Download Weeds and the monitoring of biodiversity in Australian rangelands

Austral Ecology (2004) 29, 51–58
Weeds and the monitoring of biodiversity in Australian
rangelands
A. C. GRICE
CSIRO Sustainable Ecosystems, Private Bag, PO Aitkenvale, Queensland 4814, Australia
(Email: [email protected]) and Cooperative Research Centre for Australian Weed Management,
Adelaide, South Australia, Australia
Abstract Monitoring the biodiversity of Australian rangelands has been identified as a means of informing policy
and supporting funding decisions in relation to the conservation of biodiversity. Australian rangelands are subject to
invasion by alien plants that have the potential to have major impacts on ecosystem function and biodiversity,
although there has been little quantitative documentation of these effects. Research is needed to improve our
understanding of how and to what extent alien plants affect biodiversity in Australian rangelands so that this
relationship can be considered when developing and implementing programmes to monitor biodiversity. It is also
important to consolidate existing efforts to quantify the extent of alien plant invasions and monitor their progress,
thus documenting a process that threatens biodiversity. Information on the presence and abundance of alien plant
species should be considered for inclusion as a component of biodiversity monitoring programmes that are
undertaken. Monitoring components of biodiversity can itself provide a basis for evaluating weed management
strategies.
Key words: biodiversity, invasive plants, weeds.
INTRODUCTION
An alien plant is one that is present because of intentional or accidental introduction by humans, and that
reproduces consistently and sustains populations over
many life cycles without human intervention. An
invasive plant is a naturalized plant that produces
reproductive offspring at considerable distances from
parent plants and so has the potential to spread over
considerable areas (Richardson et al. 2000). Invasion
by alien plants is widely recognized as a major
threatening process for a great variety of ecosystems
worldwide. This is certainly true in Australia (State of
the Environment Advisory Council 1996), where
1500–2000 alien species had naturalized by the mid
1990s (Humphries et al. 1991) and naturalization
continues at an average rate of 10 species per year
(Groves 1997). Numerous alien species are recognized
as environmental weeds (plants that are in some way
deleterious to the environment; Groves 1997) in that
they invade natural ecosystems, as opposed to being
restricted to, for example, areas used for agriculture.
Many occur in Australia’s rangelands (Lonsdale &
Milton 2002), which constitute approximately 70% of
the Australian continent (Young et al. 1984). Alien
plants, therefore, have the potential to influence the
structure, function and composition of rangeland eco-
Accepted for publication September 2003.
systems and so constitute a threat to the biodiversity of
those rangelands.
Currently, there is considerable interest in monitoring the biodiversity of Australian rangelands (Smyth
et al. 2004). Any programme for monitoring biodiversity must be built on an understanding of the
processes that threaten biodiversity. The present paper
discusses the significance of invasive species for
monitoring the biodiversity of Australian rangelands.
It provides a brief synopsis of the nature, scale and
mechanisms of influence of invasive plant species on
Australian rangelands and their biodiversity. It then
discusses the implications of this information for the
monitoring of Australian rangeland biodiversity.
RANGELAND WEEDS
The likely impact of alien plant species on the biodiversity of Australian rangelands can be evaluated in
terms of the number of alien species present, the
growth forms represented, the extent of their actual
and potential geographical ranges and their relative
abundances at the sites that they occupy. There is,
however, no comprehensive, quantitative picture of the
invasion of Australian rangelands by alien plant species.
The number of alien plant species in Australian
rangelands is lower than in non-rangeland parts of the
continent (State of the Environment Advisory Council
1996), but floras of rangeland regions list significant
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A. C. GRICE
numbers of alien plant species. For example, Flora of
Central Australia (Jessop 1981), which covers 30% of
the Australian continent and 43% of its rangelands,
documents 114 species of alien plants making up 6% of
the total species listed for the region. Plants of Western
New South Wales (Cunningham et al. 1981), which
covers 5% of the continent and 7% of its rangelands,
documents 400 species of alien plants making up 21%
of the total species listed for the region.
The alien flora of the Australian rangelands includes
a broad variety of growth forms. Of the exotic species
listed in Australian rangeland floras, 50–60% are
annual forbs. Other growth forms contribute smaller
proportions of exotic species: annual grasses (17%);
perennial forbs (12–19%); perennial grasses (5–8%);
shrubs (approximately 5%); and trees (1–2%) (Grice
2000). Some growth forms (e.g. smothering vines such
as Cryptostegia grandiflora R. Br.; Tomley 1998) are
likely to have greater impacts than others.
Many of the alien species present in Australian
rangelands already have extensive ranges there, as
demonstrated by maps of current distributions of
species nominated as Weeds of National Significance
(WONS). However, bioclimatic analyses predict that
for most species there remains much potential for
further range expansion (Thorp & Lynch 2000).
Although the number of species present, their
growth forms and their geographical ranges indicate
something of the threats posed by alien plant species,
their impacts at any particular site will be driven more
by their abundance, which can be expressed in terms of
plant densities or biomass per unit area. Quantitative
data on abundance of weeds across extensive areas are
not available for Australian rangelands. Qualitative,
categorical data are available for some species at State
or national scales, as presented, for example, for species
nominated as WONS (Thorp & Lynch 2000). Quantitative data on abundance are available for particular
species at individual sites. These indicate that many
alien species can reach very high densities and biomass,
to the point where they dominate the stratum of
vegetation within which they occur. For example,
densities of Prosopis pallida (Willd.) Kunth. reach at
least 1600–1800 plants per ha (Campbell & Setter
1999); dense stands of the shrub Mimosa pigra L.
typically reach 10 000–30 000 plants per ha (Lonsdale
et al. 1995); mature individuals of the annual forb
Parthenium hysterophorus L. reach densities of 14 000
plants per ha (Navie et al. 1998).
There are, therefore, numerous alien plant species
of a wide variety of growth forms in Australian
rangelands; many of them are currently widespread,
although there remains considerable scope for further
increases; and they often dominate the vegetation.
Although these factors suggest that their combined
effect on biodiversity is considerable, some species are
likely to be more important than others. Some indi-
cation of perceived relative importance of different
alien species can be gained from legislative weed
declarations. It must be acknowledged, however, that
such declarations are rather subjective and relate
strongly to particular land uses, especially for agricultural and pastoral industries. The information from
central Australia (Jessop 1981) and western New South
Wales (Cunningham et al. 1981) indicates that annual
forbs, annual grasses and perennial grasses are underrepresented in lists of declared weeds relative to their
contributions to exotic floras (Parsons & Cuthbertson
1992). In contrast, perennial forbs, shrubs and trees are
overrepresented (Fig. 1). These ratios might reflect
pastoral interests more strongly than they reflect
environmental interests. For example, the perennial
grass Cenchrus ciliaris L. (buffel grass) remains the
most widely sown perennial pasture grass in northern
Australia despite widespread concerns about its
invasive characteristics and environmental impacts in
central Australia, northern Western Australia and parts
of Queensland (Griffin 1993). The nominations of
WONS also indicate a predominance of woody species
among exotic species that have been highlighted as
most relevant to the rangelands, even though more than
half of the exotic species are annual forbs (Table 1;
Thorp & Lynch 2000). At least seven of Australia’s 20
WONS are relevant to rangelands.
As well as being subject to invasion by alien plant
species, Australian rangelands have experienced major
shifts in the composition of the native flora. Some of
these shifts involve increases in the abundance of native
shrubs and small trees (Noble 1997) that are often
referred to as invasions, although, strictly speaking, the
species often do not conform to the definition of
‘invasives’ provided by Richardson et al. (2000), but
rather are proliferations within their native ranges.
Nevertheless, these shifts no doubt have significant
consequences for biodiversity.
Fig. 1. Percentage contribution of different growth forms
to ( ) exotic plant and () declared weed lists species for
Australian rangelands based on data from Cunningham et al.
(1981), Jessop (1981) and Parsons and Cuthbertson (1992).
WEEDS AND BIODIVERSITY
MECHANISMS OF IMPACT ON
ECOSYSTEMS
Weeds affect the ecosystems that they invade in a
variety of ways, and dramatic changes are often obvious. Weeds sequester resources (water, nutrients and
light) that would otherwise be available to native plant
species. The pathways and rates of water and nutrient
cycling are then likely to be altered relative to the
situation in the uninvaded state, as is the light regime
(Standish et al. 2001). Freudenberger et al. (1997) link
increases in the abundance of native shrubs in rangeland systems with landscape dysfunction; that is, the
‘excessive flow of water and nutrients out of a landscape system’ (Tongway & Ludwig 1997). Invasive
exotic shrubs might be expected to show similar
associations.
Resources that are acquired by a plant can be (i)
retained by the individual; (ii) passed on to the
succeeding generation by being used to produce seeds;
or (iii) captured by consumers. The availability to
consumer organisms of the resources that are acquired
by plants varies depending on the plant species. For
some alien weed species, there might be few, if any,
herbivores capable of accessing the resources that have
been sequestered by the weed. Specialist herbivores,
53
most notably insects, from the weed’s native range
might not have been introduced. This forms part of the
argument in favour of biological control as a weedmanagement technique. The ramifications for the food
web could also extend to decomposers because litter
derived from weed species can differ from that
produced by native species (Pidgeon & Cairns 1981;
Schulze & Walker 1997).
Weeds also alter the structure of the vegetation that
they invade. The growth form of the weed(s) is important in this regard and the effects are likely to be most
extreme in cases where a weed is of a growth form that
does not exist, or is only a minor component in the
uninvaded community. Examples include the invasion
of Top End floodplains, which are naturally dominated
by herbaceous species, by the shrub Mimosa pigra L.
(giant sensitive plant) (Lonsdale 1992); the invasion of
Queensland Mitchell grass (Astrebla spp.) plains by the
tree Acacia nilotica (L.) Del. (Mackey 1998); and the
invasion of northern Australian riparian zones by the
shrubby vine Cryptostegia grandiflora R. Br. (Tomley
1998).
Weeds also affect ecosystems by altering fire regimes.
These effects occur when weeds influence fuel loads,
fuel distribution, fuel continuity or the timing of
curing, with repercussions for the intensity, timing
Table 1. Species that were nominated as Weeds of National Significance and that are relevant to rangelands (from Thorp &
Lynch 2000)
Scientific name
Common name
Growth form
Xanthium spinosum L.
Hyptis suaveolens (L) Poit.
Argemone ochroleuca Sweet
Xanthium occidentale Bertol.
†
Parthenium hysterophorus L.
Themeda quadrivalvis (L.) Kuntze
Bryophyllum tubiflorum Harvey
Sida spp.
Stachytarpheta spp.
Eragrostis curvula (Schrader) Nees
Sporobolus pyramidalis Beauv.
S. natalensis Th. Dur & Schinz
Pennisetum polystachion
†
Lantana camara L.
Lycium ferocissimum Miers
Jatropha gossypifolia L.
Calotropis procera (Aiton) Aiton
Bassia scoparia (L.) A. J. Scott
†
Mimosa pigra L.
Gomphocarpus fruticosus (L.) Spreng.
†
Parkinsonia aculeata L.
Senna obtusifolia (L.) H. S. Irwin & R. C. Barneby
†
Cryptostegia grandiflora R. Br.
†
Tamarix aphylla (L.) H. Karst
Ziziphus mauritiana Lam.
†
Prosopis spp.
†
Acacia nilotica (L.) Del.
Bathurst burr
Hyptis
Mexican poppy
Noogoora burr
Parthenium
Grader grass
Mother of millions
Sida species
Snakeweeds
African lovegrass
Giant rat’s tail grass
Annual forb
Annual forb
Annual forb
Annual forb
Annual forb
Annual grass
Perennial forb
Perennial forb
Perennial forb
Perennial grass
Perennial grass
Perennial grass
Perennial grass
Shrub
Shrub
Shrub
Shrub
Shrub
Shrub
Shrub
Shrub
Shrub
Shrub/vine
Tree
Tree
Tree
Tree
†
Species declared Weeds of National Significance.
Mission grass
Lantana
African boxthorn
Bellyache bush
Calotrope
Kochia
Giant sensitive plant
Narrow-leaved cotton bush
Parkinsonian
Sicklepod
Rubber vine
Athel pine
Chinee apple
Mesquite
Prickly acacia
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A. C. GRICE
and frequency of fires. Effects on fire regimes are cited
among the impacts of the bulky perennial grasses
Andropogon gayanus Kunth. (gamba grass) and Pennisetum polystachion (L) Schultes (mission grass) in the
Top End, and Cenchrus ciliaris in central and Western
Australia (Rossiter et al. 2003).
EFFECTS OF WEEDS ON BIODIVERSITY
Although some effort has been made to document
mechanisms whereby weeds influence ecosystem
structure and function, overall the impacts have not
been widely quantified. There are few data specifically
dealing with the effects of weeds on the biodiversity of
Australian rangelands.
A typical approach to examining the effects of weeds
on plant communities is that used by Mullett and
Simmons (1995). This example relates not to an
invasive exotic species but to the indigenous Australian shrub Pittosporum undulatum Vent. (sweet pittosporum) that is spreading into communities or parts of
the landscape that it did not occupy in pre-European
times, as well as apparently expanding its range in
Victoria. Mullett and Simmons (1995) quantified its
environmental impacts by estimating the cover of all
vascular plant species in 3 3-m quadrats located
along transects running through small clumps of
P. undulatum. Multivariate pattern analysis showed
that fewer plant species occurred in quadrats where
the abundance of P. undulatum was high and the
abundance and cover of the indigenous species was
inversely correlated with that of P. undulatum. A similar
approach showed negative correlations between measures of abundance of native species and the shrub
Chrysanthemoides monilifera (L.) T. Norl. (bitou bush)
that invades coastal vegetation in south-eastern
Australia (Weiss & Noble 1984).
Working in central Queensland woodlands, Fairfax
and Fensham (2000) compared floristic diversity
across boundaries between uncleared land and cleared
land with either native (exotic species < 10% of total
ground cover) or exotic (exotic species > 10% of total
ground cover) pastures. The total number of native
species was greater in uncleared land than in either
native or exotic pastures on cleared land. In cleared
land, native species richness was greater in native
pastures than in exotic pastures. One of the difficulties
with such studies is in separating the effects of invasion
by exotic pasture species from the effects of livestock
grazing, because cleared land is generally subject to
higher livestock densities. McIvor (1998) examined the
effects of various pasture-management treatments,
including over-sowing with legume/grass mixtures, on
species diversity of experimental pastures in northeastern Queensland. The presence of sown species,
including the grasses Cenchrus ciliaris and Urochloa
mosambicensis (Hackel) Dandy and the legumes Stylosanthes hamata (L) Taub. and S. scabra Vogel, was
associated with reduced numbers of native species at
plot and quadrat scales. This is also a case in which the
exotic species were deliberately sown and the effects of
their presence cannot be separated from the effects of
the sowing techniques.
Even less research has examined the effects of weeds
on animal groups, where the focus has been more
on the ecological responses of individual vertebrate
species. Again, quantification of effects has not been a
strong point of much of the work. Perhaps in contrast
to the situation with vascular plants, weeds can affect
individual vertebrate species either positively or negatively. This is evident from a list of exotic plant species
that are eaten by a variety of Australian bird species
(Loyn & French 1991). It is not apparent from this
kind of evidence how important the weeds are in the
diets of the species listed, nor does the information
indicate the net effects on bird species or communities.
Through their role as weed-dispersal agents, birds
facilitate changes in plant communities that could
continue for many decades or even hundreds of
years, with repercussions for other components of
biodiversity and ecosystem processes (Williams &
Karl 1996).
In central Australia, Griffin et al. (1989) examined
bird, reptile and plant communities in riparian zones
dominated by the native tree Eucalyptus camaldulensis
Dehnh. and the invasive Tamarix aphylla (L.) H. Karst.
Areas infested by T. aphylla supported fewer birds and
reptiles, had different avian community structures
and significantly different and less-diverse herbaceous
plant communities compared with sites dominated by
E. camaldulensis.
Published reports contain several examples of native
Australian mammal species possibly benefiting from
the presence of exotic plant species. These include the
use of Ulex europaeus L. (gorse) by the eastern barred
bandicoot (Perameles gunnii Gray), of Rubus fruticosus
L. (blackberry) by the broad-toothed rat (Mastacomys
fuscus Thomas) (Brown et al. 1991) and of Mimosa
pigra by the red-cheeked dunnart (Sminthopsis virginiae
Tarragon) (Braithwaite & Lonsdale 1987). In the first
two of these examples, the apparent associations
between weed and mammal are not quantified or
evaluated experimentally.
These examples are sufficient to indicate that native
plant species richness is generally lower in places where
one or more weed species are abundant and that
individual native animal species can respond positively
or negatively to invasion by a weed species. Causal
relationships are generally implied but not demonstrated. These conclusions fall well short of providing a
comprehensive picture of how biodiversity responds to
exotic plant invasions, although it is generally agreed
that the net effect of weed invasions is negative. In
WEEDS AND BIODIVERSITY
particular, a more specific understanding of the
relationships between weeds and biodiversity in
rangelands is not available.
Most work on the environmental impacts of weeds,
including effects on biodiversity, has examined single
weed species. However, it is common for natural ecosystems to be invaded by more than one species. In
effect, new plant communities are emerging within
Australian ecosystems as weed complexes combine
with native species that can coexist with them. The
presence of one weed species can facilitate invasion by
others. For example, a weed species that has characteristics that attract birds (e.g. fleshy fruit) could facilitate
invasion by other bird-dispersed weed species. Under
these circumstances, there will be synergistic effects of
weeds on biodiversity.
Obviously, alien plant invasions are only one of the
threats to biodiversity, and both the relative importance
of various threatening processes and the interactions
between them must be considered. Leigh and Briggs
(1992) (cited in Groves & Willis 1999) evaluated
numbers of extinct and endangered Australian plant
species threatened by various factors. They considered
that 57 species (13% of total extinct and threatened
species) were threatened by ‘weed competition’,
making this factor the third most important after
‘grazing and agriculture’ and ‘low population
numbers’.
WEEDS AND BIODIVERSITY MONITORING
Monitoring the biodiversity of Australian rangelands
has been identified as a means of informing policy and
supporting funding decisions in relation to the conservation of biodiversity (Smyth et al. 2004). There is
already sufficient information available to conclude that
invasion of Australian ecosystems by alien plants is one
of the important processes that threaten biodiversity.
What, therefore, should the links be between the
monitoring of biodiversity and work on alien plant
invasions in Australian rangelands?
First, it is useful to quantify the progress of alien
plant invasions, thus documenting a threatening process as far as biodiversity is concerned. Although
Australia does not have a national system for mapping
the occurrence of alien plants, many States do. There
would be value in at least improving the compatibility
of these systems even if a single national approach
proves elusive. Mapping at the relatively coarse scales
currently used would be sufficient to inform judgements about where biodiversity is at risk from weeds
and which weeds are involved. It is important that
weed-mapping systems include the capacity to document changes over time and incorporate at least qualitative estimates of the abundance of different weed
species. It will be necessary to consider the trade-offs
55
between the quality and quantity (scale, frequency) of
information recorded. Information on the distribution
and abundance of alien plant species in relation to
biodiversity would be used to inform decisions about
where to direct weed-management resources.
Second, there is a need for research to improve our
understanding of how and to what extent alien plants
affect biodiversity, expressly in Australian rangelands.
Many of the examples provided in the present paper
are from outside the rangelands. In particular, attention
should be given to understanding (i) how the spatial
and temporal characteristics of rangelands impinge on
the interactions between weeds and biodiversity; (ii) the
interactive effects of livestock grazing and alien species;
(iii) how alien species alter fire regimes and so impact
on biodiversity; (iv) how different taxonomic groups
respond to different types of weed invasion; and (v)
how biodiversity is affected by invasions of multiple
weed species.
It could be useful to conduct further research to
define functional groups that consist of alien plants that
affect biodiversity in similar ways (Díaz et al. 2002). A
growth-form or functional-group approach might help
predict or evaluate the effects of weeds on biodiversity.
Weed growth forms that are novel to the invaded
system are likely to be more significant for biodiversity
than growth forms for which there is a native ecological
analogue. However, there is also the possibility that
there are subtle but important differences between
species that could otherwise be regarded as being in the
same functional group or having the same growth form
as some native, non-weedy species. For example,
Cenchrus ciliaris is, in many ways, structurally and
functionally similar to the native, perennial, tussock
grasses of central Australia, but it might be less useful
to native granivores than the native grasses because of
the characteristics of its seeds. Invasion by Cenchrus
ciliaris could therefore lead to a decline in the diversity
of granivores.
Third, information on the presence and abundance
of alien plant species should be considered for inclusion as a component of biodiversity monitoring programmes that are undertaken. There are two reasons
for doing so: (i) the occurrence, distribution and abundance of weeds could be useful as an indicator of
biodiversity; and (ii) invasion by alien plants is a major
threatening process for Australian natural ecosystems.
Of course, the effort applied to incorporating alien
species as a component of biodiversity monitoring
must be determined in relation to the purpose of the
programme and the resources available to carry it out.
However, including information on alien plant species
might be a relatively small and efficient addition to a
monitoring programme that was already gathering
information on native vegetation.
Research is needed to establish the indicator value of
alien plant species, although one can expect a general
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A. C. GRICE
negative correlation between presence and abundance
of native and exotic plant species. Current evidence
indicates that the shapes of the relationships between
weed abundance and biodiversity measurements are
variable (Panetta & James 1999). It seems unlikely that
biodiversity decline is a linear function of weed abundance; rather, the impacts of weeds are likely to be
greater when weeds are in an advanced stage of
invasion. There may be thresholds such that the rate
of decline or loss of species or communities will be
much greater at high levels of weed abundance. Such
relationships and thresholds should be quantified and
effort should be made to understand the ecological
processes that underlie them. The situation is likely to
be further complicated by the fact that some species
increase in the face of weed invasions.
What measures of weed presence and abundance
could be considered as part of a monitoring programme? Weed species richness (weed species per unit
area) on its own is unlikely to be a good measure of how
severely weeds are affecting biodiversity. Some weed
species have a far greater impact than others, although
the presence of a large number of alien species might
be correlated with other forms of disruption to the
invaded ecosystem and so with the overall effects on
biodiversity.
Relative abundance of weeds at monitoring sites will
be more informative than presence/absence. Typically,
the severity of infestations of weedy trees and shrubs
are described in terms of plant density (Campbell &
Setter 1999). This is a practical means of making
intersite or temporal comparisons of abundance of
weed species whose individuals are of approximately
equal size. However, where there is a need to make
comparisons between species whose individuals differ
in size, measurement of biomass will be far more
informative. In a biodiversity monitoring programme,
site-specific data on the biomass of weeds and how that
Table 2. Alien plant species of Australian rangelands that
warrant priority consideration in monitoring programmes
Mode of impact/Species
Trees and shrubs that invade and dominate open savannas
and grasslands
Acacia nilotica
Parkinsonia aculeata
Prosopis spp.
Ziziphus mauritiana
Mimosa pigra
Jatropha gossypifolia
Perennial grasses that alter fire regimes
Cenchrus ciliaris
Pennisetum polystachion
Andropogon gayanus
Themeda quadrivalvis
Smothering vines
Cryptostegia grandiflora
biomass is distributed between species and growth
forms would generally be preferable to simple lists of
species present.
Effort to gather data on relative weed abundance
should be most heavily concentrated on those species
and growth forms that make the greatest contribution
to the biomass of the weed community or that have the
potential to do so. Such species are the ones most likely
to have significant impacts on ecosystem structure
and function. This effort should be tempered by the
possibility that peculiarities of an invasive species could
give it an impact out of proportion to its abundance.
Moreover, an alien species might not dominate the
vegetation overall, but might still have a major impact
on a particular stratum of the vegetation and so on
biodiversity and ecosystem function.
The current state of knowledge suggests that priority
species or groups should be trees and shrubs that
invade open savannas and grasslands, perennial grasses
that are likely to significantly contribute to altered fire
regimes or that are not ecological surrogates of native
tussock grasses, and smothering vines that are relevant
to more densely vegetated woodlands and forests
(Table 2). This is not to say that other species and
growth forms could not be significant transformers of
rangeland ecosystems.
Fourth, monitoring components of biodiversity can
provide a basis for evaluating weed-management
strategies. Monitoring would help assess management
strategies (as opposed to experimental treatments) in
terms of both their effectiveness against target weeds
and the side-effects of strategies on non-target components of an ecosystem. Some management techniques and strategies will produce more favourable
results than others. Weed management practices
themselves can have effects on ecosystem function
and biodiversity. For example, burning to control
riparian infestations of Cryptostegia grandiflora affects
indigenous species as well (Grice 1997). Monitoring
will always be important in cases where eradication of
alien plant species is being attempted (Groves &
Panetta 2002). In Australia, relatively little effort has
been directed at evaluating weed-management strategies.
Finally, if alien plant species are to be incorporated
into a monitoring programme, careful consideration
should be given to the timing of monitoring and the
placement of monitoring sites. Decisions should be
based on the stage and rate of weed invasion, the
phenological and demographic characteristics of the
weed(s), the distribution of the weed(s) across a region
and landscape, the status of weed management and the
purpose of the monitoring programme. Where the
emphasis is on alien plant species as a threatening
process, more frequent monitoring would be required
for species whose populations increase and whose
ranges expand rapidly than for species that increase
WEEDS AND BIODIVERSITY
and spread more slowly. Monitoring of weeds in
relation to biodiversity should be timed to take into
account the phenological cycles of the weeds involved.
This is especially important where annual weeds are
concerned.
In rangelands, riparian zones and other nutrient- and
moisture-rich parts of the landscape often support a
greater variety and abundance of weed species than
upland areas. These areas could warrant more intense
monitoring or a greater emphasis on alien species.
Riparian zones also support species and communities
that do not occur elsewhere in rangeland landscapes.
Their importance for biodiversity is disproportionate
to the area of land that they occupy, so monitoring
weeds as part of a biodiversity monitoring programme
is more important for riparian zones.
It is important to evaluate and respond to the threats
posed by invasive species in the context of the overall
threats to biodiversity and to consider how the various
threatening processes interact. Invasive species are one
threat among many. Ground-based monitoring techniques are time-consuming and expensive. Although it
is possible, for particular species and situations, to
quantify the distribution and abundance of individual
weed species by remote sensing (Abbott et al. 1999), it
is unlikely that weeds in general can be monitored by
remote techniques, particularly in the early stages of
invasion. Monitoring of weeds in rangelands can be
included in a programme for monitoring rangeland
vegetation in general, but interpreting the data
appropriately is critical. Monitoring can indicate
correlations between measures of weed abundance
(e.g. weed species richness, total weed biomass) and
measures of biodiversity, but such correlations do not
indicate the importance of weeds as causes of change
to biodiversity, or the mechanisms that might be
involved (Watson 2004). Manipulative experiments
would be a more reliable means of testing the effects of
weeds on biodiversity, although for many taxonomic
groups, the experiments would have to be long-term.
Any such work should be directed toward identifying
generalizations in relation to how different types of
weeds (species, growth forms or functional groups)
affect biodiversity in different habitats or environments, which components of biodiversity are affected,
and in what ways.
ACKNOWLEDGEMENTS
The material in the present paper was originally
presented at a workshop ‘Biodiversity Monitoring in
Rangelands’, held in Alice Springs, Australia, in
October–November 2002. I thank Alan Andersen,
Angas Hopkins, David Keith, Hugh Pringle, Jeff
Richardson and Jeremy Wallace for helpful comments
on an earlier draft of the manuscript.
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