Journal of Mammalogy, 100(3):923–941, 2019
© 2019 American Society of Mammalogists, www.mammalogy.org
Conservation of the world’s mammals: status, protected areas,
community efforts, and hunting
R.T B,* MS. B, JR. G, JL. R
Institute of Arctic Biology, Box 757000, University of Alaska Fairbanks, Fairbanks, AK 99775, USA (RTB)
Department of Biological Sciences, University of Alberta, Edmonton, Alberta T6G 2R3, Canada (MSB)
Department of Zoology and Physiology, University of Wyoming, 1000 East University Avenue, Laramie, WY 82071, USA (JRG)
Department of Fish and Wildlife Sciences, University of Idaho, 875 Perimeter Drive, MS 1136, Moscow, ID 83844, USA (JLR)
* Correspondent: firstname.lastname@example.org
Mammals are imperiled worldwide. Threats to terrestrial species are primarily from habitat loss or modication,
and in some instances from commercial, illegal, or unregulated hunting. Terrestrial species are negatively affected
throughout the tropics from deforestation. Threats to marine mammals are related to harvest, strikes in shipping
lanes, pollution, and depleted levels of food resources. Hazards to marine species are pronounced in the North
Atlantic Ocean, North Pacic Ocean, and oceans and seas anking southeastern Asia. Protected areas designed to
conserve mammals often are too small, too few, poorly delimited or isolated, and too unreliably supported. The
new conservation science proposes that human livelihoods be considered alongside traditional preservationist
perspectives. For conservation outside of protected areas to succeed, the protection of wild mammals and their
habitats should result in benet to local people, especially in rural or poor communities. Concerns about declining
populations of large mammals in North America during the late 19th and early 20th centuries resulted in the
institution of regulations that contributed to the recovery of many populations. Today, in North America and
Europe, wild populations are thriving and legal hunting is allowed for a number of mammals, something that is
less common in many developing countries, where illegal killing remains a threat to conservation. Nevertheless,
populations of large mammals are resilient to regulated hunting because of density-dependent processes that
result in increased reproduction, survival, and growth rates. Unfortunately, hunting is unregulated for cultural and
economic reasons over much of the Earth. We are beginning to see effects of climate change and invasive species
on risk of extinction for many species. The future of mammals, however, is entwined ultimately with the size,
growth, and resource demands of the human population.
Los mamíferos están en riesgo en todo el mundo. Las principales amenazas a las especies terrestres son la pérdida
y modicación del hábitat, y en algunos casos la cacería comercial, ilegal o no regulada. Las especies terrestres
son negativamente afectadas a lo largo de los trópicos por la deforestación. Las amenazas a los mamíferos
marinos están relacionadas con la extracción ilegal, colisiones con embarcaciones, contaminación y agotamiento
de recursos alimentarios. Los peligros para las especies marinas están más acentuados en el Atlántico Norte, el
Pacíco Norte y los mares y océanos que rodean el sureste asiático. Las áreas protegidas diseñadas para proteger
mamíferos frecuentemente son muy pequeñas, escasas, pobremente delimitadas o aisladas, además de contar
con un mantenimiento incierto. La nueva ciencia de la conservación propone que las necesidades humanas sean
consideradas en conjunto con las perspectivas preservacionistas tradicionales. Para que la conservación fuera
de las áreas protegidas sea exitosa, la protección de los mamíferos silvestres y sus hábitats debería redundar en
benecios para los habitantes locales, especialmente en comunidades rurales o con altos niveles de pobreza. La
preocupación por las poblaciones en declive de grandes mamíferos en Norteamérica durante nales del siglo
XIX y principios del XX, resultaron en la implementación de regulaciones que contribuyeron a la recuperación
de muchas poblaciones. En la actualidad, las poblaciones silvestres de Norteamérica y Europa son saludables, y
la cacería legal es permitida para varias especies de mamíferos, algo que es menos común en muchos países en
desarrollo, donde la cacería ilegal continúa siendo una amenaza para la conservación. Sin embargo, las poblaciones
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924 JOURNAL OF MAMMALOGY
de grandes mamíferos son resilientes a la cacería regulada debido a procesos denso-dependientes que resultan
en incrementos en las tasas de reproducción, supervivencia y crecimiento poblacional. Desafortunadamente, la
cacería no está regulada por razones culturales y económicas en la mayor parte del mundo. Estamos comenzando
a ver los efectos del cambio climático y de las especies invasoras sobre el riesgo de extinción de muchas especies.
El futuro de los mamíferos, sin embargo, está nalmente ligado al tamaño, crecimiento y demanda de recursos
de las poblaciones humanas.
Key words: conservation, density dependence, extinction, habitat, harvest, human-occupied landscapes, hunting, protected areas,
reserves, threat levels
Mammals capture our attention and engender curiosity,
thereby enhancing their intrinsic value. This likely occurs
because humans are mammals and because of our close and
long-term association with canids, felids, and other domestic
mammals as companions or livestock (Clutton-Brock 1999;
Diamond 2002), and our traditional reliance on wild mam-
mals for food and sport (Hull 1964). This afliation between
humans and mammals creates opportunities for education
about populations, ecological communities, ecosystems,
global climate, and conservation of all species. Thus, mam-
mals are the most likely group to motivate investment of
resources for conservation. Nonetheless, mammals are in
We review the conservation of mammals, and do so by
examining threats to orders of mammals, differentiating among
factors threatening terrestrial and aquatic species. We also ex-
amine strengths and weaknesses of protected areas, and how to
achieve conservation objectives in human-occupied landscapes.
We further explore the role of hunting in the conservation of
mammals, differentiating between legal, regulated harvests,
and illegal killing. Our goal is to provide an informed perspec-
tive on the threats to mammals and discuss feasible approaches
to their worldwide conservation.
Mammalian extinctions.—Mammals are imperiled world-
wide. Of 5,487 species recognized by the IUCN Red List, 76
(1.4%) have been declared extinct since 1500, with little chance
that any still exist (IUCN 2017). Extinctions are spread across
the 27 orders of mammals, and most occur in orders that are
species-rich: Primates (2 extinctions), Carnivora (5), Chiroptera
(5), Cetartiodactyla (7), Diprotodontia (7), Eulipotyphla (7), and
Rodentia (36). Slightly over one-half (55.6%) of the 27 mam-
malian orders are not known to have experienced extinctions
since 1500 (IUCN 2017). Rates of recent mammalian extinc-
tions, however, far exceed past levels, necessitating increased
efforts to conserve threatened mammals (Ceballos and Ehrlich
2002; Isaac etal. 2007; Pimm et al. 2014). Moreover, > 70%
of endangered mammals are characterized by declining pop-
ulations, which bodes poorly for their continued existence
(Ceballos etal. 2017). Also, extinctions alone fail to describe
the full extent of the threat faced by mammals.
Categories of threatened mammals.—Combining Red List
(IUCN 2017) categories of Extinct (EX), Extinct in the Wild
(EW), Critically Endangered (CR), Endangered (EN), and
Vulnerable (VU), 1,219 species of mammals (22.2%) are
ranked as Threatened or Extinct. These categories are based
on the quantitative IUCN system for classifying threatened
species (Mace etal. 2008). The median percent of Threatened
or Extinct species for the 27 orders is 19.0%; only six com-
paratively small orders (i.e., Dermoptera, Hyracoidea,
Microbiotheria, Monotremata, Notoryctemorphia,
Tubulidentata) have no species in that combined category.
Those numbers, although disturbing, likely underestimate
the true threat to mammals because an additional 323 (5.9%)
species are categorized as Near Threatened (NT), with a
median for the 27 orders of 0.6%. Chiroptera (23.8%) and
Rodentia (31.9%) have the largest percentage of NT species.
Moreover, many NT species exhibit downward trends in their
population sizes (Schipper etal. 2008; Ceballos etal. 2017).
Slightly more than one-half of mammalian species (56.7%)
are considered to be of Least Concern (LC), with a median
of 54.7% across orders. An additional 836 (15.2%) species
of unknown status are listed as Data Decient (DD), with a
median of 13.6% across the 27 orders. Orders with relatively
large numbers of Data Decient species are Primates (56),
Cetartiodactyla (62), Eulipotyphla (77), Chiroptera (204), and
Rodentia (369). Data Decient species of terrestrial mammals
are concentrated in tropical forests, and Data Decient marine
species are clustered in the Antarctic Convergence (Schipper
etal. 2008). In addition to identifying a need for further re-
search on the status of Data Decient species, this category
holds import for determining how Threat Levels are assessed
Threat Levels for orders of mammals.—The IUCN Red
List currently recognizes 5,411 species of extant mam-
mals (IUCN 2017). The Threat Level (sensu Schipper etal.
2008) for those mammals can be assessed and bracketed by
evaluating whether Data Decient species would be cate-
gorized as threatened or not threatened. Therefore, Threat
Level=[(VU + EN + CR + EW)/(Total − DD)] × 100, where
acronyms represent the number of species within those
IUCN Red List categories. The lower bound for the Threat
Level=[(VU + EN + CR + EW)/Total] × 100; the upper
bound for the Threat Level= [(VU + EN + CR + EW +
DD)/Total] × 100. The median Threat Level across 27 orders
of mammals is 23.0%, with a median for the lower bound
of 18.8%, and a median for the upper bound of 33.8%.
Davidson etal. (2017) reported that 36% of Data Decient
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BOWYER ET AL.—CONSERVATION OF MAMMALS 925
species of mammals have extrinsic and intrinsic factors
associated with a high risk of extinction in other species.
Threat Levels are extremely variable (Fig. 1), and large dif-
ferences exist in causes of potential risk of extinction among
mammalian orders (Davidson etal. 2009, 2017). Orders con-
taining large-bodied species (e.g., Cetartiodactyla, Primates,
Perissodactyla, Proboscidea, Sirenia, and Carnivora) have
comparatively high Threat Levels (Fig. 1). In addition to
taxonomic order and body size, other intrinsic characteris-
tics such as small geographic range, inhabiting islands as
opposed to the mainland, and having a slow speed of life his-
tory (e.g., low adult mortality, iteroparity, small litter size,
high maternal investment in young, long generation times,
low intrinsic rates of increase) are associated with high risks
of extinction (Davidson etal. 2017).
Geographic patterns of species richness.—Land mammals
exhibit a peak in species richness centered on the equator
(Grenyer et al. 2006; Schipper et al. 2008; Ceballos et al.
2017). Species richness is especially pronounced in the Andes
Mountains of South America, and in the mountains of Africa
and the southern Arabian Peninsula (Afromontane forests).
High species richness also occurs in Asia, particularly in south-
western China, Malaysia, and Borneo (Heaney 1986; Schipper
Marine mammals have zones of high species richness clus-
tered around 40° North or 40° South latitude. The prominent
exception is in the North Atlantic Ocean, resulting from past
exploitation by humans (Schipper etal. 2008; Davidson etal.
2012). Notable areas of high species richness for marine mam-
mals include tropical and temperate coastal benches, as well
as offshore portions of the Tasman and Caribbean Seas, and
the Southern Indian Ocean (Grenyer etal. 2006; Schipper etal.
2008; Pompa etal. 2011). Areas with threatened species of ter-
restrial and marine mammals are characterized by high species
richness, high endemism, and high human pressure (Schipper
etal. 2008; Ceballos etal. 2017).
Geographic patterns and types of threats.—Threatened land
mammals are concentrated in southern and southeastern Asia,
tropical portions of the Andes Mountains in South America, the
highlands of Cameroon, the Albertine Rift in Africa, and the
Western Ghats in India (Schipper et al. 2008; Ceballos etal.
2017). For marine mammals, threatened species occur in the
North Atlantic Ocean, North Pacic Ocean, and oceans and
seas proximate to southeastern Asia (Davidson etal. 2012).
Species of mammals with small geographic ranges are espe-
cially threatened in southeastern Asia, whereas species with
large geographic ranges are threatened in Africa, and portions
of Asia and the Arctic (Davidson et al. 2017). Fundamental
Fig. 1.—Threat Levels for 27 orders of mammals. Error bars indicate lower and upper bounds for Threat Levels. Threat Level=[(VU + EN +
CR + EW)/(Total − DD)] × 100. The lower bound for the Threat Level=[(VU + EN + CR + EW)/Total] × 100; the upper bound for the Threat
Level=[(VU + EN + CR + EW + DD)/Total] × 100 (Schipper etal. 2008). This analysis is based on the IUCN system for classifying threatened
species (Mace etal. 2008): VU=number of Vulnerable species; EN=number of Endangered species; CR=number of Critically Endangered
species; EW=number of species Extinct in the Wild; DD=number of species that are Data Decient. Data from IUCN (2017). J.Kenagy and
J.Bradley provided access to the phylogeny for inclusion in this gure.
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926 JOURNAL OF MAMMALOGY
differences exist for mammals inhabiting the Northern and
Southern Hemispheres. Endemics of the southern-temperate
zone have smaller geographic range sizes and are at a greater
risk of extinction than are their counterparts from the Northern
Hemisphere (Lamoreux and Lacher 2010). Moreover, southern-
temperate endemics are less likely to occur within protected
areas than are those from the Northern Hemisphere (Lamoreux
and Lacher 2010).
Habitat loss and degradation (40% of affected species),
and overharvesting (17%) pose the greatest threats to mam-
mals worldwide (Schipper et al. 2008). Those threats are
usually driven by human population density and climate
change, which pose increasingly greater threats to the future
of mammals (Davidson etal. 2017). Threats to land mam-
mals, however, differ markedly from those for marine spe-
cies. Terrestrial species are negatively affected throughout
the tropics from deforestation, especially in the Americas,
Asia, and Africa (Schipper et al. 2008). Negative effects
of illegal killing on land mammals are most pronounced
in Asia, but also in portions of Africa and South America.
Overharvesting may result from poaching for food, the bush-
meat trade, and use of animals for medicinal and other pur-
poses, as well as from accidental capture in snares set for
other species (Schipper et al. 2008; Harrison etal. 2016).
Overharvesting disproportionately affects larger mammals
(Cetartiodactyla, Primates, Perissodactyla, Proboscidea, and
Carnivora) compared with smaller ones, possibly because
large mammals have slower life histories and larger home
ranges compared with small mammals—characteristics that
can increase vulnerability to harvest (Cardillo etal. 2005;
Schipper etal. 2008). Nonetheless, legal and regulated sport
hunting, as practiced in much of North America and Europe,
seldom threatens and may offer conservation benets to land
mammals (Organ etal. 2010).
Disease is a threat to extinction for comparatively few (2%)
species of mammals (Schipper et al. 2008). Nevertheless,
some diseases can have catastrophic effects on particular spe-
cies. For example, a transmissible facial cancer in Tasmanian
devils (Sarcophilus harrisii) devastated populations of this
marsupial, yet because of their rapid evolutionary response,
hope exists that they may not become extinct (Epstein etal.
2016). White-nose syndrome, caused by a pathogenic fungus
(Pseudogymnoascus), has killed millions of bats (Chiroptera)
since its emergence in eastern North America in 2006 (Frick
etal. 2016). Nonetheless, a possibility exists for limiting the
spread of this disease because the fungus is extremely sensitive
to ultraviolet light, which might be applied to hibernating bats
with a few seconds of exposure from a portable light source
(Palmer etal. 2018). Chronic wasting disease, which is caused
by a prion, is uniformly fatal in cervids, and presents enormous
management difculties in coping with the spread of this infec-
tious disease (Saunders etal. 2012; Potapov etal. 2016). How
emerging infectious diseases might interact with other threats
to mammals, such as habitat loss and climate change, is uncer-
tain. For instance, recent mass die-offs of the endangered saiga
antelope (Saiga tatarica), associated with a common bacterium
(Pasteurella multocida) under conditions of warmer tempera-
tures and higher humidity, indicate that threats from diseases
are likely to increase in unanticipated ways in the future (Kock
In addition to threats posed by disease-causing microor-
ganisms, invasive animal and plant species threaten mammals
worldwide. Based on IUCN data from 2014, alien species
were listed as a causal threat for 70% (n=30) of 43 extinc-
tions (Bellard etal. 2016). Furthermore, during recent decades,
increased globalization has accelerated rates of movement and
introduction of alien species. Moreover, trade and transport of
goods and people globally have created new pathways for bio-
logical invasions and new challenges for stemming the loss of
biodiversity resulting from the spread of alien species (Hume
Invasive species can threaten mammal populations in di-
verse and interacting ways. First, invasive predators often are
effective in killing naive prey, and this is especially preva-
lent for insular populations. Arecent review identied 189
mammals representing 45 families that have been negatively
affected by invasive predators, with the greatest numbers
occurring in Australia and Central America (Doherty etal.
2016). At least 28 of Australian’s endemic mammals have
become extinct since European settlement, and extinctions
continue to occur at a rate of one to two per decade, pri-
marily as a result of predation by invasive alien predators
(Woinarski et al. 2015). Generalist predators such as feral
cats (Felis catus), canids, and rodents are among the most
challenging for island mammals. Feral cats have been re-
ported to threaten ≥ 27 mammalian species on ≥ 120 islands
(Medina et al. 2011). Efforts to remove invasive predators
from islands have met with success and offer hope for res-
toration of island fauna (Jones etal. 2016). Recent summa-
ries indicate that efforts to eradicate invasive species have
achieved > 85% success rates (Genovesi 2011). Nonetheless,
growing recognition of the complexity of multispecies inter-
actions also cautions that unintended consequences can occur
following removal or reduction of invasive species (Lurgi
Second, invasive species can compete with native ones di-
rectly for key resources and indirectly through the spread
of parasites (Dunn etal. 2012) and facilitation of predation
(i.e., apparent competition). Introduction of feral pigs (Sus
scrofa) to the California Channel Islands provided abundant
food for golden eagles (Aquila chrysaetos) that colonized
the islands and subsequently preyed heavily on the native
Channel Island foxes (Urocyon littoralis—Roemer et al.
2002). Such novel biological interactions can reduce tness
of native mammals, resulting in population declines that
lead to extirpations.
Third, plant invasions can alter community composition
with numerous potential consequences for mammals and the
ecosystems they inhabit (Vilà etal. 2011). Such effects include
reduction in availability or quality of forage for herbivores,
changes in hydrological cycles (Huxman etal. 2005), and al-
teration of historic re regimes (Brooks etal. 2004). Invasion
BOWYER ET AL.—CONSERVATION OF MAMMALS 927
of exotic annual plants, including cheatgrass (Bromus tecto-
rum), in the western United States has increased frequency
and severity of res (Balch etal. 2013), killing sagebrush
(Artimesia spp.) shrubs and precipitating loss of shrub-steppe
habitats that support sagebrush-dependent mammals such
as pygmy rabbits (Brachylagus idahoensis) and sagebrush
voles (Lemmiscus curtatus). The uncertainty associated with
effects of invasive species under changing climate regimes is
an emerging threat to mammals. We address the issue of cli-
mate change later, but emphasize here that shifts in climate
have the potential to alter consequences of species invasions
in ways that have yet to be imagined. We view this as an im-
portant area for research to support future conservation of
For marine mammals, the principal threats are from acci-
dental mortality (affecting 78% of species), including sher-
ies bycatch and vessel strikes, and pollution (60% of species).
Pollution involves chemicals, marine debris, noise, and climate
change (Schipper etal. 2008). Harvesting of marine mammals
remains a concern (52% of species) largely because of com-
mercialization, notwithstanding improvements via interna-
tional agreements (Davidson etal. 2012; Bowen and Lidgard
2013; McCaughley etal. 2015).
Comparisons of terrestrial and aquatic taxa within
Cetartiodactyla.—A positive relationship exists between
threat of extinction and body size (Purvis etal. 2000; Cardillo
et al. 2005). Evaluating threats to large mammals of the
Cetartiodactyla may offer insights into patterns that species in
other orders mayface.
For the 10 families of terrestrial species within
Cetartiodactyla, a median of 55.5% of the species in those
families are Threatened or Extinct. All species in three small
families (Girafdae, Hippopotamidae, Moschidae) are in that
category. Threatened or Extinct species compose 37.3% of the
species-rich Bovidae (142 species) and 49.0% of the Cervidae
(54 species). Comparatively few species (16) from those two
large families are Data Decient, which is a small percentage
(6.7%) of their extant species. Nevertheless, Threat Levels for
terrestrial families of Cetartiodactyla (Table 1) are high com-
pared with other mammalian orders (Fig. 1). Amedian Threat
Level of 58.3%, with large lower (53.5%) and upper (61.8%)
bounds are a clear cause for concern.
For the 11 families of aquatic species (some are not ma-
rine) within Cetartiodactyla, no species, with the possible
exception of the baiji (Lipotes vexllifer), was thought to be ex-
tinct (Turvey etal. 2007). The IUCN (2017) lists that species
as Critically Endangered with an unknown population trend.
Amedian of 33.3% of aquatic species in Cetartiodactyla were
categorized as Threatened (Table 2). This value likely is bi-
ased low because of the large number (45) of Data Decient
species, which constituted 51.7% of aquatic species. Threat
Levels to aquatic species within Cetartiodactyla (Table 2) are
high compared with other orders (Fig. 1), with a median of
50.0% (the lower bound for Threat Levels was 33.3% and
the upper bound 75%). The upper bound for Threat Levels
for aquatic families exceeded that for terrestrial families
within Cetartiodactyla, largely because of the numerous Data
Decient aquatic species.
The primary factors threatening terrestrial Cetartiodactyla are
similar across families based on the top three threats listed by
the IUCN (2017) for each family (as determined by the number
of species experiencing those threats). The category of “biolog-
ical resource use” resulting principally from hunting and trap-
ping was a primary threat for all 10 families. Nonetheless, many
of those threats pertained to historical use of species, or illegal
killing; no such reports listed legal and regulated sport harvest
of mammals as a threat. Likewise, all 10 families included the
IUCN category of “agriculture and aquaculture” as a threat,
mostly a result of livestock farming or ranching. One-half of
the families listed “residential and commercial development”
as a threat, 20% included “human intrusions and development,”
Table 1.—Threat levels with upper and lower bounds for families
of terrestrial mammals in the order Cetartiodactyla. Threat Level=
[(VU + EN + CR + EW)/(Total − DD)] × 100. The lower bound for the
Threat Level=[(VU + EN + CR + EW)/Total] × 100; the upper bound
for the Threat Level=[(VU + EN + CR + EW + DD)/Total] × 100.
VU=number of Vulnerable species; EN=number of Endangered spe-
cies; CR=number of Critically Endangered species; EW=number
of species Extinct in the Wild; DD=number of species that are Data
Decient. Data from IUCN (2017).
Antilocapridae 1 0.0 0.0 0.0
Bovidae 142 36.0 35.0 37.9
Camelidae 3 33.0 33.0 33.0
Cervidae 54 57.8 48.1 64.8
Girafdae 2 100.0 100.0 100.0
Hippopotamidae 2 100.0 100.0 100.0
Moschidae 7 100.0 100.0 100.0
Suidae 17 58.8 58.8 58.8
Tayassuidae 3 66.7 66.7 66.7
Tragulidae 10 14.2 0.0 40.0
Table 2.—Threat levels with upper and lower bounds for aquatic
families in the order Cetartiodactyla. Threat Level= [(VU + EN +
CR + EW)/(Total − DD)] × 100. The lower bound for the Threat
Level = [(VU + EN + CR + EW)/Total] × 100; the upper bound
for the Threat Level=[(VU + EN + CR + EW + DD)/Total] × 100.
VU = Vulnerable; EN= Endangered; CR =Critically Endangered;
EW= Extinct in the Wild; DD= Data Decient. Data from IUCN
Balaenidae 4 50.0 50.0 50.0
Balaenopteridae 8 60.0 37.5 75.0
Ziphiidae 21 0.0 0.0 90.5
Delphinidae 36 15.8 8.3 55.6
Monodontidae 2 0.0 0.0 0.0
Eschrichtiidae 1 0.0 0.0 0.0
Iniidae 3 100.0 66.7 100.0
Neobalaenidae 1 0.0 0.0 100.0
Physeteridae 3 100.0 33.3 100.0
Phocoenidae 7 60.0 42.9 71.4
Platanistidae 1 100.0 100.0 100.0
928 JOURNAL OF MAMMALOGY
10% listed “natural system modications,” and 10% identied
“invasive and other problematic species and diseases.” All of
those categories except hunting and disease can be considered
to be a consequence of habitat loss or modication.
The 11 families of aquatic species within Cetartiodactyla also
face common threats. Foremost among those primary threats is
“biological resource use” (90.9%), denoted mostly as harvest,
sheries bycatch, or harvesting of other aquatic resources. The
only family that did not list “biological resource use” as a threat
is the monotypic Neobalinenidae, for which no threats were listed
because the pygmy right whale (Caperea marginata) was cat-
egorized as Data Decient. “Pollution” also is a common threat
(63.6%) in aquatic families; followed by “climate change and
severe weather” (45.5%); “transportation and service corridors”
(27.3%), which involve principally risks of collisions in shipping
lanes; and “invasive and other problematic species and diseases”
Understanding patterns of extinction.—Mammals face dire
threats throughout their distribution; the continued existence of
many species, especially large-bodied, vagile taxa, is in ques-
tion. Our results, however, are extremely conservative and do
not fully address threats to mammalian species. The IUCN
Red List considers only species-level taxa in categorizing the
status of mammals, and consequently does not reect threats
to subspecies or local populations (Butchart and Dunn 2003;
Ceballos etal. 2017). For instance, the pronghorn (Antilocapra
americana) is listed as a species of Least Concern, yet the
Sonoran pronghorn (A.a.sonoriensis) is Endangered (Hosack
etal. 2002). Bighorn sheep (Ovis canadensis) also is a species
of Least Concern, but the Sierra Nevada subspecies (O.c.sier-
rae) is Endangered (Schroeder et al. 2010). The white-tailed
deer (Odocoileus virginianus) has a wide distribution in the
Americas and is a species of Least Concern, yet the Florida Key
deer (O.v.clavium) is Endangered (Haversohn etal. 2004).
Another problematic component of the IUCN database relates
to the number of extant mammalian species. For instance, the
Mammal Diversity Database (https://mammaldiversity.org)
currently recognizes 6,399 species of extant mammals, with 96
species recently extinct (Burgin etal. 2018). Moreover, a long-
term average of 25 new species has been described each year
(Burgin etal. 2018). Undescribed species probably have expe-
rienced some extinctions, which would result in further under-
estimating extinction rate. Moreover, analyses based on IUCN
(2017) data have 912 fewer mammalian species than recog-
nized by the Mammal Diversity Database. The IUCN statistics
clearly document past threats to mammals but may not track
current ones, especially for rapidly occurring changes in the
status of some mammalian species. Indeed, population status
for many species is poorly documented in developing countries
of the world, and nonexistent for the 912 mammal species that
do not occur on the IUCN list (Burgin etal. 2018).
Further, the IUCN Red List has been criticized as unscien-
tic and in need of additional or revised metrics for classifying
threatened species (Mace and Lande 1991; Nowak 2009; Joppa
etal. 2016). Nevertheless, some of those criticisms are based
on misconceptions concerning the manner in which species
are classied into categories (Collen et al. 2016). Despite
some shortcomings, we believe the IUCN Red List provides a
benchmark against which the changing status of mammal spe-
cies may be judged. The emphasis on extinction of species, re-
gardless of the database used, has resulted in underestimating
threats to mammals (Ceballos etal. 2017).
Some mammals have experienced massive contractions in
their geographic ranges during historic times (Laliberte and
Ripple 2004), and many species of Least Concern are de-
clining, which may result in local extinctions (Craigie etal.
2010; Ceballos etal. 2017). Large mammals, in particular, have
the potential to greatly inuence ecosystems that they inhabit
(McNaughton 1979; Stewart et al. 2006), but we must nd
ways to implement conservation measures that will benet a
diversity of mammalian taxa (Ford etal. 2017). Ultimately, the
fate of mammals is intertwined with the size, growth, and re-
source demands of the human population (Vitousek etal. 1997;
Czech etal. 2000; Gaston 2005; McKee etal. 2013).
Loss of habitat and habitat fragmentation (Wilcox and
Murphy 1985) are quintessential factors threatening most
mammals, especially terrestrial taxa. Habitat restoration holds
the potential to reduce rates of extinction in fragmented wood-
lands (Newmark etal. 2017). Although there has been rapid
progress in developing protected areas, those sites do not nec-
essarily protect biodiversity (Andelman and Willig 2003; Pimm
etal. 2014; Bleich 2016).
E P A
Human livelihoods versus protection.—Curbing extinction
and preserving populations are two major goals of mammal
conservation. Historically, protected areas (national parks,
reserves, and other legal designations intended to limit human
activities with the overarching goal of conserving nature) have
been a hub for conservation efforts. Recently, however, con-
servationists have come to realize that protected areas often
are too small (Woodroffe and Ginsberg 1998; Bleich 1999,
2005; Gonzalez-Maya et al. 2015), too few (Gaston et al.
2008; Newmark 2008; Durant etal. 2017), poorly delimited
or isolated (Bleich 2014, 2016), or too unreliably supported
(Andelman and Willig 2003; Caro and Scholte 2007; Craigie
etal. 2010) to meet many long-term (centuries to millennia)
conservation prospects. Consequently, some conservation-
ists are transitioning to a model in which the improvement of
human livelihoods is considered (or even prioritized) alongside
This outcome has been dubbed the “new conservation sci-
ence” (Kareiva and Marvier 2012), and it is reminiscent of the
creation of Integrated Conservation and Development Projects
(ICDPs). In the mid-1980s, and largely in response to failures
perceived with traditional conservation efforts in protected
areas (i.e., “fences and nes” approaches), the World Wide
Fund for Nature introduced 19 ICDPs in an attempt to meet so-
cioeconomic priorities of rural communities and conservation
goals simultaneously (Hughes and Flintan 2001). These ICDPs
were established across Africa and South America, normally
BOWYER ET AL.—CONSERVATION OF MAMMALS 929
in conjunction with national parks and typically funded by
western governments and nongovernment organizations, with
the goal of beneting local communities through tourism dol-
lars and job creation (Hughes and Flintan 2001). The validity of
the ICDP model for conservation rests on at least two assump-
tions. First, local peoples are hostile toward formally protected
areas, and their livelihoods pose a direct threat to biodiversity
(Hughes and Flintan 2001). Second, ICDPs offer complemen-
tary or alternative livelihoods to locals that are sustainable, will
increase living standards, and will reduce threats posed to bi-
odiversity (Newmark and Hough 2000). These assumptions
are rarely tested and are sometimes wrong. Moreover, a pos-
sible (but not universal—Salerno etal. 2014) unintended con-
sequence of ICDP establishment is the increased immigration
of people toward the outskirts of protected areas (Barrett and
Arcese 1995; Scholte 2003; Guerbois etal. 2013).
Generally, shifts from traditional conservation to empha-
sis on human livelihoods are controversial (e.g., Kareiva and
Marvier 2012; Doak etal. 2014, 2015; Vucetich etal. 2015),
and are exemplied by two quotations: “Nature preserves…are
not places to be saved to be used at a later stage when an ever-
growing human population claims more land because of lack of
economic development” (Prins 1992); and “Why should all that
land be set aside for tourists when it can be used for farming?
These white people care more about one dead elephant than
they do for a hundred black children” (Obama 1995, quoting
his sister). At the heart lies debate over the validity of two fre-
quent claims (Young 2006): conservation and improved human
livelihoods are compatible; and economic interests, particularly
those of rural communities, must be satised for conservation
to be successful (i.e., “wildlife must pay its way”). Appealing
though they may be, evidence for the two claims is equivocal
in that valid examples refuting or supporting each are readily
available. Each is true for some conservation efforts, in some
areas, some of thetime.
With these considerations in mind, we review the costs and
benets of protected areas and multi-use, human-occupied
landscapes, with particular attention to if and how each has
shaped conservation outcomes for mammals. We then consider
two examples in which on-the-ground conservation efforts are
relying on a combination of formally protected areas and land-
scapes inhabited by local communities to bolster populations
of two related, rare antelopes (Bovidae). Although these exam-
ples are broadly applicable, we focus attention on eastern and
southern Africa, which house iconic protected areas (Kruger
National Park, Serengeti National Park, Maasai Mara National
Reserve, Mara Conservancy, and other nearby conservancies),
and are occupied by people and their livestock over 75–85%
of the landscape, and thus lack formal protection (Chape etal.
2005; Newmark 2008).
Costs, benets, and logistics of protected areas for conserv-
ing wild mammals.—We follow Dudley (2008) in dening a pro-
tected area as a “clearly dened geographical space, recognized,
dedicated and managed, through legal or other effective means,
to achieve the long-term conservation of nature with associated
ecosystem services and cultural values.” By this denition, the
world’s rst protected area (the Main Ridge Forest Reserve in
Tobago) was created in 1776 via British parliamentary ordi-
nance to collect rainfall for agriculture elsewhere on the island
(UNESCO 2011). Since this time, more than 202,000 protected
areas have been established across nearly 15% of the Earth’s
terrestrial surface (U.N. Environment Conservation Monitoring
Centre 2017). Many successes in mammal conservation have
arisen precisely because of the establishment of, or targeted
efforts within, formally protected areas, which also can serve
as critical reference points for restoration (Arcese and Sinclair
1997; Berger 2008). For the 24 species of wild mammals for
which populations have been bolstered by conservation efforts
in recent decades, one-half have benetted demonstrably from
conservation in protected areas (Hoffmann etal. 2010, 2011).
In an extreme example, conservation within a protected area
(the Serengeti National Park–Maasai Mara National Reserve
complex) almost certainly prevented the conservation status of
common wildebeest (Connochaetes taurinus) from deteriorat-
ing from Least Concern to Critically Endangered (Hoffmann
Alternative views exist regarding the utility of protected areas
as ecological baselines. For example, Sarmento and Berger
(2017) demonstrate that increased rates of human visitation to
Glacier National Park have caused mountain goats (Oreamnos
americanus) to relax antipredator behaviors, thereby poten-
tially reshaping predator–prey interactions. More generally,
protected areas (even those designated as “wilderness”) often
are too small, too isolated, or both to buffer wide-ranging ungu-
lates and carnivores from human inuence outside their borders
(Woodroffe and Ginsberg 1998; Bleich 2016). Therefore, there
is some risk that our view of “pristineness” is illusory, and our
perception of ecological baselines subjective.
Conservation within protected areas requires a combination
of reduction (or exclusion) of people and the creation of eco-
nomic benets from wildlife to people around protected areas.
With regard to the former, Hilborn etal. (2006) demonstrated
that enhanced antipoaching efforts in Serengeti National Park
were sufciently strong to increase numbers of buffalo (Syncerus
caffer), African elephant (Loxodonta africana), and black rhi-
noceros (Diceros bicornis). Examples of effective enforcement
within protected areas are rare, despite their critical importance
for conservation. Some have suggested that increased funding
for antipoaching patrols, rather than increased sentences for
captured poachers, is a more economically efcient means of
deterring poaching (Dobson and Lynes 2007). Because poach-
ing reduces the effective size of protected areas, and because
protected areas display classic species–area relationships (Fig.
2), economic inputs toward enforcement should result in less
disparity between the true size of a protected area and its effec-
tive size (Leader-Williams and Milner-Gulland 1993; Dobson
and Lynes 2007).
Fencing of protected areas has been a particularly conten-
tious subject in recent years, although this debate is rooted in
discussions from the mid-1980s. The controversy (Creel etal.
2013; Packer etal. 2013a, 2013b) is related to how best to mini-
mize retaliatory killing of large carnivores (particularly African
930 JOURNAL OF MAMMALOGY
lions, Panthera leo) following livestock depredation, as well
as effects of habitat loss and degradation from livestock and
farming, and depletion of native prey because of their use as
bushmeat and competition with livestock. On one hand, pop-
ulation persistence of African lions is projected to be highest
within well-funded and fenced protected areas (where African
lions are limited primarily by density dependence), and lowest
within poorly funded, unfenced protected areas associated
with high population densities of people (Packer etal. 2013a).
Conversely, critics have argued that African lion densities are
articially high within fenced protected areas, and these are too
expensive to install and maintain in most locales (Creel etal.
2013). One point of agreement is that African lion conservation
requires a larger, sustained nancial base to be successful over
Still, protected areas are not a panacea, and often fall short as
a stopgap to prevent population declines, local extirpations, or
even extinction. This happens for a number of reasons, foremost
among them illegal hunting within protected areas that exist in
name only, but that lack adequate enforcement (so-called “paper
parks”—Gates 1999). Although they occur largely in protected
areas, West African chimpanzees (Pan troglodytes verus) are
on the brink of extinction in several countries; the same is true
for addax (Addax nasomaculatus) and dama gazelle (Nanger
dama), despite occurring in the largest protected area on the
continent (the 100,000 km2 Termit Massif Reserve in Niger). In
2000, Miss Waldron’s red colobus monkey (Piliocolobus wal-
dronae) was declared extinct in a network of national parks in
Ghana and Ivory Coast (Oates etal. 2000; but see the following
section). Additionally, extinction risk is amplied in small re-
serves in West Africa via demographic stochasticity (Brashares
etal. 2001) and in the Horn of Africa because of compromised
ability of several antelopes with small geographic ranges to
track suitable habitats as drying and warming intensies (Payne
and Bro-Jørgensen 2016).
Because protected areas generally are viewed as safeguards
in which pristine and fragile nature is physically separated from
human activities, protected areas risk being perceived as incom-
patible with human livelihoods. This observation is particularly
applicable to developing countries, where more afuent nations
are responsible for the bulk of conservation funding (Hickey
and Pimm 2011; Miller etal. 2013). Such division between the
nations that fund conservation and nations that bear the costs
of conservation can result in mistrust of government agencies
and other organizations dedicated to conservation in protected
areas. By virtue of their proximity to protected areas, local
peoples are asked to tolerate direct (e.g., livestock depredation,
crop loss) or indirect (e.g., restrictions on development, com-
petition from game species) economic losses to wild mammals
(Waylen etal. 2010; Bruskotter etal. 2017). Perceptions aside,
local peoples derive at least some benets from protected areas,
as evidenced by growing human populations on the outskirts
of national parks and reserves in Africa and Latin America
(Wittemyer etal. 2008), and increasing human pressures within
a substantial fraction of protected areas (Jones etal. 2018).
Costs, benets, and logistics of human-occupied landscapes
for conserving wild mammals.—We consider conservation in
human-occupied landscapes (areas with conservation needs
that also include people) as synonymous with community-
based conservation, in which local involvement in areas lack-
ing formal protection is aimed at enhancing conservation and
maintaining or improving living standards (Berkes 2004).
Conservation in human-occupied landscapes may complement
conservation in formally protected areas, or it may be the only
option through which conservation can occur (Western etal.
2015). The latter is particularly the case in developing nations,
where governments lack the authority or resources to advance
conservation. In such instances, the explicit consideration of
human livelihoods may be crucial for curbing extinction and
bolstering population sizes of rare or declining mammals.
Conservation failure is the norm in human-occupied land-
scapes, but it is by no means a foregone conclusion. For con-
servation outside of formally protected areas to succeed, the
protection of wild mammals and their habitats should result in
some benet to local people. Examples include direct payments
from ecotourism (Sindinga 1995) and trophy hunting (Lindsey
etal. 2006), ecosystem services in the form of enhanced forage
quality (Odadi etal. 2011), and reduced risk of tick-borne dis-
eases (Allan etal. 2017). These benets arise from the diversi-
cation of land uses, which increasingly are being implemented
to buffer against unpredictable climates and livestock markets
(Reid etal. 2014). These benets also should be tied directly to
decisions made by local residents, who should oversee conser-
vation efforts (Goodwin and Roe 2001).
Fig. 2.—Species richness of large (> 10kg) ungulates displays a clas-
sic species–area relationship (power function) for protected areas in
East Africa, implying extinction risk is a function of size of protected
areas (Brashares etal. 2001). Protected areas comprise savanna wood-
land or grassland habitats in eastern and southern Africa, and include
Amboseli National Park, Bontebuck National Park, Etosha National
Park, Kora National Reserve, Kruger National Park, Lake Manyara
National Park, Lake Nakuru National Park, Moremi Game Reserve,
Nechisar Nation Park, Queen Elizabeth National Park, Serengeti
National Park, South Luanga National Park, Tsavo National Park (East
and West combined), and West Caprivi Game Reserve. Data from the
IUCN World Database on Protected Areas (https://www.iucn.org/
BOWYER ET AL.—CONSERVATION OF MAMMALS 931
C E H-O
As with protected areas, examples of conservation success
in human-occupied landscapes are many. Black-footed ferret
(Mustela nigripes) restoration is occurring on privately owned
rangelands in Wyoming because of a recent clause of the
Endangered Species Act, affording regulatory relief for land-
owners in the form of relaxed prohibitions on take (USFWS
2014). Similar to ferrets, other species of mammals thought
to be extinct, have been rediscovered in human-occupied
landscapes, including the Philippine naked-backed fruit bat
(Dobsonia chapmani) on Cebu and Negros Islands (rediscov-
ered in 2000—Paguntalan etal. 2004), and the Miss Waldron’s
red colobus monkey in the Ehy Forest of Ivory Coast (rediscov-
ered in 2002—McGraw 2005). Engaging local people in partic-
ipatory forest management has been very effective in reducing
forest loss around Chitwan National Park, Nepal (Stapp etal.
Human–wildlife conict frequently constrains opportuni-
ties for conservation in human-occupied landscapes, particu-
larly in the absence of strong local investment in conservation
or transparent benets to individuals from conservation. This
occurs because common resources, such as rangelands, forests,
and water, are susceptible to overexploitation via the Tragedy
of the Commons (Hardin 1968). Self-governance of common
resources can benet wildlife, if major decisions are made
locally by communities (Fennell 2011). Ostrom (1990) articu-
lated eight principles that are necessary to prevent a commons
tragedy from arising, four of which are particularly germane
when considering conservation efforts in human-occupied
1) Clearly dened common resources, and effective exclu-
sion of unentitled parties;
2) Collective choice, such that users of common resources
can participate in decision-making (or elected ofcials
serve as proxies in decision-making);
3) Sanctions for those who violate community rules in
appropriating common resources; and
4) Recognition of leaders by higher-level authorities of the
The success of conservation in human-occupied landscapes
hinges strongly on those principles. Ultimately, their imple-
mentation requires strong leadership through heads, elders, or
other respected individuals in the community, who have strong
public support (e.g., Kothari etal. 2013; Hazzah etal. 2014).
Local conservation requires local knowledge, which is neces-
sary but insufcient for conservation in human-occupied land-
scapes. Local conservation also requires local leaders.
In the following section, we detail two cases in which current
conservation efforts rely on a combination of protected areas
and human-occupied landscapes to bolster population sizes of
two species of imperiled antelope. Both lean heavily on strong
local knowledge and local leadership. At rst glance, these
examples seem similar: the antelope species are close relatives,
occur in Kenya, and share the landscape with people and their
livestock. But each example highlights a truism: while conser-
vation is a global problem, the solutions are typically local,
and require sustained input over the long term (Pringle 2017).
Inevitably, such long-term dedication necessitates a focus on
place: intimate knowledge of system-specic details of the pro-
tected areas or human-occupied landscapes in which conserva-
tion challenges occur and can bexed.
Case study 1: livestock production as a tool to lessen appar-
ent competition for hartebeest.—The aim of this project is to
reverse declines of hartebeest (Alcelaphus buselaphus) inhabit-
ing the 10,000 km2 Laikipia Plateau of central Kenya over one
or more decades. Since the mid-1980s, hartebeest and several
other populations of wild ungulates in Laikipia have declined,
due to the increasing tolerance of ranchers toward large carni-
vores, particularly African lions. The rate and timing of these
declines is the same, indicating one or more common underlying
mechanisms. Georgiadis etal. (2007a) tested seven nonexclu-
sive hypotheses (including poaching, parasitism, interspecic
competition, intraspecic competition, habitat conversion, dis-
placement by humans, and exceptional rainfall patterns) for
those declines, rejecting all but predation. Further, the authors
suggested that many species of wild ungulates in Laikipia, the
hartebeest in particular, were suppressed via apparent competi-
tion with plains zebra (Equus burchelli), a wild ungulate that
has not exhibited the same steep declines characteristic of oth-
ers in Laikipia, and that is the most common prey of African
lions (Georgiadis etal. 2007b; Frank 2011).
At Ol Pejeta Conservancy in Laikipia, Ng’weno etal. (2017)
reported support for the hypothesis that African lions had sup-
pressed population growth of hartebeest. Additionally, harte-
beest were preferred by African lions over 10 wild ungulate
species with which they co-occurred, and exhibited an African
lion-mediated Allee effect (Ng’weno 2017). Risk of mortality
to hartebeest from African lion predation increased in associa-
tion with plains zebra and dense vegetation (Ng’weno etal.
2017). These observations support the idea that an increasing
African lion population in Laikipia has been subsidized by
plains zebra and caused hartebeest to decline to < 10% of their
Finding lethal control of African lions improper and extir-
pation of hartebeest populations undesirable, Ng’weno (2017)
implemented a solution to bolster numbers of hartebeest that
did not require lethal control of African lions. Within Laikipia,
ranchers corral livestock nightly in temporary, circular corrals
(“bomas”) to reduce predation (Ogada etal. 2003). Bomas are
occupied between 1 and 6months, after which they are aban-
doned and livestock are moved in a rotational scheme. Over
1year, bomas transition into nutrient-rich grazing lawns, as
dung and urine break-down and enrich the soil (Veblen 2012;
Porensky and Veblen 2015). Those grazing lawns are attrac-
tive to plains zebra, but are virtually ignored by hartebeest.
Consequently, Ng’weno (2017) used bomas to create grazing
lawns away from hartebeest territories, thereby manipulat-
ing the distribution of primary prey (zebra) on the landscape
and providing a spatial refuge from African lion predation for
932 JOURNAL OF MAMMALOGY
hartebeest. This approach more than doubled survival rates
of hartebeest, although it is too early to know whether this
effect is sufciently strong to reverse population declines of
Case study 2: engaging Somali communities to conserve
hirola antelope.—This project involves hirola (Beatragus
hunteri), a close relative of hartebeest that is classied as
Critically Endangered by the IUCN, with a global popula-
tion size of < 500 individuals (King etal. 2011). Hirola have
never been common, and are restricted to grassland habitats
east of the Tana River on the Kenya–Somalia border; such
small-ranged species are particularly susceptible to demo-
graphic stochasticity, and thus present major conservation
challenges (Caughley 1994). In the mid-1980s, geographic
range collapse (and the associated global population crash)
occurred in response to a rinderpest (Morbillivirus) outbreak
(IUCN 2008), such that hirola are now restricted to an ap-
proximate 1,200 km2 swath of land on the Kenya–Somalia
border. Nonetheless, eradication of rinderpest from the Horn
of Africa in the early 2000s did not prompt hirola recovery
(Ali etal. 2017). At approximately the same time that hirola
populations plummeted, nancial backing for the only pro-
tected area in the region, Arawale National Reserve, dwindled
because of remoteness of the area, lack of a viable tourism
industry, and lack of local involvement.
Motivated by increasing calls for in situ conservation and
local involvement, the Ishaqbini Hirola Conservancy was es-
tablished in eastern Kenya in 2005 by Terra Nuova (an Italian
nongovernmental organization) and has been overseen since
by the Northern Rangelands Trust (a Kenyan nongovern-
ment organization supported largely by USAID, The Nature
Conservancy, and other international donors). This marked
a major shift in attempts to conserve hirola, as local Somali
communities were engaged, and indeed were the driving force
behind those efforts. In 2012, a predator-proof sanctuary was
constructed within Ishaqbini to serve as a source for future
reintroductions of hirola throughout their historical range, the
rst of which is slated for 2018. Interestingly, populations of
hirola have grown within Ishaqbini at the same time that they
remained stable or declined slightly within a protected area
to which they were introduced in the mid-1990s (Tsavo East
National Park—Probert etal. 2014; Ali etal. 2018).
Despite its successes, Ishaqbini is small (72 km2) and there-
fore insufcient for the long-term persistence of hirola. At
approximately the same time as the sanctuary was being cre-
ated, a coordinated, parallel effort (the Hirola Conservation
Programme; HCP) was established. The HCP’s mission is to
“regazette” Arawale National Reserve and restore hirola pop-
ulations throughout their 7,600 km2 historical range, which
would result in the downlisting of hirola from Critically
Endangered to Endangered. In the aftermath of rinderpest erad-
ication, hirola populations have been suppressed by a com-
bination rangeland deterioration via tree encroachment and
predation by African lions, cheetahs (Acinonyx jubatus), and
African wild dogs (Lycaon pictus), each of which is an IUCN
red-listed species (Ali etal. 2017, 2018).
Because predator control is arguably undesirable and lo-
gistically impossible, efforts have centered on identifying
strategies to revert Acacia-dominated woodlands to the open
grasslands that characterized the historical range of hirola.
Local communities support restoration strategies that have
been successful elsewhere, including manual clearing of trees
and reseeding of grasses over large areas (Ali 2016). Although
it is too early to know whether range-restoration efforts have
been successful, signs are encouraging because several of
Ostrom’s (1990) criteria for combating the Tragedy of the
Commons are being met: well-dened community bound-
aries exist within the hirola’s range, rules are being created
for the provision of restored grasslands to individuals, and
participatory decision-making is being overseen by a small
number of elders who have broad community support (Ali
2016). Mammal conservationists await results with cautious
The cases of hartebeest conservation in Laikipia and hirola
conservation in eastern Kenya represent place-based, system-
specic conservation efforts. Those efforts would not be pos-
sible without elements of conservation in protected areas and
human-occupied landscapes in tandem, and neither would be
possible without strong local leadership. Such is the norm for
mammal conservation throughout the globe. Although sev-
eral contributions over the past decade have been key in iden-
tifying common challenges associated with protected areas
(e.g., Woodroffe and Ginsberg 1998; Tranquilli etal. 2014;
Venter et al. 2017) and human-occupied landscapes (e.g.,
Walpole and Thouless 2005; Chapron etal. 2014), workable
solutions often are difcult to apply generally. In the pre-
vious examples, each conservation solution arose because
of particularities inherent to the study system. In the rst,
secondary prey of African lions (hartebeest) are not attracted
to grazing lawns to the same degree as are primary prey of
African lions (zebra), thereby offering potential for physical
separation and spatial refugia. In the second, Somali pasto-
ralists typically do not hunt wild ungulates, and ascribe to
hirola a near-mythical status (because hirola are associated
with high-quality, abundant grasses Kimitei et al. 2015).
Such is the nuance of successful conservation efforts, for
which we expect both protected areas and human-occupied
landscapes to play key roles in the 21st century and beyond.
Patterns of use and exploitation.—Humans have been hunting
other mammals throughout our species’ existence (Bunn and
Gurtov 2014). The primary motivation for hunting always has
been for food, and this remains among the most common justi-
cations for hunting, whether it be for bushmeat by aboriginal
peoples in tropical areas of Asia (Harrison etal. 2016), Africa
(Fa etal. 1995), Australia (Finch etal. 2014), or South America
(Hames and Vickers 1982), or for highly regulated hunting in
North America and Europe (Responsive Management 2013).
Culling is another term for hunting, but culling is most often
used when removals are motivated to reduce population size to
BOWYER ET AL.—CONSERVATION OF MAMMALS 933
lessen conicts with agriculture or to control predators (Quirós-
Fernández etal. 2017).
When tied to economic incentives and commercial markets,
however, hunting has led to excessive harvests and declines in
wild mammal populations (Geist 1988; Harrison etal. 2016).
In some instances, this has been the consequence of market
hunting, for example, in the bushmeat trade in Asia (Scheffers
etal. 2012; Harrison et al. 2016) and Africa (Fa etal. 1995)
or early exploitation of elk (Cervus elaphus) in North America
(Geist 1988). During the 18th and 19th centuries, beavers
(Castor canadensis and C. ber) and otters (Endydra lutris,
Lontra canadensis, and Lutra lutra) were extirpated from most
of North America and Europe because of a lack of regulations
and high economic demand for fur (Ray 1974; Estes 2016).
In North America, deliberate exploitation was used to elimi-
nate bison (Bison bison) from the Great Plains during warfare
by the United States government against aboriginal people who
depended on bison for food (Allen 1954). Likewise, European
bison (B. bonasus), deer (C. elaphus and Capreolus capreo-
lus), and beavers (C.ber) were severely overharvested across
vast regions of Europe. Large carnivores have been especially
persecuted, including wolves (Canis lupus), large cats (P. leo
and Puma concolor), and bears (Ursus spp.) that were killed to
prevent livestock depredation (Ripple etal. 2014).
Declining populations and restorations.—Concerns about
declining wildlife populations in the late 19th century prompted
Theodore Roosevelt and an elite group of hunters and nature
enthusiasts to form the Boone and Crockett Club to initiate
major conservation initiatives (Organ et al. 2010; Krausman
and Bleich 2013). Regulations limiting hunting, with restricted
seasons and quotas, were implemented in many areas early in
the early 20th century, and reintroduction programs were con-
ducted to restore depleted or extirpated populations of a num-
ber of species, including bighorn sheep (Bleich etal. 2018),
elk, beavers, and otters (Geist 1995). More recently, large car-
nivores have returned to parts of their former range even though
they are still persecuted in many areas (Ripple etal. 2014).
Generally, however, these conservation measures were highly
effective for a number of mammal species, and today, in North
America and Europe, wild mammal populations are thriving
to the extent that legal hunting is allowed for a number of spe-
cies (Organ etal. 2010). Indeed, hunters in North America have
contributed billions of dollars toward conservation and wildlife
management (Southwick and Allen 2010).
Patterns of management.—In developing countries, manage-
ment of hunting varies substantially because of societal views
and economics. Artiodactyls threatened with extinction, for
example, occur most often in poor countries with unregulated
hunting (Price and Gittleman 2007). Hunting of wildlife is
banned in several countries, notably China and India, although
illegal hunting (i.e., poaching) is known to be a concern in most
places to various degrees and has been exceedingly challenging
for the black rhinoceros, white rhinoceros (Ceratotherium
simum), three species of Asian rhinoceros (Rhinocerus sondia-
cus, R.unicornis, Dicerorhinus sumatrensis), and African ele-
phant. In some countries, including Namibia and South Africa,
trophy hunting generates substantial revenues for local commu-
nities and for conservation, and justies maintaining and restor-
ing native vegetation with benets for biodiversity protection
(Lindsey etal. 2007).
The legal killing of a telemetered male African lion near
Hwange National Park in Zimbabwe prompted huge public
media discussions about the ethics of trophy hunting (Di Minin
etal. 2016; Macdonald etal. 2016). Although still controversial,
a report by the IUCN (2016) details circumstances when tro-
phy hunting can be an effective tool for conservation and offers
guidelines. The IUCN report details many examples including
the local-community management of trophy hunting to support
conservation of the Suleiman markhor (Capra falconeri) and
the Afgan urial (Ovis orientalis) in Pakistan (Woodford etal.
2004) and Tajikistan.
Sustainability of harvest.—In most places where hunting is
legal, it is regulated to ensure that harvests are sustainable, but
sustainable harvests may not ensure hunting in some instances
(Fig. 3). Hunting is sustainable, in part, because of density-
dependent survival or reproduction (Boyce etal. 1999; Kokko
2001; Bowyer et al. 2014). Hunting reduces population den-
sity, resulting in increased food availability per capita, with
consequent enhanced nutrition increasing survival or fecun-
dity, thereby compensating for animals removed by harvest
(Owen-Smith 2006). These density-dependent responses to
exploitation include increased survival, especially of juveniles
(Eberhardt 2002; Bonenfant etal. 2009), and increased growth
rates (Schmidt etal. 2007; Gamelon etal. 2017; Monteith etal.
2018). Also, life-history responses inuencing reproduction
can include larger litter sizes and reproductive output (Bowman
etal. 1999; Hanson etal. 2009; Gamelon etal. 2017), higher
pregnancy rates (Stewart etal. 2005), and earlier age at rst
reproduction (Boyce 1981). Density dependence facilitates
Fig. 3.—Grizzly bear (Ursus arctos) in central British Columbia,
Canada. Trophy hunting for grizzly bears in British Columbia was
closed in 2017 because of public opposition to hunting even though
research indicated that the hunt was sustainable. Photograph by Mark
934 JOURNAL OF MAMMALOGY
resilience to harvest removals and promotes persistence of
hunted populations of large mammals.
The hydra effect.—Although hunter harvest or culling typi-
cally reduces population size, density-dependent responses in
seasonal environments can interact with hunter harvest in a way
that can actually increase annual survival (Boyce etal. 1999)
and total population size (Jonzén and Lundberg 1999; Abrams
2009). Such overcompensation responses to hunting removals
(where harvests result in increased survival or increased popu-
lation size) have been termed the “hydra effect” (Abrams and
Matsuda 2005). Avariety of mechanisms can cause the hydra
effect in multispecies systems (Cortez and Abrams 2016).
Likewise, culling of predators resulting in trophic-level inter-
actions can create counterintuitive results (Mitchell etal. 2015;
Costa etal. 2017). For example, in a tri-trophic system on Little
Barrier Island, New Zealand, culling of feral cats (F. catus)
resulted in increased abundance of Pacic rats (Rattus exulans)
that caused Cook’s petrel (Pterodroma cookii) populations to
decline (Rayner etal. 2007). In a similar example, harbor seals
(Phoca vitulina) were culled to enhance the shery, but seal
removals allowed Pacic hake (Meluccius productus) popula-
tions to increase, which subsequently caused a decrease in the
abundance of Pacic herring (Clupea pallasii—Bowen and
Lidgard 2013). Finally, removal of dingos (Canis lupus dingo)
in Australia has resulted in population increases by red fox
(Vulpes vulpes) with resulting predation that caused a reduction
in native marsupials (Letnic and Koch 2010).
Behavioral outcomes from hunting.—Hunting also can have
complex consequences for social structure and behavior of indi-
viduals in harvested populations. In African lions and brown
bears (Ursus arctos), removal of an adult male by hunting can
result in sexually selected infanticide (Packer and Pusey 1983;
Swenson etal. 1997; Whitman etal. 2004). After a mature male
has been removed by hunting, subordinate males will move into
the area vacated by the removal, and will kill young to bring
females into estrus so that they can breed with the females in
their new home range (Packer etal. 2011).
Other behavioral consequences of hunting, such as dis-
ruption of social structure in African elephant populations
(Shannon etal. 2013) and elimination of individuals with “bold
personalities” in elk (Ciuti etal. 2012), have been documented.
Vigilance behavior of black-tailed prairie dogs (Cynomys leu-
dovicianus) is disturbed by hunting, which results in reduced
body condition and tness (Pauli and Buskirk 2007). Exposure
to hunting can inuence learning whereby older animals adopt
behaviors to reduce the chances that they will be killed by
hunters (Thurfjell etal. 2017). In this example, young North
American elk were more vulnerable to hunter harvest, result-
ing in high turnover among younger individuals in the popu-
lation. Nonetheless, because elk learn to avoid hunters (Lone
etal. 2015), by the time that elk are 9–10years old, it is highly
unlikely that they will be killed by a hunter. This outcome is
important because older individuals know migration routes,
seasonal foraging areas, and calving sites, and retaining older
individuals in the population can potentially enhance popula-
tion resilience (Thurfjell etal. 2017).
Effects of selective harvests.—Clearly, hunting can be selec-
tive, if certain individuals are more vulnerable to harvest than
others (Festa-Bianchet 2017). Yet, hunters usually are not
highly selective, often taking the rst opportunity to kill a legal
animal (Heffelnger 2018) and uncertainty remains about the
possibility that selective hunter harvests can have detrimental
effects (Mysterud 2011). An evolutionary response requires
intense selection on highly heritable traits for an extended
period (Coulson et al. 2018; Festa-Bianchet and Mysterud
2018). Clearly, hunting can alter the demography of a popula-
tion, thereby increasing population turnover, changing sex and
age ratios, or both (Milner etal. 2007; Monteith et al. 2013,
2018; Hewitt etal. 2014). Moreover, a small amount of immi-
gration from an unhunted area can be sufcient to swamp the
effect of size-selection by hunting (Tenhumberg etal. 2004).
In landscapes with private land ownership, variation in hunter
access can be sufcient to ensure that effects of hunting are
overwhelmed by spatial patterns on the landscape, because
nearby areas exist with fewer hunters and more ungulates
Conservation.—Unregulated or illegal hunting of wild mam-
mals by humans continues to be a threat to the conservation
of some mammals, especially in tropical regions with diverse
mammal faunas (Van Vliet etal. 2015). Yet, in North America,
Europe, and parts of Africa, hunting has been the basis for
effective programs to ensure conservation of essential habitats
and to restore wild populations (Organ etal. 2010). Resilience
of populations to hunting driven by density dependence has
proven to be a powerful stabilizing force that ensures sustain-
able harvests if hunting is managed. Unfortunately, unregulated
or illegal hunting continues for cultural or economic reasons
over much of the Earth. Finding ways to overcome such cultural
and economic barriers to conservation remains among those
obstacles preventing continued loss of mammalian diversity.
C A F R
Climate change.—A rapidly changing climate with accelerat-
ing risks of extinction for mammals as well as other taxonomic
groups (Urban 2015) augurs poorly for the continued existence
of many species. Life-history traits that make mammalian spe-
cies vulnerable to extinction also make them more susceptible
to a changing climate (Davidson etal. 2017). Moreover, geo-
graphical areas associated with risk of extinction for mammals
may be modied under a rapidly changing climate as a result
of altered landscapes or from the decoupling of phenology
and life-history events (Davidson etal. 2017). Indeed, threats
from climate change are prevalent for both marine and terres-
trial mammals (Simmonds and Isaac 2007; Mallory and Boyce
Large-scale climatic variability has negatively inuenced
growth, development, fecundity, and demographic trends
in northern ungulates (Post and Stenseth 1999). In addi-
tion, increases in rain or snow events may promote icing
and adversely affect cold-adapted mammals at high latitudes
(Berger etal. 2018). Changes in snow conditions inuencing
BOWYER ET AL.—CONSERVATION OF MAMMALS 935
subnivian spaces may adversely affect small mammals that rely
on those seasonal refugia (Pauli etal. 2013).
A warming climate can affect species composition and
quality of forage plants available to large herbivores (Lenart
etal. 2002). This holds demographic consequences for some
of those mammals (Burthe etal. 2011), although not all large
herbivores respond immediately to such changes (Bowyer etal.
1998). Shifts in forage phenology also may inuence diets and
behaviors for large omnivores, with unknown consequences for
the community and ecosystem (Deacy etal. 2017). Recruitment
in moose (Alces alces) near the southern extent of their range
was negatively affected by warm temperatures (Monteith etal.
2015). In addition, climate change has affected long-term pop-
ulation growth of pronghorn populations (Gedir etal. 2015).
Afree-ranging population of Soay sheep (Ovis aries) experi-
enced reduced twinning and size of neonates, as well as delayed
sexual maturation during warmer winters (Forchhammer etal.
2001). Large mammals, as well as those with nocturnal activity
patterns, were more likely to respond as expected to climate
change than were other species with regard to local extirpa-
tions, decreased abundance, range contractions and shifts, and
morphological and genetic changes (McCain and King 2014).
Only 52% of all mammalian species studied by McCain and
King (2014) responded as expected to climate change. In ad-
dition, climate change may require a rethinking of how public
lands are managed for native ungulates to lessen effects of
competition with domestic and feral herbivores (Beschita etal.
2013). Linkages between climate change and population de-
mography of mammals, including latitudinal and elevational
shifts, require additional study (Moritz et al. 2008; Meserve
etal. 2011; Baltensperger etal. 2017). Future research should
evolve from descriptions of shifting patterns to investigations
of consequences resulting from a changing climate. Species
respond to thermal landscapes in complex ways that require
information about considerably more than ambient tempera-
tures (Bowyer and Kie 2009; Long etal. 2014). For instance,
northward range shifts in the distribution of snowshoe hares
(Lepus americanus) were related to decreasing persistence of
snow cover, which created a color mismatch, and ostensibly
lead to increased mortality of white hares on a brown landscape
(Sultaire etal. 2016; Zimova etal. 2016). Similarly, Atmeh etal.
(2018) reported that a white morph of weasels (Mustela nivalis)
was declining compared with their brown counterparts because
of a shortening of days with snow cover, likely from increased
predation. Much remains to be discovered about how and when
a changing climate will affect mammals. For many species, fac-
tors that threaten them cannot be eliminated entirely, and those
species will need to be managed to ensure their persistence.
Such species are “conservation reliant” (Goble etal. 2012), and
long-term conservation of many if not most wild mammals will
require more intensive efforts.
All authors made major contributions to this paper and order of
authorship is alphabetical. We thank M.R. Willig, A.V. Linzey,
and E.J. Heske for overseeing this process, V.C. Bleich for his
thoughtful review of our manuscript, and R.Ávila Flores for
preparing the Spanish summary. The phylogeny in Fig. 1 was
modied from one created by J.Kenagy and provided courtesy
of the Burke Museum of Natural History and Culture.
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