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REVIEW PAPER

Paying for International Environmental Public Goods

Rodrigo Arriagada, Charles Perrings

Received: 21 December 2010 / Revised: 20 April 2011 / Accepted: 29 April 2011 / Published online: 2 June 2011

Abstract Supply of international environmental public
goods must meet certain conditions to be socially efficient,
and several reasons explain why they are currently under-

supplied. Diagnosis of the public goods failure associated

with

particular ecosystem services is critical to the devel-

opment of the appropriate international response. There are

two categories of international environmental public goods

that are most likely to be undersupplied. One has an
additive supply technology and the other has a weakest link

supply technology. The degree to which the collective
response should be targeted depends on the importance of

supply from any one country. In principle, the solution for

the undersupply lies in payments designed to compensate
local providers for the additional costs they incur in

meeting global demand. Targeted support may take the

form of direct investment in supply (the Global Environ-
ment Facility model) or of payments for the benefits of

supply (the Payments for Ecosystem Services model).

Keywords International environmental public goods !
Ecosystem services ! Payments for ecosystem services !
Global environmental public

INTRODUCTION

How can we best secure the provision of international
environmental public goods (IEPGs)—public goods offer-

ing benefits that span multiple national jurisdictions? It is

well understood that markets undersupply public goods,
and there is a wealth of evidence that many environmental

public goods have been systematically undersupplied over

a long period of time (Millennium Ecosystem Assessment
2005). If environmental public goods occur at the scale of

the nation state or below, the failure of markets to supply

public goods may be offset by the actions of local or

national governments. There exist many national agencies
with responsibilities for the provision of environmental

public goods such as habitat for rare and endangered spe-

cies, clean water, environmental health protection, and so
on. There also exist many offset or mitigation systems for

securing private provision of public goods at a national

level (Madsen et al. 2010). At the international level, where
there is no supranational authority to take responsibility,

the failure of markets to deliver environmental public
goods is more difficult to offset. Depending upon the

magnitude and distribution of the payoffs to public good

provision, individual countries will have a stronger or
weaker incentive to commit resources to their provision.

Doing more than that depends upon agreement between

nation states (Kaul et al. 2003a; Barrett 2007).
Many IEPGs are strictly global. Examples include the

conservation of the genetic diversity on which all future

evolution depends, the mitigation of climate change, the
control of emerging infectious diseases, and the manage-

ment of sea areas beyond national jurisdiction. Many more

are regional, such as the control of acid rain, the manage-
ment of multi-country river basins, and the protection of

international watersheds (Touza and Perrings 2011). Like

all public goods, IEPGs exhibit both consumption indi-
visibilities and non-excludability. Non-excludability means

that once the good is provided, none can be excluded from

enjoying the benefits it confers. Indivisible consumption
occurs when one country’s enjoyment of the benefits does

not diminish the amount available for others. Public goods

are said to be ‘pure’ when they are both non-exclusive and
non-rival (indivisible) in consumption. They are said to be

impure if they are either partially excludable or partially

rival—the most common form of which are local public
goods, particularly the local common pool resources

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AMBIO (2011) 40:798–806

DOI 10.1007/s13280-011-0156-2

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analyzed by Ostrom (1990). In most cases, it is not possible

for any single state to provide such goods on its own.
International public good supply depends on either inter-

national coordination or international cooperation (Anand

2004).
This article focuses on IEPGs whose benefits extend to

people in multiple countries. Such IEPGs frequently also

deliver benefits across multiple generations (Kaul et al.
1999), but we do not address this aspect of the problem. In

practice, the beneficiaries of international public goods
include national populations and their representatives,

nation states, transnational corporations and non-govern-

mental organizations, as well as a newly emerging set of
institutions. Globalization has altered the way that mem-

bers of civil society organize themselves across national

boundaries. The information revolution has also stimulated
new forms of social participation. New networks, fre-

quently built around environmental websites, enable the

exchange of ideas and implementation techniques. These
new relationships and interactions have created a ‘global

environmental public’, interested in asserting new rights

and responsibilities to the resources of the planet. Its
concerns span both the ethical responsibilities of individ-

uals, organizations, countries and corporations, and the

alternative forms of governance of the biosphere.
Following the Millennium Ecosystem Assessment, we

suppose that the benefits people obtain from biosphere

depend on a set of ecosystem services comprising:

• Provisioning services: products people obtain from
ecosystems, such as food, fuel, fiber, fresh water, and
genetic resources.

• Cultural services: nonmaterial benefits people obtain
from ecosystems through spiritual enrichment, cogni-
tive development, reflection, recreation, and aesthetic

experiences.

• Regulating services: benefits people obtain from the
regulation of ecosystem processes, including air quality

maintenance, climate regulation, erosion control, reg-

ulation of floods and droughts, regulation of human
diseases, and water purification.

• Supporting services: those that are necessary for the
production of all other ecosystem services, such as
primary production, production of oxygen, and soil

formation.

These services affect human wellbeing in many ways:

through their role in the production of consumption goods,

their support of human health and security, or the satis-
faction of peoples’ cultural and spiritual needs. A number

of these services have the characteristics of IEPGs, the

most important of which involve the regulating and sup-
porting services. Figure 1 indicates the relation between

categories of ecosystem services and components of

wellbeing identified by the Millennium Ecosystem

Assessment. Of these, only the provisioning services con-
sistently generate benefits that are both divisible (rival) and

exclusive. The other services yield benefits that are gen-

erally indivisible and non-exclusive. We focus on the group
of ecosystem services that are both public and interna-

tional. These are services that: (i) cover more than one

group of countries; (ii) benefit not only a broad spectrum of
countries but also a broad spectrum of the global popula-

tion; (iii) meet the needs of both present and future gen-
erations (Kaul et al. 1999; Anand 2004). International

public goods generated in any one county must therefore

generate spillover effects beyond a nation’s boundary
(Morrissey et al. 2002).

IEPGs can further be classified according to their

‘technology of supply’ (Sandler 2004). The standard
treatment of public goods focuses on demand (Hirshleifer

1983). However, understanding the technology of supply of

IEPGs is critical to the development of appropriate incen-
tives. Three common examples of public good supply

technologies are ‘additive’, ‘best shot’, and ‘weakest link’

technologies. As the name implies, in the additive case, the
socially available amount Y of a public good is nothing but
the ‘simple sum’ of the separate amounts, yi, produced by
each of m participating countries, the i = 1,…, m. In the
case of simple sum public goods, such as carbon seques-

tration, each unit of carbon sequestered has the same value

no matter where it occurs. In the case of weighted sum
public goods, such as habitat protection, the contribution of

each hectare protected depends on its characteristics

(Sandler 2004). For ‘best shot’ public goods, the benefit to
all countries is determined by the most effective provider.

For example, the Centers for Disease Control and Pre-

vention are funded by the U.S.A., but provide information
on infectious diseases to all countries. For ‘weakest link’

public goods, the benefits to all countries are limited to the

benefits offered by the least effective provider. The best
example of this is the control of infectious diseases. So for

HIV and tuberculosis, the level of protection available to

all countries is only as good as the control of the disease
exercised in the poorest, most densely populated, and least

well-coordinated country (Perrings et al. 2002).

Social composition functions

Y ”

X

i
yi Summation

Y ” min
i

yiY ” min
i

yi Weakest-link

Y ” max
i

yi Best-shot

Of all the Millennium Ecosystem Assessment ecosystem
services, the regulating services are most often supplied as

IEPGs. Examples include disease control, which is

frequently supplied as a weakest or weaker link public

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good, climate regulation through, e.g., carbon sequestration,

which is supplied as an additive pure public good, or

watershed protection which is generally an additive but
impure public good (Holzinger 2001; Dombrowsky 2007;

Touza and Perrings 2011). Many international public goods

are also jointly produced with local public goods.
Biodiversity in tropical forests, for example, yields a set of

private benefits in the form of timber and other products
including medicinal plants, hunting, fishing, recreation, and

tourism. At the same time, tropical forests are a source of

carbon sequestration, genetic information, hydrological and
microclimatic regulation—commonly described as co-

benefits (Perrings and Gadgil 2003).

An important feature of IEPG is that their spatial extent
depends partly on the natural hydrological and atmospheric

flows, and partly on the social linkages between coun-

tries—the flow of goods, people, and information. The
global reach of carbon sequestration is a property of the

general circulation system, but the global reach of disease

regulation is a property of the global trade and air trans-
portation systems. In fact, the closer integration of the

world economic system has rapidly increased the number

of environmental public goods that are global in reach
(Kaul et al. 2003b):

• New technologies increasingly enhance human mobil-
ity as well as the movement of goods, services, and

information around the world (e.g., case of transmission

of human diseases and air pollution as international
environmental public bads).

• Economic and political openness have provided further
impetuses to cross borders and transnational activities

(e.g., case of transmission of human diseases and air
pollution as international environmental public bads).

• Systematic risks have increased (e.g., case of climate
change as an international environmental public bad).

• International regimes are becoming more influential,
often formulated by small groups of powerful nations

yet often claiming universal applicability (e.g., case of
bio-prospecting contracts to find cure for cancer and

other human diseases).

The central problem addressed in this article is how to
secure environmental public goods that (a) are provided at

particular locations but offer benefits over a wider area, and

(b) generate local benefits that are below the local cost of
supply. These are the IEPGs that are most likely to be

undersupplied. This article is organized in four sections.

The following section reviews the fundamental problem
with IEPGs—the incentive that each country has to free

ride on the efforts of others. A third section then considers

the options for addressing the problem. This reviews the
applicability of currently popular instruments, such as

payments for ecosystem services, in terms of the charac-

teristics of the public good concerned. A final section
draws out the implications for national and international

environmental policy.

WHY ARE INTERNATIONAL ENVIRONMENTAL
PUBLIC GOODS UNDERPROVIDED?

International environmental public goods generate benefits
that spill over national borders, so that the benefits (or

Provisioning
(food, fuel , fiber, fresh water, and

genetic resources)

ECOSYSTEM SERVICES

Cultural
(spiritual enrichment, cognitive

development, reflection,
recreation, and aesthetic

experiences)

Regulating
(air quality maintenance, climate

regulation, erosion control,
regulation of floods and droughts,
regulation of human diseases, and

water purification)

Security
• Personal safety
• Secure resource access
• Security from disasters

Basic material for good life
• Adequate livelihoods
• Sufficient nutritious food
• Shelter
• Access to goods

Health
• Strength
• Feeling well
• Access to clean air and water

Good social relations
• Social cohesion
• Mutual respect
• Ability to help others

CONSTITUENTS OF WELL-BEING

ARROW’S COLOR
Potential for mediation by socioeconomic factors

Low Medium High

Supporting
(nutrient

cycling, soil
formation,

and primary
production)

Fig. 1 Linkages between
ecosystem services and human
well-being (arrow’s width
intensity of linkages between
ecosystem services and human
well-being) (adapted from
Millennium Ecosystem
Assessment, 2003)

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costs) of those goods extend beyond the country of origin.

If the marginal local benefits of public good provision are
less than the marginal local costs, there will be no incentive

to provide the public good at all. If the marginal local

benefits of public good provision exceed the marginal local
costs of provision, but benefits also accrue to other coun-

tries, there will be an incentive to produce the public good,

but unless the country is a ‘best-shot provider’ it will not be
at a level that would satisfy international demand (Ferroni

and Mody 2002; Kanbur 2003, 2004). We first of all review
the problem and then consider the options for addressing it.

We have elsewhere considered the cases where the national

incentive to produce IEPGs is sufficient to meet global
demand (Touza and Perrings 2011). In this article, we

address the case where independent local action is not

sufficient to secure efficient global supply.
Consider the conservation of endangered species. Can

we rely on the national action to produce the efficient

amount of an IEPG such as the protection of iconic spe-
cies? The key to understanding this lies in the difference

between a pure public good and a private good. For a

private good, everyone pays the same price, but is free to
consume as much or as little as they want. Consumers

adjust the quantity they consume given the market price.

For a pure public good everyone consumes the same
amount of the ‘‘good’’ but is willing to pay a different price

for it. Consumers adjust the amount they are willing to pay

for the public good given the quantity supplied (Batina and
Ihori 2005). In general, private provision of public goods

will be below the socially optimal level. Efficiency requires

that marginal benefit equals marginal cost. In the case of
conservation of endangered species (or any other public

good), the relevant measure of marginal benefits is social

marginal benefit—the sum of all countries marginal bene-

fits. We illustrate the problem in Fig. 2, in which local and
global benefit curves for species conservation in a partic-

ular country, i, are presented. Global benefits are repre-
sented by the vertical sum of the benefit curves of country
i, and all other countries.

The level of conservation in country i that maximizes
local net benefits is indicated by yi, while the level of
conservation that maximizes global benefits is indicated by

yi#. yi is given by the intersection of local supply and local
benefit curves, and yi# by the intersection of local supply
and the vertical sum of local benefits for all countries.

Since the marginal cost of provision at yi# is greater than
country i would be willing to accept on its own, socially
optimal provision of the public good depends on the

existence of a mechanism to cover the ‘incremental’ cost of
socially optimal provision to country i.

Biodiversity conservation, like many other IEPGs, is an

impure global public good. If there are many potential
providers, each generates local benefits from its conser-

vation effort, but also benefits from the conservation

actions of others. Following, Perrings and Gadgil (2003),
we characterize the problem for the individual country as

follows. Vi denotes welfare of the ith country, assumed to
depend on consumption of a vector of market goods, xi, and
global biodiversity conservation, denoted C = C(y1,
y2,…,ym), then the problem it faces is of the general form:

MaxxiyiV
i xi; yi; C y1; . . .; ym

! “! ”

In other words, it derives a direct benefit from its own

conservation efforts, yi, but also benefits from the global
conservation effort to which it has contributed, C. If the
cost of conservation effort in terms of the cost of market

Quantity of local
conservation effort

Costs and
benefits of
local
conservation
effort

Global benefits
Costs of local
conservation effort

yi yi*

‘Incremental
cost’ =
benefits to
global
interests in
excess of
costs
warranted by
local interest
of country i

Benefits to all other countries

Benefits to country i

Fig. 2 Efficient provision of
conservation effort

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goods is p(yi), and if the income available to country i is Ii,
this will be subject to the constraint that

Ii ” xi $ p

yi

! ”

yi

In the absence of cooperation, and noting that

Viyi ” oV
i=oyi; Vixi ” oV

i=oxi and Cyi ” oC=oyi, wellbeing
of the ith country will be maximized where the local level
of conservation satisfies:

ViYi
Vi
xi
” pyi %

Vic
Vi
xi
Cyi

If global wellbeing is the sum of the welfare of all
countries, V ”

P
j”1 V

j, then global wellbeing will be

maximized where

ViYi
Vi
xi
” pyi %
X

j”1

V jc
Vi
xi
Cyi

The extra terms in the summation term on the right hand
side capture the conservation benefits that the ith country
confers on all other countries. These benefits will be

neglected by the ith country unless there is a mechanism to
convert them into a direct incentive.

The failure of markets to signal the global benefit of

such public goods accordingly results in under-invest-
ment in their local provision. The benefits of protection,

management and establishment of forests provide a good

example. Apart from the loss of the valuable environ-
mental services (e.g., protection of genetic resources, air

quality maintenance, climate regulation and regulation of
human diseases), forest degradation frequently translates

into a loss of timber and non-timber forest products

important to local livelihoods (Landell-Mills & Porras
2002).

Currently, there are few measures of the underprovi-

sion of public goods. The United Nations Development
Programme (UNDP) has opted instead for measures of

‘‘adequate’’ provision that differ from one public good to

another. Such measures can, for example, correspond to
the complete elimination of global public bads. More

generally, they are measures of what is considered pos-

sible, given the current state of technology (e.g., to
control—rather than eradicate—the problem of HIV/

AIDS) and what is ‘‘fair’’ (e.g., what would emerge if all

concerned stakeholders had an effective voice in the
decision-making process) (ODS-UNC 2002). The crite-

rion of adequacy is not meant to indicate optimality—the

balancing of marginal costs with the sum of people’s
marginal willingness to pay for a particular public good

(see Samuelson 1954; Cornes and Sandler 1996). Rather,

it is meant to establish a relatively simple, yet reliable,
yardstick for measuring the present provision of a certain

good against a technical notion of adequacy.

In the case of the global public good ‘communicable

disease control’, for example, it has been possible, given
the biological characteristics of the infectious agent and

available technologies, to completely eradicate certain

diseases. In these cases, the criterion for adequate provision
is defined as complete eradication, or zero incidences in the

‘‘wild.’’ The determination of ‘‘adequate provision’’ is

based solely on technical considerations, without reference
to costs, benefits or existing preferences and willingness to

pay. Therefore, there may be cases where adequate provi-
sion may not be economically feasible. It is important to

assess the net benefits/costs of inaction against the net

benefits/costs of corrective action to determine, at least
approximately, whether meeting the technological

requirements for adequate provision is economically

desirable (UNDP 2002).

POLICY OPTION: PAYMENTS FOR ECOSYSTEM
SERVICES

In principle, the solution to IEPG problems of this form lies
in payments designed to compensate local providers for the

additional costs they incur in meeting global demand.

Indeed, that is the basis on which the Global Environment
Facility (GEF) was founded. The GEF unites 182 member

governments—in partnership with international institu-

tions, nongovernmental organizations, and the private
sector—to address global environmental issues. An inde-

pendent financial organization, the GEF provides grants to

developing countries and countries with economies in
transition for projects related to biodiversity, climate

change, international waters, land degradation, the ozone

layer, and persistent organic pollutants. These projects
benefit the global environment, linking local, national, and

global environmental challenges, and promoting sustain-

able livelihoods. The concept of incremental cost, which
notionally determines the payments made by the GEF, is

related to the difference between the cost a country would

be prepared to bear in the provision of an environmental
public good (the cost that would be warranted in terms of

the national benefits generated by the public good) and the

cost of meeting global demand for the same public good
(Pearce 2003, 2005). It is a national payment for an envi-

ronmental service that is an IEPG.

Systems of Payments for Ecosystem Services (PES)
have become popular instruments for dealing with IEPGs,

in part because they appear to satisfy the incremental cost

principle (Ferraro and Simpson 2002; Goldstein et al. 2006;
Wunder 2007; Ferraro and Kiss 2007; Pagiola 2008; Engel

et al. 2008; Wunder et al. 2008). They are not, however,

appropriate mechanisms in all cases. International PES
schemes are appropriate where non-marketed ecosystem

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services are privately supplied in one country, but offer

benefits that are public and accrue elsewhere.
To illustrate the potential pluses and minuses of PES, we

consider a particular problem: the impact of local defor-

estation on the provision of a range of IEPGs including
climate regulation (through carbon sequestration), protec-

tion of genetic diversity, and watershed protection in

addition to timber and non-timber forest products. The
Economics of Ecosystems and Biodiversity (TEEB) study

is a major international initiative to draw attention to the
global economic benefits of biodiversity, to highlight the

growing costs of biodiversity loss and ecosystem degra-

dation, and to draw together expertise from the fields of
science, economics and policy to enable practical actions

moving forward. The current assessment of TEEB on PES

has used existing studies to estimate the mean value of both
the macroclimatic regulation offered by terrestrial carbon

sequestration, and the change in provisioning and cultural

services offered by forest systems. Its findings are pre-
liminary but telling. TEEB (TEEB 2009; Kumar 2010)

suggests that the mean values of forest ecosystem services,

in US$/ha/year, are dominated by regulatory functions:
specifically regulation of climate ($1965), water flows

($1360), and soil erosion ($694). The mean value of all

provisioning services combined—timber and non-timber
forest products, food, genetic information, pharmaceuti-

cals—is $1313. This is less than the value of water flow

regulation alone. There are substantial off-site benefits to
forest conservation that are not currently captured by forest

landowners and are difficult to incorporate on PES

schemes.
Governments around the world have frequently imple-

mented forest protection policies in areas high in biodi-

versity, landscape beauty or critical for their watershed
protection. However, as pressure mounts on governments

to curtail spending and cut budget deficits, their ability to

invest directly in the provision of public goods and services
is compromised. Where public authorities have been

unable to tackle the public good problem, they have sear-

ched for ways to involve non-governmental actors. Efforts
to transfer responsibility for forest environmental services

out of the public sector have relied on a combination of

regulation and market-based approaches (Landell-Mills
and Porras 2002). Experience has shown that well-designed

market-based instruments can achieve environmental goals

at less cost than conventional ‘‘command and control’’
approaches, while creating positive incentives for continual

innovation and improvement (Stavins 2003). Examples of

such instruments in the forestry sector include stumpage
value-based forest revenue systems, financial and material

incentives, long-term forestry concessions, trade liberal-

ization, forest certification and the promotion of markets
for non-timber forest products.

The costs and benefits associated with many human

activities spill over jurisdictional boundaries, thereby
generating externalities that are often reciprocal and

quantitatively significant (Cornes 2008). Therefore, IEPGs

supply depends on either international coordination or
international cooperation. Among payment schemes to

internalize the external benefits of maintaining intact for-

ests, Reducing Emissions from Deforestation and Forest
Degradation (REDD) is an effort to create a financial value

for the carbon stored in forests, offering incentives for
developing countries to reduce emissions from forested

lands and invest in low-carbon paths to sustainable devel-

opment. REDD is an example of international coordination
in delivery of ecosystem services. Its integration into

international market-based climate change policies poses a

number of challenges both to institutional design and to
implementation. At present, for example, there are few

effective mechanisms for converting international pay-

ments to governments into incentives to on-the-ground
forest communities (Myers 2008; Sikor et al. 2010).

Indeed, creating an effective multilevel system of pay-

ments is seen as the core issue in building REDD consid-
ering that REDD goes beyond deforestation and forest

degradation, and includes the role of conservation, sus-

tainable management of forest and enhancement of forest
carbon stocks (Anngelsen and Wertz-Kanounnikoff 2008).

It is predicted that financial flows for greenhouse gas

emission reductions from REDD could reach up to US$30
billion a year. This significant North–South flow of funds

could reward a meaningful reduction of carbon emissions

and could also support new, pro-poor development, help
conserve biodiversity, and secure vital ecosystem services.

A second issue is the linkage between distinct ecosystem

services. The REDD scheme targets one important eco-
system service: carbon sequestration. However, it has the

potential to secure other services as well. These services

potentially include both habitat provision for biodiversity
conservation and watershed protection. Reaching interna-

tional agreement on an instrument to reduce emissions

from deforestation and forest degradation, while recog-
nizing the co-benefits offered by conservation, and the

sustainable management of forested watershed would both

secure global carbon sequestration services, as well as help
to maintain other valuable services provided by forests

(TEEB 2009). There is growing recognition that REDD

planning requires a broadened approach. A future REDD
mechanism should incentivize emissions reduction from

reduced deforestation, enhanced carbon sequestration and

address a number of non-carbon services. Implementation
of REDD also requires attention to the quality of forest

governance, conservation priorities, local rights and tenure

frameworks, and sub-national project potential (Phelps
et al. 2010).

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IMPLICATIONS FOR INTERNATIONAL
ENVIRONMENTAL POLICY

Globalization isoften associated withincreased privateness—

economic liberalization is associated with the growth of the
number of goods and services allocated through markets,

international market integration, and enhanced private cross-

border economic activity such as trade, investment, transport,
travel, migration and communication.However,globalization

is also about increased publicness—about people’s lives

becoming more interdependent. Events in one place of the
globe often have worldwide repercussions. Moreover, a

growing volume of international policy principles, treaties,

norms, laws, and standards is defining common rules for an
ever-wider range of activities (Kaul et al. 2003b).

Public goods are recognized as having benefits that

cannot easily be confined to a single ‘‘buyer’’ (or set of
‘‘buyers’’). Yet once they are provided, many can enjoy

them for free. A clean environment is an example. Without

a mechanism for collective action, these goods will gen-
erally be underprovided. In fact, many crises dominating

the international policy agenda today reflect the under-

provision of global public goods. With globalization,
externalities are increasingly borne by people in other

countries. Indeed, issues that have traditionally been

merely national are now global because of the greater
interconnectedness of the planet.

Kaul et al. (2003b) suggest a rethinking of three notions
underpinning the theory of public goods. First, properties

of non-rivalry in consumption and non-excludability of

current benefits do not automatically determine whether a
good is public or private. Some goods may be either public

or private. Nevertheless, it is important to distinguish

between a good’s having the potential of being public (that
is, its having non-rival and non-excludable properties) and

its being de facto public (non-exclusive and available for

all to consume). Second, public goods do not necessarily
have to be provided by the state. Many other actors can,

and increasingly do, contribute to their provision. And

third, a growing number of public goods are no longer
national in scope, having assumed cross-border dimen-

sions. Many have become global and require international

cooperation to be adequately provided.
For the most part, the theoretical and empirical literature

in economics has focused on two polar models of public

goods provision: the provision of pure public goods that
benefit all agents, and the provision of local public goods

that only benefit agents in one community (Bloch and

Zenginobuz 2007). We are concerned with cases where the
members of one community enjoy positive spillovers from

the public goods provided by other communities. In the

context of global climate regulation, the REDD scheme
will compensate tropical nations that succeed in reducing

carbon emissions from deforestation and forest degrada-

tion—source of nearly one fifth of global carbon emissions.
Since forests offer a number of benefits aside from carbon,

however, the scheme could potentially benefits to com-

munities that would otherwise be unable to afford them
(Stickler et al. 2009). If well designed and implemented,

PES schemes such as REDD have the potential to secure

provision of IEPGs that offer benefits at multiple scales,
such as the protection of water supplies, local and regional

climate regulation, and habitat provision for the protection
of biodiversity. The effectiveness of PES schemes depends

heavily on the conditionality of payments (Arriagada and

Perrings 2009), but the principles for their effective design
and implementation are well understood.

To summarize, the implications of this paper for inter-

national environmental policy are the following:

1. Diagnosis of the public goods failure associated with

particular ecosystem services is critical to the devel-

opment of the appropriate international response.
There are a number of cases where the incentive

structure is such that independent actions by nation

states will be ‘good enough’ to secure the public
interest (Touza and Perrings 2011). Where the tech-

nology of supply is ‘best shot’ or where the local

benefits are high enough to lead to a level of supply
that is close to the global optimum, then the indepen-

dent actions of nation states will be adequate. How-

ever, where local benefits lead to a level of local
supply that leaves global demand unsatisfied, then

international coordination or cooperation in the deliv-

ery of ecosystem services will be required. We note
that this largely depends on the nature and strength of

off-site effects. Local actions that generate significant

off-site benefits or costs are most likely to require
international coordination or cooperation. Off-site

effects can reflect both natural (through hydrological

or atmospheric flows) and social (through trade and
travel) transmission. Since social transmission of

effects is rapidly evolving, understanding social trans-

mission pathways is important to the diagnosis of the
public goods failure.

2. There are two categories of IEPGs that are most likely

to be undersupplied. The first involves an additive
supply technology, a high opportunity cost of supply

and transmission to a large number of other countries

through the general circulation system. Examples
include mitigation of climate change, and management

of transboundary nutrient flows, currently addressed

through the UN Framework Convention on Climate
Change and the Convention on Long Range Trans-

boundary Air Pollution. The second involves a weakest

link supply technology, and transmission to a large

804 AMBIO (2011) 40:798–806

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number of other countries through global trade,

transport and travel. Examples include the manage-

ment of infectious zoonotic diseases and the control of
invasive pest species, currently addressed through the

International Health Regulations, the Sanitary and

Phytosanitary Agreement, and the Convention on
Biological Diversity.

3. If there is a public goodfailure that demands international

coordination or cooperation, it is then important to
determine the degree to which the collective response

should be targeted. In principle, action should be targeted

to reflect the weight attaching to the supply from any one
country.So,at one extreme, international contributions to

an IEPG with an additive supply technology that is

unweighted (a simple sum), such as carbon sequestration
by forest plantations, should not be targeted at all.

Whereas international contributions to a weakest link

international environmental public good, such as infec-
tious disease control, should be targeted at the weakest

link. In practice, most ecosystem services are jointly

produced(come asa bundle),and involveanintermediate
position. Particular countries are more important for the

provision of some services than others (e.g., high

biodiversity countries contribute more to the global gene
pool than others) so most international contributions to

IEPGs should be targeted in some measure.

4. For IEPGs that are supplied in specific countries,
support may take the form of direct investment in

supply (the Global Environment Facility model) or of

payments for the benefits of supply (the Payments for
Ecosystem Services model). The fact that GEF is

under-resourced, and is only weakly targeted, suggests

that the second option may become the dominant
mechanism for assuring local provision of IEPGs. We

have elsewhere discussed the conditions that need to

be satisfied for PES schemes to be effective (Arriagada
and Perrings 2009). The most important of these is that

payments for ecosystem services should be conditional

on the supply of those services. Where PES schemes
have both income transfer/poverty alleviation and

public good supply objectives, conditionality may be

lost altogether. It is important that the design of PES
schemes fit the diagnosis of the public good problem,

and the technology of public good supply.

Acknowledgment The authors acknowledge support from the
United Nations Environment Programme.

REFERENCES

Anand, P. 2004. Financing the provision of public goods. The World
Economy 27: 215–237.

Anngelsen, A., and S. Wertz-Kanounnikoff. 2008. What are the key
design issues for REDD and the criteria for assessing options? In
Moving ahead with REDD,, ed. A. Angelse. Bogor: SUBUR
Printing.

Arriagada, R., and C. Perrings. 2009. Making payments for ecosystem
services work. Ecosystem services economics working papers.
United Nations Environment Programme, Nairobi.

Barrett, S. 2007. Why cooperate? The incentive to supply global
public goods. Oxford: Oxford University Press.

Batina, R., and T. Ihori. 2005. Public goods: Theories and evidence.
Berlin: Springer-Verlag.

Bloch, F., and U. Zenginobuz. 2007. The effect of spillovers on the
provision of local public goods. Review of Economic Design 11:
199–216.

Cornes, R. 2008. Global public goods and commons: theoretical
challenges for a changing world. International Tax and Public
Finance 15: 353–359.

Cornes, R., and T. Sandler. 1996. The theory of externalities, public
goods, and club goods. New York: Cambridge University Press.

Dombrowsky, I. 2007. Conflict, cooperation and institutions in
international water management: An economic analysis. Chel-
tenham: Edward Elgar.

Engel, S., S. Pagiola, and S. Wunder. 2008. Designing payments for
environmental services in theory and practice: An overview of
the issues. Ecological Economics 65: 663–674.

Ferraro, P., and A. Kiss. 2007. Direct payments to conserve
biodiversity. Science 298: 1718–1719.

Ferraro, P., and R. Simpson. 2002. The cost-effectiveness of
conservation payments. Land Economics 78: 339–353.

Ferroni, M., and A. Mody. 2002. Global incentives for international
public goods: Introduction and overview. In International goods
public: Incentives, measurement, and financing,, ed. M. Ferroni,
and A. Mody. Dordrecht: Kluwer Academic Publishers.

Goldstein, J., G. Daily, J. Friday, P. Matson, and R. Naylor. 2006.
Business strategies for conservation on private lands: Koa
forestry as a case study. Proceedings of the National Academy of
Sciences of the United States of America 103: 10140–10145.

Hirshleifer, J. 1983. From weakest-link to best shot: The voluntary
provision of public goods. Public Choice 41: 371–386.

Holzinger, K. 2001. Aggregation technology of common goods and
its strategic consequences: Global warming, biodiversity, and
sitting conflicts. European Journal of Political Research 40:
117–138.

Kanbur, R. 2003. IFI’s and IPG’s: operational implications for the
World Bank. In Challenges to the World Bank, IMF: Developing
country perspectives,, ed. A. Buira. London: Anthem Press.

Kanbur, R. 2004. Cross-border externalities, international public
goods and their implications for aid agencies. In Global tensions:
Challenges, opportunities in the World economy,, ed. L. Beneria,
and S. Bisnath. New York: Routledge.

Kaul, I., I. Grunberg, and M. Stern, ed. 1999. Defining global public
goods. In Global public goods: International cooperation in the
21st century. Oxford: Oxford University Press.

Kaul, I., P. Conceição, K. Le Goulven, and R. Mendoza, ed. 2003a.
How to improve the provision of global public goods. In
Providing global public goods: Managing globalization. Oxford:
Oxford University Press.

Kaul, I., P. Conceição, K. Le Goulven, and R. Mendoza, ed. 2003b.
Why do global public goods matter today? In Providing global
public goods: Managing globalization. Oxford: Oxford Univer-
sity Press.

Kumar, P. 2010. The economics of ecosystems and biodiversity.
London: Earthscan.

Landell-Mills, N., and I. Porras. 2002. Silver bullet or fool’s gold? A
global review of markets for forest environmental services and

AMBIO (2011) 40:798–806 805

! Royal Swedish Academy of Sciences 2011
www.kva.se/en 123
Author’s personal copy

their impact on the poor. London: International Institute for
Environment and Development.

Madsen, B., N. Carroll, and K. Moore. 2010. Offset and compensation
programs worldwide. Washington: Ecosystem Marketplace.

Millennium Ecosystem Assessment. 2005. Ecosystems and human
well-being: General synthesis. Washington: Island Press.

Morrissey, O., D. Te Velde, and A. Hewitt. 2002. Defining
international public goods: conceptual issues. In International
public goods: Incentives, measurement, financing, ed. M. Ferroni,
and A. Mody. Dordrecht: Kluwer Academic Publishers.

Myers, M. 2008. Policies to reduce emissions from deforestation and
degradation (REDD) in developing countries. An examination of
the issues facing the incorporation of REDD into market-based
climate policies. Resources for the future. Discussion paper RFF
DP 07-50, Washington, DC, USA.

Ostrom, E. 1990. Governing the commons: The evolution of
institutions for collective action. Cambridge: Cambridge Uni-
versity Press.

Pagiola, S. 2008. Payments for environmental services in Costa Rica.
Ecological Economics 65: 712–724.

Pearce, D. 2003. The social cost of carbon and its policy implications.
Oxford Review of Economic Policy 19: 362–384.

Pearce, D. 2005. Paradoxes in biodiversity conservation. World
Economics 6: 57–69.

Perrings, C., and M. Gadgil. 2003. Conserving biodiversity: Recon-
ciling local and global public benefits. In Providing global public
goods: Managing globalization, ed. I. Kaul, P. Conceição, K. Le
Goulven, and R. Mendoza. Oxford: Oxford University Press.

Perrings, C., M. Williamson, E. Barbier, D. Delfino, S. Dalmazzone,
J. Shogren, P. Simmons, and A. Watkinson. 2002. Biological
invasion risks and the public good: an economic perspective.
Conservation Ecology 6: 1.

Phelps, J., M. Guerrero, D. Dalabajan, B. Young, and E. Webb. 2010.
What makes a REDD country? Global Environmental Change
20: 322–332.

Samuelson, P. 1954. The pure theory of public expenditure. The
Review of Economics and Statistics 36: 387–389.

Sandler, T. 2004. Global collective action. Cambridge: Cambridge
University Press.

Sikor, T., J. Stahl, T. Enters, J.C. Ribot, N. Singh, W.D. Sunderlin,
and L. Wollenberg. 2010. REDD-plus, forest people’s rights and
nested climate governance. Global Environmental Change 20:
423–425.

Stavins, R. 2003. Experience with market-based environmental policy
instruments. In Handbook of environmental economics, ed.
M. Karl-Goran, and R.V. Jeffrey. Amsterdam: North Holland.

Stickler, C., D. Nepstad, M. Coe, D. Mcgrath, H. Rodriguez, W.
Walker, B. Soares-Filho, and E. Davidson. 2009. The potential
ecological costs and co-benefits of REDD: A critical review and
case study from the Amazon region. Global Change Biology 15:
2803–2824.

TEEB. 2009. TEEB climate issues update. Nairobi: United Nations
Environment Program.

Touza, J., and C. Perrings. 2011. Strategic behavior and the scope for
unilateral provision of transboundary ecosystem services that are
international environmental public goods. Strategic Behavior
and the Environment 1: 89–117.

United Nations Development Programme. 2002. Profiling the provi-
sion status of global public goods. New York: UNDP Office of
Development Studies.

Wunder, S. 2007. The efficiency of payments for environmental
services in tropical conservation. Conservation Biology 21:
48–58.

Wunder, S., S. Engel, and S. Pagiola. 2008. Taking stock: a
comparative analysis of payments for environmental services
programs in developed and developing countries. Ecological
Economics 65: 834–852.

AUTHOR BIOGRAPHIES

Rodrigo Arriagada (&) is a Assistant Professor in the Department
of Agricultural Economics at the Pontificia Universidad Catolica de
Chile. Dr. Arriagada is also an associated research fellow for Envi-
ronment for Development (EfD) initiative in Central America and
ecoSERVICES Group at Arizona State University. His fields of
specialization are environmental economics, program evaluation
econometrics, applied microeconomics and development economics.
Dr. Arriagada’s current research interests focus on environmental
economics and policy, the use of experimental and quasi-experi-
mental program evaluation methods on conservation interventions,
deforestation and land use, issues at the intersection of economic
development and environmental protection.
Address: Department of Agricultural Economics, Pontificia Univers-
idad Católica de Chile, Avenida Vicuña Mackenna 4860 Macul,
Santiago, Chile.
e-mail: rarriagadac@uc.cl

Charles Perrings is a Professor of Environmental Economics at
Arizona State University (ASU). Previous appointments include
Professor of Environmental Economics and Environmental Manage-
ment at the University of York; Professor of Economics at the Uni-
versity of California, Riverside; and Director of the Biodiversity
Program of the Beijer Institute, Royal Swedish Academy of Sciences,
Stockholm, where he is a Fellow. At ASU, he directs (with Ann
Kinzig) the ecoSERVICES Group within the College of Liberal Arts
and Sciences. The Group studies the causes and consequences of
change in ecosystem services—the benefits that people derive from
the biophysical environment. It analyses biodiversity change in terms
of its impacts on the things that people care about.
Address: ecoSERVICES Group, School of Life Sciences, Arizona
State University, Box 874501, Tempe, AZ 85287-4501, USA.
e-mail: Charles.perrings@asu.edu

806 AMBIO (2011) 40:798–806

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Paying for International Environmental Public Goods
Abstract
Introduction
Why are International Environmental Public Goods Underprovided?
Policy Option: Payments for Ecosystem Services
Implications for International Environmental Policy
Acknowledgment
References

1

Economic Incentives and Wildlife Conservation

Erwin H. Bult

e

Department of Economics
Tilburg University, The Netherlands

G. Cornelis van Kooten
Department of Economics
University of Victoria, Canada

Timothy Swanson
Department of Economics
University of London, United Kingdom

Draft: October 27, 200

3

________________________________________________________________________

Wildlife exploitation and conservation involves various costs and benefits, which
should all be taken into account to achieve an optimal outcome. For this to occur, it will
be necessary to develop appropriate economic instruments and incentives. Examining the
scope for his is the topic of the current study. The time and funds available to complete
this paper were extremely limited, which effectively made it impossible to complete a
thorough and detailed analysis. As a result, in the paper we focus on what can be learne

d

from standard economics. The paper lacks the level of detail and data to provide
guidance in many operational issues.

Wildlife management poses a particular challenge to the global community
because wildlife has an impact not only on people living in areas where wildlife is found,
but also on people located considerable distances away. The problem is that the costs and
benefits of wildlife exploitation facing “source” states differ substantially from those
faced by other countries. From an economist’s perspective, the main wildlife problem is
that all too often many of the costs of harvesting wildlife are not appropriately taken into
account. In particular, the values that wildlife such as elephants, tigers and rhinoceros
have for people who may someday view them in the wild and the values that such fauna
have for people who are simply delighted to know that such wildlife exist (having no
intention of ever viewing them) are ignored in most harvesting decisions. Further, when
property rights are insecure, those who harvest wildlife do not take into account the cost
of their actions on the future availability of the resource because they do not have a stake
in wildlife beyond those accessible to them today. This cost is referred to as the “user
cost” and it is typically ignored in harvest decisions unless property rights are clearly
stated, and protected. As a result, in situ wildlife is undervalued leading to their possible
overexploitation (see below).

2

In essence, there is a divergence between what is optimal from a regional,
community or individual perspective, and what is optimal from the perspective of a
country or even global society. To address this divergence, a variety of economi

c

instruments can be employed. The term “economic instrument” is used to describe any
device/method used by government to achieve an outcome contrary to (other than) the
one that occurs in the absence of any government intervention. The government
essentially has three categories of economic instruments available to it: (1) common
values and norms (threats or moral suasion in economic terms), (2) command and
control, and (3) market incentives, which are also referred to as economic incentives
(EIs). Moral suasion refers to the ability of the state to convince economic agents
(individuals or firms) to act in a fashion that is socially desirable. Voluntary instruments
(e.g., product certification/labeling by an industry association), perhaps accompanied by
threats, are one aspect, but there also exist opportunities to “convince” citizens to report
poachers, protect wildlife habitat and so on. Economic or market incentives and
command and control (i.e., regulation) are generally used in combination, often out o

f

necessity.

The objective of this study is to examine the scope of economic incentives in the
conservation of wildlife. The focus of the study is on developing countries as these host
most of the biodiversity and wildlife. The main results are as follows: While economists
often believe that, in general, the best way to conserve wildlife and their habitat is to
encourage efficient and sustainable use of these resources, the scope of EIs in such
conservation efforts as an ‘extra measure’ to regulate harvesting pressure may in some
cases be limited. Specifically, we argue that there are cases where the usual gains of EIs
may be of secondary importance. Whether or not such gains materialize depends on the
specific characteristics of a species and the parties involved in its harvesting. This should
be assessed on a case-by-case basis. If both the habitat and the harvesters are
“homogenous” (in the sense that there is little variation in the area in which the species is
harvested and the skills/technologies of those harvesting the species), then the gains from
EIs are small. These conditions may hold for (low-tech) open-access harvesting of certain
species in Africa, but not for fisheries where “firms” of various sizes from individuals to
large corporately-sponsored vessels are engaged in harvesting.

Two important qualifications are in order. First, while the role of EIs in regulation
of harvesting may (but need not) be modest, we argue that international EIs may be of
great importance when it comes to habitat conservation (indirectly contributing to
wildlife conservation). In this respect we mainly think of means to capture and channel
non-use values associated with conservation to affected parties living with (or owning)
wildlife in developing countries – an example of international transfers or subsidies.
Second, establishing property rights (or secure use rights for extended periods – that is
establishing property rights in legal or physical space) is consistently encouraged by
economists as a first step towards efficient management of resources – both of land and
the wildlife it supports. Whether this first step must be complemented by additional EIs
(tax, tradable quota) to arrive at a truly global optimum, however, is not certain.
Sometimes additional command and control measures are to be preferred, and sometimes
no additional measures are necessary (for example when external effects are small – see
the next section).

3

We begin by examining why society might wish to intervene in the protection and
provision of wildlife – the economic theory underpinning public wildlife management. In
section 2, we provide a general discussion of the types of EIs that are available for
addressing environmental spillovers, focusing on those instruments that may be useful in
wildlife management. In section 3 we compare the performance of EIs with that of
command and control instruments when addressing the two most important threats to
wildlife; overexploitation and habitat conversion. In section 4 we emphasize the
importance of the institutional context, and discuss implications for developing countries.
Section 5 summarizes and concludes, and we propose a few priorities for targeted follow-
up research that can be useful for making progress towards implementation of key issues
raised in the paper.

1. ECONOMIC EFFICIENCY AND VARIOUS FAILURES

Environmental economics has become an important subject within economics as
people have become increasingly concerned with pollution and other forms of
environmental damage. The fact that some wild fauna and flora is threatened and
endangered can be considered a special form of environmental damage. Therefore, ideas
from environmental economics are relevant to wildlife management. What then does
economic efficiency mean in a wildlife context? We consider ‘wildlife’ as broadly as
possible, encompassing all biotic resources, including timber and fish.

First off, economic efficiency refers to the maximization of the well-being or
welfare of citizens within a society. Economists measure welfare using a monetary metric
and define it in terms of the economic surpluses (or rents) that accrue to economic agents
in their capacities as consumers and producers. The surplus accruing to consumers is
given by the difference between the benefit that they get from consuming a bundle of
goods and services and what they have to pay for those goods and services. In technical
terms, the “consumer surplus” is the difference between what people are willing to pay
for goods and services and what they actually do pay. Likewise, the “producer surplus” is
defined as the difference between the revenues from the sale of goods and services and
the cost of providing them. Since fixed costs are sunk (i.e., made in the past and
unaffected by current decisions), the net benefit accruing to economic agents as owners
of factors of production is given as the difference between total revenue and total variable
costs. In essence, therefore, the total surplus or economic welfare at any time is given by
the difference between the benefit that citizens receive as “consumers” and the costs of
providing the goods and services consumed – the area between the demand and supply
curves (Figure 1).

To maximize society’s overall well being when producing a good, an additional
unit of the good should be provided as long as the benefit from this additional unit (the
marginal benefit), to whomsoever they accrue, exceed the cost of this additional unit (the
marginal cost), no matter who incurs them. Provision should stop when the benefit
received from the last unit equals the cost of supplying that unit – marginal costs equal
marginal benefits. Marginal benefits (or marginal willingness to pay) define the demand
function, and marginal costs determine the supply function (Figure 1). Then society’s

4

welfare is maximized at the price and quantity where demand intersects supply – where
marginal benefit (demand price) equals marginal cost (supply price). The area between a
falling demand function and a rising supply function, up to the point where they intersect,
represents the maximum sum of the consumer and producer surpluses, or maximum
societal welfare, as shown in Figure 1. However, environmental problems arise because
of three sorts of failure: institutional failure, market failure and policy failure. These
forms of failure are clearly interdependent. For example, property rights to resources may
be insecure (institutional failure) because governments fail to provide the legal
environment that supports them (policy failure) or because of public good characteristics
associated with the resource (market failure). One result of imperfect property rights can
be external effects (another form of market failure). We will explain these concepts
below.

Quantity per unit time

Price ($ per unit)

Supply = marginal cost

Demand = marginal benefit

Consumer
Surplus

Producer
Surplus

0

price =
marginal
cost

Q*

P*

Quantity per unit time
Price ($ per unit)
Supply = marginal cost
Demand = marginal benefit
Consumer
Surplus
Producer
Surplus
0
price =
marginal
cost
Q*
P*

Figure 1: Maximizing Social Welfare when there is No Market Failure

1.1 Institutional failure: Ill-defined and enforced property rights

First, consider the pervasive problem of institutional failure. The most relevant
manifestation of institutional failure for the case of wildlife and trade in species is
probably insecure property rights, or ‘open access’ to resources – both the species
themselves and the land upon which they live. Property rights can be understood as
characteristics that define the rights and duties associated with the use of a particular
asset or resource. Four property regimes are usually identified.

5

1. Private property In this case, the private owner has the right to utilize and benefit
from the exploitation, conservation or sale of wildlife, as long as no (socially
unacceptable) externalities are imposed on others (e.g., when shooting wildlife endangers
the lives of others). Private ownership does not imply absence of state regulation
(control), as private property cannot exist without state sanction and protection.

2. State property The state owns the wildlife and individuals may be allowed to
harvest them, but only according to rules imposed by the state or the CITES Management
Authority.

3. Common property In this case, a group owns and manages the wildlife resource,
and the group excludes those who are not members. Members of the group have specified
rights and duties, while non-members must accept exclusion. Coordination (regulation) of
management may or may not be forthcoming, depending on local circumstances.

4. No property rights (res nullius) When a property right is not assigned, open or
free access is the result. Under open-access, each potential user of the resource has
complete autonomy to utilize wildlife since none has the legal right to keep another
potential user out.

A summary is provided in Table 1. In practice, resources are often held in
overlapping combinations of these regimes, and it is possible to shift from one
(dominant) regime to another when conditions change. Failure to enforce or manage
properly a state or common property resource (which is frequent) leads to open-access,
which is the case for some endangered large-game species. The switch from common and
state regimes to open-access as a result of population growth is well documented (Murty
1994; Bromley 1999).

Table 1: Classification and Characteristics of Property Rights
Type Characteristics Implications for economic incentives

Private property Exclusive rights assigned toindividuals
Strong incentives for conservation of
resources and for investment as well

State property
Rights held in collectivity with
control exercised by CITES
authority or designated agency

Creating opportunities for attenuation of
rights; managers have incentives for personal
gains

Common property

Exclusive rights assigned to all
members of a community;
approaching private property
when coordination arises.

Creating free-riders problem and low
incentives for conservation

Open access Rights unassigned; lack ofexclusivity
Lack of incentives to conserve; often resultin

g

in resource degradation

6

Property rights do not really exist under open access, and if there is no
cooperation under communal ownership (or no enforcement under state and private
ownership), then property rights are insecure. The absence of secure property rights (or
even open-access) has resulted in excessive depletion of resources and biological assets
for the following reason. The true cost of exploiting a resource consists of two distinct
components: the private extraction costs and the unobserved opportunity cost, or the
value of the resource in situ – the user cost. The intuition behind user cost in the context
of a renewable resource is as follows: harvesting a unit of the resource today means that
this unit and the growth (including any offspring) it causes are not available for future
consumption. The (future) value of uncaught game depends on many different factors,
including the discount rate, future markets for the resource, technological developments,
reproductive features and so on. A sole private owner aiming to maximize profits will
maximize the discounted value of this rent, and treat the resource as an asset. Hence, the
value of unharvested animals and plants prevents a rational wildlife manager from over
harvesting the resource, but only as long as she expects to be the one to benefit from this
“investment”. Private property may result in a conservative harvesting policy. In the
absence of externalities and given similar discount rates, the same applies for state
ownership.

An open-access resource exists if there is no possibility to exclude firms attracted
by excess profits, with their entry competing away those profits. If there is unrestricted
access to the resource, no person can be sure of who will benefit from the value of
uncaught game. In an open-access situation, no individual harvester has an economic
incentive to conserve the wildlife, and none can efficiently conserve the wildlife by
delaying harvest. Doing so will only enhance the harvest opportunities of competitors,
which is the tragedy of open-access. One might say that the individual does not care about
escaped game, and discounts future harvests at an infinite rate (Neher 1990). New
harvesters will be attracted to the activity, or existing ones will expand their efforts so long
as they earn more than the (opportunity) cost of their effort. In bionomic equilibrium, all
rent is dissipated, and total cost equals total revenue, rather than marginal cost being equal
to marginal benefit. The situation where marginal cost exceeds marginal benefit is usually
referred to as economic overexploitation.

In terms of Figure 1, failure to account for the user cost implies that agents will not
base their harvest decisions on the supply (social marginal cost) curve as drawn, but instead
on another marginal cost curve that is below it. This is illustrated in Figure 2 (discussed
further below). As a consequence, harvested quantities in the short run will increase and
prices will fall.

1.2 Market failure: Spillover effects and public goods

Two general types of market failure may occur and undermine economic
efficiency of resource management, even when property rights are secure. First, the
supply function may not embody all of the costs of producing goods and services, in
which case market prices are no longer reliable as a measure of value. In the context of
wildlife conservation we may think of nonuse values associated with in situ conservation

7

– the utility (or well being) that people derive from knowing that certain species exist or
thrive, even though they will never “use” or view such species themselves.

One can think of wildlife as providing two sorts of products – products that result
in the “consumption” of the specimen (e.g. fiber, wool, caviar, timber, ivory, bones, gall
bladders, hides and bush meat, live export to zoos or as pets, ornamental and medicinal
plants), and non-consumptive uses like eco-tourism, bird-watching and photography
associated with the protection of in situ amenities and wildlife. A negative external effect
can occur when consumptive use of wildlife reduces their numbers, and, as a wildlife
population declines, the total economic (in situ) value to “preservers” falls. However, the
consumers of wildlife products fail to take this into account in their decisions because
there do not exist appropriate economic institutions and incentives to get “consumers” of
wildlife to regard the costs they impose on those deriving utility from conservation. This
is referred to as an externality, although the term spillover may be more descriptive of
what happens and will be used here interchangeably with externality. Externalities can be
good or bad, but their effect is that the supply price no longer reflects the true cost to
(global) society of the activity. There is a divergence between private and social costs of
provision, because one of the inputs in production, namely the environment, is not
appropriately priced; the environmental cost or damage is not taken into account. In terms
of Figure 2, agents do not base harvesting decisions on the marginal social cost (the ‘true’
supply curve), but on private marginal costs. When uncorrected, they will supply too
much and social marginal cost lies above marginal benefit.

Second, there are many situations where private provision of a good or service
does not occur because, once it is provided, no one can be excluded, and “use” or
“consumption” by one person does not diminish the amount available to others. This is
the definition of a public good. Public goods such as national defense, clean air and
water, wilderness, biodiversity, and other environmental amenities will not be supplied
privately because the provider cannot capture the benefits of so doing – once provided,
no one can be excluded, so free riding is possible. Clearly some aspects of wildlife bear
the characteristics of a public good. Wildlife contributes to global biodiversity (the “web
of life”) and enhances the well being of the majority of people (through the provision of
“non-use values”). However, no one has the appropriate incentive to provide wildlife
habitat or otherwise protect wildlife as they cannot capture the full benefits from the
needed investments. Market failure occurs because the amount of a public good is under-
provided, and thus marginal social benefits exceed marginal social costs. In this case,
more of the (public) good should be provided, but it is forthcoming only if society
subsidizes a private supplier, or provides it publicly.

1.3 Policy failure: Perverse government incentives

A final reason why wildlife may be overharvested (or why their habitat is
degraded in many regions) has to do with perverse government policies. One well-known
form of policy failure is subsidization of harvesting or habitat conversion. As will
become clear below, one way to address market failure is through implementation of a
tax or user charge. However, rather than charging users to exploit natural resources, there
are many real-world examples where exploitation of natural resources is encouraged

8

rather that restricted – policies aggravate rather than mitigate pre-existing distortions.
When use of resources is subsidized, the marginal cost curve is pushed downwards and
short-term supply expanded. This could be illustrated in Figure 2 (but is not) by adding
another marginal cost function that would lie below the ‘marginal private harvest cost
function with insecure property rights’.

Marginal private benefit
from wildlife harvest

Harvests in a given period (h)

$

0

PS

h*

P*

Marginal private harvest
costs with insecure property rights
(no user cost) and no spillovers

hS

Marginal private harvest
costs with secure property
rights (no spillovers)

Marginal social costs = marginal
private harvest costs + user costs
+ spillover costs

h1

a
b

c
e
f
g
Marginal private benefit
from wildlife harvest
Harvests in a given period (h)
$
0
PS
h*
P*
Marginal private harvest
costs with insecure property rights
(no user cost) and no spillovers
hS
Marginal private harvest
costs with secure property
rights (no spillovers)
Marginal social costs = marginal
private harvest costs + user costs
+ spillover costs
h1
a
b
c
e
f
g
d
Marginal private benefit
from wildlife harvest
Harvests in a given period (h)
$
0
PS
h*
P*
Marginal private harvest
costs with insecure property rights
(no user cost) and no spillovers
hS
Marginal private harvest
costs with secure property
rights (no spillovers)
Marginal social costs = marginal
private harvest costs + user costs
+ spillover costs
h1
a
b
c
e
f
g
Marginal private benefit
from wildlife harvest
Harvests in a given period (h)
$
0
PS
h*
P*
Marginal private harvest
costs with insecure property rights
(no user cost) and no spillovers
hS
Marginal private harvest
costs with secure property
rights (no spillovers)
Marginal social costs = marginal
private harvest costs + user costs
+ spillover costs
h1
a
b
c
e
f
g
d

Figure 2: Divergence of Social and Marginal Costs of Harvesting

1.4 Policy and land-use decisions

Figures 1 and 2 suggest that some harvesting of wildlife is typically optimal in the
sense that it maximizes welfare for society at large – this is the case when the marginal
social cost curve and the demand curve intersect for some positive harvest level h*. But
there is more. Economists have long known that allowing profitable use of resources
provides incentives for their conservation, and could lower enforcement costs associated
with conservation.

Prins et al. (2002) argue that arbitrary restrictions on use of wildlife in regions in
Africa have contributed to the demise of these species – a rather paradoxical statement

9

perhaps. The reason is that restrictions on use erode the incentive that people have to
invest in the protection of the species for potential future harvesting. People have no
incentive to invest in the conservation of the species. The result will be that base
resources on which wildlife depend for their survival (such as land) will be allocated to
other, more profitable uses. The key insight is that taking away the short-term incentive
to harvest may not be in the species’ best interests. In effect, this is another manifestation
of policy failure.

Local people make the decisions concerning land use and resource exploitation. It
is costly to enforce prohibitions against their chosen activities, and so prohibitions often
increase monitoring cost without conservation benefits. This is because the conversion of
reserve lands and incursions for wildlife poaching halt when local people want it to do so,
not when they are told to do so. The least costly policies provide incentives for the local
people to support the reserve’s designated uses, not bans against non-designated uses.
This implies that the most successful policies for the conservation of wildlife and
wildlands have been those that encourage their limited and managed development. For a
review of experiences in the context of crocodile management, roughly consistent with
this insight, refer to Hutton et al (2001).

The allowance of restricted use of wildlife encourages the local community to
view the wildlife as an asset for development. The allowance of restricted uses of reserve
lands allows the local people to receive some use of the lands while affording other uses
to the wildlife. In any event it is necessary for the benefits of wildlife conservation to be
distributed across the local community, by means of benefit sharing policies. This might
also take the form of sharing tourism revenues from the reserve with locals, or the
maintenance of a certain share of the jobs in the reserve for locals. The locals must be
encouraged by such policies to view the designation of the reserve for wildlife uses as a
specific form of local development policy for their benefit, not as a constraint on their
development of the reserved lands.

In the context of lands that are heavily used by local peoples, the designation of
protected areas and reserves simply acted as a mechanism for generating hostility
amongst the local populations. This hostility then became translated into management
costliness, as park managers attempted to enforce the restrictions on the use of the lands.
When local peoples viewed these restrictions with hostility, they simply made even
greater efforts to make use of lands that they believed to be theirs. Park managers had a
near impossible job of enforcement, and always insufficient funding to accomplish it.

An important caveat is in order here. There are exceptions to the statement that
use restrictions should not be too stringent. In certain cases zero use is economically
optimal. This is the case, for example, when full internalization of costs (including
spillover cost) would lead the market to break down – that is; the relevant supply curve
would lay above the demand curve such that there is no intersection where h*>0. In that
case, assuming there is no pressure for habitat conversion, strict preservation (possibly
supported through a trade ban) is globally optimal. Alternatively, strict preservation may
also be optimal when values derived from using the resource conflict with non-use
values. Economists typically assume that non-use values are associated with the size of

10

the wild stock, but recent research suggests that there may also be direct disutility
following from uses that are harmful to individual animals. For example, people may
care about the fact that individual whales are shot, rather than care about the fact that the
whale population becomes a bit smaller as a result. When direct disutility is sufficiently
large, global welfare is maximized by refraining from use altogether. Swanson and
Kontoleon (2003) have established that this condition holds for the black rhino, where
intrusive uses include trophy hunting and seducing rhinos to remove their horn. However,
the earlier statement about benefits sharing still applies in this context: If non-use values
are large, they should be captured (through transfer payments from the North, say) and
channeled to those who bear the burden of living with the wildlife. In the absence of such
transfers, advocating zero use may simply be non-sustainable.

1.5 Summary

Economics prescribes that wildlife should be harvested as long as the marginal
social benefits of so doing exceed the marginal social costs. Included in (marginal)
benefits are values of wildlife products (e.g., caviar, medicinal plants, ivory, bush meat,
hides) or the live specimen (if sold to a zoo or herbarium). In addition, (marginal) social
costs include (i) the loss in situ (existence, viewing) value that wildlife provide citizens
who may be located in countries other than the source country, (ii) the opportunity cost
from harvesting the wildlife today rather than waiting for a more opportune time in the
future when the specimen(s) may fetch a higher price, (iii) the lost future value of
offspring that might result from leaving the specimen(s) in place, and (iv) the opportunity
cost of the resources employed in the harvest activity. This is illustrated in Figure 2,
where failure to include all costs and benefits leads to suboptimal levels of harvest, hS,
that are likely well above those desired by global society, h*. The loss to global society
from harvesting hs rather than h* is area deg in Figure 2.

We can compare outcomes relative to the suboptimal level of harvest hs (and even
higher harvests if harvest subsidies are in place). If property rights to wildlife (and
habitat) are clearly spelled out and protected by the courts, then the harvest level would
fall to h1. In that case, global benefits would increase by area abfg in Figure 2. If in
addition it is possible to pay wildlife owners for the nonuse benefits of in situ wildlife,
harvest levels would decline further to h*, the globally optimal level of harvest. In that
case, global well-being would increase by an additional amount given by area bcef in
Figure 2. Our contention based on previous research concerning marginal willingness to
pay for increased numbers of wildlife and minimum viable populations required for
preservation of the species (e.g., see van Kooten and Bulte 2000) is that area abfg is
larger than bcef – that the benefits for wildlife protection of specifying and protecting
property rights is greater than those from attempting to subsidize “owners” of wildlife for
protecting in situ numbers. Indeed, without appropriate protection of property rights,
transfer payments to protect in situ wildlife cannot even be attempted. This is discussed
further below.

11

2. WILDLIFE AND ECONOMIC INSTRUMENTS

In this section, we examine the various economic instruments that are available to
countries, and discuss their advantages and disadvantages. As noted in the introductory
section, economists generally identify three economic instruments for addressing market
failure due to environmental externalities (or spillovers): (1) command and control
(hereafter C&C), (2) common values and norms (or more cryptically moral suasion), and
(3) market incentives. Conceptually, common values and norms are intermediary between
the “extremes” of the market and C&C (Loasby 1990; Stavins 2002). Common values
and norms develop more easily in a homogeneous society, while markets are more
appropriate in a heterogeneous society (CPB 1997, pp.42-44). However, there is much
confusion about the different instruments that are available and which are preferred (see,
e.g., Richards 2000).

One way to classify economic instruments for resolving environmental spillovers
and user cost is illustrated in Table 2, where instruments are classified according to two
dimensions – whether control of the means used to address the externality resides with
the private party or with the state, and who bears the costs. Market incentives include
subsidies, contracts, taxes and rights trading. (Rights are defined as an entitlement,
whether to a harvest quota of a wildlife species or fish stock, or the ability to develop or
conduct other activities on land, such as plowing or harvesting before a certain date.)
These give private parties complete discretion over the actions taken.

In contrast, C&C regulations generally provide much less discretion. As will
become clear below, this will lead to inefficiencies in the context of asymmetric
information between agent and regulator. At one extreme, regulations may specify
technology-based standards that regulated firms must use or, in the case of wildlife
perhaps, prescribe management standards – the “party-on-the-ground” (individual, firm,
wildlife management agency) has no degrees of freedom in decision-making.
Alternatively, regulations could provide the party-on-the-ground some degree of freedom
on how to proceed, as would be the case if the regulation only specified the number of
specimens that can be harvested each period (a quota of h* in Figure 2, say). The
regulator or CITES authority could then employ a market instrument (e.g., tradable
quota) to allocate the harvest in an efficient manner. In either case, the cost is borne by
the private party.

Table 2: Classification of Instruments for Addressing Wildlife Conservation
Private Party ControlWho Bears the

Costs? Price Based Quantity Based
Government

Control

Government/
Society

Subsidies,
transfers

Grandfathered (tradable) quota
Contracts

Public provision

Private Party Taxes, fees,charges, tariffs
Auctioned (tradable) quota C&C regulation

Harvest quota

12

The instruments included in Table 2 do not exhaust the full range of instruments for
environmental protection. For example, the literature contains discussions of liability
systems, and bond-and-deposit systems. However, neither of these types is likely to be
important for the case of wildlife conservation, and they will therefore be ignored in what
follows. We will focus on the most important economic instruments in the context of the
protection of wildlife – taxes/charges and tradable quota or rights. In addition, we consider
physical property rights, since tradable quota constitutes a legal right. Before we turn to a
discussion of these EIs, however, we will briefly evaluate the subsidy instrument.

Economists are typically critical about the use of subsidies to achieve conservation.
Consider the case where harvesters are subsidized to lower their harvest rate (this is the
logical counterpart of the literature on subsidies and pollution, where firms are paid to
lower their emissions). Assuming such agreements can be enforced, subsidies would
“work” in the sense that they tend to lower the optimal harvesting level of individual
harvesters. But there is a large potential problem with such subsidies when property rights
to the resource are imperfect: they can encourage entry into the harvesting sector that the
government aims to control. That is, even though harvesting per harvester goes down, the
number of harvesters will likely go up, compromising conservation objectives. Unless the
number of harvesters is somehow fixed such that new entry does not occur (e.g., when
property rights are secure), subsidies are a poor instrument to regulate harvesting. But,
importantly, there is another issue to consider in this context. In addition to suffering from
excess harvesting, many wildlife species are threatened by habitat conversion. Subsidies
can be an efficient, effective and equitable instrument to deal with habitat conversion. By
basing transfer flows on habitat made available by landowners, habitat conversion (and,
thus, indirectly also wildlife conservation) will be promoted. The fact that “entry” in the
habitat sector is promoted by subsidies is, of course, no problem – quite to the contrary; this
is the main intention. We return to this in section 3.3.

Following Panayotou (1994), we distinguish between property rights in physical
space (land ownership, ownership of wild fauna and flora on one’s land) and property
rights in legal space (e.g., a right to hunt or collect one or more specimens, trade live
specimens or parts or derivatives of them). The latter right specifies a narrower “bundle of
rights” to the resource than the former. Many species are migratory, so it is not possible to
establish full property rights (i.e., rights in physical space) as access will be shared with
others. But rights can still be established in legal space by defining an allowable harvest
level for individuals when the migratory species is on their land. Note that, when property
rights are established in legal space, the regulator can set the quota by taking external
effects into account. In contrast, if a private agent has the property rights to the resource in
physical space, the agent will fail to take these effects into account. When such spillovers
are important, assignment of property rights in physical space should be complemented
with other instruments in Table 2.

Establishing and enforcing property rights to resources in physical space can
provide an important impetus for sustainable use and conservation. In Figure 2, a private
agent given the right to a resource will change harvests from hS to h1, still below the
optimal harvest level h*, but movement is in the “right” direction. Other instruments are
needed to move from h1 to h*. Whether or not the change from h1 to h* can be

13

accomplished via EIs is debatable, but it may also be the case that the “effort” required to
go from from h1 to h* is not worth making: By appropriate specification of property rights
in physical space, most of the spillover problem may be overcome and the species spared
from potential extinction. Property rights depend on cultural conditions, so it may be better
in some sense to allocate them to a well-defined group or community rather than private
individuals/firms (see Table 1 and the excellent book by Baland and Platteau 1996).

The existence of property rights and the associated ‘right’ to exclude others from
using the resource implies that the user cost will no longer be ignored by those with access
to the resource. When property rights are secure, owners know that the fruits of their
‘investments’ (such as refraining from current harvesting, or postponing the decision to
convert habitat) will accrue to them. This means that they are more inclined to make such
investments. Addressing this institutional failure therefore enhances efficiency, and
comprises an important first step in enhancing efficiency and sustainability of resource
management. This is illustrated in the following case study, which illustrates the benefits of
defining property rights in legal space, and of benefits sharing.

2.1 Establishing secure property or use rights – The CAMPFIRE case study

In Southern Africa there was a widespread problem of poaching in designated
parks and reserves until government officials began to institute benefit-sharing programs.
These programs have taken many different forms. Sometimes they simply allow the local
community to set up tourist related facilities within the park (Natal’s Good Neighbour
Policy), other times they give the local community a share in the value of wildlife that
wanders onto their neighbouring lands (Zimbabwe’s CAMPFIRE program), and
sometimes the community is allotted a share of the receipts from wildlife management on
reserve lands (Wildlife Management Trusts). It is important to note how these community
funds were channeled back to the community in a manner that is widely visible
throughout the community. Sometimes this can be accomplished by means of purchasing
community goods such as schools etc. Other times it is best to send the benefits back to
the individuals in the form of jobs or money.

Zimbabwe’s approach of sustainable wildlife utilization has now been extended to all
of the communal areas by the CAMPFIRE program. Communities have been granted the
rights to manage as well as the means to capture the benefits from wildlife use. Since its
introduction, CAMPFIRE has managed to promote cooperation among village members and
has enhanced the institutional capacity of other community programs.

During the colonial times and up to 1978, legislation in Zimbabwe prohibited all
utilisation of wildlife for commercial as well as traditional hunting. Locals were even
relocated to make way for National Parks. As a result, many communities have been
disenfranchised from their natural resources and wildlife became not only valueless, but a
symbol of oppression and its destruction was encouraged. This alienation of people from
wildlife was clearly unsustainable. In 1955, the Department of National Parks and Wildlife
Management allowed commercial, (mainly white) farmers to utilize their wildlife
commercially. Consequently, farmers began to benefit from wildlife and started to look after
it. The value of wildlife products combined with the marginal economic viability of

14

conventional agriculture induced a shift from livestock to natural ecosystems
accommodating a wide range of species. While cattle could only be sold for meat, wildlife
could be photographed, sold as hunting trophies, as well as being sold as meat. At present,
some 75 percent of Zimbabwe’s commercial ranches now participate in the wildlife industry.

The first attempt to extend this system to communal areas was a program called
WINDFALL. The program involved allocating revenues from wildlife culling in National
Park and from safari hunting to district councils, but overall wildlife management remained
with the State. The results of this program were disappointing since the councils kept all the
money and local people saw few benefits. In 1975, a further step was taken which granted
councils the same rights as private landholders to appropriate the value from wildlife. In
order to increase the accountability of the councils the CAMPFIRE program was
established. The program ensured that producer communities rather than councils, managed
and benefited from wildlife.

Consider the impact of CAMPFIRE at the local level. Chikwarakwara is a small
village. Its population is exceedingly poor, largely uneducated and aging, since many of the
young people have migrated out of the area in search of work. As with many other villages,
disenfranchisement from its resource resulted in open access and over exploitation of their
wild resources. Chikwarakwara was characterised by an erosion of traditional controls on
resource use, growing population pressure, open access resources and unsustainable
resource use.

In 1989, there was a major step towards the implementation of CAMPFIRE
principles including the appropriation of wildlife revenues by the villagers. In the process,
special care was taken to ensure that villagers related the revenues they received to the
actual value of wildlife in their area. Moreover, the revenues were allocated to individuals
rather than to the community as a whole. This not only helped to increase the perception of
important individual revenues to be gained from wildlife management, but also boosted
accountability of the project.

As a result of this approach, more positive attitudes were fostered towards wildlife
and towards the management of the wildlife revenues. Villagers were able to carry out better
resource trade-offs and gained self-esteem. New institutions were created including wildlife
committees to ensure accountability and transparency. With stronger community unity, a
number of new opportunities began to open. Snaring was reduced as informal social controls
were established and strengthened. Entrepreneurial skills learned in wildlife management
were transferable to other projects such as the expansion of the irrigation system and the
management of the grinding mill.

To control levels of wildlife use, each council develops a sustainable hunting quota
in collaboration with the state departments. Middle agents who have the capital and skills
are employed to attract international clients. In order to avoid excessive monopoly power
and appropriation of the wildlife rents by the middle agents, a system of tenders was
established. Through time, the communities have improved their marketing skills, managing
to double their incomes between 1989-93. In fact, they have managed to capture better
prices than the government in key safari areas in Zimbabwe. The program has shown that

15

communities have rapidly learnt the necessary skills for natural resource management
despite the limited capacity of the state to provide technical assistance. In fact, districts with
donor support tended to be slower to develop and have suffered from excessive overhead
costs.

The philosophy of CAMPFIRE has been to set initially the conditions right for
sustainable wildlife management by local communities. The communities have started to
cooperate and build institutions for management of resources. A key insight is that allowing
use by well-defined groups (akin to establishing property rights) may go a long way towards
achieving efficiency. However, to arrive at the optimal outcome property or use rights will
generally have to be complemented with additional policies. As will become clear below,
economists usually prefer economic incentives in this case.

2.2 Preference for economic incentives

Economists generally express a preference for private party control, or market
incentives. A common feature of such incentives is that the market allocates resources,
with the role of the government or regulator restricted to providing the legal and
institutional framework, rather than interfering with the conduct of business itself, as is
often the case in C&C approaches. Economic or market incentives consist primarily of
taxes (or charges) and tradable rights.1 Before taxes can be levied or (tradable) rights
issued, the authority must have in mind some target. In the context of Figure 2, the target
is h*, which is much lower than the unregulated harvest hS. (Clearly, government
intervention is only required if h*≠hS.)

In this subsection, we present the main arguments in favor of economic
instruments found in the literature. These four arguments are often advanced to manage
polluting industries, rather than wildlife, and we will explore to what extent these insights
spill over to the realm of CITES in section 2.3.

A. Least cost approach: It is realistic to assume that agents in the economy are
heterogeneous in their ability to produce commodities – some people will be more
efficient than others in producing output, harvesting wildlife or abating emissions. It is
also realistic to assume that knowledge about the capacities of different agents is not
public; economic agents themselves have more information about their ‘true type’ than
the government, and it may not be in the interest of private parties to reveal their true type
to the regulator. If these two assumptions are met, then market incentives are more

1 The choice between a harvest tax and quota depends on uncertainty. If there is no uncertainty about the
marginal cost and marginal benefit functions (about species growth, their future value, etc.), it does not
matter whether the authority chooses a tax or quota to achieve its goal (h* in Figure 2, say). If there is
uncertainty, picking a harvest tax can lead to the ‘wrong’ level of harvests, while choosing a quantity can
result in a mistake about the forecasted ‘price’ that agents will have to pay for harvesting rights (Weitzman
1974). Such errors have social costs. However, unless high values for rights encourage ‘excessive’
poaching, it would seem that a quota is preferred from an uncertainty standpoint because, subject to the
effectiveness of policing efforts (which is also required with a tax), it guarantees that harvests do not
exceed socially desired levels.

16

efficient than across the board regulatory C&C measures. If these conditions are not met,
the regulator can ‘mimic’ the market outcome by tailoring regulations for each and every
firm/agent. However, even in such an unlikely full-information context, the
administrative requirements and costs of such regulations exceed those of market
incentives. To circumvent such costs, the regulator typically ignores heterogeneity at the
agent level and treats each agent the same (e.g., requiring each one to harvest the same
quantity). As will become clear, this comes at a cost.

The main market incentives are harvest taxes and quota trading. Harvest taxes are a
way to internalize the spillovers associated with harvesting. The optimal level of the tax is
determined by the institutional setting. When property rights are not secure, the tax should
be set at a relatively high level to account for both external effects and user costs. When
property rights are secure, resource owners already account for the user cost, such that the
optimal tax only reflects the external cost (nonuse values foregone or possibly ecological
costs) associated with harvesting at the optimal level.

Quota trading first requires the establishment of an aggregate cap (or quota) on
harvesting, followed by the ‘issuance’ of ‘rights to harvest’ (quota or permits) that can then
be traded. Tradable emission permits can be allocated to existing harvesters at no cost to
them (referred to as “grandfathering”) or sold via an auction to the highest bidders, thereby
generating revenue for the government.

The common feature of taxes and tradable rights is that, in equilibrium, harvesting
occurs by the most efficient firms/agents. With taxes, only agents with low harvesting
costs will be able to pay the tax and stay in business. The important thing to note here is
that agents themselves reveal their costs by their actions (exiting or entering the
‘industry’ by harvesting wildlife) – the regulator does not need to know anything about
individual agents. With tradable quota, harvesting rights will gravitate to firms with the
lowest marginal harvest costs. They are the ones that stand to benefit most from acquiring
these permits (as their profits per unit of harvesting are greatest), and therefore they are
able to either outbid other agents at an auction, or simply purchase quota from less
efficient firms/agents directly. Again trade among firms themselves, with superior
knowledge about their costs than the regulator, ensures that an efficient outcome
emerges. This is an important result: with taxes and quota the total harvesting costs are
minimized for a given (pre-determined) total harvest level.

B. Easy to enforce: Market incentives are no panacea. Regardless of whether the
authority chooses regulations or market incentives, or a mix of both, there is no escaping
the need to monitor agents’ behavior and enforce the regulations. Regardless of the type
of regulation chosen, agents have an incentive to ignore it and free ride on the
conservation efforts of others – over-harvest their quota, understate their catch to save on
taxes, and so on. However, it can be argued that enforcement requirements can be
lowered when property rights are created (regardless of whether they exist at the level of
the individual or the community). This will have two important effects. First, as stated
above, secure property rights will induce owners to take the user cost of extraction into
account, and therefore it diminishes the incentive to over-harvest in the short run. Second,
while access to the resource must still be enforced to restrict usage of non-owners and

17

control extraction by potential ‘cheaters’ in the case of a community-owned resource, the
costs associated with such enforcement will now to a large extent be borne by the owner
rather than the regulator. Since the resource owner likely has better knowledge about
local enforcement issues than a regulator, costs may also be lower. To some extent
similar devolution of enforcement is expected to occur when a multi-year tradable quota
system is in place, with a multi-year quota amounting to just another form of property
right. For example, when it is agreed that the aggregate cap or total harvest in any
particular season is a function of the wild stock, then it is in the interest of all quota
holders to monitor each other’s behavior and make sure that others do not over-harvest.
Allowing others to over-harvest would be to one’s own detriment as this lowers next
season’s stock and quota.

C. Dynamic incentives: Economists generally like economic incentives because
they provide agents with an incentive to adopt technical changes that lower costs.
Consider the textbook case of abating emissions. A tax or tradable emission permit
system gives an incentive to develop and adopt new and clean technologies because such
technologies will enable firms to sell permits (or avoid buying them), or avoid paying the
tax. Further, market instruments provide incentives to change products, processes and so
on, as marginal costs and benefits change over time. Because firms are always trying to
avoid the tax or paying for emission rights, they tend to respond quickly to technological
change.

D. Economic instruments may raise revenues: As mentioned under B; regulating
firms requires monitoring and enforcement – costly activities for the regulator. When
market instruments are used, some of these costs might be returned to the authority in the
form of revenue. Indeed, regulation can turn into a net revenue-raising activity, and there
is ample evidence that many environmental taxes are used exactly for this purpose in
OECD countries. It is evident that a tax system raises revenues, but creation of property
rights or tradable quota may have a similar effect. This happens when rights are
auctioned off rather than grandfathered (provided free of charge to agents already
involved in the activity). However, grandfathering of rights has a certain political appeal
since the allocation mechanism can be used to make the trading system acceptable. For
example, if it were possible to identify all harvesters of a certain wildlife species,
allocating them a certain number of harvest rights (and guaranteeing those) might cause
them to mend their ways, harvest only the allowable quota granted them, and, at the same
time, encourage them to aid in the protection of the species.

2.3 Relevance for the case of wildlife conservation

How important are these four arguments in favor of economic (market) incentives
for the particular context of wildlife conservation? First, consider the issue of least cost
harvesting. It has been demonstrated that extensive cost savings may occur in the context
of commercial fisheries management and pollution control (Weninger and Waters 2003,
Cropper and Oates 1992). But it is important to realize that commercial fishers and
polluting firms (possibly producing different goods) may constitute a much more diverse
or heterogeneous set of actors than, say, those harvesting certain wildlife species in a
developing country.

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When (i) harvesting techniques are low-tech, labor-intensive and capital-extensive,
and fairly uniform across all contributing agents, and (ii) the ecological conditions under
which the species lives, and their local densities, are rather similar across space, then the
efficiency gains in terms of equating marginal harvesting costs across agents must be
small. In other words, if marginal costs across harvesters are rather similar, the
efficiency gains from trade or EIs will be small. How “similar” are marginal harvesting
cost functions for wildlife harvesters? This should be assessed on a case-by case basis.
There are likely cases where efficiency gains from tradable quota or taxes are
considerable. When harvesting technologies vary greatly (WWI rifles versus helicopters
and high precision rifles); when habitat is diverse (common lands versus private reserves,
game ranches); or when demand for wildlife harvesting originates from different markets
(e.g. commodity versus trophy markets, nuisance harvesting), such gains can be
considerable. In contrast, when a homogenous group of people is harvesting a single
species for a common market, using similar techniques, then the gains are small. The
conclusion is that efficiency gains are context-dependent, depending on conditions of
heterogeneity. They may be smaller for the management of many wildlife species than
have been observed in many other sectors, but this is not certain.

Next, the case for dynamic efficiency (i.e., the incentive to spur technical change) is
not very compelling in the context of species harvesting. It can be debated whether
encouraging harvesting efficiency is to be applauded in this context. More important,
however, is the following observation: perhaps economic incentives will have no impact
on dynamic efficiency whatsoever. The important textbook insight discussed above
depends on the assumption that firms produce output and create pollution as a byproduct
(that can be controlled or abated at some cost). It does not spill over to the case of
wildlife conservation where the separation between product and byproduct does not exist.
By inventing more efficient harvest techniques, firms cannot sell quota (in fact, they will
presumably buy more of them), nor can they save on their tax obligations.

This implies that the main advantages of market incentives will likely be the impact
on enforcement costs and the potential to raise revenues, and possibly (in certain cases)
standard efficiency gains from adopting EIs. The former is clearly important for
conservation purposes, as emphasized above. The relevance of revenue raising is slightly
more dubious. Economists usually consider distributional issues (like transferring money
from private parties to the government) of secondary importance, but in the context of a
developing country facing difficulties in raising revenues to provide basic public goods,
this issue obviously becomes relevant as well (as already noted above). In addition, as
will become clear in section 3.2, distributional issues may be of importance when the
amount of habitat is endogenous, or chosen by local agents.

2.4 Summary

To achieve optimal use of resources typically involves regulation of users. Regulation is
necessary to internalize spillover effects and, when use or property rights are not secure,
to account for the inter-temporal user cost associated with harvesting. We argue that
defining property rights in physical or legal space is an important first step towards
optimal resource management. In this sense, EIs can often be fruitfully applied. To

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capture spillover benefits (if any) the regulator can choose either additional economic
instruments, or command and control. It has been documented in other sectors such as
commercial fisheries that adopting EIs may result in substantial efficiency gains. In this
section we argue that the scope for such additional efficiency gains may be modest in the
case of wildlife harvesting. Whether no additional regulation is preferable, or
intervention through C&C or EIs instead, should be determined at the case study level.
The costs and benefits of the various options will vary greatly, depending on
characteristics of the species, the habitat and the parties involved in harvesting.

3. ECONOMIC INSTRUMENTS: WILDLIFE HARVESTING AND HABITAT

The main threats to wildlife are introduction of exotic species (invasives),
overexploitation and habitat conversion. Trade arguably affects all three threats, for
example, by shipping species from one location to another or by changing relative prices
of factors and commodities. For trade in threatened species and/or wildlife, that is CITES,
the most important threats are overexploitation and habitat conversion. Economic
incentives may affect both the incentive to harvest species, and the incentive to convert
natural habitat for some competing purpose. We will return to these two threats in this
section.

3.1 Regulating harvesting

Under open-access no individual harvester has an economic incentive to conserve
the wildlife, and none can efficiently conserve the wildlife by delaying harvest. Doing so
will only enhance the harvest opportunities of competitors. New harvesters will be
attracted to the activity, or existing ones will expand their efforts so long as they earn
more than the (opportunity) cost of their effort. The consequence of ignoring user costs
by individuals is that all rents are dissipated, and eventually total cost equals total
revenue. Excessive hunting effort and too small resource stocks represent the
fundamental problem of open access. Various management instruments can be used to
combat rent dissipation and protect wild stocks. It will become clear that while most
instruments are theoretically able to protect stocks, only some will actually be able to
maximize resource rents.

Most textbooks on resource economics (e.g. Conrad and Clark, 1987) demonstrate
that management agencies or the CITES management authority can force harvesters to
recognize user costs by either imposing the appropriate tax on harvests (reducing
revenues) or harvesting effort (raising costs by either a license fee or effort tax). While
the resulting outcome is theoretically efficient and does not involve tedious monitoring of
effort, a few major problems remain. First and foremost is that taxes may be politically
infeasible as it transfers all of the economic rent to the government, and harvesters will
use their (political) power to prevent such a policy from being implemented. This is the
most important reason why tax policies are hardly implemented anywhere in the world to
regulate commercial fisheries (Brown 2000). Second, the authority may have difficulty in
computing the optimal tax, which depends on factors such as demand for wildlife

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products and biological processes. Taxing harvesting effort can be difficult because
fishers have an incentive to substitute types of effort that are not taxed for types that are
taxed. Finally, enforcement of a harvest tax and its collection may be difficult.

Much more common than tax schemes in actual renewable resource management
policies are quota schemes. In the case of wildlife, an annual harvest quota can be
determined from information about the species’ population dynamics and minimum
viable population, and other (economic) factors, and then allocated in some fashion. At
the national level, quota can be distributed amongst individual hunters or communities, or
hunting can remain open until the national quota is reached. While a quota system may
result in conservation of the stock and optimal harvesting levels (provided that the
authority has access to all the relevant data, and that monitoring and enforcement occur),
a quota system will not always result in efficient allocation of effort. For example, if the
hunting is opened up until the country’s quota is reached, it is possible to end up in a
situation where the wrong animals (e.g., females of child bearing age) are taken with
more effort than needed as hunters/communities rush to capture quota before others get
there first. Such rushing is likely to dissipate rents as the situation is not unlike
(controlled) open-access. The only difference is that wildlife stocks are protected from
over-exploitation by the quota.

Open-access problems can be overcome if property rights are allocated to
individuals or communities/groups. If a hunter has the right to harvest a certain quantity
in a specified time interval (say, per year), she will decide to use her effort so that harvest
costs are minimized if discounted prices are constant, for example, or that her supply is
concentrated in periods of high demand and high prices. Economic efficiency occurs at
the firm level, but from society’s point of view it is still possible to improve the
allocation of effort by allocating harvest to least-cost agents. This may be accomplished
by auctioning them off or by allowing trade in harvest rights.2 If either of these options is
implemented, the quota scheme is both efficient (maximizes resource rents) and
conserves the wild stock.

If a particular wildlife species is found in more than one country and global
management is desired (as may be the case for elephant), an overall harvest quota can be
determined and allocated among the individual countries which would then allocate quota
internally. Again each nation’s harvest rights can be traded domestically or, perhaps,
even internationally. The latter option enables society to earn further gains from trade by
exploiting international differences in harvesting cost. A condition for such a scheme to
maximize welfare, however, is that the conservation value of elephants is the same for
elephants in different countries, If this condition is violated, trading with an “exchange
rate” reflecting differences in spillover values may be introduced.

2 Quota constitutes a property right that has value. The price of quota is the value of the in situ
resource, which is simply the market price minus the marginal harvesting cost, or the scarcity
rent. Of course, enforcement of quota rights is a necessary condition for quota prices to reflect
scarcity rent. Agents with low costs will bid more for quota; likewise, if quota is tradable, low
cost ‘firms’ will buy quota from high cost ones, thereby making everyone better off. In
equilibrium, the price of transferable quota is equal to the resource rent (Anderson 1995).

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The allocation of quota can be used as a policy tool. Quota can be auctioned each
year to the highest bidders, thereby earning rents that the government can use to monitor
and enforce the scheme and fund wildlife management and habitat protection programs.
Revenue can also be used to reduce tax distortions elsewhere, or finance the provision of
other public goods. Quota can also be allocated to local communities that can then sell
the quota, harvest specimens themselves or protect them from harvest (perhaps to
enhance tourism). In this case, the local communities have a greater interest in wildlife
management than if they are left out of the system entirely.

To sum up, when backed by sufficient enforcement and monitoring effort,
economic instruments (and C&C measures alike) are capable to contribute to the
conservation of wildlife by restricting harvest effort. In theory, an ‘optimal’ or efficient
level of harvesting effort can be implemented by appropriate choice of regulatory
stringency. However, only some economic instruments (notably taxes and tradable quota)
are able to maximize the resource rent associated with harvesting.

3.2 Case study: Commercial use of The Vicuña

The vicuña (Vicugna vicugna) is one of the South American camelids along with
the guanaco (Lama guanicoe), the llama (Lama glama) and the alpaca (Lama pacos).
While vicuña and guanacos are wild, the llamas and alpacas are their domesticated
counterpart, a process of selection that appears to have started between 7,000 and 6,000
years ago. The vicuña inhabits the Andean highlands, between 3,000 and 4,600 m. Its
range currently extends over large areas of Perú (80,000), north of Argentina (23,000)
and Chile (25,000), and west of Bolivia (12,000).

Hunted for their precious wool, which is the one of the finest in the world, the
vicuña was near to extinction by the late 1960s. With the European invasion, a trade in
fibre was developed, involving the killing of the animal. The few attempts to regulate the
use of vicuñas up to this century failed and uncontrolled hunting continued until the species
reached near extinction, with just an estimated 10,000 individuals left in the 1950’s.

Vicuña wool has been long praised for its softness and fineness. Its current scarcity
also adds to the high prices commanded by the few items traded internationally. Vicuña
wool (or rather fleece) is regarded as a luxury fibre along with Alpaca, Angora, Cashmere,
Camel hair, Mohair, Musk Ox and, Yak, which are noted for their fineness, scarcity, unique
appearance and status. It is a very exclusive market, with production of all luxury fibres
representing less than 3% of annual sheepwool production by weight. Vicuña is considered
the finest and rarest of all, and its softness and colour are highly valued, commanding the
highest prices. Archaeological findings and ethnic history archives indicate three distinct
phases of interaction between the vicuña and human populations. In a first stage, in the
Arcaic era of the Central Andes (7,000–2,000 B.C.), human population in the highlands was
significant and the vicuña was popular prey of the highland hunters (Hurtado 1987).

In the second stage, from the late Arcaic area to the advent of agriculture, in parallel,
hunting and livestock rearing took place, with hunting being reduced in importance. Llamas
and alpacas, both domesticated species, provided food, wool and fuel, and the llama could

22

be used to carry loads. A more complex sociopolitical system emerged and the hunting of
vicuña was banned for religious reasons. Wool was still obtained, although this was done
through a management system imposed from political authorities. A live capture technique
called chaku was used because it allowed the shearing and release of the animal with little
impact on the population. These practices were clearly directed to the conservation and
sustainable use of resources, where the vicuña wool was only used for special robes for the
nobles and royals (Hurtado 1987).

The system, however, was affected by the European invasion, giving way to a third
phase where the planned chaku was gradually dismantled and hunting of vicuñas increased,
coupled with a regional land struggle between native communities and the Europeans. The
few attempts to regulate the use of vicuñas failed (Hurtado 1987) and uncontrolled hunting
continued until significant control measures were set in place in the 1950s by which time the
population was nearing extinction, with an estimated 10,000 individuals left (Torres 1992).

Conservation efforts to protect the vicuña started in Perú in 1969, with the creation
of the Pampa Galeras National Reserve. Subsequently, range states have coordinated
conservation efforts through several agreements. In 1969 the first agreement for the
protection of the vicuña was signed. Peru and Bolivia signed in 1969, with Argentina
joining in 1971 and Chile in 1972. The agreement banned all international and internal
trade in vicuña products and prohibited the export of fertile individuals to third parties. The
vicuña was also listed in Appendix I of CITES in 1975, ratified by all range states and
banning all international trade in the species.

These coordination efforts for conservation at the international level created a
strong base for cooperation among range states. As a result, the vicuña experienced an
impressive recovery during the last 30 years, particularly in Perú. From an estimated 6,000
over the four range countries in 1965, the vicuña reached 10,000 by 1970, 101,215 in 1983
and around 154,000 by 1992. Management areas have also increased from 248,000ha in
1965 to more than 7,289,896 ha in 1982 to some 20,800,000 ha currently under protection
status. Conservation efforts have been particularly successful in Peru and Chile, where
population levels increased significantly during the early years. However, financial and
physical requirements to effectively protect those areas have not grown at the same rate.

Although vicuñas have natural predators such as pumas and foxes, the most
important limiting factors appear to be poaching by humans and the availability of food, for
which they compete with other livestock, like llamas and alpacas. It is the first factor which
motivated governments to protect the species and ban all use; it is the second, however,
which has caused most social conflict as communities resent the competition of the vicuña
for the scarce bofedales in the highlands. Studies in Chile suggest that the vicuña population
has reached the carrying capacity of the habitat (given the existing livestock densities),
which would account for the oscillating pattern in the population levels registered since
1990 (Torres and Nuñez 1994). Every hundred vicuñas in the highlands eat the same
quantity of food as 75 alpacas, or 61 llamas or 72 sheep. The total stock of domesticated
livestock in the management zones, is estimated to be the equivalent of 51,864 heads of
vicuña. The 21,620.2ha of bofedal available in the management areas of the Parinacota
Province, is therefore estimated to be capable of supporting 25,969 vicuñas in the

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management areas of the Parinacota Province. However, in 1992 the vicuña population in
the area was estimated in 26,144, indicating that it is at carrying capacity and is in
competition with domestic livestock for food.

At the regional level, some areas show significant overstocking, as in the Lauca
National Park. As populations recovered, the competition over habitat with domestic
livestock (llamas and alpacas) increased, this being one of the factors behind the increase in
poaching. These factors made the involvement of the local communities essential for the
long term protection of the species. One way to create incentives for conservation and
protection of the vicuña at the local level was to reopen trade in vicuña wool, which can be
extracted by shearing live vicuñas with little impact on wild populations, and generating
revenue for local communities. This was the philosophy behind the second vicuña
agreement in 1979. It provided a use-based rationale for the local communities to become
interested in the conservation of the vicuna.

In 1987, Vicuña populations in the Laguna Blanca Reserve (Catamarca province)
were examined to assess their potential contribution to the indigenous peasant economy.
This is primarily a subsistence economy, with a small but increasing involvement in the
market economy. The two main sources of income are from sheep and llama spun wool.
The potential harvest of the Vicuña population was estimated using simulation techniques,
calculating the maximum sustainable yield and the carrying capacity of the area
(Rabinovich et al., 1991). If the Vicuña population were allowed to grow from its current
size of 5,000 animals to around 8,000, 15.2% of that population could be harvested each
year. The monetary value of each Vicuña is estimated at US$64: $19 for the wool, $10 for
the meat (assuming a 20kg animal fetches $0.50 per kg) and $35 for the hide. The
estimated total income that could be derived from sustainable management of the Vicuña is
US$94,464 per year. This would provide an annual household income to the peasant
community of the Laguna Blanca Reserve of almost US$1,000 if equally distributed among
the 95 families. This would clearly provide an incentive for these peasant farmers to share
their lands with the wild vicuna.

To sum up, strict conservation through use restrictions, when properly enforced,
will conserve wildlife. However, in the long run strict conservation can undermine the
stated objectives. Vicuñas and livestock compete for forage, and use restrictions on the
former remove the incentives of peasant farmers to share their land with Vicuñas.
Establishing property rights in land and wildlife provides such an incentive and therefore
represents a major step towards sustainable development. Whether tradable quota (an EI) or
non-tradable quota (a form of command and control) are implemented to regulate
management is of secondary importance. A main point of this study is that establishment of
property rights is important, and EIs or government regulation will contribute little more to
the protection of species.

3.3 Instruments and habitat conversion

The problem of wildlife conservation is intimately related to the protection of
wildlife habitat, which implies that it is bound up in land use and land ownership. In the
previous section, we examined economic instruments and incentives related to the harvest

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of wildlife. In this section, we consider wildlife habitat and land use. Of course, property
rights to wildlife, regulations concerning take and incentives to ensure that wildlife are
not over harvested affect the value of land. That is, any harvest and wildlife protection
policies that increase the value of wildlife might increase the value of land in habitat.

Economists usually consider distributional issues of secondary importance. The
focus is generally on maximizing social surplus, and whether that surplus accrues to the
regulator or private agents typically matters less. In this section, however, we argue that
distribution may be of the utmost importance for the case of wildlife conservation. The
reason is as follows. In any economy, there are agents (private or public) that decide
about land use. Supposedly such agents compare the present values of net returns from
alternative land uses – they compare the returns of habitat conservation and sustainable
resource management to those of agricultural conversion. When intervention lowers the
decision maker’s returns to habitat conservation and resource harvesting, it becomes
more likely that habitat will make place for other uses of the land.

Above we established that taxing, auctioned quota, subsidies and grandfathered
quota are equally efficient in restricting harvest effort. However, as mentioned, there is a
distributional difference. Taxing and auctions imply resource rents for the regulator,
whereas subsidies and grandfathered quota imply rents for the harvester. This translates
into different incentives to conserve habitat.

Often landowners have little incentive to protect wildlife habitat because the value
of land in habitat for agricultural producers and foresters may be very small or non-
existent. As noted earlier, wildlife and wildlife habitat are a public good and private
landowners have little if any incentive to protect wildlife habitat on their land. Indeed, as
the enactment of the Endangered Species Act in the United States has demonstrated and
as we argue further below, landowners may have every incentive to do the opposite –
convert habitat to crops. Therefore, economic instruments are required to ‘encourage’
landowners to protect wildlife habitat.

In many political jurisdictions, rural land continues to be largely publicly owned,
or, if not owned outright, agricultural and other users of rural land have ill-defined or
weak property rights. Peasants lack property rights to wildlife and often gain the right to
land only by actively farming it. Even productive forestland might be sacrificed and
wildlife habitat lost because peasants cannot demonstrate ownership of land unless they
‘develop’ it – that is, conduct cropping or grazing activities – and this inevitably results in
conflicts with wildlife. The appropriate assignment (and protection) of property rights to
undeveloped land that also serves as wildlife habitat might encourage peasants not to
develop it; peasant landowners might earn a living through sustainable, small-scale
forestry and/or harvest of non-timber forest products, including wildlife if they are given
a right to animals on their property.

However, this would require a change in the way most developing countries
allocate land and other property rights. Moreover, and importantly, the earnings from
habitat and wildlife exploitation must exceed that of agriculture. While government
policies related to wildlife (and forestry) can affect returns, it can be the case that such

25

activities cannot compete with cropping, even supposing that the ‘correct’ institutions
were in place to enable landowners the rights to all the products produced on their land.
When the social benefits of habitat conversion exceed the social benefits of conservation
(including international positive external effects), economists recommend conversion of
natural lands into alternative uses.

The most interesting case exists where habitat conservation “does not pay” from a
private perspective, but would be optimal from a social (global) perspective. In other
words, when the positive external effects associated with conservation of habitat and
wildlife are sufficiently great to topple the balance from conversion to conservation. In
this case economic instruments can be used to encourage private landowners or land
users to take into account the negative external effects of their land-use decisions on
wildlife. What instruments might be employed that directly affect land management?

Regulation

Regulations specify what landowners can and cannot do on their land. The
Endangered Species Act is an example of regulation in that it prohibits destruction of the
habitat of wildlife on private land. Regulatory approaches often entail expensive
monitoring and enforcement, and can still be ineffective if social norms and formal rules
do not coincide (Ostrom 1990; Nielson 2003). In fact, it is possible that regulations may
lead to perverse incentives that discourage conservation (‘shoot, shovel, and shut up’) if
restrictions on established property right owners are onerous (Polasky 2001).

Taxes and subsidies

Tax incentives can be designed to give farmers an incentive to protect wildlife
habitat on farmland. However, evidence from developing countries indicates that tax
policies are not, by themselves, capable of compensating rural landowners for providing a
public good (wildlife habitat) at private expense. As evidence has accumulated that
preferential tax assessments do more to subsidize farmland owners than to conserve
farmland, governments have increasingly initiated programs to purchase development
rights and conservation easements (Wiebe et al. 1996). These programs involve
separating and purchasing some but not all of an owner’s rights to a property: separated
rights might include, for example, the right to build residential or commercial buildings,
to drain sloughs, to burn associated uplands, or to remove endangered species of trees. In
the United States, most purchases have been in the form of agricultural conservation
easements that restrict residential, commercial or industrial uses, but that allow active
farming (Hardie et al. 2004).

Subsidies are perhaps better than tax incentives for protecting nature on
agricultural lands. In developed countries, subsidies are used to take land out of
production, keep extant wetlands or other critical wildlife habitat from being converted to
agriculture, or establish wildlife habitat through tree planting, plant of dense nesting
cover for migratory waterfowl, et cetera. Similar programs can be used in developing

26

countries, although financing such programs will pose a greater challenge and likely
prevent them from being implemented.

The subsidy approach most-often mentioned in the literature is that of
compensating farmers for losses from wildlife depredation. While not providing
incentives to prevent legal and illegal taking of wildlife, compensation may at least
reduce the incentives of local peasants to go out and destroy wildlife to prevent the
agricultural damage that they may cause. On the other hand, wildlife damage programs
may encourage additional conversion of habitat into cropland as they essentially amount
to a subsidy to agriculture (Rondeau and Bulte 2003).

Finally, one way to arrive at a globally optimal level of habitat conversion is
through subsidies at the international level. Fair compensation for positive external
effects of conservation implies a transfer flow from North to South. While some of this
could presumably be arranged through NGO involvement (see below) and current
opportunities provided by the Global Environmental Facility (GEF), it is an open
question whether this is enough to safeguard sufficiently large areas of nature in the long
run. The public good characteristics of nature conservation, and the implied incentives to
free ride on other’s efforts, could mean additional, cooperative efforts, should be
undertaken. One can think of large-scale programs to finance the provision of ecological
services (such as now pioneered in Costa Rica), funded through taxation in the North.

Transfer of development rights

Transferable development rights and wildlife habitat banking constitute cases
where separation of development rights can be integrated with land use planning.
Wildlife habitat banking (WHB) allows landowners to develop wildlife habitat on their
property if they have sufficient credits from investment in the completed rehabilitation of
a WHB site. Land use planning enters this program through the designation of the WHB
sites (see Fernandez and Karp 1998). Sites can be chosen that provide large high-quality
habitats with superior potential to sustain desired ecosystems. Given good choices, the
investments in the WHB can provide greater community-wide environmental benefits
than equivalent investments in the maintenance of habitat on sites that are being
developed. Good planning is crucial to obtain higher benefits, because WHB is a ‘no net
loss’ program that links area restored to wildlife habitat area removed by conversion of
habitat to agriculture (Hardie et al. 2004).

An important difference between preferential tax assessments and purchase of
development rights is the potential role of planning. Preferential tax assessments are
typically extended to all eligible landowners regardless of the location of their property.
However, purchases can be targeted to sites where the social or environmental benefits
are deemed to be particularly high, such as along a wildlife corridor or within a region
under particular agricultural pressure. While the potential for targeting exists, it generally
is not realized (see Hardie et al. 2004). Zoning-based transferable development right
(TDR) programs are initiated by dividing an area that is being opened for agricultural
conversion, or one that has already been converted, into a zone where agricultural
development is permitted and one where agriculture is limited or prohibited entirely,

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thereby protecting crucial habitat. The government partially takes private property rights
in ‘down-zoned’ area in order to protect an environmental amenity – wildlife habitat (see
Johnston and Madison 1997; Hardie et al. 2004). When the down-zoning occurs,
landowners in the affected (source) areas are granted the option to sell the separated
development rights to landowners in designated agricultural development (‘up-zoned’)
areas or sinks. It is the owners of property in the up-zoned or target areas that must
purchase the transferable development rights in order to be able to farm their land.
Landowners who lose property rights are compensated in a development rights market,
but at rates driven by the opportunity costs created by zoning instead of by willingness to
pay for cropland. Of course, governments incur costs of planning and administration of
such a TDR program, and the TDR system is only meant to make the separated zoning
politically palatable. It is unlikely that this type of instrument will work to protect
wildlife habitat in developing countries unless property rights of all kinds are made
stronger (see section 4).

One variant that might work in areas where forest concessionaires are active is to
require the forest companies to purchase TDRs from landowners who have been down
zoned. That is, a forest concessionaire would be required to purchase a certain number of
TDRs that protect wildlife habitat in exchange for the right to harvest a certain volume of
timber.

Direct purchase of conservation easements to protect wildlife habitat also
constitutes a form of property rights purchase. In this case, the state simply purchases the
right to develop land for agriculture from the landowner. Since this might be too costly
for many developing countries, one alternative is to permit NGOs (or even foreign
governments) to purchase these rights, as indicated above. Like the case of TDRs, this
option requires that economic institutions exist so that development rights can be
separated from ownership of land (and that ownership of land is well defined and
protected by the courts – see section 4 below). There may also be opposition to the idea
of selling development rights to foreigners, whether foreign governments or NGOs.

The problem is that there is no guarantee that earnings from (perhaps marginal)
agricultural land are sufficient to enable compensation to take place. If this is the case,
illegal conversion of all land capable of producing crops will still occur. It is not a simple
matter to construct a land protection scheme that includes restrictions on land use with
transferable development rights to compensate losers (see van Kooten 1993). In
developing countries, the obstacles standing in the way of implementing such a scheme
may again be too large to surmount.

Transferring income through NGO involvement

The private sector might also be relied upon to a greater extent than currently.
Environmental NGOs are perhaps the best means for transferring wildlife conservation
funds from rich to poor countries. Nonprofit private land trusts, such as the Nature
Conservancy, The Conservation Fund and the Trust for Public Land have become active
in the conservation of open space and wildlands in the United States (see Hardie et al.
2004). These organizations purchase properties or easements on lands that provide

28

environmental benefits (such as wildlife habitat) and seek to protect land slated for urban
development. Purchased land may be turned over to state and/or local governments, but
might be managed by the NGOs in order to guarantee that contributors in developed
countries receive the non-market amenity values purchased in developing countries
where the record of government management of public lands is perhaps not as good. Of
course, for this option to work, it is important that property rights are clearly delineated
and protected by the courts in the developing countries. NGOs are unlikely to purchase
property or wildlife easements on land if these property rights are non-enforceable.

Kontoleon and Swanson (2003) have shown that, in the context of giant panda
preservation (in the Wolong reserve, China), the non-use values associated with panda
conservation in the “wild” are sufficiently large to warrant setting aside extensive
stretches of land as a reserve – such that not only the flagship ‘panda’, but many other
species as well can be supported. However, when such elusive non-use values are not
backed up by true transfer flows, it will be in the interest of local people to allocate the
land to other uses. Capturing and channeling non-use values through international
transfer payments, perhaps actual purchase or lease of land by environmental NGOs, may
be one good means to protect species.

3.4 Summary

In this section we, again, demonstrate that defining property rights and benefit-sharing
programs are vital in promoting conservation of wildlife. We show that EIs are in theory
capable of maximizing resource rents, but argue that their main role could be in
promoting habitat conservation. There are various EIs that can be used to make sure that
habitat conservation occurs at the lowest cost (tradable development rights, habitat
conversion taxes). Equally important, to our opinion, will be the use of international EIs
that capture and channel nonuse values from North to South, and to promote habitat
conservation through transfers and subsidies.

4. IMPLEMENTING ECONOMIC INSTRUMENTS TO PROTECT WILDLIFE

What is the scope for adopting EIs in developing countries to promote
conservation of wildlife and enable a transition towards sustainable development? We
argue that the perspective is mixed. EIs are not a panacea, and it is an open question
whether they can be effectively employed in all contexts. Institutions and social capital
are important if economic incentives are to be used to manage and protect wildlife
populations. For example, in their review of emissions trading, Tietenberg et al. (1998)
indicate that it is impossible to institute any system of emissions trading unless the
requisite legal and other institutions are in place for monitoring, measuring, certifying
and enforcing trades, and that lack of appropriate institutions is probably the most
important obstacle to the use of market incentives for addressing climate change. For a
democratic market economy to function properly, or for market-oriented economic
policies to have effect, three criteria or factors other than markets and private property are

29

required (Fukuyama 2002). These criteria relate to economic institutions, the role of the
state, and culture.

While a full-fledged analysis of these issues is far beyond the scope of the current
study, we would like to note that it is by no means guaranteed that the current state of
economic institutions (be it formal or informal) and governments in many resource-rich
countries is sufficient to exploit the gains from employing EIs. This can be illustrated for
the case of elephant harvesting and ivory trade. In Table 3 we summarize key
institutional indicators for (i) OECD countries, (ii) Asian consuming states, and (iii) main
ivory producers. An examination of the Table suggests that the prospects of
implementing EIs in producer states are not promising. By all measures, range states are
the least capable of preventing illegal harvests and sales of ivory. They lack the required
economic institutions (courts, rule of law, government effectiveness) and social capital
(control of corruption) for enforcing and policing ivory trade. Establishing the
infrastructure to guide successful implementation of EIs comes at a cost that is unknown.

Table 3: Measures of the Effectiveness of Economic Institutions and Levels of Social
Capital in Industrial Countries, Ivory Importing States and Elephant Range States,
2000-2001

Measure Eight Industrialized
Countries

Five Major Asian
Buyer States

30 Range States
(Africa & Asia)

Voice & Accountability 1.453 0.106 -0.563

Political Stability 1.275 0.971 -0.801

Government
Effectiveness

1.586 1.048 -0.625

Regulatory Quality 1.165 0.899 -0.3

37

Rule of Law 1.628 1.073 -0.516

Control of Corruption 1.878 0.946 -0.524

Source: World Bank (2002) and calculation

5. CONCLUSIONS AND POLICY RECOMMENDATIONS

Economic instruments have great potential to address spillovers associated with wildlife
management. Economic incentives appear particularly useful for the following reasons.
First, they are theoretically able to achieve objectives at the lowest cost. They encourage
efficient use of resources and therefore have the smallest impact on economic growth and
development. The flexibility, efficiency and cost-effectiveness associated with the use of

30

EIs is presumably nowhere more important than in developing countries. Second,
instruments such as auctioned tradable quota and taxes are capable of generating
government revenues, enabling the government to provide a variety of public goods. In
countries where the scope for raising revenues is small because of limited administrative
capacity, this effect could be important. The administrative requirements of EIs are
different – for example, raising revenues through auctioning off trophy quotas is much
easier than raising funds through taxing households involved in harvesting of a species.
Third, compared to command and control measures, the information requirements of EIs
are modest. C&C requires the planner to make decisions that allocate resources across
activities. In contrast, by simply providing a setting or context, “the market” will take
care of an efficient allocation of resources when EIs are used. EIs have lower institutional
and human resource requirements than C&C – an important advantage in an information-
sparse environment.

In spite of these advantages, EIs are not a panacea that can be straightforwardly
implemented across the board. The following critical comments are a useful reminder of
the key restrictions. First off, and perhaps obviously, EIs are not a substitute for
monitoring and enforcement. Conservation of wildlife will critically depend on these
activities, regardless of whether C&C or EIs are used to allocate resources. (However, as
argued in section 2, public authorities may be able to shift the burden of enforcement and
monitoring onto private agents if property rights to resources are defined in either a
physical or legal sense.) Without effective monitoring and enforcement, economic
incentives cannot work – prices paid for tradable quota will be too low and could
approach zero (as harvesting without quota is also feasible) and tax evasion will occur on
a large scale. One potential advantage of some EIs is that they can generate the resources
that are required to support the enforcement effort that is needed to enable the EIs to
work. However, there may be circumstances where adequate enforcement and monitoring
is too costly. In that case, it may be optimal to opt for the private optimum rather than the
social one; that is, establish property rights to wildlife and ignore spillover benefits
associated with conservation. While harvest levels will be “too large” in the short run,
such that species stocks will be “too low” eventually, the costs associated with this
imperfection may be small compared to the costs of achieving the first best outcome.

Next, a key obstacle to the implementation of EIs is the availability of economic
institutions and social capital in many wildlife-rich countries. As of yet it is unclear what
“minimum” level of institutional infrastructure is necessary to successfully implement
EIs, and how this minimum level compares to actual scores in key countries.
Nevertheless, it seems plausible to argue that many countries are currently not up to the
task to implement and guide a full-fledged tradable quota or tax scheme to regulate the
use of wildlife species.

This is not to say that nothing can be done to greatly improve the efficiency and
sustainability of wildlife harvesting. While full-fledged implementation of EIs is
cumbersome and expensive, it is often possible to make substantial improvement by
making small steps forward. Specifically, by defining and protecting property rights to
land and wildlife (be it at the level of the individual or a well-defined group of users,
depending on the context), resource harvesters will be able to reach the privately optimal

31

level of resource harvesting and conservation. While inferior to the socially optimal level
of harvesting and conservation, it arguably represents a significant improvement over the
unregulated open access outcome that eventuates when property rights do not exist. To
complement the management scheme, other instruments can be applied after property
rights have been established. This would internalize any external effects. However,
whether making this additional step is warranted from a cost-benefit perspective is
something that has to be assessed on a case-by case level.

We believe, but have not analyzed, that the scope for using complementary EIs in
regulating harvest levels may be modest (but there clearly will be cases where this is not
true and where the gains from implementing EIs to regulate harvesting are large). The
efficiency gains from EIs may be modest, for example, because harvesting technologies
are sometimes fairly homogenous, suggesting little scope for gains from trade. The
greatest perspective for implementing EIs, we believe, is with respect to land use and
habitat conversion. Specifically, it seems advisable to closely consider the scope of
implementing an international transfer system from North to South to compensate for
transboundary spillover benefits from conservation. Current transfer systems are rather
ad hoc, and certainly incomplete. Whether the political will exists in the North to fund
such an effort, and whether the institutional capacity exists in the South to manage the
subsidy flows, are relevant matters that must be faced.

To reiterate our caveat from the introduction: the current study has been written
under great time pressure, and is therefore certainly incomplete, possibly even in key
respects. We would therefore like to suggest some useful directions for follow-up
research, aimed at closing the gap between what is known today and what needs to be
known before EIs can be usefully implemented.

1. Gains from implementing EIs: what are the potential efficiency gains from tax
or tradable quota systems? How much heterogeneity exists among harvesters for key
wildlife species (say: crocodiles)? How do the gains from complementary instruments
compare to the gains from securing property rights?

2. Costs of implementing EIs: What are minimum institutional requirements to
successfully implement EIs? Do many countries currently have the ability in terms of
social capital and institutions to do so (and if not, what are the associated costs of
establishing such an institutional infrastructure)?

3. Costs and benefits: how do the costs and benefits of implementing a system of
EIs compare for key wildlife species? Are there general lessons to be drawn?

4. What are the prospects for using EIs in global trade in species? How does a
wildlife quota system work across countries? What institutions are required and how
many countries satisfy institutional requirements for implementing EIs? How do
dynamics affect the usual “optimal tariff” or “optimal quota” result?

5. How important are international positive effects of wildlife conservation at the
margin (compared to domestic benefits of sustainable wildlife management)? What is the

32

scope for capturing such benefits to promote habitat conservation in the South, and how
should this be organized?

6. Operational issues: how can one define and allocate property rights, and how
can one implement a tax or tradable quota scheme? How high should the tax be (or how
large the total allowable catch) in light of many real-life uncertainties? Are there many
parallels with ITQ experiences in fisheries in developed countries, and if so: how can we
exploit them? How much income should be allocated to wildlife management? Is there a
role for eco-labeling?

6. REFERENCES

Anderson, L.G. 1995. Privatizing Open-access Fisheries: Individual Transferable Quotas.
Chap. 20 in The Handbook of Environmental Economics (pp. 453–74) edited by
D. W. Bromley. Cambridge MA: Basil Blackwell.

Barichello, R.R., R.M. Porter and G.C. van Kooten. 1995. Institutions, Economic
Incentives and Sustainable Rural Land Use in British Columbia. Managing
Natural Resources in British Columbia. A. Scott, J. Robinson and D. Cohen, eds.,
pp.7-53. Vancouver: UBC Press.

Bernstein, Steven, 2001. The Compromise of Liberal Environmentalism. New York:
Columbia University Press.

Borregaard, Nicola, Beatriz Bustos and Matthew Stilwell,2003. The Role of Economic
Instruments in the Context of Biodiversity Related Multilateral Environmental
Agreements. Discussion paper commissioned by UNEP. September. 97pp.

Bovenberg, A.L. and L.H. Goulder, 1996. Optimal Environmental Taxation in the
Presence of Other Taxes: General-Equilibrium Analysis, American Economic
Review 86(4): 985-1000.

Bromley, D.W. 1999. Sustaining development: Environmental resources in developing
countries. Cheltenham, UK: Edward Elgar.

Brown, G.M., 2000. Renewable Natural Resource Management and Use Without
Markets, Journal of Economic Literature 38: 875-914

Bulte, E.H., D.P. van Soest, G.C. van Kooten and R. Schipper, 2002. Forest Conservation
in Costa Rica: Optimal Forest Stocks under Uncertainty and Rising Nonuse
Benefits, Am. J. of Agric. Econ. 84(February): 150-60.

Clark, Colin W. 1990. Mathematical Bioeconomics. 2nd edition. New York: Wiley.

Conrad, Jon M. 1995. Bioeconomic Models of the Fishery. Chapter 18 in The Handbook
of Environmental Economics edited by D. W. Bromley. Cambridge, MA: Basil
Blackwell. .

33

Conrad, Jon M. and Colin W. Clark. 1987. Natural Resource Economics: Notes and
Problems. Cambridge: Cambridge University Press.

CPB (Netherlands Bureau for Economic Policy Analysis), 1997. Challenging
Neighbours. Rethinking German and Dutch Economic Institutions. Berlin:
Springer.

Dasgupta, Partha. 2000. Economic Progress and the Idea of Social Capital. Social
Capital. A Multifaceted Perspective. P. Dasgupta and I. Serageldin, eds., pp. 325-
424. Washington: The World Bank.

Dasgupta, P. and I. Serageldin (Eds.), 2000. Social Capital. A Multifaceted Perspective.
Washington, DC: The World Bank.

Easterly, William. 2001. The Elusive Quest for Growth. Cambridge, MA: The MIT Press.

Ekins, P. and T. Barker, 2001. Carbon Taxes and Carbon Emissions Trading, Journal of
Economic Surveys 15(3): 325-52.

Fernandez, Linda and Larry Karp, 1998. Restoring Wetlands Through Wetlands
Mitigation Banks, Environmental and Resource Economics 12(October): 323-
344.

Field, Barry and Nancy Olewiler, 2002. Environmental Economics. 2nd Canadian
Edition. Toronto: McGraw-Hill Ryerson.

Fisher, Anthony C., 1981. Resource and Environmental Economics. Cambridge, UK:
Cambridge University Press.

Fukuyama, Francis, 1995. Trust. The Social Virtues and the Creation of Prosperity. New
York: The Free Press.

Fukuyama, Francis, 1999. The Great Disruption. Human Nature and the Reconstitution
of Social Order. New York: The Free Press.

Fukuyama, Francis, 2002. Social Capital and Development: The Coming Agenda, SAIS
Review 22(Winter-Spring): 23-37.

Gelauff, G.M.M. 2003. Sociaal kapitaal in de economie, Economisch Statistische
Berichten Vol. 88, Nr. 4398, March 27, pp.3-5.

Hardie, I.W., P.J. Parks and G.C. van Kooten, 2004. The Economics of Land Use at the
Intensive and Extensive Margins. In International Yearbook of Environmental &
Resource Economics, 2003/2004 edited by H. Folmer and T. Tietenberg.
Cheltenham, UK: Edward Elgar. In press.

34

Harrison, David Jr., 2002. Tradable Permits for Air Quality and Climate Change. In
International Yearbook of Environmental & Resource Economics, 2002/2003
edited by T. Tietenberg and H. Folmer. Cheltenham, UK: Edward Elgar: 311-372.

Hutton, J., P. Ross and G. Webb, 2001. Using the Market to Create Incentives for the
Conservation of Crocodilians: A Review, Report for the IUCN/SSC Crocodile
Specialist Group

Johnston, Robert A. and Mary E. Madison, 1997. From Landmarks to Landscapes: A
Review of Current practices in Transfer of Development Rights, Journal of the
American Planning Association 63(3): 365-378.

Kinyua, Patrick I.D., 1998. Game Ranching in Machakos District, Kenya: An
Application of Mathematical Programming to the Study of Wildlife Policy.
Unpublished Ph.D. Dissertation, Department of Forest Resources Management,
University of British Columbia, Vancouver, Canada.

Knack, Stephen and Philip Keefer, 1997. Does Social Capital have an Economic Payoff?
A Cross-Country Investigation, Quarterly Journal of Economics 112: 1251-1288.

La Porta, R., F. Lopez-de-Silanes, A. Shleifer and R. W. Vishny. 1997. Trust in Large
Organizations, American Economic Review 87(May): 333-38.

La Porta, R., F. Lopez-de-Silanes, A. Shleifer and R. W.Vishny. 1999. The Quality of the
Government, Journal of Law, Economics & Organization 15(April): 222 –279.

Loasby, B.J., 1999. Knowledge, Institutions and Evolution in Economics. London, UK:
Routledge.

McKitrick, Ross and Randall M. Wigle, 2002. The Kyoto Protocol: Canada’s Risky Rush
to Judgment. C.D. Howe Institute Commentary No. 169. Toronto: C.D. Howe
Institute. October.

Menon, Vivek, 2002. Tusker. The Story of the Asian Elephant. New Delhi: Penguin
Books India.

Ménard, C. 1995. Markets as Institutions versus Organizations as Markets? Disentangling
Some Fundamental Concepts, Journal of Economic Behavior & Organization
28(October): 161-182.

Murty, M.N., 1994. Management of Common Property Resources: Limits to Voluntary
Action, Environmental and Resource Economics 4: 581-94.

Neher, Philip A., 1990. Natural Resource Economics: Conservation and Exploitation.
Cambridge UK: Cambridge University Press.

Nielsen, J.R. 2003. An analytical framework for studying compliance and legitimacy in
fisheries management, Marine Policy 27: 425-432.

35

Newell, Richard G. and William A. Pizer, 2003. Discounting the Distant Future: How
Much do Uncertain Rates Increase Valuations? Journal of Environmental
Economics and Management 46: 52-71.

North, D.C. 1990. Institutions, institutional change and economic performance.
Cambridge, UK: Cambridge University Press.

North, Douglass C. 1984. Transaction Costs, Institutions, and Economic History, Journal
of Institutional and Theoretical Economics 140(1984): 7-17.

Olson, M. Jr. 1996. Distinguished Lecture on Economics in Government: Big Bills Left
on the Sidewalk: Why Some Nations are Rich, and Others Poor, Journal of
Economic Perspectives 10(Spring): 3-24.

Ostrom, Elinor, 2000. Social capital: A Fad or a Fundamental Concept? Social Capital. A
Multifaceted Perspective. P. Dasgupta and I. Serageldin, eds., pp. 172-214.
Washington: The World Bank.

Ostrom, Elinor, 1998. A Behavioral Approach to the Rational-Choice Theory of
Collective Action, American Political Science Review 92(March): 1-22.

Ostrom, Elinor. 1990. Governing the Commons: The Evolution of Collective Action.
Cambridge University Press.

Parry, I., R.C. Williams III and L.H. Goulder, 1999. When Can Carbon Abatement
Policies Increase Welfare? The Fundamental Role of Distorted Factor Markets,
Journal of Environmental Economics and Management 37: 52-84.

Pizer, W.A., 1997. Prices vs. Quantities Revisited: The Case of Climate Change.
Resources for the Future Discussion Paper 98-02, Washington, October. 52pp.

Prins, H., J. Grootenhuis and T. Dolan 2002. Wildlife Conservation by Sustainable Use.
Dordrecht: Kluwer Academic Publishers

Polasky, Stephen. 2001. Investment, information collection, and endangered species
conservation on private land. In Protecting Endangered Species in the United
States: Biological Needs, Political Realities, and Economic Choices edited by
Jason Shogren and John Tschirhart. New York: Cambridge University Press.

Putnam, R.D. 1993, March 21. The prosperous community: Social capital and public life.
The American Prospect 4(13), 35-42.

Richards, Kenneth R., 2000. Framing Environmental Policy Instrument Choice, Duke
Environmental Law & Policy Forum 10(2): 221-285.

36

Segerson, K. and N.Li, 1999. Voluntary Approaches to Environmental Protection.
Chapter 7 in The International Yearbook of Environmental and Resource
Economics 1999/2000 (pp.273-306) edited by H. Folmer and T. Tietenberg.
Cheltenham, UK: Edward Elgar.

Shleifer, A. and R.W. Vishny. 1998. The Grabbing Hand: Government Pathologies and
their Cures. Cambridge, MA: Harvard University Press.

Stavins, Robert N., 2002. Lessons from the American Experience with Market-Based
Environmental Policies. In Harnessing the Hurricane: The Challenge of Market-
Based Governance edited by J.D. Donahue and J.S. Nye Jr. New York: Brookings
Institution Press.

Stavins, Robert N. and Scott Barrett, 2002. Increasing Participation and Compliance in
International Climate Change Agreements. FEEM Working Paper No.
94.2002/Kennedy School of Government Working Paper No. RWP02-031.
http://www.feem.it/web/activ/ _wp.html 35pp.

Tahvonen, O. and J. Kuuluvainen. 1995. The Economics of Natural Resource Utilization.
In Principles of Environmental and Resource Economics: A Guide for Students
and Decision Makers edited by H. Folmer, H. L. Gabel and H. Opschoor.
Aldershot: Edward Elgar.

Tietenberg, Tom, Michael Grubb, Axel Michaelowa, Byron Swift and ZhongXiang
Zhang, 1998. International Rules for Greenhouse Gas Emissions Trading:
Defining the Principles, Modalities, Rules and Guidelines for Verification,
Reporting and Accountability. United Nations Publication
UNCTAD/GDS/GFSB/Misc.6. As viewed on July 10, 2003 at:
http://r0.unctad.org/ghg/publications/intl_rules . 123pp. No original date or
city of publication provided.

van Kooten, G.C., 2002. Economic Analysis of the Canada-United States Softwood
Lumber Dispute: Playing the Quota Game, Forest Science 48: 712-721.

van Kooten, G.C, 1993. Land Resource Economics and Sustainable Development:
Economic Policies and the Common Good. Vancouver: UBC Press.

van Kooten, G.C. and E.H. Bulte, 2000. The Economics of Nature. Oxford, UK:
Blackwell.

Wiebe, K., A. Tegene and B. Kuhn, 1996. Partial Interests in Land: Policy Tools for
Resource Use and Conservation. Agricultural Economic Report Number 744,
Economic Research Service, United States Department of Agriculture,
Washington DC. 58pp.

Weitzman, M.L., 1974. Prices vs Quantities, The Review of Economic Studies
41(October): 477-91.

37

Weitzman, M.L., 2002. Land Fees vs Harvest Quotas with Uncertain Fish Stocks,
Journal of Environmental Economics and Management 43: 325-38.

Williamson, O.E. 1996. The Mechanisms of Governance. New York: Oxford University
Press.

Williamson, O.E. 1998. Transaction Cost Economics: How it Works; Where it is Headed
De Economist 146: 23-58.

Woerdman, Edwin., 2002. Implementing the Kyoto Mechanisms: Political Barriers and
Path Dependency. Ph.D. Dissertation, University of Groningen, The Netherlands.
620pp. As viewed on 30 July at: http://www.ub.rug.nl/eldoc/dis/jur/e.woerdman.

Woolcock, M. 1997. Social capital and economic development: Toward a theoretical
synthesis and policy framework. Theory and Society 27: 151-208.

World Bank. 2002. Data Profiles & Country at a Glance Tables. As found July 25 at:
http://www.worldbank.org/data/countrydata/countrydata.html.

Required Assignment 1
Due Date	Friday of Week 8, May 25
Conceptual Background	Read Ariagada and Perrings. 2011. Paying for International Environmental Public Goods. Ambio 40:798-806 Bulte, E., G. van Kooten, and T.Swanson. 2003. Economic Incentives and Wildlife Conservation. Working Paper.
Assignment	The intent of the assignment is to ensure that you understand the conceptual framework for the rest of the class. 4-page paper (typed, double spaced, 12 Arial font, 1” margins) discussing incentives to conserve marine biodiversity within the framework of impure public goods. Discuss what an impure public good is, the types of externalities associated with impure public goods, the technology of public good supply (best shot, weakest link, etc.), and the types of economic incentives (positive and negative) that are created for impure public goods with different technologies of public good supply.