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: peter.ﬂeming@dpi.nsw.gov.au
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 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.
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 ﬂora 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 beneﬁted
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 deﬁne invasive species and outline and
discuss some theoretical underpinnings of invasive species
science and management in agri-ecosystems, which we deﬁne as
all anthropogenically modiﬁed ecosystems used for agriculture,
both intensive and extensive. We do this to differentiate from
the term, ‘agro-ecosystems’, which is often interpreted as
highly modiﬁed 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 www.publish.csiro.au/journals/trj
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 deﬁnition; some species have invasive or colonising
qualities but are regarded as beneﬁcial, for instance those
endemic plants that recolonise after perturbations such as ﬁre,
ﬂood, soil disturbance and erosion (Bazzaz 1979). Other
examples of invasive species that are usually considered to
be beneﬁcial 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.
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 ‘prevention’through ‘quarantine’and other
biosecurity measures to ‘asset-based protection’for 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-deﬁned 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
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 inﬂicted (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 species’density 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 ﬁtted 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 beneﬁt 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 efﬁcient 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
beneﬁt 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 efﬁcacious, 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.
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 falsiﬁcation 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 ﬁrst hypothesis (H
, Table 1) is tautological, and
more a limiting statement than a predictor of invasiveness.
Surprisingly, ﬁve 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 O’Brien
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
ﬁtted niches that were not ﬁlled 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
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-
speciﬁcity of many pathogens and the fact that anthropogenically
introduced mammals have all been eutherian, is counter to
(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)reﬂects 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, Canﬁeld 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
The ﬁfth hypothesis (Table 1) is an unlikely scenario for
Australasia. Most of the Australasian ﬂora 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 ﬂourish; 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
. 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 difﬁcult to test whether disturbed
habitats are easier to invade than undisturbed habitats (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
: A species’invasion 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 species’invasion 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 efﬁciency 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
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 species’likelihood 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
: Early successional organisms are more likely to be
successful invaders than later ones
: The success of an invader is not determined by its
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
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
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 difﬁcult.
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 ﬂexible 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 falsiﬁcation
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 ﬁnal two hypotheses, but acknowledge that they
require testing for conﬁrmation. 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 inﬂuenced
by agricultural activities or previously used for agriculture).
Although only 6% of Australia’s 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
ﬂora 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 ﬂy
across Wallace’s 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 ~650–700 ybp (Matisoo-Smith
and Robins 2004).
In Australia, from 1788 until 1998, established invasions
included 27 bird species, four reptiles, seven ﬁsh 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 ﬁsh 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 ﬂora 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
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 difﬁcult 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 efﬁciency 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 efﬁcacies and cost-efﬁciencies 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 modiﬁed as better information about
system ecology, control tool efﬁcacy 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 ﬂow 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 deﬁnition, the development of equity and capacity, and
monitoring. The problem deﬁnition step is most important
because it identiﬁes 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 deﬁning 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 difﬁculty in deﬁning 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).
Redefine the issue
Step 6 Step 7
Modify & progress
the plan &
No change necessary
Fig. 5. Strategic management ﬂowchart 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 deﬁnition 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,
sufﬁcient ﬁnancial 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 ﬁnal 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
Knowledge development, in the many ﬁelds identiﬁed
(Fig. 6), is critical to deﬁning 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
There are conceptual undercurrents implied in the three themes of
the RRR conference when addressing invasive species impacts
on ecosystem processes. First let us deﬁne 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).
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. Deﬁning 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 beneﬁt
from human disturbance and activities, e.g.
farming, urbanisation, livestock
production, vegetation clearing, increased
water availability and changed ﬁre regimes
Increases and spread through the ecosystem/s.
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 beneﬁt
from human disturbance and activities
As for problem exotic species. Restoration requires their removal or
Non-problem exotic species Not native to ecosystem. Can be tolerant of or
beneﬁt from human disturbance and
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
beneﬁt from human disturbance and
Reproduce and survive in the ecosystem.
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 proﬁtable livestock production. Do we include
these components of the new agri-ecosystem in the desired state
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 O’Brien 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 ﬁnding 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 difﬁcult than
eradication of invasive animals because of long-lived seed
banks that are impossible to ﬁnd 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 O’Brien (1995) listed a further three preferred
*The animals (or plants and their propagules) can be monitored
at low densities;
*The discounted beneﬁt : cost analysis favours eradication over
ongoing suppression; and
*The socio-political environment is suitable.
The ﬁnal 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
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.
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
Vocational Education & Training
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).
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 ﬂoristic 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 difﬁcult, 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
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 beneﬁcial 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 ﬁrst 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 O’Brien 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 rectiﬁcation 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 deﬁcits outweighs
the sum of the beneﬁts, if any), Australasians must decide if the
best beneﬁt : 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
beneﬁt : cost ratio, more effective and sustainable restoration of
ecosystem processes, and better environmental, agricultural
and societal wellbeing.
Conﬂicts of interest
The authors declare no conﬂicts of interests.
This presentation and paper beneﬁted 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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