Biotic Invasions: Causes, Epidemiology, Global Consequences, and Control

Abstract

Biotic invaders are species that establish a new range in which they proliferate, spread, and persist to the detriment of the environment. They are the most important ecological outcomes from the unprecedented alterations in the distribution of the earth’s biota brought about largely through human transport and commerce. In a world without borders, few if any areas remain sheltered from these immigrations.

The fate of immigrants is decidedly mixed. Few survive the hazards of chronic and stochastic forces, and only a small fraction become naturalized. In turn, some naturalized species do become invasive. There are several potential reasons why some immigrant species prosper: some escape from the constraints of their native predators or parasites; others are aided by human-caused disturbance that disrupts native
communities. Ironically, many biotic invasions are apparently facilitated by cultivation and husbandry, unintentional actions that foster immigrant populations until they are self-perpetuating and uncontrollable. Whatever the cause, biotic invaders can in many cases inflict enormous environmental damage:
(1) Animal invaders can cause extinctions of vulnerable native species through predation, grazing, competition, and habitat alteration.
(2) Plant invaders can completely alter the fire regime, nutrient cycling, hydrology, and energy budgets in a native ecosystem and can greatly diminish the abundance or survival of native species.
(3) In agriculture, the principal pests of temperate crops are nonindigenous, and the combined expenses of pest control and crop losses constitute an onerous ‘‘tax’’ on food, fiber, and forage production.
(4) The global cost of virulent plant and animal diseases caused by parasites transported to new ranges and presented with susceptible new hosts is currently incalculable.

Identifying future invaders and taking effective steps to prevent their dispersal and establishment constitutes an enormous challenge to both conservation and international commerce. Detection and management when exclusion fails have proved daunting for varied reasons:
(1) Efforts to identify general attributes of future invaders have often been inconclusive.
(2) Predicting susceptible locales for future invasions seems even more problematic, given the enormous differences in the rates of arrival among potential invaders.
(3) Eradication of an established invader is rare, and control efforts vary enormously in their efficacy. Successful control, however, depends more on commitment and continuing diligence than on the efficacy of specific tools themselves.
(4) Control of biotic invasions is most effective when it employs a long-term, ecosystemwide strategy rather than a tactical approach focused on battling individual invaders.
(5) Prevention of invasions is much less costly than post-entry control. Revamping national and international quarantine laws by adopting a ‘‘guilty until proven innocent’’ approach would be a productive first step.

Failure to address the issue of biotic invasions could effectively result in severe global consequences, including wholesale loss of agricultural, forestry, and fishery resources in some regions, disruption of the ecological processes that supply natural services on which human enterprise depends, and the creation of homogeneous, impoverished ecosystems composed of cosmopolitan species. Given their current scale, biotic invasions have taken their place alongside human-driven atmospheric and oceanic alterations as major agents of global change. Left unchecked, they will influence these other forces in profound but still unpredictable ways.

Full text

689
Ecological Applications,
10(3), 2000, pp. 689–710
䉷
2000 by the Ecological Society of America
T
ECHNICAL
R
EPORT
Issues in Ecology
BIOTIC INVASIONS: CAUSES, EPIDEMIOLOGY, GLOBAL
CONSEQUENCES, AND CONTROL
R
ICHARD
N. M
ACK
,
1
D
ANIEL
S
IMBERLOFF
,
2
W. M
ARK
L
ONSDALE
,
3
H
ARRY
E
VANS
,
4
M
ICHAEL
C
LOUT
,
5
AND
F
AKHRI
A. B
AZZAZ
6
1
School of Biological Sciences, Washington State University, Pullman, Washington 99164 USA
2
Department of Ecology and Evolutionary Biology, University of Tennessee, Knoxville, Tennessee 37996-1610 USA
3
CSIRO Entomology and CRC for Weed Management Systems, GPO Box 1700, Canberra, ACT 2601, Australia
4
CABI BIOSCIENCE, UK Centre (Ascot), Silwood Park, Buckhurst Road, Ascot, Berkshire SL5 7TA, UK
5
School of Biological Sciences, University of Auckland, Private Bag 92019, Auckland, New Zealand
6
Biological Laboratories, Harvard University, 16 Divinity Avenue, Cambridge, Massachusetts 02138 USA
Abstract.
Biotic invaders are species that establish a new range in which they proliferate, spread, and
persist to the detriment of the environment. They are the most important ecological outcomes from the
unprecedented alterations in the distribution of the earth’s biota brought about largely through human
transport and commerce. In a world without borders, few if any areas remain sheltered from these im-
migrations.
The fate of immigrants is decidedly mixed. Few survive the hazards of chronic and stochastic forces,
and only a small fraction become naturalized. In turn, some naturalized species do become invasive. There
are several potential reasons why some immigrant species prosper: some escape from the constraints of
their native predators or parasites; others are aided by human-caused disturbance that disrupts native
communities. Ironically, many biotic invasions are apparently facilitated by cultivation and husbandry,
unintentional actions that foster immigrant populations until they are self-perpetuating and uncontrollable.
Whatever the cause, biotic invaders can in many cases inflict enormous environmental damage: (1)Animal
invaders can cause extinctions of vulnerable native species through predation, grazing, competition, and
habitat alteration. (2) Plant invaders can completely alter the fire regime, nutrient cycling, hydrology, and
energy budgets in a native ecosystem and can greatly diminish the abundance or survival of native species.
(3) In agriculture, the principal pests of temperate crops are nonindigenous, and the combined expenses
of pest control and crop losses constitute an onerous ‘‘tax’’ on food, fiber, and forage production. (4) The
global cost of virulent plant and animal diseases caused by parasites transported to new ranges and presented
with susceptible new hosts is currently incalculable.
Identifying future invaders and taking effective steps to prevent their dispersal and establishment con-
stitutes an enormous challenge to both conservation and international commerce. Detection and management
when exclusion fails have proved daunting for varied reasons: (1) Efforts to identify general attributes of
future invaders have often been inconclusive. (2) Predicting susceptible locales for future invasions seems
even more problematic, given the enormous differences in the rates of arrival among potential invaders. (3)
Eradication of an established invader is rare, and control efforts vary enormously in their efficacy. Successful
control, however, depends more on commitment and continuing diligence than on the efficacy of specific
tools themselves. (4) Control of biotic invasions is most effective when it employs a long-term, ecosystem-
wide strategy rather than a tactical approach focused on battling individual invaders. (5) Prevention of
invasions is much less costly than post-entry control. Revamping national and international quarantine laws
by adopting a ‘‘guilty until proven innocent’’ approach would be a productive first step.
Failure to address the issue of biotic invasions could effectively result in severe global consequences,
including wholesale loss of agricultural, forestry, and fishery resources in some regions, disruption of the
ecological processes that supply natural services on which human enterprise depends, and the creation of
homogeneous, impoverished ecosystems composed of cosmopolitan species. Given their current scale,
biotic invasions have taken their place alongside human-driven atmospheric and oceanic alterations as
major agents of global change. Left unchecked, they will influence these other forces in profound but still
unpredictable ways.
Key words: alien species; biological control; biotic invaders; eradication; global change; immigration; invasion;
naturalization; nonindigenous; pests; weeds.
Manuscript received 4 November 1999; accepted 4 November 1999.
Reprints of this 22-page report are available for $3.25 each. Prepayment is required. Order reprints from the Ecological
Society of America, Attention: Reprint Department, 1707 H Street, N.W., Suite 400, Washington, DC 20006.

690
R. N. MACK ET AL.
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Vol. 10, No. 3
I
NTRODUCTION
Biotic invasions can occur when organisms are trans-
ported to new, often distant, ranges where their de-
scendants proliferate, spread, and persist (sensu Elton
1958). In a strict sense, invasions are neither novel nor
exclusively human-driven phenomena. But the geo-
graphic scope, frequency, and the number of species
involved have grown enormously as a direct conse-
quence of expanding transport and commerce (Wells
et al. 1986, di Castri 1989). Few habitats on earth re-
main free of species introduced by humans (e.g., Surt-
sey [Fridriksson and Magnusson 1992]); far fewer are
so remote or display such unique environments that
they can be considered immune from this dispersal
(e.g., locales above 80
⬚
latitude). The number of species
that have entered new ranges through human agency
has increased by orders of magnitude in the past 500
years, and especially in the past 200 years (di Castri
1989), thanks to expanding human migrations and com-
merce. Nonindigenous species represent an array of
taxonomic categories and geographic origins that defy
any ready classification (Crawley 1987, Long 1981,
Holm et al. 1997).
The adverse consequences of biotic invasions vary
enormously. At one extreme, the mere presence of non-
indigenous species in a conservation reserve could be
deemed detrimental. Invaders can alter fundamental
ecological properties such as the dominant species in
a community and an ecosystem’s physical features, nu-
trient cycling, and plant productivity (Bertness 1984,
Vitousek 1990). The aggregate effects of human-
caused invasions threaten efforts to conserve biodi-
versity (Walker and Steffen 1997), maintain productive
agricultural systems (U.S. Congress 1993), sustain
functioning natural ecosystems (D’Antonio and Vitou-
sek 1992, Vitousek et al. 1996), and also protect human
health (Soule´1992). However, as a practical rather than
conceptual restriction, we do not deal here with the
invasive parasites of humans. We outline below the
epidemiology of invasions, hypotheses on the causes
of invasions, the environmental and economic toll they
take, and tools and strategies for reducing this toll.
T
HE
E
PIDEMIOLOGY OF
I
NVASIONS
Biotic invasions constitute only one outcome—in-
deed, the least likely outcome—of a multistage process
that begins when organisms are transported from their
native ranges to new locales. These immigrant organ-
isms and their descendants have been referred to as
‘‘alien,’’ ‘‘adventive,’’ ‘‘exotic,’’ ‘‘neophytes’’ (in the
case of plants),‘‘introduced,’’ and most recently, ‘‘non-
indigenous’’ (Salisbury 1961, Mack 1985, Baker 1986,
U.S. Congress 1993). These terms have been used in-
terchangeably and often without careful distinction. We
will employ ‘‘nonindigenous’’ as the most general term
for immigrant species, especially where their invasive
status is uncertain.
The fates of these organisms vary vastly. First, many,
if not most, perish en route to a new locale (e.g., prop-
agules suspended in marine ballast water). If they suc-
ceed in reaching a new site, immigrants are likely to
be destroyed quickly by a multitude of physical or bi-
otic agents (Kruger et al. 1986, Mack 1995). It is almost
impossible to obtain data quantifying the number of
species that are actually dispersed from their native
ranges, the number of emigrants that subsequently per-
ish, and the number of arrivals. But based on the num-
ber of species that have been collected only once far
beyond their native range (e.g., Thellung 1911–1912,
Ridley 1930, Carlton and Geller 1993), the local ex-
tinction of immigrants soon after their arrival must be
enormous.
Despite such wholesale destruction either in transit
or soon after arrival, immigrants occasionally survive
to reproduce. Even then, their descendants may survive
for only a few generations before going extinct locally.
Again, however, some small fraction of these immi-
grant species do persist and become naturalized. At that
point, their persistence does not depend on recurring,
frequent re-immigration from the native range (Lousley
1953). These populations’minimum size, number, and
areal extent have no commonly identified thresholds,
although a greater number and frequency of new ar-
rivals do raise the probability that a species will es-
tablish permanently (Veltman et al. 1996). Among the
naturalized species that persist after this extremely se-
vere reductive process, a few will go on to become
invaders.
A comparison is often made between epidemics
caused by parasites and all other biotic invasions be-
cause many important factors in disease epidemiology
are common to all invasions. These factors include
identity of the vectors, the parasite’s minimum viable
population size, the time course and character of its
population growth and spread, the fate of interacting
species in the new range (including their coevolution),
and mitigating (or exacerbating) effects of the new en-
vironment. All have direct parallels in studying inva-
sions, regardless of the species (Mack 1985). Below
we explore the epidemiology and the underlying mech-
anisms that allow some species to become invaders.
Humans as dispersal agents of potential invaders
Humans have served as both accidental and delib-
erate dispersal agents for millennia, and the dramatic
increase in plant, animal, and microbial immigrations
worldwide roughly tracks the rise in human transport
and commerce (di Castri 1989, U.S. Congress 1993).
Ancient human migrations and trade led to the early
spread of some domesticated species such as cereals,
dates, rice, cattle, and fowl, along with the inadvertent
spread of their parasites (diCastri 1989, Zohary and
Hopf 1993). Beginning around 1500, Europeans trans-
ported Old World species to their new settlements in

June 2000 691
BIOTIC INVASIONS
the Western Hemisphere and elsewhere. Themanifests
from Columbus’second and subsequent voyages, for
instance, indicate deliberate transport of species re-
garded as potential crops and livestock (Crosby 1972).
Global commerce has grown meteorically since the late
15th century, as indexed by the rise in shipping tonnage
(Fayle 1933); this growth has provided an opportunity
for a corresponding growth in biotic invasions. Given
the magnitude of this transport and subsequent natu-
ralizations of species in new lands, biotic invasions can
be viewed as predominantly post-Columbian events.
The human-driven movement of organisms over the
past 200 to 500 years, deliberate and accidental, un-
doubtedly dwarfs in scope, frequency, and impact the
movement of organisms by natural forces in any 500-
year period in the earth’s history. Such massive alter-
ation in species’ranges rivals the changes wrought by
continental glaciation and deglaciation cycles of past
ice ages, despite the fact that these human-driven range
shifts have occurred over much less time (e.g., Semken
1983).
The proportion of various types of organisms that
have invaded as a result of accidental vs. deliberate
movement clearly varies among taxonomic groups
(Moyle 1986, Heywood 1989). Few, if any, invasive
microorganisms have been deliberately introduced. De-
liberate microbial introductions have instead most
commonly involved yeasts for fermentation or mutu-
alists, such as mycorrhizal fungi (Read et al. 1992).
Among insects, some deliberate introductions have had
adverse consequences, including bumble bees in New
Zealand (Thompson 1922), but the majority of invasive
insects have probably been accidentally introduced. In-
troductions of marine invertebrates probably mirror in-
sects. A few deliberate introductions have been made
(e.g., the Pacific oyster [
Crassostrea gigas
] imported
from Japan to Washington State), but a growing number
of invaders such as the zebra mussel (
Dreissena po-
lymorpha
) have arrived as accidental contaminants in
ship ballast (Carlton and Geller 1993). In contrast, most
invasive vertebrates, principally fish, mammals, and
birds, have been deliberately introduced. Some of the
worst vertebrate invaders, however, have been spread
accidentally:
Rattus rattus, Rattus norvegicus,
the
brown tree snake (
Boiga irregularis
), the sea lamprey
(
Petromyzon marinus
) (Brown 1989). Some invasive
plants have been accidentally introduced as contami-
nants among crop seeds and other cargo (e.g.,
Par-
thenium hysterophorus, Rottboellia cochinchinensis
)
(Huelma et al. 1996). However, many, if not most, plant
invaders in the United States have been deliberately
introduced, including some of the worst pests:
Eich-
hornia crassipes, Sorghum halapense, Melaleuca quin-
quenervia,
and
Tamarix
spp. (R. N. Mack,
unpublished
data
).
The prominence of deliberately introduced species
that later become biotic invaders emphasizes that not
all pests arrive unheralded and inconspicuously; many
are the product of deliberate but disastrously flawed
human forethought (Fig. 1).
The transformation from immigrant to invader
The progression from immigrant to invader often in-
volves a delay or lag phase, followed by a phase of
rapid exponential increase that continues until the spe-
cies reaches the bounds of its new range and its pop-
ulation growth rate slackens (Mack 1985, Cousens and
Mortimer 1995; Fig. 2). This simplified scenario has
many variants. First, some invasions such as those by
Africanized bees in the Americas and zebra mussels in
the Great Lakes may go through only a brief lag phase,
or none at all (Crooks and Soule 1996). On the other
hand, many immigrant species do not become abundant
and widespread for decades, during which time they
may remain inconspicuous. Perhaps the most spectac-
ular example involves the fungus,
Entomophaga mai-
maiga,
introduced to the United States for control of
the gypsy moth (
Lymantria dispar
). After effectively
disappearing for 79 years, it made a reappearance in
1989 and is now inflicting substantial mortality on the
moth in the northeastern United States (Hajek et al.
1995). Brazilian pepper (
Schinus terebinthifolius
) was
introduced to Florida in the 19th century but did not
become widely noticeable until the early 1960s. It is
now established on
⬎
280 000 ha in south Florida, often
in dense stands that exclude all other vegetation
(Schmitz et al. 1997).
During the lag phase, it can be difficult to distinguish
doomed populations from future invaders (Cousens and
Mortimer 1995). Most extinctions of immigrant pop-
ulations occur during the lag phase, yet the dynamics
of such a population are often statistically indistin-
guishable from those of a future invader, which is grow-
ing slowly but inexorably larger. This similarity in the
size and range of these populations frustrates attempts
to predict future invaders while they are few in numbers
and presumably controllable.
Whether most invasions endure lag phases, and why
they occur, remain conjectural (Williamson 1996). Any
lag phase in the population growth and range expansion
for a potential invasion most likely results from several
forces and factors operating singly or in combination:
1)
Limits on the detection of a population’s growth.
A lag could be perceived simply through the inability
to detect still small and isolated but nonetheless grow-
ing populations in a new range (Crooks and Soule
1996).
2)
The number and arrangement of infestations of
immigrants.
Usually an invasion will proceed fastest
from among many small, widely separated infestations
or foci compared with a single larger one (Moody and
Mack 1988). Unless many foci arise soon after im-
migration, an unlikely event, the lag phase could be

692
R. N. MACK ET AL.
Ecological Applications
Vol. 10, No. 3
F
IG
. 1. Some invaders have widely separated new ranges, the products of repeated human dispersal and cultivation. For
example, the shrub
Lantana camara
was carried transoceanically throughout the 19th and early 20th century to many
subtropical and tropical locales where it has proliferated. Years refer to dates of introduction in widely separated locales
(Cronk and Fuller 1995).
F
IG
. 2. Many invaders occupy new ranges at an accel-
erating rate with pronounced ‘‘lag’’ and ‘‘log’’ phases of pro-
liferation and spread. Terrestrial plant invasions most com-
monly illustrate this pattern (e.g., the spread of
Opuntia au-
rantiaca
in South Africa [Moran and Zimmerman 1991]).
the result of an initial limitation in widely separated
foci.
3)
Natural selection among rare or newly created
genotypes adapted to the new range.
Strong selection
in a new range may simply destroy all but the few pre-
adapted genotypes, thus accounting in part for the very
high extinction rate among immigrant populations. Al-
ternatively, the lag phase could reflect the time for
emergence of new genotypes through outcrossing
among immigrants, although proof of this explanation
has proven elusive (Baker 1974, Crooks and Soule
1996).
4)
The vagaries of environmental forces.
The order,
timing, and intensity of environmental hazards are crit-
ical for all populations, but the consequences of con-
secutive periods of high mortality are most severe
among small populations. Thus, a small immigrant pop-
ulation could persist or perish largely as a consequence
of a lottery-like array of forces across time and gen-
erations: i.e., whether the first years in the new range
are benign or severe; whether environmental forces
combine to destroy breeding-age individuals as well as
their offspring. Immigrant populations may also be so
small that demographic stochasticity, simply the odds
that few, if any, reproductive individuals will produce
offspring as influenced by endogenous forces, can also
be important (Simberloff 1988). Much of the downward
spiral seen in the size of immigrant populations could
be attributed simply to the single and collective action
of these two forces (Mack 1995).
Clearly, some populations overcome these long odds
and grow to a threshold size such that extinction from
chance events, demographic or environmental, be-
comes unlikely (Crawley 1989). One great irony about
biotic invasions is that humans, through cultivation and
husbandry, often enhance the likelihood that nonindig-
enous populations will reach this threshold and become
established. This husbandry includes activities that
protect small, vulnerable populations from environ-
mental hazards such as drought, flooding, frost, para-
sites, grazers, and competitors. With prolonged human
effort, such crops, flocks, or herds can grow to a size
that is not in imminent danger of extinction. In fact,

June 2000 693
BIOTIC INVASIONS
the population may no longer require cultivation to
persist (Lousley 1953). At this point, the population
has become naturalized and may eventually become
invasive. Thus, humans act to increase the scope and
frequency of invasions by serving as both effective
dispersal agents and also protectors for immigrant pop-
ulations, helping favored nonindigenous species beat
the odds that defeat most immigrants in a new range
(Veltman et al. 1996).
At some point, whether after years or decades, pop-
ulations of a future invader may proceed into a phase
of rapid and accelerating growth, in both numbers and
areal spread (Fig. 2). This eruption often occurs rapidly,
and there are many first-hand accounts of invasions that
proceeded through this phase despite the concerted ef-
forts of the public to control them (Thompson 1922,
Elton 1958, Mack 1981). Eventually, an invasion
reaches its environmental and geographic limits in the
new range, and its populations persist but do not ex-
pand.
I
DENTIFYING
F
UTURE
I
NVADERS AND
V
ULNERABLE
C
OMMUNITIES
Identifying future invaders and predicting their like-
ly sites of invasion are of immense scientific and prac-
tical interest. Learning to identify invaders in advance
would tell us a great deal about how life history traits
evolve (Crawley et al. 1996) and how biotic commu-
nities are assembled (Lawton 1987). In practical terms,
it could reveal the most effective means to prevent
future invasions (Reichard and Hamilton 1997). Cur-
rent hypotheses or generalizations about traits that dis-
tinguish both successful invaders and vulnerable com-
munities all concern some extraordinary attributes or
circumstances of the species or communities. And all
are based on retrospective explanations for past inva-
sions. Evaluation of these generalizations has been dif-
ficult because they rely on post hoc observation, cor-
relation, and classification rather than experimentation
(Ehrlich 1986, Cronk and Fuller 1995, Holm et al.
1997). Probably no invasions (except some invasions
of human parasites) have been tracked closely and
quantified from their inception. Furthermore, predic-
tions of future invaders and vulnerable communities
are inextricably linked (Crawley 1987). Did a com-
munity sustain an invasion because it is intrinsically
vulnerable or because the invader possesses extraor-
dinary attributes? Do communities with few current
invaders possess intrinsic resistance or have they been
reached so far by only weak immigrants? This second
issue is confounded by the enormous bias of the op-
portunity for immigration among different locales
(Simberloff 1986, Lonsdale 1999).
Attributes of invaders
Biologists have long sought to explain why so few
naturalized species become invaders (Henslow 1879,
as cited in Gray 1879). Intriguingly, some specieshave
invaded several widely separated points on the planet
(e.g.,
Eichhornia crassipes, Imperata cylindrica, Par-
thenium hysterophorus, Avena fatua, Sturnus vulgaris,
Rattus rattus, Lantana camara,
Long 1981, Brown
1989, Holm et al. 1997), which is the ecological equiv-
alent of winning repeatedly in a high-stakes lottery.
Such repeat offenders, or winners, have sparked the
obvious question: do they and other successful invasive
species share attributes that significantly raise their
odds for proliferation in a new range (Ehrlich 1986,
Rejmanek and Richardson 1996)?
Many attempts have been made to construct lists of
common traits shared by successful invaders (e.g.,
Wodzicki 1965, Roy 1990). The hope behind such ef-
forts is clear: detect a broad list of traits that, for ex-
ample, invading insects, aquatic vascular plants, or
birds share as a group, then perhaps the identity of
future invaders could be predicted from these taxo-
nomic groups. Some invaders do appear to have traits
in common, but so far such lists are generally appli-
cable for only a small group of species, and exceptions
abound (cf. Crawley 1987, Rejmanek and Richardson
1996).
Relatives of invaders, particularly congeners, seem
to be obvious candidates for possession of shared in-
vasive attributes. Taxonomic affinities can indeed iden-
tify some potential problems: all but one of the Me-
lastomes naturalized in Hawaii, for instance, are in-
vasive (Wagner et al. 1990). Many of the world’s worst
invasive plants belong to relatively few families and
genera: Asteraceae, Poaceae,
Acacia, Mimosa, Cyperus
(Heywood 1989, Binggeli 1996, Holm et al. 1997).
Rejmanek and Richardson (1996) contend they can suc-
cessfully predict retrospectively which pines intro-
duced to South Africa are most invasive, based on a
list of morphological and ecological characteristics.
Furthermore, both Starlings (
Sturnus
) and Crows (
Cor-
vus
) have several invasive, or at least widely natural-
ized, species (Long 1981). But most biotic invaders
have few, if any, similarly aggressive relatives (e.g.,
Eichhornia crassipes
is the only
Eichhornia
that is in-
vasive [Barrett 1989]). This lack of correspondence
could simply reflect a lack of opportunities for immi-
gration rather than a lack of attributes for invasion
(Simberloff 1989). But the circumstantial evidence
suggests otherwise: guilt by (taxonomic) association
has proven imprecise at predicting invasive potential.
Many combinations of traits can apparently spell per-
sistence in a new range, but our ability so far to de-
cipher and quantify these combinations remains poor.
Community vulnerability to invasion
As stated above, attempts to predict relative com-
munity vulnerability to invasions have also prompted
generalizations, including the following.
Vacant, under- or unutilized niches
.—Some com-

694
R. N. MACK ET AL.
Ecological Applications
Vol. 10, No. 3
T
ABLE
1. Escape from native parasites and predators often translates into a huge benefitin
plant performance, including fitness.
Variable
Chrysanthemoides monilifera
Australia South Africa
Acacia longifolia
Australia South Africa
Main flowering time Apr–Aug Jun–Sep Aug–Oct Jul–Sep
Flowers/m
2
1010
⫾
170†840
⫾
136 530
⫾
30 ···
Fruit/flower 6.6
⫾
0.3 4.5
⫾
0.1 1.1
⫾
0.1 ···
Green fruit/m
2
6660‡3755 580 ···
Ripe seeds/m
2
4450
⫾
750 2160
⫾
350 364
⫾
70 2923
⫾
555
Soil seeds/m
2
Fragmented 6380
⫾
605 2352
⫾
20 25
⫾
4.2 ···
Whole 2475
⫾
560 2320
⫾
17 7.5
⫾
1.0 7600
⫾
1440
Viable 2030
⫾
460 46
⫾
28 5.6
⫾
0.8 7370
⫾
1400
Notes: Chrysanthemoides monilifera
and
Acacia longifolia
are native to South Africa and
Australia, respectively. Plants of both species display much greater flower and seed production
when grown in the other country, benefiting from the escape from native pests and little or no
attack by native pests in their new ranges (Weiss and Milton 1984).
†Values are means
⫾
SE
.
‡Calculated.
munities such as tropical oceanic islands appear to be
particularly vulnerable to invasions (Elton 1958), al-
though the evidence can be equivocal (Simberloff
1995). The vacant niche hypothesis suggests that island
communities and some others are relatively impover-
ished in numbers of native species and thus cannot
provide ‘‘biological resistance’’ to nonindigenous spe-
cies (sensu Simberloff 1986). However, manypotential
invaders arriving on islands would find no pollinators,
symbionts, or other required associates among the na-
tive organisms, a factor that might provide island com-
munities with a different form of resistance to invasion.
Yet actual demonstration of vacant niches anywhere
has proved difficult (Simberloff 1995).
Escape from biotic constraints
.—Many immigrants
arrive in new locales as seeds, spores, eggs, or some
other resting stage without their native associates, in-
cluding their usual competitors, predators, grazers, and
parasites (Elton 1958, Strong et al. 1984). This ‘‘great
escape’’ can translate into a powerful advantage for
immigrants. All aspects of performance such as growth,
longevity, and fitness can be much greater for species
in new ranges (Weiss and Milton 1984, Crawley 1987;
Table 1). According to this hypothesis, an invader per-
sists and proliferates not because it possesses a suite
of extraordinary traits but rather because it has fortu-
itously arrived in a new range without virulent or at
least debilitating associates. For example, the Austra-
lian brushtail possum (
Trichosurus vulpecula
) has be-
come an invader in New Zealand since its introduction
150 years ago (Clout 1999). In New Zealand it has
fewer competitors for food and shelter, no native mi-
croparasites, and only 14 species of macroparasites,
compared with 76 in Australia (Clark et al. 1997). Its
population densities in New Zealand forests are 10-
fold greater than those prevailing in Australia. Of
course, such a successful performance depends on an
immigrant not acquiring a new array of competitors,
predators, and parasites in its adopted community. It
is probably inevitable on continents that an invader will
acquire these foes, especially as it expands its range
and comes into contact with a wider group of native
species (Strong et al. 1984). The idea of escape from
biotic constraints is the most straightforward hypoth-
esis to explain the success of an invader, and also pro-
vides the motivation for researchers to search for bi-
ological control agents among its enemies in its native
range (De Bach and Rosen 1991).
Community species richness
.—Elton (1958) pro-
posed that community resistance to invasions increases
in proportion to the number of species in the com-
munity, its species richness. To Elton, this followed
from his hypothesis that communities are more ‘‘sta-
ble’’ if they are species-rich. This idea is a variant of
the vacant niche hypothesis; i.e., a community with
many species is unlikely to have any vacant niches that
cannot be defended successfully from an immigrant.
On land, however, resistance to plant invasion may
correlate more strongly with the architecture of the
plant community (specifically, the maintenance of a
multitiered plant canopy) than with the actual number
of species within the community. For instance, many
forest communities have remained resistant to plant
invaders as long as the canopy remained intact (Corlett
1992). Here again, exceptions abound (Simberloff
1995).
Disturbance before or upon immigration
.—Humans,
or the plants and animals they disperse and domesti-
cate, may encourage invasions by causing sudden, rad-
ical disturbances in the environment (Harper 1965,
Mack 1989). If native species can neither acclimatize
nor adapt, the subsequent arrival of preadapted im-
migrants can lead swiftly to invasions. Such biological
consequences can be provoked by fire, floods, agri-
cultural practices, or livestock grazing on land, or by
drainage of wetlands or alterations of salinity, and nu-

June 2000 695
BIOTIC INVASIONS
T
ABLE
2. Loss of the American chestnut (
Castanea dentata
) through its destruction by the
invasive fungus
Endothia parasitica
was swift.
Species
Basal area (dm
2
/ha)
1934 1941 1953
Density (no. stems/ha)
1934 1941 1953
Castanea dentata
200.53 144.67 3.38 187.82 146.98 16.67
Carya
spp. 70.68 60.56 77.82 55.93 62.24 73.81
Quercus prinus
39.43 38.12 53.35 30.71 29.86 25.00
Quercus ruba
36.95 36.39 19.47 29.87 25.45 9.99
Quercus velutina
35.97 40.04 67.44 8.58 13.92 28.81
Aesculus octandra
15.78 16.43 19.02 3.32 3.43 3.56
Quercus alba
15.75 18.50 10.34 10.72 11.41 13.43
Robinia pseudoacacia
14.66 12.86 7.74 14.89 19.62 6.67
Liriodendron tulipfera
11.05 20.83 48.80 38.57 57.39 61.55
Acer rubrum
7.51 9.29 13.99 23.07 22.70 22.86
Betula lenta
7.44 7.78 11.13 6.07 6.07 10.23
Quercus coccinea
5.63 8.16 26.09 1.90 5.71 4.41
Miscellaneous 23.38 27.45 30.15 54.92 62.12 60.86
Total 484.83 441.08 388.72 466.37 466.90 337.85
Notes:
Basal area (dm
2
/ha) and density (no. stems/ha) on Watershed 41 (Coweeta Hydrologic
Laboratory, North Carolina in 1934, 1941, and 1953 record original dominance of chestnut in
this stand and its destruction within 20 years after arrival of the parasite. Data are for all stems
ⱖ
1.27 cm (data converted to metric units from Nelson [1955]).
trient levels in streams and lakes. Novel disturbances,
or intensification of natural disturbances such as fire,
have played a significant role in some of the largest
biotic invasions, such as the extensive plant invasions
across vast temperate grasslands in Australia and North
and South America (Mack 1989, D’Antonio and Vi-
tousek 1992).
The difficulty of predicting any community’s vul-
nerability to an invasion is increased substantially by
the bias of immigration, i.e., it is nearly impossible to
test critically the relative merits of these hypotheses
because of confounding issues, such as the enormous
differences among communities in their opportunity to
receive immigrants. The likelihood that a community
will have received immigrants is influenced largely by
its proximity to a seaport or other major point of entry
and also the frequency, speed, and mode of dispersal
of the immigrants themselves (Simberloff 1989, Wil-
liamson 1996, Lonsdale 1999). For example, for more
than 300 years an ever-growing commerce has both
accidentally and deliberately delivered nonindigenous
species to the coasts of South Africa and the north-
eastern United States. Not surprisingly, the naturalized
floras in these regions are very large (Seymour 1969,
Richardson et al. 1992). In contrast, some continental
interiors, such as Tibet, have minuscule numbers of
naturalized plants and animals and few, if any, invaders
(Wang 1988). The native biota in such regions may
present strong barriers to naturalization and invasion,
but isolation alone could explain the lack of invaders.
B
IOTIC
I
NVASIONS AS
A
GENTS OF
G
LOBAL
C
HANGE
Human-driven biotic invasions have already caused
wholesale alteration of the earth’s biota, changing the
roles of native species in communities, disrupting evo-
lutionary processes, and causing radical changes in
abundances, including extinctions (Cronk and Fuller
1995, Rhymer and Simberloff 1996). These alterations
are collectively a threat to global biodiversity that is
second in impact only to the direct destruction of hab-
itat (Walker and Steffen 1997).
Biotic invaders themselves often destroy habitat, for
instance by altering siltation rates in estuaries and along
shorelines (Bertness 1984, Gray and Benham 1990). In
the past, the scope of this direct loss of habitat was
local or at most regional. However, with invasions oc-
curring at an unprecedented pace, invaders are collec-
tively altering global ecosystem processes (Vitousek et
al. 1996). Furthermore, the growing economic toll
caused by invasions is not limited by geographic or
political boundaries (U.S. Congress 1993, Sandlund et
al. 1996). Invaders are by any criteria major agents of
global change today. We provide below only a brief
sketch of the range of effects that biotic invaders cause
to biodiversity and ecological processes.
Population-level effects
Invasions by disease-causing organisms can severely
impact native species. The American chestnut (
Cas-
tanea dentata
) once dominated many forests in the east-
ern United States, especially in the Appalachian foot-
hills (Braun 1950), until the Asian chestnut blight fun-
gus arrived in New York City on nursery stock early
in this century. Within a few decades, the blight had
spread throughout the eastern third of the United States,
destroying almost all American chestnuts within its na-
tive range (Roane et al. 1986) (Table 2). The mosquito
Culex quinquefasciatus
that carries the avian malaria
parasite was inadvertently introduced to the Hawaiian
Islands in 1826. The parasite itself arrived subsequent-
ly, along with the plethora of Eurasian birds that now
dominate the Hawaiian lowlands. With avian malaria

696
R. N. MACK ET AL.
Ecological Applications
Vol. 10, No. 3
F
IG
. 3. Percentage levels of native and nonindigenous birds on Mauna Loa, Hawaii, infected with avian malaria, 1978–
1979 and mean numbers of parasites per 10 000 RBCs. As a result of the native birds’greater susceptibility, they were
largely restricted to higher elevations. Numbers in brackets or parentheses are sample sizes (Van Riper et al. 1986). RBC
⫽
red blood cells.
rampant in the lowlands, the Eurasian invaders, which
are at least somewhat resistant to it, have excluded
native Hawaiian birds, which are highly susceptible to
the disease (van Riper et al. 1986; Fig. 3).
Predation and grazing by invaders can also devastate
native species. The predatory Nile perch (
Lates nilo-
tica
), which was introduced into Africa’s Lake Victoria,
has already eliminated or gravely threatens more than
200 of the 300 to 500 species of the great evolutionary
radiation of native cichlid fishes (Goldschmidt 1996).
Feral and domestic cats have been transported to every
part of the world and have become devastating pred-
ators of small mammals and ground-nesting or flight-
less birds. On many oceanic islands, feral cats have
depleted breeding populations of seabirds and endemic
land birds. In New Zealand, cats have been implicated
in the extinction of at least six species of endemic birds,
as well as some 70 populations of island birds (King
1985). In Australia, cat predation takes its biggest toll
on small native mammals. Cats are strongly implicated
in 19th century extinctions of at least six species of
native Australian marsupials (
Pseudomys
and
Notomys
)
(Dickman 1996). The brown tree snake (
Boiga irre-
gularis
), introduced to Guam in the late 1940s from
the Admiralty Islands (Rodda et al. 1992), has already
virtually eliminated all forest birds in Guam (Savidge
1987). Goats introduced to St. Helena Island in 1513
almost certainly extinguished more than 50 endemic
plant species, although only seven were scientifically
described before their extinction (Groombridge 1992).
Invaders still extract a severe toll on St. Helena. A
South American scale insect (
Orthezia insignis
) has
recently threatened the survival of endemic plants, in-
cluding the now rare native tree,
Commidendrum ro-
bustum.
Two years after the scale infestation began in
1993, at least 25% of the 2000 remaining trees had
been killed (Booth et al. 1995).
Nonindigenous species may also compete with na-
tives for resources. The North American gray squirrel
(
Sciurus carolinensis
) is replacing the native red squir-
rel (
S. vulgaris
) in Britain by foraging more efficiently
(Williamson 1996). The serial invasion of New Zea-
land’s
Nothofagus
forests by two wasp species has
harmed native fauna, including both invertebrates that
are preyed on by wasps and native birds that experience
competition for resources (Clout 1999). For instance,
the threatened Kaka (
Nestor meridionalis
), a forest par-
rot, forages on honeydew produced by a native scale
insect. But
⬎
95% of this resource is now claimed by
invasive wasps during the autumn peak of wasp den-
sity, and as a result the parrots abandon the
Nothofagus
forests during this season (Beggs and Wilson 1991).
The native biota of the Galapagos Islands is threatened
by goats and donkeys, not only because of their grazing
but because they trample the breeding sites of tortoises
and land iguanas (Bensted-Smith 1998). Invasive
plants have diverse means of competing with natives.
Usurping light and water are probably the most com-

June 2000 697
BIOTIC INVASIONS
F
IG
.4.
Carpobrotus edulis,
a sprawling perennial plant,
invades California coastal communities. It overtops native
species, such as
Haplopappus ericoides,
and competes ag-
gressively for soil water. Its removalcoincides with a marked
increase in canopy area of
H. ericoides
; values represent
change as a percentage of initial canopy area. Error bars are
⫹
1
SE
(D’Antonio and Mahall 1991).
F
IG
. 5. Invasion of Brazilian fire ants,
Solenopsis invicta,
into woodlands and grasslands in central Texas causes a rad-
ical change in the density and species composition of the
native ant fauna, as reflected in pitfall trap records. Species
richness and numbers of native ant workers decline sharply,
while the invader’s populations are several orders of mag-
nitude greater than all ants in uninfested sites. Note the much
larger scale on the bottom graph, showing numbers of allants
combined. All values were calculated with site pitfall trap
totals summed across May, July, and October 1987 (Porter
and Savignano 1990).
mon tactics. For example, the succulent mat-former,
Carpobrotus edulis,
pervades the same shallow rooting
zone as native shrubs in California coastal communi-
ties. Its removal coincides with improved water avail-
ability for the natives, strongly suggesting that the in-
vasive
C. edulis
usurps water that would otherwise be
available for native plants’growth (D’Antonio and Ma-
hall 1991; Fig. 4).
Interference competition by invasive species is even
more easily demonstrated. For example, several widely
introduced ant species (the red fire ant [
Solenopsis in-
victa
], the Argentine ant [
Linepithema humile
], and the
big-headed ant [
Pheidole megacephala
]) all devastate
large fractions of native ant communities by aggression
(references in Williams 1994; Fig. 5). Although the
evidence is often equivocal for allelopathy, the widely
introduced agricultural pest
Agropyron repens
is one
of the few species that likely interferes with compet-
itors through release of phytotoxins (Welbank 1960).
Invasive species can also eliminate natives by mating
with them, a particular danger when the native species
is rare. For example, hybridization with the introduced
North American Mallard (
Anas platyrhynchos
) threat-
ens the existence, at least as distinct species, of both
the New Zealand Gray Duck (
Anas superciliosa
) and
the Hawaiian Duck (
A. wyvilliana
; references in Rhym-
er and Simberloff 1996). Hybridization between a non-
indigenous species and a native one can even produce
a new invasive species. For example, North American
cordgrass (
Spartina alterniflora
), carried in shipping
ballast to southern England, hybridized occasionally
with British native cordgrass (
S. maritima
). These hy-
brid individuals were sterile, but eventually one un-
derwent a doubling of chromosome number to produce
a fertile, highly invasive species,
S. anglica
(Thompson
1991). Hybridization can threaten a native species even
when the hybrids do not succeed, simply because cross-
breeding reduces the number of new offspring added
to the species’own population. For example, females
of the European mink (
Mustela lutreola
), already
gravely threatened by habitat deterioration, hybridize
with males of introduced North American mink (
M.
vison
). Embryos are invariably aborted, but the wastage
of eggs exacerbates the decline of the native species
(Rozhnov 1993).
Species can evolve after introduction to a new range.
The tropical alga,
Caulerpa taxifolia,
evolved tolerance
for colder temperatures while it was growing at the
aquarium of the Stuttgart Zoo and other public and
private aquaria in Europe. Since then it has escaped
into the northwest Mediterranean, and its new tolerance
of winter temperatures has permitted it to blanket vast
stretches of the seafloor, threatening nearshore marine
communities (Meinesz 1999). Evolution can also

698
R. N. MACK ET AL.
Ecological Applications
Vol. 10, No. 3
T
ABLE
3.
Myrica faya,
an invasive nitrogen-fixing tree in
Hawaii, radically increases the local nitrogen budget and
thus facilitates the entry of other nonindigenous species
into native communities.
Source
N input (kg/ha)
LB UB
Fixation by
Myrica faya
18.5 0.2
Native N fixation
Lichens 0.02 0.06
Litter 0.12 0.16
Decaying wood 0.05 0.03
Precipitation
NH
4
-N
⫹
N0
3
-N 1.0 1.0
Organic N 2.8 2.8
Total inputs 22.5 4.2
Note:
Annual nitrogen inputs are compared for two sites:
LB, in which
M. faya
density was
⬎
1000 plants/ha by 1987;
and UB, in which
M. faya
had only recently arrived (Vitousek
and Walker 1989).
change potential impacts in subtler ways.
Bathyplectes
curculionis,
an ichneumonid parasitic wasp imported
to the United States to control the alfalfa weevil (
Hy-
pera postica
) was originally ineffective against the
Egyptian alfalfa weevil,
Hypera brunneipennis.
Dis-
sections showed that 35–40% of its eggs were de-
stroyed by the immune response of the larval weevil.
Samples taken fifteen years later showed only 5% egg
loss (Messenger and van den Bosch 1971).
Community- and ecosystem-level effects
The biggest ecological threat posed by invasive spe-
cies is the disruption of entire ecosystems, often by
invasive plants that replace natives. For example, the
Australian paperbark tree (
Melaleuca quinquenervia
),
which at one time increased its range in south Florida
by
⬎
20 ha per day, replaces cypress, sawgrass, and
other native species. It now covers about 160000 ha,
often in dense stands that exclude virtually all other
vegetation. It provides poor habitat for many native
animals, uses huge amounts of water, and intensifies
the fire regime (Schmitz et al. 1997). Similarly,
Mimosa
pigra
has transformed 80 000 ha of tropical wetland
habitat in northern Australia into monotonous tall
shrubland (Braithwaite et al. 1989), excluding native
waterbirds. The South American shrub,
Chromolaena
odorata
or Siam weed, is not only an aggressive in-
vader in both Asia and Africa, suppressing the regen-
eration of primary forest trees, but also provides feed-
ing niches that can sustain other pests (Boppre´et al.
1992). Another highly invasive neotropical shrub,
Lan-
tana camara,
serves as habitat for the normally stream-
dwelling tsetse fly in East Africa, increasing the inci-
dence of sleeping sickness in both wild and domesti-
cated animals, as well as in humans (Greathead 1968).
Many invasive species wreak havoc on ecosystems
by fostering more frequent or intense fires, to which
key native species are not adapted.
Melaleuca quin-
quenervia
has this effect in Florida (Schmitz et al.
1997), as do numerous invasive grasses worldwide
(D’Antonio and Vitousek 1992). In general, grasses
produce a great deal of flammable standing dead ma-
terial, they can dry out rapidly, and many resprout
quickly after fires (D’Antonio and Vitousek 1992).
An invasion of Hawaii Volcanoes National Park by
a small tree,
Myrica faya,
native to the Canary Islands,
is transforming an entire ecosystem because the invader
is able to fix nitrogen and increase supplies of this
nutrient in the nitrogen-poor volcanic soils at a rate
90-fold greater than native plants (Vitousek and Walker
1989; Table 3). Many other nonindigenous plants in
Hawaii are able to enter only sites with relatively fertile
soils, so
M. faya
paves the way for further invasions,
raising the threat of wholesale changes in these plant
communities (Vitousek et al. 1987).
Myrica faya
also
attracts the introduced Japanese White-eye (
Zosterops
japonica
); the White-eye disperses
Myrica
seeds (Vi-
tousek and Walker 1989) and is believed to be a com-
petitor of several native bird species (Mountainspring
and Scott 1985).
Ecosystem transformations wrought by invaders
have been so complete in some locales that even the
landscape itself has been profoundly altered. ‘‘The
Bluegrass Country’’ of Kentucky invokes images for
most Americans of a pastoral, even pristine, setting.
But bluegrass is
Poa pratensis,
a Eurasian invader that
supplanted the region’s original vegetation, an exten-
sive open forest and savanna with
Elymus
spp. and
possibly
Arundinaria gigantea
in the understory (Dau-
benmire 1978), after European settlement and land
clearing. The European periwinkle (
Littorina littorea
),
introduced to Nova Scotia around 1840, has trans-
formed many of the coastal inlets along the northeast
coast of North America from mudflats and salt marshes
to a rocky shore (Bertness 1984; Fig. 6). Similar whole-
sale transformations of the landscape have occurred
elsewhere, including the conversion of the Florida Ev-
erglades from a seasonally flooded marsh to a fire-prone
forest of invasive trees (Bodle et al. 1994) and the
invasion of the fynbos in South Africa’s Cape Province
by eucalypts, pines,
Acacia,
and
Hakea
spp. (van Wil-
gen et al. 1996). Heavy water use by these invasive
trees in South Africa has led to major water losses
(estimated at 3
⫻
10
9
m
3
/y, Anonymous 1997
b
), and
many rivers now do not flow at all or flow only infre-
quently. This change has in turn reduced agricultural
production and also threatened the extinction of many
endemic plant species from the Cape flora (van Wilgen
et al. 1996).
Our best estimate is that, left unchecked, the current
pace and extent of invasions will influence other agents
of global change, principally the alteration of green-
house gases in the atmosphere, in an unpredictable but
profound manner (Mack 1996). The current transfor-

June 2000 699
BIOTIC INVASIONS
F
IG
.6.
Littorina littorea
(European peri-
winkle) has greatly increased the extent of rocky
shoreline along New England and the Canadian
Maritime coast through its grazing on marine
plants that once induced siltation and mud ac-
cumulation. Its removal and exclusion from ar-
eas caused a rapid resumption in sedimentation
with accompanying algal colonization. Error
bars show
⫾
1
SD
; sample sizes of sites appear
over each bar (Bertness 1984).
F
IG
. 7. Invasion of African grasses in the
Amazon Basin could eventually cause the per-
manent conversion of this vast forested carbon
sink into grassland or savanna-like areas. As
depicted schematically, fire-initiated land clear-
ing allows the entry of these grasses. The flam-
mability of their abundant litter rapidly fosters
their persistence at the expense of native woody
species. This ratchet-like conversion across
such a huge area holds important implications
for ecosystem alteration at a global scale
(D’Antonio and Vitousek 1992).
mation of ecosystems in the Amazon basin through the
burning of forests and their replacement with African
grasses provides one of the most ominous examples.
For example, in Brazil the conversion of diverse forest
communities into croplands and pastures has often in-
volved the deliberate sowing of palatable African
grasses (
Melinis minutiflora, Hyparrhenia rufa, Pani-
cum
spp., and
Rhynchelytrum repens
) (Eiten and Good-
land 1979). The spread and proliferation of these grass-
es has been fostered by fire. By 1991 cleared forest
sites that largely support grass-dominated communities
were estimated to cover 426000 km
2
in Brazil alone
(Fearnside 1993); much more of the 4
⫻
10
6
km
2
of
the multilayered forest in the Amazon basin in Brazil
is at risk of similar conversion.
These extensive human-driven grass invasions could
not only alter ecosystem-level properties in Brazil but
also have repercussions worldwide (Vitousek et al.
1996). Perhaps most significant is the fact that grass-
lands contain much less plant biomass than the native
forests and thus sequester less carbon (Kaufmann et al.
1995, Kaufmann et al. 1998). Given the extent of the
neotropical forests, continuing conversions to grass-
lands could exacerbate the buildup of carbon dioxide
in the atmosphere, potentially influencing global cli-
mate. Less evapotranspiration from grasslands com-
pared to tropical forest (Shukla et al. 1990) could also
translate into greater convective heat loss and increases
in air temperature (Walters 1979). Although fire and
other agents of land-clearing initiate these changes in
the Amazon watershed, the persistence of invasive
grasses thereafter limits any natural recolonization of
cleared areas by native forest species. Thus, invasive
African grasses are having a ratchet-like effect in the
Amazon watershed: as more of the native vegetation
is converted to pasture, these grasses prevent recolo-
nization and succession by native species (Fig. 7).
Economic consequences
Attempts to arouse public and governmental support
for the prevention or control of invasions often fail
because of a lack of understanding of the inextricable
link between nature and economy. But the threats biotic
invasions pose to biodiversity and to ecosystem-level
processes translate directly into economic consequenc-
es such as losses in crops, fisheries, forestry, and graz-
ing capacity. Yet no other aspect of the study of biotic
invasions is as poorly explored and quantified. Al-

700
R. N. MACK ET AL.
Ecological Applications
Vol. 10, No. 3
though there are ample anecdotal examples of local and
even regional costs of invaders, we consistently lack
clear, comprehensive information on these costs at na-
tional and especially global levels.
Biotic invasions cause two main categories of eco-
nomic impact. First is the loss in potential economic
output: i.e., losses in crop production and reductions
in domesticated animal and fisheries survival, fitness,
and production. Second is the direct cost of combating
invasions, including all forms of quarantine, control,
and eradication (U.S. Congress 1993). A third category,
beyond the scope of this report, would emphasize the
costs of combating invasive species that are threats to
human health, either as direct agents of disease or as
vectors or carriers of disease-causing parasites.
These costs form a hidden but onerous ‘‘tax’’ on
many goods and services. Tallying these costs, how-
ever, remains a formidable task. Pimentel et al. (2000)
attempted recently to tabulate the annual cost of all
nonindigenous species in the United States. They es-
timate that nonindigenous weeds in crops alone cost
U.S. agriculture
⬃
$27 billion per year, based on a po-
tential crop value of
⬃
$267 billion. Loss of forage and
the cost of herbicides applied to weeds in rangelands,
pastures, and lawns cause a further $6 billion in losses
each year. When they combined thesedirect losses with
indirect costs for activities such as quarantine, the total
cost of all nonindigenous species (plants, animals, mi-
croorganisms) exceeded $138 billion per year. By any
standard, such costs are a formidable loss, even for a
productive industrialized society such as the United
States.
These estimates illustrate the preliminary level of
our current understanding of the economics of inva-
sions. One solution would be a more frequent appli-
cation of economic tools such as cost–benefit analyses
when considering proposals to import species for per-
ceived economic benefit (Naylor 1996, Pannell 1999).
When it comes to future movements of species, society
needs to be able to consider results from the types of
analysis economists already provide for other projects
with potential environmental consequences, such as
construction of hydroelectric dams, canals, and air-
ports. We predict that cost–benefit analysis of many
deliberately introduced invaders would demonstrate
forcefully that their costs to society swamp any realized
or perceived benefits.
P
REVENTION AND
C
ONTROL OF
B
IOTIC
I
NVASIONS
The consequences of biotic invasions are often so
profound that they must be curbed and new invasions
prevented. This section is divided into two parts: first,
efforts to prevent the opportunity for invasions by pro-
hibiting the entry of nonindigenous species into a new
range; and second, concepts for curbing the spread and
impact of nonindigenous species, including invaders,
once they have established in a new range.
Preventing entry of nonindigenous species
The use of quarantine, which is intended to prohibit
organisms from entering a new range, has a long history
in combating human parasites (McNeill 1976). Rarely
is the saying ‘‘an ounce of prevention is worth a pound
of cure’’ so applicable as with biotic invasions. Most
invasions begin with the arrival of a small number of
individuals (Simberloff 1986, Mack 1995), and the
costs of excluding these is usually trivial compared to
the cost and effort of later control after populations
have grown and established.
The ability of a nation to restrict the movement of
biotic invaders across its borders is ostensibly governed
by international treaties, key among them being the
Agreement on the Application of Sanitary and Phy-
tosanitary Measures (SPS) (Anonymous 1994). Under
this agreement members of the World Trade Organi-
zation (WTO) can restrict movement of species that
may pose a threat to human, animal, or plant life (Anon-
ymous 1994). The International Plant Protection Con-
vention (IPPC) of 1951 deals with quarantine against
crop pests (Jenkins 1996), and the IPPC Secretariat also
coordinates phytosanitary standards (Anonymous
1994). The SPS agreement requires WTO members to
base any SPS measures on internationally agreed
guidelines (see Anonymous 1994).
Unfortunately, neither the specific wording, current
interpretation, nor implementation of these agreements
provides totally effective control against biotic invad-
ers. Nations may give variances or exceptions based
on politico-economic considerations that outweigh bi-
ological concerns. Even if a nation attempts to ban
importation of a species, its efforts may fall to inter-
national judgment if the WTO, in its regulatory ca-
pacity, rules that the ban is an unlawful or protectionist
trade barrier rather than a legitimate attempt to exclude
pests (Jenkins 1996). Thus, environmental concerns
and politico-economic interests may clash.
Within these international guidelines, some coun-
tries, including Australia and the United States, have
imposed quarantine controls that take an ‘‘innocent un-
til proven guilty’’ approach, e.g., they have allowed
entry of any nonindigenous species that are not known
to be harmful. This approach has been attacked from
two sides: some want to liberalize trade, remove non-
tariff trade barriers, and ease quarantine controls; op-
ponents argue that the precautionary principle should
apply and that a ‘‘guilty until proven innocent’’ ap-
proach should be used to tighten current quarantine
protocols (Panetta et al. 1994).
The current U.S. approach is clearly inadequate to
stem the tide of entering nonindigenous organisms, and
the U. S. Department of Agriculture’s Animal and Plant
Health Inspection Service (APHIS) is considering pol-
icy changes (Reichard and Hamilton 1997). These
might involve conducting risk assessments that would

June 2000 701
BIOTIC INVASIONS
estimate the invasive potential of a species proposed
for import (Ruesink et al. 1995). In 1997, the Australian
Quarantine Inspection Service (AQIS) adopted such a
risk assessment system for screening new plant imports
based on their biological attributes and the consequent
risk of invasiveness that they pose.
As described earlier, attempts to predict from bio-
logical attributes which species will become invasive
have had very mixed success (Perrins et al. 1992). De-
bate continues between those who maintain that quar-
antine risk assessment may be achievable (Pheloung
1995, Rejmanek and Richardson 1996, Reichard and
Hamilton 1997) and those who argue that prediction
of invasiveness will always be extremely difficult(Sim-
berloff 1989, Lonsdale 1994, Williamson 1996). Clear-
ly, much research on prediction remains to be done. If
risk assessment screening procedures are to be applied
as part of a government policy, however, more must
be considered than predictive accuracy. The low
base
rate
at which species become naturalized as well as the
base rate for becoming invaders means that the pre-
dictive power of any risk assessment must be very high
to identify invaders reliably (Smith et al. 1999). As a
consequence, screening systems are likely to produce
high rates of false positives (C. S. Smith,
unpublished
data
).
In after-the-fact assessments of previously intro-
duced plants, the screening system now adopted by
AQIS had an accuracy of
⬃
85% (Pheloung 1995). The
AQIS system rejects or recommends for further eval-
uation roughly 30% of the species proposed for import
(Pheloung 1999). It is likely that the vast majority of
these are ‘‘false positives’’ that would not have become
invasive (Smith et al. 1999). But such an exclusionary
policy risks conflict between environmentalists and
commodity groups, such as horticulturists, who ad-
vocate the liberal introduction of species. Whether this
degree of restriction on trade can be sustained remains
to be seen; globally, society is unlikely ever to prohibit
liberal movement of plants and animals in commerce.
Thus, the challenge is to identify the few potentially
harmful immigrants among an increasing throng of in-
nocuous entrants.
Eradication
Eradication of a nonindigenous species is sometimes
feasible, particularly if it is detected early and resources
can applied quickly (Simberloff 1997). Usually, how-
ever, there is insufficient ongoing monitoring, partic-
ularly in natural areas, to detect an infestation soon
after it occurs. Many regulatory agencies tend to ignore
nonindigenous species, feeling that attempts at control
are not worth the bother and expense until one becomes
widespread and invasive. Unfortunately, by that time
eradication is probably not an option (Simberloff
1997). This problem of getting agencies to take non-
indigenous species seriously is exacerbated by the pro-
longed lag times between establishment of some im-
migrant species and their emergence as invaders.
Nevertheless, some potentially damaging nonindig-
enous species have been eradicated. For example, an
infestation of the Asian citrus blackfly(
Aleurocanthus
woglumi
) on Key West in the Florida Keys was erad-
icated between 1934 and 1937 (Hoelmer and Grace
1989). This eradication project had many advantages:
there was no highway to the mainland at the time, and
the only railroad bridge was destroyed by a hurricane
in 1935. Insularity also featured prominently in an erad-
ication campaign against the screwworm fly(
Cochlio-
myia hominivorax
) by the release of sterile males. Ap-
parent success of this approach on Sanibel Island, Flor-
ida led to a similar trial on Curacao, and eradication
in that trial led to widespread release of sterile males
throughout the southeastern United States (Dahlsten
1986).
The giant African snail (
Achatina fulica
), a major
pest of agriculture in many parts of its introduced range
in Asia and the Pacific, was eradicated in sustained
campaigns against established but fairly localized pop-
ulations in south Florida (Simberloff 1997) and
Queensland, Australia (Colman 1978). Local popula-
tions of nonindigenous freshwater fishes are often erad-
icated (Courtenay 1997), and New Zealand has eradi-
cated various combinations of twelve mammal species
(ranging from rodents through feral domestic animals)
from many islands of up to 2000 ha (Veitch and Bell
1990). A few nonindigenous but not yet invasive plant
populations have been completely eradicated; these
were all from very small areas, however. For example,
Asian common wild rice (
Oryza rufipogon
) was elim-
inated from 0.1 ha of the Everglades National Park
(Simberloff 1997) and all Japanese dodder (
Cuscuta
japonica
) was apparently destroyed in a 1-ha infesta-
tion in Clemson, South Carolina (Westbrooks 1993; R.
Westbrooks,
personal communication
).
Some eradication efforts have been successful
against widespread species. For example, bacterial cit-
rus canker (
Xanthomonas campestris
pv
. citri
) was
eradicated from a broad swath of the southeastern Unit-
ed States in the early 20th century (Merrill 1989), and
a 50-year campaign succeeded in eliminating the South
American nutria (
Myocastor coypus
) from Britain
(Gosling 1989).
In all these instances, three key factors contributed
to success. First, particular aspects of the biology of
the target species suggested that the means employed
might be effective. For example, the host specificity
and poor dispersal ability of the citrus canker were
crucial to a successful eradication strategy. Second,
sufficient resources were devoted for a long enough
time. If funding is cut as soon as the immediate threat
of an economic impact lessens, eradication is impos-
sible. Third, there was widespread support both from
the relevant agencies and the public. Thus, for example,

702
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people rigorously heeded quarantines and various san-
itary measures.
Even when complete eradication fails, the effort may
well have proven cost effective and prevented sub-
stantial ecological damage. For example, a long cam-
paign to eradicate witchweed (
Striga asiatica
), an Af-
rican root parasite of several crops in the Carolinas,
has reduced the infestation from 162000 to 6000 ha
(Westbrooks 1998). The methods employed—herbi-
cides, soil fumigants to kill seeds, and regulation of
seed-contaminated crops and machinery—would have
been used anyway simply to control this invader. The
control is successful even if eradication is not com-
plete.
Other large eradication projects, however, have been
so unsuccessful that they have engendered public skep-
ticism about the entire endeavor and have, in some
instances, worsened the problem. The long campaign
to eradicate imported fire ants (
Solenopsis invicta
and
S. richteri
) from the southern United States has been
labeled by E. O. Wilson as ‘‘the Vietnam of entomol-
ogy’’ (Brody 1975) and was a $200 million disaster
(Davidson and Stone 1989). Not only did fire ants re-
invade areas cleared of ants by insecticides, but they
also returned faster than many native ant species. The
introduced range of fire ants expanded several-fold dur-
ing the 20-year campaign, and enough was known at
the time about the biology of these ants that the out-
come could have been predicted (Davidson and Stone
1989).
Maintenance control
If eradication fails, the goal becomes ‘‘maintenance
control’’ of a species at acceptable levels (Schardt
1997). Three main approaches, applied singly or in
various combinations, are widely used: chemical, me-
chanical, and biological control.
Chemical control probably remains the chief tool in
combating nonindigenous pests in agriculture. Chem-
ical controls, unfortunately, have too often created
health hazards for humans and nontarget species. For
example, problems associated with DDT are well
known. But the frequent evolution of pest resistance
(National Research Council 1986), the high cost, and
the necessity of repeated applications often make con-
tinued chemical control impossible. If the goal were to
control an invasive species in a vast natural area, the
cost of chemical methods alone would be prohibitive.
Even when there is no firm evidence of a human health
risk, massive use of chemicals over heavily populated
areas inevitably generates enormous public opposition,
as demonstrated by the heated responses to recent aerial
spray campaigns using malathion against the medflyin
California (Carey 1992).
Chemical control of plant parasites has a mixed rec-
ord, depending on the parasite and the scale of required
protection. In native forests in Australia, broadscale
chemical control of the root fungus
Phytophthora cin-
namomi
was at best only temporarily effective, while
injection of individual trees was deemed too expensive
(Weste and Marks 1987). The history of controlling
coffee rust (
Hemileia vastatrix
) in Latin America is
emblematic of the frustration of attempting to control
invasive plant pathogens. Repeatedly, each affected
coffee-growing country applied a barrage of fungi-
cides, initially attempting to eradicate the parasite and
then attempting to contain it (Hill and Waller 1982; J.
M. Waller,
personal communication
).
Mechanical methods of controlling nonindigenous
organisms are sometimes effective and usually do not
engender public criticism. Sometimes they can even be
used to generate public interest in and support for con-
trol of invasive species. In Florida’s Blowing Rocks
Preserve, volunteers helped remove Australian pine
(
Casuarina equisetifolia
), Brazilian pepper (
Schinus
terebinthifolius
), and other invasive plants and to plant
more than 60000 individuals of 85 native species
(Randall et al. 1997). Hand-picking of giant African
snails was a key component of the successful eradi-
cation campaigns in Florida and Queensland (Simber-
loff 1997 and references therein). However, equipment
expenses, the difficulty of actually finding the target
organisms, and the geographic scale of some nonin-
digenous species infestations frequently render me-
chanical control impossible.
Hunting is often cited as an effective method of
maintenance control of nonindigenous animals, and
hunting and trapping were crucial in the successful
eradication of the nutria from Britain. In the Gala´pagos
Islands, park officials have a long-established cam-
paign to eradicate nonindigenous mammals, and over
the past 30 years goats have been eliminated from five
islands (Ospina 1998). By contrast, recreational hunt-
ing alone is unlikely to serve as an effective control
on an invasive mammal. In New Zealand, hunting of
Australian brushtail possums was encouraged from
1951 to 1961 through a bounty system and harvesting
of animals for pelts. More than 1 million animals each
year were shot or trapped in the late 1950s. Neverthe-
less, the possum continued to spread (McDowall 1994).
Recreational hunting of introduced red deer (
Cervus
elaphus
) in New Zealand has also generally failed to
reduce densities enough to speed regeneration of native
forests. For both possums and red deer, widespread
control is now conducted primarily by aerial applica-
tion of poison baits, which has its own set of problems,
including lack of widespread public acceptance (Clout
1999).
Problems with both chemical and mechanical con-
trols have focused attention on biological control—the
introduction of a natural enemy of an invasive species.
In a sense, this is a planned invasion. It aims to estab-
lish in the new range at least part of the biotic control
the target species experiences in its native range. Some

June 2000 703
BIOTIC INVASIONS
biological control projects have succeeded in contain-
ing very widespread, damaging infestations at accept-
able levels with minimal costs. Examples include the
well-known control of invasive prickly pear cactus
(
Opuntia inermis
and
O. stricta
) in Australia by the
moth
Cactoblastis cactorum
from Argentina (Osmond
and Monro 1981); control of South American alligator
weed (
Alternanthera phyloxeroides
) in Florida and
Georgia by a flea beetle (Center et al. 1997), and control
of the South American cassava mealybug (
Phenacoc-
cus manihoti)
in Africa by a South American encyrtid
wasp (Odour 1996). In each of these cases, the natural
enemy has controlled the pest in perpetuity, without
further human intervention. When the pest increases in
numbers, the natural enemy increases correspondingly,
causing the pest to decline, which entrains a decline in
the natural enemy. Neither player is eliminated; neither
becomes common.
Caveats on biological control
Biological control has recently been critically scru-
tinized on the grounds that nontarget species, some of
them the focus of conservation efforts, have been at-
tacked and even driven to extinction by nonindigenous
biocontrol agents (Howarth 1991, Simberloff and Stil-
ing 1996). For example, the widespread introduction
of a New World predatory snail,
Euglandina rosea,
to
control the giant African snail led to extinction of many
endemic snail species in the Hawaiian and Society is-
lands (Civeyrel and Simberloff 1996 and references
therein). In these cases, the predators attacked many
prey species, thus preventing a mutual population con-
trol from developing between the predator and any sin-
gle prey species.
Insect biological control agents that have been sub-
jected to rigorous host-specificity testing have never-
theless been known to attack nontarget species. For
example, a Eurasian weevil,
Rhinocyllus conicus,
in-
troduced to North America to control invasive musk
thistle (
Carduus nutans
), is now attacking native non-
pest thistles. These natives include a federally listed
endangered species and narrowly restricted endemic
species in at least two Nature Conservancy refuges,
three national parks, and state lands (Federal Register
1997, Louda et al. 1997). Controversy about the extent
of such problems focuses primarily on two issues:
whether there is sufficient monitoring to detect such
nontarget impacts, and the likelihood that an introduced
biological control agent will evolve to attack new hosts.
However, the ability of
R. conicus
to attack these native
species had been detected before its release; poor leg-
islation, rather than an incomplete assessment precip-
itated the controversy (J. Waage,
personal communi-
cation
). The fact that biological control agents can dis-
perse and evolve, as can any other species introduced
to a new range, implies that their preliminary testing
should be extensive and conducted under extremely
secure circumstances.
Exclusion and control: socioeconomic issues
The difficulties of curbing biotic invasions illustrate
the problem of implementing scientifically based rec-
ommendations in an arena in which diverse segments
of society all have important stakes. At every level of
prevention and control, the thorny issues are as likely
to be socioeconomic as scientific.
A persistent problem with current methods of ex-
clusion and control is that they largely assume goodwill
and cooperation on the part of all citizens. For widely
varying reasons, large segments of entire industries are
committed to the introduction, at least in controlled
settings, of many nonindigenous species and are skep-
tical of arguments that they will escape and/or be prob-
lematic if they do escape. Thus, there is often organized
opposition to proposals to stiffen regulations relating
to introduction, and there is also frequent careless or
even willful disregard of existing laws.
The horticulture industry is often in the vanguard of
opposition to tight control of nonindigenous species.
It is a diverse multibillion dollar industry with im-
porters running the gamut from small, family opera-
tions specializing in a few species to large corporations
importing hundreds of taxonomically diverse species.
At one extreme, some horticulturists generate publi-
cations and websites scoffing at the very existence of
ecological problems with nonindigenous species. On
the other hand, many plant importers recognize the
dangers and at least support quarantine measures and
limited blacklists of species known to be invasive.
However, as a whole, through trade associations and
as individuals, horticulturists attempt to influence the
political process as it concerns regulation of nonindig-
enous species (Sray 1997). Furthermore, individuals
who purchase plants from importers are generally under
far less legal obligation and undergo little scrutiny in
their use of these plants.
Horticulturists have also been at least loosely allied
with other interest groups that desire quite unfettered
access to the world’sflora. State departments of trans-
portation, charged with landscaping highways, as well
as the U.S. Natural Resource Conservation Service,
constituted to battle erosion, have traditionally favored
nonindigenous species for these purposes (McArthur
et al. 1990). At least some state departments of trans-
portation are now moving toward use of native plants
(e.g., Caster 1994), but a long history of interaction
between these departments and private horticulturists
slows this process.
Agricultural interests and their regulatory agencies
have had a schizophrenic relationship with nonindig-
enous species. On the one hand, they promote the im-
portation of useful and profitable crop plants and live-
stock. On the other, they hope to control the influx of

704
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Vol. 10, No. 3
parasites, insect pests, and agricultural weeds. For ex-
ample, the thistle weevil discussed above as a biocon-
trol agent that attacks nontarget species was introduced
to North America by Agriculture Canada and spread
in the United States by the U.S. Department of Agri-
culture and various state agricultural agencies. The Ha-
waii Department of Agriculture introduced the carniv-
orous snail
Euglandina rosea
to the Hawaiian Islands
to control the giant African snail (Davis and Butler
1964).
The pet industry is also heavily invested in nonin-
digenous species. As with the horticulture industry, it
encompasses a tremendous range of operations in terms
of size, scope, and degree and nature of specialization,
and there is no monolithic stance toward threats posed
by nonindigenous species and the prospect of rigorous
control. As with horticulturists, through the political
and publicity activities of individuals and trade orga-
nizations, the general attitude of the pet industry toward
strict regulation of introductions has ranged from skep-
ticism to outright hostility (U.S. Congress 1993, Bul-
lington 1997).
Many domesticated or pet animals have escaped
from importers and breeders (for example, when fires
or storms destroyed cages), and some have become
invasive. In Britain, escapees from fur farms estab-
lished a feral population of nutria (Lever 1979), which
became the target of a lengthy eradication campaign
noted above. Sometimes, pet dealers or owners delib-
erately release animals. For example, some fishes are
deliberately released by aquarists (Courtenay 1997).
Again, as with horticulturists, once a pet is sold, the
dealer has no subsequent control over the owner’s ac-
tions, and the owner may be less likely than the dealer
to obey formal regulations.
Controversies over the management of feral horses
in both the United States and New Zealand illustrate
the conflicts that readily arise between various seg-
ments of society about some widely appreciated feral
domestic animals. In both countries feral horses pose
documented threats to native species and ecosystems.
Yet some groups contend the horses that escaped from
Spanish explorers in North America
⬃
500 years ago
‘‘belong’’ in the West, merely serving as replacements
for native equids that became extinct on the continent
⬃
10000 years ago. In New Zealand, however, there
were no native land mammals, except for bats, before
introductions by people began over the past 800 years.
Horses were introduced to New Zealand
⬍
200 years
ago.
In New Zealand, feral horses have occupied the cen-
tral North Island since the 1870s. Land development
and hunting progressively reduced both their numbers
and range; a 1979 census revealed only about 174 an-
imals. By 1981, however, public lobbying resulted in
creation of a 70 000-ha protected areaas the herd’s core
range. With protection, horses expanded their rangeand
increased to 1576 animals by 1994, essentially dou-
bling their population every four years (New Zealand
Department of Conservation 1995). In response to dam-
age in native ecosystems caused by this rapidly grow-
ing population, the New Zealand Department of Con-
servation (1995) recommended removal of the pro-
tected area, eradication of horses from 70% of their
range, and management to retain a herd of about 500
animals in the remaining range. The management plan,
which included shooting horses, provoked intense pub-
lic protest. This outcry eventually resulted in the over-
turning of a scientifically based management plan and
a 1997 decision to round up as many horses as possible
for sale. Sale of several hundred horses duly took place,
but the long-term fate of the growing herd remains
unresolved.
The impasse in New Zealand over feral horse control
has been mirrored in Nevada, where an intense dispute
has raged between land managers and pro-horse activ-
ists about the ecological impacts of feral horses, the
size of feral herds, and appropriate methods of popu-
lation control (Symanski 1996). At a practical level,
the removal of animals by culling would probably be
the simplest way of achieving population reduction, but
public resistance precludes this option.
The infusion of strong public sentiment into policy
for feral horses, as well as burros in the United States,
would likely serve as a mild preview of public reaction
to serious efforts to control feral cats. Ample evidence
demonstrates that feral cats are the most serious threat
to the persistence of many small vertebrates. Churcher
and Lawton (1989) estimate that domestic cats kill an-
nually at least 20 million birds in Britain; although the
toll taken by feral cats is widely disputed, this mortality
can only exacerbate the total effect of this nonindig-
enous species. The degree to which feral cats in Aus-
tralia should be eradicated and domestic cats sterilized
has already engendered vituperative debate. Similar
discussion, pitting environmentalists against the gen-
eral public, is being played out in the United States
(Roberto 1995) and Europe. Few biotic invasions in
coming decades will deserve more even-handed com-
ment from ecologists than the dilemma of feral cats.
Game and fish agencies have traditionally been major
importers of nonindigenous species, particularly fishes
(Courtenay 1997), game birds (Bump 1968), and mam-
mals (Cox et al. 1997). In Florida, for example, the
Florida Game and Fresh WaterFish Commission main-
tains a laboratory to seek out and test nonindigenous
fish species that might become attractive sport fish in
the state’s waters. The agency has imported several
species, including the peacock bass (
Cichla ocellaris
),
which is spreading, although its impacts on native spe-
cies are uncertain (Courtenay 1997). Although at least
some game and fish agencies have recently recognized
the need for more regulation of nonindigenous species
(Cox et al. 1997), the fact that they are still mandated

June 2000 705
BIOTIC INVASIONS
to import new species suggests a conflicted attitude.
Furthermore, many private individuals and organiza-
tions release game species in new locations. Some re-
leases of game fishes and other animals constitute de-
liberate flouting of laws. Groups of private individuals
in the northern Rocky Mountains surreptitiously re-
leased nonindigenous fish into isolated mountain lakes,
backpacking the fish to ensure that even the most iso-
lated alpine lakes received what these individuals
deemed as suitable biota (Ring 1995). Even apparently
innocuous actions can have ecologically catastrophic
impacts. The release of bait fishes by fishermen at the
end of the day has already led to the extinction of
species in the United States, including the Pecos pup-
fish (
Cyprinodon pecosensis
), through hybridization
(Echelle and Connor 1989).
Long-term strategies for control of biotic invaders
Effective prevention and control of biotic invasions
require a long-term, large-scale strategy rather than a
tactical approach focused on battling individual invad-
ers (Moody and Mack 1988, Anonymous 1997
b
, Sim-
berloff et al. 1997). An underlying philosophy of such
a strategy should be to establish why nonindigenous
species are flourishing in a region and to address the
underlying causes rather than simply destroying the
currently most oppressive invaders. System manage-
ment, rather than species management, ought to be the
focus.
One of the problems of taking a tactical view of
invaders, especially in a region where multiple invasive
organisms are flourishing, is the prospect of simply
‘‘trading one pest for another.’’ For example, intro-
duction of a successful biocontrol agent against only
one species may be ecologically useless unless there
is a strategy in place for dealing with the remaining
invaders. This unintended outcome may have already
occurred, possibly in the ascendance of yellow star-
thistle (
Hypericum performatum
) as a weed in Cali-
fornia as the impact of biocontrol on St. John’s wort
increased in the 1950s (Mack,
in press
), and it may
occur often. A strategic, system-wide approach is par-
ticularly appropriate for conservation areas, although
it is seldom undertaken (Luken and Thieret 1997, Storrs
et al. 1999).
In some nations, a broader strategic approach to the
control of invaders is being put into place. Australia
has recently adopted a national weed strategy aimed at
reducing the impact of plant invaders (Anonymous
1997
a
). Similarly, in a project of extraordinary scale,
South Africa is determined to clear all the invasive
woody species from its river catchments in a 20-year
program. The multispecies, multipronged strategy in-
volves manual clearing of thickets to allow native veg-
etation to reestablish, treatment of cut stumps with my-
coherbicides, and the use of biological control to pre-
vent reinvasion by exotic pines. Although this program
will cost US $150 million, it is far cheaper than alter-
natives such as massive dam-building programs to in-
sure the nation’s water supply, and it has the bonus of
creating thousands of jobs (Anonymous 1997
b
).
F
UTURE
R
ESEARCH AND
P
OLICY
P
RIORITIES
Extensive research on the ecology of biotic invasions
dates back only a few decades (Elton 1958, Salisbury
1961). Although much has been learned, too many of
the data remain anecdotal, and the field still lacks de-
finitive synthesis, generalization, and prediction. The
following include a few arenas in which research or
new policy initiatives, or both, seem particularly worth-
while.
1) Clearly, we need a much better understanding of
the epidemiology of invasions. As part of this goal we
need much better areal assessments of on-going in-
vasions, for both public policy decisions as well as
science. Few tools are as effective as time-series maps
in showing the public the course of an unfolding in-
vasion. For example, Elton’s (1958) portrayal of the
geographic scale of biotic invasions gained much visual
impact through his use of time-series maps. We also
emphasize here the need to collect in a more deliberate
manner information about the population biology of
immigrations that fail (Harper 1982), since an under-
standing of the failure of the vast majority of immi-
grants can eventually help us discern the early harbin-
gers of an impending invasion.
2) Experimentation in the epidemiology of invasions
is a logical extension of 1). So far, the most compre-
hensive data come from observing the fates of insects
released in biological control (Simberloff 1989) and
birds introduced on islands (Veltman et al. 1996). We
need to develop innocuous experimental releases of
organisms that can be manipulated to explore the enor-
mous range of chance events to which all immigrant
populations may be subjected (e.g., Crawley et al.
1993).
3) Worthwhile economic estimates of the true cost
of biotic invasions are rare and almost always involve
single species in small areas. We need comprehensive
cost–benefit analyses that accurately and effectively
highlight the damage inflicted on the world economy
by biotic invasions. The need is similar to the mandate
the World Health Organization meets by analyzing and
reporting the economic toll of human disease (e.g.,
WHO 1993).
4) Most members of society become aware of biotic
invasions only through some firsthand experience,
which usually involves some type of economic cost.
These cases often prompt action, or at least public re-
action, that is short-lived and local. We need instead a
greater public and governmental awareness of the
chronic and global effects of invasive organisms and
the tools available to curb their spread and restrict their
ecological and economic impacts. Public outreach

706
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Vol. 10, No. 3
about biotic invaders must match or exceed current
efforts that draw public attention to other ongoing
threats to global change (Bright 1998, Kaiser 1999).
C
ONCLUSIONS
Biotic invasions are altering the world’s natural com-
munities and their ecological character at an unprec-
edented rate. If we fail to implement effective strategies
to curb the most damaging impacts of invaders, we risk
impoverishing and homogenizing the very ecosystems
on which we rely to sustain our agriculture, forestry,
fisheries, and other resources and to supply us with
irreplaceable natural services. Given the current scale
of invasions and our lack of effective policies to pre-
vent or control them, biotic invasions have joined the
ranks of atmospheric and land-use change as major
agents of human-driven global change.
A
CKNOWLEDGMENTS
We thank David Tilman for his foresight in organizing the
Issues in Ecology series and the Pew Foundation for their
financial support of the project that produced this report. We
also thank Yvonne Baskin; her editing skills both improved
this technical report and produced a lucid version of this
report for the general audience. We are grateful to G. H.
Orians for his comments on an earlier draft of the manuscript.
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