Download Estimating cost functions for the four large carnivores in Sweden

EC O L O G IC A L E C O N O M IC S 6 8 ( 2 0 08 ) 51 7 –5 24
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ANALYSIS
Estimating cost functions for the four large carnivores
in Sweden
Göran Bostedt⁎, Pontus Grahn 1
Department of Forest Economics, S-901 83 Umeå, Sweden
AR TIC LE I N FO
ABS TR ACT
Article history:
The Swedish carnivore policy goal for the four large carnivores – wolverine (Gulo gulo), wolf
Received 2 July 2007
(Canis lupus), brown bear (Ursus arctos) and lynx (Lynx lynx) – is to ensure a minimum viable
Received in revised form
population on a long-term basis. To reach this goal the policy restricts population regulation
20 February 2008
activities, like hunting (prohibited for wolverine and wolf and restricted for brown bear and
Accepted 7 May 2008
lynx) in Sweden. For owners of semi-domesticated (i.e. reindeer), and domesticated
Available online 21 June 2008
(livestock) animals this policy and the existence of individuals of these four species
results in externalities associated with predation.
Keywords:
This paper presents econometric estimates of the predation and the social costs for these
Carnivores
four species, based on ecological models of functional response. The data on costs is based
Predation
on compensation provided to livestock owners by the Swedish government. The paper also
Costs
applies these econometric estimates to predict the social cost per species when the
Economics
population goals of the Swedish carnivore policy are reached. Based on out our model the
wolverine and the lynx will impose the highest marginal, as well as total costs on society,
given the current policy goals. The wolf is an efficient predator, but due to its geographical
distribution in Sweden, its social costs are less than anticipated. The brown bear is largely
omnivorous, thus resulting in relatively low social costs.
© 2008 Elsevier B.V. All rights reserved.
1.
Introduction
The conservation of large carnivores entails significant costs.
Some costs are obvious, such as predation on domesticated, or
semi-domesticated, animals such as sheep (Ovis aries) or
reindeer (Rangifer tarandus) (Bjärvall and Franzén, 1990).
However, predation on wild ungulates can also represent a
social cost to the extent that these ungulates are attractive
hunting game. Further, carnivores are known to attack and kill
hunting dogs. Species with these features can be characterized
as both environmental ‘bads’ and ‘goods’, to the extent that
there also exists a willingness-to-pay to preserve them
(Bostedt, 1999). Apart from carnivores interesting examples
include wild pigs (Tisdell, 1982) and elephants (Bandara and
Tisdell, 2003, 2004).
The four so-called “large carnivores” in Sweden – wolverine
(Gulo gulo), wolf (Canis lupus), brown bear (Ursus arctos) and lynx
(Lynx lynx) – are native Swedish species and have been so since
the last Ice Age (Bjärvall and Ullström, 1995). The size of the
populations of these carnivores is however only known from
systematic surveys for the last 20–30 years.2 The general trend
today is that these species (with the exception of the lynx) are
⁎ Corresponding author. Tel.: +46 90 786 85 11; fax: +46 90 786 60 73.
E-mail addresses: [email protected] (G. Bostedt), [email protected] (P. Grahn).
1
Present address: S. Slevgränd 108, S-906 27 Umeå, Sweden.
2
Recent surveys can be accessed (in Swedish) at www.viltskadecenter.com.
0921-8009/$ – see front matter © 2008 Elsevier B.V. All rights reserved.
doi:10.1016/j.ecolecon.2008.05.008
518
EC O LO GIC A L E CO N O M ICS 6 8 ( 2 00 8 ) 5 1 7 –5 24
increasing in numbers, due in part to the current Swedish
carnivore policy (cf. Sections 2.1–2.4 below). This policy has
specific goals concerning the population number and distribution of each species, to ensure the populations above the
level of the respective minimum viable populations. The
current policy was adopted by the Swedish Parliament in the
year 2000 (Government Bill 2000/01:57, 2000). Former policies,
especially in the beginning of the 20th century, had quite
different goals, namely to exterminate the carnivore populations. For this, bounties were paid to those who killed any of
the four large carnivores.
The aim of this paper is to estimate social cost functions for
the four large carnivores in Sweden based on regression
models of cost and population numbers founded in ecological
theory. These estimates are then used to predict the costs to
society of achieving the Swedish carnivore policy goals. It
should be emphasized that due to lack of data this analysis
ignores the social costs to hunters associated with (1)
carnivore competition with humans for game and (2) attacks
on hunting dogs.
The number of publications on this topic is very limited.
The cost study by Boman (1995) is a predecessor to the one
presented in this paper, even though that data set was shorter.
Today, the estimates of carnivore populations have greatly
improved, particularly for the brown bear. Furthermore,
Boman et al. (2003) conducted a cost–benefit analysis for the
Swedish wolf population, which showed that the spatial
distribution of the wolf population is of great importance in
explaining costs and benefits. More recently, Skonhoft (2006)
analyzed the costs and benefits of wolf re-colonization in
Norway. Costs were measured in terms of predation on, and
thereby reduced availability of, game (mainly moose) for
hunting. Benefits were measured in terms of reduced browsing damage by these game species. Furthermore, two studies
of predation and cost efficiency in the conservation of African
wild dogs (Lyacon pictus) in South Africa and Kenya, by Lindsey
et al. (2005) and Woodroffe et al. (2005), provide an interesting
comparison with a non-Scandinavian carnivore (this comparison is provided below in the Results section).
Concerning cost studies of endangered species other than
carnivores, there is a dearth of empirical studies in refereed
journals, with the well-known exception of the spotted owl
study by Montgomery et al. (1994). In addition, there was a
cost–benefit study of the rare Australian mahogany glider
(Petaurus gracilis) by Tisdell et al. (2005). This is an interesting
fact, especially when contrasted with the wealth of valuation
studies directed at endangered species. Consequently we feel
that this social cost study fills a gap in the literature, since
both benefits and costs are needed to make informed
economic decisions concerning wildlife management.
There are a significant number of publications concerning
the ecology and biology of the four large carnivores in Sweden
and Scandinavia. This literature often deals with population
ecology, genetics and animal ethology. However, of interest
here is that there are a few ecological studies that present
estimates of the diet and amount of meat required by the
predators, for instance Dahle et al. (1998) and Landa et al.
(1999). Therefore we, when necessary, rely on estimates from
countries outside Scandinavia, for instance the North American studies by Sibly and Callow (1986) and Mech (1977).
We begin the next section by outlining basic biological facts
concerning the four large carnivores, with a focus on species'
behavior, demand for nutrition and energy, and the population and spatial distribution. In the Theory section we outline
the ecological framework, which provides the basis for our
empirical estimates. The final section contains concluding
remarks.
2.
Bioeconomic background
This section presents a short biological overview of the four
large carnivores in Sweden focusing particularly on the
biological characteristics, the species' capacity to impose
externalities on private livestock, and the general population
development in Sweden in the latest decades. This analysis
excludes the Golden Eagle (Aquila chrysaetos), which is sometimes is considered the fifth large carnivore in the Swedish
fauna.
2.1.
Wolf (C. lupus)
The wolf is an efficient carnivore that usually feeds on moose
(Alces alces) and other ungulates like reindeer (R. tarandus). The
fact that wolves form packs can make them very effective in
hunting large prey (Bjärvall and Ullström, 1995; Persson, 1996).
As of 2005, the wolf population numbers approximately 100
individuals in Sweden. If Norway, with whom Sweden shares
a common wolf population, is included the population
number becomes slightly higher, approximately 100–120
individuals. The (first stage) goal of the Swedish wolf policy
is that the future population should number 20 annual
rejuvenations, which is equal to around 200 individuals. The
official policy is, somewhat simplified, that the wolf should
not be established in the reindeer herding area in the north of
Sweden, due to the potential problems it might cause to
reindeer herders (Government Bill 2000/01:57, 2000) (Fig. 1).
2.2.
Brown bear (U. arctos)
The brown bear (U. arctos) is found mainly in the northern
parts of Sweden. The Scandinavian brown bear is an
Fig. 1 – The wolf population 1971–2004. All population data
are compiled from estimates by the County Administrative
Boards, the Saami Parliament, the Swedish EPA and the
Swedish Hunters' Association, cf. Section 4.
EC O L O G IC A L E C O N O M IC S 6 8 ( 2 0 08 ) 51 7 –5 24
omnivorous species that predominantly eats non-meat food
(Bjärvall and Ullström, 1995), such as blueberries, ants and
plants (Dahle et al., 1998). However, in the spring, following
hibernation, the bear is known to seek meat in order to meet
protein demands. At this time it is more likely to predate on
moose (A. alces) and reindeer (R. tarandus), particularly
since the calves of these species are also born in the spring
and provide easy prey. A trend shown by a number of studies
suggests that the brown bear prefer a larger amount of meat
with increasing latitude (Persson et al., 2001; Kaleckaya, 1973).
As of 2005, the Swedish brown bear population is estimated
at between 1600 and 2800 individuals. This represents a
seemingly rapid increase in the population between 2000 and
2005, yet it has been argued that this change is not realistic
and is likely due to the development of estimation techniques,
rather than ecologically favorable conditions (Solberg et al.,
2006). The reason is that the brown bear population until
recently has been very hard to estimate because it does not
leave tracks during the winter. Estimation techniques have
changed and developed over the years and new DNA
techniques have made it possible to estimate the population
size with a higher degree of accuracy (Solberg et al., 2006; NFS
2004:17; NFS 2004:18).
The Swedish carnivore policy goal for the brown bear
established that the population should be larger then 1000
individuals. As of 2005, this goal has been reached (Fig. 2).
2.3.
Wolverine (G. gulo)
The wolverine (G. gulo) mainly inhabits the mountain region in
Scandinavia, with some exceptions. This species is well
adapted to the climate conditions in this area, and is an
efficient carnivore with a diet mainly consisting of reindeer.
Its advantage over this prey is due to, among other things, its
very large feet, if the overall size and weight of the animal are
considered, which gives it a great advantage over reindeer on
snow covered ground. During summer, the diet is less known,
but is thought to include smaller prey and carcass (Landa et al.,
1997, 1999). The wolverine population was estimated to
contain around 420 individuals in 2005 according to the
Swedish Environmental Protection Agency. The (first stage)
goal of the Swedish carnivore policy for the wolverine is that
the population should number at least 90 rejuvenations
annually, which is thought to be equal to about 575 individuals
(Ericsson et al., 2007) (Fig. 3).
Fig. 2 – The brown bear population in Sweden 1971–2004.
519
Fig. 3 – The wolverine population in Sweden 1971–2004.
2.4.
Lynx (L. lynx)
The lynx is the only large wild cat found in Scandinavia. The
lynx has been shown to have a significant effect on the
population numbers of roe deer (Capreolus capreolus), an
important prey in some parts of Sweden (Liberg and Andren,
2005). Due to the high populations of roe deer the largest
density of lynx is found in the middle-part of Sweden.
However, the lynx is also found in the northern parts of
Sweden. As a carnivore in the southernmost parts of
Scandinavia, the lynx prefers roe deer and small prey species
like hares (Odden et al., 2006). Lynx in the northern part of
Scandinavia also predate on reindeer and smaller mammals
like mountain hare (Lepus timidus), as well as birds like
grouses (Lagopus lagopus and Lagopus mutus) (Pedersen et al.,
1999).
Due to historic government policies that alternatively
encouraged and prohibited hunting, the lynx population has
varied over the years. Today, when limited hunting has been
allowed since the mid 1990s, the population is stable or
somewhat decreasing. The goal of the Swedish lynx policy is
that the population should exceed 1500 individuals — a goal
which has been reached as of 2005 (Fig. 4).
3.
Theory
Theoretically defining and empirically estimating the relation
between the population of a certain carnivore and the social
Fig. 4 – The lynx population in Sweden 1971–2004.
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EC O LO GIC A L E CO N O M ICS 6 8 ( 2 00 8 ) 5 1 7 –5 24
costs it imposes in terms of predation on domesticated
animals involve at least two steps:
Step 1: Estimating the relation between the population
level of the carnivore in a specific area and the expected
number of domesticated animals of different sex and age
groups that may be lost to predation to this carnivore per
time period. This will be explained below in terms of functional response models.
This step involves several sub-steps including: disentangling the effect of different carnivore species that predate on
the same prey species; estimating the prey death rate —
referred to as the functional response; consideration of a
carnivore's preference for prey species when several different
choices of prey exist; and spatial concerns, such as the effect
of dispersal of prey by the carnivore.
Step 2: Estimating the welfare loss associated with
predation of a particular animal, given its species type, sex
and age group. This will be explained below in terms of opportunity costs.
Importantly, the welfare loss associated with predation
upon a domesticated animal, such as a sheep or a reindeer,
should not always be restricted to the meat value of the animal.
For example, if the killed individual is a valuable breeder then
the welfare loss should also include the lost opportunity to
pass on its genetic information to future offspring.
3.1.
Step 1 — functional response models
Conventionally, general first-order differential equations (i.e.
Lotka–Volterra type models3) are used to describe predator–prey
dynamics (cf. Gutierrez, 1996; Abrams and Ginzburg, 2000):
dN
¼ rðNÞdN gðN; PÞdP
dt
ð1Þ
dP
¼ f ½gðN; PÞ; PdP
dt
ð2Þ
Ecological theory usually distinguishes between three
types of functional responses (cf. Abrams and Ginzburg, 2000):
Prey dependent response : Na ¼ gðNÞdP
where Na is the number of prey killed by a population of
predators. Here the prey density alone determines the
functional response. Evidently, this response is linear in P (e.g.
Lotka, 1925; Volterra, 1931; Nicholson and Bailey, 1935; Hassel
and May, 1974). The simplest of these models also assume that
for a given predator density, prey mortality increases linearly
with prey density up to some saturation level where the
.
predator simply cannot kill more prey, i.e. g(N) =sN if g(N) b Nmax
a
Predator dependent response : Na ¼ gðN; PÞdP
Here both predator and prey populations affect the
response. These functional responses were first described by
Holling (1959). In these models g(N, P) is usually assumed to
be convex in N, so that ∂g(N, P) / ∂N N 0 and ∂2g(N, P) / ∂N2 b 0. In
the limit, as N → ∞, the number of prey killed approaches the
. Another alternative is to
saturation level, i.e. Na → Nmax
a
assume that the functional response has a sigmoidal shape
in N, so that ∂g(N, P) / ∂N N 0 and ∂2g(N, P) / ∂N2 N 0 if N b N⁎ and
∂2 g(N, P) / ∂N2 b 0 if N N N⁎. Again, in the limit, as N → ∞,
the number of prey killed approaches the saturation level,
. Since predator dependent response models are
i.e. Na → Nmax
a
very difficult to estimate empirically, we will concentrate on prey
dependent response models in the following, and will further
assume that prey density is below the predator saturation level.
Functional response models can easily be expanded to
include several predator or prey species. For instance, a prey
dependent response model can be adapted for two prey types:
Multispecies dependent response : Na1 ¼ gðs1 dN1 þ s2 dN2 ÞP
where Eq. (1) is the growth rate of the prey species, N, while
Eq. (2) is the corresponding growth rate for the predator species,
P. r(N) is the per capita birth rate of the prey species, which
decreases with N in most models. Furthermore, g(N, P) is the
predator per capita consumption rate of prey, while the function
f(·) represents the predators' numerical response. To proceed
with step 1 it is necessary to focus on the function g(N, P), which
is commonly known as the predator functional response. As
demonstrated in Eqs. (1) and (2) there is a feedback effect in the
sense that the prey population influences the growth rate of the
predator, through the numerical response. However, when
multiple prey species are available this relationship will appear
less interactive, meaning that the predator will not respond
numerically to variations in the population of a particular prey
species. Furthermore, both prey and predator densities are
heavily influenced by current management regimes. The
numerical response is therefore in the following neglected and
the sizes of the predator populations are determined outside the
model.
where the lower case indices 1 and 2 indicate the two different
prey species, and s1 and s2 are constants which indicate prey
preference by the predator. For statistical purposes the
equation above shows how the population of one prey species
can contribute to explaining social costs on another prey
species. It is important to note that predators are more likely
to exhibit prey preference when food is not limiting. Consequently, all preference models may fail under conditions of
extreme hunger. As noted by Abrams and Ginzburg (2000),
very few empirical studies of functional responses in natural
settings exist, despite considerable debate around these
responses. Notable exceptions are the studies of functional
responses of wolf on moose by Messier (1994) and Vucetich
et al. (2002). In the study by Messier (1994) killing rates were
shown to be increasing asymptotically in moose population
density to a specific level and were not significantly explained
by wolf population density, i.e. a prey dependent response,
while Vucetich et al. (2002) showed that predator density
explained more variation in kill rate than did prey density.
3.2.
3
These models were initially developed by Lotka (1925) and
Volterra (1931) independently.
Step 2 — opportunity cost
The relevant cost term in this type of welfare analysis is the
opportunity cost, which is used to place a value on the inputs
EC O L O G IC A L E C O N O M IC S 6 8 ( 2 0 08 ) 51 7 –5 24
that could be used to produce other things of value to people.
With reference to Boardman et al. (1996) opportunity cost can
be defined in the following way: “The opportunity cost of using
an input to implement a policy is its value in its best alternative
use” (emphasis added). A common assumption is that the best
alternative use of a domesticated animal, other than as food
for a wild carnivore, is the market value of the animal. If the
killed individual is a valuable breeder then the lost opportunity to pass on its genetic information to future offspring
should be reflected by this market value.
Ideally, market prices for live domesticated animals times the
number of carnivore caused deaths of these animals should be
used to estimate the welfare loss, measured as the opportunity
cost. In the absence of such information government compensations can be an alternative, given that the very reason for these
schemes are to compensate the owner for the welfare loss
associated with predation. However, it should be borne in mind
that in many countries, including Sweden, livestock owners
complain that government compensations are inadequate.
In addition to domesticated animals, carnivores like wolves
and wolverines may also predate on valuable game species,
creating an opportunity cost on hunters by reducing the
harvestable population. For the carnivore populations of
concern in this analysis, the four large carnivores in Sweden,
the main relevant game species is the moose and roe deer.
Locally, certain carnivore species may have a strong impact,
although there exist no national statistics on the effect of the
“four large” on the Swedish moose, or roe deer, populations.
The opportunity cost on hunters is therefore excluded in the
empirical analysis. For this reason the opportunity cost
estimates should be seen as a lower bound measure of the
true welfare loss caused by these carnivores.
4.
Method
The data used in this paper is in the form of time series for the
carnivore populations, as well as for the annual economic
compensations for the period 1971–2003. However, the data
series contain gaps for the years 1981, 1982, 1984–86, 1988,
1989, and 1992–1996, due to unavailable data at some County
Administrative Boards that were responsible for the statistics
before 1996. The estimation of the population numbers for
wolf, wolverine and lynx is conducted by using either radio
tagged animals or the amount of animal footprints produced
by the individuals of the different species. For brown bear, the
DNA technique was used to estimate the population levels,
but in this case, the “footprints” were brown bear droppings
(NFS 2004:17; NFS 2004:18). We have chosen to use the
economic compensations to livestock owners as a representation of the welfare loss associated with predation. This
implicitly assumes that the government compensations are
correctly set, which is not known with certainty. Reindeer
herders and other livestock owners in Sweden have complained about the low level of these compensations, suggesting that they may not represent an overestimate of the welfare
loss, as we indicate above. However, it could also represent
strategic behavior on the part of herders and livestock owners.
To assess the reasonableness of using government compensation as the measure of opportunity cost in this analysis,
521
we conduct a rough comparison between compensations per
killed reindeer and the slaughter (i.e. meat) value of an average
reindeer. In 2006, SEK 42.5 million was paid in compensations
(Viltskadecenter, 2006), and approximately 30,000 reindeer were
killed by carnivores during this period4 (exact figures are
unavailable), resulting in an average compensation of approximately SEK 1400/reindeer. Recent (2004) reindeer meat prices are
around SEK 40/kg (including SEK 12/kg in government price
subsidy) (Svensson and Nergård, 2005). With an average
slaughter weight of 31.58 kg (Bostedt, 2001), this gives a slaughter
value of SEK 1263/reindeer, suggesting that compensations
approximately cover the loss in slaughter value. However, we
note that the livestock killed by carnivores might not be the ones
intended for slaughter (e.g. valuable breeders), and that the
existence of carnivores results in higher livestock management
costs (e.g. surveillance and rounding up scared animals).
The current Swedish system for damage compensation to
owners of semi-domesticated and domesticated animals can
be divided into two separate parts: one part for domesticated
animals, such as sheep, cows, horses, dogs, etc., and one part
for the reindeer, which is considered a semi-domesticated
animal in Sweden. The compensation to reindeer herders
today is administrated by the Saami Parliament, a Swedish
government body that gives the Saami a limited amount of
autonomy. Before 1993 this compensation was based on the
actual damage done by the carnivores.5 However, since 1996
the compensation system for reindeer is related to the number
of carnivores in the sense that the number of carnivores and
carnivore rejuvenations is the base for the amount of money
that is transferred to the reindeer herders. The compensation
for predation on other types of livestock is administrated by
the County Administrative Boards, and is still based on the
actual number of carnivore-killed animals. The data on the
economic compensation in this analysis is collected from the
Saami Parliament and the Wildlife Damage Center in Grimsö.
5.
Results
The results presented in this section describe our econometric
estimates of the social costs of the Swedish carnivore policy.
As shown in the Theory section the predation, and thus costs,
are likely to be explained by the number of carnivores, as well
as the amount of prey. Furthermore, the shape of the cost
functions can be nonlinear. Given the unknown, but likely
multiplicative form of the predator functional response, g(N,
P), we have chosen to consequently take the natural logarithm
of both the dependent and independent variables. For all cost
functions an interaction term, ln(N) · ln (P), is also tested. For
each species, three different time series have been chosen to
represent N, the population of prey, namely the Swedish
reindeer population, the annual cull of roe deer (used as a
proxy for the population of roe deer), and the annual cull of
4
According to the homepage of the Swedish Saami Parliament,
www.sametinget.se.
5
The actual damage in terms of found livestock or reindeer
killed or injured by a large carnivore.
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EC O LO GIC A L E CO N O M ICS 6 8 ( 2 00 8 ) 5 1 7 –5 24
Table 1 – Ordinary least squares estimates of cost
functions for the “four large” carnivores in Sweden
(t-values within parenthesis)
ln
(Bear
cost)
Constant
ln(Bear pop.)
ln(Wolf pop.)
ln
(Wolf
cost)
− 2.138 7.838
(−.778) (.883)
1.661
(5.58)⁎⁎⁎
17.111
(2.526)⁎
ln(Wolverine pop.)
ln
(Wolverine
cost)
− 13.049
(−3.122)
1.280
(6.769)⁎⁎⁎
1.035
(6.119)⁎⁎⁎
ln(Roe deer)
.446
.231
(1.496) (.280)
ln(Reindeer)
ln(Wolf pop.) ⁎ ln(Moose)
Degrees of freedom
F-test value
Adjusted R2
Mean cost (SEK)
− 5.095
(−2.863)⁎
1.349
(5.796)⁎⁎⁎
ln(Lynx pop.)
ln(Moose)
ln
(Lynx
cost)
1.796
(4.582)⁎⁎⁎
−1.420
(− 2.401)⁎
18
17
20.50 12.17
.661
.626
868,089 174,796
18
48.78
.827
5,961,663
18
63.81
.862
5,950,406
⁎Significant at 5% level, ⁎⁎significant at 1% level, ⁎⁎⁎significant at .1%
level.
moose (used as a proxy for the population of moose).6 The
results in Table 1 present the models with the best F-test
values for the four carnivore species. All estimates have been
corrected for heteroscedasticity using White's consistent
estimator, see White (1978), and all cost data have been
deflated to the year 2000.
As is evident from Table 1, different prey species contribute
to different carnivore cost functions. Note that human
hunting of wild ungulates such as roe deer and moose in
Sweden results in great quantities of slaughter offal (animal
remains) being left in the forest, which is thought to be an
important food source for the four large carnivores. This could
contribute to carnivore population growth and thus indirectly
to social cost through predation on livestock. For the brown
bear cost function the annual moose cull – as a proxy for the
moose population – was used, but did not contribute
significantly to explain cost (probability of rejecting the null
hypothesis that the β-coefficient is equal to zero was 84.8%).
The marginal cost of an increase in the bear population,
evaluated at the 2003 national population level of bears (2000)
and moose cull (103,185) is approximately SEK 5120.7 However,
should the recent brown bear population increase be over-
6
Roe deer and moose are included as substitute preys for the
predator, which intuitively has an effect on the predation on
livestock and reindeer, and thus, the economic compensation.
7
Using the parameter estimates in Table 1 the cost function is:
C = e[− 2.138 + 1.661 ⁎ ln(Bearpop) + .446 ⁎ ln(Moosecull)]. Taking the first derivative of the cost function and evaluating at the 2003 levels give the
marginal cost estimate. The other marginal cost estimates were
calculated in a corresponding way.
estimated (due to changes in estimation techniques discussed
in Section 2.2) this figure would likely be an underestimate.
The cost function for wolves shows signs of heteroskedasticity8 when regressed on the wolf population and
the annual cull of moose. This can partly be explained by the
high dispersal capacity of wolves combined with their
efficiency as a predator. A few wolves in the “wrong” place
can cause large social cost. Thus, the geographical distribution
of wolves is of great importance, which was also highlighted in
Boman et al. (2003). The moose cull variable is not significant
by itself but contributes significantly through an interaction
term. The marginal cost of an increase in the wolf population,
evaluated at the 2003 population level of wolves (85) and
moose cull (103,185) is SEK 7480, which is more than 2.5 times
the marginal cost of bears. This should come as no surprise
however, given the omnivorous nature of the brown bear. The
wolf marginal cost figure can also be compared with estimated
costs of conserving African wild dogs (another pack-hunting
canine carnivore) in northeastern South Africa from Lindsey
et al. (2005). Estimates ranged from 18 to 27 packs per USD
100,000 or – with an average of 7 wild dogs per pack – USD 793
to 1020 per wild dog. Using an exchange rate of 7 SEK per USD
this gives approximately SEK 5500 to 7000 per wild dog, which
is comparable to the estimated cost figure for wolves in
Sweden. Note, however, that the Lindsey et al. (2005)
estimates are based on prey profiles composed entirely of
wild prey (i.e. lost recreational hunting opportunities) and
that they are average cost estimates while the carnivore
cost function approach in this paper allows for marginal cost
estimates.
The wolverine cost function includes both the wolverine
and reindeer populations, which are both significant at the 1%
level. The marginal cost of an increase in the wolverine
population, evaluated at the 2003 population levels of
wolverines (375) and reindeer (245,000), is a staggering SEK
109,700. This relatively large figure can be attributed to the
natural habitat and present population distribution of the
Swedish wolverine population, which brings it in direct
conflict with the reindeer herding industry. Without doubt is
wolverine conservation one of the greatest challenges facing
Swedish wildlife management.
Finally, the lynx cost function gives the best fit and
includes the lynx and roe deer populations. The marginal
cost of an increase in the lynx population, evaluated at the
2003 population levels of lynx (1550) and roe deer cull
(162,000), is SEK 15,120, which is considerably lower than the
marginal cost for wolverines but is twice the marginal cost for
wolves. Table 1 also gives the mean value of the compensation
costs used as dependent variable. This shows that wolverine
and lynx compensations are approximately 34 times higher
than the compensations for wolf predation. This is due to the
fact that wolf population in Sweden has been found mainly
outside the reindeer herding area and outside other areas with
extensive livestock management, e.g. free-ranging sheep.
At this stage it is of interest to use the regression models
from Table 1 to forecast the social costs of the four large
8
The null hypothesis of homoscedasticity can be rejected at the
13% level through the Breusch–Pagan test.
EC O L O G IC A L E C O N O M IC S 6 8 ( 2 0 08 ) 51 7 –5 24
Table 2 – Forecasts of social costs for the “four large”
carnivores in Sweden (SEK)
523
carnivores in terms of predation when the current Swedish
carnivore policy population goals are reached. Our estimates
are presented in Table 2 and include a comparison with the
current (2003) costs. All figures have been deflated to the year
2000.
Note that the figures for brown bear and lynx in the
rightmost column are negative. This is because the 2003
population figures for these species are actually above the
government population goal; i.e. reaching the population goal
implies reducing the populations. In total, reaching the
government goal would imply a 35% increase in social costs.
The table also illustrates that the social cost of reaching the
government population goal is almost solely attributable to
one species: the wolverine. However, this conclusion depends
to a large extent on the current population distribution of
the Swedish wolf population. A permanent establishment of
wolves in the reindeer herding area in northern Sweden – from
which it is now excluded – would likely imply a dramatic
increase in predation costs.
costs are not expected to be incurred. In spite of the incessant
public debate about the expanding Swedish wolf population
the marginal costs for increasing the population are relatively
modest by comparison. However, these estimates are based
on the 1971 to 2003 geographical population distribution of
wolves. It is important to note that the costs of an expansion of
the wolf population into the reindeer herding area in northern
Sweden cannot be forecasted using the cost functions
estimated in this paper. Furthermore, it should be noted that
political pressure might result in greater relative compensation for the wolf, being a “politically sensitive” carnivore,
which might slightly bias findings for this carnivore.
In total, expanding the Swedish carnivore population to
reach the population goals decided upon by the Swedish
Parliament would imply a more than 50% increase in social
costs, compared with the 2003 level. This cost is almost solely
caused by the wolverine population. For this increase in
carnivore populations to be socially worthwhile in an economic sense this cost should be weighed against the possible
benefits in terms of use and non-use values of these
carnivore populations. Recent Swedish attempts at estimating non-market benefits of carnivore conservation can
be found in Ericsson et al. (2007, in press) and Bostedt et al.
(2008).
This paper has shown that theoretical models from population ecology – in this case functional response models – can
be applied to estimate welfare costs associated with wildlife
conservation. This type of interdisciplinary effort is likely to
be fruitful given the challenge of making informed policy
decisions on the benefits and costs of conserving our megafauna, despite the real externalities associated with them
(e.g. carnivores in Scandinavia and North America or elephants
in Africa).
6.
Acknowledgements
Carnivore Population Cost when
species
goal
population
goal is
reached (A)
Brown bear
Wolf
Wolverine
Lynx
Sum
1000
200
575
1500
1,952,000
1,639,000
54,279,000
17,559,000
75,429,000
Current
(2003)
cost (B)
Difference
(A − B)
6,174,000
886,000
30,493,000
18,312,000
55,866,000
−4,222,000
752,000
23,785,000
−753,000
21,194,000
Conclusion
Carnivore conservation and, in some countries, reintroduction, bring with them significant social costs. This paper
estimates individual cost functions for carnivore species
based on time series data on populations of the four large
carnivores in Sweden, and costs of predation (measured by
government compensations to livestock owners), and
together with ecological models of functional response. In
some cases, notably in the estimation of the wolf cost
function, there are problems with heteroscedasticity, which
can largely be explained by variations in the spatial distribution. For example, a few wolves may migrate into the reindeer
herding areas in one year, but then disperse from this area in
the following year. During years when they are inside the
reindeer herding area, costs are high due to the wolves'
efficiency as a predator and to the fact that reindeer herding is
the most widespread form of extensive livestock management
in Sweden.
Marginal cost estimates based on these functions show
that an expansion of the Swedish wolverine population comes
with a large price tag attached. The marginal cost for
expanding the lynx population is also high, but given that
the current Swedish carnivore policy this is less of an issue
since the target lynx population level is already reached, these
We acknowledge funding from the MountainMistra (“FjällMistra”)
research program. We appreciate comments by Mattias Boman,
Southern Swedish Forest Research Centre, Swedish University
of Agricultural Sciences, Alnarp, and Jens Persson, Dept. of
Conservation Biology, Swedish University of Agricultural
Sciences, Grimsö. We are also grateful for English editing by
Scott Cole.
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