332 marcel rejmánek ET AL.
13
Plant invasions and invasibility
of plant communities
Marcel Rejmánek, David M. Richardson
and Petr Py1ek
13.1 Introduction
Some 2500 yr ago, Heraclitus of Ephesus said that `All things change . . . and you
cannot step twice into the same stream'. Today, ecologists would not only say the
same about streams but also about vegetation. Plant communities change with time
due to changes in the environment (Chapter 7), biotic interactions (Chapter 9) and
invasions of alien species and genotypes, introduced intentionally or accidentally by
humans. Invasions have received detailed attention only recently. There have always
been migrating taxa, but now the rate of human-assisted introductions of new taxa is
several orders of magnitude higher. In California, for example, more than 1000 alien
plant species were intentionally or non-intentionally introduced and established viable
populations over the last 250 yr. In the Galápagos Islands, over 3 million yr of their
history, only one new plant species arrived with birds or sea currents every 10,000 yr.
However, over the last 20 yr the introduction rate has been c. 10 species per yr, or
some 100,000 times the natural arrival rate (Tye 2001).
Three basic questions arise:
1 What kind of ecosystems are more (or less) likely to be invaded by alien plants?
2 What kind of plants are the most successful invaders and under what circumstances?
3 What is the impact of the plant invaders?
13.2 Definitions and major patterns
Unlike natives (taxa that evolved in the region or reached it from another area where
they are native without help from humans), aliens (`non-native' or `exotic') owe their
presence to direct or indirect activities of humans. Most of them occur only temporarily
and are not able to persist for a long time without human-assisted input of
diaspores; these are termed casual. Naturalized taxa form sustainable populations
without direct human help but do not necessarily spread; the ability to spread characterizes
a subset termed invasive taxa. This distinction is critical because not all
naturalized taxa reported in floras and checklists are invasive. Not all naturalized
plant taxa, and not even all invaders, are harmful invaders ­ the last-mentioned should
plant invasions and invasibility of plant communities 333
Fig. 13.1 Total number of alien plant species, percentage of alien plant species, and
number of alien plant species per log(area) along the Pacific coast of Americas. `Alien
species' here are plants growing in individual areas without cultivation. Not all of them are
fully naturalized and even fewer are invasive. Nevertheless, numbers of naturalized and
invasive species are proportional to numbers of `alien species' in this diagram. Primary
data or references are in Kartesz & Meacham (1999) and Vitousek et al. (1997).
200
150
100
50
0
Alienspecies/log(area)
Alaska
BritishColumbia
Washington
California
BajaCalifornia
ChamelaBay(Mex.)
Chiapas(Mex.)
Panama
Choco(Colombia)
Peru
Chile
TierradelFuego
25
20
15
10
5
0
%ofalienspecies
1400
1200
1000
800
600
400
200
0
Numberofalienspecies
334 marcel rejmánek ET AL.
rather be called exotic weeds or exotic pest plants (Booth et al. 2003; Richardson
et al. 2000a).
Weeds are both native and alien and the alien element in weed floras varies over
the world. Most weedy taxa in southern Europe, Malaya, Mexico or Taiwan are native,
whereas most weedy taxa in Australia, the USA, Chile, South Africa, New Zealand,
Hawai'i, and many other islands are non-native. There may be inherent differences
in invasibility of different parts of the world. Uneven representation of alien, mostly
naturalized, plant species in regional floras along the Pacific shore of the Americas
illustrates this point (Fig. 13.1). These differences are certainly partly due to the
history of human colonization and trade. Nevertheless, similar patterns can also be
recognized on other continents (Rejmánek 1996; Lonsdale 1999). For instance, areas
with mediterranean climates (with the exception of the Mediterranean Basin itself)
seem to be more vulnerable and the tropics appear more resistant to plant invasions.
This should not be generalized, however. Savannas and especially disturbed
deforested areas in the Neotropics are very often dominated by African grasses such
as Hyparrhenia rufa and Melinis minutiflora, while similar tropical habitats in Africa
and Asia are dominated by Neotropical woody plants, e.g. Lantana camara and
Opuntia spp. The absolute number of alien species, therefore, is not necessarily the
best indicator of ecosystem invasibility, at least at this scale. Undisturbed tropical
forests, however, harbour only a very small number of alien plant species and most of
them do not spread beyond trails and gaps (Rejmánek 1996). It is probably not the
extraordinary species diversity of tropical forests that is important but simply the
presence of fast-growing multilayered vegetation that makes undisturbed tropical
forests resistant to invasions.
At the regional scale enormous differences in presence and abundance of invaders
among different communities (ecosystems) within one area seem to be the rule. An
overview is now available for Central Europe (Table 13.1). Alien species are concentrated
mostly in vegetation of deforested mesic habitats with frequent disturbance
(Pysek et al. 2002a,b). In general, native forests harbour a low number and proportion
of both archaeophytes (introduced before 1500) and neophytes (introduced
later); alien species are completely missing from many types of natural vegetation,
e.g. bogs, natural Picea abies forest, and rare in many natural herbaceous communities.
Herbaceous communities of extreme habitats and/or with strong native clonal
dominants (Nanocyperion flavescentis, Phragmition, Nardion) seem to be most resistant
to invasions of both archaeophytes and neophytes. In general, Californian lowland
communities (Fig. 13.2) are more invaded than corresponding communities in Europe.
However, there are some important similarities. Open and disturbed communities
are more invaded, while undisturbed forests are less invaded.
Data from California (Fig. 13.2) suggest that proportions of alien species numbers
are reasonably well correlated with their dominance (cover). This is probably attributable
to a simple sampling effect: with increasing proportion of alien species, there is
an increasing chance that one or more of them will dominate the community. While
there appears to be a general agreement between the proportion of alien species
numbers and their actual importance (cover and biomass), some exceptions are quite
remarkable. While alien species number in Chelidonio-Robinion woodland is certainly
not exceptionally high (Table 13.1), the dominant Robinia pseudoacacia is an alien
plant invasions and invasibility of plant communities 335
Table 13.1 Numbers of alien species, classified according to the time of introduction into
archaeophytes and neophytes, in representative plant communities of the Czech Republic on
the phytosociological alliance level. Within each vegetation group, alliances are ranked
according to decreasing total number of alien species. Data from Pysek et al. (2002a).
Number of Number of % invasive
archaeophytes neophytes among
neophytes
Ruderal vegetation
Sisymbrion officinalis tall-herb comm. of annuals on 96 106 9.4
nitrogen-rich mineral soils
Aegopodion podagrariae nitrophilous fringe comm. 16 76 36.8
Arction lappae nitrophilous comm. of dumps and 36 45 31.1
rubbish tips
Balloto-Sambucion shrub comm. of ruderal habitats 18 34 41.2
Matricario-Polygonion arenastri comm. of trampled sites 20 20 15.0
Potentillion anserinae comm. of salt-rich ruderal 12 20 10.0
habitats
Convolvulo-Agropyrion comm. of field margins and 24 16 31.3
disturbed slopes
Onopordion acanthii thermophilous comm. of village 34 8 12.5
dumps and rubbish tips
Weed communities of arable land
Veronico-Euphorbion weed comm. of root crops on 47 28 21.4
basic soils
Panico-Setarion weed comm. of root crops on 28 15 40.0
sandy soils
Caucalidion lappulae thermophilous weed comm. on 79 11 0.0
base-rich soils
Aphanion weed comm. on acid soils 41 8 12.5
Sherardion weed comm. of cereals on medium 47 7 14.3
base-rich soils
Grasslands
Arrhenatherion mesic meadows 15 56 25.0
Festucion valesiacae narrow-leaved dry grasslands 12 12 0.0
Bromion erecti broad-leaved dry grasslands 6 8 0.0
Nardion subalpine grasslands 0 1 0.0
Helianthemo cani-Festucion pallentis rock-outcrop 2 0 ­
vegetation
Forests
Alnion incanae ash-alder alluvial forests 4 15 40.0
Carpinion oak-hornbeam forests 6 14 14.3
Chelidonio-Robinion plantations of Robinia 5 10 60.0
Genisto germanicae-Quercion dry acidophilous oak 1 11 36.4
forests
Tilio-Acerion ravine forests 5 8 37.5
Luzulo-Fagion acidophilous beech forests 0 4 50.0
Quercion pubescenti-petraeae thermophilous oak forests 1 2 0.0
Quercion petraeae acidophilous thermophilous 0 2 50.0
oak forests
Salicion albae willow-poplar forests of lowland rivers 0 2 50.0
336 marcel rejmánek ET AL.
Table 13.1 (cont'd)
Number of Number of % invasive
archaeophytes neophytes among
neophytes
Alnion glutinosae alder carrs 0 2 0.0
Fagion beech forests 0 1 100.0
Betulion pubescentis birch mire forests 0 0 Piceion
excelsae spruce forests 0 0 Aquatic
and wetland vegetation
Lemnion minoris macrophyte vegetation of eutrophic 0 3 0.0
and mesotrophic still waters
Cardamino-Montion forest springs without tufa 0 2 50.0
formation
Phragmition reed beds of eutrophic still waters 1 1 0.0
Magnocaricion elatae tall-sedge beds 0 1 0.0
Nanocyperion flavescentis annual vegetation on 1 0 wet
sand
tree from North America. On the other hand, there are many alien species in some
grassland communities (Festucion valesiaceae, Bromion erecti), but dominants are exclusively
native and aliens are rarely invasive.
13.3 Invasibility of plant communities
Can we say anything conclusive about differences in invasibility (vulnerability to
invasions) of particular ecosystems? Analyses of ecosystem invasibility based just on
one-point-in-time observations (a posteriori) are usually unsatisfactory (Rejmánek
1989). In most of the cases we do not know anything about the quality, quantity
and regime of introduction of alien propagules. Nevertheless, available evidence
indicates that only very few non-native species invade successionally advanced plant
communities (Rejmánek 1989; Meiners et al. 2002). Here, however, the quality of
common species pools of introduced alien species ­ mostly rapidly growing and
reproducing r-strategists ­ is probably an important part of the story. These species
are mostly not shade-tolerant and many of them are excluded during the first 10
or 20 yr of uninterrupted secondary succession (Fig. 13.3), or over longer periods
of primary successions. However, some r-strategists are shade-tolerant. e.g. Alliaria
petiolata, Microstegium vimineum and Sapium sebiferum. Such species can invade successionally
advanced plant communities and, therefore, represent a special challenge
to managers of protected areas.
Plant communities in mesic environments seem to be more invasible than communities
in extreme terrestrial environments (Rejmánek 1989). Apparently xeric environments
are not favourable for germination and seedling survival of many introduced
species (abiotic resistance) and wet terrestrial habitats do not provide resources mainly
light ­ for invaders because of fast growth and high competitiveness of resident
plant invasions and invasibility of plant communities 337
100
80
60
40
20
0
Cover(%)
Invasivespecies
Annual
grassland
Relative
Quercus
douglasii
savanna
Adenostoma
fasciculatum
chaparral
Burned
2 yr ago
Unburned
35
30
25
20
15
10
Number(%)
16
14
12
10
8
6
4
Number(per100m2
)
30
28
26
24
22
Number(per100m2
)
Total
Quercus
berberidifolia
chaparral
Salix spp.
riparian
woodland
Quercus
wislizenii
forest
Nativespecies
Fig. 13.2 Native and invasive species in seven plant communities of the Stebbins Cold
Canyon Reserve, North Coast Ranges, California (150­500 m above sea level). Each
column represents a mean from three 100-m2
plots. `Relative cover' of invaders is their cover
with respect to the cumulative vegetation cover in all strata (herbs, shrubs and trees).
Comparing means for individual vegetation types, the only significant correlation is between
percentage of invasive species and total cover of invasive species (r = 0.75; n = 7; p = 0.05).
Rejmánek (unpublished data).
338 marcel rejmánek ET AL.
10
8
6
4
2
0
Numberofspecies/m2
0 10 20
Exotic
30 40
Years since abandonment
Native
Fig. 13.3 Effect of time since abandonment on the mean species richness of native and
exotic species over 40 yr of old-field succession in New Jersey (Meiners et al. 2002). Decline
of the mean percentage of exotic species is even more dramatic, decreasing from 58 to 28%.
Total and relative cover of exotic species declines significantly as well, but there can be a
temporary increase during the first 5 or 10 yr of succession; see also Rejmánek (1989).
species (biotic resistance). We have to be cautious, however, in the interpretation of
these patterns. When the `right' species are introduced, even ecosystems that have been
viewed as invasion-resistant for a long time may turn out to be susceptible, for instance
the Mojave and Sonoran deserts are facing recent invasions by Brassica tournefortii
and Pennisetum ciliare. Open water is notoriously known as open to all kinds of exotic
aquatic plants. In general, disturbance, nutrient enrichment, slow recovery rate of
resident vegetation, and fragmentation of successionally advanced communities promote
plant invasions (Rejmánek 1989; Hobbs & Huenneke 1992; Cadenasso &
Pickett 2001). In addition, the increasing CO2 level will probably accelerate invasions
in arid ecosystems (Smith et al. 2000).
A general theory of invasibility was put forward recently by Davis et al. (2000):
intermittent resource enrichment (eutrophication) or release (due to disturbance)
increases community susceptibility to invasions. Invasions occur if/when this situation
coincides with availability of suitable propagules. The larger the difference between
gross resource supply and resource uptake, the more susceptible the community to
invasion. This was anticipated by Vitousek & Walker (1987) (Fig. 13.4) and expressed
more rigorously by Shea & Chesson (2002). Davis & Pelsor (2001) experimentally
manipulated resources and competition in a herbaceous community, and showed that
fluctuations in resource availability of as little as one week in duration could greatly
enhance plant invasion success (survival and cover of alien plants) up to one year
after such events.
Experiments on invasibility of different types of ecosystems have been gaining
momentum in recent years (Hector et al. 2001; Fargione et al. 2003). Crawley et al.
(1999) and Davis et al. (2000) suggested that there is no necessary relationship between
invasibility of a plant community and number of species present in that community.
plant invasions and invasibility of plant communities 339
Resourcesupplyordemand
Time since disturbance
Disturbance
Supply
Demand
Fig. 13.4 Changes in supply and demand of resources after disturbance in terrestrial
ecosystems. Resource availability is generally at its maximum shortly after disturbance,
although conditions of bare ground can inhibit seedling establishment in some sites.
Modified from Vitousek & Walker (1987).
Other studies show that such a relationship exists: positive at the landscape scale
(e.g. Stohlgren et al. 1999; Sax 2002) and negative at scales of < 1 m2
(neighbourhood
scales sensu Levine 2000). Wardle (2001) provided a valuable methodological criticism
of many of these studies. Nevertheless, Kennedy et al. (2002) concluded that in
herbaceous communities neighbourhood species richness (within 5­15 cm radius)
represents `an important line of defence against the spread of invaders'. Hubbell et al.
(2001) found that in an undisturbed forest in Panama neighbourhood species richness
(within 2.5­50 m radius) had a weak but significantly negative effect on focal
tree survival. Is there a generalization emerging from studies on neighbourhood
scales? This would not be surprising as vascular plants are sedentary organisms and
actual interactions are occurring among neighbouring individuals.
The experimental studies mentioned above relate the number of resident plant
species to the number and abundance of alien plant species that establish or become
invasive. But, the diversity of organisms at other trophic levels in the receiving environment
may well be as important, if not more important, than the number of plant
species. We can expect that diverse assemblages of mutualists (pollinators, seed dispersers,
microbiota that form symbioses with plant roots) would promote invasibility
(Simberloff & Von Holle 1999; Richardson et al. 2000b). Recent experiments by
Klironomos (2002) on species from Canadian old-fields and grasslands showed that
native rare plant species accumulate soil plant pathogens rapidly, while invasive species
do not. This result has potentially very important consequences. When introduced
340 marcel rejmánek ET AL.
Propagule
input of
native taxa1
Identity of
exotic species
Identity of
resident
plant taxa
?
+/­
+/­
­
­
­ ­
+/­
+
+
+
+
+
+
+
+
+/­
+/­
+/­
+/­+/­
­?
­?
Historical isolation vs.
exposure to other biota
Life-span of
resident plants
Input of
resources1
(eutrophication)
Functional group
diversity of
residents
Species diversity
of residents
Disturbance1
Invasibility1
ACTUAL
INVASION
RATE IN
A COMMUNITY1
Post-disturbance
recovery rate
of residents
Resident herbivores
and pathogens
of exotic taxa
Actual amount
of available
resources1
Propagule
input of
exotic taxa1Long-term
character of local
environment*2
Non-specific
mutualists
Fig. 13.5 Causal relationships between factors and processes which are assumed responsible
for invasions of exotic species into plant communities. The most important relationships are
indicated by thick arrows. * = Spatial heterogeneity, (micro)climate and long-term regime
of available resources and toxic compounds. 1 = Time scale: days­years. 2 = Time scale:
years­centuries. The key components are in boxes.
outside of their native territories, plants are often liberated from their enemies,
including soil pathogens. This is a clear advantage that makes natives and aliens, at
least temporarily, different (but see Colautti et al. 2004). At the same time, however,
as many mycorrhizal fungi can associate with a broad range of plant taxa, root
beneficial symbionts are likely to be always available to many alien plants.
A conceptual cause-effect diagram (Fig. 13.5) captures all the fundamental aspects
of the ongoing debate on the issue of invasibility. The fact that both invasibility and
species diversity of residents is regulated in a similar way by the same set of factors
­ (micro)climate, spatial heterogeneity, long-term regime of available resources explains
why there are so many reports of positive correlation between numbers of
native and non-native species when several different communities or areas are compared.
Fast post-disturbance recovery of residents may be a key factor making the
wet tropics more resistant to plant invasions ­ measured as number of invading
species per log(area) (Rejmánek 1996).
However, there is very likely one extra factor that is currently poorly understood: the
historical and prehistoric degree of exposure of resident taxa to other biota (Fig. 13.5).
Is this the reason why islands are more vulnerable and Eurasia least vulnerable to
invasions? Is instability of so many man-made monocultures a result of the `lack of
plant invasions and invasibility of plant communities 341
any significant history of co-evolution with pests and pathogens' (May 1981)? Actual
species richness may not be as important as the complexity of assembly history.
In addition to mathematical models and computer simulations (Law 1999) relevant
experiments with plant communities will have to be designed. Artificial experimental
plant communities that are so often used for invasibility experiments have a clear
advantage of homogeneous substrata and microclimates. However, assembly processes
here are very short and/or artificially directed via arbitrary species pool selection,
weeding, reseeding, etc. The existence of well-established phytosociological
associations and the fact that plant species are combined in highly non-random
patterns within their natural communities (Gotelli & McCabe 2002) indicate that
historical assembly processes cannot be substituted by arbitrary mixtures of species.
Finally, longevity/persistence of resident plants is a distinct component of resistance
to invasions (Von Holle et al. 2003), especially in forest communities, resulting
in `biological inertia', including allelopathic chemicals produced by living or dead
residents.
13.4 Habitat compatibility
Identity of exotic taxa (Fig. 13.5) is important for two reasons. First, they may or
may not survive and reproduce in habitats where they are introduced. Second, they
may or may be not spread and become invasive. Recipient habitat compatibility is
usually treated as a necessary condition for all invasions. The match of primary (native)
and secondary (adventive) environments of an invading taxon is not always perfect
but usually reasonably close (e.g. Beerling et al. 1995; Rejmánek 2000; Widlechner
& Iles 2002). In North America, for example, latitudinal ranges of naturalized
European plant species from the Poaceae and Asteraceae are on average 15­20°
narrower than their native ranges in Eurasia and North Africa. These differences
essentially reflect the differences in the position of corresponding isotherms and major
biomes in Eurasia and North America. Major discrepancies between primary and
secondary ranges have been found for aquatic plants where secondary distributions
are often much less restricted than their primary distributions. Vegetative reproduction
of many aquatic species seems to be the most important factor. Obviously, secondary
ranges, if already known from other invaded continents, should be employed in any
prediction of habitat compatibility.
As for plants introduced (or considered for introduction) from Europe, several
useful summaries of their `ecological behaviour' are available. Especially the combination
of Ellenberg indicator values (Ellenberg et al. 1992) with Grime's functional
types (strategies) (Grime et al. 1988) can be a powerful tool for predictions of habitat
compatibility of European species. The strength of affiliation with phytosociological
syntaxa (section 1.4.2) is well known for almost all European taxa. Environmental
conditions (climate, soil, disturbance, management) of all syntaxa are available and
potential habitat compatibility of taxa can be extracted from the European literature.
Knowledge of this `phytosociological behaviour' of taxa allows predictions about
compatibility with analogous (vicarious) vegetation types, even if these will not always
be correct.
342 marcel rejmánek ET AL.
`Open niches', habitats that can support life forms that are not present in local
floras for historical and/or evolutionary reasons, deserve special attention. Dramatic
invasions have occurred in such habitats, e.g. Ammophila arenaria (a rhizomatous grass)
in coastal dunes in California, Lygodium japonicum (a climbing fern) in bottomland
hardwoods from Louisiana to Florida, Pinus spp. and Acacia spp. in South African
shrubby fynbos, Opuntia spp. (Cactaceae) in East African savannas, Rhizophora mangle
(mangrove) in tree-less coastal marshes of Hawai'i, and the tree Cinchona pubescens
(Rubiaceae) in mountain shrub communities on Santa Cruz Island, Galápagos. The
explanation of such invasions is confirmed by experiments showing that the competitive
inhibition of invaders increases with their functional similarity to resident abundant
species (Fargione et al. 2003).
13.5 Propagule pressure and residence time
The notion of habitat compatibility includes all factors embraced in the concept of
habitat. Most effort in assessments of habitat compatibility has been devoted to climatic
and substrate compatibility, although it is well known that many other factors influence
range limits. Invasions result from an interplay between habitat compatibility and
propagule pressure (Fig. 13.5). This is illustrated by the invasion dynamics of the
New Zealand tree Metrosideros excelsa (Myrtaceae) in South African fynbos (details
in Richardson & Rejmánek 1998). Multiple regression of the number of Metrosideros
saplings on a potential seed rain index (PSRI) and soil moisture revealed that, in this
case, both factors are about equally important (Fig. 13.6). This example clearly shows
that classification of habitats or communities into `invasible' and `non-invasible' cannot
be absolute in many situations. Habitats that are currently unaffected (or only
slightly affected) by plant invasions may be deemed resistant to invasion. However, as
populations of alien plants build up and propagule pressure (Foster 2001) increases
outside or within such areas, invasions could well start or increase. Another aspect is
the propagule pressure of native species: if propagules of natives are not available, as
for instance on abandoned fields in California, the `repairing' function of ecological
succession (Fig. 13.3) does not work.
Residence time ­ the time since the introduction of a taxon to a new area ­ represents
another dimension of propagule pressure. As we usually do not know exactly
when a taxon was introduced, we use a `minimum residence time' based on herbarium
specimens or reliable records. Nevertheless, the number of discrete localities of naturalized
species is significantly positively correlated with minimum residence time
(Fig. 13.7). One trivial but important conclusion is that the earlier an exotic pest
plant taxon is discovered, the better is the chance of its eradication.
13.6 What are the attributes of successful invaders?
The identity of introduced species certainly matters (Fig. 13.5). Extrapolations based
on previously documented invasions are fundamental for predictions in invasion
ecology. With the development of relevant databases, this approach should lead to
plant invasions and invasibility of plant communities 343
Fig. 13.6 The dependence of the sapling density of Metrosideros excelsa on potential seed
rain index (PSRI) and moisture in fynbos of the Western Cape, South Africa. PSRI =
SUM(1/di), where di is distance to the i-th mature tree in metres within the radius 300 m.
The first ordination axis (below) serves as a surrogate for moisture gradient. Standardized
partial regression coefficients (st. part. regr. coeff.) of the multiple regression are almost
identical. Therefore, both independent variables ­ environment and propagule pressure are
equally important in this case. M. Rejmánek & D.M. Richardson (unpublished data).
Seed rain (PSRI)
160
120
80
40
0
X2 X1
0
1­5
6­20
> 20
Moisture (1st ordination axis)
Myrica serrata
Metalasia muricata
Stoebe incana
Elegia filacea
Erica
ericoides
Cliffortia hirsuta
Osmitopsis
asteriscoides
Gleichenia
polypodioides
Number of
Metrosideros
saplings
per 100 m2
NumberofMetrosideros
saplingsper100m2
Erica perspicua
R2
= 0.39
b
1 (st. part. regr. coeff. for X1) = +0.42, p < 0.005
b
2 (st. part. regr. coeff. for X2) = +0.39, p < 0.005
344 marcel rejmánek ET AL.
Totalnumberofreportedlocalities
Selected naturalized
vascular plant species
in the Czech Republic
log (Y) = 1.41 + 0.0074X
n = 63, R2
= 0.38, p < 0.001
1000
100
10
1
Naturalized grasses in Venezuela
log (Y) = -0.015 + 0.017X
n = 111, R2
= 0.36, p < 0.001
1000
100
10
1
0 50 100
Minimum residence time (yr)
150 200 250
Fig. 13.7 The dependence of the total number of reported localities on the minimum
residence time (yr since the first record) of selected naturalized species in the Czech
Republic and Venezuela. P. Pysek & M. Rejmánek (unpublished data).
immediate rejection of imports of many invasive taxa (prevention) and prioritized control
of those that have already been established. Such transregional, taxon-specific
extrapolations are very useful in many situations, but our lack of understanding
makes them intellectually unsatisfactory. Understanding how and why certain biological
characters promote invasiveness is very important since even a global database
will not cover all potentially invasive taxa. In New Zealand, for example, Williams
et al. (2001) reported that 20% of the alien weedy species collected for the first time
in the second half of the 20th century had never been reported as invasive outside
New Zealand. For these reasons, several attempts have been made to find differences
in biological characteristics of non-invasive and invasive taxa or, at least, between
native taxa and non-native invasive taxa in particular floras. Major predictions made
by an emerging theory of plant invasiveness were summarized recently by Booth
et al. (2003), Myers & Bazely (2003) and Rejmánek et al. (2004):
plant invasions and invasibility of plant communities 345
1 Fitness homoeostasis (the ability of an individual or population to maintain relatively
constant fitness over a range of environments) promotes invasiveness.
2 Genetic change can facilitate invasions, but many species have sufficient phenotypic
plasticity to exploit new environments.
3 Several characters linked to reproduction and dispersal are key indicators of
invasiveness.
4 Seed dispersal by vertebrates is implicated in many plant invasions.
5 Low relative growth rate of seedlings and low specific leaf area are good indicators
of low plant invasiveness in many environments.
6 Large native range is an indicator of potential invasiveness. However, several
important exceptions are known (Impatiens parviflora, Pinus radiata).
7 Vegetative reproduction is responsible for many plant invasions, especially in aquatic
and wetland environments.
8 Alien taxa are more likely to invade a continental area if native members of the
same genus (and family) are absent, partly because many herbivores and pathogens
cannot switch to phylogenetically distant taxa (Agrawal & Kotanen 2003; Mitchell
& Power 2003). However, invaders on islands seem to exhibit the opposite tendency
(Duncan & Williams 2002).
9 The ability to utilize generalist mutualists greatly improves an alien taxon's chances
of becoming invasive.
10 Efficient competitors for limiting resources are likely to be the best invaders in
natural and semi-natural ecosystems (Tilman 1999; Shea & Chesson 2002).
11 Characters favouring passive dispersal by humans greatly improve an alien plant
taxon's chance of becoming invasive.
Points 3, 4 and 5 are particularly relevant. Reproduction and dispersal are key
issues. Consistent seed production in new environments is usually associated with
rather simple or flexible breeding systems. For example, rare and endangered taxa in
the genus Amsinckia (e.g. A. furcata, A. grandiflora) are heterostylic, while derived
invasive taxa (A. menziesii, A. lycopsoides) are homostylic and self-compatible. Selfpollination
has been consistently identified as a mating strategy in colonizing species.
Nevertheless, not all sexually reproducing successful invaders are selfers. Pannel &
Barrett (1998) examined the benefits of reproductive assurance in selfers versus outcrossers
in model metapopulations. Their results suggest that an optimal mating system
for a sexually reproducing invader in a heterogeneous landscape should include the
ability to modify selfing rates according to local conditions. In early stages of invasions,
when populations are small, plants should self to maximize fertility. However,
later, when populations are large and pollinators and/or mates are not limiting,
outcrossing will be more beneficial, mainly due to increasing genetic polymorphism.
Invasiveness of woody taxa in disturbed landscapes is associated with small seed
mass (< 50 mg), a short juvenile period (< 10 yr), and short intervals between large
seed crops (1­4 yr) (see Fig. 13.8 and Rejmánek & Richardson 1996, 2003).
These three attributes contribute, directly or indirectly, to higher values of three
parameters which are critical for population expansion: net reproduction rate, reciprocal
of mean age of reproduction, and variance of the marginal dispersal density.
For wind-dispersed seeds, the last parameter is negatively related to terminal velocity
of seeds which is positively related to (seed mass) (Rejmánek et al. 2004). Because
346 marcel rejmánek ET AL.
25
16
9
4
1
900
Juvenileperiod(yr)
576
324
144
36
0 7 6 5 4 3 2
1
NON-INVASIVE
cemo
cema
pal
sab
cou
rig
edu
tor
res
fle
pon
nig
ell
mursyl
ban
pin
pat
hal
rad
con
str
Seed mass (mg)
Interval between
large seed crops (yr)
INVASIVE
lam
K
r
Fig. 13.8 Distribution of 23 frequently cultivated Pinus species in a space created by three
biological variables critical in separating invasive and non-invasive species. The K­r selection
continuum running from the upper left to the lower right corner of the diagram also
represents the direction of the discriminant function (Z) separating non-invasive and
invasive Pinus species. Z = 23.39 - 0.63M - 3.88J - 1.09S, where M = mean seed mass
(in mg), J = minimum juvenile period (in yr), and S = mean interval between large seed
crops (in yr). Pine species with positive Z scores are invasive and species with negative
Z scores are non-invasive. Species abbreviations: ban = banksiana; cema = cembra; cemo =
cembroides; con = contorta; cou = coulteri; edu = edulis; ell = elliottii; eng = engelmannii;
fle = flexilis; hal = halepensis; lam = lambertiana; mur = muricata; nig = nigra; pal = palustris;
pat = patula; pin = pinaster; pon = ponderosa; rad = radiata; res = resinosa; sab = sabiniana;
str = strobus; syl = sylvestris; tor = torreyana.
of the trade-off between seed number and mean seed mass, small-seeded taxa usually
produce more seeds relative to biomass. Invasions of woody species with very small
seeds (< 3 mg), however, are limited to wet and preferably mineral substrata (Rejmánek
& Richardson 1996). Based on invasibility experiments with herbaceous species, it
seems that somewhat larger seeds (3­10 mg) extend species habitat compatibility
(Burke & Grime 1996). As seed mass seems to be positively correlated with habitat
shade, large-seeded aliens may be more successful in undisturbed, successionally
more mature plant communities.
plant invasions and invasibility of plant communities 347
Seed dispersal by vertebrates is responsible for the success of many invaders in
disturbed as well as `undisturbed' habitats (Binggeli 1996; Rejmánek 1996; Richardson
et al. 2000b; Widlechner & Iles 2002). Even some very large-seeded alien species like
Mangifera indica can be dispersed by large mammals. The proportion of naturalized
plant species dispersed by vertebrates seems to be particularly high in Australia: over
50% (Tables 2.1 and 7.3 in Specht & Specht 1999). The assessment of whether
there is an opportunity for vertebrate dispersal is an important component of the
screening procedure for woody plants (Table 13.2). However, vertebrate seed dispersal
in relation to invasions is complicated (Richardson et al. 2000b).
Table 13.2 General rules for detection of invasiveness of woody seed plants based on values
of the discriminant function Z*, seed mass values, and presence or absence of opportunities
for vertebrate dispersal. Modified from Rejmánek & Richardson (1996).
Opportunities for vertebrate dispersal
Absent Present
Dry fruits and Likely1
Very likely2
seed mass > 3 mg
Z > 0 Dry fruits and Likely Likely
seed mass < 3 mg in wet habitats3
in wet habitats3
Fleshy fruits Unlikely4
Very likely5
Not Possibly7
Z < 0 unless dispersed
by water6
*Z = 23.39 - 0.63M - 3.88J - 1.09S; where M = mean seed mass (mg); J = minimum
juvenile period (yr); S = mean interval between large seed crops (yr). Z was derived on the
basis of a priori defined groups of invasive and non-invasive Pinus species. The function was later
successfully applied on other gymnosperms and, as a component of this table, even on woody
angiosperms. Note that parameters in this discriminant function are somewhat different from
those in Rejmánek & Richardson (1996). This is due to exclusion of Pinus caribaea from the data
set used for estimation of the parameters. This species, that is in general non-invasive in many
countries, is highly invasive in New Caledonia.
1
For example Acer platanoides, Cedrela odorata, Clematis vitalba, Cryptomeria japonica, Cytisus
scoparius, Pinus radiata, Pittosporum undulatum, Pseudotsuga menziesii, Robinia pseudoacacia, Senna spp.,
Tecoma stans.
2
Species with large arils (Acacia cyclops) are dispersed by birds.
3
For example Alnus glutinosa in New Zealand, Eucalyptus globulus in California, Melaleuca
quinquenervia in southern Florida, Tamarix spp. in the south-western USA, Cinchona pubescens in
Galápagos and Baccharis halimifolia in Australia.
4
Feijoa sellowiana and Nandina domestica are frequently cultivated but non-invasive species in California.
The second species, however, is dispersed by birds and water in the south-eastern USA.
5
Berberis spp., Clidemia hirta, Crataegus monogyna, Lantana camara, Lonicera spp., Myrica faya,
Passiflora spp., Psidium guajava, Rubus spp., Schinus terebinthifolius, Solanum mauritianum.
6
Nypa fruticans is spreading along tidal streams in Nigeria and Panama. Thevetia peruviana can be
dispersed over short distances by surface run-off in Africa.
7
Examples of invasive species in this group are Pinus pinea, Melia azedarach, and Maesopsis eminii
in Africa, Quercus rubra in Europe, Mangifera indica in the Neotropics, and Persea americana in
Galápagos.
348 marcel rejmánek ET AL.
Table 13.3 Differences between means of growth related variables for non-invasive,
unclassified and invasive Pinus species. Same superscript letters for each variable denote
means that are not significantly different (p > 0.05, Scheffé test). From Grotkopp et al.
(2002).
Variable Non-invasive Unclassified Invasive
n = 8 n = 8 n = 8
Relative growth rate (RGR, mgg-1
d-1
) 23a
33b
37b
Net assimilation rate (NAR, mgcm-2
d-1
) 0.505a
0.559a
0.572a
Leaf area ratio (LAR, cm2
g-1
plant) 50a
67b
73b
Specific leaf area (SLA, cm2
gleaf) 79a
101b
111b
Relative leaf production rate (leafleaf­1
d­1
) 0.014a
0.022b
0.024b
Many ecologists assume that a high relative growth rate should be an important
characteristic of invasive plant taxa in disturbed or open areas, especially in resourcerich
environments. Only a few studies have demonstrated this experimentally (Pattison
et al. 1998; Grotkopp et al. 2002). An analysis of seedling growth rates for 29 Pinus
species (Table 13.3) revealed that: (i) relative growth rate (RGR) of invasive species
is significantly higher than that of non-invasive species; (ii) differences in RGR are
primarily determined by leaf area ratio (LAR; leaf area per plant biomass); and (iii)
LAR is primarily determined by specific leaf area (SLA; leaf area per leaf biomass).
Consequently, invasive species have significantly higher specific leaf area. Moreover,
there is a highly significant (R2
= 0.685; p < 0.001) positive relationship between
RGR and invasiveness of the Pinus species (Fig. 13.8, Table 13.2). Identical results
were obtained using phylogenetically independent contrasts (Grotkopp et al. 2002).
High SLA was implied as an important factor associated with invasiveness of grasses
and other plants. In general, SLA < 90 cm2
g-1
(dry leaf mass) seems to be a good
indicator of non-invasive or, at least, less-invasive evergreen woody plants and SLA
< 150 cm2
g-1
probably means the same for other vascular plant taxa.
Basic taxonomic units used in plant invasion ecology are usually species or subspecies.
However, genera are certainly worth considering as well. Species belonging
to genera notorious for their invasiveness or `weediness', e.g. Amaranthus, Echinochloa,
Ehrharta, Myriophyllum, should be treated as highly suspicious. On the other hand,
a continuum from highly invasive to virtually non-invasive species is also common
in many genera, e.g. Acer, Centaurea, Pinus. Recently, some attention has been paid
to taxonomic patterns of invasive plants (Daehler 1998; Pysek 1998; Rejmánek &
Richardson 2003). In terms of relative numbers of invasive species, some families
seem to be over-represented: Amaranthaceae, Brassicaceae, Chenopodiaceae, Fabaceae,
Gramineae, Hydrocharitaceae, Papaveraceae, Pinaceae and Polygonaceae. Among
the larger families, the Orchidaceae is the only under-represented one.
13.7 Impact of invasive plants, justification and prospects
of eradication projects
Numerous studies have documented the wide range of impacts caused by invasive
plants. Many invasive taxa have transformed both the structure and function of eco-
plant invasions and invasibility of plant communities 349
systems by, for example, changing disturbance- or nutrient-cycling regimes (D'Antonio
et al. 1999). In many parts of the world, impacts have clear economic implications
for humans, e.g. as a result of reduced stream flow from watersheds in South African
fynbos following alien tree invasions (Van Wilgen et al. 2001), or through disruption
to fishing and navigation after invasion of aquatic plants such as Eichhornia
crassipes.
It is important to stress, however, that the impact of invasive plants on biodiversity
is much less dramatic than impact of exotic pathogens, herbivores or predators. It
seems that most of the naturalized/invasive plant species have hardly detectable
effects on biotic communities (Williamson & Fitter 1996; Meiners et al. 2001).
There are at least 2000 naturalized plant species in North America and more than
1000 of them are invasive. However, not a single native plant species is known to
have been driven to extinction due to interactions with alien plants alone. Even on
islands, where numbers of exotic plant species are often increasing exponentially,
extinctions of native plant species cannot be attributed to plant invasions per se (Sax
et al. 2002). Also, the often reported correlation between numbers of native and
exotic plant species on the landscape scale can be interpreted as a lack of mechanisms
for competitive exclusion of native plants by exotic ones. Nevertheless, we should be
careful with conclusions ­ many invasions are quite recent and extinction takes a
long time.
Considerable progress has recently been made in developing methodologies for
making biological, ecological and economic assessments. In attempting to quantify
the value of ecosystem services of South African fynbos systems and the extent to
which these values are reduced by invasions, Higgins et al. (1997) showed that the
cost of clearing alien plants was very small (< 5%) as compared to the value of
services provided by these ecosystems. Their conclusion was that pro-active management
could increase the value of these ecosystem services by at least 138%. The most
important ecosystem service was water, and much work has been done on developing
models for assessing the value (in monetary terms) of allocating management
resources to clearing invasive plants from fynbos watersheds. Among the most dangerous
invaders in riparian areas within the USA are species of the Old World genus
Tamarix (salt cedar). An instructive economic evaluation of Tamarix impacts is
provided by Zavaleta (2000).
As we showed earlier, the most reliable predictions based on biological characters
are limited to invasiveness (likelihood of species establishment and spread). Predictions
of potential impacts will always be less reliable. Because decline in native species
richness is dependent on cover of invaders (Fig. 13.9; Richardson et al. 1989; Meiners
et al. 2001), indices based on a ratio of cover to frequency should be tested as impact
predictors for individual taxa. Other obvious impact indicators may be biological
characters of plants that are known to have ecosystem consequences (e.g. high transpiration
rates or nitrogen fixation).
Invasiveness and impact are not necessarily positively correlated. Some fast-spreading
species, like Aira caryophyllea or Cakile edentula, exhibit little (if any) measurable
environmental or economic impact. On the other hand, some relatively slowly spreading
species, e.g. Ammophila arenaria or Robinia pseudoacacia, may have far-reaching
environmental effects (stabilization of coastal dunes in the first case and nitrogen soil
enrichment in the second).
350 marcel rejmánek ET AL.
50
40
30
20
10
0
Numberofnativespecies
0.1 1 10 100 1000
Area (m2
)
Fynbos with > 25%
cover by aliens
Y = 5.04 + 5.531 log (X)
R2
= 0.87
(n = 6)
Fynbos without aliens
Y = 15.6 + 10.4 log (X)
R2
= 0.595
(n = 20)
Fig. 13.9 Species-area relationships for native vascular plant species in South African
fynbos areas densely infested (squares) by alien woody plants and in uninfested areas
(circles). Elevations of the two regression lines are significantly different (p < 0.001).
Sources of the data are acknowledged in Richardson et al. (1989).
There is a need for a universally acceptable, and objectively applicable, term for
the most influential invasive plant taxa within given regions, or globally. A potentially
useful term to use in this regard is `transformer species' (Richardson et al.
2000a). Such species, comprising perhaps only about 10% of invasive species, have
profound effects on biodiversity and clearly demand a major allocation of resources
for containment/control/eradication. Several categories of transformers may be
distinguished:
1 Excessive users of resources: water ­ Tamarix spp., Acacia mearnsii; light ­ Pueraria
lobata and many other vines, Heracleum mantegazzianum, Rubus armeniacus; water
and light ­ Arundo donax; light and oxygen ­ Salvinia molesta, Eichhornia crassipes;
high leaf area ratio, LAR, of many invasive plants (discussed earlier) is an important
prerequisite for excessive transpiration;
2 Donors of limiting resources: nitrogen ­ Acacia spp., Lupinus arboreus, Myrica faya,
Robinia pseudoacacia, Salvinia molesta;
3 Fire promotors/suppressors: promotors ­ Bromus tectorum, Melaleuca quinquenervia,
Melinis minutiflora; suppressors ­ Mimosa pigra;
4 Sand stabilizers: Ammophila spp., Elymus hymus spp.;
5 Erosion promotors: Andropogon virginicus in Hawai'i, Impatiens glandulifera in
Europe;
6 Colonizers of intertidal mudflats ­ sediment stabilizers: Spartina spp., Rhizophora
spp.;
plant invasions and invasibility of plant communities 351
7 Litter accumulators: Centaurea solstitialis, Eucalyptus spp., Lepidium latifolium, Pinus
strobus, Taeniatherum caput-medusae;
8 Soil carbon storage modifiers: promotor ­ Andropogon gayanus; suppressor Agropyron
cristatum;
9 Salt accumulators/redistributors: Mesembryanthemum crystallinum, Tamarix spp.
The potentially most important transformers are taxa that add a new function,
such as nitrogen fixation, to the invaded ecosystem (Vitousek & Walker 1989).
Many impacts, however, are not so obvious. For example, invasive Lonicera and
Rhamnus change vegetation structure of the forest, affecting nest predation of birds
(Schmidt & Whelan 1999), and Lythrum salicaria and Impatiens glandulifera can
have negative impacts on pollination and reproductive success of co-flowering native
plants (Grabas & Laverty 1999; Chittka & Schürkens 2001).
It follows from the discussion on impacts of exotic plants that careful prioritization
is needed before starting often very expensive and time-consuming eradication projects.
Maintenance of biodiversity is dependent on the maintenance of ecological processes.
Our priority should be protection of ecological processes. Attempts to eradicate
widespread invasive species, especially those that do not have any documented
environmental impacts (including suppression of rare native taxa), may be not only
useless but also a waste of time and resources. Exotic taxa with large-scale environmental
impacts (transformers) are usually obvious targets for control and eradication.
But when is complete eradication a realistic goal?
There are numerous examples where small infestations of invasive plant species
have been eradicated. There are also several encouraging examples where widespread
alien animals have been completely eradicated. Can equally widespread and difficult
alien plants also be eradicated? On the basis of a unique data set on eradication
attempts by the California Department of Food and Agriculture on 18 species and 53
separate infestations targeted for eradication in the period 1972­2000 (Table 13.4),
it is shown that professional eradication of exotic weed infestations smaller than 1 ha
is usually possible. In addition, about 1/3 of infestations between 1 and 100 ha and
1/4 of infestations between 101 and 1000 ha have been eradicated. However, costs of
eradication projects increase dramatically. With a realistic amount of resources, it is
very unlikely that infestations larger than 1000 ha can be eradicated (Table 13.4).
Table 13.4 Areas of initial gross infestations (at the beginning of eradication projects) of
exotic weeds in California, numbers of eradicated infestations, numbers of on-going projects,
and mean eradication effort for five infestation area categories. The data include 18 noxious
weedy species (two aquatic and 16 terrestrial) representing 53 separate infestations. From
Rejmánek & Pitcairn (2002).
Initial infestation (ha) < 0.1 0.1­1 1.1­100 101­1000 > 1000
No. of eradicated infestations 13 3 5 3 0
NO. of on-going projects 2 4 9 10 4
Mean eradication effort per infestation (work hours)
Eradicated 63 180 1496 1845 On-going
174 277 1577 17,194 42,751
352 marcel rejmánek ET AL.
Early detection of the presence of an invasive harmful taxon can make the difference
between being able to employ offensive strategies (eradication) and the necessity
of retreating to a defensive strategy that usually means an infinite financial
commitment. Nevertheless, depending on the potential impact of individual invaders,
even infestations larger than 1000 ha should be targeted for eradication effort or, at
least, substantial reduction and containment. If an exotic weed is already widespread,
then species-specific biological control may be the only long-term effective method
able to suppress its abundance over large areas (Myers & Bazely 2003).
Regardless of their environmental and/or economical effects, plant invasions provide
unique chances to understand some basic ecological processes that are otherwise
beyond the capacity or ethics of standard ecological experiments. We are just beginning
to fully appreciate these opportunities. However, we have a long way to go to
achieve a more complete understanding and more rational decision making.
Acknowledgements
We thank Jennifer Erskine and Jennifer Sedra for comments on a draft of this
chapter.
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