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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
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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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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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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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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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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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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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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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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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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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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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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Intertidal Shores
Rocky intertidal shores are quite susceptible to damage by oil spills depending on the amount and characteristics of the oil to which they are exposed. The 1985 Oil in the Sea report adequately characterized the damage of shorelines to spills and stressed the critical role of geomorphology in the recovery of these shorelines. We reiterate here the importance of the interactions of wave and tidal energy with shoreline geomorphology in determining recovery and punctuate this with lessons from more recent studies.
The persistence of the oil and the time to recovery are a function of the energetic fluxes where oil is deposited. If the initial oiling from a spill is an outer, exposed coast, and the rocky substrate is continuous without substantial low energy interstices, then oil will not persist long and recovery will be relative quick (e.g., see Chan, 1977 for an account of recovery on heavily oiled rocky coasts after the San Francisco Bay spill). If the shoreline is relatively sheltered or there are significant interstices where the oil can enter and be sheltered from the energetic fluxes of waves and tides, then oil will persist and recovery may take substantially longer.
The degree of impact and recovery from a spill on the rocky intertidal is very much a function of the circumstances of a spill. Not only is the aforementioned geological structure of the shoreline important, but the type of oil, the weather conditions following the spill, the thickness and lateral continuity of the slick, the time of year, and the recent history of disturbance of the biological communities are all important factors affecting severity of impact. One example of how low energy environments can retain oil and effects can persist is a southern ocean spill at Macquarie Island. In this spill, most intertidal components appear to have recovered within several years after the spill occurred, but in the holdfasts of kelp, which is an environment not unlike mussel beds, oil was retained for years and the fauna of this microhabitat has not recovered (Smith and Simpson, 1995).
By far the greatest acute injury to intertidal communities as a whole arises from direct contact with oil. Heavy deposits of oil essentially smother intertidal organisms. Toxicity also occurs from elevated concentrations of the soluble components of oil in small pools of water, in wetted surfaces and in the water of rising tides. The common organisms found on rocky intertidal shores of North America—Fucus, mussels, periwinkles, starfish, and barnacles—are all susceptible to the toxic effects of oil (Chan, 1977; Stekoll et al., 1993). Recovery of these components can be quite substantial within a year or two, or nearly complete. Subtle long-term effects are possible, however (Peterson, 2001). In the Exxon Valdez spill, the aggressive washing of the intertidal rock shores resulted in loss of a significant amount of silt from the rock interstices and the associated bivalve fauna has not been fully re-established and may not be until these sediments have been replenished by natural processes (Driskell et al., 1996).
The above caveats about the nature of the oil, the thickness and extent of the slick and the weather conditions determining impact also apply to softer substrates. Of particular note is the stranding of oil in protected, low-energy environments, such as bays and harbors. If oil arrives in one of these otherwise low-energy environments under storm conditions and gets worked into the substrate, it will likely be there for years and possibly decades. Two examples are the Florida spill in West Falmouth, Massachusetts in 1969 (Burns and Teal, 1979) and some areas affected by the Amoco Cadiz spill in France in 1978 (Dauvin and Gentil, 1990). It was clear at the time of the Oil in the Sea report (NRC, 1985), that the combinations of circumstances resulting in acute effects can also result in recovery times of years and even decades.
In the last 17 years there has been more focus on chronic contamination by PAH, the sensitivity of meiofauna, and indirect effects mediated by changes in predator-prey relationships, as well as by the direct toxic impacts. In particular, chronic exposure of fauna and potential effects have been studied over more realistic time scales and concentrations. Microcosm experiments where realistic doses of PAH are maintained in sediments to provide a chronic exposure regime have been particularly valuable. For example, in salt marsh sediments in Louisiana concentrations of high molecular PAH (up to 16 ppm) were found to decrease the biomass of epibenthic diatoms and cyanobacteria after 4-day exposures, with some indications that snails from high exposure treatments lost weight after initial gains (Bennett et al., 1999). Such experimental results point to the need to examine more closely estuarine food webs where concentrations of PAH in this range can be found.
The spatial scale of the affected sand or mud shoreline area will determine the rebound of the affected area. A practical example of this is the impact of the Amoco Cadiz oil spill on benthic crustaceans. Failure to recover in some subtidal habitats was due to the fractionated distribution pattern of favored habitat by some species of amphipods (Dauvin and Gentil, 1990). Nevertheless, the populations were able to recover; densities on the impacted site attained high values similar to those found before the spill within 15 years (Dauvin, 1998).
Subtidal Areas
Oil can arrive in the subtidal by two mechanisms. Surface oil can weather, lose buoyancy and eventually sink, and it can associate with particulate matter suspended in the water and eventually sink, thereby affecting the benthic community (Elmgren et al., 1983). A second route of oil to the benthos is the transport of oil or contaminated particles from nearby oiled beaches.
As with the intertidal fauna, the most sensitive organisms in the subtidal benthos appear to be the crustaceans. Major effects on the crustacean fauna were documented in the
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Tsesis spill (Elmgren et al., 1980), the Florida barge spill (Sanders et al., 1980), the Amoco Cadiz spill (Dauvin and Gentil, 1990), the Exxon Valdez spill (Jewett et al., 1999), and the 1996 North Cape oil spill where 8 to 9 million American lobsters were killed subtidally from a fuel oil spill (McCay, 2001). In addition, the rhepoxiniid amphipods, which appear to be particularly susceptible, are one of the few severely depressed faunal components in the benthic communities in areas of moderate petroleum seepage in the Santa Barbara Channel (Davis and Spies, 1980).
Not all spills demonstrated adverse effects in subtidal habitats. A study of the possible effects of tar residues from the Haven oil spill in Italy revealed no discernable differences between tar-affected and non-affected benthic communities (Guidetti et al., 2000). Exxon Valdez oil was generally not discernable below 40 meters in most portions of Prince William Sound and was never found in measurable quantities below 100 m depth. It is not surprising then that a study of deep benthic communities found no differences between various areas that could be attributed to oil from the spill (Feder and Feder, 1998).
During the Braer spill off the Shetland Islands, 84,700 tonnes of a light Gullfaks crude oil were released from the grounded vessel during hurricane-force winds, and an estimated 35 percent of the oil was deposited on the seabed in water depths from 2-100 m in an area of 4,000 km2 (Kingston, 1999). The sedimented oil provided a long-term pathway for exposure to benthic fisheries. For example, burrowing Norway lobster (Nephrops) remained contaminated for over five years, whereas epibenthic lobsters (Homarus) eliminated petroleum contaminants to background levels of PAH in one month (Kingston, 1999).
Effects Associated with Various Sources
As discussed in Chapter 3, petroleum enters the marine environment from a variety of sources, at different rates, and in diverse settings. Understanding how the environment responds to releases associated with specific sources is an important aspect of understanding the overall impact of widespread extraction, handling, and use of petroleum hydrocarbons.
Lessons from Natural Seeps
Natural petroleum seeps occur in many parts of the ocean, and can be utilized to understand the effects of oil contamination (Spies et al., 1980). As petroleum enters the ocean from the seabed, it is relatively unweathered in comparison to many other sources of oil that reaches the bottom (Reed and Kaplan, 1977; Steurmer et al., 1982). There are some significant consequences to this difference that limit the usefulness of oil seeps as effects models for other sources of oil in which weathering occurs before the oil is deposited in bottom sediments. Also, the possibility must be kept in mind that, with a history of thousands of years, animals living near seeps might have unique adaptations. Biological studies of seeps have concentrated on the extensively contaminated benthos (Spies and Davis, 1979; Spies et al., 1990).
There are two aspects to the effects of fresh seeping petroleum on benthic ecosystems. First, fresh petroleum, being a highly reduced source of energy, is readily oxidized by microbes (Bauer et al., 1988), which, in turn, can serve as a supplementary food source for benthic food webs in shallow water (Spies and DesMarais, 1983; Bauer et al, 1990). In the case of seeps in deep water, it can be a nearly exclusive carbon source. Second, at sufficiently high concentrations, the aromatic components of seep petroleum are toxic to marine organisms (Davis et al., 1981). There is also an interaction between toxicity of oil and microbial metabolism of petroleum. The decrease in oxygen in the surface layers of the sediments that results from microbial metabolism of petroleum is a limiting factor to benthic organisms. The oil, while supporting microbial growth that acts as a food source, may also be toxic to other organisms or indirectly decrease habitat quality through oxygen deficiency (Spies et al., 1989; Steichen et al., 1996). Microbial transformations of aromatic hydrocarbons may alter hydrocarbon composition and various oxidized products may be formed (Bartha and Atlas, 1987). Natural biogeochemical tracers indicate that both the petroleum carbon, particularly the lighter fractions, and sulfur from sulfide is incorporated into benthic meiofauna and macrofauna (Spies and DesMarais, 1983; Bauer et al., 1990). Circumstantial evidence for damage to gill tissues in bottom-feeding surf perches are linked to oil exposure through cytochrome P450 1A induction and aromatic petroleum metabolites in bile (Spies et al., 1996).
The most detailed investigations of petroleum seepages have been carried out in the Santa Barbara Channel off the coast of southern California. The following summarizes the findings of studies conducted at a depth of 20m in one of these oil seep areas, the Isla Vista seep. Starting with the fresh oil and gas in the sediments of a petroleum seep, several related phenomena occur. Bacterial populations, as measured by ATP content or by direct microscopic counts, are elevated several fold over surrounding sediments (Spies et al., 1980; Bauer et al., 1988). The sediments are highly reducing, oxygen is undetectable in sediments below a very thin surface layer, sulfate oxidizing activity is markedly elevated, hydrogen sulfide is abundant, and sulfide-oxidizing bacteria (Beggiatoa) are abundant at the surface of sediments, often forming prominent white mats.
The heavy seepage areas where the Beggiatoa mats form support a low-diversity benthic community consisting of large numbers of nematodes, a few polychaete worms (e.g., Capitella capitata), some oligochaete worms, and a limited number of harpacticoid copepod species (Spies et al., 1980; Montagna et al., 1987, 1989, 1995). Porewater concentrations of aromatic hydrocarbons within a few centimeters of an active seep were approximately 1 ppm. The nematodes
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form a halo of high abundance, often only a few centimeters from the most active sources of seepage. Within several meters of the very active seeps, and where a small amount of seepage is still found, a diverse benthic community occurs of mainly detrital feeders. This community is similar in composition to the surrounding community that occupies much of the inner continental shelf in southern California. There are some key differences: oligochaetes are a significant component of the community, and some rhepoxiniid amphipods that are particularly sensitive to oil are missing but are found outside the seepage area (Davis and Spies, 1980). Also, in comparison to a nearby station, this community has a consistently larger number of organisms per unit area of sandy bottom.
Based on these findings the following is a conceptual model of the shallow water seep system in the Santa Barbara Channel. The lighter fractions of seeping petroleum are metabolized by microbes, and the energy in the petroleum is partially converted into microbial biomass. Rapid utilization of oxygen by microbial petroleum oxidizers shifts sediment biochemistry to sulfate oxidation and the resulting product, hydrogen sulfide, is utilized by sulfide oxidizers (Beggiatoa). At some distance from areas of active seepage, where pore water aromatic hydrocarbon and sulfide concentrations drop to tolerable levels and oxygen can penetrate further into the surface layer of the sediment, the diversity of organisms increases. Bacteria that assimilate and oxidize petroleum are utilized by meio- and macrofauna as an additional source of energy to supplement the usual contemporary photosynthetic sources of carbon for nearshore benthic communities. This pattern within the benthic community is broadly similar to that described for other sources of organic enrichment in the ocean, e.g., sewage and paper mill wastes (Pearson and Rosenberg, 1978).
In very deep water hydrocarbon seeps, most of the carbon utilized by mussels can originate from petroleum, because little surface water primary production is available. Various microbial chemosynthetic endosymbionts may be present in association with macrofauna of deepwater seeps. These microbes can oxidize either hydrogen sulfide or methane. Methane oxidizers then become carbon sources for their hosts. Methane oxidation has been little investigated in shallow water benthic systems, but could be an important process.
Since the shallow water seeps in southern California occupy only a small fraction of the continental shelf and seep benthic communities are composed almost predominantly of species with wide dispersal mechanisms (particularly in the larval stage), the studies of these “open” communities leave unresolved the issue of multi-generational effects of oil exposure.
Production Fields as Tutoring Grounds
Production fields offer an opportunity for uncovering long-term, chronic effects of petroleum hydrocarbons in the marine environment. Production fields are common in U.S. estuaries, primarily in the northwestern Gulf of Mexico, but are also increasing globally. In the United States, offshore oil and gas production has been limited primarily to continental shelf waters but continues to progress into deeper waters seaward of the shelf break. Regulations regarding the use of oil-based drilling fluids differ globally, as do regulations concerning the discharge of produced waters within U.S. state and federal waters and globally, so that practices that would contribute to the long-term accumulation of contaminants differ regionally as well as with time as regulations change. Still, practices that are currently disallowed in many coastal environments may continue to be allowed in other areas, and as oil and gas production expands globally into developing countries permitting regulations may allow practices there that are currently disallowed in U.S. waters. Thus, lessons learned from various practices, even if currently disallowed, provide the tutoring ground for further production globally.
Because of the continued inputs of drilling fluids and produced water discharges into estuarine, coastal and marine receiving waters, production fields are likely to illustrate long-term, low-level, chronic biological effects in sedimentary habitats resulting from the persistence of metals and medium and high molecular weight aromatic hydrocarbons, heterocyclics and their degradation products. Chronic contamination may result from continuous or intermittent discharges (produced waters, drilling fluids, deck washings) or from repetitive, accidental spills (numerous small spills and/ or a small number of major spills during the life of a field). The effects of these effluents are complicated by the addition of drilling discharges accumulated during field development (as opposed to exploratory drilling).
Drilling Fluids and Produced Waters
The major discharges associated with exploratory and development drilling are drilling cuttings and drilling fluids. Drill cuttings are particles of crushed sedimentary rock produced by the action of the drill bit as it penetrates the formation. Their accumulation on the ocean floor alters the benthic sedimentary environment. Drilling fluids are mixtures of natural clays and/or polymers, weighting agents and other materials suspended in a water or oil-based material. Oil-based drilling fluids have never been permitted for discharge to U.S. state or federal waters. Discharges of oil-based mud cuttings were permitted to waters of Canada, the North Sea countries, Australia, and several other offshore regions in the world until recently. Their ocean discharge has been banned in most of the world. Several metals of environmental concern found in drilling fluids are arsenic, barium, chromium, cadmium, copper, iron, lead, mercury, nickel, and zinc.
During production of oil or gas from a platform, produced water from the formation may be discharged to the environment. Besides being more saline than sea water, produced
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waters contain elevated concentrations of radionuclides, metals, volatile organic aromatic compounds, monoaromatic hydrocarbons, light alkanes, higher molecular weight aromatic hydrocarbons, ketones, phenols, and organic acids. The environmental effects that may result from oil and gas production in a field depend greatly on the characteristics of the receiving environment. For example, there was a decreased abundance of fouling organisms, particularly barnacles, from the surface to a depth of about 3 m on a platform leg immediately below the produced water discharge located 1 m above the water surface (Howard et al., 1980). Produced water discharges, however, are usually dispersed to some degree. If discharged into the ocean, the produced water dilutes rapidly so that no impacts are ascribed to salinity. In more confined estuarine waters, produced water discharges form dense, saline plumes that move along the bottom sediments, but the resulting elevated water column and interstitial sediment salinity levels are within the range of tolerance of euryhaline estuarine organisms. In shallow, more confined areas with high suspended sediment loads or fine-grained sediments, medium molecular weight hydrocarbons and metals can absorb to particles and be deposited. Measurable effects are most likely in shallow waters, areas of restricted flow and dispersion, water with a high concentration of suspended particulates, and areas of fine-grained anaerobic sediments.
Effects of Production Discharges in Estuarine Waters
U.S. regulations now prohibit most discharges of produced waters from platforms to state waters of Texas and Louisiana, although the phase-out is not yet complete and some exceptions are provided, for example the highly dispersive distributaries of the Mississippi and Atchafalaya Rivers. The discharge of treated produced water from several offshore platforms at shore-based facilities is still permitted in Upper Cook Inlet, Alaska. The discharge of produced waters into estuaries and shallow coastal waters continues globally in developing fields (e.g., Nigeria, Angola, China, Thailand), and the effects of produced water discharges may still linger where the practice has been discontinued (Rabalais et al., 1998).
The effects of produced water discharges in estuaries have been studied extensively in Texas and Louisiana. For example Mackin (1971) surveyed estuarine benthic communities in eight Texas bays receiving produced water effluents. He reported no effects in two bays, minor localized effects in several other bays, and a zone of severely depressed fauna up to 106 m from submarine outfalls in Trinity Bay, Texas and a zone of enhanced faunal abundance and diversity down-current from there. Mackin (1971), however, conducted no chemical analyses. Armstrong et al. (1979) repeated these studies in Trinity Bay to correlate the benthic community effects with the distribution of hydrocarbons in sediments. In shallow waters of 2-3 m, they demonstrated the impacts of high concentrations of hydrocarbons, in this case sediment naphthalene concentrations of 4 to 8 ppm up to 1200 m from the platform, with corresponding severely depressed benthic fauna.
It was not until the mid to late 1980s that more extensive, systematic studies of the effects of produced water discharges in estuarine waters were conducted. While most surface water disposal was terminated on January 1, 2000, except within the distributary channels of Louisiana’s major rivers (the Mississippi and Atchafalaya), it is prudent to review the results of these studies for several reasons. First, this disposal method was practiced in coastal Louisiana and Texas and at one time accounted for 2,500,000 bbl/d of discharges into estuarine waters with the potential in some areas for long-term accumulation of contaminants and subsequent reintroduction to the environment (Boesch and Rabalais, 1989b; Rabalais et al., 1991a,b; Rabalais et al., 1998). Second, accidents associated with current disposal methods (pipelines and barges) will have similar results. Third, surface water disposal in estuarine waters still occurs elsewhere in the world.
Boesch and Rabalais (1989a), Neff et al. (1989), St. Pé (1990), Rabalais et al. (1991a,b), Steimle & Associates, Inc. (1991), Mulino et al. (1996) studied the effects of produced water discharges in estuarine waters of Louisiana. Where suitable measurements were made, the eventual fate of the dispersed produced water and the effects on benthic infauna could be explained by the volume of the discharge, the concentration of the various constituents, and the sedimentary regime, physical structure, and hydrology of the receiving environment (Boesch and Rabalais, 1989; Rabalais et al., 1991a). Dilution of water-soluble contaminants was influenced primarily by the volume of the receiving waters, the current velocity, and the potential for resuspension of sediments. Dispersion of sediment-adsorbed contaminants was influenced by the bed shear stress, sedimentation rates, and the grain size distribution of the surface sediments. The dilution potential of the environment was high for erosional environments with high current speeds and low for depositional environments, with intermediate potential for environments with periodic resuspension (storm-related) and deposition. There were no documented effects on the benthic community due to elevated salinities, because the overlying water and sediment interstitial salinities were within the range of the euryhaline organisms found in these habitats. Volatile hydrocarbons in the water column density plume that disperses across the sediment bed varied from nil to as high as 130 mg/L; alkylated PAH in bottom sediments reached concentrations from 2 to 40 ppm with one value of 100 ppm (Rabalais et al., 1991a,b). The potential for accumulation to depth in depositional environments exists (some sediments contained 30 ppm alkylated PAH at 35 cm depth) (Rabalais et al., 1991a). Produced water source contaminants persisted in surface sediments for two years after cessation of the effluent, as did benthic community effects, and persisted for as
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long as nine years in vertical sediment cores (Rabalais et al., 1998).
The effects of produced water discharges in estuarine systems include toxicity to various organisms and, at the community level, the reduction of infaunal abundance and diversity (reviewed by St. Pé, 1990; Rabalais et al., 1991a). The persistent elevation of sediment hydrocarbon and metal concentrations and modification of benthic communities occurred from within a few hundred meters to up to a kilometer from the discharge.
In the Lake Barre field, Louisiana (one out of five such studies), oysters placed in trays adjacent to a produced water discharge suffered mortality as far as 23 m from the outfall and showed decreased growth rates between 23 and 46 m from the outfall (Menzel, 1950; Menzel and Hopkins, 1951, 1953). Similarly deployed oysters near produced water discharges in Pass Fourchon and Bayou Rigaud, Louisiana, resulted in mortality to oysters, reduced growth in others, and bioaccumulation of alkylated PAH and total hydrocarbons 3 to 18 times above background level (Rabalais et al., 1992). The potential for oysters to take up and accumulate contaminants originating in produced water occurred both in close proximity to the discharge and to great distances (350 m at Bayou Rigaud and 1000 m at Pass Fourchon).
Production Effects in Continental Shelf Waters
Several production field studies have been conducted in U.S. offshore waters, in the Norwegian and British sectors of the North Sea field, and on the Dutch continental shelf. [Continental shelf is defined from subtidal barrier shoreface to shelfbreak, usually 200 m water depth.] Because continental shelf ecosystems are complex, open and dynamic, there are fundamental problems in identifying the nature and extent of environmental effects and in determining causality. Any studies conducted in production fields must be designed and interpreted within the context of natural variability and other environmental factors that are important in shaping the physical environment and biological communities. In addition, there is a range of biological, chemical and statistical techniques that can be applied to any of these studies, often suitably, but just as often not. The complicating factors of natural variability have plagued many studies of oil and gas production effects in the northern Gulf of Mexico where one would be most likely to expect to find demonstrable effects on marine ecosystems. Two studies—the Offshore Ecology Investigation (OEI) and the Central Gulf Platform Study—were conducted in coastal Louisiana where extensive oil production activity might make this area a worse case scenario for effects. The locations just west of the Mississippi River delta, however, confounded possible oil-related effects with ecological factors of salinity variability, turbidity, high organic loading, sediment type, and seasonal bottom-water hypoxia (low dissolved oxygen concentrations).
Following 25 years of oil production in both Timbalier Bay and the area directly offshore on the continental shelf in 5 to 25 m water depth, the OEI study (reviewed by Neff, 1987; Spies, 1987) was designed to examine both localized platform effects, and the overall “health” of the ecosystems in the study area. The conclusions of the study (Menzies et al., 1979) were that there were no effects and that (1) natural phenomena in the area were a greater impact than petroleum-related activities, (2) petroleum contamination in the area was low and could not be tied to platform sources, and (3) the region was in “good ecological health.” The OEI study was criticized by Sanders (1981) who pointed out faults in study design, laboratory procedures, insufficient contaminant data, and inappropriate statistical treatment of benthic data. A second group of scientists independent of the OEI effort concluded that chemical contaminant data were insufficient to draw conclusions and that the benthic data indicated that the communities were more likely controlled by estuarine- and riverine-influenced salinity and turbidity than by adverse effects of oil (Bender et al., 1979).
The Central Gulf Platform study (water depths of 9 to 98 m; reviewed by Neff, 1987 and Spies, 1987) corrected many of the shortfalls of the OEI study by employing a larger variety of analyses, but was again situated in an area influenced by the Mississippi River, with confounding effects of turbidity, fluctuating salinity, seasonal hypoxia, and potential additional anthropogenic inputs of petroleum (Bedinger et al., 1981). In addition, a major tropical storm likely affected the benthic communities mid-way through sampling. Meiofaunal and macrofaunal benthic data analyzed with clustering techniques identified groups of fauna that were most similar with regard to depth, salinity, distance from shore (= sedimentary characteristics) and dissolved oxygen. Chemical analyses revealed both low and high molecular weight hydrocarbons often in very high concentrations, but there were no consistent patterns with regard to the amount of production from the platform, the age of the platform, or distance from the platform. Conclusions that the Mississippi River was the probable principal source of hydrocarbons to the study area were not supported by the data. The implications in the study conclusions that hydrocarbons were having a chronic sublethal effect on the fauna of the study area were criticized by Spies (1987), because of extrapolation of laboratory toxicity literature to concentrations of hydrocarbons found in the field. That is, most of the toxic effects documented in the literature were from relatively low molecular weight aromatic hydrocarbons, and the sediment hydrocarbons in the Central Gulf Platform study were dominated by highly-weathered mixtures.
Rabalais et al. (1993) tested specifically for the differences in hypoxia versus effects of petroleum production at two production platforms in 20 m water depth. Hydrocarbon concentrations, in general, were low for both study sites (even where sediments were silty) and were characterized as weathered petrogenic or biogenic in nature. There were no
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consistent patterns of benthic community structure with distance from either discharge nor were there any relationships with petroleum indicators. As found in other studies, dissolved oxygen concentration, bottom water temperature and salinity, and sediments were the important environmental factors that explained the variation in benthic community parameters of species richness and abundance.
In a shallower area of the continental shelf (2 m) just offshore of the lower end of Atchafalaya Bay where uniformly silty sediments dominated and the environment was expected to be dispersive, a clear signal of produced water-associated contaminants and effects on benthic biota were observed to at least 200 m and 300 m, respectively (Rabalais et al., 1991a). The shallow water column at this site and flushing potential of Atchafalaya River discharge were expected to dilute and transport the produced water effluents away from the area. The high silt content of the sediments, the large volume of produced water discharged (21,000 bbl/d), and the high concentrations of volatile hydrocarbons may have been factors in the pronounced produced-water-effect at this station. By contrast, a nearby station in 8-m water depth, where the discharge of produced water was an order of magnitude less, was contaminated only within 20 m of the discharge where benthic fauna were also impacted (Neff et al., 1989).
Several production platforms in southern California were assessed for oil and metal contamination and affected marine communities (18 to 30 m water depth; reviewed by Neff, 1987). There were some elevated levels of production contaminants in sediments directly under and adjacent to platforms, but no concentrations of metals and petroleum hydrocarbons in selected fish and mussels. The platforms, piles of cuttings, and biofouling organisms both on the platforms and those sloughed to the bottom functioned as artificial reefs, providing habitats for a wider variety of marine animals than occurred on nearby hard and soft bottoms. In a study of a produced water discharge outfall in a high energy subtidal (10-12 m) environment off southern California, Osenberg et al. (1992) indicated that benthic infaunal community effects were localized within 100 m of the outfall.
In offshore waters around production platforms in the Gulf of Mexico, there was little evidence of bioaccumulation of produced water contaminants in edible tissues of resident fishes and invertebrates (Continental Shelf Associates, Inc., 1997). For the southern California produced water outfall in a high energy subtidal zone, Osenberg et al. (1992) found that the effects on outplanted mussels were more widespread than on the benthic infauna, between 500 and 1000 m from the outfall, as opposed to within 100 m. The observed effects on the mussels were reduction in growth, condition, and tissue production and varied inversely with relative exposure of the mussels to the produced water plume.
Two major studies have been conducted on the Texas continental shelf to examine the ecological effects of chronic contamination as well as sublethal impacts. These studies were conducted away from the influence of the Mississippi River and focused on near-field effects with a closer link between biological and chemical analyses. The Buccaneer Gas and Oil Field study (20 m depth; Middleditch, 1981) documented persistent accumulation of sediment hydrocarbons only within about 100 m of the platforms, but did not provide a thorough analysis of the chemical constituents present. A widespread effect on the benthos, including reduced numbers of individuals and species around the platforms, was apparent, but there were also areas well away from the platforms with similar benthic communities.
The Gulf of Mexico Offshore Operations Monitoring Experiment (GOOMEX) was designed to test and evaluate a range of biological, biochemical and chemical methodologies to detect and assess chronic sublethal biological impacts in the vicinity of long-duration activities associated with hydrocarbon production (Kennicutt et al., 1996b). The study was located in a gas field in the western Gulf of Mexico continental shelf and as removed as possible from confounding effects of Mississippi River discharge. The three platforms were in progressively deeper water, 29, 80 and 125 m. Sediments close (< 100 m) to the three platforms studied were enhanced in coarse-grain materials primarily derived from discharged muds and cuttings. Hydrocarbon and trace metal (Ag, Ba, Cd, Hg, Pb, and Zn) contaminants were associated with these coarse-grain sediments (Kennicutt et al., 1996a). Contaminants were asymmetrically distributed around each platform in response to the prevailing currents. The positive relationship between sand content and contaminant levels is contrary to the view of contaminants being associated with finer-grain sediments (Peterson et al., 1996).
The hydrocarbons occurred in concentrations that seemed too low to be important contributors to the observed toxicological effects. PAH were generally less than 100 ng/g, which was an order of magnitude lower than what Spies (1987) suggested was needed to induce biological response. At a few locations close to one platform, trace metal (i.e., Cd, Hg, Pb, and Zn) concentrations exceeded levels thought to induce biological effects. In deeper water (> 80 m), sediment trace metal contaminant patterns were stable over time frames of years. A few metals (Pb, Cd) exhibited evidence of continued accumulation in sediments over the history of the platform at the deeper water sites (> 80 m) immediately after cessation of drilling cf. 5-10 years after the last discharges. The chemical contaminants principally originated from the original drilling mud discharge and perhaps from produced waters during production (Kennicutt et al., 1996b).
Sediment chemical analyses and porewater toxicity tests with sea urchin fertilization and embryological development assays from the GOOMEX study (Carr et al., 1996) indicated toxicity near four of the five platforms, the majority collected within 150 m of a platform and those with the highest concentrations of contaminants. There was agreement among results of porewater tests with three species (sea urchin embryological development, polychaete reproduction,
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and harpacticoid nauplii survival). Samples from the deepest site (> 80 m, HI-A389 near the Flower Gardens), which contained the highest contaminant concentrations, were the most toxic samples of the sites. Repeatability of toxicity between seasons demonstrated the persistence of the toxicity.
The meiofauna and macrofauna effects (Montagna and Harper, 1996) were localized within 100-200 m from the platforms (Table 5-7). The patterns of community change were increases in deposit-feeding polychaetes and nematodes that indicated organic enrichment, while density declines of harpacticoid copepods and amphipods indicated toxicity. The increase in annelids closer to the platforms occurred despite the steep gradient in sand content; total anne
TABLE 5-7 Responses of Biological and Ecological Indicators to Distance from Platforms in the Gulf of Mexico (compiled by P. Montagna)
Platform
Indicator
MAI-686 29 m depth
MU-A85 80 m depth
HI-A389 125 m depth
Macroinfauna
Total Abundance
0
↑70 percent
↑3-5×
Polychaete (P) Abundance
0
↑90 percent
↑3-5×
Amphipod (A) Abundance
↓4-12×
↓4-5×
↓3-5×
P:A ratio
↑6-10×
↑8-10×
↑20-30×
Meiofauna
Total Abundance
↓2-3×
0
↑50 percent
Nematode (N) Abundance
↓2×
↓2×
↑60-80 percent
Harpacticoid (H) Abundance
↓2×
0
↓3×
N:H ratio
0
0
↑4-5×
Nematode Production
↓49×
↓135×
↓2-3×
H. Gravid Females
↑3×
0
↑3×
H. Clutch Size
0
↓10 percent
↓14 percent
H. Genetic Diversity
↓2×
↓1-5×
↓2-3×
Megafauna
Catch Per Unit Effort
0
0
↑
Size
0
↓
↑
Histopathology
0
↑
0
Sex Ratio
0
0
↑Female
Toxicity & Biomarkers
Pore Water Toxicity
↑
0
↑
Invertebrate AHH
0
0
0
Fish: EROD, PAH in Bile, P4501A mRNA
0
0
0
Summary of results for macrofauna and meiofauna (Montagna and Harper, 1996), nematode production (Montagna and Li, 1997), genetic diversity (Street and Montagna, 1996), harpacticoids reproduction (Montagna, unpublished data), megafauna (Ellis et al., 1996), toxicity (Carr et al., 1996) and biomarkers (McDonald et al., 1996). Table symbols represent percent increase (↑), no change (0), or decrease (↓) from near-field (< 100 m) to far-field stations (100 m to 3 km). 1x = 100 percent.
lids would be expected to be more abundant in finer sediments, not coarser. In contrast with annelids and oligochaetes, amphipod abundances were depressed around all platforms, with effects confined to 50 to 100 m. This was also consistent with literature on modest pollution, and is suggestive of a toxic response. Sea stars were reduced near the platform, but that pattern did not hold for ophiuroids. Changes in meiofaunal responses were most noticeable within 50 m of platforms. Harpacticoid abundance, community diversity, genetic diversity, reproductive success and survivability declined nearer the platforms with an increasing contaminant gradient at all study sites. On the other hand, total nematodes were enhanced. Patterns were absent at the shallowest site (29 m, MAI-686) where the relatively high-energy physical environment has led to more extensive dispersion of materials discharged. The other sites were in 80 m (MU-A85) and 125 m (HI-A389). They concluded that patterns of response to sedimentary contamination were detectable at higher taxonomic levels, and that these responses were driven by intrinsic physiological and ecological characteristics of higher taxa. Crustaceans (especially amphipods and harpacticoid copepods) and echinoderms are sensitive to toxics whereas polychaetes, oligochaetes, and nematodes (especially nonselective deposit feeders) are enhanced by organic enrichment (either from hydrocarbons or biologically produced materials falling from the platform structure). They concluded that metals drove the toxicity effects, and that the dual effects of toxicity and organic enrichment resulted in readily detectable responses in benthic meiofauna and macrofauna to 100-200 m.
The GOOMEX studies also focused on chronic, sublethal effects. Various physiological (McDonald et al., 1996) and genetic results (Montagna and Harper, 1996) provided evidence that crustaceans around the platforms were exhibiting sublethal responses to contaminant exposure. The percentage of gravid female harpacticoid copepods was greater and the percentage of juveniles was reduced within 50m of the platforms. In addition, reproductive effort for female harpacticoids carrying eggs was reduced. These responses could be explained as sublethal physiological reactions of these organisms to stress related to exposure to toxicants. The demonstration in multiple species of harpacticoids that genetic diversity was significantly reduced near the platforms and associated with increased contaminants as compared with the far-field sites suggested detection of a sublethal response of these sensitive organisms to some aspect of the platform-associated environment.
Several studies have been conducted in the North Sea (reviewed by Neff, 1987; Spies, 1987). The Ekofisk Oilfield study (70 m water depth) was designed to assess the impacts of ballast water discharge from a one-million barrel oil storage tank placed on the sea bottom, as well as impacts from production platforms. The Ekofisk study was complicated by a well blowout, and it is not known whether drilling employed oil-based fluids. Results indicated elevated concen
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trations of oil constituents around some platforms at Ekofisk. Some faunal change around the platforms could be related to elevated hydrocarbon content, but were more readily explained by the changed sedimentary conditions and associated total organic carbon content surrounding the platforms (Spies, 1987). The Forties Oilfield, developed using only water-based drilling muds, is situated in 100 to 125 m water depth and represents deeper conditions than most studies on the Louisiana and Texas continental shelves. Impacts of oil production activities there were localized, primarily within 450 m of the platforms, and were of low magnitude. Only water-based drilling fluids were used in the Buchan field, and no produced water was discharged. No biological impacts of the platform or production activities were detected.
Oil-based drilling fluids were used extensively in the North Sea, and the amounts of petroleum hydrocarbons discharged with drill cuttings and their subsequent accumulation have been documented by several authors, (reviewed by Davies et al., 1983; summarized by Neff, 1987). There are four zones of chemical and biological impact around platforms discharging contaminated cuttings. Zone I, extending out to 250 m and exceptionally to 500 m from the platform, is characterized by hydrocarbon concentrations 1000 times above background and severely impoverished and modified benthos. Zone II extends from 200 to 2000 m from the platform with sediment concentrations 10-700 times above background, modified species diversity, and increased abundance of opportunistic polychaetes. Zones III and IV have normal benthic communities and decreasing gradients of hydrocarbon contamination. At the time of the Davies et al. (1983) review, none of the areas had been studied long enough to determine benthic recovery.
In contrast to the implication of metals over the long-term being the agent of benthic effects in a production field offshore Texas, U.S., the implicated contaminants where oil-based drilling fluids were used over many decades in oil and gas production in the North Sea are the hydrocarbons in those drilling fluids (Grant and Briggs, 2002). In toxicity studies of sediments from around the North West Hutton platform in the North Sea, sediments from 600 m from the platform remained acutely toxic to the amphipod Corophium and acutely toxic to the same organism at 100 m from the platform when 3% contaminated sediment was mixed with clean sediment. Metals toxicity was only a factor immediately adjacent to the platform.
Production Fields
There are clear effects of produced water discharges on waters, sediments, and living resources in estuarine production fields where the receiving environment is not conducive to the dispersion of the effluent plume. In shallow shelf waters, hydrocarbons from produced water accumulate in bottom sediments and benthic fauna may be depressed up to 300 m from the outfall. Measurable effects occur around offshore platforms, but except for artificial reef effects, sedimentary changes or changes brought about by a cuttings pile, such effects are usually localized. Beyond some contamination of organisms by petroleum, there is little convincing evidence of significant effects from petroleum around offshore platforms in deeper water. Where oil-based drill cuttings are discharged, there are readily evident effects of sediment contamination and benthic impacts to much greater distances from the platforms (up to 1 to 2 km).
While directed studies have identified some specific sublethal effects of long-term oil and gas development, the most significant unanswered questions remain those regarding the effects on ecosystems of chronic long-term, low-level exposures resulting from discharges and spills caused by development activities. Ultimately, we must determine whether the potential for effects from production fields are significant with regard to the geographic scale, what the cumulative effects are, whether ecosystem integrity is compromised, and whether there are significant impacts to resources that humans value, such as fisheries, marine mammals, endangered species, or rare or aesthetically pleasing environments.
Deep Sea Communities
Unfortunately, our knowledge base for the effects of chemicals or habitat perturbation is the most meager for the deep sea. It is unexpected, however, that ecological processes in the deep sea are fundamentally different from those of the continental shelf. There are additional environmental and physical parameters at work in the deep sea that make populations and communities there unique. What is not known are the sensitivities to contaminants; rates and mechanisms for population control, biological interactions, and recruitment; or rates or potential for recovery from impact.
Unique features of the deep sea and the fauna make them more susceptible to certain types of chemical spills. Increased turbidity from a spilled chemical such as a drilling mud could impact animals adapted to low light (including possibly, those with bioluminescent capabilities) by increased turbidity from deep plumes of low transmission water and indirectly through biological light interactions. Spills of chemicals with labile carbon may alter the local balance of oxygen consumption and result in hypoxia or anoxia, especially in oxygen minimum zones. Microhabitat diversity is a key to deep-sea diversity, and any chemical spill that alters deep habitats will likely have an impact. Chemical spills that disrupt the accessibility of fluxed detrital material for the dominant deposit-feeding organisms will affect feeding and subsequently the health of the organism(s). Chemicals that affect mortality, population levels, biological interactions, recruitment, growth rates, through either acute or chronic, sublethal toxicity or habitat alteration or both are likely to affect soft-bottom benthos in the deep sea similarly to continental shelf organisms. Basic biological information for most deep-sea organisms (e.g., feeding type, reproduc
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tion, life span, growth rates, predators, and community ecology), however, is nonexistent.
Chemosynthetic seep communities are considered prevalent between 300 and 1000 m water depth on the northern Gulf of Mexico slope. Commonality, however, is not a reason for relaxing criteria for acceptable impact without knowledge of the ability of undamaged or damaged fauna to ultimately repopulate any impacted areas. Some organisms that inhabit the cold seep communities may be extremely old, and damaged communities would be slow or unlikely to recover. Hard bottom communities with highly diverse biogenically-structured communities are afforded protection from drilling operations in the Gulf of Mexico, and any chemical spills that approximate these types of effects would be expected to produce similar harm to live-bottom communities.
Summary
Since the compilation of the 1985 NRC report, Oil in the Sea, great progress has been made in identifying the toxic effects of petroleum hydrocarbons in a wide variety of organisms. We have also gained considerable knowledge of the effects of oil on various marine habitats through laboratory experiments, mesocosm experiments and practical experience with spills. Our knowledge of the effects of produced waters has expanded for inshore and offshore production fields and for multiple mixtures of oil and other contaminants in confined water bodies such as harbors. We now have first-hand experience with spills in coral reefs, mangroves, seagrass beds, and high-latitude cold-water environments. We are now in a better position to assess risks to individual organisms and habitats from the production, transport and consumption of petroleum than we were in 1985.
Assessing the effects of any particular spill and recovery from its effects has proven more complicated than was anticipated in 1985. We know that the natural variability of marine ecosystems and the open nature of marine communities, in which recruitment of young may be dependent on planktonic larvae transported from great distances, creates a substantial challenge in assessing both the effects of a spill and recovery from those effects. Although we now know much more about the toxicity and sublethal effects of petroleum hydrocarbons to organisms, we still have great difficulty in assessing the population, community, or ecosystem effects of pollution events. To assess the effects of oil in the sea and recovery from impacts, we need new information on the population structure of these marine organisms that is critical for the function of their communities. In addition, appreciation of the influence of decadal-scale and longer climate change means that we cannot expect communities or ecosystems to return to the state in which they were at the time of a pollution incident. Given the various time-scales of ecosystem change, before-after and control-impacted (BACI) designs for assessing damage are valuable, but they are no substitute for an up-to-date time series from a well-designed monitoring program. Based on an improved understanding of change in the marine environment, there is great value in having time series for detecting change and for pointing to processes critical for understanding change.
The effects of oil in the sea depend greatly on the season, place, and the types of organisms present. Although for a given habitat at a given time, a large amount of oil is likely to create more damage than a small amount, small amounts in sensitive environments or where there are populations at risk can have devastating effects.
Reducing the Threat to the Marine Environment
Ecosystems and their components vary at time-scales from seasons to decades and longer. Therefore, in the absence of on-going monitoring it is exceedingly difficult to quantify the effects of oil in the sea, or to establish when recovery from a pollution event is complete. Establishment of monitoring programs in selected regions with an elevated risk of petroleum spills or discharges would enhance the ability to determine effects and recovery, and to understand the processes controlling ecosystem responses to pollution. Existing databases on the distribution, frequency and size of petroleum spills and existing petroleum and distribution routes could be used to identify locations most appropriate for monitoring. Federal agencies, especially the USGS and EPA, should work with state and local authorities to establish or expand efforts to monitor vulnerable components of ecosystems likely to be exposed to petroleum releases.
There are demonstrable effects of acute oiling events at both small and large spatial-scales. These effects result from physical fouling of organisms and physiological responses to the toxic components of oil. Although there is now considerable information on the toxicological effects of individual components of oil, there is a lack of information about the synergistic interactions within organisms between hydrocarbons and other classes of pollutants. This problem is particularly acute in areas subject to chronic pollution, e.g., urban runoff. Research on the cumulative effects of multiple types of hydrocarbons in combination with other types of pollutants is needed to assess toxicity and organism response under conditions experienced by organisms in polluted coastal zones. Federal agencies, especially the USGS, NOAA, and EPA, should work with industry to develop or expand research efforts to understand the cumulative effects of multiple types of hydrocarbons in combination with other types of pollutants on marine organisms. Furthermore, such research efforts should also address the fates and effects of those fractions that are known or suspected to be toxic in geographic regions where their rate of input is high.
There are demonstrable sublethal physiological effects of long-term, chronic releases of hydrocarbons into the marine environment. These have been found in areas affected by
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urban runoff, in areas where oil has been incorporated in sediments and is then released back to the water column, and in production fields. Chronic sources of hydrocarbon pollution remain a concern, and their effects on populations and ecosystems need further assessment. Federal agencies, especially the USGS, EPA, and NOAA should work with state and local authorities and industry to implement a comprehensive laboratory and field based investigation of the impact of chronic releases of petroleum hydrocarbons.
Biogenically-structured habitats, such as salt marshes and mangrove forests, are subject to destruction or alteration by acute oiling events. Because the structure of these habitats depends upon living organisms, when these are killed, the structure of the habitat, and sometimes the substrate on which it grows, is lost. Depending upon the severity of oiling and particularly if oil is incorporated in the sediments or structure of the habitat, recovery of the habitat and the organisms dependent on it may be exceptionally slow. In areas of sensitive environments or at-risk organisms, federal, state, and local entities responsible for contingency plans should develop mechanisms for higher level of prevention, such as avoidance, improved vessel tracking systems, escort tugs, and technology for tanker safety.
Although there is now good evidence for the toxic effects of oil pollution on individual organisms and on the species composition of communities, there is little information on the effects of either acute or chronic oil pollution on populations or on the function of communities or ecosystems. The lack of understanding of population-level effects lies partly in the fact that the structure of populations of most marine organisms is poorly known because of the open nature of communities and the flow of recruits between regions. Also, in some populations, (e.g., bony fish), the relationships between numbers of juveniles produced and recruitment to the spawning adult population are unknown. The U.S. Departments of Interior and Commerce should identify an agency, or combination of agencies, to develop priorities for continued research on:
the structure of populations of marine organisms and the spatial extent of the regions from which recruitment occurs,
the potential for cascades of effects when local populations of organisms that are key in structuring a community are removed by oiling, and
the basic population biology of marine organisms may lead to breakthroughs in understanding the relationship between sublethal effects, individual mortality, and population consequences.
There is a tremendous need for timely dissemination of information across state, federal, and international boundaries about the environmental effects of oil in the sea. Although the United States has experience that might benefit the international community, the United States might benefit greatly from lessons learned in other countries with offshore oil production, heavy transportation usage, and diffuse inputs of petroleum from land- and air-based sources. Therefore, the federal agencies identified above, in collaboration with similar international institutions, should develop mechanisms to facilitate the transfer of information and experience.
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