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Compiled by Leon P. Zann
Great Barrier Reef Marine Park Authority, Townsville Queensland
Ocean Rescue 2000 Program
Department of the Environment, Sport and Territories, Canberra, 1995
ISBN 0 642 17399 0
G.P. Jones & U.L. Kaly
Department of Marine Biology
James Cook University, Townsville Qld 4811
The primary goal of the conservation legislation of most countries is to prevent the extinction of rare and endangered species, whether this is by providing programs to rehabilitate species on the edge of extinction or by protecting the habitats or ecosystems on which they depend (Franklin 1993). With the exclusion of high profile marine organisms (mammals, reptiles, birds) most of the concerns and conservation practices undertaken have been based on terrestrial or freshwater species. The effects of habitat loss and fragmentation, the introduction of exotic organisms, exploitation and pollution are well documented (Soule 1991, Allan & Flecker 1993). There are long lists of recent extinctions and species threatened with extinction that can be directly related to these factors. The principles of establishing and maintaining viable populations of endangered species in terrestrial environments are being formulated and applied with some success (Soule & Wilcox 1980, Soule 1986, 1987, Simberloff 1988).
Recognition of the threats to marine biodiversity is recent (Ray & Grassle 1991, Thorne-Miller & Catena 1991). There is widespread public concern over the world-wide decline of coral reefs (Bunley-Williams & Williams 1990, Veron 1993), changes to temperate kelp bed communities (Estes, Duggins & Rathbun 1989), decline in seagrass beds (Thayer, Wolfe & Williams 1975) and loss of saltmarshes and mangroves (Hatcher, Johannes & Robertson 1989). These all appear to relate directly or indirectly to either coastal development and pollution (Veron 1993) or exploitation by humans (Aronson 1990, Hughes 1993), but the processes are complex and any predictions are made with great uncertainty. Although there are obvious examples of marine mammals and birds that have either become extinct (great auk, Steller's sea cow) or are considered endangered (eg blue, fin, humpback, sei and southern right whales), little is known of this problem for the vast majority of marine animals - including fish and invertebrates (Vermeij 1989). Although there has
been only one recorded post-Pleistocene extinction of a marine invertebrate (the limpet Lottia alveus was last recorded in the eelgrass beds off the eastern coast of North America in the 1920s: Carlton et al. 1991), we have no idea whether there are many extinctions that have gone unnoticed, whether such extinctions are likely to become a major problem, and whether or not the current approaches to dealing with the conservation of fauna and flora will be appropriate for marine organisms.
In focusing attention on potentially threatened marine species, we address a number of important questions in this article. They are:
In Australia, most public attention and research into conservation has focused on the unique bird and mammal fauna (Common & Norton 1992). With 18 or so species becoming extinct over the last century and approximately 120 thought to be threatened (Kennedy 1990) there is considerable cause for alarm. The lack of recorded marine extinctions should not fuel complacency about our present and future impact on marine ecosystems however. There is evidence of overfishing and collection of a wide variety of marine organisms, including offshore pelagic fishes (eg southern bluefin tuna, Thunnus maccoyii: Caton, McLoughlin & Williams 1990), deepwater fishes (eg gemfish, Rexea solandri in New South Wales: Bureau of Rural Resources 1992), shallow water reef fishes (eg three species of coral trout, Plectropomus Goeden 1979, Craik 1979, Rigney 1990) and giant clams, Tridacna (International Union for the Conservation of Nature 1983, Braley 1987). Also potentially important is the over-collecting of black corals (Antipathidae), cowries (Cyraea species), cone shells (Conus species) and tritons (Cymatiidae). Coastal development and industrial pollution (Cambridge & McComb 1984) and destructive fishing methods (Hutchings 1990) are having significant impacts on shallow marine habitats in some areas. The long-term prognosis for individual species - particularly those of great ecological and social significance - needs to be assessed.
There has been no systematic species-level approach to the conservation of Australian marine organisms. Apart from the Whale Protection Act 1980, which protects all whales, dolphins, seals, sea lions and dugong, individual species protection legislation does not generally consider marine organisms as animals, and only a handful of species have been given 'endangered' species status (Bates 1992). The threat of extinction of tridacnid clams was recognised by the International Union for the Conservation of Nature (IUCN) in 1983. All seven species of tridacnid clam, including the giant clam Tridacna gigas, are endangered. Local extinctions of giant clams have been reported throughout Micronesia (Crawford, Lucas & Munro 1987) and the species has been overfished on the Great Barrier Reef (Braley 1987). Three of the largest grouper species in the Great Barrier Reef complex (Epinephelus tukula, potato cod; E. tauvina, estuary grouper; E. lanceolatus, giant grouper) are so few in number that now they are protected completely during their adult phase. The Queensland National Parks and Wildlife Service also lists several other fish species as being endangered.
In a recent review of Australian endangered species, Kennedy (1990) listed only five marine fish, all of which have temperate and subtropical distributions. They are the great white shark (Carcharodon carcharias), the grey nurse shark (Carcharias taurus), the sand tiger shark (Odontaspis ferox), the black cod (Epinephelus damelii) and the southern bluefin tuna. Although Kennedy (1990) considered these species 'vulnerable' rather than 'endangered', he did not define the classifications. Fry and Robinson (1986) listed over 230 mollusc taxa that are potentially vulnerable and in need of monitoring to determine their status. Their list included 44 tropical cowries, 23 cone shells, six tritons and few representatives of numerous other tropical families. Fry and Robinson considered that only the Queensland cowrie, Cypraea queenslandica, is endangered.
As can be seen, the approach to marine species conservation in Australia so far has been inconsistent. The approach highlights deficiencies in both the scientific theory and information being used to manage marine species.
It is important to draw a distinction between 'rare' and 'threatened' species (Edgar et al. 1991). The vast majority of marine organisms are, for whatever reason, naturally rare in that they have exceedingly low local abundance (sensu Rabinowitz, Cairns & Dillon 1986). However, they may have wide geographical distributions and overall numbers of individuals in the species may be high. For example, most of the 1500 or so fish species on the Great Barrier Reef have a broad Indo-Pacific distribution (Randall, Allen & Steene 1990), yet a large fraction of them would be considered rare on the basis of average density. Although naturally rare species may be more vulnerable to extinction than abundant ones (particularly if they have narrow geographic ranges or are highly specialised), not all rare species are (necessarily) threatened (Mace & Lande 1991). Recent studies are showing that - at least for terrestrial species - rarity can be associated with particular life history traits: for example, low reproductive effort, asexual reproduction and poor dispersal (Kunin & Gaston 1993).
In contrast to rarity, endangered or threatened species are those for which 'survival of the species' is unlikely if the causal factors (threats) continue operating (International Union for the Conservation of Nature 1988). Just as rare species are not necessarily at risk, species at risk are not necessarily rare (Mace & Lande 1991). In a first attempt to cope with this variability in status, the IUCN recognised a range of categories including Extinct, Endangered, Vulnerable, Rare and Indeterminate.
Unfortunately, the classification of species is highly subjective and no standards can be applied across taxa or habitats. This lack of standardisation is particularly a problem if the categories of threat are to be used to set priorities for research and management. While there are many clear examples of endangered terrestrial insects, amphibians, reptiles, birds and mammals, the same is not true for fauna occupying the marine environment. Nevertheless, whether that statement applies because endangered marine species do not exist or because we have not detected them, remains to be established. It is possible that at this stage, human impact on marine species has not been as devastating as it has been in the terrestrial environment - perhaps because we are not as efficient species exploiters or habitat destroyers when it comes to the sea.
The majority of marine organisms have life histories and population characteristics that make global extinction unlikely on ecological time scales. The following series of events illustrates this point. A dispersive larval stage in the majority of teleost fish and 70% of marine invertebrates is associated with generally broad geographic ranges and low historic extinction rates (Jablonski 1986). Larval dispersal links isolated subpopulations of adults and can replenish local areas where extinction occurs (Fairweather 1991). The magnitude of recruitment varies in relation to planktonic processes and can be to a large extent independent of a falling adult stock size or processes impacting on the adult population.
On the other hand, marine populations have characteristics that confound any attempt to detect dangerous levels of depletion. Fluctuation in recruitment and breeding population size can obscure any long-term trends and an imminent population crash may be impossible to forecast (the collapse of fisheries for Peruvian anchovy and south-east Australian gemfish are good examples). Patchy distributions can make reliable estimates of density or population size difficult to obtain, even for very common species: only quantum changes in numbers may be detected (Schroeter et al. 1993). Added to these, the methodologies for detecting and determining trends in the abundance of rare species are themselves scarce (but see Gerrodett 1987, Green & Young 1993) and are needed urgently.
Although the global extinction of marine organisms appears to have been rare in the last 200 years, there are other forms of extinction which can have a major impact on the structure of marine assemblages and the functioning of marine ecosystems. Estes, Duggins and Rathbun (1989) recognised the importance in certain species of detecting 'local' extinctions (the disappearance of a species from part of it range). To illustrate their point, they cited observations on the Californian sea otter, Enhydra lutris. Where the sea otter has become locally extinct, benthic communities are dominated by sea urchins and the abundance of kelp is reduced by grazing. This situation is reversed where populations of sea otters have re-established.
Such shifts in community structure can occur without local extinction. 'Ecological' extinction arises when a species is reduced to such low abundance that, although still present, it no longer plays the ecological role it used to (Estes, Duggins & Rathburn 1989). For example, in North America the overfishing of lobsters - once important predators in shallow kelp forest systems - has led to alternative benthic community structures even though lobsters are still present though in low numbers. Overfishing of triggerfish on Kenyan reefs also has not led to local extinction. However, because these fish had a natural predatory role, the reduction in their number had a major impact on coral reef communities (McClanahan & Muthiga 1988). It is our view that, in the marine environment, we must be more concerned about the local or ecological extinction of species playing 'keystone' ecological roles (sensu Mills, Soule & Doak 1993) than about the unlikely global extinction of the vast number of essentially 'redundant' species (sensu Walker 1992).
It is clear that much of the current theory developed in conservation biology cannot be uncritically applied to marine species and habitats. There are four areas in which we see the theory as deficient.
The minimum viable effective population size for terrestrial animals is usually based on the '50-500' rule, ie an effective population size (Ne) of greater than 50 being the critical limit to avoid inbreeding problems, and an Ne of greater than 500 being the critical limit to avoiding loss of genetic variation through genetic drift. Both processes can lead rapidly to extinction in 'closed' populations. An Ne of 500 may equate with an actual population size of up to 2500 individuals depending on the species, because it depends on the number of mature, breeding individuals in the population (Shaffer 1981, Nunney & Campbell 1993). The 50-500 rule has a chequered history and is not universally accepted even for terrestrial organisms (Simberloff 1988). One thing for certain is that such genetic factors are not problems for small 'open' populations of marine organisms with high levels of gene flow among subpopulations (Slatkin 1981).
Despite the chequered history, minimum viable population estimates lie at the heart of the most recently published threatened species categories (Mace & Lande 1991). These are:
CRITICAL: effective population size (Ne) <50;
(or actual population size (N) <250)
ENDANGERED: Ne <500; N <2500
VULNERABLE: Ne <2000; N <10 000
We know of no marine fish or invertebrate in this range. Nor are there any for which the data could be obtained, except perhaps for endemic species with exceedingly restricted geographic ranges. Simple thresholds in population size are unlikely to apply to the majority of marine organisms which characteristically exhibit extremely high juvenile mortality, fluctuating recruitment and a poor relationship between recruitment and adult numbers. Marine mammals appear to be classified arbitrarily as 'endangered' when numbers are reduced to four digit estimates (Kennedy 1990).
The more general approach to the factors affecting the persistence of small populations is called Population Viability Analysis (Gilpin & Soule 1986). Preliminary models are available to assess population persistence under different extinction pressures which include demographic variation, stochastic environmental factors, metapopulation structure and increasing fragmentation (see Mace & Lande 1991). While this approach is promising in the long-term, there are few models that realistically can be applied to open marine metapopulations. Increasing habitat fragmentation in terrestrial environments is seen as a major threat to population persistence, but a high level of fragmentation is the natural condition for marine species and may not represent an extinction threat at all. Large numbers of subpopulations in combination with high levels of dispersal theoretically make the extinction of the metapopulation unlikely (Hanski 1989). On the other hand, if large-scale catastrophic processes lead to correlated population trends among subpopulations, then extinctions are more likely and may occur even if metapopulations are extremely large.
No doubt increasing habitat fragmentation will occur as seagrass beds and coral reefs decline. However, until we have realistic measurements of the numbers of subpopulations, the degree to which trends in different subpopulations are correlated, the level of larval and adult exchange and knowledge of any source-sink relationships, it will not be feasible to base management strategies using Population Viability Analysis per se.
Terrestrial ecologists have also been pre-occupied with the community and ecosystem responses to habitat fragmentation (see Saunders, Hobbs & Margules 1991). This pre-occupation has also led to discussions on reserve size, design and shape which culminated in the so-called SLOSS ('single large or several small') debate (see Simberloff 1988 for review). Reserved habitat patches are treated as 'islands' and island biogeography theory is applied to determine the optimal size and arrangement of reserves for maximising colonisation and minimising extinction. The basic assumption that reserves of a particular configuration have an equilibrium number of species has not been supported by long-term studies of community composition on natural reef isolates (eg Sale 1991). Rates for colonisation and extinction appear to be largely independent of variation in the number of resident species, patch size and shape, and proximity to the source because of the unpredictable nature of larval supply and adult mortality. While most marine reserves have been set up on an ad hoc basis, even very small, isolated reserves have had measurable effects on the abundance of highly mobile exploited species (Jones, Cole & Battershill 1993).
The pre-occupation of terrestrial conservation biologists with 'corridors' to connect habitat fragments or reserves (see review by Hobbs 1992) has no strictly marine analogue.
At this stage in our knowledge, any attempt to grade marine species according to the degree of threat (critical, endangered and vulnerable) is likely to prove futile. An alternative yet still subjective approach is to recognise the characteristics of species that are at least 'potentially threatened' by extinction, and develop management strategies as a precautionary measure.
Many species exhibit combinations of these characteristics and those species may be most at risk. We stress however, that species exhibiting these characteristics are not necessarily endangered. For the majority of species we simply do not have information on such matters as distribution, recruitment and population variability, susceptibility to stress and degree of overfishing sufficient to make judgements about the likelihood of local, ecological or global extinction. Arbitrary thresholds could be used to create a list of 'potentially vulnerable' species (eg geographic range less than 'x' km, organisms growing to 'x' m, or living to more than 'x' years). However, such arbitrary thresholds should not be used to establish a hierarchy of threatened species categories nor establish research priorities.
One of the oldest debates in ecology - whether research effort should favour the species by species approach or focus on communities/ ecosystems - is being revisited by conservation biologists and managers (Franklin 1993). It is perhaps a lesson that the debate was never resolved (we still have population and ecosystem ecologists!), yet it is the diversity of approaches that will in the long term maximise our chance of protecting species within natural systems. The sheer number of marine species (many undescribed), our inability to recognise the endangered among them and the potentially high level of ecological redundancy, favour the 'ecosystem' perspective (Walker 1992). The 'species' approach has never served small, low profile but ecologically important trophic groups such as detritivores and decomposers.
However, Noss (1990) identified five types of species deserving special conservation status because of their relevance to the ecological communities of which they are members. They are:
The best way to ensure some degree of protection of widely distributed but rare marine organisms is to establish a network of marine reserves encompassing a proportion of all marine habitats and biogeographical regions (Pressley et al. 1993). There is no doubt that marine reserves have had considerable beneficial effects on excessively exploited species in tropical and temperate regions (Roberts & Polunin 1991) and can shift communities back toward a more 'natural' community structure (Jones, Cole & Battershill 1993). However, their effectiveness as a management tool for rare or endangered marine species has yet to be evaluated and may be a problem for highly mobile organisms. Protecting communities and ecosystems does not guarantee species survival, and a 'safety net' of species-orientated approaches will also be necessary (Walker 1992).
As an additional precautionary measure, marine protected areas should also be chosen to contain populations or breeding areas of 'potentially vulnerable' species, particularly those endemics with small geographic ranges, localised breeding sites and unusual aggregations of long-lived, large species. Rare habitats supporting endemic species will be high priority. An example is the Shark Bay region in Western Australia. This 2.1 million ha area was nominated for world heritage listing in 1990, and includes the famous stromatolites at Harmelin Pool, the extensive Wooramel seagrass banks and the habitats of several rare species (Rigney 1990). Another area, the Solitary Islands Marine Reserve in northern New South Wales, was gazetted in 1990. This area is a unique mixture of coral and kelp-dominated reefs and contains unusual aggregations of subtropical endemic species with restricted ranges. The Solitary Islands Marine Reserve is similar only to the Houtman Abrolhos in the west, where another suite of subtropical endemics are found (Veron 1993).
Other conservation tools that could be applied to protect particular species include complete closure to collecting or fishing and strict pollution or development controls in areas where sensitive marine organisms are concentrated (eg coral reefs). The possibility of enhancing excessively exploited marine species also needs to be explored. The technology for captive breeding and outplanting of juveniles has been developed for some invertebrates (eg giant clams: Crawford, Lucas & Munro 1987). For those species which cannot be reared through the larval stage, the opportunities for recruitment enhancement should be evaluated. Many reef fish species, for example, have very specific habitat requirements at settlement. Once these are known, artificial substrata may be developed and deployed on a scale that will enhance recruitment. The 'last resort' approach for terrestrial species is to transfer remaining individuals to island refuges containing suitable habitat that is free of exotic competitors or predators. This practice may not work for marine species with dispersive larvae, although the evidence is there: introdu