Oil in the sea from anthropogenic sources, whether from spills or chronic releases, is perceived as a major environmental problem. Major oil spills occur occasionally and receive considerable public attention because of the obvious attendant environmental damage, including oil-coated shorelines and dead or moribund wildlife, especially oiled seabirds and marine mammals. Acute effects may be of short duration and limited impact, or they may have long-term population- or community-level impacts depending on the timing and duration of the spill and the numbers and types of organisms affected. Oil also enters the sea when small amounts are released over long periods, thus creating chronic exposure of organisms to oil and its component chemical species. Sources of chronic exposures include point sources, such as natural seeps, leaking pipelines, offshore production discharges, and non-point runoff from land-based facilities. In these cases, there may be a strong gradient from a high to a low oil concentration as a function of distance from the source. In other cases, such as with land-based runoff and atmospheric inputs, the origin of the oil is a nonpoint source, and environmental concentration gradients of oil compounds may be weak. Chronic exposures may also result from the incorporation of spilled oil into sediments in which weathering of oil is retarded, and from which nearly-fresh oil may be released to the water column over extended periods. In recent years, it is the long-term effects of acute and chronic pollution that have received increasing attention (Boesch et al., 1987).
What separates short-term from long-term effects is open to debate. Boesch et al. (1987) suggested that effects of duration longer than two years should be considered as long-term. These can be either effects that persist after an initial insult, or effects that result from persistent pollution. We do not know the upper bound for the potential length of a long-term effect. It is likely to be at least the length of a generation of the affected organisms, and it may be longer. An effect can be either direct damage to a resource or damage to the ability of an environment to support a resource. An effect can be said to be over when complete recovery has taken place. The quantification of both effects and recovery are difficult, particularly when they must be measured against a changing marine environment (Figures 5-1A and B) (Wiens,
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Oil in the Sea III: Inputs, Fates, and Effects 5 Biological Effects of Oil Releases HIGHLIGHTS This chapter focuses on: The complexity of determining effects of petroleum hydrocarbons in the marine environment within the background of highly complex natural variables. The advances in our understanding of acute and chronic effects of petroleum hydrocarbons in the marine environment made since the 1985 NRC Review Oil in the Sea. The advances in modeling for assessing oil impacts in the marine environment. The advances in our understanding of how communities respond to petroleum discharges especially biogenically structured communities. The unique aspects of production fields and natural seeps in understanding the long-term effects of petroleum discharges in the marine environment. The identification of important information gaps that still exist in our understanding of the effects of petroleum hydrocarbons on populations of marine organisms and ecosystems and the time course of recovery. Oil in the sea from anthropogenic sources, whether from spills or chronic releases, is perceived as a major environmental problem. Major oil spills occur occasionally and receive considerable public attention because of the obvious attendant environmental damage, including oil-coated shorelines and dead or moribund wildlife, especially oiled seabirds and marine mammals. Acute effects may be of short duration and limited impact, or they may have long-term population- or community-level impacts depending on the timing and duration of the spill and the numbers and types of organisms affected. Oil also enters the sea when small amounts are released over long periods, thus creating chronic exposure of organisms to oil and its component chemical species. Sources of chronic exposures include point sources, such as natural seeps, leaking pipelines, offshore production discharges, and non-point runoff from land-based facilities. In these cases, there may be a strong gradient from a high to a low oil concentration as a function of distance from the source. In other cases, such as with land-based runoff and atmospheric inputs, the origin of the oil is a nonpoint source, and environmental concentration gradients of oil compounds may be weak. Chronic exposures may also result from the incorporation of spilled oil into sediments in which weathering of oil is retarded, and from which nearly-fresh oil may be released to the water column over extended periods. In recent years, it is the long-term effects of acute and chronic pollution that have received increasing attention (Boesch et al., 1987). What separates short-term from long-term effects is open to debate. Boesch et al. (1987) suggested that effects of duration longer than two years should be considered as long-term. These can be either effects that persist after an initial insult, or effects that result from persistent pollution. We do not know the upper bound for the potential length of a long-term effect. It is likely to be at least the length of a generation of the affected organisms, and it may be longer. An effect can be either direct damage to a resource or damage to the ability of an environment to support a resource. An effect can be said to be over when complete recovery has taken place. The quantification of both effects and recovery are difficult, particularly when they must be measured against a changing marine environment (Figures 5-1A and B) (Wiens,
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Oil in the Sea III: Inputs, Fates, and Effects FIGURE 5-1 Hypothetical examples show how the impact of an oil spill and subsequent recovery can be assessed when the system under study undergoes natural variations (solid line). In (A), the system varies in time, but the long-term mean remains unchanged. In (B), there is a long-term decline in the state of the system (e.g., population size). Dashed lines indicate a “window” of normal variation about the mean (e.g., a 95 percent confidence interval). Operationally, “impact” occurs when the system is displaced outside this “window” (from Wiens, 1995, American Society for Testing and Materials). 1995; Spies et al., 1996; Peterson, 2001). Perhaps more difficult than detecting an effect is determining its significance (Boesch et al., 1987) (Figure 5-2). The spatial extent, persistence and recovery potential are all important, as is the perceived or monetary value of the affected resources. All else being equal, damage to a large area is more significant than damage to a small area of similar habitat. Damage to a small area that contains a highly valued resource can be of greater significance than damage to a much larger area devoid of valued resources. These issues are hotly contested after major pollution incidents. DETERMINING EFFECTS IN A VARIABLE ENVIRONMENT Oil can kill marine organisms, reduce their fitness through sublethal effects, and disrupt the structure and function of marine communities and ecosystems. While such effects have been unambiguously established in laboratory studies (Capuzzo, 1987; Moore et al., 1989) and after well-studied spills (Sanders et al., 1980; Burns et al., 1993; Peterson, 2001), determining the subtler long-term effects on populations, communities and ecosystems at low doses and in the presence of other contaminants poses significant scientific challenges. Multiple temporal and spatial variables make deciphering the effects extremely difficult, especially when considering the time and space scales at which marine populations and ecosystems change. Marine ecosystems change naturally on a variety of time scales, ranging from hours to millennia, and on space scales ranging from meters to that of ocean basins. There are many causes of ecological change aside from oil pollution, including human disturbance, physical habitat alteration, other pollution, fishing, alteration of predation patterns, weather, and climate. Time scales at which oil affects the ocean range from days to years or even decades for some spills; chronic pollution occurs over years to decades. Oil spills affect the oceans at spatial scales of tens of square meters to thousands of square kilometers; chronic oil pollution can affect areas as small as a few square centimeters and as large as thousands of square kilometers. Climatic changes can complicate the interpretation of contaminant impacts, especially if they have different effects on control and impact stations in an experimental design, or if a long time series of data is used to establish the “norm.” Considerable scientific attention has been directed to understanding how climatic forcing affects marine ecosystems and fisheries (Beamish, 1993; Hare and Francis, 1995; McFarlane et al., 2000). Climate change can be cyclical, e.g., the Southern Ocean Oscillation the Pacific Decadal Oscillation (Barnston and Livesy, 1999), the North Atlantic Oscillation (Trenbreth and Hurrell, 1994; Hare and Mantua, 2000), or can be secular e.g., gradual rise in upper ocean temperature. The biological effects of oil pollution are often referred to as acute or chronic. Spills are commonly thought of as hav
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Oil in the Sea III: Inputs, Fates, and Effects FIGURE 5-2 Schematic representation of oil spill influences on seabirds. The three primary avenues of effects, on population size and structure, reproduction and habitat occupancy, are highlighted (from Wiens, 1995, American Society for Testing and Materials). ing short-term effects from high concentrations of petroleum. Chronic pollution, such as might occur from urban runoff into coastal embayments, may have continuous effects at low exposures. Not all oil pollution is clearly separable into these two categories. For example, exposure and effects are known to occur for long periods after some spills (Vandermeulen and Gordon, 1976; Sanders et al., 1980; Spies, 1987; Teal et al., 1992; Burns et al., 1993), and chronic exposures can be quite high, as is the case near petroleum seeps (Spies et al., 1980; Steurmer et al., 1982). The reader should bear this in mind during the ensuing discussion of the effects of acute and chronic exposure to oil. Additionally, this report generally focuses on the effects to benthic and wildlife populations, which were found to be most at risk from oil (Boesch et al., 1987). It is within this complex multi-scale, spatial, and temporal environment that we are challenged to detect change caused by oil in the sea, and to assess the damage at the level of individuals, populations, communities, and ecosystems. Difficulty of detection increases with level of biological organization, with spatial and temporal scales of the affected system, and with the inherent variability of the system. Similarly, determination of complete recovery is complicated by this inherent variability. The complex mosaic of change in the ocean has two aspects with regard to detecting the effects of oil pollution. First, it poses strategic challenges to determining the impact of oil through gathering observational data, as inevitably we make assumptions about the variability in the ecosystem and that variability can obscure large and continuing impacts. Second, the actual impact of the oil may be more complex than we realize if it interacts with spatially or temporally constrained phenomena. In the closing decades of the twentieth century it was commonly held that the “balance of nature” has been severely altered by human actions. Consequently, much of our public policy was directed toward maintaining the status quo or returning ecosystems to a more pristine condition. While there is little doubt that human activities have had considerable impact in oceanic ecosystems, there has not been an equally widespread appreciation of how ecosystems change without human interference. The occurrence of several well-developed El Niños in the 1980s and 1990s made strong impacts on the public consciousness about longer-term cycles in the oceans. In Alaska, which has a strong resource-based economy, the rise and fall of salmon stocks in concert with the Pacific Decadal Oscillation (Beamish, 1993; Francis et al., 1998; Beamish et al., 1999) is now well known in the general population. Because public appreciation of ecosystem change seems to be following the growing scientific attention to long-term change in the oceans, the expectation that recovery of a polluted site will result in the return of an ecosystem to the state that it was in at the time of a pollution event is changing. The observational framework for quantifying impacts involves determining differences based on sets of observations
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Oil in the Sea III: Inputs, Fates, and Effects PHOTO 19 Oil from the Lake Barre spill, May 1997, spill formed a narrow band on the marsh stems, and there was little oiling of the soils. Also, the oil is highly degradable. Thus, most of the marsh vegetation survived. (Photo courtesy of Jacqui Michel, Research Planning, Inc.) at impacted and putatively non-impacted areas, or at one or a series of sites where before-and-after impact observations are available. Ideally, before-after and control-impact (BACI) observations can be made (Stewart-Oaten et al., 1992; Wiens and Parker, 1995; Peterson et al., 2001). The inherent assumption is that the variability of the ecosystem is sufficiently controlled (in the experimental sense) by these designs, which may or may not be correct. Controlling for impact by comparison of sites that have been affected and not affected allows for a variety of potentially important non-oiling variables to influence the system—such as differences in water temperature, salinity, or substrate type. For example, see Bowman (1978) for a case where high temperatures were documented to have a differential effect on intertidal invertebrate mortality, that might have otherwise been attributed to oil or dispersant toxicity. Usually an attempt is made to find study sites that are as similar as possible in factors suspected to be important. When effects are determined based on comparisons of before-impact conditions and after-impact conditions, it is possible that the ecosystem has changed in ways unknown to the observer. The chances of making errors can be lessened when: (1) multiple sites are used in each of the impacted and non-impacted sites, (2) multiple times are used in the time series, or, even better, (3) when both multiple sites and multiple times are available. Nevertheless, unreported factors not related to oil can interfere with ecosystem processes in ways that disguise the effects of pollution. Of course, with each additional kind of impact that is measured, the chance of making an error (Type I) rises. At the same time, the mosaic of complex interactions and the resultant changes in ecosystems makes it possible to miss an impact that occurs (Type II error). For example, if an oil spill occurs when the pelagic larval stages of a fish species are developing near the sea surface, many or most of these larvae may die. If these larvae were to be the foundation of what would otherwise have been a strong year class for that fish species and whose population is maintained by infrequent large year classes, then the impact could be much larger than otherwise supposed. That would be a disproportional effect on a process that is temporally constrained. There are also examples of potential impacts on processes that are disproportionate because they are spatially constrained. For instance, a small spill around a seabird habitat where a large proportion of a population is gathered for breeding could have a disproportionately large impact. A good example of this occurred when an estimated 30,000 oiled seabirds washed up along the coasts of the Skagerrak following a small release of oil from one or two ships (Mead and Baillie, 1981). At the other extreme, the wreck of the Amoco Cadiz off the coast of Brittany, France, resulted in the release of 230,000 tonnes of crude oil into coastal waters and the death of less than 5,000 birds (Hope-Jones et al., 1978). These examples help illustrate that the volume of oil is only one factor determining mortality of birds and the
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Oil in the Sea III: Inputs, Fates, and Effects weak empirical relationship between spill volume and bird mortality points out the need to better understand the other sources of uncertainty (such as spill timing). Assessing recovery after a pollution event is perhaps even more challenging than assessing initial damage. Recovery is further removed in time from the acute phase of the damage, and thus may be occurring in a different environmental framework than that which existed at the time of the accident. If there is variation in time, but the long-term mean remains stable, recovery might be judged by some to have been complete when the environmental variable of concern returns to within the normal range of variation (see Fig 5-1A, Wiens, 1995). In contrast, if the long-term environmental mean is changing, then recovery would occur when the variable of concern returns to within a range of variation around a short-term mean that will be quite different from that when the perturbation occurred (Fig. 5-1B). To assess recovery quantitatively requires either a well designed BACI approach, or one that compares measurements of the environmental variable of interest along a gradient of perturbation (Wiens, 1995). This gradient can be in space or time. One must be certain that, when numbers of organisms are being compared for assessment of recovery, attributes such as age or reproductive potential be taken into account. For example in marine birds, young, inexperienced animals do not have the same value to the population as experienced breeding adults. The natural variability inherent in estimates of populations introduces considerable uncertainty in assessing impact and recovery from pollution events. Confidence limits in excess of 20 percent of the mean size are usual in wildlife censuses. Such variability in the estimated mean makes it certain that population changes will be difficult to detect without a high degree of replication spatially and temporally before and after an event. More importantly, under some circumstances estimates of recovery based on the population returning to a “window” of natural fluctuation could minimize the time to true recovery. Other important considerations in evaluating oil pollution effects are the roles that laboratory studies, mesocosms and impact modeling play in complementing, or, in some cases, replacing the field observations discussed above. Laboratory studies avoid the aforementioned problem of lack of control, but their improved precision disallows the wide range of possible interactions and indirect effects that can occur in complex ecosystems. Such indirect effects might be substantial. For example, in the Exxon Valdez and Torrey Canyon oil spills, destruction of the algal cover had indirect impacts on limpets and other invertebrates (Southward and Southward, 1978; Peterson, 2001). Such successional, reverberating or cascading indirect effects in a complex ecosystem may be very important, but are not captured by laboratory studies. The bulk of laboratory studies have examined oil impacts on organism mortality and health using dissolved oil or seawater suspensions. Most experiments are conducted for short durations (Capuzzo, 1987), which does not take into account long-term effects. Field observations and laboratory experiments, as ways of knowing effects, represent two ends of a spectrum. Field observations allow little or no control of interactions between the full complement of ecosystem variables; laboratory experiments allow control of the interaction of single components that have been removed from the ecosystem. Taken together they still may not tell the whole story of oil impact. As a result, efforts have been made to bridge the gap between these two ends of the experimental control-field complexity continuum. Intermediate approaches include: laboratory experiments with multiple species, or communities that include environmental components (micro-and mesocosms); and field experiments, for example that put oiled sediments into the environment to be colonized by natural populations of animals and plants. The modeling of the impacts of oil spills and their potential effects provides another route for predicting the potential effects of spilled oil. Oil spill impact modeling, which was originally applied to predicting the fate of oil in the environment, has recently been extended to prediction of effects (McCay, 2001). In this chapter, we provide a brief review of progress in addressing the research recommendations of the 1985 Oil in the Sea report (NRC, 1985). We then examine the acute and chronic effects of oil at the organism, population and community/ecosystem levels. In the review, we single out marine birds and mammals for special attention because of their high visibility in spills and the great public concern for their welfare. It has been our intent to focus on the significant advances in knowledge and perceptions of the effects of oil in the sea, rather than to provide a detailed examination of the many research papers that have been published since the completion of the NRC (1985) or the Boesch and Rabalais (1987) reviews. Progress Since 1985 Report Since the major review of oil in the sea conducted by the National Research Council and published in 1985, there have been thousands of individual studies contributing to our overall understanding of the acute and chronic toxicity of oil in the marine environment and the restoration and recovery of oiled habitats. The major recommendations of the 1985 report were: To expand studies of effects of low concentrations of petroleum hydrocarbons on marine organisms, especially larval and juvenile stages; To examine the apparent coincidence of petroleum hydrocarbon exposure with increased prevalence of pollution-related disease in marine organisms; To examine the impacts of petroleum hydrocarbons in polar and tropical habitats; To better integrate laboratory studies with field investigations;
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Oil in the Sea III: Inputs, Fates, and Effects PHOTO 20 (A) Julie N spill of IFO 380 coated the intertidal marshes of the Fore River near Portland, Maine. Photo taken in September 1996. (B)Photo Same area, one year post spill, September 1997. Most of the vegetation had completely recovered. Factors leading to recovery were: the plants were already in senescence when oiled, little or no sediment contamination occurred; large tidal range with good flushing. (Photos courtesy of Jacqui Michel, Research Planning, Inc.)
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Oil in the Sea III: Inputs, Fates, and Effects To assess the potential effects of petroleum hydrocarbons at population and ecosystem levels, especially for fish stocks and critical habitats such as mangroves and coral reefs. Many of the studies conducted since 1985 have addressed these recommendations and have led us to a better understanding of the vulnerability of different habitats and different life history stages of a variety of marine organisms. Field and laboratory investigations have integrated studies of chemical fate and biological effects so that an improved understanding of the recovery process has been defined. In addition, oil spills have been monitored for longer periods of time and across wider far-field conditions to examine the chronic, long-term effects of spills. In their synthesis volume, Long-Term Environmental Effects of Offshore Oil and Gas Development, Boesch and Rabalais (1987) identified several important areas of research needs that complemented those identified in the Oil in the Sea report. Based on detailed consideration of the probability and severity of effects and the potential for resolution of uncertainties, they identified ten categories of potential long-term environmental effects. These were: High Priority Chronic biological effects resulting from the persistence of medium and high molecular weight aromatic hydrocarbons and heterocyclic compounds and their degradation products in sediments and cold environments. Residual damage from oil spills to biogenically structured communities, such as coastal wetlands, reefs and vegetation beds. Effects of channelization for pipeline routing and navigation in coastal wetlands. Intermediate Priority Effects of physical fouling by oil of aggregations of birds, mammals, and turtles. Effects on benthos of drilling discharges accumulated through field development rather than from exploratory drilling. Effects of produced water discharges into nearshore rather than open shelf environments. Lower Priority Effects of noise and other physical disturbances on populations of birds, mammals, and turtles. Reduction of fishery stocks due to mortality of eggs and larvae as a result of oil spills. Effects of artificial islands and causeways in the Arctic on benthos and anadromous fish species. Many of these concerns have now been fully addressed and are detailed in several synthesis reports written since 1987 (Box 5-1). Those topics not covered in synthesis reports will be addressed in this report. Toxic Effects of Petroleum Hydrocarbons The responses of organisms to petroleum hydrocarbons can be manifested at four levels of biological organization: (1) biochemical and cellular; (2) organismal, including the integration of physiological, biochemical and behavioral responses; (3) population, including alterations in population dynamics; and (4) community, resulting in alterations in community structure and dynamics. Impairment of behavioral, developmental, and physiological processes may occur at concentrations significantly lower than acutely toxic levels; such responses may alter the long-term survival of affected populations. Thus, the integration of physiological and behavioral disturbances may result in alterations at the population and community levels. The effects of petroleum hydrocarbons in the marine environment can be either acute or chronic. Acute toxicity is defined as the immediate short-term effect of a single exposure to a toxicant. Chronic toxicity is defined as either the effects of long-term and continuous exposure to a toxicant or the long-term sublethal effects of acute exposure (Connell and Miller, 1984). Acute and chronic toxicity of petroleum hydrocarbons to marine organisms is dependent upon: concentration of petroleum hydrocarbons and length of exposure, persistence and bioavailability of specific hydrocarbons, the ability of organisms to accumulate and metabolize various hydrocarbons, the fate of metabolized products, the interference of specific hydrocarbons (or metabolites) with normal metabolic processes that may alter an organism’s chances for survival and reproduction in the environment (Capuzzo, 1987), and the specific narcotic effects of hydrocarbons on nerve transmission. Many of the early studies of acute toxicity focused on the toxicity of individual compounds to marine organisms or the differential toxicity of crude and refined oils (Anderson, 1979). The findings from these types of studies can be summarized as follows: The acute toxicity of individual hydrocarbons is largely related to their water solubility. The acute toxicity of a specific oil type is the result of the additive toxicity of individual compounds, especially aromatic compounds. Narcotic effects of individual petroleum compounds are an important component of acute toxicity and are most closely related to low molecular weight volatile compounds (Donkin et al., 1990). Sublethal effects following acute or chronic exposure to petroleum hydrocarbons include disruption in energetic processes; interference with biosynthetic
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Oil in the Sea III: Inputs, Fates, and Effects BOX 5-1 Recent National Research Council Synthesis Reports Addressing Oil in the Sea and Offshore Oil and Gas Development The following list reflects the extensive attention the NRC and government agencies have placed on the effect of petroleum in the environment. Assessment of the U.S. Outer Continental Shelf Environmental Studies Program: I. Physical Oceanography, 1990. Assessment of the U.S. Outer Continental Shelf Environmental Studies Program: II. Ecology, 1992 Assessment of the U.S. Outer Continental Shelf Environmental Studies Program: III. Social and Economic Studies, 1992. Assessment of the U.S. Outer Continental Shelf Environmental Studies Program: IV. Lessons and Opportunities, 1993. The Adequacy of Environmental Information for Outer Continental Shelf Oil and Gas Decisions: Georges Bank, 1991. The Adequacy of Environmental Information for Outer Continental Shelf Oil and Gas Decisions: Florida and California, 1989. Oil Spill Risks From Tank Vessel Lightering, 1998. Environmental Information for Outer Continental Shelf Oil and Gas Decisions in Alaska, 1994. Improving the Safety of Marine Pipelines, 1994. Tanker Spills: Prevention by Design, 1991. Double-Hull Tanker Legislation: An Assessment of the Oil Pollution Act of 1990, 1998. Managing Troubled Waters: The Role of Marine Environmental Monitoring, 1990. Using Oil Dispersants on the Sea, Committee on Effectiveness of Oil Dispersants, 1989. Contaminated Sediments in Ports and Waterways: Cleanup Strategies and Technologies, 1997. processes and structural development; and direct toxic effects on developmental and reproductive stages (Capuzzo et al., 1988). Weathering processes are extremely important in altering the toxicity of an oil spill. Neff et al. (2000) demonstrated rapid loss of monocyclic aromatic hydrocarbons (e.g., benzene, toluene, ethylbenzene, and xylene) from evaporation and a reduction of acute toxicity of the water-accommodated fraction (WAF) with loss of these compounds (see Box 5-2). With weathering processes and loss of the monoaromatic compounds, the polycyclic aromatic hydrocarbons become more important contributors to the toxicity of weathered oils. Other factors that may contribute to alterations in toxicity include photodegradation and photoactivation (Garrett et al., 1998; Boese et al., 1999; Mallakin et al., 1999; Little et al., 2000). Barron et al. (1999) examined the chemistry and toxicity of water-accommodated fractions, from three environmentally-weathered middle distillate oils differing in aromatic content to test the hypothesis that the aromatic components of oil are the most toxic fraction. Using short-term growth and survival tests with the mysid, Mysidopsis bahia, they demonstrated that the oil with the lowest aromatic content (expressed as PAH concentration or naphthalene concentration in WAF) had the greatest toxicity. The toxicity of the three weathered oils was consistent with the reported toxicity of unweathered middle distillates tested under similar conditions (Anderson et al., 1974; Markarian et al., 1995) and were more similar to one another when reported as total petroleum hydrocarbons. Therefore, heterocyclic compounds and other soluble components in the water-accommodated fraction of weathered oil may contribute to acute toxicity. The importance of PAH to weathered oil toxicity depends on the concentrations present, presence of other toxic components, and the degree to which the weathered oil has been degraded by microbial and photooxidation. Neff et al. (2000) provided an estimate of the contribution of different hydrocarbon classes to the toxicity of several Australian oils that had been weathered by evaporation in the laboratory (no microbial or photodegradation). Shelton et al. (1999) showed the importance of microbial degradation on weathered crude oil toxicity. Barron and Ka’aihue (2001) argued that photoenhanced toxicity could contribute to the toxicity of crude oil in the field. Although a large volume of literature existed in 1985 on the effects of petroleum hydrocarbons on marine organisms in laboratory studies, the majority of studies conducted prior to 1985 were carried out at concentrations higher than is environmentally realistic. Those studies contributed to our understanding of the range of effects that could occur following an oil spill and the potential for long-term consequences, but they could not be used to develop realistic scenarios of the linkages between recovery of organisms and habitats and the degradation/disappearance of hydrocarbons from the habitat. Much progress has been made since the 1985 report addressing these issues. Some of the best examples of acute and chronic toxic effects of oil to marine organisms have been derived from observations in the field following oil spills and in laboratory studies designed to replicate the exposure field of actual spill conditions.
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Oil in the Sea III: Inputs, Fates, and Effects BOX 5-2 Benzene, Toluene, Ethyl Benzene, and Xylenes (BTEX) BTEX is the collective name for benzene, toluene, ethyl benzene, and xylenes, the volatile aromatic compounds often found in discharges, and petroleum oils and products (Wang and Fingas, 1996). The behavior of the four compounds is somewhat similar when released to the environment and thus they are usually considered as a group. Most light crude oils contain BTEX usually from about 0.5 up to 5% or more. Gasoline can contain up to 40% BTEX. BTEX compounds are volatile and, if discharged into the sea, will rapidly volatilize into the air, and there is, in fact, a net loss of BTEX compounds. Because of this behavior, the discharges of BTEX were not considered in this study. BTEX compounds are acutely toxic to aquatic organisms if contact is maintained. BTEX compounds are relatively soluble in water, the solubility of benzene is about 1400 mg/L and xylenes about 120 mg/L. Because of the volatility of BTEX, the time exposure to aquatic organisms may be short enough to avoid toxic effects. BTEX are generally neurotoxic to target organisms. Benzene, in particular, has also been found to be carcinogenic to mammals and humans. Gasoline contains large amounts of BTEX. The bulk solubility of gasoline has been found to vary from 100 to 500 ppm, depending on the specific type of gasoline and its constituents. The aquatic toxicity of gasoline is relatively high. The fifty-percent lethal concentration to test organisms over a 48-hour period has been found to be 10 to 50 mg/L for Daphnia magna, the water flea, 5 to 15 mg/L for Artemia, small brine shrimp, and 5 to 10 mg/L for rainbow trout larvae. Produced waters contain a variety of volatile hydrocarbons, including the BTEX series (Rabalais et al., 1991a,b). Produced waters generally have concentrations of dissolved salts much higher than sea water and therefore sink through the water column into which they are disposed. BTEX compounds in produced water discharged to well-mixed open ocean waters are diluted rapidly. Twenty meters down-current from a production platform discharging 11 million L/d of produced water containing an average of 6,410 μg/L total BTEX to the Bass Strait off southeast Australia, the average concentration of BTEX was 0.43 μg/L, a dilution of 14,900-fold (Terrens and Tait, 1996). In well-flushed, dispersive and deeper water environments of the Louisiana coast, the BTEX chemical contaminant signal may be negligible as close as 50-100 m from the point of discharge (Rabalais et al. 1991a,b). In shallower, less dispersive environments the produced water plume along with the BTEX spreads in a thin dense plume across the surface sediments of the receiving environment, and the chemical signature of the produced waters can be detected up to 1000 m from the point of discharge (Rabalais et al., 1991a, b). BTEX were detected in the water overlying the sediment surface near estuarine and coastal environments that were categorized as less dispersive or where the concentration of the BTEX was high in the discharge. Produced waters vary considerably in BTEX concentrations, but produced waters discharged into surface waters of Louisiana ranged from 26—4,700 μg/L benzene, 11—1,300 μg/L toluene, 2.1—75 μg/L ethylbenzene, and 8.8—520 μg/L xylenes. BTEX persisted in the density plume that dispersed across the sediment surface in poorly flushed Louisiana study areas in concentrations up to 86 μg/L benzene, 32 μg/L toluene, 2.3 μg/L ethylbenzene, and 17 μg/L xylenes; in more dispersive environments, they were not detected. BTEX in the overlying water column, if present, along with the more persistent polynuclear aromatic hydrocarbons in the sediments, likely contributed to the mortality of the benthic infauna where diminished benthic communities were documented adjacent to produced water discharges. The mortality could not be attributed to high salinity, because the salinity of the interstitial waters of the sediments examined were within the tolerance range of the euryhaline benthos found in the study area. Data gathered from several spills that occurred in the 1970s and 1980s demonstrated that the medium and higher molecular weight aromatic compounds, such as the alkylated phenanthrenes and alkylated dibenzothiophenes, are among the most persistent compounds in both animal tissues and sediments (Capuzzo, 1987). The half-lives of these compounds in marine bivalves following spill conditions can be quite long compared to the relatively rapid decline in monoaromatic compounds and unsubstituted phenanthrenes and naphthalenes (Oudot et al., 1981; Farrington et al., 1982; Anderson et al., 1983; Burns and Yelle-Simmons, 1994). The degree to which the persistence of these compounds in tissues interferes with normal metabolic processes that affect growth, development and reproduction has been the focus of much debate and research. Sublethal effects from hydrocarbon exposure can occur at concentrations several orders of magnitude lower than concentrations that induce acute toxic effects (Vandermeulen and Capuzzo, 1983). Impairment of feeding mechanisms, growth rates, development rates, energetics, reproductive output, recruitment rates and increased susceptibility to disease and other histopathological disorders are some examples of the types of sublethal effects that may occur with exposure to petroleum hydrocarbons (Capuzzo, 1987). Early developmental stages can be especially vulnerable to hydrocarbon exposure, and recruitment failure in chronically contaminated habitats may be related to direct toxic effects of hydrocarbon contaminated sediments (Krebs and Burns, 1977; Cabioch et al., 1980, Sanders et al., 1980; Elmgren et al., 1983). Several studies have demonstrated the potential for oil residuals on beach sediments to have significant toxic effects on fish eggs and embryos. Heintz et al. (1999) reported embryo mortality of pink salmon with laboratory exposure to aqueous total PAH concentrations as low as 1 ppb total PAH derived from artificially weathered Alaska North Slope crude oil. This is consistent with the field observations of Bue et al. (1996) of embryo mortality of pink salmon in streams traversing oiled beaches following the spill from the
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Oil in the Sea III: Inputs, Fates, and Effects Exxon Valdez. Carls et al. (1999) exposed Pacific herring eggs for 16 days to weathered Alaska North Slope crude oil and observed that exposure to initial aqueous concentrations as low as 0.7 ppb PAH caused developmental malformations, genetic damage, mortality, decreased size at hatching, and impaired swimming. Concentrations as low as 0.4 ppb caused premature hatching and yolk-sac edema. Exposure to less weathered oil produced similar results but at higher exposure concentrations (9.1 ppb). Other investigators have observed developmental effects on fish and invertebrates exposed to low concentrations of petroleum hydrocarbons (Capuzzo et al., 1988). The high toxicity of weathered oil reported by Heintz et al. (1999) and Carls et al. (1999), however, suggests that higher concentrations of one or more constituents in weathered fractions relative to total PAH contribute to the increased toxicity. Bioavailability, Bioaccumulation, and Metabolism The concept of bioavailability is extremely important in understanding and describing the environmental fates and biological effects of petroleum in the marine environment. A concise definition of what is meant in this context by bioavailability is essential. In aquatic toxicology, bioavailability usually is defined as the extent to which a chemical can be absorbed or adsorbed by a living organism by active (biological) or passive (physical or chemical) processes. A chemical is said to be bioavailable if it is in a form that can move through or bind to the surface coating (e.g., skin, gill epithelium, gut lining, cell membrane) of an aquatic organism (Kleinow et al., 1999). Accumulation of petroleum hydrocarbons by marine organisms is dependent on the biological availability of hydrocarbons, the length of exposure, and the organism’s capacity for metabolic transformations. There are two aspects of petroleum hydrocarbon bioavailability that are important in understanding the behavior of oil in the environment: environmental availability, and biological availability. Environmental availability is the physical and chemical form of the chemical in the environment and its accessibility to biological receptors. Generally, chemicals in true solution in the ambient water are considered more bioavailable than chemicals in solid or adsorbed forms. Petroleum hydrocarbons of the types found in the marine environment may be present in true solution, complexed with dissolved organic matter and colloids, as dispersed micelles, adsorbed on the surface of inorganic or organic particles, occluded within particles (e.g., in soot, coal, or tar), associated with oil droplets, and in the tissues of marine organisms (Readman et al., 1984; Gschwend and Schwarzenbach, 1992). The hydrocarbons in the different phases are exchangeable but, at any given moment, only a fraction of the total hydrocarbons in water, sediments, and biota is in bioavailable forms. The dissolved hydrocarbons are the most bioavailable, followed by those in tissues of marine organisms (if the organisms are eaten) or associated with liquid, unweathered oil droplets. Thus, bioavailability of PAH from sediments and food is less than that from solution in the water (Pruell et al., 1987). Particulate PAH associated with soot or weathered oil particles (e.g., tarballs) have a low bioavailability (Farrington, 1986; Gustafsson et al., 1997a,b; Baumard et al., 1999). As oil weathers, its viscosity and average molecular weight increase, decreasing the rate of partitioning of higher molecular weight PAH from the oil phase into water in contact with the oil, decreasing the accessibility of these PAH to aquatic organisms (McGrath et al., 2001). Soot-associated PAH are not bioaccumulated in the tissues of aquatic animals. Maruya et al. (1996) showed that sediment-associated animals in San Francisco Bay, CA, were not able to bioaccumulate PAH from the very fine-grained particles (identified as soot) in the sediments. Pruell et al. (1986) showed that the bioaccumulation of PAH from contaminated sediments by mussels correlated with the concentration of dissolved but not particulate PAH in the sediments. The other aspect of environmental availability is accessibility. Petroleum hydrocarbons that are buried deep in sediments or sequestered in solid, highly weathered oil deposits on the shore are not accessible to marine and terrestrial organisms and, therefore have a low bioavailability. Biological availability depends on the rate at which a chemical is assimilated into the tissues of the organism and accumulates at the sites of toxic action in the organism. This depends on the physical/chemical properties of the chemical in contact with the organism, the relative surface area of permeable epithelia in the organism, and the ability of the organism to excrete or detoxify the chemical. Nonpolar (hydrophobic) organic chemicals such as petroleum hydrocarbons, have a low aqueous solubility and a high lipid solubility. Hydrocarbons in solution in water diffuse down an activity or fugacity gradient from the water phase into lipid-rich tissues of marine organisms in contact with the water. According to equilibrium partitioning theory (Davies and Dobbs, 1984; Bierman, 1990), when an aquatic animal is exposed to a nonpolar organic chemical dissolved in the ambient water, the chemical partitions across permeable membranes into tissue lipids until an equilibrium, approximated by the octanol/ water partition coefficient (Kow) for the chemical is reached. At equilibrium, the rates of absorption into and desorption from the lipid phase of the organism are equal. Toxic responses in the organism occur when the concentration of nonpolar organic chemicals in the tissues reach a critical concentration (McCarty and Mackay, 1993). The log Kow of PAH increases with increasing molecular weight (Neff and Burns, 1996). However, bioavailability, measured as log bioconcentration factor (BCF: concentration in tissues/concentration in water at equilibrium), does not increase in a linear fashion with increasing PAH log Kow (Baussant et al., 2001a,b). The sediment organic carbon-water coefficient, Koc is also useful in predicting uptake of sediment-associated hydrocarbons (Fisher, 1995; Meador et al. 1995; DiToro
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Oil in the Sea III: Inputs, Fates, and Effects et al., 2000; ). The higher molecular weight PAH are less bioavailable than predicted by equilibrium partitioning theory because of limitations on their uptake rates by organisms, their lower solubility in tissue lipids, and rapid metabolism of higher molecular weight PAH in some marine animals. Bioaccumulation factors for pyrogenically derived hydrocarbons are much less than predicted based on Koc and suggest that an additional estimate of the fraction of compound available for equilibrium partitioning may be needed (McGroddy and Farrington, 1995; McGroddy et al., 1996). Biotransformation is an important factor in examining tissue burdens and biological effects. An organism’s capacity for biotransformation of hydrocarbons has been used in many instances as an estimate of exposure in the absence of measurable hydrocarbon concentrations. Vertebrates have a high capacity for metabolizing aromatic hydrocarbons including PAH through cytochrome P450 1A mediated oxidation (Stegeman, 1989; Stegeman and Lech, 1991; Spies et al., 1996). Elevation of cytochrome P450 1A levels in fish may indicate exposure to some aromatic hydrocarbons, even though tissue levels do not show elevated concentrations. There is a large literature that links elevated P450 1A levels in fish tissues to aromatic contaminants in marine sediments (e.g., Stegeman and Lech, 1991), but it is theoretically possible for some other natural compounds to induce these enzymes as well. Measurement of hydrocarbon metabolites in tissues where elevated cytochrome P450 1A is observed provides further evidence of the relationship of hydrocarbon exposure, metabolism and cytochrome P450 1A activity (Stein et al., 1992; Collier et al., 1993; Wirgin et al., 1994). Metabolism of hydrocarbon mixtures may result in excretion of some compounds but also activation of some compounds to toxic metabolites including DNA adducts (Wirgin et al., 1994). Long-Term Effects on Benthic Populations Chronic toxicity of petroleum hydrocarbons after an oil spill is associated with the persistent fractions of oil and individual responses of different species to specific compounds. Alterations in bioenergetics and growth of bivalve molluscs following exposure to petroleum hydrocarbons appear to be related to tissue burdens of specific aromatic compounds (Gilfillan et al., 1977; Widdows et al., 1982, 1987; Donkin et al., 1990). Widdows et al. (1982) demonstrated a negative correlation between cellular and physiological stress indices (lysosomal properties and scope for growth) and tissue concentrations of aromatic hydrocarbons with long-term exposure of Mytilus edulis to low concentrations of North Sea crude oil. Recovery of mussels following long- PHOTO 21 Oil penetrated deeply into burrows in the muddy sediments on tidal flats and marshes along the Persian Gulf. Note the liquid oil draining out of a burrow in 1993, two years after the spills. (Photo courtesy of Jacqui Michel, Research Planning, Inc.)
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