ArticlePDF Available

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


Abstract and Figures

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.
Content may be subject to copyright.
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.
JThe Rangeland Journal P. J. S. Fleming et al.
Allison, S. K. (2011). The paradox of invasive species: do restorationists
worry about them too much or too little? In:Invasive and Introduced
Plants and Animals: Human Perceptions, Attitudes, and Approaches
to Management. (Eds I. D. Rotherham and R. Lambert.) pp. 265275.
(Earthscan Press: London, UK.)
Andrews, T. S., Whalley, R. D. B., and Jones, C. E. (1996). Seed production
and seedling emergence of giant Parramatta grass on the North Coast
of New South Wales. Australian Journal of Experimental Agriculture
36, 299308. doi:10.1071/EA9960299
Australian Bureau of Statistics (2012). 7121.0 Agricultural Commodities,
Australia, 201011. Available at:
(accessed 15 March 2017).
Barnes, T. S., Hinds, L. A., Jenkins, D. J., and Coleman, G. T. (2007).
Precocious development of hydatid cysts in a macropodid host.
International Journal for Parasitology 37, 13791389. doi:10.1016/
Barrett-Lennard, E. G., Frost, F., Vlahos, S., and Richards, N. (1991).
Revegetating salt-affected lands with shrubs. Journal of Agriculture
Western Australia 32, 124129.
Bazzaz, F. A. (1979). The physiological ecology of plant succession. Annual
Review of Ecology and Systematics 10, 351371. doi:10.1146/annurev.
Bird, P. R., Bicknell, D., Bulman, P. A., Burke, S. J. A., Leys, J. F., Parker,
J. N., van der Sommen, F. J., and Voller, P. (1993). The role of shelter
in Australia for protecting soils, plants and livestock. In:The Role
of Trees in Sustainable Agriculture: Review papers presented at the
Australian Conference, The Role of Trees in Sustainable Agriculture.
Albury, Victoria, Australia, October 1991. (Ed. R. T. Prinsley.) pp. 5986.
(Springer: Dordrecht, The Netherlands.)
Bird, P., Mutze, G., Peacock, D., and Jennings, S. (2012). Damage caused
by low-density exotic herbivore populations: the impact of introduced
European rabbits on marsupial herbivores and Allocasuarina and
Bursaria seedling survival in Australian coastal shrubland. Biological
Invasions 14, 743755. doi:10.1007/s10530-011-0114-8
Bomford, M., and Hart, Q. (2002). Non-indigenous vertebrates in Australia.
In:Biological Invasions: Economic and Environmental Costs of Alien
Plant, Animal and Microbe species. (Ed. D. Pimentel.) pp. 2544. (CRC
Press: New York.)
Bomford, M., and OBrien, P. (1995). Eradication or control for vertebrate
pests? Wildlife Society Bulletin 23, 249.
BourdôT, G. W., and Hurrell, G. A. (1989). Cover of Stipa neesiana Trin.
&Rupr. (Chilean needle grass) on agricultural and pastoral land near
Lake Grassmere, Marlborough. New Zealand Journal of Botany 27,
415420. doi:10.1080/0028825X.1989.10414122
Braysher, M. (2017). Managing Australias Pest Animals: A Guide to
Strategic Planning and Effective Management.(CSIRO Publishing:
Caneld, P. J., Hartley, W. J., and Dubey, J. P. (1990). Lesions of
toxoplasmosis in Australian marsupials. Journal of Comparative
Pathology 103, 159167. doi:10.1016/S0021-9975(08)80172-7
Caughley, G. (1980). Analysis of Vertebrate Populations.Reprinted with
corrections. (John Wiley and Sons: London, UK.)
Caughley, G., and Birch, L. C. (1971). Rate of increase. The Journal of
Wildlife Management 35, 658663. doi:10.2307/3799769
Chapple, R. S., Ramp, D., Bradstock, R. A., Kingsford, R. T., Merson, J. A.,
Auld, T. D., Fleming, P. J. S., and Mulley, R. C. (2011). Integrating
science into management of ecosystems in the Greater Blue Mountains.
Environmental Management 48, 659674. doi:10.1007/s00267-011-
Cheal, D., and Coman, B. (2003). Pest plants and animals. In:Ecology: an
Australian Perspective. (Eds P. Attiwill and B. Wilson.) pp. 457473.
(Oxford University Press: South Melbourne.)
Chejara, V. K., Kristiansen, P., Sindel, B. M., Johnson, S. B., Whalley,
R. D. B., and Nadolny, C. (2015). The biology of Australian weeds 64.
Hyparrhenia hirta (L.) Stapf. Plant Protection Quarterly 30,211.
Chivers, I. H., and Raulings, K. A. (2009). Australian Native Grasses:
A Manual for Sowing, Growing and Using Them.(Native Seeds P/L:
Claridge, A. W., Mills, D. J., Hunt, R., Jenkins, D. J., and Bean, J. (2009).
Satellite tracking of wild dogs in south-eastern mainland Australian
forests: Implications for management of a problematic top-order
carnivore. Forest Ecology and Management 258, 814822. doi:10.1016/
Clarkson, C., Jacobs, Z., Marwick, B., Fullagar, R., Wallis, L., Smith, M.,
Roberts, R. G., Hayes, E., Lowe, K., Carah, X., Florin, S. A., McNeil, J.,
Cox, D., Arnold, L. J., Hua, Q., Huntley, J., Brand, H. E. A., Manne, T.,
Fairbairn, A., Shulmeister, J., Lyle, L., Salinas, M., Page, M., Connell, K.,
Park, G., Norman, K., Murphy, T., and Pardoe, C. (2017). Human
occupation of northern Australia by 65,000 years ago. Nature 547,
306310. doi:10.1038/nature22968
Clout, M. N. (2002). Ecological and economic costs of alien vertebrates in
New Zealand. In:Biological Invasions: Economic and Environmental
Costs of Alien Plant, Animal and Microbe Species. (Ed. D. Simberloff.)
pp. 185193. (CRC Press: New York.)
Cook, G. D., and Dias, L. (2006). It was no accident: deliberate plant
introductions by Australian government agencies during the 20th century.
Australian Journal of Botany 54,601625. doi:10.1071/BT05157
Darwin, C. (1859) (republished 1998). On the Origin of Species by Means
of Natural Selection or, the Preservation of Favoured Races in the
Struggle for Life.(Wordsworth Editions: Ware, Hertfordshire, UK.)
Davis, M. A., Chew, M. K., Hobbs, R. J., Lugo, A. E., Ewel, J. J., Vermeij,
G. J., Brown, J. H., Rosenzweig, M. L., Suding, K. N., Ehrenfeld, J. G.,
Grime, J. P., Mascaro, J., and Briggs, J. C. (2011). Dont judge species on
their origins. Nature 474, 153154. doi:10.1038/474153a
Department of Environment and Primary Industries Victoria (2017).
Protecting Victoria from Pest Animals.(The State of Victoria:
Melbourne, Vic.)
Doupé, R. G., Mitchell, J., Knott, M. J., Davis, A. M., and Lymbery, A. J.
(2010). Efcacy of exclusion fencing to protect ephemeral oodplain
lagoon habitats from feral pigs (Sus scrofa). Wetlands Ecology and
Management 18,6978. doi:10.1007/s11273-009-9149-3
Edwards, G. P., Pople, A. R., Saalfeld, K., and Caley, P. (2004). Introduced
mammals in Australian rangelands: future threats and the role of
monitoring programmes in management strategies. Austral Ecology 29,
4050. doi:10.1111/j.1442-9993.2004.01361.x
Elton, C. S. (1958). The Ecology of Invasion by Plants and Animals.
(Methguen: London, UK.)
Fleming, P., Corbett, L., Harden, B., and Thomson, P. (2001). Managing the
Impacts of Dingoes and Other Wild Dogs.(Bureau of Rural Sciences:
Canberra, ACT.)
Fleming, P. J. S., Allen, B. L., Allen, L. R., Ballard, G., Bengsen, A. J., Gentle,
M. N., McLeod, L. J., Meek, P. D., and Saunders, G. R. (2014).
Management of wild canids in Australia: free-ranging dogs and red foxes.
In:Carnivores of Australia: Past, Present and Future. (Eds A. S. Glen
and C. R. Dickman.) pp. 105149. (CSIRO Publishing: Melbourne.)
Gamage, B. (2011). The Biggest Estate on Earth: How Aborigines made
Australia.(Allen & Unwin: Sydney, NSW.)
Gardener, M. R., Whalley, R. D. B., and Sindel, B. M. (2003). Ecology of
Nassella neesiana, Chilean needle grass, in pastures on the Northern
Tablelands of New South Wales. I. Seed production and dispersal. Crop
&Pasture Science 54, 613619. doi:10.1071/AR01075
Gardener, M. R., Cordell, S., Anderson, M., and Tunnicliffe, R. D. (2010).
Evaluating the long-term project to eradicate the rangeland weed
Martynia annua L.: linking community with conservation. The Rangeland
Journal 32, 407416. doi:10.1071/RJ10029
Gause, G. F. (1934). The Struggle for Existence.(Williams and Wilkins:
Baltimore, MD.)
Invasive species impacts and restoration solutions The Rangeland Journal K
Gongora, J., Fleming, P., Spencer, P. B. S., Mason, R., Garkavenko, O.,
Meyer, J. N., Droegemueller, C., Lee, J. H., and Moran, C. (2004).
Phylogenetic relationships of Australian and New Zealand feral pigs
assessed by mitochondrial control region sequence and nuclear
GPIP genotype. Molecular Phylogenetics and Evolution 33, 339348.
Gormley, A. M., Holland, E. P., Pech, R. P., Thomson, C., and Reddiex, B.
(2012). Impacts of an invasive herbivore on indigenous forests. Journal
of Applied Ecology 49, 12961305. doi:10.1111/j.1365-2664.2012.
Groves, R. H. (2002). The impacts of alien plants in Australia. In:Biological
Invasions: Economic and Environmental Costs of Alien Plant, Animal
and Microbe Species. (Ed. D. Simberloff.) pp. 1123. (CRC Press:
New York.)
Groves, R. H. (2006). Are some weeds sleeping? Some concepts and reasons.
Euphytica 148, 111120. doi:10.1007/s10681-006-5945-5
Gunderson, L. (1999). Resilience, exibility and adaptive management-
antidotes for spurious certitude? Conservation Ecology 3, 7. www.
Hand, S. J., Novacek, M., Godthelp, H., and Archer, M. (1994). First Eocene
bat from Australia. Journal of Vertebrate Paleontology 14, 375381.
Hart, Q., and Edwards, G. (2016). Guest Editorial: outcomes of the
Australian feral camel management project. The Rangeland Journal
38,iiii. doi:10.1071/RJ16028
Hayward, M. W., and Kerley, G. I. H. (2009). Fencing for conservation:
Restriction of evolutionary potential or a riposte to threatening
processes? Biological Conservation 142,113. doi:10.1016/j.biocon.
Higgins, S. I., and Richardson, D. M. (1999). Predicting plant migration rates
in a changing world: The role of long-distance dispersal. American
Naturalist 153, 464475.
Holling, C. S. (1978). Adaptive Environmental Assessment and
Management.Wiley International Series on Applied Systems Analysis,
3. (John Wiley & Sons: London, UK.)
Hone, J. (1994). Analysis of Vertebrate Pest Control.(Cambridge
University Press: Cambridge, UK.)
Hone, J. (2007). Wildlife Damage Control.(CSIRO Publishing:
Howald, G., Donlan, C., Galván, J. P., Russell, J. C., Parkes, J., Samaniego, A.,
Wang, Y., Veitch, D., Genovesi, P., Pascal, M., and Saunders, A. (2007).
Invasive rodent eradication on islands. Conservation Biology 21,
12581268. doi:10.1111/j.1523-1739.2007.00755.x
Jackson, S. M., and Groves, C. P. (2015). Dingo. In:Taxonomy
of Australian Mammals. (Eds S. M. Jackson and C. P. Groves.)
pp. 287289. (CSIRO Publishing: Melbourne.)
Jackson, S. M., Groves, C. P., Fleming, P. J. S., Aplin, K. P., Eldridge,
M. D. B., Gonzalez, A., and Helgen, K. M. (2017). The wayward
dog: Is the Australian native dog or Dingo a distinct species? Zootaxa
4317, 201224. doi:10.11646/zootaxa.4317.2.1
Jarman, P. (1990). Bird pest research: the gap between research and
application. In:National Bird Post Workshop Proceedings. University
of New England, Armidale, 89 February 1990. (Eds P. Fleming,
I. Temby and J. Thompson.) pp. 712. (NSW Agriculture & Fisheries:
Glen Innes, NSW.)
Johnson, C. (2006). Australias Mammal Extinctions: a 50000-year History.
(Cambridge University Press: Cambridge, UK.)
Johnson, C. N., and Letnic, M. (2014). Introducing a new top predator,
the dingo. In:Invasion Biology and Ecological Theory: Insights from
a Continent in Transformation. (Eds H. H. T. Prins and I. J. Gordon.)
pp. 414428. (Cambridge University Press: Cambridge, UK.)
Kharrat-Souissi, A., Siljak-Yakovlev, S., Brown, S. C., Baumel, A., Torre, F.,
and Chaieb, M. (2014). The polyploid nature of Cenchrus ciliaris L.
(Poaceae) has been overlooked: new insights for the conservation
and invasion biology of this species a review. The Rangeland Journal
36,1123. doi:10.1071/RJ13043
Krebs, C. J. (2014). Ecology: The Experimental Analysis of Distribution
and Abundance.6th edn. (Pearson Education Limited: Harlow, UK.)
Landsberg, J., and Stol, J. (1996). Spatial distribution of sheep, feral goats
and kangaroos in woody rangeland paddocks. The Rangeland Journal
18, 270291. doi:10.1071/RJ9960270
Lodge, G. M., and Whalley, R. D. B. (1989). Native and natural pastures on
the Northern Slopes and Tablelands of New South Wales a review and
annotated bibliography.Technical Bulletin No. 35. (NSW Agriculture
and Fisheries: Orange, NSW.)
Lunt, I. D., Jansen, A. M. Y., Binns, D. L., and Kenny, S. A. (2007). Long-
term effects of exclusion of grazing stock on degraded herbaceous
plant communities in a riparian Eucalyptus camaldulensis forest in south-
eastern Australia. Austral Ecology 32, 937949. doi:10.1111/j.1442-
Masters, P., Duka, T., Berris, S., and Moss, G. (2004). Koalas on Kangaroo
Island: from introduction to pest status in less than a century. Wildlife
Research 31, 267272. doi:10.1071/WR03007
Matisoo-Smith, E., and Robins, J. H. (2004). Origins and dispersals of
Pacic peoples: evidence from mtDNA phylogenies of the Pacic rat.
Proceedings of the National Academy of Sciences of the United States
of America 101, 91679172. doi:10.1073/pnas.0403120101
McGregor, M., and Edwards, G. (2010). Guest Editorial: managing the
impacts of feral camels. The Rangeland Journal 32,iiii. doi:10.1071/
McNeely, J. A. (2001). The great reshufing: human dimensions of invasive
alien species.(IUCN, Biodiversity Policy Coordination Division: Gland,
Switzerland.) Available at:
(accessed 15 March 2017).
McNeely, J. A. (2011). Xenophobia or conservation: some human
dimensions of invasive alien species. In:Invasive and Introduced
Plants and Animals: Human Perceptions, Attitudes and Approaches
to Management. (Eds I. D. Rotherham and R. A. Lambert.) pp. 1936.
(Earthscan: London, UK.)
Medd, R. W., Kemp, D. R., and Auld, B. A. (1987). Management of weeds
in perennial pastures. In:Temperate Pastures, their Production, Use and
Management. (Eds J. L. Wheeler, C. J. Pearson and G. E. Robards.)
pp. 253261. (CSIRO Publishing: Melbourne.)
Mitchell, M. L., Norman, H. C., and Whalley, R. D. B. (2015). Use of
functional traits to identify Australian forage grasses, legumes and
shrubs for domestication and use in pastoral areas under a changing
climate. Crop & Pasture Science 66,7189. doi:10.1071/CP13406
Morton, S. R., Stafford-Smith, D. M., Friedel, M. H., Grifn, G. F., and
Pickup, G. (1995). The stewardship of arid Australia: Ecology and
landscape management. Journal of Environmental Management 43,
195217. doi:10.1016/S0301-4797(95)90402-6
Nathan, R., and Muller-Landau, H. C. (2000). Spatial patterns of seed
dispersal, their determinants and consequences for recruitment. Trends in
Ecology & Evolution 15,278285. doi:10.1016/S0169-5347(00)01874-7
Norton, D. A. (2009). Species invasions and the limits to restoration:
learning from the New Zealand experience. Science 325, 569571.
Nugent, G., Fraser, W., and Sweetapple, P. (2001). Top down or bottom up?
Comparing the impacts of introduced arboreal possums and terrestrial
ruminants on native forests in New Zealand. Biological Conservation
99,6579. doi:10.1016/S0006-3207(00)00188-9
Parkes, J. P. (1993). Feral goats: designing solutions for a designer pest.
New Zealand Journal of Ecology 17,7183.
Pejchar, L., and Mooney, H. A. (2009). Invasive species, ecosystem services
and human well-being. Trends in Ecology & Evolution 24, 497504.
Pimentel, D. (2002). Introduction: non-native species in the world. In:
Biological Invasions: Economic and Environmental Costs of Alien
LThe Rangeland Journal P. J. S. Fleming et al.
Plant, Animal and Microbe Species. (Ed. D. Pimentel.) pp. 38. (CRC
Press: Boca Raton, FL.)
Price, J. N., Berney, P. J., Ryder, D., Whalley, R. D. B., and Gross, C. L.
(2011). Disturbance governs dominance of an invasive forb in
a temporary wetland. Oecologia 167, 759769. doi:10.1007/s00442-011-
Prins, H. H. T., and Gordon, I. J. (2014a). Testing hypotheses about
biological invasions and Charles Darwins two-creators rumination. In:
Invasion Biology and Ecological Theory: Insights from a Continent
in Transformation. (Eds H. H. T. Prins and I. J. Gordon.) pp. 119.
(Cambridge University Press: Cambridge, UK.)
Prins, H. H. T., and Gordon, I. J. (2014b). A critique of ecological theory
and a salute to natural history. In:Invasion Biology and Ecological
Theory: Insights from a Continent in Transformation. (Eds H. H. T. Prins
and I. J. Gordon.) pp. 497516. (Cambridge University Press:
Cambridge, UK.)
Robinson, S. A., and Copson, G. R. (2014). Eradication of cats (Felis catus)
from subantarctic Macquarie Island. Ecological Management &
Restoration 15,3440. doi:10.1111/emr.12073
Robley, A., Gormley, A., Forsyth, D. M., Wilton, A. N., and Stephens, D.
(2010). Movements and habitat selection by wild dogs in eastern Victoria.
Australian Mammalogy 32,2332. doi:10.1071/AM09030
Rotherham, I. D., and Lambert, R. A. (2011). Balancing species history,
human culture and scientic insight: introduction and overview. In:
Invasive and Introduced Plants and Animals: Human Perceptions,
Attitudes and Approaches to Management. (Eds I. D. Rotherham and
R. A. Lambert.) pp. 318. (Earthscan: London, UK.)
Russell, B. G., Letnic, M., and Fleming, P. J. S. (2011). Managing feral
goat impacts by manipulating their access to water in the rangelands. The
Rangeland Journal 33, 143152. doi:10.1071/RJ10070
Ruttledge, A., Whalley, R. D. B., Reeve, I., Backhouse, D. A., and Sindel,
B. M. (2015). Preventing weed spread: a survey of lifestyle and
commercial landholders about Nassella trichotoma in the Northern
Tablelands of New South Wales, Australia. The Rangeland Journal
37, 409423. doi:10.1071/RJ15010
Sarre, S. D., Aitken, N., Adamack, A. T., MacDonald, A. J., Gruber, B., and
Cowan, P. (2014). Creating new evolutionary pathways through
bioinvasion: the population genetics of brushtail possums in New
Zealand. Molecular Ecology 23, 34193433. doi:10.1111/mec.12834
Scott, J. K. (2000). Weed invasion, distribution and succession. In:Australian
Weed Management Systems. (Ed. B. M. Sindel.) pp. 1938. (R.G. and
F.J. Richardson: Melbourne.)
Seki, M., Kamei, A., Yamaguchi-Shinozaki, K., and Shinozaki, K. (2003).
Molecular responses to drought, salinity and frost: common and different
paths for plant protection. Current Opinion in Biotechnology 14, 194199.
Sheppard, A. W. (2000). Weed ecology and population dynamics. In:
Australian Weed Management Systems. (Ed. B. M. Sindel.) pp. 3960.
(R.G. and F.J. Richardson: Melbourne.)
Sibly, R. M., and Hone, J. (2002). Population growth rate and its determinants:
an overview. Philosophical Transactions of the Royal Society of London.
Series B, Biological Sciences 357, 11531170. doi:10.1098/
Simberloff, D. (2011). The rise of modern invasion biology and American
attitudes towards introduced species. In:Invasive and Introduced
Plants and Animals: Human Perceptions, Attitudes and Approaches to
Management. (Eds I. D. Rotherham and R. A. Lambert.) pp. 121135.
(Earthscan: London, UK.)
Sindel, B. M. (2000). The history of integrated weed management.
In:Australian Weed Management Systems. (Ed. B. M. Sindel.)
pp. 253265. (R.G. and F.J. Richardson: Melbourne.)
Sindel, B. M., Kristiansen, P. E., Wilson, S. C., Shaw, J. D., and Williams,
L. K. (2017). Managing native plants on sub-Antarctic Macquarie Island.
The Rangeland Journal 39, In press.
Talonia, L. R., Reid, N., Gross, C. L., and Whalley, R. D. B. (2017).
Germination ecology of six species of Eucalyptus in shrinkswell
vertosols: moisture, seed depth and seed size limit seedling emergence.
Australian Journal of Botany 65,2230. doi:10.1071/BT16155
Thomson, P. C. (1992). The behavioural ecology of dingoes in north-western
Australia: IV. Social and spatial organisation, and movements. Wildlife
Research 19, 543563. doi:10.1071/WR9920543
Tracey, J., and Fleming, P. (2008). Regent parrots and almonds: balancing
conservation and production issues. Australian Nutgrower 22,2930.
Villamagna, A. M., and Murphy, B. R. (2010). Ecological and socio-
economic impacts of invasive water hyacinth (Eichhornia crassipes):
a review. Freshwater Biology 55, 282298. doi:10.1111/j.1365-2427.
Vitousek, P. M., DAntonio, C. M., Loope, L. L., Rejm, X. C., Nek, M., and
Westbrooks, R. (1997). Introduced species: A signicant component
of human-caused global change. New Zealand Journal of Ecology
Walters, C. J., and Hilborn, R. (1978). Ecological optimization and adaptive
management. Annual Review of Ecology and Systematics 9, 157188.
Walters, C. J., and Holling, C. S. (1990). Large-scale management
experiments and learning by doing. Ecology 71, 20602068. doi:10.2307/
Waters, C. M., Whalley, R. D. B., and Huxtable, C. H. A. (2000). Grassed
Up (A guideline for revegetating with Australian native grasses).(NSW
Agriculture: Dubbo, NSW.)
Watt, L. A., and Whalley, R. D. B. (1982a). Establishment of small-seeded
perennial grasses on black clay soils in North-western N.S.W. Australian
Journal of Botany 30, 611623. doi:10.1071/BT9820611
Watt, L. A., and Whalley, R. D. B. (1982b). Effect of sowing depth and
seedling morphology on establishment of grasses seedlings on cracking
black earths. Australian Rangeland Journal 4,5260. doi:10.1071/
Webb, C. J., Sykes, W. R., and Garnock-Jones, P. J. (1988). Flora of
New Zealand Vol. 4. Naturalized Pteridophytes, Gymnosperms,
Dicotyledons.(Landcare New Zealand: Christchurch.) Available
at:, http://oraseries.landcare (accessed 15 March 2017).
Westoby, M., Walker, B., and Noy-Meir, I. (1989). Opportunistic
management for rangelands not at equilibrium. Journal of Range
Management 42, 266274. doi:10.2307/3899492
Whalley, R. D. B. (1994). State and transitions models for rangelands. 1.
Successional theory and vegetation change. Tropical Grasslands 28,
Whalley, R. D. B., Price, J. N., Macdonald, M. J., and Berney, P. J. (2011).
Drivers of change in the Social-Ecological Systems of the Gwydir
Wetlands and Macquarie Marshes in northern New South Wales,
Australia. The Rangeland Journal 33, 109119. doi:10.1071/RJ11002
Whisson, D. A., Holland, G. J., and Carlyon, K. (2012). Translocation of
overabundant species: implications for translocated individuals. The
Journal of Wildlife Management 76, 16611669. doi:10.1002/jwmg.401
Williams, K., Parer, I., Coman, B., Burley, J., and Braysher, M. (1995).
Managing Vertebrate Pests: Rabbits.(Bureau of Resource Sciences
and CSIRO Division of Wildlife and Ecology, Australian Government
Publishing Service: Canberra, ACT.)
Invasive species impacts and restoration solutions The Rangeland Journal M
... However, the motivation for this type of killing is distinct from other forms of animal killing. Animals might be killed by humans simply because they are exotic or "not from here" (van Eeden et al. 2020), but they are usually killed because their invasive characteristics and traits (Elton 1958) raise concern that they will cause subsequent issues that will require further and otherwise avoidable animal killing (Fleming et al. 2017;Callen et al. 2020). These concerns include the protection of human health and safety (Section 2.2), agricultural production (Section 2.3), threatened species protection (Section 2.6), or the prevention of ecosystem collapse or shifts characterised by the mass killing and loss of many local animals. ...
... Many invasive and overabundant animals create real and perceived undesirable impacts on the environment, human economies, and on social or cultural values (e.g. Witmer and Proulx 2010;Castorani and Hovel 2015;Doherty et al. 2016;Fleming et al. 2017;Diagne et al. 2021). These impacts include the harm, killing and death of relatively large numbers of other animals that could otherwise be alleviated and avoided by killing relatively small numbers of invasive and overabundant native animals (e.g. ...
Full-text available
Killing animals has been a ubiquitous human behaviour throughout history, yet it is becoming increasingly controversial and criticised in some parts of contemporary human society. Here we review 10 primary reasons why humans kill animals, discuss the necessity (or not) of these forms of killing, and describe the global ecological context for human killing of animals. Humans historically and currently kill animals either directly or indirectly for the following reasons: (1) wild harvest or food acquisition, (2) human health and safety, (3) agriculture and aquaculture, (4) urbanisation and industrialisation, (5) invasive, overabundant or nuisance wildlife control, (6) threatened species conservation, (7) recreation, sport or entertainment, (8) mercy or compassion, (9) cultural and religious practice, and (10) research, education and testing. While the necessity of some forms of animal killing is debatable and further depends on individual values, we emphasise that several of these forms of animal killing are a necessary component of our inescapable involvement in a single, functioning, finite, global food web. We conclude that humans (and all other animals) cannot live in a way that does not require animal killing either directly or indirectly, but humans can modify some of these killing behaviours in ways that improve the welfare of animals while they are alive, or to reduce animal suffering whenever they must be killed. We encourage a constructive dialogue that (1) accepts and permits human participation in one enormous global food web dependent on animal killing and (2) focuses on animal welfare and environmental sustainability. Doing so will improve the lives of both wild and domestic animals to a greater extent than efforts to avoid, prohibit or vilify human animal-killing behaviour.
... Measures have been introduced for the prevention and early detection of invasive species, but management tends to be reactive once the pest arrives and an outbreak is discovered. The first management practices are usually aimed at eradication, but if this is unsuccessful, the pest establishes and strategies switch to population control and slowing down the spread of the invasive species (Fleming et al. 2017;Harris et al. 2018;Robertson et al. 2020). ...
Full-text available
Non-native invasive arthropod species threaten biodiversity and food security worldwide, resulting in substantial economic, environmental, social and cultural costs. Classical biological control (CBC) is regarded as a cost-effective component of integrated pest management programmes to manage invasive arthropod pests sustainably. However, CBC programmes are traditionally conducted once a pest has established in a new environment, and invariably all research needed to achieve approval to release a biological control agent can take several years. During that time, adverse impacts of the pest accelerate. A pre-emptive biocontrol approach will provide the opportunity to develop CBC for invasive pests before they arrive in the country at risk of introduction and therefore enhance preparedness. A critical aspect of this approach is that risk assessment is carried out in advance of the arrival of the pest. Implementing pre-emptive biocontrol risk assessment means that natural enemies can be selected, screened in containment or abroad and potentially pre-approved prior to a pest establishing in the country at risk, thus improving CBC effectiveness. However, such an approach may not always be feasible. This contribution defines the fundamental prerequisites, principles, and objectives of pre-emptive biocontrol risk assessment. A set of guidelines and a decision framework were developed, which can be used to assess the feasibility of conducting a pre-emptive risk assessment for candidate biological control agents against high-risk arthropod pests.
... Invasive mammalian predators are global drivers of biodiversity decline and extinction, as well as livestock loss (Doherty et al. 2016;Fleming et al. 2017). Such predators thrive in novel landscapes and challenge ecosystems through predation, competition, hybridisation and disease (Daniels et al. 2001;Norbury 2001;Harris and Macdonald 2007;Wyatt et al. 2008;Spencer et al. 2017). ...
... Some studies have highlighted the plant characteristics, or the mechanisms used by naturalized woody plants to successfully establish and spread (Table 1). Invasive woody species are therefore a subset of naturalized plants that generally have a superior ability to occupy and maintain space, as well as the potential for rapid growth, particularly in non-resource-limited environments, establishing self-regenerating populations and have spread beyond their area of introduction (Richardson et al. 2000;Kairo et al. 2003;Leishman et al. 2007;Fleming et al. 2017;Shiferaw et al. 2019). Many of these invasive species can spread by vegetative propagation, an aspect that (a) secures colonization of unoccupied ground, (b) increases competitive power of the species within the community, and (c) increases survival rate in marginal habitats (Jeník 1994). ...
... Some studies have highlighted the plant characteristics, or the mechanisms used by naturalized woody plants to successfully establish and spread (Table 1). Invasive woody species are therefore a subset of naturalized plants that generally have a superior ability to occupy and maintain space, as well as the potential for rapid growth, particularly in non-resource-limited environments, establishing self-regenerating populations and have spread beyond their area of introduction (Richardson et al. 2000;Kairo et al. 2003;Leishman et al. 2007;Fleming et al. 2017;Shiferaw et al. 2019). Many of these invasive species can spread by vegetative propagation, an aspect that (a) secures colonization of unoccupied ground, (b) increases competitive power of the species within the community, and (c) increases survival rate in marginal habitats (Jeník 1994). ...
The occurrence and spread of invasive woody species are a truly global phenomenon, but tropical regions seem to be particularly vulnerable due to high rates of soil degradation in combination with climate change, and limited resources for containment. There is increasing awareness that complete eradication programs are often not efective. The existence of many “controversial species,” i.e., species with both negative and positive impacts, renders decision-making processes for management exceedingly complex. By providing a very extensive overview of the current state of knowledge on impacts and containment strategies of invasive woody species, we aim to help underpin such decisions. We discuss both negative impacts and potential benefts of invasive woody species, focusing on the two most important ones, namely animal fodder production and the positive impacts on soil functioning and soil quality. Invasive woody species can positively impact livestock production (1) indirectly by improving pasture quality because of improved soil quality and functioning, and (2) directly by supplying a high-quality protein component for animal fodder. Invasive woody species increase soil carbon sequestration and nitrogen and phosphorus availability depending on the density of the invader, its capacity to fx nitrogen, the quantity and quality of its litter, and the direct interactions between its roots and the soil microbial community. The balance between potential benefts and risks depends to a large extent on the interaction with the local environment (climate, soil, vegetation, and animals) and the socioeconomic context of each region. When an invasion process starts because there is no local predator, then management can target eradication or very strict containment. If the invasion is the result of strong disturbance of the ecosystem, then intensive but well-thought management of the invasive species would be the choice to be made, as this may help to restore the ecosystem.
... Vertebrate pests such as feral cats (Felis catus), European red foxes (Vulpes vulpes), feral pigs (Sus scrofa) and wild dogs (Canis familiaris) cause major negative impacts to agriculture and the environment (Braysher 1993;Fleming et al. 2017). Controlling introduced pests requires the integration of methods such as aerial and ground baiting to distribute toxic baits (Newsome 1990;Fleming et al. 2014;Ballard et al. 2020), trapping (Meek et al. 1995;Fleming et al. 1998;Meek et al. 2022) and the use of other devices (e.g. ...
... 'Compassionate Conservation' may offer compassion to some individuals of a limited group of taxa, but ultimately consigns many more individuals to an uncompassionate demise . It is a thinking that has only recently received critical attention (Driscoll and Watson, 2019;Fleming and Ballard 2018;Rohwer and Marris, 2019;Russell et al., 2016). ...
The 'Compassionate Conservation' movement is gaining momentum through its promotion of 'ethical' conservation practices based on self-proclaimed principles of 'first-do-no-harm' and 'individuals matter'. We argue that the tenets of 'Compassionate Conservation' are ideological-that is, they are not scientifically proven to improve conservation outcomes, yet are critical of the current methods that do. In this paper we envision a future with 'Compassionate Conservation' and predict how this might affect global biodiversity conservation. Taken
Plant invasion is the biggest challenge for ecologist that affects biodiversity and environmental health. Forecasting of invasive plant species, its identification, early detection and distribution mapping are necessary for making plan of actions against negative consequences of alien invasive species. An invasive plant affects biodiversity along with ecosystem health and services. However, very few studies are available on plant invasion dynamics and its impacts on ecosystem. Invasive plant invades natural ecosystem including forest and agriculture which affects soil, food and climate security. Human and animals are also affected by dynamic intervention of invasive species. Mostly, invasive species also fix atmospheric carbon (C) through C sequestration process which helps in mitigating C footprint and climate change issues. However, many invasive species change its distribution and other mechanisms under changing climate scenario. Climate change amplifies the population dynamics and diversity of invasive species which is a major ecological risk. However, it is needed for accurately prediction of invasive plant distributions and its varying impacts on desire species that can be changed under projected climate change scenarios. This study helps in understanding effective control and preventive measures against spreading of plant invasions. A sound scientific strategy and policy framework are required for invasive plant management which would be helpful in conservation of desired natural resources. A link must exist between local and global policy networks to address plant invasions in changing climate. Therefore, an effective policy framework and plan of actions are employed to control and prevent plant invasions which build ecological stability and environmental sustainability.
Full-text available
What sets this book apart is its emphasis on eco-friendly technology as a cornerstone of peatland restoration efforts. Readers will delve into an array of innovative methods, from state-of-the-art remote sensing and monitoring techniques to the application of sustainable land management practices. Through these advancements, the restoration process becomes not only more effective but also environmentally responsible, ensuring that our efforts to heal peatlands do not cause unintended harm elsewhere. "The eco-friendly Technology 4N concepts" is not just a theoretical treatise. It also serves as a practical guide for policymakers, conservationists, and environmental practitioners seeking to make a tangible impact on the ground. The book presents a novel restoration strategy by incorporating the 4N concept: No plastic, No burning, No fertilizer, and No exotic and invasive species, which was developed by the TMI project. As we collectively strive to create a more sustainable future, the restoration of peatlands stands as a beacon of hope and a symbol of our commitment to the well-being of our planet. This book comes at a critical moment when decisive action is needed, and its insights will undoubtedly inspire and empower readers to take up the challenge.
Full-text available
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.
Full-text available
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.