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Biological invasions are a widespread and significant component of human-caused global environmental change. The extent of invasions of oceanic islands, and their consequences for native biological diversity, have long been recognized. However, invasions of continental regions also are substantial. For example, more than 2,000 species of alien plants are established in the continental United States. These invasions represent a human-caused breakdown of the regional distinctiveness of Earth's flora and fauna - a substantial global change in and of itself. Moreover, there are well-documented examples of invading species that degrade human health and wealth, alter the structure and functioning of otherwise undisturbed ecosystems, and/or threaten native biological diversity. Invasions also interact synergistically with other components of global change, notably land use change. People and institutions working to understand, prevent, and control invasions are carrying out some of the most important - and potentially most effective - work on global environmental change.
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New Zealand Journal of Ecology (1997) 21(1): 1-16 ©New Zealand Ecological Society
Humans move species beyond their native ranges
both deliberately and inadvertently, and many of
these species become established and spread in their
new habitat. The list of established introduced
species grows annually, as does the number of them
that cause significant economic and ecological
effects. One recent and notorious example in North
America is the Eurasian zebra mussel - which like
many other aquatic organisms entered in the ballast
water of ships, and like many others spread rapidly
once it arrived. The invasion of zebra mussels is
unusual in the magnitude of its economic
consequences; the mussels grow and reproduce
rapidly, covering river and lake bottoms and
municipal and industrial water inlets. The cost of
clearing blocked intake pipes has been calculated to
be approximately US$2 billion (Office of
Technology Assessment, 1993). Zebra mussels also
alter populations of algae and the concentrations of
nutrients in whole ecosystems (Caraco et al., 1997),
and they are continuing to spread in rivers, lakes,
and canals throughout North America.
We suggest that biological invasions by
notorious species like the zebra mussel, and its many
less-famous counterparts, have become so
widespread as to represent a significant component
of global environmental change. This point has been
made before (eg Elton, 1958), but is not widely
appreciated, even by the global change research
community or by those who study and/or work to
control biological invasions. In part, this lack of
appreciation reflects the fact that our perception is
limited spatially - it is possible to document the
presence and importance of biological invasions
almost anywhere, but more difficult to perceive that
invasions are almost everywhere. In part, it may also
reflect a narrow view of global environmental
change, one that emphasizes climate change (global
warming) at the expense of other, equally significant
components of human-caused global change.
In this paper, we place biological invasions in
context with other human-caused global environ-
mental changes; briefly describe the global extent of
biological invasion; illustrate the consequences of
particular invasions as they affect human health and
wealth, and/or the functioning and biological
diversity of natural ecosystems; discuss interactions
between biological invasions and other components
of global change; and describe ways that society can
prevent, manage, and/or cope with invasions.
Human-caused global
environmental change
Our perspective on global environmental change is
summarized in Fig. 1, in which the third level lists
Department of Biological Sciences, Stanford University, Stanford, California 94305 USA.
Department of Integrative Biology, University of California, Berkeley, California 94720 USA.
Pacific Islands Ecosystem Research Center, Haleakala National Park Field Station, P.O. Box 369, Makawao, Hawaii
96768 USA.
Department of Botany, University of California, Davis, California 95616 USA
Noxious Weed Program, Animal and Plant Health Inspection Service, P.O. Box 279, Whiteville, North Carolina 28472 USA
Summary: Biological invasions are a widespread and significant component of human-caused global
environmental change. The extent of invasions of oceanic islands, and their consequences for native
biological diversity, have long been recognized. However, invasions of continental regions also are
substantial. For example, more than 2,000 species of alien plants are established in the continental United
States. These invasions represent a human-caused breakdown of the regional distinctiveness of Earth’s flora
and fauna - a substantial global change in and of itself. Moreover, there are well-documented examples of
invading species that degrade human health and wealth, alter the structure and functioning of otherwise
undisturbed ecosystems, and/or threaten native biological diversity. Invasions also interact synergistically
with other components of global change, notably land use change. People and institutions working to
understand, prevent, and control invasions are carrying out some of the most important - and potentially most
effective - work on global environmental change.
Keywords: Biological invasion; Invasions into parks and preserves; Invasion and biological diversity;
Invasion and ecosystems; Land-use change; Introduced pests and pathogens.
six relatively well-documented global changes: the
increasing concentration of CO
in the atmosphere,
alterations to the global biogeochemical cycle of
nitrogen and other elements, the production and
release of persistent organic compounds such as the
chlorofluorocarbons, widespread changes in land use
and land cover, hunting and harvesting of natural
populations of large predators and consumers, and
biological invasions by non-native species. All of
these clearly represent ongoing global changes, and
all are clearly human-caused.
These changes are driven proximately by the
industrial and agricultural enterprises of humanity,
and ultimately by the explosive growth over the past
two centuries of both the human population and per
capita resource use. The six well-documented
changes in turn cause other global changes; some
drive global climate change by enhancing the
greenhouse effect, and some drive loss of biological
diversity by causing the extinction of species and
genetically distinct populations. The importance of
biological invasion as one of these global changes is
described here.
The scope and distribution of
How widespread are biological invasions?
The importance of biological invasions to oceanic
island ecosystems has long been recognized.
Invasions also are frequent in many continental
areas, where they represent a substantial component
of the flora and fauna of most countries. Table 1
summarizes the pattern and number of plant
invasions in many regions. On continents, there is an
increase in the number of invading species per log
(area) from north to south until one reaches dry
subtropical regions; invasions are relatively low in
the tropics, then increase again in south temperate
areas. Heavily-visited islands are invaded to a
greater extent (per log area) than continents or less-
trafficked islands. The information in Table 1
confirms patterns illustrated by Rejmánek and
Randall (1994). A more general point of Table 1 is
that invasions are everywhere, on continents as well
Figure 1: Components of global environmental change. Growth in the size of and resource use by the human population is
expressed through growing industrial and agricultural (including forestry, grazing, etc) activity. These have caused a set of
relatively well-documented global environmental changes (well-documented both in the sense that they are occurring, and
in that they are human-caused), including increasing concentrations of carbon dioxide in the atmosphere, the production
and distribution of novel and persistent compounds such as chlorofluorocarbons (with their attendant effects on
stratospheric ozone) and PCBs, global-scale alteration of the biogeochemical cycles of nitrogen, sulfur, and other
elements, changes in land use and land cover, the removal of top predators from most terrestrial and many marine
ecosystems, and biological invasions by exotic species. These components of change interact; they will also drive changes
in global climate, and losses of biological diversity. After Vitousek (1994).
Table 1: Established alien vascular plants in selected continental and island floras. Species richness of alien floras is
expressed as: (1) The total number of naturalized species. (2) The percentage of naturalized species in the flora. (3) The
number of naturalized species/log(area); there is generally an approximately linear relationship between the numbers of
species in an area and the log(area). Species that are not established beyond cultivation or which have not been confirmed
in this century are not included.
Number of Number of Percentage of Number of
Area native established established alien species
Region/country (km
)species alien species alien species per log(area) Sources
Russian Arctic 3,500,000 1,403 104 6.9 15.9 1
Europe 10,382,000 11,820 721 5.7 102.8 2
Western and central Sahara 4,000,000 830 <28 <3.3 <4.2 3
Tropical Africa 22,300,000 23,500 536 2.2 72.9 4
Southern Africa 2,693,389 20,573 824 3.9 128.1 5
Alaska 1,528,200 1,229 144 10.5 23.3 6
Canada 9,976,139 3,270 940 22.3 134.3 7
Coterminous U.S.A. 7,844,400 ca 17,300 ca 2,100 ca 10.8 ca 304.6 8
Peru 1,285,200 17,900 314 1.7 51.4 9
Chile 756,600 4,437 678 13.3 115.3 10
Australia 7,686,848 15,638 1,952 11.1 283.5 11
Murmansk area 120,000 983 82 7.7 16.1 12
Finland 338,145 1,250 247 16.5 44.7 13
Norway 323,878 1,195 580 32.7 105.3 14
Poland 312,680 2,250 275 10.9 50.1 15
France 549,619 4,350 480 9.9 83.6 16
Egypt 1,000,250 2,015 86 4.1 14.3 17
Djibouti 23,000 641 44 6.4 10.1 18
Uganda 236,040 4,848 152 3.1 28.3 19
Rwanda 26,338 2,500 93 3.6 21.1 20
Namibia 824,293 3,159 60 1.9 10.1 21
Swaziland 17,366 2,715 110 3.9 25.9 22
Cape region 90,000 8,270 441 5.1 88.9 23
NW Territories (Canada) 3,380,000 1,055 53 4.8 8.1 24
British Columbia 948,600 2,048 547 21.1 91.5 25
Ontario 1,068,587 2,056 805 28.1 133.5 26
Minnesota 217,136 1,618 392 19.5 73.5 27
New York 137,795 1,940 1,083 35.8 210.7 28
Missouri 174,242 1,920 634 24.8 121.1 29
California 411,020 4,844 1,025 17.5 182.6 30
Central Florida 68,738 1,746 440 20.1 90.9 31
Texas 692,400 4,498 492 9.9 84.2 32
Baja California 143,700 2,480 183 6.9 35.5 33
Valle de Mexico 7,500 1,910 161 7.8 41.5 34
Chiapas (Mex.) 74,211 6,650 206 3.1 42.3 35
Panama 77,082 7,123 263 3.6 53.8 36
Choco (Colombia) 42,205 3,818 48 1.2 10.4 37
Guaianas 469,234 8,030 287 3.5 50.6 38
Monte Video area 664 843 180 17.6 63.8 39
Buenos Aires area 80,000 1,369 363 21.1 74.1 40
Northern Territory (Australia) 1,331,900 3,293 262 7.4 42.8 41
Queensland 1,707,520 7,535 1,161 13.3 186.3 42
Perth region 10,500 1,510 547 26.6 136.1 43
New South Wales 792,150 4,677 1,253 21.1 212.4 44
Victoria 224,983 2,773 1,190 30.1 222.3 45
Table 1 continued over
Table 1: continued
Number of Number of Percentage of Number of
Area native established established alien species
Region/country (km
)species alien species alien species per log(area) Sources
Devon Is. (Canada) 58,000 115 0 0 0 46
Jan Mayen 380 57 4 6.5 1.6 47
Greenland 326,000 427 86 16.8 15.6 48
Queen Charlotte 9,200 469 116 19.8 29.3 49
British Isles 244,872 1,255 945 42.9 175.4 50
Sakhalin 75,370 1,081 92 7.8 18.9 51
Newfoundland 144,890 906 292 24.4 56.6 52
San Juan Islands (USA) 390 546 283 34.1 109.2 53
Angel Is. (Calif.) 3 282 134 32.2 280.9 54
Santa Cruz (Calif.) 244 462 157 25.4 65.8 55
Crete 8,700 1,586 92 5.5 23.4 56
Canary Islands 7,252 1,254 680 35.2 176.2 57
Bermuda 54 165 303 64.7 174.9 58
Bahamas 14,500 1,104 246 18.2 59.1 59
Hormoz \ Qeshm 1,290 230 49 17.6 15.8 60
Cuba 114,500 5,790 376 6.1 74.3 61
Hawaii 16,764 1,143 891 43.8 210.9 62
Cayman Is. 259 536 65 10.8 26.9 63
Puerto Rico 8,897 2,741 356 11.5 90.1 64
Guadalupe & Martinique 2,620 1,668 360 17.8 105.3 65
Guam 583 327 185 36.1 66.8 66
Ascension 94 25 >120 >82.8 >60.8 67
Galapagos 7,870 604 260 30.1 66.7 68
Rodrigues 40 132 305 69.8 190.4 69
Tristan da Cunha 102 58 119 67.2 59.3 70
Lord Howe 550 206 173 45.6 63.1 71
New Zealand 268,575 2,449 1,623 39.9 298.9 72
Marion \ Prince Edward 330 21 10 32.3 4.1 73
Auckland 450 187 41 17.9 15.5 74
Falklands 11,900 163 83 33.7 20.4 75
Tierra del Fuego 48,700 417 128 23.5 27.3 76
Macquarie Is. 90 44 5 10.2 2.6 77
Southern Shetland Islands 1,390 2 0 0 0 78
Sources: 1. Tolmachev (1960-1987), Gorodkov and Poyarkova (1953-1966); 2. Tutin et al.(1964-1980), Tutin et al.(1993),
Clement and Foster (1994), Rejmánek (unpublished); 3. Ozenda (1991); 4. Lebrun and Stork (1991-1995), Rejmánek
(unpublished); 5. Arnold and de Wet (1993); 6. Welsh (1974); 7. Boivin (1968); Scoggan (1978-1979); 8. Kartesz (1994), Morin
(1993), Shetler and Skog (1979), U.S. Department of Agriculture (1982); 9. Barko and Zarucchi (1993), Tryon and Stolze (1989-
1994); 10. Marticorena and Quezada (1985), Aroyo (unpublished); 11. Hnatiuk (1990); 12. Gorodkov and Poyarkova (1953-
1966); 13. Tutin et al. (1964-1980), Tutin et al. (1993), Ahti and Hämet-Ahti (1971), Suominen (1979); 14. Fremstad, Elven
and Tømerås (1994); 15. Kornas (1990); 16. Tutin et al.(1964-1980); Tutin et al.(1993); Jovet (1971); 17. Täckholm (1974);
18. Lebrun, Audru and Cesar (1989); 19. Rejmánek (unpublished); 20. Troupin (1978-1988); 21. Merxmüller (1966-1972),
Roessler and Merxmüller (1976); 22. Kemp (1983); 23. Arnold and de Wet (1993), Bond and Goldblatt (1984); 24. Porsild and
Cody (1980); 25. Douglas, Straley and Meidinger (1990-1994); 26. Morton and Venn (1990); 27. Ownbey and Morely (1991);
28. Mitchell (1986); 29. Yatskiewych and Turner (1990); 30. Hickman (1993), Rejmánek and Randall (1994); 31. Wunderlin
(1982); 32. Johnson (1990); 33. Wiggins (1980), Gould and Moran (1981); 34. Rzedowski and Rzedowski (1989); 35. Breedlove
(1986); 36. D’Arcy (1987); 37. Forero and Gentry (1989); 38. Boggan et al.(1992); 39. Lombardo (1982-1984); 40. Cabrera
and Zardini (1978); 41. Hnatiuk (1990); 42. Hnatiuk (1990); 43. Marchant et al. (1987); 44. Hnatiuk (1990); 45. Carr (1993);
46. Barrett and Teeri (1973); 47. Lid (1964); 48. Porsild (1932), Bøcher et al.(1978), Bay (1993); 49. Calder and Taylor (1968);
50. Clement and Foster (1994), Ryves et al.(1996); 51. Vorobiev et al.(1974); 52. Rouleau and Lamourex (1992); 53. Atkinson
and Sharpe (1985); 54. Ripley (1980); 55. Wallace (1985), Junak et al. (1995); 56. Barclay (1986); 57. Kunkel (1980); 58. Britton
(1918); 59. Correll and Correll (1982); 60. Kunkel (1977); 61. Borhidi (1991); 62. Wagner, Herbst and Sohmer (1990), Wilson
(1996); 63. Proctor (1984); 64. Liogier and Martorell (1982), Francis and Liogier (1991); 65. Fournet (1978); 66. Stone (1970),
Lee (1974); 67.Duffey (1964), Conk (1980); 68. Lawesson (1990); 69. Strahm (unpublished); 70. Dean et al.(1994); 71. Pickard
(1984); 72. Atkinson and Cameron (1993); 73. Gremmen (1982); 74. Meurk (1982); 75. Moore (1968);76. Moore (1983); 77.
Selkirk, Seppelt and Selkirk (1990); 78. Komárková, Poncet and Poncet (1990)
as islands, and in the tropics as well as temperate
regions. The continental United States and Australia
both support ~2,000 species of established alien
plants! While the absolute number of species
generally is less, introduced plants on some islands
make up half or more of the flora.
Biological invasions by fishes and birds are not
as frequent as invasions by plants. However, some of
the same patterns are evident (Table 2). Isolated
islands often support more introduced than native
fish species. Even many continental sites (for
example, California, Europe and Brazil) have
relatively large numbers of non-native fish species.
The lack of data on numbers of fish introductions in
Africa does not imply that they are unimportant - for
example, introduction of Nile perch (Lates nilotica)
and tilapia (several species in 3 genera) into Lake
Victoria has led to dramatic species loss and
ecosystem change in a matter of a few decades
(Goldschmidt, 1996).
Introduced birds have established wild
populations in most countries where data are
available, and in some areas (Hawaii, New Zealand)
they comprise a substantial proportion of the
avifauna. Outside of urban areas, numbers of
introduced bird species are relatively low in most
continental regions (compared to islands). However,
individual species can be quite abundant in
continental habitats - witness the widespread and
abundant European house sparrow (Passer
domesticus) and starling (Sturnus vulgaris) in North
Invasion Into U.S. Parks and Reserves
Biological invasions are particularly prominent in
disturbed areas, leading some to consider invasions to
be primarily consequences of disturbance rather than
a component of change in their own right. Parks and
biological preserves generally represent the least-
altered areas of land - a former director of the U.S.
National Park Service championed the concept of the
parks as a national analogue of “miners’ canaries”,
relatively pristine sites where the pervasiveness of
environmental deterioration might be evaluated. What
do parks and reserves in the United States tell us
about pervasiveness of invasions?
Vascular plants: Floristic lists for a large
sample of U.S. reserves have 5 - 25% non-native
species. However, the majority of introductions are
indeed confined to disturbed areas and appear to
Table 2: Native and exotic freshwater, inland fish, and breeding bird species in selected regions and countries around the
Region Area (km
)native exotic reference native exotic reference
Europe 10,400,000 74 1 514 27 11, 12
California 411,020 76 42 2, 3
Alaska 1,528,200 55 1 4
Canada 9,976,139 177 9 5
Mexico 1,958,200 275 26 5
Australia 7,686,848 145 22 5 32 12
South Africa 3,500,000 107 20 5 900 14 12, 13
Peru 1,285,200 12 6
Brazil 8,512,000 517 76 7 1,635 2 14
Bermuda 54 6 12
Bahamas 14,500 288 4 12, 15
Cuba 114,500 10 6 3 12
Puerto Rico 8,897 3 32 8 105 31 16
Hawaii 16,764 6 19 9 57 38 9
New Zealand 268,575 27 30 10 155 36 17
Japan 372,197 13 6 248 4 12, 18
a Includes only freshwater, inland species.
b Includes only permanent and breeding non-permanent species.
Sources:1. Holcik (1991); 2. McGinnis (1984); 3. Courtenay et al.(1984); 4. Moyle (1986); 5. Macdonald, Kruger and Ferrar
(1986); 6. Welcomme, R.L. (1981); 7. Nomura (1984); 8. Erdman (1984); 9. Stone and Stone (1989); 10. McDowall (1984);
11. Jonsson (1993); 12. Long (1981); 13. Roberts (1985); 14. Sick (1993); 15. Paterson (1972); 16. Raffaele (1989); 17. Kinsky
(1980); 18. Higuchi, Minton and Katsura (1995).
pose little or no threat to native species or
ecosystems (Loope, 1992). In continental areas, the
most important exceptions to this generalization
include invasions into otherwise little-altered semi-
arid areas by grasses, most notably the annual
cheatgrass (Bromus tectorum), and invasion into
riparian habitats and wetlands (i.e. tamarisk in the
Southwest, Melaleuca in the Florida Everglades, and
purple loosestrife in northeast and midwest) (Loope,
1992). Hawaiian reserves, where the percentage of
non-native species in the flora reaches 50 - 70 % and
plant invasions clearly threaten native biota, confirm
the vulnerability of islands to invasion.
Ungulates: Feral pigs may be the single most
damaging introduction in national parks and reserves
of the United States. Singer (1981) found that they
inhabit 13 areas in the National Park system, in
southeastern U.S., Hawaii, and California. Effects of
pigs on otherwise undisturbed areas are severe and
pervasive in Great Smoky Mountains National Park
and in Haleakala and Hawaii Volcanoes National
Parks. In Hawaii, pigs are major dispersers and
facilitators of plant invaders (Stone, 1985). Other
particularly damaging invasive ungulates in parks
include feral goats in Hawaii (now largely removed
from Hawaiian parks); feral burros in Death Valley
(now largely removed), Grand Canyon, and other
southwestern parks; and mountain goats (Oreamnos
americanus) in Olympic National Park.
Aquatic and wetland ecosystems: Invasions of
aquatic and wetland ecosystems of continental U.S.
are fully as severe as island invasions. In Sequoia-
Kings Canyon National Park, for example,
intentionally introduced brook trout (Salvelinus
fontinalis) and brown trout (Salmo trutta) have
displaced native rainbow trout (Onchorhynchas
mykiss) in many streams (G. Larson, personal
communication). Brook and rainbow trout
introduced in waters previously barren of fish have
greatly reduced native invertebrate organisms and
amphibians. In Great Smoky Mountains National
Park, the introduced rainbow trout threatens native
brook trout populations with local extirpation (G.
Larson, personal communication). Even the
relatively pristine waters of Glacier National Park
have been largely (84 %) compromised by past fish
introductions (Marnell, 1995).
Forest pathogens and insects: White pine blister
rust (Cronartium ribicola) and the balsam woolly
adelgid (Adelges piceae) illustrate the devastating
effects of introduced forest “pests ”, even in
undisturbed parks and preserves. Both were brought
to the U.S. 80—100 years ago on European nursery
stock, and (after years of harm to commercial forest
concerns) both are now affecting U.S. parks and
White pine blister rust attacks five-needled
pines; it is now causing increasing mortality of sugar
pine (Pinus lambertiana) in forests of Yosemite and
Sequoia-Kings Canyon National Parks (L. Bancroft,
pers. comm.). Whitebark pine (P. albicaulis) also is
being hit hard; fewer than one tree in 10,000 is rust
resistant, and large die-offs are expected to occur
through the range of whitebark pine. Since
whitebark pine seeds are an extremely important
food of the grizzly bear and other animals (Kendall,
1995), decline of the tree may have severe
consequences in Glacier, Yellowstone and Grand
Teton National Parks.
Balsam woolly adelgid attacks true firs of the
genus Abies , causing mortality within 2—7 years
through feeding and chemical damage to vascular
tissue. This small cottony insect is particularly
damaging to Fraser fir (Abies fraseri), a species
found only in the southern Appalachian Mountains,
where it occurs primarily within the high-elevation
spruce-fir forest of Great Smoky Mountains National
Park. Since its discovery in 1963 in the park, the
adelgid has killed nearly all adult (cone-bearing) fir
trees in the park (Langdon and Johnson, 1992).
Consequences of invasions
The number and variety of species introductions
makes clear that it is no exaggeration to say
biological invasions are breaking down the
biogeographic barriers that have created and
maintained the major floral and faunal regions of
Earth. In other words, invasions are blurring the
regional distinctiveness of Earth’s biota. However,
while all human-caused biological invasions
represent environmental change, we are not equally
concerned about the consequences of all of them.
Many invasions are reflections of other changes,
rather than being themselves drivers of change. For
example, invading plants that only occupy roadside
areas cannot now be regarded as serious threats to
native biological diversity; they are a consequence
of land-use change (which may itself threaten
diversity). Moreover, some introduced species
clearly are beneficial to humanity; for example, it
would be impossible to support the present
population of the United States entirely on native
foods. However, some invading species degrade
human health and/or wealth directly; others affect
the structure and functioning of ecosystems, and/or
the maintenance or restoration of native biological
diversity. We will discuss an example of each of
these, to illustrate some of the consequences of
current invasions. For each that we discuss, there are
many others that are at least as well documented and
at least as damaging.
Human Health
Most infectious diseases themselves are human-
transported biological invaders over most of their
range. Several centuries ago, the indigenous people
of North America could have cited smallpox as a
devastating Old World invader (Crosby, 1986) - just
as modern Americans can point to HIV.
Introduced species themselves can act as vectors
of disease. One recent example is the Asian tiger
mosquito Aedes albopictus . Its larvae were brought
into the United States as hitch-hikers in used car and
truck tires imported for retreading and resale
(Craven et al. , 1988). Two earlier introductions of
A. albopictus in shipments of military tires had
failed to establish - but with the growth of
commercial importation, A. albopictus and other
mosquitoes have been imported more frequently (6.8
tires/10,000 were found to be infested in 1986), and
over a much wider area (Craven et al. , 1988). A.
albopictus became established in the U.S. in the
1980s, and as of 1992 occurred in 25 states. It can
feed on most mammals and birds; in its natural
range, it is a known vector of dengue fever and other
human arbroviruses. Perhaps most importantly, in
the U.S. it is a documented vector for eastern equine
encephalitis, an often-fatal viral infection of people
as well as horses (Craig, 1993).
The zebra mussel invasion mentioned above is a
recent invasion that has been expensive for North
American cities and industries. Other invasions
affect crops, rangelands, and commercial forests,
costing many millions of dollars annually in lost
yields and control efforts. Invasions can also be
costly to developing economies, where the margin
for dealing with additional costs is less. One
example is the golden snail (Pomacea canaliculata)
in Asian rice ecosystems. The snail was brought
from South America to Taiwan to provide a
supplemental source of protein and export income to
small rice farms. Its benefits were illusory - local
people find the snail distasteful (a recipe calling for
“washing in a vinegar solution repeatedly to remove
mucus and slime” may help to explain why), and the
export market was closed by health concerns (Food
and Agricultural Organization, 1989). At the same
time, the costs of golden apple snail importation
were high - the snail has rapid population growth,
spreads rapidly through irrigation canals, and
voraciously consumes young rice plants. When the
costs became clear, the entrepreneurs who imported
the snail simply exported it to other countries; it has
now spread throughout east and southeast Asia
(Fig. 2).
The economic costs of this invasion have been
evaluated carefully in the Philippines (Naylor,
1996). In 1990 alone, the total cost to farmers was
$27.8—45.3 million, split among costs of control
with molluscides and handpicking, replanting costs,
and yield losses (despite control and replanting).
This amounted to 25—40 % of what the Philippines
spent on rice imports in 1990; it represents just one
year’s damage in one of many infested countries.
Ecosystem effects
Invaders that alter ecosystem processes such as
primary productivity, decomposition, hydrology,
geomorphology, nutrient cycling and/or disturbance
regimes do not simply compete with or consume
native species - they change the rules of existence
for all species. Invaders that affect each of these
processes are known; we cannot discuss all of them
here, but one dramatic example is the invasion of the
nitrogen-fixing tree Myrica faya into Hawaii
Volcanoes National Park. Seeds of Myrica are
dispersed by a variety of native and introduced birds,
Figure 2: Distribution and spread of the golden apple snail
through Asian rice-growing countries. From Naylor
and thereby readily reach young sites created by
volcanic eruptions. Studies in Hawaii Volcanoes
National Park show that: 1) plant growth in young
volcanic sites is profoundly limited by low N
availability in soil; 2) colonization by Myrica
increases total inputs of N by more than 4-fold; 3)
the N fixed by Myrica cycles rapidly through Myrica
and to biologically available pools in the soil
(Vitousek and Walker, 1989); and 4) the N that
Myrica adds to sites alters community composition
of other plant species, and of soil organisms - in both
cases towards dominance by other non-native
organisms. In essence, invasion by one species
changes the composition and dynamics of the entire
ecosystem (Vitousek and Walker, 1989).
Effects on Biological Diversity
The eastern deciduous forests of North America
represent a diverse continental ecosystem, one that
might be expected to be as resistant as any to
biological invasions. These forests were cleared
extensively in the 1800s, but have recovered
substantially in this century. The scientific community
has put a great deal of effort into determining current
and probable future effects of climate change,
increased CO
concentrations, acid rain, and oxidant
air pollution on these forests. However, by far the
greatest perturbations to these ecosystems in this
century have involved the invasion of wave after wave
of introduced pests and diseases (Sinclair, Lyon and
Johnson, 1987, Campbell and Schlarbaum, 1994,
Niemelä and Mattson 1996). Some of these pests,
such as gypsy moth (Lymantria dispar), consume a
variety of species, and their effects on forest diversity
are not yet known. Other more specialized pathogens
have eliminated the American chestnut, Castanea
dentata (once a dominant component of eastern
forests) and American elm, Ulmus americana from
the eastern forest. Other tree species undergoing major
decline due to non-native diseases or insects include
American beech (Fagus grandifolia), mountain ash
(Sorbus americana), butternut (Juglans nigra), eastern
hemlock (Tsuga canadensis), flowering dogwood
(Cornus florida), and sugar maple (Acer saccharum)
(Langdon and Johnson, 1992, Campbell and
Schlarbaum, 1994) - in addition to the Fraser fir
discussed above. We suspect that invasions will
continue to represent the most important factor
reducing diversity of these forests for the foreseeable
Interactions With Other Global Changes
In addition to being a component of global change,
biological invasions interact with the other major
components of change (Huenneke, 1996). We
discuss interactions with two of these - land use
change and extinction/loss of biological diversity.
Land Use Change
Biological invasions interact with land-use change in
several ways. The most obvious of these is through
human alteration of disturbance regimes. The
association between disturbance and invasion was
noted above - and humans are now the premier
agents of disturbance on the planet. Moreover, we
have not merely increased the frequency and/or
intensity of disturbance; in many cases we have
created types of disturbances that are unlike
anything in the evolutionary history of many species.
These alterations have promoted invasion, often by
species that are associated with similar disturbances
within their original range (Hobbs and Huenneke,
The interaction between land use change and
invasion is not a one-way street. Both introduced
plants and animals can alter the disturbance regime
of sites they invade (D’Antonio et al. , in press). For
example, introduced fire-promoting grasses have
invaded many arid or semi-arid ecosystems, and in
so doing have increased the frequency, size and/or
intensity of fires. A recent literature review
concluded that non-native, fire-promoting grasses
are common in the Americas, Australia and Oceania,
where they threaten the maintenance of remaining
seasonally-dry tropical forests in some areas, and
represent a major impediment to the restoration
(even reforestation) of cleared lands (D’Antonio and
Vitousek, 1992). The dynamics of the introduced
grass/fire cycle are summarized in Fig. 3. In this
scenario, initial disturbance such as land clearing
(which often utilizes fire) allows the invasion of
introduced grasses. These grasses then create
microclimate and fuel conditions that favor an
increased frequency of fire. Fire in turn selects
against many native species and further promotes
fire-adapted grasses, resulting in a positive feedback
that perpetuates low diversity grassland or savanna.
External disturbance is not always required to
set this feedback in motion - at least in some cases,
grass invasion in and of itself is sufficient to enhance
fuel loading and increase the probability of fire. It is
even possible for grass species to promote human-
caused land use change. For example, the ready
availability of forage grasses that withstand grazing
and drought conditions has lead to the conversion of
millions of hectares of Sonoran desert woodland to
near monocultures of African buffel grass (Cenchrus
ciliaris , also called Pennisetum ciliare). (Yetman
and Burquez, 1994). Likewise, in Central and South
America dry and mesic forests have been replaced
by grazing tolerant (and fire responsive) African
pasture grasses (Parsons, 1972).
Perhaps the most dramatic and well documented
example of an introduced grass/fire cycle is the
invasion of the intermountain west in North America
by the European cheatgrass (Bromus tectorum). This
annual species invaded shrub/steppe habitat in the
Great Basin that was previously dominated by native
shrubs and native perennial grasses. After cheatgrass
invasion, fire frequency has increased from an
estimated once every 60—110 yr to once every 3—5
yr. Almost 5 million hectares of land in Idaho and
Utah are now nearly monospecific stands of
cheatgrass (Whisenant, 1990).
The suppression of disturbance can also
promote invasion by introduced species, particularly
in aquatic ecosystems where reproduction and
recruitment are often synchronized with disturbance
cycles. Indeed, the damming and impoundment of
most of the rivers in the U.S. has been correlated
with the invasion of rivers, streambanks and
floodplains by introduced species, and with the rapid
conversion of diverse, native riparian forests to low
diversity stands of introduced species. For example,
prior to the construction of the large network of
dams that control the Colorado river, its floodplain
forests were dominated by native cottonwood and
willow species. With dam construction, groundwater
tables have dropped, scouring floods have ceased
and cottonwood and willow have declined - and
been replaced by nearly monospecific stands of the
introduced saltcedar (Tamarix sp.) (Ohmart,
Anderson and Hunter, 1988).
The fragmentation of wildland habitat resulting
from agricultural or urban development has also
affected the spread of introduced species. Urban
forests and parklands represent an increasing
percentage of our remaining near-natural habitats.
Because they are subjected to pollution stresses and
because of their proximity to sites of introduction and
their (often) large ratio of edge to interior habitat,
they are prime habitat for introduced plant or animal
pests which can then spread into less urban habitat.
Gypsy moths, for example, first became established
in an urban forest and subsequently became a major
pest species throughout the eastern United States (see
Liebhold et al. , 1995 for a history of the gypsy moth
outbreak). Outbreaks of introduced fungal pathogens
have also been found to be more common in forest
fragments that are close to urban areas (Castello,
Leopold and Smallidge, 1995).
Invasion and Extinction
A greatly enhanced rate of extinction of species and
of genetically distinct populations is the least
reversible of the many ongoing global environmental
changes (Vitousek, 1994) - and there is good
evidence that biological invasions contribute
substantially to extinction. As of 1991, 44 species of
freshwater fish in the continental United States were
threatened or endangered by the introduction of non-
native fish. Of the 40 species of fish known to have
gone extinct since 1890, 27 were negatively affected
by introduced fish (Wilcove and Bean, 1994).
While most extinctions in which introduced
species are known to have played a major part have
been on islands or in aquatic systems, the potential
for invasion-driven extinctions in continental
systems is substantial. At a global scale, this impact
can be estimated using species - area curves. These
Figure 3: Land clearing and grass invasions can interact in the initiation and maintenance of a grass-fire feedback system
that prevents forest regeneration over large areas of Earth. From D’Antonio and Vitousek (1992).
summarize the relationship between the size of an
area (an island or isolated patch of habitat) and the
number of species it supports. Preston (1960) plotted
the number of species of breeding birds in different
habitats against log of the area supporting them, and
found a linear relationship. Extrapolating that
relationship to the area of Earth’s land surface yields
a total number of bird species that is substantially
less than the actual number. The difference comes
about because areas that are isolated from each other
support wholly different bird faunas - in other
words, because regional distinctiveness begets
global diversity.
Westbrooks and colleagues applied this
approach to calculate directly the potential for
extinction resulting from biological invasion; Wright
(1987) had earlier carried out a similar analysis. For
example, a plot of the number of mammalian species
on each continent versus log area yields a straight
line with r
= .94 ; extending this relationship to the
land area of Earth, a single supercontinent would
support ~2,000 species (Fig. 4). Earth’s continents in
fact support 4,200 species, due to isolation of
distinct faunas in different regions. This analysis
implies that if invasions were so widespread as to
cause a complete breakdown of the biogeographic
barriers separating different regions, a substantial
number of Earth’s mammalian species would
(ultimately) be driven to extinction.
We believe that this analysis is as solid as
estimates of potential extinction rates based on
habitat loss and fragmentation (Wilson, 1992).
Moreover, this approach is supported by
paleobiological evidence. Two or three million years
ago, the Isthmus of Panama connected North and
South America, and allowed a massive exchange of
their biota (at least, of that portion able to survive in
the tropics) (Simpson, 1980). The result was
asymmetrical - while some South American
mammals (notably the opossum) spread and thrived
in North America, many more North American
mammals spread through South America. This
invasion by North American mammals corresponded
with a significant increase in the extinction rate of
South American mammals (Marshall et al. , 1982).
What Can Be Done?
In discussing biological invasions with other
scientists and the public, we run into two major
concerns. The first is a belief that invasions
represent a natural process that has always been
with us; the second is the feeling that the ease of
travel and the increasing global nature of the
economy make it impossible to prevent invasions
for long.
For the first, it is of course true that invasions
(like extinctions) have always been with us. What
differs now is the increased rate of invasions,
resulting from the extraordinary mobility of
humanity and our goods - an increase in the rate of
invasions that is so large as to represent a difference
in kind rather than degree. For example, the
complete insect fauna of the Hawaiian Islands
resulted from a successful colonization (followed by
evolutionary radiation) every 50,000—100,000 years
- but recently, 15—20 insect species per year have
become established there (Beardsley, 1979).
Similarly, detailed paleoecological studies of Eastern
North America indicate that there was one
prehistoric instance (in the past 10,000 years) in
which a tree species (eastern hemlock) declined
precipitously in a pattern consistent with pathogen
attack throughout most of its range (Davis, 1981).
This contrasts with devastation of several American
tree species by pathogens in the past century.
Figure 4: A species/area curve for mammals. The number
of species on a continent is tightly correlated with the size
of that continent - but extrapolating that relationship to the
land area of Earth (reuniting Gondawanaland) yields less
than half the total number of species that actually occur on
these continents. Much of the diversity of mammalian
species globally is due to the isolation of separate biotic
regions. Analysis prepared by A. Launer of the Center for
Conservation Biology, Stanford University.
For the second concern, we have framed the
problem of biological invasions as a fundamental
component of human-caused global change,
important in driving global losses of biological
diversity as well as (in many cases) undesirable from
a purely anthropocentric viewpoint (health, wealth).
As with other types of human-caused global change,
stemming the tide of biological invasions poses a
huge challenge to the ingenuity of humankind. A
large part of the task is convincing our colleagues,
students, and the public that it is a problem worthy
of our best efforts, and giving them sufficient
understanding that they can respond in a positive
way. Government-directed efforts are not going to
work without widespread support from citicens. Our
experience suggests that such citizen support in
beginning to arise in the United States, in the areas
that have been hardest hit by invasions. In Hawaii
and Florida, invasion stories are front-page items in
local newspapers, County Councils have been
known to provide funds for emergency invasive
species control projects to assure protection of
biodiversity, local lifestyles, and tourism.
Several of us attended an international
conference on alien species in Norway in mid-1996
and were encouraged by the high level of concern
accorded the problem in many countries and by a
number of serious efforts being initiated to confront
it at international, national, and local levels. (See
Sandlund, Schei and Viken 1996.) Better legal
frameworks are being sought in several countries.
New Zealand’s Biosecurity Act of 1993 and
Hazardous Substances and New Organisms Act of
1996 are recognized as outstanding examples of
progressive legislation.
The challenge of slowing invasions may prove
to be as rewarding as - but less threatening to
economic growth and lifestyles than - slowing fossil
fuel combustion. Existing national laws and policies
can be enforced and strengthened, and intelligent
new approaches can be devised, given reasonable
public support. Moreover, concerned and informed
citizens can participate personally in recognizing
incipient invaders and preventing them from
spreading. The concept of thinking globally but
acting locally applies extremely well to stopping
invasions. Perhaps with no other form of global
change can educated and dedicated individuals have
such an opportunity to make a lasting difference.
We thank John Randall and Colin Townsend for
suggstions and critical comments on the manuscript,
Alan Launer for carrying out the analysis in Fig. 4,
and Cheryl Nakashima for preparing the manuscript
for publication. John Katzenberger and the Aspen
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... A wide range of studies on ecosystems found that climate change already had or is predicted to have far-reaching effects (Vitousek et al. 1997;Walther et al. 2002). Documented biological responses to climate change include shifts in species ranges (Aronson et al. 2001 andParmesan et al. 2003), phenological patterns (Parmesan et al. 2003;Root et al. 2003), and community structure due to varied climate effects at different trophic levels (Voigt et al. 2003). ...
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Coral reefs are an essential part of the earth’s ecosystem. They are sensitive indicators of the health of marine environments and are important economically, providing people around the globe with food, jobs, coastal protection, and recreational opportunities. Many of the world’s reefs have been severely damaged over the past few decades due to a combination of factors including habitat destruction, land-based pollution, sedimentation, overfishing, vessel groundings, coastal development, disease, and climate change. A few previous researches based on multispectral satellite images with high resolution, climatic and ocean color satellite images with moderately resolution studied the coral reef habitat mapping, land use/land cover, bathymetric mapping, change detection and climate change at Egyptian Red Sea. The studied cities were chosen as an appropriate site for this study for these reasons: 1. Offshore reefs degradation increased at an alarming rate. 2. Fast touristic and industrial developments 3. Increasing of diving and water sports. 4. Absence of a global warming unit. 5. Increasing of overfishing at coral reef areas. 6. Lack of awareness of individuals and their personal behavior. 7. Absence of annually monitoring plans for coral ecosystem. The main objective of this study is to use the remote sensing tools for time series analysis of the climatic parameter and correlation with the hard coral reef cover at the Egyptian Red Sea. The specific objectives of this study are: 1) Change detection analyses of coral reef cover and Land use at the studied cities (Red Sea) during the period of 42 years extending from 1973 to 2015. 2) Annual bleaching observations of hard corals cover from 2013 to 2015 at the four studied cities. 3) To compare the annual sea surface temperature with current coral bleaching observations from 2013 to 2015. 4) Bleaching events of the thermal stress anomalies through the time series analyses of SST from 2002 to 2015. 5) Time series analysis of SST, Chl-a, PAR, Kd, SSH and SSS to investigate weekly, monthly and annual variations means, climatology and anomalies at the studied cities Red Sea (Ras Ghareb, Hurghada, Qusier and Marsa Alam). 6) Correlation and regression between bleaching observations and other climatic parameters. These mentioned goals were achieved by observation data, Landsat images and climatic data, to estimate the change detection, hard coral cover and climate change by using highly resolution Landsat (MSS, TM, ETM+ and OLI_TIRS+) from 1973 – 2015 about 36 images, a daily moderately MODIS-A from 2002 to 2015 and HYCOM from 2009 to 2015 with about more than 20000 images. Based on these images four studied cities with sixteen sites were chosen to make the survey and habitat ground truthing. Landsat images geometrically, radiometrically, atmospherically, water column corrected, unsupervised, supervised classified and change detection was performed using the following software’s, ENVI, ERDAS, CPCe and Arc GIS. The climatic data (MODIS –Aqua) was preprocessed from Level 1A to Level 3 using (SeaDAS) bio-optical algorithm for SST and Chl-a. then all daily climatic data SST, Chl-a, PAR, Kd, SSS and SSH was processed using Matlab programming language (code scripts) to perform the time series analyses equations on the daily satellite images such as the average, climatology, anomalies and standard deviations to calculate the week, month and annual. On the other hand, the SST is a limiting factor for coral reef growth, we also running sone equations on the SST such as HotSpot, DHW, TSA to determine the bleaching events. The SPSS software used to perform the statistical analysis. Field survey was planned to represent the validation of the unsupervised classes and conducted by selecting random points for each habitat in the area of the study around the studied cities. Ground truthing data collection was conducted over a period of 25 days in 2015. About 150 reference points for every studied city were selected for each water classes using GPS points for the image data validation, data validation was gathered using georeferenced photo transects collected at a set distance from the bottom, each photo was logged by Guno Trimble GPS and interpreted using Coral Point Count (CPCe 4.0) software. Locations of the logged photo were labeled based on common benthic habitat cover classes scheme. Bleaching observations were surveyed in the high warm season (early July to late September) from 2013-2015 using SCUBA diving and snorkeling. At each city 108 transect during three years, 36 per year and 9 at each site, random, replicated Line Intercepted Transect, video transects (100 m) and Quadrate (1 m) were recorded on the upper and lower reef slope 2 to 15 m. In addition to, pictures were also taken by underwater housing camera at a set distance from the bottom, each photo was logged by Guno Trimble GPS and interpreted using Coral Point Count (CPCe 4.0) software, for cross reference and species identification. Also, different pictures were taken to record the effects of degradation of coral reefs. The important results of the study: 1) The urbanization total change of the studied citied was gradually increased during 42 years about 4.15 Km2 for Ras Ghareb, 13.4 Km2 for Hurghada, 8.4 Km2 for Qusier and 11.7 Km2 for Marsa Alam. 2) The present study indicated that the total change of shore line landfilling during 42 years about 2.35 Km2 for Ras Ghareb, 3.50 Km2 for Hurghada, 3.58 Km2 for Qusier and 5.18 Km2 for Marsa Alam. While the total change of shore line erosion about 4.69 Km2 for Ras Ghareb, 1.03 Km2 for Hurghada, 2.62 Km2 for Qusier and 4.53 Km2 for Marsa Alam. 3) The total change of coral reef area degradation during 42 years about 11.3 Km2 for Ras Ghareb, 6.21 Km2 for Hurghada, 10.2 Km2 for Qusier and 6.7 Km2 for Marsa Alam. 4) The study indicated that there are a four coral bleaching events occurred during 2003, 2007, 2010 and 2012 at the four studied cities. 5) The present study obtained that the three years of the total observed bleaching was highly positively correlated with sea surface temperature at all studied cities.
... (3) Departamento de Conservación y Educación Ambiental, División Conservación y Manejo, Parque Nacional Los Alerces, Parques Nacionales. [correspondencia:] La introducción de especies se reconoce mundialmente como una de las mayores amenazas a la biodiversidad (Vitousek et al. 1997;Mack et al. 2000;Simberloff et al. 2005). El visón americano Neogale vison (Schreber, 1777) representa un ejemplo per- ...
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The American mink Neogale vison, introduced in Argentina since the beginnings of the XX century, is known locally as a mayor threat to biodiversity. Within the existing control methods, eradication and population control are implemented. The latter has been carried out in Los Alerces National Park since 2019, with live-capture traps and calculation of costs. Sixteen minks were captured, with different costs according mostly to the difficulty of access. We propose to continue the management to assess its efficiency and to include the costs in future programs
The absences of large carnivores from many ecosystems, human‐induced landscape changes, and resource supplementation have been theorized to increase the abundance of small carnivore species around the world. Overabundant and/or unconstrained small carnivores can have significant effects on specific prey species that, in some cases, can cascade through entire ecosystems. Here, we review the effects of small carnivores on threatened species. We focus on four well‐studied families (Procyonidae, Mephitidae, Mustelidae, and Herpestidae) and emphasize that this is a global conservation issue with consequences for biodiversity. We review and compare the impacts that small carnivores can have on a variety of prey taxa including small mammals, nesting avian and reptilian species, and rare invertebrates. We differentiate between native and exotic small carnivores because this is often an important distinction in terms of the impact severity and range of effects. In addition to direct lethal effects (i.e. predation), small carnivores can also impact threatened species as disease vectors and through competition or overexploitation, which can disrupt communities via ecological release or extinction. Furthermore, we explore other case studies in which small carnivores have had positive effects on threatened species and discuss studies that reveal other taxa responsible for exerting stronger negative effects on threatened prey. We offer some concluding remarks about global small carnivore conservation and emphasize the need for decision‐analytic approaches and robust analyses that can improve our assessment of how populations of threatened species can be affected. To date, indirect effects are especially difficult to measure in the field and many studies have provided only anecdotal or correlative results, signalling a need for improving our scientific methodologies and management approaches.
Successful species introductions are not homogeneously distributed over the globe, which points to the need to understand why some have succeeded, yet others failed. We summarized information on small carnivore introductions worldwide and assessed whether introduction outcomes (success or failure) supported one or more of the following hypotheses: climate‐matching, propagule pressure, inherent superiority, island susceptibility and Darwin's naturalization hypotheses. Using the literature, we summarized: number of individuals released, mean body size, mean litter size, consumer type, latitude difference, ecoregions difference, congener presence, and mainland or island release. We generated generalized linear models and ranked them using Akaike's Information Criterion and Akaike's weights. We identified 253 documented introduction events of 24 species from five families, with two thirds of them involving the northern raccoon, Procyon lotor , the American mink, Neovison vison , and the small Indian mongoose, Urva [= Herpestes] auropunctata . Overall introduction success was high, with a success rate > 70% for four of the five represented families. We found support for climate‐matching, inherent superiority, and Darwin's naturalization hypotheses. Likelihood of success increased with matching climatic conditions that allow survival, a greater body size together with smaller litter size, a carnivorous diet, and the absence of congeners in the area of introduction. Islands were not more susceptible than the mainland, and the number of individuals introduced did not influence success. As biological invasions become increasingly widespread, understanding the biological and environmental factors affecting introduction success is important for conservation and management.
Riparian zones are among the most threatened habitats due to invasive species. Traditionally, studies on invasive alien plants (IAPs) of riparian zones have focused on their ecological implications. Studies assessing the impacts of IAPs on ecosystem services (ES) and ecosystem disservices (EDS) through people's perceptions are lacking in this habitat. This study was conducted in the riparian zone of the Ganga River in Varanasi, India. We used a questionnaire‐based survey to assess people's perceptions of IAPs and their potential to provide ES and EDS. Results indicated that residents were aware of IAPs and perceived these both positively (providing ES) and negatively (providing EDS). Soil stabilization and loss of native plant diversity were the most commonly perceived ES and EDS respectively. Socio‐demographic variables such as age and education shape people's perceptions of IAPs. Younger and educated respondents were more likely to perceive IAPs negatively. Respondents most commonly used fire as a practice to manage IAPs. They placed Parthenium hysterophorus L. as the highest priority based on its impacts and management needs. This study highlights the need to incorporate public perceptions into the policy, planning, and management of IAPs, particularly in the riparian zone.
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The invasive species are of global concern, and the Invasive American Bullfrog (IAB; Lithobates catesbeianus) is one of the worst invasive amphibian species worldwide. Like other countries, South Korea is also facing challenges from IAB. Although many studies indicated impacts of IAB on native anurans in Korea, the actual risk at the specific level is yet to evaluate. Considering the putative invasiveness of IAB, it is hypothesized that any species with the possibility of physical contact or habitat sharing with them, will have a potential risk. Thus, we estimated and observed their home range, preferred habitats, morphology, behavior, and ecology. Then, comparing with existing knowledge, we assessed risks to the native anurans. We found a home range of 3474.2 ± 5872.5 m² and identified three types of habitats for IAB. The analyses showed at least 84% of native anurans (frogs and toads) were at moderate to extreme risks, which included all frogs but only 33% of toads. Finally, we recommended immediate actions to conserve the native anurans based on our results. As this study is the first initiative to assess the specific risk level from the invasiveness of L. catesbeianus, it will help the managers to set conservation priorities and strategies.
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Invasive plants often modify soil biotic communities through changes in soil physicochemical characteristics or the amount and/or quality of litter inputs. We assessed the impacts of Gunnera tinctoria invasions on soil and the earthworm community, on Achill Island, Co. Mayo, Ireland. We compared replicated (n = 5) areas invaded by G. tinctoria with uninvaded semi-natural grasslands, as well as with areas subjected to mechanical removal or herbicide treatment. Modifications in physiochemical properties included lower soil temperatures and higher soil pH during the summer in invaded areas, yet little effect on C and N stocks, or soil moisture. Marked differences in litter were observed, however, with invaded areas having c. 20-fold higher (above-ground) litter input than uninvaded ones, as well as lower C:N ratio (17 vs. 29). This was associated with a significantly higher overall abundance and biomass of earthworms in invaded plots (375 individuals m –2 , 115 g biomass m –2 ), compared to the uninvaded control (130 individuals m –2 , 45 g biomass m –2 ), with removal treatments having intermediate values. Earthworm communities comprised 10 species, typical for Irish grasslands, dominated by the common endogeic species Allolobophora chlorotica, Aporrectodea caliginosa and Aporrectodea rosea . Both earthworm species richness and Shannon diversity were significantly higher in invaded areas, but only in spring samples. Based on this new information, plant invaders may increase the abundance and diversity of earthworms, mainly due to much larger litter inputs, increased soil pH and possibly lower soil temperatures in the summer.typical of Irish grasslands
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Nutria (Myocastor coypus) were introduced to the eastern shore of Chesapeake Bay, USA in the 1940s. They reached peak densities in the late 1990s, causing massive wetland loss. Beginning in 2002, a systematic plan to eradicate nutria from the 1.7M ha Delmarva Peninsula was implemented. Since that time the nutria population has been effectively reduced, and no nutria have been detected since May 2015. A lack of detection does not equate with complete absence. We address the following three questions. (1) What is the expected probability of nutria eradication from the Delmarva Peninsula as of the end of 2020? (2) If the probability of eradication is below the management target of 0.95, how much more surveillance is required? (3) How sensitive is the estimated probability of eradication to varying levels of public surveillance and modelled population growth rates? These questions were addressed by employing a stochastic spatially-explicit surveillance model that uses data in which no nutria were detected to quantify the probability of complete absence (PoA) over the entire Delmarva Peninsula. We applied an analytical framework that decomposes the spatial risk of survivors and data into management zones, and took advantage of low-cost public reporting of nutria sightings. Active surveillance by the eradication program included detector dog and tracker surveys, shoreline surveys, detection with ground and water platforms (with hair snares), and camera traps. Results showed that the PoA increased with time and surveillance from a beginning PoA in May 2015 of 0.01 to a mean of 0.75 at the end of 2020. This indicates that the PoA on the Delmarva was well below the target threshold of 0.95 for declaring eradication success. However, given continued surveillance without detection, a PoA of 0.95 would be achieved by June 2022. This analysis provides an objective mechanism to align the expectations of policy makers, managers and the public on when eradication of nutria from the entire Delmarva Peninsula should be declared successful.
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This book tends to investigate various aspects of freshwater non-native fish species in six chapters. In the first chapter, definitions and introductory concepts in this field are discussed. Undoubtedly, one of the most critical aspects of investigating these species’ impacts is recognizing their ecological and socio-economic consequences and is the second chapters’ central objective. In this chapter, several examples of the conducted research on the ecological concepts of alien fishes have been investigated. For sure, dealing with these species will only be possible by setting laws, legislation, and regulations. The third chapter outlines the existing international laws and regulations and national laws and regulations in the United States of America, New Zealand, Germany, Italy, and Iran. Risk assessment protocols are one of the most significant concerns for non-native species, and species distribution modeling is embedded as a tool into them. All of them are discussed in chapter four. Moreover, there have been some efforts to identify different exotic fish species, assess their distribution in Iran, and determine their biological characteristics. Now, the main concern is how to manage these species so that there could be low impacts on natural ecosystems. Chapter five is the subject of the discussion and investigates the different methods of controlling non-native fishes introduced in freshwater ecosystems. Finally, the last chapter is devoted to non-native fishes introduced in the freshwater ecosystems of Iran. In this chapter, different characteristics of these species include taxonomy, identification and morphological characteristics, distribution in the world and Iran, various biological aspects such as habitat, nutrition and reproduction, ecological and socio-economic impacts, and their current status in the country are discussed. An important feature of this chapter is the review of studies on various aspects of these species in the country. ماهیان غیربومی، گونه‌هایی هستند که تصادفی یا عامدانه توسط انسان به مناطقی در بیرون از دامنه طبیعی پراکنش خود وارد شده و در آنجا زیست می‌کنند. برخی از این گونه‌ها به دلیل توانایی‌هایی مانند قدرت تولیدمثل بالا، همه‌‌چیزخواری و تحمل دامنه وسیعی از شرایط محیطی، وقتی به محیط جدیدی وارد می‌شوند باعث ایجاد تغییرات و پیامدهای اکولوژیک گسترده شده و با به مخاطره انداختن گونههای بومی و ایجاد خسارتهای اقتصادی‌-اجتماعی فراوان، به گونه آفت در آن منطقه تبدیل می‌شوند؛ در این صورت به آن‌ها، گونه‌های غیربومی مهاجم گفته می‌شود. رودخانه‌ها و تالاب‌های ایران با عواملی مانند خشکسالی و آلودگی رو‌به‌رو هستند و علاوه بر آن، ورود گونه‌های غیربومی ازجمله ماهیان نیز شرایط را برای این اکوسیستم‌ها بحرانی‌تر کرده است. کتاب حاضر با هدف شناخت هر چه بیشتر ماهیان غیربومی اکوسیستم‌های آب شیرین، بررسی اثرات اکولوژیکی و اقتصادی‌-‌اجتماعی و کنترل و مدیریت این گونه‌ها، به رشته تحریر درآمده است.
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Exotic invasive plants are considered major threats to biodiversity globally; however, our understanding of the long-term dynamics of invasion remains limited. Over time, invasive plants can accumulate novel pathogens that may be capable of causing population declines because invaders have a greater chance of encountering such pathogens as they spread, and native pathogens can adapt to use invasive plants as a resource over time. However, reports typically focus on individual species and a framework capable of predicting pathogen vulnerability in plant invaders is not available. Pathogen resistance and tolerance may be related to plant traits, which we suggest can contribute to a framework for understanding and predicting the vulnerability of invasive plants to novel pathogens. We reviewed which traits of invasive plant species can aid in understanding the long-term dynamics of invasions due to associations between such traits and the vulnerability to novel-pathogen attack. We then extracted data from the literature to which we applied a multivariate model to associate plant traits with pathogen response to predict pathogen vulnerability of invasive plants. Finally, we provide directions for future research. There were 31 published tests of novel pathogen effects on invasive plant species in the introduced range. We found that together plant height, leaf nitrogen content, seed mass, specific leaf area, and leaf dry matter content can be useful in determining pathogen effects when included as part of a broader framework. We propose that more trait data for invasive plants are now needed to keep refining the predictive capacity of the proposed framework. Considering the emergence of trait-based approaches and comprehensive databases, advances in our understanding of invasive plant-pathogen interactions can lead to breakthroughs both at fundamental and management decision-making levels.
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The first comprehensive botanical field guide to the San Juan Islands situated in the Cascadia rain shadow ecotype. Includes an annotated checklist of all vascular plants recorded for San Juan County, Washington. On these hilly, rocky islands, at the northern end of the California Mediterranean zone are prairie, savanna, chaparral, soft and soft sage communities [and beauty beyond measure]. The field guide, illustrates and describes 190 species of wildflowers, shrubs and trees [& exotics] The guide celebrates common, unusual an intriguing species [cactus, manzanita, oak, juniper, madrone]. This handy trail companion is useful in nearby arid insularity including the Gulf Islands and southern Vancouver Island, Canada. Structured around plant communities[meadows, rocky outcrops, woodlands, maritime, fresh water, disturbed sites & Mt. Constitution] the species are listed by type, then color. The text also includes information on how the climate, topography and geology of the Islands affect plant life. Of special interest to serious students is the only annotated checklist in print of all vascular plants recorded for San Juan County.
Ninety-three species of vascular plants are recorded from a 16 sq. mile coastal lowland on the northern coast of Devon Island, Northwest Territories. The following taxa are apparently new records for Devon Island: Cystopteris fragilis, Woodsia alpina, Equisetum variegatum, Poa alpigena, Carex amblyorhyncha, Draba oblongata, Saxifraga tenuis, Epilobium arcticum, Hippuris vulgaris, Pedicularis lanata, Puccinellia vaginata var. paradoxa. These additions bring the total known flora of Devon Island to 115 species. The Truelove flora is part of the High Arctic biogeographic element of the Canadian Arctic Archipelago. However, a distinct element of species of more southerly distribution is present probably due to the moderating influence of the lowland environment.