Early View (EV): 1-EV
megafauna have been moved to new regions and between
continents. Introductions of megafauna worldwide may
have inadvertently provided refuge for threatened mega-
fauna, increased regional large herbivore species richness, and
restored or added ecological functions. Acknowledgement
of this possibility is being fostered by the burgeoning con-
cept of ‘ rewilding, ’ which includes eﬀ orts to proactively
introduce species in order to provide refuge and to restore
lost ecological processes (Donlan et al. 2006, Svenning
et al. 2016). However, much remains unknown about the
contribution of already introduced populations to global
Given that introduced populations are often unwanted
and considered components of anthropogenic harm, the
existence of populations that are simultaneously introduced
and threatened or extinct in their native ranges has been
highlighted as a conservation paradox (Marchetti and
Engstrom 2015). Indeed, the considerable redistribution
of biota that characterizes the Anthropocene may be a
countercurrent to the extinction crisis by providing refuge
and new opportunities for threatened species (Wallach et al.
2015). However, comprehensive analyses of the interaction
between the processes of extinction and redistribution have
not been conducted.
To assess the potential conservation values of introduced
megafauna we compiled current information on their threat
statuses and population trends in their native ranges, their
relative population sizes in and out of their native ranges,
and their functional roles. To understand how introduced
megafauna have potentially rewilded the world, we assessed
Ecography 40: 001–010, 2017
© 2017 e Authors. Ecography © 2017 Nordic Society Oikos
Subject Editor: Jacquelyn Gill. Editor-in-Chief: Hanna Tuomisto. Accepted 24 July 2017
Terrestrial herbivorous megafauna are undergoing severe
declines around the world. Of 74 extant large terrestrial her-
bivorous mammal species with body masses ⱖ 100 kg, 44
( ∼ 60%) are threatened with extinction (Ripple et al. 2015).
e decline of this functional group began 10 000 – 50 000 yr
ago, most likely due to overhunting by humans during the
late Pleistocene (Barnosky et al. 2004, Bartlett et al. 2015).
Large ( ⱖ 100 kg) herbivorous megafauna (henceforth
‘ megafauna ’ ) perform distinct roles that contribute to the
functioning of ecological systems. Megafauna consume
ﬁ brous vegetation, which can beneﬁ t smaller herbivores,
reduce ﬁ re risk, accelerate rates of nutrient cycling by orders
of magnitude, and shift plant community structure by
facilitating coexistence between diﬀ erent plant functional
types. Due to their large size, these organisms cause physi-
cal disturbance and disperse large seeds and nutrients great
distances (Ripple et al. 2015). e considerable loss of this
functionality at the end of the Pleistocene had dramatic eﬀ ects
on plant community structure, ﬁ re regimes, nutrient and
mineral cycling across landscapes, and community assembly
(Gill et al. 2009, Ripple and Van Valkenburgh 2010, Smith
et al. 2015, Bakker et al. 2016a, Doughty et al. 2016a, b,
c, Malhi et al. 2016). Modern declines have similar conse-
quences for terrestrial ecosystems and community dynam-
ics (Ripple et al. 2015) and have led to broad international
calls for immediate action to conserve the world ’ s remaining
mammalian megafauna (Ripple et al. 2016, 2017).
Less well considered is the role of megafauna introductions
on their conservation and on ecosystem function. Since the
advent of the Anthopocene, particularly in the past 200 yr,
Introduced megafauna are rewilding the Anthropocene
Erick J. Lundgren , Daniel Ramp , William J. Ripple and Arian D. Wallach
E. J. Lundgren (http://orcid.org/0000-0001-9893-3324) (email@example.com), Arizona State Univ., School of Life Sciences, Tempe, AZ, USA.
– D. Ramp, A. D. Wallach and EJL, Centre for Compassionate Conservation, School of Life Sciences, Univ. of Technology Sydney, NSW, Australia.
– W. J. Ripple, Global Trophic Cascades Program, Dept of Forest Ecosystems and Society, Oregon State Univ., Corvallis, OR, USA .
Large herbivorous mammals, already greatly reduced by the late-Pleistocene extinctions, continue to be threatened with
decline. However, many herbivorous megafauna (body mass ⱖ 100 kg) have populations outside their native ranges. We
evaluate the distribution, diversity and threat status of introduced terrestrial megafauna worldwide and their contribution
towards lost Pleistocene species richness. Of 76 megafauna species, 22 ( ∼ 29%) have introduced populations; of these eleven
(50%) are threatened or extinct in their native ranges. Introductions have increased megafauna species richness by between
10% (Africa) and 100% (Australia). Furthermore, between 15% (Asia) and 67% (Australia) of extinct species richness,
from the late Pleistocene to today, have been numerically replaced by introduced megafauna. Much remains unknown
about the ecology of introduced herbivores, but evidence suggests that these populations are rewilding modern ecosystems.
We propose that attitudes towards introduced megafauna should allow for broader research and management goals.
the contribution of introduced megafauna to continental
assemblages, and the contribution of introduced megafauna
to Anthropocene richness relative to the Holocene and
We searched for introduced populations of herbivorous
megafauna (mammals only) with body masses ⱖ 100 kg
based on Ripple et al. (2015) using Long (2003) and supple-
mented with online searches (Google Scholar and Google)
using the terms ‘ feral ’ , ‘ introduced ’ , ‘ invasive ’ , ‘ exotic ’ and
‘ non-native ’ . We used grey literature (e.g. government
reports) and journalism sources (e.g. e New York Times)
alongside peer-reviewed literature to identify megafauna
populations outside their native ranges. Data collection con-
cluded in July, 2017. While some native megafauna popu-
lations live in fenced and managed conditions (e.g. Kruger
National Park), only free-roaming wild introduced popula-
tions were included because it was not clear if fenced/man-
aged introduced populations are ecologically viable in their
To understand to what extent introduced megafauna
represent the taxonomic diversity of the world ’ s remaining
megafauna, we calculated the number of large herbivore
families represented by introduced species, the number of
genera of each family represented by introduced species, and
the percentage of species with introduced populations within
each taxonomic family.
To determine the potential conservation value of intro-
duced megafauna as refuge populations, we compiled
IUCN (2017) Red List threat statuses and trends in each
species ’ historic native ranges and the proportion of each
population that is currently outside of its native range
(Supplementary material Appendix 2 Table A1). Wild
post-domestic species were assigned the threat status of
their pre-domestic ancestor. For example, introduced wild
dromedary camels Camelus dromedarius originate from the
domesticated form of an extinct camel species (possibly
C. thomasi ), and were therefore considered extinct in the
wild in their native range.
To understand to what geographic extent introduced
megafauna have rewilded the world, we calculated mega-
fauna species richness by Taxonomic Databases Working
Group level 3 countries (henceforth TDWG), which are
bio-geographic units deﬁ ned by political (nation, state, prov-
ince, or district) boundaries at a biologically relevant scale
(Brummitt 2001). Inter- and intra-continental introduc-
tions were included in this comparison. e distributions
of introduced megafauna were determined from literature
and Google searches (Supplementary material Appendix 1
Data A1). Geographic ranges for native megafauna were
downloaded from the IUCN (2017) Red List. e percent-
age of each TDWG country ’ s megafauna assemblage that is
introduced was calculated and compared between continents
to understand how introductions have altered continental
We assessed how Anthropocene megafauna richness com-
pares to those of past geological epochs. For each continent,
we compared megafauna species richness and conservation
status between the late Pleistocene (50 000 – 10 000 BP),
Holocene ( ⬍ 10 000 BP), and Anthropocene (past ∼ 200 yr)
epochs. Only inter-continental introduced megafauna were
Pleistocene species were classiﬁ ed as ‘ extinct ’ , ‘ extir-
pated ’ or ‘ survived ’ based on their fate through the late-
Pleistocene extinction. Pleistocene megafauna presence was
based on Sandom et al. (2014) and body masses ( ⱖ 100 kg)
were conﬁ rmed through literature searches. e Holocene
included species from the end of the Pleistocene until the
Anthropocene. Holocene species included ‘ survived ’ taxa,
natural immigrants, and species that went extinct during
the Holocene (e.g. aurochs Bos primigenius and dromedary
camel). Anthropocene species included ‘ survived ’ , ‘ survived,
threatened ’ , ‘ introduced ’ , and ‘ introduced, threatened ’ spe-
cies, reﬂ ecting their current IUCN (2017) threat statuses
(Supplementary material Appendix 3 Table A2).
To describe the range of functional traits of introduced
megafauna, we reviewed their average body masses, habitat
types, dietary types (grazer, browser, or intermediate), and
other unique traits using the IUCN (2017) and published
Twenty-two (32%) of the 76 extant megafauna species have
established wild populations outside their native ranges
(Supplementary material Appendix 2 Table A1). Sixteen are
inter-continental introductions, two are intra-regional but
overcame oceanic barriers, and four are intra-continental.
By including post-domesticates of extinct heritage, an addi-
tional two species (the dromedary camel and cattle Bos tau-
rus ) are added to the 74 remaining native megafauna. Six
additional species were excluded from analysis: three species
because they appear to be conﬁ ned to game ranches, one
because introduced populations are described as semi-wild,
and two because of uncertain taxonomic relation to already
Six (55%) of the eleven families containing megafauna
species have established populations outside their native
ranges. Introduced species represent between 29% (Equidae)
and 56% (Cervidae) of the megafauna species within their
families (Fig. 1). Likewise, introduced populations repre-
sent between 50% (Camelidae) and 100% (Equidae) of the
megafauna genera within their families.
Of the 22 species with introduced populations, eleven
(50%) are threatened or extinct in their native ranges (Fig. 2).
is includes four ( ∼ 18%) Vulnerable non-domesticated
species, three (14%) post-domestics whose progenitors are
Endangered, one ( ∼ 5%) Endangered non-domesticated spe-
cies, two (9%) post-domestic species whose wild progenitors
are Extinct, and one ( ∼ 5%) post-domestic whose progeni-
tor is Critically Endangered. All six post-domestic species are
extinct or threatened in their native ranges. Of the remain-
ing eleven introduced megafauna, three (14%) are Near
reatened, and eight are ranked as Least Concern in their
native ranges, of which 50% have stable population trends,
22% are increasing, and 11% are declining (Fig. 2). Of the
20 introduced species with surviving native populations,
eleven (55%) are declining in their native ranges, ﬁ ve (25%)
are stable, and four (20%) are increasing (Supplementary
material Appendix 2 Table A1). In all, 64% of introduced
megafauna are threatened or declining in their native ranges
On average, over 38% (ranging between ⬍ 1 and 100%)
of megafauna populations are outside of their native
ranges. Whereas two species have relatively small (possibly
∼ 100 individuals) populations outside their native ranges
(hippopotamus Hippopotamus amphibius , and Asian ele-
phant Elephas maximus ), twelve populations are estimated in
the thousands and up to over 1 million individuals (Fig. 3,
Supplementary material Appendix 2 Table A1).
By including introduced megafauna, the worldwide
distribution of megafauna species richness increases sig-
niﬁ cantly (Fig. 4). Introduced megafauna have sub-
stantially increased continental megafauna richness and
TDWG-country-scale species richness within each conti-
nent: 62% of South American (mean ⫾ SD, 37% ⫾ 34%),
57% of North American (24% ⫾ 37%), 33% of European
(36% ⫾ 33%), 11% of Asian (17% ⫾ 34%), and 11% of
African (10% ⫾ 27%) megafauna are introduced. Introduced
megafauna comprise at least 75% of the megafauna assem-
blages of 56 of the 369 (15%) TDWG countries.
Strikingly, the entire continental megafauna assemblage
of Australia is composed of introduced species. Australia lost
all megafauna species during the Pleistocene extinctions,
yet has become home to eight introduced species in the
Anthropocene, including the Endangered banteng Bos javan-
icus , the world’s only population of wild dromedary camel,
the Vulnerable sambar deer Rusa unicolor , and the water
buﬀ alo Bubalus bubalis , the descendant of the Endangered
water buﬀ alo B. arnee. Wild donkeys Equus asinus , whose
progenitor, the African wild ass E. africanus is Critically
Endangered, and Endangered horses E. ferus caballus , have
also found refuge in Australia, as well as in North America,
South America, and Europe.
Percent of family with introduced
populations by IUCN threat status
Least Concern Near−Threatened
Critically Endangered Extinct
Figure 1. reatened megafauna species are ﬁ nding refuge outside
their native ranges. Percentage of megafauna in each family with
introduced populations, colored by IUCN threat categories in their
native ranges. Number within parentheses indicates total number
of megafauna within each family.
Number of introduced species
Decreasing Stable Increasing N/A
Figure 2. e number of introduced megafauna species by IUCN
(2017) threat status and population trends in their native ranges.
e majority (59%) of introduced megafauna are threatened or
have declining populations in their native ranges.
Percent of total populations
Least Concern Near−Threatened
Critically Endangered Extinct
Figure 3. Percent of global populations of megafauna that are
introduced. Color indicates IUCN (2017) status. Bars indicate
high and low estimates if multiple estimates were found. Includes
only species with known population sizes in native and non-native
* indicates post-domestic species.
Native megafauna richness
Introduced megafauna richness
All megafauna richness
Figure 4. Contribution of introduced megafauna to TDWG-country species richness. (a) Native megafauna species richness (b) introduced
megafauna species richness, (c) all megafauna species richness, and (d) percent contribution of introduced species to TDWG-country
megafauna assemblages. Inter- and intra-continental introductions were included. Native richness was derived from IUCN (2017) species
distribution data. Introduced species distributions are available in Supplementary material Appendix 1 Data A1.
introduced megafauna in the Anthropocene, so that there
are currently more megafauna species per continent than
at the end of the Holocene. Introduced megafauna have
numerically replaced extinct species richness in Australia by
67%, in South America by 21%, in North America by 26%,
in Europe by 33%, in Asia by 15%, and in Africa by 31%
(Fig. 5, Table 1).
Megafauna are likely to have signiﬁ cant functional
roles in their introduced ranges. eir average body
masses ranges from 109 to 3270 kg (median ⫽ 256 kg,
mean ⫽ 526 kg, SD ⫽ 697 kg) (Table 2), which is rep-
resentative of the native megafauna body mass distribu-
tion ranging from 100 to 3825 kg (median ⫽ 238 kg,
mean ⫽ 496 kg, SD ⫽ 666 kg). Introduced megafauna are
primarily grazers (45% of species) or intermediate grazers
and browsers (41% of species), and three species (14%) are
primarily browsers (Table 1). Introduced megafauna are
adapted for habitats ranging from Arctic tundra (muskox
Ovibos moschatus ) to tropical forest (sambar deer) and
deserts (dromedary camels) (Table 1). Although there is
little known about the speciﬁ c ecological functionalities
of several introduced megafauna, many introduced species
are known for unique traits, such as the ability to drink
brackish water and consume halophytic plants (dromedary
camel) or to survive without surface water (gemsbok Oryx
gazella ) (Table 2).
Introduced megafauna represent a signiﬁ cant proportion of
the remaining taxonomic diversity of their functional group
and are themselves signiﬁ cantly threatened in their historic
native ranges. is raises the question of how to assign con-
servation value in an era of extinction and redistribution.
Conservation biology is a ﬁ eld driven by a plurality of values,
which oﬀ er various visions at diﬀ erent scales and times
(Sandbrook et al. 2011). Many current schools of thought
prioritize the conservation of species considered to be native
at the local and regional scale. However, given the ongoing
global extinction process, more research and dialogue is
needed to understand when these values may undermine
other conservation goals and values.
While many introduced populations were formerly
domesticated, they may still eﬀ ectively represent their wild
relatives. Introduced populations of Endangered banteng
in northern Australia have maintained high genetic ﬁ delity
Late Pleistocene losses of megafauna species (100% for
Australia, 89% for South America, 89% for North America,
53% for Europe, 41% for Asia, and 27% for Africa) and
Holocene losses (14% for Europe, 5% for Asia, and 3% for
Africa) were substantial. Following the Pleistocene, North
American species richness increased from 4 to 6 due to
immigration of wapiti Cervus canadensis and moose Alces
alces from Eurasia concurrent with the arrival of the ﬁ rst
humans to the continent (Hundertmark et al. 2002, Meiri
et al. 2014). Reductions in species richness on all continents
since the Pleistocene have been counteracted by gains from
Africa Asia Europe North
Epoch by continent
Number of megafauna species
Extinct Extirpated Immigrated
Survived Survived, threatened Introduced
Figure 5. Megafauna species richness per epoch by continent.
‘ Extinct ’ indicates species that went extinct in the wild on all
continents; ‘ extirpated ’ are species that survived elsewhere; ‘ immi-
grated ’ are species that immigrated without human intervention;
‘ introduced ’ indicates species introduced by humans; ‘ introduced,
threatened ’ are introduced species threatened in their native ranges;
‘ survived ’ are species that were still present into the following epoch;
‘ survived, threatened ’ are threatened native species (Supplementary
material Appendix 3 Table A2).
Table 1. Changes in megafauna species richness from the Pleistocene to the Anthropocene. In column 2, percent survived is the percent of
megafauna to survive the late Pleistocene extinctions; in column 3, percent lost/gained is the percent change in Holocene species richness
due to extinction/immigration during the Holocene; in column 4, percent replaced is the percent of all extinct megafauna richness (Pleistocene
and Holocene) to be numerically replaced by introductions in the Anthropocene.
* indicates natural immigration from Eurasia to North
America during the early Holocene.
Continent Pleistocene species richness
Holocene species richness
Africa 44 32 (73%) – 1 ( – 3%) 35 (31%)
Asia 61 36 (59%) – 2 ( – 6%) 38 (14%)
Australia 12 0 (0%) N/A 8 (67%)
Europe 15 7 (47%) – 1 ( – 14%) 9 (33%)
North America 35 4 (11%) ⫹ 2 ( ⫹ 33%)
* 14 (26%)
South America 44 5 (11%) 0 (0%) 12 (18%)
Table 2. Functional traits of introduced megafauna. ABM is average body mass (Jones et al. 2009); foraging type is ‘ B ’ is browser, ‘ G ’ is grazer, and G/B are intermediate; habitats are derived from IUCN
Redlist species accounts IUCN (2017) Red List.
Species Common name ABM Type Habitat Known or potential unique ecological functions
Alces alces Moose 541 B Woodlands, tundra, montane forests Browse at heights up to 2 m, affecting stand height and canopy composition (Pastor et al.
Bison bison Bison 625 G Grasslands, open forests Create wallows that become ephemeral pools, serve as ﬁ re breaks, and increase landscape
scale plant diversity (Knapp et al. 1999).
Bos javanicus Banteng 636 G/B Open dry forests
Bos taurus Cattle 613 G Numerous
Boselaphus tragocamelus Nilgai 182 G/B Open grasslands Open trails in dense shrubland, capable of jumping 2.5 m high, potentially sustaining seed/
nutrient dispersal in fenced landscapes (Leslie 2008).
Bubalus bubalis Water buffalo 919 G/B Moist grasslands, marshes Used for conservation grazing to maintain open water habitat for birds and ﬁ sh (BBC News
Camelus dromedarius Dromedary camel 488 B Desert scrub Salt-tolerant (Root-Bernstein and Svenning 2016); large home ranges (Spencer et al. 2012), may
redistribute sodium (Doughty et al. 2016a).
Cervus elaphus Red deer 241 G/B Generalist
Connochaetes gnou Black wildebeest 157 G Short-grass grasslands
Elephas maximus Asian elephant 3270 G/B Tropics Ecological engineer in native range by dispersing large seeds and removing trees (Donlan et al.
Equus asinus Donkey 180 G/B Deserts Digs wells used by other species.
Equus caballus Horse 400 G Grasslands, open forests Feeds on coarse, abrasive grasses (Naundrup and Svenning 2015).
Hippopotamus amphibius Hippopotamus 1536 G Aquatic daytime refuge; grasslands Maintain grazing meadows, fertilize riparian systems (Bakker et al. 2016b), unstudied in
Hippotragus niger Sable antelope 236 G/B Woodland edges
Kobus ellipsiprymnus Waterbuck 204 G Savanna woodlands Riparian grazer, likely inﬂ uences riparian vegetation and river geomorphology (Naiman and
Rogers 1997, IUCN 2017).
Oryx gazella Gemsbok 188 G Desert scrub, desert grassland Dig wells used by other species (Hamilton et al. 1977).
Ovibos moschatus Muskox 313 G Arctic tundra Few other herbivores adapted to extreme arctic environment (Schmidt et al. 2015).
Ovis ammon Argali 114 G Steep, rocky environments
Rangifer tarandus Reindeer 109 G/B Mountains, arctic tundra Grazing can alter arctic albedo, causing temperature reductions that may counteract climate
change (te Beest et al. 2016). Uniquely capable of digesting lichens (Palo 1993).
Rucervus duvaucelii Barasingha 171 G Forests, riparian grasslands Riparian grazer, likely inﬂ uences riparian vegetation and river geomorphology (Naiman and
Rogers 1997, IUCN 2017).
Rusa unicolor Sambar 178 G/B Generalist
co-evolutionary history versus ecological context in deter-
mining species coexistence and ecosystem function (Wallach
et al. 2015).
Introduced megafauna vary in body mass considerably,
which inﬂ uences their ability to open thickets and digest
coarse ﬁ brous vegetation and thus their relation to plant
communities and other herbivores. Introduced megafauna
also possess unique functional adaptations that may be of
ecological signiﬁ cance in their new ranges. For example,
introduced camels are capable of ingesting brackish water
and consuming halophytic plants (Root-Bernstein and
Svenning 2016), which in conjunction with their large
home ranges (Spencer et al. 2012) may contribute to the
megafaunal redistribution of terrestrial salts (Doughty et al.
2016a). Likewise, the ability of gemsbok Oryx gazella to
survive without surface water (Hamilton et al. 1977) likely
allows it to occupy novel niches in the North American
deserts in which it now lives.
ere is substantial and growing evidence that intro-
duced species can perform signiﬁ cant and desirable eco-
logical roles (Schlaepfer et al. 2011). Bighorn sheep forage
more eﬃ ciently, with less time invested in vigilance behav-
iors in mixed herds with introduced wild horses (Coates and
Schemnitz 1994). Giant tortoises introduced onto oceanic
islands as substitutes for extinct species are dispersing large-
seeded endemic plants and shaping plant communities
through grazing (Hansen et al. 2010). Intentional introduc-
tions of horses and cattle in the Oostvaardersplassen nature
reserve in the Netherlands have created Pleistocene-like
savanna conditions in a temperate deciduous forest envi-
ronment (Vera 2009). In North America and Australia, the
drying and constriction of desert springs and the extinc-
tion of several endemic ﬁ sh populations was linked to
the removal of wild introduced megafauna whose grazing
appeared to maintain open-water habitat (Kodric-Brown
and Brown 2007).
Likewise, our own ongoing research is yielding simi-
larly surprising observations. For example, in the Sonoran
Desert of North America, wild donkeys ( ‘ burros ’ , E. asi-
nus ) dig groundwater wells of more than a meter in depth
(Supplementary material Appendix 4 Movie A1). ese wells
are common wherever groundwater approaches the surface,
have been recorded in use by more than thirty mammal and
bird species, and in certain conditions become nurseries for
riparian trees (Fig. 6). It is possible that by creating new water
sources across the landscape, maintaining access to receding
water-tables during droughts, and providing conditions ideal
for the germination of riparian trees, wild donkeys play a
facilitative role, one that may improve the resilience of these
arid ecosystems to climate change. Furthermore, given the
ubiquity of taxa whose contemporaries dig wells, such as
Proboscideans (Ramey et al. 2013) and other equids (Feh
et al. 2002) in the North American Pleistocene, it is likely
that introduced donkeys have restored a functionality lost
from these landscapes.
Unfortunately, little more is known about the ecological
functions of megafauna outside their native ranges because the
majority of studies are conducted on the premise that intro-
duced species are harmful and should be suppressed or eradi-
cated. Future research on the ecological functions of introduced
megafauna, under varying ecological contexts (e.g. predator
to their pre-domestic ancestors (Bradshaw et al. 2005).
Likewise, domesticated horses retain a substantial component
of the genetic diversity of extinct Holarctic horse lineages
(Lippold et al. 2011). Given that the closest wild relatives of
all six post-domestic megafauna are Endangered or extinct,
it appears that domestication has provided a crucial bridge
for certain species from the pre-pastoral wild landscapes of
the early Holocene to the post-industrial wild landscapes of
Evolutionary and ecological change has also been witnessed
in post-domestic populations. Wild goats Capra aegragus on
Aldabra Atoll regularly drink saltwater when freshwater is
absent (Burke 1990). Wild sheep Ovis aries show higher
resistance to local parasites than sympatric domestic sheep.
Wild Ossabaw island pigs Sus scrofa have unique lipid struc-
tures (Van Vuren and Hedrick 1989). Wild cattle in Mexico
do not linger in riparian areas like their sympatric domestic
cousins due to altered predation threats (Hernandez et al.
1999). Native Torresian crows Corvus orru appear to have
developed a mutualistic grooming behavior on introduced
banteng in Australia (Bradshaw and White 2006).
Like all herbivores, introduced megafauna can exert
strong grazing or browsing pressure to the detriment of
other species, most notably where apex predators are extir-
pated or continue to be persecuted (Wallach et al. 2010).
Unfortunately, much of the research to document these
eﬀ ects has ignored the ecological context of predator con-
trol, which is to ignore an important explanatory variable
for the density-dependent eﬀ ects of all herbivores. Indeed,
wild horses in the United States may be limited by mountain
lions (Turner and Morrison 2001) and dingoes appear to
suppress populations of wild donkeys in Australia (Wallach
et al. 2010). e potential to inﬂ uence the ecologies of intro-
duced megafauna by protecting or restoring large predators
is an important topic for further research.
In the Pleistocene, the ecological inﬂ uences of herbivo-
rous megafauna on disturbance regimes, seed dispersal,
nutrient cycling, and community structure were ubiquitous.
Introduced megafauna have potentially augmented this lost
functional and taxonomic diversity across most continents,
particularly in those regions most depleted: Australia, North
America, and South America (Fig. 4). Asia and Africa have
retained many Pleistocene megafauna and have fewer intro-
duced species. Several of these introductions restore taxo-
nomic analogues to extinct Pleistocene species. For example,
introduced donkeys are morphologically similar to conge-
neric extinct North American and South American stilt-
legged horses, and the modern wild horse is the same species
as the horse of the Holarctic Pleistocene (Weinstock et al.
e late Pleistocene extinctions in Australia included all
megafauna and many browsing herbivores, the loss of which
appears to have led to increased ﬁ re frequency and altered
plant community structure (Miller et al. 2005, Rule et al.
2012). Introduced megafauna, especially browsers such as
dromedary camels, may reverse these ecological state shifts.
However, determining how introductions of taxonomically
dissimilar species restore or add new functionalities within
insular ecosystems (there are no surviving taxonomic ana-
logues to Australia ’ s Pleistocene marsupial megafauna)
requires further research into the relative importance of
and without potential predators in the novel ecosystems of
the Anthropocene will be essential in reconciling the con-
cerns of local managers with global conservation eﬀ orts and
will bring new attention to the emerging eco-evolutionary
trajectories of these populations.
Acknowledgements – We thank M. Sluk for his research assistance,
and J. Stromberg, F. Horgan, C. Sandom, V. K. Harris, and two
anonymous reviewers for helpful feedback on earlier drafts.
Conﬂ icts of interest – e authors declare no conﬂ icts of interest.
Bakker, E. S. et al. 2016a. Combining paleo-data and modern
exclosure experiments to assess the impact of megafauna extinc-
tions on woody vegetation. – Proc. Natl Acad. Sci. USA 113:
847 – 855.
Bakker, E. S. et al. 2016b. Assessing the role of large herbivores in
the structuring and functioning of freshwater and marine
angiosperm ecosystems. – Ecography 39: 162 – 179.
Barnosky, A. D. et al. 2004. Assessing the causes of late Pleistocene
extinctions on the continents. – Science 306: 70 – 75.
Bartlett, L. J. et al. 2015. Robustness despite uncertainty: regional
climate data reveal the dominant role of humans in explaining
global extinctions of Late Quaternary megafauna. – Ecography
39: 152 – 161.
BBC News 2004. ‘ Buﬀ alo improve wildlife habitat ’ . – BBC News,
15 February 2004.
control, landscape connectivity), will be essential to understand
the novel megafaunal communities of the Anthropocene.
Reassessing conservation attitudes towards introduced
megafauna may ﬁ nd synergy with other conservation goals.
Introduced megafauna are likely vulnerable to similar threats
as native megafauna as they require large tracts of land and
may be vulnerable to exploitation. Valuing introduced mega-
fauna as umbrella or ﬂ agship species in eﬀ orts to expand
protected areas or establish movement corridors would con-
tribute to important conservation goals. Broadening the
range of wildlife valued and protected by conservation prac-
titioners could also help form alliances with public advocates
of introduced megafauna, who are often alienated by proj-
ects that treat these species as pests. Conﬂ icts between these
groups and conservation professionals erode trust and under-
mine conservation eﬃ cacy (Crowley et al. 2017), yet these
groups are natural allies in their concern for the welfare and
persistence of non-human life (Bruskotter et al. 2017). It is
likely that incorporating broader value systems towards these
organisms would oﬀ er a range of practical beneﬁ ts towards
conservation objectives and could strengthen the diversity
and inclusiveness of the conservation community.
e introduced megafauna of the world have restored
species richness across many continents to levels approach-
ing the Pleistocene, contribute fascinating and potentially
important ecological functions, and are an important refuge
for their functional group. We propose that further research
and dialogue on how introduced megafauna interact with
Figure 6. Wild donkeys Equus asinus increase surface water availability in the Sonoran Desert. (a) Wild donkey digging well to water table
( ‘ burro well ’ ), (b) troop of javelina Pecari tajacu bathing and drinking in burro wells, and (c) several-year-old Fremont’s cottonwood Popu-
lus fremontii growing in an abandoned burro well on a high channel bar.
Long, J. L. 2003. Introduced mammals of the world: their history,
distribution & inﬂ uence. – CSIRO publishing.
Malhi, Y. et al. 2016. Megafauna and ecosystem function from the
Pleistocene to the Anthropocene. – Proc. Natl Acad. Sci. USA
113: 838 – 846.
Marchetti, M. P. and Engstrom, T. 2015. e conservation
paradox of endangered and invasive species. – Conserv. Biol.
38: 434 – 437.
Meiri, M. et al. 2014. Faunal record identiﬁ es Bering isthmus
conditions as constraint to end-Pleistocene migration to the
New World. – Proc. R. Soc. B 281: 20132167.
Miller, G. H. et al. 2005. Ecosystem collapse in Pleistocene
Australia and a human role in megafaunal extinction. – Science
309: 287 – 290.
Naiman, R. J. and Rogers, K. H. 1997. Animals and system-
level characteristics in river corridors. – BioScience 47:
521 – 529.
Naundrup, P. J. and Svenning, J. C. 2015. A geographic assessment
of the global scope for rewilding with wild-living horses ( Equus
ferus ). – PLoS One 10: e0132359.
Palo, R. T. 1993. Usnic acid, a secondary metabolite of lichens and
its eﬀ ect on in vitro digestibility in reindeer. – Rangifer 13:
39 – 43.
Pastor, J. et al. 1988. Moose, microbes, and the boreal forest.
– BioScience 38: 770 – 777.
Ramey, E. M. et al. 2013. Desert-dwelling African elephants ( Loxo-
donta africana ) in Namibia dig wells to purify drinking water.
– Pachyderm 53: 66 – 72.
Ripple, W. J. and Van Valkenburgh, B. 2010. Linking top-down
forces to the Pleistocene megafaunal extinctions. – BioScience
60: 516 – 526.
Ripple, W. J. et al. 2015. Collapse of the world’s largest herbivores.
– Sci. Adv. 1: e1400103.
Ripple, W. J. et al. 2016. Saving the world’s terrestrial megafauna.
– BioScience 66: 807 – 812.
Ripple, W. J. et al. 2017. Conserving the World’s megafauna
and biodiversity: the ﬁ erce urgency of now. – Bioscience 67:
197 – 200.
Root-Bernstein, M. and Svenning, J.-C. 2016. Prospects for
rewilding with camelids. – J. Arid Environ. 130: 54 – 61.
Rule, S. et al. 2012. e aftermath of megafaunal extinction:
ecosystem transformation in Pleistocene Australia. – Science
335: 1483 – 1486.
Sandbrook, C. et al. 2011. Value plurality among conservation
professionals. – Conserv. Biol. 33: 285 – 374.
Sandom, C. et al. 2014. Global late Quaternary megafauna
extinctions linked to humans, not climate change. – Proc. R.
Soc. B 281: 20133254.
Schlaepfer, M. A. et al. 2011. e potential conservation value of
non-native species. – Conserv. Biol. 25: 428 – 437.
Schmidt, N. M. et al. 2015. Long-term patterns of muskox ( Ovibos
moschatus ) demographics in high arctic Greenland. – Polar
Biol. 38: 1667 – 1675 .
Smith, F. A. et al. 2015. Unraveling the consequences of the
terminal Pleistocene megafauna extinction on mammal
community assembly. – Ecography 39: 223 – 239.
Spencer, P. B. S. et al. 2012. Identiﬁ cation and management of a
single large population of wild dromedary camels. – J. Wildl.
Manage. 76: 1254 – 1263.
Svenning, J. C. et al. 2016. Science for a wilder Anthropocene:
synthesis and future directions for trophic rewilding research.
– Proc. Natl Acad. Sci. USA 113: 898 – 906.
te Beest, M. et al. 2016. Reindeer grazing increases summer albedo
by reducing shrub abundance in Arctic tundra. – Environ. Res.
Lett. 11: 125013.
Turner, J. W. Jr and Morrison, M. L. 2001. Inﬂ uence of predation
by mountain lions on numbers and survivorship of a feral horse
population. – Southwest. Nat. 46: 183 – 190.
Bradshaw, C. J. A. and White, W. W. 2006. Rapid development of
cleaning behavior by Torresian crows Corvus orru on non-
native banteng Bos javanicus in northern Australia. – J. Avian
Biol. 37: 409 – 411.
Bradshaw, C. J. A. et al. 2005. Conservation value of non-native
banteng in northern Australia. – Conserv. Biol. 20:
1306 – 1311.
Brummitt, R. K. 2001. World geographical scheme for recording
plant distributions, 2nd ed. – Hunt Inst. for Botanical
Bruskotter, J. T. et al. 2017. “Animal rights and wildlife conservation
conﬂ icting or compatible?” – e Wildlife Professional July/
Burke, M. G. 1990. Seawater consumption and water economy of
tropical feral goats. – Biotropica 22: 416 – 419.
Coates, K. P. and Schemnitz, S. D. 1994. Habitat use and behavior
of male mountain sheep in foraging associations with wild
horses. – Great Basin Nat. 54: 86 – 90.
Crowley, S. L. et al. 2017. Conﬂ ict in invasive species management.
– Front. Ecol. Environ. 15: 133 – 141.
Donlan, C. J. et al. 2006. Pleistocene rewilding: an optimistic
agenda for twenty-ﬁ rst century conservation. – Am. Nat. 168:
660 – 681.
Doughty, C. E. et al. 2016a. Interdependency of plants and animals
in controlling the sodium balance of ecosystems and the
impacts of global defaunation. – Ecography 39: 204 – 212.
Doughty, C. E. et al. 2016b. Global nutrient transport in a world
of giants. – Proc. Natl Acad. Sci. USA 113: 868 – 873.
Doughty, C. E. et al. 2016c. Megafauna extinction, tree species
range reduction, and carbon storage in Amazonian forests.
– Ecography 39: 194 – 203.
Feh, C. et al. 2002. Status and action plan for the Asiatic wild ass
( Equus hemionus ). – In: Moehlman, P. D. (ed.), Equids: zebras,
asses and horses. Publication Services Unit, Cambridge, UK,
pp. 62 – 71.
Gill, J. L. et al. 2009. Pleistocene megafaunal collapse, novel plant
communities, and enhanced ﬁ re regimes in North America.
– Science 326: 1100 – 1103.
Hamilton, W. J. III et al. 1977. Intersexual dominance and
diﬀ erential mortality of gemsbok Oryx gazella at Namib Desert
waterholes. – Madoqua 10: 5 – 19.
Hansen, D. M. et al. 2010. Ecological history and latent conserva-
tion potential: large and giant tortoises as a model for taxon
substitutions. – Ecography 33: 272 – 284.
Hernandez, L. et al. 1999. A note on the behavior of feral cattle
in the Chihuahuan Desert of Mexico. – Appl. Anim. Behav.
Sci. 63: 259 – 267.
Hundertmark, K. J. et al. 2002. Mitochondrial phylogeography of
moose ( Alces alces ): late Pleistocene divergence and population
expansion. – Mol. Phylogenet. Evol. 22: 375 – 387.
IUCN 2017. e 2017 IUCN Red List of threatened species.
– < www.iucnredlist.org > accessed April 2017.
Jones, K. E. et al. 2009. PanTHERIA: a species - level database of
life history, ecology, and geography of extant and recently
extinct mammals. – Ecology 90: 2648 – 2648.
Knapp, A. K. et al. 1999. e keystone role of bison in North
American tallgrass prairie: bison increase habitat heterogeneity
and alter a broad array of plant, community, and ecosystem
processes. – BioScience 49: 39 – 50.
Kodric-Brown, A. and Brown, J. H. 2007. Native ﬁ shes, exotic
mammals, and the conservation of desert springs. – Front.
Ecol. Environ. 5: 549 – 553.
Leslie, D. M. Jr 2008. Boselaphus tragocamelus (Artiodactyla:
Bovidae). – Mamm. Species 813: 1 – 16.
Lippold, S. et al. 2011. Whole mitochondrial genome sequenc-
ing of domestic horses reveals incorporation of extensive
wild horse diversity during domestication. – BMC Evol.
Biol. 11: 328.
Wallach, A. D. et al. 2015. Novel trophic cascades: apex
predators enable coexistence. – Trends Ecol. Evol. 30:
146 – 153.
Weinstock, J. E. et al. 2005. Evolution, systematics, and phyloge-
ography of Pleistocene horses in the New World: a molecular
perspective. – PLoS Biol. 3: e241.
Van Vuren, D. and Hedrick, P. W. 1989. Genetic conservation in
feral populations of livestock. – Conserv. Biol. 3: 312 – 317.
Vera, F. W. M. 2009. Large-scale nature development – the
Oostvaardersplassen. – Br. Wildl. 20: 28 – 36.
Wallach, A. D. et al. 2010. Predator control promotes invasive
dominated ecological states. – Ecol. Lett. 13: 1008 – 1018.
Supplementary material (Appendix ECOG-03430 at ⬍ www.
ecography.org/appendix/ecog-03430 ⬎ ). Appendix 1 – 4.