BookPDF Available

Reintroduction Biology: Integrating Science and Management


Abstract and Figures

This book aims to further advance the field of reintroduction biology beyond the considerable progress made since the formation of the IUCN/SSC Re-introduction Specialist Group. Using an issue-based framework that purposely avoids a structure based on case studies the book's central theme is advocating a strategic approach to reintroduction where all actions are guided by explicit theoretical frameworks based on clearly defined objectives. Issues covered include husbandry and intensive management, monitoring, and genetic and health management. Although taxonomically neutral there is a recognised dominance of bird and mammal studies that reflects the published research in this field. The structure and content are designed for use by people wanting to bridge the research-management gap, such as conservation managers wanting to expand their thinking about reintroduction-related decisions, or researchers who seek to make useful applied contributions to reintroduction.
Content may be subject to copyright.
Animal Translocations: What Are They
and Why Do We Do Them?
Philip J. Seddon1, W. Maartin Strauss2and John Innes3
1Department of Zoology, University of Otago, New Zealand
2Department of Environmental Sciences, UNISA, South Africa
3Landcare Research, Hamilton, New Zealand
‘Translocation is now well entrenched as a conservation tool, with the numbers
of animals being released in reintroduction and re-enforcement projects increasing
almost exponentially each year.’
Page 23
For as long as people have been moving from one place to another, which is as
long as humans have been ‘human’, animals and plants have been moved with
them, often hidden, unnoticed or ignored, but also as valued cargo. These
so-called ‘ethnotramps’ include economically and culturally favoured species
such as deer, macaque, civets, wallabies, cassowaries and wild-caught songbirds
that were commonly carried around with humans (Heinsohn, 2001).
The variety of animals shown to have been translocated by prehistoric human
colonists has been described as ‘astonishing’, with archaeological evidence of
numerous and widespread human-mediated introductions as far back as tens
of millennia, during the Pleistocene (Grayson, 2001). For example, it has
been shown that people moved wild animals from the New Guinea mainland
Reintroduction Biology: Integrating Science and Management. First Edition.
Edited by John G. Ewen, Doug P. Armstrong, Kevin A. Parker and Philip J. Seddon.
©2012 Blackwell Publishing Ltd. Published 2012 by Blackwell Publishing Ltd.
2Philip J. Seddon, W. Maartin Strauss and John Innes
to and between islands to the east and west over at least the past 20 000
years, for food and trade items as humans expanded their distribution and
sought to retain access to animals whose habits were already known to them
(White, 2004). It was during the Holocene (from 11 000 years before the
present), however, that the translocation of non-domesticated animals into
novel habitats became one of the most significant human impacts on native
animal populations (Kirch, 2005).
Clearly there are many reasons to translocate animals and some broad-scale
classifications have been proposed, for example to distinguish between con-
servation translocations and those for commercial or amenity values (Hodder
& Bullock, 1997), and along the way the terminology relating to transloca-
tions has become confused, contradictory and ambiguous. In this chapter
we provide a framework for classifying the different motivations for animal
translocation. We propose a simple decision tree that will enable conservation
managers to categorize easily the different types of translocation, from rein-
troductions to assisted colonizations, and standardize the terminology applied
in the species restoration literature. Throughout this chapter terms given in
italics are defined in Box 1.1.
Box 1.1 Glossary and definitions
Analogue species Closely related form that could be used as an
ecological replacement for an extinct species
(Parker et al., 2010)
Assisted colonization Translocation of species beyond their natural
range to protect them from human-induced
threats, such as climate change (Ricciardi &
Simberlof, 2009a)
Assisted migration Synonym for assisted colonization
Augmentation Synonym for re-enforcement
Benign introduction Synonym for conservation introduction
Biological control Intentional use of parasitoid, predator, pathogen,
antagonist or competitor to suppress a pest
population (Hoddle, 2004)
Classical biocontrol The introduction of exotic natural enemies to
control exotic pests (Thomas & Willis, 1998)
Animal Translocations 3
Conservation introduction An attempt to establish aspecies,forthe
purposes of conservation, outside its
recorded distribution but within an
appropriate habitat and ecogeographical
area (IUCN, 1998)
Ecological replacement Conservation introduction of the most suitable
extant form to fill the ecological niche left
vacant by the extinction of a species
(Seddon & Soorae, 1999)
Ecological restoration The process of assisting the recovery of an
ecosystem that has been degraded,
damaged or destroyed (SER, 2004)
Establishment Survival and successful breeding by founder
individuals and their offspring; this is a
prerequisite for, but not a guarantee of,
population persistence
Follow-up translocation Where one or more additional translocations
are conducted to supplement an initial
population established by reintroduction
(Armstrong & Ewen, 2001)
Introduction Intentional or accidental dispersal by a
human agency of a living organism outside
its historically known native range (IUCN,
Managed relocation Synonym for assisted colonization
Marooning Translocation to a predator-free offshore
Persistence The likelihood of population decline or
extinction over some appropriate
taxon-specific time frame
Re-enforcement Addition of individuals to an existing
population of conspecifics (IUCN, 1998)
Re-establishment Synonym for reintroduction that implies the
reintroduction has resulted in establishment
(IUCN, 1998)
4Philip J. Seddon, W. Maartin Strauss and John Innes
Rehabilitation The managed process whereby a displaced,
sick, injured or orphaned wild animal
regains the health and skills it requires to
function normally and live self-sufficiently
(IWRC, 2009)
Reintroduction Intentional movement of an organism into a
part of its native range from which it has
disappeared or become extirpated in historic
times (IUCN, 1987)
Reintroduction biology Research undertaken to improve the outcomes
of reintroductions and other translocations
(Armstrong & Seddon, 2008)
Relocation Synonym for translocation
Restocking Synonym for re-enforcement
Restoration ecology The science upon which the practice of
ecological restoration is based (SER, 2004)
Species restoration The application of any of a wide range of
management tools, including translocation,
that aim to improve the conservation status
of wild populations
Subspecific substitution Asubsetofecological replacement where the
replacement taxon is a subspecies (Seddon &
Soorae, 1999)
Supplementation Synonym for re-enforcement
Translocation Movement of living organisms from one area
with free release in another (IUCN, 1987)
Transplantation Synonym for translocation
The translocation spectrum
Seddon (2010) defined a conservation translocation spectrum, ranging from
reintroductions through to forms of conservation introduction.Figure1.1
broadens the scope and provides a framework for considering all motivations
for moving wild animals. The first, simple, bifurcation divides movements
into those that are accidental or incidental and those that are intentional
Is release intentional?
Is conservation of the target species the
primary objective?
Accidental translocations
Non-lethal control
Rehabilitation release
Biological control
Animal rights liberation
Yes No
Yes No Yes No
Are animals to be released
within the historical range?
Have conspecifics been extirpated
at the release site?
Reintroduction Re-enforcement Ecological replacement
Translocation spectrum
Assisted colonization Novel ecosystems
Is the aim to fill an available ecological niche left
vacant by the extinction of an original form?
Conservation translocation
Conservation introductionPopulation restoration
Figure 1.1 The translocation spectrum.
6Philip J. Seddon, W. Maartin Strauss and John Innes
(Figure 1.1). Strictly speaking, accidental movements of wild animals are not
translocations in the sense intended by the 1987 definition (IUCN, 1987).
This IUCN definition, however, lacks mention of intent contained within
a later, but confusing, redefinition (IUCN, 1998) that erroneously appears
synonymous with re-enforcement and wild-to-wild movements. We take it
that translocations are the deliberate and mediated (IUCN, 1998) movement
of organisms, from any source, captive or wild, from one area to free release
in another (IUCN, 1987). Thus translocation is the overarching term.
Not all translocations relate to the conservation of the species being moved.
The next division in our framework therefore asks the question: is conservation
of the target species the primary concern? (Figure 1.1). A split between conser-
vation and non-conservation is in some senses simplistic and naive. Multiple,
sometimes indirect, conservation benefits may accrue through translocations
for, for example, recreational, commercial or wildlife rehabilitation motiva-
tions, not the least being opportunities for increased public engagement with
nature and the enhanced public support for conservation measures that can
arise from this engagement. Nevertheless, it is useful to make the distinction
around primary concerns as many translocations may have multiple objectives
and not uncommonly enhancement of the conservation status of the species
may exist as a secondary goal.
Non-conservation translocations
There are at least seven types of translocation for which conservation is
not the primary aim (note that species conservation may be an associated
aim and protection of individual animals of endangered species may be a
primary aim): non-lethal management of problem animals, commercial and
recreational, biological control, aesthetic, religious, wildlife rehabilitation
and animal rights activism. One of the characteristics of many of the non-
conservation translocations is that they are introductions, with the sometime
exception of non-lethal management and wildlife rehabilitation.
Non-lethal management of problem animals
As urban, suburban and agricultural landscapes spread, and where nat-
ural populations of wildlife species recover or expand, the potential for
humanwildlife conflicts increases. In the United States, for instance, many
Animal Translocations 7
states have Nuisance Wildlife Control Operator (NWCO) programmes in
response to increased complaints of urban wildlife conflicts (O’Donnell &
DeNicola, 2006). The most common forms of conflict are predation of
livestock (Bradley et al., 2005), attacks on humans and their domestic pets
(Goodrich & Miquelle, 2005) and damage to property (Gammons et al., 2009;
Herr et al., 2008). In large part due to public attitudes, lethal management
of so-called problem wildlife is not a favoured option and instead a standard
method of dealing with problem individuals is to capture and translocate
them away from the focal point of conflict. The numbers of animals involved
can be significant; for example, in 1994 in Illinois alone NWCO permittees
moved >18 000 animals, including >13 000 raccoons (Procyon lotor), squirrels
(Sciurus spp.) and bats, and throughout the US some hundreds of thousands
of animals are moved to mitigate conflicts (Craven et al., 1998). By far the
majority of problem animals are mammals, with the larger species most com-
monly carnivores including ursids (Ursus spp.), felids (Panthera spp., Felis
concolor, Lynx lynx), wolves (Canis lupus), and mustelids (Mustela spp.), but
translocation of raptors, including the golden eagle (Aquilia chrysaetos), black
eagle (A. verreauxii), crowned eagle (Stephanoetus coronatus)andmartialeagle
(Polemaetus bellicosus), has taken place in the United States and South Africa
to reduce livestock predation (reviewed in Linnell et al., 1997). Translocation
of hen harriers (Circus cyaneus) has been considered to mitigate the impact
of harrier predation on red grouse (Lagopus lagopus scotica) on UK moorland
managed for grouse shooting (Watson & Thirgood, 2001). The priorities for
problem animal translocations are primarily solving the specific conflict and
secondarily the welfare of the individual animal. It is a public perception that
translocated problem animals ‘live happily ever after’ (Craven et al., 1998).
In the few instances where post-release monitoring of such translocations has
taken place, the reality is very different. Translocated problem animals typically
show high post-release mortality due to the stress of capture, transport and
release, aggression by territorial conspecifics, poaching, disorientation, unsuit-
able habitat, long-distance dispersal and disease (Craven et al., 1998; Fischer &
Lindenmayer, 2000). Carnivores in particular show strong homing behaviour
(Bradley et al., 2005), with many individuals able to return to the capture site
and resume nuisance behaviour from release sites between 150 km (Lenain &
Warrington, 2001) and up to nearly 500 km (Linnell et al., 1997) away. Even
when translocated animals do not return to the focal point of conflict, they tend
to settle away from the release site and may become a new problem elsewhere
(van Vuren et al., 1997; Le Gouar et al., this volume, Chapter 5). Consequently,
8Philip J. Seddon, W. Maartin Strauss and John Innes
translocation of nuisance animals is often ineffective at achieving either of the
two objectives. This, in conjunction with challenges to the assumption that
problems are caused by a few problem animals (Linnell et al., 1999), have led
to increased efforts to manage potential conflict proactively using such things
as habitat modification, exclusions and repellents. In South Africa cheetah
(Acinonyx jubatus) are protected under legislation and consequently, instead
of trapping and shooting animals that prey on livestock, there is a programme
of translocation to fenced reserves, possibly for future ecotourism, although
post-release survival rates can be low, especially in the presence of competing
predators at release sites (Marnewick et al., 2009). There are also attempts to
shift public attitudes to accept lethal control by moving away from a focus
on individual animal welfare and towards an appreciation of the need for
population level conservation (Linnell et al., 1997).
Commercial and recreational
Commercially and recreationally motivated translocations are often indis-
tinguishable, with introductions to support recreational hunting or derived
from the pet trade having a significant commercial element. One of the most
significant threats to the conservation of freshwater fishes is the introduction
of alien invasive sport fish, with three of the hundred ‘World’s Worst Invasive
Alien Species’ being fish introduced solely for sport (Cambray, 2003). One of
the earliest examples of fish translocation for stocking a lake was in Turkey
by Murat the III, Sultan of Ottoman between 1546 and 1595, and since the
1950s in Turkey a total of 25 exotic fish species have been introduced and 14
native fish species translocated to habitats outside their natural range (Innal
& Erk’akan, 2006). There is a globalization of alien fish for sport; for example
rainbow trout (Onchorhynchus mykiss) are now established in 82 countries
(Cambray, 2003). Mammals too are translocated for sport across national
and international boundaries. For example, there is evidence of the illegal
translocation of invasive feral pigs (Sus scrofa) by recreational hunters in
Australia to supplement existing populations and to create new populations
(Spencer & Hampton, 2005). Red foxes (Vulpes vulpes), grey foxes (Urocyon
cinereoargenteus)andcoyotes(Canis latrans) are illegally translocated to stock
hunting enclosures or ‘fox-pens’ (Davidson et al., 1992), and in one year pri-
vate hunting clubs in Kentucky, USA, translocated >2 300 raccoons (Nettles
et al., 1979). Ungulate introductions have taken place in well over 50 countries,
with the USA and South Africa having the highest numbers of introduced
Animal Translocations 9
ungulates globally (Spear & Chown, 2009a). To date, however, with only
limited evidence of negative impacts, efforts to restrict ungulate introductions
may be constrained in the face of the economic gains through trophy hunting
and ecotourism (Spear & Chown, 2009b).
The pet trade per se does not fit the definition of translocation because the
end point is not intended to be the ‘free release’ of animals, but it is worth
acknowledging that accidental releases and intentional releases by owners
can be an endpoint. It has been estimated that in the United States 38 % of
exotic bird species are established or establishing from pets that have escaped
(Temple, 1992), and two of three invasive bird species in Hong Kong originate
from the pet trade (Shieh et al., 2006). Since 1994, 290 exotic species of pet
birds have been imported into Taiwan, and of these 93 species have escaped
and at least 28 of these breed in the wild (Shieh et al., 2006).
Biological control
Biological control, or biocontrol, uses living organisms as pest control agents.
Classical or traditional biological control involves the introduction of exotic
natural enemies to control exotic pests under the simple premise that an exotic
organism becomes a pest partly because it has been released from population
regulation by natural enemies (Hoddle, 2004). As a form of introduction,
therefore, all the risks and uncertainties of potential invasive species apply
(Simberloff & Stiling, 1996). Opponents of biological control point to examples
where the introduced agent has had an impact on non-target native species,
through competition and predation, and other more complex interactions
(Simberloff & Stiling, 1996). Proponents of biological control point out that
the greatest problem arose in early programmes involving the introduction of
generalist vertebrate predators, such as mosquito fish (Gambusia affinis), now
in about 70 countries worldwide, cane toads (Bufo marinus) to target various
invertebrate crops pests in Australia, and red foxes and mustelids to control
rabbits (Oryctolagus cuniculus) in Australia and New Zealand, and that since
then increased regulation has reduced risks (Thomas & Willis, 1998; Hoddle,
2004). Increasingly, biological control is being considered for conservation as
well as for agriculture (Henneman & Memmott, 2001) and there are examples
of proposed development of biological control, such as genetically engineered
viral and bacterial diseases, to target invasive arthropods that threaten native
flora and fauna (Hoddle, 2004).
10 Philip J. Seddon, W. Maartin Strauss and John Innes
Although less common now, the introduction of exotic species, mostly birds,
by Western colonists in the 18th and 19th centuries was a major form of species
translocation. Settlers from Europe attempted to create a huntable resource
of familiar species, or sought to create new populations of exotic songbirds
and other species that were well known to them for purely aesthetic reasons
(Duncan et al., 2003). Around 70 % of bird introductions (953 events) have
been to islands, and over half to Pacific Islands (271 events) and Australasian
regions (216 events) (Blackburn & Duncan, 2001). In New Zealand as many
as 137 exotic bird species were introduced before 1907, with 284 releases
of 43 species occurring mainly between 1861 and 1885 (Duncan, 1997); 28
(20%) of these established populations persist to the present day (Veltman
et al., 1996). Translocation of animals to support both consumptive and non-
consumptive nature-based tourism addresses several motivations, including
aesthetic, recreational, education and advocacy, and commercial, as well as
conservation (Cousins et al., 2008; Mbaiwa, 2008; see Box 1.2).
Box 1.2 Game ranching in Southern Africa: commercial or
conservation translocations?
Traditionally, the conservation of species, ecosystems and their under-
lying functions has been accomplished by setting land aside for
conservation purposes the so-called Yellowstone paradigm. In sub-
Saharan Africa, as elsewhere, the main shortcoming of the Yellowstone
paradigm was the fact that it ignored the vast majority of land that falls
outside National Parks (Child, 2000). The generic term ‘game ranch-
ing’, which here includes private game reserves, embodies what has
been referred to as a ‘second conservation paradigm’, which is based on
sustainable use, wildlife ownership and pricing (Child, 1996, in Child,
2000). Game ranching, as practised in South Africa (see Figure 1.2), Zim-
babwe and Namibia, is defined as the managed, extensive production of
free-living animals on large tracts of land that are fenced or unfenced
for purposes of live animal sales, hunting, trophy hunting, venison pro-
duction, tourism or other uses (Bothma, 2002). Officially proclaimed
conservation areas in South Africa total 5.8 % of the country’s surface
Animal Translocations 11
area, with private game ranches estimated to contribute an additional
13 % (ABSA 2003, in Carruthers, 2009).
Figure 1.2 A southern reedbuck from the Drakensberg region of Kwazulu/
Natal, South Africa. This species is commonly used in game ranching. (Photo:
W. Maartin Strauss).
The pivotal role that early ranchers in South Africa played in the con-
servation of ungulates such as the bontebok Damaliscus dorcas dorcas,the
black wildebeest Connochaetes gnou and the Cape mountain zebra Equus
zebra zebra is well documented. Although the game ranching industry
has expanded significantly in recent decades, the conservation role that
it plays today is not clear-cut. In an attempt to increase local diversity
and thereby economic viability of game ranches, managers now fre-
quently apply artificial selection in breeding indigenous ungulates. The
so-called white and black springboks (Antidorcas marsupialis)are,for
example, popular animals for translocation and at game auctions their
monetary value has, on average, increased by 38.4 % and 20.3 % per year
over a 10 year period (van der Merwe et al., 2008). Nevertheless, there
is an assumption that game ranching contributes more to biodiversity
12 Philip J. Seddon, W. Maartin Strauss and John Innes
conservation than other forms of agricultural land use (Bond et al.,
2004), despite profit being the primary reason for their establishment
(Hearne & McKenzie, 2000). There are no official figures available,
but du Toit (2007) estimated that up to 70 000 animals are captured
and translocated annually in South Africa, resulting in an estimated
turnover, including the value of the captured animals, of between R750
million (US$101 million) and R900 million (US$121 million). Prior
to the enactment of recent (2004) laws aimed at regulating and con-
trolling the translocation and introduction of large mammals in South
Africa, exotic and/or extralimital ungulate species were translocated and
introduced into game ranching areas across the country, resulting in
South Africa having the second highest number of introduced ungulates
globally (Spear & Chown, 2009b).
Game ranching has undoubtedly contributed to the increase in the
number of wild animals in South Africa during the last few decades.
Conservation in South Africa is, however, at a crossroads, as the game
ranching industry pushes for exemption from all nature conservation
regulatory control.
Approximately 30 % of people of all religions in East Asia believe they can
accrue merit by freeing captive animals during ceremonies termed ‘prayer
animal releases’. These ceremonies are organized by temples using local and
exotic animals, mostly birds, supplied by pet stores (Severinghaus & Chi,
1999). With the practice prevalent in Taiwan, Malaysia, Thailand, Cambodia,
Vietnam, Honk Kong and Korea, and with some temples organizing as many
as 24 release ceremonies per year, the scale of translocations is huge. It was
estimated, for example, that 128 000 birds were released in only one year in
Taichen City, Taipei (Severinghaus & Chi, 1999).
Wildlife rehabilitation
Capture, care and release of wildlife is a significant and growing practice
internationally; for example in Britain some 30 000 to 40 000 wild animal
Animal Translocations 13
casualties end up in wildlife hospitals, the most common species being the
European hedgehog (Erinaceus europaeus)(Molonyet al., 2006). The most
frequent causes of injuries to terrestrial species include collisions with vehicles
and domestic animal attacks (Hartup, 1996). For seabirds oiling has been a
major threat ever since the start of large-scale transportation of petroleum
products by sea, with large-scale seabird mortality due to dumping of tanker
waste oil from 1917 and oil spills in 1937 off San Francisco (Carter, 2003)
and during WWII (Mezat et al., 2002). Following significant oil spills in the
late 1960s, rehabilitation efforts for oiled seabirds developed in the United
States (Carter, 2003) and South Africa (Nel et al., 2003). Rehabilitation of
oiled seabirds involves capture, transport, cleaning and release, and has been
characterized by high failure rates, with in some cases only 120 % of birds
surviving the first year post-release (Mead, 1997). Post-release survival rates
vary, however, with the type and degree of oiling, but also with the species. The
highest success rates have been achieved with African penguins (Spheniscus
demersus), with up to 84 % of processed penguins being released (Nel et al.,
2003) and up to 65 % of released penguins being resighted within two years
(Underhill et al., 1999). Nearly comparable survival rates have been achieved
for little penguins (Eudyptula minor)(Goldsworthyet al., 2000).
Although conservation of endangered species is one of the reasons cited for
the rehabilitation of marine mammals (Moore et al., 2007), the greatest risks
involved in the release of rehabilitated animals is that of disease transmission
from captivity to wild populations (Quakenbush et al., 2009). There is general
agreement by authorities that the health of wild populations should be a
greater concern than the welfare of an individual animal, thus euthanasia is
often the best option, but one that carries a significant negative image and
risks loss of public support for rehabilitation efforts (Moore et al., 2007).
Consequently, there may be public pressure to sustain rehabilitation efforts
even for species that have healthy populations for which the rehabilitation
of a single animal has no conservation value (Moore et al., 2007) but which
poses significant risks to wild populations (Quakenbush et al., 2009). In
contrast, the numbers of individual birds involved in major oil spills can be
significant at a population conservation level; for example oil spills off the
South African coast from the MV Treasure in 2000 (Parsons & Underhill,
2005) and the Apollo Sea in 1994 (Underhill et al., 1999) resulted in the
processing of 10 000 and nearly 20 000 oiled African penguins, respectively,
from a world total population around that time of <60 000 pairs (Birdlife
International, 2008).
14 Philip J. Seddon, W. Maartin Strauss and John Innes
Animal rights activism and animal liberations
While conservation biologists are rightly concerned with animal welfare and
the reduction of unnecessary suffering, there is a difference between an indi-
vidual animal welfare perspective that may motivate activities such as animal
rehabilitation and the conservation management of wildlife populations. For
the most part these differences do not create problems; for example welfare
concerns will dictate that any release of captive animals must ensure that each
animal has the skills and behaviours necessary for survival in the wild (Waples
& Stagoll, 1997). Different perspectives become problematic, however, where
animal welfare becomes animal rights. Animal rights activists are committed,
inter alia, to the total abolition of use of animals in science, commercial agri-
culture and sport hunting, and consider as fundamentally wrong any system
that views non-human animals as resources to be used by humans (Regan,
1983). There exists a challenging incompatibility between a conservation ethic
and animal rights, which some see as a ‘highly reductionist view’ that focuses
exclusively on individual sentient animals (Hutchins, 2008). This can lead
to illegal liberations of captive animals that effectively expand the range of
introduced species and have detrimental impacts on native fauna (Lewis
et al., 1999). Furthermore, the released animals can suffer. The liberation of
captive-bred furbearers such as mink (Mustela vison) from fur farms provides
one example. Accounts in the media indicate multiple liberations of groups of
up to 6000 captive mink from farms in Canada, USA, UK, Ireland, Finland,
the Netherlands and Greece over the last decade. Many of the liberations of
mink are attributed or claimed to be the actions of the Animal Liberation
Front (ALF). Inevitably, released mink start to die in large numbers soon after
release, before survivors can be recovered. Nevertheless, liberationists claim
that outside their cages mink have a fighting chance of survival. This is despite
overwhelming evidence that most freed mink face a slow death in the wild
versus a humane end in captivity.
Conservation translocations
Where conservation of the target species is the primary objective we can
consider conservation translocations (Hodder & Bullock, 1997) and ask a new
question: ‘are animals to be released within the historical distribution range of
the species?’ Releases within the documented natural range may be classified
Animal Translocations 15
as translocations for population restoration, the implication being that the goal
is to recover populations of the species back to some past target state. Releases
outside the historical range, but with population conservation as the primary
objective, are termed conservation introductions (IUCN, 1998) (Box 1.1).
Population restorations
Reintroduction is the release of an organism into an area that was once part of its
range but from which it has been extirpated (IUCN, 1987) (Box 1.1). In broadly
stated terms the objective of a reintroduction is to re-establish a self-sustaining
population of a species within its historic range (Griffith et al., 1989), and
ideally that population will have a high probability of persistence with minimal
or no intervention (Seddon, 1999). Despite some early reintroduction success
stories, such as Arabian oryx (Oryx leucoryx) in Oman (Stanley Price, 1989)
and peregrine falcon (Falco peregrinus) in North America (Cade & Burnham,
2003), the failure of other, less well-conceived, reintroduction attempts meant
that reintroduction project success rates were low (Griffith et al., 1989; Wolf
et al., 1996). The situation was not helped by a lack of post-release monitoring,
which meant that the timing and causes of failures was not known (Seddon
et al., 2007a). In response to the perceived problems the World Conservation
Union (IUCN) Reintroduction Specialist Group (RSG) was formed in 1988
under the auspices of the Species Survival Commission (Stanley Price & Soorae,
2003). One of the first actions of the RSG was the formulation of Guidelines for
Reintroductions (IUCN, 1998) in order to improve reintroduction practice;
for example the guidelines place emphasis on the identification of release sites
within the historic range of the species and acknowledge a need to ensure that
previous causes of decline have been addressed, both factors having been shown
to strongly influence project outcomes (Fischer & Lindenmayer, 2000). In part
due to the actions and outputs of the RSG, improved pre-release planning, care
over the selection of founders and the composition of founder groups, release
site preparation and detailed post-release monitoring have improved project
success rates, at least in the short term (Soorae, 2008). Although assessment
of reintroduction success is not straightforward it is useful to think of any
project needing to progress through two phases (Armstrong & Seddon, 2008):
population establishment, which requires survival of founders, and breeding
by founders and their offspring; and population persistence, which may be
assessed for taxonomically relevant time frames using population modelling
tools (Seddon, 1999).
16 Philip J. Seddon, W. Maartin Strauss and John Innes
Meta-analyses of factors contributing to reintroduction success indicate
individuals released (Germano & Bishop, 2009; Griffith et al., 1989; Wolf
et al., 1996, 1998). The effects of habitat quality on reintroduction success
have been confirmed in recent experimental studies (Moorhouse et al., 2009).
The number of animals released, however, is often correlated with several
other factors that may be important prerequisites of success. For example,
projects that release the most individuals are usually those that are well
funded and well resourced, and that are perceived apriorito have the
greatest chance of success. In contrast, only few founders are released in
short-term projects that do not have significant institutional and community
Re-enforcement (IUCN, 1998), also termed restocking (IUCN, 1987) and
supplementation (IUCN, 1998), or augmentation (Maguire & Servheen, 1992),
involves the release of individuals into an existing population of conspecifics
(Box 1.1), in order to increase population size and reduce the risks of genetic
or demographic collapse due to stochastic effects. Translocations for re-
enforcement are used to overcome barriers to natural dispersal from other
free-ranging populations (e.g. Gusset et al., 2009), to speed up population
growth, or to enhance genetic diversity and avoid inbreeding depression
(Jamieson et al., 2006). In some cases ongoing re-enforcement may be required
to sustain non-viable free-ranging populations until natural productivity is
sufficient to support population growth and persistence. For example, kaki or
black stilt (Himantopus novaezelandiae) are sustained in the wild through the
release of captive-reared birds while habitat restoration measures are being
trialled (e.g. Keedwell et al., 2002).
Seddon (2010) poses the question of when does a reintroduction become
re-enforcement? While seemingly trivial semantics, this question does relate
to a more significant one – that of when to stop releases. There is a substantial
body of literature that discusses evaluation of reintroduction success (Fischer
& Lindenmayer, 2000; Griffith et al., 1989; Seddon, 1999; Wolf et al., 1996,
1998), and there is now widespread use of population modelling to set
re-establishment goals, to define optimal reintroduction strategies and to
assess population persistence (Armstrong et al., 2002, 2006; Rout et al., 2009;
Schaub et al., 2009). Pre-release target setting considers the number, size and
composition of founder cohorts and the efficacy of single versus multiple
releases. Nevertheless, post-release monitoring will enable refinement of
pre-release models (Armstrong & Davidson, 2006; Armstrong et al., 2007;
Animal Translocations 17
Wakamiya & Roy, 2009) and may indicate a low probability of population
persistence that could be addressed through the release of more individuals.
Such post hoc secondary releases have been termed follow-up translocations
(Armstrong & Ewen, 2001) and could be seen as supplementation of the
re-established free-ranging population, but should strictly be considered part
of the original, but not yet successful, reintroduction attempt (Seddon, 2010).
Conservation introductions
Mediated movement of organisms outside their native range constitutes a
species introduction (IUCN, 1987) and if the goal is the establishment of a new
population explicitly and primarily for conservation, then such a translocation
is regarded as a conservation,orbenign (in intent at least), introduction (IUCN,
1998). The current IUCN guidelines consider conservation introductions to be
justified ‘when there is no remaining area left within a species’ historic range’
(IUCN, 1998). This limited rationale marks conservation introductions as a
somewhat reactive, stop-gap measure, in some cases perhaps to mark time until
appropriate habitat restoration can take place within the historical distribution
range of the target species. However, more pro-active interventions are now
being considered by natural resource managers, and we can broadly define
two types of conservation introduction: ecological replacement and assisted
Ecological replacement is the release of species outside their historic range in
order to fill an ecological niche left vacant by the extinction of a native species.
Extinction removes the option of reintroduction through the release of either
wild or captive individuals and may mean the loss of critical or otherwise
desirable ecological functions. One option is therefore to restore lost ecological
function through the establishment of a viable population of an ecologically
similar species (Atkinson, 2001). The most readily acceptable approach will be
the release of a subspecific substitute, such as using the North African subspecies
of ostrich (Struthio camelus camelus) as a replacement for the extinct Arabian
subspecies S. c. syriacus (Seddon & Soorae, 1999). The recent use of other
analogue species includes yellow-crowned night heron (Nycticorax violacea)for
an extinct endemic Nyctocorax species in Bermuda, tundra musk ox (Ovibos
moschatus)fortheextinctO. palantis in Sibera (review in Parker et al., 2010)
and North Island kokako (Callaeas wilsoni) for the extinct South Island form
C. cinerea (see Box 1.3). While subspecific substitutes may be expected to be
the most appropriate ecological replacements, other forms may potentially be
18 Philip J. Seddon, W. Maartin Strauss and John Innes
better functional equivalents. For example, Parker et al., (2010) make a case for
the replacement of the extinct New Zealand quail (Coturnix novaezelandiae),
not with its closest extant relative, the Australian stubble quail C. pectoralis,but
with the more distantly related but ecologically better suited Australian brown
quail C. ypsilophora. It may not be the case that the analogue species is rare or
threatened in its natural range and thus its conservation may not be a primary
objective of its introduction as an ecological replacement, necessitating a
broader interpretation of the earlier dichotomy between conservation and
non-conservation translocations to include the conservation objective beyond
the target species (Figure 1.1).
Box 1.3 North Island kokako translocation to the South Island as an
example of an ecological replacement
In October 2008, 10 North Island kokako Callaeas wilsoni were translo-
cated from Mapara in the central North Island of New Zealand to
8 140 ha Secretary Island on the southwestern corner of the South
Island, 1 000km south (see Figure 1.3). South Island forests were previ-
ously occupied by a southern kokako species Callaeas cinerea,declared
extinct in 2004. The intention of the release was to restore the ecolog-
ical functions of kokako into a South Island forested ecosystem. It is
inherently experimental.
IUCN guidelines (1995) and Seddon & Soorae (1999) suggested that
an ecological substitute should be selected from extant subspecies or
races (rather than species) to avoid fundamental differences in habitat
preferences between the original and substitute taxa. However, North
and South Island kokako were regarded as subspecies until recently
(Holdaway et al., 2001). Plumage of the two is the same although South
Island birds had orange wattles (small fleshy appendages arising from
the gape and lying against the throat) while North Island wattles are
blue (Higgins et al., 2006). Subtly different behaviours, such as more
ground-feeding, may have led to the early decline and extinction of the
South Island form due to predation by introduced pest mammals, but
this is unclear (Clout & Hay, 1981; Holdaway & Worthy, 1997). The
extinction of C. cinerea prevents further comparison of behaviours of
the two taxa.
Animal Translocations 19
Figure 1.3 A North Island kokako nestling being banded for monitoring
purposes on Tiritiri Matangi Island, New Zealand. (Photo: John G. Ewen).
Diet and behaviour of C. wilsoni are very well known (Higgins
et al., 2006), whereas C. cinerea was never studied in detail. C. wilsoni
mainly eat leaves and fruits, and some insects (Higgins et al., 2006).
Kokako were very abundant in both islands before human settlement,
and their ecological roles included herbivory, pollination and fruit
dispersal, as well as being prey for New Zealand’s original predators
(raptors), some of which are also extinct. The demise of native birds
has in turn impaired pollination and perhaps seed dispersal of trees
and shrubs (Kelly et al., 2010). Kokako are quite large (38 cm; 230 g)
with 13 mm gapes, capable of dispersing fruits of several structurally
important large-seeded plants (Clout & Hay, 1989; Kelly et al., 2010).
Reintroducing kokako to Secretary Island is a small part of the biotic
restoration planned there. While this translocation meets most objectives
of a reintroduction listed by IUCN (1995) – to enhance the long-term
20 Philip J. Seddon, W. Maartin Strauss and John Innes
survival of a species; to re-establish a keystone species (in the ecological
or cultural sense) in an ecosystem; to maintain and/or restore natural
biodiversity; to provide long-term economic benefits to the local and/or
national economy; to promote conservation awareness – it is primarily
an attempt at ecological restoration of a lost biotic community, as first
championed in New Zealand by Atkinson (1988).
Monitoring survival of some of the kokako released in October
2008 with transmitters revealed only one death, due to falcon (Falco
novaeseelandiae) predation. The monitored birds settled in the general
area of the island where they were released. A further 17 kokako released
in 2009 were sourced from two additional source populations – Kaharoa
and Rotoehu – to increase the genetic representation of the North Island
species and minimize future inbreeding.
Assisted colonization, also referred to as assisted migration (McLachlan
et al., 2007) and managed relocation (Richardson et al., 2009) has been best
defined as ‘translocation of a species to favourable habitat beyond their native
range to protect them from human induced threats’ (Ricciardi & Simberlof,
2009a). Recent interest in this form of conservation introduction has been
driven by predicted habitat change due to rapid climate change (Hoegh-
Guldberg et al., 2008), but assisted colonization could be and has been used to
mitigate a variety of threats, including agricultural expansion and urbanization
(Ricketts & Imhoff, 2003), filling of hydroelectric reservoirs (Richard-Hansen
et al., 2000) and the threats posed by (other) deliberately introduced species
(Vitousek et al., 1997). Specific examples include the translocation of slow-
worm (Anguis fragilis) from sites for future housing development in the UK
(Platenberg & Griffiths, 1999) and giant land snails (Powelliphanta augusta)
from sites designated for coal mining in North Westland, New Zealand
(Walker et al., 2008), both of which involved releases outside the species’
known historical range.
The debate around assisted colonization has focused on the risk of impacts of
introduced species (Mueller & Hellmann, 2008; Ricciardi & Simberlof, 2009a,
2009b; Sax et al., 2009; Seddon et al., 2009; Vitt et al., 2009) and is assumed
by many commentators to mark a major shift in conservation translocation
Animal Translocations 21
practice. However, assisted colonization is a well-established (if previously
unnamed) conservation tool in some parts of the world. For example, in New
Zealand the extinction threats to endemic birds, herptiles and invertebrates
posed by introduced mammalian predators have been addressed with some
success through ‘marooning’, whereby species are translocated to predator-
free offshore islands (Saunders & Norton, 2001). In many cases these islands
were not within historically documented parts of the species range. In fact
one of the earliest examples of assisted colonization could be the pioneering
conservation work undertaken by Richard Henry in New Zealand (Jones &
Merton, this volume, Chapter 2).
During the 1890s Richard Henry was caretaker of Resolution Island in
remote and rugged Fiordland on the west coast of New Zealand’s South
Island. A keen naturalist, he noted with dismay the impact on native birds of
the arrival of recently introduced stoats (Mustela erminea)astheyinvadedthis
last corner of New Zealand. In a desperate attempt to protect populations of the
flightless kakapo (Strigops habroptilus) and little spotted kiwi (Apteryx oweni)
between 1894 and 1900 he translocated hundreds of individuals from the
mainland on to Resolution Island (Saunders & Norton, 2001). Unfortunately
Resolution was too close to the mainland and stoats invaded in 1900 and
Henry’s efforts were in vain (Clout, 2006). Nevertheless, the technique of
marooning vulnerable species on predator-free islands that may or may not
have been occupied by the species in the past became a vital tool to avert
extinctions in the face of predation by introduced mammalian predators in
New Zealand (Saunders & Norton, 2001).
We can therefore envisage a simple dichotomy between ecological replace-
ment and assisted colonization (Figure 1.1). If the aim is to fill an available
ecological niche left vacant by the extinction of the original form, then the
type of conservation translocation is an ecological replacement. In the case of
marooning and other translocations to move members of a species outside
their range to avoid some threat, the primary aim is not to fill an available
ecological niche but rather to sustain a viable population, perhaps until appro-
priate habitat restoration has taken place within their core distribution. In
the specific case of climate change mitigation, where a suitable habitat has
opened up outside the historic distribution range, then assisted colonization
may be used to fill a newly available niche that has not been created by an
extinction and is thus similarly distinguished from the release of ecological
substitutes for an extinct form (Figure 1.1). In all cases assisted colonization
22 Philip J. Seddon, W. Maartin Strauss and John Innes
is the human-mediated movement of individuals of a species that would
otherwise be unable to survive current or anticipated threats within its current
Human dimensions in animal translocations
It is now well understood that intensive conservation interventions, such as
reintroductions, cannot hope to succeed without some level of engagement
with local and national government, non-governmental agencies and profes-
sionals, and, critically, the public (Kleiman, 1989). This is seen nowhere more
strongly than in the restoration of populations of large carnivores, where
local community attitudes and behaviours can be significant determinants
of project outcomes (Hayward et al., 2007; Hunter et al., 2007; Jule et al.,
2008; Lohr et al., 1996; Meadow et al., 2005; Nilsen et al., 2007; Zahniser &
Singh, 2004).
Public engagement with a translocation project can come at different levels
and stages, including the provision of funding (and conversely through eco-
nomic benefits that accrue to local communities (Kleiman, 1989) and labour,
support for approvals, advocacy and lobbying for political and legislative
change, and wider attitudinal changes through education and interaction
(Williams et al., 2002). In many cases community engagement is not just a
useful part of translocation planning, nor even just a prerequisite for success;
rather it is one of the desirable outcomes.
Translocation projects provide a means to engage with the public, to make
them collaborators in the programme (Kleiman, 1989) and to potentially
change negative or biased public views of scientists and resource managers.
The opportunity to learn about conservation projects, to see wild animals
up close and even to participate in their liberation into new areas provides
a powerful means to counter the view that conservation is preservation and
entails the locking up of resources. Meaningful and positive contact with native
species and natural areas can mitigate the alienation from the natural world
and the extinction of experience that is a feature of increasingly urbanized
societies. It has been suggested that advocacy for the natural world may be
the main role of conservation biology (Brussard & Tull, 2007). Parker (2008)
proposes that we view translocations as a vehicle to enhance linkages between
scientists, managers and the general public, and that meaningful community
participation should be considered one of the main outputs of a translocation,
along with its management and scientific objectives.
Animal Translocations 23
Concluding comments
Translocation is now well entrenched as a conservation tool, with the num-
bers of animals being released in reintroduction and re-enforcement projects
increasing almost exponentially each year. The more conservation transloca-
tion activity there is, the greater the proliferation of terms and concepts being
used to describe the various actions, resulting in a mass of synonyms and
variations that may end up obscuring meaning. What we have tried to do in
this chapter is to provide a standardized and justified terminology to describe
the full spectrum of conservation translocation activities. Our hope is that
these will enable practitioners and researchers to be very clear about what they
are doing or what they propose to do, not only for themselves, but also in
discussions or debate with others.
Resource managers have been forced to deal with critical population declines
and extinctions, ecosystem degradation and changing habitat conditions due
to global climate change (Sekercioglu et al., 2008); consequently new forms
of conservation interventions are being explored. These include ecological
replacements, whereby functionally equivalent taxa fill a niche left available
by the extinction of the original form, and assisted colonization, where species
are moved into areas not previously occupied in order to avoid some human-
induced threat. These conservation introductions are perhaps not as radical
as they first appear, but they do mark the start down an interesting pathway
leading away from strict reintroductions towards more controversial decisions
about which species we want to have where. There is a clear convergence in
thinking between the disciplines of reintroduction biology and restoration
ecology (Lipsey & Child, 2007; Seddon et al., 2007b), whereby historical
restoration targets are seen as arbitrary and unrealistic, and increasingly there
is talk of futuristic restoration (Choi, 2004) and novel (Figure 1.1) or designer
ecosystems (Temperton, 2007; Seddon, 2010). The debates will no doubt rage
for decades as we consider what our future natural world could or should look
This chapter was improved by the comments of Doug Armstrong, Tim
Blackburn, John Ewen, Ian Flux, Richard Maloney, Ollie Overdyck, Kevin
Parker, Francois Sarrazin, Yolanda van Heezik and Megan Willans.
24 Philip J. Seddon, W. Maartin Strauss and John Innes
Armstrong, D.P. & Davidson, R.S. (2006) Developing population models for guiding
reintroductions of extirpated bird species back to the New Zealand mainland.
New Zealand Journal of Ecology, 30, 73 85.
Armstrong, D.P. & Ewen, J.G. (2001) Assessing the value of follow-up translocations:
a case study using New Zealand robins. Biological Conservation, 101, 239 247.
Armstrong, D.P. & Seddon, P.J. (2008) Directions in reintroduction biology. Trends
in Ecology and Evolution, 23, 20 25.
Armstrong, D.P., Castro, I. & Griffiths, R. (2007) Using adaptive management
to determine requirements of reintroduced populations: the case of the New
Zealand hihi. Journal of Applied Ecology, 44, 953 962.
Armstrong, D.P., Davidson, R.S., Dimond, W.J. et al. (2002) Population dynamics of
reintroduced forest birds on New Zealand islands. Journal of Biogeography, 29,
Armstrong, D.P., Raeburn, E.H., Lewis, R.M. et al. (2006) Estimating the viability of
a reintroduced New Zealand robin population as a function of predator control.
Journal of Wildlife Management, 70, 1020 1027.
Atkinson, I.A.E. (1988) Presidential address: opportunities for ecological restoration.
New Zealand Journal of Ecology, 11, 1 12.
Atkinson, I.A.E. (2001) Introduced mammals and models for restoration. Biological
Conservation, 99, 8196.
BirdLife International (2008) Species factsheet: Spheniscus demersus. Downloaded
from on 9/9/2009.
Blackburn, T.M. & Duncan, R.P. (2001) Determinants of establishment success in
introduced birds. Nature, 414, 195197.
Bond, I., Child, B., Harpe, D. de la et al. (2004) Private land contribution to conser-
vation in South Africa. In Parks in Transition: biodiversity, rural development and
the bottom line, ed. B. Child. IUCN, SASUSG and Earthscan, London, Sterling,
Bothma, J. du P. (2002) Game Ranch Management. Van Schaik Publishers, Pretoria,
South Africa.
Bradley, E.H., Pletshcer, D.H., Bangs, E.E. et al. (2005) Evaluating wolf translocation
as a nonlethal method to reduce livestock conflicts in the Northwestern United
States. Conservation Biology, 19, 1498 1508.
Brussard, P.F. & Tull, J.C. (2007) Conservation biology and four types of advocacy.
Conservation Biology, 21, 2124.
Cade, T.J. & Burnham, W. (2003). Return of the Peregrine. The Peregrine Fund, Boise,
Cambray, J.A. (2003) Impact on indigenous species biodiversity caused by the global-
isation of alien recreational freshwater fisheries. Hydrobiologia, 500, 217230.
Animal Translocations 25
Carruthers, J. (2009) ‘Wilding the farm or farming the wild?’ The evolution of scientific
game ranching in South Africa from the 1960s to the present. Transactions of the
Royal Society of South Africa, 63(2), 160181.
Carter, H.R. (2003) Oil and California’s seabirds: an overview. Marine Ornithology,
31, 1–7.
Child, B. (2000) Making wildlife pay: converting wildlife’s comparative advantage
into real incentives for having wildlife in African savannas, case studies from
Zimbabwe and Zambia. In: Wildlife Conservation by Sustainable Use, eds H.H.T.
Prins, J.G. Grootenhuis & T.T. Dolan, pp. 335387. Kluwer Academic Publishers,
Choi, Y.D. (2004). Theories for ecological restoration in changing environments:
Toward ‘futuristic’ restoration. Ecological Restoration, 19, 7581.
Clout, M.N. (2006) A celebration of kakapo: progress in the conservation of an
enigmatic parrot. Notornis, 53, 1–2.
Clout, M.N. & Hay, J.R. (1981) South Island kokako (Callaeas cinerea cinerea)in
Nothofagus forest. Notornis, 28, 256259.
Clout, M.N. & Hay, J.R. (1989) The importance of birds as browsers, pollinators
and seed dispersers in New Zealand forests. New Zealand Journal of Ecology,12
(supplement), 2733.
Cousins, J.A., Sadler, J.P. & Evans, J. (2008) Exploring the role of private wildlife
ranching as a conservation tool in South Africa: Stakeholder perspectives. Ecology
and Society, 13, 43 [online] URL:
Craven, S., Barnes, T. & Kania, G. (1998) Toward a professional position on the
translocation of problem wildlife. Wildlife Society Bulletin, 26, 171 177.
Davidson, W.R., Appel, M.J., Doster, G.L. et al. (1992) Diseases and parasites of red
foxes, gray foxes, and coyotes from commercial sources selling to fox-chasing
enclosures. Journal of Wildlife Diseases, 28, 581–589.
Duncan, R.P. (1997) The role of competition and introduction effort in the success
of passeriform birds introduced to New Zealand. American Naturalist, 149,
Duncan, R.P., Blackburn, T.M. & Sol, D. (2003) The ecology of bird introductions.
Annual Review of Ecology and Evolutionary Systematics, 34, 7198.
du Toit, J.G. (2007) Role of the private sector in the wildlife industry. Report, Tshwane,
Wildlife Ranching SA/Du Toit Wilddienste. 87 pp.
Fischer, J. & Lindenmayer, D.B. (2000) An assessment of the published results of
animal relocations. Biological Conservation, 96, 111.
Gammons, D.J., Mengak, M.T. & Conner, L.M. (2009) Translocation of nine-banded
armadillos. Human–Wildlife Conflicts, 3, 64 –71.
Germano, J.M. & Bishop, P.J. (2009) Suitability of amphibians and reptiles for
translocation. Conservation Biology, 23, 715.
26 Philip J. Seddon, W. Maartin Strauss and John Innes
Goldsworthy, S.D., Giese, M., Gales, R.P. et al. (2000) Effects of the Iron Baron oil
spill on little penguins (Eudyptula minor). II. Post-release survival of rehabilitated
oiled birds. Wildlife Research, 27, 573582.
Goodrich, J.M. & Miquelle, D.G. (2005) Translocation of problem Amur tigers
Panthera tigris altaica to alleviate tigerhuman conflicts. Oryx, 39, 454457.
Grayson, D.K. (2001) The archaeological record of human impacts on animal popu-
lations. Journal of World Prehistory, 15, 168.
Griffith, B., Scott, J.M., Carpenter, J.W. et al. (1989) Translocation as a species
conservation tool: status and strategy. Science, 245, 477 480.
Gusset, M., Jokoby, O., Muller, M.S. et al. (2009) Dogs on the catwalk: modelling
re-introduction and translocation of endangered wild dogs in South Africa.
Biological Conservation, 142, 27742781.
Hartup, B.K. (1996) Rehabilitation of native reptiles and amphibians in DuPage
County, Illinois. Journal of Wildlife Diseases, 32, 109112.
Hayward, M.W., Adendorff, J., O’Brien, J. et al. (2007) Practical considerations
for the reintroduction of large, terrestrial, mammalian predators based on
reintroductions to South Africa’s Eastern Cape Province. Open Conservation
Biology Journal, 11, 11.
Hearne, J. & McKenzie, M. (2000) Compelling reasons for game ranching in Maputa-
land. In Wildlife Conservation by Sustainable Use, eds H.H.T. Prins, J.G. Grooten-
huis & T.T. Dolan, pp. 417 438. Kluwer Academic Publishers, Massachusetts.
Heinsohn, T.E. (2001) Human influences on vertebrate zoogeography: animal
translocation and biological invasions across and to the east of Wallace’s Line. In
Faunal and Floral Migrations and Evolution in SE Asia– Australia, eds I. Metcalfe,
M.B.J. Smith, M. Morwood & I. Davidson, pp. 153–170. Balkema, Rotterdam.
Henneman, M.L. & Memmott, J. (2001) Infiltration of a Hawaiian community by
introduced biological control agents. Science, 293, 1314 1316.
Herr, J., Schley, L. & Roper, T.J. (2008) Fate of translocated wild-caught and captive-
reared stone martens (Martes foina). European Journal of Wildlife Research, 54,
Higgins P.J., Peter J.M. & Cowling S.J. (2006) Handbook of Australian, New Zealand
and Antarctic Birds,vol.7,Boatbill to Starlings. Oxford University Press.
Hodder, K.H. & Bullock, J.M. (1997) Translocations of native species in the UK:
implications for biodiversity. Journal of Applied Ecology, 34, 547 565.
Hoddle, M.S. (2004). Restoring balance using exotic species to control invasive exotic
species. Conservation Biology, 18, 38 49.
Hoegh-Guldberg, O., Hughes, L., McIntyre, S. et al. (2008) Assisted colonization and
rapid climate change. Science, 321, 345 346.
Holdaway R.N. & Worthy T.H. (1997) A reappraisal of the late quaternary fossil
vertebrates of Pyramid Valley Swamp, north Canterbury, New Zealand. New
Zealand Journal of Zoology, 24, 69 121.
Animal Translocations 27
Holdaway R.N., Worthy T.H. & Tennyson, A. J. D. (2001) A working list of breeding
bird species of the New Zealand region at first human contact. New Zealand
Journal of Zoology, 28, 119 187.
Hunter, L.T.B., Pretorius, K., Carlisle, L.C. et al. (2007) Restoring lions Panthera leo
to northern KwaZulu-Natal, South Africa: short-term biological and technical
success but equivocal long-term conservation. Oryx, 41, 196 204.
Hutchins, M. (2008) Animal rights and conservation. Conservation Biology, 22,
Innal, D. & Erk’akan, F. (2006) Effects of exotic and translocated fish species
in the inland waters of Turkey. Reviews in Fish Biology and Fisheries, 16,
IUCN (World Conservation Union) (1987) IUCN position statement on the transloca-
tion of living organisms: introductions, re-introductions, and re-stocking. IUCN,
Gland, Switzerland.
IUCN (World Conservation Union) (1995) Guidelines for re-introductions. Annex 6
to the minutes of the 41st meeting of council. Gland, Switzerland.
IUCN (World Conservation Union) (1998) Guidelines for re-introductions. IUCN/
SSC Re-introduction Specialist Group, IUCN, Gland, Switzerland and Cambridge,
IWRC (International Wildlife Rehabilitation Council) (2009) What is wildlife reha-
bilitation? Retrieved 7 September 2009.
Jamieson, I.G., Wallis, G.P & Briskie, J.V. (2006) Inbreeding and endangered species
management: is New Zealand out of step with the rest of the world? Conservation
Biology, 20, 3847.
Jones, C.G. & Merton, D.V. (2011) A tale of two islands: the rescue and recovery
of endemic birds in New Zealand and Mauritius. In Reintroduction Biology:
integrating science and management, eds J.G. Ewen, D.P. Armstrong, K.A. Parker
& P.J. Seddon, Chapter 2. Wiley-Blackwell, Oxford, UK.
Jule, K.R., Leaver, L.A. & Lea, S.E.G. (2008) The effects of captive experience on rein-
troduction survival in carnivores: a review and analysis. Biological Conservation,
141, 355–363.
Keedwell, R.J., Maloney, R.F. & Murray, D.P. (2002) Predator control for protect-
ing kaki (Himantopus novaezelandiae): lessons from 20 years of management.
Biological Conservation, 105, 369374.
Kelly D., Ladley J.J., Robertson A.W. et al. (2010) Mutualisms with the wreckage of an
avifauna: the status of bird pollination and fruit-dispersal in New Zealand. New
Zealand Journal of Ecology, 34, 66 85.
Kirch, P.V. (2005) Archaeology and global change: the Holocene record. Annual
Review of Environmental Resources, 30, 409440.
Kleiman, D.G. (1989) Reintroduction of captive mammals for conservation. Bioscience,
39, 152–161
28 Philip J. Seddon, W. Maartin Strauss and John Innes
Le Gouar, P., Mihoub, J.-B. & Sarrazin, F. (2011) Dispersal and habitat selection:
behavioural and spatial constraints for animal translocations. In Reintroduction
Biology: integrating science and management, eds J.G. Ewen, D.P. Armstrong,
K.A. Parker & P.J. Seddon. Wiley-Blackwell, Oxford, UK.
Lenain, D.M. & Warrington, S. (2001) Is translocation an effective tool to remove
predatory foxes from a desert protected area? Journal of Arid Environments, 48,
Lewis, J.C., Sallee, K.L. & Golightly Jr, R.T. (1999). Introduction and range expansion
of nonnative red foxes (Vulpes vulpes) in California. American Midland Naturalist ,
142, 372–381.
Linnell, J.D.C., Annes, R., Swenson, J.E. et al. (1997) Translocation of carnivores as a
method for managing problem animals: a review. Biodiversity and Conservation,
6, 1245–1257.
Linnell, J.D.C., Odden, J., Smith, M.E. et al. (1999) Large carnivores that kill livestock:
do ‘problem individuals’ really exist? Wildlife Society Bulletin, 27, 698 705.
Lipsey, M.K. & Child, M.F. (2007) Reintroduction biology and restoration ecology:
are two disciplines better than one? Conservation Biology, 21, 1387 1388.
Lohr, C., Ballard, W.B. & Bath, A. (1996) Attitudes toward gray wolf reintroduction
to New Brunswick. Wildlife Society Bulletin, 24, 414420.
MacLachlan, J.S., Hellmann, J.J. & Schwartz, M.W. (2007) A framework for debate
of assisted migration in an era of climate change. Conservation Biology, 21,
Maguire, L.A. & Servheen, C. (1992) Integrating biological and social concerns
in endangered species management: augmentation of grizzly bear populations.
Conservation Biology, 6, 426434.
Marnewick, K., Hayward, M.W., Cilliers, D. et al. (2009) Survival of cheetahs relocated
from ranchland to fenced protected areas in South Africa. In Reintroduction of
Top-Order Predators, eds M. Hayward & M. Somers. Wiley-Blackwell Publishing,
Oxford, UK.
Mbaiwa, J.E. (2008) The success of consumptive wildlife tourism in Africa. In Tourism
and the Consumption of Wildlife, ed. B. Lovelock, pp. 141 –154. Routledge, London
and New York.
Mead, C. (1997) Poor prospects for oiled birds. Nature, 390, 449 –450.
Meadow,R.,Reading,R.P.,Phillips, al. (2005) The influence of persuasive
arguments on public attitudes toward a proposed wolf restoration in the southern
Rockies. Wildlife Society Bulletin, 33, 154163.
Mezat, J.A.K., Newman, S.H., Gilardi, K.V.K. et al. (2002) Advances in oiled bird
emergency medicine and management. Journal of Avian Medicine and Surgery,
16, 146–149.
Molony, S.E., Dowding, C.V., Baker, P.J. et al. (2006) The effect of translocation
and temporary captivity on wildlife rehabilitation success: an experimental study
Animal Translocations 29
using European hedgehogs (Erinaceus europaeus). Biological Conservation, 130,
Moore, M., Early, G., Touhey, K. et al. (2007) Rehabilitation and release of marine
mammals in the United States: risks and benefits. Marine Mammal Science, 23,
Moorhouse, T.P., Gelling, M. & Macdonald, D.W. (2009) Effects of habitat quality
upon reintroduction success in water voles: evidence from a replicated experiment.
Biological Conservation, 142, 5360.
Mueller, J.M. & Hellmann, J.J. (2008) An assessment of invasion risk from assisted
migration. Conservation Biology, 22, 562 567.
Nel, D.C., Crawford, R.J.M. & Parsons, N. (2003) The conservation status and impact
of oiling on the African Penguin. In Rehabilitation of Oiled African Penguins: a
conservation success story, eds D.C. Nel & P.A. Whittington, pp. 1 7. Birdlife
South Africa and the Avian Demography Unit, Cape Town, South Africa.
Nettles, V.F., Shaddock, J.H., Sikes, R.K. et al. (1979) Rabies in translocated raccoons.
American Journal of Public Health, 69, 601 602.
Nilsen, E.B., Milner-Gulland, E.J., Schofield, L. et al. (2007) Wolf reintroduction to
Scotland: public attitudes and consequences for deer management. Proceedings of
the Royal Society B, 274, 9951002.
O’Donnell, M.A. & DeNicola, A.J. (2006) Den site selection of lactating female
raccoons following removal and exclusion from suburban residences. Wildlife
Society Bulletin, 34, 366–370.
Parker, K.A. (2008) Translocations: providing outcomes for wildlife, resource
managers, scientists, and the human community. Restoration Ecology, 16,
Parker, K.A., Seabrook-Davison, M. & Ewen, J.G. (2010) Opportunities for non-
native ecological replacements in ecosystem restoration. Restoration Ecology, 18,
Parsons, N.J. & Underhill, L.G. (2005) Oiled and injured African penguins Spheniscus
demersus and other seabirds admitted for rehabilitation in the Western Cape,
South Africa, 2001 and 2002. African Journal of Marine Science, 27, 1 –8.
Platenberg, R.J. & Griffiths, R.A. (1999) Translocation of slow-worms (Anguis frag-
ilis) as a migration strategy: a case study from south-east England. Biological
Conservation, 90, 125–132.
Quakenbush, L., Beckmen, K. & Brower, C.D.N. (2009) Rehabilitation and release of
marine mammals in the United States: concerns from Alaska. Marine Mammal
Science, 25, 994999.
Regan, T. (1983) The Case for Animal Rights. University of California Press, Berkeley,
Ricciardi, A. & Simberlof, D. (2009a) Assisted colonization is not a viable conservation
strategy. Trends in Ecology and Evolution, 24, 248253.
30 Philip J. Seddon, W. Maartin Strauss and John Innes
Ricciardi, A. & Simberlof, D. (2009b) Assisted colonization: good intentions and
dubious risk assessment. Trends in Ecology and Evolution, 24, 476 477.
Richard-Hansen, C., Vie, J.-C. & de Thoisy, B. (2000) Translocation of red howler
monkeys (Alouatta seniculus) in French Guiana. Biological Conservation, 93,
Richardson, D.M., Hellmann, J.J., McLachlan, J.S. et al. (2009) Multidimensional
evaluation of managed relocation. Proceedings of the National Academy of Science,
106, 9721–9724.
Ricketts, T. & Imhoff, M. (2003) Biodiversity, urban areas, and agriculture: Locating
priority ecoregions for conservation. Conservation Ecology, 8(2), 1 [online] URL:
Rout, T.M., Hauser, C.E. & Possingham, H.P. (2009) Optimal adaptive manage-
ment for the translocation of a threatened species. Ecological Applications, 19,
Saunders, A. & Norton, D.A. (2001) Ecological restoration at Mainland Islands in New
Zealand. Biological Conservation, 99, 109119.
Sax, D.F., Smith, K.F. & Thompson, A.R. (2009) Managed relocation: a nuanced
evaluation is needed. Trends in Ecology and Evolution, 24, 472 473.
Schaub,M,Zink,R.,Beissmann, al. (2009) When to end releases in reintroduction
programmes: demographic rates and population viability analysis of bearded
vultures in the Alps. Journal of Applied Ecology, 46, 92 –100.
Seddon, P.J. (1999) Persistence without intervention: assessing success in wildlife
re-introductions. Trends in Ecology and Evolution, 14, 503.
Seddon, P. J. (2010). From reintroduction to assisted colonization: moving along the
conservation translocation spectrum. Restoration Ecology, 18, 796 802.
Seddon, P.J. & Soorae, P. (1999) Guidelines for subspecific substitutions in wildlife
restoration projects. Conservation Biology, 13, 177 184.
Seddon, P.J., Armstrong, D.P. & Maloney, R.F. (2007a) Developing the science of
reintroduction biology. Conservation Biology, 21, 303 312.
Seddon, P.J., Armstrong, D.P. & Maloney, R.F. (2007b) Combining the fields
of reintroduction biology and restoration ecology. Conservation Biology, 21,
Seddon,P.J.,Armstrong,D.P.,Soorae, al. (2009) The risks of assisted colonization.
Conservation Biology, 23, 788789.
Sekercioglu, C.H., Schneider, S.H., Pay, J.P. et al. (2008) Climate change, elevational
range shifts and bird extinctions. Conservation Biology, 22, 140 150.
SER (Society for Ecological Restoration) International Science and Policy Working
Group (2004) The SER International Primer on Ecological Restoration.Societyfor
Ecological Restoration International, Tucson.
Severinghaus, L.L. & Chi, L. (1999) Prayer animal release in Taiwan. Biological
Conservation, 89, 301304.
Animal Translocations 31
Shieh, B.-S., Lin, Y.-H., Lee, T.-W. et al. (2006) Pet trade as sources of introduced bird
species in Taiwan. Taiwania, 51, 8186.
Simberloff, D. & Stiling, P. (1996) Risks of species introduced for biological control.
Biological Conservation, 78, 185192.
Soorae, P.S. (ed.) (2008) Global re-introduction perspectives: re-introduction case
studies from around the globe. IUCN/SSC Re-introduction Specialist Group,
Abu Dhabi, UAE.
Spear, D. & Chown, S.L. (2009a) Non-indigenous ungulates as a threat to biodiversity.
Journal of Zoology, London, 279, 117.
Spear, D. & Chown, S.L. (2009b) The extent and impacts of ungulate translocations:
South Africa in a global context. Biological Conservation, 142, 353363.
Spencer, P.B.S. & Hampton, J.O. (2005) Illegal translocation and genetic structure of
feral pigs in Western Australia. Journal of Wildlife Management, 69, 377 384.
Stanley Price, M.R. (1989) Animal re-introductions: the Arabian oryx in Oman.
Cambridge University Press, Cambridge, UK.
Stanley Price, M.R. & Soorae, P. (2003) Re-introductions: whence and wither?
International Zoo Yearbook, 38, 6175.
Temperton, V.M. (2007) The recent double paradigm shift in restoration ecology.
Restoration Ecology, 15, 344347.
Temple, S.A. (1992) Exotic birds: a growing problem with no easy solution. Auk, 109,
Thomas, M.B. & Willis, A.J. (1998) Biocontrol – risky but necessary? Trends in Ecology
and Evolution, 13, 325329.
Underhill, L.G., Bartlett, P.A., Baumann, L. et al. (1999) Mortality and survival of
African Penguins Spheniscus demersus involved in the Apollo Sea oil spill: an
evaluation of rehabilitation efforts. Ibis, 141, 29–37.
van der Merwe, P., Saayman, M. & Krugell, W. (2008) Factors that determine the price
of game. Koedoe, 47(2), 105113.
van Vuren, D., Kuenzi, A.J., Loredo, I. et al. (1997) Translocation as a non-lethal alter-
native for managing California ground squirrels. Journal of Wildlife Management,
61, 351–359.
Veltman, C.J., Nee, S. & Crawley, M. (1996) Correlates of introduction success in
exotic New Zealand birds. American Naturalist, 147, 542–557.
Vitousek, P.M., D’Antonio, C.M., Loope, L.L. Rejmanek, M. & Westbrooks, R. (1997)
Introduced species: a significant component of human-induced global change.
New Zealand Journal of Ecology, 21, 1 16.
Vitt, P. Havens, K. & Hoegh-Guldberg, O. (2009) Assisted migration: part of an
integrated conservation strategy. Trends in Ecology and Evolution, 24, 473 474.
Wakamiya, S.M. & Roy, C.L. (2009) Use of monitoring data and population viability
analysis to inform reintroduction decisions: Peregrine falcons in the Midwestern
United States. Biological Conservation, 142, 17671776.
32 Philip J. Seddon, W. Maartin Strauss and John Innes
Walker, K.J., Trewick, S.A. & Barker, G.M. (2008) Powelliphanta augusta,anewspecies
of land snail, with a description of its former habitat, Stockton coal plateau. Journal
of the Royal Society of New Zealand, 38, 163–186.
Waples, K.A. & Stagoll, C.S. (1997) Ethical issues in the release of animals from
captivity. Bioscience, 47, 115121.
Watson, M. & Thirgood, S. (2001) Could translocation aid hen harrier conservation
in the UK? Animal Conservation, 4, 3743.
White, J.P. (2004) Where the wild things are: prehistoric animal translocations in the
New Guinea Archipelago. In Voyages of Discovery: the archaeology of islands,ed.
S.M. FitzPatrick, pp. 147–164. Praeger Publishers, Westport, Connecticut.
Williams, C.K., Ericsson, G. & Heberlein, T.A. (2002) A quantitative summary of
attitudes toward wolves and their reintroduction (1972– 2000). Wildlife Society
Bulletin, 30, 575–584.
Wolf, C.M., Garland, T. & Griffith, B. (1998) Predictors of avian and mammalian
translocation success: reanalysis with phylogenetically independent contrasts.
Biological Conservation, 86, 243–255.
Wolf, C.M., Griffith, B., Reed, C. et al. (1996) Avian and mammalian translocations:
update and reanalysis of 1987 survey data. Conservation Biology, 10, 1142 –1154.
Zahniser, A. & Singh, A. (2004) Return of the wolves to Yellowstone National Park,
USA: a model of ecosystem restoration. Biodiversity,5,37.
... Connectivity among reintroduced populations within a network of suitable habitat or other extant or reintroduced populations is an important condition to mitigate reintroduction failures (Richardson et al. 2015). Management actions in reintroduction biology have focused largely on more immediate threats to population persistence, such as habitat suitability, emigration, and predation (Armstrong and Seddon 2008;Ewen 2012). Only recently have studies begun to incorporate explicit models and evaluations of the permeability of landscapes for reintroduced individuals (Kramer-Schadt et al. 2004;La Morgia et al. 2011;Cianfrani et al. 2013;Richardson et al. 2015;Howell et al. 2016;Jarchow et al. 2016;Torres et al. 2017). ...
... We collected contemporary samples for Chequamegon (2012Manlick et al. 2017), Nicolet (2015Grauer et al. 2017), and Iron County (2014) during winter (December through April) using noninvasive hair-snare techniques (Pauli et al. 2008;Manlick et al. 2017). Hair snares consisted of a 38-cm length of 10-cm diameter PVC pipe with steel wire brushes inserted at a perpendicular angle to the pipe approximately 5 cm from each end. ...
Species reintroductions are successful when established populations maintain both demographic stability and genetic diversity. Such a result may be obtained by ensuring both structural habitat connectivity and genetic connectivity among reintroduced and remnant populations. Nevertheless, prezygotic barriers such as assortative mating can prevent the flow of genetic material between populations, even when migration between populations is high. Limited gene flow may be particularly relevant for reintroductions that were sourced either from captive-bred populations or from disparate locations in the wild. American martens (Martes americana) have been reintroduced repeatedly in the Upper Midwestern United States in an effort to establish self-sustaining populations. We quantified levels of genetic diversity within and spatial genetic variance among four marten populations during two time periods separated by 10 years. Spatially informed and naïve discriminant analysis of principal components were used to assign individuals to populations. Results indicate that heterozygosity declined and inbreeding coefficients increased between the two collection periods, while genetic structure among populations also increased. Data are consistent with assortative mating contributing to reapportioning of genetic variation. Population assignment tests show that migration among populations is apparent, but admixture (based on cluster membership probabilities) is low and declined over time. Specifically, martens may be successfully dispersing between populations but a lack of admixture indicates a lack of reproductive contributions to genetic diversity by migrants. Because marten reintroductions in this region are well-documented and well-monitored, lessons can be derived from results to inform future reintroductions. We encourage a careful balance of supplementing genetic diversity via augmentation while avoiding translocation of animals from disparate populations that may result in reproductive isolation of migrants. In combination with the maintenance of a functionally connected landscape, this strategy would maximize the likelihood of a successful reintroduction in terms both of demography and genetics.
... This raises the stakes for conservation, but it also presents an opportunity to use interventions targeted at single species to achieve broader ecological objectives (Simberloff, 1998). Conservation translocations typically focus on species recovery but can also promote ecosystem recovery by restoring lost mutualisms, reintroducing keystone species, or introducing ecological replacements (e.g., Ewen et al., 2012;Seddon, 2010;Seddon et al., 2014). For example, seed dispersal has been restored through reintroduction of brown howler monkeys (Alouatta guariba clamitans) (Genes et al., 2018) and red-rumped agoutis (Dasyprocta leporina) (Mittelman et al., 2020) in Brazil and ecological replacement of extinct giant tortoises by extant species in Mauritius (Griffiths et al., 2011) and the Galápagos (Hunter et al., 2013). ...
... The difficulty of predicting, and then monitoring, one species' contributions to the ecosystem may explain why translocation programs often rely on proxies to measure function. Yet, obtaining accurate measures of function is becoming more essential as ecological restoration becomes a higher priority in many translocation programs, including reintroductions, reinforcements, and ecological replacements (often as part of rewilding) (e.g., Seddon, 2010;Ewen et al., 2012;Seddon et al., 2014). We, therefore, encourage translocation programs to monitor this objective directly when possible and to select metrics carefully so success can be evaluated and improved through adaptive management . ...
Full-text available
Conservation translocation is a common method for species recovery, for which one increasingly frequent objective is restoring lost ecological functions to promote ecosystem recovery. However, few conservation translocation programs explicitly state or monitor function as an objective, limiting the ability to test assumptions, learn from past efforts, and improve management. We evaluated whether translocations of hihi (Notiomystis cincta), a threatened New Zealand passerine, achieved their implicit objective of restoring lost pollination function. Through a pollinator-exclusion experiment, we quantified, with log response ratios (lnR), the effects of birds on fruit set and seed quality in hangehange (Geniostoma ligustrifolium), a native flowering shrub. We isolated the contributions of hihi by making comparisons across sites with and without hihi. Birds improved fruit set more at sites without hihi (lnR = 1.27) than sites with hihi (lnR = 0.50), suggesting other avian pollinators compensated for and even exceeded hihi contributions to fruit set. Although birds improved seed germination only at hihi sites (lnR = 0.22–0.41), plants at sites without hihi had germination rates similar to hihi sites because they produced 26% more filled seeds, regardless of pollination condition. Therefore, although our results showed hihi improved seed quality, they also highlighted the complexity of ecological functions. When an important species is lost, ecosystems may be able to achieve similar function through different means. Our results underscore the importance of stating and monitoring the ecological benefits of conservation translocations when functional restoration is a motivation to ensure these programs are achieving their objectives.
... Increased competition between large carnivore species through artificially high population densities (or rapid population growth) and the associated high degrees of territory overlap because of restricted space (Palomares and Caro 1999;Hayward and Kerley 2008) are points of concern and require monitoring and research (Ewen et al. 2012). This is particularly true in South Africa, where a large diversity of carnivore species is found throughout various fenced nature reserves throughout the country (Wentzel et al. 2021). ...
Context: The spatio-temporal partitioning of large carnivores is influenced by interspecific competition and coexistence within small, enclosed reserves. Lions (Panthera leo), spotted hyaenas (Crocuta crocuta) and leopards (Panthera pardus) are the three largest African carnivores and have the greatest potential for intra-guild competition, particularly where space is limited. Aim: To investigate the spatio-temporal partitioning between lions, spotted hyaenas and leopards in a small (~75 000 ha), enclosed nature reserve, Madikwe Game Reserve (Madikwe), South Africa. Methods: We deployed 110 camera traps (baited n = 55 and unbaited n = 55) across Madikwe from 26 August 2019 until 6 May 2020. Von Mises kernel density plots were used to investigate daily temporal partitioning among the three species. A multiple-species, single-season occupancy model was used to investigate daily space use patterns. Key results: We found both temporal and spatial exclusion between lions and spotted hyaenas on Madikwe. However, no evidence was found of spatio-temporal partitioning between lions and leopards, and spotted hyaenas and leopards. Conclusions: Exploitative and interference competition on Madikwe might be high enough to warrant spatio-temporal partitioning between lions and spotted hyaenas to avoid the negative effects of intra-guild competition. Contrastingly, patterns observed between leopards and both lions and spotted hyaenas preclude the possibility of top-down control by superior carnivores. Implication: These findings call for an adaptive management approach, where both carnivore and prey species compositions are constantly monitored. Management strategies such as these will allow for the conservation of valuable resources (i.e. prey species) to ensure the persistence of large carnivore populations across African ecosystems.
... The small population on Stirling appeared resilient to the relatively high level of male mortality in 2013 and continued to grow through 2017. Continued monitoring of this fisher population will provide the best opportunity to investigate population persistence and to identify any potential drivers of population change in the future (Ewen et al., 2012). Our implementation of a SCR population model allowed us to incorporate unique aspects of the work completed on Stirling that have implications for other monitoring efforts. ...
Full-text available
Understanding the role of landscapes managed for timber production in the conservation of forest-obligate species is a priority for preserving ecological integrity and fostering socioeconomic wellbeing. The forest characteristics generally associated with the survival, reproduction, and persistence of forest-obligate species (e.g., large-diameter trees, standing dead trees, understory vegetation, downed logs) are often believed to be at odds with timber production. One such species is the fisher (Pekania pennanti), a mesocarnivoran (member of the order Carnivora) associated with mature forest characteristics whose range has decreased substantially since the mid-1800s. Fishers exemplify the perceived conflict between forest-obligate species and timber production because they generally require areas exhibiting complex forest structure including multiple canopy layers, old trees and standing dead trees with cavities, logs, and understory vegetation that provide sufficient prey, escape cover, and structures to support reproduction and parturition. Consequently, understanding if fishers can persist in landscapes managed for timber production can provide a critical test of the compatibility among forest-obligate species and forest management. We reintroduced 40 fishers (24 females, 16 males) between November 2009 and December 2011 onto a landscape managed for timber production to establish a new fisher population and to evaluate the viability of fisher populations on a forest managed for timber production. We studied this reintroduced population of fishers for 8 years following the reintroduction using annual live-captures and year-round tracking with radio telemetry. Using population modeling with spatial capture-recapture methods, we estimated this population of fishers to be growing during the 7-year study period. The density of the reintroduced fisher population in 2017 (10.8 fishers/100 km²) was within the reported range of fisher densities across the western United States. The reintroduction of fishers to previously occupied portions of their range is an important component of fisher conservation and will play a role in the recovery of the species in western portions of the fisher's range. Our results suggest that forests managed for timber production with landscape conditions similar to our study area may be important for future fisher reintroductions and species recovery.
Movement ecology and environmental factors are topics of paramount importance to consider when planning conservation programmes for target species. Here we discuss this topic by reviewing the available information related to the Egyptian Vulture Neophron percnopterus, with reference to the remnant breeding population of Southern Italy, of high conservation concern and subject of a long-term captive-breeding re-stocking programme. We describe how adverse wind conditions over the Central Mediterranean Sea make the sea-crossing challenging with detrimental effects on the survival of inexperienced birds, and coupled this information with count data of migrating Egyptian Vultures. Furthermore, we indicate how low population size and scarce opportunities in meeting migrating conspecifics could potentially lead juvenile Egyptian Vultures to follow unfavourable migratory routes, with possible repercussions on survival. We postulate how these concomitant factors could be indirectly influencing the long-term survival of this small population, principally affected by anthropogenic threats. We also discuss how the same factors could actually be affecting captive-bred young individuals released in late summer in southern continental Italy, in the framework of the restocking programme. An integrative approach with tailor-made release methods, which also takes into account the age of released birds and geographical and environmental factors, would likely be useful for a more goal-oriented and long-lasting conservation outcome, for the preservation of this endangered scavenger.
Common species play a disproportionate role in shaping ecosystem structure and function, but are currently under‐represented in conservation translocation initiatives. This represents a missed opportunity because common species are typically easier to source for restoration projects, and larger numbers of common species can feasibly be translocated without substantially impacting source populations. Reintroduction of common species is an important first step in the faunal restoration of severely impacted habitats, such as urban spaces. Common species typically retain higher genetic diversity than threatened species, but this also means that they may have more to lose via population bottlenecks that can occur from translocation. To inform efforts to translocate common species, we assessed genetic impacts of a reintroduction of the common native bush rat Rattus fuscipes to an urban reserve at North Head, Sydney (Australia). Using single‐nucleotide polymorphism diversity, we found that differentiation between source populations was low. Nevertheless, admixture during reintroduction and follow‐up translocations initially increased standardized observed heterozygosity of North Head‐born bush rats and population NE, with a subtle corresponding decrease in within‐population kinship. For 3 years following the last translocation, we detected a small decline in genetic diversity in the North Head population, although final statistics remained similar to the source populations. Our results indicate that no short‐term interventions are necessary to further promote bush rat genetic diversity at North Head, but that continued genetic monitoring will be important to determine whether a trend in declining diversity continues as the population stabilizes. We conclude that translocation of a large number of individuals from multiple sources presents a suitable option for restoring an extirpated small mammal population whilst minimizing genetic effects typically associated with such management actions. Common species present viable candidates for translocations aiming to return biodiversity to disturbed or fragmented urban ecosystems. Native Australian bush rat Rattus fuscipes identified during a recent survey at North Head, Sydney. Using genomic analysis of the reintroduced bush rat population at the site, our study shows that translocations of common species can successfully maintain population genetic diversity, an important outcome as urban restoration improves ecosystem services and increases opportunities for engagement with nature. Photo credit Angela Raña / Australian Wildlife Conservancy.
Although numerous predictions have been developed regarding climate change, many current tools and methods represent a reductionist approach and are insufficient to fully assess the ecological impact of climate change, and the adaptive capacity of species to mitigate these impacts. A complex systems approach is required: assessing the impacts of climate change on species is flawed without consideration of variation in behaviour and social organisation. The aim of this study is to evaluate the utility of using simulations to understand how the full range of behavioural variation, and social organisation, contributes to the adaptive capacity of species to respond to climate change. A high-resolution agent-based model was developed to simulate behavioural variation within a community of antelope species in Mokala National Park, South Africa, subject to different climate change scenarios. The model was then extended to evaluate and compare the effectiveness of two separate management interventions (habitat management and park expansion), in conserving the community under climate change. Comparisons were based on individuals’ energy levels and whether there were selection pressures for specific behaviour or social groups. The results suggest both types of intervention show promise for mitigation of climate change effects but both schemes also selected for specific behaviour types and social groups. This may impact on populations ability to adapt to change and may affect the social cohesion of these populations.
Full-text available
This chapter provides an introduction to "Conserving Biodiversity on Military Lands – A Guide for Natural Resources Managers, 3rd Edition". The chapter describes what biodiversity is, how it is faring across the United States, and the role that military lands play in conserving the nation's biodiversity. The chapter also describes DoD's evolving approach to conservation and natural resource management and how effective biodiversity conservation is key to sustaining military readiness. The chapter ends with a discussion of recent trends in military biodiversity conservation, including a focus on broader landscape collaborations, innovation in endangered species recovery, and climate change adaptation and resilience.
Full-text available
The genetic composition of an individual can markedly affect its survival, reproduction, and ultimately fitness. As some wildlife populations become smaller, conserving genetic diversity will be a conservation challenge. Many imperiled species are already supported through population augmentation efforts and we often do not know if or how genetic diversity is maintained in translocated species. As a case study for understanding the maintenance of genetic diversity in augmented populations, I wanted to know if genetic diversity (i.e., observed heterozygosity) remained high in a population of gray wolves in the Rocky Mountains of the U.S. > 20 years after reintroduction. Additionally, I wanted to know if a potential mechanism for such diversity was individuals with below average genetic diversity choosing mates with above average diversity. I also asked whether there was a preference for mating with unrelated individuals. Finally, I hypothesized that mated pairs with above average heterozygosity would have increased survival of young. Ultimately, I found that females with below average heterozygosity did not choose mates with above average heterozygosity and wolves chose mates randomly with respect to genetic relatedness. Pup survival was not higher for mated pairs with above average heterozygosity in my models. The dominant variables predicting pup survival were harvest rate during their first year of life and years pairs were mated. Ultimately, genetic diversity was relatively unchanged > 20 years after reintroduction. The mechanism for maintaining such diversity does not appear related to individuals preferentially choosing more genetically diverse mates. Inbreeding avoidance, however, appears to be at least one mechanism maintaining genetic diversity in this population.
Full-text available
A summary of re-introduction projects in a standardized format from around the world totalling 62 case-studies. These case studies cover the following taxa as follows: invertebrates (4), fish (7), amphibians (3), reptiles (8), birds (17), mammals (13) and plants (10). I
In response to increases in human-wildlife conflicts in urban and suburban environments, many states have established nuisance wildlife control operator (NWCO) programs. Criticism has been leveled at such programs for insufficient emphasis on nonlethal means, particularly the exclusion and translocation of nuisance small and medium-sized mammals. We determined den site selection of lactating raccoons (Procyon lotor) after exclusion and subsequent release from a human residence. Fifteen of 20 females extracted from suburban residences returned to another house at least once. Of the den sites selected 2 months postrelease, 59% were in man-made structures. Our data suggest that most female raccoons removed and excluded from human dwellings will select another house if released on site. We contend this behavior is facilitated by suburban homeowners who fail to take the initiative to prevent entry of wildlife into their homes. (WILDLIFE SOCIETY BULLETIN 34(2):366–370; 2006)
Restoration of gray wolves (Canis lupus) to their original range depends not only on a sound ecological basis but also on public acceptance. We samplest 4 special interest groups in New Brunswick about a hypothetical reintroduction to this area. Two white-tailed deer (Odocoileus virginianus) hunter groups and 2 naturalist groups were sampled by questionnaire to test the hypothesis that deer hunters would have more negative attitudes and be less, willing to reintroduce wolves to New Brunswick than would members of naturalist groups. Deer hunters in northern New Brunswick, where deer bunting was closed due to low numbers of deer, were more negative about a reintroduction than southern deer hunters (deer seasons open) and members of naturalist groups. None of the groups were willing to reintroduce wolves to New Brunswick. Positive attitude anti greater willingness to support reintroduction were correlated with higher education, not having previously hunted big game, and less fear of hiking in the woods knowing wolves were present. Knowledge-of-wolf scores for all groups were low. The most common reason given for opposing wolf reintroduction was that it would result in a deer population decline. If wolf reintroduction were ever to be contemplated for New Brunswick, education programs would be necessary to placate public fear of deer population declines.
This chapter is about establishing mechanisms that price wildlife, how these mechanisms work and how they can be valuable for promoting economic development and conservation simultaneously. It uses several examples, mainly from Zimbabwe, to describe how wildlife was converted from a public good with little or even a negative value to landholders, into a private good which landholders or communities have a positive incentive to produce. It explains why wildlife has a comparative economic advantage and is often a better use of agriculturally marginal savannahs than more conventional livestock monocultures, and provides data from the private ranching sector in Zimbabwe to support this argument. The central assertion in the chapter is that both wildlife conservation and economic development are best served in much of savanna Africa by converting wildlife into a commercial asset. This is achieved by modifying macro-economic institutions and legislation so that mechanisms develop to ensure prices more closely reflect scarcity or value, and resources are allocated more efficiently. This would ensure that where wildlife has a comparative advantage, it would be reflected in incentive structures and landholders would produce wildlife rather than livestock which owes much of its past prominence to fiscal and environmental subsidisation.
The context of modern conservation management of kakapo is introduced and a brief overview of the research presented in this special issue of Notornis is provided.
The last decade has seen a considerable increase in the area of land devoted to wildlife in South Africa. In this chapter we argue that this has been driven by economic rather than conservation reasons. A strong local and overseas market within the right political and legal framework have made game ranching an economically attractive proposition for many landowners in certain areas of South Africa such as Maputaland. An obstacle to the creation of game ranches is the large sum of initial capital required for the acquisition of game and the erection of fences and other infrastructure. This chapter provides four case studies on how this problem was overcome. Shareblock schemes provide one method for raising the required capital. This method has had a significant impact on the growth of game ranching in South Africa and so the schemes are discussed in some detail. The use of mathematical models have contributed to the optimal utilisation of capital in the set-up phase of a game ranch and in determining management strategies to maximise sustainable income. A case study is provided where this approach was used to calculate an optimal mix of species. Another mathematical model then generated the sequence of purchases and acquisitions of game that minimise total net capital outflow during the set-up period. This minimisation enabled the optimal sustainable populations to be attained within three years taking into account ecological constraints and the limited availability of some species. The model output also includes a detailed cash flow statement for the set-up period.