PART 4
DIVERSITY AT RISK: THE GLOBAL PERSPECTIVE



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BioDiversity PART 4 DIVERSITY AT RISK: THE GLOBAL PERSPECTIVE

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BioDiversity Aerial view of a coral reef in the Capricorn Group at the southern end of the Great Barrier Reef, Australia. Photo courtesy of G.Carleton Ray.

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BioDiversity CHAPTER 17 LESSONS FROM MEDITERRANEAN-CLIMATE REGIONS HAROLD A.MOONEY Professor of Biological Sciences, Department of Biological Sciences, Stanford University, Stanford, California Discussions on the loss of biological diversity are correctly focused on tropical regions because of the massive, rather recent alterations in the structure of these extensive biotic communities. The consequences of these alterations are many. There are of course no landscapes on Earth that have not been modified to some extent by the human species. Many of these landscapes have been totally altered from their prehuman configuration and functioning, and others appear less affected; however, none are protected from the types of global changes that are resulting from human-induced alterations of the Earth’s atmosphere. This section focuses on the nature and some of the consequences of alterations of nontropical biogeographic regions. The discussions are selective, concentrating on selected processes and organisms within a few systems. In Chapter 18, Franklin deals with temperate and boreal forests, which occupy 16% of Earth’s land surface—an area equivalent to that covered by tropical forests (Waring and Schlesinger, 1985)—and which have provided to a large degree the timber and in part the fuel to support the growing human population. In the next chapter, Risser discusses the impact of humans on biological diversity in grasslands, the biome that has largely provided, either directly or indirectly, the food for the world’s human population. Finally, in Chapter 20, Vitousek details the kinds of biotic changes that have resulted from human settlement on Hawaii and on oceanic islands in general—systems that have proven to be particularly susceptible to losses and additions of species. Each chapter emphasizes somewhat different points. Franklin focuses on the consequences of structural diversity loss in forest ecosystems, drawing examples

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BioDiversity from the magnificent coniferous forests of the Pacific Northwest. Risser notes the low loss of species in the high-impact North American grasslands and the potentially high loss of ecotypes. He also discusses the variable consequences of different land-use patterns on species diversity. Vitousek relates the apparent devastating effects of species invaders on the endemics of the Hawaiian Islands, noting that although species diversity has actually increased, ecosystem types have been lost. As an introduction to these chapters on threats to diversity in nontropical systems, I first compare the community diversity of tropical systems with those of temperate regions, providing plants as examples. I then focus more specifically on Mediterranean-climate (cool wet winter, dry summer climate) regions to balance the presentations on forests and grasslands. Mediterranean-climate regions, of which there are five in the world, are of special interest for two reasons: they rival tropical regions for their biological richness, and because they have had very different histories of human settlement, they serve as interesting comparison areas in studies to determine the human impact on biotic diversity. COMMUNITY DIVERSITY IN TROPICAL AND TEMPERATE REGIONS The fact that tropical regions are biologically richer than temperate regions has been stated repeatedly: for example, Raven (1976) has noted that 65% of the world’s 250,000 flowering plants are found there. Until recently, the tropics, particularly the lowland wet tropics, have remained one of the last areas that has not been subjected to extensive human exploitation. In temperate regions of the world, many of the natural ecosystems have been massively altered by human settlement and activities. By looking at some of these disturbed regions, we can assess the consequences of human activities on biological diversity and, to some extent, learn what we should expect in the tropics in the future. If we were to pick only one biome type to serve as a model of comparison, it should be the Mediterranean-climate regions of the world. These regions are remarkably diverse by any measure. Gentry (1979) reported that the number of plant species he encountered in 0.1-hectare plots increased as he moved from dry tropical to wet tropical forests (Table 17–1). In his most diverse sites in Panama he encountered more than 150 species of woody plants thicker than 1 inch in diameter at breast height. In contrast, only 21 woody species were found in a temperate forest in Missouri. Data on total species counts in tropical forests have not been available. However, Whitmore (1986) reported the results of a survey in which 236 species of vascular plants were counted in a 0.01-hectare plot in Costa Rica; he estimated that “one man decade would be required to enumerate one hectare” (Whitmore, 1986). Counts of all the vascular plants in sample plots in other climatic regions are available for comparison. In the Mediterranean-climate region of Israel, Naveh and Whittaker (1979) found sites that included as many total species as woody species found by Gentry in Panama. The richest sites were those with some degree of current disturbance. Mediterranean-climate sites of the same size in other parts of the world also have relatively high species counts in comparison to counts of temperate-zone vegetation (Whittaker, 1977).

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BioDiversity The bases for the high diversities among the different Mediterranean-type vegetations differ. In Israel, the diversity is accounted for mostly by herbaceous species, principally annuals, and is the result of human-driven “relatively rapid evolution under stress by drought, fire, grazing and cutting” (Naveh and Whittaker, 1979). In contrast, the high diversity of the South African fynbos (Mediterranean-climate scrubland) vegetation consists of woody species, of which there are few annuals. This type of vegetation has not been subject to a long history of human disturbance. The data thus indicate that tropical systems are probably among the world’s richest in terms of local, or alpha, diversity, but that the vegetation of Mediter- TABLE 17–1 Mean Numbers of Species per 0.1-Hectare Sample Area (Non-Mediterranean Sites Include Only Data for Woody Plants over 1 Inch in Diameter at Breast Height) Sample Area Mean No. of Species Dry Tropical Foresta   Costa Rica upland, Guanacaste 41 Costa Rica riparian, Guanacaste 64 Venezuelan Llanos, Calabozo 41 Venezuelan coastal, Boca de Uchire 67 Moist Tropical Foresta   Panama Canal Zone, Curundu 88 Brazil, Manaus 91 Panama Canal Zone, Madden Forest 125 Wet Tropical Foresta   Panama Canal Zone, pipeline road 151 Ecuador, Rio Palenque 118 Costa Rica, near La Selvab 236 Temperate Zonea   Missouri, Babler State Park 21 Temperate Zonec   Australia, forests and woodlands 48 Tennessee, Great Smoky Mountains 25 Oregon, Siskiyou Mountains 26 Arizona, Santa Catalina Mountains 21 Colorado, Rocky Mountain National Park 32 Mediterranean Zoned   Israel, grazed woodlands 136 Israel, open shrubland 139 Israel, closed shrubland 35 California, grazed woodlands 64 California, closed shrubland 24 Chile, open shrubland 108 Australia, heath 65 South Africa, fynbos 75 aData from Gentry, 1979. bData from Whitmore, 1986, for a 0.01-hectare plot. cData from Whittaker, 1977. dData from Naveh and Whittaker, 1979.

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BioDiversity ranean-climate regions is also quite rich. In Mediterranean-climate regions the basis for the localized diversity can differ with the pattern of disturbance. In some systems with a long history of association with human activities, diversity has actually increased (Naveh and Whittaker, 1979). Data on diversity at a given site indicate its structural dynamics as related to both evolutionary history and pattern of disturbance. We are just now beginning to appreciate the role of both natural disturbances and the impacts of humans in controlling community structure, including its diversity (Bazzaz, 1983). Such knowledge is essential for understanding and hence managing a given level of diversity. MEDITERRANEAN-CLIMATE FLORISTIC DIVERSITY Data on local diversity are an indication of disturbance pattern and evolutionary history leading to niche diversification. Another view of the biotic richness of an area is the degree of endemism of the biota. Data on species numbers and degree of endemism for Mediterranean-climate regions form the basis for identifying them as critical sites for conservation. An indication of the diversity and uniqueness of Mediterranean-climate plant life is given below for South Africa, California, and the Mediterranean basin—areas that share unusually high biotic diversities but have dissimilar histories of human impact. For example, South Africa has large tracts of land dominated by the original species-rich shrubland, and the Mediterranean basin contains predominantly herb or shrub degradation forms of the original vegetation. The diversity of South Africa is threatened by development and the invasion of alien species; the Mediterranean basin diversity, by changes in land-use patterns. South Africa The Mediterranean-climate region (fynbos biome) of South Africa covers 75,000 square kilometers. This area includes 8,550 vascular plants (Macdonald and Jarman, 1984), three-quarters of which are endemic (Jarman, 1986). According to estimates by Hall (1978), the flora indigenous to the South African Cape, which is found in an area of 46,000 square kilometers, contains at least 6,000 higher plant species—a species richness three times that found in tropical regions of similar areas. This subregion has been considered one of the world’s six distinctive floristic regions. In the fynbos biome, 1,585 plant species are considered rare and threatened (Macdonald and Jarman, 1984), and 39 have recently become extinct (Jarman, 1986). Although the fynbos region occupies less than 1% of southern Africa, it contains 65% of the threatened plant species (Hall, 1979). Much of the vegetation in this region has been destroyed by human activities, but not to the extent it has occurred in other Mediterranean-climate areas. In the lowland regions, only about 30% of the original vegetation remains, whereas in the mountains, approximately 80% of the vegetation remains intact. Overall, about 67% of the natural fynbos vegetation remains (Jarman, 1986). One threat to the native flora is the presence of alien, generally woody species, which have invaded

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BioDiversity about one-fourth of the native vegetation (Jarman, 1986). Of 70 critically threatened or recently extinct taxa, 23% are threatened by invading acacias, 8% by pines, and 2% by hakeas (Hall, 1979). In summary, the South African Mediterranean-climate vegetation is as rich as any found on Earth. This richness is being threatened by human development, as everywhere, but also by a rather remarkable invasion of woody plants that are altering the basic functioning of these systems (Macdonald and Jarman, 1984). California There is rather complete information describing the biotic richness of the State of California, most of which falls within a Mediterranean-type climate. Although not as rich as South Africa in plant species, it certainly is one of the world’s most biotically diverse areas. In an area of 411,000 square kilometers, there are more than 5,046 native vascular plant species, 30% of which are endemic. (In comparison, there are about 20,000 vascular plant species in the continental United States.) About one-tenth of the flora in these regions of California has recently become extinct or endangered. This represents 25% of all the extinct and endangered species of the United States as a whole (Raven and Axelrod, 1978). California has suffered great losses of natural communities through human development of agriculture, industry, and housing, especially in coastal and valley regions. Entire ecosystems have evidently been irrevocably lost. One of the most spectacular examples of this is the native perennial grassland of the Central Valley and north coastal regions, which has been replaced by an annual grassland dominated by species mostly inadvertently introduced from the Mediterranean basin (Burcham, 1957). Raven and Axelrod (1978) estimate that more than 10% of the flora in these regions is now composed of naturalized aliens. Thus California, like other Mediterranean-climate regions, has an unusually diverse biota that is being threatened by human activities. But to a greater extent than in other regions, substantial areas of the state have been set aside as parks and preserves. The Mediterranean Basin The entire Mediterranean basin encompasses more than 2 million square kilometers and may include as many as 25,000 higher plant species, about half of which are endemic (Quezel, 1985). Of 2,879 species endemic to individual Mediterranean countries (excluding Syria, Lebanon, Turkey, and the Atlantic islands), 1,529 are rare (1,262) or threatened, and 300 are not categorized. If the Atlantic islands (Azores, Madeira, and the Canaries) are included, these figures increase to 3,583 endemics and 1,968 rare or threatened plant species (Leon et al., 1985) In contrast to California and South Africa, where large areas of climax vegetation remain, much of the Mediterranean basin has been completely transformed from its native state. Naveh and Dan (1973, p. 387) reported that the region as a whole “is composed of innumerable variants of different degradation and regeneration stages.” Since the impact of humans in this region has been so extensive for a long time, it is believed that the Mediterranean endemic has evolved under conditions of frequent disturbance or in depauperate microsites, such as rock outcrops (Gomez-Campo, 1985). Greuter (1979, p. 90) observed that “the rare threatened taxa are

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BioDiversity seldom members of the characteristic vegetation units as defined by the plant sociologists: they are marginal creatures living on the borderline of biota….” This general viewpoint has led to the following conclusion of Ruiz de la Torre (1985, p. 197): “Unlike the tropical rain forest, where most of the indigenous species can be conserved with climax formations under conditions of maximum stability, the Mediterranean region has been severely influenced by man and various other factors and is still very rich in species. Very few of these species are known to be part of Mediterranean climax vegetation. Most of them correspond to successional stages affected by either natural or artificial exploitation, and they should be conserved under the prevailing conditions of relative instability.” INCREASING BIOTIC DIVERSITY—THE INVADERS As indicated above, plant diversity in Mediterranean-climate regions is among the world’s richest in terms of numbers of species, but there have been losses of species and continuing threats of extinction to many others. However, there have also been additions of new species to these and other regions of the world. As shown in Table 17–2, the floras of certain islands, ranging from subarctic to tropical, have been enriched half again by species from other biographic regions. In mainland Mediterranean-climate regions such as California, and even to a greater extent in South Australia, there are also substantial numbers of invading species that have become naturalized, many maintaining large and dominating populations. In these regions, as elsewhere, these invading species are not distributed uniformly in the landscape but are generally associated with ecosystems that have experienced human impact. Organisms other than plants are also being enriched by the addition of species in these climates. In California, for example, 49 species have been added to the 132 indigenous inland fishes (Moyle, 1976). Thus in some cases, human disturbance can actually enrich biotic diversity. However, species counts in a given area give us little understanding of ecosystem functioning and how the invasions affect it. Some invaders may become the dominant species in the host-region ecosystem. Examples of this include a species of oat (Avena fatua) in the grasslands of California (Burcham, 1957) and brome grass (Bromus tectorum) in the intermountain West (Mack, 1986). Many of the invaders are pest species of one sort or another and may cause economic havoc. These species of course receive considerable attention, and their biology and community role is generally well known. However, we generally know little about the effects of most invaders on the ecosystem or, for that matter, the effects of most species on natural communities. Are these invaders enriching biotic diversity? They are when considered in absolute numbers of species. In many cases, however, they are impoverishing the biota by leading to species exclusions (Race, 1982) or even to extinctions. The invaders are generally symptoms of an abused landscape, one that has been disturbed and has generally lost some of its original productive capacity. The successful introduction of exotic mammals has often resulted in greatly perturbed ecosystem function and losses of indigenous species. In general, new community types are being added to the original ones that in turn are being reduced in extent. The

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BioDiversity TABLE 17–2 The Plant Invaders Region Area (thousands of square kilometers) Indigenous Species Endemic Species Naturalized Species Type of Planta Reference Non-Mediterranean-Climate Islands   Seychelles 0.3 153 69 165 F Proctor, 1984. Status of 130 species unknown Faeroe Islands 1.4 370 0 30 F Hansen and Johansen, 1982 New Zealand 268.0 1,996 1,618 500 V Godley, 1975 Hawaiian Islands 16.7 ca. 1,250 ca. 1,180 600 F Wagner et al., 1985; Smith, 1985 Mediterranean-Climate Regions   South Australia 984.4 2,380 NRb 654 F Specht, 1972 California 411.0 5,046 1,517 674 V Raven and Axelrod, 1978 Canary Islands 7.3 1,050 550 700 V Kunkel, 1976 aV, vascular plants; F, flowering plants. bNot reported.

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BioDiversity landscapes are becoming more complex. Yet, when viewed on a more global scale, the biota is becoming less interesting because of homogenization. For example, geographically separate and distinctive biological regions are often invaded by the very same weedy species. As a result, regions such as parts of California and Chile, which once had only a few plant species in common, now share hundreds. The maintenance of a diverse landscape, rich in community types and species, requires knowledge of the dynamics of ecosystems as well as the ecology of individual species. Since this information is generally lacking, attempts to conserve individual species or populations are still filled with surprises, even in preserves. REFERENCES Bazzaz, F.A. 1983. Characteristics of populations in relation to disturbance in natural and man-modified ecosystems. Pp. 259–275 in H.A.Mooney and M.Godron, eds. Disturbance and Ecosystems. Springer-Verlag, New York. Burcham, L.T. 1957. California Rangeland. California Division of Forestry, Sacramento, Calif. 261 pp. Gentry, A. 1979. Extinction and conservation of plant species in Tropical America: A phytogeographical perspective. Pp. 110–126 in I.Hedberg, ed. Systematic Botany, Plant Utilization and Biosphere Conservation. Almqvist and Wiksell International, Stockholm. Godley, E.J. 1975. Flora and vegetation. Pp. 177–229 in G.Kuschel, ed. Biogeography and Ecology in New Zealand. Monographiae Biologicae, Vol. 27. W.Junk, The Hague, the Netherlands. Gomez-Campo, C. 1985. The conservation of Mediterranean plants: Principles and problems. Pp. 3–8 in C.Gomez-Campo, ed. Plant Conservation in the Mediterranean Area. W.Junk, Dordrecht, the Netherlands. Greuter, W. 1979. Mediterranean conservation as viewed by a plant taxonomist. Webbia 34:87–99. Hall, A.V. 1978. Endangered species in a rising tide of human population growth. Trans. R. Soc. S. Afr. 43:37–49. Hall, A.V. 1979. Invasive weeds. Pp. 133–147 in Fynbos Ecology: A Preliminary Synthesis. South African National Scientific Programmes Report No. 40. Cooperative Scientific Programmes, Council for Scientific and Industrial Research, Pretoria, South Africa. 166 pp. Hansen, K., and J.Johansen. 1982. Flora and vegetation of the Faeroe Islands. Pp. 35–52 in G.F. Rutherford, ed. The Physical Environment of the Faeroe Islands. Monographiae Biologicae, Vol. 46. W.Junk, The Hague, the Netherlands. Jarman, M.L. 1986. Conservation Priorities in the Lowland Regions of the Fynbos Biome. South African National Scientific Programmes Report No. 87. Cooperative Scientific Programmes, Council for Scientific and Industrial Research, Pretoria, South Africa. 55 pp. Kunkel, G. 1976. Notes on the introduced elements in the Canary Islands flora. Pp. 249–266 in G. Kunkel, ed. Biogeography and Ecology in the Canary Islands. Monographiae Biologicae, Vol. 30. W. Junk, The Hague, the Netherlands. Leon, C., G.Lucas, and H.Synge. 1985. The value of information in saving threatened Mediterranean plants. Pp. 177–196 in C.Gomez-Campo, ed. Plant Conservation in the Mediterranean Area. W.Junk, Dordrecht, the Netherlands. Macdonald, I.A.W., and M.L.Jarman. 1984. Invasive Alien Organisms in the Terrestrial Ecosystems of the Fynbos Biome, South Africa. South African National Scientific Programmes Report No. 85. CSIR Foundation for Research Development, Council for Scientific and Industrial Research, Pretoria, South Africa. 72 pp. Mack, R. 1986. Alien plant invasion into the Intermountain West: A case history. Pp. 191–213 in H.A.Mooney and J.Drake, eds. The Ecology of Biological Invasions into North America and Hawaii. Springer-Verlag, New York. Moyle, P.B. 1976. Inland Fishes of California. University of California Press, Berkeley. 405 pp. Naveh, Z., and J.Dan. 1973. The human degradation of Mediterranean landscapes in Israel. Pp. 373–390 in F.de Castri and H.A.Mooney, eds. Mediterranean Type Ecosystems: Origin and Structure. Springer-Verlag, Berlin.

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BioDiversity Naveh, Z., and R.H.Whittaker. 1979. Structural and floristic diversity of shrublands and woodlands in northern Israel and other Mediterranean areas. Vegetatio 41:171–190. Procter, J. 1984. Floristics of the granitic islands of the Seychelles. Pp. 209–220 in D.R.Stoddart, ed. Biogeography and Ecology of the Seychelles Islands. Monographiae Biologicae, Vol. 55. W.Junk, The Hague, the Netherlands. Quezel, P. 1985. Definition of the Mediterranean region and the origin of its flora. Pp. 9–24 in C. Gomez-Gampo, ed. Plant Conservation in the Mediterranean Area. W.Junk, Dordrecht, the Netherlands. Race, M.S. 1982. Competitive displacement and predation between introduced and native mud snails. Oecologia 54:337–347. Raven, P.H. 1976. Ethics and attitudes. Pp. 155–181 in J.B.Simmons, R.I.Beyer, P.E.Brandham, G.Lucas, and V.T.H.Parry, eds. Conservation of Threatened Plants. Plenum, New York. Raven, P.H., and D.I.Axelrod. 1978. Origin and Relationships of the California Flora. Univ. Calif. Pub. Bot. 72:1–134. Ruiz de la Torre, J. 1985. Conservation of plant species within their native ecosystems. Pp. 197–218 in C.Gomez-Gampo, ed. Plant Conservation in the Mediterranean Area. W.Junk, Dordrecht, the Netherlands. Smith, C.W. 1985. Impact of alien plants on Hawai’i’s native biota. Pp. 180–250 in C.P.Stone and J.M.Scott, eds. Hawai’i’s Terrestrial Ecosystems: Preservation and Management. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu, Hawaii. Specht, R.L. 1972. The Vegetation of South Australia. A.B.James, Adelaide, Australia. 328 pp. Wagner, W.L., D.R.Herbst, and R.S.N.Yee. 1985. Status of the native flowering plants of the Hawaiian Islands. Pp. 23–74 in C.P.Stone and J.M.Scott, eds. Hawai’i’s Terrestrial Ecosystems: Preservation and Management. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu, Hawaii. Waring, R., and W.H.Schlesinger. 1985. Forest Ecosystems: Concepts and Management. Academic Press, New York. 340 pp. Whitmore, T.C. 1986. Total species count on a small area of lowland tropical rain forest in Costa Rica. Bull. Br. Ecol. Soc. 17:147–149. Whittaker, R.H. 1977. Evolution of species diversity in land communities. Evol. Biol. 10:1–67.

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BioDiversity Risser, P.G. 1986. Preservation status of true prairie grasslands and ecological concepts relevant to management of prairie preserves. Pp. 339–344 in D.L.Kulhavy and R.N.Connor, eds. Wilderness and Natural Areas in the Eastern United States: A Management Challenge. Papers presented at a symposium: Wilderness and Natural Areas in the East, held in Nacogdoches, Texas, on May 13–15, 1985. Center for Applied Studies, Stephen F.Austin State University, Nacogdoches, Tex. 416 pp. Risser, P.G., E.C.Birney, H.D.Blocker, S.W.May, W.J.Parton, and J.A.Wiens. 1981. The True Prairie Ecosystem. Hutchinson Ross, Stroudsburg, Penn. Steiger, T.L. 1930. Structure of prairie vegetation. Ecology 11:170–217. Sampson, F.B. 1980. Island biogeography and the conservation of prairie birds. Pp. 293–299 in C. L.Kurera, ed. Seventh North American Prairie Conference, Proceedings. Southwest Missouri State University, Springfield, Mo. Transeau, E.N. 1935. The prairie peninsula. Ecology 16:423–437. Weaver, J.E. 1954. North American Prairie. Johnson Publishing Company, Lincoln, Nebr. 348 pp. Whitcomb, R.F., J.Kramer, M.D.Coan, and A.L.Hicks. In press. Ecology and evolution of leafhopper-grass host relationships in North American prairies, savanna and ecotonal biomes. Curr. Top. Vector Res. 3.

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BioDiversity CHAPTER 20 DIVERSITY AND BIOLOGICAL INVASIONS OF OCEANIC ISLANDS PETER M.VITOUSEK Associate Professor, Department of Biological Sciences, Stanford University, Stanford, California To date, human-caused species extinctions are more an island-based than a continental phenomenon. Of the 94 species of birds known to have become extinct worldwide since contact with Europeans, only 9 were continental (Gorman, 1979). Currently, more endemic Hawaiian bird species are officially listed as endangered or threatened than are listed for the entire continental United States. Where information is available on other groups of animals, it indicates that human-caused extinctions are invariably more frequent on islands. Heywood (1979) summarized the causes of extinction on islands as deforestation and fire, the introduction of grazing mammals, cultivation, and the introduction of weedy plants. All these factors can be important on continents as well, but species introductions (deliberate or accidental) are disproportionately important on islands (Elton, 1958). Isolated islands and archipelagos often lack major elements of the biota of continents, and their native species often lack defenses against grazing or predations. Biological invasions are not the only factor leading to elevated extinction rates for island species. Extinction rates are also higher on islands because island species generally have small populations, restricted genetic diversity, and narrow ranges prior to human colonization, and because human alterations of land through use destroy an already-limited critical habitat. The plant and animal hitchhikers and fellow travelers who accompany humans to isolated islands interact with these other causes of extinction, however, and biological invaders endanger native species in reserves and other protected lands.

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BioDiversity The fact that biological invasions decrease diversity on islands is paradoxical, because, as pointed out by Lugo in Chapter 6, the introduction of alien species generally increases the total number of species on an island, often spectacularly. However, most of the introduced species are cosmopolitans that are in no danger of global extinction, whereas most species on isolated islands are endemic. Biological invasions can therefore cause a net loss of species worldwide and a homogenization of the biota of Earth (Mooney and Drake, 1986). SCOPE OF THE PROBLEM The disproportionate effects of human colonization and attendant biological invasions on island ecosystems are well known (Carlquist, 1974; Darwin, 1859; Elton, 1958; Wallace, 1880); they can be demonstrated even on large islands such as Madagascar and Australia (Carlquist, 1974). The most severe consequences are experienced on old, isolated, mountainous, tropical, or subtropical islands or archipelagos. Islands located near continents receive organisms from those continents and rarely develop unique species. Truly oceanic islands have rates of evolution and speciation greater than those of immigration; hence, their biota contains many endemic species. Low islands (such as atolls) lack the range of environments that permits evolutionary radiation, while islands at high latitudes are subjected to strong climatic fluctuations (Bramwell, 1979), which prevent radiation. Together these factors suggest that the Hawaiian Islands, the most isolated archipelago in the world, should have a large number of exotic species and a large potential for loss of endemic species as a consequence of biological invasions. The very large number of endemic species on these islands is well documented (Carlquist, 1974); the importance of biological invasions can also be demonstrated. For example, a survey of exotic plants on National Park Service lands (Loope, in press a) shows that island parks have a much larger proportion of alien species in their flora than do continental parks (Table 20–1). Moreover, in most continental parks alien species are largely confined to roadsides and areas occupied by humans before the park was established. In contrast, Channel Islands National Park in California, Everglades National Park (an island of tropical vegetation at the tip of the Florida peninsula), and the Hawaiian parks contain alien species that establish themselves in otherwise undisturbed native ecosystems and change the nature of the sites they occupy (Ewel, 1986; Stone and Scott, 1985; Stone et al., in press a). The problems in the Hawaiian parks reflect in part the overall abundance of exotic species in Hawaii. As many as 1,765 native species of vascular plants (probably fewer as taxonomic revisions take hold) existed in the islands when the Polynesians arrived, and 94 to 98% of them were endemic (Kepler and Scott, 1985). Polynesians brought additional species, perhaps 30 of them (Nagata, 1985), when they colonized Hawaii and journeyed among the Pacific islands. The advent of more rapid transportation from distant areas and especially the occupation of Hawaii by people from diverse western and eastern cultures, each with its distinctive food, medicinal, and ornamental plants, greatly increased the number of species present. More than 4,600 species of introduced vascular plants are now known to

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BioDiversity TABLE 20–1 Proportion of Alien Plants in the Vascular Flora of Selected U.S. National Parksa National Park Alien Species (% of total) Sequoia-Kings Canyon 6–9 Rocky Mountain 7–8 Yellowstone 11–12 Mount Ranier 12–14 Acadia 21–27 Great Smoky Mountain 17–21 Shenandoah 19–24 Channel Islands 16–19 Everglades 15–20 Haleakala 47 Hawaii Volcanoes 64 aFrom Loope, in press a. grow in Hawaii, and at least 700 of these are reproducing successfully and maintaining populations in the field (Smith, 1985; Wester, in press). At the same time, more than 200 endemic species are believed to be extinct, and another 800 are endangered (Jacobi and Scott, in press). Most sites below 500 meters elevation, and many higher ones, are entirely dominated by alien species (Moulton and Pimm, 1986). Similar patterns of introduction of alien insects, mammals, reptiles and amphibians, and birds have been described (Carson, in press; Moulton and Pimm, 1986). The birds are probably the best documented (Moulton and Pimm, 1986; Olson and James, 1982), although mammals are the most spectacular (from 1 native bat to at least 18 species of alien mammals). At least 86 species of land birds are known to have been present in Hawaii 2,000 years ago, and at least 68 of them were endemic passerines. Forty-five species, including 30 passerines, disappeared around the time of Polynesian colonization; another 11 have disappeared since Europeans arrived; and several more are on the verge of extinction (Moulton and Pimm, 1986; Stone, 1985). In contrast, at least 50 species of alien passerines have become established since 1780. Even casual observers of lower-elevation birds in Hawaii have noted a kaleidoscope of shifting dominance by different species of alien birds over the past 30 years; the one constant has been the near absence of natives. This pattern of successful invasion by cosmopolitan species and the decline of certain native species is not unique to Hawaii. A similar conversion of native-dominated to alien-dominated ecosystems occurs on isolated islands in all the oceans—from the Galapagos to New Zealand to Diego Garcia to Tristan da Cunha and St. Helena (Bramwell, 1979; Carlquist, 1974; King, 1984; Wace and Oilier, 1982). In many cases, the successful invaders are identical—goats (Capra hircus) and guava (Psidium guajava and P. cattleianum) are problems in Hawaii, the Galapagos, and the Rodrigues Islands in the Indian Ocean.

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BioDiversity WHY ARE ISLANDS SUSCEPTIBLE? The reasons why biological invasions are disproportionately successful on islands, and why island species seem more likely to become extinct, have long been debated. Loope (in press b) summarized this discussion with seven possible explanations for the observed patterns: Reduced competitive ability due to repeated “founder effects,” i.e., chance events during colonization by small initial populations Disharmony of functional groups and relative lack of diversity Small populations and genetic variability; restrictive specialization Relative lack of adaptability to change; loss of resistance to consumers and disease Loss of essential co-evolved organisma Relative lack of natural disturbance, especially fire, in the evolutionary history of many island biotas Intensive exploitation by humans He also pointed out that the apparent lack of vigor of island species can be overstated, sometimes with negative consequences. For example, Lyon (1909) interpreted a decline of native óhiá (Metrosideros polymorpha) in Hawaii as reflecting that species’ inability to survive in the modern world, and spearheaded the introduction of many alien species to replace it. In fact, periodic diebacks of natural populations of Metrosideros are a natural feature of forest dynamics in Hawaii and elsewhere in the Pacific (Mueller-Dombois, 1983), and Metrosideros naturally recolonizes most of these areas. More generally, many native island species maintain themselves quite successfully in mixed native/exotic ecosystems (Mueller-Dombois et al., 1981). At the other extreme, it has been argued that alien species are merely temporary components of island ecosystems, certain to be replaced by natives in the course of ecological succession (Allan, 1936; Egler, 1942). In fact, some aliens invade intact native ecosystems, whereas others alter the course of succession in already disturbed sites (Smith, 1985) and seem capable of persisting in those altered sites. Although biological invasions clearly have contributed to the extinction of native species on islands, the importance of direct competition between native and exotic species in causing these extinctions is uncertain. Habitat destruction by humans and feral animals, alterations in basic ecosystem properties caused by newly introduced species, grazing and predation pressure from introduced consumers, and exotic animal diseases (such as avian pox and malaria) appear to be at least equally important. The importance of grazing and predation by alien animals deserves special emphasis. Most isolated oceanic islands originally lacked whole groups of organisms; mammals were especially sparse. Even ants were nearly or entirely absent on some islands, including Hawaii (Medeiros et al., 1986). The introduction of mammals has had enormous effects on island ecosystems throughout the world. Comparisons of islands with introduced ungulates and those without such animals in widely separated Pacific island groups (the Hawaiian

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BioDiversity Islands, the Cook Islands, and the Kermadec Islands) demonstrate that native communities often hold their own in the absence of mammals but that invasions by plants are much more common and disruptive of native communities on heavily grazed islands (Merlin and Juvik, in press). WHAT CAN BE DONE? Biological invasions of oceanic islands appear to be an immense and largely unmanageable problem. Of the approximately 4,600 species of alien plants on Hawaii, more than 700 reproduce in the wild and 86 are considered serious threats to native ecosystems (Smith, 1985). At present, there are neither the resources nor the will to attack a problem of this magnitude. Moreover, while interception and quarantine systems can slow the further introduction of additional exotic species and stop a few indefinitely, the sheer volume and pace of transport by jet aircraft may overwhelm most controls. Finally, any inspection system detailed enough to be broadly effective would necessarily hinder and annoy tourists that are the major economic support of many oceanic islands. Moreover, many island residents have strong reasons for importing or protecting introduced species as agricultural, timber, or forage crops, medicinal or ornamental plants, watershed protection, domestic livestock, pets, agents of biological control, or targets of sport or commercial hunting or fishing. These economic or cultural attachments to alien species mean that there is little chance of developing broad-based, politically effective support for controlling alien species that are not regarded as weeds in the classical (economic) sense. There are nevertheless several steps that can be taken to reduce the effects of biological invasions and protect some of the native biological diversity on isolated oceanic islands: identification of the aliens most likely to threaten native ecosystems and concentration of control efforts on those species; selection of critical habitat areas from which most or all species of aliens are excluded; protection of areas from further habitat destruction; and study of biological invasion and species extinction on islands to learn how these same processes may affect continents. IDENTIFICATION OF PROBLEM SPECIES Identification of the invading species most likely to disrupt native ecosystems requires some understanding of the biology of both the invader and the invaded community. Research designed to obtain that information is now being conducted, and its results are being used in management decisions on many islands. The most disruptive species (not necessarily in order of importance) include herbivorous mammals, vertebrate and invertebrate predators, species that can alter ecosystem-level characteristics of invaded areas, and species that can invade otherwise undisturbed native ecosystems.

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BioDiversity Herbivorous Mammals Grazing and browsing mammals effect islands in such pervasive ways that it is difficult to see how native ecosystems can be protected unless they are eliminated. Studies of whole islands and of exclosures have clearly demonstrated that ungulate populations affect erosion, soil fertility, and the success of invasions by alien plants (Loope and Scowcroft, 1985; Merlin and Juvik, in press; Mueller-Dombois and Spatz, 1975; Vitousek, 1986). Island plants often lack defenses, such as thorns and toxic chemicals, against herbivores, and herbivority reduces total plant cover and selects for better defended alien plants. Moreover, feral pigs (which are widespread on many oceanic islands) directly disrupt soil structure in the course of their feeding. Efforts to eliminate mammals are expensive and difficult, but they have been highly successful in a number of areas (Bramwell, 1979; Stone et al., in press b). In many cases, the removal of grazing animals has been followed by the recovery of native plants and even by the discovery of entirely new species of native plants (Bramwell, 1979; Mueller-Dombois and Spatz, 1975). Predators Alien vertebrate and invertebrate predators can have significant effects on island ecosystems both directly, by eliminating natives, and indirectly, by altering community structure. For example, rats and feral cats affect the breeding success of ground-nesting birds in many areas (Clark, 1981; King, 1984; Wace, 1986). Alien ants altered invertebrate communities in the Hawaiian lowlands years ago, and other ant species are now threatening to do so at high elevations (Medeiros et al., 1986). Invertebrate predators are particularly problematic in that they may eliminate important native pollinators from island faunas. Ecosystem-Level Effects Any alien species that alters ecosystem-level characteristics (such as primary productivity, nutrient availability, hydrological cycles, and erosion) of the area it invades alters the living conditions for all organisms in that area (Vitousek, 1986). It may also alter the kind or quality of the services that natural ecosystems provide to human societies (Ehrlich and Mooney, 1983). Alien animals clearly alter ecosystem properties in a number of ways (as described above), and it is becoming clear that alien plants can do so as well. In Hawaii, for example, the exotic nitrogen-fixing fire tree (Myrica faya) increases the availability of the soil nitrogen in nitrogen-limited volcanic ash deposits (Vitousek, in press). Similarly, the alien grasses Andropogon virginicus and A. glomeratus provide fuel for fires and also sprout rapidly following fires, thereby greatly increasing both their abundance and the overall frequency of fires to the detriment of native species not adapted to fire resistance (Smith, 1985). Invasion of Intact Native Ecosystems Alien animals are frequently (not invariably) able to invade intact native ecosystems, but plants species that can do so are not common. Most often, alien plants invade undisturbed native ecosystems in association with alien animals. In Hawaii,

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BioDiversity alien birds and mammals consume and disseminate the fruit of the aggressive alien plants strawberry guava (Psidium cattleianum) and banana poka (Passiflora mollissima) throughout native forest areas. Interactions between feral pigs and these invading plants are particularly severe: pigs disseminate seeds of these fleshy-fruited aliens, mix them with organic fertilizer, and deposit them into seedbeds, which are cleared by the pigs’ rooting activity. The pigs’ descendants then use fruit of the daughter plants as a major food source (Smith, 1985; Stone, 1985). Similar interactions between cattle and common guava (Psidium guajava) occur in the Galapagos (Bramwell, 1979). These interactions between alien plants and animals further illustrate why control of alien animals is fundamental to protecting the native ecosystems of islands. IDENTIFICATION OF CRITICAL HABITATS A second strategy for limiting the effects of biological invaders is to control manageable alien species in selected critical habitats. This process is expensive and time-consuming, but it does lead to the maintenance of areas as close to their natural state as possible (although birds, flying insects, and microorganisms are of course difficult or impossible to control). Management in “Special Ecological Areas” of Hawaii Volcanoes National Park has been designed to protect areas that represent the major ecosystems in the park by minimizing the influence of alien species. These areas can then act as refugia for threatened native biota and as areas for ecological study and education (Stone et al., in press a; Tunison et al., 1986). HABITAT DESTRUCTION Control over habitat destruction is also essential to protecting biological diversity on oceanic islands. Land clearing or fire in native systems can both destroy individuals of threatened native species and lead to the establishment of alien-dominated successional ecosystems. Conflicts in achieving this objective are inevitable; most islands are neither museums nor biological preserves, and one person’s “habitat destruction” will certainly be another’s source of food or income. Destruction of critical habitat on islands is perhaps most severe on Madagascar, but it is not a problem confined to developing countries. Nearly half of Hawaii’s largest native-dominated lowland rain forest was cleared during 1984 and 1985 in a subsidized endeavor to generate electricity from wood chips. ECOLOGICAL RESEARCH ON ISLANDS Controlling the effects of biological invasions on islands is paramount, but there is also a great deal to be gained from studying their effects carefully. The relative simplicity of the biota of many islands perhaps enables invading species to have greater effects on native communities than they would in continental areas; it certainly facilitates a much more complete evaluation of those effects. Better understanding of biological invasions and their consequences for biological diversity on islands will contribute to the development and testing of basic ecological theory

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BioDiversity on all levels of biological organization. Few of the effects of biological invasions described here are unique to islands; they are only more highly developed and occur most rapidly there, as demonstrated by the invasion of European wild boars into Great Smoky Mountains National Park (Singer et al., 1984). An understanding of the effects of invasions on biological diversity in rapidly responding island ecosystems may give us the time and the tools needed to deal with similar problems on continents; it may even contribute to the prediction and evaluation of the effects of environmental releases of genetically altered organisms. REFERENCES Allan, H.H. 1936. Indigene versus alien in the New Zealand plant world. Ecology 17:187–193. Bramwell, D., ed. 1979. Plants and Islands. Academic Press, London. 459 pp. Carlquist, S. 1974. Island Biology. Columbia University Press, New York. 660 pp. Carson, H.L. In press. Colonization and speciation. In A.H.Gray, M.Crawley, and P.J.Edwards, eds. Colonization and Succession. Blackwell Scientific, Oxford. Clark, D.A. 1981. Foraging patterns of black rats across a desert-montane forest gradient in the Galapagos Islands. Biotropica 13:182–194. Darwin, C.R. 1859. On the Origin of Species by Means of Natural Selection, or the Preservation of Favored Races in the Struggle for Life. 1st edition. John Murray, London. 502 pp. (Facsimile of the first edition published by the Harvard University Press, Cambridge, 1964.) Egler, F.E. 1942. Indigene vs. alien in the development of arid Hawaiian vegetation. Ecology 23:14–23 . Ehrlich, P.R., and H.A.Mooney. 1983. Extinction, substitution, and ecosystem services. BioScience 33:248–253. Elton, C.S. 1958. The Ecology of Invasions by Animals and Plants. Methuen Co., London. 181 pp. Ewel, J.J. 1986. Invasability: Lessons from south Florida. Pp. 214–230 in H.A.Mooney and J. Drake, eds. The Ecology of Biological Invasions of North America and Hawaii. Springer-Verlag, New York. Gorman, M. 1979. Island Ecology. Chapman and Hall, London. 79 pp. Heywood, V.H. 1979. The future of island floras. Pp. 431–441 in D. Bramwell, ed. Plants and Islands. Academic Press, London. Jacobi, J., and J.M.Scott. In press. An assessment of the current status of native upland habitats and associated endangered species on the Island of Hawai’i. Pp. 3–22 in C.P.Stone, C.W.Smith, and J.T.Tunison, eds. Alien Plant Invasions in Hawaii: Management and Research in Near-Native Ecosystems. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu. Kepler, C.B., and J.M.Scott. 1985. Conservation of island ecosystems. ICBP Tech. Pub. 3:255–271. King, C.M. 1984. Immigrant Killers: Introduced Predators and the Conservation of Birds in New Zealand. Oxford University Press, Aukland. 224 pp. Loope, L.L. In press a. An overview of problems with introduced plant species in national parks and reserves of the United States. In C.P.Stone, C.W.Smith, and J.T.Tunison, eds. Alien Plant Invasions in Hawaii: Management and Research in Near-Native Ecosystems. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu. Loope, L.L. In press b. Haleakala National Park and the “island syndrome.” In L.K.Thomas, ed. Proceedings of a Symposium on Ecology and Management of Exotic Species. Conference on Science in the National Parks, Ft. Collins, Colorado, July 1986. U.S. National Park Service and The George Wright Society, Washington, D.C. Loope, L.L., and P.G.Scowcroft. 1985. Vegetation response within exclosures in Hawai’i: A review. Pp. 377–402 in C.P.Stone and J.M.Scott, eds. Hawai’i’s Terrestrial Ecosystem: Preservation and Management. Cooperative National Park Resources Study Unit, University of Hawaii, Honolulu. Lyon, H.L. 1909. The forest disease on Maui, Hawaii. Plant. Rec. 1:151–159. Medeiros, A.C., L.L.Loope, and F.R.Cole. 1986. Distribution of ants and their effects on endemic biota of Haleakala and Hawaii Volcanoes National Parks: A preliminary assessment. Pp. 39–

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BioDiversity 51 in Proceedings of the Sixth Conference in Natural Sciences, Hawaii Volcanoes National Park, June 1986. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu. Merlin, M.D., and J.O.Juvik. In press. Relationships between native and alien plants on oceanic islands with and without wild ungulates. In C.P.Stone, C.W.Smith, and J.T.Tunison, eds. Alien Plant Invasions in Hawaii: Management and Research in Near-Native Ecosystems. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu. Mooney, H.A., and J.Drake, eds. 1986. The Ecology of Biological Invasions of North America and Hawaii. Springer-Verlag, New York. 321 pp. Moulton, M.P., and S.L.Pimm. 1986. Species introductions to Hawaii. Pp. 231–249 in H.A. Mooney and J.Drake, eds. The Ecology of Biological Invasions of North America and Hawaii. Springer-Verlag, New York. Mueller-Dombois, D. 1983. Canopy dieback and successful processes in Pacific forests. Pac. Sci. 37:317–325. Mueller-Dombois, D., and G.Spatz. 1975. The influence of feral goats on the lowland vegetation of Hawaii Volcanoes National Park. Phytocoenologia 3:1–29. Mueller-Dombois, D., K.W.Bridges, and H.L.Carson, eds. 1981. Island Ecosystems: Biological Organization in Selected Hawaiian Communities. Hutchinson-Ross, Stroudsburg, Pa. 583 pp. Nagata, K.M. 1985. Early plant introductions in Hawai’i. Hawaii. J. Hist. 19:35–61. Olson, S.L., and H.F.James. 1982. Fossil birds from the Hawaiian Islands: Evidence for wholesale extinction by man before western contact. Science 217:633–635. Singer, F.J., W.T.Swank, and E.E.C.Clebsch. 1984. Effects of wild pig rooting in a deciduous forest. J. Wildl. Manage. 48:464–473. Smith, C.W. 1985. Impact of alien plants on Hawai’i’s native biota. Pp. 180–250 in C.P.Stone and J.M.Scott, eds. Hawai’i’s Terrestrial Ecosystems: Preservation and Management. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu. Stone, C.P., 1985. Alien animals in Hawai’i’s native ecosystems: Towards controlling the adverse effects of introduced vertebrates. Pp. 251–297 in C.P.Stone and J.M.Scott, eds. Hawai’i’s Terrestrial Ecosystems: Preservation and Management. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu. Stone, C.P., and J.M.Scott, eds. 1985. Hawai’i’s Terrestrial Ecosystems: Preservation and Management. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu. 584 pp. Stone, C.P., C.W.Smith, and J.T.Tunison, eds. In press a. Alien Plant Invasions in Hawaii: Management and Research in Near-Native Ecosystems. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu. Stone, C.P., P.K.Higashino, J.T.Tunison, L.W.Cuddihy, S.J.Anderson, J.D.Jacobi, T.J. Ohashi, and L.L.Loope. In press b. Success of alien plants after feral goat and pig removal. In C.P. Stone, C.W.Smith, and J.T.Tunison, eds. Alien Plant Invasions in Hawaii: Management and Research in Near-Native Ecosystems. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu. Tunison, J.T., C.P.Stone, and L.W.Cuddihy. 1986. SEAs provide ecosystem focus for management and research. Park Sci. 6(3):10–13. Vitousek, P.M. 1986. Biological invasions and ecosystem properties: Can species make a difference? Pp. 163–176 in H.A.Mooney and J.Drake, eds. The Ecology of Biological Invasions of North America and Hawaii. Springer-Verlag, New York. Vitousek, P.M. In press. Effects of alien plants on native ecosystems. In C.P.Stone, C.W.Smith, and J.T.Tunison, eds. Alien Plant Invasions in Hawaii: Management and Research in Near-Native Ecosystems. Cooperative National Park Resources Studies Unit, University of Hawaii, Honolulu. Wace, N.M. 1986. Control of rats on islands—research is needed. Oryx 20:79–86. Wace, N.M., and C.D.Oilier. 1982. Biogeography and geomorphology of South Atlantic Islands. Pp. 733–758 in National Geographic Society Research Reports. National Geographic Society, Washington, D.C. Wallace, A.R. 1880. Island Life. Macmillan, London. 526 pp. Wester, L.L. In press. Alien plants and their status in Hawaii. In C.P.Stone, C.W.Smith, and J.T.Tunison, eds. Alien Plant Invasions in Hawaii: Management and Research in Near-Native Ecosystems. Cooperative National Park Studies Unit, University of Hawaii, Honolulu.

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