TL: Whales in a Changing Ocean (GP) SO: Greenpeace International DT: 1994 Keywords: oceans endangered species whales atmosphere global warming climate change toxics diseases greenpeace reports gp / Whales in a changing ocean (GP) the impacts and implications of global change NOTE: All references omitted here; unscannable. TABLE OF CONTENTS 1. Executive Summary 2. Introduction 3. Whales and climate change 4. Whales and ozone depletion 5. Whales and organochlorine pollution 6. Whales and a decreasing food supply 7. Whales and disease 8. Whales in a changing ocean 9. Conclusion Edited by Lesley Riley With thanks to Kieran Mulvaney, Bill Hare, Isabel McCrea, Louise Bell and Michael Earle Executive Summary The marine environment, home to the world's great whales, is under a sustained threat from the consequences of human activities, including climate change, increased UV-B radiation from a diminishing ozone layer, pollution by organochlorine compounds, and increasingly intensive fisheries. The complex cumulative and interactive effects of these changes are largely unknown. The evidence, however, suggests that these multiple stresses upon ocean ecosystems are already affecting marine life. Most immediately at risk from these human threats are the plankton, organisms that form the basis of all oceanic food webs. The productivity of phytoplankton is being affected by climate change in the central North Pacific (Section 3) and by increased UV-B radiation in the Southern Ocean (Section 4); fish larvae are vulnerable to the effects of toxic, persistent, bioaccumulative organochlorines in the marine microlayer (Section 5). Changes in the abundance and composition of plankton communities are likely, in turn, to affect the stability of populations of organisms at higher levels in food webs. Cetaceans as high-level consumers will be particularly vulnerable to changes in plankton productivity in the oceans. More direct effects on whales are also probable. Declines in some marine mammal populations have already been linked to food shortages brought about by the intensification of commercial fishing. That whales have died from entanglement in fishing gear is well documented (Section 6). Organochlorines have been shown to cause reproductive failure in marine mammals and to be potent suppressors of mammals' immune systems. Hence it is possible that physiological stresses brought about by changes in the availability of food or by exposure to toxic chemicals may further decrease the cetaceans' ability to resist and combat disease (Section 7). Some of these factors will have a cumulative effect on the world's whales and, when they act together, can be expected to cause even greater harm. The impacts are likely to be extremely hard to identify and quantify. A time lag between the onset of chronic stress and its measurable result is probable (Section 8). What is certain is that the cetaceans, even the oceanic species, no longer live in a pristine environment. The moves towards a resumption of commercial whaling for certain species suggest that little account has been taken of the increasing threat to their survival from the degradation of the marine environment. Yet the impact of changes documented in this review could be catastrophic for the health and stability of cetacean populations even with the current moratorium on whaling in place. Anything less than a precautionary approach is unacceptable in the effort to conserve whales. The benefit of doubt must be given to the environment. A failure to do so may extinguish any hope of long-term protection for the world's whale populations. Whaling is widely held responsible for the extinction of the North Atlantic gray whale (Eschrichtius robustus) (Mitchell, 1973) and has substantially reduced the sizes of almost all other whale populations (Braham, 1984). While the stock definition and status of the populations of most of the great whales is uncertain, there is sufficient information to conclude that populations of humpback (Megaptera novaeangliae), right (Balaena glacialis), bowhead (B. mysticetus), fin (Balaenoptera physalus), sei (B. borealis) and blue (B. musculus) whales are severely depleted, in some cases to below 25% of their original sizes. Some species have probably ceased to play any functional role in the ecosystem. At least six populations are on the immediate verge of extinction, probably numbering less than 5% of their original sizes. These include the western North Pacific gray whale, North Atlantic and North Pacific right whale, North Atlantic bowhead, and Southern Ocean blue and humpback whales. Virtually no information exists on the status and size of populations of Bryde's whale (Balaenoptera edeni) (NMFS, 1988; Braham, 1984). Minke whales (Balaenoptera acutorostrata) are reduced in the North Atlantic and Pacific Oceans as well as in some sectors of the Southern Ocean. The present concern, however, is whether or not the recovery-- even the survival--of the world's remaining whale populations will be prevented or impeded by a wider range of human activities, the most important of which are discussed in this document. In greatly reducing most populations, whaling may have increased the whales' vulnerability to both human-induced changes and natural fluctuations in their environments. Severely depleted species have a low reproductive potential, reduced genetic diversity and, perhaps, reduced contact between reproductively active members. Our ability to assess the potential effects of such changes is hampered in part by insufficient knowledge and understanding of whale populations themselves. Where estimates of population size do exist, most are no more accurate than plus or minus 50%. Trends in populations are virtually impossible to detect over a short period of time. Recovery of a heavily depleted population, following decades of complete protection from hunting, has been observed in a few cases such as the Eastern Pacific gray whale, but not in others, such as the Southern Ocean blue whale. Since 1990 the International Whaling Commission's Scientific Committee has attempted to assess the effects on a few stocks, mostly of minke whales, of the moratorium on commercial whaling, which came into force in 1986. In all cases, including the southern hemisphere minkes for which the best population estimates exist, the Committee has been unable to do so because, by its own reasoning, of the short length of time the moratorium has been in place, the lack of precision in population estimates and surveying techniques, and the slow growth rate of whales. Scientists have so far been unable to determine the existence of discrete biological populations of most species of whale, or to what extent such populations mix, particularly in their feeding grounds. Basic information on the migration routes of most species, on the social structure especially of baleen whale populations, and on reproduction is also lacking. There is also limited understanding of the habitat requirements and of the factors governing the natural regulation of whale populations. No serious attempt has been made, for example, to determine the ecological importance of whales. Coinciding with the precarious situation of many populations of whales are numerous visible signs of widespread deterioration and change in the marine environment. Chemical contaminants are known to be toxic in a variety of ways at all levels of the food web (Waid, 1986). Mismanagement of fisheries and over- exploitation of fish stocks on a global scale are considered to have resulted in a significant reduction in sustained yields (Gulland, 1984). Furthermore, commercial fisheries have dramatically altered the abundance of numerous species of fish (MacCall, 1986). Increased UV-B radiation, as a result of ozone depletion in the stratosphere, is already known to cause a reduction in phytoplankton productivity in the Antarctic (Smith et al., 1992). Coastal degradation from chemical and nutrient pollution, development, river impoundment, introduced species, and a variety of other causes, has also been widely reported (Linden, 1990; GESAMP, 1991). Major disturbances in the marine environment appear to be increasing in frequency and recently include coral-reef bleaching (Glynn, 1991), mass mortalities among marine mammals (Simmonds, 1991), epidemics of fibropapilloma disease in green turtles (Balazs and Pooley, 1991), ulcerative syndromes in coastal finfish (Sindermann, 1988), and the spectacular increase in noxious and novel phytoplankton blooms (Smayda, 1990). Damaging human activities are now so widespread and pervasive that their impacts on the biosphere are global in nature. It is in the light of this global environmental change that the protection of the world's great whales and their environment must be viewed. A precautionary approach The only effective way to safeguard the future of whales, their habitat and the wider environment is through a 'precautionary' approach to their protection. This principle has gained currency in the past decade and stems from the knowledge that we can no longer wait for conclusive scientific evidence that a human action is harmful to the environment before taking measures against it; with the precautionary approach, if there is reason to assume that an action may be detrimental, then it should be stopped or prevented from taking place at all. The precautionary approach first emerged, and has been most often applied, in the context of marine pollution. It arose from the recognition that past approaches to environmental protection had failed. Before the 1980s, policies to protect the marine environment were based almost universally on the assumption that the sea had an 'assimilative capacity' for pollution. It is now widely accepted, however, that such a permissive approach is not only inappropriate when applied to substances that are toxic, persistent or bioaccumulative; it is also dangerous and irresponsible. It is now also increasingly apparent that such is the diversity of the chemical compounds entering the marine environment and so complex is the network of species that live there that it is impossible to assess their impacts with any degree of certainty. The principle of precautionary action has been enshrined in an increasing number of legal instruments and conference agreements or declarations (see box). Encouragingly, the international community has moved towards acknowledging the need for precautionary t philosophies in relation to living creatures. The United Nations General Assembly Resolution 44/225 (1991), for example, recognized the role of the precautionary approach in recommending a moratorium on large-scale pelagic driftnets. The incorporation of a precautionary approach into the 1992 Earth Summit's Declaration on Environment and Development demonstrates clearly that it is appropriate not just to marine pollution, but to the entire spectrum of environmental policy making and to all types of impact that humans may have on the environment. More recently, at the first session of the UN Conference on Straddling Fish Stocks and Highly Migratory Fish Stocks held in New York in July 1993, key governments (e.g., New Zealand, Sweden, Chile, Canada, the Philippines and the US) were among a growing chorus of proponents for the precautionary approach as a means of protecting the environment and marine species. All the above recognize the many interrelated problems that must be addressed if we are to secure a future for the world's oceans. History has demonstrated clearly the need to protect whales from the threat of commercial whaling: for this reason alone, the indefinite whaling moratorium must be maintained. Such action by itself, however, will not be enough to protect the oceans in which whales live. Only by bringing an end to the wide variety of other human activities altering marine environments can we insure the continued survival of all species that live within them. Accepting the Precautionary Principle Since 1984 the principle of a precautionary approach to protecting the environment has been enshrined in no fewer than 16 conference declarations or agreements: -1984: Ministerial Declaration of the International Conference on the Protection of the North Sea Bremen, 1 November 1984. -1987: Montreal Protocol (to the Vienna Convention) on substances that deplete the Ozone Layer (adjusted and amended, London, 1990). -1987: Ministerial Declaration of the Second International Conference on the Protection of the North Sea, London, 25 November 1987. -1989: Parcom Recommendation (22 June 1989) of the Parties to the Paris Conventions for the prevention of marine pollution from land-based sources on the Principle of Precautionary action. -1989: UNEP Governing Council 15th Session (25 May 1989), Decision 15/27 on the precautionary approach to marine pollution, including waste dumping at sea. -1989: Decision on Dumping of Sixth meeting of Parties to Barcelona Convention for the Protection of the Mediterranean Sea against Pollution, Athens, October 1989. -1990: Ministerial Declaration of the Third International Conference on the North Sea, The Hague, 8 March 1990. -1990: UNEP Governing Council Decision SS.11/4B, A Comprehensive Approach to Hazardous Waste, Special Session, 3 August 1990. -1990: UN Economic Commission for Europe (ECE) Declaration on Sustainable Development in the ECE Region, and Joint Agenda for Action, Bergen, May 1990. -1991: Organization for African Unity (OAU) Bamako Convention on the Ban of the Import into Africa and the Control of Transboundary Movement and Management of Hazardous Wastes in Africa, Bamako, Mali, 30 January 1991. -1991: UN Economic and Social Commission for Asia and the Pacific (ESCAP), Ministerial Declaration on Environmentally Sound and Sustainable Development in Asia and the Pacific, Bangkok, October 1991. -1991: UN General Assembly Resolution 46/215 establishing a global moratorium on all large-scale pelagic driftnet fishing operations on the high seas, effective 1 January 1993, December 1991. -1992: Framework Convention on Climate Change, UNCED, Rio de Janeiro, June 1992. -1992: Rio Declaration on Environment and Development, UNCED, Rio de Janeiro, June 1992. -1992: Agenda 21, UNCED, Rio de Janeiro, June 1992. -1992: Paris Convention for the Protection of the Marine Environment of the North East Atlantic, 22 September 1992. Whales and climate change by Malcolm MacGarvin & Mark Simmonds Introduction It has been predicted that the world's climate will change if 'greenhouse' gases continue to increase in the atmosphere (Melillo et al., 1990). This change has the potential to affect ocean ecosystems drastically. In order to explore the possible consequences for cetaceans, we must consider the physical changes that climate change could cause in the world's oceans as well as the effects it may have on biological cycles. A number of predictions have been made about the probable responses of the world's oceans to climate change, including a rise in sea level and sea surface temperatures, and a change in ocean circulation patterns (Table 1). The circulation changes predicted may be compared with other events such as the Southern Oscillation (El Nino) in the Pacific (Mann and Lazier, 1991). The average surface temperature anomaly in the area of the Pacific most affected by El Nino is only 1.6C (Rasmussen and Carpenter, 1982) but even this has an important influence on coastal upwelling regions, where nutrients come to the surface and normally support highly productive marine ecosystems. Any change in climate may intensify the weather that creates the upwelling zones: namely high atmospheric pressure offshore and low pressure over the adjacent hot land mass (Bakun, 1990). The resulting alongshore winds drive the surface water offshore, to be replaced by nutrient-rich water from greater depths. Such current systems exist adjacent to northern California, west of the Iberian peninsula and north-western Africa, off south- western Africa and off the coast of South America. This increase in alongshore winds might compensate for a more general reduction in wind speed that others have predicted (Table 1), but the extent to which this might happen is just one of the imponderables in predicting the consequences of climate change. Detailed studies in the North Sea show the potential effects of climate change in shallow coastal seas (Karas and Kelly, 1989). They include a rise in sea level of 20cm by the year 2030; changes in tidal amplitude; changes in wave height, which are exacerbated if local wind speed increases due to a shift in the track of weather systems; changes in currents; alteration of areas of coastal erosion and deposition; the possible loss of tidal flats and wetland areas; higher temperatures, especially in shallow waters; and greater penetration of saline water into estuaries and rivers. Effects on marine life (a) Marine communities There are many examples of the limited ability of marine organisms to respond to changes in their physical or chemical environment. Estuarine species, for example, affected by increasing salinity or temperature, are unable to escape because they cannot pass through full salinity water to new habitats (Kennedy, 1990). Offshore there are also distinct bodies of water that harbour distinctly different marine communities. There are numerous examples of unusually warm weather temporarily displacing their boundaries towards higher latitudes. The most famous of these is the 'Russell cycle' in the western English Channel, where the advances and retreats of a southern warm-water community (typified by the pilchard) displace a more northerly community (typified by the herring). These fluctuations can be traced over hundreds of years (Cushing, 1982). Similar changes between sardines/pilchard and herring have been identified at numerous sites in the Pacific (Kawasaki, 1991). Climate change threatens to make such shifts to higher latitudes permanent. Boundary displacement caused by climate change has especially serious implications for Arctic communities, which are unable to retreat. As the International Panel on Climate Change (IPCC) Scientific Assessment noted (Melillo et al., 1990), warming is expected to diminish both the area covered by ice and the length of time that it occurs, with some models predicting an ice-free Arctic. Reduction of ice cover in the Arctic could have significant effects on the phytoplankton community that lives beneath the ice, a community that, in turn, supports an important food web including fish, seabirds and marine mammals (Gulliksen and Lonne, 1989). Away from the ice, the phytoplankton community in open water is also likely to be significantly affected, as a result of greater inputs of fresh water. A decrease in productivity here has implications for species throughout the food web, ranging from scavenging crustaceans to gray whales (Eschrichtius robustus). On the surface, polar bears (Ursus maritimus) depend on sea ice for hunting and breeding, while the loss of open areas within the ice would result in the reduction or elimination of yet another habitat supporting a rich and diverse fauna. (b) Marine productivity Changes in the base of the marine food web could result from changes in water temperature, surface salinity, turbulence, or nutrient concentrations. One ecosystem where sensitivity to climate change has reportedly been demonstrated is that of the central North Pacific. Here, phytoplankton biomass has been increasing since 1968, possibly because of an increase in winter winds and a fall in sea surface temperature (Venrick et al., 1987). However, an increase in phytoplankton should not necessarily be seen as a good thing (Bakun, 1990): while it may result in increased yields for the fishing industry, it could just as well result in ecosystem degeneration (perhaps via eutrophication or deoxygenation of the seabed). There are firm reasons to expect changes in species' composition of phytoplankton as well, since different species have different requirements (Harris, 1986). Major changes in the structure of the phytoplankton community around the British Isles since the 1940s have been cautiously attributed to fluctuations in climate. The data up to the early 1980s indicate that blooms of some species (diatoms) had become scarcer in these waters, while others (microflagellates) had increased (Robinson, 1983; Harris, 1986; Reid, 1990). An important factor controlling phytoplankton productivity is the existence (in all but shallow water), for at least part of the year, of a thin surface layer of less dense water that is vertically isolated from the rest. This effectively cuts off the phytoplankton from deeper water. This layer (the thermocline) can initially help to maintain the phytoplankton near the surface but it does make it more difficult for nutrients to be supplied from deeper waters. Typically, the phytoplankton eventually exhaust the supply of nutrients 'locked' in with them in the surface layer, thus limiting overall productivity. The depth and stability of this layer and the passage into it of nutrients are all affected by weather, temperature and influxes of freshwater. The consequences of climate change have been studied in some detail for this layer in the Arctic (Muller-Karger and Alexander, 1987; Alexander, 1988). The freshening of high- latitude seas by freshwater inputs and meltwater could be expected to lengthen the duration of both the thermocline and the halocline (a layer separating brackish water from denser, saltier water below); it could also increase the depth of the haloclirie and exacerbate the salinity gradient. It is predicted that the longer period of stratification will disadvantage larger phytoplankton such as diatoms and increase the numbers of smaller species. This will lengthen the food chain between primary producers and larger consumers, effectively reducing the biomass of the latter (Mann and Lazier, 1991). The greater depth of the halocline means that all phytoplankton will spend more time in less than ideal illumination, while the strengthened salinity gradient will make it more difficult for nutrients to enter the surface layer from below. Both of these factors are expected to reduce overall productivity. Overall, the IPCC (1990) noted that 45% of all marine production is in the zones of oceanic and coastal upwellings and subpolar regions. A change in productivity in these regions will have a major effect throughout the marine ecosystem. (c) The implications for top predators, including whales Changes in the abundance and community structure of phytoplankton will have knock-on effects through the food web. Copepods, for example, provide important food for other zooplankton, plankton-feeding fish and, indirectly, other parts of the food web. Research has shown that copepods prefer to feed on certain phytoplankton and protozoan ciliates (Kleppel et al., 1991; Lancelot et al., 1987; Daro, 1985), whose community structure may be affected by climatic change. Other protozoal zooplankton are also known to feed on particular species (Taniguchi and Takeda, 1988; Strom and Welschmeyer, 1991), which may be affected by climate change. Similarly, both commercial and non-commercial fish species can be expected to be affected by a reduction in productivity and by changes in the species' composition of plankton. Larval fish also feed on particular species of phytoplankton or zooplankton. Small cod (less than 5mm) rely on diatoms and young copepods; fish above this size feed on copepods alone (Last, 1978a); and small plaice feed almost exclusively on the larvae of the zooplankton Oikopleura (Last, 1978b). Unnatural shifts in phytoplankton composition and productivity will create effects that are passed up to the top predators, such as seabirds and marine mammals. When considering the implications for whales, four other factors should also be borne in mind: the apparent rate of climate change (3-4øC in high latitudes in 50 years--i.e. within one generation for some whale species) is outside the evolutionary experience of marine species. populations of many whale species are presently at extremely low levels. many whale species follow pre-determined migration routes, using specific bodies of water in particular locations for specific purposes. whales are also being affected by many other factors, as detailed elsewhere in this document. Therefore, whales could well be unable to adapt to the pace of the predicted changes. Conclusion Much of the evidence associating species' changes with climate change is based on correlation and extrapolation, but for the whales, and for other marine animals, there are rational grounds for expecting extremely serious, even catastrophic, effects to result if the climate changes as predicted. Table 1 1. Temperature change A doubling of atmospheric carbon dioxide could, in 50 years, result in an increase of 2-3øC in sea surface temperatures in any oceans, and of 3-4øC in high latitude seas. Winter sea surface temperatures already appear to be at their highest point in records stretching back 1.6 million years. Small, but possibly significant, exceptions where surface temperatures are predicted to fall are two areas of Arctic waters; the Irminger Sea, south-west of Greenland, and the Antarctic Weddell Sea. 2. Sea level rise Sea levels are predicted to rise, due to a combination of the thermal expansion of the oceans and the partial melting of land ice. The greatest increases are predicted to occur in the Northern Atlantic and on the northern edge of the Antarctic circumpolar current. Sea levels already appear to be at their highest point in records stretching back 175,000 years. 3. Ice cover and rainfall In northern high latitudes at least, the thickness, area and persistence of ice cover are all expected to be reduced. Precipitation is expected to increase in high northern latitudes, which will also exacerbate the salinity gradient to the sub-tropics. US and UK naval data from recent years suggest significant thinning of Arctic ice and reduction in the extent of sea ice. 4. Water circulation changes River run-off will increase due to the rise in precipitation. This will reinforce vertical layering in coastal waters and strengthen buoyancy-driven currents such as the Labrador current. The reduction in the temperature gradient between low and high latitudes may result in a reduction of wind stress. This would alter currents such as the North Atlantic Gulf Stream and the North Pacific Kuroshio current. In the Labrador region of the North Atlantic, the increased buoyancy of the more brackish water, coupled with reduced wind speed, is expected to reduce the rate at which water sinks. This will reduce the area's contribution to global oceanic circulation. In the Greenland and Norwegian Seas (the other sources of down- welling North Atlantic water), freshwater inputs are expected to increase less, while a possible retreat of the ice may result in a local increase in wind strength. In the north-east Pacific the northerly shift in the Aleutian low pressure weather system, and the increase in south-westerly winds, may result in significant current changes in the Gulf of Alaska, reversing the north- and west-flowing Alaska and Kenai currents. 5. Other changes An increase in atmospheric carbon dioxide could cause an increase in sea-water acidity of about 0.3 pH units. This may affect chemical processes including the concentrations of nutrients and heavy metals and cause increased toxic impacts of humus substances on marine organisms. (Sources: Mikolajewicz et al., 1990; Chappell and Shackleton, 1986; Fairbanks, 1989; Ruddiman and Raymo 1988; Raymo et al., 1989; Kawasaki, 1991; Mann and Lazier, 1991; Wright et al., 1986; Sibley and Strickland, 1985) Whales and ozone depletion by Alan Pickaver Introduction The ozone layer in the stratosphere shields the earth from damaging ultraviolet radiation in the 280-320nm wavelength range (UV-B). However, the layer is being destroyed by chlorine- and bromine-based chemicals such as chlorofluorocarbons (CFCs). Global decreases in stratospheric ozone are now well recorded and give real cause for concern (Waters et al., 1993). The rate of ozone depletion has consistently been underestimated (UNEP, 1991a). Recent satellite observations of stratospheric chemistry and ozone indicate that 'chlorine depletion of stratospheric ozone is of greater concern than previously thought' (Waters et al., 1993). Indeed, in 1992 the global average of total ozone was 23% lower than in the previous year- more than twice the loss predicted by the scientific models (at least 1.5% lower) (Gleason et al., 1993). In recent years ozone depletion during the southern spring has regularly resulted in ozone falling to half its normal level over Antarctica. In 1992, the greatest-ever depletion of the ozone layer was recorded, extending to the southern tip of Chile. Ozone depletion in northern latitudes is also increasing. In spring 1992 levels were down by 20% over the Arctic and in Europe (Harding 1992). Waters et al. report that levels for 1993 were even worse: in late winter ozone levels were 10% lower than the previous year; in some regions they were 20% lower. No similar trend in ozone depletion has yet been observed in tropical regions (UNEP, 1991a). However, because levels of UV-B normally reaching these regions are very high (Sharma and Srivastava, 1992), even small reductions in ozone will lead to substantial increases in total UV-B (UNEP, 1991a). Recent research indicates that the processes that result in loss of ozone at the poles are contributing to larger and longer term losses the world over (Waters et al. 1993). This work supports the theory that ozone-rich air circulates from the tropics to the poles, where the ozone is destroyed, before flowing back, ozone poorer, to lower latitudes. In addition, it has been argued that increased carbon dioxide (see Section 3) will enhance ozone depletion over the Arctic in winter (Austen et al., 1992). On the basis of ozone losses already observed, cumulative annual UV-B doses could be increasing by 5-10% per decade in the northern hemisphere and by 10-40% in the southern hemisphere (Madronich et al., 1991). It is calculated that absolute UV doses rise between 30 and 60ø in both latitudes at the summer solstice (Madronich, 1992). At McMurdo Station, Antarctica, 3-6 fold increases in UV irradiance associated with ozone depletion have been observed (Stamnes et al., 1992). Reported UV levels rose by more than 200% in southern Chile in October 1992 (World Meteorological Office, pers. comm.). The effects of increasing UV-B levels on marine organisms UV-B radiation is able to penetrate to depths of more than 25m in the waters around Antarctica (Smith and Baker, 1989) and to more than 30m in clear open oceanic water (Jokiel, 1980; Herndl et al., 1993). The activity of UV-B has now been measured to a depth of 30m in Belize (Herndl et al., 1993). Increases, therefore, are potentially damaging to phytoplankton-primary producers, fundamental to food webs-which are largely restricted to the upper layers of the ocean. UV-B has been shown to damage DNA and adversely affect the processes or apparatus involved in photosynthesis, including photosystem II and chlorophyll pigment production (Palenik et al., 1991), as well as ATP production (Vosjan et al., 1990); enzymes, nitrogen metabolism, ability to move and growth rate are also affected (Palenik et al., 1991; Cullen and Lesser, 1991). The peak of ozone depletion in spring coincides with the period of greatest primary productivity, which may mean that the total productivity of the oceans could be reduced and/or the species' composition of communities be altered (Hader and Worrest, 1991). This in turn would lead to changes in the availability and nutritional value of primary food sources (Karentz, 1991). Whales, being at the top of the marine food chain, will be especially vulnerable: both reproduction and growth could be affected because the whales would have to spend more time feeding (Voytek, 1990). (a) Productivity The first findings that productivity is affected by ozone depletion were published in 1992, following a six-week survey in the marginal ice zone of the Bellingshausen Sea during the southern spring of 1990 (Smith et al., 1992). The survey showed that phytoplankton productivity was reduced by at least 6-12%. This was associated with a thinning of the ozone layer, an increase in the ratio of UV-B to total irradiance and a related reduction in photosynthesis. Effects were noted to a depth of 25m. The authors calculated that there was a 2-4% annual decrease in production in the marginal ice zone and concluded that the ecological consequences of such a loss 'remain to be determined'. In temperate regions, growth and production of bacteria in surface waters has been shown to be limited by UV-B (Voytek, 1990). Bacteria are a significant food source at the base of the food web and play a critical role in vital recycling processes, such as the nitrogen and sulphur cycles. Voytek (1990) believes that changes in bacterial populations might lower recruitment rates of krill (Euphausia superba) and even change nutrient distribution among life in the ocean. Herndl et al. (1993) have also suggested that increased UV-B will influence the cycling of organic matter in the sea, and they have shown that marked suppression of bacterioplankton activity occurs to a depth of more than 10m. Calkins and Thordardottir (1980), observing phytoplankton ecosystems in Iceland, reasoned that solar UV radiation is a significant ecological factor because there was no capacity to withstand increased doses of UV in the habitat. While organisms may be able to avoid solar UV, or adapt to it, any such action reduces the energy available for other purposes, such as growth, and it is possible that current levels of UV-B are already stressful for phytoplankton communities. An increase in UV-B levels has also been shown to affect marine fish larvae (Hunter et al., 1982). The translucent, pelagic larvae of many of the world's most important fish--e.g. tuna (Thunnus spp.), mackerel (Scomber japonicus), flatfishes, pilchards (Sardina spp.), anchovies (Engraulis spp.) and cods-- are highly sensitive to UV-B radiation. Studies with anchovy showed that growth rate and survival of eggs and larval fish up to adulthood are particularly susceptible as they remain in the surface layers, migrating to deeper waters later on. Hunter et al. (1982) concluded that, in their natural environment, between 12 and 24% of the larvae in the water column would be exposed to damaging doses of UV-B. This study looked only at mortality and not at the long-term effects on larvae that survived exposure. Chapman and Hardy (1988) concluded that large increases in larval mortality could be expected if ozone were reduced by 16% of its pre-1970 levels; reductions greater than this are already being seen at spawning times. Worrest and Hader (1989) also concluded that fish losses could be as high as 9%, or more than 60 million tonnes of fish per year. Such losses are likely to have serious effects on the upper food chain of which the larvae are an integral part (Strickland et al., 1985). Studies of the impact of UV-B on phytoplankton in tropical waters recognized that multicellular organisms are far more vulnerable than unicellular ones, because in single-celled organisms any lethal effect is diminished as the cell reproduces (Helbling et al., 1992). This bears out previous results that show that many species of sponges, bryozoans and tunicates are sensitive to UV-B (Jokiel, 1980). UV-B has also been linked to coral bleaching on the Great Barrier Reef, the Pacific coast of Panama, and in the Caribbean Sea (Lesser et al. 1990). Increased exposure to UV-B causes loss of pigment and productivity, and reduced numbers of the symbiotic algae (zooxanthellae) within each coral polyp. In the past, reef-building corals, and other marine organisms, in tropical waters were believed to be resistant to UV-B radiation, due to the normally high ambient concentrations in these regions (Dunlap et al., 1986). Now it is thought that increased levels of UV-B are likely to have serious effects on coral reefs. One of the consequences of a loss of phytoplankton productivity could be that less carbon dioxide is taken up from the world's oceans; this would lead to increased levels of carbon in the atmosphere, thus adding to global warming. (b) Species' composition Species' composition may also be altered by UVB. Helbling et al. (1992) demonstrated that UV-B inhibits some planktonic species more than others and even induces resting spores of the diatom Chaetoceros. Smith et al. (1992) verified that different species grow at different rates on exposure to UV-B, and discovered that variations also occur among species at different depths. Cullen and Lesser (1991) demonstrated variations among species relating to nutrient status: a nitrate-limited culture of a diatom, Thalassiosira spp., proved to be nine times more sensitive to UV-B than a nutrient-replete culture. Mitchell and Karentz (1990) also showed that, within species, both sensitive and resistant, the ability to withstand UV-B light varies considerably. The implications of such changes in the composition of the phytoplankton community will not be restricted to those species directly involved, but may also affect predators, many of which have marked feeding preferences. (c) Metabolism Different organisms respond in different ways to UV-B radiation. Dohler (1992) has shown that populations of the marine alga Phaeocystis pouchetii from the Wadden Sea are very sensitive to UV-B when assimilating nitrogen compounds to produce amino acids, the building blocks of proteins. After UV-B exposure, uptake of Nnitrate, and to a lesser extent N-ammonium, was very low or totally blocked. The concentration of total amino acids also decreased, leading to a reduction in overall protein production. However, Dohler et al. (1991) showed that other communities of algae responded by taking up less N-ammonia than N-nitrate; other phytoplankton, such as Ceratium, Coscinodiscus and Noctiluca, also behaved differently (Dohler, 1992). (d) Photoadaptation Organisms vary in their ability to move away from damaging levels of UV-B or to protect themselves from it in other ways. Bollens and Frost (1990) have shown that Acartia hudsonica, a planktonic copepod, cannot control its diurnal migration to surface waters and, thus, cannot avoid UV-B radiation. Herndl et al. (1993) also concluded that bacterioplankton do not have any adaptive mechanisms against UV-B. Hader and Worrest (1991) have shown that, in Euglena, response to both light and gravity is impaired by exposure to UV-B, causing the organisms to swim randomly. If the ability to orientate is damaged, organisms may not be able to develop strategies to move away from the damaging radiation. In any case, the authors conclude that few organisms capable of movement possess a sensor for UV-B radiation and thus most are unable to escape. The same species in different ecosystems will also respond differently to increased UV-B. In Antarctic waters Phaeocystis pouchetii produces substances that absorb UV radiation, but the same species in northern waters shows no such response (Marchant et al., 1991). However, the production of UV-absorbing substances is not without cost, since it diverts energy from other processes. Phytoplankton depend upon light to photosynthesize, so moving away from UV-B is not a viable option. Conclusion In the southern spring of 1992-93 and the northern spring of 1993, stratospheric ozone depletion reached the highest levels ever recorded. With atmospheric chlorine and bromine levels continuing to rise, stratospheric ozone depletion rates are expected to become significantly worse, perhaps doubling, during the next decade (UNEP, 1991b). Thus, it appears that all the inhabitants of the earth, including those living in surface sea and oceanic waters, and particularly those living from 30ø polewards, will be subjected to increasing concentrations of UV-B. Research has shown that UV-B can damage all groups of organisms. Those at the bottom of the food web are the most vulnerable and the most important for the total productivity of the oceans and for the diversity of life within them. Decreasing primary productivity, already measurable in Antarctic waters, could reduce fish stocks and food available for other species higher in the food web, including marine mammals. Those baleen whales that feed primarily on plankton may suffer more directly from a shortage of food. The different ways in which marine organisms respond to UV-B indicate how easily species' composition could alter in a marine environment exposed to a greater intensity of UV-B, since many organisms have no ability to adapt to withstand increased levels of the radiation. The effects will be global in nature. Ozone depletion, once thought to be a phenomenon affecting only southern waters, is now being observed in northern latitudes. In the tropics only slight decreases in ozone levels will be needed to increase UV-B concentrations substantially, due to the high ambient UV-B levels normally found there. The potential for large-scale disruption of life in open oceanic waters is very real, and early indications are that it may even have started. Whales and organochlorine pollution by Paul Johnston and Mark Simmonds Introduction Of all the indirect threats facing whales and other marine mammals, their contamination with synthetic organochlorines (chemicals containing carbon and chlorine) has received the most attention. Research has shown that organochlorines have the potential to disrupt a variety of biochemical and physiological processes in mammals and that marine mammals have a limited capacity to detoxify these chemicals. Although some organochlorine compounds are produced naturally, their functions remain largely unknown (Gribble, 1992). It seems likely that many act as chemical messengers and defence agents (Hay and Fenical, 1988). The global organochlorine budget is now dominated by the industrial production, use and disposal of these compounds. Industrial production of chlorine began in the late 1920s and is currently around 40 million tonnes per year. Either directly or indirectly, this chlorine is converted to thousands of organochlorine chemicals, the majority unknown in nature, which have become global pollutants. The effects of chlorofluorocarbons (CFCs) on the ozone layer are now well known. As described in Section 4, loss of ozone leads to increased UV-B radiation reaching the earth's surface; this can interfere with the productivity of phytoplankton and hence have a profound impact on aquatic ecosystems. Other organochlorines accumulate in specific environmental compartments. High levels concentrate, for example, in the marine microlayer: a thin film on the sea's surface that is rich in natural fats and oils. This is a site of intense primary production (Hardy, 1982). The development of eggs and larval stages of fish also takes place in association with the microlayer (Roesijadi and Spies, 1987). Organochlorines may therefore affect cetacean populations indirectly, by changing the structure of food webs: for example, fish embryos exposed to a contaminated microlayer may be killed or fail to develop normally. Alternatively cetaceans may be affected directly, when they eat contaminated food: marine top predators appear to be particularly vulnerable to the effects of organochlorines. Persistent organochlorines Some organochlorine compounds, in particular the polychlorinated biphenyls (PCBs), DDT and its metabolites, have been well studied. Because production of these persistent chemicals has been banned in most developed countries, there is a tendency to regard them simply as historical problems that are now under control. In fact, DDT and other organochlorine pesticides are still in use in developing countries (Johnston and Stringer, 1992) and around 60% of the estimated 2 million tonnes of PCBs produced since 1929 are still in use or in dumps and landfills (Tanabe, 1988a). Hence, these compounds are still entering the environment, and are slowly being distributed across the planet (Ballschmiter et al., 1989). Another group of organochlorines that have been extensively researched are the highly toxic dioxins, which are produced as unwanted byproducts of chlorine chemistry and combustion (Rappe and Buser, 1989). These chemicals too are now known to have a global distribution. Recently, a new group of organochlorines have been identified as widespread contaminants of marine mammals. The compounds tris(4- chlorophenyl) methanol (t-CP-OH) and tris(chlorophenyl)-methane (t-CP-H) have been found in fatty tissues from animals in both the northern and southern hemispheres (Jarman et al., 1992). The source of these chemicals has not yet been identified, although it is suspected that they arise from the production of dyes and polymers using chlorine. Distribution and uptake In general, cetaceans in the northern hemisphere carry higher body burdens of organochlorines than those in the south; in fact, the waters of the north Atlantic Ocean are regarded as the single biggest global reservoir of PCBs (Larsson, 1985). The semi-volatile organochlorines, including DDT and its metabolites, PCBs and dioxins, are transported through the atmosphere (Standley and Hites, 1991). In the North American Great Lakes, direct inputs of organochlorines have been reduced, but the constant exchange between sediments, water and atmosphere means that these chemicals are still being released from the lakes into global ecosystems (Swackhamer and Eisenreich, 1991). Because organochlorines dissolve readily in fat and resist degradation they build up in fatty tissue; levels of bioaccumulated organochlorines increase as they pass up the food chain. Marine mammals are at the top of short direct food webs and are thus particularly likely to accumulate high levels of these contaminants. Organochlorines have become ubiquitous contaminants of cetaceans, which have large fat reserves in the form of blubber. In general, pollutant levels rise according to the level of the food web at which the animal feeds (Borrell, 1993). Baleen whales, for example, which feed on zooplankton, tend to have lower concentrations of contaminants in their tissues than porpoises or dolphins, which feed on fish and squid nearer to the top of the food web. Metabolism of organochlorines Marine mammals appear to be less able to metabolize persistent organochlorine chemicals than terrestrial mammals or birds (Boon et al., 1992). No data have been reported for cetaceans, but in seals highly persistent derivatives of PCBs are formed. Although the effects of these metabolites have not been evaluated, one persistent DDT metabolite may cause widespread hyperplasia of the adrenal cortex in seals (Brandt et al., 1992), leading to a range of abnormalities. Effects of organochlorine contamination Quantitatively, PCBs and DDT and its t metabolites are the most abundant chlorinated aromatic compounds contaminating aquatic ecosystems. Dioxins, which are present at very much lower levels are, however, more toxic and they too may present a significant hazard to marine mammal populations (Tanabe, 1988b). Work in the laboratory has established that organochlorines are toxic to a variety of organisms, and abnormalities in wildlife populations have also been linked to high organochlorine levels. However, in the field, precise relationships between cause and effect are difficult to determine, because of the interactions between the many compounds now contaminating natural populations (Gilbertson, 1989). Data on organochlorines in cetaceans are sparse but high levels of these contaminants are frequently found. Cetaceans themselves, however, are difficult to study and their population structures are poorly understood, which makes it extremely difficult to determine the exact impact organochlorine pollutants have on them. However, the available information does indicate that cetaceans show effects similar to those recorded in other animals exposed to organochlorines. a) Effects on reproduction Organochlorines have been found to cause a range of effects in laboratory animals including reproductive failure, alterations to the immune system, and carcinogenesis. Similarly, in the wild, high levels of organochlorines have been associated with widespread reproductive failure in Wadden and Baltic Sea seals and in Californian sea lions (Zalophus californianus) (Gilbertson, 1989). At the highest levels of contamination in Baltic grey seals (Halichoerus grypus), deformities of the skull have been found (Bergman et al., 1992) together with changes leading to obstruction of the reproductive tracts (Helle et al., 1976). The relationship between organochlorine pollution and reproductive failure in seals has been confirmed in the laboratory: harbour seals (Phoca vitulina) fed on fish from the Wadden Sea that were contaminated with organochlorines produced fewer pups than another group fed on less-contaminated fish (Reijnders, 1986). Pollution-induced reproductive and other problems are not as well-documented in cetaceans. However, in the population of beluga whales (Delphinapterus leucas) from the St Lawrence River and Estuary, Canada, high levels of contamination are associated with population decline, low reproductive success and a high prevalence of various tumours (Beland et al., 1991). These observations are consistent with the knowledge that organochlorines can interfere with hormonal systems that control reproduction and development (Colborn and Clement, 1992) and that are also important elements of homoeostatic mechanisms in adult animals (Goldstein and Safe, 1989). Dioxins, for example, can increase the metabolism of both testosterone and oestrogen, thus impairing important functions at the cellular level and in the organism as a whole. Reduced levels of testosterone in Dall's porpoise (Phocoenoides dalli) have been linked to elevated levels of PCBs and DDE in the blubber (Subramanian et al., 1987). This evidence of disruption of cetacean hormonal systems by organochlorines is consistent with effects found in other mammals. Cetaceans are generally considered to be highly vulnerable to reproductive disturbance as a result of exposure to organochlorines (Tanabe & Tatsukawa, 1991). Levels of organochlorines in cetaceans appear to increase with age until sexual maturity. Even then, in males, they continue to rise; but in females they decline markedly (Cockcroft et al., 1989). It has been estimated that 80% of the mother's organochlorines, mobilized from the blubber to her milk, may pass to the first-born dolphin calf during feeding. Even in relatively uncontaminated dolphins from the southern hemisphere, it has been suggested that the concentration of organochlorines passed on in mother's milk would be sufficient to kill the firstborn calf (Cockcroft et al., 1989). Some circumstantial evidence exists of this taking place in Atlantic species: a bottlenose dolphin calf (Tursiops truncatus) found dead in the UK carried extremely high levels of organochlorines (Morris et al. 1989). Therefore, direct and indirect effects upon reproduction, a primary determinant of population dynamics, can be anticipated in cetaceans. The threshold (if any) for effects remains unknown. The significance of interference with hormonal pathways is extremely difficult to determine from the existing information. More worrying still, organochlorines are still spreading across the planet and, as they do so, general exposure to them is likely to rise. b) Effects on the immune system Chlorinated aromatic compounds have been shown to cause immune dysfunction in a variety of animals. The precise biochemical mechanism is not clear (Vos and Luster, 1989) but it is thought to involve the Ah receptor system. The immune system provides a defence against invasion by disease-causing microorganisms and by parasites, hence any disruption of the system has potentially serious consequences. This has been thrown into sharp relief by recent mass mortalities in marine mammal populations in various regions. In 1988 an estimated 18,000 harbour seals died in the North Sea and adjoining waters. It is now known that a virus precipitated this event, but it is believed to have been exacerbated by organochlorine contaminants, although this has not been unequivocally established. In the Baltic, Olsson et al. (1992) stated that there were no indications that organohalogen compounds studied were at higher levels in infected animals. However, a recent review of the die-off (Heide-Jorgensen and Harkonen, 1992) did not rule out pollution as a contributing factor. Indeed, Simmonds et al. (1993) found that in UK harbour seal populations, the highest mortalities were among those seals with the highest organochlorine burdens, while Hall et al. (1992) found higher organochlorine levels in seals that died than in those that survived. Similarly, the mass mortality of bottlenose dolphins on the east coast of the US in 1987-1988 was originally attributed to the natural algal poison brevitoxin. Antibodies to a morbillivirus, however, were detected in some affected dolphins (Geraci, 1989) and very high levels of organochlorines in blubber were also found (Kuehl et al., 1991), leading workers to recommend that the precise role of these contaminants be evaluated further. The mass mortality of Mediterranean striped dolphins (Stenella coeruleoalba) in 1990 and 1991 was also associated with high levels of organochlorine contamination and infection with a morbillivirus (Aguilar and Raga, 1991). Organochlorine levels in dolphins that died were higher than those in animals sampled both before and after the event (Borrell and Aguilar, 1992). It is difficult to determine exactly whether organochlorine contamination affected the severity of the disease in each case. Nonetheless, the observations to date are consistent with organochlorines exacerbating the effects of pathogenic organisms. Conclusion Global contamination of the environment with organochlorine compounds has the potential to affect cetacean populations adversely, both directly and indirectly. Effects on primary production through increased levels of UV-B radiation or contamination of specific habitats, such as the marine microlayer, may affect the cetaceans' food resources. More directly, contamination of the food chain, through which cetaceans accumulate organochlorines, may have an impact on their biochemical systems. As a result, impairment of reproduction and the immune system could profoundly affect individuals, with wide-ranging effects on cetacean populations. Whales and a decreasing food supply by Matthew Gianni Introduction The amount of fish caught in the world's oceans has increased dramatically in recent years, from approximately 20 million tonnes per year in the early 1950s to almost 90 million tonnes per year today. This increase has come about largely as a result of the extensive industrialization of the world's fisheries- advances in the technology to harvest, process, store and transport fish and fish products-and has often been touted as an answer to world hunger-the so-called 'blue revolution'. Now the United Nations Food and Agricultural Organization (FAO) estimates that most commercially important stocks of fish are either fully or over-exploited (FAO, 1992). There have been numerous instances of fish stock depletion and collapse as well as adverse effects on a range of marine species in specific areas and specific fisheries, e.g. North Sea herring (Clupea harengus) and Barents Sea capelin (Mallotus villosus). On the whole, however, there is an enormous lack of information on the long-term impact that the tremendous increase in fishing over the last several decades is having on marine ecosystems and biodiversity. As stocks of slower-growing, higher-value fish such as cod (Gadus morhua), haddock (Melanogrammus aeglefinus) and other demersal species become fully or over-exploited, fishing effort is often re-directed to lower-value, shorter-lived pelagic species. Five such species--Japanese pilchard (Sardinops melanosticta), Peruvian anchoveta (Engraulis ringens), walleye pollock (Theragra chalcogramma), South American pilchard (Sardinops sagax), and Chilean jack mackerel (Trachurus murphyi)--accounted for well over half of the increase in global catches throughout the 1980s (FAO, 1992). Many of these species are lower on the marine food chain than those previously targeted, and if the trend continues, it is likely to have major implications for top predators such as marine mammals. The consequences of over-exploited fisheries Food availability is one of the major factors that limit the distribution and abundance of marine predators. Population crashes in seabirds and some marine mammals have been associated with changes in the availability of the fish they eat. Seabird crashes have been extensively documented and studied. Good historical data exist for the British Isles, which have more breeding seabird colonies than any other European country; here, populations of several seabird species show declines that follow those of fish stocks (Furness, 1989). Elsewhere in the world, Peruvian seabird numbers have changed in a way that can be related to changes in anchoveta stocks, and pigeon guillemots (Cepphus columba) show large population declines in parallel with reduced catches of pollock. Marine mammals, and in particular whales, are far more difficult to monitor and have not been so well studied in this respect. Nonetheless, there is some evidence to suggest that both seals and whales are affected by changes in the abundance of fish. There has been a massive build-up in the fishing industry in both the Bering Sea and the Gulf of Alaska, as well as changes in the abundance, both absolute and relative, of several fish stocks. Populations of herring, Pacific Ocean perch (Sebastes alutus), Atka mackerel (Pleurogrammus monopterygius) and other rockfishes have declined, while those of various flatfish and pollock have fluctuated considerably: some are now more abundant than 15-20 years ago (Alverson, 1992). The area extending from California across the Pacific Rim to northern Japan is the breeding range of Steller sea lions (Eumetopias jubatus). The species was estimated to number 245-290,000 animals in the late 1950s, but a range-wide survey conducted in 1989 accounted for only 116,000 sea lions, less than one half of the earlier estimate (Loughlin et al., 1992). In the area of Alaska between the Kenai Peninsula and Kiska Island, repeated counts show the population to have decreased by 61% between 1985 and 1991 (Merrick et al., 1992). Other pinnipeds in Alaska have also declined. The number of Northern fur seal (Callorhinus ursinus) pups born on the Pribilof Islands has been falling since the 1950s (Fowler, 1990). Harbour seals (Phoca vitulina) on the largest rookery in the Gulf of Alaska declined by 85% between 1976 and 1988, and declines have been reported for other areas as well (Pitcher, 1990). A workshop held in Alaska in 1991 to answer the question 'Is It Food?' concluded that the declines are indeed food related (Anon., 1991a). This food shortage has manifested itself in a lower rate of survival for juvenile Steller sea lions, and possibly for breeding females as well. A simple model constructed by the working group suggested that this could be the result of a lack of young walleye pollock, one of the main sources of food for Steller sea lions. The abundance of young pollock has been declining since the 1970s (Springer, 1992). With fur seals, too, the available data indicated a poor rate of survival for juveniles. For this species, the cause appears to be insufficient food to support young seals during their migration through the Gulf of Alaska. (The possible impact of food shortages of harbour seals in Alaska could not be evaluated, since very little is known of these seals, other than the fact that they have declined rapidly [Anon., 1991a]). The capelin fishery in the Barents Sea yielded in excess of a million tonnes annually in the 1970s. Then, in 1986, the stock virtually disappeared. This may have been due to over-fishing (quotas had been frequently exceeded) or to some sort of natural change, but in any case the result was immediate and dramatic. The Barents Sea also supports one of the three major stocks of harp seal (Phoca groenlandica) breeding in the White Sea. The stock is thought to have numbered some three million in the early part of this century, but hunting reduced it to less than one million by 1980 (ICES, 1990). The seals move into the Barents Sea in the summer to feed. While their diet there has been little studied, research on harp seal diet in the north- west Atlantic suggests they normally feed on a wide variety of fish and crustaceans, including capelin, polar cod (Arctogadus glacialis), herring and shrimp (Wallace and Lavigne, 1992). In the winter following the collapse of the capelin stock in the Barents Sea there was the largest influx of seals ever recorded on the coast of Norway. Whereas relatively low numbers of seals had drowned in coastal fishing nets in the years up to 1986, almost 60,000 were recorded in 1987 and over 20,000 in 1988; real mortality has been estimated at 2.5 times that number. Also, seals caught before 1986 were of all ages and in normal condition, while the seals from 1987 were mostly young seals in very poor condition (ICES, 1990). Russian scientists have conducted aerial surveys of harp seals on the breeding patches in the White Sea for many years. While the data cannot be used to estimate total population size, they do provide an indication of relative abundance. Between the surveys in 1985 and 1988, the number of breeding females decreased by about 50%, from 140,000 to 71,000 animals (ICES, 1990). It is difficult to escape the conclusion that the harp seals were suddenly deprived of one of their major sources of food following the crash of the capelin stock in 1986. As a result, they were forced to search elsewhere for food, and many ended up in the coastal waters of Norway. The resulting mortality was a significant proportion of the total population. Capelin also form an important part of the diet of humpback (Megaptera novaengliae), fin (Balaenoptera physalus) and, particularly, minke (B. acutorostrata) whales in the North Atlantic. Capelin is the target of a very large commercial fishery in this area (ICES, 1993). The decline of capelin off Newfoundland in 1978, possibly related to fishing, changed the distribution of humpback whales in that area (Whitehead, 1987) and may have lowered their reproductive rate (Whitehead, 1982). North Atlantic stocks of herring, some of which have also experienced substantial fluctuations as a result of commercial fishing, are important prey for fin and minke whales (ICES, 1993). Other examples of fish important to whales as food and subject to a commercial fishery are horse mackerel (T. symmetricus), anchovy (E. capensis), and pilchard (S. ocellatus) for Bryde's whales (B. edeni) off South Africa and in the North Pacific, and mackerel (Scomber japonicus) for sei whales (B. borealis) in the southern North Pacific. Krill (Euphausia superba) plays a central role in the Antarctic ecosystem and any increase in the fishery that affects the abundance of the species could have an impact on baleen whales (ICES, 1993). Entanglement Another serious threat to many cetacean species is entanglement in fishing gear. The most extensive survey to date was reported at a scientific workshop in 1990, sponsored in part by the IWC. It reviewed entanglement of cetaceans in passive fishing gear, including gillnets, traps and driftnets (IWC, 1990). The workshop concluded that, for seven populations of small cetaceans, the numbers killed were so high that the populations would no longer be able to sustain themselves. For several other populations, including right whales (Balaena glacialis) in the North Atlantic and North Pacific, mortality was described as 'possibly not sustainable'. Overall, for some 60% of the populations considered, it was not possible to estimate the impact of entanglement in fishing gear, although some level of mortality was known to occur. Among the baleen whales, this was the case for southern right whales off New Zealand, pygmy right whales (Caperea marginata) off South Africa, minke whales in several areas, Bryde's whales off Brazil, and humpback and fin whales in the Indian Ocean. But, even if mortality from entanglement seems to be at a low level, it must be evaluated in conjunction with other causes of death resulting from human activity. Kraus (1990) used a variety of sources to estimate mortality of North Atlantic right whales. He suggested that one third of the numbers of deaths in the first four years of life could be due to a combination of entanglement and collisions with boats, and that such events could be a significant factor in the failure of this population to recover. Clearly, even relatively low levels of mortality from a single cause can become significant when a species is affected by other causes of death, especially in a small population. The issue of high-seas driftnet fishing has been the subject of considerable international attention in recent years. Efforts to assess its impact on whale species have led to some disturbing conclusions. Many whale species are known to inhabit 'driftnet fishing areas of the north Pacific. A major review of the impact of driftnet fishing expressed concern that, although some species of whales had been observed entangled in the nets, 'most large whales caught in driftnet fisheries would not be observed' because 'whales that become entangled on the high seas probably tow pieces of driftnet away from the fishing area and die at a later time' (Anon, 1991b). There is little reason to doubt that the difficulties in observing whale mortality in the large-scale driftnet fisheries of the north Pacific hold true in other areas of the world's oceans. Although the United Nations acted decisively in calling for a moratorium on all large-scale pelagic driftnet fishing on the high seas by the end of 1992 (United Nations resolution 46/215), it remain's to be seen whether the moratorium will be fully and effectively implemented. Even if it is, a decade or more of extensive high seas driftnet fishing may have taken a heavy toll. Conclusions Fisheries in the world's oceans continue to expand and have already resulted in declines in a number of fish stocks, including several of importance to marine mammals (FAO, 1991). As this trend continues, it is only to be expected that marine mammal populations will be vulnerable to food shortages in the years ahead. Seal and sea lion populations in both the Barents Sea and the Bering Sea/Gulf of Alaska have suffered dramatic declines during the past two decades, and the evidence points to food shortages being a major factor in the declines. At the same time, there have been large increases in commercial fisheries in each area. Entanglement in fishing gear may also cause a substantial mortality rate in some populations, which may leave them unable to sustain their numbers. While the relationship between the build-up of fisheries and declines in marine mammals is unlikely to be simple, these examples certainly suggest that the impact of huge and rapidly expanding fisheries is not limited to the target stock, but has consequences for a wide variety of species, including whales. Whales and disease by Sue Mayer Introduction Population biologists have tended to believe that the critical factors controlling the size of an animal population are predator-prey interactions and the availability of food. The role of disease has been considered less important. Intuitively it might seem that, to survive, a species must come into balance with all the various diseases it faces. However, if there are pressures that reduce a species' ability to resist an infection, or a new disease is suddenly introduced into a ecosystem, the balance may be disturbed; as a result, a population may go into sudden decline, from which recovery is difficult. The mass mortalities of marine mammals around the world in recent years raise the question of whether whale populations could also be affected by such events and whether they would recover in the aftermath. These 'die-offs' include the death of over 17,000 harbour seals (Phoca vitulina) in the North Sea in 1988, several thousand striped dolphins (Stenella coeruleoalba) in the Mediterranean during 1990-91, and perhaps half the bottlenose dolphin (Tursiops truncatus) population of the US east coast in 1987 and the Gulf of Mexico in 1990 (Simmonds, 1992; Harwood and Hall, 1990). Many other pressures such as pollution and competition for food are now having an impact on marine mammals, and whether they will be able to withstand old or new diseases in these new circumstances requires consideration. There has been little study of wildlife diseases and their impact on the regulation of populations, even in terrestrial species (Plowright, 1988a; 1988b). Studying marine mammals is, in this respect, both more difficult and less advanced. Diagnosing causes of death There is little information about the causes of death in whales. When whales die their bodies may sink and never be seen, or they may reach land but be in a state of decomposition that makes post-mortem diagnosis of death impossible. There are also practical difficulties in performing post-mortem examinations on large whales. For example, 32 dead sperm whales (Physeter illacrocephalus) were found stranded or seen floating in northern European waters during 1988. This compares to a total of 14 recorded strandings in the Baltic and North Sea region between 1900 and 1969 (Christensen, 1990). Yet it was not possible to determine the cause of death of any of these animals, or the significance of the increased numbers stranded. Indeed, those stranded may have been only a tiny proportion of those dying in that period. Most work dealing with the effects of disease on wildlife populations has concentrated on infectious diseases (those caused by parasites or bacteria and viruses, for example) as they are often easier to characterize and are less sporadic in nature than non-infectious diseases (conditions such as tumours and heart failure). This does not mean that other causes of mortality are not important. For example, it has been suggested that the death of a number of humpback whales (Megaptera novaengliae) around Cape Cod in 1987-88 was caused by ingesting toxins produced by algal blooms (Geraci et al., 1989). However, in this case, as in many of the other recent die-offs, the role of pollutants in reducing the animals' ability to resist infection may be more important than has been acknowledged (Simmonds, 1992). Natural mortality and population size Population biologists, reluctant to accept the importance of disease in regulating population numbers, have argued that since there is a gradual evolution of a species in parallel with the pathogens affecting it, there will be little effect on the dynamics of the population (May, 1986). For instance, natural selection could favour those individuals with more resistance to a disease or with behaviour patterns that reduce the likelihood of contracting the disease. However, this ignores the possibility that much more rapid mutations may be occurring in the infectious organism, which may not necessarily allow the host population time to evolve resistance to it. Or a pathogen may suddenly make a jump across species, as is thought to have happened when canine distemper virus caused the death of seals (Phoca sibirica) in Lake Baikal in 1988 (Visser et al., 1990). It also ignores examples such as rinderpest (a disease caused by a virus similar to that which caused the mass mortality of seals in northern Europe and striped dolphins in the Mediterranean) in wildebeest (Connochaetes taurinus) in East Africa, where the regulatory effect of the disease in the 1950s has been well documented, although only by retrospective analysis of data (Plowright, 1988b). When an animal is infected with a virus or bacteria there are two possible outcomes: it may die, or recover and develop some level of immunity to re-infection. The degree to which a population will be regulated by an infectious agent depends upon the mortality rate it causes or the extent to which it interferes with reproduction (Anderson, 1991). For instance rabies, which causes high mortality, significantly depresses the numbers of foxes (Vulpes vulpes) in Europe, well below the number their habitat could support (Anderson et al., 1981). Our understanding of infectious disease processes in whales and their significance to these creatures is poor. Parasitic infections in cetaceans are extremely common (Dailey, 1986; Baker, 1992) and their impact can be significant. In one population of spotted dolphin (Stenella attenuata) a nematode worm (Crassicauda sp.) was considered responsible for 10-14% of all natural deaths (Perrin and Powers, 1980). Another study found that 90-95% of fin whales (Balaenoptera physalus) examined were infected with the giant kidney nematode (Crassicauda boopis) and the infection was considered severe enough in over 85% of animals to cause eventual death from kidney failure (Harwood and Hall, 1990). The evolution of an immune system in vertebrates has probably been one of the most important factors enabling populations to survive all the infectious agents they encounter (Anderson, 1991). Such a strategy is more important for long-lived species with low rates of reproduction, such as whales, than for species that have high reproductive rates and are generally short-lived, such as insects and fish. If the disease agent itself impairs the immune system or has sophisticated mechanisms for avoiding it, there may be very damaging results for the population in the future, which could completely change predictions of population growth made before a disease became evident. For instance, the impact of HIV, the virus that causes AIDS, in humans may reverse the signs of population growth in certain countries (Anderson, 1991). In this case one of the reasons for such a devastating impact is the low population growth rate of humans (late maturing, low number of offspring) coupled with an infection that does not stimulate a natural immunity, has a long latency and from which recovery is unlikely. While there is no evidence that such an infection exists in whales, they share some common characteristics with humans, indicating that disease could play a role in regulating their population size. A previously unrecognized viral infection in seals that caused the deaths of many thousands of harbour seals in the North Sea during 1988 (Kennedy, 1990) highlighted the fact that marine mammals are vulnerable to infectious diseases. Similar, but different, morbilliviruses are considered to have caused the death of hundreds of seals in Lake Baikal, Russia, in 1988 and thousands of striped dolphins in the Mediterranean in 1990-91 (Domingo et al., 1992; Van Bressem et al., 1991). Morbilliviruses have also been found in harbour porpoises (Phocoena phocoena) in Europe (Kennedy et al., 1992) and antibodies to them have been detected in bottlenose dolphins in the US (Geraci, 1989) as well as in other species of seal such as crabeater seals (Lobodon carcinophagus) in the Antarctic (Bengtson et al., 1991). Post-mortem signs of morbillivirus have also been seen in a common dolphin (Delphinus delphis) (Baker, 1992). Responses to disease The apparent increase in mass mortalities during the last decade raises questions about whether environmental conditions may influence the course of a disease (Simmonds, 1992). Many factors influence a population's or an individual's ability to recover from a disease. First, nutritional status is important. In humans, this is one of the main reasons for a difference in survival between children in less-developed countries and those in developed countries when exposed to the same infectious agent (Anderson and May, 1979). The same principle will apply to whales. A shortage of food may also reduce a population to such small numbers that an outbreak of disease may then result in the population's extinction. This is what appears to have happened to black-footed ferrets (Mustela nigripes) in the US in the 1980s. Their numbers had fallen dramatically following a decline in their prey, the prairie dog (Cynomys gunnisoni). Then there was an outbreak of canine distemper among an apparently recovering population of ferrets, which resulted in their virtual extinction in the wild (Plowright, 1988b; May, 1986). Recovery will also be determined by the development of an effective immune response. If the immune response is hindered by other infections, toxin accumulation or the disease process itself, recovery may be delayed or made impossible. Equally importantly the ability to reproduce and replace lost individuals will determine the long-term response of a population to a disease. If the reproductive system is compromised by the disease itself, or by the effects of a toxin then, again, recovery may be delayed or not possible. For example, following the seal die-off in 1989, scientists studied the extent of organochlorine contamination in the seals, which may suppress the immune system or cause reproductive disorders (see Section 5). Levels of organochlorines were found to be higher in animals dying from the disease than in those surviving (Hall et al., 1992) and highest among those populations that suffered the highest mortality (Simmonds et al., 1993). It is also possible that a combination of factors may influence the way an animal responds to an infection. In isolation, none of the factors may seem critical. When combined, however, they may result in a population being 'stressed', becoming more susceptible to disease and less likely to recover from it. Conclusion Diseases may play an important role in regulating marine mammal populations, including those of whales. Some characteristics of whales make their populations especially vulnerable to events that result in the sudden death of large numbers of individuals (Harwood and Hall, 1990). Whales are long-lived and therefore sensitive to changes in adult survival rates. Their patterns of social organization involve groups gathering to breed and feed, so the introduction of a pathogen at this time could lead to large numbers of deaths. These behaviour patterns also mean that even if only a few animals are involved, they will encounter one another frequently, thus providing many opportunities for the disease to spread. Low rates of population growth also contribute to whales' vulnerability, especially when other factors may be affecting a population and undermining its ability to recover. Therefore, nutritional, reproductive and immune status, and losses by other causes may be critical in determining the long- term impact of a disease. Because whales now live in an environment where there is increasing pressure from pollution, climate change and fisheries, they may be less able to recover from disease than in the past. Without knowing the importance of disease in marine mammal populations, it is difficult to predict the likely significance of any disease, or even a mass mortality, in population terms. However, scientists agree that disease outbreaks could contribute to the extinction of threatened populations of marine mammals. The Mediterranean monk seal (Monachus monachus) is one such population that would be at risk should the morbillivirus infection of harbour seals or striped dolphins spread to this endangered species (Osterhaus et al., 1992; Harwood and Hall, 1990). Currently there is insufficient understanding of the role disease plays in regulating whale populations. Because of the possibility of sporadic mass mortalities, current survey intervals may not detect sudden fluctuations in whale abundance, and may allow continued exploitation of a species at a time when it should be protected to ensure its recovery. Understanding and unravelling the influence of disease on whales and dolphins demands longterm study, but this is a commitment that has to be made. Whales in a changing ocean by Bruce McKay Introduction Although whaling has been largely responsible for the decline of whale populations, the impacts of other human activities on them may be--or may become--substantial. Climate change, stratospheric ozone depletion, chemical contamination, reductions in commercial fish stocks, alterations in physical habitat, and other changes related to human activities, will all stress the environment. Their effects are currently poorly understood, and determining the significance of any one of these on whale populations is complicated by the simultaneous occurrence and impact of many. Any one of them can set in motion or promote a cascade of effects, that may ultimately cause or accelerate large-scale alterations in populations, communities and ecosystems. Of concern, too, are the effects of repeated, sequential, and simultaneous human activities. The interactions of combined stresses may exacerbate a problem; for example, an organism's tolerance of one stress is reduced when other stresses are operating at the same time (Esch et al., 1975). Alternatively, different stresses may combine to produce an effect that is greater than the sum of the individual components (Rapport et al., 1985). Because of their reduced numbers, whale populations are already more susceptible to environmental change. The most likely effect of the array of known, and possible, impacts from human activities is to cause increased stress in individuals and, if the conditions are pervasive, decline in the size of populations. Marine mammals and stress Stress: a disturbance that tends to extend any stabilizing process beyond its normal limit, at any level of biological organization, and that renders the individual, population or ecosystem more vulnerable to further environmental change. A disturbance to a system can be (a)foreign to that system or (b) natural to it, yet applied at an excessive level. (Adapted from Bayne, 1975 and Barrett et al., 1976) There are clearly difficulties in generalizing about the effects of various stresses, cumulative or otherwise, on all species and populations. Their impact depends on many factors, including: the genotype and stock; population size; reproductive rates; behaviour and diet; habitat requirements; genetic fitness; immune status; the magnitude of stress from which recovery is possible; and the recovery rate following a disturbance. These, however, are modified by: the type of stress, including its novelty, timing, intensity and frequency; the intensity of different stresses occurring at the same time; the cumulative relationships of simultaneous, repeating or successive stresses; and residual effects of past stresses. An animal will always try, physiologically, to counteract stress but this, in itself, is metabolically costly and may reduce the ability to respond adequately to simultaneous or subsequent stresses. An animal's health will decline if its defences, such as immune or behavioural responses, are unable to withstand or counteract the stress(es). This may lead to a reduction in growth and ability to reproduce, or in survival rates, for example because the animal has become more vulnerable to disease (Koehn and Bayne, 1989). The final effect may be the extinction of a population or the failure of a depleted population to recover. Detecting chronic stress on populations is difficult largely because of delays between the onset of the stress and its measurable result. The effects of environmentally persistent chemicals such as DDT and the PCBs on birds and mammals provide the best example of this. The experience with contaminants has also made it clear that by the time the effects of a physiological stress are measurable at the population level, the changes it has induced are occurring at a rate that is unacceptably rapid (Gilbertson, 1989). For many large animals there is a significant risk of extinction from chance demographic or environmental events if the population numbers fewer than 1,000 individuals (Belovsky, 1987). Even a few incidental deaths in small populations of animals that have long reproductive cycles may have a major impact. Furthermore, small populations may suffer a loss in genetic diversity, leading to reduced birth rates and increased mortality rates (Gilpin and Soule, 1986). Even populations of marine mammals that, although depleted, have recovered from overexploitation could still be particularly vulnerable to changes in their environment since they may lack the genetic diversity needed to adapt to changing conditions (Bonnell and Selander, 1974). Causes of stress to whales and whale populations Although whales are well adapted to their environment they face numerous natural stresses; the most important of these are likely to involve predation, food availability and disease, although little information exists on the extent to which they govern populations. Lambertsen (1990; 1991) suggests that the recovery of fin (Balaenoptera physalus), blue (B. musculus) and humpback whale (Megaptera novaeangliae) populations may be challenged by disease caused by the highly invasive giant nematode Crassicauda boopis. Recent mass mortalities among various pinniped and cetacean species dramatically illustrate how disease may reduce the size of marine mammal populations (see Section 7). The only possibly significant predator of the great whales is the orca (Orcinus orca) and attacks on all species have been recorded (Jefferson et al., 1991). Even low levels of predation could have a serious impact on rare or depleted species such as the bowhead (Balaena mysticetus). Populations of more abundant species may also be affected if they coincide for significant periods of time with a regionally abundant population of predatory orca. For example, 16 of 39 gray whales (Eschrichtius robustus) found stranded on St Lawrence Island Alaska, had been killed by orca (Fay et al. 1978, cited in Rice et al., 1984a). Shark attacks on the large cetaceans, although apparently much rarer, have also been reported (e.g. NMFS, 1991). Bowhead and humpback whales are known to die as a result of being trapped in ice, although exactly how often this happens is unknown (Mitchell and Reeves, 1982; NMFS, 1991). The El Nino current has resulted in significant mortality among numerous pinniped and seabird populations in the east Pacific; one component of this is starvation because of a reduction in available food (Trillmich et al., 1991; Duffy, 1993). Although associated mortality in cetaceans has not been reported, alterations in food distribution or declines in its abundance will exert additional stress on foraging animals. In that regard, the long-term decline in populations of some zooplankton species in the north-east Atlantic and North Sea (Coolbrook, 1985), for example, may be having an impact on baleen whales in the region. There is ample evidence to suggest that the removal of massive numbers of different species of fish by commercial fisheries (see Section 6) can produce a series of effects in exploited environments. These include changes in the relative abundance of species of fish and in the structure of the planktonic community (Sissenwine, 1986; McQueen and Post, 1988). There may also be a reduction in the amount of food available for some species of marine mammals and seabirds, which may, in turn, lead to population declines or altered distribution (Payne et al., 1990; Alverson, 1992; Parsons, 1992). The decline of capelin (Mallotus villosus) off Newfoundland in 1978, which was probably related to fishing (Whitehead, 1987; but see Carscadden, 1983) changed the distribution of humpback whales in that area (Whitehead, 1987) and may have lowered their rate of reproduction (Whitehead, 1982). Such effects were compounded because the whales' redistribution brought them into greater contact with fishing gear, and more animals than usual became entangled in nets and subsequently died (Whitehead (1987). Geraci et al. (1989) reported the fatal poisoning of 14 humpback whales off the coast of Massachusetts, USA, by saxitoxin, a type of neurotoxin produced by some phytoplankton species. This event and others like it (e.g. O'Shea et al., 1991; Gilmartin, 1987) suggest that biotoxins may have played a part in at least some other marine mammal die-offs that have so far been unexplained. The apparent global increase in the frequency, intensity, magnitude, duration and geographic extent of harmful algal 'blooms' (Smayda, 1990), some of which may result in biotoxins contaminating zooplankton, fish and other components of the food web (Smayda, in press; Fritz et al., 1992), further suggests that marine mammal populations may be increasingly susceptible to poisoning in the future. The results may be catastrophic when, for example, contaminated zooplankton occur in the same waters at the same time as endangered North Atlantic right whales (Balaena glacialis); both are found in the Bay of Fundy during summer (White, 1984; Gaskin, 1987). The apparent changes in algal bloom dynamics, as noted above, have been related to various disturbances in coastal environments caused by human activities, including: alterations in grazing structure as a result of fisheries (Graneli et al., 1989); nutrient enrichment (Paerl, 1988); accidental introduction of species (Hallegraeff and Bolch, 1992); and chemical pollutants (Graneli et al., 1989; Sieburth, 1989). Change in upwelling patterns, which bring colder, nutrient-rich water to the surface, is a possible consequence of climate change, and some (e.g. Fraga and Bakun, 1991) have suggested that this may foster increases in blooms of dinoflagellates. Many of these produce biotoxins (Steidinger, 1983). Other causes of stress related to human activities include environmental contaminants, notably the highly toxic and persistent halogenated compounds (see Section 5). Some, such as the DDT-related chemicals and the PCBs, have been found in virtually all the great whale species (Hutchinson and Simmonds, 1991; Woodley et al., 1991; Varanasi et al., in press). In a review of cetaceans and oil, Wursig (1990) felt that, as a group, the baleen whales may be the most vulnerable to the effects of oil spills because of their generally low numbers combined with their particular feeding strategies and tendency to use specific sites. While some suggest that there is little evidence that short-term exposure to oil, through ingestion, contact with baleen or skin, or inhalation, poses a problem for cetaceans (e.g. Geraci and St Aubin, 1982; St Aubin et al., 1984), it may be the indirect impacts of oil spills that are the most significant. Many species of euphausids and various other crustaceans, and their eggs and larvae, as well as the eggs and larvae of fish, are highly susceptible to the toxic effects of oil (Rice et al., 1984b). Large oil spills such as that from the Exxon Valdez could cause major disruptions in the productivity and distribution of whale prey. Marine traffic and coastal activity are thought to play a part in altering distribution or use of habitat for at least some whale populations, including bowhead whales in the Beaufort Sea (Richardson et al., 1987), humpback whales in Hawaiian waters (NMFS, 1991) and Alaska (Beach and Weinrich, 1989), and orca in British Columbian waters (Reeves, 1992). If an animal is repeatedly interrupted while foraging and forced to flee, the associated stress may lead to a general deterioration in health. Ultimately, if disturbed too often, a population may abandon a region or migration route, which will add further stress to individual animals (Richardson et al., 1991). In some cases, animals do appear to become accustomed to human disturbance and activity (e.g. Watkins, 1986) although, as in the case of right whales, this may mean that they are more liable to be in collision with boats (Reeves, 1992). However, the fact that an animal does not abandon a disturbed habitat may not necessarily prove it has learned to tolerate it; rather, this may show simply that it has no alternative. Conclusions The cumulative effects of human activity clearly pose a threat to whales and whale populations. Whaling has reduced virtually all populations, some to the edge of extinction. The past, current and future use of persistent organohalogen chemicals will ensure contamination and consequent biological effects in generations of whales to come. The encroachment of human activities on coastal whale habitat continues to occur through coastal development, shipping, fishing and minerals' extraction. Meanwhile, a growing human population will increase competition with some whale species for dwindling commercial fish stocks. Fisheries, habitat destruction, coastal nutrient enrichment, toxic chemical contamination, and the reduction in marine mammal populations themselves, have altered the structure of some food webs and have disrupted the natural function of numerous marine environments. Global climate change and stratospheric ozone depletion will result in further cascading effects (see Sections 3 and 4). These and other activities present a dizzying array of known and possible effects on whale populations, not only directly, but indirectly through long-term change of their environment. Although there has been an increase in research on the cumulative effects of multiple human activities on the marine environment (CEARC, 1987; Dayton, 1986), little has been done to incorporate this experience into environmental planning, protection and impact assessment programmes. Where management and conservation options are available, it is essential to adopt an approach that takes into account uncertainty and the cumulative effects of past, present and likely future human activity, not only to give the world's great whales the very best chance to survive and recover, but to protect marine environments as a whole. Conclusion When the International Convention for the Regulation of Whaling (ICRW) was signed in 1946, the thoughts of those who drafted and concluded it were on the exploitation and management of whale populations. If any consideration was given to other human activities that might affect whale populations, then it was in isolation from the immediate problem of direct commercial hunting. Experience and increased knowledge has taught us that such a separation of approaches to human actions is not valid. Ecosystems interlink and are interactive in ways that we are only beginning to understand. Furthermore, the cumulative effects of human activities discussed in this document can be expected to be greater than the simple addition of effects from each individual impact. We are now discovering that our impact on the marine environment is in many ways greater than we had previously believed. The indications are that we are entering a period of potential environmental crisis of truly global proportions. This is manifesting itself in ways that, only a matter of years ago, were scarcely even imagined: marine mammal populations suffer mass mortalities; high levels of environmental contaminants are found in wildlife as far from industrialized centres as Antarctica and the Arctic; the thinning of the ozone layer increases year by year; the apparent rate of climate change is outside the evolutionary experience of marine species. The conclusion is clear. The biosphere--and, it would seem, the marine community in particular--is facing a wide and probably unprecedented range of stresses. We have gone beyond the point where particular populations or even species are under threat. There is evidence that whole ecosystems may be at risk. Despite this global change, it has been argued by apologists for whaling that a directed take of a few hundred or even a few thousand whales should be a relatively minor concern. Regardless of the legitimacy of grading environmental problems, which is in itself contentious, such an argument is fundamentally flawed and ignores one basic reality. Direct killing cannot be considered in isolation from the effects of ozone depletion any more than the effects of climate change can be regarded as separate from the effects of marine pollution. However, some nations are still continuing to lobby hard for a resumption of full-scale commercial whaling. To argue for renewed exploitation at a time of increased environmental stress can at best be regarded as misguided, at worst irresponsible. The situation we face is an urgent one. The threats are serious, the impacts are real. The effects of human activities are being felt, directly and indirectly, in the oceans and beyond. All the indications are that things are likely to get worse before they improve. A precautionary approach is required, one that advocates safety and security over speculation and chance. Such an approach to whale conservation demands an indefinite halt to commercial whaling and an end to the other damaging human activities discussed in this document. There is no reason why the International Whaling Commission should not adopt the precautionary principle. Its commercial whaling moratorium was, after all, a tacit acknowledgement of the principle's validity. A failure to accept the precautionary principle now may extinguish any hope of longterm protection for the world's whale populations. Greenpeace International Keizersgracht 176 1016 DW Amsterdam Netherlands ISBN 90-73361-02-8