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Invasive species and their impacts on agri-ecosystems: Issues and solutions for restoring ecosystem processes


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Humans are the most invasive of vertebrates and they have taken many plants and animals with them to colonise new environments. This has been particularly so in Australasia, where Laurasian and domesticated taxa have collided with ancient Gondwanan ecosystems isolated since the Eocene Epoch. Many plants and animals that humans introduced benefited from their pre-adaptation to their new environments and some became invasive, damaging the biodiversity and agricultural value of the invaded ecosystems. The invasion of non-native organisms is accelerating with human population growth and globalisation. Expansion of trade has seen increases in purposeful and accidental introductions, and their negative impacts are regarded as second only to activities associated with human population growth. Here, the theoretical processes, economic and environmental costs of invasive alien species (i.e. weeds and vertebrate pests) are outlined. However, defining the problem is only one side of the coin. We review some theoretical underpinnings of invasive species science and management, and discuss hypotheses to explain successful biological invasions. We consider desired restoration states and outline a practical working framework for managing invasive plants and animals to restore, regenerate and revegetate invaded Australasian ecosystems.
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Invasive species and their impacts on agri-ecosystems:
issues and solutions for restoring ecosystem processes
Peter J. S. Fleming
, Guy Ballard
, Nick C. H. Reid
and John P. Tracey
Vertebrate Pest Research Unit, New South Wales Department of Primary Industries,
Orange Agricultural Institute, 1447 Forest Road, Orange, NSW 2800, Australia.
Ecosystem Management, School of Environmental and Rural Science, University of New England,
Armidale, NSW 2351, Australia.
Vertebrate Pest Research Unit, New South Wales Department of Primary Industries, Allingham Street,
Armidale, NSW 2350, Australia.
Corresponding author. Email:
Abstract. Humans are the most invasive of vertebrates and they have taken many plants and animals with them to
colonise new environments. This has been particularly so in Australasia, where Laurasian and domesticated taxa have
collided with ancient Gondwanan ecosystems isolated since the Eocene Epoch. Many plants and animals that humans
introduced beneted from their pre-adaptation to their new environments and some became invasive, damaging the
biodiversity and agricultural value of the invaded ecosystems. The invasion of non-native organisms is accelerating
with human population growth and globalisation. Expansion of trade has seen increases in purposeful and accidental
introductions, and their negative impacts are regarded as second only to activities associated with human population
growth. Here, the theoretical processes, economic and environmental costs of invasive alien species (i.e. weeds and
vertebrate pests) are outlined. However, dening the problem is only one side of the coin. We review some theoretical
underpinnings of invasive species science and management, and discuss hypotheses to explain successful biological
invasions. We consider desired restoration states and outline a practical working framework for managing invasive plants
and animals to restore, regenerate and revegetate invaded Australasian ecosystems.
Additional keywords: adaptive management, biological invasions, removal, rate of increase.
Received 17 May 2017, accepted 3 October 2017, published online 28 November 2017
Biological invasions of native ecosystems are pervasive and
often degrading, requiring effort for restoration through removal,
revegetation and regeneration (Elton 1958). Humans are the
most invasive of vertebrates and they have taken many plants and
animals with them to colonise new environments (Vitousek
et al.1997; Rotherham and Lambert 2011). This is particularly
so in Australasia, where Old World Laurasian and domesticated
taxa have collided with ancient and geographically isolated
ancient Gondwanan ecosystems, the fauna and ora of which
Darwin (1859) noted, were utterly dissimilar in form but
analogous in function and trophic position. Plants and animals
have also been introduced to New Zealand from Australia, for
example, red necked wallabies (Notamacropus rufogriseus)
and brush-tailed possums (Trichosurus vulpecula), the latter
having devastating impacts on native biota, cattle production
and the economy (Nugent et al.2001).
Many plants and animals that humans introduced beneted
from pre-adaptation to their new environments and some
became invasive, damaging the biodiversity and agricultural
value of the invaded ecosystems. The risk of invasion of
non-native organisms is accelerating with human population
growth and globalisation (McNeely 2011). Despite excellent
quarantine services in both Australia and New Zealand,
expansion of trade has seen increases in purposeful and
accidental introductions, and their negative impacts are regarded
as second only to activities associated with human population
In this paper, we dene invasive species and outline and
discuss some theoretical underpinnings of invasive species
science and management in agri-ecosystems, which we dene as
all anthropogenically modied ecosystems used for agriculture,
both intensive and extensive. We do this to differentiate from
the term, agro-ecosystems, which is often interpreted as
highly modied ecosystems affected by agronomic practices.
We also discuss existing and new hypotheses to explain
successful biological invasions and review some conceptual
issues affecting regeneration, revegetation and the restoration
of invaded agri-ecosystems. A practical working framework for
managing invasive plants and animals is outlined.
Journal compilation Australian Rangeland Society 2017 Open Access CC BY-NC-ND
The Rangeland Journal Review
What are invasive species?
There are two types of invasive plant and animal species. Most
readily understood to be invasive are those alien species that,
when introduced, become established and harm human and
environmental values (Pimentel 2002; Prins and Gordon 2014a).
However, there are also native species that, when conditions
are anthropogenically changed, harm those same values, for
example, the large macropodids, red kangaroos, Osphranter
rufus, and eastern grey kangaroos, Macropus giganteus. Even
the iconic koala, Phascolarctos cinereus, has invasive impacts
on vegetation in southern Victoria and on Kangaroo Island
(Whisson et al.2012). The impact component is important
in this denition; some species have invasive or colonising
qualities but are regarded as benecial, for instance those
endemic plants that recolonise after perturbations such as re,
ood, soil disturbance and erosion (Bazzaz 1979). Other
examples of invasive species that are usually considered to
be benecial are exotic pasture plants that effectively persist
as a productive part of naturalised swards (e.g. Phalaris
aquatica) and non-endemic plants used to stabilise soils (e.g.
Dactyloctenium australe).
Some theory and hypotheses about biological invasions:
three fundamental curves, a rate & 13 hypotheses
Generalised invasion curve
The most frequently presented conceptual function describing the
procession of activity classes and returns on investment for
managing biological invasions is a sigmoidal curve (Fig. 1,
Department of Environment and Primary Industries Victoria
2017), often termed the generalised invasion curve. There are
four phases from preventionthrough quarantineand other
biosecurity measures to asset-based protectionfor established
and widespread biological invaders (Braysher 2017).
Perusal of the generalised invasion curve reiterates the
importance of prevention and eliminating small invasions early
before establishment. This is when such actions are more likely
to be logistically feasible. If the pest or weed, having escaped
biosecurity measures, becomes established, focus can be shifted
to containment in regions of establishment to limit the impacts
to only those areas. The curve also implies that once an invasive
species has become established, investment should be wound
back to target the protection of high-value assets. Often, the
impacts of the established pest are such that investment must be
continuous to protect the assets (Williams et al.1995; Fleming
et al.2001; Braysher 2017). Although appropriate for
established pests and weeds with well-dened ranges, focal
distributions and static or slow-moving invasion fronts, this
simple concept may be unsuitable for wide-ranging (e.g. wild
dogs, Thomson 1992; Claridge et al.2009; Robley et al.2010),
migratory or widely dispersing species such as plants with seeds
that are dispersed by wind (Higgins and Richardson 1999;
Nathan and Muller-Landau 2000) or by wide-ranging animals
(Cheal and Coman 2003). This is because control programs for
these three groups of invasive species require community
effort and cannot be adequately addressed by protecting assets
at a smaller scale than the home ranges of the pest or the
dispersion of their propagules. Solely protecting focal assets
in these cases will likely result in broader distribution and greater
cost to individuals and the community.
Density : damage functions
Before the economics of managing invasive species can be
determined, the shape of the underlying response function to the
Relative economic return
Area occupied
Asset protection
Fig. 1. A generalised invasion curve depicting the four phases of invasion and their descriptions, appropriate management
actions and the relative returns on control investment in each action. Up arrow is the point of invasion (adapted from Department
of Environment and Primary Industries Victoria 2017).
BThe Rangeland Journal P. J. S. Fleming et al.
density of the weed or pest is required (Hone 1994,2007).
These density : damage and density : yield curves describe the
incremental increase in impacts or decrease in yield for each
incremental increase in pest or weed density (Fig. 2). Multiple
hypothetical density : damage functions exist. The simplest,
a monotonic linear function, is often implicitly assumed in
economic analyses (Hone 1994). More common underlying
relationships are curvilinear functions that reach a plateau level
over which no further damage is inicted (e.g. 100% or
unharvestable levels), a similar curvilinear function but with
a threshold density below which no damage is evident, and
sigmoidal curves similar to the generalised invasion curve.
To establish these response curves requires measurement of
invasive speciesdensity and the associated response in yield
loss. Achieved yields are measured along a gradient of replicated
invasive species densities, where those densities are obtained
by adding or removing individuals (Caughley 1980), and the
data are subsequently analysed by regression. Environmental
impacts can also be tted to density : damage functions where
the damage, represented on the y-axis, is the population density
of an affected species or another community or ecosystem
measurement. The main benet of these functions is to enable
managers to identify break-even points for investment: when to
invest, how much effort or expenditure is required to achieve
a desired response, and when to stop (Hone 1994).
Rate of population increase:
Practical management of invasive plants and animals involves
manipulating the population dynamics through the rate of
increase of the invader/s to minimise their adverse impacts
(Sibly and Hone 2002). This is often manifest as reducing
populations to below the break-even points of density : damage
Rate of population increase is a fundamental concept required
to understand population dynamics, and management of invasive
species. In the absence of predation or harvesting, the rate at
which a population increases is determined by the interaction
of its life-history strategy and the quality of environmental
conditions (Caughley and Birch 1971). This is the intrinsic rate
of increase (r
), the exponential rate at which a demographically
stable population grows without resource limits.
There are several other measures that track losses from
a population (annual mortality and emigration) and additions
to a population (annual births and immigration), including the
observed rate of increase (
r), which is the exponential rate at
which a population grows over a given period of time (Caughley
1980). The observed rate of increase has greater utility for
recording the change in density of invasive species when subject
to management. In general when:
r= 0, the population is stable;
r>0, the population is growing and, for invasive species,
potentially expanding, and
r<0, the population is declining. That is, if the sum of births
and immigration are less than the sum of mortality and
emigration, the population declines.
Control of invasive species is about achieving and
r0 until a density proportional to an acceptable
level of damage is reached.
Invasion population dynamics curves
Although the generalised invasion curve (Fig. 1) conceptualises
invasion progression and the relative return for effort of
management strategies at different phases, it has several implicit
assumptions that may not prevail. Many restoration projects
require that the invasive population size as well as the area
invaded be reduced below the break-even point of investment
for most efcient management and ecosystem recovery (Hone
2007). Another implicit assumption is that investment in
a maintenance phase during asset protection from an established
invasive species will be cheaper than suppressing invasive
populations to much lower levels. Also implicitly, the shape of
the density : damage curve is assumed to be linear (ain Fig. 2),
which for invasive predators is unlikely (e.g. there is no simple
relationship between wild dog density and livestock predation
costs: Fleming et al.2014), and for many invasive plants and
animals it is unknown (Hone 1994; Braysher 2017).
If the y-axis in Fig. 1is replaced with population size
or density, then an invasion population dynamics curve results
(Fig. 3, after Caughley 1980).
The traditional generalised invasion curve (Fig. 1), which
emphasises the investment returns for management effort as
invasive species move from incursive to established pests,
should be viewed with reference to density : damage curves
(Fig. 2) and population dynamics curves (Fig. 3). Where
density : damage functions establish break-even points for
investment or thresholds above which investment fails to
achieve productivity responses, these can also be considered
as the desired stocking rate of the pest or weed. The marginal
benet or return on investment for an established pest may be
greatest after the regional population is suppressed and returns
to levels where containment is possible (Fig. 4).
Suppression to a new, lower population dynamic is a possible
objective when control tools and strategies are efcacious, but
requires buy-in from the human community and often landscape-
level application. Community buy-in may be the limiting factor
and will usually require considerable investment in time and
expertise to achieve a level of coverage to align with the size of
Density (animals or plants ha–1)
Fig. 2. Some hypotheticaldensity : damage functions describing; (a)asimple
linear function; (b) a curvilinear function that reaches a plateau at maximum
damage; and (c) a similar curvilinear function but with a threshold density
below which no damage is evident (adapted from Hone 1994).
Invasive species impacts and restoration solutions The Rangeland Journal C
effective management units (Braysher 2017). This alternative
economic model of invasion requires further investigation.
Explanatory hypotheses
Prins and Gordon (2014a) summarised hypotheses that have
been proffered to explain successful biological invasions. These
hypotheses have been usually expressed as conditions where
invasive species can invade, for example Hypothesis 1: a species
will not be able to invade an area that has abiotic conditions that
are outside its physiological tolerances(Prins and Gordon
2014a). Importantly, some of these hypotheses are alternatives
whereas others only need one falsication to be legitimate
explanations of successful invasion and establishment. To
permit these hypotheses to be tested, in order to establish their
validity against alternatives, we have rewritten them in a null
format (Table 1).
The rst hypothesis (H
, Table 1) is tautological, and
more a limiting statement than a predictor of invasiveness.
Surprisingly, ve of the articles summarised in Prins and
Gordon (2014b) rejected the hypothesis that a species can only
invade within its physiological tolerances. If an organism can
survive and reproduce outside its physiological tolerances in
a new environment, then its presumed tolerances were ipso
facto incorrectly delimited.
The conclusion of those that rejected the hypothesis or found
no support for it (4/16, Prins and Gordon 2014b) implies
a confusion by some authors of physiological experiences with
physiological tolerances. A species must be physiologically
pre-adapted to the novel environment to survive, let alone
reproduce and successfully invade. This does not mean that
the source and destination environments have to be similar as
implied by bio-climatic matching (e.g. Bomford and OBrien
1995). For example, physiological features that enable plants to
survive in salty air at sea level can decrease frost susceptibility
at higher elevations (Seki et al.2003). Human interferences are
also important drivers of vegetation change (Whalley et al.2011)
and such anthropogenic preparation of novel environments
can be critical for weed invasions (Ruttledge et al.2015).
Australasia has varied environments and the preadaptation of
exotic animals to the abiotic conditions experienced when
they were introduced could be expected. For example, rabbits
(Oryctolagus cuniculus) are derived from Mediterranean stock
and the original Australian and New Zealand propagules were
sourced from the United Kingdom, so the spread of rabbits
across Australasia would be expected.
The animals that have been introduced have in some instances
tted niches that were not lled when they arrived. For example,
cane toads (Rhinella marina) have no Australian amphibian
competitors of even approximate equivalence. Coupled with
the anthropogenically changed environments into which they
were introduced, their success as an invader was almost certain,
particularly given that the neotropical abiotic conditions into
which they were introduced matched their primary (South
America) and secondary (Hawaii) sources (H
, Table 1). New
Zealand had no arboreal mammals before the introductions
and invasion of brush-tailed possums (Trichosurus vulpecula)
from Tasmania, Victoria and possibly New South Wales (Sarre
et al.2014), that is, these were vacant niches that had not been
occupied until the invasive species was anthropogenically
introduced. These examples seem to support H
and H
(Table 1).
The inverse of H
is the competitive exclusion principle
(Gause 1934), which suggests that invasive species that occupy
a native niche dissimilar to any occupied niche in the new
environment are likely to be successful. For examples, feral cats
(Felis catus) and red foxes (Vulpes vulpes) were larger than any
quoll (Dasyurus) species at the time of their introductions,
implying that they occupy niches that were between those of the
largest extant quoll (D. maculatus) and dingoes.
The preponderance of marsupials in Australia, the host-
specicity of many pathogens and the fact that anthropogenically
introduced mammals have all been eutherian, is counter to
hypothesis H
(Table 1). Indeed, the novelty of the introduced
hosts to extant pathogens could equally result in the novel
ecosystem (i.e. the novel host) being unsuitable for the pathogens.
For Australasia, the vast difference in extant parasites, pathogens
and predators to those found in the places of origin of the
invasive species could mean that it is less likely that the invasive
Density (animals or plants ha–1)
Fig. 3. Hypothetical population dynamics of invasion and establishment.
The dark curve (a)reects the generalised invasion curve where the invader
reaches carrying capacity (k
) and levels off. The lighter curve (b) represents
the invasion when the resource base is degraded by the invader such that
carrying capacity (k
) is reduced and the population enters a dynamic
equilibrium at lower density (adapted from Caughley 1980).
Density (animals or plants ha–1)
Fig. 4. Hypothetical control curve (c) showing the return on investment for
moving established pests from expensive and ongoing maintenance for
protecting assets (curve b) to more cost-effective containment at lower
density. Higher return for investment and lower costs result from initially
suppressing the population from pest-degraded carrying capacity (k
a lower threshold or stocking rate (d) where losses are minimal or acceptable.
DThe Rangeland Journal P. J. S. Fleming et al.
species will be affected by them. Alternatively, the invasive
microbes might be more harmful to the extant biota. This is
supported by evidence that microbes and pathogens that have
entered as passengers of domestic animals can have major
impacts on native wildlife (e.g. toxoplasmosis causal agent
Toxoplasma gondii on marsupial reproduction, Caneld et al.
1990; hydatidosis causal agent Echinococcus granulosus on
brush-tailed rock-wallabies, Petrogale penicillata, Barnes et al.
2007; and bovine tuberculosis causal agent Mycobacterium
tuberculosis in New Zealand brush-tailed possums, Nugent
et al.2001).
The fth hypothesis (Table 1) is an unlikely scenario for
Australasia. Most of the Australasian ora and fauna evolved
separately in isolation from eutherian carnivores and ungulates,
so Australia and New Zealand had no co-evolved prey or forage
species for the invaders when they were introduced. Any
co-evolved life-cycle-essential species likely immigrated with
them, for instance within their guts or on their backs. Co-evolved
plants were not necessary for the ungulate livestock and pests or
lagomorphs to ourish; they just needed to be similar structurally
and nutritionally to the co-evolved species, which supports the
null hypothesis H
The applicability of the sixth hypothesis depends upon H
and H
. If there are no pathogens or predators, and no niche
equivalents in the novel ecosystem, there is no reason to expect
rarity in the native environment to preclude invasiveness in the
novel environment. For example, koalas are usually uncommon
or rare, and are endangered in some native environments.
However, koalas can become invasive and destructive when
introduced into ecosystems that are naive to them (e.g. Kangaroo
Island, Masters et al.2004). An example from the plant world is
Cenchrus ciliaris L. which is becoming endangered in its native
Tunisia, but is an environmental weed in parts of Australia and
in Texas, USA (Kharrat-Souissi et al.2014).
Although Hypothesis 7 (Table 1) is intuitively satisfying, its
inverse is also likely. That is, when competing for a limiting
resource, an introduced species is more likely to prevail because
it may have competitive advantages aligned with other
hypotheses, such as resistance to pathogens and absence of
predators that affect the native species, advantageous life-history
characteristics, or temporal niche separation.
Given that anthropogenic activities, such as urbanisation,
forestry, agriculture and commercial grazing, usually disturb
environments at a similar time to the introduction of new plants
and animals, it is sometimes difcult to test whether disturbed
habitats are easier to invade than undisturbed habitats (H
see H
, Table 1). There are many examples in the weed literature
where disturbance favours invasive plant species, for example,
Table 1. Some hypotheses used to explain successful biological invaders and invasions
Hypothesis Null hypothesis Source
: A species will not be able to invade an area that has
abiotic conditions outside its physiological tolerances
: The physiological tolerances of a species are not
predictors of its success in a new environment
Prins and Gordon (2014a)
: The presence of fewer competitors enables successful
invasion and greater spread
: The extent of an invasion is neither positively or
negatively correlated with species diversity of functional
guild competitors in the invaded environment
Prins and Gordon (2014a)
: Existing equivalent niche occupants preclude
successful invasion
: A speciesinvasion success does not differ in the
presence or absence of a species that occupies an equivalent
niche and is in all other ways equivalent
Prins and Gordon (2014a)
: Extant diseases and predators that are novel to the
invader preclude successful invasion
: Previously un-encountered pathogens and predators do
not affect a success a speciesinvasion success
Prins and Gordon (2014a)
: Absence of co-evolved, lifecycle-essential species
precludes successful invasion
: There is no difference in the success of invasion for
a species invading with or without co-evolved species
necessary for its persistence in its native habitats
Prins and Gordon (2014a)
: Rare species in their native range are unlikely to be
invasive in new ecosystems
: The density of a species in its native range is not
a predictor of its density in a new environment
Prins and Gordon (2014a)
: A species cannot invade where a similar native species
has a competitive advantage through better efciency in
using a limited critical resource
: There will be no difference in the success of an
introduced and a native species competing for a limiting
Prins and Gordon (2014a)
: Disturbed habitats are easier to invade that undisturbed
: There will be no difference in the rate of invasion of
disturbed and undisturbed areas
Prins and Gordon (2014a)
: Invasive species are more likely to displace those of
older lineage that occupy similar niches
: The lineage age of extant species is not a predictor of its
displacement from its niche by an invasive species
Prins and Gordon (2014a)
: A more r-selected species is only able to invade where
extant niche occupants are less r-selected
: Life history strategy differences between extant and
novel species do not affect the likelihood of invasion by the
novel species
Prins and Gordon (2014a)
: The likelihood of a species being invasive is
unpredictable, happening by chance
: There is no null hypothesis that can be formulated from
this hypothesis. It is not possible to formulate reliable
predictors of a specieslikelihood of becoming invasive
Prins and Gordon (2014a)
: More generalist species are more likely to invade than
: There is no difference in the success of invasion for
generalist species over specialist species
This paper
: Early successional organisms are more likely to be
successful invaders than later ones
: The success of an invader is not determined by its
successional order
This paper
Invasive species impacts and restoration solutions The Rangeland Journal E
altered disturbance situations in wetlands affected the invasion of
native water couch communities (Paspalum distichum L.) by the
introduced forb lippia (Phyla canescens (Kunth) Greene, Price
et al.2011).
Seedling establishment of many species of plants depends
on disturbance of the existing vegetation to allow the
germination of seed and the establishment of the resultant
seedlings. This principle is well established (Scott 2000;
Sheppard 2000) and the elimination of establishment niches is
an important principle of weed management in pastures and
crops throughout the world. It is also clear that the rate of
invasion of weeds will be more rapid in disturbed than in
undisturbed areas (H
)(Scott2000;Sheppard2000). Also,
many weeds of pastures invade more rapidly under disturbance
by grazing because their low acceptability by livestock
increases their competitive advantage over native or sown
acceptable grassland species (Medd et al.1987). However, this
may be due to the life-history characteristics of the invaders or
coincidence of the introduction and the preparation for its
introduction, particularly when considering the invasiveness
of pasture plants and the ungulates that eat them.
Australia, being ancient and isolated for extended geological
time, provides a site to test the ninth hypothesis that species of
older lineage are more likely to be displaced than newer species
(Table 1). However, separating the lineage factor from other
likely coincident factors, such as those associated with niche
overlap and separation, would be difcult.
A more r-selected species is only able to invade where extant
niche occupants are less r-selected (H
, Table 1). This hypothesis
may explain why bilby species (Macrotis spp.) were usurped
by expanding populations of rabbits, which have a greater
reproductive capacity. Additionally, we suggest that a species
with a exible life-history, that is, one that can be r-selected or
K-selected (Krebs 2014) depending on circumstances, is more
likely to be invasive. Among plants, those with back-up
reproductive strategies, such as cleistogenes in Chilean needle
grass, Nassella neesiana, (Trin. & Rupr.) Barkworth, a pasture
invader in Australasia (BourdôT and Hurrell 1989; Gardener et al.
2003), or high fecundity, such as giant Parramatta grass,
Sporobolus fertilis (Stued.) Clayton (Andrews et al.1996), are
likely to have a competitive advantage after disturbances.
The likelihood of a species being invasive is unpredictable,
happening by chance (H
, Table 1). Prins and Gordon (2014a)
suggested this as a possible null hypothesis for all the others, and,
hence, formulating a null H
is problematic. Any falsication
of this hypothesis indicates that one or more of the previous
hypotheses must explain the likelihood of invasion by a species.
In addition to the hypotheses of Prins and Gordon (2014a), we
suggest the nal two hypotheses, but acknowledge that they
require testing for conrmation. We propose these hypotheses in
the light of our observations while studying vertebrate pests and
weeds: for animals, species with generalist tendencies are more
likely to invade than specialists (H
, Table 1), and for plants,
early successional plants are more likely to be successful invaders
than later ones (H
, Table 1). The latter hypothesis is likely
whether successional vegetation change is assumed to be a linear
process or a state and transition model (e.g. Westoby et al.1989)
applies, as is the case for Australian grassland systems (Lodge and
Whalley 1989; Whalley 1994).
Biological invasions of Australasia
Most ecosystems in Australasia have been affected by human
activities, including the facilitation, whether purposeful or
accidental, of biological invasions. Many of these are agri-
ecosystems (i.e. those that are currently used for or inuenced
by agricultural activities or previously used for agriculture).
Although only 6% of Australias land mass is arable, 53% of the
total area is currently used for extensive livestock production
(Australian Bureau of Statistics 2012) and much of the remainder
has, at one time, had sheep or cattle grazed there. Some
conservation reserves, even world heritage areas such as Kakadu
National Park, have long-established populations of feral
ungulates (e.g. feral goats, Capra hircus, Parkes 1993; Russell
et al.2011; feral water buffalo, Bubalus bubalis; feral horses,
Equus caballus; banteng Bos javanicus, Edwards et al.2004)
and feral camels, Camelus dromedarius (McGregor and
Edwards 2010; Hart and Edwards 2016). Aboriginal and Maori
people also managed the land to increase productivity of desirable
ora and fauna, and, in consequence, most of the Australia and
New Zealand landscapes have been and are affected by human
activities and the ecosystem dynamics changed (Morton et al.
1995; Gamage 2011).
Australasian invasive animals
Four groups of eutherian mammals successfully invaded
Australia before 1788. First of these were the bats, which could y
across Wallaces Line (Hand et al.1994). These were followed
by rodent members of the Muridae during the late Miocene,
leading to the evolution of 66 native species present in 1788.
Humans arrived at least 65 000 years ago (Clarkson et al.2017)
and what invasive passengers and chattels they brought with
them is unknown. People were also responsible for the
introduction of dingoes, an ancient breed of Canis familiaris
introduced from South-East Asia ~4500 ybp (Jackson and
Groves 2015; Jackson et al.2017), that subsequently became
feral (Fleming et al.2014) and invaded most Australian
environments (Johnson and Letnic 2014). The Maori people
brought the now extinct Polynesian dog and Polynesian rats
(Rattus exulans) to New Zealand ~650700 ybp (Matisoo-Smith
and Robins 2004).
In Australia, from 1788 until 1998, established invasions
included 27 bird species, four reptiles, seven sh and one
amphibian, the cane toad (Bomford and Hart 2002). Jackson and
Groves (2015) list 33 established species of mammals and
Bomford and Hart (2002), 28. Europeans brought 28 mammals,
nine sh and six birds that have become established pests in
New Zealand over the past 200 years (Clout 2002).
Australasian naturalised invasive plants
The number of invasive native and alien ora species of
Australia (Groves 2002) and New Zealand (Webb et al.1988)
are close to parity in both countries. Some 2681 plant species
are recognised as naturalised in Australia and ~2200 in
New Zealand (Norton 2009). Not all naturalised species are
aggressively invasive but all affect Australasian ecosystems
to some degree. Most alien plants were deliberately introduced:
grasses and forbs for livestock production (Cook and Dias
2006) and many ornamental plants. In addition, others were
FThe Rangeland Journal P. J. S. Fleming et al.
introduced accidentally as passengers of immigrant people
and visitors.
Introduced plant species may remain in low abundance in
their new host country before suddenly increasing in abundance
and becoming a problem to manage. These have been given the
name sleeper weeds(Groves 2006) and it is often difcult to
predict whether a particular species will become a weed in the
future (Groves 2006).
So what can we do about invasive species?
A strategic approach: passive adaptive management
The most practical approach that has been brought to bear on the
seemingly intractable issues of managing invasive species is
the implementation of adaptive management (Holling 1978;
Walters and Hilborn 1978; Walters and Holling 1990). Adaptive
management permits the managers of pests and weeds to start with
a working model then iteratively improve it by systematically
acquiring reliable information to assist in decision making.
There are three types of adaptive management (Walters and
Hilborn 1978; Walters and Holling 1990). In order of increasing
information, inference and efciency these are evolutionary,
passive adaptive and active adaptive management. The latter
involves a formal experiment that is imposed on the managed
system. For example, a weed control strategy for an aquatic
system might involve a multifactorial experimental design where
treatments and their combinations are applied to replicate dams or
waterways and contrasted with untreated invaded ecosystems.
The relative efcacies and cost-efciencies of each tool and
strategy are measured to provide better information about how to
cost-effectively manage the problem. Importantly, the required
budget that is necessary to make a substantial reduction in
negative impacts, such as oxygen depletion and subsequent
change in aquatic faunal communities, can be determined.
Without such information the risk is that, in an effort to be doing
something, managers spend inadequate money to achieve real
improvements (e.g. water hyacinth, Eichhornia crassipes:
Villamagna and Murphy 2010).
In Australasia, the passive form of adaptive management
(Walters and Hilborn 1978; Walters and Holling 1990) is more
common than other types (Braysher 2017). It typically involves
the use of historical data to form a single working model of
management, which is modied as better information about
system ecology, control tool efcacy and application comes
to hand from a quasi-experimental framework. The system
is considered passive because a formal, controlled randomised
experimental intervention is often logistically or socially
impossible (e.g. wild canid management, Fleming et al.2014).
A major advantage of the process is increased ownership of
the invasive species issue and its solutions by stakeholders such
that their on-going involvement in the process is more likely
(Chapple et al.2011; Braysher 2017). A further development
of adaptive management principles for weeds is Integrated
Weed Management (Sindel 2000).
Fleming et al.(2014) and Braysher (2017)presented a simple
explanatory ow chart to lead managers through the principles
of strategic management in a passive adaptive management
framework (Fig. 5). Fleming et al.(2014) also provided detail
that is pertinent to practical management of invasive species,
including the crucial human dimension (McNeely 2001; Pejchar
and Mooney 2009).
The critical components of passive adaptive management are
problem denition, the development of equity and capacity, and
monitoring. The problem denition step is most important
because it identies what the problem is, where it is, where it
comes from, who has the problem, when it occurs and how critical
it is. A useful way of dening the issue of invasiveness to be
managed is to regard the positive, neutral and negative impacts
of a species and the situation in which it occurs (Jarman 1990;
Table 2after Allison 2011).
A major difculty in dening the issue for invasive plants
is that some introduced species only become invasive after
many years, i.e. sleeper weeds (Groves 2006). For example,
Coolatai grass (Hyparrhenia hirta (L.) Stapf) was introduced
into Northern NSW in the 1890s, was planted along roadsides for
erosion control in the 1940s and only became recognised as an
invasive weed in the late 1980s (Chejara et al.2015).
Step 1
Revise actions
Revise objectives
Raise capacity
Redefine the issue
Step 2
Step 3
Step 4
Step 5
Step 6 Step 7
equity &
Develop a
plan of
Evaluate the
Modify & progress
the plan
the plan &
No change necessary
Fig. 5. Strategic management owchart for preparing plans of action for
invasive species, starting at the top left and following the arrows. Rectangles
are steps in the planning process and ovals identify the levels at which
adaptations can occur on evaluation and review. In active adaptive
management, the experimental design is included in the plan of action (Step 4)
(adapted from Fleming et al.2014).
Invasive species impacts and restoration solutions The Rangeland Journal G
Determining the human capacity of stakeholders is an
important part of the issue denition stage. Human capacity to
act effectively is multi-faceted but three common determinants
are: knowledge of the problem and what to do to solve it,
sufcient nancial resources to do something useful about
the problem, and the time to enact the practical aspects of
management. If any of these three elements is missing,
management will be suboptimal and the invasive species will
continue to prosper. As an example, a comprehensive survey
of landholders revealed that the majority of residential
professional farmers in the Northern Tablelands were well aware
of the problem of serrated tussock (Nasella trichotoma (Nees)
Hack. Ex Arechav.) invasion and could recognise the species
(Ruttledge et al.2015). However, the majority of non-
professional or absentee farmers generally did not have this
knowledge or ability. In addition, the latter group almost
universally did not adopt biosecurity precautions to control the
spread of seeds by livestock, vehicles and machinery. The
prognosis, therefore, was that without effectively engaging
these people and increasing the knowledge component of their
capacity, the continued spread of this species on the Northern
Tablelands of NSW is inevitable (Ruttledge et al.2015).
Monitoring is the nal essential ingredient in effective
management of invasive species. Without monitoring of
management actions and the responses to them, no evidence of
regeneration, revegetation or restoration of ecological processes
and ecosystems can be demonstrated.
Invasive species research and management progress
The strategic approach outlined in Fig. 5encompasses iteratively
increasing knowledge about the invasive species issue and its
solutions. The study and management of invasive species, and,
indeed, the study of their management, broach many disciplines
(Fig. 6).
Knowledge development, in the many elds identied
(Fig. 6), is critical to dening the issue (Step 1, Fig. 5), building
capacity (Step 2, Fig. 5), as well as determining the appropriate
technologies and their application for restoring ecological
processes where invasive species have been detrimental (Steps 3
and 4, Fig. 5). Ecology, invasion biology, agricultural sciences,
conservation science, and human dimensions are central to this.
However, all the other disciplines are important, which indicates
that collaboration is essential both for useful research and
effective management of invasive species.
Managing invasive species to restore ecological processes
Conceptual undercurrents
There are conceptual undercurrents implied in the three themes of
the RRR conference when addressing invasive species impacts
on ecosystem processes. First let us dene the three themes:
restoration is the repairing of an ecosystem by moving it to a prior
desired state; regeneration means to generate again by re-
establishment of the desired state by generating it from
propagules within, and to revegetate is to reinstate vegetation with
propagules from outside the system (i.e. an active process).
Issues: restoration
The primary questions here are to which state is the agri-
ecosystem to be restored, and is restoration possible? Australasian
ecosystems before Aboriginal, Maori and European invasions
were dynamic (Johnson 2006). They now have new dynamics,
with abiotic drivers that are essentially similar to those that have
prevailed since the conclusion of the last Ice Age. However,
climate change is affecting temperatures, rainfall totals, intensity
and patterns, and will have consequences for Australasian
Some changes to biota have been and are desirable for human
wellbeing since European settlement. Agricultural production
primarily uses animal and plant species that are novel to
Australasian ecosystems. There were no ungulates in either
Australia or New Zealand until Europeans brought them,
although it is possible that Asian wild boar (Sus scrofa) were
introduced from South East Asia via Cape York before European
introduction of Asian and European domestic pigs (Gongora
et al.2004). Ungulates were behaviourally, reproductively and,
most importantly, morphologically different from any animals
Table 2. Dening invasive species according to their origin, impacts and effects on restoration of ecological processes (adapted from Allison 2011)
Species type Situation Characteristics and impacts
Problem exotic invasive
Not native to ecosystem Tolerant of or benet
from human disturbance and activities, e.g.
farming, urbanisation, livestock
production, vegetation clearing, increased
water availability and changed re regimes
Increases and spread through the ecosystem/s.
r0 Offspring
numerous and/or easily dispersed. Causes decline in desirable
species through competition, predation or parasitism. Can cause
changes to abiotic properties and hydrology, energy and nutrient
cycles. Restoration requires their removal or density reduction
Problem native invasive
Native to ecosystem. Tolerant of or benet
from human disturbance and activities
As for problem exotic species. Restoration requires their removal or
density reduction
Non-problem exotic species Not native to ecosystem. Can be tolerant of or
benet from human disturbance and
activities. Non-invasive
Reproduce and survive in the ecosystem. r0orr<0 Populations
exhibit normal dynamics or decline. Populations may facilitate
ecosystems but generally do not lead to decline of the ecosystem.
Restoration does not require their removal or density reduction.
Non-problem native species Native to ecosystem. Can be tolerant of or
benet from human disturbance and
activities. Non-invasive.
Reproduce and survive in the ecosystem.
r<0 Populations
exhibit normal dynamics or decline. Populations may facilitate
ecosystems but generally do not lead to decline of the ecosystem.
Restoration may require their encouragement and density
augmentation through regeneration or revegetation
HThe Rangeland Journal P. J. S. Fleming et al.
in Australasia. This led to impacts on soils and vegetation
(e.g. piospheres of adverse impacts around permanent waters,
Landsberg and Stol 1996) that were not evident from the
largest extant marsupial herbivores. The interaction of ungulate
domesticates with purposely introduced and naturalised forage
species, as well as preferred native forage plants, is an essential
component of protable livestock production. Do we include
these components of the new agri-ecosystem in the desired state
for restoration?
Eradication of invasive species is rare and mainly
demonstrated on islands (e.g. rodents on islands, Howald et al.
2007; feral cats and rabbits on Macquarie Island, (Robinson and
Copson 2014; Sindel et al.2017) where the scale is logistically
tractable. No established invasive vertebrate or plant has been
eradicated from a continent because efforts usually fail to meet
three essential criteria (Bomford and OBrien 1995). These are:
*The rate of removal must be greater than
rat all population
*Immigration must be zero; and
*All reproductive animals (or plants and their seeds) must be
at risk of control tool/s and strategies.
The nding and removal of the last few plants of a target
invasive plant species (i.e. a weed) in complex vegetation is
often impossible or prohibitively expensive (Gardener et al.
2010). The eradication of weeds can be more difcult than
eradication of invasive animals because of long-lived seed
banks that are impossible to nd and remove or render unviable.
In addition, the impact of the weed species on ecological
processes within the region of occurrence must be determined,
and if these effects are not an apocalyptic threat to biodiversity,
then the cost of eradication is hard to justify (Davis et al.2011).
Bomford and OBrien (1995) listed a further three preferred
*The animals (or plants and their propagules) can be monitored
at low densities;
*The discounted benet : cost analysis favours eradication over
ongoing suppression; and
*The socio-political environment is suitable.
The nal criterion is also critical for gaining acceptance of
weed and pest animal management technologies and strategies.
Apart from occasional lay and scientist contrarians (so described
by Simberloff 2011), it is likely that the prevailing societal
attitudes towards and perceptions of invasive species are
negative. If true, this attitude will aid acceptance of the need for
population reductions and eradications of invasive species for
ecological restorations.
If eradication is essential for restoration of ecological
processes, then in such cases restoration may be impossible. For
example, the regeneration of drooping she-oak, Allocasuarina
verticillata, and sweet bursaria, Bursaria spinosa, in arid
ecosystems in South Australia cannot occur while rabbits occur
at densities of 0.5 ha
(Bird et al.2012). Without recruitment
of the structural backbones of the ecosystem, it is destined for
degradation and local extinction of dependent native fauna.
Even livestock, as arguably desirable components of the new
ecosystem, might be negatively impacted upon because of loss
of shade and browse forage.
Issues: regeneration
Regeneration is a simpler concept; it is the encouragement
of generation to increase populations within the degraded
ecosystem. Invasive incursions by animals degrade systems
and often reduce seedbanks or reduce recruitment by preying on
juveniles. Invasive plants can outcompete and increase the
relative abundance of their propagules at the expense of the
extant natives or desirable introduced pasture species. In such
Invasion Biology
Invasive Species
Computer Science
Remote Sensing
Citizen Science
Media Studies
Environmental Psychology
Vocational Education & Training
Conservation Science
Restoration Ecology
Climate Science
Animal Science
Welfare Science
Fig. 6. The interrelationships between science and humanities disciplines associated with improving
invasive species management. Management of the dingo (pictured) is an example of a wicked
invasive species issue with many dimensions.
Invasive species impacts and restoration solutions The Rangeland Journal I
cases, removal or reduction of the population of the invader is
required to halt further degradation and permit regeneration.
Exclusion plots (e.g. Lunt et al.2007) and exclusion fencing
(e.g. Doupé et al.2010) also allow regeneration of both plant
and animal communities from remnant propagules, but have
received some criticism (Hayward and Kerley 2009). These
exclosures often demonstrate that invasion must be prevented for
regeneration to occur and that contemporaneous removal of
the invader is an essential component of regeneration. However,
it could be pragmatic to accept the new, altered state and less
regeneration of desired ecosystem components (i.e. containment
and asset protection in Fig. 1).
Issues: revegetation
This is also a deceptively simple concept; that is, the
reinstatement of desirable vegetation where it has been removed
or outcompeted. Revegetation can be prevented by invasive
native and alien animals, and invasive plants. In New Zealand, for
example, the oristic assemblage of native forests is altered by
selective browsing on preferred species by brush-tailed possums
and regeneration of preferred species is only possible when
considerable and regular control effort is exerted on the possums
(Gormley et al.2012). In those systems, continuous control
effort will be required to suppress possum populations to
maintain the forest species mix and dynamics (Gormley et al.
2012): such investment may not be possible in the long term
and the new forest structure, assemblage and dynamics after
selective browsing by possums might be inevitable.
The residual negative impacts of invasions, i.e. ecosystem
degradation, can also affect the likelihood of successful
regeneration. The re-establishment of Australian native plant
species in revegetation programs, whether trees, shrubs or
herbaceous species following landscape degradation, is often
very difcult, particularly on shrink-swell vertosols common in
parts of the continent (Watt and Whalley 1982a,1982b; Waters
et al.2000; Chivers and Raulings 2009; Mitchell et al.2015;
Talonia et al.2017). The seeds of many native Australian
grasses and other herbaceous species germinate readily but
the seedlings are initially slow growing and are very susceptible
to weed competition (Barrett-Lennard et al.1991; Waters et al.
2000; Chivers and Raulings 2009). In addition, these soils
provide an inhospitable environment for seedling establishment
as do the soils in many other degraded landscapes (Watt and
Whalley 1982a,1982b; Barrett-Lennard et al.1991; Talonia
et al.2017).
Additionally, introduced plants can provide the same
ecosystem services as extant natives: trees can provide shade
(Bird et al.1993), forage (e.g. nut crops for endangered regent
parrots; Tracey and Fleming 2008), nest sites and nutrient
recycling; shrubs can provide forage and cover; and grasses
and forbs provide forage eaten by native animals, insects and
livestock. Effectively, we have created new agri-ecosystems
consisting of native and naturalised species with neutral
effects, or that are benecial for agricultural production, and of
invasive native and introduced plants that are detrimental for
production and environmental values. The determination of
what states are most desirable requires value judgements: how
much regeneration is to be encouraged, is revegetation to be
active or indirect through suppression of invasive species, and
what ecosystem is to be restored?
Restoring ecological processes through regeneration and
revegetation encapsulates conservation activities and conservation
science is based on principles of population and community
ecology. The conservation of a population, community or
ecosystem that is potentially threatened by invasive species
requires rst that its persistence be evaluated, either by measuring
and monitoring changes, or by making observations of like
systems previously exposed to the biological invader.
Because the requirements for eradication are rarely met except
for plants and animals detected early on during an incursion
(Fig. 1, Bomford and OBrien 1995), most invasive species are
here to stay. Therefore, management to remove or reduce
impacts of invasive species will be ongoing. An adaptive
management approach provides a workable framework that is
readily adopted by stakeholders because it will involve them
in applying the best practice and lessons about how to better
deal with the problem animal or plant through the process.
We must decide whether the impacts of an invasive
species require rectication or acceptance as a new part of
the biota, naturalised and adding the ecological equivalent of
multiculturalism. These are new agri-ecosystems with naturalised
plants and animals that have a range of impacts from positive
through neutral to negative. As a rule of thumb, positive values
and impacts should be encouraged, negative impacts should
be discouraged (usually through population suppression and
exclusion) and no action is needed for neutral effects.
The human dimensions must be measured, evaluated and
encapsulated in any invasive species management program or
we will always fail (Gunderson 1999; Chapple et al.2011). If the
impacts are net detrimental (i.e. the sum of the decits outweighs
the sum of the benets, if any), Australasians must decide if the
best benet : cost ratio is achieved by protecting environment,
cultural and agricultural assets, as is suggested in the oft-cited
generalised invasion curve (Fig. 1), or by applying the substantial
investment required to push the state of the invasion from asset
protection back to eradication or containment as we suggest
(Fig. 4). We suspect that the marginal gains of this latter strategy,
which changes the investment from costly suppression to
manageable maintenance, will result in the best long-term
benet : cost ratio, more effective and sustainable restoration of
ecosystem processes, and better environmental, agricultural
and societal wellbeing.
Conicts of interest
The authors declare no conicts of interests.
This presentation and paper beneted from discussions with many people
over the years, including but not exclusive to: Jim Hone, Mike Braysher,
Glen Saunders, Andrew McConnachie, Kerinne Harvey, Mary Bomford, Bob
Harden, Quentin Hart, Ben Russell, Guy Robinson, Wal Whalley, Paul Meek,
Peter Caley, Tony Pople, John Fleming and the late David Choquenot.
The comments of the editor and two anonymous reviewers improved
the manuscript. The authors receive funding from the Invasive Animals
Cooperative Research Centre.
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Invasive species impacts and restoration solutions The Rangeland Journal M
... However, the voice of invasive animal ecology contended that the expectation of entirely "solving" an invasive species problem is unrealistic; they cannot be eradicated [69]. Additionally, the search for such a "solution" completely ignores the biological reasons for these animals being successful invaders [70,71]. Nevertheless, in most instances, instead of trying to exterminate or eliminate a particular species, people should aim to reduce the negative effects of those animals [72]. ...
... The voice of invasive animal ecology proposed that, as a concept, invasive animal management is quite simple, and that which actions are taken depend on the situation [75]. If an introduced animal is invasive, then it will invade, take over areas and suppress other species [70,76]. The management of a detrimental invasive species would involve taking certain decisions to suppress their numbers or exclude them. ...
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A public forum can reveal a wide range of perspectives on the ethical treatment of animals. This article describes how a panel of experts navigated through a discussion on the many and varied challenges of attempting to manage invasive and native fauna in Australia. The panel acknowledged the variety of these fauna, their effects on others and the consequences of control measures for three parties: animals, humans and the environment. The One Welfare concept has been developed to guide humans in the ethical treatment of non-human animals, each other and the environment. The forum accepted the need to consider this triple line, and exemplifies the merits of a One Welfare approach to discussions such as this. We used a series of questions about past, present and anticipated practices in wildlife control as the core of the panel discussion. We revealed five different but intersecting perspectives: conservation action, wildlife research, invasive animal ecology, mainstream animal protection and compassionate conservation. This article shows how understanding of lines of contention on various core topics can provide a framework for further discourse that may bear fruit in the form of One Welfare solutions.
... 7 Climate change and invasive alien plants: Distribution, changes, and impacts on agriculture Alien plant invasion is a growing concern in many parts of the world and can have a tremendous impact on global agricultural production (Fleming et al., 2018). Plant introduction is not a new phenomenon; the majority of the world's major crops important to agriculture have exotic origins and are beneficial to the people (Pimentel et al., 2001). ...
... Some introduced plants remain as key crops that contribute profoundly to global food production and food security, i.e., rice (Oryza sativa L.), corn (Zea mays L.), and wheat (Triticum spp.) (Ziska et al., 2011). However, some other introduced plants escape and establish in natural ecosystems and become invasive, with frequent harmful impacts to biodiversity and agricultural wealth (Fleming et al., 2018;Simberloff et al., 1997). Invasive alien plants (IAPs) are fast-spreading alien plants whose introduction and spread threatens biodiversity, ecosystem services, agricultural production, and human wellbeing. ...
Climate change is affecting many facets of our lives and livelihoods, and food production is one of them. While the world population continues to increase, agricultural land and food production are being impacted by climate change at an ever-increasing rate. This chapter looks at climate change and its impacts on agri-food systems and food production. It briefly looks at the science of climate change, some projections, including rainfall and temperature projections to the end of the 21st century, and then is followed by a discussion of impacts of several climate-change–related stressors on agri-food production. Some of the stressors discussed include extreme weather events, such as droughts, floods, cyclones, and heat waves; sea-level rise, including inundation and salinity; invasive alien plant species; pests and pathogens; and neglected and underutilized crop species.
... Unfortunately, eradication success stories are few (e.g., [3]), particularly if invasion is advanced or occupying a broad range of biomes (e.g., [4]). Some invasive species have come from locations where their physiological tolerances are well adapted to their new environment and have occupied niches that were not filled when they arrived, such as by the cane toad (Rhinella marina) in Australia [5]. Similarly, some introductions can prosper by slotting into niches between existing native species [6]. ...
... This includes the feral domestic cat (Felis catus), which is a ubiquitous predator throughout the globe [7]. In Australia, feral cats could fill a niche between quolls (Dasyurus spp.) and dingoes (Canis familiaris) [5]. Determining strategies for their management requires knowledge of their preferred habitat and, for animals, of the extent and frequency of locations they utilise for their everyday activities [8]. ...
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Feral cats are one of the most damaging predators on Earth. They can be found throughout most of Australia’s mainland and many of its larger islands, where they are adaptable predators responsible for the decline and extinction of many species of native fauna. Managing feral cat populations to mitigate their impacts is a conservation priority. Control strategies can be better informed by knowledge of the locations that cats frequent the most. However, this information is rarely captured at the population level and therefore requires modelling based on observations of a sample of individuals. Here, we use movement data from collared feral cats to estimate home range sizes by gender and create species distribution models in the Pilbara bioregion of Western Australia. Home ranges were estimated using dynamic Brownian bridge movement models and split into 50% and 95% utilisation distribution contours. Species distribution models used points intersecting with the 50% utilisation contours and thinned by spacing points 500 m apart to remove sampling bias. Male cat home ranges were between 5 km2 (50% utilisation) and 34 km2 (95% utilisation), which were approximately twice the size of the female cats studied (2–17 km2). Species distribution modelling revealed a preference for low-lying riparian habitats with highly productive vegetation cover and a tendency to avoid newly burnt areas and topographically complex, rocky landscapes. Conservation management can benefit by targeting control effort in preferential habitat.
... We also tried to assess the degree of invasion of the species using the example of the Emirate of Fujairah, compile a preliminary point map of the distribution of the species, and assess the main routes of invasion in its territory. According to our preliminary rough assessment of the presence of Prosopis in places of distribution in the Emirate of Fujairah, the species is at stage 3 of invasion according to the standards of the Australian Department of the Environment and Primary Industries of Victoria (Fleming et al., 2017); Prosopis can be completely eradicated and its populations can be monitored. ...
... A more detailed substantiation of the technique and examples of its possible application are planned in the publication of the article "Counteracting the Invasion of Prosopis juliflora (Mimosaceae) in Fujairah (UAE)" and will not be presented further in this article. This assessment methodology differs both from the methodology of the Australian Department of Environment and Primary Industries of Victoria (Fleming et al., 2017) and from the methodology for assessing invasion by the cover of a species determined using drones (Bekele et al., 2018). ...
The article analyzes the secondary area in the Emirate of Fujairah, as well as the peculiarities of seed dispersing, seed germination and early seedling development of Mesquite, or Prosopis juliflora - the alien species of Mimosaceae, which appeared on the territory of the Emirates in the twentieth century, and one of the first collected herbarium specimen is dated 1983. In the secondary area most often, the Mesquite is found in anthropogenic habitats: in gardens, as well as near roads, on waste grounds, less frequent on the streets of settlements, garden fences etc. In regions represented by herbarium collections, P. juliflora is successfully naturalized and creates stable self-sustaining populations. It has a complex of specific helio-mesomorphic features that allow it to take root successfully in relatively open moderately wet, and even dry or saline habitats and compete with native species of acacia ( Acacia tortilis, A. ehrenbergii ) and local prosopis ( P. cineraria ). Characteristics such as good germination, significant morphological, dimensional and temporal variability of premature individuals of P. juliflora , identified in this work, undoubtedly contribute to the successful naturalization of the species in the secondary range and its wide distribution throughout the Emirate. Due to the high aggressiveness of Mesquite, it is necessary to develop a method of dealing with this plant in the UAE, which will stop its uncontrolled settlement in the region. We have compiled a map of Mesquite distribution in Fujairah and surrounding areas, which clearly shows the scale of the disaster. A method for assessing invasion on a five-level scale based on reproductive success has been developed and applied. The structure of ecotopes at an early and middle stage of penetration of P. juliflora was analyzed using large wastelands (2 and 1.2 ha) in the village Mirbah and the city of Fujairah on the coast of the Gulf of Oman.
... Consequently, considerable focus has been afforded to the monitoring and management of this pest in recent years. As with all pest management, the monitoring of cat population change in response to control efforts provides key information for conservation agencies and groups about the success or otherwise of their treatments and enables improvements, planning and reporting (Braysher 2017;Fleming et al. 2017). ...
... Implications of such introductions in the case of land gastropods were studied for several taxa [1][2][3][4] regarding the possible impact on human health. A number of studies showed that alien species as newly established components of fauna can have an unpredictable impact on the local floras and faunas including parasites [5][6][7][8][9][10]. Thus, Cameron [11] attributed the decline of local gastropod fauna to the introduction and spread of alien gastropod species. ...
Purpose: The present study investigates the origin of Arion vulgaris slugs in the parks of Moscow city and their parasites. Methods: Snails and slugs inhabiting green areas of Moscow city were collected in the summer season of 2020 and examined on the presence of gastropod-associated nematodes and trematodes using morphological and molecular methods. Results: The presence of the alien slug species, Arion vulgaris, was recorded in several locations, and the mitochondrial gene-based analysis has shown that slug populations inhabited Moscow parks originated from West and Central Europe. Out of a total of 15 gastropod species examined, A. vulgaris was the only species infected by the nematode Alloionema appendiculatum Schneider, 1859, a larval parasite of molluscs. It is the first record of this nematode from the territory of the Russian Federation. COX1 mtDNA sequences of A. appendiculatum obtained from 3 populations of infected slugs were identical with those from Western and Central Europe similarly to their gastropod hosts thus indicating that the nematodes travelled with their hosts. No parasites dangerous for humans or animals were found. Conclusion: The complex life cycle of A. appendiculatum includes a free-living stage in soil which offers a source of infection for other potentially susceptible gastropod species but the capacity of A. appendiculatum to change hosts in local conditions needs to be further investigated. The particular susceptibility and tolerance of A. vulgaris to nematodes in our study was in concordance with earlier data while in contradiction with the enemy release hypothesis.
... Globalization has brought about the increased transboundary movement of alien invasive pests from their native ranges to completely new areas where they have since successfully established (Early et al. 2016;Fleming et al. 2017;Meyerson and Mooney 2007). In the absence of coevolved natural enemies and knowledge systems on their management, they often reproduce exponentially thus overwhelming the invaded areas. ...
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Fruit production in Sub-Saharan Africa is of paramount importance both socially and economically. Millions of farmers derive livelihoods from mango, avocado, citrus, cashew, and coconut farming, but native and alien invasive species constrain production The region’s capacity to contain invasives is weak due to the absence of national and institutional support systems for early detection, containment, eradication, or management of the pests. Climate change is expected to play a huge role in the influx of more alien invasive species and the shift of ecological requirements of some native species. Though a fair share of pre-and post-management pest management techniques for several insect pests has been developed, adoption and adaptation of the options are limited. Data on economic and social implications are largely lacking, making it challenging to implement informed policy decisions. The existence of the “Strategy for Managing Invasive Species in Africa 2021–2030” promises a paradigm shift in the management of invasives, from reactive thinking to coordinated proactive approaches. The uncoordinated deployment of management measures in the region and the lack of funding, play a negative role in managing the pests effectively. Prospects for enhanced future research are wide, and efforts are currently being channeled to Area-Wide-Integrated Pest Management in a bottom-up approach with stakeholders owning the process. Participatory development of technologies is also taking centre stage, paving the way for increased adoption and adaptation. Postharvest technologies promise to provide the adequate phytosanitary assurance required by countries importing fruit from Sub-Saharan Africa.
... For example, at early stages of an invasion, or at low population densities, costs or benefits are likely limited or may not yet be apparent and the cost-benefit ratio may change significantly once spatial spread or density increases (e.g. Shackleton et al. 2007;Wise et al. 2012;van Wilgen and Richardson 2014;Fleming et al. 2017;Ahmed et al. 2021). This may also include, for example, cultural connections to the IAS which at an early stage of the invasion likely have yet to be realized (Gaertner et al. 2016). ...
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In addition to being a major threat to biodiversity and ecosystem functioning, biological invasions also have profound impacts on economies and human wellbeing. However, the threats posed by invasive species often do not receive adequate attention and lack targeted management. In part, this may result from different or even ambivalent perceptions of invasive species which have a dual effect for stakeholders—being simultaneously a benefit and a burden. For these species, literature that synthesizes best practice is very limited, and analyses providing a comprehensive understanding of their economics are generally lacking. This has resulted in a critical gap in our understanding of the underlying trade-offs surrounding management efforts and approaches. Here, we explore qualitative trends in the literature for invasive species with dual effects, drawing from both the recently compiled InvaCost database and international case studies. The few invasive species with dual roles in InvaCost provide evidence for a temporal increase in reporting of costs, but with benefits relatively sporadically reported alongside costs. We discuss methods, management, assessment and policy frameworks dedicated to these species, along with lessons learned, complexities and persisting knowledge gaps. Our analysis points at the need to enhance scientific understanding of those species through inter- and cross-disciplinary efforts that can help advance their management.
... Moseby & Read 2006). Lethal control must remove enough cats to cause their rate of increase to become negative over time if it is to be effective and financially sustainable (Fleming et al. 2017). If the reduction in exposure of threatened animals to feral cats is consistently insufficient, all the investment is wasted and feral cats have been excluded or killed for no conservation gain. ...
The feral cat (Felis catus) is a key threat for many Australian native critical weight range animals (i.e. species of intermediate body mass between 35 and 5,500 g that are particularly susceptible to introduced predators) and estimates of cat abundance are required for assessing changes in population size. Camera trapping is a much used tool for monitoring and estimating population sizes, including with mark–resight techniques, for which the more robust estimators require individual identification. Many feral cats are individually marked, which potentially makes them suitable for such monitoring programmes. We sought to determine what proportion of cat images captured during a commonly used field deployment of camera traps could be individually identified, and whether aspects of camera trap deployment affected the rate of individual identification. Camera trap arrays were established in four conservation areas in south-west New South Wales, Australia, during 2017 (range 39–50 camera traps per site). The unlured camera traps were continuously deployed over 26 months, with five or 10 images captured per trigger. Where possible, cats were individually identified based on phenotypic characteristics. Over the deployment period (95,413 camera trap nights; CTN), we obtained 2.25 million images, of which 13,845 contained feral cats. Feral cat events (i.e. a series of images taken <5 minutes apart on the same camera trap) ranged from 0.004 to 0.047 events per CTN across the four conservation areas, with 85 individual cats identified. Depending on camera settings, few images could be assigned to a known individual (12.2–27.4% of feral cat events per site were of identifiable individuals). Minimum number known alive were 10–46 feral cats per site, with resultant quarterly densities ranging from 0.01 to 0.16 cats/km². With our current deployment, individual identification of feral cats was insufficient for estimating abundance or survival using individual mark–resight methods. Such deployment deficits limit the ecological conclusions that can be drawn from ours and similar studies.
The economics of plant and animal health protection influence country policies through rapidly evolving benefit-cost tradeoffs that are difficult to forecast. Increased threat of infestation by invasive species following novel trade pathways is one recent trend, being counteracted by advances in data analytics to target interventions on higher risk pathways. The availability of increasingly large, complicated datasets generated from daily enforcement of regulations are available to safeguarding analysts. These data resources used to monitor and evaluate pathways are increasingly available electronically with shorter time lags. But the efficacy of increased analytic capabilities requires a clear objective of what is optimal. Economic frameworks can help focus the analytics. For example, increased protection that costs more than the benefit generated is not efficient. Economic theory provides a systematic method with which to develop policy or to assess existing programs. This chapter provides basic economic concepts and examples relevant to biosecurity safeguarding.
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The taxonomic identity and status of the Australian Dingo has been unsettled and controversial since its initial description in 1792. Since that time it has been referred to by various names including Canis dingo, Canis lupus dingo, Canis familiaris and Canis familiaris dingo. Of these names C. l. dingo and C. f. dingo have been most often used, but it has recently been proposed that the Australian Dingo should be once again recognized as a full species—Canis dingo. There is an urgent need to address the instability of the names referring to the Dingo because of the consequences for management and policy. Therefore, the objective of this study was to assess the morphological, genetic, ecological and biological data to determine the taxonomic relationships of the Dingo with the aim of confirming the correct scientific name. The recent proposal for Canis dingo as the most appropriate name is not sustainable under zoological nomenclature protocols nor based on the genetic and morphological evidence. Instead we proffer the name C. familiaris for all free-ranging dogs, regardless of breed and location throughout the world, including the Australian Dingo. The suggested nomenclature also provides a framework for managing free-ranging dogs including Dingoes, under Australian legislation and policy. The broad principles of nomenclature we discuss here apply to all free-roaming dogs that coexist with their hybrids, including the New Guinea Singing Dog.
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The time of arrival of people in Australia is an unresolved question. It is relevant to debates about when modern humans first dispersed out of Africa and when their descendants incorporated genetic material from Neanderthals, Denisovans and possibly other hominins. Humans have also been implicated in the extinction of Australia’s megafauna. Here we report the results of new excavations conducted at Madjedbebe, a rock shelter in northern Australia. Artefacts in primary depositional context are concentrated in three dense bands, with the stratigraphic integrity of the deposit demonstrated by artefact refits and by optical dating and other analyses of the sediments. Human occupation began around 65,000 years ago, with a distinctive stone tool assemblage including grinding stones, ground ochres, reflective additives and ground-edge hatchet heads. This evidence sets a new minimum age for the arrival of humans in Australia, the dispersal of modern humans out of Africa, and the subsequent interactions of modern humans with Neanderthals and Denisovans.
Pest animals are but one of many factors that influence the desired outcome from managing natural resource based systems, whether for production or conservation purposes. Others include diseases, weeds, financial resources, weather and fire management. To be effective, an integrated and systematic approach is required, and the principles and strategic approach outlined in this book can also be used to plan and manage the damage due to other factors. Managing Australia's Pest Animals includes case studies of successful and unsuccessful pest management strategies and covers a range of topics, including the history of pest management, current best practice principles, and guidelines for planning and applying strategic pest management approaches to effectively reduce pest damage. This book is the first clear and comprehensive guide to best practice pest management in Australia and will benefit students and trainers of pest managers, landholders, people involved in natural resource management, and industry and government pest management staff.
The types of damage caused by wildlife are many and varied, and can be costly and far-reaching. Until now, there has been little effort to identify and evaluate generalities across that broad range of species, methods and topics. Wildlife Damage Control promotes principle-based thinking about managing impact. It documents and discusses the key principles underlying wildlife damage and its control, and demonstrates their application to real-life topics – how they have been used in management actions or how they could be tested in the future. It synthesises the wide but diffuse literature dealing with the impacts of vertebrate pests and encourages readers to adopt a more theoretical framework for thinking about pest impacts and ways to manage them. The book is organised around key principles that apply across species, rather than looking at individual species, and is damage-based not pest animal-based. Within each chapter there are exercises designed to help readers learn and evaluate key principles. Conservation biologists, ecologists and others involved in wildlife management will find the sections covering principles in biodiversity conservation, of production such as agriculture, and in human and animal health of real value.
The Antarctic region is one of the most inhospitable frontiers on earth for weed invasion. On Australia's world heritage sub-Antarctic Macquarie Island only three species of invasive weeds are well established (Poa annua L., Stellaria media (L.) Vill. and Cerastium fontanum Baumg.), although isolated occurrences of other species have been found and removed. These weed species are believed to have initially been introduced through human activity, a threat which is likely to increase, although strict biosecurity is in place. All three weeds are palatable and may have been suppressed to some extent by pest herbivore (rabbit) grazing. Given the high conservation value of Macquarie Island and threats to ecosystem structure and function from weed proliferation following rabbit eradication, well targeted invasive plant control management strategies are vital. We propose that a successful restoration program for Australia's most southerly rangeland ecosystem should integrate both control of non-native plants as well as non-native herbivores. Of the non-native plants, S. media may most easily be managed, if not eradicated, because of its more limited distribution. Little, however, is known about the soil seed bank or population dynamics after rabbit eradication, nor the effect of herbicides and non-chemical control methods in cold conditions. A current research project on this non-grass species is helping to fill these knowledge gaps, complementing and building on data collected in an earlier project on the ecology and control of the more widespread invasive grass, P. annua. With an interest in off-target herbicide impacts, our work also includes a study of the movement and fate of herbicides in the cold climate Macquarie Island soils. Research in such a remote, cold, wet and windy place presents a range of logistical challenges. Nevertheless, outcomes are informing the development of effective, low-impact control or eradication options for sub-Antarctic weeds.
We examined the potential of direct-seeding Eucalyptus species to revegetate the vertosol ('cracking clay') soils that characterise the floodplains of north-western New South Wales. We investigated the influence of sowing depth (0, 6, 12 and 20mm) and three soil-moisture scenarios (dry, moist and flooded) on seedling emergence of seedlings of six species of Eucalyptus with a range of seed sizes (E. blakelyi, E. camaldulensis, E. melanophloia, E. melliodora, E. pilligaensis and E. populnea). We used cracking clay soil from the region in a glasshouse environment. Seedling emergence was low despite high seed viability and provision of optimum temperatures and soil moisture conditions. All six species exhibited greatest emergence when sown at 0-6-mm depth, with seed size being less important than moisture (except under dry conditions) and proximity to the surface. Species responded differently to the three watering treatments. Eucalyptus melanophloia exhibited greatest emergence in the 'dry' watering treatment. The floodplain species, E. camaldulensis, E pilliganesis and E. populnea, had the greatest emergence under flood conditions. Eucalyptus blakelyi and E. melliodora exhibited intermediate emergence in relation to all three soil-moisture regimes. Although the direct seeding of these species in vertosol soils in the region may be successful on occasion, windows of opportunity will be infrequent and the planting of seedling tubestock will be more reliable for revegetation.