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CHAPTER 9
Biological Invasions and
the Homogenization of
Faunas and Floras
Julian D. Olden1, Julie L. Lockwood2, and Catherine L. Parr3
1
School of Aquatic and Fishery Sciences, University of Washington, Seattle, USA
Ecology, Evolution, and Natural Resources, Rutgers University, New Brunswick, NJ, USA
3
Environmental Change Institute, School of Geography and the Environment, University of Oxford, Oxford, UK
2
9 . 1 T HE B I OGE OGR AP HY OF
S PECI E S I NV AS I ONS
In considering the distribution of organic beings
over the face of the globe, the first great fact which
strikes us is, that neither the similarity nor the dissimilarity of the inhabitants of various regions can
be accounted for by their climatal and other physical
conditions … A second great fact which strikes us in
our general review is, that barriers of any kind, or
obstacles to free migration, are related in a close and
important manner to differences between the productions of various regions.
(Charles Darwin, 1859, pp. 395–396)
9.1.1 The invasion process
One of the fundamental elements of life on Earth is
change. Species appear through time via evolution and
disappear by the natural actions of environmental
change (e.g. volcanic eruptions, changing sea levels,
glaciation). Species have also regularly shifted their
geographical ranges in response to biological and physical forces, sometimes becoming less common and
other times becoming more widespread. In general,
however, the large majority of species are not distributed broadly, because individuals of most species have
limited dispersal capabilities.
These limitations on dispersal ability have produced
the interesting phenomenon that many, perhaps even
most, species do not occupy all of the areas of the world
in which they could quite happily thrive. Instead, they
are restricted to certain regions, where they are able to
interact with only those species with which they cooccur. The limited geography of species is responsible,
in part, for the fantastic array of diversity that presently
carpets the Earth, as it provides opportunity for convergent evolution in disparate unconnected regions.
With the range expansion of modern humans, initially out of Africa, then across the globe, came the
possibility of human-mediated dispersal of a large
variety of other species. By this, we mean that humans
provided the conduit for individuals of some species to
disperse much farther abroad than they could naturally. Species were moved within, or on, humans as
parasites or disease organisms, in their household
goods as hitchhikers, as their livestock or working
animals, as their crop plants, as their pets, and as commodities themselves.
There is written evidence that intentional movements of species by humans traces back to ancient
Conservation Biogeography, First edition. Edited by Richard J. Ladle and Robert J. Whittaker.
© 2011 by Blackwell Publishing Ltd. Published 2011 by Blackwell Publishing Ltd.
Conservation planning in a changing world
times, such as the introduction of the tamarind tree
(Tamarindus indica) into China by way of commerce
along the Shu-Yan trade route that linked China to
India 8,000 years ago (Yan et al., 2001). Some species
apparently have nearly circumglobal distributions
because of ancient trade activities, with many of these
examples only recently coming to light thanks to the
power of molecular analyses to locate the evolutionary
origins of now very widespread species (e.g. Wares
et al., 2002).
There is ample historical evidence that the number
of species that were moved out of their native ranges
and introduced to somewhere novel via human actions
increased as the world began to become ever more
interconnected (Elton, 1958). As this number grew,
the need to understand how this process occurs, and
to differentiate natural species’ range expansions from
those mediated by humans, became critical. Without
making this distinction, it becomes difficult to untangle
the mechanisms that are driving historical biodiversity
changes, to understand the role of new arrivals in
driving evolutionary dynamics and, more practically,
to stem the flow of species that cause ecological or
economic harm (see below).
Before continuing, however, it is very important to
recognize that a multitude of names have been given
to species that are introduced to a novel location via
human actions – such as ‘exotic’, ‘invasive’ or ‘alien’
species (Lockwood et al., 2007). We use the term ‘invasion’ to refer to the process whereby species expand
their geographical distribution outside of their natural
dispersal range via the actions of humans, while we
refer to populations that have become otherwise established outside the bounds of their native ranges as
‘non-native’.
A more lucid understanding of the invasion process
may be achieved if it is considered as a stepwise progression of events, whereby individuals of some species
are moved out of their native ranges, released into a
novel location, establish self-sustaining populations
there and then spread to new locations (Figure 9.1;
Sakai et al., 2001).
Fundamental to this process is that not all individuals successfully pass through all these stages. The tens
rule of Williamson (1996) states that only ≈10 per
cent of transported individuals are released into a
foreign location, ≈10 per cent of these introduced
species will go on to survive and successfully breed (i.e.
establish a new population) and ≈10 per cent of these
established species will expand their geographical
225
Figure 9.1 Generalized stages common to all species
invasions. A species must successfully transition through
each sequential stage, and the proportion of species that
proceed from one stage to the next is less than the previous
one (depicted by arrow width).
ranges and become pests. These estimates were based,
in large part, on non-native animals and plants of
Britain. More recently, Jeschke and Strayer (2005)
investigated all freshwater fish, mammal and bird
species native to Europe or North America that have
been introduced outside their native range. They found
that the frequencies of transitions across all three of
the above stages averaged 6.1 per cent, 56.0 per cent
and 59.7 per cent, respectively.
Regardless of the specific percentages for each stage,
it is apparent that only a fraction of the species that are
moved by people, either on purpose or by accident, will
complete all stages of the invasion process. A considerable amount of research within invasion biology has
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Biological invasions and the homogenization of faunas and floras
therefore focused on attempts to understand which
factors differentiate between those species that successfully progress through all invasion stages and those
that do not (Lockwood et al., 2007).
9.1.2 Human-assisted versus
prehistoric invasions
A valid and persistent question is the extent to which
modern trends in species invasions differ from those
that occur naturally. This question is especially relevant to students of biogeography because range expansions are a very clear component of palaeoecological
and historical biodiversity patterns (Vermeij, 2005).
Do modern invasions warrant the attention currently
given to them by scientists? How different are the
mechanisms, spatial patterns and rates of modern
versus prehistoric invasions? Can we use prehistoric
trends to help predict the consequences of modern biological invasions?
Human-assisted dispersal of non-native species
differs from natural dispersal events in several important aspects (J.R.U. Wilson et al., 2009). Ricciardi
(2007) detailed the differences between prehistoric
and human-assisted invasions, which we summarize
below and in Table 9.1.
The most obvious differences are in the number
and frequency of ‘dispersal’ events. Natural dispersal
events are characteristically rare, both in the number
of species being transported and in the temporal frequency with which species disperse. By contrast,
modern human-assisted dispersal events happen constantly and involve a wide variety of species, which
show a much wider array of biological traits than
those species that are likely to experience natural longdistance dispersal.
The rate at which non-native populations are establishing around the world is consistently several orders
of magnitude larger than fossil-derived estimates for
natural dispersal events at the same locations. For
example, the invasion rate of terrestrial species for the
Hawaiian Islands was approximately 30 species per
million years (0.00003 per year) prior to human settlement, but increased to 20,000 species per million
years (0.02 per year) after the arrival of the Polynesians
and to approximately 20 per year during the past two
centuries (Ricciardi, 2007). In other words, contemporary rates of biological invasions are nearly one million
times higher than the prehistoric rate for Hawaii before
human influence.
The number of individuals of each species being
transported is also vastly different between natural and
human-assisted invasion events. Natural dispersal
events typically involve a few individuals of a species
finding their way out of the native range and attempting to establish a self-sustaining population in the
novel locale. Occasionally the number of individuals in
these natural events can be quite high – as for instance,
during biotic interchanges involving episodic events of
mass dispersal. For example, the opening of the transpolar corridor between the Pacific and Atlantic oceans
and the formation of the Panamanian land bridge
between North and South America during the Great
American Interchange permitted a massive flux of
species between formerly isolated regions (Vermeij,
2005; Lomolino et al., 2006). By contrast, humanassisted dispersal events are commonly characterized
Table 9.1 A comparison of key characteristics of prehistoric versus human-assisted invasions. Modified from Table 1
of Ricciardi (2007).
Characteristics
Prehistoric invasions
Human-assisted invasions
Frequency of long-distance dispersal event
Number of species transported per event
Propagule size per event
Number of mechanisms and routes of dispersal
Temporal and spatial scales of mass transport events
Degree of homogenizing effect
Potential for interactions with other stressors
Very low
Low*
Small*
Low
Episodic (short-distance)
Regional
Low
Very high
High
Potentially large
High
Continuous (long-distance)
Global
Very high
* Except during biotic interchange events.
Conservation planning in a changing world
by the release of hundreds to thousands of individuals
of a species into one novel locale, although there is
much variation around this number.
Finally, human-assisted invasions serve to connect
two or more locations that are geographically very
distant from one another, whereas natural dispersal
events tend to link sites that are comparatively close
together or otherwise linked naturally. Quite simply,
patterns of modern dispersal unite parts of the world
solely by social and economic ties, as opposed to biophysical pathways such as prevailing wind directions,
jet streams or ocean currents, as would happen for
natural dispersal events (Box 9.1).
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9.1.3 Economic and ecological impacts
of invasion
The interest in human-assisted invasions has grown
rapidly over the past two decades, which is attributable
to three factors (Lockwood et al., 2007):
• First, as the world economy globalizes, there are
increased trade and social connections between geographical localities, and along with these connections
come the introduction of non-native species (Perrings
et al., 2005; Hulme, 2009). Thus, the sheer number of
non-native populations establishing worldwide has
increased substantially in recent times.
Box 9.1 The human imprint on modern day species dispersal patterns
The Earth is now better connected via human transport than ever before. In recent decades, human
activities have greatly increased the frequency and spatial extent of species introductions across
the globe through both intentional and unintentional actions. These include ballast-water discharge
from international shipping; bait-bucket releases associated with recreational fishing; the global pet
trade; intentional translocations of wildlife for recreation purposes; biological control; and inadvertent
releases from aquaculture and horticulture activities. The following two case studies illustrate how
modern biotas are connected via social and economic networks and by sea and air.
Ship traffic
In marine and estuarine systems, the dominant invasion pathway worldwide is the ballast water of
commercial ships (Carlton & Geller, 1993; Drake & Lodge, 2004). Ocean-going vessels must achieve
proper stability to minimize drag (and thus maximize speed) and to reduce the likelihood of capsizing
in rough seas. To achieve this, early ships strategically filled ballast compartments within the hull
with soil, rocks or scrap metal – essentially, anything with some weight that could be easily loaded
into a ship at dock. Today, ships pump water into ballast tanks, and a typical commercial bulk vessel
might carry over 30,000 metric tonnes of ballast water during an inter-oceanic voyage. Ballast water
is usually taken from the harbour in one port and subsequently may be discharged in a recipient
port through openings in the ship’s hull.
The number of non-native species that are transported via ship ballast has increased with the
rise in global commerce and the consequent upsurge in the number of ships travelling the world’s
oceans and major waterways (Figure B9.1a). Current estimates suggest that a global fleet of approximately 35,000 commercial vessels transports an annual volume of about 3.5 × 109 metric tonnes of
ballast water, containing some 7,000–10,000 species (mostly marine) at any one time (Wonham
et al., 2005). Even if only a small fraction of these species establish non-native populations, it is
easy to see that ballast water is a primary mechanism by which aquatic invasions are occurring.
By tracking the number of ships that visit ports worldwide, Drake and Lodge (2004) were able to
map ‘hotspots’ of marine invasions and, via network modelling, to determine which ports are likely
to have increased rates of invasions in the coming years (Figure B9.1a). These hotspots are clearly
the product of economic and social influences on global trade and are in marked contrast to what
we might expect given natural dispersal patterns of marine species via oceanic currents.
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Biological invasions and the homogenization of faunas and floras
Figure B9.1a (a) The frequency of commercial shipping traffic along shipping routes around the world, ranging
from low (blue) to high (red). From Halpern et al. (2008). (b) Global hotspots for biological invasion from ballast
water, ranging from low (blue) to high (red). From Drake and Lodge (2004). (See Plate B9.1a for a colour version of
these images.)
Airline traffic
International air travel has been recently pinpointed as a significant factor in the movement of
economically damaging pest species and infectious diseases (Tatem, 2009). Among others, the
Mediterranean fruit fly Ceratitis capitata has been consistently imported in airline baggage (Liebhold
et al., 2006), plant pathogens are often found in air cargo (McCullough et al., 2006) and diseasecarrying mosquitoes have survived long haul flights in aircraft cabins (Lounibos, 2002). Far-removed
regions with similar climates have now been suddenly linked by a busy flight schedule, which has
resulted in an elevated risk of foreign invasions.
This risk, however, depends greatly on the time of year. Tatem and Hay (2007) identified an ‘invasion window’ across the global air network from June to August, when climatic conditions in regions
linked by long-haul routes are most similar to one another and the higher number of flights increases
the chances of exotic species hitching a ride to somewhere new. With expected increases in global
trade and travel (Perrings et al., 2005; Hulme, 2009), opportunities for such extreme hitchhiking
through the world airline transportation and shipping network look set to increase further (see trend
in Figure B9.1b).
Figure B9.1b Trends in global shipping cargo volumes and air freight, 1970–2005. From Hulme (2009).
Conservation planning in a changing world
• Second, as the number of non-native populations
increases, scientists find it increasingly hard to ignore
them. It is important to recognize that many of these
species present unique opportunities to test various
ecological, evolutionary and biogeographical concepts
and theories. Certainly the basic insights gained from
the study of modern invasion events are substantial
(Sax et al., 2007).
• Third, some of the non-native populations that have
established have gone on to impart substantial economic and ecological cost (Simberloff et al., 2005;
Pimentel et al., 2006).
As detailed above and shown in Figure 9.1, not all
species that are dispersed via human actions have
negative impacts within their new environment. The
definition of what constituents ‘impact’ is somewhat
problematic for at least two reasons:
1 There are scientific and societal influences on the
perception of impact (not to mention that the effects of
invasive species are often subtle and difficult to observe).
2 After impact is perceived, there is a variety of ecological factors that determine the level of impact produced (Lockwood et al., 2007).
Let us move past this issue by simply conceding that
human perception and valuation are an integral part
of the integration stage of the invasion process (Figure
9.1). It is important to recognize that the proportion of
species that do cause harm as compared to those that
are simply moved out of their native range is quite low.
Nevertheless, these few species will eat, parasitize and
compete with native species, often driving the latter
extinct or into very low population numbers (Elton,
1958; Clavero & García-Berthou, 2005; Strayer et al.,
2006). Some non-native populations invade natural
areas such as parks or wildlife reserves and disrupt
native species communities (Simberloff et al., 2005). In
these instances, the value of the natural area in terms
of its ability to conserve biodiversity may be reduced if
the non-native is not controlled or eradicated.
Many species threaten human economic interests,
notable examples including the zebra mussels (Dreissena
polymorpha) that clog utility companies’ water intake
valves (MacIsaac, 1996); emerald ash borers (Agrilus
planipennis) that devastate urban and commercial
forests (Poland & McCollough, 2006); and monk parakeets (Myiopsitta monachus), whose bulky nests can
cause electric power line failures (Avery et al., 2002).
A substantial number of non-native species have
adverse impacts on human health by transmitting diseases (Lounibos, 2002; Tatem, 2009), the most obvious
229
of which is the widespread distribution of Norway rats
(Rattus norvegicus). This rodent species was regularly,
and inadvertently, transported with human colonists
as they expanded across the globe. They serve as the
reservoir and vector for a variety of particularly troublesome human diseases, the most well know being
bubonic plague.
In general, scientists reserve the term ‘invasive’ for
these few non-native species that cause ecological or
economic harm. It is an open question as to whether
these few invasive species have characteristics that
make them unique amongst the world’s species, but
there is a clear need to be able to identify them as
potentially harmful long before they have the chance
to become invasive.
9. 2 B I OT I C H OMOGEN I Z AT I ON
The regional connectivity of the world is stronger and
more varied than ever before and, consequently, there
are very few places where non-native species have not
become established. Looking back over human history,
it is apparent that changes in species diversity are
frequently the result of the widespread invasion of
ubiquitous non-native species into areas containing
rare, and often unique, native species (Elton, 1958,
Ricciardi, 2007).
If the same non-native species are being introduced
to multiple locations, then there is potential for disparate regions to become more similar in their species
composition through time, a process known as biotic
homogenization. There are certainly well-known
invaders that can be found nearly everywhere. These
days, for example, you can land at nearly any airport
in the world and, while waiting for your next flight,
watch house sparrows (Passer domesticus) cavorting on
the tarmac. This species is native to Eurasia, but it has
realized a very broad geographical distribution via
human-mediated introductions.
For many years, the biodiversity crisis has been
focused on the loss of species through global extinction. Although this is clearly of prime importance, at
sub-global scales the loss of populations through local
extirpation, combined with the invasion of already
common non-native species, may be the more dramatic reconfiguration of modern biodiversity. In fact,
changes in diversity patterns at fine and coarse scales
of analysis can be either concordant or, alternatively,
can be decoupled and even conflicting.
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Biological invasions and the homogenization of faunas and floras
For example, Pautasso (2007) conducted a metaanalysis of the relationship between human population size and change in the plant and animal species
richness of study areas. The study reported negative
changes in richness at small spatial scales of analysis
(or small extent) but positive changes at larger spatial
scales. The introduction of non-native species by
humans is typically integral to such changes. In
essence, anthropogenic changes driving habitat loss,
fragmentation, species invasions and ecosystem transformation may result in declining local richness but,
across larger landscapes and regions, relatively few
native species may become entirely extinct, while nonnatives boost the richness above natural baseline levels.
Changes such as these, in the inventory richness of
smaller areas nested within larger regions, may also be
accompanied by changing patterns in differentiation
diversity, i.e. in the degree of compositional turnover
between localities – also known as ‘beta diversity’. A
change in beta diversity can, in fact, occur either
through a reduction in the total number of species in
the region (regional species richness or sometimes
‘epsilon diversity’) or through a change in the species
similarity between areas. Basically, if a similar suite of
species is shared across the areas in a region, beta
diversity will be quite low. If very different species
occur in different areas, beta diversity will be high.
Biotic homogenization is thus a term describing the
process of reducing differentiation diversity between
regions, but it may be accompanied by varying
patterns of change in inventory richness at different
scales of analysis. See Box 1.2 for an explanation of
terminology.
Put another way, biotic homogenization is described
as the process by which regionally distinct native communities are gradually replaced by locally expanding,
cosmopolitan, non-native communities (McKinney &
Lockwood, 1999). Some have likened the process of
biotic homogenization to the now global distribution of
fast-food restaurants, coffee houses and big-box retailers (Olden et al., 2005). The more connected we are as
a society, the more likely we are to see the trans-global
distribution of both species and businesses. In circumstances where invasive species impact negatively on
locally co-occurring native species, rare and endemic
native species may be lost, resulting in rapid loss of
differentiation diversity. However, it is also important
to recognize that the reverse can also occur and that,
in cases, the combined effects of invasions and extirpations can be to increase the mean differentiation
diversity across a study region, a phenomenon termed
‘biotic differentiation’ by Olden and Poff (2003).
9.2.1 The process of biotic homogenization
In the simplest sense, human activities that increase
rates of species invasions and extirpations are the ultimate cause of biotic homogenization. However, biotic
homogenization can arise when only invasions occur
without the concurrent loss of species, or conversely
where only species extirpations occur. In other words,
species additions or replacements need not occur for
regions to become homogenized or even differentiated
over time (Olden & Poff, 2003).
To illustrate this point, we provide a simple graphical
example showing how the number and manner in
which non-native species establishment and native
species extirpations occur may lead to very different
levels of homogenization or differentiation (Figure
9.2). In the absence of any extirpation, the establishment of the same non-native species at two separate
localities will lead to increases in the similarity of the
invaded communities. Conversely, the establishment of
a different non-native species at each locality will
decrease community similarity. Although this example
is useful to illustrate the simplest way biotic homogenization can occur, both empirical data and theoretical
modelling suggests that the process is both complex
and sensitive to the spatial and temporal scale of investigation (Olden, 2006).
9.2.2 Different manifestations of biotic
homogenization
Biotic homogenization is considered an overarching
process that encompasses either the loss of taxonomic,
genetic or functional distinctiveness over time (Olden
et al., 2004). Taxonomic homogenization, which we
used to introduce the concept of homogenization
above, has been the primary focus of previous research
and is commonly referred to as biotic homogenization.
However, imposing a narrow definition of biotic homogenization does not truly reflect the multidimensional
nature of this process. Consequently, it is useful to
think of biotic homogenization as a broader ecological
process by which formerly disparate biotas lose biological distinctiveness at any level of organization, including in their genetic and functional characteristics.
Conservation planning in a changing world
231
Figure 9.2 Illustration of how species invasions and extinctions can cause either biotic (taxonomic) homogenization in
scenario 1 or differentiation in scenario 2, depending on the identity of the species involved. A pair of communities (shaded
ovals) for each scenario is illustrated, where extirpation events are represented by the disappearance of a species icon over a
time step, whereas introduction events are represented by the arrow and appearance of a species icon. Importantly, both
scenarios share the same species pool (6 native butterflies, 2 introduced butterflies) and species richness through time is
identical for both scenarios. From Olden and Rooney (2006).
Let us spend a moment exploring these two additional
ways in which biotic homogenization can be
manifested.
Genetic homogenization refers to a reduction in
genetic variability within a species or among populations of a species. It can occur through at least three
mechanisms:
• First, the intentional translocation of populations
from one part of the range to another enhances the
potential for intraspecific hybridization (i.e. hybridization between different sub-species within a species),
with the end result being the assimilation of gene pools
that were previously differentiated in space (Stockwell
et al., 1996).
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Biological invasions and the homogenization of faunas and floras
• Second, introductions of species outside of their
original range(s) increases the likelihood of a founder
effect and reduced levels of genetic variability, as well
as setting the stage for interspecific hybridization (i.e.
hybridization between different species within the
same genus) (Rhymer & Simberloff, 1996).
• Third, if extirpations were a cause for faunal homogenization, then one consequence might be bottleneck(s)
in local populations of the impacted species, along with
lowered effective population size(s) (Lee, 2002).
Functional homogenization refers to an increase in
the functional similarity of biotas over time resulting
from the replacement of ecological specialists by the
same widespread generalists. It occurs primarily
because patterns of species invasions and extirpations
are not random, but instead are related to particular
biological traits that commonly predispose native
species to extirpation and non-native species to successful establishment. The end result is an increase in
the functional convergence of biotas over time associated with the establishment of species with similar
‘roles’ in the ecosystem (e.g. high redundancy of functional forms or traits) and the loss of species possessing
unique functional ‘roles’ (McKinney & Lockwood,
1999; Olden et al., 2004).
For example, Winter et al. (2008) examined how the
presence of non-native plant species in Germany
affected the distribution of a genetic trait, namely
ploidy level (referring to the number of homologous
sets of chromosomes in a biological cell), at two spatial
scales. It is commonly accepted that polyploidy species
should have a greater ability to colonize or invade new
habitats due to greater genetic variability. Interestingly,
this study found evidence for functional differentiation
at fine spatial scales (<130 km2) due to more heterogeneous ploidy levels of non-native plants compared to
native plants, whereas, at a coarser spatial scale, more
homogeneous ploidy levels of non-native species lead
to functional homogenization.
9. 3 PAT T E R NS OF B I OT I C
H O MO GE NI Z AT I ON
Many scientists, including ourselves, have argued that
we are entering a period characterized by widespread
faunal and floral homogenization, fittingly dubbed the
‘Homogecene’, in a place appropriately called the ‘New
Pangaea’ (the original Pangaea being the global supercontinent of approximately 250 million years ago).
Although the jury is still out on this, it is clear that the
study of biotic homogenization represents a unique
challenge because it is a multifaceted process, encompassing both species invasions and extirpations, which
requires the explicit consideration of how the identities
of species (not just species richness) change over both
space and time.
A simple perusal of the literature shows that the
majority of research to date has focused on quantifying
patterns of taxonomic homogenization, whereas the
processes of genetic and functional homogenization
have received considerably less attention. Moreover,
even estimates of taxonomic homogenization are
sparse and highly variable within and between taxonomic groups. Despite this trend, tremendous progress
has been made in recent years to better understand
and quantify patterns of taxonomic homogenization
(Table 9.2). We review the taxonomic groups (fishes,
birds, plants and mammals) that have received the
most attention next.
9.3.1 Fishes
The homogenization of freshwater fish faunas has
received the greatest attention thus far. In a landmark
study, Rahel (2000) compared the species similarity of
US states between present-day and pre-European settlement time frames and found that pairs of states averaged 15.4 more species in common now than they did
in the past. On average, fish faunas became more
similar by 7.2 per cent, with the highest increases in
similarity observed in western and north-eastern states
(Figure 9.3a). The high degree of biotic homogenization is best illustrated by the fact that the 89 pairs of
states that historically had zero similarity (no species
in common) now share an average of 25.2 species,
resulting in an average present-day similarity of 12.2
per cent. Patterns of fish homogenization were primarily the result of non-native species establishment associated with fish stocking for recreational purposes (e.g.
brown trout (Salmo trutta), rainbow trout (Oncorhynchus
mykiss) and smallmouth bass (Micropterus dolomieu) or
aquaculture (e.g. common carp, Cyprinus carpio), and
to a smaller degree the extirpation of endemic species
(harelip sucker, Lagochila lacera).
Taylor (2004) found a similar pattern among
Canadian provinces and territories, where average
faunal similarity increased from 27.8 per cent to 29.1
per cent – a trend driven in large part by the differential
Conservation planning in a changing world
233
Table 9.2 Review of the published studies that report estimates of community similarity change between two time
periods in the context of biotic (taxonomic) homogenization. Change in similarity refers to mean pair-wise difference
between historical and extant community similarity across all sites, unless otherwise noted. Positive values indicate
homogenization and negative values indicate differentiation. Note that this table only includes studies for which
estimates of per cent change in community composition were reported.
Taxonomic group
Location
Change in
similarity
Spatial extent
Unit
Country-wide
Basin divisions
North-eastern
coastal
Coastal watersheds
−1.4%
Country-wide
Provinces/territories
1.3%
British Columbia
Aquatic ecoregions
−3.5%
Country-wide
Major basins
2.2%
Iberian Peninsula
and France
Major basins
17.1%
Country-wide
States
California
Zoogeographic
provinces
Reference
Freshwater fishes
Australia
Canada
Europe
USA
Watersheds
South Dakota
3.0%
7.2%
20.3%
Olden et al. (2008)3
Taylor (2004)1
Leprieur et al. (2008)1
Clavero & García-Berthou
(2006)1
Rahel (2000)1
Marchetti et al. (2001)1
−10.7%
Hoagstrom et al. (2007)1
Geomorphic provinces
8.0%
Watersheds
2.4%
Minnesota
Lakes
9.0%
Kansas
Streams
−0.2%
Eberle & Channell (2006)1
Florida
Select counties
−0.8%
Smith (2006)1
Canada & USA
Country-wide
States and provinces
1.2%
Canada & USA
Select regions
States and provinces
−0.6%
Chile
Country-wide
Administrative regions
0.3%
Castro & Jaksic (2008)1,7
South-eastern
Pacific
Islands
2.0%
Castro et al. (2007)1
Europe
Germany
Grid cells (130 km2)
3.9%
Kühn & Klotz (2006)5
Europe & USA
Select regions
Forest stands
3.9%
Vellend et al. (2006)6
Great Britain
Country-wide
Grid cells (1 km2)
−1.0%
Smart et al. (2006)1
United States
Select regions
Parks and local areas
0.8%
McKinney (2004)1
Countries
0.5%
Schwartz et al. (2006)3
Forest stands
2.6%
Rooney et al. (2004)2
Radomski & Goeman (1995)1
Amphibians and reptiles
USA
Terrestrial plants
Wisconsin
Qian & Ricklefs (2006)1
Rejmánek (2000)1
234
Biological invasions and the homogenization of faunas and floras
Table 9.2 Continued
Taxonomic group
Location
Change in
similarity
Spatial extent
Unit
Reference
Canada & USA
Country-wide
Transects
Netherlands
Country-wide
Grid cells (5 km)
2.8%
Van Turnhout et al. (2007)2
Global
Atlantic Ocean
Oceanic Islands
0.9%
Cassey et al. (2007)3
Terrestrial birds
−2.0%
La Sorte & Boecklen (2005)4
−0.9%
Caribbean Ocean
Indian Oceans
1.8%
Pacific Ocean
−0.2%
Terrestrial mammals
Global
Select countries
Country
2.1%
South Africa
Country-wide
Grid cells (0.25 degree)
Spear & Chown (2008)1
−1.3%
Grid cells (1 degree)
4.2%
Grid cells (2 degrees)
8.1%
Taxonomic similarity based on: 1 Jaccard’s Similarity Index, 2 Bray-Curtis Similarity Index, 3 Sörensen’s Similarity Index,
4
Beta-sim Index, 5 Simpson’s Index, 6 Raup and Crick Index of beta diversity, 7 Mean values based on a published range.
invasion of 48 non-native fishes over the past century
(Figure 9.3a).
Similar broad-scale efforts have been conducted in
other parts of the world. Recent evidence points to the
homogenization of Australian fish faunas in response
to human-mediated species introductions (Olden et al.,
2008). Fish compositional similarity among major
drainages increased 3.0 per cent, from a historical
similarity of 17.1 per cent to a present-day similarity
of 20.1 per cent. Sometimes, the degree of faunal
similarity between drainages doubled or even tripled
with time. This trend was particularly obvious in the
southern corners of the continent – areas which are
highly populated relative to other regions of Australia
(Figure 9.3b). Similar to the United States and
Canada, fish faunal homogenization in Australia was
the result of the widespread introduction and subsequent escape/spread of non-native fishes for recreation
(rainbow trout), aquaculture (common carp) and mosquito control (western mosquito fish, Gambusia affinis),
and from the ornamental/aquarium trade (goldfish,
Carassius auratus; guppy, Poecilia reticulata).
Recent efforts in Europe have shown that exotic and
translocated native species generate distinct geographical patterns of biotic homogenization because of their
contrasting effects on the changes in community similarity (Leprieur et al., 2008). Although biological invasions have resulted in an overall increase in faunal
similarity on the order of 2.2 per cent (Figure 9.4a),
this research found that translocated native species (i.e.
species introduced by humans into regions where they
were not historically found) promoted homogenization
among basins (+5.0 per cent: Figure 9.4b), whereas
exotic species (i.e. species originating from outside
Europe) tended to decrease their compositional similarity (−1.6 per cent: Figure 9.4c). This finding is highly
consistent with patterns in floral homogenization (discussed in Section 9.3.3), suggesting that differences in
the geographical distribution of exotic and translocated species may play an important role in shaping
patterns of homogenization.
Clavero and García-Berthou (2006) used distributional data for freshwater fish in four time periods to
assess the temporal dynamics of biotic homogenization
among river basins in the Iberian Peninsula. They
found strong evidence for biotic homogenization, with
faunal similarity among rivers basins increasing by
17.1 per cent from historical times to the present day.
Changes in faunal similarity were highly dynamic in
time. The establishment of non-native species in 1995
Conservation planning in a changing world
235
Figure 9.3 Fish faunal homogenization of: (a) states and provinces in the United States and Canada (data from Rahel
(2000) and Taylor (2004), respectively); (b) major drainage divisions of Australia (from Figure 2 of Olden et al., 2008).
resulted in slight differentiation, but by 2001 the range
expansion of previously established non-native species
caused biotic homogenization in some regions and the
continuing addition of new non-native species led to
biotic differentiation in others.
9.3.2 Birds
Avifaunal homogenization has been another area of
recent focus, although the number of studies are
limited compared to fishes. In the Netherlands, Van
Turnhout et al. (2007) evaluated changes in breeding
bird composition over a 25-year period and found that
regions exhibited a 2.8 per cent increase in community
similarity. Significant spatial variation in patterns of
homogenization existed. Low-lying western regions
exhibiting low historical species richness showed the
greatest increase in resemblance by converging towards
those avifaunas more characteristic of eastern regions.
Based on breeding bird surveys for North America
(exclusive of Mexico), La Sorte and Boecklen (2005)
showed substantial change in the diversity structure
of avian assemblages at the local scale in non-urban
areas from 1968 to 2003. However, there was little
evidence that overall similarity in species composition
was increasing – in fact, the general trend was towards
a two per cent level of biotic differentiation. Despite
236
Biological invasions and the homogenization of faunas and floras
Figure 9.4 Fish faunal homogenization of major river drainages in Europe based on: (a) all non-native species; (b) only
translocated native species; (c) only non-native species originating from outside Europe. Adapted from Figure 1 of Leprieur
et al. (2008).
this, their study did find that more highly populated
regions located closer to the Atlantic and the Pacific
coasts of the United States experienced the strongest
patterns of homogenization.
At the global scale, Cassey et al. (2007) explored patterns of invasion and extirpation and their influence
on the similarity of oceanic island bird assemblages
from the Atlantic, the Caribbean, and the Indian and
Pacific Oceans. The authors found that patterns of
homogenization differed significantly between and
among archipelagos but, in general, avian assemblages
tended to show increased similarity over time to other
islands within their archipelago, compared with islands
outside their archipelago. Islands in the Indian Ocean
exhibited the greatest homogenization, whereas biotic
differentiation occurred for most islands in the Atlantic
Ocean. However, although avifaunal homogenization
was apparently the rule rather than the exception for
islands in the Indian Ocean, the authors found that the
relationship of this change to initial similarity was
Conservation planning in a changing world
scale-dependent. At smaller spatial scales (islands
within archipelagos), the expected pattern of low
initial similarity leading to greater homogenization
was observed, whereas this relationship reversed at the
larger spatial scale of islands between archipelagos.
This study illustrates that the spatial extent of investigation and the evolutionary history of the region
under consideration can influence patterns of taxonomic homogenization and differentiation within and
across what appear to be equivalent spatial units (i.e.
ocean basins).
9.3.3 Plants
Evidence for floral homogenization comes from studies
conducted in many countries at a variety of spatial
scales. However, evidence to date suggests that levels
of floral homogenization are considerably smaller than
those observed for freshwater fishes (Table 9.2).
Within the United States, McKinney (2004) found
that non-native plant species contributed significantly
to floral homogenization of 20 parks and local conservation areas, although the magnitude was relatively
low and sometimes negative (indicating differentiation). Cosmopolitan plant species most responsible for
the observed homogenization included curly dock
(Rumex crispus), dandelion (Taraxacum officinale) and
bluegrass (Poa annua). Similarly, Schwartz et al. (2006)
found that the county floras of California, USA, have
shown slight homogenization. The establishment of
noxious weeds played a central role in shaping patterns
of homogenization, but the authors suggest that the
greatest potential for future homogenization may come
from extirpations of extant native populations within
counties.
At a finer spatial scale, Rooney et al. (2004)
re-surveyed 62 upland forest stands in northern
Wisconsin, USA, to assess the degree of floral homogenization of under-storey communities between 1950
and 2000. By incorporating changes in both species
occurrence and abundance, the authors found that
two-thirds of the sites had become more similar in their
composition as a result of declines in rare species and
increases in already regionally abundant native and
non-native species. Interestingly, levels of homogenization were greatest in areas without deer hunting, suggesting that selective grazing by overabundant deer
populations was acting as a key driver of floral
homogenization.
237
Distributional patterns of native and non-native
species may vary in such a way that they will have
opposing effects on patterns of homogenization. For
example, Qian and Ricklefs (2006) evaluated changes
in differentiation diversity of vascular plants across
North America (excluding Mexico) and found that nonnative species tended to homogenize floras in distant
areas whose native plant species differ greatly, but
differentiate neighbouring areas that exhibited more
closely related native floras (Figure 9.5). Because few
native species have yet been extirpated from state and
provincial floras, these authors reason that the pattern
of homogenization and differentiation probably reflects
the haphazard introduction and establishment of nonnative species with respect to suitable habitats. At play
is also the natural and human-assisted spread of nonnative species with no regard to the ecological constraints acting on native species.
Smart et al. (2006) used botanical data for flowering
plants in Great Britain to test the hypothesis that plant
communities have become taxonomically and functionally more similar over the past 20 years in humandominated landscapes. Although little evidence was
found for the taxonomic homogenization of plant communities, this study revealed that plant traits related to
dispersal ability and canopy height increased in their
occurrence across the communities over time. The
authors suggest that environmental change has caused
different plant communities to converge on a narrower
range of winning trait syndromes (i.e. functional
homogenization), while species’ identities remained
relatively constant.
Similarly, Castro and Jaksic (2008) reported that
the compositional similarity of the continental flora
of Chile has not shown significant modifications
over time. Interestingly, this result is not shared for
oceanic island floral assemblages off the coast of Chile,
in which present-day islands share a greater number
of species compared to the pre-European condition
(Castro et al., 2007).
9.3.4 Mammals
Interest in the process of biotic homogenization has
thankfully expanded beyond fishes, plants and birds in
recent years. In a compelling study, Spear and Chown
(2008) examined the effects of ungulate introductions
on biotic similarity across four spatial scales, at three
spatial resolutions within South Africa and among
238
Biological invasions and the homogenization of faunas and floras
Figure 9.5 Distribution of homogenization indices (H) among pairs of state- and province-level floras of the United States
and Canada. The pairs of floras are grouped by degree of native plant similarity (Jnative). Native floras portrayed in the left hand
side panel are more distant (Jnative = 0.00–0.20) than those in the middle (Jnative = 0.20–0.40 and 0.40–0.60) and right-hand
side (Jnative > 0.60) panels. Jtotal refers to floral similarity based on native and non-native species composition. Floral similarity is
based on Jaccard’s coefficient of similarity (J), which ranges from 0 (no species in common) to 1 (all species in common). From
Figure 1 of Qian and Ricklefs (2006).
41 nations located worldwide. They found that between
1965 and 2005, ungulate assemblages had become
two per cent more similar for countries globally and
eight per cent more similar at the coarsest resolution
within South Africa.
Interestingly, species introduced from other continents, as opposed to those introduced from within
Africa, were found to have different effects on patterns
of homogenization. Homogenization was most affected
by translocations of species from neighbouring localities (extra-limital species) (4.6 per cent increase in
similarity), whereas introductions of ungulates from
more distant areas (extra-regional species) tended to
differentiate assemblages (3.8 per cent decreased in
similarity). Quite simply, non-native species introduced
from distant regions are more likely to establish in only
a few localities, resulting in differentiation.
Similar findings have also been reported for plants
and freshwater fishes in the United States (McKinney,
2005; LaSorte & McKinney, 2006). Levels of homogenization were found to increase with increasing
resolution (see Table 9.2) and with time. In the South
African study, from 1971 to 2005, homogenization
by extra-limital introductions increased rapidly after
initially having a smaller homogenizing effect than the
differentiating effect of extra-regional introductions
(Figure 9.6).
9. 4 EN V I R ON MEN T AL AN D H U MAN
DR I V ER S OF B I OT I C
H OMOGEN I Z AT I ON
Environmental change ultimately promotes the geographical expansion of some species and the geographical reduction of others, leading to biotic homogenization
(McKinney & Lockwood, 1999). Habitat loss, pollution,
climate change or other sources of disturbance often
precede, and in a sense prepare, the environment for
changes in beta diversity over time. The research highlighted above, in addition to a number of other studies
in the literature, has provided compelling evidence
linking human-induced environmental change to
biotic homogenization across taxonomic groups.
Collectively, this research has shown that human activities on the landscape are often characterized by greater
increases in taxonomic similarity, suggesting that
Conservation planning in a changing world
239
Figure 9.6 Temporal trends in ungulate homogenization as a result of extra-regional and extra-limital introductions in
South Africa, at the quarter-degree grid cell resolution, between 1971 and 2005. Redrawn from Figure 4 of Spear and Chown
(2008).
humans are playing a central role in promoting the
homogenization process by introducing new species
and favouring the persistence of non-native species
over native species.
For freshwater ecosystems, Scott and Helfman
(2001) reported that cosmopolitan species’ richness
increased and endemic species’ richness decreased in
response to increased watershed deforestation and
density of buildings and roads in Tennessee, USA. At a
larger spatial scale, Marchetti et al. (2001) observed
that measures of human occupancy and aquatic
habitat alteration, including the density of dams and
aqueducts in the watershed, were associated with
increased similarity of zoogeographical provinces in
fish communities in California, USA. However, at a
finer spatial scale, Marchetti et al. (2006) found a negative relationship between change in community similarity and the proportion of the watershed in
development (including commercial, industrial, urban
and suburban) – or, in other words, more developed
watersheds showed greater biotic differentiation. Olden
et al. (2008) found that geographical patterns of
homogenization in Australia were highly concordant
with levels of disturbance associated with human
settlement, infrastructure and land use. These results
suggest that human settlement may directly increase
the likelihood of intentional or accidental non-native
species introductions, and disturbance associated with
physical infrastructure and land-use change may
promote the establishment of these species by disrupting environmental conditions.
Wetland degradation has also led to the homogenization of aquatic and invertebrate communities in
Michigan, USA (Lougheed et al., 2008). Specifically,
habitat homogenization at both the local and landscape scales were found to shift community structure
from a species-rich and spatially heterogeneous community dominated by floating-leaved plants in undeveloped wetlands, to nutrient-rich wetlands dominated
by ubiquitous duckweed (Lemnaceae).
Urban/rural gradient studies have provided important insights into associations between urbanization
and bird and plant homogenization. Blair (2004) found
that temporal changes in bird community composition
varied in a similar fashion along an urban/rural gradient in the oak woodlands of northern California
and the eastern broadleaf forests of Ohio, USA. The
degree of taxonomic overlap in the bird communities
increased from approximately 5 per cent in the least
developed sites to approximately 20 per cent in the
240
Biological invasions and the homogenization of faunas and floras
most urbanized sites – an outcome of the replacement
of local endemic species (often urban-sensitive species)
by ubiquitous non-native species (urban-adapted
species).
By contrast, Clergeau et al. (2006) found that avifaunal similarity of town centres in Europe was actually
lower than in less urbanized habitats – a result that
may have been connected to the larger size of towns
and, thus, greater types of potential habitat in this
study system. The results from this study also suggested that urbanization might cause homogenization
by decreasing the abundance of ground-nesting bird
species and bird species that preferred bush/shrub
habitats. Schwartz et al. (2006) reported floristic
homogenization of urbanized counties in southern
California, whereas they found no change in more
rural areas of northern California. The study of Kühn
and Klotz (2006), on the other hand, found no overall
relationship between patterns of homogenization and
urbanization across Germany.
In summary, although urbanization undoubtedly
plays a role in shaping patterns of biotic homogenization, the exact nature and generality of this relationship is still unclear (McKinney, 2006).
Environmentally mediated interactions between
species may also be an important driver of biotic
homogenization. Holway and Suarez (2006) examined
native ant communities in scrub and riparian habitats
of mediterranean California to test the hypothesis that
the invasion of Argentine ant (Linepithema humile) has
caused biotic homogenization. By comparing invaded
and un-invaded sites across similar habitats, the
authors showed that sites invaded by Argentine ants
have lower beta diversity compared to un-invaded sites.
Specifically, functional homogenization of ant communities occurred via shifting community dominance to
smaller-bodied workers with lower thermal tolerance
and a reduced diversity of behaviours (i.e. nesting
habits, dispersal strategies and foraging behaviours).
Because Argentine ant abundance in seasonally-dry
mediterranean environments is positively correlated
with soil moisture, the authors hypothesized that the
homogenizing effects of the Argentine ant are facilitated by inputs of urban and agricultural water run-off
that acts to create mesic soil conditions. This observation supports the notion that anthropogenic modifications to the environment indirectly cause biotic
homogenization by creating opportunities for the invasion of the Argentine ant, as opposed to threatening
the persistence of native ants directly.
9. 5 B I OT I C H OMOGEN I Z AT I ON AN D
CON S ER V AT I ON
Biotic homogenization is an important dimension of
the modern biodiversity crisis, with significant ecological, evolutionary and social implications (McKinney &
Lockwood, 1999; Olden et al., 2004; 2005). It extends
beyond the narrow focus on elevated extinction rates
to incorporate the other side of the equation: the establishment of non-native species. Biotic homogenization
conjures the prospect of Kunstler’s (1993) The
Geography of Nowhere, in which biotic distinctiveness is
gradually dissolving over time. Consequently, a major
challenge within conservation biogeography is to identify and understand present-day patterns of biotic
homogenization to guide policy aimed at mitigating its
future effects (Rooney et al., 2007).
Clearly, the most effective conservation of biodiversity involves reducing and, where possible, preventing
the two processes generating biotic homogenization –
species invasions and extinctions. The conundrum is
determining the best way to achieve this goal. Because
the key factors facilitating homogenization include
people and habitat transformation (through extinctions or the establishment of non-native species), a first
step towards achieving biodiversity conservation goals
is to focus efforts in areas subject to human activities
and to reduce human-related impacts.
Unfortunately, there is a strong correlation between
human population density and species richness, and
the areas of high biotic diversity that are under the
greatest threat are often in the most populated areas
(Chown et al., 2003; McDonald et al., 2008). Indeed, at
a finer scale of analysis, designated conservation areas
may often attract people to them through perceived
benefits of employment, market access and foreign aid
(Wittmeyer et al. 2008). The increased external threat
from accelerated human population growth does not
bode well for the native biota in these areas, which consequently face the risk of increased homogenization.
In the past, purposeful homogenization was undertaken within countries such as Australia and some
Pacific island territories by acclimatization societies
within colonist human societies who, for a variety of
reasons, wanted to surround themselves with familiar,
colourful or (regarding birds) tuneful species. Even
today, some conservation organizations encourage the
intentional movement or translocation of species,
which may also have the unintended consequence of
promoting homogenization.
Conservation planning in a changing world
This act is a problem when species are introduced
and become established outside of their historical distribution, or where the genetic consequences (e.g.
interspecific hybridization) are not considered. For
example, in parks across southern Africa there has
been a trend to introduce the same suite of species
across nature reserves. Fuelled by tourism and the public’s desire to see large mammals (especially predators),
spotted hyena (Crocuta crocuta), wild dog (Lycaon pictus)
and antelope such as roan (Hippotragus equinus) have
been introduced and have established within areas
where they did not historically occur, or to areas that
are now unsuitable due to small park sizes.
In fact, Spear and Chown (2008) demonstrated that
it is extra-limital introductions that are driving the
homogenization of ungulate assemblages in South
Africa (Figure 9.6). They warn that the potential for
changes in local diversity and ecosystem functioning as
a consequence of translocations should not be underestimated. These concerns contrast with other conservation actors arguing for various forms of rewilding, or
for assisted migrations of species as a climate-change
mitigation strategy (see, e.g. Chapter 3; Donlan, 2007).
The concept of biotic homogenization and differentiation may provide a useful tool in conservation
planning (Rooney et al., 2007). Much attention in conservation has focused on reserve selection and choosing the best network of reserves to maximize
biodiversity coverage. Such efforts have largely focused
on species number, endemism and complementarity as
the metrics that should be optimized (Chapters 6 and
7; Pressey et al., 1993).
Complementarity exists when an area has some biodiversity components that are unrepresented in other
areas. It may thus be possible to use biotic homogenization to monitor whether complementarity goals are
being met. For example, if a network of reserves
becomes more similar over time due to the loss of
unique species, this reduces complementarity (Rooney
et al., 2007).
Importantly, any assessment of complementarity
related to conservation planning should be restricted
to indigenous species only. The inclusion of non-native
species could show increased biotic homogenization
when, in reality, the full set of native species that the
reserve network was designed to conserve still occur.
This idea has much potential, but there are a few
caveats. For example, when dealing with a minimum
set complementarity (each area contains distinctive
species) goal, all areas may lose the same number of
241
unique species over time, and neither complementarity
nor the level of biotic homogenization would change,
yet the true state of biodiversity loss will not be reflected.
9. 6 N OV EL AS S EMB LAGES
Novel assemblages, sometimes referred to as novel or
emerging ecosystems, are communities that consist of
extant species which have not occurred previously in
the same combinations found today (Hobbs et al.,
2006). Increased homogenization of biotas associated
with the massive and accelerating movement of species
within and between regions/provinces is likely to contribute substantially to the creation of novel or noanalogue assemblages.
Although, technically, any area that has lost native
species or gained non-native species is novel in some
respect, some current assemblages have been transformed to such an extent that they are verging on
becoming entirely new assemblages (Williams &
Jackson, 2007). Certainly, in terms of system functioning, many ecosystems have already become ‘novel’.
One of the best examples comes from the San
Francisco Bay, California, which has the dubious distinction of being the most invaded aquatic region on
Earth, with more than half its fish and most of its
bottom-dwelling organisms representing non-native
species (Cohen & Carlton, 1998). The total dominance
(number of species and biomass) of non-native species
has transformed the bay from a pelagic (mid-water)
system to a benthic (bottom) one and productivity has
declined. Invasive species such as Corbula amurensis
(Asian clam), Sphaeroma quoyanum (a burrowing
isopod from Australia and New Zealand) and Spartina
alterniflora (smooth cordgrass) have become among the
most important species in the bay in terms of both
biomass and their role in controlling biological processes in the bay (Cohen & Carlton, 1998).
Although the process of homogenization can create
novel assemblages, global climate change is increasingly likely to magnify this effect. Thus, any prediction
of where novel assemblages will form needs to take into
account not only non-native species introductions, but
also global climate change and the individualistic
responses of species (native and non-native) to environmental change (Chapters 4, 7). Recent models
suggest there will be substantial regions of the world
with novel climates by 2100 (particularly in tropical
and sub-tropical regions) and also that some extant
242
Biological invasions and the homogenization of faunas and floras
Figure 9.7 A conceptual diagram showing how nonanalogue combinations of species arise in response to novel
climates. The set of climates in existence at two periods are
represented as open ellipses. Novel climates are the portions
of the 21st century envelope that do not overlap 20th
century climates, and disappearing climates are the portions
of the 20th century envelope that do not overlap 21st
century climates. Species co-occur only if their fundamental
niches simultaneously intersect with each other and the
current climatic space. Future climate change may cause a
variety of ecological responses, including shifts in species’
distributions (species 1–3), community disaggregation
(species 1 and 3), new communities forming (species 2 and
3), and extinction (species 4). From Figure 1 of Williams
et al. (2007); copyright (2007) National Academy of
Sciences, USA.
create novel environmental conditions or, as Saxon
et al. (2005) refers to them, ‘environmental domains’.
The disappearance or contraction of present environmental domains and the appearance of new domains
will have profound consequences for most species and
the identity of communities today. Climate change is
expected to alter the effectiveness of environmental
filters; to alter the likelihood of species establishing; to
change pathways of species introductions; and to
affect the impact of non-native species (Rahel & Olden,
2008). The combination of novel assemblages and
altered biophysical conditions will result in new
systems that have unknown functional characteristics,
and whose processes and interactions are hard to
predict (Hobbs et al., 2006).
Given the dynamic nature of species’ distributions,
current homogenization patterns and trends are likely
to change too. It is very difficult to predict the make-up
of novel assemblages, given that it is almost impossible
to know which species will co-occur, whether they
will interact and how altered climatic regimes will
influence any interaction. Importantly, many of these
communities may be more, or less, similar across
locations than the native assemblages they replaced.
In other words, homogenization is not the only outcome of the massive movement of species across the
globe.
Perhaps the only certainty is that conservation
efforts will have to intensify to tackle the threat of
anthropogenically-assisted novel assemblages, and
society will be faced with some tough decisions as to
what biodiversity it values.
FOR DI S CU S S I ON
climate types will have disappeared (Williams et al.,
2007).
Because climate is a primary control on species’
distributions and ecosystem processes, novel 21st
century climates may promote the formation of novel
species associations and other ecological surprises. On
the other hand, the disappearance of some extant
climates increases the risk of extinction for species with
narrow geographical or climatic distributions, as well
as the risk of disruption of existing communities
(Figure 9.7).
Of greater concern, perhaps, is the combined effect
of altered climate and other abiotic environmental
characteristics (such as topography or soil type) which
1 How do natural patterns of species invasion differ
from anthropogenically assisted species invasions, and
with what consequences?
2 In the light of social demands and economic
development, what are the most likely timescales and
scenarios of introduction, establishment and spread
of non-native species in the future?
3 What are the ecological consequences of faunal and
floral homogenization?
4 What are the temporal dynamics of taxonomic and
functional homogenization?
5 What are the primary environmental and biological
drivers of biotic homogenization at different spatial
and temporal scales?
Conservation planning in a changing world
6 How will rates and patterns of biotic homogenization respond to shifting pathways of species introductions and future environmental change?
7 What novel species assemblages are likely to emerge
in response to climate change?
8 What might be the consequences of novel ecosystems for biodiversity, ecosystem functioning, and
human societies?
S U G G EST E D R E AD I NG
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Norton, D., Ojima, D., Richardson, D.M., Sanderson, E.W.,
Valladares, F., Vilà, M., Zamora, R., & Zobel, M. (2006)
Novel ecosystems: theoretical and management aspects
243
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McKinney, M.L. & Lockwood, J.L. (1999) Biotic homogenization: a few winners replacing many losers in the next mass
extinction. Trends in Ecology & Evolution, 14, 450–453.
Olden, J.D. (2006) Biotic homogenization: a new research
agenda for conservation biogeography. Journal of
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Rahel, F.J. (2002) Homogenization of freshwater faunas.
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