biodegradation of spilled fuel oil in norwegian marine ... · terrestrial and marine, m...

SINTEF Ocean AS Fate and Effects 2018-01-10

OC2017 A-218- Unrestricted

Biodegradation of Spilled Fuel Oil in Norwegian Marine Environments A Literature Review Author(s) Jane Helén Carlsen Øksenvåg, Kelly McFarlin, Roman Netzer, Odd Gunnar Brakstad, Bjørn Henrik Hansen, Trond Størseth

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KEYWORDS:

Biodegradation

Fuel Oil

Crude Oil

Seawater

Sea ice

Sediments

Ecotoxicology

KYSTVERKET

Report

The Fate of Spilled Fuel Oil in

Norwegian Marine Environments A Literature Review

VERSION

Final version

AUTHOR(S)

DATE

2018-01-10

Jane Helen Carlsen Øksenvåg, Kelly McFarlin, Roman Netzer, Odd Gunnar Brakstad, Bjørn

Henrik Hansen, Trond Størseth

CLIENT(S)

The Norwegian Coastal Administration

PROJECT NO.

302003240

ABSTRACT

CLIENT'S REF.

Rune Bergstrøm

NUMBER OF PAGES/APPENDICES:

58 + Appendices

This report summarizes the current state of knowledge concerning the biodegradation of

fuel oils in cold marine environments, with an emphasis on the coastline of Norway and

the Arctic region dose to Svalbard. The biodegradation potential of oil in seawater, sea

ice, and coastal sediments was reviewed. To date, the majority of research pertaining to

the biodegradation of oil in cold environments has focused on crude oils. Here we bu ild

upon this existing knowledge to understand the potential for fuel oil to biodegrade in

cold marine environments. Recommendations for future biodegradation studies are

described and include experiments with the new generation fuel oils and traditional fuel

oils at various weathering stages, also in the presence of difference ice conditions.

REPORT NO.

OC2017 A-218

ISBN

978-82-7174-322-2

CLASSIFICATION

Unrestricted

CLASSIFICATION THIS PAGE

Unrestricted

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PREPARED BY

Jane H.C. Øksenvåg

CHECKED BY

Julia Farkas

APPROVED BY

Mimmi Throne-Holst

PROJECT NO. 302003240

REPORT NO. OC2017 A-218

VERSION Final version

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Document history VERSION DATE VERSION DESCRIPTION

Draft version 2017-10-23 Draft

Final version 2018-01-10 Final




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Table of contents EXECUTIVE SUMMARY ........................................................................................................................... 5

SAMMENDRAG ...................................................................................................................................... 6

1 Background ................................................................................................................................... 7

2 Introduction .................................................................................................................................. 7 2.1 Marine fuel oils – classification and properties ............................................................................. 7 2.2 Weathering of fuel oils and further fate ...................................................................................... 11

2.2.1 Oil composition ................................................................................................................ 11 2.2.2 The importance of the fuel oil physical and chemical properties ................................... 13 2.2.3 Fate and behaviour of oil spills at sea ............................................................................. 14 2.2.4 Weathering properties of new diesel and fuel oil products used in the Arctic ............... 18

3 Biodegradation of oils in seawater ............................................................................................... 22 3.1 Microbiology of the Arctic oceans ............................................................................................... 22 3.2 Biodegradation of oil in seawater ................................................................................................ 24

3.2.1 Oil sedimentation processes ........................................................................................... 28 3.2.2 The deepwater Horizon oil spill and the fate of the oil ................................................... 30

3.3 Bioremediation strategies at sea ................................................................................................. 31 3.3.1 Dispersibility of bunker fuel oils ...................................................................................... 32

3.4 Summary and recommendations – biodegradation at sea ......................................................... 34

4 Biodegradation of oil in sea ice .................................................................................................... 35 4.1 Marine ice and associated microorganisms ................................................................................. 36 4.2 Oil in the ice – different scenarios ............................................................................................... 37 4.3 Oil-degrading microorganisms in sea ice ..................................................................................... 38 4.4 Degradation of oil in sea ice ......................................................................................................... 39 4.5 Bioremediation strategies in ice .................................................................................................. 40 4.6 Summary and recommendations – biodegradation of oil in ice.................................................. 40

5 Biodegradation of oil in sediments ............................................................................................... 41 5.1 Shoreline sediments ..................................................................................................................... 41 5.2 Bioremediation strategies in sediments ...................................................................................... 42 5.3 Seafloor sediments ...................................................................................................................... 44 5.4 Summary and recommendations – biodegradation of oil in sediments ..................................... 45

6 Ecotoxicity of biodegraded oil ...................................................................................................... 47 6.1 Future Studies of Ecotoxicity ....................................................................................................... 47

7 References .................................................................................................................................. 48




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Appendix ............................................................................................................................................. 58

A Abbreviations .............................................................................................................................. 58




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EXECUTIVE SUMMARY Summer sea ice coverage in the Arctic has reached the lowest extent on record and continues to open new opportunities for many industries (Comiso, 2008). Areas that were once covered in ice year-round are now accessible for shipping, tourism, and exploration. This projected rise in activity increases the risk of oil spills and generates the need to advance environmental research to address data gaps regarding the fate and effects of oil. The majority of oil biodegradation research in cold marine environments has focused on crude oil; however, as shipping continues to increase in Arctic waters research must also determine the fate of fuel oils. The objective of this literature review is to access the state of knowledge of fuel oil biodegradation in Norwegian marine environments. Not much research has been done on weathering and further fate of fuel oils. In order to understand the potential for fuel oil to biodegrade in cold environments, it has therefore been important to build upon existing knowledge on crude oil and use their physical and chemical similarities as a basis. This report discusses the potential for oils, including marine fuel oils where information has been available, to biodegrade in seawater, sea ice, and marine sediments. Data gaps are identified and recommendations for future research are included to provide a framework for relevant research for the years to come. Oil-degrading microorganisms have been discovered from pole to pole and are thought to be ubiquitous (Schneiker et al., 2006; Head et al., 2006; Yakimov et al., 2007). In a variety of environments, both terrestrial and marine, microorganisms have evolved over time to utilize petroleum hydrocarbons as a source of carbon and energy (Prince et al., 2010). Natural oil seeps have been discovered throughout the World’s oceans, including in the Arctic (NRC, 2005), and these seeps continue to enrich oil-degrading microorganisms (Prince & Clark, 2004). Oil-degrading bacteria in sub-Arctic marine sediments were studied for many years in the aftermath of the Exxon Valdez oil spill, and recently, the biodegradation of oil in Arctic seawater by indigenous bacteria has been quantified. In contrast, relatively little is known concerning the ability of bacteria to biodegrade oil in and around sea ice. Researchers are now trying to understand the limitations surrounding the availability of oil to oil-degrading microorganisms in the presence of diverse types of ice. Understanding the biodegradation of oil in different ice conditions remains incredibly important and its fate and effects remain unknown. Oil in ice biodegradation studies in ice are highly recommended using various fuel types, from heavy fuel oils to light fuel oils. Further studies should also focus on the biodegradation of new hybrid fuel oils at different stages of weathering. While the aerobic biodegradation in marine sediments is documented, knowledge concerning anaerobic biodegradation is lacking, especially within deep-sea sediments. Future studies are also necessary to understand the effects of oil-related marine snow on deep-sea organisms and on deep-sea oil biodegradation. Limited information exists on the toxicity of oil degradation products (i.e. metabolites) to organisms in seawater, sea ice, and marine sediments. Numerous studies in the aftermath of the Prestige (2002) and Exxon Valdez (1989) oil spill have shown that biodegradation of oil in coastal sediments can be strongly stimulated by addition of nutrients/fertilizers. However, reliable data on a systematic evaluation of bioremediation products available in the market is still missing. This operational relevant knowledge gap should be filled by experiments testing the effect of selected bioremediation products on biodegradation of oil in marine sediments. To sum up, further studies are recommended to focus on oil biodegradation processes, rates and the toxicity during the process (acute and chronic effects), preferably on fuel oils used today at different weathering degrees. These studies should be performed in the presence and absence of dispersants, or with the use of biostimulation products, at sea, in ice and in sediments.




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SAMMENDRAG Isdekningsgraden I Arktis har for sommerhalvåret blitt registrert til å nådd sitt laveste, noe som gjør at mulighetene for flere industrier utvider seg (Comiso, 2008). Områder som tidligere var dekt med is året rundt, er nå tilgjengelig for skipsfart, turisme, borrevirksomhet og utforskning. Denne økningen i aktivitet, øker risikoen for oljesøl og genererer et behov for videre miljørettet forskning for å adressere kunnskapshull knyttet til skjebne og effekt av oljesøl. Hovedsakelig har forskning på biodegradering av olje omhandlet råolje, nå som skipsfarten fortsetter å øke i Arktiske farvann, må forskning fremover ha et større fokus på skjebne og effekter av marint drivstoff. Målet med denne litteraturstudien er å sammenfatte tilgjengelig litteratur på biodegradering av marint drivstoff i det marine miljø i Norge. Det er gjort lite forskning på forvitring og videre skjebne av marint drivstoff. For å få en bedre forståelse av marint drivstoffs evne til å biodegradere under Arktiske betingelser, har det derfor vært viktig å bygge på eksisterende kunnskap om råoljer, samt og bruke fysiske og kjemiske likheter mellom disse. Denne rapporten diskuterer potensialet for olje, inkludert marint drivstoff der informasjon har vært tilgjengelig, å biodegradere i sjøvann, sjøis og marine sedimenter. Kunnskapshull er identifisert og anbefalinger for videre forskning er diskutert for å fremheve nødvendig forskning i årene som kommer. Oljedegraderende mikroorganismer er blitt påvist fra pol til pol og er ansett for å være tilstedeværende over alt i det marine miljø (Schneiker et al., 2006; Head et al., 2006; Yakimov et al., 2007). Mikroorganismer har, både i det marine og terrestrisk miljø, utviklet seg til å nyttiggjøre seg hydrokarboner fra petroleumsindustrien som en kilde til karbon og energi (Prince et al., 2010). Kilder med naturlige utslipp av olje har blitt oppdaget i hav over hele verden, inkludert i Arktis (NRC, 2005), og disse utslippene fortsetter å berike oljedegraderende mikroorganismer (Prince & Clark, 2004). Oljedegraderende bakterier i sub-Arktiske marine sedimenter ble studert i mange år i etterkant av oljesølet Exxon Valdez, mens biodegradering av olje i Arktisk sjøvann av naturlig tilstedeværende bakterier er nylig blitt kvantifisert. Det er derimot lite kunnskap om bakteriers egenskap til å biodegradere olje i og rundt sjøis. Forskere jobber nå med å kartlegge muligheter og begrensninger for mikroorganismer å bryte ned råolje ved forskjellige isbetingelser. Å få en bedre forståelse av biodegradering av olje under forskjellige isbetingelser er viktig og oljens skjebne og effekt er fortsatt i stor grad uviss. Forskning på biodegradering i is av forskjellige typer marint drivstoff, ved forskjellige er derfor høyst relevant. Det vil videre være spesielt viktig å fokusere på toksisitet av oljedegraderingsprodukter (d.v.s. metabolitter) på organismer i sjøvann, sjøis og marine sedimenter. Mens aerob biodegradering i marine sedimenter er dokumentert, mangler det fortsatt kunnskap om anaerob biodegradering av olje, spesielt i dypvannssediment. Flere studier på dypvanns oljebiodegradering, samt effekten av oljebefengt marin snø på dypvannsorganismer vil også være viktig. Begrenset informasjon er tilgjengelig på Flere studier, utført I etterkant av oljesølet med Prestige (2002) og Exxon Valdez (1989), har vist at biodegradering av olje i strandsediment kan bli stimulert ved påføring av næringsstoffer/gjødsel, men det mangler fortsatt systematisk forskning knyttet til effektiviteten av remedieringsprodukter tilgjengelig på markedet i dag. Fremtidig forskning bør fokusere på å få fylt dette operasjonelle kunnskapshullet. For å oppsummere vil det for videre studier være anbefalt å fokusere på biodegraderingsprosesser, -rater og toksisitet som følge av denne prosessen (akutt og kroniske effekter), fortrinnsvis på marint drivstoff ved forskjellige forvitringsgrader. Disse studiene burde bli utført med og uten bruk av dispergeringsmiddel, eller ved bruk av biostimulerende produkter, både på sjø, i is og i sedimenter.




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1 Background Shipping traffic is increasing due to the decline of Arctic sea ice. This increased activity in the Norwegian Arctic and surrounding areas poses a risk to the environment through the potential release of marine fuel oil from commercial ships and holding tanks. Understanding the fate of fuel oil is crucial to understanding its impact on marine life. Previous oil biodegradation studies have demonstrated the ability of Arctic and sub-Arctic microorganisms to biodegrade oil. This report summarizes what is known to date concerning the biodegradation of marine fuel oils in open seawater, marine ice and coastal sediments. This important summary was initiated by the Norwegian Coastal Administration and may assist spill responders, fate modelers, and regulators as they seek to understand the fate of spilled fuel oil in and around the Norwegian coast, including Svalbard and other cold environments.

2 Introduction Oil-degrading microorganisms have been discovered from pole to pole and are thought to be ubiquitous (Schneiker et al., 2006; Head et al., 2006; Yakimov et al., 2007). In a variety of environments, both terrestrial and marine, microorganisms have evolved over time to utilize petroleum hydrocarbons as a source of carbon and energy (Prince et al., 2010). Biodegradation is regarded as the main process that removes oil from the environment (NRC, 2005). A community of microorganisms work together to biodegrade oil and these organisms will continue to obtain energy from oil until it is completely broken down into carbon dioxide and water. Some bacteria can only use oil compounds for energy (i.e. hydrocarbonoclastic bacteria) while other bacteria can use a variety of carbon sources, including oil (Prince et al., 2004). Typically, the abundance of oil-degrading bacteria in the natural environment is limited by the presence of oil, but once oil is present their abundance dramatically increases (Prince et al., 2005). The two largest oil spills in marine waters include the Exxon Valdez oil tanker spill in Prince William Sound (1989) and the recent Deepwater Horizon deep-sea blowout (2010). These spills have drawn attention to the importance of biodegradation and the use of this natural process as an oil spill response strategy is highlighted in a review by Atlas and Hazen (2011).

2.1 Marine fuel oils – classification and properties Marine fuel oils are obtained from crude oils after a complex refining process involving atmospheric distillation and the refining of distillates. Marine fuels consist of a combination of distillate oil and residual oil and the amount of each determines its classification. The International Standard Organisation (ISO) reports quality specifications for marine fuel oils. Marine fuels are defined in the ISO 8217 and are designated by a group of letters, which constitute a code. The categories of fuels consist of three letters (e.g. DMX or RMG). The first letter identifies the oil as either a distillate “D” or a residual “R”. The second letter designates the fuel oil as marine “M”. The third letter does not represent a specific word, but is used to identify an oil with different properties. Marine Fuel oils and can be divided in the following classes (Moldestad and Daling, 2006):

• Marine Gas Oil (MGO) (In Norwegian: "Marine gassoljer") – is considered as a light gas oil due to its high content (~60%) of aromatic hydrocarbons. MGOs are commonly classified as DMX, DMA, DMB, and DMZ according ISO 8217 “Petroleum Products – Fuel (class F)” (Table 2.1). It may be important to note that DMB fuel oil also can contain a small proportion of heavy fuel oil, therefore DMB is not a pure distillate and thus not a “real” marine gas oil. DMB is classified as a marine diesel oil.




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• Marine Special Distillate (MSD) (In Norwegian: "Tyngre destillater") – is made from heavier distillate of crude oils and is often referred to as Heavy Gas Oil (HGO), Wide Range Oil (WRO), and Special Distillate Marine (SDM). This type of marine fuel can contain a large amount of wax, and usually contains an additive to reduce its viscosity. The termination "IF-10" is often used to describe MSD and it is classified as DMB or DFB according the ISO 8217 standard (Table 2.1).

• Marine Residual Fuel Oil (RFO) (In Norwegian: "Bunkersoljer/tungoljer") – This fuel group is a group identified with a high viscosity and high density. These are made from a mix of residue and cutter stock (distillate diluent, e.g. marine diesel oil or marine gas oil) blended to the desired combination. The properties, both physical and chemical, will thus vary based on the quality or properties of the crude oil used in the distillation process, the variation of distillate added to give the required viscosity and the differences in refinery processes (Moldestad and Daling, 2007; Lewis 2002). ISO 8217 classifies RFO as RMA, RMB, RMD, RME, RMG, or RMK (Table 2.2).

ISO 8217 has existed since 1987 and a major update was introduced in 2017. The purpose of ISO 8217 is to define the requirements for petroleum fuels used in marine diesel engines and boilers as a guide to interested parties (i.e. equipment designers, suppliers, and distributers). These standards can also be useful to spill responders and environmental regulators, since marine fuel oils are classified based upon their chemical and physical characteristics. For spill responders, properties as viscosity, density, flash point, and pour point would be of most interest, while the content of sulphur and hydrogen sulphide maybe of interest to environmental regulators. The physical and chemical characteristics of classified distillate marine fuel oils are listed in Table 2.1 and the characteristics of residual marine fuel oils are listed in Table 2.2.




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Table 2.1 Classification criteria for distillate marine fuels according ISO 8217 (2017, 6th edition) (https://www.iso.org)

Characteristics Unit Limit Category ISO-F- DMX DMA DFA DMZ DFZ DMB DFB

Kinematic viscosity at 40 °C mm2/s Max. 5.500 6.000 6.000 11.00 Min. 1.400 2.000 3.000 2.000

Density at 15 °C kg/m3 Max. — 890.0 890.0 900.0 Cetane index — Min. 45 40 40 35 Sulphur mass % Max. 1.00 1.00 1.00 1.50 Flash point °C Min. 43 60 60 60 Hydrogen sulfide mg/kg Max. 2.00 2.00 2.00 2.00 Acid number mg KOH/g Max. 0.5 0.5 0.5 0.5 Total sediment by hot filtration mass % Max. — — — 0.10 Oxidation stability g/m3 Max. 25 25 25 25 Fatty acid methyl ester (FAME) volume % Max. — — 7.0 — 7.0 — 7.0 Carbon residue: micro method on the 10 % volume distillation residue mass % Max. 0.30 0.30 0.30 —

Carbon residue: micro method mass % Max. — — — 0.30

Cloud point winter °C Max. -16 report report — summer °C Max. -16 — — —

Cold filter plugging point winter °C Max. — report report — summer °C Max. — — — —

Pour point (upper) winter °C Max. — -6 -6 0 summer °C Max. — 0 0 6

Appearance — — Clear and bright Water volume % Max. — — — 0.30 Ash mass % Max. 0.010 0.010 Ash mass % Lubricity, corrected wear scar diameter (wsd 1.4) at 60 °C µm Max. 520 520 520 520




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Table 2.2 Classification criteria for residual marine fuels according ISO 8217 (2017, 6th edition) (https://www.iso.org)

Characteristics Unit Limit Category ISO-F- RMA RMB RMD RME RMG RMK 10 30 80 180 180 380 500 700 380 500 700

Kinematic viscosity at 40°C mm2/s Max. 10 30 80 180 180 380 500 700 380 500 700

Density at 15 °C kg/m3 Max. 920 960 975 991 991 1010

CCAI Max. 850 860 860 860 870 870 Sulphur mass % Max. Statitory requirements Flash point °C Min. 60 60 60 60 60 60 Hydrogen sulfide mg/kg Max. 2 2 2 2 2 2 Acid number mg KOH/g Max. 2.5 2.5 2.5 2.5 2.5 2.5 Total sediment - Aged mass % Max. 0.1 0.1 0.1 0.1 0.1 0.1 Carbon residue: micro method mass % Max. 2.5 10 14 15 18 20

Pour point (upper)

winter °C Max. 0 0 30 30 30 30 summer °C Max. 6 6 30 30 30 30

Water volume % Max. 0.3 0.5 0.5 0.5 0.5 0.5 Ash mass % Max. 0.04 0.07 0.07 0.07 0.1 0.15 Vanadium mg/kg Max. 50 150 150 150 350 450 Sodium mg/kg Max. 50 100 100 50 100 100 Aluminium plus silicon mg/kg Max. 25 40 40 50 60 60 Used lubricating oil (ULO): Calcium and zinc Or Calsium and phosphorus

mg/kg Calcium >30 and zinc >15 or Calcium >30 and phosphorus >15

In 2014, Intermediate Fuel Oil (IFO) 180 and IFO 380 (Table 2.3) was reported as the most common bunker fuel oils used on ships along the Norwegian coast (Statlig dispergering, 2014). The density, and content of distillates and residuals of different IFOs are shown in Table 2.3. IFO 180 is considered a Medium Fuel Oil (MFO) and IFO 380 is classified as a Very Heavy Fuel Oil (VHFO). Table 2.3 IFO grade of marine residual fuel oils (Moldestad and Daling, 2006).

IF-grade Density [kg/L] Distillate (“flux”) [Vol%] Heavy residue Vol%] Light Fuel Oil (LFO) IF30 0.93 35-40 60-65 Medium Fuel Oil (MFO) IF80 0.93-0.96 18-30 70-80

IF180 0.94-0.97 5-20 80-92 Heavy Fuel Oil (HFO) IF240 0.96-0.98 3-12 90-95 Very Heavy Fuel Oils (VHFO/LAPIO) IF380 0.97-0.99 0-10 90-100

In EU, the Baltic Sea and the North Sea, Sulphur Emission Control Areas (SECA) has been introduced to limit the sulphur content to 0.1 % and were effectuated January 2015 (PAME 2016). Further, a global limit of 0.5% sulphur is expected to come in to force in 2020 or 2025. The current sulphur content for residual oils is stated in ISO 8217 and defined by statutory requirements, e.g. national or international emission requirements (e.g. EU Sulphur Directive).




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Due to these regulations, today the most common marine fuels also include marine gas oil (DMA) and marine diesel oil (DMB). In addition, new regulations are introduced in order to protect national parks and nature reserves on Svalbard. The new regulations for use and loading of fuel oils came in to force January 1, 2015 and require a higher quality of fuel oils for ships traveling in the three natural reserves on the east side of Svalbard and in the national parks in the west https://www.sysselmannen.no/en/Shortcuts/Ban-on-heavy-fuel-oil/. Other requirements to reduce air pollution from fossil fuels have also been effectuated. The MARPOL Annex VI (International Convention for the Prevention of Pollution from Ships) was revised and strengthened to reduce the global emissions of NOx, SOx, and Particulate Matter (PM) and to introduce Emission Control Areas (ECA) to further reduce air pollution in designated areas (http://www.imo.org/en/OurWork/Environment/PollutionPrevention/AirPollution/Pages/Air-Pollution.aspx). Ships built after 2015 are required to reduce their NOx emissions by 75 %, compared to current emission standards for international shipping (http://www.pbl.nl/en/publications/2012/assessment-of-the-environmental-impacts-and-health-benefits-of-a-nitrogen-emission-control-area-in-the-north-s) in Nitrogen Emission Control Areas (NECA). To meet these requirements, a new generation of fuel oils (also called hybrid fuel oils) has been developed and is produced and used at increasing rates as bunker fuel, both within and outside the SECA area today. The sulphur content of these hybrid fuel oils is less than 0.1 % and often referred to as Ultra Low Sulphur Fuel Oil (ULSFO) or Hybrid Wide Range Gas oil (HDME 50). Hybrid fuel oils are not easily classified within the ISO 8217 standards, and exhibit properties of both light and heavy distillate fuels. However, low sulphur crude oils, used as feed oil in the refinery process, tend to be rich in waxes (Moldestad et al., 2007), resulting in higher pour points for the fuel products. Oils high in waxes are challenging to remove from the marine environment. Therefore, this new generation fuel oils may introduce a challenge in relation to oil spill response. Hybrid fuel oil are likely to become more common in years to come (The Norwegian Coastal Administration, 2017), it is therefore becoming increasingly important to study their fate in the marine environment.

2.2 Weathering of fuel oils and further fate

2.2.1 Oil composition The composition of petroleum oils is extremely complex and can contain thousands of different compounds. Oil compounds are basically separated into linear or cyclic alkanes, aromatic hydrocarbons, asphaltenes, and resins; and the distributions of these compound groups vary considerably within oils. Most environmental studies of petroleum-derived chemicals have focused on effects related to specific hydrocarbons such as n-alkanes, BTEX (Benzenes, Toluenes, Ethylbenzenes, Xylenes) and polycyclic aromatic hydrocarbons (PAHs). These compounds are typically easy to identify and quantify using standard analytical chemistry techniques such as Gas Chromatography Mass Spectrometry (GC-MS) or by using Flame Ionizing detector techniques (GC-FID). However, in the case of environmentally weathered samples, most oil compounds appear as Unresolved Complex Mixtures (UCM) in gas chromatograms, and are often referred to as the “hump”, see Figure 2.2. These are compounds that are not separable by the traditional chromatographic instruments. 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

https://www.sysselmannen.no/en/Shortcuts/Ban-on-heavy-fuel-oil/

https://www.sysselmannen.no/en/Shortcuts/Ban-on-heavy-fuel-oil/

http://www.imo.org/en/OurWork/Environment/PollutionPrevention/AirPollution/Pages/Air-Pollution.aspx

http://www.imo.org/en/OurWork/Environment/PollutionPrevention/AirPollution/Pages/Air-Pollution.aspx

http://www.pbl.nl/en/publications/2012/assessment-of-the-environmental-impacts-and-health-benefits-of-a-nitrogen-emission-control-area-in-the-north-s

http://www.pbl.nl/en/publications/2012/assessment-of-the-environmental-impacts-and-health-benefits-of-a-nitrogen-emission-control-area-in-the-north-s




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1990). Figure 2.1 shows the compositional distribution of a selection of oils, illustrating that the UCM is larger than the known fraction, constituting as much as 50 % of the compounds detected by GC in fresh oils. In addition, oil is comprised of ~20% of compounds that are not analysable by GC at all.

Figure 2.1 Compositional data for 16 different oils. The GC Unknowns is the UCM components. The "Not possible by GC fraction" are compounds that are not analysable (detectable) by GC.

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 (Brakstad et al., 2017), highlighting the persistence of these compounds. Due to their persistence, these compounds may reach Arctic regions both through local oil releases and as long-range transported pollutants. 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 significantly reduced biodegradation rates, and currently nothing is known about the abilities of psychrophilic (cold lowing) or psychrotrophic (cold tolerant) bacteria to degrade these compounds. 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. Currently, little is known about the identity of compounds within the UCM. As UCM compounds are known to be biodegradation metabolites with potential toxicity, it becomes imperative to update traditional analytical techniques to assist in their identification. By using the novel analysing technique two-dimensional quadrupole-time-of-flight mass spectrometry (GCxGC-QTOF MS), the UCM "hump" can be resolved in 2 dimensions so that individual components may be studied. This technique enables detailed studies of biodegradation patterns and may separate persistent UCM components from intermediate biodegradation products. Figure 2.2 shows an oil analysed by GC-FID, and with GCxGC-QTOF MS. The UCM fraction also seem to include uncharacterized compounds causing chronic effects. In a study where water-accommodated fractions (WAFs) were separated on polarity, the polar fractions associated with the UCM resulted in effects related to different chronic endpoints, while the non-polar fraction, containing the BTEXs, naphthalenes and PAHs, was associated acute toxicity (Melbye et al., 2009). Further research on the




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relations between biodegradation, effects and analytical identification of UCM compounds are therefore important as a part of the risk evaluation of oil spills.

Figure 2.2 Comparison of 1D GC-FID with 2D GCxGC-MS. Figure to the left: The UCM is located under the blue curve and is limited by the red line. Figure to the right: The components of the UCM are separated in two dimensions enabling the study of individual UCM components.

2.2.2 The importance of the fuel oil physical and chemical properties The physical and chemical properties of a fuel oil vary depending upon its origin and classification (i.e. how it is refined), and determine its fate and behaviour during a spill. In addition, effective response options are a factor of the oils current properties. For example, if the oil is highly weathered and emulsified, the application of chemical dispersants may not be the best response option. Important oil parameters include: viscosity, density, pour point, and content of asphaltenes, resins or waxes. For safety reasons, it is also important to know the oil’s flash point and concentration of volatile compounds.

Viscosity is a measure of the resistance to flow. Therefore, oil with a high viscosity moves slower than oil with a low viscosity. This movement is dependent upon the temperature. The viscosity of oil increases with decreasing temperature, as e.g. waxes (and related compounds) crystalize at lower temperatures.

Pour point refers to the temperature at which the oil starts to flow and will thus provide information about the fuels behaviour (e.g. solid or not) at different sea temperatures. Solidification typically arises when the pour point of the oil is 10-15°C above the sea temperature. Each oil type has a specific pour point. Fuel oils with an origin in waxy and paraffinic crude oils, will most likely have a high pour point (Moldestad and Daling 2006).

Density measurements provide information to determine where the fuel can be found; on the surface, just below the surface (submerged) or if has sunken to the sea floor. Density is increased with an increase in weathering degree and the oil loses its lighter components.

Knowledge about the content of asphaltenes, resins or waxes in the fuel is important, as these compounds influences on the stability of the water-in-oil emulsification (Faksness 2008). If the oil’s viscosity, pour point, and density are known, one could infer the content of asphaltenes, resins or waxes.




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2.2.3 Fate and behaviour of oil spills at sea Weathering processes change the physical–chemical characteristics of the oil and influence the fate and behaviour of surface oil spills. These weathering processes are influenced by many factors, such as temperature, waves, currents, wind, sun light and the presence of marine ice. Figure 2.3 illustrates the various fates of oil spilled at sea in conditions expected in Norway waters and around Svalbard. In cold marine environments, oil may be spilled in open water and/or drift into ice, or it may be spilled on top of already formed ice.

Figure 2.3 Schematic illustrating the fate of spilled oil in seawater and marine ice (Modified from Daling et al. (1990)

Immediately after a fuel oil spill in open water, the oil will spread to form an oil slick of about 1mm thick, and the most volatile compounds will start to evaporate. Surface evaporation processes result in losses of short branched alkanes (C5 – C10) and monoaromatic compounds like BTEX. Some oil components are water soluble, and marine distillate oils will likely contain more water-soluble components than residual fuel and crude oils. However, these compounds will also evaporate. Evaporation of marine distillate fuel oils is quite rapid, and can lead to a substantial loss of the total oil. Over 75% of distillate fuel oils can be lost due to evaporation, compared to 20-50% of crude oil and 10% of residual fuel oils within the same time frame (NRC, 2005, 2014). The presence of ice can reduce evaporation and therefore delay the weathering process, which may increase the efficiency of various oil spill response options. Brandvik and Faksness (2009) reported that evaporation of light oil at air temperatures ranging from -15°C to approximately -5°C was higher in open water (30% loss) than in light ice coverage (25% loss) or heavy ice coverage (19% loss). The percent of oil that evaporates is dependent upon its thickness and content of VOCs (Volatile Organic Compounds), as well as environmental parameters (i.e. wind speed and temperature).




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When oil is spilled on the water surface, it will incorporate water and form oil-in-water emulsion. This process changes the physical properties of the oil by increasing its viscosity and volume, and will be accelerated by mixing energy from waves. As with all weathering processes, the type of oil impacts the extent of weathering. Distillate fuel oils will likely emulsify at slower rates than residual fuel oils and give a wider window of opportunity for oil spill response options, since in situ burning and chemical dispersants are less effective on emulsified oils. Photo-oxidation by natural sunlight is an additional process that contributes to the degradation and transformation of spilled oil (NRC, 2005). In Norwegian waters, light intensity near the water surface will be lower than waters located near the Equator due to the low angle of incidence experienced at northern latitudes (NRC, 2014). Although, due to the refraction of sunlight on snow and the long duration of sun exposure during the summer months, photo-oxidation of oil may be significant in Arctic environments, but the magnitude of the effect in marine environments remains unknown. Natural oil-in-water dispersion (mixing of small oil droplets into the water column) will take place if there is sufficient energy on the sea surface, i.e. if there are breaking waves present. The waves will break the oil slick into droplets of diameters, typically 1 µm - 1 mm, which are then mixed into the water mass. The largest oil droplets will resurface and form a thin oil film (typically <50 µm) behind the oil slick. This thin oil film will be rapidly dispersed again by breaking waves as smaller droplets into the water column. The natural dispersion rate depends highly on the oil type and can be one of the main processes that determine the lifetime of an oil slick on the sea surface. Natural o/w dispersion will gradually decrease since the evaporation of the lighter compounds will increase the viscosity of the remaining oil. Dispersion of oil, is the most important physical process for biodegradation, increasing the surface/volume ratios, and thereby increasing the bioavailability. The susceptibility of the fuel oils to disperse is therefore quite important. Seawater contains an abundance of micro-organisms that can break down all types of oil components. The various micro-organisms prefer specific oil components as their energy source. Bacteria can only degrade oil in contact with water and depend on the water/oil interface area. Moldestad and Daling (2006) predicted the fate and behaviour of five different fuel oils: Gasoil, Marine Diesel, Wide Range Gasoil, IF-30 and IFO-360. In separate graphs, Figure 2.4 illustrates the various fuel oils property to naturally disperse, evaporate and emulsify. In addition, predicted lifetime and the increase in viscosity over time is illustrated.




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Figure 2.4 Predicted lifetime, natural dispersion, evaporative loss, emulsion viscosity and water content of an oil slick at 0°C sea temperature and 10 m/s wind.

The figures show a quite short lifetime on the sea surface for gas oil and marine diesel (only a few hours), compared to the Wide Range Gas oil (WRG) having a sulphur content of 0.25 %, IF 30 and IFO 380 which are predicted to persist much longer (days – to weeks). This is mainly due to high evaporation rates and effective dispersion processes of the gas oil and diesel. WRG was also predicted to leave a residue with a high wax content. The marine gas oil and marine diesel showed no ability to emulsify. Biodegradation is expected to be low if the oil emulsifies. In contrast, dispersion of oil will increase the biodegradation process. The fuel oils ability to disperse is therefore important when predicting the oils ability to be biodegraded. Few oils are dense enough to sink in seawater, but sinking of oils may be caused by Oil–Mineral Aggregation (OMA) or by incorporation of oil in aggregates of zooplankton feces or mucoid particles of phytoplankton and bacteria (Lee et al., 1996; Passow, 2016). This will be further discussed in Chapter 3.2.1.

2.2.4 Weathering properties of new diesel and fuel oil products used in the Arctic In 2015, the Norwegian Coastal Administration initiated a project studying the physical, chemical, and emulsifying properties of five different diesel products (Sørheim and Daling, 2015). The chosen fuels are currently utilized or transported in and around the Arctic. The diesel products showed quite different degrees of evaporation and emulsifying properties, but all showed low pour points. Solidification at sea is therefore not an issue for these diesels, even in the Arctic. Marine gas oil (MGO) showed the greatest evaporation degree and lowest density. Wide range diesel oil (WRD) showed to be the "heaviest" diesel product, containing no components with a boiling point below 250 °C. The fresh sample and the evaporated residue was therefore assumed to have the same physical and chemical properties. WRD was also the only diesel quality having emulsifying properties, in which it rapidly consumed water (89-90 %


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volume). Although, the emulsion was unstable and broke when energy stopped. The study was performed at 2°C. In 2017, the Norwegian Coastal Administration extended the initial study. Weathering properties of six different marine fuel oils were tested to determine the efficiency of different oil spill response methods (Hellstrøm, 2017). The tested marine fuel oils included: Gas Oil (GO), MGO, Rotterdam diesel, WRG (0.05 % sulphur content – from 2017 WRG must have a sulphur content of < 0.05 % to be allowed used as a marine diesel), Heavy Distillate Marine ECA 50 (HDME 50), and Shell ULSFO. The results are shown in Figure 2.4. Response operations such as application of oil spill dispersants, mechanical recovery and in-situ burning (ISB) were evaluated. Three of the fuel oils were classified as DMA-quality according to ISO 8217:2017: GO, MGO, and Rotterdam diesel. DMA-classified fuels are allowed in the nature reserves and national parks around Svalbard, while WRG is currently not allowed. Although, all six fuel oils are commonly used in Arctic regions and along the Norwegian coast. ExxonMobil produce HDME 50 and Shell produce ULSFO, which are both hybrid fuel oils. The testing was performed at temperatures that represented Arctic winter conditions (2 °C) and summer conditions in the North Sea (10-13 °C). A screening of six different oil spill dispersants was performed for all the fuel oils, using low energy conditions. In addition, the natural dispersion (no dispersant) of the DMA-qualities and WRG were investigated with the same methodology. The results showed little variation in the physical properties of GO, MGO, and Rotterdam diesel. None of the DMA oils had emulsifying properties. They do have similar pour points and densities, which suggest that they would have similar weathering properties, and thus dispersibility and ignitability. The WRG showed significant differences compared to the DMA-qualities, both in physical and chemical properties. The WRG oil formed unstable emulsions with low viscosities that were not easily dispersible at low temperatures. Results from the hybrid fuel oils testing showed that both HDME 50 and ULSFO solidified at low temperatures, and that high pour points strongly influenced the oil behaviour at 2°C. The amount of oil (%) predicted to remain on the sea surface was calculated for the six fuel oils at different wind conditions, see Figure 2.5. Due to evaporative loss and natural dispersion, the oil on the sea surface will gradually be reduced. The high evaporative loss a natural dispersion of the DMA-qualities is shown as a rapid removal from the sea surface at all wind conditions, compared to the heavier WRG and hybrid fuel oils.




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Figure 2.5 Predicted volume of remaining surface oil (%) for the six tested fuel oils.




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Different response operations can be relevant at different temperatures. In cold temperatures (2°C), emulsification may occur at a slow rate, and solidification and high emulsion viscosities will limit the effectiveness of dispersants. Repeated application of dispersant showed to be more effective, but solidified lumps were not affected. In warmer temperatures (13 °C+), emulsification began immediately and chemical dispersion had low effectiveness on emulsions, but repeated dispersant application increased the effect. Since weathering properties of the new hybrid fuel oils are mainly unknown, it is uncertain how they will behave in the natural environment. Individual weathering studies could provide insight into their potential to evaporate, emulsify, and photo-oxidize. These results can then assist in determining efficient spill response strategies for hybrid fuels.




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3 Biodegradation of oils in seawater The recently performed Arctic Response Technology program (2012-2017), managed by the International Association of Oil and Gas Producers (IOGP) carried out a literature review summarizing available knowledge on biodegradation of oil in the Arctic, and was finalised in 2013 (Word et al., 2013). Some of this chapter will be retrieved from that review and supplemented by resent research, mainly performed by SINTEF. Several studies have been performed to study biodegradation of crude oils over the years, while knowledge on biodegradation of fuel oils, especially knowledge covering the new generation fuel oils, is still lacking. The following literature review will therefore also include literature on crude oils and discussed on a general basis. A report by Rasmussen et al., covering the fate of heavy fuel oils spilled in marine waters will be released by the end of this year (2017). This report also summarizes biodegradation studies performed on crude and fuel oils in seawater, in sediments and in sea ice at different temperatures relevant for the Arctic.

3.1 Microbiology of the Arctic oceans The Arctic and Antarctic marine environments are characterized by seasonal extremes of photoperiod, spatial variability in salinity and temperature, as well as generally colder temperatures compared to the temperate latitudes. The temperature of the Arctic Ocean normally does not exceed 4-5°C, with cold (close to 0°C) low-salinity surface water in the upper part of the sea column, down to 50 m depth, and warmer high-salinity water (up to 5°C) below 50 m depth. The marine organisms in the Arctic are therefore consistently exposed to low temperatures (Word et al., 2013). In contrast, due to the Atlantic current, sea temperatures along the Norwegian coast fluctuate between 5°C during wintertime and 13-15°C during summer time. These differences may result in different expectations about the rate of oil degradation. They also result in different expectations about the indigenous populations of oil degrading microorganisms (Word et al., 2013). The population structures of bacteria in Arctic seawater are comparable to those in seawater from temperate regions, with the predominance of Alphaproteobacteria, Bacteroidetes (mostly Flavobacteria), Gammaproteobacteria and Verrucomicrobia, constituting more than 90% of the communities (Comeau et al., 2011; Teske et al., 2011). Studies of psychrophilic and psychrotolerant marine bacteria have indicated that hydrocarbon degradation is mainly associated with Gammaproteobacteria (Bowman and McCuaig, 2003; Yakimov et al., 2004; Deppe et al., 2005; Gerdes et al., 2005; Brakstad and Bonaunet, 2006; Brakstad et al., 2008; Bagi et al., 2014; McFarlin et al., 2014; Lofthus et al., 2015; Garneau et al., 2016). Not all 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 (Word et al., 2013). Figure 3.1 gives an overview of the taxonomy of Arctic and Antarctic oil-degrading bacteria.

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Table 3.1 Taxonomy of Arctic and Antarctic oil degrading bacteria's (Brakstad et al., 2017). Class Family Genus Source

Alphaproteobacteria Rhodobacteraceae Loktanella Arctic, Seawater

Sulfitobacter Arctic, Seawater

Sphingomonadaceae Sphingopyxis Arctic, Seawater Sphingomonas Antartic, seawater

Gammaproteobacteria Alteromonadaceae Alteromonas Seawater Glaciecola Arctic, Seaice Marionobacter Arctic, Antarctic, Seawater,

Seaice Colwelliaceae Colwellia Arctic, Antarctic, Seawater,

Seaice, Sediment Thalassomonas Seawater

Moritellaceae Moritella Arctic, Seawater, Seaice, Sediment

Pseudoalteromonadaceae Algicola Arctic, Seaice Pseudoalteromonas Arctic, Antarctic, Seawater,

Seaice, Sediment Psychromonadaceae Psychromonas Arctic, Seawater Shewanellaceae Shewanella Arctic, Antarctic, Seawater,

Seaice, Sediment Alcanivoracaceae Alconivorax Arctic, Seawater, Sediment Oceanospirillaceae Marinomonas Arctic, Antarctic, Seawater,

Seaice, Sediment Oleispira Arctic, Antarctic, Seawater,

Seaice Halomonadaceae Halomonas Arctic, Antarctic, Seawater,

Seaice, Sediment Moraxellaceae Psychrobacter Arctic, Seawater Pseudomonadaceae Pseudomonas Arctic, Antarctic, Seawater,

Seaice, Sediment Piscirickettsiaceae Cycloclasticus Arctic, Seawater, Sediment

Epsilonproteobacteria Campylobacteraceae Arcobacter Arctic, Antarctic, Seawater Bacteroidetes Cytophagales Cytophagia Antarctic, Seawater Flavobacteriia Flavobacteriaceae Ulvibacter Arctic, Seawater

Polaribacter Arctic, Seawater, Seaice Actinobacteria Nocardiaceae Rhodococcus Antarctic, Seawater

Microbacteriaceae Agreia Arctic, Seaice, Seawater Arthrobacter Antarctic, Seawater




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3.2 Biodegradation of oil in seawater Natural oil seeps have been discovered throughout the World’s oceans, including in the Arctic (NRC, 2005), and these seeps continue to enrich oil-degrading microorganisms (Prince and Clark, 2004). Oil-degrading microorganisms are diverse and widespread, including in the Arctic (McFarlin et al., 2014; 2017) and other cold environments (Deppe et al., 2005; Gerdes et al., 2005; Siron et al., 1995). Oil compounds can be degraded by a variety of prokaryotic (i.e. bacteria) and eukaryotic (i.e. fungi and algae) organisms (Prince et al., 2010). Despite the range of organisms that can act on oil, bacteria are responsible for the majority of oil biodegradation in seawater (Prince et al., 2010). The extent to which crude oil biodegrades in Arctic seawater has been reported (McFarlin et al., 2014) and laboratory experiments have shown the biodegradation of individual hydrocarbon compounds in Arctic seawater at near-freezing temperatures (Brakstad and Bonaunet 2006). In addition, bacterial taxa and functional genes known to biodegrade oil have been found in situ throughout the water column in an offshore Arctic oil release area (McFarlin et al., 2017). Biodegradation rates are thought to be slower in Arctic than temperate regions (Margesin et al., 2003; Michaud et al., 2004) due to the influence of temperature on metabolic processes. Although, some microorganisms are adapted to low temperatures (Feller, 2003), which may explain why some have reported similar oil biodegradation rates in cold and temperate environments (Braddock and McCarthy, 1996; Gibb et al., 2001; Margesin and Schinner, 1997). In general, biodegradation of oil compounds is expected to follow the order straight-chain n-alkanes > branched isoalkanes > cyclic alkanes > cyclic naphthenes > aromatics> resins > asphaltenes (Perry, 1984). In cold seawater, the same order is expected, although degradation will be highly influenced by the physico-chemical characteristics of the oil and weathering degree. Other important factors influencing the biodegradation rate are the nutritive supply that contain nitrogen and phosphorus, and the oxygen supply. The low temperature affects both dissolution from the non-aqueous (oil) to the aqueous phase (Schluep et al. 2001), and evaporation of volatile compounds (Word et al., 2013). For fuel oils containing a high amount of wax, resulting in high pour points; evaporation, dilution, and dispersion may be reduced due to solidification. The precipitated wax may build a matrix which limits the internal mixing of the oil and act as a diffusion barrier between the oil and the water. 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 3.1, showing the mineralization 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 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.




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Figure 3.1 Mineralization of 14C-labelled hydrocarbons spiked in crude oil in seawater at 0°C. No mineralization was measured in sterile controls (Brakstad and Bonaunet, 2006).

At sea, the formation of oil droplets by natural or chemical enhanced dispersion will increase the biodegradation rate in the water mass by 10 to > 100 times compared to surface oil due to increased water/oil interfacial area, and it has been shown that n-alkanes are biodegraded within 2-4 weeks at North Sea conditions (Brakstad and Lødeng, 2005). PAHs dissolved in water can be degraded within a few days (Brakstad and Faksness, 2000). Other higher molecular-weight oil compounds are biodegraded more slowly and some very high molecular-weight compounds (equivalent to the heavy residues in crude oil that are used to make bitumen) may not biodegrade to any significant degree. Previous and recent studies have shown how the microbial community dynamics changes during biodegradation of oil in cold seawater and ice. A few hydrocarbonoclastic bacteria initially dominate in the degradation, metabolizing bioavailable n-alkanes and volatile compounds, including Alteromonadales and Oceanospirillales (e.g. Oleispira) (Yakimov et al. 2003; Head et al. 2006; Hazen et al. 2010; Dubinsky et al. 2013; Lofthus et al. 2015; Ribicic et al. 2015). The alkane degraders are succeeded by more diverse consortia of bacteria degrading PAHs and more complex hydrocarbons, including members of the families Piscirickettsiaceae (e.g., Cycloclasticus), Alteromonadaceae (e.g., Marinobacter), Pseudoalteromonadaceae, and Shewanellaceae (Dubinsky et al. 2013; Brakstad et al. 2015; Ribicic et al. 2015). In addition, some bacteria have complete pathways for both aliphatic and aromatic hydrocarbon degradation, like members of the genus Colwellia, known to be abundant in cold environments (Brakstad et al. 2004; Methe´ et al. 2005; Dubinsky et al. 2013; Mason et al. 2014). As a part of a risk assessment performed on fuel oils in Antarctic and sub-Antarctic waters, a study describing partitioning of hydrocarbons from three fuels (Special Antarctic Blend diesel (SAB), Marine Gas Oil (MGO), and Intermediate Fuel Oil 180 (IFO 180) into seawater at 0 and 5 °C was performed (Brown et al., 2016). The subsequent depletion of the fuel oils were analysed over 7 days. The results showed that the initial total hydrocarbon content (THC) of water accommodated fraction (WAF) in seawater was highest for SAB, see Figure 3.2. Rates of THC loss and proportions in equivalent carbon number fractions differed between fuels and over time. THC was most persistent in IFO 180 WAFs and most rapidly depleted in MGO WAF. Depletion for SAB WAF was strongly affected by temperature.




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Figure 3.2 Relative loss of total hydrocarbon content (THC) from 100% water accommodated fractions of Special Antarctic Blend diesel (SAB), Marine Gas Oil (MGO) and Intermediate Fuel Oil 180 (IFO 180). The natural petroleum hydrocarbon degrading capacity of the Archipelago Seawater in S-W Finland was studied in a microcosm experiment (Reunamo et al., 2013). This study showed that bacteria with capacity for petroleum hydrocarbon degradation exist in the water of brackish Baltic Sea, in the archipelago of SW Finland. The diesel microcosms were dominated by Actinobacteria, followed by Alphaproteobacteria and Cyanobacteria. These results differ from studies of oceanic bacteria, where the main oil degrading groups often belong to the Gammaproteobacteria. The quantity of C10-C40 hydrocarbons in each microcosm was measured on the first day after 1 h and at the end of the experiment. At the end of the experiment, the amount of hydrocarbons had diminished by 7–40% compared to the first measurements, see Figure 3.3.

Figure 3.3 Diesel removal in diesel exposed microcosms. Amount (%) of diesel in one hour of addition (day 1) and in the end of the experiment (day 22). A: pristine site, P: recreational harbour, R: previously exposed site. Error bars represent standard deviations. In 2017, dispersibility and biodegradability of chemically dispersed crude oils and emulsions with different physical-chemical properties were studied at seawater temperature of 13°C, relevant for North Sea and Norwegian Sea summer conditions (Brakstad et al., 2017).




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All oils were chemically dispersible, using well-known dispersants, showing median oil droplet sizes of 18-47 µm. Oil properties affected dispersibility slightly. The most viscous oil and the oil with incorporated water (emulsion) resulted in dispersions with the highest median droplet sizes. Despite the oil/emulsions differences in physico-chemical oil properties, biodegradation half-lives of oil compound groups of the different oils and the emulsion used in the study were comparable. Extractable Organic Carbon (EOC) was degraded by > 70 % in all dispersions (average 75.0 ± 7.5 %) at the end of the experiments, n-Alkanes were biotransformed by > 90 % in all dispersions after 14 days. PAHs were also biotransformed fast, being depleted by > 80 % after 14 days. The data from this study therefore supported the use of generic rather than oil-specific biodegradation data for predictions related to the fate of chemically dispersed oil and emulsions at otherwise similar environmental conditions. This will have implications for fate and exposure models like OSCAR, which require biodegradation rates as input data (Reed et al., 2001). Using empirical data in the model will strengthen the predictions of the fate of the oil after oil spill dispersant treatment. However, further studies are required, and are under way, to compare biodegradation kinetics at different temperatures and with different seawater sources. The Total Extractable Organic Carbon (TEOC) concentrations, quantification of C10-C36, and distribution between other relevant components is shown in Figure 3.4. As shown in the figure, the UCM fractions became completely dominant during the biodegradation period.




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Figure 3.4 TEOC concentrations and distribution between n-alkanes, PAH and UCM in the extracts during biodegradation of different crude oil types (fresh and emulsified).

3.2.1 Oil sedimentation processes Oil biodegradation in seawater may result in the generation of macroscopic aggregates. These aggregates, or “flocs,” have been observed during experimental biodegradation studies with dispersed oil (Macnaughton et al. 2003; Bælum et al. 2012), see Figure 3.5. In our experience, these “flocs” are fragile structures, and generation of turbulence like magnetic stirring rapidly disintegrates the structures. Aggregates of oil and bacteria were reported in the deepwater oil plume during the Deepwater Horizon (DWH) oil spill (Hazen et al. 2010), and it was suggested that these types of aggregates resulted in fallout of oil from the plume to the Gulf of Mexico seabed (Valentine et al. 2014). Analyses of “flocs” from field studies or laboratory experiments showed distribution of microorganisms, diverse polysaccharides and extracellular polymeric substances (EPS), oil, and oil degradation products in the typical “floc” structure




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(Hazen et al. 2010; Bælum et al. 2012). Time-related analyses during biodegradation studies suggested that the oil tended to initially concentrate in the “floc” material, but was subsequently degraded, while the biological material in the “flocs” increased (Bælum et al. 2012), and the “floc” is therefore considered to be an active site of oil biodegradation.

Figure 3.5 A “floc” generated during biodegradation of a paraffinic oil at 1–2 °C in Arctic seawater (Svalbard) (scale 0.5 cm). Photo: Emlyn Davies, SINTEF. Sinking of oils may also be caused by Oil–Mineral Aggregation (OMA), or by the formation of Oil-Related Marine Snow (ORMS). OMA is primarily a near-shore process, typically occurring in surf zones, near river outflows, melting glaciers, and sea ice (Daly et al. 2016), while ORMS is formed by incorporation of oil in aggregates of zooplankton feces or mucoid particles of phytoplankton (Lee et al., 1996; Passow, 2016). ORMS typically contains high concentrations of bacteria capable of hydrocarbon metabolization and thus, biodegradation of oil compounds in ORMS may be an important part of total oil biodegradation after an oil spill. ORMS is close connected to OMA, but the definition of marine snow is that the aggregate is greater than 0.5 mm. This means that OMA can form to ORMS when bacteria, algae etc. attach to the OMA, forming a aggregate greater than 0.5 mm. OMA and ORMS processes may cause the oil to achieve higher density than the seawater, sink through the water column, and end up in the seabed sediments, with subsequent impacts on the benthic environment. These processes can therefor play a key role in the vertical transport of petroleum hydrocarbons after a marine oil spill. Oil residues sinking to the seabed will be subject to aerobic or anaerobic biodegradation in the sediments. Biological processes like bioturbation are important for venting and oxygenation of the sediments. Important bioturbators in marine sediments include typically polychaetes, bivalves, burrowing shrimp, and amphipods. Bioturbation contributes to increased availability of oxygen and organic matter in sediments, which in turn, results in increased activity of aerobic biodegradation processes, including oil biodegradation. Since oil biodegradation is supposed as the major process for mineralization of oil, bioturbation is also highly relevant for oil transported to the seafloor in form of ORMS. The research on ORMS processes and the fate of these aggregates is still in its infancy, and relevant studies have not been conducted under conditions relevant for the Norwegian and Arctic regions. To date, the majority of reported ORMS studies are related to temperate environments associated with the DwH oil spill (Valentine et al., 2016). However, due to complex bio-physico-chemical processes contributing to ORMS formation, statistically reliable and reproducible data suitable for implementation in numeric models, are still missing. SINTEF has recently established experimental systems for studying the formation processes, characterization and biodegradation processes of ORMS in cold water environments. The project is funded by the industry and titled "Marine Snow Flocculation and Sedimentation in relation to Oil Spill Response".




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3.2.2 The deepwater Horizon oil spill and the fate of the oil The Deepwater Horizon (DWH) oil spill in the Gulf of Mexico was one of the largest spills in the history, with a subsea blowout lasting from April 10 to July 14, 2010. During this period, approximately 4.9 million barrels (780 million litres) of a light paraffinic oil was spilled from the Mississippi Canyon Block 252 (MC252), located 77 km offshore the coast of the Mississippi delta. The responders began injecting the chemical dispersant Corexit 9500A at the wellhead in early May 2010 to reduce the amount of oil that formed surface slicks and subsequently reached the Gulf of Mexico (GoM) coastlines (Atlas and Hazen, 2011, Kujawinski et al., 2011, Lehr et al., 2010). It was estimated that most of oil reaching the surface consisted of oil droplets of millimetres in diameter (Ryerson et al., 2012). The use of subsea dispersant injection resulted in a deepwater plume at a depth of 800-1100 m. Measurements of volatile aromatic hydrocarbons in water samples during the release showed rapid decrease with distance from the release point, being < 1 ppb within 15-20 miles (24-32 km) from the release source, and biodegradation was suggested to play an important role in the depletion of oil (Atlas and Hazen, 2011). During the spill, it was suggested that some oil from the deepwater plume reached the seafloor. Based on 17 α(H), 21β(H)-hopane (Valentine et al., 2014), a well-known oil biomarker persistent to biodegradation (Prince et al., 1994), sediment analyses from over 3000 sediment samples from 534 locations suggested that 4-31 % of the oil from the DWH incident ended up in the deep ocean (Valentine et al., 2014). However, these estimations are highly uncertain, and they did not take into account the dissolution and biodegradation of oil compounds during the sinking period. Later studies of sediment core samples collected in March-October 2011 demonstrated that the oil found on the seafloor (1.6 km from the Macondo wellhead) was highly weathered by dissolution loss of partly water-soluble oil compounds and severe biodegradation (Stout and Payne, 2016). Bacteria associated with hydrocarbon degradation, like Alcanivorax, Marinobacter and Colwellia, showed higher activity in the surficial than the deeper sediment (Yeargau et al., 2015). The highly weathered and biodegraded condition of the deposited oil residue was further emphasized in a study describing sediment analyses up to 4 years after the spill. After 160 days post start of the spill period, most short-chain n-alkanes and 2- to 3-ring PAH were completely degraded, while larger n-alkanes (> C28-C29) and PAH (> 17 carbons) still persisted. However, biodegradation of these larger hydrocarbons continued 4 years after the spill (Bagby et al., 2017). A conceptual model for oil sedimentation was suggested by Stout and Payne (2016), as shown in Figure 3.6, although this is speculative. This model proposes that most of the oil sedimentation from the DWH spill came from the deepwater plume in an area of more than 8 km from the ruptured well, while some contribution may also have come from oil-related marine snow formed in surface waters. Despite data from field sampling and several laboratory studies, the contribution of oil sedimentation from the DWH spill is still under discussion and many data gaps remain.




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Figure 3.6 Conceptual model for Macondo oil deposition on the deep-sea floor (Stout and Payne, 2016).

3.3 Bioremediation strategies at sea As an alternative to mechanical removal of oil from the sea surface, e.g., use of oil booms and skimmers, remediation by use of chemical dispersants, addition of nutrients (fertilizers stimulating indigenous microbes to enhance degradation) or the combination of these two, could be used with the aim to accelerate the biodegradation processes. Dispersants are used primarily to remove oil from the water surface in order to reduce the impacts on seabird and mammal populations close to an oil spill, but this treatment has also been suggested to improve hydrocarbon biodegradation in the seawater column. Dispersants are mixtures of surface-active agents reducing the surface tension of the oil, resulting in the formation of small oil droplets which will rapidly disperse in the water column. The surface/volume ratios will be increased, and thereby increasing the bioavailability of the dispersed oil. The ability of the dispersants to generate small oil droplets will be the most important factor for the biodegradation process. It is important to point out that efficient dispersant treatment of surface oil spills will bring more dispersed oil into the water column, and this may have effects on sensitive organisms, for instance vulnerable life stages of fish species. In low-temperature seawater, exposure periods may be prolonged due to slow biodegradation. Environmentally acceptable technologies for the improvement of biodegradation in seawater is therefore a relevant research area (Word et al., 2013). Efforts to stimulate crude oil biodegradation in seawater and ice have not been investigated to the same extent as for stranded oil. However, dispersants have been used in several oil spill operations such as the Deepwater Horizon blowout in 2010 to reduce oil surfacing (Atlas and Hazen, 2011; Kujawinski et al., 2011). The use of chemical dispersants has, based on these experiences, become an important operational tool for treatment of surface or subsurface oil discharges in the marine environment.




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In addition to the use of chemical dispersants, accelerated biodegradation by adding fertilizer formulations in order to improve the oil biodegradation capabilities of the natural microbial populations has been investigated. Fertilizers have showed to stimulate biodegradation of crude oils in cold seawater under controlled experimental conditions. The slow-release fertilizer Inipol EAP 22 was added to Antarctic seawater contaminated with crude oil in a mesocosm study, resulting in enhanced abundance of heterotrophic and oil-degrading bacteria, and increased rates of biodegradation during a 40-day experiment, both in ice-covered and ice-free seawater (Delille et al. 1998). Biostimulation is a “promising” technology for treatment of polluted marine environments, and is especially suited for remote locations, as the Arctic. This treatment is cheap, labor-effective, generates no harmful by-products, and is especially suited for secondary clean-up, in combination with other technologies. Novel tools for microbial and chemical analysis will also be important for determining restitution of oil-polluted environments during remediation actions (Brakstad et al., 2017). Bioaugementation Bioaugmentation involves the addition of bacteria to enhance biodegradation rates in natural environments. This technology has been proposed for soil and sediments, but can also be used on water, often as a supplement to biostimulation treatments. Several bioaugmentation studies from marine environments have been reported, although none of these are from cold waters. A number of commercial products exist, which include microbial inocula. These products are often lacking essential information about the bacterial content, making national authorities sceptical about using bioaugementation as a response action (Brakstad et al., 2017). Bioaugmentation has often proved inferior to biostimulation. One plausible explanation for this may be that the introduced bacteria will have an immediate effect due to the biomass added, but these exogenous microbes may gradually be outcompeted by the indigenous microbes adapted to the local environment (Word et al., 2013).

3.3.1 Dispersibility of bunker fuel oils Through the project "Statlig dispergeringsberedskap" (2014), the Norwegian Coastal Administration (NCA) evaluated and developed an implementation plan for a governmental oil spill dispersant strategy in Norway. Parts of the objective of this project were to determine which type / categories of bunker fuel oils are dispersible, and further recommend the most suitable dispersant to estimate the "time window" after release with use of dispersants (Sørheim et al., 2014). The laboratory study was performed by SINTEF and was an important contribution to NCA's total decision-making for the establishment of a dispersant contingency of bunker fuel oils. Testing was performed on four bunker fuel oils; IFO-80LS, IFO-180LS, IFO-180NS and IFO-380LS, at 5 and 13 °C, reflecting typical winter and summer conditions in the North Sea / Norwegian Sea. The bunker fuel oils were weathered in the laboratory to produce 150 °C +, 200 °C+ and 250°C + residues (corresponding to approximately to 0.5-1 hours, 0.5 -1 days and up to 5 days of weathering at sea). Testing of dispersibility of IFO-80 was also performed under Arctic conditions (0 °C degrees) in conjunction with NCA's contingency analysis for Svalbard and Jan Mayen. A screening study of a total of eight different dispersants was performed. It was essential that the recommended dispersant should meet requirements according to the Norwegian Pollution Regulations for use of dispersant, Chapter 19, i.e. fulfil requirements relating to toxicity and efficiency with the use of standard bioassays (algae test, Skeletonema). The results indicated that bunker fuel oils may be dispersible at higher viscosity limits than previously expected. Furthermore, the results indicated that oils with viscosities less than 10 000 cP could be dispersible with only one application (provided a dosage to emulsion ratio (DER) of 1:25), while fuel oils




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having viscosities less than 20 000 -25 000 cP, needed two applications with dispersant for effective dispersibility. A dispersibility test on marine fuel oils were performed in 2017 (Hellstrøm). As stated in Section 2.2.4, calm conditions (< 5 m/s wind speed), spreading of oil, and formation of thin oil film thicknesses; limit the time window for dispersant application. A minimum film thickness > 50 - 100 μm can be considered for application of oil spill dispersants. According to OWM predictions, DMA classified fuel oils may, in calm weather conditions, have potential for application of oil spill dispersant within the first 1-2 days (24-48 hr) after release. In breaking wave conditions, the lifetime of DMA-diesels on the sea surface is considered to be limited (< 1 day) due to natural dispersion and evaporative loss. Furthermore, application of oil spill dispersant to WRG, HDME 50 or Shell ULSFO should be performed while emulsion viscosities are low. The lowest viscosities are expected directly after the spill has occurred for WRG, regardless of temperature. For HDME 50 and Shell ULSFO, the lowest emulsion viscosities are expected directly after the release in summer temperatures, while in winter temperatures the initial viscosities may become high a short time after release due to the high pour point of these oils. However, these emulsion viscosities may decrease slightly with time. When taking response time for response operation vessels into consideration, application of dispersant should be performed as soon as possible, regardless of temperature. Repeated application of dispersant will be more effective than a single high-dose application for these fuel oils. Application of additional energy (e.g. use of high dosage water flushing) in low energy conditions will likely enhance the dispersion of the treated oil/emulsion.




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3.4 Summary and recommendations – biodegradation at sea In summary, studies have shown slower biodegradation by lowering of the temperature, but results also show that biodegradation at low seawater temperature is considerable. The physico-chemical characteristics and weathering conditions of marine fuel oils at temperatures typical for the Arctic region may vary considerably, having significant impacts on biodegradation efficiencies. Not many studies have been performed studying biodegradation on weathered marine fuel oils, but based on the literature available, the biodegradation is expected to be low if the oil emulsifies and will strongly be dependent on the oils ability to disperse, increasing with an effective dispersion process. Recommendations for future research:

• 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.

• The bioavailability oil droplets will depend on the physical properties of oil (e.g. larger droplets at

lower temperatures would decrease the surface area-to-volume ratio). Marine fuel oil characteristics should therefore be addressed in more detail. This should be done by studying biodegradation of dispersed fuel oil types in use today, at different weathering degrees, at several seawater temperatures.

• Efficient dispersant treatment of surface oil spills will bring more dispersed oil into the water

column, and this may have effects on sensitive organisms, for instance vulnerable life stages of fish species. In low-temperature seawater, exposure periods may be prolonged due to slow biodegradation. Environmentally acceptable technologies for the improvement of biodegradation in seawater should therefore be a relevant research area.

• UCM dominates the composition of oil. Today, compositional changes in this major fraction during

biodegradation is understudied when compared to the known fractions. This knowledge gap should be filled using state of the art methodologies now available. What is defined as UCM using the standard GC-FID and GC-MS methods used routinely today is not necessarily unresolved when modern techniques such as GCxGC (2D GC) and high-resolution mass spectrometry is used to analyzed the same samples. Future degradation studies of oil compounds in cold environments should therefore consider these environmentally persistent and toxic UCM-related compounds.

• Novel tools for microbial and chemical analysis will be important for determining restitution of oil-

polluted environments during remediation actions.




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4 Biodegradation of oil in sea ice Summer sea ice coverage in the Arctic has reached the lowest extent on record (Comiso and Nishio, 2008) and its decline continues to open new shipping routes and opportunities throughout the Northeast Passage. This increased activity poses an environmental risk through the potential release of fuel oil and crude oil from cargo ships, oil tankers, pipelines, and exploration. Prevailing seawater currents in the Norwegian Arctic (Figure 4.1 A) may facilitate the drifting of spilled oil into ice (Figure 4.1 B). Understanding the fate of oil in ice is crucial to understanding the impacts of a spill on the marine ecosystem.

Figure 4.1 A: Prevailing currents in the Arctic Ocean (source: Woods Hole Oceanographic Institution; https://www.whoi.edu/main/topic/arctic-ocean-circulation), and B: maximum monthly ice coverage in the period 1984-2013 (source: The Norwegian Polar Institute). The risk of oil spills reaching the marginal ice zone (MIZ) is particularly relevant in the winter when sunlight is absent and ice coverage is at a maximum. Environmental conditions and remoteness provides additional challenges associated with Arctic oil spill response in ice covered waters. Understanding the dynamic aspects of marine ice, in relation to its physical and biological characteristics, is important when identifying and describing oil spill response options and related data gaps in cold environments.

A

B




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4.1 Marine ice and associated microorganisms Marine ice formation, deformation and evolution/melting are complex processes, depending on weather conditions. During calm weather conditions, the ice will form and grow from a dark or light nilas (Figure 4.2 B; <5 to 10 cm thick) to young ice (0.1 to 0.3 m thick), and further to first-year ice and ultimately to old (second and multi-year) ice. Under wave-conditions, frazil ice (Figure 4.2 A) will develop into larger units (pancake ice) by ice consolidation (Figure 4.2 C). The evolved marine ice may further be subject to snow deposition, depression below sea level with flooding and snow-ice formation. Spring/Summer melting will lead to surface melting, and rotten ice floes, with eventual complete melting of the floes (Petrich et al., 2012).

Figure 4.2 Growth of Marine ice, represented by frazil ice (A), nilas (B), and pancake ice (C). Brine channels are formed when ice grows and the ice-water interface advances downward. Salt concentrations build up ahead of the advancing ice/water interface, increasing the salinity in thin layers (a few millimetres) of ice over time. In ordinary columnar sea ice, the lamellar interface consists of sub-millimetre blades of ice, separated by narrow films of brine (Petrich and Eicken, 2017). A microscopic visualization of these brine channels is shown in Figure 4.3 These brine channels are present as liquid environments in first-year ice and are essential for biological activity in marine ice.

Figure 4.3 Scanning electron microscopy (SEM) image of the brine channel properties of sea ice (source NOAA; http://oceanexplorer.noaa.gov/explorations/02arctic/background/sea_ice/brine_220.jpg). The high salinity in the brine channels will provide liquid niches that enable motility and respiration of microorganisms at sub-zero temperatures (Junge et al., 2003, 2004, 2006). The brine channels also act as a matrix for the transport of hydrocarbons (Fingas and Hollebone, 2003; Faksness and Brandvik, 2005). Under conditions with ice temperatures of -20°C and salinity conditions high as 200 ‰, laboratory experiments have shown that the bacterium Colwellia psychroerythraea is able to incorporate amino acids into proteins (Junge et al., 2006). In addition, the ability to grow at sub-zero temperatures has been demonstrated by the

A B C

http://oceanexplorer.noaa.gov/explorations/02arctic/background/sea_ice/brine_220.jpg

https://www.google.no/imgres?imgurl=https://upload.wikimedia.org/wikipedia/commons/thumb/c/c0/Frazil_ice_in_Yosemite_Creek.png/1200px-Frazil_ice_in_Yosemite_Creek.png&imgrefurl=https://en.wikipedia.org/wiki/Frazil_ice&docid=Y20hi4KVhnhHCM&tbnid=dp9-ieqdEI_zBM:&vet=10ahUKEwiwlbaOgMXWAhXjAJoKHX1gCNgQMwglKAAwAA..i&w=1200&h=675&bih=860&biw=1536&q=frazil%20ice&ved=0ahUKEwiwlbaOgMXWAhXjAJoKHX1gCNgQMwglKAAwAA&iact=mrc&uact=8

https://www.google.no/imgres?imgurl=https://upload.wikimedia.org/wikipedia/commons/thumb/0/05/Nilas_Sea_Ice.JPG/1200px-Nilas_Sea_Ice.JPG&imgrefurl=https://fr.wikipedia.org/wiki/Nilas&docid=IQrAV-bwOyA9jM&tbnid=zBSYoM0IYQ_5sM:&vet=10ahUKEwi325_igMXWAhVkLZoKHfyEB7QQMwglKAAwAA..i&w=1200&h=800&bih=860&biw=1536&q=nilas&ved=0ahUKEwi325_igMXWAhVkLZoKHfyEB7QQMwglKAAwAA&iact=mrc&uact=8

https://www.google.no/imgres?imgurl=http://blogs.egu.eu/geolog/files/2013/11/Pancake-ice.jpg&imgrefurl=http://blogs.egu.eu/geolog/2013/12/30/imaggeo-on-mondays-a-feast-of-pancakes/&docid=w6bljA3t4-bJIM&tbnid=otZwFjxX39owTM:&vet=10ahUKEwja-uaGgcXWAhVJIpoKHV-5D30QMwgnKAIwAg..i&w=570&h=379&bih=860&biw=1536&q=pancake%20ice&ved=0ahUKEwja-uaGgcXWAhVJIpoKHV-5D30QMwgnKAIwAg&iact=mrc&uact=8

http://www.google.no/url?sa=i&rct=j&q=&esrc=s&source=images&cd=&cad=rja&uact=8&ved=0ahUKEwjt-ceFlsXWAhWrFZoKHQnECvwQjRwIBw&url=http://oceanexplorer.noaa.gov/explorations/02arctic/background/sea_ice/sea_ice.html&psig=AFQjCNHjORPRUEiuSSXfjVV8ZW6NI8fkaw&ust=1506594447853670




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bacterium Psychromonas ingrahamii, which has been cultured at temperatures as low as -12°C (Breezee et al., 2004). Although data from several studies showed that most bacterial and archaeal phylotypes were common between Arctic and Antarctic seawater and sea ice (Bano et al. 2004; Brinkmeyer et al. 2003), recent investigations have suggested that sea ice assemblages may differ from the assemblages in the seawater underneath the ice. Studies have shown that sea ice microbial communities have a higher dominance of certain Flavobacteriia compared to seawater, where Alphaproteobacteria are abundant (Boetius et al. 2015; Hatam et al. 2014). The implications of these differences for oil biodegradation in the ice are not known. Bacterial communities have been characterized in both first-year and multi-year ice, and first-year ice and multi-year ice are dominated by the same bacterial classes; however, first-year ice has a higher diversity than multi-year ice (Hatam et al. 2016).

4.2 Oil in the ice – different scenarios Oil spills may reach the ice either by directly spilling onto the ice or by drifting into the ice. Direct spills are associated with ships colliding into ice bergs are a rare occurrence, but should nonetheless be considered as a likely scenario. A more plausible spill situation would be oil drifting into the ice from surface water, either as crude or fuel oil. For crude oils, spilled oil from a leaking pipe or wellhead will rise-up through the water column and contact the ice at the ice-water interface. Fresh oil spilled in open water and drifting on the surface would become weathered as it travels and would likely reach the ice as a water-in-oil (w/o) emulsion. When oil reaches first year or multi-year ice, it may become frozen and thus entrapped into the ice and follow the ice as it moves, making it difficult to track during the winter. When oil accumulates in ice, the oil will tend to move through the ice using its naturally formed brine channels (Faksness, 2008a). Of course, the type of ice has an impact on oil movement, but research has suggested that the age of multi-year ice may also be important. Karlsson et al. (2011) reported that oil migrated slower in ice collected in May than in ice collected from other months. Some of the oil may escape the ice in the spring as the ice deteriorates by two general processes: a) vertical rise of the oil through the brine channels in the ice, and b) ablation of the ice surface down to the oil lens within the ice (Fingas and Hollebone 2003). Some of the processes associated with oil in the ice are shown in Figure 4.4. This may result in a secondary discharge of the spilled oil to a new and often unexpected location. In the spring and summer season, chemical photooxidation of the oil, either on the ice surface, or after secondary discharge, may become an important degradation process (Faksness, 2008b).




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Figure 4.4 Weathering of oil spilled in ice-infested environment (AMAP, 1998). Behaviour of oil encapsulated in ice is dependent upon the type of oil and its degree of weathering or degradation. At low temperatures, oil with high contents of wax and asphaltenes have temperature-dependent oil viscosities, and will likely solidify. However, oils with low pour points may maintain some fluidity in the brine channels, depending on the temperature. While the bulk of the oil trapped in the ice may be considered to be inert, water-soluble compounds may be released into the brine channels and transported to the underlying water. These compounds may then impact viable organisms in the brine channels and in the underlying seawater. These organisms may include typical ice edge phytoplankton, and proto- and metazoans (Grandinger et al., 2005; Marquardt et al., 2011), but also bacteria with potential to degrade oil compounds (Garneau et al. 2016). In addition, a large oil spill in ice may decrease the abundance of ice-algae and meiofauna, which would impact important aspects of the Arctic food web, such as sea ice-pelagic coupling (Kramer, 2010).

4.3 Oil-degrading microorganisms in sea ice The life in the brine channels of marine ice represents harsh conditions of low temperature and high salinity. Microbes with the ability to tolerate such conditions may be considered to be psychrophilic (adapted to cold temperatures) and also halophilic (adapted to high salt conditions). Sea ice naturally has a high dominance of these psychrophilic and/or halophilic bacteria, such as Flavobacteriia (Boetius et al. 2015; Hatam et al. 2014). Oil-degrading bacteria as well as oil-degradation genes are thought to be ubiquitous and have been found in a variety of pristine environments with no known exposure to petroleum compounds (Head et al., 2006; Kostka et al., 2011; Yakimov et al., 2007). Hydrocarbon-degrading bacterial taxa have been isolated from Arctic and Antarctic sea ice. For example, Yakimov et al. (2003) isolated Oleispira antarctica from Antarctic sea ice and showed its ability to biodegrade oil. In contrast to the diverse microbial community that is known to dominate marine ice, studies of psychrophilic and psychrotolerant marine bacteria have indicated that hydrocarbon degradation is mainly associated with Gammaproteobacteria, as shown in Table 3.1 (Brakstad et al. 2008; Deppe et al. 2005; Garneau et al. 2016; Gerdes et al. 2005). Other genera with identified hydrocarbon-degraders in Arctic sea ice include members of Flavobacteriia, Polaribacter, Pseudoalteromonas, and Oleispira (Brakstad et al., 2008; Garneau et al. 2016; Gerdes et al. 2005).




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4.4 Degradation of oil in sea ice Very few studies have been performed to investigate hydrocarbon biodegradation in sea ice. Some studies from oil-polluted Arctic soil indicated biodegradation at sub-zero temperatures (Børresen and Rike 2007; Rike et al. 2003), while other studies have shown that oil in ice stimulates bacterial growth (Brakstad et al. 2008; Delille et al. 1997). In a study performed in marine ice at Svalbard, oil was applied to sea ice in the Van Mijen fjord, and biodegradation of hydrocarbons followed over a period of 4 months. No n-alkane depletion was recorded, but the results indicated slow biotransformation of soluble naphthalenes dissolving into the ice brine channels (Brakstad et al., 2008). Garneau et al. (2016) conducted mesocosm experiments with melted sea ice and found that indigenous microbial communities were able to biodegrade 48% of the initial hydrocarbons within 15 days at -1.7C. In addition, Gerdes et al. (2005) performed biodegradation experiments with specific bacterial strains isolated from Arctic sea ice and reported a 20-50% loss of [14C] hexadecane. These experiments indicate the potential for hydrocarbon biodegradation in sea ice, but exemplify the need and the challenge to perform these experiments with intact sea ice. An Arctic Oil Spill Response Technology JIP titled 'Environmental Effects of Arctic Oil Spills and Arctic Oil Spill Response Technologies', recently conducted in situ experiments and reported the fate of oil in the upper sea-ice layer and its effects on microbial communities (Boccadoro, 2016). The effects of oil spill treatments (dispersants and in situ burning) on microbial communities and on petroleum biodegradation was measured. Experiments were conducted 0.8-2.5 km from shore in Svalbard, Norway. Crude oil, oil + dispersant, and burnt oil residue was placed on the surface of established ice within the in situ mesocosms. Ice cores were collected and analyses confirmed the transfer of water soluble oil compounds through the ice and the abundance and activity of oil-degrading bacteria in all layers of the sea ice and in the seawater. The presence of dispersant was shown to increase the dissolution and migration kinetics of the lightest PAHs, as well as the abundance of oil-degrading microorganisms (e.g. Oleispira and Colwellia). In 2016 a 3-year project, financed by The Norwegian Research Council and the oil industry was initiated, investigating the fate and behaviour of weathered oil drifting into an ice edge or scattered ice conditions (FateIce). Results from this study will serve as a foundation for establishing robust oil spill response technologies, strategies and operations for such spill scenarios. One of the objectives is to obtain better biodegradation data and improve our understanding of microbial processes with respect to natural attenuation and response strategies for different oils interacting with the ice edge. Studies will be conducted with natural seawater as microbial sources in mixtures of seawater and ice (slush, and ice floes). This will be used to update model framework to better describe biodegradation of different oils in the MIZ. To provide new knowledge to oil spill responders and the response industry, SINTEF Ocean is currently performing experiments to increase the understanding of the fate and behaviour of different oils when drifting into the ice edge or in scattered ice conditions. Mesocosm experiments with either frazil ice or pack ice are currently being performed with dispersed oil. Frazil ice is a mixture of seawater and ice (similar to slush) and it is formed as a precursor to first-year sea ice. Biodegradation of weathered-dispersed oil in the presence of ice will be reported as chemical loss and the change in microbial community structure over time. Additionally, this project will evaluate the effectiveness of dispersants and in situ burning at the ice edge as alternatives or supplements to mechanical recovery. Together these data will provide important updates to models that describe transport, weathering, entrainment, dispersibility, ignitability and biodegradation of different oils in the marginal ice zone.




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4.5 Bioremediation strategies in ice Some attempts have been made to increase in situ oil bioremediation rates in sea ice. If biodegradation of crude oil could be stimulated in ice, especially for the most toxic compounds migrating out of the ice through the brine channels, this would benefit organisms inhabiting and surrounding the polluted ice. Studies have shown that fertilizers can stimulate biodegradation of crude oils in cold seawater under controlled experimental conditions (Delille et al. 1998). The slow-release oleophilic fertilizer Inipol EAP 22 was added to Antarctic seawater contaminated with “Arabian light” crude oil in a mesocosm study. The experiment was completed over 5 weeks during the Austral summers of 1992/1993 and 1993/1994. In both ice-covered and ice-free seawater, the addition of the fertilizer enhanced both the concentrations of heterotrophic and hydrocarbon-degrading bacteria and increased the rate of biodegradation during the experiments, measured as n-C17/Pristane and n-C18/Phytane ratios. A winter field experiment was conducted at Svalbard in 2004 as part of the ARCOP programme. Crude Statfjord oil with and without fertilizers (mixture of Inipol EAP 22 and fish meal) was placed in fjord ice (Van Mijen Fjord, Svea) for a period of 6 months (December 2004-June 2005). At sub-zero temperatures, no significant degradation of oil occurred with the addition of nutrients, but melt-pool samples (0°C) fertilized with inorganic nutrients showed a significant change in bacterial diversity (Gerdes and Dieckmann, 2006). Importantly, many of the available slow-release fertilizers are not suitable for use in Polar Regions as they will solidify if used in ice at very low temperatures. For example, the pour point of the Inipol EAP 22 is 11°C which makes it difficult to use effectively in Arctic conditions. Therefore, if bioaugmentation is desired, slow-release fertilizers will require reformulation or new products will need to be developed specifically for use at very low temperatures.

4.6 Summary and recommendations – biodegradation of oil in ice Many data gaps exist concerning oil biodegradation in natural sea ice. We are beginning to understand the biodegradation of naphthenic oil in frazil ice using laboratory mesocosms and experiments have been conducted in situ that mimic an oil spill on the surface of solid ice; however, more experiments with intact ice are necessary to understand the fate of oil flowing into and through different types of ice (ex: first-year vs. multi-year sea ice). Future oil in ice experiments should focus on weathered oils with different properties (e.g. paraffinic, naphthenic and asphaltenic crude oil and fuel oils), oil in the presence and absence of dispersants, and the potential for biostimulation. To determine if biostimulation with fertilizer is feasible, products suitable for use at freezing temperatures should be researched. In addition, experiments designed to understand the chemical fate of oil in ice should utilize intact sea ice when logistically possible, instead of melted sea ice. Regarding the effects of oil on ice biota, future research is necessary to understand how oil compounds and their metabolites affect the abundance of primary and secondary producers in and around sea ice. Photooxidation of different oils during spring ice melt may increase oil degradation, but further research is necessary to quantify this pathway and determine the potential to produce toxic products.




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5 Biodegradation of oil in sediments Accidental oil spills in the marine environment is not only directly affect the water column in close proximity, but also sediments often remotely from the origin of the spill. Oil spilled offshore can be transported to shorelines and pollute tidal sediment zones. Oil fractions not stranded on-shore will undergo bio-physico-chemical reactions, such as weathering and biodegradation, and finally be transported to and interact with seafloor sediments. The Exxon Valdez (1989) and the DWH (2010) oil spills have been the worst in U. S. history. Even though completely different in nature, they are prominent examples for the fate of spilled oil in different sediments (Atlas and Hazen 2011).

5.1 Shoreline sediments Biodegradation of oil compounds in tidal shoreline sediments has been studied in detail after the Exxon Valdez oil spill. The spill was caused by the Exxon Valdez oil tanker that released an estimated 42 million liters of heavy crude oil near shore after running aground in Prince William Sound, Alaska. Approximately 778 km of shoreline in Prince William Sound and 1309 km of shoreline in the Gulf of Alaska were polluted by the resulting oil slick (Owens and Teal, 1990). Results for sediment samples, collected and analyzed in 1989, indicated that about 25-30% of the total hydrocarbon (HC) in the oil originally stranded on Prince William Sound shorelines had been lost within the first days to weeks after the spill. Indigenous biodegradation rates of oil were estimated to be 1.3 g oil/kg sediment/yr for surface oil and 0.8 g oil/kg sediment/yr for subsurface oil (Bragg et al. 1994). Concentrations of naturally occurring oil-degrading bacteria during this period were 1-5 x 103cells/mL of seawater or about 1-10% of the total heterotrophic bacterial population. In late 1989, oil-degrading bacterial abundances had increased to about 1 x 105 cells/mL and represented about 40% of the heterotrophic bacteria in oiled shoreline pore waters. Several years after the oil spill, approximately 60 - 100 tons of subsurface oil were still thought to be present in polluted beaches (Short et al., 2004, Short et al., 2006). Similar findings were also reported by Taylor and Reimer (2008) and Li and Boufadel (2011). Studies performed 12 years after the spill found that the persistent oil had undergone substantial weathering, but high concentrations of potentially toxic and mutagenic contaminants remained (i.e. chrysenes and other polycyclic aromatic hydrocarbons, PAHs), suggesting that the oil continued to pose an substantial environmental threat (Short et al., 2004). Polluted beaches in the area are characterized by a high-permeable sediment layer on top of a low-permeable layer, as lingering oil was found approximately 10 cm below the interface (Xia and Boufadel, 2011). Oil-contaminated sediments were anoxic with dissolved oxygen (DO) concentrations around 1.0 mg/L and had a high ratio of ammonia to nitrate (as nitrogen-N) (Li and Boufadel, 2011; Sharifi et al., 2011). In contrast, similar non-polluted sediments showed oxic conditions (DO > 3 mg/L) and a moderate ammonia to nitrate ratio. These findings strongly support the generally accepted perception that oil biodegradation is oxygen-limited in oil-contaminated sediments. In addition, the concentration of nutrients measured in the pore-water of contaminated sediments was < 0.5 mg N/L and < 0.04 mg P/L. These values are significantly lower than the nutrient concentrations reported to support near-maximum oil biodegradation rates, which are 2-10 mg N/L and ≥ 0.2 mg P/L, with an approximate optimal ratio of N/P around 10 (Atlas and Bartha 1972; Atlas 1981; Wrenn and Venosa 1996; Smith et al., 1998; Boufadel et al., 1999; Du et al., 1999; Wrenn et al., 2006). These concentrations are in good agreement with field study data from 1994 performed in Delaware Bay which reported that nitrogen concentrations of 3-6 mg N/L in the interstitial pore water stimulated hydrocarbon biodegradation by 2- to 3-fold over a moderately high natural attenuation rate, where the average interstitial nutrient concentrations averaged 0.8 mg N/L (Venosa et al. 1996).




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Biodegradation of crude oil requires about 0.04 g N/g oil (Atlas and Bartha, 1972) and about 3 g O2/g oil (Atlas, 1981). This rule of thumb can be used to estimate the required oxygen and nitrogen demand for complete biodegradation of a given amount of crude oil. For example, Venosa et al. (2010) calculated that 0.6-1.3 g N/kg sediment and 45-99 g O2/kg are needed (based on an average oil concentration of 15±18 g oil/kg sediment) for complete mineralization of shoreline oil derived from the Exxon Valdez spill. Local seawater contained only about 0.2 mg N/L and about 8-9 mg O2/L, and only about 0.2-0.3 L seawater/kg sediment is available in the pore water. These numbers clearly show the limitations for biodegradation processes under natural, non-stimulated conditions. Concentrations of primarily nitrates and phosphates, but also iron, typically limit oil biodegradation kinetics. Having an adequate supply of these rate limiting nutrients when massive quantities of petroleum hydrocarbons have been spilled into the marine environment is critical for stimulating biodegradation. Bioremediation strategies, which were used extensively after the Exxon Valdez spill, involved adding fertilizers containing nitrogen nutrients to speed up the rates of oil biodegradation (Atlas and Hazen, 2011). Finally, it is important to emphasize that biodegradation does not remove all of the hydrocarbons in an oil spill. Some compounds are recalcitrant to microbial attack (such as asphaltenes), but, due to low bioavailability, these compounds are also suggested to have a low eco-toxicological impact.

5.2 Bioremediation strategies in sediments In the aftermath of the Exxon Valdez oil spill, strategies for optimizing oil biodegradation have been studied in detail. Currently applied conventional oil spill response actions include mechanical containment and recovery of spilled oil, addition of chemical dispersants and physical cleanup of shorelines. Although, even under optimal conditions, mechanical and physical methods typically do not recover more than 10-15 % of the oil after a major spill (Zahed et al., 2011), and the use of chemicals may pose an additional environmental hazard. Bioremediation describes eco-friendly processes that can lead to complete mineralization of complex organic pollutants to CO2 and H2O by microorganisms. Remediation strategies based on biodegradation may be more efficient and cost less than other response options, especially in shorelines with porous media, where oxygen is available (Diaz, 2004). Biodegradation has also been demonstrated to significantly remove and restore both superficial (Gallego et al., 2007) and buried oil in contaminated beaches (Pontes et al., 2013). Bioremediation processes include natural attenuation, biostimulation, and bioaugmentation techniques, the latter ones aiming to increase the rate or extent of in situ biodegradation. Natural attenuation is dependent upon the complex metabolic network of indigenous microbial communities and their interactions with other organisms such as algae (McGenity et al., 2012). In seawater, one of the main factors limiting oil biodegradation is low hydrocarbon solubility. By producing biosurfactants and modifying cell membrane hydrophobicity, microorganisms increase and modulate HC bioavailability according to environmental conditions. Oxygen availability determines the pathway of HC breakdown. In the water column, under aerobic conditions, microbial enzymes activate oil compound molecules by incorporating oxygen atoms, generating corresponding alcohols, which are further oxidized to carboxylic acids that are degraded through β-oxidation (Rojo, 2009). In general, the performance of bioremediation depends on various physical, chemical, and biological conditions in the contaminated environment. Main requirements for efficient bioremediation are the presence of pollutant-degrading microorganisms and conditions that allow these microorganisms to




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proliferate, i.e. appropriate nutrient and oxygen conditions. In addition, pH and temperature are important parameters for microbial growth. Suboptimal composition of these and other parameters may result in low biodegradation rates. As for organic pollutants in general, aerobic conditions increase the extent and the rate of biodegradation (Das and Chandran, 2011). Oxygen typically penetrates only a few millimeters into coastal sediments (Militon et al., 2015) and thus, is not available for biodegradation of oil trapped in deeper layers. Although, biodegradation of petroleum HCs under anoxic conditions has also been reported, but aerobic bioremediation is generally faster and thus, metabolically preferred over its anaerobic counterpart (Cruz Viggi et al., 2015). Bioaugmentation Bioaugmentation involves, as mentioned in Chapter 3.3, the introduction of cultivated microorganisms specialized for biodegradation of target pollutants to enhance bioremediation in a given environment (Thompson et al., 2005; D'Annibale et al., 2006). Laboratory and field studies have shown that bioaugmentation can enhance oil biodegradation rates and reduce adaptation time (Mukherjee and Bordoloi, 2011; Szulc et al., 2014). For instance, in a field study, the introduction of selected microorganisms enhanced the total petroleum hydrocarbon removal by 76% in 180 days compared with the mere 3.6% removed in the control samples. However, the failure of the bioaugmentation technique due to the inability of exogenous microorganisms to compete with indigenous microorganisms has also been reported (Hosokawa et al., 2009). Biostimulation Biostimulation is the addition of nutrients or biosurfactants (Whang et al., 2008) to enhance microbial activity and growth in polluted areas. Due to the hydrophobic properties of oil, corresponding HCs cannot be accessed by all microorganisms in a community which are capable of degrading petroleum compounds. Similar to seawater, applying biosurfactants to oil-contaminated sediments can reduce the oil-water interfacial tension and thus increase the bioavailability of oil compounds for enhanced biodegradation (Szulc et al., 2014). Whang et al. (2008) found that by adding a rhamnolipid biosurfactant to diesel-contaminated sediments, the total petroleum hydrocarbon concentration was reduced from 7 to 0.2 g/kg within 88 days. Most biostimulation activities have focused on stranded oil, with application of fertilizers to increase natural degradation by the indigenous microbial flora. Biostimulation treatment is often combined with mechanical treatment to improve oxygen and nutrient availability. In marine environments, some growth- and biomass-stimulating factors are essential for oil biodegradation. Nitrogen and phosphorus is frequently applied to contaminated environments to increase the abundance of oil-degrading communities and thus increase bioremediation. Balancing nutrient concentrations is important, since high nutrient concentrations can inhibit microbial activity in some cases (Braddock et al., 1997). It is therefore important to avoid excess nutrients, which can cause detrimental effects, such as eutrophication (dense growth of algae's and bacteria's). During biostimulation, molar carbon/nitrogen/phosphorus (C/N/P) ratios of 100/10/1 have often been used (Obbard et al., 2004; Beolchini et al., 2010). Although, results from laboratory studies have also shown that certain microbial populations may require more than one N/P ratio for optimal degradation of different hydrocarbons (Smith et al., 1998). Various nutrient products for bioremediation are commercially available in different formulations, such as briquettes, granules or liquids. Liquid inorganic fertilizers have proven effective but require frequent application, and therefore oleophilic slow-release nutrient formulations have been developed, which promote hydrocarbon-degraders. For improved results, bioremediation may be combined with other clean-up procedures.




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Field biostimulation studies have been performed on beaches at Svalbard in the 1980’s with the slow-release oleophilic fertilizer Inipol EA22. The results of these studies indicated that application of the fertilizer to oil in beach sediments resulted in increased oil biodegradation in coarse sediments, but not in finer sediments (Sveum and Ladousse, 1989). A full-scale trial of several remediation processes at Svalbard in 1997, with a fuel oil in mixed intertidal shorelines, included remediation methods like sediment relocation (surf washing), mixing (tilling) and bioremediation (Guénette et al., 2003; Sergy et al., 2003). Over a one-year period, the addition of soluble or slow-release fertilizers to oiled sediments approximately doubled the biodegradation rate, compared to non-treated oiled sediments, and no acute toxicity was associated with the bioremediation treatment (Prince et al., 2003). Mixing of the sediment showed to increase sediment permeability and result in increased microbial activity (Owens et al., 2003). Biostimulation field experiments have also been conducted in Antarctic environments, where mesocosms 'enclosures' were placed in intertidal sandy beaches on the main island of the Kerguelen Archipelago (Pelletier et al., 2004). A weathered crude oil was added to the mesocosms and different fertilizers were added to the top of the oil (i.e. slow-release Inipol EAP 22 and various mixtures of dry fish compost with or without supplements of urea, phosphate and surfactants). Oil was depleted after 300 days in both untreated and treated sediments at seawater temperatures 3 - 4°C, and the various fertilizers accelerated the biodegradation rates (Pelletier et al., 2004). It was also observed that the treatment containing the lipidic neutral surfactant reduced the toxicity of the oil during the last three months of the experiment (Pelletier et al., 2004). Bioremediation employing the fertilizers Inipol EAP 22 and Customblen (slow-release granulated fertilizer) was also used extensively after the Exxon Valdez oil spill. Approximately 50,000 kg nitrogen and 5000 kg phosphorus were applied to the shorelines over the summers of 1989-1992 (Bragg et al., 1994). For a low-energy beach containing both surface and subsurface oil and treated with both fertilizers, it was estimated that the fertilizers enhanced oil biodegradation by 3 to 5 times over untreated beaches (Bragg et al., 1993).

5.3 Seafloor sediments During the Deepwater Horizon oil spill in the Gulf of Mexico in 2010, about 4.1 million barrels of light crude oil and gas were uncontrolled released from the Macondo well (MC252) over a period of 87 days. In addition, about 37,000 barrels of the chemical dispersant Corexit 9500 was applied on the sea surface and at the well at 1,500 m depth. The subsurface application of dispersants directly at the well resulted in the formation of a small oil droplet plume and contributed to the formation of complex aggregates consisting of oil residues, bacteria, and EPS referred to as oil-related marine snow (ORMS). The generation of ORMS has been used to explain the sedimentation of the Macondo oil to the seabed, and even used to calculate the amount of sedimented oil from the spill (Chanton et al., 2014; Valentine et al., 2014). A fallout of oil from the deepwater plume was also suggested, based on sediment analyses of 17 α(H), 21β(H)-hopane (Valentine et al., 2014), a well-known oil biomarker persistent to biodegradation (Prince et al., 1994). Based on measurements from > 3000 sediment samples from 534 locations it was suggested that 4-31 % of the oil from the DWH incident ended up in the deep ocean (Valentine et al., 2014). This shows that seafloor sediments represent an important sink for petroleum HCs after off-shore oil spills. Upon sedimentation, oil typically penetrates the upper 1-30 cm sediment layers (Bagby et al., 2017). The consequences of oil deposition on the seabed could impact the benthic community in general, or more specifically, perform an impact on coral communities/riefs/corals. The residue that remained on the surrounding seabed after the DWH oil spill was severely biodegraded and consisted of compounds that were too insoluble and too resistant to biodegradation to be further depleted, i.e. they were not bioavailable to oil-degrading microorganisms.




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As described for shoreline sediments, in situ bioremediation is regarded as efficient and sustainable strategy to clean up contaminated seafloor sediments. Different approaches have been proposed to stimulate naturally occurring microbial communities that degrade petroleum HCs including the subsurface addition of limiting-nutrients, electron acceptors and (bio)surfactants (Venosa et al., 2010; Hazen et al., 2016). The typical terminal electron acceptor in aerobic biodegradation processes is oxygen, which provides optimal biodegradation. However, in anoxic sediments, other electron acceptors, such as nitrate, nitrite, sulfate or metals, can substitute oxygen and facilitate HC degradation by microbes. Nitrate and Phosphate are often limiting nutrients in sediments, and supplementation with corresponding salts is known to stimulate biodegradation processes. Biosurfactants are surface active compounds produced by many bacteria to increase the bioavailability of HC, resulting in elevated oil biodegradation. However, the interaction between the bioavailability of electron acceptors (e.g. oxygen, nitrate, sulfate, Fe3+/Mn4+) and oil compounds is probably the most critical factor affecting the efficacy of sediment bioremediation systems. Since oil compound biodegradation occurs rapidly under aerobic conditions different engineered systems have been proposed to deliver oxygen into the sediments. Genovese et al. (2014) have recently developed a modular slurry system, which performs in situ aeration of the contaminated sediments, while minimizing the risk of spreading the contamination away from the treatment zone. Although the system was highly effective in stimulating the metabolism of aerobic oil-degrading bacteria, operating this system on an industrial scale will be highly labor and energy intensive. Other methods include the employment of oxygen-releasing compounds, such as calcium peroxide-based chemicals; however, in an anoxic environment many reactions are starved for oxygen. Challenges associated with controlling the rate at which oxygen is released and the subsequent consumption of oxygen by reactions with reduced chemical species (e.g. Fe2+, S2−) may limit the amount of oxygen utilized by oil-degrading microorganisms (Mapelli et al., 2017). Recently, a novel bioelectrochemical approach was proposed to stimulate aerobic biodegradation of petroleum HCs in anoxic marine sediments (Cruz Viggi et al., 2015). The so-called “Oil-Spill Snorkel” system consists of a conductive graphite rod positioned for electrochemically connecting two spatially segregated redox zones: the anoxic oil contaminated sediment and the overlying O2 rich water. The portion of the snorkel positioned in the polluted sediment serves as an electron acceptor (anode) for the oxidation of organic substrates in the sediment. Due to oxygen's high redox potential, the electrons flow through the rod up to the part exposed to the O2 rich water (the cathode). The movement is driven by the existing redox gradient and the electrons eventually combine with oxygen and protons to form water as a by-product. Besides serving as a virtually inexhaustible respiratory electron acceptor in the oxidation of petroleum HCs, the snorkel was suggested to indirectly stimulate oil biodegradation by sulfate-reducing bacteria via the scavenging of toxic sulfide diffusing from the bulk of the sediment (Daghio et al., 2015). Feasibility of the system has only been demonstrated in the laboratory, but due to low-maintenance and energy needs the system has potential for long-term remediation of contaminated sediments in remote open sea areas (Bagby et al., 2017).

5.4 Summary and recommendations – biodegradation of oil in sediments In contrast to oil present in the water column after a spill at sea, oil polluting coastal sediments is typically poor on easily biodegradable light fractions (e.g. short n-alkanes) and mainly composed of larger aromatic compounds, which is a general challenge for bioremediation strategies. Due to the porous structure of sediments, polluting oil can penetrate deep layers which are poor on oxygen and nutrients. Biodegradation processes in marine sediments are strongly influenced by abiotic factors such as temperature, oxygen, pH




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and concentrations of carbon and nutrients. Because of the high carbon content of oil and the low level of other nutrients essential for microbial growth, the rate and extent of biodegradation are limited by the low availability of nitrogen and phosphorus, but also oxygen in subsurface layers. Another rate-limiting factor is the increased viscosity of remaining oil resulting in a slow diffusion of degradable oil compounds to the oil-water interface and thus, reduced availability for microbial metabolization. Consequently, oil biodegradation can be strongly enhanced by addition of nutrients and also surface-active compounds (such as biosurfactants) to increase bioavailability. However, since commercially available fertilizers will be quickly diluted and washed away from intertidal shorelines, supply of nutrients and biosurfactants to subsurface oil residues is an operational challenge. Numerous studies in the aftermath of the Prestige and Exxon Valdez oil spill have shown that biodegradation of oil in coastal sediments can be strongly stimulated by addition of nutrients/fertilizers. However, reliable data on a systematic evaluation of bioremediation products available in the market is still missing and would be necessary to assess the potential of bioremediation from an operational perspective. Bioremediation studies on shorelines with moderately to heavily weathered oil have not been reported. This operational relevant knowledge gap should be filled by lab-experiments testing the effect of selected bioremediation products on biodegradation of weathered oil in marine sediments. Such products may include fertilizers, dispersants and biosurfactants. Commercially available bioremediation products (and modified formulations thereof) can be applied and impact on biodegradation can be assessed by analyzing following parameters: • Microbial community structure analyses: The impact of the spilled oil on the local microbial community

should be studied by using 16S amplicon sequencing on an Illumina platform. Changes in the community composition will give valuable information on biodegradation performance and when biodegradation processes need to be re-stimulated. It is expected that successful bioremediation will result in a qualitative reconstitution of the original microbiota. Detailed monitoring of the microbial community can thus also indicate when bioremediation efforts have reached their limits and can be terminated.

• Microbial activity: Total biological activity in sediment samples can be easily measured by respirometry using the OxiTop system (WTW). Monitoring the oxygen consumption rate will provide information on biodegradation performance and indicate time points for re-stimulation or termination of bioremediation efforts.

• Chemical analyses: Chemical oil compound dynamics in the sediment and tidal water run-off should be

monitored using high resolution GC-MS-MS analysis. This will allow detailed monitoring of biodegradation processes on oil compound level. Resulting data can also be used to determine to which extend the polluting oil has been degraded and how the perspectives for bioremediation under given conditions are expected to be.

• Ecotoxicity: Standard microtox tests can be used to assess acute toxicity during the degradation process.

In combination with chemical and microbiological data, microtox tests will be an important component to assess the status of bioremediation efforts. Moreover, this method can also be used to evaluate the toxicity of bioremediation products.




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6 Ecotoxicity of biodegraded oil The degree of toxicity of oil in the environment during a spill and in the aftermath of a spill depends highly on the properties of the oil and the environment in which it has been released. Where the oil is located at any given time, determines which biotopes and organisms will be affected by it. Exposure in the pelagic environment is a very different type of exposure than exposure in ice or sediments. In the pelagic environment, dissolved oil components are considered the main driver for toxicity due to their bioavailability, but weathered oil residue may adhere to organisms as well, as observed for haddock (Sørensen et al., 2017). Small oil droplets may be filtered by filter-feeding organisms (e.g. copepods and mussels) living in open water, on ice and on sediments (Nordtug et al., 2015; Andreassen et al., 2013). Detritus-feeding organisms living in the sediments will mainly be exposed to residual oil through digestion of contaminated sediments, but toxic hydrocarbons can continue to diffuse from the residual oil which can affect pelagic biota as well as for sediment-living organisms. In sea ice, entrapped oil underneath ice will be sheltered from evaporation and photo-oxidation, and biodegradation rates may be lower at lower temperatures. Thus, organisms living on and under the ice can be exposed to more hydrocarbons over longer periods of time when oil is entrapped in adjacent ice compared to when a weathered oil slick floats into ice. Water-soluble components from oil entrapped in ice have been shown to leak through adjacent brine channels. Limited studies exist concerning the changes in oil toxicity due to microbial degradation, and most experiments have been conducted with soil and groundwater (Wang et al., 1990; Belkin et al., 1993; Tiehm et al., 1997). In marine environments, it is generally assumed that weathering of oil causes it to become less toxic because volatile toxic components are lost to the atmosphere and oil is diluted in a large water volume. Dissolution of oil components from residual oil is a continuous process, but is reduced over time as the components with highest water-solubility are removed from the weathered oil residue. The most highly water-soluble oil components like the BTEX and C0-C1 naphthalenes drain rapidly from the oil into water and are biodegraded, whereas the more oleophilic components like heavier PAHs are released over a longer timer span. Thus, the bioavailability of heavier components may limit microbial metabolism. Biodegradation, however, does not cause instant disappearance of oil components, as the majority of hydrocarbons are degraded into intermediate metabolites prior to their complete mineralization. Depending upon the type of hydrocarbon, mineralization (the complete conversion of oil components into CO2 and H2O) can be a time-consuming process and intermediate metabolites may display toxicity. Little information is available on the toxicity of intermediate metabolites. For oil spills that have been compartmentalized in sediments over long time, it is hard to evaluate when the oil had biodegraded enough not to cause any harm to organisms. Even in the water phase, very limited information exists on the toxicity of degradation products (metabolites) from microbial metabolism of oil components.

6.1 Future Studies of Ecotoxicity Future experiments should combine biodegradation studies and ecotoxicity effect studies using different substrates (water, sediment and ice). This should include adding oil to a substrate (e.g. ice, water, sediments) and conducting subsequent toxicity tests of the substrate over time. These types of experiments will allow to study toxicity dynamics as a function of time. A suggestion using seawater as a substrate is to generate a dissolved fraction of oil and leave it for biodegradation over a long period (weeks to months) where samples for chemical characterization and ecotoxicity tests are taken at different time points. Ecotoxicity testing should be done on relevant organisms for the Norwegian cold-water environment, like early life stages of fish, which are sensitive to oil exposure. Similarly, ecotoxicity tests can be performed on oil-contaminated sediments, which has been biodegraded over time using sediment-species like the amphipod Corophium volutator.




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Appendix

A Abbreviations MGO - Marine Gas Oil MSD - Marine Special Distillate RFO - Residual Fuel Oil HGO - Heavy Gas Oil WRO - Wide Range Oil SDM - Special Distillate Marine MFO - Medium Fuel Oil ULSFO - Ultra Low Sulphur Fuel Oil VHFO - Very Heavy Fuel Oil HDME - Hybrid Wide Range Gas oil IFO - Intermediate Fuel Oil WRG - Wide Range Gas oil GO - Gas Oil WRD - Wide Range Diesel oil HDME 50 - Heavy Distillate Marine ECA 50 PM - Particulate Matter GC-MS - Gas Chromatography Mass Spectrometry GC-FID - Gas Chromatography Flame Ionizing Detector GCxGC-QTOF MS - Two-dimensional Gas Chromatography Quadrupole-Time-Of-Flight Mass Spectrometry UCM - Unresolved Complex Mixtures BTEX - Benzenes, Toluenes, Ethylbenzenes, Xylenes OMA - Oil–Mineral Aggregation ORMS - Oil-Related Marine Snow HC - HydroCarbon THC - Total HydroCarbon EOC - Extractable Organic Carbon TEOC - Total Extractable Organic Carbon WAF - Water Accommodated Fraction EPS - Extracellular Polymeric Substances

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