Coastal marshes were heavily impacted by the Deepwater Horizon (DWH) oil spill in 2010, with approximately 90% of shoreline impacts occurring in Louisiana's coastal wetlands. Spilled crude oils impact an environment through four major mechanisms: ecosystem exposure to reactive and toxic aromatic compounds; covering and smothering that hinders normal plant and animal physiology; depletion of dissolved oxygen; and disruption of the aquatic food web. Crude oil's ability to cause environmental harm depends upon its composition, which is a very complex mixture of many thousands of reduced carbon compounds made from the degradation of plant material deposited deep underground. This study reviews the results from the chemical characterization of petroleum hydrocarbons, at various weathering stages, in >2000 marsh surface sediments and select sediment cores samples collected from various sampling locations in Terrebonne Bay, Grand isle, and northern Barataria Bay from 2010 to 2018. The sediment samples were analyzed for target saturated alkanes, polycyclic aromatic compounds, and the forensic biomarker (hopane and sterane) compounds. The chemical characterization of the compositional changes of target compounds in DWH oil, from its pre-stranding stage just offshore in the Louisiana Bight, through stranding on marshy shorelines and through its degradation and weathering over eight years has given insights into the complexity of oil residues and potential for impacts in these varying environmental conditions. Stranded oil initially had two prominent fates: settling on surface sediment/soils of the marshes, and subsurface deposition primarily by means of settling into fiddler crab burrows. Both initial fates affected shorelines and 10–20 meters inward. Over time, surface oil residues were spread beyond initially impacted areas by Tropical Storm Isaac in 2012 and other weather events, and oil residues were quickly degraded. Subsurface stranded oil was degraded much more slowly under anaerobic conditions and some was re-released as fairly fresh oil during the coastal erosions caused by DWH surface oiling damage to the marsh plants. However, these re-releases were relatively slow and were quickly aerobically degraded once the stranded oil reached marsh surfaces. There was also evidence of anaerobic degradation of heavily weathered surface oil residues during the 2015 to 2018 timeframe. This eight-year study establishes a very complex narrative between the physical and chemical properties of stranded oil and its interactions with coastal marsh environments.

Crude oil is a very complex mixture of many thousands of reduced carbon compounds (i.e., hydrocarbons) made from the degradation of plant material deposited deep underground (Overton et al., 2020). Oil weathering involves a continuous series of compositional changes of these hydrocarbons when oil enters marine and coastal environments. These compositional changes are caused by physical processes such as evaporation, dissolution, emulsification, sedimentation, and by chemical oxidations due to sunlight and biotic interactions (Aeppli et al., 2014; Atlas et al., 2015; Tarr et al., 2016). Weathering, by changing the composition of the original spilled oil, changes the oil's physical, chemical and toxic properties (Overton et al., 2020; Tarr et al., 2016). Fresh oil is more volatile, contains more water-soluble components, floats, is not very viscous, and can easily spread out from the release point within coastal waters. All of these characteristics contribute to the impacts of fresh oil in the environment. As oil weathers, it initially loses volatile components, which are also the more water-soluble components, and the oil residue becomes more viscous and more likely to glob or aggregate together as opposed to spreading out in a thin film (Aeppli et al., 2014; Albers, 2003). During its transit from the source into the coastal environments, oil is generally mixed with water by wind and wave energy to form a more viscous water-in-oil emulsion that is fairly resistant to rapid weathering. Consequently, emulsification greatly slows down the weathering processes, and so the weathered oil will stay in the environment longer than non-emulsified liquid oil (Tarr et al., 2016). Further, emulsified oil is also somewhat more difficult to remediate by skimming, dispersing, or burning. Fortunately, emulsified oil is generally less harmful to the environment in terms of chemical toxicity, however it becomes more sticky to cause damage by covering or smothering as opposed to toxic interactions. Moreover, if emulsified oil is ingested through, for example, preening of feathers, it can have significant toxic effects on internal organs. Even though emulsified oil weathers slowly, various weathering processes continue to change the composition of the oil until it has been degraded, leaving behind only small quantities of residue known as tarballs or surface residue balls. These oil residues are composed of mostly the resins and asphaltenes from the oil and some oxidation products formed during the weathering process. Since these residues are very insoluble, they are persistent and their environmental impacts are not thought to be significant except for the physical disruption of marsh surfaces from these tar like asphaltenic residues and mats.

When crude oil is released into the marine environment in quantities exceeding the local environment's ability to assimilate and degrade the oil, it can cause harmful impacts via several mechanisms. First, some of the hydrocarbons in crude oil, typically the aromatic compounds in the molecular weight range from 78 to about 300 g/mole, which constitutes ~4% of the weight of the initial surface oil (Reddy et al., 2012), have the potential to be toxic to a wide variety of organisms. In addition to photochemical and microbial oxidations, these molecules are oxidized by organisms that contain Mixed Function Oxidase (MFO) enzyme systems, thus causing potential toxic consequences. Second, crude oil is a hydrophobic material that will coat or stick to plant materials and animal body parts (e.g., feathers). This coating/smothering can disrupt the normal functions of the coated/smothered material and frequently cause organism death. Third, virtually all compounds in crude oils can be degraded by water born microorganisms that catalyze the oxidation and mineralization of petroleum compounds. This oxidation creates a biological oxygen demand (BOD) that can deplete the natural levels of dissolved oxygen and may cause normally aerobic aquatic environments to become devoid of oxygen (anoxic), thus causing death to local organisms. Fourth, the quality and quantity of microbes change as they degrade oil, which may disrupt the non-microbial community composition, including plankton blooms that have cascading impacts. Therefore, environmental impacts from oil spills are influenced not only by the chemical compositions of toxic compounds in oils, but also by the physical properties of the spilled material, and the distribution of weathered oil residues in various chemical and biological components of coastal marshes. Further, the weathering process changes all of the fundamental implications associated with these mechanisms for causing environmental damage (toxicity, routes of exposure, bioavailability, carbon enrichment) and, in general, are thought to reduce opportunities to cause environmental damage. The impacts of the eventual concentration of highly refractory, insoluble oxygenated crude oil residues in coastal marshes remains debated, as does the residue of biomass formed from oil degradation (Overton et al., 2020; Tarr et al., 2016).

Coastal marshes were heavily impacted by the Deepwater Horizon (DWH) oil spill in 2010, with approximately 90% of shoreline impacts occurring in Louisiana's coastal wetlands. Most coastal marsh environments are aerobic near the marsh surface but quickly become anaerobic within a centimeter or two of the surface. Thus, once oil is stranded on coastal marsh surfaces, it can exist in both aerobic and anaerobic environments, and because of shading from coastal grasses, is generally not highly exposed to sunlight, especially when compared to sunlight exposures in floating oil slicks. This means that the primary degradation process of oil in coastal marshes is thought to be by microbial, not photo-oxidative pathways. That is, the oxidation of petroleum residues in marsh environments is primarily from enzymatic oxidations as opposed to photo induced hydroxyl radical oxidations. This means that the oxidation follows a pattern known to be associated with microbial degradation as opposed to photodegradation (Atlas et al., 2015).

Under ideal aerobic conditions, most petroleum hydrocarbons are rapidly biodegradable (Prince and Walters, 2007) and generally follow a clear degradation pattern: n-alkanes>branched alkanes>low-molecular weight aromatics>high-molecular-weight aromatics and cyclic alkanes (Wang and Fingas, 2003, Gros et al., 2014, Olson et al., 2016). The most readily biodegradable compounds are the normal alkane hydrocarbons and the one- to three-ringed polycyclic aromatic hydrocarbons (PAHs). For larger PAHs (> 4 aromatic rings) and their associated alkyl homologs, the parent compound is typically the first to enzymatically degrade, while the alkyl homologs are slower to degrade (i.e., increasing alkylation slows biodegradation) (Wang and Fingas, 2003). Photo-oxidative processes tend to degrade the higher carbon number alkyl homologs in preference to the parent and C1 alkyl homolog PAHs (Radovic et al., 2014).

There are hundreds of species of bacteria, archaea, and fungi that are widely distributed throughout the environment and that are fully capable of using high-energy reduced petroleum hydrocarbons as a carbon or nutrient source (Atlas et al., 2015; Prince and Walters, 2007). However, the mechanism for how biodegradation affects the composition of oil in the environment is highly dependent on the physical and weathered properties and amount of oil spilled, as well as other environmental factors like redox conditions, nutrient availability, temperature, salinity, and wave/mixing energy. These factors greatly influence the microbial ecology and petroleum hydrocarbon degradation dynamics; therefore, the biodegradation rates of oil residues are site-specific (Atlas et al., 2015).

Anaerobic or anoxic biodegradation of petroleum hydrocarbons is accomplished by gram-negative proteobacteria, mycobacteria, and some methanogenic archea (Atlas et al., 2015). In comparison to aerobic biodegradation, anaerobic biodegradation is a much slower process due to the fact that the rates of hydrocarbon degradation decrease with decreasing oxygen reduction potential (Atlas, 1981). As a result, petroleum hydrocarbons can remain relatively unaltered in reduced environments, such as sediments, for long periods of time so that the buried oil may even appear as “fresh” oil compared to the same oil exposed at the surface conditions (Atlas et al., 2016).

Select hydrocarbon compounds referred to as ‘petroleum biomarker compounds' are considered to be recalcitrant in most conditions due to their large molecular weights, their molecular structures, and their very low water solubility (Peters et al., 2005). The compound C30-hopane is even used as a conserved internal marker by which all other compounds in oil can be normalized to in an effort to quantify degradation (Prince et al., 1994). However, Wang et al., 2001, Aeppli et al., 2014) and Atlas et al., 2015 demonstrated that even hopane is susceptible to biodegradation, although these buried petroleum compounds are the last in the succession of petrogenic compound to be fully biodegraded.

There are several indicators used to quantify oil biodegradation in marsh sediment samples (Tarr et al., 2016; Meyer et al., 2018; Overton et al., 2020). One measures the decrease in the ratio of heptadecane (n-C17) to pristane (a branched unsaturated alkane, or isoprenoid) or octadecane (n-C18) to phytane (also an isoprenoid). A n-C17/pristane and C18/phytane ratio less than in the initial spilled oil represents a biodegraded oil residue (Meyer et al., 2018; Prince and Walters, 2007). The other biodegradation clue is the formation of a “hump”, an unresolved complex mixture (UCM), in the chromatographic data. As the normal alkanes are rapidly weathered by microbial attack, they leave behind the less degradable branched and cyclic alkanes, aromatics, and resins, which leave behind a disproportionately higher amount of these chromatographically un-resolved compounds (Gros et al., 2014). This hump in the chromatographic baseline is called a “complex unresolved mixture”. Some of the compounds in the hump may also be oxidized hydrocarbons from the weathering processes (Mohler et al., 2020). Another indication of microbial weathering is the loss of the parent and C1 alkyl homologs of petrogenic PAH compounds relative to the C2 to C4 alkyl homolog isomers. Further, it also appears that specific isomers of the various alkyl homologs will be degraded in preference to other isomers of the same molecular weights.

In order to study oil degradation and weathering in coastal marshes, we revisited previously reported (Adhikari et al., 2016; Meyer et al., 2018; Tarr et al., 2016; Turner et al., 2014) as well as data from over 2000 marsh surface sediment samples and select sediment cores collected from various Louisiana coastal locations from 2010 to 2018. The samples were collected from various locations along the coastal marsh shorelines in northern Barataria Bay Louisiana (see map in Figure 5) that were impacted by oiling during the summer of 2010. Control samples were also collected from areas well beyond the impacted sites. The top 5 cm were samples and stored immediately on ice for transport to the lab, where they were frozen and stored until analyses. Several deeper 15–20 cm cores were collected that contained oil residues in fiddler crab borrows and these were stored by freezing. All samples were analyzed by using standard GC/MS analytical methods, including extraction with methylene chloride and selected ion monitoring GC/MS analysis of the extracts using standard QA/QC procedures (Adhikari et al., 2016; Olson et al., 2017). All samples were analyzed for target alkanes from C10 to C35 plus pristane and phytane, target 2 to 6 ringed PAH aromatic compounds and their respective alkyl homologs, dibenzothiophene and naphtylbenzothiophene and their alkyl homologs, and the hopane (ion 191), sterane (ions 217and 218) and triaromatic steroid (ion 231) petroleum biomarkers. The chemical characterization of the compositional changes of target compounds in DWH oil, from its pre-stranding stage just offshore in the Louisiana Bight, through stranding on marshy shorelines and through its degradation and weathering over eight years has given insight into the complexity of oil residues and impacts in varying environmental conditions.

The DWH Oil was transported into marsh environments as slicks of emulsified and evaporatively weathered and photo-oxidized oil residues (Ward et al., 2018). It was initially distributed into and onto marshes mostly along the marsh-open water interface and as far inland as wave action normally can take the floating slick during high tide, typically 10 to 30 meters inland. Unusual weather events, such as tropical storms, hurricanes and cold fronts can produce unusually high tides and waves which can spread the stranded oil further into the marsh. However, these processes also dilute the oily residues over a much larger area and can generally speed up weathering by breaking solid oil residues into much smaller particles. This exposes the dispersed particles of oil residue to aerobic environments where microbial degradation can proceed at a faster rate than when oil residues are lumped together or buried.

Figure 1 outlines the composition of the MC252 source oil released during the DWH oil spill. These data show the target alkane and target petrogenic aromatic compound distribution in fresh MC252 source oil, as well as the distribution of hydrocarbons in three distinct ranges of molecular size ( very volatile compounds from n-C1 to n-C10, readily weathered compounds from n-C10 to n-C24, and compounds heavier that than n-C24) . It is important to note at the simulated distillation of the source oil showed that approximately twenty percent of the liquid oil was made up of volatile organic compounds (VOC) with a carbon content ranging from C5 to C9. Sixty-eight percent of the hydrocarbons had a molecular carbon structure containing C10 to C44 molecules, and fourteen percent was comprised of non-gas chromatograph amenable high MW residues including asphaltenes and resins. Interestingly, 70% of the liquid oil mass was comprised of compounds that eluted before n-C25 molecules, including all of the benzenes and 2 to 3 ringed PAHs and their respective alkyl homologs. This represents the chemical composition that is most readily lost during weathering (Tarr et al, 2016).

Figure 1:

Figure showing the composition, from simulated distillation data, of the riser fluid and fresh surfacing liquid oil, estimating that portion of the oil that was composed of natural gas and VOCs, GC amenable hydrocarbons, and high molecular weight non-GC amenable hydrocarbons. Target compounds that are typically measured in oil samples comprised approximately 13% of the liquid oil by mass. Estimates are that the composition of two and three ringed aromatics and alkanes below n-C24 comprise approximately 70% of the mass of liquid floating oil.

Figure 1:

Figure showing the composition, from simulated distillation data, of the riser fluid and fresh surfacing liquid oil, estimating that portion of the oil that was composed of natural gas and VOCs, GC amenable hydrocarbons, and high molecular weight non-GC amenable hydrocarbons. Target compounds that are typically measured in oil samples comprised approximately 13% of the liquid oil by mass. Estimates are that the composition of two and three ringed aromatics and alkanes below n-C24 comprise approximately 70% of the mass of liquid floating oil.

Close modal

Figure 2 shows the gas chromatographic and PAH compositional data for 4 floating oil samples collected just off the Louisiana coast near the Mississippi River Birds Foot delta, and 4 samples of recently stranded oil residues collected from Bay Jimmy in the summer of 2010. By examining these data, we can conclude that the oil that impacted most marsh coastlines had been weathered by evaporation, some dissolution during its assent from the wellhead, and by photooxidation on the sea surface, losing the alkanes below ~n-C13, as well as most of the lower molecular weight polycyclic aromatics such as the naphthalenes and fluorenes with the parent and C1 homologs being lost in preference to the C2 to C4 homologs. Isoprenoid to normal alkane ratios of floating oil residues were unchanged when compared to the source oil, indication little or no biodegradation. This floating oil material also contained higher levels of insoluble oil residue material (non-gas chromatographically detectable material) produced from photooxidation that was in addition to the resin and asphaltene components (Ward et al., 2018).

Figure 2:

Outline of the chemical composition of typical oil residues collected from the just offshore of the MR Birds Foot delta and from the shoreline of Bay Jimmy Louisiana during the 2010 time frame, highlighting hopane calculated depletion of the hydrocarbon residues compared to liquid oil samples collected from the riser pipe (red) in stranded samples and ratio of the sum of the aromatic to total targeted aromatic and alkane hydrocarbons (blue). This ratio is around 14.6 in the riser oil sample. The C17/pristane and C18/phytane ratios showing little biodegradation are bordered in black.

Figure 2:

Outline of the chemical composition of typical oil residues collected from the just offshore of the MR Birds Foot delta and from the shoreline of Bay Jimmy Louisiana during the 2010 time frame, highlighting hopane calculated depletion of the hydrocarbon residues compared to liquid oil samples collected from the riser pipe (red) in stranded samples and ratio of the sum of the aromatic to total targeted aromatic and alkane hydrocarbons (blue). This ratio is around 14.6 in the riser oil sample. The C17/pristane and C18/phytane ratios showing little biodegradation are bordered in black.

Close modal

Stranded oil residues were slightly more weathered than floating residues and this was caused by microbially degradation as evidenced by changes in the normal to isoprenoid ratios. Interestingly, stranded oil residues also contained unusually low levels to the target aromatics as compared to the sum of target aromatics and target alkanes (shown in blue in Figure 2), as well as being hopane depleted of the aromatics by 70% to 90%. These data suggest that dissolution of stranded oil played an important role in removal from oil residues of the 2 and 3 ringed parent and C1 alkyl homolog aromatics in preference to removal of the normal alkanes in the 13 to 17 carbon number range (also see Figure 3). This is clearly contrary to the conventional wisdom concerning degradation of alkanes prior to aromatics during oil weathering.

Figure 3:

Chromatographic and PAH compositional data as well as methyl and dimethyl Chrysene isomeric data (selected ions 242 and 256) from representation samples collected from Bay Jimmy during the 2010 to 2012 time frame, together with data from the source oil and a laboratory weathered source oil after 3 days of weathering. Most methyl and dimethyl chrysene isomers had shown extensive weathering when compared to the isomer profile of the source and lab weathered oil.

Figure 3:

Chromatographic and PAH compositional data as well as methyl and dimethyl Chrysene isomeric data (selected ions 242 and 256) from representation samples collected from Bay Jimmy during the 2010 to 2012 time frame, together with data from the source oil and a laboratory weathered source oil after 3 days of weathering. Most methyl and dimethyl chrysene isomers had shown extensive weathering when compared to the isomer profile of the source and lab weathered oil.

Close modal

Weathering of stranded oil in the 2010 to 2011 time period demonstrated that the various alkyl homolog isomers were not being lost at the same rates, with certain isomers being degraded in preference to other isomers within a family of PAH compounds. This is shown in Figure 3 for the methyl and dimethyl isomers of the 228 molecular weight (MW) PAH compounds (benzoanthracenes, chrysene and triphenylene isomers) in representative stranded surface marsh samples collected early in the spill (2010 to 2012). Interestingly, the predominant methyl chrysene isomer (MW 242) in most residues is methyl triphenylene, not methyl chrysene, as shown by high resolution GC separations of the chrysene and triphenylene isomers (in preparation 2020). Notice that lab microbial weathering did not show similar weathering patterns for the 242 C1 chrysenes homolog and 256 C2 chrysenes homolog isomers of these PAH compounds (Olson et al., 2017). We believe this supports the microbial degradation of the alkyl chrysenes in preference to degradation of methyl-triphenylene in coastal environments, which has a symmetrical structure compared to the chrysenes with an open structure, thus allowing enzyme mediated degradation. Also, the sulfur containing PAHs seems to be degraded at a faster rate than the similar sized hydrocarbon PAH compounds. These follow a microbially mediated weathering pattern (Hazen et al., 2016).

Visible oil residues stranded on sandy beaches and in coastal marshes impacted by the oil spill were degraded to below ~10% of their peak concentration by 2012/13. However, the concentrations of hydrocarbons remained ~10 times higher than pre-spill concentrations in the surface of the coastal marsh (Turner et al., 2019).

Oil in and along coastal marshes was distributed as a viscous liquid and in a very nonhomogenous or patchy manner in terms of its mass distribution on and in marshes. Further, it's weathering rates varied dramatically with mass deposited and with time. As is shown in Figure 4, large globs or chunks of oil weathered more slowly than if this same mass has been distributed in small patches or thin layers. Oil residues were mostly found in and on the top 5 cm of the coastal marsh with little mixing deeper into the soil/sediments, with the exception of liquid oil entrapped into fiddler crab borrows 8–10cm deep. This material remained a liquid and was degraded very slowly under anaerobic conditions. Figure 4 shows the alkane and PAH distributions in 8 representative samples (2 per year) collected from surface oily residues in the Bay Jimmy marsh area of Barataria Bay in years 2010/11, 2012, 2014 and 2016. The oil residues were progressively degraded over this time frame, showing losses of the 2 to 3 ringed PAHs compared to the 4 to 6 ringed PAHs over time. We also categorized detectable oil residues in marsh samples as having a biomarker profile pattern similar to the source oil, known as Pattern A, a weathered Pattern A profile known as Pattern AB, and a highly weathered sterane pattern known as Pattern B (Figure 4). Biomarker weathering was used as another gauge to show weathering and degradative removal of stranded oil residues. Virtually all samples that showed biomarker weathering Pattern B had been heavily depleted of normal alkanes and PAH compounds including complete loss of the 2 and 3 ringed PAHs and their C1 to C4 alkyl homologs. Interestingly, in samples with only small amounts of oil residue, even the sterane biomarker compounds were weathered starting as early as the summer of 2010, producing biomarker the Pattern B. Visible oil residues stranded on sandy beaches and in coastal marshes impacted by the oil spill were degraded to below 10% of their peak concentration by 2012/13. Small amounts of the oil residues were entrained in fiddler crab borrows in anaerobic conditions, and these showed little degradation even into the 2016 and 2018 time frames, and these residues maintained a biomarker Pattern A.

Figure 4:

Outline of the chromatographic and PAH compositions from the top 5 cm sediment cores taken from coastal marsh samples impacted by stranded oil residues from the DWH oil spill over time from 2010–11 to 2016. Small amounts of oil was trapped in fiddler crab borrows in 2010 and underwent slow anaerobic degradation over time as compared to surface sediments.

Figure 4:

Outline of the chromatographic and PAH compositions from the top 5 cm sediment cores taken from coastal marsh samples impacted by stranded oil residues from the DWH oil spill over time from 2010–11 to 2016. Small amounts of oil was trapped in fiddler crab borrows in 2010 and underwent slow anaerobic degradation over time as compared to surface sediments.

Close modal

Figure 5, on the other hand, shows the average PAH compositional profile and quantitative analytical data from marsh surficial sediments, collected in even years from 2010 to 2018 for samples containing detectable forensic biomarkers (shown in blue) and samples (shown in green) that did not contain detectable levels of forensic biomarker compounds. The average PAH concentrations jumped two orders of magnitude in 2010 between samples containing oil biomarkers compared to those that did not. These quantities had returned to approximately the same levels in samples with and without biomarker compounds by 2012/4. Interestingly, Tropical Storm Isaac impacted the Louisiana coastal marsh in 2012, and average PAH levels in samples with and without forensic biomarkers were essentially similar thereafter, even considering the compositional profiles from 2014 onward. In 2016 and 2018, average PAH concentrations and their profiles showed a distinctive pyrogenic PAH pattern, mixed with very low levels of petrogenic 2 to 3 ringed PAHs. The source of these 2 to 3 ringed PAHs in years 2016 to 2018 is difficult to understand because these light PAH compounds are known to be rapidly degraded during weathering.

Figure 5.

Sampling location (top left corner), and average PAH distributions in the top 5 cm of sediment cores collected from locations in the Louisiana coastal marshes that were impacted by the DWH oil spill in 2010 to 2016

Figure 5.

Sampling location (top left corner), and average PAH distributions in the top 5 cm of sediment cores collected from locations in the Louisiana coastal marshes that were impacted by the DWH oil spill in 2010 to 2016

Close modal

The concentration of hydrocarbons remained ~10 times higher than pre-spill concentrations in the surface of the coastal marsh (Turner et al., 2019). Hydrocarbons released into the environment from oil residues are commonly called petrocarbon, meaning these specific hydrocarbons were derived from petroleum. Interestingly, petrocarbon's 14C isotopic composition is zero, compared to carbon in plants and animals recently collected from the environment. Therefore, carbon from oil spills can be differentiated from carbon in terrestrial plants and animals using their respective 14C isotopic abundances. This Figure 6 is a conceptual drawing that illustrates the quantity of petrocarbon put into the environment during spills when compared to the quantity of oil and oil residues detectable after the spill. Weathering degraded oil and most of the spilled material is mineralized into either CO2 or biomass produced from oil metabolism. Oil and natural gas released in offshore waters was rapidly degraded during the summer of 2010. Some of the floating offshore oil washed ashore and was stranded on coastal beaches and in marshes. This stranded oil was degraded more slowly than floating oil, with concentrations reduced by ~90% after 2012/13 period. However, some small amount of liquid oil, trapped in below ground burrows, remain after 10 years as readily identifiable liquid oil undergoing only very slow anaerobic degradations. Solid insoluble oil residues remain in marsh areas impacted by the spill in 2010.

Figure 6:

The figure outlines the general quantities of detectable (both analytically and visually) oil residues over time during the Deepwater Horizon oil spill. Offshore oil residues were 90% depleted by September 2010 time frame as shown from sediment trap data collected near the wellhead in slope and deep waters (Stout and German, 2019). Oil residues that were stranded on beaches and in marshes were depleted by over 90% of the target alkanes and PAHs during the 2011/12 time frame (Turner et al., 2019). This figure highlights the transformation of petrocarbon in oil from petrogenic compounds like alkanes and PAHs into CO2, biomass, and heavily degraded and insoluble oil residues. Petrocarbon remains in the environment as CO2, biomass and insoluble and heavily degraded oil residues, whereas the liquid oil in offshore environments was degraded beyond recognition by Fall of 2010, and in stranded oil residues by 2012/13, except for small residues that was entrained in fiddler crab borrows.

Figure 6:

The figure outlines the general quantities of detectable (both analytically and visually) oil residues over time during the Deepwater Horizon oil spill. Offshore oil residues were 90% depleted by September 2010 time frame as shown from sediment trap data collected near the wellhead in slope and deep waters (Stout and German, 2019). Oil residues that were stranded on beaches and in marshes were depleted by over 90% of the target alkanes and PAHs during the 2011/12 time frame (Turner et al., 2019). This figure highlights the transformation of petrocarbon in oil from petrogenic compounds like alkanes and PAHs into CO2, biomass, and heavily degraded and insoluble oil residues. Petrocarbon remains in the environment as CO2, biomass and insoluble and heavily degraded oil residues, whereas the liquid oil in offshore environments was degraded beyond recognition by Fall of 2010, and in stranded oil residues by 2012/13, except for small residues that was entrained in fiddler crab borrows.

Close modal

Visible and detectable oil residues in offshore waters, biota and sediments were rapidly degraded. Visible oil residues stranded on sandy beaches and in coastal marshes impacted by the oil spill were degraded to below ~10% of their peak concentration by 2012/13. The concentrations of hydrocarbons remained 10 times higher than pre-spill concentrations in the surface of the coastal marsh (Turner et al., 2019). Heavily weathered oil residues, essentially asphaltene like debris can still be detected on marsh shoreline and in some deep-water sediments. Further, while there are petrocarbon resides in detectable levels in these environments, the potential for their causing significant environmental damage is very small due to their insolubility and/or conversion into biomass.

We thank the many oil spill research colleagues for consultation, field assistance and general support. This research was made possible by funding from the Gulf of Mexico Research Initiative to the Coastal Waters Consortium. The financial sources had no role in the design or execution of the study, data analysis, decision to publish, or manuscript preparation. The data are publicly available through the Gulf of Mexico Research Initiative Information & Data Cooperative (GRIIDC) at https://data.gulfresearchinitiative.org (doi: 10.7266/N7028PZZ; 10.7266/n7-czff-sm91; 10.7266/n7-tph2-3e25; 10.7266/n7-kdeg-nj49).

Adhikari,
P.L.,
Maiti,
K.,
Overton,
E.B.,
Rosenheim,
B.E.
and
Marx,
B.D.,
2016
.
Distributions and accumulation rates of polycyclic aromatic hydrocarbons in the northern Gulf of Mexico sediments
.
Environmental pollution
,
212
, pp.
413
423
.
Albers,
P.H.,
1995
.
Petroleum and individual polycyclic aromatic hydrocarbons
.
Handbook of ecotoxicology
,
2
, pp.
330
355
.
Aeppli,
C.,
Nelson,
R.K.,
Radović,
J.R.,
Carmichael,
C.A.,
Valentine,
D.L.,
and
Reddy,
C.M.
2014
.
Recalcitrance and degradation of petroleum biomarkers upon abiotic and biotic natural weathering of Deepwater Horizon oil
.
Environmental Science and Technology
,
48
:
6726
6734
.
Atlas,
R.M.
1981
.
Microbial degradation of petroleum hydrocarbons: an environmental perspective
.
Microbiological Reviews
,
45
(
1
):
180
209
.
Atlas,
R.M.
and
Hazen,
T.C.
2011
.
Oil biodegradation and bioremediation: a tale of the two worst spills in U.S. history
.
Environmental Science and Technology
,
45
:
6709
6715
.
Atlas,
R.M.,
Stoeckel,
D.M.,
Faith,
S.A.
Minard-Smith,
A.,
Thorn,
J.R.,
and
Benotti,
M.J.
2015
.
Oil biodegradation and oil-degrading microbial populations in marsh sediments impacted by oil from the Deepwater Horizon well blowout
.
Environmental Science and Technology
,
49
(
14
):
8356
8366
.
Hazen,
T.C.,
Prince,
R.C.,
and
Mahmoudi,
N.,
2016
.
Marine Oil Biodegradation
.
Environmental Science and Technology
,
50
,
2121
2129
.
Meyer,
B.M.,
Adhikari,
P.L.,
Olson,
G.M.,
Overton,
E.B.
and
Miles,
M.S.,
2018
.
Louisiana coastal marsh environments and MC252 oil biomarker chemistry
.
In
Oil Spill Environmental Forensics Case Studies
(
pp.
737
756
).
Butterworth-Heinemann.
Olson,
G.M.,
Gao,
H.,
Meyer,
B.M.,
Miles,
M.S.
and
Overton,
E.B.,
2017
.
Effect of Corexit 9500A on Mississippi Canyon crude oil weathering patterns using artificial and natural seawater
.
Heliyon
,
3
(
3
),
p.
e00269
.
Overton,
E.B.,
Wetzel,
D.L.,
Wickliffe,
J.K.
and
Adhikari,
P.L.,
2020
.
Spilled Oil Composition and the Natural Carbon Cycle: The True Drivers of Environmental Fate and Effects of Oil Spills
.
In
Scenarios and Responses to Future Deep Oil Spills
(
pp.
33
56
).
Springer, Cham
.
Peters,
K.E.,
Walters,
C.C.,
and
Moldowan,
J.M.
2005
.
The Biomarker Guide
, 2nd Edition.
Cambridge, UK
:
Cambridge University Press
.
Prince,
R.C.,
Elmendorf,
D.L.,
Lute,
J.R.,
Hsu,
C.S.,
Haith,
C.E.,
Senius,
J.D.,
Dechert,
G.J.,
Douglas,
G.S.,
and
Butler,
E.L.
1994
.
17α(H),21β(H)-hopane as a conserved internal marker for estimating the biodegradation of crude oil
.
Environmental Science and Technology
,
28
:
142
145
.
Prince,
R.C.
and
Walters,
C.C.
2007
.
Biodegradation of oil hydrocarbons and its implications for source identification
.
In
Oil Spill Environmental Forensics: Fingerprinting and Source Identification
,
Wang,
Z.
and
Stout,
S.A.
(
eds.
),
Burlington, MA
:
Academic Press
,
pp.
349
379
.
Reddy,
C.M.,
Arey,
J.S.,
Seewald,
J.S.,
Sylva,
S.P.,
Lemkau,
K.L.,
Nelson,
R.K.,
Carmichael,
C.A.,
McIntyre,
C.P.,
Fenwick,
J.,
Ventura,
G.T.
and
Van Mooy,
B.A.,
2012
.
Composition and fate of gas and oil released to the water column during the Deepwater Horizon oil spill
.
Proceedings of the National Academy of Sciences
,
109
(
50
), pp.
20229
20234
.
Tarr,
M.A.,
Zito,
P.,
Overton,
E.B.,
Olson,
G.M.,
Adhikari,
P.L.
and
Reddy,
C.M.,
2016
.
Weathering of oil spilled in the marine environment
.
Oceanography
,
29
(
3
), pp.
126
135
.
Turner,
R.E.,
Rabalais,
N.N.,
Overton,
E.B.,
Meyer,
B.M.,
McClenachan,
G.,
Swenson,
E.M.,
Besonen,
M.,
Parsons,
M.L.,
Zingre,
J.,
2019
.
Oiling the continental shelf and coastal wetlands over eight years after the 2010 Deepwater Horizon oil spill
.
Environmental Pollution
,
252
(
Part B
):
1367
1376
.
Wang,
Z.,
Fingas,
M.F.,
Owens,
E.H.,
Sigouin,
L.,
and
Brown,
C.E.
2001
.
Long-term fate and persistence of the spilled Metula oil in a marine salt marsh environment: degradation of petroleum biomarkers
.
Journal of Chromatography A
,
926
:
275
290
.
Wang,
Z.
and
Fingas,
M.F.
2003
.
Development of oil hydrocarbon fingerprinting and identification techniques
.
Marine Pollution Bulletin
,
47
:
423
452
.