Models have been developed that adequately address the transport and fate of surface oil spills (e.g. ASA 2011 and McKay 2009).  These models take into account the effects of wind and surface currents and are useful predictions of the movement of oil at the surface and into the water column.  They predict the concentrations of oil within a volume of water which permits the comparison of those exposure values to toxic effects based measurements of adverse effects observed in laboratory studies (Gallaway et al., in review). When oil is released at the surface of the water or at depth it begins to spread, thin and dilute into 2 or 3 dimensions.  For surface spills the oil begins to spread primarily in a horizontal direction based on currents and wind-driven transport with the more soluble components diluting into the third or vertical dimension, which includes the water under the surfaced oil and release of the more volatile components to the atmosphere. Subsurface releases generally have an initial momentum that provides rapid dilution by entrainment of surrounding waters during the initial rise of the fluids from the bottom, driven by any jet turbulence from the simultaneous release of natural gasses and the natural buoyancy of the oil.  This rapid dilution is followed by continued vertical transport based on the specific gravity or density and size of oil droplets created during the initial release including entrained diffusion of the more soluble components into the water column accompanied by horizontal dispersion by subsurface currents.  Other processes that occur and which may influence the bioavailability of oil components include biodegradation processes and physical/chemical changes that may occur due to pressure changes (e.g. clathrate formation) or emulsion formations by incorporation of water into the oil mass.  In addition to the physical contact of organisms with oil (e.g. fouling on birds) organisms also respond to the bioavailable chemical components that are present in water and food.

Another component of the fate evaluation for models is the change in chemical composition and bioavailability of oil that occurs during physical weathering and biological activity, mineralization and biodegradation.  The relative contribution and total concentrations of oil compounds changes through time and the rate of change varies depending on the environmental compartment in which the oil resides.  Surface oil that emulsifies undergoes less efficient biodegradation and appears to be less available for biological uptake into tissues than oil that is dispersed into the water column as small droplets.  Oil that is stranded on beaches or within interstitial waters between grains of sediment also undergoes slower rates of biodegradation as demonstrated by the ‘lingering oil’ effects seen in Prince William Sound decades after its stranding.  These important concepts are further developed in Sections 3 and 5 (fate of oil and biodegradation processes, respectively).

The transport and fate of a released quantity of oil has the potential to affect the population abundance, standing crop/production, increase daily mortality coefficients, and response to stressor exposure within different environmental compartments.  Variation in these attributes for each environmental compartment are expected and assessments can be generated based on assumptions that are meant to provide a maximum level of effect using the most conservative estimates or to best characterize the distinctions between the compartments.

The bioavailable fractions of oil derived from the soluble compounds that are present during the initial release of the oil or in locations where oil becomes re-concentrated (e.g. at convergence zones, pycnoclines, shorelines, air/water and ice/water interfaces).  As dilution of these compounds and biodegradation processes occur the relative and total concentrations may be reduced.  Eventually, the concentrations of the soluble compounds decreases to the point that the acute effects may be replaced by more subtle, sublethal doses resulting in longer term exposure to lower but more constant concentrations of the remaining compounds.  Measurements made during accidental and experimental spill events have shown that the higher total petroleum hydrocarbon (TPH) concentrations observed in the upper 3 m of the water column range from ~2 to 37 mg/L for physically dispersed oil and between 36 and 527 mg/L for chemically dispersed oil (Trudel et al., in press; Brandvik et al. 1995).  The chemical dispersants introduced more oil into the water column but the effects based concentrations for measured petroleum compounds were less per unit of oil for the chemical dispersions (Gardiner et al., 2013).  These field evaluations also showed reduced TPH concentrations of 50% within an hour and approximately 2 orders of magnitude differences in concentration within the upper 10 m of the water column for both physically and chemically dispersed oil. 

Empirical studies of the acute toxicity of oil and its components or modeling the effects of complex mixture exposure and uptake are alternative methods for estimating the effects of oil in different compartments.  For acute responses to oil exposure, including mortality and avoidance behaviors toward soluble components, toxicity testing is appropriate.  The exposure regimen for this type of assessment has been developed to emulate field conditions observed during spill events and controlled experimental spill events that were described in the previous paragraph.  The methods employ spiked concentrations of oil followed by dilutions with clean water to simulate the dispersion/dilution processes that occur in the field.  There are a significant number of experiments that have been performed using this protocol and it provides a good characterization of the acute effects that might be expected from dynamic exposure concentration from an oil spill at the surface or a depth (refer to Section 6 on Ecotoxicology for further toxicological information). The longer term, chronic and sublethal responses of organisms can be modeled by using constant exposure to lower concentrations of oil that might occur with organisms drifting with the oil after the initial dispersion or those inhabiting contaminated sediments.  There are also models that have been built to estimate the potential effects of the uptake of the bioavailable components into the tissues of organisms resulting in chronic responses to oil components.  These models are based on the relative narcosis effect of many PAH compounds (Di Toro et al. 2007).

Modeling population responses that can occur as a result of exposure to oil should include a representation of the temporal and spatial distribution of age classes of the VECs.  The spatial evaluation needs to take into account the vertical and horizontal distribution of the abundance of each VEC and its relationship to key environmental components.  Examples of environmental components include river discharges, lagoon occupation, annual and permanent ice, propensity to cluster at zones of convergences (e.g. currents, pycnoclines, upwelling and downwelling) and different types of shorelines and bottom types for each of the VEC age classes.  The populations of VECs undergo different natural stressors within each of these habitats ranging from environmental exposure conditions, inter and intra-annular differences in environmental conditions, differences in spawning success and variable predation rates.  As a result, the influence of an additional stressor will influence populations at different rates.  These natural and ongoing impacts produce natural mortality rates that can be determined.  For fish the daily mortality rates can be determined by examining daily growth rings on otoliths and the relative abundance of individuals at comparable ages.  Invertebrates generally lack these types of structures and their mortality rates are based on the relative abundance of different size classes within a population.  Once the natural mortality rates are estimated, the impact of another stressor can be added to determine what the combined mortality will be. 

For example, a recent assessment of the acute effects of chemically dispersed oil into Arctic pelagic open ocean, ice-free waters on larval Arctic cod (Boreogadus saida) was modeled to determine the adult female equivalents that would be lost as a result of a large oil spill (Gallaway et al., in review).  The fisheries model obtained daily mortality rates from otolith assessments and applied toxicity information obtained by Gardiner et al. (2013) and assumed that a large oil spill would be at the center of the maximum population sizes observed.  Comparing the increased mortality rates of the oil indicated that number of adult female equivalents impacted by the oil were a fraction of the natural range in abundance of this species.  Based on these results the resiliency of the population of Arctic cod would be sufficient to recover during the following spawning year.  Had the oil not been treated with dispersants it would have been transported toward shore where juvenile Arctic cod are present in very large schools numbering in the millions, and the Arctic cod population would be subjected to greater stress. 

The following sections include technical information on copepod, amphipod, and fish population ecology and discuss the types of information necessary to assess the potential population effects of oil on two groups of species, copepods and fish.