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CHAPTER 9

Biological Invasions and the Homogenization of Faunas and Floras

Julian D. Olden 1 , Julie L. Lockwood 2 , and Catherine L. Parr 3

1 School of Aquatic and Fishery Sciences, University of Washington, Seattle, USA 2 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

9.1 THE BIOGEOGRAPHY OF SPECIES INVASIONS

In considering the distribution of organic beings over the face of the globe, the fi rst great fact which strikes us is, that neither the similarity nor the dis-similarity 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 pro-ductions of various regions.

(Charles Darwin, 1859 , pp. 395 – 396)

9.1.1 The i nvasion p rocess

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 phys-ical forces, sometimes becoming less common and other times becoming more widespread. In general,

however, the large majority of species are not distrib-uted 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 co - occur. 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 con-vergent evolution in disparate unconnected regions.

With the range expansion of modern humans, ini-tially 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 natu-rally. 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 com-modities themselves.

There is written evidence that intentional move-ments 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 225

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 fi sh, 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 specifi c 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 consider-able amount of research within invasion biology has

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 diffi cult 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 fl ow 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 ‘ inva-sion ’ 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 estab-lished 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 pro-gression 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 individu-als 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

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).

226 Biological invasions and the homogenization of faunas and fl oras

show a much wider array of biological traits than those species that are likely to experience natural long - distance dispersal.

The rate at which non - native populations are estab-lishing 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 set-tlement, 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, contempo-rary rates of biological invasions are nearly one million times higher than the prehistoric rate for Hawaii before human infl uence.

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 fi nding their way out of the native range and attempt-ing 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 trans-polar corridor between the Pacifi c and Atlantic oceans and the formation of the Panamanian land bridge between North and South America during the Great American Interchange permitted a massive fl ux of species between formerly isolated regions (Vermeij, 2005 ; Lomolino et al. , 2006 ). By contrast, human - assisted dispersal events are commonly characterized

therefore focused on attempts to understand which factors differentiate between those species that suc-cessfully progress through all invasion stages and those that do not (Lockwood et al. , 2007 ).

9.1.2 Human - a ssisted v ersus p rehistoric i nvasions

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 rele-vant to students of biogeography because range expan-sions 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 bio-logical invasions?

Human - assisted dispersal of non - native species differs from natural dispersal events in several impor-tant 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 fre-quency with which species disperse. By contrast, modern human - assisted dispersal events happen con-stantly and involve a wide variety of species, which

Table 9.1 A comparison of key characteristics of prehistoric versus human - assisted invasions. Modifi ed from Table 1 of Ricciardi (2007) .

Characteristics Prehistoric invasions Human - assisted invasions

Frequency of long - distance dispersal event Very low Very high Number of species transported per event Low * High Propagule size per event Small * Potentially large Number of mechanisms and routes of dispersal Low High Temporal and spatial scales of mass transport events Episodic (short - distance) Continuous (long - distance) Degree of homogenizing effect Regional Global Potential for interactions with other stressors Low Very high

* Except during biotic interchange events.


Box 9.1 The h uman i mprint on m odern d ay s pecies d ispersal p atterns

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 fi shing; 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 t raffi c 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 fi lled 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 fl eet of approxi-mately 35,000 commercial vessels transports an annual volume of about 3.5 × 10 9 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 infl uences on global trade and are in marked contrast to what we might expect given natural dispersal patterns of marine species via oceanic currents.

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 bio-physical pathways such as prevailing wind directions, jet streams or ocean currents, as would happen for natural dispersal events (Box 9.1 ).

9.1.3 Economic and e cological i mpacts of i nvasion

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 geo-graphical 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.


Airline t raffi c International air travel has been recently pinpointed as a signifi cant factor in the movement of economically damaging pest species and infectious diseases (Tatem, 2009 ). Among others, the Mediterranean fruit fl y 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 disease - carrying mosquitoes have survived long haul fl ights in aircraft cabins (Lounibos, 2002 ). Far - removed regions with similar climates have now been suddenly linked by a busy fl ight 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) identifi ed an ‘ inva-sion 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 fl ights 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 B9.1a (a) The frequency of commercial shipping traffi c 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.)

Figure B9.1b Trends in global shipping cargo volumes and air freight, 1970 – 2005. From Hulme (2009) .


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 trou-blesome 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 BIOTIC HOMOGENIZATION

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 dis-parate 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 fl ight, 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 extinc-tion. 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 dra-matic reconfi guration of modern biodiversity. In fact, changes in diversity patterns at fi ne and coarse scales of analysis can be either concordant or, alternatively, can be decoupled and even confl icting.

• Second, as the number of non - native populations increases, scientists fi nd 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 eco-nomic 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 defi nition of what constituents ‘ impact ’ is somewhat problematic for at least two reasons: 1 There are scientifi c and societal infl uences on the perception of impact (not to mention that the effects of invasive species are often subtle and diffi cult to observe). 2 After impact is perceived, there is a variety of eco-logical factors that determine the level of impact pro-duced (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 & Garc í a - 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 para-keets ( 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 dis-eases (Lounibos, 2002 ; Tatem, 2009 ), the most obvious


diversity across a study region, a phenomenon termed ‘ biotic differentiation ’ by Olden and Poff (2003) .

9.2.1 The p rocess of b iotic h omogenization

In the simplest sense, human activities that increase rates of species invasions and extirpations are the ulti-mate 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 establish-ment 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 homogeni-zation can occur, both empirical data and theoretical modelling suggests that the process is both complex and sensitive to the spatial and temporal scale of inves-tigation (Olden, 2006 ).

9.2.2 Different m anifestations of b iotic h omogenization

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 defi nition of biotic homo-genization does not truly refl ect 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 biologi-cal distinctiveness at any level of organization, includ-ing in their genetic and functional characteristics.

For example, Pautasso (2007) conducted a meta - analysis of the relationship between human popula-tion 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 trans-formation may result in declining local richness but, across larger landscapes and regions, relatively few native species may become entirely extinct, while non - natives 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 com-munities 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 retail-ers (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 circum-stances 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 extirpa-tions can be to increase the mean differentiation


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 butterfl ies, 2 introduced butterfl ies) 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 popula-tions 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 intraspecifi c hybridization (i.e. hybridiza-tion 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 ).


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, encom-passing 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 taxo-nomic 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 (fi shes, birds, plants and mammals) that have received the most attention next.

9.3.1 Fishes

The homogenization of freshwater fi sh 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 set-tlement time frames and found that pairs of states aver-aged 15.4 more species in common now than they did in the past. On average, fi sh faunas became more similar by 7.2 per cent, with the highest increases in similarity observed in western and north - eastern states (Figure 9.3 a). The high degree of biotic homogeniza-tion 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 fi sh homogenization were prima-rily the result of non - native species establishment asso-ciated with fi sh 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

• 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 interspecifi c hybridization (i.e. hybridization between different species within the same genus) (Rhymer & Simberloff, 1996 ). • Third, if extirpations were a cause for faunal homog-enization, 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 suc-cessful establishment. The end result is an increase in the functional convergence of biotas over time associ-ated with the establishment of species with similar ‘ roles ’ in the ecosystem (e.g. high redundancy of func-tional 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 fi ne spatial scales ( < 130 km 2 ) due to more heteroge-neous 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 PATTERNS OF BIOTIC HOMOGENIZATION

Many scientists, including ourselves, have argued that we are entering a period characterized by widespread faunal and fl oral homogenization, fi ttingly dubbed the ‘ Homogecene ’ , in a place appropriately called the ‘ New Pangaea ’ (the original Pangaea being the global super-continent of approximately 250 million years ago).


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 Spatial extent Unit

Change in similarity Reference

Freshwater fi shes

Australia Country - wide Basin divisions 3.0% Olden et al. (2008) 3

North - eastern coastal

Coastal watersheds − 1.4%

Canada Country - wide Provinces/territories 1.3% Taylor (2004) 1

British Columbia Aquatic ecoregions − 3.5%

Europe Country - wide Major basins 2.2% Leprieur et al. (2008) 1

Iberian Peninsula and France

Major basins 17.1% Clavero & Garc í a - Berthou (2006) 1

USA Country - wide States 7.2% Rahel (2000) 1

California Zoogeographic provinces

20.3% Marchetti et al. (2001) 1

Watersheds − 10.7%

South Dakota Geomorphic provinces 8.0% Hoagstrom et al. (2007) 1

Watersheds 2.4%

Minnesota Lakes 9.0% Radomski & Goeman (1995) 1

Kansas Streams − 0.2% Eberle & Channell (2006) 1

Amphibians and reptiles

USA Florida Select counties − 0.8% Smith (2006) 1

Terrestrial plants

Canada & USA Country - wide States and provinces 1.2% Qian & Ricklefs (2006) 1

Canada & USA Select regions States and provinces − 0.6% Rejm á nek (2000) 1

Chile Country - wide Administrative regions 0.3% Castro & Jaksic (2008) 1,7

South - eastern Pacifi c

Islands 2.0% Castro et al. (2007) 1

Europe Germany Grid cells (130 km 2 ) 3.9% K ü hn & Klotz (2006) 5

Europe & USA Select regions Forest stands 3.9% Vellend et al. (2006) 6

Great Britain Country - wide Grid cells (1 km 2 ) − 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

Wisconsin Forest stands 2.6% Rooney et al. (2004) 2


contrasting effects on the changes in community simi-larity (Leprieur et al. , 2008 ). Although biological inva-sions have resulted in an overall increase in faunal similarity on the order of 2.2 per cent (Figure 9.4 a), 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 9.4 b), whereas exotic species (i.e. species originating from outside Europe) tended to decrease their compositional similar-ity ( − 1.6 per cent: Figure 9.4 c). This fi nding is highly consistent with patterns in fl oral homogenization (dis-cussed in Section 9.3.3 ), suggesting that differences in the geographical distribution of exotic and translo-cated species may play an important role in shaping patterns of homogenization.

Clavero and Garc í a - Berthou (2006) used distribu-tional data for freshwater fi sh 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

invasion of 48 non - native fi shes over the past century (Figure 9.3 a).

Similar broad - scale efforts have been conducted in other parts of the world. Recent evidence points to the homogenization of Australian fi sh 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 9.3 b). Similar to the United States and Canada, fi sh faunal homogenization in Australia was the result of the widespread introduction and subse-quent escape/spread of non - native fi shes for recreation (rainbow trout), aquaculture (common carp) and mos-quito control (western mosquito fi sh, Gambusia affi nis ), and from the ornamental/aquarium trade (goldfi sh, Carassius auratus ; guppy, Poecilia reticulata ).

Recent efforts in Europe have shown that exotic and translocated native species generate distinct geograph-ical patterns of biotic homogenization because of their

Table 9.2 Continued

Taxonomic group Location Spatial extent Unit

Change in similarity Reference

Terrestrial birds

Canada & USA Country - wide Transects − 2.0% La Sorte & Boecklen (2005) 4

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

Caribbean Ocean − 0.9%

Indian Oceans 1.8%

Pacifi c Ocean − 0.2%

Terrestrial mammals

Global Select countries Country 2.1% Spear & Chown (2008) 1

South Africa Country - wide Grid cells (0.25 degree) − 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 S ö rensen ’ 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.


regions exhibited a 2.8 per cent increase in community similarity. Signifi cant 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

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 fi shes. In the Netherlands, Van Turnhout et al . (2007) evaluated changes in breeding bird composition over a 25 - year period and found that

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 ).


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

this, their study did fi nd that more highly populated regions located closer to the Atlantic and the Pacifi c coasts of the United States experienced the strongest patterns of homogenization.

At the global scale, Cassey et al . (2007) explored pat-terns of invasion and extirpation and their infl uence on the similarity of oceanic island bird assemblages from the Atlantic, the Caribbean, and the Indian and Pacifi c Oceans. The authors found that patterns of homogenization differed signifi cantly between and

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) .


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 non - native species tended to homogenize fl oras in distant areas whose native plant species differ greatly, but differentiate neighbouring areas that exhibited more closely related native fl oras (Figure 9.5 ). Because few native species have yet been extirpated from state and provincial fl oras, these authors reason that the pattern of homogenization and differentiation probably refl ects the haphazard introduction and establishment of non - native species with respect to suitable habitats. At play is also the natural and human - assisted spread of non - native species with no regard to the ecological con-straints acting on native species.

Smart et al . (2006) used botanical data for fl owering plants in Great Britain to test the hypothesis that plant communities have become taxonomically and func-tionally more similar over the past 20 years in human - dominated landscapes. Although little evidence was found for the taxonomic homogenization of plant com-munities, 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 fl ora of Chile has not shown signifi cant modifi cations over time. Interestingly, this result is not shared for oceanic island fl oral 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 fi shes, 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

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 inves-tigation and the evolutionary history of the region under consideration can infl uence patterns of taxo-nomic homogenization and differentiation within and across what appear to be equivalent spatial units (i.e. ocean basins).

9.3.3 Plants

Evidence for fl oral homogenization comes from studies conducted in many countries at a variety of spatial scales. However, evidence to date suggests that levels of fl oral homogenization are considerably smaller than those observed for freshwater fi shes (Table 9.2 ).

Within the United States, McKinney (2004) found that non - native plant species contributed signifi cantly to fl oral homogenization of 20 parks and local conser-vation areas, although the magnitude was relatively low and sometimes negative (indicating differentia-tion). Cosmopolitan plant species most responsible for the observed homogenization included curly dock ( Rumex crispus ), dandelion ( Taraxacum offi cinale ) and bluegrass ( Poa annua ). Similarly, Schwartz et al . (2006) found that the county fl oras 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 fi ner spatial scale, Rooney et al . (2004) re - surveyed 62 upland forest stands in northern Wisconsin, USA, to assess the degree of fl oral homog-enization 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 homogeniza-tion were greatest in areas without deer hunting, sug-gesting that selective grazing by overabundant deer populations was acting as a key driver of fl oral homogenization.


initially having a smaller homogenizing effect than the differentiating effect of extra - regional introductions (Figure 9.6 ).

9.4 ENVIRONMENTAL AND HUMAN DRIVERS OF BIOTIC HOMOGENIZATION

Environmental change ultimately promotes the geo-graphical expansion of some species and the geograph-ical 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 high-lighted 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 activ-ities on the landscape are often characterized by greater increases in taxonomic similarity, suggesting that

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 conti-nents, 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 locali-ties (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 fi ndings have also been reported for plants and freshwater fi shes in the United States (McKinney, 2005 ; LaSorte & McKinney, 2006 ). Levels of homo-genization 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

Figure 9.5 Distribution of homogenization indices ( H ) among pairs of state - and province - level fl oras of the United States and Canada. The pairs of fl oras are grouped by degree of native plant similarity ( J native ). Native fl oras portrayed in the left hand side panel are more distant ( J native = 0.00 – 0.20) than those in the middle ( J native = 0.20 – 0.40 and 0.40 – 0.60) and right - hand side ( J native > 0.60) panels. J total refers to fl oral similarity based on native and non - native species composition. Floral similarity is based on Jaccard ’ s coeffi cient 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) .


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 disrupt-ing environmental conditions.

Wetland degradation has also led to the homogeni-zation of aquatic and invertebrate communities in Michigan, USA (Lougheed et al. , 2008 ). Specifi cally, habitat homogenization at both the local and land-scape scales were found to shift community structure from a species - rich and spatially heterogeneous com-munity dominated by fl oating - leaved plants in unde-veloped wetlands, to nutrient - rich wetlands dominated by ubiquitous duckweed (Lemnaceae).

Urban/rural gradient studies have provided impor-tant 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 gradi-ent 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

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 fi sh communities in California, USA. However, at a fi ner spatial scale, Marchetti et al . (2006) found a nega-tive relationship between change in community simi-larity 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

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) .


9.5 BIOTIC HOMOGENIZATION AND CONSERVATION

Biotic homogenization is an important dimension of the modern biodiversity crisis, with signifi cant ecologi-cal, 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 estab-lishment 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 iden-tify 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 biodiver-sity 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 extinc-tions or the establishment of non - native species), a fi rst 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 fi ner scale of analysis, designated conservation areas may often attract people to them through perceived benefi ts 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 con-sequently face the risk of increased homogenization.

In the past, purposeful homogenization was under-taken within countries such as Australia and some Pacifi c 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.

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 avifau-nal 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 sug-gested 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 fl oristic homogenization of urbanized counties in southern California, whereas they found no change in more rural areas of northern California. The study of K ü hn 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 homogeniza-tion, the exact nature and generality of this relation-ship 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. Specifi cally, functional homogenization of ant commu-nities 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 facili-tated by inputs of urban and agricultural water run - off that acts to create mesic soil conditions. This observa-tion supports the notion that anthropogenic modifi ca-tions to the environment indirectly cause biotic homogenization by creating opportunities for the inva-sion of the Argentine ant, as opposed to threatening the persistence of native ants directly.


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 refl ected.

9.6 NOVEL ASSEMBLAGES

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 con-tribute substantially to the creation of novel or no - analogue 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 trans-formed to such an extent that they are verging on becoming entirely new assemblages (Williams & Jackson, 2007 ). Certainly, in terms of system function-ing, many ecosystems have already become ‘ novel ’ .

One of the best examples comes from the San Francisco Bay, California, which has the dubious dis-tinction of being the most invaded aquatic region on Earth, with more than half its fi sh 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 alternifl ora (smooth cordgrass) have become among the most important species in the bay in terms of both biomass and their role in controlling biological proc-esses in the bay (Cohen & Carlton, 1998 ).

Although the process of homogenization can create novel assemblages, global climate change is increas-ingly 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 envi-ronmental 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

This act is a problem when species are introduced and become established outside of their historical dis-tribution, or where the genetic consequences (e.g. interspecifi c 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 pub-lic ’ 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 under-estimated. These concerns contrast with other conser-vation 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 differen-tiation may provide a useful tool in conservation planning (Rooney et al. , 2007 ). Much attention in con-servation has focused on reserve selection and choos-ing 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 bio-diversity components that are unrepresented in other areas. It may thus be possible to use biotic homogeniza-tion 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


create novel environmental conditions or, as Saxon et al . (2005) refers to them, ‘ environmental domains ’ . The disappearance or contraction of present environ-mental 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 fi lters; 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 diffi cult 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 infl uence 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 out-come 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 DISCUSSION

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 fl oral 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?

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

Figure 9.7 A conceptual diagram showing how non - analogue 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.


6 How will rates and patterns of biotic homogeniza-tion respond to shifting pathways of species introduc-tions 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 eco-systems for biodiversity, ecosystem functioning, and human societies?

SUGGESTED READING

Elton , C.S. ( 1958 ) The ecology of invasions by animals and plants . Methuen , London .

Hobbs , R.J. , Arico , S. , Aronson , J. , Baron , J.S. , Bridgewater , P. , Cramer , V.A. , Epstein , P.R. , Ewel , J.J. , Klink , C.A. , Lugo , A.E. , Norton , D. , Ojima , D. , Richardson , D.M. , Sanderson , E.W. , Valladares , F. , Vil à , M. , Zamora , R. , & Zobel , M. ( 2006 ) Novel ecosystems: theoretical and management aspects

of the new ecological world order . Global Ecology and Biogeography , 15 , 1 – 7 .

McKinney , M.L. & Lockwood , J.L. ( 1999 ) Biotic homogeniza-tion: 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 Biogeography , 33 , 2027 – 2039 .

Rahel , F.J. ( 2002 ) Homogenization of freshwater faunas . Annual Review of Ecology and Systematics , 33 , 291 – 315 .

Riccardi , A. ( 2007 ) Are modern biological invasions an unprecedented form of global change? Conservation Biology , 21 , 329 – 336 .

Sax , D.F. , Stachowicz , J.J. , Brown , J.H. , Bruno , J.F. , Dawson , M.N. , Gaines , S.D. , Grosberg , R.K. , Hastings , A. , Holt , R.D. , Mayfi eld , M.M. , O ’ Connor , M.I. , & Rice , W.R. ( 2007 ) Ecological and evolutionary insights from species inva-sions . Trends in Ecology & Evolution , 22 , 465 – 471 .

Strayer , D.L. , Eviner , V.T. , Jeschke , J.M. , & Pace , M.L. ( 2006 ) Understanding the long - term effects of species invasions . Trends in Ecology & Evolution , 21 , 645 – 651 .

biological invasions and the homogenization of faunas and ...depts.washington.edu › oldenlab ›...

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