Essays
Effect of Hunting in Source-Sink Systems
in the Neotropics
ANDRES J. NOVARO,*§ KENT H. REDFORD,† AND RICHARD E. BODMER‡
*Department of Wildlife Ecology and Conservation, 303 Newins Ziegler Hall, University of Florida, Gainesville, FL
32611, U.S.A.
†Wildlife Conservation Society, 2300 Southern Boulevard, Bronx, NY 10460, U.S.A.
‡Center for Latin American Studies, 319 Grinter Hall, Gainesville, FL 32611, U.S.A.
Abstract: Previous studies of the sustainability of wildlife hunting in the Neotropics have not considered the
potential dispersal of animals into hunted areas. A literature review of studies of subsistence hunting in the
Neotropics suggests that hunting is often conducted in areas adjacent to relatively undisturbed habitat that
may act as sources of animals for the hunted sites. We compared studies of tapir ( Tapirus terrestris) hunting
at different sites to illustrate the potential bias of sustainability evaluations based on local productivity. The
limited information available suggests that dispersal could have a key role in rebuilding animal populations
depleted by hunting. Thus, factors that strongly affect dispersal—such as spatial distribution and size of areas
with and without hunting, population size in source areas, and social behavior—should be considered when
the sustainability of hunting is evaluated in areas with heterogeneous hunting pressure. We suggest the application of two models that use spatial controls (recognizing the potential source-sink nature of some hunted
systems and protecting unhunted refugia) to avoid wildlife overexploitation when biological data and enforcement capabilities to regulate harvests are limited. This approach may produce more reliable evaluations
of sustainability, provide information on the dynamics of hunting systems, and help local communities and
policymakers conserve key areas (including protected areas) that may act as game sources.
Efectas de la Cacería en Sistemas Fuente-Sumidero del Neotrópico
Resumen: Las evaluaciones previas de sustentabilidad de la cacería de fauna en el neotrópico no han considerado la dispersión potencial de animales hacia las áreas de caza. Una revisión bibliográfica de estudios
de caza de subsistencia sugiere que ésta se efectúa frecuentemente en áreas adyacentes a hábitats poco perturbados, las que podrían ser fuentes de animales para los sitios con caza. Comparamos estudios sobre caza
de tapires ( Tapirus terrestris) en diferentes sitios, para ilustrar el sesgo potencial de las evaluaciones de sustentabilidad basadas en la producción animal local. La escasa información disponible sugiere que la dispersión tiene un papel crucial en la recuperación de poblaciones de animales con abundancia reducida por la
caza. Por lo tanto, se deberían considerar los factores que influyen fuertemente en la dispersión (estructura
espacial y tamaño de áreas con y sin caza, tamaño poblacional en áreas fuente y comportamiento social) al
evaluar la sustentabilidad en áreas con presión de caza heterogénea. Se sugiere la aplicación de dos modelos
que usan controles espaciales (reconociendo la dinámica fuente-sumidero en algunos sistemas de caza y protegiendo refugios sin caza) para evitar la sobreexplotación de fauna cuando hay escasez de información biológica y de mecanismos de regulación de cosecha. Este enfoque produciría evaluaciones de sustentabilidad
más confiables e información sobre la dinámica de sistemas de caza y ayudaría a las comunidades locales y
a quienes elaboran las políticas a conservar áreas clave (incluyendo áreas protegidas) que podrían servir de
fuentes de fauna.
§ Current address: Laboratorio Ecotono, Universidad Nacional de Comahue, Unidad Postal Universidad, Bariloche, 8400 Río Negro, Argentina,
email [email protected]
Paper submitted September 8, 1998; revised manuscript accepted September 22, 1999.
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Volume 14, No. 3, June 2000
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Hunting in Source-Sink Systems
Introduction
Subsistence and commercial hunting are important to
the livelihood of rural people throughout Central and
South America (Redford & Robinson 1991; Bodmer et al.
1994). Hunting activities also have profound, direct effects on wildlife populations and indirect effects on ecosystems (Redford 1992). Thus, evaluating the sustainability of hunting is a key step to ensuring both the
livelihoods of rural people and the conservation of wildlife populations.
Many attempts have been made in recent years to
measure sustainability of hunting in tropical areas of the
Americas (reviewed by Robinson & Redford 1994a and
Robinson & Bodmer 1999). These attempts have included evaluations of population trends of hunted species, comparisons of age structures and hunting yields
across space and time (which are limited to measuring
relative levels of sustainability), and the use of sustainability models. Models developed to test sustainability
estimate population production with data on densities
and reproductive rates (Bodmer 1994) or estimate maximum population growth with data on reproduction
(Robinson & Redford 1991; Alvard et al. 1997) and survival (Slade et al. 1998). Estimations of population production or growth are then compared to actual harvest
rates to determine if harvest is sustainable.
Sustainability models have been applied with the implicit assumption that the majority of animal production
that is harvested by hunters is derived from reproduction within the catchment area (territory where hunting
is conducted). The reasons for making this assumption
are twofold. First, researchers may assume that hunter
mobility is high and that catchment areas are large in relation to the average dispersal distances of most game
species; however, hunters who live in permanent villages in tropical regions of the Americas and hunt on
foot usually do not travel much farther than 10 km to
hunt (Stearman 1990, 1992; Vickers 1991; Townsend
1995; Alvard et al. 1997). There are exceptions to this
rule among some groups that make long hunting treks
(Hill & Hawkes 1983; Werner 1983). Second, the role of
dispersal into catchment areas has been ignored, perhaps because of lack of information on dispersal rates of
game species in the Neotropics (Redford & Eisenberg
1992; Eisenberg 1999) and/or because of concentrated
attention by researchers on processes that occur at small
spatial and temporal scales (Wiens 1989).
If hunted areas are surrounded by or adjacent to areas
with abundant game populations and/or relatively undisturbed habitat, dispersal of game from these areas into
the catchment area can occur. Dispersal can contribute
significantly to recruitment of animal populations that
are intensively hunted on sites adjacent to unharvested
areas, as has been documented for furbearer populations
in North America (Allen & Sargeant 1993; Slough & Mowat
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Novaro et al.
1996) and Patagonia, Argentina (Novaro 1997). Net animal movement from unhunted to hunted areas is likely if
densities of the undisturbed populations are near carrying capacity (source areas, sensu Pulliam 1988) and if
densities on hunted areas are greatly reduced below carrying capacity, which is common in Neotropical forests
where hunting occurs (Redford 1992). Thus, dispersal
from source areas—and not just local reproduction in
hunted or sink areas (Pulliam 1988)—should be considered when evaluating the sustainability of hunting.
Furthermore, if unhunted areas are typically adjacent
to hunted areas, game dispersal into hunted areas could
be a widespread process, and evaluations of sustainability conducted to date could be seriously biased (but see
Hill & Padwe 1999). This bias would be due to the assumption that most hunted animals are offspring of resident populations when they are actually dispersers from
nearby sources. If this were the case, protecting nearby
sources ( Joshi & Gadgil 1991; McCullough 1996) could
be as important in achieving sustainable hunting as regulating the size of the harvest.
We sought (1) to determine if hunted areas in most
studied Neotropical sites are adjacent to potential source
areas for game populations and thus if game dispersal
from sources into hunted areas could be common in the
Neotropics; (2) to present an example of the potential
role of adjacent unhunted areas in the sustainability of
hunting by comparing studies of lowland tapir (Tapirus
terrestris) in different sites; (3) to suggest which ecological patterns and processes should be considered when
evaluating sustainability of hunting in areas with heterogeneous hunting pressure and to analyze alternative,
preliminary models to evaluate sustainability; and (4) to
propose an approach to conservation efforts that considers the potential role of protected areas as unhunted
sources of game.
Hunted Areas and Potential Sources
We conducted a review of studies of wildlife hunting by
indigenous people and colonists throughout the Neotropics to determine if hunted areas are commonly adjacent to potential source areas. We defined potential
sources as areas adjacent to hunted (catchment) areas
that have low human population densities, low levels of
or no hunting, relatively undisturbed habitat similar to
the catchment area, and area similar to or larger than
that of the catchment area. We assumed that areas with
these characteristics would support game populations
near carrying capacity and thus would be potential
sources of game. We reviewed vegetation maps of studied areas and consulted with authors to determine the
presence of potential game sources.
Our review indicated that most (16 of 21) of the intensively hunted areas are surrounded by or adjacent to po-
Novaro et al.
Hunting in Source-Sink Systems
tential source areas for game populations (Table 1). In
study sites in Venezuela and Peru, the potential sources
are protected areas, whereas in most other sites potential sources are uninhabited or lightly hunted areas.
We conclude that dispersal of game from sources into
hunted areas could be prevalent throughout the Neotropics. Thus, evaluation of the sustainability of hunting
must be conducted at a landscape scale large enough to
incorporate hunted areas and adjacent potential sources.
If studies of hunting sustainability include only the
hunted areas, their results could be biased. The following example illustrates this problem.
Tapir Hunting and the Effect of Unhunted Areas
Because of its low density and low reproductive rate,
the lowland tapir may be one of the species most
susceptible to overharvest in the Neotropics (Bodmer
1995a). Also, it is the largest terrestrial mammal in the
region and is a preferred target of hunters (Redford &
Robinson 1991). Thus, tapir hunting is a good model for
evaluating the potential effect of game dispersal from areas adjacent to hunted areas on the sustainability of
hunting and the appropriateness of models.
Table 1.
Several recent studies have estimated the sustainability of tapir hunting in different Neotropical sites (Robinson
& Redford 1991, with data from Stearman 1990, 1992;
Vickers 1991; Bodmer 1994; Townsend 1995; Alvard et
al. 1997) (Table 2). Vickers concluded that tapir hunting
was sustainable by comparing harvests per unit of effort
among recent years. The other authors applied Robinson and Redford’s (1991) and Bodmer’s (1994) models
and concluded that the numbers of tapirs removed by
hunters were much larger than those that would allow a
sustainable harvest (e.g., twice as much each year at Alvard et al.’s site and 140% of local annual reproduction
at Bodmer’s site). Under these harvest levels, tapirs
should have been locally extirpated within a few years
unless local production was grossly underestimated or
there was another source of recruitment (e.g., dispersal). By the end of the studies, however, tapirs were
still present at all but the site studied by Stearman (Table
2), and hunting of tapirs had been maintained for several
years (up to 20–30 at Alvard et al.’s site), with tapirs still
providing a significant proportion of the bushmeat.
We suggest that the main reason the predictions of the
models did not fit the observations on tapir hunting (except at the site studied by Stearman) was the presence
of nearby unhunted or lightly hunted areas (Table 2).
Vickers reported that most hunting was done in a “core”
Neotropical sites where the effect of hunting by indigenous people or colonists on wildlife populations has been evaluated.
Site (name of people)
Indigenous people
Shushufindi (Siona-Secoya)
Yomiwato (Machiguenga)
Diamante (Piro)
Taren Baurú (Pemón)
Ibiato (Siriono)
Yuqui camp (Yuqui)
Pimental Barbosa (Xavante)
Mbaracayu (Ache)
Colonists
Tahuayo
Lago da Fortuna
Jaraqui
Riozinho
Sao Domingos
Ponta da Castanha
Guatopo
Rancho Grande
El Jaguar
La Urbana
Rio Grande
Paraiso
Pipeline Rd.
Location
A “large,a” slightly,
or nonhunted area
adjacent
(name)b
Reference
northeastern Ecuador
southeastern Peru
southeastern Peru
eastern Venezuela
eastern Bolivia
central Bolivia
central Brazil
eastern Paraguay
yes (section of reserve)
yes (Manu)
yes (Manu)
yes (Canaima)
yes (private ranches)
no
yes (section of reserve)
yes (section of reserve)
Vickers 1991
Alvard et al. 1997
Alvard et al. 1997
Silva & Strahl 1991
Townsend 1995
Stearman 1990, 1992
Leeuwenberg 1993a, 1993b
Hill & Padwe 1999
northeastern Peru
western Brazil
western Brazil
western Brazil
western Brazil
western Brazil
northern Venezuela
northern Venezuela
northern Venezuela
central Venezuela
southern Belize
Panama
Panama
yes (Yavari)
no
yesc
yesc
yesc
yes (Acaituba)
yes (Guatopo)
yes (Pittier)
no
yes (Caura)
yesc
no
no
Bodmer et al. 1994
Peres 1990
Peres 1990
Peres 1990
Peres 1990
Johns 1986
Silva & Strahl 1991
Silva & Strahl 1991
Silva & Strahl 1991
Silva & Strahl 1991
Fragoso 1991
Glanz 1991
Glanz 1991
a
“Large” is an area at least the size of the catchment area.
Name of area with reduced hunting or with no hunting.
Name of area not available.
b
c
715
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Hunting in Source-Sink Systems
Table 2.
Novaro et al.
Studies of sustainability of hunting of lowland tapirs in the Neotropics.
Location
Northern Ecuador
Central Bolivia
Southeastern Peru
Northeastern Peru
Northeastern Bolivia
Tapir hunting
sustainablea
Tapirs present at
end of study
Potential
source-area
adjacentb
Reference
yes
no
no
no
no
yes
no
yes
yes
yes
yes
no
yes
yes
yes
Vickers 1991
Stearman 1990, 1992
Alvard et al. 1997
Bodmer 1994
Townsend 1995
a
b
As estimated in the studies of hunting sustainability cited.
A large (at least the size of the catchment area), slightly, or nonhunted area with habitat similar to that of the catchment area.
area around the village, that in surrounding areas (of size
similar to that of the core area) hunting was done less
frequently (12% of hunting days), and that both areas
were surrounded by a third, only occasionally hunted
area 17% larger than both other areas combined. Similarly, the areas studied by Bodmer, Townsend, and Alvard et al. were all adjacent to large areas that were subject to low or no hunting and that might act as sources
of tapirs for the hunted areas. Alvard’s sites were located
inside and next to Manu National Park, Bodmer’s site
was located within a large uninhabited portion of Departamento Loreto in the Peruvian Amazon, and Townsend’s
site was surrounded by large cattle ranches where the
habitat was relatively undisturbed and hunting pressure
was low (Townsend 1995). Conversely, the conclusions
of Robinson and Redford (1991) about tapir hunting by
Yuqui people apparently were correct, because tapirs
disappeared from the catchment area a few years after
Stearman’s first study. By then, however, there had been
extensive colonist settlement in the areas surrounding
the Yuqui territory (Robinson & Redford 1994b), and
they had probably overharvested or disrupted game
movement from potential source populations.
Thus, the spatial context in which catchment areas
are located, and particularly the presence of unhunted
areas, could be significant to the persistence of tapirs at
hunted sites and to the sustainability of tapir hunting.
The adjacent unhunted areas could help rebuild depleted populations by acting as sources of dispersing tapirs. There is, however, no information on dispersal
rates of any tapir species and only preliminary data on
movement patterns of lowland tapirs (Da Silva & Rodrigues 1999; Medice & Padua 1999). Thus, we can only
speculate about the contribution of dispersers to the
harvest within catchment areas. Demographic characteristics of tapirs (Bodmer 1995a) probably determine their
low dispersal ability compared to that of other Neotropical game species. The persistence of tapirs at sites
where potential sources of dispersers exist agrees with
the prediction that even low levels of dispersal may be
sufficient to maintain populations in sink areas (Pulliam
1988) and suggests that persistence is more likely for
species with higher dispersal rates.
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Volume 14, No. 3, June 2000
The Role of Dispersal
For unhunted areas to act as sources of game, animal
movement must occur from the unhunted to hunted areas. Numerous studies document the rebuilding of populations by dispersal in carnivores (Pyrah 1984; Knick
1990; Allen & Sargeant 1993; Slough & Mowat 1996),
small mammals (reviewed by Stenseth & Lidicker 1992,
but for an example where dispersal is absent see Boutin
et al. 1985), and birds (Ellison 1979; Myrberget 1985; Little et al. 1993). The role of dispersal may be reduced in
some primates by behavioral constraints (Ramirez 1984),
but the fast recovery after removal of other primate populations (Glander et al. 1984) and the overall high rates
of dispersal in primates (Pusey 1992) suggest that rebuilding could be important in some cases.
Little information is available about the dispersal of ungulates and other large mammals (Sinclair 1992), which
are usually the primary game species. Caughley and
Krebs (1983) reviewed the data available and suggested
that dispersal was not an important regulatory mechanism in large mammals, indicating that dispersal into
hunted areas may be uncommon. Dispersal rates, however, vary greatly among ungulates, principally according to their social organization (reviewed by McCullough
1985 and Sinclair 1992). Thus, dispersal could play an
important role in some ungulates (e.g., caribou [Rangifer tarandus]; Bergerud 1988]) and probably accelerated repopulation of much of the western United Sates
by elk (Cervus elaphus) from the Yellowstone-area source
population (Houston 1982). Dispersal was probably significant in the past for African elephants (Loxodonta africana) and white rhinoceros (Ceratotherium simum),
and their ability to repopulate depleted areas led to the
management practice of establishing fixed culling areas
to reduce habitat degradation in African parks (OwenSmith 1988).
The role of dispersal also may depend on the level of
disturbance in areas surrounding heavily hunted populations. Dispersal from populations that are near the carrying capacity (K ) of the environment (“saturation” dispersal) is more common than dispersal from populations
below K (“presaturation” dispersal; Lidicker 1975) among
Novaro et al.
carnivores, primates, ungulates, and noncycling small
mammals (Sinclair 1992). Only populations in completely unhunted or lightly hunted areas can be near K
(Robinson & Redford 1991). Thus, at least among mammals, dispersal may help rebuild hunted populations
only if surrounding areas are completely unhunted or
have low hunting pressure.
Dispersal distance is another key element determining
the role of dispersal. Within-population (intrademic) dispersal is more common than between-population (interdemic) dispersal in all animals studied (Bekoff 1977;
Dobson 1982; Packer 1985; Sinclair 1992; Stenseth &
Lidicker 1992). Furthermore, dispersal frequency is usually a decreasing function of distance from the natal area
(Wolfenbarger 1949; Kitching 1971). Thus, dispersal is
more likely to rebuild hunted populations near unhunted areas than populations farther from unhunted areas. The magnitude of these dispersal distances will vary
greatly among species.
Perhaps most important, the role of dispersal also is affected by landscape structure. Landscape structure is determined by the physiognomy (or spatial arrangement of
habitat patches), composition (size and type of patches),
and connectivity (ease of animal movement) among habitat patches (Dunning et al. 1992; Taylor et al. 1993).
Therefore, the structure of the landscape is likely to
have varying effects on game species. Knowledge about
dispersal mechanisms of different species in different
landscapes is necessary to evaluate the role of unhunted
areas as sources of game animals for hunted sites in the
Neotropics.
The Need for New Models to Evaluate Sustainability
Evaluation of the sustainability of hunting needs to be
carried out with improved models that incorporate the
spatial complexity of hunted areas and the role of animal
movement across the landscape. The sustainability of
hunting depends upon demographic rates (survival and
fecundity) within hunted and adjacent unhunted sites
and the rates of game dispersal among sites. There are
two approaches to including the effect of game dispersal in evaluations of hunting sustainability. The first
approach estimates the dispersal rate from unhunted
populations, and the second approach involves a preliminary model using both the proportion of unhunted area
in the vicinity of the catchment area and basic demographic data to indicate which species may be hunted
sustainably at each site.
Game Productivity and Dispersal
If hunted sites are adjacent to unhunted ones, the dynamics of hunted and unhunted populations can be
modeled as that of sources and sinks (Lidicker 1975).
Hunting in Source-Sink Systems
717
Under the source-sink model (Pulliam 1988), the total
productivity (P) of a hunted area that acts as a sink population can be estimated as P ⫽ LP ⫹ DI (equation 1),
where LP is the rate of local animal production and DI is
the rate of dispersal from nearby sources into the sink.
The models proposed by Robinson and Redford (1991)
and Bodmer (1994) can be used to estimate LP within
hunted areas. The rate of dispersal can be estimated by
modifying an equation given by Pulliam (1988) for habitat patches of different quality, such that IM ⫽ Ksource
( ␭source ⫺ 1) (equation 2), where K and ␭ are the carrying capacity and the discrete rate of population increase
of the source, respectively. It is assumed that, because of
harvest in the adjacent sink, reproduction is higher than
mortality on population sources (thus they have ␭s ⬎ 1)
even though their densities may be at or near carrying
capacity (Pulliam 1988). Estimating K and ␭, however,
requires long-term studies to determine population size,
reproduction, and survival rates (Caughley 1977).
Alternative means of estimating dispersal rates between areas with different hunting pressure include radiotelemetry (Knick 1990; Slough & Mowat 1996; Novaro 1997) and tagging studies (Allen & Sargeant 1993).
Other methods are population-genetic studies (Little et
al. 1993), particularly those involving DNA probes (Dias
et al. 1996), that can significantly reduce time and effort
in the field. Knowledge of the dispersal patterns of the
main game species at each site could help determine the
appropriate spatial arrangement and size of source and
sink populations.
Finally, other demographic data can aid in the study of
dispersal and population productivity. In particular,
comparisons of age and sex composition of hunted and
unhunted populations can assist in determination of
whether saturation or presaturation dispersal (Lidicker
1975) predominates among the main game species. This
knowledge can influence decisions about the level of
protection that needs to be given to source areas. Comparisons of age and sex composition have been used
also to estimate productivity and impact of hunting on
different populations (Collett 1981; Bodmer 1995a). The
validity of these comparisons is limited, however, by differences in vulnerability to hunting between ages and
sexes and by density-dependent effects (Shaw 1985).
A Model of Game Sources or Refugia
Until more information is available on dispersal of game
populations, other approaches are needed to help with
decisions about the sustainability of hunting. Recently
Joshi and Gadgil (1991) and McCullough (1996) proposed methods to achieve high, sustainable harvests by
means of spatial controls. Spatial controls may be appropriate to achieve sustainability of hunting given limited
field information and an inability to regulate harvest in
hunted areas, which is common throughout the Neotrop-
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Hunting in Source-Sink Systems
ics and other regions. Joshi and Gadgil used a logistic
growth model to show that if a portion of the catchment
area is set aside in refugia and if dispersal among refugia
and hunted areas is intense (complete mixing), the total
harvest could reach the same maximum sustained yield
that would be attained in the absence of refugia. The
main advantage of this model is that, if the refugia are
large enough, the resource population is protected from
extinction regardless of the level of harvest effort ( Joshi
& Gadgil 1991). Until information on demographic rates
(including dispersal) and the effect of landscape structure on hunted systems is available, this model can be
used to help approximate the size of refugia or source
areas that are needed to minimize the probability of local extinction of animal populations.
According to Joshi and Gadgil’s discrete model, the
following equation reflects the minimum proportion of
the catchment area (␣) that needs to be set aside as refugia for a given species with a maximum discrete rate of
increase ␭max: ␣ ⬎ 1/␭max (equation 3). Protecting an ␣
proportion of the catchment area would prevent the decline of the game species even if all individuals were removed annually from the hunted areas ( Joshi & Gadgil
1991).
Joshi and Gadgil’s deterministic model may be modified by adding a term to account for the environmental
and demographic stochasticity that affect rates of increase (Brillinger 1986) and that may increase the size of
refugia needed. Temporal series of estimates of ␭max,
which are missing for all Neotropical game species, are
needed to assess environmental stochasticity of ␭max.
Small sample sizes used to estimate ␭max values for most
game species (Robinson & Redford 1991; Fa et al. 1995)
determine that standard deviations of ␭max estimates
may incorporate the effects of demographic stochasticity (Akcakaya 1991). Thus, standard deviations (SD) of
␭max estimates may provide a preliminary measure of
stochastic variation of ␭max. Assuming a normal distribution of ␭max, the model for game refugia is then ␣ ⬎ 1 /
( ␭max ⫹/⫺ 1.96 SD␭max) (equation 4) for a probability
level of 95%.
Furthermore, under the assumption of complete mixing ( Joshi & Gadgil 1991) and adding a term to account
for stochasticity, the following ␣ level could maximize
the harvest: ␣max ⫽ 2/[1 ⫹ ( ␭max ⫹/⫺ 1.96 SD␭max)]
(equation 5). If mixing is not complete, as is likely the
case for most species, the sustained yield would be
lower than the maximum. A large number of small refugia would allow better mixing (more dispersal) than a
few large ones (Gadgil & Vartak 1976). Other considerations need to be made regarding the size of refugia,
however, such as the minimum size of areas required by
species with larger home ranges and higher risk of population decline (Wright 1990; Pimm 1991).
McCullough (1996) suggested a system of trial and error to determine the optimum number, size, and ar-
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Novaro et al.
rangement of refugia which would maximize the harvest. This method involves monitoring the harvest through
time and making small adjustments to the size or location of the refugia. If the harvest increases, further adjustments can be made in the same direction. If the harvest declines, the adjustments must be reversed. Apart
from the constraints of area requirements of the more
vulnerable and mobile game species (McCullough 1996),
cultural and land-tenure constraints may limit the availability and location of refugia at most sites. Therefore,
a combination of the systems proposed by Joshi and
Gadgil (1991) and McCullough (1996) may be useful to
guide initial management and conservation decisions
and could approximate an adaptive management scheme
(Walters 1986). Joshi and Gadgil’s model (1991) can be
used to determine which species can be harvested with
reasonable safety given the size of refugia available or
that can be created. If there is enough flexibility in the
location and size of refugia and hunted areas, McCullough’s method (1996) can be used to attempt to
maximize the harvest, perhaps using equation 5 to estimate a starting point for the main game species. Experimentation with size and location of refugia, supported
by close monitoring of harvests and preferably of game
populations, could lead to increased knowledge about
the combined effects of hunting and landscape structure
on game population dynamics.
To provide an example, we used equation 4 to calculate the proportions of area in refugia needed to prevent
a population decline of the main game mammals in the
Peruvian Amazon (Table 3). In the case of the tapir, 63–
100% of the land needs to be protected to prevent a decline. A preliminary analysis by Bodmer (2000) suggests
that in the Reserva Comunal Tamshiyacu-Tahuayo, Peru,
the proportion of unhunted areas adjacent to lightly and
heavily hunted sites is approximately 48%. Thus, tapirs
should not be harvested in this area unless the size of
the harvest can be strictly controlled or more refugia
can be set aside. Species that can be harvested without
risking a regional population decline are collared peccaries (Tayassu tajacu) and two caviomorph rodents, because their ␣ ranges are smaller than 48%. All the monkeys and most ungulates are at risk of decline. These
conclusions agree broadly with demographic data obtained by Bodmer (1995b) and Bodmer et al. (1997),
which suggest that tapirs and most monkeys are overharvested, whereas the other species are not (Table 3).
Deer and small primates are not overharvested according to field data, but they are expected to be according
to the model (Table 3). This discrepancy could be
caused by low vulnerability of deer and low hunter preference for small primates (Bodmer 1995b), making harvests of these species sustainable in spite of the small
proportion of unhunted area. No species predicted by
the model to be harvested sustainably was actually overharvested according to field data (Table 3). Thus, Joshi
Novaro et al.
Hunting in Source-Sink Systems
719
Table 3. Minimum percentage of area required to be in refugia (␣) to prevent a population decline of the main game mammals
in the Peruvian Amazon.
Common name
White-lipped peccary
Collared peccary
Red brocket deer
Grey brocket deer
Lowland tapir
Black agouti
Green achouchy
Woolly monkey
Howler monkey
Red uakari monkey
Brown capuchin
White-fronted capuchin
Monk saki monkey
Titi monkey
Spider monkey
Squirrel monkey
Scientific name
␭maxa
␣b
Overharvested c
Tayassu pecari
Tayassu tajacu
Mazama americana
Mazama gouazoubira
Tapirus terrestris
Dasyprocta fuliginosa
Myoprocta pratti
Lagothrix lagothrica
Alouatta seniculus
Cacajao calvus
Cebus apella
Cebus albifrons
Pithecia monachus
Callicebus cupreus
Ateles paniscus
Saimiri sciureus & S. boliviensis
2.32
3.49
1.49
1.63
1.22
3.00
4.22
1.15
1.19
1.13
1.15
1.19
1.13
1.26
1.07
1.27
33–61
22–41
52–95
47–87
63–100
26–47
18–34
67–100
65–100
69–100
67–100
65–100
69–100
61–100
72–100
61–100
no
no
no
no
yes
no
no
yes
no
no
yes
yes
no
no
yes
yes
a
Maximum finite rate of increase calculated from rmax (maximum exponential rate of increase). Values obtained from Robinson and Redford
(1991) and Bodmer et al. (1997) with equation ␭ ⫽ er, where e (2.72) is the base of natural logs (Caughley 1977).
b
The ␣ range (%) estimated with equation 4 (see text). Because no estimates of standard deviations of ␭max were available (Bodmer et al.
1997), we assumed a 15% coefficient of variation of ␭max for all species. The ␣ upper range values that were larger than 100% were truncated
to 100%.
c
Overharvest was evaluated according to the magnitude of the change in abundance between persistently and infrequently hunted sites
(Bodmer 1995b; Bodmer et al. 1997).
and Gadgil’s model seems to provide a safe (conservative) first step in determining which species can be harvested sustainably according to the size of the unhunted
area available.
Conclusions
Management Applications
Evaluations of the sustainability of game hunting in the
Neotropics should consider whether adjacent unharvested populations exist, and, if they do, they should incorporate the habitat of these populations in the management plan of catchment areas. Current programs to
manage wildlife throughout the Neotropics should not
abandon attempts to regulate harvest composition and
size within the hunted sites. Rather, at sites where unhunted areas exist, wildlife managers should attempt to
regulate harvests and at the same time protect as much
unhunted area as possible. An adaptive management approach combining the methods proposed by McCullough
(1996) and Joshi and Gadgil (1991) may maximize harvests, reduce the risk of population decline, and provide
information on the effects of game dispersal and landscape on the sustainability of hunting. Whenever possible,
management decisions should be made that complement
the conclusions obtained by applying other sustainability models that estimate game productivity within catchment areas.
Sources of Game, Protected Areas, and
Community-Based Conservation
Many of the potential sources for game populations we
identify occur in protected areas. The need to preserve
these areas as potential sources of game has also been
advocated recently by other authors (Hill & Padwe
1999; Robinson & Bennett 1999), including some working outside the Neotropics (Fimbel et al. 1999). Unfortunately, local communities that hunt game for subsistence or commercial purposes view protected areas as
restricting their rights to use natural resources. This is a
consequence to a large extent of top-down systems of
protected areas designed by conservationists and governments with little or no involvement of local communities (Robinson & Redford 1994b). The issues of determining and allocating rights over the use and management
of such game populations, both protected source populations and harvested adjacent populations, are complicated and are just beginning to be examined (e.g.,
Naughton-Treves & Sanderson 1995). Clearly, resolution
of such issues is critical if the approaches we advocate
are to succeed.
The recognition of protected areas as vital sources of
game would help change the negative regard in which
many local people hold these areas. If local communities
recognize the value of setting aside unhunted areas as
sources for game populations, they could be actively involved in the protection of those areas. This protection
should involve both the populations of game and their
habitat. Furthermore, in areas where community-based
Conservation Biology
Volume 14, No. 3, June 2000
720
Hunting in Source-Sink Systems
conservation is implemented, the protection of unhunted source areas could help resolve some of the contradictions of the community-based approach (Robinson
& Redford 1994b; Redford et al. 1995). In particular, it
could help maintain the spatial heterogeneity of extractive reserves by setting aside unhunted areas where game
population densities and ecosystem diversity (Robinson
1993) are not reduced as a result of human use.
We propose combining the community-based approach to wildlife conservation with the understanding
that many hunted systems in the Neotropics may operate as source-sink systems. This combination should result in more sustainable uses of wildlife in the region and
should help in the conservation of biological diversity by
(1) increasing the number of fully protected areas; (2)
making protected areas socially acceptable among local
communities; and (3) making them economically viable
through active protection by those communities.
Acknowledgments
We thank J. G. Robinson, J. E. Rabinovich, L. C. Branch,
R. S. Walker, C. S. Holling, G. C. White, and three anonymous reviewers for helpful comments on the manuscript. A.J.N. was supported by a Compton fellowship at
the University of Florida and a postdoctoral fellowship
from Consejo Nacional de Investigaciones Científicas y
Técnicas of Argentina.
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