was observed in situ in a deep plume (circa 1,400 m) in the Gulf of Mexico under aerobic conditions (Hazen et al. 2010).  Furthermore, during weathering in subsurface petroleum reservoirs, alkyl chains on substituted soluble PAHs such as alkane-substituted naphthalenes may be transformed even more rapidly than linear alkanes (Jones et al. 2008).  Whether this phenomenon, observed in anaerobic subsurface reservoirs, would occur in the presence of petroleum hydrocarbons released into the deep sea, remains unknown. 

The specific bacteria known to accomplish alkane degradation are numerous (Whyte et al. 1997; Rabus et al. 1999, Hara et al. 2003, van Beilen et al. 2004, Throne-Holst et al. 2006, Feng et al. 2007, Throne-Holst et al. 2007, Wentzel et al. 2007, Rojo 2009, Teramoto et al. 2009, Wasmund et al. 2009, Tapilatu et al. 2010, Alonso-Gutierrez et al. 2011, Teramoto et al. 2011).  Among these, many were characterized from high-latitude marine environments.  Specifically, the Pseudomonas strains isolated by Whyte et al. (1997) from Arctic soils may be transported to the marine environment via runoff.  Alcanivorax species are known to be widespread in marine environments exposed to oil (Hara et al. 2003, van Beilen et al. 2004) and, if not prevalent in the Arctic environment, might be expected to be present because of currents.  Thus, bacteria capable of alkane degradation are expected to be present in the Arctic oceans.

The ability to degrade aromatic hydrocarbons and, in particular, polynuclear aromatic hydrocarbons typically is considered to be less widespread than the ability to degrade alkanes.  For example, some organisms have diverse pathways that confer the ability to degrade polynuclear aromatic hydrocarbons, e.g., Mycobacterium vanbaalenii (Kweon et al. 2011) and various Pseudomonas spp. (Whyte et al.1997). The distribution of these genes among bacteria in Arctic marine environments remains unknown.       

In general, biodegradation of oil compounds is expected to follow the order n-alkanes > branched alkanes > low molecular weight aromatics > cyclic alkanes (Perry 1984). In cold seawater the same order is expected, although degradation will be highly influenced by the physico-chemical characteristics of the oil. The low temperature affects both dissolution from the non-aqueous (crude oil) to the aqueous phase (Schluep et al. 2001), and evaporation of volatile compounds, as described above.  

At temperatures above the freezing point of seawater (approximately -1.8°C) biodegradation of crude oil hydrocarbons is well documented. This is exemplified in Figure 5-2, showing the mineralization curves of ¹⁴C-labelled naphthalene, phenanthrene and hexadecane in seawater at 0°C when the compounds were spiked into crude paraffinic oil. Degradation of the n-alkane (hexadecane) was faster than for the aromatic compounds, and a smaller aromatic (naphthalene; 2-ring) degraded faster than larger aromatics (phenanthrene; 3-ring). This pattern followed the generally accepted order of crude oil compound biodegradation described above.

Figure 5-2. Mineralization in seawater at 0°C of 14C-labelled hydrocarbons spiked into crude oil (from Brakstad and Bonaunet 2006). No mineralization was measured in sterile controls (not shown).

One of the first attempts to study oil biodegradation in Arctic seawater at low temperatures (2-11°C) showed that shifts in microbial populations towards more oil-degrading bacteria, that abiotic oil losses were lower than expected, and that various classes of hydrocarbons (saturates, mono-, di- and polyaromatics) were subject to biodegradation (Horowitz and Atlas 1977). Several studies have compared oil biodegradation in seawater or with bacterial cultures at different temperatures, and results from some of these including temperatures relevant for the Arctic are summarized in Table 5-1.

In summary, these and most other relevant studies (e.g. MacNaughton et al. 2003) show slower biodegradation by lowering of the temperature, but the results also show that biodegradation at low seawater temperature is considerable. In a recent study with low concentrations (2.5 mg/L) of Alaska North Slope oil with Atlantic seawater, 80% was biodegraded (saturates, 2- to 4-ring aromatics) after 60 days at 8°C (Prince et al. 2012). While laboratory studies indicate that biodegradation in Arctic seawater may be slower than in temperate seawater, these results have not been confirmed by field studies. Seasonal biodegradation data and comparison of oil biodegradation from different geographic areas with the same oils and analytical procedures may be necessary to test these assumptions. Oil characteristics should also be addressed in more detail, for instance, by comparison of dispersed oil biodegradation of different oil types and weathering degrees at several seawater temperatures.  The physical properties of oil may decrease bioavailability of oil (e.g. larger droplets at lower temperatures would increase the surface area-to-volume ratio).

Table 5-1. Summary of selected biodegradation studies performed at different seawater temperatures 

^A) k1 is first-order rate coefficient; t_½ is half-life (0.69/k1) ^B) Reduction determined by comparison to sterile controls

Several biodegradation studies of oil in Arctic sediments have been conducted, most of these to investigate the potential for bioremediation of stranded oil in the Arctic (see later chapter). Studies on oil pollution of Arctic and Antarctic beaches has demonstrated the presence of indigenous hydrocarbon-degrading bacteria in these pristine environments (e.g. Grossman et al. 2000; Delille and Delille 2000; Powell et al. 2005). Oil removal from beach sediments may be attributed to several processes, including physical removal, photooxidation and biodegradation. For instance, significant depletion of total hydrocarbon concentrations and mineralization of radiolabelled hexadecane have been measured in Canadian Arctic soils at 4°C (Greer 2008). Anaerobic biodegradation has also been measured in the Arctic. Low-temperature degradation of PAH-compounds was reported from Arctic soils under anoxic and nitrate-reducing conditions at 7°C (Eriksson et al. 2003). In the Arctic winter the upper parts of marine sediments become frozen.  Whether biodegradation stops or continues at very slow rates under these conditions is not known, although microbial activity at subzero temperatures has been demonstrated (Doyle et al. 2012).  Several studies with oil-contaminated freeze-thaw cycled soil or permafrost have shown that microbial respiration takes place even at subzero temperatures and hydrocarbon degradation was observed (Rike et al. 2003; Børresen et al. 2007; Chang et al. 2011).  These studies therefore demonstrate that the lower limit for biodegradation can be below the freezing point.

If oil spills reach the marginal ice zone, the ice may become oil-infested. Once trapped within the ice, ocean currents can transport the oil over large distances.  A secondary discharge situation occurs during the spring melt season and, if the ice has been transported from the original spill site, this can result in contamination of new locations. In the spring and summer seasons, chemical alteration of the crude oil through photooxidation may also become an important process (Refer to Section 3).  Although the immediate impact of oil spills in ice has been studied (e.g. Fingas and Hollebone 2003) and is fairly well understood, little is known about the long-term fate and effects of such pollutants on ecosystems in polar environments.  To date, few studies have attempted to determine the transport and fate of individual water-soluble oil components in sea ice.  However, data from some recent studies have shown that the more water-soluble compounds (mainly naphthalenes, phenanthrenes and dibenzothiophenes) migrate through the brine channels in the ice (Figure 5-3).  As a result, such compounds come into contact with sea ice microbes in the brine and the underlying water (Faksness and Brandvik 2008a; Faksness and Brandvik 2008b).

In line with the results from studies with Arctic soils one should expect that biodegradation may also take place in marine ice at subzero temperatures. As described earlier, microbial metabolism and motility have been measured in the brine channels of marine ice (Breezee et al. 2004; June et al. 2002; Junge et al. 2003; Junge et al. 2004; Junge et al. 2006; Mykytzuk et al. 2013).  However, biodegradation of oil in marine ice has not yet been fully investigated. In a winter field study (February to June) performed on Svalbard with crude oil frozen into fjord ice, a slow reduction in the ratio between naphthalene and phenanthrene was measured in the parts of the ice with downward migration of soluble compounds, while no significant change in n-C₁₇/Pristane was measured, as shown in Figure 5-3 (Brakstad et al. 2008). However, the bulk oil stimulated bacterial biomass, including a few bacterial genera expected to be oil-degraders (Brakstad et al. 2008). The results from another field study performed at Svalbard showed that no significant degradation of oil hydrocarbons occurred in the ice at subzero temperatures, but at 0°C melt pool oil samples fertilized with inorganic nutrients showed a significant change in bacterial diversity (Gerdes and Dieckmann 2006). Marine ice represents an extreme environment for life.  The combination of low temperature and high salt content in the brine channels require that microbes be both halo- and psychro-tolerant.  Extremely halophilic or halotolerant microbes able to degrade oil have been reported (e.g. Diaz et al. 2002, Al-Mailem et al. 2010), but not so far in cold environments.  However, as described above, it has been demonstrated that oil pollution in marine ice may stimulate the growth of a few specific bacteria (Brakstad et al. 2008), but the ability to degrade oil compounds needs to be clarified. In addition, most oils will also be solidified under these conditions, but the migrating water-soluble compounds may be relevant target compounds for oil-degrading bacteria in this environment. If this is true, bacteria able to degrade small aromatics may be more relevant than alkane-degrading bacteria. 

Figure 5-3. Oil migration and degradation in marine ice. [The middle core shows the migration of different oil components through an ice core with the chromatograms of the components in the left panel and the relative ratios of specific components in the right panel (from Brakstad et al. 2008)].

Several oil spill models have been developed during the last decades.  Most of these are physical models which can be separated into oil weathering models, trajectory models (predicts the route of an oil spill), or stochastic models (describing an impact area of an oil spill).  Examples of well-known models are the Oil Spill Contingency and Response (OSCAR) and the OILMAP models (www.sintef.no/Materialer-og-kjemi/Marin-miljoteknologi/Miljomodellering/Modellverktoy/OSCAR-Oil-Spill-Contingency-And-Response/; http://www.asascience.com/software/oilmap/index.shtml). Models have also been presented for predictions of oil behavior in ice-infested water (Drozdowski et al. 2011). Most of these models are physical models, but the OSCAR model also incorporates biodegradation of 25 pseudo oil compound groups, in addition to descriptions of the physical environment, physical-chemical fate processes and ecotoxicity (Aamo et al. 1997; Reed et al. 2000). In the OSCAR model, which is an industry standard in Norway, biodegradation is one of the fate processes together with physico-chemical processes like advection, spreading, evaporation, dispersion, dissolution, particle adsorption/dissolution, volatilization from water column, and seabed contamination. An example of vertical oil concentrations and mass balance after a simulated 60-day blowout is shown in Figure 5-4. However, biodegradation as part of the mass balance may be overestimated in the model, since degradation is determined on the bases of biotransformation, not complete biodegradation, and only compounds determined by gas chromatography-mass spectrometry (GC-MS) analyses are included.  

Figure 5-4. Simulation of a deep water blowout by the OSCAR model. [The left figure shows the oil concentration vertically in the water masses during a simulated 60-d blowout from 1600 m depth with a light paraffinic oil. The right figure shows the mass balance between different fate processes during the blowout period (from Brakstad et al. 2011).]

In the oil spill models biodegradation must be predicted at different temperatures. Oil biodegradation data in Arctic environments with cold seawater are limited, since most published studies have been performed at higher temperatures than relevant for these environments. In order to transform degradation data between different temperatures, plots have been used to transpose results of bacterial metabolism and growth at different temperatures. Temperature-related bacterial growth rates may be estimated by using modified Arrhenius plots (Arrhenius, 1889).  Ideally, Arrhenius plots should show temperature-related linearity.  However, in a study with psychrotrophic toluene-degrading strains of Pseudomonas putida grown on toluene or benzoate, growth rates had to be fitted using two linear segments at a temperature range of 4-30°C: one segment above and one below 17-20°C (Chablain et al. 1997).  When using Arrhenius plots, the temperature range should therefore not be too broad.  For water-soluble compounds, temperature-dependent biodegradation has been suggested to follow a Q₁₀-value, which is a relationship describing the degradation rate increases when temperatures are raised by 10°C increments.

Equation 1

Where R is the general gas constant (8,314·10^-3 kJ/mol·K), E_a is the activation energy (kJ/mol), T₁ is the reference temperature in Kelvin and T₂ is the actual temperature in Kelvin.  According to this approach, the degradation rates should double for every 10°C increase, resulting in an ideal Q₁₀ of 2.0.  The Q₁₀–values for the biodegradation of oil hydrocarbons in seawater were determined with a heavy fuel oil (Bunker C), and with winter or summer water samples from the North Sea.  When incubation temperatures of 4-18°C were used, Q₁₀–values of 2.4 and 2.1 were determined for waters in winter and summer, respectively, where biodegradation was measured as biological oxygen demand (Minas and Gunkel 1995).  Calculations of Q₁₀-values from a variety of studies have shown that the rule-of-thumb value (Q₁₀ = 2.0) is a fairly good approximation in a temperature range of 5–27°C (Andrea Bagi, personal communication).  However, for the narrow range and freezing temperatures of the Arctic the expectation that calculated Q₁₀ would remain close to 2 may not be valid.  For instance, a calculation of Q₁₀ in immobilized oil films (Statfjord B oil) based on data for 5 and 0°C (Brakstad and Bonaunet 2006) showed a value of 16.2 (Andrea Bagi, personal communication).  This may be caused by changes in the physico-chemical characteristics of this oil at these temperatures.  Thus, changes in oil characteristics at low seawater temperatures may affect the biodegradation models, and therefore predictions of oil degradation rates in Arctic seawater will require closer examination. 

Since oil consists of thousands of different compounds (Marshall and Rogers 2003) measurements of individual compounds is a challenge.  Bulk oil biodegradation may be determined by traditional gravimetric analyses (e.g. Horowitz and Atlas 1977), while broader groups of oil components (saturates, aromatics, resins and asphaltenes = SARA) may be determined by Iatroscan thin-layer chromatography with flame ionization detection (TLC-FID; Stevens 2004).  Using this method, crude oil components are determined according to their polarity.  The saturate fraction consists of nonpolar material including linear, branched, and cyclic saturated hydrocarbons (paraffins). Aromatics, which contain one or more aromatic rings, are slightly more polarizable.  The remaining two fractions, resins and asphaltenes, have polar substituents.  Additional bulk oil analytical methods include Fourier Transform Infrared (FTIR) spectroscopy and Nuclear Magnetic Resonance (NMR) spectroscopy.  FTIR is an absorption technique that uses infrared (IR) electromagnetic radiation to examine the identity of chemical bonds within the substance of interest.  As microbial degradation of the oil is expected to result in the addition of oxygen atoms into the structure of oil compounds this method may be a method for measuring bulk changes in composition, although the resolution and sensitivity is poor compared to other methods.  NMR is a nondestructive technique that is well-suited for identifying and quantifying different hydrocarbon classes and can provide information on the relative content of aliphatic, olefinic, and aromatic components.  Studies have shown that NMR spectra in conjunction with multivariate statistical analysis can be correlated to a number of physicochemical properties and standard distillation cut yields (Molina et al. 2007).  Mass spectrometry (MS) has become one of the most important detection principles in modern analytical chemistry.  The principle behind MS is that molecules can be identified through their molecular weight and fragmentation patterns.  MS is very often connected to a separation step, usually gas (GC) or liquid (LC) chromatography.  These methods may be used to identify and quantify targeted oil compounds or for fingerprinting of complex chemical mixtures.  To separate between different oil compound groups gas chromatographic analyses (GC-FID and GC-MS) are the standards today, but these methods favor detection of nonpolar compounds.  The common use of these methods therefore limits our knowledge of oil biodegradation, mainly to some compound groups, such as the C10-C40 saturates, cyclic saturates (decalines), BTEX, phenols, 2-6 ring PAHs, and a variety of biomarkers.  LC-MS analyses may therefore be an important supplement to the gas chromatographic analyses for more polar compound groups.  In addition, biodegradation studies of compounds like naphthenic acids have been of interest in specific areas like Canada.  Several high-resolution instruments, like time-of-flight mass spectrometers (ToF-MS) coupled to GCxGC systems (GCxGC-ToF-MS)(e.g. Tran et al. 2010) and Fourier transform ion cyclotron resonance (FT-ICR) mass spectrometer (e.g. Hughey et al. 2008) provide powerful techniques for the analytical separation of complex mixtures combined with methods for characterizing the resolved compounds. Minor components hidden in the large background can be detected by these instruments, and both resolution and sensitivity allow for searching of spectra from very narrow peaks.  For instance FT-ICR MS can separate masses of <0.002 Dalton of compounds that contain heteroatoms such as N, O, S and other elements, identifying oil compounds by mass and molecular formula at high resolution.

Advances in microbial sampling capabilities, in particular sampling of the ocean in drilling areas, came with advances in drilling technology.  The Ocean Drilling Project (ODP) and subsequent Integrated Deep Ocean Drilling Program (IODP 2003-2013) and planned International Ocean Discovery Program (IODP 2013-2023) provide a framework for these activities (Edwards et al. 2012).  Each of the named programs includes or will include a sampling component for microbial ecology research.  In particular, the 2003 IODP included extensive evaluations of seafloor and sub-seafloor microbial communities (Cyranoski 2003).  In conjunction with new microbiological techniques, these samples provided new perspectives on deep ocean microbial community composition and function (D'Hondt et al. 2004; Schippers et al. 2005; Inagaki et al. 2006; Biddle et al. 2008; Kobayashi et al. 2008; Forschner et al. 2009; Lomstein et al. 2012).  Understanding native populations in Arctic drilling fields requires sampling such as has been carried out in these programs.

Much of the sampling that is associated with drilling activities focuses on the deep subsea floor while water column and sediment samples can be collected with remote samplers or can also  be collected by autonomous underwater vehicles (AUVs).  The Chemosynthetic Ecosystem Science (ChEss) project of the Census of Marine Life (2002-2010) was one such project that generated a substantial amount of new information about marine microbial communities.  Much of the success of the ChEss project was attributed to the development of improved deep-ocean AUVs (German et al. 2011) that allowed systematic exploration of previously understudied areas, including cold seeps.  Modern AUVs are capable of rapid deployment and operation at a range of depths.  They have been effectively deployed to sample in response to events such as the Deepwater Horizon spill of 2010 (Camilli et al. 2010).  These vehicles contribute to the ability to observe natural processes and conduct in situ experiments, particularly at depth in harsh marine environments.

Another strategy is to employ microbial observatories in marine environments that incorporate real-time sensors, time-lapse cameras, and other experimental devices.  These observatories, along with autonomous and cabled sensors, allow direct measurement of microbial processes in the deep ocean.  In particular, beginning some 20 years ago, circulation obviation retrofit kits (CORKs) came into use to study connectivity of hydraulics and biogeochemistry at the interface of the ocean bottom and open water (www.corkobservatories.org; Cowen et al. 2003) rather than relying on extrapolation from controlled laboratory experiments (e.g. Bartlett 2002; Tapilatu et al. 2010) or inference from population composition (Simonato et al. 2006) as is more commonly done.  Another type of observatory, the Microbial Methane Observatory for Seafloor Analysis (MIMOSA), is an autosampler that collects and archives microbial material for later recovery and analysis.  Two of these devices recently were deployed in the Gulf of Mexico to evaluate petroleum seeps and spills as they affect microbial population structure (Balinski 2012).  This type of observatory may be useful to implement in situ experiments to monitor biodegradation rates and processes and further advance knowledge of petroleum hydrocarbon degradation processes in the deep ocean environment.

Finally, another tool that is directly relevant to petroleum biodegradation and carbon utilization is the so-called “bug trap,” in which hydrophobic beads or woven matrix is dosed with petroleum hydrocarbons to evaluate in situ degradation potential and analyzed to characterize degrading community composition.  Because petroleum-degrading microorganisms can be chemotaxic to suitable substrates, these experimental devices can be used to attract and study degraders in the laboratory (Brakstad and Bonaunet 2006) and in situ experiments (Raloff 2010; DeAngelis et al. 2011).

In standard laboratory studies, oil degradation is usually determined by comparison of depletion in normal seawater or cultures to depletion in sterile (killed) controls. In this way processes like evaporation, wall effects, dissolution of compounds from the oil phase etc. may be accounted for and separated from the biodegradation process. However, in field and meso-/large-scale studies biodegradation is determined by normalization of degradable compounds to less degradable (recalcitrant) compounds. Common compounds for this internal normalization are pentacyclic triterpane biomarkers (e.g.C₃₀17α(H),21β(H)-hopane) and the isoprenoids pristane and phytane (Prince et al. 1994; Douglas et al. 1996; Page et al. 1996). The isoprenoids have proven to be biodegradable themselves, although at slower rates than their corresponding n-alkanes (e.g. Douglas et al. 1996). Hopanes also have limitations if used to determine biodegradation of compounds with low boiling points, since it may be difficult to separate biodegradation from evaporation. In addition, determination of biodegradation as ratios between biodegradable and more persistent compounds has also been suggested using other compounds, like 2-methylphenanthrene/1-methylphenanthrene, and C3-phenanthrene/C3-dibenzothiophene (Fedorak and Westlake 1981; Christensen and Larsen 1993; Wang et al. 1998; Lamberts et al. 2008).

Most environmental studies of petroleum-derived chemicals have focused on effects related to specific and easily identified hydrocarbons such as n-alkanes, BTEX and PAHs. However, in the case of environmentally weathered samples, most oil compounds appear as an unresolved complex mixture (UCM in gas chromatograms, and are often referred to as the “hump” (see Figure 5-5). Having undergone a variety of weathering processes (e.g. evaporation, biodegradation and photooxidation), this residual UCM is comprised of thousands of environmentally persistent compounds (Gough and Rowland 1990, Killops and Al-Jaboori 1990). In fact, it has been established that natural biodegradation of spilled crude oil leads to a significant increase in the UCM  concentration relative to other crude oil components (e.g. Meredith et al. 2000), highlighting the persistence of these compounds. Fractionation and subsequent characterization studies have shown that both non-polar (e.g. aliphatic and aromatic) and polar (resin and asphaltene) compounds contribute to crude oil UCMs.  Both the aromatic hydrocarbon and polar UCMs have been shown to bioaccumulate in marine organisms and elicit ecotoxicological responses and impaired health (e.g. Farrington et al. 1982; Widdows et al. 1995; Barron et al. 1999; Smith et al. 2001; Rowland et al. 2001; Donkin et al. 2003). The polar UCM fractions comprise compounds containing highly polar N, S, and O atoms in their structures (so-called NSO compounds).  Many of these compounds (e.g. phenols and naphthenic acids) are thought to be homologous in structure to compounds present in the non-polar fraction of the UCM, hence their resistance to biodegradation.  Due to their persistence these compounds may reach other locations. 

Figure 5-5. Gas chromatogram of the aromatic hydrocarbon UCM fraction extracted from mussels (Mytilus edulis) from a petroleum impacted site (Southend) on the UK coast (From Booth et al., 2007a).

The biodegradation potential of these seemingly persistent UCM compounds is further complicated by the environmental conditions prevalent in Arctic regions.  Lower ambient temperatures will result in reduced biodegradation rates, and currently nothing is known about the abilities of psychrophilic or psychrotrophic bacteria to degrade these compounds. In temperate regions, microbial communities from previously oil-impacted sites have been shown to partially degrade model UCM compounds, such as alkylcyclohexyltetralins, alkylcyclohexylnaphthalene and naphthenic acids (Scott et al. 2005; Booth et al. 2007b; Frenzel 2008).  Recent studies using naphthenic acids showed that it was the molecular structure rather than the number of carbon atoms that was important for determining biodegradation.  Specifically, the most recalcitrant compounds included those with relatively high degrees of alkyl branching (Han et al. 2008).  During future degradation studies of oil compounds in cold environments it is therefore of major importance to consider these environmentally persistent and toxic UCM-related compounds.