Oil in the Sea: Inputs, Fates, and Effects (1985)

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s Effects INTRODUCTION A vast amount of data and literature has accumulated s ince the 1975 NRC report. Much of this accumulation has been brought together at confer- ences (e.g ., Amer loan Petroleum Institute, 1975, 1977, 1979, 1981; Wolfe, 1977), at specialized symposia (American Institute of Biological Sciences, 1976, 1978; Fisheries Research Board, 1978; Cowan et al., 1981), and in two major reviews (Malins, 1977; Sprague et al. , 1981~ . This has come about in response to increased funding since the ear ly 1970s, partly due to increasing human concerns over oil in the marine environment, and partly as a natural outcome of continued spills and accidental discharges. One interesting and encouraging development has been a noticeable change in research emphasis, from descriptive to more process-or tented research, as in studies of physiological impact and ecological change (Table 5-1~. During the early days of oil pollution research following the Torrey Canyon accident, most research was aimed at quantifying toxicity thresholds. At the same time there was little scientific consistency, in that researchers developed their own exposure methodology and analytical preferences. AS a result, intercomparison of laboratory data was difficult. This began to change in the mid-1970s with a redirection of research interest toward understanding the mechanisms of hydrocarbon toxicity and the sites of toxic action. This effort was paralleled by concerted efforts of var. ious workers to standardize analytical methods, using certain reference oils set aside by the Amer ican Petroleum Institute (API). As a result, more meaning and comparability have come into the f ield of toxic e f feats of petroleum, and a type of data is being produced with which many members of the scientific community can agree upon (Rice et al., 19771. This change in emphasis in recent years represents a significant advancement since the 1975 NRC report. The f ield of oil pollution impact presents unusual and ma jor difficulties to the researcher in that at virtually every turn of study new techniques and analytical and sampling methods have to be devised. This is due to the newness of this research area, having come into its own only since 1967 with the breakup of the Torrey Canyon. Prior to that event most of the research interest with respect to petroleum concerned its physicochemical aspects, and the analytical methods arid 369

370 TABLE 5-1 emphasis of Oil-Pollution-Related Study Reports for the Temperate and Nor thern Mar ine Environment Between 1967 and 1977 Bnphas is Pre-1974 Post-1974 1. Oil--physical-chemical changes, fate and distr ibution, env ironmental concentr ations 2 . Gross biological ef feats: mortality, toxicity Physiological, developmental, and ecological change Microbiology: hydrocarbon- utiliz ing banter ia 35% (83) 33% (107) 4396 (100) 31% (99) 6% (15) 2296 (70) 16% (37) 14% (44) NOTE: Numbers in parentheses denote number of studies. SOURCE: Environmental Protection Agency (1977 ~ . expertise reflected these interests. However, with the Torrey Canyon a new scientific discipline was required, including an understanding of the behavior of oil in water, sediment, and even in tissues, and requir- ing analytical methods capable of resolving petrogenic compounds in unfamiliar environmental samples . This has called for a much mor e interdisciplinary approach, with constant and new exchange of expertise __A .A~_~ Ion_ "~ ha: `:£~_~. A.~.~1 ;~e. ~ ~ , ~ _, _ _ _ _ I -I I ~ ~- - - ~ - ~ - ~ -~~ - =- -..-- ~ Sc ientis ts in th is f ield are often competent in several areas, combining organic chemistry with biochemistry or biology, and often a more than passing acquaintance with microbiology or geology. Environmental teams of scientists have developed, working with integrated effort. This is not to say that the problem of understanding oil pollution in the mar ine environment is now well in hand . Ma jor inroads have been made in analytical capabilities and in understanding petroleum effects on two levels--physiological and ecological. However, generally there is a good appreciation of oil effects in temperate and northern temperate waters. At the same time the area of subcellular effects has received less attention. Ecological studies have been done pr imar fly in the field, using spills of opportunity and, more recently, large f ield enclosures (mesocosms) with known dosages of oil. At spills of opportunity, studies have been done mainly in the soft sediment areas such as salt marshes and shallow embayments. Such areas have been documented as sites where spilled oil will persist for long periods (years to

371 decades) (see Chapter 3) and have provided the basis for most of what is known about how oil affects on mar ine populations and communities . The mesocosm studies have provided exper imental evidence of similar perturbations occurring in planktonic and benthic communities. Investigation into the physiological (e.g., photosynthesis, respiration, growth, neurotransmission) effects of oil is largely through laboratory investigations, although some fundamental work has been done in the field. One reason for the laboratory emphasis is the need to control experimental conditions. The coverage of this level of investigation has been uneven. AS in mammalian physiology, there are preferred invertebrate and algal species. Certain bivalves (Mya arenaria, Mytilus sp., Macoma spp.) and crustaceans (Cancer spp., Uca pugnax' Crampon spp., Penaeus aztecus) J by virtue of their accessibil- ity and ease of culture, are far easier to work with than benthic or pelagic organisms available only seasonally and/or by dredging or trawling. The same thing holds for the marine algae. It seems prefer- able, however, to attempt to understand in good detail the toxicology of petroleum in two or three well-studied representative organisms, rather than to attempt to establish simple toxic tolerance levels across all the phyla. (Although it should be noted that some of these may not be the most vulnerable.) Study at the subcellular level has received far less attention, despite recent concerns over certain hydrocarbons interacting with cellular macromolecules such as nucleic acids. We discuss primarily the impact of petroleum hydrocarbons, leaving the possible impact of other contaminants or compounds--either contained in oil or in some manner by-products of petroleum- or gas-related activities--to other more specialized discussions. Such contaminants would include, for example, trace metals, chemical dispersants, and drilling muds. Again, the available literature on petroleum impact alone is so vast that to include detailed discussions on these other materials, beyond merely mentioning them, would inevitably lead to an unmanageable exercise. The approach taken in this chapter is to discuss and review the impact of petroleum on marine biota and communities, by proceeding from one level ho the next--from effects on processes (cellular), through a discussion of effects on the marine foodchain (organismic), to the effects on communities (ecosystem). Inevitably this leads to some repetition or duplication, but this approach makes the most sense in unravelling and describing an extremely complex problem, involving a complex pollutant and the complexities of marine life. We hope that the index to this report will aid the reader in finding his or her way through it. Inevitably in a task such as this, some studies and reports will not have been referenced in the writing of this chapter because of the limited space available. Throughout we tried to refer to those studies that illustrated a particular point or aspect of petroleum pollution most aptly or most concisely. In other instances we referenced those studies that would lead the interested reader in turn to other studies. In an appendix to this chapter we have included a discussion of some well-known oil spills and oil seep problems, largely to add some

372 real dimensions to the at times very detailed discussions in the main body of the chapter. The examples were selected for their general appropriateness and because they represented in each instance a particular spill type under certain conditions. Finally, any report tries to be up to date in its coverage, but inevitably there has to be a cut-off date. For various reasons, editor ial and technical, the review process for this chapter was very lengthy. We have tried to maintain as current a reference list as possible, but have not been able to go much beyond 1982-1983. We regret therefore having missed much excellent new literature and newer findings published in the last year. Toxicity In its most general sense, toxicity can be defined as the imparting of a deleterious effect, whether lethal or sublethal, to an organism, population, or community. The toxic effect can result in a permanent perturbation or change, for example, the crooked-back syndrome in larval fish, stunted growth, deformed shell formation in mollusks, and changed population patterns. However, not all effects are disruptive, and there exist adaptive mechanisms, both at the cellular level (e.g., detoxifying enzyme systems) and at the population and community levels (Capuzzo, 1981~. Toxic effects from petroleum exposure vary widely and for reasons that are not well understood. Certainly these have to do with the complexity of its chemical composition, with different products or even crude oils differing markedly in their chemical makeup. Another factor is the var lability in sensitivity to oil found among mar ine organisms, differing not only with the species (Figure 5-1) but even for life- cycle stages (Figure 5-21. While it is generally true that the younger stages of organisms are more sensitive to petroleum hydrocarbons, there are exceptions. Unfortunately, not many studies have compared the sensitivity of organisms at various life stages under identical experi- mental conditions for any one species. While the absolute toxicity of petroleum hydrocarbons appears to be greater for the higher-molecular-weight compounds (for example, 3- and 4-ring aromatics), most of the toxic effect of petroleum in water is thought to be due to the lower-molecular-weight (C12-C24) e-paraffin compounds and to the monoaromatic fraction, for the simple reason that these compounds are the most water-soluble (Chapter 3, Table 3-4~. From examinations of the concentrations present in water-soluble fractions (WSF), it is clear that the contribution of compounds higher in molecular weight than the alkyloaphthalenes is very small and may be insignif icant in terms of producing acute toxicity. Bioassay tests have been used to a considerable extent to determine the toxicities of var ious crude oils and of refined products. Most of these tests have used mortality as the index of toxicity, expressed for example, as LC5g (the lethal concentration yielding 50% mortality over pre-determ~ned exposure time, for example, 24, 48, or 96 hours). In practice, however, their usefulness as a research technique is

PR] 26 24 22 20 18 ~ 16 IS 14 t`' 1 2 8 4 3 3 373 3 n I Crago 2 Neanthes 3 Cancer 4 Salmon fry 5 Cyprinodon 6 Copepod 7 Palaemonetes 8 Amphipod  9 Striped bass 10 Penaeus aztecus 11 Grass shrimp · 24 hour LC50 5 ~ ~ -I - - ~ Jo ; ~ '' '''aft " ·~1~111151 ' Ill~llllt~ly ~J ~ ~ 7~ _~ n ,~e, ~ 0~ ~ ~ ~ `~ AQUA O`~ Opt O`= 6~ O'er ~ O~ 10 ~ ~ it_ ~~s Cal FIGURE5-1 Acute toxicity (24- and 96-hour LC50 static tests ~ of some aromatic hydrocarbons for selected mar ine macroinver tebrates and fish. SOURCES: Caldwell et al. (1977), Benville and Korn (1977), Neff et al. (1976), R.E. Thomas and Rice (1979), Young (1977), Rossi and Neff (1978), Ott et al. (1978), and W.Y. Lee and Nicol (1978).

374 A; ~ I: , ~ ~ ~ At. FIGURE S-2 Effects of oiling on f ine structure of surf smelt embryo retinas. (Left) Retinal cells of an unoiled embryo (x5000~. Inset is enlargement of a synaptic junction X18,000. {Right) Retina' cells of an embryo exposed to 113 ppb Cook Inlet crude oil (xSOOO). vesiculation is evident in the myoid regions (asterisk) of the receptor cells. Note also necrotic neurons (arrows). Synaptic junction (viz., inset, x18 '000) appears normal. Ellipsoid region (e) of inner segment of receptor cell' nucleus {m) of receptor cell' outer segment (o) of receptor cell, and synaptic (s) junctional complex are indicated on figure. (Photo by J. Hawkes.) limited in that they provide no data except on mortality. Toxicity tests are subject to several variables such as complex mixture of the oil, test parameters, and various biological factors such as age, sex, and contamination history of the organism. For these reasons they are somewhat imprecise measures of toxicity, and many researchers feel that their output, the LC50, has little relevance to what may happen to an organism as a result of a spill. By its very nature the LC50 gives no indication of sublethal toxic problems that the organism may be experiencing, and gives no measure of any long term impacts that may be occurring, measuring only the death of the organism. Instead, acute lethal bioassays serve best as tests to compare the relative toxicities of complex, unknown toxicants, or the comparison of relative sensitivities of species or life stages.

375 Laboratory Versus Field Studies Most studies of effects have been developed in the laboratory for the simple reason that field studies often depend on spills of opportunity, which are highly unpredictable . ALSO, f ield studies are often expen- sive and difficult to carry out. On the other hand, laboratory studies have frequently been criticized because experimental conditions do not simulate field conditions. Also, concentrations of oil or hydrocarbons frequently exceed those encountered in the field. Both criticisms are probably correct. However, in recent years attempts have been made to bring laboratory conditions closer to field conditions by simulating the hydrocarbon composition and concentrations more closely through the use of flow-through systems and by careful management of the test organisms chosen. Most promising in this respect have been various studies carried out in "mesocosms," large enclosures that allow control of some environmental variables under near-open ocean conditions (for example, Marine Ecosystems Research Laboratory (MERL), CEPEX, Loch Ewe, viz., Table 5-~. In this respect, natural oil seeps also offer certain opportunities to study the impact of petroleum under more open-ocean conditions. As for the criticism of high experimental dosages, there are situations where seemingly high concentrations of oil or hydrocarbon are warranted, for example, in the initial establishment or detection of certain toxic effects and in the analysis of metabolic pathways. High initial concentrations of toxicants are frequently necessary to establish a toxic effect that otherwise might be indistinguishable from the background "noises or masked by other changes. Using high concen- trations allows the experimenter to better define the toxic effect or response. Again, the detection and identification of primary or secondary metabolites, or of short-lived intermediates, often requir e unusually large doses of toxicant. In these instances the object is not so much to determine a toxic effect, as it is to better understand certain aspects of hydrocarbon metabolism for which low dosages would not elicit a measurable response. However, care has to be taken in , work ing with h igh dosages and in interpreting results because of the possibility that extraordinary metabolic pathways may be expressed. In the end, both laboratory and field study have merit. Although field studies are fraught with uncontrolled and interfering factors, they nonetheless are the ultimate testing ground. For that reason, spill sites should be visited, and revisited, whenever possible. On the other hand, laboratory studies support field studies by providing the opportunity to investigate an effect in detail and to study its under- lying mechanism. Factors Affecting Impact of Oil When an oil spill occurs, many factors determine whether that spill will cause heavy, long lasting biological damage; comparatively little or no damage; or some intermediate degree of damage. An example of the variability that exists among the effects of oil spills on the mar ine

376 biota is outlined by C.T. Mitchell et al. (1970) in their description of the widely different effects resulting from the Tampico Maru and the Santa Barbara oil spills. Nonetheless, there are some patterns begin- ning to emerge that are useful in identifying those physical and biological features that can influence the ultimate impact of a spill or chronic pollution (e.g. , Michael et al. , 1978; Gundlach and Hayes, 1978; Owens and Robilliard, 1981; Vandermeulen, 1977, 1982) . Geographic Location In many reports, organisms from any one geographic location are apparently no different from any other location in terms of their vulnerability to petroleum hydrocarbons. While there are of course species-specific genetic differences, arctic fish or invertebrates do not appear to differ physiologically from similar organisms at lower or tropical latitudes in terms of lethal toxic concentration thresholds or toxic r esponses . However, there are a number of physical features related to geo- graphic location--mainly temperature and ice cover, together with differences in community diversity, which is latitude dependent--that will influence both impact and biolog i Cal recovery. Temperature, for example, plays a significant role in the solubility of hydrocarbons in the water column and in the rate of their degradation through microbial activity. Similarly, community diversity at different latitudes (low diversity in polar regions, high diversity in tropical environments) can lead to differences in both times and patterns of biological recovery following an oil impact. Oil Dosage and Impact Area If the spill occurs in a small, confined area so that the oil is unable to escape, damage will be greater, almost without exception, for a given volume and type of oil spilled than if that same volume were released in a relatively open area. For example, at the Arrow spill site in Chedabucto Bay, Nova Scotia, about 2.5 million gallons of Bunker C fuel were spilled in an embayment, whereas the Argo Merchant spilled about 7.7 million gallons of No. 6 fuel oil into the open ocean of f Nantucket Island, Massachusetts. Although a considerable amount of Arrow spill eventually was swept out to sea, the confined nature of the oil during the first days resulted in nearly uniform oiling of the entire bay coastline and in considerable damage to the associated fauna and flora. However, this generalization is not inflexible. The spill of the super- tanker Amoco Cadiz occurred offshore, but prevailing winds were such that the oil was kept near shore of North Brittany for several weeks and continuously driven onto shore. Similarly, there are differences resulting from the manner of the spillage--whether low level but chronic or consisting of a sudden accidental release. The former is covered in greater detail elsewhere (for example, see Chronic Oiling section), but

377 generally the impact from such chronic releases differs both in sever ity and in k ind from accidental spills, where the spilled oil will eventu- ally disappear with time due to physical-chemical processes and micro- b ial and other biological degradation . On the other hand, in the case of chronic releases the spilled oil becomes a continuing irr itant or toxicant to which the community must adjust, for example, in selection for hydrocarbon-utiliz ing species . Oceanograph ic Conditions Currents, sea state, coastal topography, and tidal action all combine to influence the impact of a given spill. Currents and wave action--in open water or open bays--act to break up the oil into smaller slicks, and also act to disperse some of the oil into the water column. In areas of large tidal ranges the oil can become distr ibuted over a broad range of the intertidal zone, arid can be deposited far above the high tide mark by extreme "spr ing" tides coinciding with high winds and strong tidal flow . Coastal topography plays a large role in the residual impact of a spill, with low energy environments (salt marshes, lagoons, estuaries, embayments) acting as long term hydrocarbon "sinks." Impact on biota in such systems is usually long lasting. Meteorological Conditions Normally, storms increase wave action and wind speed and thereby aid in evaporation of the lower-molecular-weight, more volatile toxic compo- nents . On occasion, however, wave action may intensify the problems, as apparently occurred at ache Flor Ida No. 2 fuel oil spill near West Falmouth. Soon after this spill, the surf drove the oil ashore into the sediments and the surrounding marshland (Sanders, 1978) . The oiled marshland and sediments then became a long term reservoir of oil with persistence in some areas to this day. A1SO storm-induced resuspension of subtidal sediments probably brought these sediments into contact with oil from more intertidal areas. Similar events occurred following the breakup of the Amoco Cadiz, where the winter storms drove the oil deep inland up the nearby estuar ine tidal r ivers (Hess, 1978 ~ . Season Season is particularly important in terms of the biota that might lie in the path of a spill or in the vicinity of a chronic oiling s~tua- tion. For example, if a spill occurs in an area where seabirds are feeding or nesting, bird mortality might be in the thousands; at some other time of the year the mortality might be much lower. Similarly the coincidence of a spill with f ish spawning events or hatching and development of larval f ish migration might result in higher than normal larval mar tall ties .

378 Seasonal changes in the phys decal parameter s of the mar ine environ- ment might also influence the potential impact of a spill, as for example, seasonally timed changes in local circulation patterns that might lead to local containment of slicks. Oil Type Oil type determines both the short term and the long term impact. Immediate impact can be very high from such highly toxic oils as diesel and jet fuel. However, these dissipate readily and leave relatively little residue, unlike the crude and Bunker oils, which can persist in certain sediments for up to several decades. This aspect is not as simple and clear cut as it seems, however. Traces of No. 2 fuel oil, a relatively volatile product, still persist in sediments of Falmouth, Massachusetts, 13 years after the spill of the Flor Ida . Oil Metabolites and Photochemical Reaction Products This subject deserves separate mention as it was not raised in any detail in the 1975 NRC report, but has become of interest in more recent years, following observations that some petroleum metabolites or intermediate products can be quite toxic and may even have mutagenic proper ties . Oil metabolites can be formed from the parent oil by biological conversion of compounds taken up by marine biota (including bacteria), and by photochemical processes. The relatively few data available on either method come mainly from laboratory studies. The formation of compounds by photochemical processes has been addressed earlier (Chapter 4 ~ . In general, irradiated samples of petroleum or water-soluble preparations appear to be more toxic than the parent compounds. For example, Lacaze and Villedon de Naide (1976) cite field studies sug- gesting that the irradiated water-soluble fractions (WSF) of Kuwait crude oil were 3 times as toxic, depressing C-fixation, as nonirradiated WSF after 40- and 64-hour exposure of the alga Phaeodactylum cornutum. In similar studies, Scheier and Gominger (1976 ~ examined the toxic effects of irradiated versus nonirradiated No. 2 fuel oil, using a Sylvania sunlamp, and compared the results with solar-irradiated WSF. They observed that (1) sunlight was nearly 10 times more effective than sunlamp exposure in raising the toxicity of the irradiated WSF, as indicated by the anthracene-dianthracene conversion ratio, and (2) both significantly increased the toxicity of the WSF due to the irradiation. There is 1 ittle or no information on the potential toxicity of bio, ogical metabolites of petroleum compounds, and any conclusion is difficult, for metabolites have been demonstrated in only a few instances (e .g ., Corner and Harr is , 1976 ; Sanborn and Mal ins , 1977, 1980; Varanasi and Gmur, 19801. There is no evidence to date that the bulk of the petroleum hydrocarbon metabolites formed by biological activity are any more toxic than their parent compounds e However, a small proportion of petroleum compounds do give rise to mutagenic

379 intermediates and to metabolites capable of binding with nucleic acids (Varanas~ and Gmur, 1980; Varanasi et al., 1980, 1982~. This potential appears to be limited only to the polycyclic aromatic hydrocarbons. The evidence to date is sparse but does not indicate that a mutation load has been introduced generally into the mar ine environment by this mechanism as a direct result of petroleum spillage or chronic spillage. However, this possibility cannot be ruled out in isolated incidents. Remedial Measures A great deal of effort continues to be expended, on countermeasures and various cleanup and control methods. These generally fall into one of two categories--mechanical and chemical--and because of their nature they inevitably leave some traces on the landscape, be it some form of physical disruption following mechanical cleanup or the risk of chemical alteration following application of chemical methods. As these are an almost automatic response to oiling incidents, a brief discussion of their potential effects on the marine environment seems appropriate. Mechanical Containment and Cleanup This category includes those methods which focus on the actual removal of oil or oiled debris, as by bulldozing or hosing with water under pressure. Most of this activity involves the intertidal zone. Offshore oiling incidents rarely are suitable for mechanical cleanup except by surface skimmers or possibly the cropping of oiled kelp using mechanical aquatic weed cutters. Neither of these is very likely to have much of an adverse effect on the environment. However, the problem becomes more serious in the intertidal zone, largely due to the physical disrup- tion of habitats. Rocky coastlines present the least problem in terms of cleanup. Oiled rocky surfaces are cleaned most often with either flushing, steam cleaning, sand blasting, or manual scraping. None of these is likely to alter the substrate to any extent, and the main damage is the removal of fauna and flora. The biological recovery process may take several years, but nevertheless, recovery will occur. As the settling surfaces have probably not been chemically or physically altered in the cleanup process to any great extent, the only limitations to recovery are biological ones. Of course, the rerelease of the stranded and flushed oil into the water column may pose additional problems. The problem becomes greater with the oiling of finer-grained sediments such as cobble-boulder beaches or the fine silt sediments of lagoons and marshes. Because of the penetration of oil into such sediments, removal of oiled sediments often accompanies cleanup. Excessive removal can result in the disturbance of physical and ecological equilibrium. Excessive removal of beach sediments can lead to beach retreat or beck shore (cliff) erosion. This was observed following the Arrow disaster, where a 20-m landward movement of pebble- cobble beach was recorded following large scale removal of oiled cobble

380 from the shorelines (Owens and Drapeau, 1973~. Pebble-cobble beaches present a particular problem in that oil is likely to penetrate rapidly and deeply. Thus, if cleanup is advised, it necessarily involves large scale removal of the beach material with heavy machinery. Pebble, cobble, and boulder sediments usually are replaced only slowly by natural coastal processes (Owens, 1973~. Another physical consequence of excessive sediment removal is that of habitat alteration, which in turn leads to long term ecological perturbation. An extreme example is the Ile Grande salt marsh in North Brittany, France, which was cleaned following the 1978 Amoco Cadiz oiling (Hess, 1978; Long and Vandermeulen, 1979, in press; Vandermeulen et al., 19811. Large volumes of the marsh sediment were removed along with oiled vegetation. The marsh surface and stream bed were severely disturbed by movement of heavy machinery. Dredging of the adjacent marsh area for a marina aggravated the conditions. The marsh has since undergone a series of degradative stages, including deposition of sand over the finer marsh silts, undercutting of secondary and tertiary marsh tidal channels, and erosion of the marsh surfaces, all as a direct result of the increased water flow through the marsh following bulldoz- ing. This sediment removal has resulted in net erosion of the shore- line, which ranges between 6.5 and 17 m/year in some parts of this marsh. Marsh recovery, in this extreme case, is thought to run into decades if not centuries. Cleanup of lagoons and coastal marsh systems (salt marsh, mangrove swamp) presents problems in that both the physical (habitat) and the biological aspects of these systems are easily damaged. In such areas, natural cleaning may frequently be the most effective, appropriate, and least damaging option (e.g., Baker, 1975; Lindstedt-Siva, 1979, 1980, 1981; Tramper et al., 1981), although cosmetically it is the least appealing one. Some progress is now being made in studies of replant- ing marsh vegetation and mangroves in oiled and cleaned sediments, and this line of investigation appears useful for extreme cases such as the Ile Grande oiling. A problem with mechanical cleanup is the ultimate disposal of oiled debris, including oiled seaweed, sediment, and shoreline material. These often can be in volume or mass far greater than that of the stranded oil itself. An understanding of both the physical and the ecological environ- ment is necessary for the development of efficient, effective, and least damaging countermeasure techniques. Considerable progress has been made in this area in recent years, mostly through combined input from ecologists, geologists, coastal sedimentologists, and physical oceanographers. Chemical Control and Cleanup The use of chemicals for the control and cleanup of oil spills appears to be an alternative to mechanical removal of the oil, especially in certain situations such as in offshore spills and in polar regions. However, the use of these chemicals continues to be a subject of

381 concern and debate and has generated much research (for recent review, viz . Wells, 1984) . A range of chemical means has been tr fed on oil slicks, with varying results. These include sinkants, gelling agents, herders, chemical dispersants, and deemulsifiers. Of these the last two have seen the most widespread use. Sinkants were used widely during the early days of oil pollution, for example, dur ing the Torrey Canyon spill . Herders remain largely in an experimental stage. Deemulsifiers are now used extensively in petroleum production systems but are also being studied with an eye on f ield application on "mousse ~ slicks and for transfer of collected oil to storage systems. Except for sinkants, use of these products in the field has been limited. The use ~ ~ ~ ~ of chemical dispersants as a spill response tool remains controversial. The formulations, essentially mixtures of chemical surfactants and stabilizers in a carrier solvent, are specifically designed to reduce the interracial tension between oil and water and thus result in a break up of the oil slick into smaller droplets. These droplets can then be distr ibuted into the water column by the natural actions of surface and subsurface turbulence. Much of the success of conventional dispersants depends on the parallel application of physical dispersive energy, such as by the use of high speed "breaker boards, n which increase surface turbulence. The more recently developed "concentrate" dispersants do not require additional dispersive energy in any but the calmest of waters. Concentrate dispersants have a "self- mix" action when appl fed to the oil because they con ta in h igh concentr a- tions of surfactant molecules per unit. Chemical dispersants first gained international prominence following their use at the Torrey Canyon spill. Much of their notoriety stems from the high toxicity of early formulations as well as improper appli- cation techniques used at the spill site. In fact, the high acute toxicity of those products was due pr imar i' y to the car r ier solvents (aromatic hydrocarbons ~ and not the dispersant sur factants themselves . Considerable advances have been made in refining their chemical composi- tion and design and in understanding their potential toxicity (Table 5-2 ~ . Today ' s formulations ar e developed for speci f ic purposes and consist of compositions that retain their dispersant effectiveness but r equire less mixing energy and are at reduced toxicity. Wile the newer formulas may be inherently less toxic, their increased effectiveness results in greater amounts of oil being put into the water column and thereby becoming available for contact with the pelagic and benthic communities (Swedmark et al., 1973; Doe and Wells, 1978~. However, there exist no field data from recent years to - suggest that damage from dispersants has been greater than if the oi had been left alone. Thus, no evidence of damage due to the use of dispersants has come from U.K. coastal waters, where the use of chemi- cal dispersants is common in oil slick control. More studies are needed to look at this aspect of oil pollution. To be fair, it should be pointed out that the simultaneous monitor ing of offshore pelagic and benthic communities for significant effects is extremely difficult. One argument is that use of chemical dispersants simply represents the introduction of yet another contaminant into the mar ine environment.

382 TABLE 5-2 Acute Lethal Toxicity of Some Oil Spill Dispersants to Marine Organisms--A Selection of Current Data Threshold Concentrations Expressed as Four-Day Species/Stage Dispersant LC50 's, mg/1.! . Inver tebrates Stony coral (Madracis mirabilis) Shell dispersant LTX°162 (1 day) Oligochaete (Marionina subterranea) Corexit 7664 Finasol OSR-2>100 0 Finasol OSR-5 Intertidal limpet (Patella vulgata) BPllOOX3700 (approx. ) BP1100WD 270 (appr ox. ) Crustaceans Amphipods (Gammarus spp. ) Mysids (Neomysis sp. ) Amph ipod (Ganunarus ocean icus ) Brown shr imp (Crangon crangon ) Grass shr imp (Palaemonetes pugio) Fish Fish larvae (Pleuronectes platessa, Solea solea) Gobies (Chasmichthys, Luciogobius) Stickleback (Gasterosteus aculeatus ) Dace (Phoxinus phoxinus) Coho salmon tOncorhynchus k isutch) .Killif ish (adult) (Eundulus heteroclitus water-based dispersants >10000 petroleum-based dispersants 200 ~ 130 wa ter -based d isper san ts > 4 50 0 pe troleum-based d isper sants -150 AP oil dispersant 10-100 ( 1.5 days ) 10 conventional disper sants 3300->10000 ( 2 days ) 7 concentrated dispersants 2800->10000 (2 days) ( unnamed ) Corexi ~ 7664 Atlantic-Pacif ic Gold Crew Nokomis-3 >104 (27°C), nontoxic (17°C) 1000 (27°C), 1800 (17°C) 150 (27°C), 380 (17°C) 140 (27°C), 250 (17°C) Corexit 7664 400 Shell dispersant LT 440-480 wa ter -based d isper san ts 9 5 0 + 2 5 0 petroleum-based dispersants >10000 water-based dispersants 1400 + 200 BPllOOX 1700 AP oil dispersant approx . 100 ( 2 days), (GEC Chemical Co. ) 50-100 ( 3 days ) aUnless otherwise noted.  SOURCE: From Wells (1984 ) . This argument is answered, in part, by the design of less toxic and more biodegradable dispersants . From the available evidence there would seem to be little point in the application of chemical dispersants in far offshore spills except to control the onshore movement of slicks, to protect critical seabird populations and other such marine biota aggregations, and to prevent oil foulings of ships and pleasure craft transiting the area. Their use may well be more attractive in inshore and polar waters. Appro- priate application of chemical dispersants in the former could avert wholesale oiling of the intertidal region, as seen with the Amoco Cadiz spill. It is difficult to assess, however, the potential impact had that volume of oil been mixed into the water column instead. Chemical d ispersants are being considered ser iously for use in polar spills, many of which would be expected to be inaccessible by traditional

383 countermeasures. Their effectiveness and potential toxicity, both alone and ~ n combination with oil, are now under study. Indirect Effects Not all impact, whether from a spill or from chronic oiling, is due directly to a specific toxic effect of the oil. For example (see Behavior section), mortality of bivalves during oil spillage often is due directly to suffocation or toxicity. However mortality may also result out of the habit of bivalves in oiled sediments to live closer to the surface, thereby becoming easier prey for seabirds and other predators. Following the Santa Barbara oil spill, it was observed that initial settlement by larvae of the intertidal barnacle Chthamalus fissus was heaviest on solidly oiled black surfaces (Straughan, 1976~. However, subsequent survival of larvae was lowest on such surfaces, apparently because of the heat-absorbing capacity of the black sub- strate dur ing exposed tidal per iods. Tarred sediments forming e ither layers or near-permanent pavements have been observed after the Arrow (Owens , 1976), the Metula (Gundlach , 19791 . and the Amoco Cadiz (Vandermeolen et al. . 1979: d'Ozouville et al., 1981) spills. Just how these oily or tarry layers attect the physical environment is not clear, but studies of an exper imentally oiled sandy beach suggest that heavy concentrations of stranded oil can cause measurable changes in beach water flux and in interstitial fauna (McLachlan and Harty, 1981, 1982) . Also the normal processes of sedimentation and vegetative propagation in marshes by tarry surface crusts clearly are inhibited (Levasseur and Jory, 1982~. Similarly, the immobilization of beach surface substrate by heavy residual stranded tar as seen in Chedabucto Bay (Owens, 1981) alters the usual seasonal beach dynamics, with implications for the stability of the back-beach community. EFFECTS ON BIOLOGICAL PROCESSES Developmental Problems Chromosomal Aberrations Damage by oil to chromosomes or chromosome functions is only poorly understood, and research efforts to date on oil-induced aberrations in marine organisms are 1 imited . Even so, while there exists no broad threat to chromosome function, there is indirect and direct evidence suggesting that measurable increases in the burden of chromosome mutation can occur, at least in fish, under certain conditions of petroleum contamination. Studies of the mutagenicity of whole oils or refined products or their water-soluble products are few, even with mammalian tissues. The most recent study available, an examination of chromosome mutagenicity of heavy oil extracts, demonstrated the capacity of this material to

384 induce chromosome aberrations in hamster cell cultures at 0.06 and 0.03 mg/mL (Matsuoka et al., 1982) . Subsequent fractionation of the oil showed the mutagenic potential to be associated with a fraction containing neutral or weak basic nitrogen-containing compounds . Most of the research to date has focused instead on individual mutagenic compounds or their mutagenic metabolites. Results from studies with mammalian and other tissues indicate that several aromatic hydrocarbons, metabolites of the polycyclic aromatic hydrocarbons, and heavy metals found in crude petroleum and in refined products can impair fidelity of DNA synthesis, increase sister-chromatic exchange and chromosome mutation, and/or cause abnormalities in chromosome number (Table 5-3~. These include benzene, phenanthrene, naphthalene, chrysene, benzota~pyrene, 3-methylcholanthrene, benzoanthracene, 7,12- dimethylbenzofaJanthracene, benzotb~fluoranthrene, benzote~pyrene, dibenzanthracene, phenol, nickel, arsenic, vanadium, and lead. Fish and most marine invertebrates are now known to have the requisite enzymatic systems for converting promutagens, such as the polycyclic aromatic hydrocarbons, into gene-active metabolites capable of inducing gene and chromosome aberrations (e.g., Payne and May, 1979; R.F. Lee and Singer, 1980; R.F. Lee, 1981; Stegeman, 19801. ALSO, fish are known to be particularly susceptible to irradiation as a mutagenic agent, producing a range of responses with harmful effects on reproduction not unlike those found in mammals. One might suspect then that mutagenic agents, such as petroleum hydrocarbons, might provoke generally similar responses in fish chromosomes . Chromosomal aberrations (chromosome breaks, chromatic breaks, gaps, and interchanges) were all observed in gill cells of a freshwater tooth carp (Notobranchius rachow) exposed for 4-6 days to a 10-4 mg/L concentration of benzota~pyrene (Hooftman and Vink, 1981~. In another study , Engl ish sole (Parophrys vetulus ), collected from a relatively unpolluted site in Puget Sound, when injected intraperitoneally with benzota~pyrene, were found to have enhanced sister-chromatic exchanges in kidney cells when compared with nontreated sole from the same area (Stromberg et al., 1981~. S~milarly, cultured cells of rainbow trout gonadal tissue, incubated with either 3-methylcholanthrene (1-16 ug/mL) or with benzota~pyrene {0.1-1 ug/mL) showed increased incidence and severity of chromosome abnormalities over control cultures (Rocan et al., 19811. Fertilized eggs of English sole exposed to O .1-4. 2 ug/L benzo {a~pyrene (B (a) P) revealed significantly increased incidences in chromosome abnormalities of their yolk-sac membranes (Hose et al., 1982~. Treatment of the sperm and eggs of the purple sea urchin (Strongylocentrotus purpuratus) for 15 and 30 minutes with concentrations of B(a)P of 1.0-50 ~g/L yielded embryos with significant increases in the number of mitotic abnormalities tHose et al., in press). Cytologic-cytogenetic analyses have been appl ied, in one instance, to samples taken directly from the scene of the Argo Merchant spill on the Nantucket Shoals (Longwell, 1977, 1978) . Cod and pollock species eggs from the two most severely contaminated sampling stations showed markedly greater cytological deter ioration than eggs from less severely oiled stations. Abnormal differentiation or dedifferentiation of the

385 embryonic cells, grossly malformed embryos, and fouling of the outer egg membrane with tar were str ik ing at these two stations. There was also indication of effects on the embryonic mitotic index in these samples. Cod eggs collected from all five oiled sampling stations showed greater incidence of mitotic abnormalities than similar laboratory-spawned cod eggs. Twice as many pollock eggs as cod eggs experienced mitotic difficulties, possibly because pollock eggs were higher in the water column. Many pollock eggs were contaminated with oil on the chorion. The overall heredity risk factors to marine populations from oil are difficult to assess because of unknown natural mutation rates, the occurrence of natural mutagens, and a multiplicity of unrelated muta- genic pollutants. To this must be added varying species sensitivity and acclimation, var. table cell and tissue sensitivities, and variable tissue accumulation and metabolism. The enormous number of aromatic hydrocarbons in petroleum probably act synergistically or antagonisti- cally with one another, and with other classes of marine contaminants to increase or decrease mutagenic potential (Longwell and Hughes, 1980~. Increased mutation reduces genetic fitness of individuals and populations, but demonstrations of mutagenic potential in highly sensitive laboratory tests cannot be taken as evidence of similar genetic damage affecting the population under normal in vivo exposure in nature. Enormously fecund marine species and those with par- ticularly short generation times can tolerate greater mutation frequency and gametic wastage than can less fecund ones with longer generation times. However, the effect of Genetic damage on localized populations of some marine species is a posse. contaminated coastal waters by commercial fish stocks for oreea~ng, spawning, and nursery grounds increases their genetic risk. The risk in these specialized cases is enhanced by the greater susceptibility and sensitivity of the embryonic and juvenile stages. The regularity with which these abnormalities can occur, even after relatively short exposures, suggests this as a possibility. For these reasons, further attention to research in this area seems worthwhile. _ _ - . _ . -- _ · . · · . ~ _ The use of Reproduction and Development Both laboratory and field studies have shown that a broad range of reproductive and developmental processes can be affected by hydrocarbon exposure. As successful reproduction and early development are essential to the survival of species and populations, there is much-needed research emphasis in this area, particularly in light of their sensitivity to petroleum contamination. Most results have been obtained in laboratory experimental studies, with fewer confirming data from field situations. For the most part laboratory exposures have been with single hydrocarbon compounds or with known mixtures in either water-soluble or water-dispersed phases, over var ious time per iods. Observed field exposures have been with dispersed crude and fuel oils over more extended durations.

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390 Gametogenesis Petroleum compounds and whole oils are known to inter- fere with sex pheromone responses of algal gametes, but the concentra- tions at which this occurs (0.02-1.0 mg/L) are at least 3 times above the levels of naturally occurring pheromones (Derenbach et al., 1980; Derenbach and Gerek, 1980~. Algal (Steele, 1977) and molluskan (Renzoni, 1975) sperm are sensitive to oil, but there are differences in sensitivity. Thus algal sperm demonstrate sensitivity down to 0.2 ~g/L oil in water, but echinoid sperm are apparently much less sensitive (Lonning, 19771. The development of gonadal tissue has been found to be sensitive to petroleum hydrocarbons, but the mechanism is not understood. Thus, corals exposed to the water-soluble fractions of crude oils created by floating oil (nominally 3,000 mg/L) on the surface of a flowing water system for 2 months resulted in lower numbers of female gonads per polyp (Rinkevich and Loya, 1979~; in another study, exposure to No. 2 fuel oil caused degenerated ova and abnormal gonadal development (Peters et al., 1980a,b). Larvae can also be prematurely released from oiled corals {Loya and Rinkevich, 19791. Hydrocarbons may commonly be transferred from the gonads to early developmental stages, as shown with polychaetes (Ross) and Anderson, 1977) and scaphopod mollusks {Koster and Vanden Biggelaar, 1980~. Early embryogenesis with a wide variety of species is a particularly sensitive stage of development (seas, 1974; Donahue et al., 1977; Ernst et al., 1977; Lonning, 1977; Kuhnhold , 1978; Linden, 1978 ; Vashchenko, 1980; among others), which has been demonstrated in detail with echinoids and teleosts. Embryological effects among birds exposed internally or externally to oil have been shown to occur (Coon et al., 1979; Grau et al., 19771. The hatching success of teleost and avian eggs exposed to hydrocarbons is markedly reduced, often but not always at low levels (less than 1 mg/L) (J.W. Anderson et al., 1977; Ernst et al., 1977; Sharp et al. , 1979; Kuhohold, 1978; Albers, 1977; Szaro and Albers, 1977; White et al., 1979~. Development The general sensitivity of larvae to hydrocarbons, especially of crustaceans and teleosts, is now well recognized, but concentrations causing effects vary between life stages and can be influenced markedly by physiological condition. The delayed develop- ment of some larvae, notably of decapods (Wells and Sprague, 1976; Caldwell et al., 1977; Laughlin et al., 1978, among others) , often occurs at in~tial total hydrocarbon levels well below 1 mg/L. In a series of studies, Lonning and colleagues (Lonning and Falk-Petersen, 1982; Falk-Petersen et al., 1982; Kjoravik et al., 1982) have demonstrated deleterious effects on cod eggs and on sea urchin eggs and embryos when exposed to a variety of aromatic hydrocarbons (naphthalene, methylnaphthalenes, benzene, phenanthrene, xylene). Short term exposure (3-6 hours) of cod eggs to water solutions of xylene showed a range of reactions, from irregular cell cleavage (at 2-7 ug/L) to no cleavage and poor accumulation of cytoplasm (16-35 ug/L). Treatment of sea urchin embryos with benzene (10-2 M) or phenanthrene (10-4 M) resulted in abnormal differentiation and abnormal larvae. Methylnaphthalenes (1- and 2-methylnaphthalenes)

391 elicited a high incidence of skeletal abnormalities in sea urchin larvae, including aberrant skeletal forms or the formation of an extra skeleton rod. Laboratory studies have also shown that some reproductive and developmental processes are fairly resistant to hydrocarbon exposure, but these are fewer in number relative to those processes found to be vulnerable. For example, no effects were noted among oysters chroni- ca~ly exposed to crude oil, even though their gonads accumulated high concentrations of total hydrocarbons (Vaughan, 19731. It seems that sperm of echinoderms are quite tolerant of oil dispersions or extracts, and fertilization among echinoderms is seldom prevented by oil exposures (Allen, 1971; DeAngelis and Giordano, 1974; L5nning, 1977; NiCo1 et al., 19773. Field Studies Some marked effects of hydrocarbons on reproduction and early development have been observed in the field. The fecundity of coral populations in chronically oiled areas of the Red Sea was sharply reduced (Rinkevich and Loya, 1977), and mussels at the West Falmouth oil spill were sterilized for at least one season (Blumer et al., ~ 970) . Reproductive problems, including a reduced ratio of females to males, reduced juvenile settlement, and long term (>7 years) inhibition of recruitment and low population densities, have been noted for the fiddler crab Uca nuqnax and have been directly related to high oil-in sediment content in salt marshes contaminated by the Flor Ida oil spill (Krebs and Burns, 1977 ~ . Seabirds are very susceptible to direct oiling or oiling of their egg clutches. R.G.B. Brown (1982) states that oil may act indirectly by affecting the laying rate, the Matchability of eggs, and the growth of chicks--sublethal effects which in the long term could reduce sur- vival of individual birds and breeding populations. However, very little direct information exists for this problem. Apparently not all populations in the field are affected seriously by hydrocarbons, and indeed some appear to be notably tolerant even during their reproductive and early developmental phases. Thus beds of kelp (Laminaria digitata) at the Amoco Cadiz spill site in North Britanny, France, apparently were not affected by the oil, at least judging from the few studies and surveys that were carried out at that time. Reproduction in 1978 and densities in 1979 were also believed to be normal (Raas, 19811 . Meiofaunal copepods in experimental beach plots treated chronically with oil reportedly increased in reproductive activity (Feder et al., 1976~. Reproduction among barnacles and mol- lusks near oil seeps or at spill sites appears unimpaired (Straughan, 1971, 1977a,b), and settlement of larvae is not always unaffected (Woodin et al., 1971) . Conclusion The above work, most of it conducted in the last decade, - shows that sensitivities to hydrocarbons vary between phyla and across and between developmental stages and covers processes from the matura- tion of gonads to the survival and settlement of larvae. This feature of var table sensitivities of 1 if e stages, organisms, and processes is particularly clear from studies of crustacea, echinoids, and teleosts

392 and is now well recognized by aquatic toxicologists. The reported threshold concentrations for the adverse developmental effects, based on initial measured hydrocarbons by spectroscopy or chromatography, are well below 1 mg/L and even down to 1 ug/L, for acute exposures in the laboratory. In contrast, some processes appear tolerant to much higher hydrocarbon levels. It is important to reemphasize that significant reproductive impair- ment in oiled field conditions has seldom been observed, although few field studies have been performed. Based on available studies at the population level, annelids, gastropods, and copepods seem to suffer no long lasting damage. Macrophytes, barnacles, and birds may sometimes be affected. However, corals, bivalves, and decapod crustacea can suffer marked and sometimes long term (years) reproductive damage at oiled sites. There is sufficient information now available on effects at rela- tively low concentrations to cause concern about spills and chronic discharges of oil into protected or enclosed coastal waters. Of equal concern, perhaps, is the frequent absence of threshold data for many processes in groups of organisms when exposed to realistic concentra- tions of hydrocarbons. Pathological Consequences Laboratory studies have shown that individual aromatic hydrocarbons, whole petroleum, and fractions of petroleum induce a variety of cel- lular and subcellular alterations in marine teleosts and invertebrates (Malins, 19821; however, the concentrations of petroleum components used were often substantially higher than those found in petroleum- contaminated marine environments. In teleosts the effects include abnormalities of the eyes, e.g. alterations in lens fiber cells (J.W. Hawkes, 1977) and of the chloride cells in the gills (Engelhardt et al., 1981) (Table 5-4~. In inver- tebrates, the effects include bronchial, e.g., pigment body formation, and renal lesions, e.g., necrosis (Gardner et al., 1975), and changes in the cellular and lysosomal structure of digestive cells (Lowe et al., 1981~. Embryos and larvae appear to be particularly susceptible to petroleum exposure, sometimes at lower concentrations than those inducing morphological changes in mature organisms. For example, petroleum-induced morphological changes in teleost embryos include retinal damage, e.g., necrotic neurons (J.W. Hawkes and Stehr, 1982} (Figure 5-2) and the failure of fins to differentiate from the lateral 1 ine body wall (R.L. Smith and Cameron, 1979~. Overall, the laboratory findings have not as yet provided a thorough insight into how the petroleum-induced pathological changes relate to the degree or time of exposure, or to a variety of other experimental conditions. Moreover, real difficulties exist in trans- lating the findings into the field situation, where, for example, there are compounding problems arising from the complexity of ecosystems and the influence of other contaminants (Malins and Collier, 1981~.

393 TABLE 5-4 Summary of Gill Morphology in Rainbow Trout (Salvo gairdner i) Following 7 Days of Oil Treatment, Compar ing Dif ferent Modes of Exposure Degree of Epithel ial Chlor ice-Type Cell Exposure Mode Fusing Separation Exposure Location Vacuoles ~ _ _ Freshwater Control none none none basal rare to none emulsion in paraffin none none rare basal none E3nuls ion in NW some some extens ive lengthy extens ive Emulsion in VEN extensive extensive extensive length some WSF of NW none none rare basal few to none WSF of VEN none none little basal none i.p. with NW little none some (basal) b rare Seawater Control rare none none basal none Emulsion in NW some none little (basal) some Emuls ion in VEN var. table none some (basal ) common i.p. with NW rare none none basal none NOTES: NW, Norman Wells crude oil, weathered 3 days to 200 pL/L; VEN, Venezuelan crude oil, weathered 3 days to 200 pL/L; WSF, water-soluble fraction; i.p., intraperitoneal in Section, 100 tlL/kg f ~shweight per day. Ocelot ice-type cells found along length of secondary lamellae . Chloride-type cells found in basal region and extending part way up secondary lamellae (from Engelhardt et al., 1981) . Contaminant-induced lesions or tissue abnormalities (e.g., tumors) also are difficult to distinguish from those caused by some other agent, such as viruses , e.g., over fan neoplasmic responses in clams from a Maine oil spill site (Yevich and Barszcz, 1977; R.S. Brown et al., 1977) . Field studies, notably those conducted after the Amoco Cadiz spill, have shown that gross pathological conditions, e.g., fin erosion in teleosts (Haensly et al., 1981) r and cellular/subcellular alterations, e.g., abnormal proliferation of chloride cells in invertebrates (Lopez et al., 1981), can be found in association with spilled petroleum. Again, it is difficult to link these pathological abnormalities directly to the spilled oil, either because of absence of proper controls or because of the presence of other contaminants or interfering conditions. Also, the data provide little understanding about how these petroleum- induced morphological changes can affect the viability of important resource species. For example, certain petroleum-related morphological changes, such as the hepatocellular lipid vacuolization in teleosts

394 (McCain et al., 1978) are questionable indicators of altered organism health. Further, relationships between chemically and biologically induced transformations in petroleum components in the f ield and deleterious effects at the cellular/subcellular level remain largely unknown (Malins et al., 1980~. Also, only a limited understanding of teleost and invertebrate pathology has been attained, especially with regard to the significance and frequency of lesions associated with petroleum exposure. Despite the various limitations , such as our inability to describe dose-response relationships with any degree of precision or, in field studies, our inability to separate effects of polycyclic aromatic hydrocarbon from those due to other pollutants, evidence is accumulating that petroleum exposure can cause gross and cellular abnormalities in marine organisms. The recognition that alterations in cellular and subcellular structures do result from petroleum exposure is an important first step in understanding relations between petroleum and effects on the biological structure of marine organisms. In this respect it is regrettable that more work has not been done on spill sites or in regions (e.g., coastal waters) receiving chronic petroleum inputs. Growth and Metabolism Since the early 1970s, research interests have shifted demonstrably toward studies of the mechanisms of petroleum toxicity, and toward an understanding of petroleum hydrocarbon toxicity both at the cellular and at the organismic level. This is reflected in the nature of the large amount of data that have been assembled since the National Research Council (1975) report and in the quality of current research Petroleum hydrocarbons are now known to affect nearly all aspects of physiology and metabolism, although the ultimate site or sites of toxic action within the cell or organism still are not understood. Plants The literature dealing with petroleum effects on the metabolism and physiology of marine plants is not extensive, and is restricted mainly to work with phytoplankton species. Processes studied most often are those of growth, photosynthesis, and to a lesser extent, respiration. Thus, there is little information available that might lead to general izations about the comparative effects of petroleum hydrocarbons on a taxonomic basis (see Macrophytes section below). Photosynthesis Phytoplankton exhibit widely varying responses to oil and to oil products, sometimes showing, for example, enhanced photo- synthesis and at other times inhibited processes. Species also differ widely in their vulnerability (Pulich et al., 1974~. In general, the green algae have been found to be the most sensitive, blue-green algae next, and the diatoms most tolerant to petroleum (Winters et al. , 1977) . How much of this is due to experimental procedures, however, is not

395 known. Thus, R. F. Lee et al . (1978 ~ observed the opposite, with diatoms h ighly susceptible to petroleum. In a ser. ies of exper zments compar ing the phytotoxicity of several No. 2 fuel oils, differences in toxicity vrere ascribed to the chemical composition of the water-soluble fraction (WSF) (Batterton et al., 1978a). The presence of p-toluidine in this WSF was found to be cor- related with the increasing phytotoxicity observed. Indeed, low concentrations of p-toluidine (50 ug/L) would arrest growth in a blue-green alga (Batterton et al., 1978b) . Cur Piously, in another study (Winters et al., 1977), light was found to influence the effect of a constituent of No. 2 fuel oil, per inaphthenone, on growth of two green algae. For unknown reasons a shift from white to yellow light resulted in an increased inhibitory threshold from 0.125 to 5 mg/L "Winters et al., 1977~. There is some suggestion that the inhibitory effect, at low concen- trations of the toxicant, may be reversible. Cells of the unicellular alga Monochrysis lutheri, after inhibition of photosynthesis on exposure to naphthalene, were found to recover their photosynthetic 14C fixation when transferred to clean nonnaphthalene containing culture medium (Vandermeulen and Ahern, 1976~. Cellular Mechanisms Little is known of the mechanisms by which petro . leum affects either photosynthesis or metabolism in marine plants. Studies with freshwater unicellular algal species suggest a correlation between toxicity and Leakage n of several ions from the exposed cells (Hutchinson et al., 19791. Other studies indicate disruption of intra- cellular macromolecule pools, including an alteration of the ATP/ADP balance in unicellular algae (e.g., Vandermealen and Ahern, 1976), and of DNA and RNA synthesis in certain macroalgae (Davavin and Yerokhin, 1979~. Comparatively little is known of metabolism of petroleum hydro- carbons by mar ine algae, such as by the mixed function oxidase system based on the cytochrome P450 complex as found widely throughout the animal phyla. However, recent studies by Cerniglia et al. (1980a,b, 1981a, 1982) indicate that blue-greens (Cyanobacter ia ~ and a number of green algae, as well as a red and a brown alga can metabolize naphthalene and aniline (Cerniglia et al., 1981b). Whether the process is P450 based remains to be seen. Abnormal Growth Cer ta in polynuclear aromatic hydrocarbons appear to be able to affect growth form in macroalgae, possibly through al tering normal apical growth of the plants (Boney, 1974~. Tumor-like growths found in Porphyra tenera, in areas of industr ial pol lotion, would seem . to correlate with the occurrence of several polycyclic aromatic bydro- carbons in the sediments (Ishio et al., 1971 , 1973) . In this instance , however, as is often the case, it is di fficult to separate the hydrocar- bon effects from those of other pollutants or from other sources.

396 Animals Evaluation of the effects of petroleum hydrocarbons on growth and metabolism of animals is more extensive, with a sizable literature dealing with feeding, respiration, growth, and enzymatic detoxification systems. Feeding Of the various responses that animals are capable of with respect to toxicant exposure, feeding seems to be the first or one of the first to be affected, generally negatively. Although feeding rate is, of course, ultimately important to the wel1-being of the organism, it is only one factor in the total energy budget. Therefore, the direct consequences of an altered feeding rate are difficult to assess. They are probably most serious in planktonic animals with little food reserves and those with short life cycles or durations in the plankton (e.g., meroplankton and microzooplankton). Effects on feeding have been observed in most phyla, at concentra- tions approaching those measured initially under oil spill situations (e.g., Berman and Heinle, 1980; Elmgren et al., 1980a). The effects are found both in the planktonic and pelagic groups, e.g., lobster larvae (Wells and Sprague, 1976) and adult copepods, and in benthic organisms , e .g ., lugworms (Augenfeld, 1980 ~ . The consequences of reduced feeding rates can be further manifested in reduced production rates (Ott et al., 1978) and in the case of benthic systems, in reduced rates of sediment rework ing, e .g . ~ Arenicola mar ina (Gordon et al ., 1978 ~ . Respiration Respiration rates are very labile and can be affected in both directions by many factors other than petroleum. Again, respira- tion is only one factor in an organism's total energy budget. Nonethe- less, respiration is easily measured by a range of laboratory methods and can serve as an indicator of the organism's well-being, provided the proper controls are included. For that reason it has formed the basis for many oil-related studies over the past 10-15 years. Depression of respiration has been observed in a range of marine organisms, including crustacea, mollusks, and fish (for example, Malins, 1977; G~lfillan and Vandermenlen, 1978; Fisheries Research Board, 1978; Thomas and Rice, 1979~. Again, the results are variable, and in some instances respiration has been found to increase in response to certain oils. To date, there is no clear relationship between oil type and the observed response. In general, petroleum levels of 1 mg/L, or higher are required for effects to be visible in juvenile and adult crustacea and mollusks. In fish and planktonic crustacea the effects are observed in the laboratory at concentrations of petroleum less than 1 mg/L, approaching those levels observed under oil spill conditions. Growth Growth in animals bears directly on the capacity of the organism to survive in its environment. Measured in a variety of ways (changes in length or weight, scope for growth), growth has been found to be markedly affected by exposure to petroleum in a vat iety of marine organ- isms. Reduced growth in oiled field situations has been observed,

397 reproducibly, in both mollusks and fish (Gilfillan et al., 1976, 1977a,b; Gilfillan and Vandermeulen, 1978; Desauney, 1981; J.W. Anderson et al., in press) . Parti cularly susceptible are those benthic organ- isms, such as bivalves, inhabiting chronically oiled sediments. The degree of reduction in growth seems to be a function of the ani'Ta1's feeding mode; thus in mollusks the deposit-feeding species appear to be more affected than the filter-feeding species (Augenfeld, 1980; Roesijadi and Anderson, 1979~. In fish, growth depression is most pronounced in such species as flatfish, which live in continued intimate contact with oiled bottom sediments (e.g., McCain et al., 1978~. Transfer of hydrocarbons from such oiled sediments to associated benthic organisms occurs readily, as has been demonstrated under controlled simulated conditions. ~ , , ~ Deification Systems Little information is available on the direct interaction of petroleum hydrocarbons with the more fundamental processes of metabolism such as enzyme activity and ion transport. One area that has received considerable attention in the past 7 years is that of detoxification mechanisms, mainly via a mixed function oxidase (MFO) system involving cytochrome P450. This enzyme system has been found in all mar ine animals investigated, including both ver tebrates and invertebrates, and has been found to be readily inducible (i.e., will r ise to elevated levels) in organisms on exposure to oil or certain other compounds. The MFO activity appears to be affected by several factors, including seasonality, sex, and maturity of the organism, as well as its pollution history (Stegeman, 1980~. The latter becomes important in assessing MEO capability, particularly in coastal waters with chronic low level pollution. The inducibility of the enzyme system appears to be in response to the polycyclic aromatic hydrocarbon content of the oil, particularly the 3-to-5-ring aromatics, such as benzofa~pyrene, chrysene, and benzanthracene (viz., Chapter 4~. Several field studies report elevated MFO activity in fish taken from contaminated areas (Burns, 1976; Bend et al., 1978, 1979; Bend, ~980; Iwaoko et al., 1977; Vandermeolen et al., 1978; Payne et al. , 1978b; Spies et al., 1982) . The only exception to this inducibility appears to be the bivalve mollusks, which even after several years In chronically oiled sediments did not show elevated MEO levels in their tissues (Vandermoulen and Penrose, 19787. Because of the correlation between enhanced MFO activity and environmentally available aromatic hydrocarbons, e.g., MFO in English sole versus sediment benzofa~pyrene levels (Dunn, 1980), there has been interest in using this system as an environmental indicator of oil pollution (Dunn and Stich, 1975, 1976; Payne, 1976) . However, many factors can clearly influence the degree of inducibility of the MFO system. In addition, there are fundamental differences between the MEO system in f ish and that found in invertebrates . Also, there are many other compounds besides those derived from oil that may induce MEO activity in mar ine organisms, including some organic compounds of b iogenic or igin .

398 Metabolism Various studies have shown oil impact on such metabolic functions as lysosome stability (M.N. Moore et al., 1978; M.N. Moore, 1979), taurine/glycine ratios in bivalves (Roesijadi and Anderson, 1979), oxygen to nitrogen ratios and lipid utilization in zooplankton (Vargo, 1981; Capuzzo and Lancaster, 1981), hepatic lipogenesis in certain fish (Stegeman and Sabo, 1976; Sabo and Stegeman, 1977), and reduced pi asma copper (Dillon, 1981) . Changes in plasma chloride have been noted in f ish exposed for 6 months to floating oil in simulated environments (Payne et al. , 19781 . Together , these and other observa- tions indicate the general toxicity of petroleum at the metabolic level of cell and organ function. Most of these observations come from laboratory studies, but interesting findings come from recent studies on changes in digestive enzyme levels in zooplankton from Amoco Cadiz impacted waters (Samain et al., 1979~. Analyses of zooplankton taken from the English Channel during and following oiling from the tanker showed changes both in the individual levels and in the ratio of trypsin to amylase. While there is, of course, again the question of relating the observed effects to the spilled oil in the water column, this approach would appear to be a useful avenue to explore further. Behavior Behavior may be considered the first line of defense for dealing with an environmental perturbation (Slobodkin, 1968), and the response may be avoidance by movement or other changes that would tend to reduce exposure. Oil, at sublethal concentrations, can significantly alter the behavior of marine organisms, including both microorganisms (bacteria, motile phytoplankton) and invertebrates, fish, and larger organisms. Changes in this behavior ultimately can reflect on such processes as feeding, reproduction, and larval settling. In microorganisms the behavior patterns are primarily those of changes in motility. In higher organisms they include more complex patterns of general activity, avoidance, burrowing, feeding, and reproduction. Microorganisms Sublethal concentrations of oil can change the motile behavior of unicellular organisms or alter important metabolic processes that are closely allied with the motility apparatus. Several ecological consequences have been suggested, including inhibition of bacterial- mediated nutrient regeneration and pollutant removal, disruption of intermicrobial predation and of alga-bacterial interactions, and prevention of phenomena mediated by the settling of mobile microbes on surfaces (Mitchell and Chet, 19781. Bacteria are repelled by several known components of petroleum, including benzene, aniline, and phenol; thresholds for detection average 10 4 M (Young and Mitchell, 1973; Tso and Adler, 1974~. Most repel- lents are cytotoxic at concentrations well above those which produce

399 negative chemotaxis. Among eucaryotic cells, as in the algae, important attractants include the terpene hydrocarbon derivatives, which act as sexual pheromones for the male gametes of the marine brown algae FUCUS, ~argassum, and D~ctyota (Kochert, 1978; Ra~iwara et al., 1980; Muller et al., 1981~. More recently, Vandermsulen et al. (in press) have descr ibed altered strutting patterns in the unicellular phytoplankton Monochrysis lutheri in response to both whole oil and individual hydro- carbons, at concentrations we' 1 below those eliciting effects on photosynthesis and other physiological processes (<500 ug/L Kuwait and naphthalene). The mechanism of inhibition or blockage is not understood. Bacteria detect chemical stimuli via specific protein chemoreceptors, some of which double as active transport enzymes for the substrates with which they combine . The resulting "signal " is transduced to the flagellar appar atus v ia separate membr ane-bound chemotaxis prose ins (MacNab, 19783. There is circumstantial evidence for the existence of highly specific membr ane-bound chemor eceptor s in algal gametes as well. Chemotaxis can be inhibited by blocking chemoreception, signal transduc- tion, or the normal functioning of the flagellar apparatus. These var ions processes are dependent upon the normal functioning of the cell membrane, which therefore provides a highly accessible target for the action of various petroleum hydrocarbons. Positive chemotaxis functions to maintain bacterial cells in a nuts itionally favorable environment (Bell and Mitchell, 1972) . Negative responses serve to remove cells from potentially toxic conditions. Prevention of normal chemotactic behavior can inhibit what is in fact an important contribution to the general homeostatic mechanism of the bacterial cell and thus adversely affect microbial activity. Interfer- ence with sexual pheromone reception in algal gametes has obvious implications for reproductive success. Additional ecological studies are needed before we can adequately assess the effects of sublethal concentrations of petroleum hydrocarbons on native microbial populations. An oil spill is likely to result In an initial reduction or even inhibition of many aspects of native micro- bial activity, including chemotaxis (Bertha and Atlas, 1977) . However, oil pollution creates a new set of intensely selective environmental conditions which rather quickly result in the development of a hydrocarbon-based microbial ecosystem (Barsdate et al., 1980~. This may bring with it the development of certain resistances to the other- wise toxic effects of the hydrocarbons as , for example, carriage of plasmids conferring the ability to metabolize components of oil (Hada and Sizemore, 19811. Bacteria isolated in the presence of petroleum hydrocarbons have been found to exhibit normal chemotactic responses in the presence of these compounds, unlike those in non-oil-exposed cells, suggesting an underlying cellular resistance to their effects (Bitton et al., 1979~. Such bacteria may be more representative of the microbial populations which develop after oil spills.

400 Higher Organisms Petroleum hydrocarbons may cause large alterations in the behavior of mar ine invertebrates and f ishes . However, or itical compar ison of results is difficult because of the variation in the specific behavior measured, the diversity of the species studied, and in most cases, the insufficient data on exposure conditions. The influence of petroleum hydrocarbons on behavior of invertebrates and fish is less well understood than it is at the microbial level. While a small number of studies have shown that behavior, which is mediated by chemoreception, can be affected by oil, it has yet to be demonstrated whether effects are manifested at the sensory or at the behavioral level. A similar lack of understanding of the underlying causes of a behavioral alteration caused by oil is characteristic of most of the behavioral studies on the higher organisms published to date . Avoidance Invertebrates and fish and possibly other marine organisms - (viz., marine mammals, seabirds) can avoid polluted waters, although only a very small number of exper imental studies have dealt specifically with avoidance of oiled waters. This is regrettable, especially in view of the many claims that pelagic adult fish, for example, have the ability to avoid spills. In fact, there is little evidence to support this claim. Wherever avoidance has been studied, however, it always is dependent on the species and on such factors as concentration and type of petroleum (Percy, 1977), the aquatic environment, the season (Rice, 1973), as well as on the internal state and ecological requirements of the species. Even when an animal does avoid oil successfully, it is not neces- sar fly the appropriate response. For example , in the process of avoiding, other critical resources may be denied, such as food or shelter. This is shown in a study of the effect of oiled sediment on the burrowing behavior of the littleneck clam Prototheca staminea. In this case, experimental clams remained closer to the surface in oiled sediments than did individuals residing in clean sediments (Pearson et al., 1981), thereby avoiding the deeper oiled sediments but at the same time becoming more susceptible to predation. As the tests included a natural predator, the Dungeness crab, predation rates were observed to increase significantly. Similar impact on oiled hard clams (Mercenarza mercenaria} has been described by Olla et al. (19833. Thus risk factors can be shifted from those associated with petroleum toxicity to those associated with habitat, e.g., predator-prey interaction. Chemoreception Blumer (1969) was possibly the first to suggest that petroleum hydrocarbons could produce serious consequences by interfering with chemoreception. As chemical senses play a major role in mediating critical aspects in the behavior of mar ine organisms, including feeding , r eproduction, habitat selection, and predator recognition, the implica- tions of this phenomenon may be far reaching. Indications of this possibility have come out of the work of Kittredge et al. (1974), who described alterations of sexual/mating behavior in the shore crab

401 Pachygrapsus exposed to oil, and received confirmation from the studies on altered lobster feeding behavior in the presence of crude oil, carried out by Atema (e.g., Atema et al., 1973; Atema and Stein , 19741 . To this list must be added decreased reproductive success of certain eucaryotic macroalgae through interference with pheromone chemorecep- tion by male gametes (Derenbach and Gereck, 1980~. As yet there is no certainty that these alterations were at the sensory level. The most convincing evidence thus far, without the use of neurophysiological techniques, is the observation that the chemosensory antennular response in the Dungeness crab Cancer magister is impaired by low concentrations (less than 1 ppm) of petroleum hydrocarbons (Pearson et al., 1981~. Feeding Feeding and behavior associated with feeding are well known to be affected by sublethal concentrations of petroleum as low as a few ug/L. This includes reduction in the capability to respond to a food source (Jacobson and Boylan, 1973; Pearson et al., 1981), recognition of food (Atema et al. , 1973) , and behavior associated with location of food (Atema and Stein, 19741. Measurements of amounts of food ingested have also shown effects of petroleum, both in animals residing in the water column (Berdugo et al., 1977; Berman and Heinle, 1980) and in the sediment (Prouse and Gordon, 1976; Gordon et al., 1978~. Recurrent throughout these various studies is the observation that organisms vary widely in their response to petroleum or its components. This was shown perhaps most conclusively in the response differences to oil of two arctic marine amphipods and two isopod species (Percy, 19771. The one amphipod species, Onisimus affinis, avoided oil-contaminated sediment. In contrast, this response in a second amphipod, Corophium clarencense, and of two isopods, Mesidotea entomon and M. sibirica were either totally absent or much reduced. This illustrates the point that generalizations regarding behavioral responses to oil and regarding certain common species are not reasonable. It also underscores the risk in the use of Vindicator species n or Vindicator organisms n in assessing and predicting generalized effects of oiled ecosystems (Olla, 1974 ; Olla et al., 1980a,b). EFFECTS ON THE: MARINE FOOD WE 13 Food Web Microbes The effect of oil on marine microorganisms depends on the type and amount of oil spilled, the physical nature of the area (e.g., open ocean or estuarine marsh), nutritional status, oxygen concentration, and previous history of the impacted area with regard to hydrocarbon exposure. Such pr for impact may be reflected in changes in number and types of microorganisms, as well as in their chemical composition and changes in microbial activities. Light crude oils and refined products tend to be more toxic and to affect biological activities more than the heavy crude oils. Chronic low level spills, which allow time for selection of naturally occur r ing microbial degradation of oil. Both types of spills produce changes in

402 the community structure and activities of the microbial population, the chronic one less dramatically, the catastrophic one very quickly. The net result is that those sensitive to hydrocarbons will be killed or their growth suppressed, whil e those with the genetic potential to utilize hydrocarbons as a carbon and energy source will grow and increase in numbers and/or biomass. The best documented microbial response to the intrusion of oil in marine systems is the increase in proportion of oil-utilizing bacteria to the total heterotrophic bacteria and is summarized by Atlas (1981~: n In unpolluted ecosystems, hydrocarbon utilizers generally constitute less than O.19e of the microbial population; in oil polluted ecosystems, they can constitute up to 100% of the viable microorganisms. The degree of elevation above the unpolluted compar ison reference sites appears to quantitatively reflect the degree or extent of exposure of that eco- system to hydrocarbon contaminants. n Such increases usually refer to bacteria, but Ahearn and Meyers (1972) reported a selective effect of oil on the incidence of mar ine hydrocarbon degrading yeast and/or fungi. Hence, the fate of spilled petroleum and, therefore, an oil- contaminated environment, lies in the microbes' ability to use hydrocarbons as sources of carbon and energy. Because of the problems in the enumeration and identification of marine heterotrophic bacterial populations (Staley, 1980), little information is available on the effect of oil on species diversity. Species of oil-utilizing bacteria that have been isolated varied with the geographic site of the sample and differed between pelagic and benthic samples taken from a given site (Colwell and Walker, 1977~. An increase in the number of plasmid-containing (extrachromosomal genetic material that enhances biochemical activities including hydrocarbon oxidation) strains of Vibrio spp. and a greater number of plasmids per strain were found in samples from an active oil field in the northwest- ern Gulf of Mexico (Hada and Sizemore, 1981) . The authors suggested that oil field discharges might be responsible for this increased plasmid incidence. Petroleum has been shown to have an effect on the uptake and mineralization of low-molecular-weight compounds such as glucose, glutamate, acetate, and glycoxolate. Griffiths et al. (1981) reported that mineralization of glutamic acid and glucose was inhibited by Cook Inlet crude oil and related products, the pelagic populations being more sensitive to the effect of oil than benthic ones. Regions with a history of chronic hydrocarbon input showed less sensitivity to the effects of oil than pristine ones. The addition of oil to environmental samples has resulted in a decrease in microorganisms showing proteolytic, chitinolytic, and cellulytic activities (Walker and Colwell, 19757. Similarly, Griffiths et al. (1982b) found that Cook Inlet crude oil reduced the levels of cellulose and chitinase activities observed in sediments but stimulated the activities of enzymes involved in degradation of starch and alg in . Nitrogen fixation (acetylene reduction) in northern mar ine sediments has been reported to be significantly reduced by exposure to ~fresh" but not "weathered" Prudhoe Bay oil (Griffiths et al., 1982b) .

403 Effects on Mar ine Plankton Phytoplank ton There are ample observations that petroleum hydrocarbons can have immediate and marked effects, in terms of minutes or hours, on the rate of photosynthesis of natural phytoplankton assemblages (Shiels et al., 1973 ; Gordon and Prouse , 1973 ; Lacaze , 1974 ; Le Pemp et al ., 1976 ; R. F. Lee and Takahasi, 1977; Brooks et al., 1977; Hsiao et al., 1978; Federle et al., 1979~. Laboratory observations with unialgal cultures and pure cultures have provided a similar picture (Pulich et al., 1974; Soto et al., 1975a; Parsons et al., 1976; Vandermeolen and Ahern, 1976; Rusk, 1978; Karydis, 1979) . Both laboratory and field studies clearly show that hydrocarbons can inhibit algal growth, although at the lower con- centrations of oil , occasionally an enhancement is noted (e .g ., Gordon and Prouse, 1973~. Unfortunately little more is available to present a better under standing of both the toxic e f feats and the impact on natur al populations under spill conditions. - Growth and Metabolism Most of the research to date into the effects of oil on phytoplankton teas dealt with either the effects on culture growth or on photosynthes is . Presumably that is because of the ease with which these can be assayed, either in the laboratory or on ship- board. Far less attention has been directed to research on effects on the more fundamental aspects of metabolism or cellular fine structure. And indeed most of what is known of the latter comes from work with freshwater species. Growth of phytoplankton is readily depressed by a wide range of petroleum hydrocarbons, including both whole oils as well as specific compounds . This includes blue-green a' gee, green algae, diatoms, dino- flagellates, and chrysophytes. Effects on growth vary widely, depending on the oil or compound used and on the algal species, but generally growth lags or lethality has been noted in the range of 1-10 mg/L (e.g., Mommaerts-Billiet, 1973; PUlich et al., 1974; Soto et al., 1975b; Prouse et al., 1976; Batterton et al., 1978b; Hsiao, 1978; Mahoney and Haskin, 1980~. Recent work suggests that algal sensitivities may not only be species specific, but also clone specific depending on the environmental origin of the clone (Eppley and Weller, 1979~. The growth response var. ies considerably and is somewhat dose respon- dent. Thus at low concentrations the main response consists of a lag period, after which culture growth occurs normally, growing through the usual log phase and reaching the plateau stage. Cell size under these conditions does not differ from normal control cultures. At higher concentrations the inhibition becomes increasingly marked, with little or no growth at the highest concentrations used. Interestingly, at very low concentrations (<0.1 mg/L) growth enhancement has been observed in both laboratory and field collections (e.g., Prouse et al., 1976). Photosynthesis is equally depressed in phytoplankton exposed to petroleum. Enhancement, at very low hydrocarbon concentrations, as seen for growth has not been noted. One study has examined the poten

404 tial recovery of a phytoplankton species after exposure to naphthalene (Vandermeulen and Ahern, 1976~. On return to noncontaminated medium, the culture, Monochrysis lutheri, showed a partial recovery of photo- synthesis following initial inhibition. Jordan et al. (1978) reported that nitrogen fixation in the epiphytic blue-green alga Nostic sp. failed to recover for ~ year following a single dose of ASA 90 crankcase oil. Effect on Populations To date, no mass toxicity to phytoplank ton has been reported from the field, either from a spill or for chronic input conditions. In part this is due to the fact that very few field s tudies involving phytoplankton have been done dur ing an oil spill . One data set that is available comes from a sampling program carried out fortuitously when the tanker Kurdistan broke up off Nova Scotia (O' Boyle, 1980~. No effect of any sort was observed on phytoplank ton of the Scotia Shelf waters. However, the spill occurred in a prebloom period, and although oil patches and tar were reported from the Scotia Shelf (Vandermeulen, 1980), there is little indication that the algae in fact encountered appreciable oil concentrations. Observations on phytoplankton biomass and primary productivity carried out following the Tsesis spill (in Sweden, 1977, No. 5 fuel oil) revealed no signifi- cant differences between noncontaminated and contaminated areas (Johansson et al., 1980~. In fact, if anything, for a brief postspill period both primary productivity and cell numbers were found to be slightly higher in the contaminated areas, perhaps because of reduced grazing by zooplankton. This points out the difficulty of studying phytoplankton in isolation from the rest of the ecosystem. Even if a large number of algal cells were affected during a spill, regeneration time of the cells (9-12 hours), together with the rapid replacement by cells from adjacent waters, probably would readily obliterate any major impact on a pelagic phytoplankton community. The situation may not be as saJutory, however, in more chronically oiled waters such as inshore and coastal embayment systems, where flushing rates may be low and where concomitant oil contamination may be higher and prolonged. Under those conditions, per iodic events such as plankton blooms which are critical to oceanic planktonic processes may well be af fected . Zooplank ton The impact of oil on zooplankton has been studied extensively since the 1975 NRC report, primarily because of the importance of zooplankton in marine ecosystems as secondary producers. Work has been carried out in both the laboratory and at spill sites, as well as in ~ number of mesoscale" field enclosures (~mesocosms n ~ . The responses to petroleum exposure are numerous (e.g., Kuhohold, 1977; Corner, 1978~. As individuals, most zooptankters studied to date in acute and chronic exposure exper iments in the laboratory and the field appear to be highly vulnerable to dispersed and dissolved petroleum constituents, and less so to floating oils. The acute lethal

405 toxicity of dispersions and WSF, mostly expressed as 96-hour LC5n using initial measured concentrations range between 0.05 and 9.4 mg/L, with a few higher values (Table 5-5~. These values, based on measure- ments of actual rather than nominal concentrations, are very close to the lethality thresholds predicted, 0.1-10 mg/L, of soluble hydrocarbons for fish eggs, larvae, and pelagic crustaceans, derived from an earlier intensive survey of the literature (S.F. Moore and Dwyer, 1974~. It must be noted that sublethal deleterious effects can set in well before these high concentrations are achieved. Many components of the zooplank ton should, therefore, be considered sensitive to dispersed and solubilized hydrocarbons in seawater, based on these laboratory studies and on the combination of observations in field enclosures and at oil spill sites. On the positive side, concen- trations of hydrocarbons in open waters may not persist long enough to always cause many of the toxic effects, either lethal or sublethal (McAuliffe, 1977; also Chapter 4 of this report). Based on the available data on lethality of WSF (Table 5-5), there may be no marked differences in vulnerability among the planktonic ctenophores, mollusks, crustacea, and teleosts ~ This compar ison should be considered very tentative, as the data were not collected in one laboratory or under one set of conditions. Enclosure exper iments with zooplank ton (Lytle, 1975; R. F. Lee and Anderson, 1977; R.F. Lee and Takahashi, 1977; R.F. Lee et al., 1977, 1978a,b; Davies et al., 1980; Elmgren et al., 1980a; Vargo, 1981) (see also Chapter 3, Biological Methods section) point out the different capabilities within natural but confined zooplankton communities for accommodating the presence of hydrocarbons. These experiments also show the type and variability of toxic responses (lethality, lowered feeding and reproduction, community changes) that mi ght be expected among natural zooplank ton communities when continuously exposed to low levels of oil-der ived hydrocarbons . Field observations on zooplank ton have now been made at several spills and chronically polluted sites. Collectively, these studies have shown that biological effects and changes have been detected (Spooner, 1978; Johansson et al., 1980) , but these appear to be short lived; there are seldom significant prolonged changes in biomass or standing stocks of zooplankters in the open water near spills. Individual organisms in spills have been affected in a number of ways: direct mortality (fish eggs, copepods, mixed plankton), external contamination by oil (chor ion of f ish eggs, cuticles and feeding appen- dages of crustacea), tissue contamination by aromatic constituents, abnormal development of fish embryos, possibly temporary inhibition of feeding in copepods, and altered metabolic rates (Longwell, 1977; Gilfillan et al., 1983; Samain et al., 1980, 1981) . In addition, ingested oil has been seen in copepods on some occasions. But zoo- plankton populations and communities exper fencing spills or chronic d ischarges in open waters appear to recover eventually and maintain themselves, due largely to their wide distr ibution and rapid regenera- tion rates (Michael, 1977~. This type of recovery may not occur in enclosed waters . In fact, Sanborn (1977 ~ has suggested that the lowered zooplankton production in the Caspian Sea may be a reflection

406 · a, o Z . a.) 3 _1 `4 ~ ,~ O S~ o tn  .< o CO 3 o o U] C o .,4 JJ ~0= _ n ~ 3 54 o U] G) 1 _' `4 ~ ~3 4J O ; ~U) 3 ~n U) - a C) ,4 o a~ o - 4~ o c) 0 al a _ ~ ~ 0 ~ 4~ s Y 4~ ~: o ~ 0 1 c~ ~ ~: m ~: ~ ~D · Q) Z · 0 3 X 1  · 0 3 Z 3 _1 t-) O o C, o <~ - q:5 ~_ ^ ~ _ _ ~_e ~0 0 C~ ~ l ~ _ _ _ _~ 0 r~ ~o ~0 · . ~ ~ ~O - O~D U~_ __ ~ ~o. ~_ -O S ~ ^tD _ ~1 ~ ~- __ ~ _ _' ~ ~_ a·~ -t ~-- -1 ~ ~ U. ~ O U~ · · ~ ~· ~ O ~ O ~ ~- ~ · . - ~ c~ ~- ~ v ~ f _ ~ -- _ co ~a -u ~tn s c~ c~ · u, ~o `: v ~ ~ o ~ ~ ~·- ~o o ~'m a ~1 1 1 a~ c ~-~ ~u ~_~ ~ _ ~ u~ U, 1 · 1 ~V C) ~Y O ~ - , ~ ~'  · · - ~· ~1 · 1 ~ O ~J ^ a 0 0 ~v ~ a' ~ ~ ~ ~ ~. - ~ - ~ - ,\5 o ~3 a~- ~: C) ~· ~ ~ a, v ~ ~- ' :, O ~' ^ · o. - _ :^ ~ 3 S:: ~ O (D · ~ ~ ' ~ · CO s ~: ~ ~ ~ - U) I: - C - ~ - <D .,4 ~ ~ ~ o4 · eq ~ ~ _ . ~ 3 4,) u' ~ ~ ~ r" u~ V ~_~ Ld C: ~ _~ 1 O ~ ~ - - - U~ C ~C) .~ · ~ - (V ~tt, o 3 a) ~ u,- · tn :' .~ ~ _ o _ _3 ~1 ~' ~ ' E~ o ~ _ - V _ . s~ o ~ ~ ~ ~ _ s 4~ 0 ~ ~ a) 0 _ ~ ~ ~ ~ ~ ~ ~ ~n 0 < ~0 ~o ~m ~-~ - 0 a' · ~,,, · - ,o4 ,- tD0 0 ~ ~_ ~ 0 eQ - ~ ~ o ~· ~ _ ~ Ll S - ~ O - _~ _ r~ e, _I ~ ~ a)-~1~ ~ _ o · ~ U ~t- `: Z ~ 1- _ _ m~ :~ O ~c ~_ ~ ~ x r~l u, ~ ~ ~ r~ q r ~0 r~ 0 ~ C t~ _ ~U, >1 JJ ~ ~ ~ 1" r~ _ _ C-) s ~0 ~ r~ ~ ~0 e ~ ~_ ~ E! 0 ~ ~ - ~ r~ O ~ ~ O ~ 1 ~L) z r~ v O _' ~ ~ ~0 ~· ~ ~ ~ ~ r~l s · ~r~/ ~ 3 0 r~ .0 V ~ · o ~4 ~ ~ r~ t- ~_ _ ~ ~ r~ ~ ~ JJ _ 34 - ° _ _ t- U ~.4 ~4 3 ~ a' ~ O ~ . ~=~ r~ r- ~64 ~ ~· @~ ~ ~ 3 _ · O · O ~' r ~O ~' ~) O ~' J~ `,) ~ ~ ~ ~ r~l t~ c ~o 1 - 1 X _ cc, _ 0 -~ ~ - 0 ~ ~ tn ra 0 0 ~t~ ~ O ~r~ ~' co r~ O ~O ~ ~ ~ _ r~ r ~r~1 ~· ~u r ~ a~ ~s~ r 5 _ r~ ~r~1~ O 0 0 tv ~ `4 v r~l m v ~ ~ O ~ 3 ~s ~ ~ ~ 1p ~1 _ t~, 0 _ u, t) ~ r~l ~ ~ ~ :> ~0 ~ ns 0 _ 0 ~ ~ t- ' ~ ~ ~ = m ~ 54 ~0 E! 0C V _ v 03 r~l r ~ O ~0 rO 3 0 I_ ~ tO U) 0 0 _I J ~l ~3 r ~C r ~ E ~ rO 0 ~O B ~tJ ~5 41) 0 ~4 ~U1 P ~ltS -I r~1 ~Q tU m ~ ~ ~r~  V ~S 18 0 05 r~l rO 0 ~rO ~0 ,= ~(L) r~ _ ~-.~ ~ ~O ~ (U C.) O ~O r ~O O ~0 0 0 ~O rO ~I J) O O ~S V [Q ·Q ~4 54 0 C) ~ ~-~ ~ ~ ~ P4 .:1 ~ ~ V ~ dJ 0 0 ~ 0 -~-S ~ r~ Q~ U] ¢: E ~ ~ ~ ~ ~ ~ s ~ ~ s ~ ~ ~ ~ ~r~ _ · - · ~ ~ C O -~ O ~ ~S ~ ~ ~| ~ ~ ~ ~ Ch r ~C) U) C) 0 O C~ ~· ~ ~ t~ r~ r~ ~ ~ S V V C r~1 r~ V 0 ~ O E 0 ~ 0 t~ ~ O3 ~ 0 ~ ~ r~l ~ t- 3 3 O ~ 0 r~ C.) Pt ~ m ~ ~ ~ £ r~ ~ ~ E~ ~ 45 ~ 0~ 0 cn C) C ~ ~ O ~ 61 O ~ 0) r~l J V r~l ~ od P4 C) £ t) E~ Z ~ ~ _ -~ ~ __

407 of this. Likewise, if eggs or larvae of a species are highly concen- trated in the area of a large spill or chronic discharge, individual and population responses may be greater, and recruitment may become affected (Food and Agricultural Organization, 1977~. Such an area is exemplified by the under-ice environment of the poles, where the under-ice communities appear very vulnerable to spilled oils (see Polar Environments section). The main routes of zooplankton contamination by oil are direct uptake from the water, uptake from food, and direct ingestion of oil particles. Metabolic capabilities to metabolize and detoxify hydro- carbons appear to vary somewhat among zooplankton. Thus in scyphozoans and ctenophores they are discharged chemically unchanged, while in crustaceans and ichthyoplank ton they are discharged as metabolites. Depuration of the zooplankton tissues is often incomplete, even after 1-month exposure to clean seawater, as in some copepods, euphausiids, and shrimp larvae (R.F. Lee, 1975; Lee and Anderson, 1977; Lee and Takahashi, 1977~. Adverse sublethal effects have been observed among oil-exposed zooplankton at concentrations often well below 1 mg/L (total measured hydrocarbons). The exposure times involved varied from several hours {behavior, physiological perturbations), to days (developmental abnor- malities), to weeks (growth, developmental, and reproductive problems). Sublethal responses of many important groups of zooplankters have not yet been adequately studied, or given any attention at all, with the bulk of the studies focused on a few groups. The high sensitivities of certain developmental stages, especially before and during fertilization and during early embryonic developments hatching, and larval phases, suggest that substantial damage may occur to localized populations of zooplankton during oil spills. Indeed, some acute damage has been observed at spills. Particularly vulnerable would be those stages that float or swim weakly near or on the surface. Less vulnerable would be the more advanced stages and adult zooplankton capable of diet migration. Much of the susceptibility of zooplank ton populations will depend on the persistence of oil in the water column after a spill or under chronic discharge conditions. For example, the susceptibility of local populations of surface-dwelling zooplankton to significant sublethal damage would be high if total hydrocarbon concen- trations greater than 0.05-0.3 mg/L persist for days or a few weeks. Lethal effects would be expected if concentrations between 0.5 and 1.O mg/L persist. However, such concentrations rarely persist for longer than a few days following a spill, and only in limited areas (see Chapter 4 and Appendix A). The recovery of an oil-impacted zooplankton community is probably fairly rapid, mainly due to recruitment from other areas and to intrinsic biological character istics of wide distr ibution, large numbers, short generation times, and high fecundity. Little is known of the physiological recovery of oiled zooplankton, but this may be of little importance in open water in view of the other recovery factors. In more enclosed waters, however, where recruitment from outside becomes less important, these intrinsic factors may well become limiting to recovery. Chronic pollution may also pose long term problems.

408 TABLE 5-6 Experimental Conditions of Three Field Enclosure Systems or Mesocosms MARL, Loch Ewe, CEPEX, Saanich Inlet, Condition Narragansett Baya Scotlan ~ B.C., CanadaE . Volume of 13.1 m3 304 m3 60 m3 (?) water column Number of columns 14 3 Mixing yes no no Flow-through/ 30 days no no turnover time Oil No. 2 fuel North Sea Prudhoe Bay crude (Forties Field) No. 2 fuel Concentration Mean=181 ppb 100 ppb 10, 20, and 40 and 93 ppb in ppb initial two experiments Several additions concentration over 2 weeks to Oil-water reach initial dispersion 2/week concentration; for 163-122 declined to days 25 ppb over 5 Single additions weeks Total added 414 and 153 mL 30 g Major effect Reduced zooplankton Loss of calanoid Changes in biomass, and a t 190 ppb; copepods due to zooplank ton and Reduced benthic increase in phytoplankton fauna at 90 and predatorse communitiesf 190 ppbd . AElmgren et al. (1980). tDavies et al. (1980). SR. F. Lee et al . (1977, 1978a,b) . ~Oviatt et al. (1982). ~Giesy (1980 ) . tR. F. Lee and Takahash i (1977) . Studies in Field Enclosures (Mesocosms ~ Difficulties in studying the effects of oil on biota in the water colu~, in open water, and the problems in o~rercoming extrapolation from the laboratory results to the field, have led to the development of several "mesocosm" experiments. These have been in the form either of large tanks situated on land and inoculated with a volume of nearby bay water (e.g., Gordon et al., 1976; MERL, Narragansett Bay) or have consisted of large, suspended bag systems placed in the f ield {Loch Ewe, Scotland; CE:PEX, Saanich Inlet, B.C., Canada) (Table 5-6) . These enclosure experiments allow simultaneous testing of the effects of relat~vely Jow levels of oil on both phytoplank ton and zooplankton and their interactions. They also permit simulation of both the effects of transitor fly elevated concentrations in the water after an o'1 spill by using single oil additions and the effects of

409 chronic low level additions to mimic the input from refineries, terminal activities, sewage effluents, and drainage in the coastal zone. The advantage of the mesocosm approach is that this scale lies midway between the laboratory setup and the open field. Mesocosms allow one to run controls and to have known concentrations of pollutants and yet still retain many or most of the species interactions, animal-sediment interactions, etc., which occur in natural systems. The strength of the mesocosm studies has been the ability to work at low concentrations (less than 100 ug/L) and to use environmentally realistic marine population assemblages. With these various schemes, detrimental effects were found at approximately 90 ug/L in the MERL studies, the lowest concentration tested there (Elmgren and Fr ithsen, 1982 I, and at 20 ug/L in the CEPEX exper iments {R. F. Lee and Takahashi, 1977) . The effects on the enclosed plankton and microbial communities differed from experiment to experiment, but in each instance there were profound changes in the balance between species in the zooplank ton and phytoplank ton communities and in the biomass of the compartments. These changes included shifts in dominant species (e.g., a sharp decline in the abundance of the dominant diatom Ceratulina bergonii, CEPEX), radical changes in phytoplankton species composition and marked increase in biomass, possibly due to reduced grazing pressure by zooplankton and benthic filter feeders (MERL), decimation of a calanoid copepod population and increased numbers of predators (siphonophores and ctenophores) (Loche Ewe) (Giesy, 1980~. The Loch Ewe experiment also resulted in adverse developmental effects in copepod eggs and nauplii larvae (Figure 5-3~. Clearly, the results of these three mesocosm experiments were not identical, and replication of experiments in any one mesocosm showed low precision. The results probably reflect the starting components, conditions of the various systems, and the feeding links between the systems' components. At the same time these large scale studies demonstrated that communities can experience dramatic changes and shifts due to low level oiling and that a wide range of changes can be expected at surprisingly low concentrations of oil in water. The experiments also demonstrated the usefulness of this approach and provided a much-needed intermediate step between the laboratory and studies of spills of opportunity. Macrophytes--Intertidal and Subtidal Little is known about oil effects on macrophytes, including the intertidal species and the larger offshore species. What there is has come largely from observations made during spills, with very little direct follow-up work and virtually no laboratory studies. The habitats involved here include both high energy rocky shore- lines (e.g., Fucus spp.) and low energy mud flats and salt marshes (e.g., Spar tina spp.) as well as the nearshore subtidal environment (Laminaria spp.~. Of these, the plants growing in the intertidal environment are most vulnerable to oiling and suffer the more severe impact during a spill. ThuS, during the Arrow spill in Chedabucto Bay,

410 NAUPLII 1980 ~ 3 O ~ ) ~ ~ an,, ~ .. . . . . . . . . . ~ . . .. .. ...... ~ ~DATE 3. aL 0a 23.4 ~ · S CALANOIDS 1980 lo _ 283 44 ~ t4.S "7E t6 ~ aL on 23.4 ~ 1~ FIGURE 5-3 Schematic diagram of the offshore "mesocosm" study at Loch Ewe, Scotland, and changes in zooplankton populations observed with the introduction of petroleum. SOURCE: Adapted from Davies et al. (19811. Nova Scotia, masses of oil and algae, mostly Fucus serrates, washed ashore (M.L.H. Thomas, 1973), presumably having been broken away from their rocky substrate by the weight of the oil clinging to the fronds. Similar losses of inter tidal algal cover have been descr ibed for the Amoco Cadiz spill where much of the inter tidal zone was cover ed by newly arrived oil slicks for periods up to 2 or 3 weeks {e.g., Hess, 1978~. However, recovery of intertidal algae appears to occur quite readily. Thus, investigations of the impact of Amoco Cadiz oil on Fucus species from nine oiled sites along the north Brittany coast showed growth rates to fall within the normal expected ranges (Topinka and Tucker, 1981~. Some decrease in population density was observed in the red alga Chondrus crispus, economically important for its carrageenan content, at two heavily oiled sites fully a year after the spill (Kaas, 1981~. However, such drops in density are not that unusual, and the intrinsic growth of the species was not affected.

411 Impact on more subtidal or submerged flora, while they are perhaps less vulnerable to oiling due to the depth of the water column, none- theless does occur . For example, a chronic oil situation, resulting from a grounded troop ship, the General MaCe Meigs on the northwest coast of Washington, provided observations on oil damage to a wide range of mar ine flora (R. C. Clark and Finley, 1973) . Damage included loss of fronds in the subtidal Laminar ia andersonii and bleaching of tissues, sometimes complete, in several red algae and in the false eel grass Phylospadix sp. These observations have led to some concern for potential oil impact on sea grass communities, especially where they represent either a commercial/economic significance (as in Southern California or North Brittany) or where they provide the basis for a valuable and complex invertebrate and vertebrate community, as in Florida (e.g., Zieman, 1982~. Data on oiling impact of such submerged sea grass communities are scarce. Where it has been observed and evaluated, the impact has been varied, ranging from little (Amoco Cadiz, Den Hartog and Jacobs, 1980; Topinka and Tucker, 1981; Maurin, 1981) to severe with high mortalities (Zoe colocotronis, Nadeau and Berqouist, 1977). It would seem that the impact depends in part on the depth of the water column, the type of oil released, and the local mixing conditions. One seemingly positive consequence of oiling is the often-observed Proliferation of certain algal species, such as EnteromorPha so. . Ulva sp. , and some Porphyra spp., following a spill (e.g., J.E. Smith, 19681. However, such proliferation is invariably a direct result of the elimi- nation, by the oil, of the naturally occurring grazers-limpets and other intertidal herbivores (e.g., North et al., 1964; Southward, 1982~. With the elimination of these consumers other algae proliferated, and by virtue of their spreading green cover gave the appearance of rapid recovery by the oiled vegetation. In fact, long term recovery of oiled macrophytes is variable, depending in part on their intrinsic tolerance. Both laboratory tests and field observations suggest that the various species of bladder wrack (Fucales) are resistant to moderate levels of petroleum and to brief exposures (Ganning and Billing, 1974; Ravanko, 1972; Notini, 1978; Percy, 1981~. There has been one suggestion that this resistance may be due to the inability of the oil to cling to the mucilagenous coating of the plant wall {Nelson-Smith, 1973~. Another possibility is that the wracks are not rooted in sediments, which absorb oil readily, but live attached to a rocky substrate above the oiled sediments and are therefore subject to oil concentrations much lower than found in the water (Vandermsulen and Gordon, 19761. While generally speaking the reestablishment of a floral community may be quite rapid, especially following cleaning, imbalances in the recover ing community can persist for several years . Final return to a stable floral community similar to nonoiled communities probably requires a decade or more (M.L.H. Thomas, 1978; Southward and South- ward, 1978) . In one instance, the Arrow spill, one species of bladder wrack, Fucus spiralis, that had disappeared during the spill in 1970 had not yet reappeared at the time of a 1976 follow-up survey (M.L.H. Thomas, 1978~.

412 One interesting feature of oil impact and biological recovery is delay in the decline of certain species following the initial oiling incident. Thus some species do not necessarily exhibit maximum impact until a year or two after the spill, e.g., Spar tina sp. (M.L.H. Thomas, 1973~. Too few data are available to date to determine if this is a significant feature of macrophyte recovery following oiling, but if so, this could have some significance for assessing spill impact. One area that has been totally neglected is that of spill impact on tropical or warm-water microphytes and macrophytes. Of particular concern in this respect are the giant algal flats that constitute important components of tropical trophic systems and barr ier reef systems. These are highly vulnerable to oiling because of their shallowness. Except for a single follow-up study (Lopez, 1978), nothing is known of their recovery potential following oiling. Benthic and Intertidal Invertebrates Considerable work has been done studying the effects of oil on the macroinvertebrates, mainly the intertidal species, from both physio- logical and population viewpoints. It is in this group of organisms also that most advances have been made in determining much more precisely the relationships between oil-hydrocarbon concentration, exposure time, and toxic responses, mainly because of the growing realization that a description of effects without mention of these relationships is of little value. The intertidal species have been the subjects of choice: they are obvious victims of oil spills, several are of economic significance, and several serve well as experimental subjects in laboratory studies. Hence, a good deal of information has been assembled on bivalves {Mya arenaria, Mytilus edulis, Crassostrea sp., Littorina sp.), crustaceans (crabs, lobster, amphipods, and isopods), and annelids (Arenicola spy.. Results of these studies can be found in several reviews and sources published over the last 5 years (Wolfe, 1977; Malins, 1977; Fisheries Research Board, 1978; Neff, 1979; Sprague et al. , 1981; Sampson et al., 1980; Connell and Miller, 1980) . Generalizations about the vulnerability of the benthic and inter- t idal invertebrates are difficult because there exists a great deal of variation among the genera and species, as well as among the various life cycle stages of any one species (Figure 5-4~. However, intertidal invertebrates, although highly vulnerable to oiling, may exhibit some slightly greater tolerance to petroleum hydrocarbons than do offshore benthic or pelagic species. Also, juvenile and molting stages tend to be more sensitive than the mature adult stages. However, these two statements must be made with considerable care, and each instance should be interpreted in its own context. Vulnerability Most susceptible are those species inhabiting the intertidal zone, especially those found in lagoons, embayments, estuaries, marshes, and

413 10 LO An 6 Q ~ 4 o J I CM CM __ 0 8 U i m Cal - ~N -a.) ·= a) ~- _ E~ (in o E 0 Lo Cal CM a)  BROWN SHRIMP ( PENAEUS AZTECUS ) _` ~ ^ o0 `_ . a, ~ ~ A_ ~ a' ~ _ in - _= WH ITE SHRI M P ( PENAEUS SETI FERUS ) an a) a, n An a) >I AL J - ~ r 1 GRASS SH Rl M P ( PALAEMONETES PUGIO ) in a) O POLYCHAETA  ( NEANTHES ARENACEODENTATa ) FIGURE 5-4 Toxicity of No. 2 fuel oil to life-cycle stages of selected mar ine shr imp and polychaetes . Or segmen t number . Life-cycle stages are indicated by size SOllRO:: Adapted from Neff et al. (1976) and Rossi and Anderson (1976) .

414 tidal flats. This risk der ices from two factors: high oil concentra- tions and shallow depth of the water column. Oil coming ashore during spills becomes highly concentrated in a very narrow band along the shoreline, while the shallowness results in high concentrations through- out the water column down to the sediments. These factors combine to raise the hydrocarbon concentrations in the intertidal zone beyond the tolerance thresholds of most organisms. These same low energy areas are also prime targets for chronic pollution, either from catastrophic spills or from chronic input, for the fine sediments characterizing these environments tend to act as sorbents or "sinks n for the spilled petroleum components (e.g., Blumer et al., 1973; Sanders et al., 1980; Vandermeulen and Gordon, 1976~. Besides the physiologically toxic effect of oil, there is also a physical hazard of spilled oil, and in this respect, intertidal organ- isms are especially vulnerable. For example, large numbers of inter- tidal invertebrates were killed during the Arrow Bunker C spill simply by smother ing (M.L.H. Thomas, 1973~. Sessile species such as barnacles are easy prey to smothering by heavier oils (Straughan, 1972), and the more mobile invertebrates become immobilized and thus more susceptible to either the toxic effects of the hydrocarbons or fall to predators, as was also observed during the Arrow spill (e.g., M.L.H. Thomas, 1973~. Following a Bunker C spill in Puerto Rico, oiled tree crabs were found stuck to the mangrove branches by dried oil (Gundlach et al., 1979~. Intertidal species also are exposed to volatile components of spilled petroleum. To what extent this happens in the field is not known , but laboratory studies have shown that such volatiles also can readily narcotize Littorina sp. (Straughan, 1979), rendering them susceptible to being washed off the rocky intertidal, either being washed out to sea or becoming easy prey. This is not to say that benthic organisms living subtidally are safe from oil impact, protected by an overlying water column. With vertical mixing of hydrocarbons into and throughout the water column, the impact can be felt subtidally, as was learned dur ing the 1978 Amoco Cadiz spill, in which numerous subtidal razor clams and heart urchins were killed during the first weeks following the spill (e.g., Hess, 1978) (Figure 5-5~. In a 1-year follow-up study, Cabioch and colleagues have monitored the disappearance of a benthic amphipod species offshore from North Brittany, correlating with increasing hydrocarbon levels in 30 m (Cabioch et al., 1980; Beslier et the benthic sediments down to al., 1980~. Also Addy et al. (1978) have demonstrated that chronic oil pollution in relatively deep, offshore waters can have deleterious effects on benthic invertebrates in a limited area around an open bottom oil storage structure. In contrast, benthos at the Arrow were contaminated but apparently unaffected (Anon, 19701. Speculation that arctic or polar species may be more sensitive to oiling than more temperate species has been made in the literature. There is, in fact, no evidence to date that intrinsic differences in oil sensitivity exist between polar and temperate species. About the only generalization that can be made is that polar organisms exhibit the same variability in sensitivity to petroleum hydrocarbons, over much the same concentration range, as is found at other latitudes

a I_ ~ - - ~ ~ - ~Ei3 _ ~- Fly ME ~ Act, .. a FIGURE 5-5 Subtidal invertebrates such as sea urchins and razor clams proved to be highly vulnerable to the oil spilled from the Amoco Cadiz tanker, off the coast of France, 1978. Shown here are numbers of dead urch ins and clams typically found in 1 m2 areas during the days following the spill. SOURCE: Hess (1978~. (Percy, 1981; see also Polar Environments section). However, there is one physical factor, that of reduced evaporation at the lower polar temperatures and therefore slower loss of the toxic lower-molecular- weight compounds, that may make a difference regarding the potentially acute impact of oil spilled in the Arctic or Antarctic. A similar effect presumably occurs in temperate regions during the winter. Lethal and Sublethal Effects Sublethal effects are seen in all phyla examined, ranging from alterations in respiration, growth, reproduction, and behavior to the more specific processes of calcification, molting, ion transport, and enzyme function. Much of the work so far has come from laboratory studies, which enable precise control of various experimental con- ditions. When the 1975 NRC report was written, the concentrations of hydrocarbons used were often unrealistically high. This excess is now being corrected, and more recently, attempts are made generally to use realistic concentrations in such experimental studies. There are very

416 few studies of metabolic or physiological perturbations, due to oiling, that have been made at actual oil spills or at oil pollution sites (e.g., Fisheries Research Board, 1978; Centre National pour l'Exploita- tion des Oceans, 1981) . This omission is unfortunate in that such studies can suggest new approaches or new areas of research. In an Of for t to close this gap in our knowledge, a number of simulated f ield experiments have been des igned and carried out during the last few years (e.g. , Roesi jadi et al. , 1978; Roesi jadi and Anderson, 1979; J.W. Anderson et al., in press; Bieri and Stamoudis, 1977; R.F. Lee and Ryan, 1983~. The broad range of effects that oil or its components elicit can be seen from studies done with one typical benthic invertebrate, the soft-shell clam MYa arenaria. They include mortality due to smothering (M.L.H. Thomas, 1973, 1977~; long term mortality (M.L.H. Thomas, 1977~; altered population composition (Gilfillan end Vandermeolen, 1978~; altered metabolic rates and feeding rates (J.W. Anderson et al., 1974; Avolizi and Nuwayhid, 1974; Stainken, 1978~; reduced filtration and C-assimilation rate (Gilfillan et al., 1976, 1977b; Gilfillan and Vandermaulen, 1978), reduced survival, condition index, and levels of tissue amino acids (Roesijadi and Anderson, 1979~; changes in tissue structure (5tainken, 19761; restriction in growth (Gilfillan and Vandermeulen, 1978; M.L.H. Thomas, 19781; and alteration in shell formation (Gilfillan and Vandermeulen, 19781. These, and other effects, not yet looked for in the soft-shell clam, are found as well in other invertebrates exposed to oil or to oil fractions. Included are effects both at the subcellular level, e.g., genetic alteration (Cole, 1978) as well as long term perturbations at the population and community levels (e.g., Elmgren et al., 1981~. An apparent enigma observed at both spill sites and in experimental studies is the puzzling survival of some benthic macroinvertebrates in highly oiled sediments. For example, several studies have noted the lack of significant mortality during exposures to concentrations of oil in sediments in excess of 1,000 ug/g (Wells and Sprague, 1976; J.W. Anderson et al., 1977; Roesijadi et al., 1978; Vandermeulen and Gordon, 1976; Gordon et al., 1978) . Krebs and Burns (1977) remarked on the adverse responses of Uca sp. apparently surviving in sediments contain- ing 1,000-7,000 ng/g petroleum hydrocarbons. Similarly, Shaw and Cheek (1976) made measurements on the clam Macoma balthica from sediments containing 640-3,890 Gig (dry weight) hydrocarbons. File this phenomenon is by no means fully understood, there are several aspects for which some answers now exist. First, the very toxic lower-mo~ecular-weight components (see Toxicity section) rarely persist in highly oiled sediments in the field. Second, the feeding habits of certain of these invertebrates may preclude uptake directly from contaminated sediments because they feed instead on the lesser contaminated waters above the sediments. Third, adult stages generally show a relative resistance to the hydrocarbon toxic effects. Another possibility is that of influx of mature organisms into these oiled areas from nonoiled "seed populations." In this context, J.W. Anderson et al. {1978) describe near-normal recruitment of a range of macroinver- tebrates (polychaetes, bivalves, gastropods, crustaceans) into experi

417 mentally oiled sediment plots set out in trays in the field. On the other hand, judging from field studies of biological recovery, larval recruitment under such high oil concentrations appears to be very sen- sitive, frequently with poor survival (e.g., Gilfillan and Vandermeulen, 1978; M.L.H. Thomas, 1978; Sanders, 19781. Effects at the Population Level There is ample evidence that oil and oil spills can perturb entire invertebrate populations. As alluded to above, Mya arenaria popu- lations in an oiled lagoon of Chedabucto Bay, 6 years after the Arrow spill, still experienced recruitment problems (M.L.H. Thomas, 1978; Gilfillan and Vandermealen, 1978~. In another study, isozyme patterns in a population of snails, Urosalpiox cinera, 6 years after the Wild Harbor marsh oiling, showed a persistently greater variation than that found in nonoiled populations (Cole, 1978), including a persistent imbalance in genetic patterns. Long term recruitment inhibition and low population densities in the salt marsh crab Uca pugnax were found 8 years after oiling (Krebs and Burns, 19771. Similar patterns of long term impact on benthic invertebrates are beginning to unfold following the more recent spills, e.g., Amoco Cadiz (Cabioch et al., 1980; Beslier et al., 1980; Glemarec and Hussenot, 1982) and Tsesis (Elmgren et al., 1981~. Fish Fish can be affected directly by petroleum, either by ingestion of oil or oiled prey, through uptake of dissolved petroleum compounds through the gills and other body epithelia, through effects on fish eggs and larval survival, or through changes in the ecosystem which support fish. In addition, the commercial fisheries as well as the sports and subsistence fisheries can be affected through contamination of gear, closure of fishing seasons, or buyer resistance to tainted or suspect products. The potential impact of spilled oil on f ish populations has stimu- lated many studies to determine the lethal and sublethal effects of oil on fish. Most of the studies to date have been done in the laboratory, mainly because both the test organism and the test conditions can be controlled. Field studies are far fewer. As a result, the effects of oil on fish in the field are only poorly understood. Natural stresses may predispose fish to increased sensitivity to hydrocarbon toxicity. These stresses may include changes in salinity, temperature, and food abundance, as well as competitive or potentiating stress from other pollutants. For example, Moles et al. (1979) found Prudhoe Bay crude oil WSF as well as benzene, although at admittedly high concentrations (>1 mg/L), were twice as toxic to several salmonid species in seawater as in fresh water, this at the stage of their life cycle when they would normally migrate to seawater. Lev~tan and Taylor (1979) found that euryhaline killifish acclimated at extremes of

418 salinity were less tolerant to nap}~=alene than fish exposed in dilute seawater . Rainbow trout also, when acclimated to seawater, showed greater susceptibility to crude oil-in-water dispersions (Englehardt et al., 1981 ~ . Temper atur e as a natur al var table becomes impor tent in r elation to hydrocarbon pollution, affecting its toxicity by affecting the persis- tence of hydrocarbons in water and by imposing physiological stress on the fish at both ends of the temperature range. The interaction of these stresses can be complex and severe. For example, in colder waters, hydrocarbons will tend to persist longer in the water column, while at the same time, hydrocarbon metabolism and clearance rates in the fish may become reduced. This would explain the increased sen- sitivity of pink salmon exposed to toluene at low temperatures (Korn et al., 1979) e Salmon exposed to toluene at low temperatures had greater stress responses than salmon exposed at high temperatures (R.E. Thomas and Rice, 19791. Also, the tissue burden at lower temperatures may persist longer than at the higher temperatures. It would appear that fish, in comparison to invertebrates, might be more sensitive to short term, acute exposures, requiring relatively shorter exposure periods to absorb lethal quantities of hydrocarbons. The sensitivity of fish to oil WSF varies widely. Pelagic species seem to be more tolerant than benthic species, while fish inhabiting intertidal areas appear to be the most tolerant (Rice et al., 1979~. Of these, the benthic species are of special concern because of their lifelong association with benthic sediments, which are known to become hydrocarbon traps in the event of oil spillage. L i fe-Cycle Stages The most vulnerable portion of a f ish life cycle to environmental/ pollution stress probably occurs during the development of the germ-line primordia, early embryo (pregastrulation), and especially the larval transition to exogenous food sources (Rosenthal and Alderdice, 1976) (Tables 5-7 and 5-8~. And indeed, from the various data available on the vulnerability of gametes, developing eggs, larvae, juveniles, and adult fish (e.g., Kuhnhold et al., 1978; Strohsaker et al., 1974), some stages of eggs and most stages of larvae probably are more vulnerable to oil than are juveniles and adults. There are, however, exceptions and nuances to the generalization that all eggs are sensitive. For example, Moles et al. (1979) found salmon eggs (postgastrula) to be extremely tolerant to benzene and oil WSF. ALSO' fish have a variety of reproductive strategies that cause further variations and differ- ences; egg incubation times vary from days to a year, and yolk content var ies tr emendously . Damage to an embryo may not become apparent until after hatching of 'che eggs into the larvae. For example, Kuhnhold (1978 ~ found that the effect of oil on the eggs was far worse if one used larval deformities as an index instead of egg survival. Many of the larvae, after hatch- ing, was found to be deformed and incapable of swimming. These abnormalities would not be included or recognized if only the survival

419 TABLE 5-7 Ef feats of Oil on Eggs and Embryos of Mar ine Fish Threshold Concentrations for Type of Response Response tmg/L) 1 Reference Le thal Increased mortality 5.1 (TH) (l-d LC50) 1.3 - 2.2 (TH), 0.28-0.47 (TN)15 <0.68 (TH) 0.1->44 (TH) - 2 (4-d LC50)9 0.02 - 0.2 (TH) (6-d LC50) 0.01-0.5 (TH)17 1.5(TH), 0.4(TN) (4-d LC50) 21 17,18 Suble Uhal Fertilization decrease >14.1 (TH) 14 Uptake, accumulation of hydrocarbons Contamination char ion <0.25 (TH) 17 Uptake, metabol ism, and retention 0.01-2.1 (B) 5 Phys iology Hear tbeat changes 0.01-0.1 (TH) 10 <3.1 - 11.9 (TH) 14 5.0 - 6.0(TH), 1-1.2(TN) 1 <2.1(TH), 0.54(TN) 21 Respiration affected <2.1 (B) 2.4 - 8.0 (TH) 3 Behavior E0ryon ic movemen ts DeveloE~nent Pathological tissues Morphological anomalies Rate changes Growth Length, h e igh t changes Tissue Ha tch ing Occurrence affected Premature or delayed Success decreases Duration changes <3.1-11 9 (TH) 1.1-2.2 (TH), O .6-1.1 (TN) <0.1 (TH) 0.0001 (B (a) P) <1.7 (TH), 0.38 (TN) 2.1 (B) 14 6 11 29 16 4 <1.9 (TH) 26  2.2-4.4 (TH), 1.1-1.3 (TN) 6 <0.1 (TH) 11 <3.1-11.9 (TH) 14 <2.1 (TH), 0.54 (TN) 21 <0.8 (B) 24 0.01-0.1 (TH) 11 S-6 (TH), 1.0-1.2 (~) 1 0.0001 (B (a) P) 29 <1.3 (TH), 0.28 (TN) 15 >0.68 (TH) 22 NOTE: TH, total hydrocarbons; TN, total naphthalenes; B. benzene B (a ) P. Benzo ( a ) pyr ene. REFERE~CES: 1. J. . W . Ander son r'enn, ~ Davenport et al. 5. Eldr idge et al . (1978) . 6. Ernst et al. (1977) . 9. Kuhnhold (1974). 10. Kuhnhold (1978). 11. Kuhnhold et al. (1978). 12. Linden (1975). 14. Linden (1978). 15. Linden et al. (1979). 16. Linden et al. (1980). 17. Longwell (1977). 18. Longwell (1978). 21. Sharp et al. (1979) . 22. R. L. Smith and Cameron (1979) . 24. Strohsaker (1977) . 26. Vuor inen and Axell (1980) . 27. Whipple et al. (1981) . 29. Hose et al. (19B2). Also see Johnson et al. (1979), Kuhnhold (1972), Linden (1976), Lonning (1977b), Mironov (1967), Stoss and Haines (1979), Struhsaker (1977), Strahsaker et al. (1974). et al. (1977). 2. Cameron and Smith (1979). 4. Eldridge et al. (1977). Ernst et al. (1977). 9. Kuhnhold

420 TABLE 5-8 Su''unary of Effects of Oil on Marine Fish Larvae Type of Response Threshold Concentration for Response (mg/L) Reference Lethal Increased mortality 43(TH) (4-d LC50) Voorinen and Axell (1980) <0.8(B) Strahsaker (1977) l.S(TH), 0.4(TN) 4dLC50 Sharp et al. (1979) Sublethal Uptake, accumulation of hydrocarbons 0.01-2.1(B) Eldridge et al. (1978) Physiology Respiration changes 2.4-8.0(TH) Davenport et al. (1979) <0.85(TH), 0.21(TN) Sharp et al. (1979) Development Delays <0.68(TH) R.L. Smith and Cameron (1979) Tissue pathologies <0.1-2.2(TH) Vuorinen and Axell (1980) Morphological anomalies <3.1-ll.9(TH) Linden (1978) <0.68(TH) R.L. Smith and Cameron (1979) <8.0 (TH) Davenport et al. (1979) <3.0(TH) Eye pigmentation Irregular cytolysis Growth Length decreases <5.4-~.8(TH) Linden (1978) 0.1-2.2(TH) Vuorinen and Axell (1980) Length increases <0.68 R.L. Smith and Cameron (1979) <2.1(B) Eldridge et al. (1977) NOTE: Threshold concentrations are based on direct hydrocarbon measurements. TH, total hydrocarbons; B. Benzene; TN, total naphthalenes. of the eggs were considered. Linden (1978) a' so noted that many oil-exposed Baltic herring eggs that had appeared to develop normally ultimately failed to hatch. Typical f ish larval responses to toxic concentrations of petroleum include a brief increase in activity, followed by reduced activity, sporadic twitching, narcosis, and ultimately, death (Struhsaker et al., 1974; Linden, 19751. Morphological and physiological effects include deformed spinal column, tissue destruction (particul arly of the fish fins), and reduced growth (Linden, 1975, 19763. They seem generally to become more sensitive to oil as their yolk sac becomes used up (Kuhohold , 1972 ; Rice et al., 1975 ~ . Sublethal Effects Although fish can accumulate hydrocarbons from contaminated food, there is no evidence of food web magnification in fish. Fish have the capability to metabolize hydrocarbons and can exorete both metabolites and parent hydrocarbons from the gills and the liver. There is some evidence that metabolites can persist in the tissues longer than parent

421 hydrocarbons but their toxicity is not known (see Chapter 4, ~nverte- brates and Vertebrates section and Rates of Petroleum Biodegradation in the Marine Environment section). There are indications, however, that some metabolites or intermediates may be toxic or even mutagenic. Oil effects in fish can occur in many ways: histological damage, physiological and metabolic perturbations, and altered reproductive potential. Histological studies have documented damage to liver (McCain et al., 1978; Sabo and Stegeman, 1977), gill (Ernst et al., 1977; Englehart et al., 1981), gut (J.W. Hawkes et al., 1980), vertebrae (Linden et al. , 1980), eye lens (J.W. Hawkes, 1977), stomach (Wang and Nicol, 1977), brain (DiM'chelle and Taylor, 1978; Cameron and Smith, 1980), and olfactory organs (Gardner, 1975~. Physiological changes include increased heart beat (Wang and Nicol, 1977; J.W. Anderson et al., 1977; Linden, 1978), increased coughing (Rice et al., 1977; Barnett and Toews, 1978), ionic and osmotic imbal- ances (McKeown and March, 1978; Englehardt et al., 1981) changes in respiration (Barrett and Toews, 1978; J.W. Anderson et al., 1974; R.E. Thomas and Rice, 1979), changes in blood parameters (Fletcher et al., 1979; P. Thomas et al. , 1980), decreased energy reserves (Stegeman and Sabo, 1976; Sabo and Stegeman, 1977; J.W. Hawkes, 1977; Kovaleva, 1979), and changes in gill enzymes (Wong and Englehardt, 19821. Long Term Effects Many sublethal effects of oil on fish are symptomatic of stress and may be transient. Others may persist longer but may be only slightly debilitating. However, because all repair or recovery requires some energy, these sublethal effects can ultimately lead to increased vul- nerability to disease or to decreased growth and reproductive success, even though the individual may continue to live for some time. Several studies have indicated a correlation between hydrocarbon stress and increased vulnerability to disease. For example, hydrocarbon exposure has been found to be associated with increased fin erosion or fin rot (Minchew and Yarbrough, 1977; Giles et al., 1978), reduction in external bacterial flora (Giles et al. , 1978), and reduction in the rate of tissue repair or regeneration (Fingerman, 1980) . There also appears to be some relationship, albeit poorly understood, between hydrocarbon exposure and parasitism. For example, juvenile coho infested with parasites were more sensitive to oil, toluene, and naphthalene than were uninfested fish (Moles, 1980~. Decreases in growth after hydrocarbon exposure have been observed in severe' studies: herring embryos (Linden, 1978; R.L. Smith and Cameron, 1979), killifish larvae (Sharp et al., 1979; Linden et al., 1980), cutthroat trout juveniles (D.F. Woodward et ale, 1981), pink salmon alevins (Rice et al. , 1975), English sole (McCann et al., 1978), and coho salmon {Moles et al. , 1981~ . Several of these studies were of considerable length, and involved flow-through systems where exposures may be comparable to field exposure conditions. Growth decreases may be a highly significant factor with respect to the fisher ~ s, for the entry of fish into the fishery may be delayed.

422 Low levels of oil contamination (less than 1 ppm) can also affect the ability of individuals to reproduce, either by causing malformation of gonads or gametes or by simply decreasing the energy that the f ish has left to invest in growth. Such concentrations are found in the f ield under spill conditions, and damage to gonadal tissues has been noted at very low hydrocarbon concentrations. For example, starry flounder, exposed to 50-100 ~g/L monoaromatic hydrocarbons, showed changes in gonadal maturation (Whipple et al., 1978~. Also Kuhohold et al . (1978 ~ have demonstrated that exposure of adult f ish dur ing gonad maturation to 10-ug/L WSF can result in reduced survival of the larvae hatched from the eggs laid by these f ish. Vulnerability and Avoidance It is difficult to state definitely whether fish will avoid hydrocarbons in the f ield. Observations are few and circumstantial . When Atlantic s liver side were exposed to crude oil, they lost their normal schooling behavior, possibly because the olfactory organs and lateral lines of the f ish were blocked or damaged (Gardner, 1975) . On the other hand, the appeal of easy prey may attract f ish into an oiled area. Thus other factors in the fish life habit may override avoidance of oiled waters. For example, in a study of adult salmon returning to a home stream, about 50% of the salmon avoided a contaminated f ish ladder containing 3.2 mg/L of monoaromatic hydrocarbons, indicating that a considerable proportion of the salmon could avoid much lower concentra- tions. But it was also clear that an equally large proportion would pass through even higher concentrations that approached acutely toxic levels (Weber et al., 1981~. In contrast, pelagic eggs and larvae at or near the surface clearly are unable to avoid contaminated surface waters and are more vulnerable to oil pollution. Effect on Fish Stocks Oil has interfered with or affected the fishery in several ways. Some fish have become tainted through ingestion of oil or through contact with oil or oiled gear, and are therefore unmarketable (e.g., vander- msulen and Scarratt, 1979~. Oil spills have also interfered more indirectly through fouling of gear and closing of harbors and fishing seasons . In many instances, f ish even slightly suspect of tainting or thought to originate from oiled waters, have been found to be unmarket- able . However, a direct impact on f ishery stocks has not been observed , nor has it been looked for directly in most cases, although close Inspection is made of f ish catch statistics . The problem of tainting has been reviewed by Howgate et al. (1977), Stansby (1978), and Whittle {1978) . Tainting , or the presence of off f favor s in f ish meat, is a troublesome problem in that often the source of the off flavors is not known, and off flavors can be due to a number of causes. To make matters worse, not infrequently can different

423 sources give rise to similar off flavors. For example, a kerosene-like flavor in a catch of tainted mullet could be linked to the presence of crude-oil-derived hydrocarbons later found in the fish tissues (Shipton et al., 1970~. However, in another apparently similar case, the kerosene-like flavor in a different batch of tainted mullets arose from thermal decomposition of naturally occurring components in fish (Vale et al., 1970~. Tainting has occurred in nearly all types of oil spillage, including diesel fuel in brown trout (Mackie et al., 1972) and unidentified petroleum refiner effluent (Nitta et al., 1965), and have been fre- quently reported for spill accidents such as the Juliana (Motohiro and Inoue, 1973), the Torrey Canyon, the Ekofisk blowout {Mackie et al., 1978), and the Amoco Cadiz (Chasse, 1978~. In two instances the taint- ing compounds may have entered the fish from contaminated sediments instead of through the water column (Cornell, 1971, 1974; Nitta et al., 1965). Invertebrate tissues also are liable to tainting, including lobsters (Wilder, 1970; Scarratt, 1980), mussels (Nelson-Smith, 1970; Brunies, 1971), and clams (A.L. Hawkes, 1961~. Oysters from beds near drilling operations offshore from Louisiana were found to contain oily taints, correlating with the proximity to the drilling operations (Mack in and Sparks, 1962; St. Amant, 1958; Menzel, 1947, 1948~. The precise route of entry of the tainting substances is uncertain but probably includes both the respiratory process and uptake through tainted food. Little is known of the actual substances responsible for the tainting of fish meat. No unequivocal identifications of substances in crude oil and distillates exist which exactly match the taint profile in contaminated fish products (Whittle, 1978), but it is thought that the more polar components (aromatic hydrocarbons, substituted benzenes and naphthalenes, naphthenic acids, organosulphur compounds, and olefins) are involved. Representatives of these compounds have all been identified in contaminated marine products (Ogata and Miyake, 1973; Lake and Her shner, 1977; Ogata et al., 1977; Shipton et al., 1970). It seems that under natural conditions, tainting is unlikely to occur at the sort of petroleum hydrocarbon concentrations found in the water column after a spill, barring a major catastrophe such as the Amoco Cadiz. It is likely, however, in more restricted spill incidents, as in bays, or under chronic oiling conditions. There are two major factors due to pollution that severely compli- cate the detection of impacted fishery stocks. First, there are extremely wide variations in the recruitment to fisheries that are caused by both natural environmental factors and by current overfishing (e.g., Ware, 19821. Predictions of fish abundance often deviate from the observed for no apparent cause, and our present capability to assess a standing fish stock is very limited, due mainly to sampling methodology (e.g., Sinclair, 1982~. Moreover, natural environmental factors, often poorly understood, can cause fluctuations in year-class abundance which bear little relation to the size of the parent stock. It would therefore be very difficult to quantitatively assess the

424 impact of an oil spi, 1 on populations that already possess broad and unpredictable changes in year-class s ize . The second factor is that massive fish kills during oil spills probably have not occurred. Some mortalities have been observed at a number of spills, but generally only in limited areas, and then not in large amounts. Fish have the ability to move away from an impacted area, either laterally or by moving to a greater depth (whether in fact this occurs is still not known). If any large mortalities do occur, they probably occur in the egg and larval stages found in the surface waters. Being more sensitive than the adult stages, eggs and larvae may have been killed in large numbers during spills. However, such kills are extremely difficult to document, simply because these fragile life stages are difficult to sample, since the dead fall out of the water column and decompose within hours, becoming unrecognizable tissue debris. In this respect the available evidence from the Argo Merchant spill, although open to or iticism for lack of statistical corrobora- tion, suggests a k ill of cod and pollock . Marine Mammals Information on oil impact for marine mammals is still limited. There are a number of documented or inferred oil spill fouling incidents but only few definitive experimental studies. Despite this paucity of information, because of their life habits, the potential exists for many marine mammal species to come into contact with spilled petroleum in the seas. As a large proportion of the world's marine mammals spends part or all of its life in ice-infested waters, this risk seems to be especially great in the polar regions. Adding to this is the grooming habit of several species that adds petroleum toxicity through ingestion of oil. Recent reviews, including indicated research needs, have been completed by Geraci and St. Aubin (1980), Smiley (1982), and Englehardt {1983~. Vulnerability The dependence of pelagic seals and whales on air and the amphibious habit of polar bear {Ursus maritimus) enhance the possibility of unavoidable contact with spilled oil on the sea sur face . This Is readily visualized for ice-covered waters where the diving mammal can surface only in restricted areas of open water. Ice will modify the pattern of spreading and disappearance of an oil spill (Glaeser , 1971; Snow and Scott, i975), tending to move or concentrate oil in leads and breathing holes. In addition, wind may "herd. oil between moving ice floes or up against ice edges (Ayers et al., 1974~. In this way, preferred travel routes of many of the marine mammals could become highly contaminated with oil. For example, the preferred travel route for narwhal (Monodon monocerus ~ is the ice edge. The situation for bowhead whales {Balaena myst~cetus) may be particularly serious because this endangered species uses ice leads

425 extensively during its arctic migration (Braham et al., 1980~. The entire population of bowhead whales (Bering, Chukchi, Beaufort stocks), comprising only some 3 ,000-4 ,000 animals, travels from the Bering Strait to Point Barrow, Alaska, through a rather well-defined ice-lead system. Ringed seals (Phoca hispida) are another good example of a special- ized ice dependency and potential oiling vulnerability. Birth and postnatal care occur in this species in subnivean (under snow) birth lairs, with access through the ice surface (T.G. Smith and Stirling, 1975), and it has been suggested that ringed seals use under-ice air pockets on extended subice travels (Bertram, 1940 ; Milne, 1974) . In both instances, coating and inhalation toxicity may be a problem, for oil would concentrate in these areas. Also, the colonial breeding habit of most seals and of certain other marine mammals creates a particular vulnerability of these animals to oil spills within the breed ing per iod . Oiling Incidents Documented cases of oil pollution incidents involving mar ine mammals are few, even if one includes news media accounts. In most instances these are not case histories but are obituary accounts that only implicate but do not define oil as the cause of death. One of the most detailed of these occurred during the 1969 blowout in the Santa Barbara Channel in southern California. Nelson-Smith (1970) causally related this event to the deaths of gray whales (Eschrichtius gibbons), a dolphin, northern fur seals (Callorhinus ursinus), California sea lions (Zalophas californianus), and northern elephant seals (Mirounga angustirostris). However, the mortality link was subsequently inter- preted as tenuous (Simpson and Gilmartin, 1970; Brownell and Le Boeuf, 1971; Le Boenf, 19711. A number of other incidents have been associated with deaths of marine mammals. These have included gray seals (Halichoerus grypus) on the coast of Wales (J.E. Davis and Anderson, 1976) and seal mortalities during the Torrey Canyon (Spooner 1967) and Arrow spills (Anon, 1970~. Harp seals (Phoca groenlandica) were found dead and coated with Bunker l C oil after a spill in the Gulf of Saint Lawrence, Canada (Warner, 1969~. Gray and harbor seal (Phoca vitulina) mortalities were associ- ated temporally with coating by oil from the Kurdistan Bunker C spill (Parsons et al., 1980) (Figure 5-6~. Other reports describe oil fouling in various species without being identified with a spill incident or necessarily with mortality. Oil-fouled individuals have been reported in the Arctic (Hess and Trobaugh, 1970 ; Morr is , 1970 ; Muller-Willie, 1974), in the Antarctic (Lillie, 1954), and in European waters (J.L. Davis, 1949; Spooner, 1967; Van Haaften, 1973~. Oil Adherence to Body Sur faces The extent of oil adhesion to the skin or pelage of marine manuna~s depends at least on the following : texture of the exposed body sur face,

426 ~:i I-~ . _! _< ..~ ; . . ~ ~ ~ ~,, ~:.: ~., f ~ I . *- Air >_7-ran ~ ,; ~ it's-' ~ At', ' i A,, 4~ ji~:~ _ ~ i.' .- In. FIGURE 5-6 (Top) Oiled and (bottom) unoiled harbor seal pups seen on Sable Island, Nova Scotia, Canada, in June 1979. Causes of oiling was unknown. Note the clean ring, result of tearing, around the eyes of both animals. (Photo by J. Parsons.) SOURCE: Parsons et al. (1980~. frequency and duration of exposure, and character istics of the oil. A heavy, viscous oil would seem more likely to adhere to skin or pelage, and marine mammals possessing well-developed pelage would be expected to have oil adhere readily.

427 This assertion is supported by laboratory studies involving r inged seals (T.G. Smirch and Geraci, 1975; Geraci and Smith, 1976), sea otters (Williams , 1978), and polar bears (Or itsland et al., 1981 ; Engelhardt , 1981~. Similar evidence comes from oil-fouled animals such as harp seals (Warner, 1969), gray seal pups (J.E. Davis and Anderson, 1976), and elephant seal pups (Le Bocuf, 1971) taken from spilled oil areas. The pers istence of adherent oil var. fed among the species examined . Ringed seal pelage cleaned itself from complete coating by 1 day In seawater. Captive sea otters and polar bears groomed their oil-fouled hair readily, although this resulted in oil ingestion and subsequent illness or death. In an examination of 58 free-ranging elephant seal pups, in which initially more than 75% were coated with spilled oil, all but one were clean when examined 1 month later (Le Boenf, 19717. Marine mammals with a poorly developed or no pelage (e.g., walrus, sirenians, cetaceans) would seem to be less likely to have oil adhere to them. Thus there have been no substantiated repor ts of oil fouling of cetacean skin, although whether due to the inability of oil to adhere to the ir gener ally smooth sk in or due to avoidance by the mammals of oiled areas is not known. The sk in of most cetaceans is generally quite smooth over the entire body sur face, although there are excep- tions. These include the gray whale, with large numbers of attached barnacles, right whales with their prominent rostral callosities (Slijper, 1979), and the bowhead whale with dozens to hundreds of eroded areas, particularly involving the skin of the head (Albert, 19811. Such roughened body areas would seem to increase the likelihood of oil adherence. Spilled oil may also be expected to interfere with baleen func- tioning, for the inner aspect of the baleen plates presents a very roughened surface. This is supported from laboratory studies (Braith- waite, 1981) which demonstrated that the filtering efficiency of bowhead whale baleen was reduced by about 109 when coated with Prudhoe Bay crude oil and by 85% when coated with oil with a higher wax content. Geraci and St. Aubin {1982) reported transient changes in water flow through oil-exposed baleen from f in (Balaeno; ?tera physalus ~ and gray whales, and suggested that the temporary inhibition of baleen function would not be important in the long term feeding strategies of baleen whales. The consequences of a possibly increased ingestion of oil from this coated f Liter ing apparatus cannot be evaluated at this time . Avoidance of Oiled Water s Little is known of the capability or the willingness of marine ma~runals to detect or avoid oil-contaminated waters. Multiple observations of o il-coated seals suggest that these species do not actively avoid oil . Limited observations on captive sea otters (Williams, 1978) and seals (Spooner, 1967) also showed that they did not necessar fly avoid oil- co~rered water. In the Canadian north, free-ranging polar bears have repeatedly been reported to consume small quantities of petroleum oils from dump sites of Arctic camps.

428 A recently released report indicates that a broad range of cetaceans do not actively avoid oil. Observations in an oil slick resulting from the Regal Sword spill (Goodale et al., 1981) of Bunker C and No. 2 fuel oil showed cetaceans swimming and feeding in both oil-covered and oil- free waters. Both surface and below-surface feeding was observed in humpback whales (Megaptera novaeangliae), fin whales and in white-sided dolphins (Lagenorhynchus acutus). Another species, probably right whales {Eubalasna glacialis), was also observed in the oil slick. Their behavior was apparently normal when observed in the oil slick. Part of the gray whale migration along the California coast goes through the oil seep waters off Coal Oil Point, California, apparently in clear disregard of the contaminated waters (Geraci and St. Aubin, 19821. Yet in an experimental setting, captive bottlenosed dolphins {Tursiops truncatus) were able to detect and chose to avoid oil on the water surface both visually and apparently by echolocation (Geraci et al. 1983). Although these observations do not answer the question of detection ability, they at least suggest that under normal conditions, species observed may not actively avoid oil-covered waters. Thermoregulation and Metabolism A survey study of stranded cetaceans has shown the presence of petroleum hydrocarbons, particularly in blubber tissue (Geraci and St. Aubin, 1982). Oil uptake and excretion in marine mammals have been examined experimentally in ringed seals (Engelhardt et al., 1977; Engelhardt, 1978) and in polar bears (Engelhardt, 1981; Oritsland et al., 1981~. Major sites of accumulation of petroleum residues in the seals were in the blubber and liver, with excretion of petroleum residues and of metabolites via the urine and bile. In the polar bears, ingestion of oil was brought on primarily by grooming of the oiled pelage. There was evidence of some long term contamination of internal tissues, as shown by continued hydrocarbon concentrations in the serum some weeks after initial oiling. This may have been due to continued uptake of oil through grooming, or through transfer of contaminated fatty tissue from fat stores to excretory routes. In the polar bears, highest hydrocarbon levels were found in the kidney, brain, and bone marrow. Effects on thermoregulation and metabolic stress have been studied in some detail in the sea otter (Rooyman et al., 1977; Costa and Kooyman, 1980~. Coating of the fur resulted in major conductance changes, increasing with the degree of oil coverage. In comparison, isolated skins from seals and sea lions showed less or little con- ductance change when oiled. A consequence of the increased conductance is a greater heat flow across the body surface. Loss of heat from the body core must be compensated for by an increase in metabolism. Although core temperatures did not change, the oiled sea otters showed a 5°-10°C decrease in subcutaneous temperatures below the selectively oiled areas when they were In the water. The lack of core temperature change was attributed to a near doubling of body metabolism. Increases

429 in metabolism also showed a dose response in relation both to the amount of oil applied and to the amount of body surface covered. Failure of an increase in metabolism to compensate for heat loss may lead to hypothermia, thought to be the cause of death in the case of one oiled sea otter (Williams, 1978~. Studies with polar bears suggest similar implications from skin oil fouling. Oil coating of isolated polar bear fur led to as much as a tripling in the conductance of heat across the skin (Hurst et al., 1982 ; Oritsland et al., 1981) . This was increased even further In the presence of wind or by increasing the viscosity of the test oil. The consequences of exposing bears experimentally to oil were changes in temperature and metabolism. Subcutaneous temperatures increased as a result of reactive vasodilation in the skin, leading to a major heat loss promoted by the increased conductance. Resting deep body tem- peratures decreased slightly but metabolism nearly doubled. Such thermal and metabolic responses in polar bears can eventually lead to hypothermia, particularly in the presence of wind or in instances of poor health. Unlike the case in sea otters and polar bears, oil-coated r inged seals and late stage harp seal pups showed no core temperature response to crude oil (T.G. Smith and Geraci, 1975~. In this case the thick insulative blubber layer was thought to serve as an adequate thermal barrier. Newborn phocid seals, having little or no blubber (Engelhardt and Ferguson, 1980) may be more seriously affected, as they rely on fur (lanugo) and metabolic activity for their thermal balance (Oritsland and Ronald, 1973 ; Blix et al., 1979) . Pathological Consequences Mar ine mammals may be expected to show clinical and toxicological responses to petroleum hydrocarbons similar to those in other manuals. Aberrations in hematological parameters, in diagnostic enzymes, and in tissue structure have been recorded, although not consistently. No indications of clinical abnormalities were found in the oiled sea otters (Costa and Kooyman, 1980; Williams, ~ 978) . Ringed seals showed only a limited toxic response (T.G. Smith and Geraci, 1975; Geraci and Smith, 1976) . Twenty-four hours of oil exposure resulted in complete coating of the body, including the eyes, and caused eye damage such as conjunctivitis and corneal lesions. The eye involvement sub- sided, however, after removal from the oil and was considered reparable. Oil-induced eye damage has also been suggested by Nelson-Smith (1970) and Morris (1970~. However, eye damage is by no means uncommon in normal seal populations (King, 1964; Ridgway, 1972) and any circumstan- tial link to oil spill events should be treated with caution. Although petroleum exposure in seals, either by immersion or by ingestion, has led to petroleum uptake and to its accumulation in tissues (Engelhardt et al., 1977 ; Engelhardt, 1978), there was no evidence in either r inged or harp seals of ma jor haematological, plasma chemical, or histological/clinical changes which coul d be associated with the oiling (Geraci and Smith, 1976) . Similarly, a study of the

430 effect of hydraulic oil (Caldwell and Caldwell, cited by Geraci and St. Aubin, 1982) found no clinical pathology. From this it would seem that at least phocid seals are not likely to show a lasting toxic effect from short term exposures to crude oil, even at high doses. The same conclusion cannot be drawn, however, for long term exposures. Aside from the clinical damage, long exposures may lead to adrenal steroid exhaustion. In r inged seals for example, oil ingestion was found to lead to a greatly elevated plasma cortisol level, while cortisol breakdown rates were nearly doubled (Engelhardt, 19821. The situation is different in the case of oiling of polar bears. Exposure, and presumably continued exposure as well, to unknown quantities of oil by short term immersion and ingestion led to severe clinical pathological abnormalities (Oritsland et al., 1981; Engelhardt, 1981~. Predominant were extensive anemia, caused by peripheral hemolysis and erythropoietic dysfunction, and renal abnormalities associated with a buildup of nitrogenous metabolites and hydromineral imbalance. Adrenocortical, pulmonary, and skin changes were also evident. Renal failure was proposed as the ultimate cause of death of two bears {duck, in Oritsland et al., 1981~. Summary and Conclusion Although information of oil effects on mar ine mammals is scanty at best, the sum of the results indicates that both habit and habitat make the various species highly vulnerable to oil pollution at sea. Experi- mental studies have shown that seals, sea otters, whales, and polar bears differ in their sensitivities to oil exposure. Fur-insulated mar ine mammals respond to oil contact by a compromised ability to thermoregulate. Continued contact may result in skin and eye lesions. Both seals and polar bears can absorb oil readily, distributing it through the body tissues, including the fatty reserves. Seals showed endocrine stress responses but few other tissue problems, while polar bears were severely affected in blood and renal functions. Cetaceans were little or only transiently affected by oil exposure. Odontocetes appear to be able to detect oil under captive circumstances, but do not necessarily avoid slicks at sea. Birds Birds are probably the most conspicuous casualties of oil pollution in the sea. Death of seabirds from oil pollution receives great publicity and in several countries attracts a public reaction such as the death of few other animals does. In addition, probably because of its visual impact, the death of birds from oiling evokes an emotional reaction stronger than would their death from other pollutants. These reactions are not susceptible to scientific assessment, but their strength cannot be ignored. Much of the information about the effects of oil on seabirds has been reviewed fairly thoroughly in recent years (e.g., Bourne, 1976;

431 Group of Experts on the Scientific Aspects of Marine Pollution, 1977; Holmes and Cronshaw, 1977; Royal Society for the Protection of Birds, 1979; R.G.B. Brown, 1982~. Many of the data relate to west European waters, and it is largely to this area that we must turn to assess the short term and long term effects of oil pollution on seabirds. For- tunately, due to recent data gathering in western Atlantic waters, the interpretations can be applied to the North American waters with some degree of confidence. Effect of Oil on Individual Birds The direct effect of oil on a bird is to clog the fine structure of its feathers, which is responsible for maintaining water-repellence and heat insulation (Holmes and Cronshaw, 1977~. The plummage absorbs water as the bird sinks and drowns. Unfortunately, it is not known to what extent this effect occurs at sea or just how many birds are lost in this way. The loss of thermal insulation produces a more immediate response. It results in greatly increased metabolic activity to maintain body temperature (Hartung, 1967~. As a result, fat and muscular energy reserves are rapidly exhausted, leading to mortality (Croxall, 1977~. For these reasons it may be anticipated that birds are more likely to succumb from oiling in colder climates than in warmer climates (R.G.B. Brown, 1982) or after prolonged stormy weather when feeding has been limited or energy reserves are Jowl Birds also ingest oil, probably mainly from preening their oiled plumage. Autopsies of oiled seabirds have revealed, in addition to wasting of fat and muscle tissues, abnormal conditions in the lungs, adrenals, kidneys, liver, nasal salt gland, and gastrointestinal tract, and a reduction in white blood cell count (Croxall, 1977; R.G.B. Brown, 1982~. Few analyses are available from oiled birds collected live, but one has shown elevated levels of the mixed function oxidase system (MFO) in the presence of hydrocarbon contamination of liver, kidney, and muscle tissues (Vandermeulen et al., 1978~. However, whether pathological conditions are related to petroleum hydrocarbons or to generalized stress is uncertain. Nor is there evidence to suggest that any of the observed abnormalities were a primary cause of death, which in the majority of cases, is likely to have been drowning or hypothermia. A variety of physiological changes have been recorded in experimental studies involving ingested oil. Osmoregulator and hormone changes have been found (Holmes, 1975; Peakall et al., 1981), including retardation of weight gain of young birds (Miller et al., 1978), induction of hepatic enzymes (Goraline et al., 1981), and generalized pathological effects (Holmes et al., 19781. However, other workers have reported conflicting findings (i .e., Gorman and Singes, 1978; McEwan and Whitehead, 1978~. In some cases these differences can be ascribed to different oils (Peek all et al., 1982), and differences in species and dosing regimes are also important. Relatively small amounts of ingested oil can cause a temporary depression of egg laying and

432 reduce the hatching success of those eggs that are laid (Ainley et al., 1981~. The importance of all of these biochemical and physiological change in the wild are unknown. Ingested crude oils may inter fere with water and sodium ion trans- fer in the intestine and with excretion of salt by the nasal gland in some species, and may retard the growth of young birds, but the evidence is conflicting (R.G.B. Brown, 19821. Better evidence exists to show that relatively small amounts of ingested oil can cause a temporary depression of egg laying and can reduce the hatching success of those eggs that are laid (e.g., Ainley et al., 1981) . Small quantities of oil applied to the sur face of the egg are known to destroy the embryo at certain stages of development in the laboratory (e.g., R.G.B. Brown, 1982 ~ and the f ield (Birkhead et al ., 1973 ~ . There is no evidence , however, to suggest that this happens in practice on a widespread scale. Seabird Mortalities Due to Oiling An accurate estimate of the number of seabird casualties from oil pollution is not now and may never be poss ible . The only f irm f igures available are from counts of oiled birds found on shore surveys, but these are subject to severe limitations imposed by the intensity of the search, accessibility of the shore, repor ting eff iciency by cleanup crews , etc., and there is often doubt about the proportion of corpses found on beaches that were in fact oiled after death. A very large and probably signif icant source of error is the unknown proportion of oiled birds that die at sea but which never reach the coast, thereby escaping the shore surveys. The evidence suggests that, in fact, 30% or fewer of bird corpses drift ashore (Hope-Jones et al., 1978~. Estimates of actual losses in major incidents are therefore usually little more than informed guesses. They do probably indicate the orders of magnitude , however, whether hundreds, thousands , or exceptionally, tens of thousands (Bourne, 1976; Holmes and Cronshaw, 1977; see also Table 5-91. There is even less cer tainty about the number of casualties from oil pollution in regions such as the North Sea where frequent small oil slicks are formed as a consequence of the discharge of oily bilge and ballast water from ships. Pollution from this source approaches the chronic and is suspected of accounting, in aggregate, for at least as many seabird deaths as those resulting from more spectacular incidents (Croxall, 1977~. The estimate by Tanis and Morzer-Broyns (1968) that 150,000-450,000 seabirds annually are killed by oil pollution in the North Sea and North Atlantic has a slender factual base but may well indicate the order of magnitude of losses. In view of these var. ious uncertainties and in the differences in oil spills (oil type, weather, etc. ~ and available bird communities ~, it i s therefore not surprising that there is little relationship between the size of an oil spillage and the number of seabird deaths (Table 5-9) . One of the largest k ills of seabirds by oil on record was in the Skaggerrak, Denmark, in January 1981, when some 30 ,000 oiled birds appeared on neighboring beaches (Mead and Baillie , 1981~; this was

433 TABLE 5-9 Relationship Between Amount of Oil Spilled in an Incident and Number of Dead Birds Found Dead Birds Dead Birds Incident Tons Spilled Found per Kilometer Waddensee, Netherlands, Feb. 1969aunder 1,00014,56440 Poole, U.K. Jan. 1961a300487. Seestern, MedwaY, U.K., Tank Duchess, Tay, U.K., - Feb.-March 1968"871,368? Loch Indaal, U.K., Oct. 1969a11544925 Sept. 1966a1,7002,772 Torrey Canyon, English Channel April 1967a119,3287,815 Hamilton Trader, Ir ish Sea, _ ~ ~ May 1969~7004,092 ? San Francisco Jan. 19712,7007,380 ? Arrow. Cane Breton, . - _ ~ _ Feb.-March 1970~10,400567 26 Irvina ~ale. SE Newfoundland, Feb. 1970tunder 30625 47 Kurdistan, Cape Breton, March-April 1979D7,9001,697 26 Amoco Cadiz, Br ittany, March ~ 97~200,0004 ,572 ? NOTE: The actual numbers killed, allowing for extrapolation for known length of contaminated shoreline, and for the sinking of dead birds before they can come ashore, would be at least an order of magnitude greater in each of these cases. Knead birds per kilometer" is a rather more meaningful figure than "dead birds found n for compar isons between spills . Melbourne (1976, Table V). r own and Johnson (1980, Table 3). - ~Hope-Jones e t al . ( 19 7 8 ) .

434 caused by relatively small amounts of oil discharged by perhaps two ships (Royal Commission on Environmental Pollution, 1981~. Heavy casualties are sustained when floating oil encounters concentrations of seabirds on the water. The risk is greater in heavily traveled sea lanes and near oil industry operations but is critically influenced by seasonal and climatic factors. Species Vulnerability The species most commonly oiled are well known and well documented (Bourne, 19761. For obvious reasons the most susceptible birds are those which are gregarious, spend most of their time on the water, and dive rather than fly up when disturbed. Auks, especially murres (Uria sp.) and dovekies (Alle alle), form large concentrations on the water both at their breeding colonies in the summer and in their winter ing areas, and they have suffered very heavy casualties from oil pollution. Their counterparts in the southern hemisphere are the penguins, and of these, the jackass penguin (Spheniscus demersus) suffers regular losses from oil pollution, being close enough to shipping routes along the coast of South Africa. Diving ducks such as scoters (Melanitta spp.), oldsquaw (Clangula hyemalis), scaup and canvasback (Aythya spp.), and mergansers (Mergus spp.) suffer heavy casualties when concentrated on their winter feeding grounds. Common eiders (Somateria mollissima) appear to be vulnerable at most times of the year. Grebes (Podiceps spp.) and loons {Gavia spp.) tend to concentrate in coastal waters in winter, when they may be oiled; the numbers affected are rarely large, but in vow of their small world populations, even small losses may be significant. Phalaropes (Phalaropus spp.) congregate in winter at sea at narrow convergence fronts (R.G.B. Brown, 1982), and they would be vulnerable there to offshore spills, though such mortality has not as yet been reported. Many waterfowl and shorebirds flock on salt marshes and mud flats, and would be vulnerable to the destruction of their feeding habitat by oil spills. This would be especially significant for species such as the greater snow goose (Anser caerulescens atlanticus) which are highly concentrated on a very small stretch of shoreline in the Saint Lawrence estuary and Chesapeake Bay, and the semipalmated sandpiper Ereunetes pusillus in the Bay of Fundy. Unlike the foregoing, which are all from temperate/cold regions, seabirds of tropical and subtropical waters are less vulnerable because the species occurring there usually do not feed by pursuing their prey under water (e.g., Ashmole, 19713. The one exception is the cormorants (Phalacrocorax spp.), which do feed under water and occur in tropical seas; in one incident, several hundred of these birds were killed by a leakage of fish oil off Namibia (Berry, 19763. Impact on Seabird Populations Despite these various concerns and consider ing the large losses of seabirds from oil pollution, there may not be a material impact on Me

435 total population of a given species. This apparently is true of the auks of the British Isles and probably also for the diving ducks. In contrast, the northern (Arctic) auk populations are already in serious decline from other forms of human interference, and increased exposure to oil pollution is likely to affect them more seriously. The reproductive strategy adopted by auks and most other primary seabirds (as opposed to diving ducks ~ is a low reproductive turnover combined with great longevity and a long adolescent Period (e .g . , Ashmole, 1971 ; R.B . Clark, 1969) . This allows them to cope with erratic natural catastrophes such as fluctuations in their food supply. However, such a strategy makes the species slow to recover from addi tional and persistent man-induced mortalities due to excessive hunting, drowning in fishing nets, and oil. For these reasons most auk popula tions in the Atlantic have declined sharply in the course of this century (e.g. Cramp et al., 1974; Norderhaug et al., 19771. The surpr is ing exception to th is is in the BE itish I sles, where the most recent census figures show that 25% of the colonies of common murres (Ur ia aalge) and razorbills {Alca tor de) have maintained stable numbers - over the last decade, and 50% have actually increased. Colonies of Atlantic puffins (Fratercula arctica) in western and northern Scotland seem to be following the same pattern (Herr is and Murray, 1981) . Despite the chronic oil pollution in the North Sea and adjacent waters, which is estimated to kill several hundred thousand seabirds a year, many of them auks (Tanis and Morzer-Broyns, 1968), it indicates that the pollution has not been severe enough to check the regeneration of the populations following protection or perhaps climatic amelioration (e.g., Burton, 1981) . On the other hand, auk populations elsewher e in the Atlantic (with the exception of dovekies) continue to decline. These declines are probably mainly due to overhunting and drowning in fishing nets at present, though Arctic auks have suffered mortality from oil spills on their winter quarters off Scandinavia and eastern Canada (e.g., R.G.B. Brown and Johnson, 1980~. The risks there will undoubtedly increase with the development of offshore oil production off Alaska, northern Canada, and Newfoundland, along with associated shipping transport. There are reasons to believe that these northern seabirds will be more seriously affected by oil spills than auks and ocher venerable species in the more temperate waters. Oil spilled in these cold waters may remain unweathered for a longer period than in tropical waters due to slow or no evaporation of the more volatile fractions. Oil frozen into the sea ice and released during thaw may further prolong the effects of a spill. Added to this is the fact that the Arctic auks tend to breed in few very large colonies of cat 100,000-plus birds, rather than in small to medium colonies as occurs further south (e.g., Cramp et al., 1974 ; R.G.B. Brown et al., 1975) . These facts taken together suggest a real potential for oil on the water close to an Arctic auk colony to cause greater mor tality to the nor thern populations. The diving ducks, the other principal seabird group vulnerable to oil because of their diving habit, are probably less sensitive to oiling because of a different breeding strategy. They have a high

436 reproductive potential so that adult losses can be replaced rapidly. In recent years there have been large losses of common eiders in a number of oil pollution incidents, but the numbers in the breeding colonies appear to recover within a year or two (e.g., Leppakoski, 1973~. There may, however, be short term economic losses in areas such as Greenland, where these birds are a staple food for hunting com- munities (Salomonsen, 1967~. Remedial and Preventive Measures There is no doubt that the death of birds from oil pollution is regret- table. On this account there are continuing efforts to devise methods of reducing the number of casualties through rescuing and cleaning those birds that are oiled (Research Unit on the Rehabilitation of Oiled Seabirds, 1972a,b). While from a humanitarian point of view the cleaning and rehabilitation of oiled birds are commendable, from a conservation point of view the practice seems to have little value (R.B. Clark, 1978~. The number of birds saved and returned to the wild by this practice would be entirely too few to make any difference to breeding colonies. A more positive conservation measure is the restocking of depleted colonies, as has been practiced with apparent success with Atlantic puffins on the coast of Maine {Kress, 1980~. However, the cost, difficulty, and protracted nature of such a program suggests that, although it is probably practical, the restocking of colonies can make only a very local contribution to seabird conservation. For polar colonies the logistics would become awesome. The very limited scope for remedial measures has led conservation bodies to look instead to the prevention of seabird losses from oil pollution. This includes chemical dispersion of the slicks, scaring birds from the path of threatening of' slicks, and the protection of sensitive colonies by the use of booms (Rosk i and Richardson, 1976; Royal Society for the Protection of Birds, ~ 979) . However, experience suggests that these also have only very limited applicability and in the end will add very little to the saving of birds from oil pollution. Regardless of safeguards and remedial measures, a small amount of oil in the wrong place and at the wrong time can kill a disproportionately large number of birds. EFFECTS ON COMMUNITIES ~ ECOSYS~S The earlier sections on biological effects have been concerned prin- cipally with the impact of oil spills and chronic pollution on processes and on individual species and populations of similar life habit. Beyond these specific effects there is the further and more difficult question of how perturbations of individual populations affect other members of the biological association and the overall balance of the ecosystem. These ecosystem effects most commonly operate by altering either food web interrelations or competition for available space, although some

437 other effects have been noted. The problems are complex, and few areas have been studied In sufficient detail or for a long enough time to pro- vide satisfactory information about the whole ecosystem. Conclusions ar e meager, and many problems deserve fur ther attention . Wetlands, Saltmarshes, and the Intertidal Zone The margins of the sea are particularly susceptible to the impact of oil pollution. They are subject to heavy oiling when a large spill drifts ashore, with a fraction of the of' becoming sequestered in sediments and persisting in some cases for years. This is in marked contrast to conditions in the open sea, where currents and dif fusion usually rapidly reduce the concentration of petroleum, making it less toxic and most likely more amenable to degradation processes. The immediate effects of heavy oiling of the shore zone are obvious. There is widespread death of plants and animals due to smother ing and toxic effects. In the longer term the effects are more variable and subtle, and a f irst step In their evaluation is a knowledge of the behavior and persistence of the petroleum. Southward (1982), Vander- meolen (1982) , and Teal and Howarth (1984) reviewed this matter as part of a general account of long term effects observed after some major oil spills . As has been discussed in Chapter 4, oil behaves differently in the different coastal environments, dependent very much on the porosity of the sediments and the wave-erosion activity acting on them. In high energy environments, mainly rocky shores, the stranded oil coats the rocks and gradually hardens by weathering into a tough, tarry ~skin." The oil is gradually removed by wave erosion, although the rate of removal declines as weathering progresses. As much as 50% is lost within a year and a half to two years, although pools of oil are likely to collect in hollows among the rocks, protected by a skin of weathered oil, and may remain essentially unchanged for a long time. On cobble and sandy beaches the oil can sink more deeply into the sediments and can remain longer than on bare rocks. Wave erosion becomes less effective, and microbial degradation assumes a more impor- tant role. However, as the oil is mobile in these porous systems, some of it is gradually returned to the overlying water, where it is more subject to dissipation but may also have tonic effects on the organisms. A general hypothesis has emerged which appears to be applicable to the low energy systems: sandy beaches (Long et al., 1981a; Gundlach et al., 1982), marshes (Vandermeulen, 1980), and tidal rivers (Vandermenlen et al., 19817. Tidal pumping is the active factor causing penetration into the sediments. Sediment grain size controls the rate of penetra- tion (Owens and Robilliard, 1981~. In muddy sediments, penetration is minimal, and only the upper few centimeters are affected. However, because these are low energy environments with little physical weather- ing, stranded oil can persist here for up to decades, frequently becoming bound up in the soft organic sediments. All these variations are important in shaping ecosystem responses to oil spills. The macrophytes in wetlands--mainly mangrove swamps in

438 TABLE 5-10 Summary of Known Long Term Biological Effects of Some Major Marine Oil Spills Spill Effect Reference Torrey Canyons 1967 Continued community pertur bations during recovery, 1967-1978. Southward and Southward (1978) Searsport 1971 C-flux abnormalities in Gilfillan et al. gyp, 1976. (1976) Suppressed recovery of Bye Mayo et al. tl978) population, 1976. Arrow 1970 Intertidal species abnor- M.L.H. Thomas (1978) malities, 1976. Population abnormalities Gilfillan and in Mya, 1976. Vandermeolen (1978) C-flux depression in Mya, 1976. Abnormal shell formation in gyp, 1976. Florida 1969 Community abnormalities. Sanders (1978) Sanders et al. (1980) Changes in genetic structure Cole (1978) of Urosalpinx populations, 1976. Mussel sterility, 1970. Blumer et al. (1970) Long term inhibition of Krebs and Burns (1977) recruitment and low popula- Michael et al. (1978) tion densities in Uca pugnax. Behavioral disorders. Metula 1974 Slow marsh recovery. Gundlach et al. (1982) Alteration of total Colwell et al. (1978) microbial ecology. Bouchard 1974 Impaired reseeding and rhizome Hampson and Moull growth in salt marsh vegeta- (1978) tion, reduced interstitial fauna, increased marsh erosion. Argo Merchant 1974 No known long term effects. the tropics and salt marshes in h igh latitudes--are broadly susceptible to a variety of hydrocarbons (Baker, 1979, 1981; Bender et al., 1980; Getter et al., 1981; Golley et al., 1962 ; Hampson and Moul , 1978) . The e ffect of e ither physical or chemical stress from an oil spill in these systems almost invar iably is a sever e reduction in population and growth rate, amounting in some cases to complete obliteration (Baker, 1979; Hershner, 1977; Her shner and Moore, 1979; Hershner and Lake, 1980; Chan, 1977) . However, in the absence of continued stress there is likely to

439 TABLE 5-10 (continued) Spill Effect Reference Amoco Cadiz 197~: Tsesis 1977 Reproductive effects leading Elmgren et al. (1980, to reduced Pontoon reia 1981) population, 1980. Persistent disturbed community composition, 1980. Benthic offshore sublittoral community perturbations, 1979. Elevated mortalities in plaice, 1980 Persistence of benthic infauna perturbations. Elmgren et al. (1980b, 1981) Cabioch et al. (1980) Friha and Conan (1981) Glemarec et al. (1980), LeMoal (1981), Glemarec and Hussenot (1982) EDispersants used. tDispersants used only offshore. SOURCE: Adapted from Vandermeulen (1982). be some degree of recovery within a time span of one generation. This var ies from 1 year for some marsh grass species to a decade for man- groves. Probably recovery is facilitated in part because the oil does not penetrate deeply enough into the mud to k ill all of the extensive system of underground rhizomes. In this report the marsh grass (Zostera sp. ), growing in oiled sediments, has been found to have a higher level of contaminate ion than, for example . the rockweed Fucus (Vandermealen and Gordon, 1976), which was growing in a contaminated shore zone but, of course, was attached to rocks above the oiled sediments rather than being in direct contact with them. On the other hand, the persisting hydrocarbon load in such oiled marshes is now known to have long term impact on growth character istics of such marsh vegetation as Juncus maritimus, a species common to both shores of the North Atlantic. Studies in an Amoco Cadiz oiled marsh have shown chances in reproductive capacity, growth abnormality es, abnormalities in seed formation, and reduction in plant spikes in samples collected in August 1981, 3 years after the still (Levasseur and JorY, 1982) . These k inds of impact, ~ _,, lead to continued imbalances in recovery patterns at the level. These may be exacerbated with persistence of oil r Bounce In a dominant member of a community, must inevitably  commun ity in the associ- ated sediments. On the other hand, vegetative recovery is possible, although of the order of decades, once the source of the contaminant is reduced and eliminated, and natural weathering processes reduce the more toxic components of residual oil (e.g., Dicks and Iball, 1981) . Petroleum hydrocarbons can apparently serve as a supplemental energy source for microbial populations in wetlands systems (L. R. Brown and Tischer, 1969; Herwig, 1978; Kator and Herwig, 19771, and there is at least a temporary increase in the number of bacter ia and yeasts that are capable of degrading oil. Nevertheless, in the few cases that have

440 been followed for as much as 10 years, considerate' e quantities of petroleum compounds can remain in the soft sediments. Rapid decimation of the animal population is the i~runediate effect of a ma jor spill . The recovery per iod is slow. Table 5-10 summarizes toxicity effects that have been observed during the recovery period. There are abnormalities in the development and recruitment of individual species and, in some cases, large scale community perturbations. In general, species diversity and total abundance are reduced, hut resis- tant and opportunistic species often increase dramatically, temporarily mak ing use of vacated space and then crash ing from overproduction . Aside from individual species differences in tolerance, the life habit can be important in survival. For example, epibenthic f liter feeders would seem to be sub jected to less direct contamination than sedimen t f ceder s . This recovery process has been examined in detail for intertidal infauna by Glemarec and colleagues following the Amoco Cadiz spill in north Br ittany (Aelion and LeMoal, 1981; Glemarec, 1981; LeMoal and Quillien-Monot, 1981; Bodin and Boucher, 1981; Glemarec and Hussenot, 1981, 1982; LeMoal, 1981, 1982) . Based on a scheme of species gradients developed for domestic sewage-polluted marine coastal wetlands (Glemarec and Hily, 1981), they describe progressive temporal fluctuations within the infaunal communities of two coastal inlets. After the complete destruction of the original communities, recolonization following the Amoco Cadiz spill went through a series of transitory faunal species, before proceeding through intermediate phases to the development of stable population character istics of the unpolluted communities of the area (Figure 5-7~. Glemarec and colleagues were able to recognize particular groups of infaunal species (polychaetes, bivalves ~ according to the extent of their sensitivity to petroleum. Appearance and dis- appearance of the var. ious transitory or intermediate species were closely tied to the persistence of oil in the sediments, in turn related to the intensity of hydrodynamic (nweatbering") processes in these coastal sys tems . Desp' te a growing amount of descr iptive information about ecosystem perturbations, there are big gaps in the data and uncertainties about interpretation. Toxicity is generally blamed for the overall decrease in the animal population, but pr imary productivity is reduced too, and its role is undefined. The fluctuations in benthic populations probably affect the fishes that normally feed upon them, but little is known about that. A1SO, the effects of effluents from oiled wetlands on adjacent intertidal and subtidal waters have not been studied. Coastal and Offshore Waters and Sediments Broadly def ined, Coastal n and "offshore " are any areas seaward of the low tide level and include bays, open coastal waters, and the deep ocean. Oil spills in the open ocean do not appear to have as severe an effect on the biota as oil in coastal water or in the shore zone. The latter also, of course, are sub ject to ser. ious effects from chronic pollution.

441 a b 6 5 4 3 I<: ,d,::~,..~:: I: ^~//~////~ 6 4-2 4 _ :a 1 ~ ~! ~ ~ .~ ye ~ ~ f ~ .~\ _ , - ' ~ :\ ~. .. ~ ~ ~ . ..~. ~ ::::: :~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~: ~ ~ by_ : .~ . ',,., ~ 6-2 2-6 2-4 2 1 O C }~ FIGURE 5-7 (a) Schema of ecological succession of benthic infauna species after oiling of intertidal sediments (stage 61. (b) Establishment of opportunistic species (groups IV and V) after stabilization of pollution. (c) Regression of opportunistic species and reestablishment of more vulnerable species (group III). Group I. Sensitive species, dominant under normal conditions (e.g., Bathyporeia spp., Ampelisca spy.. Group II. Tolerant species r normally present in small numbers (e.g., Nephtys, Glycera, and Platynereis spy.. Group III. Species vulnerable to hydrocarbon pollution which reestablish rapidly on cleanup (e.g., Phyllodoce, Spio, and Nereis spy.. Group IV. Opportunistic species, fewer in number, but which occur in high densities in maximally polluted environments (e.g., Capitella capitata Scolelepis fuliginosa, oligochaetes). SOURCE: Adapted from Glemarec and Hussenot (1982) .

442 Coastal and Of fshore Water s The effects of oil in offshore waters have proven difficult to study and have in fact received little investigation. Thus far no evidence has been found that plankton populations are significantly altered by oil spills on the high seas. However, in near-shore waters, impact has been documented. For example, after the Tsesis spill in the Baltic Sea there was a decrease in zooplankton in the vicinity of the wreck (Johansson et al., 1980~. The quantity of phytoplankton, consisting mostly of microflagellates, increased briefly. There was no increase in the rate of production per unit of biomass, and Johansson therefore concluded that the change was due to a decrease in the amount consumed by zooplankton. The only analyses of oil in the water column gave values of 50-60 ~g/L at a depth of 0.5-1 m at a distance of 5 km from the wreck after 2-5 days of weathering of the sl ick . Quite similar results were obtained in long term oiling experiments in the 13 m3 MERL mesocosms (Elmgren et al., 1980a; Elmgren and Fr ithsen , 1982 ; Oviatt et al ., 1982 ; see also Studies in Field Enclosures section). The concentration of No. 2 fuel oil ranged from 60 to 350 ug/L, averaging 180 fig. This average probably was equivalent to about 75 ug/L of total aromatics (Gearing et al., 19791. Phytoplankton biomass declined initially but later increased to several times the values in control tanks. Species composition in the oiled tanks differed radically from the controls, shifting toward excessive dominance of nanoplankton. At the same time zooplankton declined, while bacterioplankton increased in number, thereby Paralleling the field observations made following the Tsesis still. Although the Tsesis is the only spill where effects of this sort have been noted, and indeed one of the few where such effects have been looked for, chronic pollution in some bays and estuaries may be suf- ficient to produce similar perturbations of the plankton part of the ecosystem. Values as high as 74 ng/L have been measured in inshore waters around Br itain, although typical values offshore are 1-3 ug/L ( Law, 1981 ) . In fact, the type of perturbation that has been described may be a common result of various kinds of pollution effects. Thus, in Long Island Sound (Riley, 1956), where the pollution load includes sewage and mirror amounts of oil and trace metals, pr imary production is maintained at a high level, with a much larger proportion of nanoplank- ton than in open coastal waters. The copepod population is small, r elatively speak ing; the cause might be e ither sublethal toxic ef feats or inefficient feeding on small phytoplankton species. Poor utilization of phytoplankton results in an increase in the quantity of nonliving par ticulate organic matter and in bacter ioplank ton. In short, both in Long Island Sound and in oil-polluted environments, there is a tendency to shift from the typical grazing pattern of coastal waters toward a detritus food web. The most obvious causal factor is competitive exclus ion of some of the larger phytoplank ton spec ies by hardy and opportunistic nanoplankton; however, the microzooplankton, usually the main predators on nanoplankton, did not increase correspondingly either . . -

443 in the MERL tanks or in the aftermath of the Tsesis spill. As they usually respond quickly to an abundance of food, a toxic effect on these organisms was suspected (e.g., Oviatt et al., 19821. Experiments indicate that fish eggs and larvae can be affected by exposure to petroleum hydrocarbons in water at levels similar to those found in the more polluted marine areas discussed above (Kuhnhold et al., 1978; Rice et al., 1979~. This has been a matter of concern, particularly when oil operations are conducted in an area that supports an important commercial fishery. Particularly vulnerable in this respect are both that fraction of the plankton population, the neuston, which normally lives in the immediate surface lamer, and the eggs and larvae of some of the commercially important fishes and invertebrates spending part of their life cycle in or near the surface layer. There are a few reports in the literature of deleterious effects, but new and better devices need to be developed for sampling under slicks before this problem can be evaluated properly. However, eggs and larvae are subjected also to many natural hazards, and survival of most of these species depends upon the production of enormous numbers of eggs. During the feeble plankton~c stage, storm- generated currents can carry them off into deep-water areas that offer no support for developing juveniles. Along the continental slope of the eastern seaboard and the southern edge of Georges Bank, warm-core eddies from the Gulf Stream can have the same effect. These natural catastrophic events generally involve larger water masses than are affected by the largest oil spills observed to date. In addition, eggs and larvae are fed upon by a variety of predators throughout the developmental period. The number surviving long enough to enter the commercial fishery is literally one in a million for cod and perhaps as much as 5 x 10-4 for herring and mackerel. Because of the varm ty of hazards, some sudden and catastrophic and others requiring long and careful observation, attempts to predict year-class fluctuations have not been very successful. Little connection has been demonstrated between the success of egg and larval stages and the success of the year class when it eventually enters the fishery (Hennemuth et al., 1980~. Thus, observed mortality arising from an oil spill does not necessarily imply that recruitment into the commercial stock will be reduced, and to prove that there was a significant effect because of oiling, among all the other var tables, would be virtually impossible. On the other hand, this cannot be taken to mean that such an impact does not exist. Benthic Subtidal Environments As has been described earlier (see Chapter 4), oil spilled onto the surface of the water column can be transferred to bottom sediments in variety of ways: sorption on clay particles and subsequent sinking, sinking of dead organisms, uptake and packaging as fecal pellets by zooplankton, and direct mixing to the bottom in shallow water. Thus even the deepest ocean bottoms are potentially liable to pollution, although shallow bottoms are matters of more practical concern because

444 of possible pollution effects on f isher ies. Deposit feeders, living in intimate contact with pore waters, can be expected to show a rapid uptake from sediments, while filter feeders and carnivores are more likely to derive oil from food or from water overlying the sediments (J.W. Anderson et al., 1978, 1979) . The MERL exper iments discussed earlier provided detailed information on the effects of oil on a benthic system. At the end of 20 weeks there were 109 ug/g of oil in the upper 2 cm of sediments in the bottom of the tanks. Dur ing the exper imental per lad the macrofauna decreased drastically. There was an average of 325 individuals with a total weight of 845 g in the control tanks and 95 animals weighing 78 g in oiled tanks . There was also a signif icant decrease in metazoan meio- fauna populations (especially harpacticoids and ostracods, with a smaller reduction in nematodes), although benthic diatoms, ciliates, and foraminifera increased signif Scantly. These increases were believed to be due to decreased bioperturbation, predation, and grazing as a result of reduction in the quantity of larger animals. The observed changes could not have been predicted on the teas is of knowledge of the effects of oil on the individual populations involved (Grassle et al., 1981; Elmgren et al., 1980a) . As for ecosystem interactions, so many possibilities exist that observed results are seldom intuitively obvious. The MERL tanks, despite certai n limitations and in part because of them, have proven very useful in descr thing these inter- actions. An unexpected ecosystem effect was found in studies of the Tsesis spill. The reproduction of her r ing was significantly reduced in the oiled area, and Nellbring et al. (1980) reported that it was not due to a direct effect of oil on the eggs, but rather was a consequence of a decrease in amphipod populations which ordinal fly graze fungi growing on the fish eggs, thus preventing fungal damage. Significant oiling of the sediments frequently has a deleterious effect of one sor t or another on ground fish populations. For example, following the Amoco Cadiz shill in North Br ittany, populations of sole (Solea vulgarism from the oiled plaice (Pleuronectes platessa ~ and Aver Benoit r iver system showed a number of degradative features including reduction in growth and increased incidence of fin and tail rot (Conan and Friha, 1981~. These and other toxic defects are now well known (R. D. Anderson and Anderson , 1976 ; Fletcher et al., 1981; Jackson et al., 1981; McCain et al., 1978~. Although some fishes may avoid contaminated areas, this is not a universal type of behavior, as illustrated by the results of Weber et al. (1979) with juvenile English sole . I n fact, f launder may be a par ticular ly susceptible group . A majority of the species live in shallow, inshore waters, exhibit little or no migratory behavior, and spend a considerable amount of time lying on the bottom or even partially buried in the sediments. In the final analysis, the ben'chic co''ununities appear to be most vulnerable to oiling, although different spill conditions will lead to different degrees of impact. This confirms The conclusion of the 1975 NRC report, which was based on a few isolated studies. Two examples come from the Amoco Cadiz and the Tsesis spills which point out very ~ ~ ~ - ~ occur, albeit unexpectedly. r emar kably the sor ts of impacts that

445 In the case of the Amoco Cadiz, studies by Cabioch and coworkers (Cabioch et al., 1980; Beslier et al., 1980) have demonstrated the destruction of an entire population of the amphipcd Ampelisca. Of particular significance in this case is that r epopulation is expected to be difficult if not unlikely because of the absence of other such populations upstream on the southern side of the English Channel. Studies following the Tsesis spill (e.g., Elmgren et al., 1980b, 1981 , 1983) show rather lucidly the results that may arise from the different sensitivities of mar ine organisms to oiling. Three years after the accident, the population of amphipods, Pontoporeia, remained depressed below prespill levels, partly due to heavy mortality of the amphipods by oiling, and partly because the expected repopulation from other nonoiled areas had not occurred. At the same time, the bivalve Macoma balthica population persevered and increased, apparently because it is , less sensitive to spilled oil (No. 5 fuel oil plus some Bunker fuel), and possibly because of heavy recruitment during the time when Pontoporeia was eliminated. Further imbalance has been added to the community by Macoma's long life span, relative to the amphipod. Current studies indicate that full recovery is likely to require more than 5 years and possibly up to a decade (Elmgren et al., 19831. Chronic Oiling While the discussion has dealt primarily with spills, because they have provided the main source of information, chronic pollution in fact is a more serious problem statistically (see Chapter 2), and deserves more attention. It results from the continuous long term release of low concentrations of petroleum hydrocarbons as may occur from refinery effluent and general petroleum activities. Bays and estuaries are par- ticularly subject to damage. As for coastal waters and offshore fishing grounds, few adequate investigations have taken place. Those that have been published show that such regions are not immune to the dangers of chronic pollution. The development of the Ekofisk oil field provided opportunity to monitor environmental response in an area of potential long-term oiling. Observations were made in the vicinity of Platform B and a nearby storage tank. The initial survey (ricks, 1975) was conducted in 1973 at about the time production began, and further studies were made in 1975 and 1977 (Addy et al., 1978~. During this period there was a decrease in total abundance and number of species of benth ic fauna and an increase in the hydrocarbon content of the sediments in the immediate vicinity of the operation, but the effect was limited to a radius of a few kilometers. Thus the impact was minor and may have been due to the open-bottom design of the storage tank (a situation since corrected) rather than due to the platform operation. A different kind of long term oiling potential is found in the waters off Louisiana and eastern Texas, following initial drilling there in the 1930s. Production in the area became significant, and by the 1970s, navigational charts showed more than 2,600 platforms extend- ing from east of the mouth of the Mississippi River to a point beyond

446 the Louisiana-Texas border . Because of the scale of dr illing and the controversy surrounding some of the environmental studies, it is worth- while to consider this situation in some detail. In the early 1970s a major effort was launched to determine whether environmental quality had been affected by drilling and production operations, using sites in the estuarine waters of Timbalier Bay and on the adjacent continental shelf off central Louisiana. This program, known as Offshore Ecology Investigation by the Gulf Universities Research Consortium (OEI/GURC), produced a series of reports by 23 principal investigators from 14 research institutes and university s and was funded by oil companies and oil-related industries. The Project Planning Council released a consensus report (Morgan et al., 1974), which stated that conditions near the platforms were not very different from control stations and concluded that little if any damage had been done. The original reports were microfilmed, and some were later published in the Rice University Studies, together with a general review by the editors. In the meantime the conclusions of the consensus report were challenged on the basis of data from the original report (Sanders, 19817. Major points in the critique were that the so-called control stations were too contaminated to serve as adequate controls and that the collection and analysis of bottom fauna had been improperly conducted. However, this controversy is put into a larger context by another survey conducted in 1978-1979 by the Southwest Research Institute under the auspices of the U.S. Bureau of Land Management. The series of reports coming from this survey was summarized by Bedinger (1981) and revealed a situation much more complex than was apparent in the earlier investigation. A portion of the outflow from the MiSSiSSippi River moves westward as a coastwise density current, and particularly during periods of peak runoff it distributes a heavy sediment load and a variety of contami- nants including trace metals and hydrocarbons. In addition, vertical stability associated with the reduction of salinity in the surface layer leads to depletion of oxygen in the bottom waters. The net result was that large areas of "dead bottom" were found, extending as far as 300 km west of the main Mississippi tributaries. However, this was a varying situation in which storm mixing tended to restore healthy bottom conditions and permitted repopulation by benthic fauna. Bedinger con- cluded That the entire Louisiana OCS is experiencing chronic contami- nation from the Mississippi River and probably production activities, but that the periodic flooding of the river and irregular but relatively frequent tropical cyclones cause such serious effects that they mask any platform related effects." The quotation refers to the area in general; biotic effects and increased hydrocarbon content in the sedi- ments were noted in the immediate vicinity of some of the platforms. Another Gulf Coast survey has been described in a book of composite authorship edited by Middleditch (1981~. It was conducted in the Buccaneer Gas and Oil Fi eld, about 50 km south of Galveston, Texas . Environmental investigations of possible oil impact from this operation did not begin until some 15 years after start-up of the field activi

447 ties. It is therefore possible that the studies missed the period of biotic transition that might have occurred with an environmental perturbation. The Louisiana coastal current continues through this area most of the year, with maximum reduction in salinity 1-2 months after the peak outflow from the MiSSiSSippi, but there is no evidence that significant hydrocarbon contamination is carried that far from its source. Currents commonly are of the order of 10-30 cm/s and occasion- ally are much stronger dur ing storms and are able to resuspend and disperse soft sediments as well as to remove dissolved materials. Because of this rapid dispersion, hydrocarbon contamination appeared to be negligible except in the biofouling community on the platform. The Buccaneer and Ekofisk surveys suggest that thus - ~ tion her; nosed only localized environmental hazard of a far oil produc- chronic nature ~ , ~ , ~ in open waters. However, there is concern that chronic releases in some coastal and continental shelf areas when coupled with restrictive circulation of water or mesoscale gyres could result in adverse effects over a period of years. While this concern appears to us justifiable, the current data related to this concern are not conclusive. Whatever ache expected activity, baseline surveys of any inshore area pr for to petroleum operations should include a thorough study of flushing rates and existing pollution levels. Another example of chronic oil impact concerns a saltmarsh near Southampton, U.K., which was continuously receiving refinery and petrochemical effluents between the years 1953 and 1971. Studies at the time indicated considerable ongoing damage to marsh biota, especially because of repeated light oiling of the vegetation, with considerable concern for long-term recolonization (Baker, 1970, 19711. One of the remarkable consequences accompanying this regression in marsh quality was the progressive die-back of Spartina between 1950 and 1970 to the point where much of the impacted marsh was denuded of this grass, and consisted largely of areas of bare mud. With the early 1970s an active effluent improvement program, on the part of the refinery/petrochemical industry involved, led to a reduction in effluent hydrocarbons, while at the same time an ecological monitor- ing program documented a steady recolonization of the mud areas, both by marsh grasses and by fauna (ricks and Iball, 1981~. The recovery showed a number of interesting features: (1 ) Regrowth by the dominant marsh plants, Salicornia and Spartina, paralleled the decrease in effluent hydrocarbon concentration, following the effluent remodeling program, and could be directly related to this. (2) This vegetative recovery did lag behind the effluent hydrocarbon reduction by about two years. (3) Vegetative recovery consisted of a graded process, with Salicornia the primary recolonized and Spar tina the secondary recolon- izer, while in particularly polluted mud areas near the effluent outfalls a si milar graded progression of algae were documented, coinciding with distance from the outfalls. (4) Vegetative recovery did proceed despite long-term hydrocarbon pollution of the marsh sediments, apparently correlating with considerable weathering of the aliphatic fraction. These observations indicate that saltmarshes, despite their obvious vulnerability and sensitivity, can recover from oiling impact. One

448 must be careful, however, in extrapolating scheme results and this recovery to situations involving more massive oiling, such as in the Ile Grande Marsh following the Amoco Cadiz, where much of the marsh was inundated by a thick layer of whole crude oil and mousse. A different kind of chronic pollution is found in areas subject to repeated oil spillage, such as the English Channel, scene of several large spills beg inning with the Torrey Canyon . In these cases the environmental impact of one spill may overlap on the impact of a previous spill, with the result potentially exceeding that of either of the individual spills . Little is known of e ither the cumulative impact or of the long term biological recovery after such repeated spillage, partly because of the lack of sufficient follow-up studies. Thus there are inherent difficulties in distinguishing ecological perturbations due to a previous spill from those resulting from a subsequent accident. The ecological surrey method, using benthic infauna species, developed by Glemarec and coworkers (viz., Wetlands, Marshes, and the Inter tidal Zone section, Figure 5-7 ~ seems one useful approach . In a comparison of benthic infauna from six North Brittany sites (three known oiled by the 1978 Amoco Cadiz, and three by the 1980 Tanio), their data suggest that one s ite , the bay of Ste . Anne (Tregastel), was already in ecological imbalance at the time of the Tanio oiling, pos- s ibly as a result of the Amoco Cadiz oiling 2 years earlier (LeMoal, 1982) . Effects on Communities and Ecosystems Despite a growing amount of descr iptive information about ecosystem perturbations resulting from oil pollution, there are big gaps in the data and uncertainties about interpretati on . Mere there have been heavy spills, and severe damage still exists after some years, further monitoring is needed to examine the whole history of the recovery process. In the cases that have now reached a late stage of recovery, investigation is hampered by lack of sufficient knowledge about normal, unstressed ecosystems. There are virtually no surveys that have been continued long enough to provide a standard for compar ing long term var lability in stressed and unstressed environments . In Table 5-10, continued community perturbations have been noted in the vicinity of the Torrey Canyon disaster despite the fact that the beaches were thoroughly cleaned with dispersants (Southward and Southward, 1978~. M. L. H. Thomas (1978) has also found long term effects in rocky intertidal areas where self-cleaning was rapid after the Arrow grounding. These are presumed to be rel axation responses from the original perturbation. However, there are perturbations in natural systems, too (Jones, 1982; Ware, 1964; Bowman, 1978) and we do not know the extent of their effect on the ecosystem as a whole. Year-class f fluctuations in some of the commercial f ishery stocks are well known, and there are similar variations in recruitment of some of the long-lived benthic invertebrates.

449 SPECIAL PROBLEM AREAS Tropical Regions Coral Communities At the writing of the 1975 NRC report, little was known of the potential impact of oil on coral reefs or on associated organisms, and indeed the general consensus at that time was that There appears to be no con- clusive evidence that oil floating above reef corals damages them" (Johannes, 1975~. Studies carried out since that time, principally those of Loya, Rinkevich, and coworkers, have established very clearly the vulnerability and sensitivity of the hermatyp~c or reef-building corals and their communities to oil and oil components. In addition, their importance, both as major marine ecosystems and in terms of human use, ranks them near the top in terms of spill concerns. Data on the fate and effects of petroleum hydrocarbons on coral systems are still limited, despite the vigorous efforts of a few workers (for recent reviews, e.g., Loya and Rinkevich, 1980; Ray, 1981; Rnap et al., 1983; Vandermoulen and Gilfillan, 1984~. Unfortunately, the variety of field and laboratory methodologies used, the various exposure conditions, and the differing hydrocarbon preparations and concentrations all add to the difficulty in assessing the problem. In fact, in several studies, reliable estimates of oil composition or concentration during the experiment are not available. Fortunately, more information is now beginning to become available, although much more is needed to assess properly both the vulnerability and sensitivity of reef systems, and their recoverability following a spill (Figure 5-8~. A wide range of responses to oil has been observed for hermatypic corals, including decreases in reproduction and colonization capacity, and effects on feeding and behavior (Table 5-11~. Oiling can cause problems, particularly with the reproductive process, which in turn has consequences for colonization and recolonization. A comparison, for example, of populations of the reef coral Stylophora pistillata, from chronically oiled and clean reefs and from laboratory populations, has shown a higher mortality rate in the oiled colonies. The oiled reefs also exhibited smaller numbers of breeding colonies, a decrease in the average number of ovaria per polyp, smaller numbers of planula larvae produced per coral head (fecundity was 4 times higher in the clean reef), and a lower settlement rate of planulae on artificial substrates (Rinkev~ch and Loya, 1977~. Subsequent long term laboratory studies (2-6 months), using an aquarium-type flow-through system with surface oiling to simulate the field oiling conditions, showed a significant decrease in the number of female gonads per polyp in 75 % of the polluted Stylophora colonies as compared to nonoiled control specimens (Rinkevich and Loya, 1979) . Similar results , due to No. 2 fuel oil, were descr ibed for the Bahamian coral Manicina areolata by Peters et al. (19811. In this case, degeneration of the coral's ova and lack of gonadal development were noted, under both static and flow-through bioassay procedures. .

450 __ _ . ~ ~ ~,. 4 ~ -2' ~ ,a,, - I' =. ~ .~ - ~:F _== W~ ~.~ - ,. ~ ;.~ _= _ ._. 9 i ~ . FIGURE 5-8 Oil pollution at coral natur e reserve of Eilat, Red Sea . The oil is coming in to shore as broad, 3-m-wide slicks and can be seen along the shore up to the high water mark. (Photo by Y. Loyal) Premature expulsion of the coral planula larvae appears to be a co~runon feature of oil pollution of reef systems examined to date. Larval extrusion due to sublethal concentrations of a crude oil, after 72-hour exposure, were reported for the soft coral Heteroxenia £uscescens (Cohen et al ., 1977) . Loya and Rinkevich (1979) descr ibed immediate mouth opening in the coral Stylophora pistillata in response to sublethal concentrations of Iranian crude oil, followed by premature extrusion of planulae larvae. Under natural conditions, S. pistillata releases its planula only during the night. In ache presence of water-soluble fractions of this crude oil the shedding was immediate, regardless of time of the day, decreasing their chances of survival and settl ing . Detr imental effects on the feeding response and tactile stimuli have been reported by J.B. Lewis (1971) for four Caribbean corals. The effects included a marked decrease in the number of tentacles expanded dur ing feeding, and increased extrusion of septal f filaments. Reimer (1975a) has descr ibed abnormal feeding reactions in four scleractinian corals and in one zoanthid coral, Palythoa sp., due to oil floating over the coral (Reimer, 1975b). Other effects, such as thinning of the coenosarc (Peters et al., 1980a,b) and changes in the pulsation rate of the polyps of Heteroxenia fuscescens in response to Iranian crude  (Cohen et al., 1977) have also been reported. Most corals produce excessive mucus on exposure to oil. This probably is a generalized protective mechanism whereby the coral cleans itself of irritating foreign material. In fact, abundant mucus producers such as Fungia and

451 TABLE 5-11 Summary of Effects of Oil Spills on Coral Reef s - Amoun t Repor ted Spill Spilled Effects Reference World War II, unknown, Many porous rocks and Johannes (197S) Japtan Island (Enewetak) boulders near the remains of the ship still heavily tarred in 1974 WW II, several tankers, Dennis (1959), U.S. Gulf of Mexico and Coast Guard (1959, Car ibbean Sea 1969) WW I I, Dry Tor togas Young mangroves of 4-5 years Odum and Johannes were killed (1975) 1967, Argea Prima 10 ,000 tons Mortalities of adult and Diaz-Piferrer Puerto Rico crude juvenile lobster, crabs, (1964) sea urchins, sea stars, sea cucumbers, gastropoda, octopus, and fish; Thalassia beds degenerated, rock y areas denuded of algae; extens ive damage to mangrove swamp habitat 1966, Br itish Crown, 25,000 tons Persian Gulf Qater crude Beynon (1971), Nelson Smith (1973) 1967, R.C. Staner, 22,000 tons About 2,500 kg reef fish Gooding (1971) Wake Island, Pacific gasoline, killed and stranded; large Ocean jet fuel, mortalities snails and turbine fuel, sea urchins diesel oil, Bunker C 1968, General 4, 500 tons Spooner and Spooner Colocotron is crude ( 19 6 8 ) Installation, Eleuthera-Bahamas 1968, Ocean Eagle, Many mortalities among Cerame-Vives (1968) San Juan, Puerto Rico intertidal organisms due to oil and emulsifier, including f ish, mollusks, and algae; recovery good 1968 Witwater Galeta 20 000 barrels Harmful effects to meiofauna, Rutzler and Sterrer , . . . Island, Canal Zone diesel oil, mangroves, fiddler crabs, (1970) Bunker C elimination of algae; reef corals least af fected 1970 Ennerdale,20,000 tonsNelson-Smith (1973) Seychellesfuel 1970 Pipeline break, Tarut Bay, Saudi Arab ia  1970 Ocean ic Gr andeL , Torres Strait, N. Great Bar r ier Ree f cr ude 100,000 barrels Mortalities among crabs, Arabian light bivalves, gastropods, f ish; mangrove trees less affected; no detrimental effects on corals and associated fauna; good subsequent recovery ',-1,100 tons Heavy mortalities of pearl crude oysters Spooner (1970) Smithsonian Institute (1970a)

452 TABLE 5-11 (continued) Amount Reported Spill Spilled Effects Reference 1970, unknown tanker Unknown; slick No apparent effects; no Smithsonian Institute near Pennekamp coral 75 mi long, apparent stranding of oil (1970b) reef, Florida Keys 0.5 mi wide 1971 MV Solar Trader, 520 tons fuel Numerous dead lobsters and Smithsonian Institute West Fayu, Caroline I and lubri- clams; survey 8 months (1970c) cating oils afterward reported large algal growth on corals in the area 1974 Sygma, Stockton 400 tons heavy 13 km beaches affected; no Hughes (1974) Bight, east coast of fuel damage reported to marine Australia life 1975 MV Lindenbank, 10,000 tons Mortalities of fish, crustacea, D.J. Russell and Fanning Atoll, Pacific copra palm and mollusks; afterward Carlson (1978) Ocean oil, coconut extensive growth of oil, cocoa Enteromorpha and ulva; beans reportedly complete recovery of original coralline algal community after 11 months 1969-1979, two oil Many small Decrease in coral and fish Fishelson (1973, 1977), terminals, Eilat, scale oil diversity; lack of Loya (1975, 1976), Red Sea, Israel spills, colonization by hermatypic Rinkevich and Loya various corals in reef areas (1977, 1979), Soya and · tankers chronically polluted by Rinkevich (1979) (crude oil) oil; damage to reproductive system of corals SOURCE: After Loya and Rinkevich (1980). Symphyllia showed good survival when directly contaminated with oil (Johannes et al., 19721. There appears to be some relation between mucus production and associated bacterial growth (Mitchell and Chet, 1975), but this is by no means clearly understood. Some observations suggest that corals may ingest oil, but nothing further is known of the fate of such ingested oil, or indeed of any metabolism that might be involved (Reimer, 1975b; Cohen et al., 1977~. Certain species would seem to have a greater affinity for oiling, possibly because of their morphology and surface texture. For example, J.B. Lewis {1971) noted that the branching species (e.g., Acropora) had a greater affinity for oil than the encrusting species (e.g., Agaricia). Summary Recent quantitative field studies have considerably advanced our knowledge of both coral reef ecology and on the potential effects of petroleum in these ecosystems. However, there are many gaps (see also Chapter 3, Biological Methods section). One item missing from most field studies involving oil is reliable information on the actual concentrations or composition of petroleum hydrocarbons found in the water, particularly during exposure times. Clearly this information is needed for cr itical comparison of data. Even more fundamental is the need for a better understanding of the reproductive physiology and

453 metabolism of reef corals, which for many genera is totally lacking. Without this basic information and understanding the mar ine toxicologist is work ing in the dark . At present the potential hazard of petroleum to reef corals and reef communities can best be estimated from the sensitivity of the larval stages. Despite their awesome mass and apparent robustness, reef corals are very sensitive organisms, forming the matrix for an extensive and highly intricate food web. In many parts of the southern hemisphere the coral reef is also the basis for human economy. Con- siderations of evidence to date rank this marine tropical system high in terms of potential impact of oil and the need for fur ther research (e.g., Gundlach et al., 1979) . Mangroves Most of the information on oil impact on mangrove systems comes from studies done on spills of opportunity (see Table 5-123. Unfortunately, in many cases the studies were done only sporadically and shortly after their occurrence, but without subsequent follow-up work. In others, several years had elapsed before work was begun, in the absence of data on the on-scene spill impact data. To make matters more difficult, very little of the available information is in the primary scientific literature, and has to be sought from conference reports, workshop proceedings, etc. Nonetheless, the survey of work assembled for this report shows that the mangrove system represents a unique problem regarding oiling and oil impact, exceeding that of the salt marsh of the temperate zone. Mangroves ar e essential to the tropical mar ine environment , repre- senting a major component in the productivity!of tropical coastal systems. They are found in most tropical areas of the world, occurring along an estimated 75% of the coastlines between 25°N and 25°S latitude. In Australia they extend to 39°S, and in Japan and Bermuda to 32°N. In the United States, mangroves are represented by four species: the red mangrove (Rh~zophora mangle), black mangrove (Avicennia germinans), white mangrove (Laguncularia racemosa Gaertn), and Bruguiera gymnorrhiza. The later is found only in Hawaii as an introduced species. Mangrove systems, in terms of the marine environment, provide two essential functions. They act to protect coastal systems against storm and current erosion through trapping and stabilizing sediments and debris (e.g., Socffin, 1970; Carlton, 1974; Teas et al., 1975) . They also provide food and shelter for a large number of invertebrate and vertebrate species through a complex detrital food web (e.g., Blasco, 1982) . Many commercially impor ten t species depend on mangroves for par t of their life cycles, including the spiny lobster, snapper, drum, sea trout, crabs, shrimp, mullet, and menhaden. Based on analyses of juvenile fish species present in mangroves, Carter et al. (1973) concluded that the most important nursery grounds in the marine environ- ment were the mangrove-fringed bays. A similar relationship between

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456 mangroves and prawn and finfish catches has been argued for portions of the Indo-Pacific. In Florida the mangrove habitat is considered a nursery ground for such commercially valuable crustaceans and fish as the pink shrimp (Penaeus duorarum), blue crab (Callinectes sapidus), mullet (Mugil cephalus) , gray snapper (Lubjanus griseus), red drum (Sciaenops ocellata), and sea trout (CYnoscion nubulosus) (Odum and Heald, 1972) . In many tropical regions where they occur, mangroves also represent an important source of wood and other products and are a significant feature and staple of local culture and co~erce (Teas, 1979; Blasco, 1982~. The vulnerability of the mangrove system to oil residues resides in two features, both unique to this ecosystem--the aer ial root system and the permeability of the mangrove swamp to tidal water and therefore to oil (Figure 5-91. The roots of mangroves are highly adapted to anaero- b ic soils or muds, emerging above the surface as aer ial prop roots (red mangrove) or pneumatophores (e.g., black mangrove). The surfaces of these structures are marked by numerous small pores termed lenticels, through which oxygen passes into the air passages within the root system. While it is a remarkable adaptation to an otherwise anaerobic environment, this aerial root system is also the Achilles heel of the mangrove in the event of oiling, for the aerial roots are highly susceptible to oiling, with clogging of the lenticels and inner al r respiratory system (e.g., Bacon, passages, eventually choking off the 1970; Odum and Johannes, 1975~. The second problem is that of permeability of this coastal system. Mangroves are the tropical equivalent of the more temperate salt marshes (Teas, 1979) and share many of the physical features that make salt marshes highly sensitive to oiling--low wave energy, large numbers of small channels, fine sediments. Their highly organic detritus-derived sediments make them especially susceptible to oil entrapment. Most of what is known about oil impact on mangroves comes from opportunistic investigations at the time of an oiling incident (for reviews viz. Baker, 1982; Vandermealen and Gilfillan, 1984~. Few of these observat~ons have been reinvestigated or confirmed in a rigorous way, either in the field or in the laboratory. Mangrove seedlings show considerable sensitivity to oil. Oiled propagules show reduced rooting rates in comparison to nonoiled controls. Young trees generally appear to be more sensitive than the mature trees. A common feature of oiling of a mangrove forest is leaf loss, and severely impacted areas have seen complete defoliation, frequently accompanied by either root or leaf exposure to the oil. Partial defoliation usually occurs in any spill impact areae Recovery from leaf loss is a varying phenomenon. Totally defoliated trees have not been known to recover. In partially defoliated trees, recovery was found to be initiated within 4 months. In others, recovery was still occur r ing only slowly 7 years after the spill incident. The leaves appear to be particularly sensitive to direct oiling, possibly because the trichomes on the lower surface of the leaves are damaged, resulting in disruption of the plant's ability to regulate water loss through its stomates. The wilting and desiccation often seen in oiled trees supports this suggestion.

~- :: - - - FIGURE 5-9 Effects of oil on a mangrove forest. (Top) Aer ial view (lighter regions along r iver drainage system show r iver banks through leafless trees) and (bottom) close-up view. (Photo by E. Gilfillan. ~

458 Recovery has been seen in most spill areas, although ache few rates that have been estimated vary in their magnitude. Microbial degradation of petroleum in these tropical sediments appears to be more rapid than in temperate climates (Gilfillan et al.' 1981), but very little else is known of the capacity of the associated microbial community to degrade the spilled and residual oil. Oil, once buried within the anaerobic sediments below the surface zone of biological activity, may be effec- tively isolated from such activity. The level of activity of macro- organisms in these sediments, for example, burrowing crabs, would have considerable influence on the long term fate of the impacted area. Similarly, estimates of biological recovery are uncertain. Totally devastated areas have appeared to be on climax cycles of approximately 20 years. However, functional recovery probably occurs over a shorter per Sod. Canopy recovery occurs within 6-7 years, and deposition of the rootmat detritus cover is well established after 10 years (H.J. Teas, unpublished data, 1983~. Some efforts at artificially reseeding damaged areas have been successful, and more work in this direction is continuing. Summary Mangrove regions represent a large unknown for the tropical areas of the world's ocean in terms of petroleum impact. Both the mangroves and the dependent communities appear vulnerable to oiling, with mangroves in particular highly sensitive to oil impact. Further r esearch in all aspects is recommended. Polar Environments Dunbar's (1968) timely warning that oil represents a serious threat to the arctic marine ecosystem still appears justified today. The rapid pace of petroleum activities in the Arctic during the past decade has aroused intense pUbli c concern about the effects of spilled oil, and the potential for oil spillage will increase even more as proposed production and transportation networks spread more widely across the Arctic. Unfortunately, the nature and magnitude of the threat have been difficult to assess in specific terms, and predicted effects range from the virtually negligible to the cataclysmic . Dur ing the past decade, however, sufficient basic and applied research has been under- taken (viz ., Tab1 e 5-13; also Arctic Mar ine Oil Spill Program, 1978- 1982) to at least place the problem in a more realist) c perspective and to identify areas of particular concern. Considerable efforts are being expended and some progress seems to have been made on developing spill countermeasures. In contrast, petroleum development in the Antarctic is not imminent, and as a result, little attention has been paid to the effects of oil on the fauna and flora of the southern polar ocean. The many pronounced physical and biological differences between the two polar marine areas (e.g., Knox and Lowry, 1977; George, 1977) will make it difficult to utilize the information gathered in the Arctic for anticipating any ecological impact of oil in the Antarctic.

459 TABLE 5-13 Summary of Studies on the Sublethal Effects of Petroleum Hydrocarbons on Polar Marine Organisms Species Location Sublethal Effect Reference Microbiota Beaufort Sea Heavy oils more resistant Atlas (1974) Bacteria coast to biodegradation than lighter oils, light volatiles toxic to bacteria at low tempera tures, biodegradation enhanced by nutrients Bacteria Point Barrow Increase in oil-degrading Atlas (1977) microorganisms after Atlas et al. contamination, biodegra- (1976, 1978) cation slow at low temperature Bacteria Point Barrow, Alkane metabolism by Arhelger et al. Arctic Ocean surface microflora wide- (1977) spread in Arctic Ocean Bacteria Beaufort Sea Ability of mixed cultures Bunch and Harland to degrade oil at 0°C in (1976) laboratory Phytoplank ton Beau for t Sea Growth of diatoms and Hsiao (1978) green flagellates inhibited by exposure to high concentrations Phytoplankton Beau for t Sea Increasing inhibition of Hsiao et al. primary production with (1978) increasing oil concentration Phytoplankton Kachemak Bay, Naphthalene metabolism, Cerniglia et al. (ice-edge stations) Alaska changed growth characteristics (1982), Van sensitivity to crude oil Baaken and Gibson (1982)

460 TABLE 5-13 (continued) Species Location Sublethal Effect Reference Phytoplank ton Cape Parry Slight enhancement of Adams (1975) primary production in water under oiled ice even though light reduced 50% Phytoplank ton Cape Parry Growth and photosynthesis NORCOR (1975) inhibited at high concentrations Macroflora Laminaria Beaufort Primary production Hsiao et al. saccharine Sea inhibited (1978) Phyllophora truncate Marsh vegetation Cape Parry Chlorophyll content and NORCOR (1975) photosynthetic capacity reduced Coelenterates Halitholus Beaufort Sea Locomotory activity Percy and Mullin cirratus impaired (1977) Cyanea Cape Parry Locomotory activity NORCOR (1975) capillata impaired unidentified medusse Ctenophores Pleurobrachia Cape Parry Oil exposure, tolerated NORCOR (1975) pileus behavior unaffected P. pileus Fletcher's Benzo(a)pyrene taken up R.F. Lee (1975) ice island but not metabolized

461 TABLE 5-13 (continued) Species Location Sublethal Effect Reference Polychaetes: Pectinaria Point Barrow Attraction to oil- Atlas et al. hyperborea contaminated sediments (1978); Nephythys Busdosh (1978) longosetosa Spio sp. Acanthostephisia Point Barrow No recolonization of oil- Atlas et al. behrengiensis contaminated sediments (1978); Busdosh (1978) Nereis vexillosa Alaska Sensitivity Rice et al. (1979) Harmothoe imbricata Mollusks Macoma balthica Gulf of Alaska Forced to surface when Taylor and exposed to dissolved or Karinen (1977) sediment-absorbed oil Mytilus edulis Gulf of Alaska Exposure to hydrocarbons Rice et al. resulted in lower rate of (1978) byssal thread formation Wide range of Alaska Sensitivity Rice et al. (1979) mollusks Echinoderms Strongyl-ocentrotus Spitzbergen Uptake of lower weight Carstens and droebachiensis aromatic hydrocarbons Sends tad (1979) Opiopholis aculeata after spill Cucumaria vega Alaska Sensitivity Rice et al. (1979) _ droebachiensis Leptasterias hexactis Eupentacta quinquesimita Crustaceans Copepods Calanus finmarchicus Spitzbergen Uptake of lower weight Carstens and l C. glacialis aromatic hydrocarbons Sends tad (1979) after spill C. hyperboreus Fletcher's Uptake and depuration of R.F. Lee (1975) ice island benzo(a)pyrene Davis Strait Depression of feeding by Gilfillan et al. seep oil (1984)

462 TABLE 5-13 (continued) Species Location Sublethal Effect Reference Isopods Mesidotea entomon Beaufort Sea Inhibition of growth and Percy (1977b) molting only at high concentrations _. entomon Beaufort Sea Neutral response to Percy (1976, M. sibirica presence of oil, oil- 1977a) tainted food, and cont aminated sediments _. entomon Point Barrow Readily recolonized Atlas et contaminated sediment al. (1978) M. entomon Baltic Sea Abortions in gravid Oksama and - females exposed to phenol ~ristoffersson and 4-chlorophenol, exposed (1979) animals more aggressive Amphipods Boeckosimus affinis Point Barrow Food search success Busdosh . . Gammarus zaddachi reduced, recovery occurs; (1978) . no effect on respiration; reduced burrowing in contaminated sediments; reduced locomotor y activity Unidentified Cape Parry Reduced activity following NORCOR amphipods prolonged exposure (1975) _. affinis Beau for t Sea Reduction in respiration Percy (1977b) at low oil concentrations but increase at high concentrations B. affinis Beau for t Sea Avoid oil, oil-tainted Percy (1976) Cammarus oceanicus food, and contaminated sediments B. affinis Beau for t Sea Reduction in locomotor y Percy and - activity Mullin (1977)

463 TABLE 5-13 (continued) Species Location Sublethal Effect Reference Corophium cIarencense Beau for t Sea Neutral behavioral Percy (1977a) response to contaminated sediments Pontoporeia femorata Point Barrow Aceroides latipes Melita Formosa Gammaracanthus - loricatus MonoculodR5 ':D_ Little or no recolonize- Atlas et al. tion of contaminated (1978) sediments within 60 days Range of crustacea Alaska Sensitivity Rice et al. (1979) Teleosts MYoxocephalus Bering Sea Exposure to napthalene, DeVries verrucosus rapid uptake, loss of (1979) equilibrium, cessation of feeding, no effect on ionic and osmotic regula tion, decrease in hemato crit, reduced respiration, no effect on protein synthesis by liver Range of fish Alaska Sensitivity Rice et al. (1979) (including salmon Moles and Rice herring, flounder) (1983) Oncorhynchus kisutch Alaska Reduced growth Moles et al. (Coho salmon) in salmon fry (1981) Mammals Phoca hispida Cape Parry, Seals immersed in floating T.G. Smith and NET, Canada oil; immediate signs of Geraci (1975), stress; severe eye irrita- Geraci and tion, recovery occurs; no Smith evidence of haematologic, (1976, 1977) biochemical, or physiolo gical disturbances; hydro carbons in urine and bile Callorhinus ursinus . Erignathus barbatus Phoca groenlandica Ieptonychotes weddelli Odobenus rosmarus . Various areas Small amounts of crude oil Kooyman have great effect on thermal conductance of fur bearing pelts, but little effect on non-fur-bearing pelts et al. (1976)

464 TABLE 5-13 (continued) Species Location Sublethal Effect Reference _. ursinus Bering Sea Light oiling of pelt Kooyman surface, increased et al. metabolic rate when animal (1976) immersed in cold water, effect lasts at least 2 weeks Thalarctos maritimus Churchill Ingested oil affected the Oritsland Manitoba, hematopoietic system et al. Canada culminating in renal (1981) failure, oil on the skin and fur increased heat loss and reduced insulation drastically Polar Studies Much of the useful information about effects of oil on temperate environments has come from studies on oil released into the sea accidentally, experimentally, or naturally (as from seeps). In polar seas, little of ecological significance has been learned from the many minor spills there (Keevil and Ramseier, 1975; E.L. Lewis, 1979~. This is due to a combination of logistic difficulties, interference from cleanup activities, and lack of prespill baseline information. Thus, there has been a continuing interest in the use of small-scale experi- t'~en~a' splays. ~nese nave Been particularly useful in increasing our understanding of the behavior and fate of oil spilled on or under ice (Glaeser and Vance, 1971; McMinn, 1972; McMinn and Golden, 1973; Adams, 1975; NORCOR, 1975; Martin, 1979; Arctic, 1978) but have rarely been effectively utilized for the examination of biological effects (Busdosh and Atlas, 1977; Atlas et al., 19781. The Baffin Island oil spill (BIOS) program now nearing completion in the Canadian Arctic involves the first moderate-sized experimental spill devoted in part to an integrated and multidisciplinary study of biological effects (Blackall and Serov, 1981 1983). _, ~_: a, ~ ~= __ _ ~ _ _ ~ _ ~ . _ . . - - , Although oil seeps occur in many areas of the Arctic, the total quantity of oil released is probably low, and elevated hydrocarbon concentrations in the water column are detectable only in their immediate vicinity (Levy and Ehrhardt, 1981~. Little attempt has been made to examine the effects of these seeps on the surrounding biota, except for one study now in press (Gilfillan et al., 1983; see also Effects of Natural Seeps section). With these few local exceptions, petroleum hydrocarbon concentrations in seawater and biota are generally low (~50 ng/L) in both the Arctic (Straw and Cheek, 1976; C.S. Wong et al., 1976; R.C. Clark and Finley, 1982) and the Antarctic (Swinner ton and Lamontagne, 19747.

465 Vulnerability Features unique to polar waters and relevant to oil pollution include ice leads, polynyas (open-water areas), and the subice and ~ce-edge habitat. These are heavily utilized at certain times of the year by many species of birds (Vermeer and Anweiler, 1975; Cross, 1980; Bradstreet and Cross, 1980) and by marine mammals (Stirling et al., 1975, 1977; Fraker et al., 1978; Stirling, 1980~. Leads and polynyas are generally well defined (e.g., Stirling and Cleator, 1981) and appear to be essential for the feeding, migration, and reproduction of many species (e.g., R.G.B. Brown and Nettleship, 19811. These habitats are also especially susceptible to oil contamination. Thus their poten- tial for severe oil impact is considerable. This may be of particular significance in the case of the bowhead whale which, by virtue of its low population numbers and its annual migration through the lead systems in the Bering and Beaufort seas, could be very much threatened by spilled oil. The ice-water interface presents special problems in the event of a submarine blowout. Associated with this interface in the spring is a well-developed under-ice community based on a highly productive algal layer, often extending several centimeters into the ice (Meguro et al., 1967; Homer, 1977~. While there remains much uncertainty about the trophic relationships and about the quantitative contribution of subice primary and secondary production to the annual values for the marine ecosystem as a whole, there is general agreement that timing rather than magnitude is of critical importance to certain species. It supple- ments the abbreviated season of water column production at a time when · . many species are releasing young. The intertidal zone, often subjected to severe oil impacts in lower latitudes, tends to be extremely depauperate in the Arctic, a conse- quence of low air temperature, freshwater runoff, and severe ice abra- sion. The ecological impact of oil in this zone is likely to be · · minimal. On the other hand, the shallow coastal lagoons found in many parts of the Arctic present greater problems in terms of oil impact. Although their total areal extent is small, they are heavily utilized by bird (Johnson and Richardson, 1981, 1982; Richardson and Johnson, 1981) and f ish populations at certain times of the year (Craig and Haldor son, 1980 ) . Physiological Sensitivity Data on sensitivities of polar organisms to oil are too sparse and fragmentary to permit many useful generalizations. There is, however, little evidence that polar species are sensitive to petroleum hydro- carbons than are comparable species from more temperate areas. In some instances the lower temperatures may enhance the apparent toxicity of oil by altering the physicochemical characteristics of its components or by prolonging the residence time of toxic fractions in the water column (Rice et al., 1977~. However, we cannot be too dogmatic. Thus

466 Van Baalen and Gibson (1982) and Cerniglia et al. (1982), in studying the capacity of Bering Sea ice algae to metabolize petroleum hydrocar- bons (Figure 5-10), observed that their particular cold-adapted diatoms proved highly sensitive to Cook Inlet and Prudhoe Bay crude oils. This vulnerability together with the very slow growth rates at near 0 °C, led these authors to suggest that such arctic diatoms will generally prove more sensitive to any accidental crude oil spills in or around the ice edge than might be expected from toxicity data alone. A number of sublethal effects, similar to those occurring in tem- perate species, have been reported for polar organisms (Table 5-13~. It is likely that certain types of physiological disturbances, par- ticularly those interfering with carbon flux (Gilfillan, 1975) and energy storage tW.Y. Lee et al., 1981), could have more serious ecological consequences for polar species, many of which experience extreme seasonal oscillations in food supply. Any interference with the very abbreviated period of summer production may be critical for overwintering (Littlepage, 1964) and reproductive success (Dunbar, 1968). Habitat and Biological Recovery There is little doubt that the rate of recovery following a severe spill will be much slower than at lower latitudes. Oil cleanup in ice-infested waters presents many problems (Barber, 1970; Milne and Smiley, 1976, 1978; Wadhams, 19801. The low temperature and presence of ice will probably result in much of the oil remaining unweathered for extended periods (Mackay, 1977; NORCOR, 1975), especially when coupled with less rapid microbial degradation than in temperate zones (Atlas and Bartha, 1972; Atlas, 1977; Bunch and Harland, 1976; Arhelger et al., 1977~. Although psychrophlic bacteria can grow rapidly (Morita, 1975), biomass can be reduced at polar temperatures (Bunch and Harland, 1976; Colwell et al., 19781. Biological recovery will also be delayed because of the reduced fecundity, dispersal, and growth rates of many polar species (Clarke, 1979; Dunbar, 1968~. Natural self-cleaning of oiled intertidal sediments, however, can be very rapid, judging from preliminary results obtained at the BIOS study site (Owens et al., 1983~. This was found to depend greatly on the degree of shoreline exposure and the amount of fetch offshore. On the other hand, in very low energy areas, persistence of stranded oil was estimated at longer than 10 years, similar to that recorded at lower latitudes. Summary Information on oil impact on polar environments is still fragmentary, with large knowledge gaps making spill impact assessment more guesswork than sound appraisal. Underlying much of the uncertainty is the absence of data about the basic biology of many important polar marine species.

467 ~ . ~ ax' :~::) at: . .~ . _ Hi..: ~ ~.:. ..~ ryes r  : . I: ~ ~  : "I'm ~ i:: :~ C~:,~ :,, I: :: IS. Lye. ~:~ ~,'~ :, ,.: ^..:S..~ by. :~.~:<S:~. Nitzschia Navicula Chaetocerous spy. spy. spy. Naphthalene Naphthalene FIGURE 5-10 Radioautogram of metabolites formed from (1-14C) naphthalene by ice-edge diatoms isolated from the Ber ing sea. SOURCE:: Adapted from Van Baalen and Gibson (1982) .

468 Thus studies are needed, not only on the effects of oil, but perhaps more so on ecological relationships and on the precise ecological significance of such aspects as the several unique polar habitats-- leads, polynyas, ice edge, and the under ice . EFFECTS OF NATURAL SEEPS Seep Stud ies Natural petroleum seeps in many respects offer unique advantages not readily available in other spill situations: (1) seeps often occur in coastal areas where other pollutants appear to be insignificant, therefore allowing in situ study of ecosystem-level effects of petroleum alone; ~ 2) they provide an opportunity for a study of really long term chronic contamination; (3) they can be revisited year after year , allowing detailed planning and refining of experimental approaches; and (4) they offer a realism unobtainable in laboratory studies. On the other hand, the seepage oil often becomes highly weathered in its passage through the crustal rocks and sediments and r therefore, can differ considerably from the oil released in spills, whether from tanker, production platform, or other sources. Also, there is often no knowledge of the past history of the seep area, of its biota and biotic communities (see also Appendix A). WE ile submar ine petroleum seeps are available for study and some are found on continental shelves (Fischer, 1978; Nelson et al., 1978), only a few efforts have been made to understand seep ecology. Of these, most of the work has centered on a group of very active seeps near Coal Oil Point in the Santa Barbara Channel, California. They include research on sublethal effects on local organisms by Straughan (1976) and a general study of seep ecology {Spies et al., 1980~. Elsewhere the only other studies are those in Tamiahua Lagoon, Mexico (Giammona, 1980), part of a large program on naturally occurring hydrocarbons in the Gulf of Mexico {Geyer and Giammona, 1980) , and an exploratory study program in Scott Inlet, Baffin Bay, Canada (Levy, 1981; Levy and Ehrhardt, 1981; Gilfillan et al., 1983~. In California, in areas of heavy seepage, such that the oil oozes out and forms tar mounds and asphaltic coatings on the sediment, this is often accompanied by low oxygen tension and high sulfide content. Where the sea floor has a significant sediment overburden, it often is compl etely anaerobic near the largest sources of seeping oil. White mats of Beggiatoa spp., the sul fide oxidizing bacterium, may be found together with large populations of nematodes. The polychaete Capitella capitata also occurs occasionally, but few other infaunal invertebrates appear able to exist in such heavily oiled sediments. The starfish Patiria miniata, Astropectin spp., and the surf perch Pheneroden furcata are sometimes found in these isolated areas of heavy oil accumulation Espies and Davis, 1979). No direct measurements on hydrocarbon utilization by microbes in such sediments have been made, but it is thought to be vigorous. In addition to the visible mats of Beggiatoa, ATP measurements indicate

469 that the microbial biomass is 2-3 times higher than in background sediments and there is an abundance of isotonically light H2S (Spies et al., 1980), indicating a microbial origin. Also the oil droplets and streamers emerging from the sediment have a low n-alkane content, characteristic of microbial degraded oils. Thus the oil entering the sedimentary environment is already partially degraded, and a complex microbial community of sulfur bacteria, sulfide oxidizers, and hydrocar- bon metabolizers is associated with it. These are the kinds of changes that generally accompany large sources of organic matter in the sedi- ments (Stanley et al., 1978) and probably represent an intensification of processes that normally occur in coastal sediments (cf. Novitsky and Kepkay, 1981~. Interestingly, despite the loss of many of the lighter fractions and the considerable weathering that this particular seep oil has undergone in its passage through the sediments, in one case at least, its toxicity, as measured by the diminution of growth in starfish embryos in laboratory bioassays, was comparable to or greater than several crude oils (Spies and Davis, 1982~. Biodegradation in buried reservoir source rocks may be responsible for the heavy, naphthenic nature of shallow oil deposits that give rise to seeps (Phillipi, 1977). Evidence from seeps in a quarried Gulf Coast salt dome also links H2S production with biodegradation processes (Sassen, 1980~. Low oxygen tension and high sulfide content presumably are not the only reasons for the impoverished macrofauna. Concentrations of dissolved hydrocarbons in interstitial water (sediment-bound water) as high as 1,300 ug/L have been determined in some seep areas. One such sample was composed almost exclusively of relatively toxic monoaromatic and diaromatic compounds (Stuermer et al., 1982~. This concentration and composition are probably sufficient to inhibit colonization by most infaunal species. Areas with more moderate seepage show correspondingly less impact, and fewer differences from control area. A detailed 24-month study of benthic macrofauna in the Coal Oil Point area showed that the benthic population, despite some differences, was generally representative of the shallow water extension of the Nothria/Tellina community described by Jones (1969) for large portions of the southern California shelf. In such areas at least a few drops of oil were found in every sediment core taken, and interstitial water samples contained 45-117 ug/L of dissolved hydrocarbons. Nearby control areas contained no free oil, although there was some weathered oil in the sediments, and the pore water contained 0.2-5 ~g/L of hydrocarbons. The species in the seep and in the comparison stations were both representative of this same community, but some deposit feeders, especially oligochaetes, were more common in the seep, and phoxocephalid amphipods were less common. In general, the populations in the seep area were significantly greater and tended to fluctuate more rapidly. However, the community structure generally was nearly constant and identical in the two areas (Spies and Davis, 1979; P.H. Davis and Spies, 1980~. Closer examination of the incorporation of bacterial carbon into some benthic polychaetes suggests that microbial activity, at least in areas of moderate seepage, and the microbial use of seep hydrocarbons

470

471 Sublethal Effects and Adaptation Most of the work on possible sublethal effects was done by Straughan (19761. In addition to examining a large number of invertebrates for malformations, there were comparisons of growth and reproduction in Coal Oil Point organisms with those at several control sites in the Channel Islands well offshore. Coal Oil Point abalone, mussels, and nonstalked barnacles appeared to grow and reproduce as might be expected in unpolluted environments. However, variations were observed at the different control sites, so that there was difficulty in discriminating between possible effects of petroleum and from other environmental factors. A comparison of oiled and unoiled goose barnacles, Pollicipes polymerus, at Coal Oil Point showed that differences in brooding rates between oiled and unoiled animals correlated with oiling, apparently due to thermal effects associated with enhanced heat absorption by tar on the car apaces of the oiled animals. Spies et al. (1980) reported that the starfish Patriria miniata at Coal Oil Point had consistently lower gonadal indices and a shorter breeding season than a population at Naples Reef, about 8 km away. However, no relationship has been found between ache gonadal index and hydrocarbon levels in the tissues, so that again there is a possibility of other causitive factors. Malformations have not been reported for animals from Coal Oil Point. Thousands of benthic organisms have been collected by Straughan (1976) and P.H. Davis and Spies (1980), and no gross malformations were seen, but neither study included detailed studies of tissues for evidence of pathological conditions. Straughan and Lawrence (1975) examined bryozoans for ovicell hyperplasia, which was reported from areas exposed to coal-tar derivatives and petroleum (Powell et al., 1970~. Again the results were negative, although the account did not state the neither of colonies that were examined. Whether seep organisms are adapted to petroleum is still an unanswered question. Laboratory experiments have given mixed but generally negative results. In tests with sea urchin embryos (Straughan, 1976) , starfish embryos (Spies and Davis, 1982), and several adult invertebrates (the mollusks Acanthia punctata and Tegula funebralis, the isopod Excirolana linguifrons, the sea anemone . Anthonleura eleeantissim~ ~ . seen Copulations have not Proved to be mor e resistant than organisms from nonoiled areas. On the other hand, a study of southern California mussels showed that a Coal Oil Point population was better able to withstand the rigors of an oil-covered aquarium (Straughan, 1976) than were nonseep mussels. However, it is not clear that there has been much opportunity for the sort of genetic selection through series of generations that might lead to significant adaptation in the organisms examined. Most of these species that have been studied have pelagic young, and inevitably there has been scat- tering and intermixture of inshore and offshore populations over the years with each breeding period. A different line of evidence developed to assess possible adaptation by seep organisms to the oil was an examination of the levels of hydrocarbon-metabolizing enzymes (see Chapter 3, Part B for a general

472 review of this subject). Hepatic aryl hydrocarbon hydroxylase (AHH) activity in the ambiotocid fish Phenerodon furcatus taken from a seep was significantly higher than in specimens from a nearby comparison area (Spies et al., 1980) . Similarly, the flatfishes Citharictys sordidus and C. stigmaeus had elevated AHH levels, and in related work, l AHH activity was induced in both species by giving them food contami- nated with oil (Spies et al., 19821. The general interpretation of these limited experiments is difficult because of the variety of organisms tested and because the dosage was much larger than the concentrations that seep populations would normally experience. The high doses were applied in order to get a measurable toxic effect in a short period of time. However, the fact that the results show little evidence of superior resistance to acute toxicity in seep organisms does not necessarily have a bearing on the degree of resistance to sublethal doses. Thus it is possible that these "ppm" doses overwhelmed adaptive mechanisms that might be effective at more environmentally realistic "ppb n and lower concentrations. HUMAN HEALTH During the last decade there have been various suggestions and warnings that petroleum hydrocarbons entering the marine environment may constitute a health hazard to humans. These concerns have been amplified as a result of several major maritime transportation and production accidents over the last 5 years. Contact with the human population following such accidents may be acute or chronic. Cleanup crews, members of the investigating scientific co~ranunity, and coastal r es idents may be sub jected to acute exposures following a local spill . In such cases, human uptake may be by inhalation, sk in contact, or even by ingestion of the petroleum or refined products. Exposure may also occur by direct contact with hydrocarbon components in the pelagic tars that are washed ashore frequently, such as the tar balls now found commonly on many of the world's beaches. Chronic nonoccupational exposure may occur through the accumulation and transfer of potentially harmful hydrocarbons from contaminated seafoods (Figure 5-11 ~ . The effects of the majority of petroleum components on human health are unknown. Studies with human subjects in the main are lacking. Most of the available information on the toxicity of natural crude oils and compounds pertains primarily to animals and not to humans. However, in most instances the animal studies serve as models for the human situation, and the data obtained have implications for human responses . Toxicity of Hydrocarbons to Humans--Acute Exposures Because of the concerns surrounding petroleum pollution to humans, some realistic and others more emotional, it is worthwhile to briefly review what is known of petroleum hydrocarbon toxicity generally. The direct ingestion of a variety of petroleum distillates will cause a number of toxic symptoms. The organ systems affected include the lung, gastro

' ~: . ~ ~, ~:~ - -I -# 473 .x 3~- ~ i- ~ =Z~, _ -q . ._ -. ~:: - _ _ ~_ a_ ~_ ~r..~ ._~ a. ~ · ~ ~ ~ ~ ~- ~-~ ~ ~ f . ~ ~. . .~^ . , - ~ ~ hi ,,~ Am' ~ By_ it__ - , - * :=ZF~ - a_ _ _ _ ~ Hi;_ it_ Why ~ I_ - _ ~ it' _ -at ~ -~ .; _, -.t . ~ i. .~ ~ ~ . . . . , .. .~ . .i,~ ~ .. . _ _, :~ ~. _ ~ · ^ -.~ · ~ At_ _ .* ~ .~q ~I ~1 _~.^ ~ ~ Dot-~ ~ -' ~ !_ :~: +, ~` ~ * ~ J~ : '' :~ - ~ ~ ~ - ~ ma: ~ ~ : :~h,- ~ - ~ ~  a: ~ ~ rl~ 1 ',: Ha. " FEZ ma. _ - r r - - ._ ^~ J . _ ~8 ~ . . ~...~ .. .. it_ ,. _ .. ~ ^~ ~ ~d ~_. r ~ r __ ~ Ha_ _ __ _~ at_ ~ :~ __ _ FIGURE 5-11 One potential route of oil contact with man is through eating of oiled seafood, as by this North Brittany coastal resident digging for clams in sediments oiled by the Amoco Cadiz tanker spill 1978. (Photo by J.H. Vandermenlen.) In

474 intestinal tract , liver , kidney , central nervous system {CNS), and the hematopoietic system (Zieserl, 1979; Vaziri et al., 1980; Poklis and Burkett, 1977) . In such cases, CNS symptoms are generalized weakness, lethargy, dizziness, convulsions, and coma. A greater risk of acute CNS toxicity is associated with those refined petroleum products having high concentrations of the more volatile aromatic hydrocarbons. Symptomatic involvement of the respiratory tract is the most common complication. If sufficient material is ingested, death will result presumably from aspiration of ingested materials into the lungs. Knowledge of human response to acute exposure comes primarily from studies with benzene-containing solvents and gasoline. Thirty-minute exposures to the solvents "Stoddard~ and "70,~ containing 22% and 74% alkylbenzene, respectively, resulted in eye, nose, and throat irrita- tions. However, there were no observable differences in eye blink, swallowing, or respiratory rates {American Petroleum Institute, 1976~. Other inhalation studies using rats, dogs, and the sensory response of human subjects with Stoddard solvent suggested a hygienic exposure level of 1.2 mg/L air for humans (Carpenter et al., 1975b). A few studies have been reported for the acute toxicity of gasoline to humans (Poklis and Burkett, 1977~. Thirty- to sixty-minute exposures to 500-1,000 ppm of gasoline vapor produced eye, nose, and throat irritation and dizziness. Exposures to higher concentrations for the same per iods of time resulted in varying degrees of nausea, headache, numbness, and anaesthesia. At 10,000 ppm, deep anaesthesia was achieved in 4-10 minutes for all subjects. The acute toxicities of a series of other petroleum hydrocarbon products have also been reported from inhalation studies with both human subjects and with experimental animals. Recommended human hygienic standards from these studies are summarized in Table 5-14. From a human health point of view, the volatile aromatic hydrocar- bon benzene occupies a unique position in that it is one of the few hydrocarbon compounds that have been established as human carcinogens. Normally the benzene content in fuels ranges from as low as 0.1% in some crude oils (H.M. Smi'ch, 1968) to a high of 1696 in some refined products, although it apparently constitutes 30-40% of oil discharged in formation waters in Alaska. Acute toxicity to benzene can be induced very rapidly via inhalation exposure. The most prominent effect is CNS stimulation, followed by depression and respiratory fa' lure (Leong, 1977) . Subacute and chronic exposures as low as 44 ppm can lead to a sequence of hematopoietic tissue changes. Lymph old tissues and the myeloid bone marrow itself, are affected, resulting in anemia and leukopenia. The adverse effect of prolonged exposure to benzene on myeloid and lymphoid tissues also can result in a deteriora- tion of the immunological defense mechanisms. There is also clinical evidence that hyperplastic bone marrow leukemia is associated with benzene exposure. A recent clinical study of tank-cleaning personnel chronically exposed to petroleum vapors suggested an exposure-connected relationship between vapors and chromosomal aberrations in bone marrow cells (Hogstedt et al., 1981~. For these reasons, benzene occupational exposure limits have been set at approximately 10 ppm.

475 TABLE 5-14 Suggested Inhalation Exposure Standards in Humans for a Variety of Petroleum Hydrocarbons Suggested Hygiene Standard for Humans Petroleum Vapor (mg/L) Reference . Mixed xylenes 0.46 (110 ppm)b Carpenter et al. tl975c) Rubber solvent 1.7 (430 ppm) Carpenter et al . (1975b) Varnish maker 's and Carpenter et al. painter's naphtha 2.0 t430 ppm) {1975a) Toluene concentrate 1.9 (480 pm) Carpenter et al . (1976d) Deodor ized kerosene 0.1 (14 ppm) Carpenter e t al . (1976b) "60 Solvent" O.44 (90 ppm) Carpenter et al . (1975d) ~80 Thinners 0.45 (100 ppm) Carpenter et al . (1976a) "40 Thinner" 0.15 (25 ppm) Carpenter et al . (1976c) "High aromatic solvents 0.15 (26 ppm) Carpenter e t al . (1977a) "High naphthenic solvent" 2.1 (380 ppm) Carpen ter et al . (1977b) Abased on sensory response in humans. Thor conversion to ppm, see U. S. Department of Health, Education and Welfare (1973) . Direct Sk in Contact Acute dermal tests have been done primarily with laboratory animals. These usually produce slight irr itation but do not necessar fly result in systemic toxicity. However, subacute dermal testing using 8 mL/kg of No . 6 f Mel oil (Bunker C) produced dermal irr itation as well as other dose-related responses . Histopathologic observations con f irmed both dermal and hepatic tissue toxicity (Beck and Hepler, 1980a) . In

476 another study, diesel fuel was found to be extremely irritating to the skin of rabbits when it was allowed to stay in contact with the skin for 24 hours (Beck and Hepler, 1980b). Subacute dermal testing with 4 and 8 mL/kg dosage levels produced treatment-related responses. Skin in the test area became necrotic, and weight loss occurred. There was also a 67% mortality rate in the 8-mL/kg dose group. Liver and kidney involvement was also observed. A similar set of toxicity tests has been conducted for motor oil, where 24-hour exposures produced only slight primary skin irritation (Beck and Hepler, 1980c). Acute and subacute dermal toxicity tests showed skin irritation and dermal corrosion, respectively. However, no signs of systemic toxicity were reported. Pregnancy and Development A series of studies has been carried out to determine the influence of particular products on pregnant laboratory animals and their offspring. These included kerosene (106 and 364 ppm), diesel fuel (101 and 401 ppm), and n-hexane (93 and 408 ppm) airborne concentrations, applied on days 6-15 of gestation (Mecler and Beliles, 1979a,b; Ament et al., 1979~. Only decreases in food consumption, at the higher exposure levels of diesel fuel were observed. There were no indications of compound-induced abnormal growths, variation in sex ratios, embryo toxicity, or inhibition of fetal growth and development. Several biochemical and pathological alterations are known to occur within hours or days following exposure to certain specific PAHs (Zedek, 1980~. Huggins et al. (1961) observed a delay in growth for several days, and the development of leukopenia after the second day, following a single oral dose of 7,12-dimethylbenzofa~anthracene in rats. The inhibition of DNA synthesis and mitoses in proliferating hepatic cells of rats has been reported (Marquardt and Philips, 19701. In this case, however, the impact lasted only 24 hours and was rever- sible. Hematopoietic functions in bone, spleen, and thymes, as well as the cell renewal in gonads and the intestinal lining, are found to be especially sensitive to acute exposure. PAHs can also be toxic to the developing rat fetus (Currie et al., 1970~. Human Exposure During Spills Field observations are few and mostly qualitative and anecdotal. Nonetheless there are several reports of human response to acute exposure during spills. Symptoms characteristic of acute toxicity to petroleum vapors were reported following the Amoco Cadiz oil spill along the coast of France (Menez et al., 1978~. An estimated 40,000 metric tons of light hydrocarbons may have been released into the atmosphere of the coastal area during the spill, creating a potential health hazard to the inhabitants, as well as the personnel in spill cleanup activities. Exposure to the workers was increased by mists and aerosols resulting from the high-pressure projection of water and steam

477 during various phases of the cleanup operation. Such activities con- tributed to increased dermal contact and even ingestion of small amounts of petroleum by the cleanup personnel. Among the symptoms reported by workers and coastal inhabitants, as well as by some of the scientific groups studying aspects of the spill, were headaches, dizziness, nausea, sensation of inebriation, vomiting, and abdominal pains. Workers coming in direct contact with the oil also reported skin irr Stations and erythema on the hands and limbs. Biochemical tests of blood samples taken from cleanup personnel who had worked at least 15 days in the vicinity of maximum impact, and from inhabitants of a local community, revealed, however, no significant changes in blood chemistry or enzymatic activity. Atmospheric samples taken adjacent to the Amoco Cadiz cleanup activities revealed the presence of many volatile aromatic and aliphatic hydrocarbons (Dowty et al., 1981~. However, the concentrations of benzene, toluene, and alkylbenzenes were substantially lower than those measured for a nearby urban center. The reverse was true for naphtha- lenes, which were substantially higher in the spill-impacted areas. Carcinogenic Potential in Humans It has been recognized for many years that specific hydrocarbon con- stituents commonly found in natural Prudes, refined products, and other related fossil fuel sources can result in the induction of cancer in humans and animals. Thus the introduction of large quantities of these materials, whether accidentally from spills, land runoff, or natural sources, into the marine environment becomes of concern, particularly where such materials may become incorporated into common seafoods. The ingestion of hydrocarbon-contaminated seafoods over a long period of time then may conceivably result in humans obtaining a carcinogenic insult. While the potential exists for ingestion of these carcinogenic components from the marine environment, the actual likelihood of their inducing cancer in humans is from small to negligible. Carcinogens in Oil Crude oils, several of the refined products, and such products as coal tar, crude shale oil, and furnace tar have long been known to have carcinogenic or mutagenic potential (e.g. , Bonser, 1932; Leitch, 1922; Passey , 1922; Cook et al., 1933 ; Twort and Twort, 1931; Woodhouse and Irwin, 1950; gingham et al., 1965~. For an extensive review of the literature see gingham et al. (1979~. This carcinogenic potential is invariably associated with the PAR fraction, the distillate fractions of crude oils with boiling point above 350°C. The PAR content of oils varies quite widely, both in amount and composition of PAHs. Content varies from a low of 0.2-7.4% in a range of crude oils (Gilchrist et al., 1972) up to 7.7-14.6% in three synthetic crude oil samples (Woodward et al., 1976~.

478 TABLE 5-15 Benzofa~pyrene Content of Selected Petroleum Products B (a) P Content Product (ppm) Reference Crude oil Libya 1.32 Graf and Winter (1968) Venezuela 1. 66 Graf and Winter (1968) Persian Golf 0.4 Graf and Winter (1968) Wilmington 2 . 5-2 .7 Tomkins et al. (1980 ~ Gasoline 0 . 21-0 .48 Walcave et al. (1971) No. 2 fuel oil 0 .6 Pancirov and Brown (1975, 1977) Bunker C 44 Pancirov and Brown (1975, 1977) Asphalt 27 Walcave et al. {1971) Shale-derived oil 4 Weaver and Gibson (1979) There are 13 PAHs listed as carcinogenic by the International Agency for Research on Cancer (1973), including the commonly detected compounds 7 , 12-dimethylbenzo (a) anthracene , benzo (a) pyrene (B (a) P), dibenz (ahlanthracene, and 3-methylcholanthrene, as well as the Bta)P metabolite 7-hydroxymethyldimethylbenzota~pyrene. B(a)P and related isomers plus the other carcinogenic hydrocarbons, despite their notoriety, are often present in crude oils in extremely small quantities relative to other PAHs (Table 5-15~. However, the used or spent petroleum products appear to be enriched in Bta)P and other PAR content. For example, Graf and Winter (1968) observed that the Bta)P content increased by as much as 200-fold in used oils. A potential future source of PAHs are the synthetic oils derived from oil shale and coal. A number of studies have found that overall mutagenicity of these products is greater than that observed for natural crude sources (Guerin et al., 1981~. They are also more active sk in carcinogens (Holland et al ., 1979) . While presently these products are transported primarily on land, accidental discharges relating to bulk transportation and handling may allow synthetic fuel products to find their way into the marine environment. Besides the potential mutagenic/carcinogenic risk from the synthetic parent products, there is also the possible increase in their mutagenicity after use, as has been found in Ache case of used motor oils (Payne et al., 1978~. Judging from the range of PAHs and der ivatives identif fed so far in these ~ ~ ~ al., 1981; products (Weaver and Gibson, 1979; Robbins, 1980; Ho et al., 1981; Guer in et al., 1981), the future use and discharge of these synthetic products should be monitored with care.

479 PAHs in the Mar ine Environment PAHs enter into the marine environment from a wide range of sources, including surface runoff from land, atmospheric fallout and rainout, as well as from the more traditional sources of spills and discharges of petroleum. A major contributor of PAHs is the direct combustion of fossil fuels , i .e ., gasoline- and diesel-powered vehicles ~ elects ical and heat-generation operations, catalytic cracking of crude oils in refining and related industrial processes, and refuse burning. Most significant in the formation of PAHs are those processes that utilize high temperature pyrolysis of organic material. Thus, it is estimated that from forest and agricultural fires alone, circa 420 t (metric tons, or tonnes) enter the atmosphere annually (Suess, 1976~. Clearly, the direct con tr ibution to the mar ine PAH load from petroleum spills and discharges is only one factor. Neff (1979), using a ser ies of assumptions, estimated that the total PAH input into the global aquatic environment would be 230,040 t/yr. Of that, some 170,000 t/yr were estimated to be derived from petroleum spillage. However, when viewed instead from the point of view of B (a) P input, an estimated 697 t B (a) P/yr enter the aquatic environment, of which only 20-30 t are due to petroleum spillage. This matches closely the calculations of about 10-20 tons of B (a)P entering the oceans as a direct result of natural crude and oil petroleum product discharges (Sullivan, 1974; quoted in O' Conner et al., 1981~. Risk to Humans The main concern regarding the risk to humans is the known carcino- genicity of several of the oil components. Because of their lipophilic nature, hydrocarbons will accumulate in seafoods and can potentially be passed on to man. The existence of PAHs in marine animals has been known for some time, and considerable documentation has been assembled over the past 5-10 years (Zechmeister and Roe, 1952; Group of Experts on the Scientific Aspects of Marine Pollution, 1977; Neff, 1979~. Certain mar ine organisms are also known to accumulate PARS from the environment very readily, without losing them rapidly over time either by metabolism or simple depuratione Mollusks, for example, are known to have low or little enzymatic activity to degrade PAHs. However, they readily accumulate these compounds, and one can find a strong r elationsh ip between the level of PAH in the tissues of mussels and their proximity to anthropogenic hydrocarbon inputs (Dunn and Stich, 1976; Dunn and Young, 1976; Farrington et al., 1983) . These PAHs in seafood are not the only source of potentially carcinogenic material to humans. Other sources include the full range of foods, from plant products to meats (Table 5-16), Over 100 PAHs have been reported in environmental and food sources. But of these, only 11 have been shown to be carcinogenic to test animals. In addition to exposures from other foodstuffs, humans are exposed to PAHs from a variety of other sources {Table 5-17~. While the nature of personal habits, such as cigarette smoking, will vary exposure to

480 TABLE 5-16 PAH Levels in Foodstuffs (~.g/kg Wet Wei ght) PAH Foodstuffs Benzo (a~pyrene Benzo (a) anthracene Meats Fresh _ _ Cooked 0.17-4.2 0.2-1.1 Charcoal broiled 2. 6-11. 2 1. 4-31 Smoked 0 . 02-14 .6 up to 12 Fish Cooked 0 . 9 up to 2 . 9 Smoked 0 . 3-60 0 . 02-2 .8 Grains and cereal products Grains 0 . 2-4 .1 0 . 4-6 .8 Flour and bread 1.1-4 .1 0 . 4-6 . ~ Baker's dry yeast 1.8-40.4 2.9-93.5 Fruits and vegetables Soybean 3.1-12.8 - Salad 2.8 4.6-15.4 Spinach 7.4 16.1 Kale 12.6-48.1 43.6-230 Apples 0.1-0.5 - SOURCE: Group of Experts on Scientific Aspects of Marine Pollution (1977~. PAHs and to Bta)P considerably, nonetheless, uptake from seafood appears not to be unusually high. There is no epidemiologic evidence for human cancer from intake of PAH-contaminated food (Lo and Sandi, 1978~. The preparation of seafood by smoking or grilling may be a factor in some isolated cases. However, for the continental United States and for other counts ies with similar food habits, combinations of source foods with high PAH content may be a factor along with occupational and other forms of exposure. There may be exceptions in those isolated areas were certain seafoods, especially If smoked, constitute a ma jor portion of the staple diet. The limited information available on the metabolism of PAHs by humans suggests that the majority of these compounds are rapidly absorbed and excreted and do not tend to accumulate in humans. Small quantities of PAHs can accumulate in body fat, adrenals, and over ies up to 8 days, but those that do remain are apparently rapidly metabolized. For example, in one study with laboratory rats 70-8096 of the B(a)P injected was metabolized to binary and secondary oxidation products within 6 hours (Falk et al., 1962~. There does exist some concern over the fate of these secondary products of metabolites, because several of them may possess either mutagenic and/or carcinogenic activity, while

481 TABLE 5-17 Estimated Human Exposure to Benzo (a)pyrene (B (a) P) Through Respiratory and Gastrointestinal Intake Dally Estimated Annual Source Consumption Intake of B (a)P (lag) Respiratory intakea Air 0 . 05-500 Cigarette smok ing 20 cigarettes 15-90 0 Gastrointestinal intake Or ink ing water 2 . 5 L 6-70 Food Normal diet 250-500 Smoked food diet 1.S kgb 550-3000 Potential seafood con tr ibution 100 gd 36 . 5-182 5 Contaminated seafood burdens 24-48 g 263-920 Respiratory intake is assumed to be 5 ,000 m3/person/yr . tFor 0 .5 fig B (a) P/Kg of food. SAssumed to be contaminated with 1-5 ug/kg B(a)P. Assumed 8ta)P levels to be from 1 to 50 ug/kg. group of Experts on the Scientific Aspects of Mar ine Pollution (1977) . SOURCE:: Connell and Miller (1980 ~ . Others have been shown to bind positively to nucleic acids (e.g., Varanasi and Gmur, 1980; Varanasi et al., 1980, 1982) . Conclus ion There are recognized human biochemical and physiological responses associated with acute exposure to natural crudes or their refined products. In general these responses appear to be transient and short lived unless ache exposure levels are unusually high. Prolonged subacute exposures, however, can result in tissue damage. In situations where the potential for human contact exists, efforts should be made to limit exposure through the use of respiratory and protective equipment.

482 The quantities of PAHs found in seafoods are for the most part equal to or less than the levels reported in other food sources. The major exceptions to this are seafoods harvested in the vicinity of municipal outfalls, creosote pilings, or local petroleum hydrocarbon sources. However, in most instances the natural intake, deputation, and metabolic processes are such that the PAH tissue levels will drop rapidly if the source is removed. Thus at present there is no demonstrated relationship that chronic exposures through eating petroleum-derived PAH-contaminated seafood are related to the incidence of cancer or other diseases in humans (King , 1977; Cowell, 1976; Group of Experts on the Scientific Aspects of Marine Pollution, 19777. Exceptions to these conclus ions may ar ise in local ized ar eas, as in the case of isolated fishing villages where seafood constitutes a major portion of the annual diet. No data are available, however, for these cases. SUMMARY Petroleum in the marine environment can elicit a broad range of toxic responses, at low concentrations (less than 1 mg/L), to many marine organisms, both plant and animal. However, in sediments and in the water column there is no compelling evidence to date indicating permanent damage to the world's ocean resources or even to a particular part of it. Nor is there yet evidence of increased pathological abnormalities in marine biota, due to petroleum hydrocarbons alone. Prior to the 1973 workshop, which formed the basis for the 1975 NRC report, much of the focus was on establishing toxicity and lethality thresholds or on the assessment of hydrocarbon concentrations in environmental samples. Since that time, research activity has broad- ened to include work on the site of action of toxicity, regarding petroleum toxicity as a dynamic process affecting the living organism at various levels: enzymatic, metabolic, ultrastructural, molecular. This has become evident from the breadth of studies brought together for this chapter, ranging from effects studied at the ecosystem level down to the effects observed and measured at the chromosomal level. There is also increased ability to predict, in the event of oil spills and of chronic oiling, the vulnerable and sensitive components of ecosystems, and to identify those parts of the ecosystem where petroleum hydrocarbons are likely to persist. One can see an increased direction of the research effort, with more integration between different disciplines as well as between different research teams . Research needs are being identif fed more clearly, and after considerable critical discussion with peers. New research areas are springing up, and the research effort is using more novel and critical approaches and technology . The last few years have seen an increased recognition of the importance of intercalibration, both of methods and of intercomparison of experimental results. There has also been a better understanding of the variability that exists in

483 different organisms or different life-cycle stages. The past 8 year s or so of effects-related research can be marked by two main advances-- the increasingly strong shift toward understanding ache physiological impact of petroleum toxicity, and the recognition that both organisms and life-cycle stages vary widely in their sensitivity and responses. As a result there is an encouraging trend toward establishing the study of petroleum pollution on a solid basis of interrelated chemical and b iological data . Evidence of this change in petroleum-pollution-related research is the increasing quality of published scientific papers that have come out of this work, and their appearance in the ranks of first-rate scientific literature. Mention has been made of the particular prob- lems in the s Judy of oil in the mar ine environment--the extreme com- plexity of petroleum and its refined derivatives, the state of analytical capability which is still far behind the requirements, and the complexity of the marine environment itself which is only poorly under stood, as well as the lack of knowledge of other toxic contaminant ef feats in the sea that could have served as models . Impact of Petroleum Without reiterating the individual f indings and conclusions set out in the preceding pages, it is nevertheless useful to focus on some specif ic observations in order to develop an understanding of how petroleum does affect the marine environment. First, no one marine organism has been found capable of actively excluding petroleum hydrocarbons from its tissues, be it plant or animal. In fact, all marine biota appear to be readily permeable to hydrocarbons, and readily accumulate them from their environment e ither directly from the water column or from pore water, or through their food. Also petroleum hydrocarbons can affect and can cause changes in a broad range of organisms and at all levels--cellular, organismic, and community. However, much of this information has come from laboratory studies, and just how these data apply to the field, in either a spill or with chronic input, is the subject of much discussion and much needed research. Part of the problem lies with the range and variation of laboratory studies, and the difficulties encountered in cougar ing different laboratory data. Part of the problem also lies in the lack of corroborative fie Id data. There are difficulties in applying results of the necessar fly simple laboratory systems to the complex, real ecosystems. Nonetheless, there are enough pieces of quantitative information available now that some broad outlines beg i n to emerge . For example, there is such a wide range of data available for petroleum impact on fish that we can at least suggest the potential impact of a spill or of chronic input of petroleum on a stock or a region (e.g., Longhurst, ~ 982) . Thus, we have considerable information on the effects of petroleum at very low levels (less than 100 ug/L), i.e. , there is the knowledge that oil exposure can enhance susceptibility to disease, that there exists a differential sensitivity of the various life-cycle stages and a greater susceptibility of larval stages, that

484 there can occur genetic effects (although not documented in all of its forms, yet indicative of a problem area), and that there is a wide range of deleterious effects on metabolism. Together these observa- tions frame the understanding of a potential and real impact on fish stock, both of lethality in the short term and deleterious and debili- tating impact in the long term, under certain environmental conditions. Our present inability to measure several of these effects in the fin Id in a fish population does not contradict the possibility of an impact. A similar conception of toxic impact due to petroleum can be sketched out for other marine biota, although for some the information is as yet slim. Relatively little attention has been paid, for example, to marine flora, either phytoplankton and macroalgae. The information for macroalgae especially is far from complete. Surprising also is that the data base for larval and juvenile fish is not nearly as plenti- ful and complete as it is for adult forms. Probably best understood are the macroinvertebrates, especially the intertidal forms of commer- cial interest (bivalves, crustaceans). Marine organisms differ widely in their sensitivity to petroleum. For example, arctic amphipods are highly susceptible to relatively low concentrations of crude oil, but isopods from the same sediments appear to be totally unaffected. Similar differences in susceptibility can be found throughout the phyla. Nonetheless there are some common features beginning to emerge from the very large volume of data assembled since 1975. There are suggestions that intertidal biota may be somewhat more resistant to petroleum than are the offshore benthic or pelagic forms. Whether this is related to some intrinsic higher tolerance, having to do with their highly variable intertidal habitat, is not known. Another generality is that larval fish are more sensitive than are juveniles and adults. Even certain stages in the development of fish eggs apparently are more resistant. However, again one encounters great variability between growth stages and between species, and while it is tempting to draw a sensitivity curve for fish life-cycle stages, in fact, this still cannot be done with certainty. Where this is needed, as for the commercially important species, the sensitivities will have to be determined separately for each species. Impact on Processes and Organisms Studies of oil impact on development at present seem to raise more questions than answers. The developmental process appears to be particularly sensitive to petroleum, and even relatively low concen- trations of petroleum (less than 1 mg/L; see Table 5-18) can result in measurable abnormalities, including spinal abnormalities in larval fish, symptomatic not only of petroleum but also of other contaminant pollution. The interaction between petroleum hydrocarbons and the chromosomal/genetic fraction of the cell is a relatively new area of research. We know that polynuclear hydrocarbons can bind with nucleic acids, and in other ways can perturb the normal meiotic and mitotic processes, resulting in subsequent abnormalities in development.

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486 Certain of the petroleum hydrocarbons have a histopathologic and/or mutagenic potential, although the research effort in these areas is still limited to only a few workers. Tumor formation in a range of organisms has been linked to oil exposure on several occasions, although the relationship has never been firmly established for field-collected samples. Similarly, such pathologic features as fin rot in groundfish appear to be more than casual correlates of oiling, being reported for some oil spills or chronic pollution sites. The mutagenic potential of petroleum stands on somewhat firmer grounds, albeit mainly from labora- tory studies, being associated definitely with its polycyclic aromatic hydrocarbon fraction. To date, however, this mutagenic potential has been demonstrated principally with certain microbes and is only now being extended to higher marine organisms. As most marine invertebrates and vertebrates alike possess the metabolic conversion mechanism (cytochrome P450 dependent) capable of producing the genotox~c inter- mediates from these PAHs, these intermediates are probably produced following oil spillage. Whether persistent mutagenic intermediates occur, in invertebrates or vertebrates, is totally unknown. However, any serious potential mutagenic threat to marine biota will probably come, not directly from oil spillage into the oceans, but more so from chronic inputs such as combustion products, street runoff, and used lubricating oils entering marine coastal waters. Of all the processes examined, the perturbation of normal behavior at very low concentrations of petroleum (as low as 10 ug/L) suggests a particular concern. The continuance of normal behavior underlies and is absolutely critical to larval sets' ing, feeding, reproduction, substrate recognition, and homing. In this context a change in or cessation of feeding is one of the first indications of oil pollution in many test animals. Yet most of the data available are largely anec- dotal, and at least for the higher organisms, the effect of petroleum on behavior is poorly understood. The last 5 years have seen a dramatic increase in the number of studies investigating the effects on various metabolic and physio- logical processes and have demonstrated a broad range of effects on those fundamental biological processes that govern normal growth and development. However, as yet no site of action has been determined for even a single petroleum hydrocarbon, with only gross perturbations noted. While such processes as respiration, photosynthesis, ATE production, carbon assimilation, and lipid formation are known to be affected by single hydrocarbons (e.g., naphthalene), the ultimate site or sites of action have not yet been determined. In fact, there is still uncertainty over the mechanism of uptake of hydrocarbons, whether active or passive, and whether it involves the lipid component of the cell membrane or enters the cell in some other manner. The lipid portion of cell membranes is thought by many investigators to aid in lipophilic hydrocarbon transport through the membrane into the cell, based mainly on some early studies on hydrocarbon Volubility in artificial lipid bilayers.

487 Impact on Communities Damage from oil can be extensive and catastrophic, affecting several hundred kilometers of shoreline as in the case of the super tanker Amoco Cadiz breakup. Entire communities have been impacted or even elimi- nated. However, with time such communities do recover. The recovery time var. ies, depending on the degree of oiling, the physical conditions of the ecosystem, and the nature of the community. Recovery to something approaching prespill conditions begins within a matter of months, and general prespill appearances will return within a year or two . However, there will be local pockets or "hot spots " of particularly heavy oiling and residue where impact may persist for 15 Years or longer. Also. certain community perturbations may continue ~ _ _ , for a decade or longer before return to a stable situation. In Chedabucto Bay, for example, the main portion of the oiled shorelines had returned to near-normal conditions within 2-3 years. At the time of this writing there is no visual evidence of any impact over 99% of the coastline, even though trace levels of petroleum hydrocarbon persist in many of the coastal sediments. The recovery process is a function of the degree of self-cleaning due to wave action and the proportion of soft-sediment low-energy systems (lagoons, estuaries, marshes) in the oiled areas. Thus an oiled rocky coast will be mostly self-cleaned within several months to a year, whereas an oiled lagoon or salt marsh can retain stranded of' residue for several years. For these reasons the time Period of recovery differs for different oiled environments. one course o' recovery, i.e., the pattern of biological recovery, differs also, depending in part of the biomass composition at the time of the spill, the climate, and other factors. Thus the survival of a dominant predator can influence the recovery process significantly over several years. Again, elimination of a key species, without a ready nearby available source for reintroduction, can result in a long-term altered community. The matter of "recovery to prespill conditions" has been the subject of some discussion, particularly as it relates directly to spill impact assessment. The notion of prespill conditions, of course, implies return to the ecosystem function and structure that existed prior to the spill. In reality, that is neither likely nor possible, for eco- systems and communities are dynamic assemblages, forever undergoing change and cycles of composition. A coastal community or benthic assemblage is never static, and what may have been its composition in one year becomes a different composition 5 years hence. Therefore, the best one can hope for is a return to the sort of community composition, in terms of biomass and species diversity and their cycles, character- istic of that particular environment. Recovery can thus be reasonably addressed only by comparison with what would have occurred in an undisturbed but otherwise similar ecosystem in the same time period. Depending on the degree of impact, the recovery process will proceed through a series of fluctuations, eventually to return to some stability. , mad ~.._~ ~¢

488 Polar and Tropical Reg ion s Special attention was given, in discussions and the workshop leading up to this report, to trop, Cal and polar regions , areas largely ignored in the 1975 NRC report. Since 1975 considerable attention has been diverted to problems of oil pollution in the arctic regions, largely as a direct response to increasing oil exploration activities in the Arctic. Much of this work originated with the Beaufort Sea studies of the early 1970s and has since been expanded by Canadian, Norwegian, and U.S. researchers. The Arctic presents special problems because of nearly year-round ice cover and inaccessibility. These are compounded by the large gaps in the data base on arctic biology, there existing only slim under- standing of biological events during the brief summer open-water season and virtually no understanding of winter events. The Arctic possesses unique features such as marine mammals, under-ice algal/crustacean communities, and seabird nesting areas. On the other hand, the threat of ice and ice scouring, the apparently slower degradation of stranded oil by arctic hydrocarbon-utilizing microbes, and inaccessibility place this region near the top of environmental concern. Some encouragement may be gained from the knowledge that the arctic marine environment is not nearly as fragile and pristine as it was once popularly thought to be. The northern marine biota are no more fragile than their more temperate counterparts. Natural oil seeps form part of the arctic substrate, and petroleum hydrocarbons have probably been part of the Arctic seas for unknown eons just as they have been for the rest of the world's oceans. However, those hydrocarbons probably have existed at very low concentrations, largely beyond the present detec- tion limits, and the potential impact of a major oil spill on an arctic ecosystem can presently not be estimated with confidence. The arctic (polar) ecosystem experiences special problems because of the sharply reduced ice-free season, and the particular conditions imposed on polar biota critically in tune with this climate. The tropics with their coastal mangrove swamps and coral reefs pose an entirely different kind of problem. Our present knowledge suggests that these systems may be highly vulnerable to oil. They are common coastal systems and highly porous in terms of oil penetration. Damage is not limited to the local biological component of these systems but has further ramifications in that these systems also support local fisheries and in much of the tropics form part of the human economic resources (mangrove timber, atoll human settlements}. Unfortunately, little research has been carried out on the effects of oil on mangroves or reef corals, or on their associated biota, and the amount of work is significantly less than that done in the temper- ate zones (North America, Europe, USSR). This is especially para- doxical in view of the vast tonnage of crude oil and refined products that annually are shipped through these environments. Certainly oil spills are not a rarity in the tropics.

489 Impact on Human Health Concern for human health centers mainly on the ingestion toxic or mutagenic components of oil, either directly or seafood . Inhalation of these components is less of a concern, and direct contact with oil as on contaminated holiday resort beaches in some parts of the world is viewed more as a socioeconomic concern. There is no evidence to date of a deleterious impact of petroleum via the marine environment on human health, although admittedly there exist fee? direct data. However, examination of circumstantial evidence from industrial hydrocarbon inhalation studies and from hydrocarbon levels found in marine biota suggests that any contact with either toxic or mutagenic oil components via the seas is far smaller than that affecting human health from other sources (atmospheric input, smoked probably the absence of food _ the food chain. But there exists the possibility that under certain conditions, as in chronically oiled areas or in situations where man is highly depen- dent on a certain seafood, man may well come into contact with fractions of oil that are deleterious to health, in concentrations higher than those encountered by the average world resident. _ foods, etc. ~ e One contributing factor is web magnification of hydrocarbons through Petroleum and Other Chemical Contaminants of e ither via tainted Many research workers have expressed concerns over the potential syner- gistic effects of petroleum, acting in concert with other contaminants, as in waters adjacent to highly industrialized areas (Puget Sound, North Sea, New York Bight). The phenomenon is poorly understood, but there are strong indications that the presence of one contaminant can enhance the toxic effect of a second. Whether this is related directly to some sort of molecular potentiation or whether it is simply a matter of lowering resistance to contaminants is as yet not known. Whatever the mechanism, the net effect is that while the potential damage of one compound may be minimal, the potential can become intensified in the presence of a second compound. Several workers feel strongl y that this aspect of oil pollution deserves more emphasis and that the potential impact through synergistic processes is much more of a problem than that of oil alone. CONCLUSION A great deal of productive and relevant work has been done since the 1975 NRC report. One outcome has been a shift from concerns over the cataclysmic tanker spill to more long term chronic input of petroleum = m_ : ~ ~ ~^~~^ en using present-day assessment techniques, that tanker spills have unalterably changed the world ' s oceans or mar ine resources. Studies from such spill sites have shown that oiled environments do recover with time. However, i t has become equally evident that petroleum can affect local into the marine environment. There is no evidence to date, . .

490 environments, where under certain conditions, oil may persist for several decades. In this respect, we find that in contrast to offshore situations where impact may be minimal and transient (although informa- tion is scanty), there are greater immediate concerns for coastal waters (for example, biologically productive estuaries) receiving on the average a greater proportion of discharged petroleum. Much more impressive, and fundamentally more disquieting, is the broad range of biological processes that can be affected negatively by petroleum hydrocarbons, particularly changes that can be elicited by some hydrocarbons in the genetic framework of marine biota. Petroleum hydrocarbons have demonstrably deleterious effects in laboratory and mesocosm experiments at concentrations as low as a few parts per billion, and such concentrations are now found over wide areas in the coastal oceans. Unfortunately we do not yet know the extent of any deterioration or damage, if any, from these low concentrations that may have occurred from oil hydrocarbon pollution in nature. This is because of the difficulty in separating the effects of oil from other kinds of pollution, from overfishing, or from natural changes and perturbations. In this respect, petroleum may well presage potential hazards due to other, more persistent and less biodegradable chemical contaminants entering our world's oceans, particularly along industrialized coastlines. RESEARCH RECOMMENDATIONS To give further depth to our understanding of the effects of petroleum in the marine environment, more work in certain areas is strongly recommended. Some of these areas are already subject to research, while others are relatively new. 1. Mutagenicity/Tumorigenicity. While recognized as a problem area, insufficient study has been directed toward solving this aspect of petroleum effects. Information is scant on invertebrate and vertebrate marine animals and marine plants. 2. Interference of Behavior by Petroleum. The perturbation of normal behavior at very low concentrations of petroleum (less than 0.1 ug/L) is a matter of particular concern. Change in or cessation of feeding is one of the first indications of oil pollution in many test animals. Yet most of the available data are anecdotal, and at least for the higher organisms, effects on behavior are poorly understood. 3. Mechanisms of Toxicity. The present focus on research into perturbations of physiological processes should be encouraged. While such processes as respiration, photosynthesis, ATP production, carbon assimilation, and lipid formation are known to be affected by single hydrocarbons (e.g., naphthalene), the ultimate site or sites of action have not yet been determined. 4. Polar and Tropical Environments. The polar environment presents special problems because of its nearly year-round ice cover and inac- cessibility. These are compounded by large gaps in the data base on polar biology. Data are scanty even for the brief summer season, and

491 there is virtually no understanding of winter events. The potential intact of a major oil spill on an arctic ecosystem cannot now be estimated with confidence. With respect to tropical regions, there also only a minimal amount of information on the effects of oil on mangroves, coral reefs, and their associated biota. 5 . Synergistic Toxicity. Interaction of various petroleum com- pounds is poorly understood, as is their interaction with other con- taminants. Further work along these lines is badly needed, for chronic pollution of inshore waters commonly involves a number of contaminants. 6. Ecosystem Effects. Population changes caused by an oil spill or chronic pollution inevitably alter food-web relations and interspecific competition in the ecosystem as a whole. Each oil spill is different, and the effects are sometimes quite unexpected. Continued study of ache history of recovery from oil spills is essential in order to assess their biological and economic signs icance . 7. Chronic Pollution. Much effort has been directed toward under- standing the impact of petroleum hydrocarbons from episodic events such as tanker spills, pipeline breaks, or leaks from shore-based facilities Much less has been done on incidents involving chronic pollution, where the concentration of petroleum is frequently low but released continu- ously over extensive periods of time. We encourage research toward a better understanding of the impact of chronic and accidental oil input at the ecosystem level. We recognize that this requires a much better understanding of both the natural processes occurring in ecosystems and the interaction of oil with other anthropogenic influences. 8. Pollution Indices. Further attention must be directed toward developing capability to assess petroleum impact at sea, In the pelagic environment, involving especially zooplankton and larval f ish . The single most significant difficulty is transferring information obtained from laboratory studies to predicting and/or evaluating potential impact of petroleum on living marine resources in the field, especially in the case of spill impact on such commercially important stocks as fish and shellfish. REFERENCES Adams, W.A. 1975. Light intensity and primary productivity under sea ice containing oil. Beaufort Sea Project Technical Report 29. Depar tment of Environment, Victor ia, B. C. 156 pp. Addy, J.M., D. Levell, and J.P. Hartley. 1978. Biological monitoring of sediments in Ekofisk oil field, pp. 515-539. In Proceedings, Conference on Assessment of Ecological Impacts of Oil Spills. Amer ican Institute of Biological Sciences, Arlington, Va. Aelion , M., and Y. LeMoal . 1981. Impact ecologique de la mar~ee noire du Tanio sur les Places de Tregastel (Bretagne nord-occidentale). Rapport de contrat CNEXO. 80 65 96 . Institut d' etudes Marines, Univ. de Bretagne Occidentale, Brest, France . 30 pp.

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