Changes in the ecosystem services provided by forests and their economic valuation: A review

Abstract

In this chapter we discuss the trends in forest change and the associated drivers, the economic value of forests, the principles and challenges in evaluating the economic value of forests, and the role of valuation in informing decision-making. We address current major forest conservation initiatives at different scales and the mechanisms involved, whether supported by economic valuation or not. Today, 30 % of the world's forests are designated for productive functions, 24 % for multiple uses, 11.5 % for biodiversity conservation, 8.2 % for protective functions, and 3.7 % for social functions. The remaining 22.6 % are designated for other uses or remain unclassified. Global trends indicate that although the area of intensively managed forest continues to expand, the global extent of conservation and protective forests is also increasing as a result of political efforts to preserve and restore the ecological functions of forests. Forest management practices are potentially better supported by extended cost-benefit analyses that require an economic valuation of the whole array of benefits, whether market or non-market, provided by forests. Although we acknowledge other values and decision-making and support tools, the focus of the chapter is on the economic valuation approach. Our review in this chapter was guided by the goal of updating previous reviews of these topics. We have provided additional evidence that forests contribute to human well-being in many ways, and use the concept of ecosystem services as a building block to better understand, frame, and assess the economic benefits we derive from well-functioning forests. © 2014 Springer Science+Business Media New York. All rights reserved.

Full text

107
J.C. Azevedo et al. (eds.), Forest Landscapes and Global Change:
Challenges for Research and Management, DOI 10.1007/978-1-4939-0953-7_5,
© Springer Science+Business Media New York 2014
Abstract In this chapter we discuss the trends in forest change and the associated
drivers, the economic value of forests, the principles and challenges in evaluating
the economic value of forests, and the role of valuation in informing decision- making.
We address current major forest conservation initiatives at different scales and the
mechanisms involved, whether supported by economic valuation or not. Today,
30 % of the world’s forests are designated for productive functions, 24 % for
multiple uses, 11.5 % for biodiversity conservation, 8.2 % for protective functions,
and 3.7 % for social functions. The remaining 22.6 % are designated for other uses
or remain unclassifi ed. Global trends indicate that although the area of intensively
managed forest continues to expand, the global extent of conservation and protec-
tive forests is also increasing as a result of political efforts to preserve and restore
the ecological functions of forests. Forest management practices are potentially bet-
ter supported by extended cost–benefi t analyses that require an economic valuation
of the whole array of benefi ts, whether market or non-market, provided by forests.
Although we acknowledge other values and decision-making and support tools, the
focus of the chapter is on the economic valuation approach. Our review in this chapter
Chapter 5
Changes in the ecosystem services provided
by forests and their economic valuation:
a review
Cristina Marta-Pedroso , Lia Laporta , Vânia Proença ,
João C. Azevedo , and Tiago Domingos
C. Marta-Pedroso (*) • J. C. Azevedo
CIMO—Centro de Investigação de Montanha, Escola Superior Agrária, Instituto Politécnico
de Bragança , Campus de Santa Apolónia, Apartado 1172 , 5301-855 Bragança , Portugal
e-mail: cmartapedroso@gmail.com
L. Laporta
IN+, Center for Innovation, Technology and Policy Research, Environment and Energy
Section, DEM, Instituto Superior Técnico, Universidade Técnica de Lisboa ,
Avenida Rovisco Pais, 1 , 1049-001 Lisbon , Portugal
CIBIO, Centro de Investigação em Biodiversidade e Recursos Genéticos, Universidade do
Porto , Campus Agrário de Vairão , 4485-601 Vairão , Portugal
V. Proença • T. Domingos
IN+, Center for Innovation, Technology and Policy Research, Environment and Energy
Section, DEM, Instituto Superior Técnico, Universidade Técnica de Lisboa ,
Avenida Rovisco Pais, 1 , 1049-001 Lisbon , Portugal

108
was guided by the goal of updating previous reviews of these topics. We have provided
additional evidence that forests contribute to human well-being in many ways, and
use the concept of ecosystem services as a building block to better understand,
frame, and assess the economic benefi ts we derive from well-functioning forests.
5.1 Forest ecosystem services
Ecosystem services are broadly defi ned as the benefi ts that people obtain, directly or
indirectly, from ecosystems (MEA 2005 ), and that contribute to human well- being
when combined with other factors such as education, health care, and social equity.
The interdependency between ecological and social systems can be seen as a
feedback loop: human well-being depends on the delivery of ecosystem services,
but the capacity of ecosystems to deliver services depends on ecosystem conditions,
which in turn are affected by society’s choices about how to use ecosystem services
and manage the ecosystems that provide them. These choices are greatly infl uenced
by the level of human well-being and by the way society perceives and values eco-
system services.
Forests provide many ecosystem services, including supporting a large percent-
age of the world’s biodiversity and contributing to human well-being at local (e.g.,
wood production), regional (e.g., groundwater recharge), and global (e.g., climate
regulation) scales. The most easily understood and most quantifi able source of ben-
efi ts derived from well-functioning forests pertains to the provision of goods and
materials, even if their provision is not directly observed. Water, a basic and valu-
able good required for human existence, is a suitable fi rst example. Though it is not
obvious to the untrained eye, forests interact closely with and affect the hydrologi-
cal cycle through evapotranspiration and their ability to increase infi ltration into the
soil by decreasing runoff; thus, forests are a key source of freshwater resources
(Wang and Fu 2013 ). For example, about 80 % of the freshwater resources in the
United States at the turn of the century originated from forests, which covered, at
that time, about one-third of the country’s surface area (USDA 1999 ). Human use
and management of forest ecosystems can change the level of ecosystem services
delivery and induce the production of one service to the detriment of others. This is
the case for productive forests, which are planted and managed to produce timber,
and the case for protective forests, which are planted or managed to prevent or
reduce soil erosion.
Less quantifi able benefi ts that are often inadequately addressed include the ben-
efi ts people obtain from forests through abstract concepts such as esthetic , spiritual ,
and inspirational values—which are called cultural services. Unlike timber production
or soil erosion control, these benefi ts are not physically measurable. Instead, they
take the form of experiences people can obtain from forests (Kareiva et al. 2011 ).
Because they are intangible, communicating this category of benefi ts is more diffi cult,
even when attempts are made to express the values in monetary terms. For forest
ecosystems, a signifi cant part of these benefi ts relates to recreational opportunities.
C. Marta-Pedroso et al.

109
In southern Africa, for instance, trees play a crucial role in the cultural and spiritual
lives of local communities (Sileshi et al. 2007 ), despite any hypothetical benefi t they
provide as tourist attractions. The inherent complexity of valuing people’s experi-
ences is well acknowledged in the literature (Boyd and Banzhaf 2007 ). We examine
the methods for valuing ecosystem services more closely in Sect. 4 .
5.2 Classifi cation of ecosystem services
and conceptual approaches
Exhaustively listing the whole array of benefi ts people obtain from ecosystems can
be a challenging task, and some sort of labeling and operationalization of the
concept was required at the beginning of efforts to conceptualize ecosystem ser-
vices. Multiple classifi cation systems for ecosystem services have evolved, and this
variety has been justifi ed by the premise that classifi cation systems should focus on
the purpose and context of the study (Costanza 2008 ). Notwithstanding, one of the
most generally used classifi cations was described by the Millennium Ecosystem
Assessment (MEA 2005 ), which distinguishes among four categories of services,
which MEA defi nes as “the benefi ts that people derive from ecosystems”: provi-
sioning services (products obtained from ecosystems), regulating services (benefi ts
obtained from the self-regulation of ecosystems), cultural services (non-material
benefi ts obtained from ecosystems), and supporting services (services that are
necessary for the production of other ecosystem services). Moreover, in this classi-
fi cation, biodiversity is understood as not only underpinning the ecosystem services
but also as an ecosystem service itself; for example, medicinal plants are a provi-
sioning service, whereas bird-watching is a cultural service.
Although this classifi cation is still in use and it is generally accepted, it has some
drawbacks. In particular, the consideration of supporting services as a separate
category often leads to overlapping estimates and double-counting; this issue is a
particular concern if economic valuation is to be undertaken, as we discuss later in
the chapter. Another ecosystem service classifi cation emerged from a more recent
global initiative, The Economics of Ecosystems and Biodiversity (TEEB 2010 ).
Although TEEB is similar to the MEA classifi cation, TEEB considers the support-
ing services only as ecological processes, and introduces a new category called
“habitat” services (TEEB 2010 ) to highlight the importance of ecosystems in pro-
viding habitat for migratory species and as gene-pool protectors. The TEEB classi-
fi cation and its approach differ from the MEA approach because TEEB explicitly
aims to incorporate an economic analysis of changes in ecosystem services.
The Common International Classifi cation of Ecosystem Services (Haines-Young
and Potschin 2010 ) does not aim to replace existing classifi cations, but rather
provides a framework that enables translation among different classifi cations and
links to other classifi cation systems that are used in economic and environmental
accounting. Each classifi cation has its own purposes, drawbacks, and advantages.
The MEA and TEEB approaches are directed more at assessment and valuation of
5 Changes in the ecosystem services provided by forests and their economic…

110
ecosystem services, whereas the CICES approach was conceived as a system
compatible with the design of integrated environmental and economic accounting
methods (Maes et al. 2013 ). Those who advocate the use of CICES have pointed out
that this classifi cation at least potentially helps to overcome the problem of double-
counting. This topic has been widely addressed in the literature (e.g., Boyd and
Banzhaf 2007 , Fisher et al. 2009 , Mace and Bateman 2011 ), with debate focusing
on the need to distinguish between services and benefi ts when economic valuation
or environmental accounting is the purpose of the study. In essence, and regardless
of the specifi c nomenclature adopted by each author in explaining their rationale,
valuation should only be applied to things that are directly consumed by benefi ciaries
given that the values of ecological processes are already embedded in that fi nal
output. Some argue (Bateman et al. 2010 ) that if other input capitals are used to
generate a benefi t, they should be subtracted from the estimated value of the benefi t
to provide the net benefi t.
Perhaps more important than fi nding a sovereign and unifying classifi cation system
or approach, we argue in this chapter that a deep understanding of the ecological
dynamics of forest ecosystems is necessary to generate powerful insights into details
of the chain of benefi t delivery, and can therefore help managers to identify the best
management options based on a more fully informed economic valuation.
5.3 Global trends and drivers of forest ecosystem services
5.3.1 Past, current, and future trends for forest systems
Human activity has caused the loss of about 40 % of the planet’s original forests
since preagricultural times, starting ca. 8000 years ago (Shvidenko et al. 2005 ).
Temperate regions, such as Europe and North America, were particularly affected,
losing more than 50 % of their natural forest cover before the mid-twentieth century
(Fig. 5.1 ; Kaplan et al. 2009 , MEA 2005 ). In tropical regions, the loss of forest
cover has been less severe, but became a pervasive trend during the last half-century
and will probably continue during the twenty-fi rst century (Fig. 5.1 ; FAO 2012 ,
MEA 2005 ).
In temperate and boreal regions, laws and policies to protect forests and to
reverse deforestation emerged as a response to the shortage of timber and fuelwood
and to the degradation of the forest’s protective functions (Farrell et al. 2000 , Rudel
et al. 2005 ). Moreover, rural abandonment due to economic growth, improvements
in agricultural effi ciency, and the replacement of wood by fossil fuels as a source of
energy also decreased pressures on forests (FAO 2012 , Kaplan et al. 2009 ).
Reforestation initiatives in the twentieth century, but also a few centuries ago, and
natural forest regeneration following land abandonment restored much of the forest
cover and helped to halt forest decline in temperate and boreal regions (Hobbs and
Cramer
2007 , Keenleyside et al. 2010 , Rudel et al. 2005 ).
C. Marta-Pedroso et al.

111
Globally, the overall extent of natural forests continues to decline, and the
expansion of new forests does not compensate for the loss of natural ones (Butchart
et al. 2010 , FAO 2011 ). Moreover, the value lost with the degradation or deforesta-
tion of old-growth forests cannot be fully replaced by new forests because planted
and regenerated forests differ from natural stands in many characteristics (Rey
Benayas et al. 2009 ). First, restored forests do not support the same biotic commu-
nities as old-growth forests (Hobbs and Cramer
2007 , Rey Benayas et al. 2009 ).
Second, most new forests are located in temperate regions and cannot replace the
biodiversity lost in highly diverse tropical regions; that is, the creation of forests in
one region may not compensate for the destruction of forests in another region.
Third, many planted forests are grown and managed for industrial purposes, so their
contribution to biodiversity conservation and to the delivery of regulating and cul-
tural services is modest or even negative (Kanowski
2003 ; Proença et al. 2010a , b ).
For instance, when continuous forest plantations replace traditional landscape
mosaics, there is a loss of landscape heterogeneity and a decline, or even local
Figure 5.1 Past forest losses and projected future losses in the world’s main forest systems. The
proportion of the forest lost before 1950 was estimated based on the potential distribution of each
forest system based on soil and climatic conditions. Projections of forest loss correspond to the
average value of projections obtained for the four Millennium Ecosystem Assessment future sce-
narios; error bars indicate the range of values for the four future scenarios. Adapted from the
Millennium Ecosystem Assessment (MEA
2005 )
5 Changes in the ecosystem services provided by forests and their economic…

112
extinction, of species associated with open habitats such as grasslands and meadows
(Poyatos et al. 2003 , Reino et al. 2009 ).
The global area of forest will probably continue to decline if countries fail to
provide adequate incentives to halt deforestation. Global policy choices infl uence
society’s choices and may play a critical role in determining the selection of land- use
options. A recent study by Wise et al. ( 2009 ) explored the effect of different carbon
taxation policies on global land-use changes. The authors found that imposing a
global carbon tax covering anthropogenic carbon emissions from all sectors, includ-
ing emissions from land-use change, would promote the protection and expansion of
forests, leading to an increase in forest cover (Fig.
5.2 ; the MiniCAM B scenario).
However, taxing only fossil fuel and industry emissions may prompt the expan-
sion of biofuels and lead to a drastic loss of forest cover worldwide (Fig. 5.2 ;
MiniCAM C scenario). Previous scenarios, which were also based on socioeco-
nomic drivers, projected less drastic changes in forest cover (Fig. 5.2 ; MEA, GBO2,
and GEO4 scenarios). The narrower range of variation in forest area projected by
these scenarios is in part explained by compensatory mechanisms in the underlying
Figure 5.2 Projected changes in global forest cover until 2050 under various global scenarios: the
Millennium Ecosystem Assessment (MEA) scenarios (Sala et al.
2005 ), the Global Biodiversity
Outlook 2 scenarios (ten Brink et al.
2006 ), the Global Environmental Outlook 4 scenarios (UNEP
2007 ), the representative concentration pathway scenarios (Hurtt et al. 2009 ), and the MiniCAM
scenarios (Wise et al.
2009 ). For each set of scenarios, we have only shown the two most contrast-
ing results. The wider envelope for the MiniCAM projections, compared to the envelope of sce-
narios with the IMAGE model (IMAGE-team
2001 ), suggests that there are opportunities for
action to reverse the global trend of forest decline, but also that wrong policy choices can exacer-
bate the loss of forest cover compared with the other scenario assessments. Sources : Leadley et al.
(
2010 ), Pereira et al. ( 2010 )
C. Marta-Pedroso et al.

113
socioeconomic scenarios, which lead to a convergence of the trend lines for changes
in forest cover area. For instance, the option for biofuels in the “environmentally
friendly” scenarios implies the replacement of forests by biofuel crops (Leadley
et al. 2010 , Pereira et al. 2010 ).
Forest use and deforestation in the northern hemisphere are historically associ-
ated not only with socioeconomic development but also with ecosystem degradation,
a shortage of forest products, and environmental disasters, such as fl ooding, which
later motivated forest restoration and sustainable forest management in these regions
(FAO 2012 ). Today, tropical forests are the ones most exposed to deforestation and
forest degradation. The unsustainable use of forest resources jeopardizes socioeco-
nomic development and human well-being in these regions (Rodrigues et al. 2009 ),
and its negative impacts may also be felt at larger spatial and temporal scales
(Leadley et al. 2010 ). Halting unsustainable use trends requires action from local to
global levels and the adoption of socioeconomic development pathways that will
ensure the sustainable use of forests, including the management of both tangible and
non-tangible services, and sustainable support for human welfare. The management
of forests and forest ecosystem services, and particularly non-provisioning services,
should be based on polycentric and diverse governance systems that ensure an
equitable representation of users and that promote knowledge sharing, collective
decision-making, and enforcement of management decisions (Ostrom 2009 ).
5.3.2 Global trends in the use of forest ecosystem services
From a utilitarian perspective, natural forests have by default a multifunctional
nature in the sense that they can simultaneously deliver several benefi ts to people,
including forest goods (e.g., fuelwood, medicinal herbs, bush meat), regulating
services (e.g., climate regulation, soil protection), and cultural benefi ts (e.g., esthetic
pleasure, sacred groves). Still, some forests are designated and managed for a par-
ticular function, such as industrial plantations and forests in protected areas. Today,
30 % of the world’s forests are primarily designated for productive functions
(production of wood, fi ber, biomass, and non-wood forest products), 24 % for
multiple uses (forests managed to deliver a range of benefi ts without the dominance
of a particular function), 11.5 % for the conservation of biodiversity, 8.2 % for
protective functions (conservation of ecosystem functions and processes underlying
the delivery of regulating services), and 3.7 % for social functions (recreation, edu-
cation, and conservation of cultural heritage) (FAO 2011 ). The remaining 22.6 %
are designated for other uses or remain unclassifi ed.
The annual rate of growth between 2000 and 2010 was particularly high in
regions with a low proportion of conservation forests compared with the global
average (e.g., a 3.3 % increase in East Asia) but also in Europe (3.9 %, excluding
the Russian Federation) and in South America (4.8 %). Most regions have already
set aside 10 to 20 % of their forest for biodiversity conservation purposes (FAO
2011 ).
The region with the highest proportion of conservation forest is Central America
5 Changes in the ecosystem services provided by forests and their economic…

114
(47 %), whereas East Asia, West and Central Asia, and the Russian Federation have
designated less than 10 % of their forest for conservation. Globally, 463 × 10
6 ha of
forest have been primarily dedicated to biodiversity conservation (Fig. 5.3a ).
Nevertheless, protective forests are now emerging as a tool to conserve ecosys-
tem functioning and manage the delivery of regulating services. For instance, sev-
eral Asian countries have reported a high proportion of protective forest (Fig. 5.3b ).
This is particularly the case for China, where a large area of forest has been planted
with the main purpose of controlling desertifi cation (Cao et al. 2011 ), and for sev-
eral western Asian countries in arid zones, such as Turkmenistan and Uzbekistan,
where water is a critical resource (FAO 2011 ). The same pattern is found for African
countries in arid zones, such as Libya and Kenya.
Globally, the area of forest designated primarily for productive functions
(Fig. 5.3c ) has decreased at an annual rate of 0.2 % from 2000 to 2010, and cur-
rently covers 1200 × 10 6 ha (FAO 2011 ). This reduction in area is in part explained
by the increase in the area of forest dedicated to intensive forestry but also because
some areas previously classifi ed as productive forests were reclassifi ed as multiple-
use forests (FAO 2011 ). Europe is the region that has reported the largest proportion
of areas designated primarily for productive functions (52 % or 57 % excluding the
Russian Federation), and North and South America have reported the lowest proportions
Figure 5.3 Proportions of the forest area designated for ( a ) biodiversity conservation, ( b ) water
and soil protection, ( c ) production of forest products, and ( d ) multiple uses in 2010, and recent
trends. Proportions are indicated by the color gradient, with darker tones indicating higher pro-
portions. Trends are indicated by color, with blue indicating an increase in the designated area
from 2000 to 2010 and red indicating a decrease. Note that the proportion does not indicate the
extent (area) of forest in a country, and that the trends do not indicate the rate of change. Source :
FAO (
2011 )
C. Marta-Pedroso et al.

115
(14 %, but with heterogeneity within the region; for example, 1 % in Canada
compared with 29 % in the United States) (FAO 2011 ). Despite the global decrease
in the area designated for productive functions, the pattern is heterogeneous, show-
ing a pattern of interspersed areas with increasing, decreasing, and stable trends.
In addition to areas specifi cally designated for productive purposes, many forest
products are obtained illegally or informally from areas that are not classifi ed as
productive. This implies that the real area used for the extraction of forest products
is much larger than the area that is formally designated as productive forest.
Moreover, multiple-use forests (Fig. 5.3d ) also encompass productive functions.
Currently, the area designated for multiple uses totals 949 × 10
6 ha globally, and
increased by 10 × 10
6 ha between 1990 and 2010 (FAO 2011 ). Global and regional
trends are heterogeneous, refl ecting different types of transitions, including shifts in
the classifi cation from productive to multiple use and vice versa but also shifts from
undesignated to multiple use.
Social forests, which are primarily used for recreation, environmental education,
and preservation of cultural heritage, are still infrequent, despite the widespread use
of forests for outdoor activities and their cultural role as natural heritage sites. The
social function of forests is usually associated with conservation forests or with
multifunctional forests. Today, only 3.7 % of the world’s forests are designated
primarily for this purpose, but available data suggests that this proportion is
increasing (FAO 2011 ). Brazil has the largest area of social forest, at 119 × 10
6 ha
(i.e., more than 75 % of the global area of social forest), and this forest is designated
for the protection of indigenous peoples and their culture.
The statistical data used in this section was reported by individual countries and
gathered together for the Global Forest Resources Assessment 2010 (FAO 2011 ),
which is the most comprehensive assessment to date. However, two main sources of
uncertainty should be considered when comparing countries or regions. First, there
are disparities among the reporting countries in terms of data availability, either due
to real data gaps or due to differences in national forest inventory methodologies.
Second, the criteria used to defi ne forest categories and functions are subject to dif-
ferent interpretations by the reporting countries. Also note that the designation of a
forest for a particular purpose does not imply the existence of sustainable manage-
ment practices or even of a management plan for that forest.
5.3.3 Drivers of change and impacts on forest ecosystem
services
The Millennium Ecosystem Assessment identifi ed fi ve main direct drivers of bio-
diversity and ecosystem change (MEA 2005 ): habitat change, climate change, inva-
sive species, overexploitation, and pollution. The impacts of these drivers and their
trends vary across the globe, and affect forest biomes differently. Direct drivers are
often shaped by social demand for provisioning services, including both forest
provisioning services and farmland services when agriculture replaces forest use.
5 Changes in the ecosystem services provided by forests and their economic…

116
For example, population growth can cause a higher demand for food and fi ber and
can therefore lead to production activities that cause deforestation. On the other
hand, global policies for climate mitigation can encourage forest conservation, and
technological advances can improve the effi ciency of forestry and agricultural
production, thereby lessening the pressure on natural forests.
Pollution became an important driver in the last century in many forms, and
particularly in the forms of excessive nutrient loading in production systems and of
industrial emissions (Shvidenko et al. 2005 ). Boreal forests have been particularly
badly affected by air pollution from industrial sources during the last century, with
reported events of signifi cant tree damage and mortality (Shvidenko et al. 2005 ).
When combined with climate change, pollution is expected to have a serious impact
on the condition of these forests during the twenty-fi rst century. Climate change not
only will affect tree physiology and phenology but will also affect the fi re regime by
increasing the frequency and severity of wildfi res as a consequence of drier and
hotter summers (Soja et al. 2007 , Stocks et al. 1998 ).
Habitat change and overexploitation were the main drivers of forest change in
temperate regions during the last century. Today, the effect of these drivers is declin-
ing as new forests are planted and regenerate in abandoned fi elds and pastures.
On the other hand, tropical forests have been particularly affected by land-use
change during the last century, and the impact of this driver is expected to increase
in the twenty-fi rst century as forests are replaced by pasture and cropland (in part to
respond to international demand for food) but also by infrastructure and urban areas.
Overexploitation of forest goods is also expected to intensify due to population
growth in these regions, as well as logging driven by international demand (Davidson
et al. 2012 , Lambin et al. 2003 ). Overall, the impacts of climate change will increase
during the twenty-fi rst century in all biomes (Leadley et al. 2010 , MEA 2005 ). The
impact of invasive species is also expected to increase due to global trade and travel,
as well as the impact of pollution, in particular due to a signifi cant intensifi cation in
the fl ow of reactive nitrogen into the environment (MEA 2005 ).
The effects of drivers are often synergistic. Changes caused by a driver or by a
set of drivers may create the conditions for triggering, intensifying, or maintaining
other drivers, rendering the control of their impacts diffi cult (Lambin et al. 2003 ,
Leadley et al. 2010 ). In some situations, these interactions lead to regime shifts,
with strong impacts on ecosystem structure and functioning. Although researchers
can identify tipping-point changes and their potential risks, their dynamics are com-
plex and diffi cult to predict (Leadley et al. 2010 ). Tipping points can be broadly
defi ned as events that occur when an ecological threshold is passed, leading to shifts
in ecosystem functioning that signifi cantly affect biodiversity and ecosystem
services. Tipping-point changes tend to be fast due to reinforcing feedbacks that
amplify the effects of drivers or due to abrupt shifts when thresholds are crossed.
They also tend to be diffi cult to reverse due to feedback loops that trap systems in
undesirable stable states and long lag times between a driver’s action and its impacts,
which hamper policy decisions (Leadley et al. 2010 , in press). The Amazonian,
Mediterranean, and boreal forests present important examples of potential regime
shifts in forest systems.
C. Marta-Pedroso et al.

117
In the Amazon basin, forest conversion coupled with climatic changes may lead
to a regime shift that will have impacts from local to global scales (Davidson et al.
2012 , Nobre et al. 2010 ). Deforestation, logging, and forest fi re are inducing
regional climate changes, including less rainfall and increased frequency and sever-
ity of drought, which increase the susceptibility of forests to fi re, thereby creating a
feedback loop that sustains fi re occurrence, promoting further forest damage and
fragmentation. In addition, projections from climate models indicate that long-term
global climate change will amplify drought in the Amazon region due to a combina-
tion of climate warming and less precipitation. At moderate to high rates of defor-
estation, the interaction between land-use change, fi re, and climate change may lead
to a feedback loop that will be diffi cult to control and that may cause extensive
forest loss (Davidson et al. 2012 , Vergara and Scholz 2011 ). Carbon release due to
this deforestation will also contribute to global climate change and will aggravate
climate-change impacts at the regional scale.
Consequences for ecosystem services and biodiversity will be severe. The
Amazon is one of the world’s largest carbon pools and carbon sinks, with the excep-
tion of dry years, when forests becomes a carbon source (Davidson et al. 2012 ,
Phillips et al. 2009 ). The shift from a carbon sink to a carbon source will contribute
to global warming and cause negative impacts at a global level. The Amazon is also
a biodiversity center (Pereira et al. 2012 ), and loss of Amazonian forest will result
in a severe loss of biodiversity at a global scale. In addition, there is the risk of losing
species, many still unknown, that have medical and pharmacological value, and
consequently a risk of losing the opportunity to fi nd and develop new medications
and vaccines. At local and regional scales, local communities will be affected by the
loss of forest goods, including food, fi ber, fuel, and medicinal plants; by the loss of
regulating services, such as climate regulation, fi re regulation, and fl ood regulation;
and by the loss of cultural services, since the forest environment is a major compo-
nent of the cultural heritage and way of life of local peoples.
In the Mediterranean region of southern Europe, land-use change, fi re distur-
bance, and climate change are interacting to create conditions suitable for a shift in
ecosystem composition (Proença and Pereira 2010 ). Rural abandonment is driving
land-use change in marginal areas of farmland, through the regeneration and
encroachment of natural vegetation and the expansion of fi re-prone forest planta-
tions, thus promoting fuel continuity in the landscape. The accumulation of biomass
coupled with frequent (anthropogenic) fi re ignition is causing a change in the fi re
regime, with more frequent and severe fi res. This situation is further aggravated by
climate change, in particular by hotter and drier summers. All these factors cause an
increase in fi re risk and promote the expansion of fi re-prone communities, such as
shrublands, which then create the conditions for the establishment of a feedback
loop that inhibits the progression of natural succession towards regeneration of nat-
ural forests. Under some circumstances, alien invasive species gain competitive
advantages in the burned areas, letting them replace native species and impoverish-
ing natural communities (Keeley et al. 2003 , 2005 ). This may eventually lead to a
compositional shift that will be hard to reverse.
5 Changes in the ecosystem services provided by forests and their economic…

118
Boreal forests provide a third example of regime shifts. In this case, climate
change is leading to a warming trend that is moving northward, creating an environ-
ment unsuitable for boreal species. On the one hand, these species may not be able
to respond to this change because their natural rate of propagation is too slow for
them to migrate north and also because tundra sites into which the boreal species
will be forced to migrate might be unsuitable for their establishment and growth
(Lloyd et al. 2011 , Soja 2007). Chapter 2 of this book discusses these issues in more
detail. But on the other hand, where tundra sites are suitable for boreal species,
changes in soil albedo, the melting of snow cover, and retreating permafrost will
allow tree establishment and forest invasion into the tundra (Soja 2007). The main
mechanism underlying this change is an amplifi er feedback loop driven by increas-
ing summer temperatures (Fernandez-Manjarrés and Leadley 2010 ). Warmer
temperatures lead to earlier melting of the snow cover, exposing soil with a lower
albedo that traps more solar radiation over a larger period. Snow cover creates an
insulating effect that is critical for the maintenance of permafrost; the loss of snow
cover causes permafrost degradation and increases the warming effect, thereby pro-
moting further snow melting. These changes will have impacts on the lives of local
people, who will have to adapt to a changing landscape (e.g., travel routes may
become unsafe due to decreasing ice stability), but will also have impacts at a global
level due to the release of large quantities of carbon and methane stored in the per-
mafrost, with consequences for climate change (Fernandez-Manjarrés and Leadley
2010 , Schaefer et al. 2012 ). Moreover, some parts of the boreal forest are being
increasingly affected by fi res driven by climate change, which also causes the
release of carbon stored in the trees and soil in addition to the loss of old-growth
forest and other social and ecological losses (Soja et al. 2007 ).
As the demand for non-provisioning services (e.g., soil protection from erosion,
water purifi cation, recreation) increases, it counteracts the economic bias towards
provisioning services. Direct drivers will be gradually affected by indirect drivers in
response to this demand. This may include national to global policies but will also
include changes in markets. In the past, and to a certain extent, still today, economic
choices tended to disregard non-market services and promote the expansion and
overexploitation of productive forests or the replacement of forest by more profi t-
able land uses. The incorporation of the benefi ts delivered by non-provisioning
services in economic choices is likely to reshape market demand and its effect on
the direct drivers of change, thereby promoting forest conservation and restoration
to preserve forest’s regulating services and cultural value.
5.4 Economic valuation of forest ecosystem services
Forests can provide multiple benefi ts to society other than wood, with the whole
array of benefi ts depending on the characteristics of the forest and the prevailing
management strategies (Duncker et al.
2012 ). This understanding is a prominent
feature of the current literature and is usually associated with the concept of
C. Marta-Pedroso et al.

119
multifunctional forests (e.g., Carvalho-Ribeiro et al. 2010 , Gustafsson et al. 2012 ).
One possible approach to capture the contribution of forest ecosystems to humans
is through an improved understanding of the linkage between the functioning of the
ecological system, which is perceived as a composite of processes and structures,
and the functioning of the socioeconomic system. The crucial role that natural
systems play in underpinning economic activity and human well-being is of grow-
ing concern (Bateman et al. 2010 ). Thus, economic valuation of ecosystems and
their services has been receiving increasing attention in the literature.
Economic valuation is not the only approach to assigning a value to nature, nor
is it necessarily the best approach; for examples of other forms of valuation, see
Oksanen ( 1997 ), Martín López et al. ( 2012 ), and TEEB ( 2010 ). As Kareiva et al.
( 2011 ) pointed out, it is important to emphasize that an economic valuation does
not replace or ignore the intrinsic value of nature, nor does it reduce the moral
imperative to conserve nature. Following the logic of Martín-López et al. ( 2009 )
and Mace and Bateman ( 2011 ), we note the importance of combining economic
and other valuation approaches to provide a more holistic picture of the value of
forests. Nonetheless, in this chapter we will focus on the economic valuation
approach. The primary role of economic analysis is to assist decision-making
(Daily et al. 2000 , Pearce et al. 1989 , Tietenberg 1996 ). In the context of forest
management, the high rate of deforestation we are facing globally—13 × 10
6 ha per
year (FAO 2007 )—and the rise of international concern about the consequences of
deforestation together mean that economic valuation of forest ecosystem services
has an important role to play.
Before jumping into the principles and methodological details of economic
valuation, we will briefl y illustrate how economic valuation of a forest ecosystem
can restrain deforestation. As we noted earlier, forests provide many non-market
goods, such as watershed protection. Landowners seek profi t maximization, and in
the absence of other mechanisms, they rely on existing markets to pursue this goal.
Existing markets defi ne their costs and revenues. Hence, even though we know that
clearing the forest would increase problems such as downstream fl ooding and sedi-
mentation, these costs do not accrue to the landowner who will decide whether to
harvest the forest; thus, these costs are not factored into the landowner’s decision.
This is clearly a market failure from a larger perspective. Economic valuation can
mitigate this problem if the analysis allows for an extended accounting of benefi ts
and costs and, based on this more complete picture, fosters mechanisms such as
subsidies, taxes, direct payments, and payments for ecosystem services that can
prevent the market failure and reduce the likelihood of deforestation. For a concise
review of market-based mechanisms, see Pagiola et al. ( 2002 ). These mechanisms
aim to fully internalize the benefi ts and costs that do not accrue directly to landown-
ers but rather that affect other groups in society. In Sect. 4.1 , we further explain the
occurrence of externalities and market failure from a conceptual point of view.
At this point, and before we begin discussing the principles of economic valu-
ation, we want to emphasize that the value of forest ecosystem services refl ects
the different ways in which they satisfy human needs. This can be considered
from the perspective of the total economic value (TEV) taxonomy (Pearce
1993 ).
5 Changes in the ecosystem services provided by forests and their economic…

120
This taxonomy defi nes the different sources of values that people may attach to
the different services provided by a given ecosystem. Note that this taxonomy
relies on whether ecosystem services satisfy human needs directly or indirectly.
Economic value, then, is a measure of the degree of satisfaction provided by these
services. The TEV approach and terminology are not uniform across the litera-
ture, but TEV generally includes the following value components: direct use,
indirect use, option, and non-use. The fi rst three categories are generally referred
to together as use values, and the non-use values often aggregate values such as
bequest and existence values.
Among the use values, direct use values include services that are used directly,
and include provisioning services (e.g., forest goods) and cultural services (e.g.,
recreation opportunities). Indirect use values include services that are indirectly
used, such as the benefi ts derived from regulating services (e.g., climate regulation).
Non-use values are divided into bequest and existence values, and are almost
entirely associated with cultural services. Bequest values represent the value that an
individual assigns to an ecosystem or species due to its relevance to the well-being
of future generations. Existence value, on the other hand, represents the value that
an individual assigns to an ecosystem or species due to its personal relevance at the
present time. In other words, it is the satisfaction this individual derives from know-
ing that a certain species or ecosystem exists.
Option values include all values (both use and non-use) that are expected to be
enjoyed in the future (e.g., provision of genetic resources, maintenance of a gene
pool for bioprospecting, cultural heritage). Note that the option and bequest values
both refl ect the importance that people give to maintaining or restoring ecosystems
in order to ensure the delivery of ecosystem services in the future.
5.4.1 Principles of economic value estimation
The economic value of an ecosystem service refers to the contribution of a certain
ecosystem functional dynamic to human well-being. Many ecosystem services are
only obtained because of other capital inputs; for instance, agricultural production
of food implies the use of machinery and labor together with the use of natural
resources and ecosystem processes. Hence, as pointed out by Bateman et al. ( 2010 ),
estimating the economic value of ecosystem services requires isolation of the eco-
system function’s contribution before the value can be converted into a monetary
metric. This suggests that it is also necessary to clarify how economic analysis dif-
fers from fi nancial analysis: the former examines society as a whole, whereas the
latter focuses on particular groups within society. Hence, when estimating the eco-
nomic value of an ecosystem service, we must account for the costs (private and
external) of producing the service and for the benefi ts (private and external) gener-
ated by it. Here, “external” refers to externalities, whether benefi ts or costs, that are
generated as unintended by-products of an economic activity that do not accrue to
the parties involved in the activity, and for which no compensation is provided.
C. Marta-Pedroso et al.

121
Depending on its impact on a third party, an externality may be positive (e.g., the
creation of a forest landscape) or negative (e.g., the creation of fragmentation).
We fi rst approach the economic foundation of ecosystem service valuation by
considering a well-defi ned market in which ecosystem services can be traded and in
which there are no external costs or benefi ts. There are two building blocks in the
process of estimating economic value: consumer and producer surpluses, with the
social surplus equaling the sum of these two surpluses. See Mankiw ( 2008 ) for a
discussion of this topic. These measures are illustrated in Figure 5.1 for the case of
an ecosystem service for which there is a market, such as timber (a forest provision-
ing service), based on the assumption of a perfectly competitive market. The timber
market is in equilibrium when demand ( D ) equals supply ( S ) at price P . The demand
curve shows consumer marginal willingness to pay (WTP), which represents the
consumer’s WTP for each additional unit of a product. The supply curve shows the
marginal costs of harvesting timber, which represents the producer’s marginal will-
ingness to accept (WTA) a given price for their product.
Figure 5.4 tells us that buyers who value the good more than the price (repre-
sented by the line segment AB ) choose to buy the good and receive a surplus of
benefi t: the area of the triangle ABP defi nes the magnitude of the consumer surplus.
This represents the amount a buyer is willing to pay for a good, minus the amount
the buyer actually pays for it, or, in different words, the benefi t that buyers receive
from participating in the market. Buyers who value the good less than the price
(represented by the line segment BE ) choose not to buy the good or receive its ben-
efi ts. Symmetrically, on the production side, those sellers whose costs are less than
the price (represented by the line segment CB ) choose to produce (in this case, to
harvest) and then sell the good (wood). Sellers receive a surplus given by the area of
Figure 5.4 Social surplus
[ ABC ] for a forest good such
as timber under perfect
market conditions. D demand
curve, S supply curve, P price
where D = S at point B , Q p
the
quantity at price P
5 Changes in the ecosystem services provided by forests and their economic…

122
the triangle PCB ; this represents the amount a seller is paid, minus the cost of
production. The producer’s surplus measures the benefi t sellers receive from partici-
pating in the market. Sellers whose costs are greater than the price (represented by
the line segment BD ) do not sell the good or receive benefi ts from the sale. The
social benefi t (i.e., the overall surplus) in this case equals the private benefi t, which
equals the sum of the consumer and producer surpluses (i.e., the area of triangle
ABC ). The social surplus is of interest in economic analysis because it concerns the
net benefi ts that society as a whole derives from the good. Mathematically, the total
or social surplus can be expressed as follows:
Social surplus value to buyers amount paid by buyers
amount rec
=
()
+
–
eeived by sellers costs beared by sellers–
()
é
ë
êù
û
ú
Thus far, we have analyzed a situation in which the social benefi t equals the
private benefi ts. However, when there are external costs and benefi ts, the private
surpluses do not equal the social surpluses. Let’s again consider a timber market,
but in the presence of external benefi ts, based on the example provided by Hanley
and Barbier ( 2009 ). Consider a sustainable timber harvester, with sustainability
here defi ned as a state of non-declining well-being, as defi ned by Tietenberg
( 1996 ; pp. 33–34). This harvester manages their land in a wildlife-friendly manner,
thereby improving the ecological quality of their woods and overall forest health by
(among other things) creating many habitats for birds and butterfl ies. They also
harvest timber for sale. The market rewards them for their timber production, since
they can sell the timber to interested buyers. But the market is unlikely to reward
them for their “production” of wildlife habitats, even though these habitats might be
valued by society. Although this is not the forum for further discussion of this topic,
these types of services fall into the category of public goods (non-excludable and
non- rival in consumption). See Boardman et al. ( 2001 ) for further explanation.
In Figure 5.5 , D m is the market demand curve for timber, S is the supply curve for
Figure 5.5 Social surplus
( ABDF ) for a good (e.g.,
timber) in the presence of a
positive external effect (e.g.,
provision of wildlife habitat).
S supply curve, D m market
demand for timber, D s
society’s demand for timber
plus its external benefi ts, Q m
the quantity at price P m , Q s
the quantity at price P s . Grey
area ( BDE ) represents
welfare lost under perfect
market conditions (e.g.,
without government
intervention)
C. Marta-Pedroso et al.

123
timber, and D s is society’s demand for timber plus its external benefi ts (in this example,
“production” of wildlife habitats). The market reaches equilibrium at point B .
At this point, both timber consumers and the rest of society receive a benefi t (an
external benefi t whose magnitude equals the vertical distance between D m and D s )
which is represented by the line segment BD . The social surplus obtained when
quantity Q m is produced therefore equals the area of the polygon ABDF .
Notwithstanding, the social optimum would be reached at point E , where the
marginal social cost (which, as no negative externality is being considered, is the
same as the marginal private cost) is equal to the marginal social benefi t. Under
perfect market conditions (e.g., without government intervention), quantity Q s will
not be supplied because harvesters are not rewarded for producing such quantity.
The area of triangle BDE (grey shade) represents the welfare loss under these
market conditions. Many ecosystem services are externalities, in the sense that the
benefi t or cost they represent to society is generated as a consequence of standard
ecosystem management but it is not intentionally produced, and it does not accrue
benefi ts or costs to the producer. This means that, for instance, the value of timber
does not refl ect the array of benefi ts that may be jointly provided by forests to
society as a whole. Often, in the literature, these externalities are referred to as
non-market ecosystem services. As we discuss in the next section, several methods
have been developed to estimate the value of such ecosystem services.
5.4.2 Economic valuation methods
Our purpose is not to fully review all the available valuation methods, but rather to
provide a concise overview of such methods while illustrating the objective and
context of their application. In the previous section, we focused on consumer and
producer surpluses as the measures of interest and explained how these measures
relate to WTP and WTA. Bearing in mind that these are the measures of interest, we
should also note that the focus of economic valuation is to estimate such measures
for a well-defi ned change. This implies estimating the changes in the consumer and
producer surpluses and the change in their sum (Freeman 2003 ) by considering
changes in the welfare of both consumers and producers.
The methods used to value ecosystem services can be grouped into three main
categories: direct market valuation approaches, revealed preferences approaches,
and stated preferences approaches. Direct market valuation approaches rely on the
use of data that can be readily obtained from existing markets (such as prices,
demanded quantities, and production costs), and include three main approaches:
approaches based on market prices, costs, and production functions. Market price
approaches rely on the use of market prices as a proxy for value. Although this
appears to be the most straightforward approach, there are several aspects that
should be emphasized about its application. Under the general case of perfect com-
petitive markets, prices are defi ned by the interaction of supply and demand; as a
result, prices are acceptable or starting point approximations of the marginal value.
5 Changes in the ecosystem services provided by forests and their economic…

124
If this holds true, and the change being analyzed is suffi ciently small that prices
remain constant, application of the method is straightforward: we just multiply the
change in the number of units (for instance, the increase or decrease of the available
m
3 of water) by the associated marginal price. When the changes are large enough
to change prices, then the changes in consumer and producer surplus must be
estimated. Even if prices can be taken as a proxy for the marginal value, price dis-
tortions created by subsidies and taxes should be taken into account; the cost of
making the good available should be subtracted from the price in some cases, since
labor and transportation costs involved in making the benefi t available represent
opportunity costs that could be transferred to generate alternative goods and values;
in addition, prices generated by supply and demand refl ect scarcity, not value, as is
often illustrated using the relative prices of water and diamonds (a paradox origi-
nally posed by Adam Smith), since water is vital to support life (unlike diamonds)
but because it is generally abundant, it is cheaper than diamonds.
Cost-based approaches include the avoided cost, replacement cost, and mitigation
or restoration cost methods, and are used to estimate the costs that would be incurred
to artifi cially provide the benefi t instead of using ecosystem services. In the context
of forest ecosystem services valuation, the avoided cost method could be used (for
example) to estimate the value of fl ooding protection provided by a forest based on
the costs of building protection infrastructures to generate the same benefi t; for
other applications of the avoided cost method, see Nowak et al. ( 2006 ) and van
Kooten ( 2007 ). The replacement cost method could be applied (for example) to
estimate the value of soil protection based on the costs to restore the storage capacity
of downstream dams after siltation of the reservoir. For other applications of this
method, see Chopra and Kumar ( 2004 ) and Rodríguez et al. ( 2007 ). The restoration
costs may, for instance, be useful in determining the value of water purifi cation or
infi ltration based on the investments made to reverse degradation of the service. For
other examples of the restoration cost approach, see Birch et al. ( 2010 ).
The last of the approaches based on market valuation is the production function
method, which Barbier ( 2007 ) referred to as “valuing the environment as input”.
Behind the method’s application is the idea that several ecosystem services (e.g.,
regulation services, biodiversity) enhance the production of market goods. Hence,
if changes in these services affect the marketed product, then the effects of these
changes will be visible through the price system. For instance, if the purifi cation
capacity for water decreases and this generates additional costs for the producers
of bottled water, then the price of the water would increase. An example of this
method in the context of forest ecosystem service valuation is provided by
Nahuelhual et al. ( 2006 ).
In the revealed preferences approach, the main methods are the travel cost and
hedonic pricing methods. These methods use consumption behavior in markets that
are related to the non-market goods and that therefore serve as proxies for those
goods. The travel cost method is the most commonly used method, and has been
widely applied to infer the value of forests for recreation (e.g., Badola et al. 2010 ,
Bowker et al.
2007 ). This method uses visitation rates and the distance traveled to
infer the demand for such a benefi t. The observed variation in visitation rates and
C. Marta-Pedroso et al.

125
travel costs (used as a proxy of price) describes the changes in demand for the site,
and the demand function allows researchers to determine the consumer surplus. The
hedonic pricing method uses the differences in the price of a benefi t that refl ect its
inherent properties to infer the value of non-market attributes of ecosystems. A
recent application of the method was provided by Sander and Haight ( 2012 ).
In the stated preferences approach, contingent valuation is the most well-known
method. This method involves directly asking a representative sample of a popula-
tion to defi ne their WTP and their WTA for a well-defi ned change in the provision
of a certain ecosystem service (for instance, a change in water quality). Researchers
can then use compensating variation or equivalent variation to estimate the eco-
nomic value. Both are exact welfare measures, and may not be identical to the
consumer surplus for market goods. Instead, these measures estimate the change in
income that is needed to maintain a certain level of utility (welfare, satisfaction).
Note that along an ordinary demand curve, utility is not constant if income is kept
constant. For further explanation of these measures, see Freeman ( 2003 ) and Zerbe
and Bellas ( 2006 ). Choice modeling is a questionnaire-based method that gained
relevance with practitioners of economic valuation of ecosystem services. The
method consists of presenting individuals with two or more alternatives defi ned by
a set of attributes regarding the ecosystem services under valuation, and it is
designed to elicit the WTP for having that alternative. The levels of the set of attri-
butes vary among the alternative sets that individuals must choose among or rank.
Both methods have been applied to estimate the value of several forest ecosystem
services. For examples of contingent valuation applications, see Sattout et al. ( 2007 )
and Barrio and Loureiro ( 2010 ); for examples of choice modeling, see Rolfe et al.
( 2000 ) and Brey et al. ( 2007 ). Individual-based questionnaires aggregated to repre-
sent a socially relevant unit (e.g., a community) might be appropriate when the
services being valued are purely enjoyed on an individual level (e.g., valuing forests
for timber), but have limited applicability in the cases of more communal services.
For example, the value of forests to a community whose social system is intimately
dependent on them is more than the sum of the independent personal values (Farber
et al. 2002 ). Hence, another stated preferences method, group valuation, is gaining
relevance. Although a stated preferences method, its focus is not on valuing indi-
vidual preferences but rather on collecting social preferences. Wilson and Howarth
( 2002 ) and Chan et al. ( 2012 ) provide a detailed discussion of this method.
5.4.3 Challenges in estimating economic value
Estimating the economic value of ecosystem services faces several challenges, and
regardless of the objective of the economic valuation, whether to inform macro-
economic policies or to evaluate programs (Bateman et al. 2010 ), estimation of
fl ows of ecosystem services is often necessary. A fl ow estimation is usually an
estimate of money per unit area obtained for a certain period, usually on an annual
basis. Although fl ow estimations provide valuable information, they are not, per se,
5 Changes in the ecosystem services provided by forests and their economic…

126
relevant to inform land-use decisions because few interventions would result in an
entire loss of the fl ows of ecosystem services. Instead, management often results in
incremental small changes. What is needed is an understanding of how land-use
changes would affect societal well-being, so the focus of the economic valuation is
on valuing the incremental or marginal changes in the fl ows of services. This is
often done by means of scenario analysis, in which researchers compare the conse-
quences of two or more scenarios. Valuing such changes implies a deep under-
standing of the ecological dynamics of the system and how the system responds to
perturbation.
Though the economic valuation approach is remarkably valuable because of its
ability to provide more objective comparisons of alternatives, it has not yet over-
come signifi cant challenges to tackling such complexity (Robertson 2011 ). This
problem has been pointed out by several authors under the headings of uncertainty,
ecological thresholds, and irreversibility (Morse-Jones et al. 2011 ), and in the
contexts of weak or strong sustainability (Olschewski and Klein, 2011 ). Because
the valuation focuses on estimates of marginal changes, caution is needed with
the valuation itself because the marginal value may not be constant. This is clearly
illustrated by the example provided by Bateman et al. ( 2010 ), who examined the
recreational value of an urban green space (a park). They found that increasing the
area of this space altered the recreational marginal value, with the fi rst increases in
area being highly valued, but subsequent increases becoming less valued.
There are other problems related to the assumption of a constant marginal value
that suggest a need for caution. Ecosystems and their services are not spatially
homogeneous and thus may not provide the same fl ow throughout the system’s
spatial extent (Fisher et al. 2009 ). Moreover, even when ecosystems provide the
same fl ow of services from different areas, the marginal values of these fl ows may
not be the same. We can illustrate this again using the value of a green space for
recreation. An urban forest area of a given size may have a higher recreation value
when it is near an urban area than when it lies in a region that is not accessible to
urban residents. The issue of spatial variability of ecosystems and ecosystem ser-
vices suggests the need to perform economic valuation on a spatially explicit basis.
In addition, the effect of scale is a challenging topic that has not been fully tackled.
This affects the discussion of ecosystem services valuation because the scale at
which benefi ts might be provided ranges from local to global. For example, a forest
might provide recreational opportunities (local), downstream fl ood prevention
(regional), and climate regulation (global) when considered from the supply side.
This variability also holds for the demand side. For instance, endemic species may
have benefi ciaries very far from the location of their occurrence (as in the case of
residents of developed nations placing value on endangered species in developing
nations). The issue of scale has been extensively debated (EEA 2010 , Hein et al.
2006 ), and the advantages of spatial analysis in tackling the issue are making scale-
explicit analyses increasingly relevant. Failing to properly address the issue of scale
may complicate or bias the design of ecosystem services payments, which is an
emerging mechanism to ensure the provision of non-market ecosystem services.
C. Marta-Pedroso et al.

127
5.4.4 Economic estimates of forest ecosystem services around
the world
As we have stressed in the previous sections of this chapter, awareness of the
importance of forest ecosystems and the vulnerability of their valuable services is
increasing around the world. This awareness is ultimately at the base of recent eco-
nomic valuation efforts that targeted forest ecosystem services. In this section, we
provide an overview of recent studies in which forest ecosystems from around the
world were monetarily valued.
Previously, we briefl y introduced the concept of TEV as a commonly used tax-
onomy to determine the aggregate economic value of all the benefi ts people obtain
from ecosystems. Several researchers have attempted to obtain the TEV of forests
(e.g., Adger et al. 1994 , Merlo and Croitoru 2005 , Pearce 2001 , Thompson et al.
2011 ) and Ferraro et al. ( 2012 ) provide a review of this topic.
Targeting specifi c areas for the estimation of TEV is a common practice.
One recent attempt to assess TEV focused on the Hoge Veluwe Park (Hein 2011 ), a
protected area in the Netherlands. Hein estimated the economic value of ecosystem
services provided by the park’s more than 5000 ha of pine and deciduous forests to
be around 10.7 million € per year, of which 2.1 million € per year were due to air
pollution removal and 1.9 million € per year were due to groundwater infi ltration,
for example. By combining land-cover mapping with benefi t-transfer calculations,
Vorra and Barg ( 2008 ) estimated the aggregate economic value of ecosystem ser-
vices provided by the Canadian UNESCO World Heritage Site of Pimachiowin Aki
to be approximately C$130 million per year, mostly due to its provision of pure
water and fi sh. Ingraham and Foster ( 2008 ) used a similar approach, but their goal
was even more ambitious: to determine the aggregate economic value of the ecosys-
tem services provided by a large network of protected forests (the U.S. National
Wildlife Refuge System) in the 48 contiguous states, which turned out to be around
$26.9 billion per year. Although this was a fi rst approximation, their research high-
lighted the need for further and more rigorous examination of the value of ecosystem
services to assist management and policy decision-making, and their results empha-
sized that the TEV of protected areas exceeds that of its pure recreational value.
Bolder initiatives to assess the aggregate monetary value of global forest ecosys-
tem services have also taken place in recent years. By using ad hoc value-transfer
protocols, Chiabai et al. (
2009 ) estimated the economic value of a comprehensive
set of ecosystem services (timber and non-timber forest products, carbon storage,
and recreation and tourism) for all forest biomes around the world. Benefi t (or
value)-transfer uses economic information captured at one place and time to make
inferences about the economic value of environmental goods and services at another
place and time, as discussed by Wilson and Hoehn (
2006 ). An interesting aspect of
this research is that it also presents potential estimates of total economic losses by
the year 2050 due to policy inaction. They identifi ed major economic losses of 78
billion €, mostly due to the loss of the forests’ provisioning and carbon sequestration
services. Their results underlined current production scenarios; for example, they
5 Changes in the ecosystem services provided by forests and their economic…

128
estimated the marginal value of provisioning services from tropical forests in Africa
at around US$1800 per ha per year as a result of the high production value of
fuelwood in Africa.
Even though valuing nature as a whole through the TEV taxonomy might be
pertinent for current conservation agendas, valuation studies are most commonly
performed on a case study basis, with a particular ecosystem service or a particular
subset of services being targeted. Focusing on a particular ecosystem service might
help to solve specifi c challenges and support policy development at a relatively
small scale, as it is a far simpler task than estimating TEV and may provide more
practical outputs. For example, to provide insights into the protection of woodlands
as a climate-change mitigation measure, Brainard et al. ( 2009 ) performed a cost–
benefi t analysis to assess the value of the carbon sequestration services provided by
the woodlands of Great Britain. Their results, which depended strongly on the
discount rate and the social value of sequestered carbon being considered, ranged
from US$82 to US$853 million annually. The variation of estimations due to the
application of different discount rates is not uncommon, and imposes an additional
challenge for the valuation of nature (Freeman and Groom 2013 ). In fact, the choice
of the discount rate to be applied in environmental appraisal is not exempt from
problems and criticisms, and there is no single discount rate to be applied, so sensi-
tivity analysis may be generally performed for a range of discount rates (Boardman
et al. 2001 ).
Appraising the value of the carbon sequestration services of forests as part of the
tactics for mitigating human-origin CO
2 emissions has been receiving increasing
attention in response to growing awareness of climate change. In Brazil, Guitart and
Rodriguez ( 2010 ) have assessed the value of potential carbon sequestration services
provided by two commercial eucalyptus plantations. Based on their results, they
suggested a minimum annuity of US$18.8 per ton of stored carbon to be paid to the
owner of the forests in order to stimulate and justify the adoption of silvicultural
regimes that would increase carbon sequestration.
Though globally relevant nowadays, the ability to sequester carbon is not the
only valuable service that forests can provide. Olschewski et al. ( 2012 ) determined
the WTP for the avalanche protection services that the forests of the Swiss Alps
provide to the population of the Swiss municipality of Andermatt, in Canton Uri.
Their results ranged from US$20 per household (a one-time payment due to avoid-
ance costs) to more than US$300 per household (a one-time payment due to risk
reduction). In Africa, the risk of snow avalanches is obviously not a primary con-
cern. Instead, Schaafsma et al. ( 2012 ) estimated the total fl ow of benefi ts from char-
coal production in the tropical forests of southern Kenya to be worth around US$14
million per year, as charcoal provides an important source of income to local house-
holds and supplies around 11 % of the charcoal used in the major cities of Kenya
and Tanzania. For timber forest products, Ojea et al. ( 2012 ) estimated the potential
value of wood provision in sustainably harvested Mediterranean forests in Spain at
around 500 € per ha per year based on the sustainable provision of timber over 30
years at a discount rate of 2 %. Their results indicated that, in some cases, non-
sustainable forests provided higher returns over shorter time spans, but that sustainable
C. Marta-Pedroso et al.

129
harvesting provided the highest overall returns over long terms. Regardless of the
ecosystem service targeted, studies at a regional level bring a very policy- oriented
perspective into the valuation exercise.
5.5 Initiatives and policy responses
Today, most large-scale initiatives and agreements regarding forest management
revolve around the concept of sustainable forest management to prevent forest
degradation and to promote the development of multifunctional forest systems.
The REDD+ mechanism ( Reducing emissions from deforestation and forest deg-
radation in developing countries and the role of conservation, sustainable manage-
ment of forests and enhancement of forest carbon stocks in developing countries ) is
currently the most promising tool designed to support the conservation of forests,
with a particular emphasis on carbon-regulating services (FAO 2012 , http://www.
un-redd.org/ ). The mechanism was established under the United Nations Framework
Convention on Climate Change ( http://unfccc.int/2860.php ) and included in the
global climate-change agenda in 2007 that was defi ned at the climate summit in Bali
(Angelsen and Rudel 2013 ). The mechanism has been implemented through several
initiatives, such as the UN-REDD program ( http://www.un-redd.org/ ) or the Forest
Carbon Partnership Facility hosted by the World Bank ( http://www.forestcarbonpart-
nership.org/ ). REDD+ is a fi nancial mechanism designed to reduce carbon emissions
caused by forest losses and degradation in developing countries while at the same
time creating conditions for sustainable forest management and promoting sustain-
able development programs in the participating countries (Angelsen et al. 2012 ,
IUCN 2009 ). In brief, the underpinning idea is to implement payment for ecosystem
services (PES) schemes in which the international community pays forest users in
developing countries to adopt policies and programs aimed at conserving forests,
improving forest stocks, and reducing forest degradation (Angelsen et al. 2012 ).
Despite the overall support for this mechanism, and its ongoing application in
several countries, it has also been criticized based on several important issues. These
include the need to guarantee sustainable sources of funding, the diffi culty in moni-
toring the outcomes of implemented projects, and more importantly, the lack of a
clear understanding and a legal framework for land tenure and carbon rights in
many countries, which can be a barrier to the implementation of PES schemes, par-
ticularly if this promotes inequity and disregard for the rights of forest communities
and indigenous peoples (Angelsen et al. 2012 , FAO 2012 ). In addition, there is some
apprehension concerning the subordinate role of biodiversity in relation to carbon
storage and sequestration, which constitute the main focus of REDD+. Unclear tar-
gets for biodiversity and other ecosystem services may allow the occurrence of
trade-offs instead of achieving the envisioned synergies, which are expected to arise
as positive externalities from activities directed towards carbon storage and seques-
tration (Visseren-Hamakers et al.
2012 ). For instance, there is a risk of leakage (i.e.,
intensifi cation of activities in areas not covered by REDD+ projects) and of inadequate
5 Changes in the ecosystem services provided by forests and their economic…

130
implementation of REDD+ activities, such as the establishment of plantations of
exotic species (Visseren-Hamakers et al. 2012 ). Also, because there is some discon-
nection between the global distribution of carbon stocks and the associated biodi-
versity, the outcomes of REDD+ projects may be less effective for protecting
biodiversity and other ecosystem services (Visseren-Hamakers et al. 2012 ).
Other PES mechanisms are emerging at a regional scale and are aimed at con-
serving forest ecosystem services. For instance, in the European Union, under the
current European Agricultural Fund for Rural Development ( http://ec.europa.eu/
agriculture/cap-funding/budget/index_en.htm ), a payment scheme has been imple-
mented to support the development of multifunctional forests and the adoption of
good management practices.
The rationale behind PES mechanisms lies in their attempt to internalize market
externalities. As described above, many ecosystem services are not tradable in a
market; thus, producers are unable to introduce their value into the price of the prod-
ucts they supply. Due to this market externality, producers will always be underre-
warded for the services they provide in the absence of production incentives or
subsidies. Through PES, central governments or private users and consumers make
payments for the ecosystem services provided by landowners, producers, and other
entities such as environmental agencies or nongovernmental conservation organiza-
tions. Examples involving the industry sector refl ect how industries can often play a
leading role as benefi ciaries or buyers of ecosystem services, as in the case of the
Nestlé Waters Programme ( www.nestle-waters.com/ ).
Although the EU Forest Strategy is implemented at a regional scale, it nonethe-
less provides a good example of intergovernmental action to promote and support
sustainable forest management through the coordination of forest policies by the
member states and through community policies (CEC 2005 ). The strategy also
acknowledges the multifunctional role of forests and their multiple services, and
their relevance for the well-being of society. The Ministerial Conference on the
Protection of Forests in Europe ( http://www.foresteurope.org/ ) constitutes the polit-
ical process at a pan-European level for the establishment of sustainable forest man-
agement. The Forest Europe strategy for 2020 developed by Ministerial Conference
on the Protection of Forests in Europe has been signed by 46 countries, and lists
among its main targets the valuation of multiple forest services and raising society’s
awareness of the importance of forests to human well-being. This strategy will fos-
ter cooperation among countries to develop and update their forest policies so as to
secure and promote sustainable forest management.
Targeting different scales, the certifi cation or information labeling of agrofor-
estry products might also be a way to communicate environmental and other attri-
butes that are not directly visible in the products, with the goal of promoting
sustainable management of forestry resources. The rationale behind certifi cation
mechanisms is simple: if there is a market demand for differentiated agroforestry
products, meaning that consumers are willing to pay for the price difference listed
by producers due to their compliance with environmental standards and the conse-
quent delivery of external benefi ts, then a market-based solution for sustainable
rural development becomes possible. Countries like the Netherlands, Germany, and
C. Marta-Pedroso et al.

131
the UK are clear examples of major markets for certifi ed forestry products, and
specifi cally for timber. The most widely acknowledged example of certifi cation
mechanisms are the regional and national standards developed by the Forest
Stewardship Council ( https://ic.fsc.org/ ). Other less widely known initiatives include
the Programme for the Endorsement of Forest Certifi cation Schemes ( http://www.
pefc.org/ ), which is now a recognized label in Europe and in a few other countries
such as Brazil and the United States. FAO ( 2007 ) estimated that around 270 × 10
6 ha
of forests around the world, amounting to roughly 7 % of the world’s forests, are
certifi ed for sustainability through an independent labeling organization. However,
in less developed countries, the costs of complying with such environmental stan-
dards in addition to the costs of the certifi cation process itself represent a great chal-
lenge for producers that want to enter a certifi cation market.
At the local level, initiatives to manage forest ecosystem services can take many
forms that depend on several factors, including land ownership, size and gover-
nance, forest location and condition, and (of course) the targeted services. Local
initiatives can be independent or can derive from initiatives at larger scales, such as
the global initiatives discussed above. Other examples of local actions include
sustainable forest management and land-use planning, measures to improve forest
resilience to disturbance (Fernandes 2013 ), and actions to restore degraded forests
and deforested land, which can help to restore ecosystem services and biodiversity
(Chazdon 2008 , Rey Benayas et al. 2009 ).
Despite the scale of implementation, the success of these initiatives to manage
forests and their multiple services will depend on the existence of governance struc-
tures and legal frameworks that safeguard access to the forest resources and respect
the rights of all forest users, thereby avoiding inequity in the access to benefi ts (FAO
2012 ). It will also be very important to invest in building capacity to create condi-
tions suitable for the implementation of sustainable forest management programs
and to provide local peoples with the knowledge and tools they require to participate
in the design and implementation of those programs (FAO 2012 ) .
Literature cited
Adger WN, Brown K, Cervigni R, Moran D (1994) Towards estimating total economic value of
forests in Mexico. Centre for Social and Economic Research on the Global Environment,
University of East Anglia, Norwich, UK, and University College London, London.
http://mil-
lenniumindicators.un.org/unsd/envaccounting/ceea/archive/Forest/TEV_Mexican_Forest.PDF
Angelsen A, Rudel TK (2013) Designing and implementing effective REDD+ policies: a forest
transition approach. Rev Environ Econ Policy 7:91–113
Angelsen A, Brockhaus M, Sunderlin WD, Verchot LV (2012) Analysing REDD+: challenges and
choices. Center for International Forestry Research (CIFOR), Bogor, Indonesia.
http://www.
cifor.org/online-library/browse/view-publication/publication/3805.html
Badola R, Hussain SA, Mishra BK, Konthoujam B, Thapliyal S, Dhakate PM (2010) An assess-
ment of ecosystem services of Corbett Tiger Reserve, India. Environmentalist 30:320–329
Barbier EB (2007) Valuing ecosystem services as productive inputs. Econ Policy 22:177–229
Barrio M, Loureiro ML (2010) A meta-analysis of contingent valuation forest studies. Ecol Econ
69:1023–1030
5 Changes in the ecosystem services provided by forests and their economic…

132
Bateman IJ, Mace GM, Fezzi C, Atkinson G, Turner K (2010) Economic analysis for ecosystem
service assessments. Environ Resour Econ 48:177–218
Birch JC, Newton AC, Aquino CA, Cantarello E, Echeverria C, Kitzberger T, Schiappacasse I,
Garavito NT (2010) Cost-effectiveness of dryland forest restoration evaluated by spatial analy-
sis of ecosystem services. Proc Natl Acad Sci U S A 107:21925–21930
Boardman AE, Greenberg DH, Vining AR, Weimer DL (2001) Cost-benefi t analysis: concepts and
practice, 2nd edn. Prentice Hall, Upper Saddle River, NJ
Bowker JM, Bergstrom JC, Gill J (2007) Estimating the economic value and impacts of recre-
ational trails: a case study of the Virginia Creeper Rail Trail. Tourism Econ 13:241–260
Boyd J, Banzhaf S (2007) What are ecosystem services? The need for standardized environmental
accounting units. Ecol Econ 63:616–626
Brainard J, Bateman IJ, Lovett A (2009) The social value of carbon sequestered in Great Britain’s
woodlands. Ecol Econ 68(4):1257–1267
Brey R, Riera P, Mogas J (2007) Estimation of forest values using choice modeling: an application
to Spanish forests. Ecol Econ 64:305–312
Butchart SHM, Walpole M, Collen B, van Strien A, Scharlemann JP, Almond RE, Baillie JE,
Bomhard B, Brown C, Bruno J, Carpenter KE, Carr GM, Chanson J, Chenery AM, Csirke J,
Davidson NC, Dentener F, Foster M, Galli A, Galloway JN, Genovesi P, Gregory RD, Hockings
M, Kapos V, Lamarque JF, Leverington F, Loh J, McGeoch MA, McRae L, Minasyan A,
Hernández Morcillo M, Oldfi eld TE, Pauly D, Quader S, Revenga C, Sauer JR, Skolnik B,
Spear D, Stanwell-Smith D, Stuart SN, Symes A, Tierney M, Tyrrell TD, Vié JC, Watson R
(2010) Global biodiversity: indicators of recent declines. Science 328:1164–1168
Cao SX, Chen L, Shankman D, Wang CM, Wang XB, Zhang H (2011) Excessive reliance on
afforestation in China’s arid and semi-arid regions: lessons in ecological restoration. Earth Sci
Rev 104:240–245
Carvalho-Ribeiro SM, Lovett A, O’Riordan T (2010) Multifunctional forest management in north-
ern Portugal: moving from scenarios to governance for sustainable development. Land Use
Policy 27:1111–1122
CEC (2005) Reporting on the implementation of the EU Forestry Strategy. COM (2005)85.
Commission of the European Communities, Brussels
Chan KMA, Satterfi eld T, Goldstein J (2012) Rethinking ecosystem services to better address and
navigate cultural values. Ecol Econ 74:8–18
Chazdon RL (2008) Beyond deforestation: restoring forests and ecosystem services on degraded
lands. Science 320:1458–1460
Chiabai A, Travisi C, Ding H, Markandya A, Dias Nunes P (2009) Economic valuation of forest
ecosystem services: methodology and monetary estimates. FEEM working paper (12).
http://
papers.ssrn.com/sol3/papers.cfm?abstract_id=1396661
Chopra K, Kumar P (2004) Forest biodiversity and timber extraction: an analysis of the interaction
of market and non-market mechanisms. Ecol Econ 49:135–148
Costanza R (2008) Ecosystem services: multiple classifi cation systems are needed. Biol Conserv
141:350–352
Daily GC, Söderqvist T, Aniyar S, Arrow K, Dasgupta P, Ehrlich P, Folke C, Jansson AM, Jansson
BO, Kautsky N, Levin S, Lubchenco J, Mäler KG, Simpson D, Starrett D, Tilman D, Walker B
(2000) The value of nature and the nature of value. Science 289:395–396
Davidson EA, de Araujo AC, Artaxo P, Balch JK, Brown IF, Bustamente MMC, Coe MT, DeFriess
RS, Keller M, Longo M, Munger JW, Schroeder W, Soares-Filho B, Souza CM Jr, Wofsy SC
(2012) The Amazon basin in transition. Nature 481:321–328
Duncker PS, Raulund-Rasmussen K, Gundersen P, Katzensteiner K, De Jong J, Ravn HP, Smith
M, Eckmüllner O, Spiecker H (2012) How forest management affects ecosystem services,
including timber production and economic return: synergies and trade-offs. Ecol Soc 17(4):50.
doi:
10.5751/ES-05066-170450
EEA (2010) Scaling up ecosystem benefi ts: a contribution to The Economics of Ecosystems and
Biodiversity (TEEB) study. European Environmental Agency.
http://www.eea.europa.eu/publi-
cations/scaling-up-ecosystem-benefi ts-a
C. Marta-Pedroso et al.

133
FAO (2007) Forest products annual market review, 2005/2006. United Nations, Food and
Agriculture Organization, Rome, and Economic Commission for Europe, Geneva.
http://www.
unece.org/fi leadmin/DAM/timber/docs/fpama/2007/FPAMR2007.pdf
FAO (2011) Global forests resources assessment. United Nations, Food and Agriculture
Organization, Rome.
http://www.fao.org/forestry/fra/en/
FAO (2012) State of the world’s forests. United Nations, Food and Agriculture Organization,
Rome
Farber SC, Costanza R, Wilson MA (2002) Economic and ecological concepts for valuing ecosys-
tem services. Ecol Econ 41(3):375–392
Farrell EP, Führer E, Ryan D, Andersson F, Hüttl R, Piussi P (2000) European forest ecosystems:
building the future on the legacy of the past. For Ecol Manage 132:5–20
Fernandes PM (2013) Fire-smart management of forest landscapes in the Mediterranean basin
under global change. Landsc Urban Plan 110:175–182
Fernandez-Manjarrés J, Leadley PW (2010) Arctic tundra (Appendix 1). In: Leadley P, Pereira
HM, Alkemade R, Fernandez-Manjarrés JF, Proença V, Scharlemann JPW, Walpole MJ (eds)
Biodiversity scenarios: projections of 21st century change in biodiversity and associated eco-
system services. Technical series no. 50. Convention on Biological Diversity, Montreal,
Canada, pp 53–66
Ferraro PJ, Lawlor K, Mullan KL, Pattanayak SK (2012) Forest fi gures: ecosystem services valu-
ation and policy evaluation in developing countries. Rev Environ Econ Policy 6:20–44
Fisher B, Turner RK, Morling P (2009) Defi ning and classifying ecosystem services for decision
making. Ecol Econ 68:643–653
Freeman AM (2003) The measurement of environmental and resource values: theory and methods.
Resources for the Future, Washington, DC
Freeman MC, Groom B (2013) Biodiversity valuation and the discount rate problem. Account
Auditing Account J 26(5):715–745
Guitart AB, Rodriguez LCE (2010) Private valuation of carbon sequestration in forest plantations.
Ecol Econ 69:451–458
Gustafsson L, Baker SC, Bauhus J, Beese WJ, Brodie A, Kouki J, Lindenmayer DB, Lõhmus A,
Martínez Pastur G, Messier C, Neyland M, Palik B, Sverdrup-Thygeson A, Volney WJA,
Wayne A, Franklin JF (2012) Retention forestry to maintain multifunctional forests: a world
perspective. BioScience 62:633–645
Haines-Young R, Potschin M (2010) Proposal for a common international classifi cation of ecosys-
tem goods and services (CICES) for integrated environmental and economic accounting (V1).
Report prepared for the European Environment Agency.
http://www.nottingham.ac.uk/CEM/
pdf/UNCEEA-5-7-Bk1.pdf
Hanley N, Barbier EB (2009) Pricing nature: cost-benefi t analysis and environmental policy.
Edward Elgar, Cheltenham, UK
Hein L (2011) Economic benefi ts generated by protected areas: the case of the Hoge Veluwe forest,
the Netherlands. Ecol Soc 16(2):13.
http://www.ecologyandsociety.org/vol16/iss2/art13/
Hein L, van Koppen K, de Groot RS, van Ierland EC (2006) Spatial scales, stakeholders and the
valuation of ecosystem services. Ecol Econ 57:209–228
Hobbs RJ, Cramer VA (2007) Old fi elds: dynamics and restoration of abandoned farmland, 1st
edn. Island Press, Washington, DC
Hurtt GC, Chini LP, Frolking S, Betts R, Feedema J, Fischer G, Goldewijk KK, Hibbard K, Janetos
A, Jones C (2009) Harmonization of global land use scenarios for the period 1500–2100 for
IPCC AR5. Integrated Land Ecosystem-Atmosphere Processes Study (iLEAPS) Newsletter
7:6–8
IMAGE-team (2001) The IMAGE 2.2 implementation of the SRES scenarios. CD-ROM publica-
tion 481508018. National Institute for Public Health and the Environment, Bilthoven, The
Netherlands
Ingraham MW, Foster SG (2008) The value of ecosystem services provided by the US National
Wildlife Refuge System in the contiguous US. Ecol Econ 67:608–618
5 Changes in the ecosystem services provided by forests and their economic…

134
IUCN (2009) Scope and options for the role of forests in climate change mitigation strategies.
International Union for Conservation of Nature, Washington, DC
Kanowski P (2003) Challenges to enhancing the contributions of planted forests to sustainable
forest management. In: UNFF Intersessional Experts Meeting on the role of planted forests in
sustainable forest management. United Nations Forum on Forests, Washington, DC, pp 24–28
Kaplan JO, Krumhardt KM, Zimmermann N (2009) The prehistoric and preindustrial deforesta-
tion of Europe. Quat Sci Rev 28:3016–3034
Kareiva P, Tallis H, Ricketts TH, Daily GC, Polasky S (2011) Natural capital: theory and practice
of mapping ecosystem services, 1st edn. Oxford University Press, Oxford
Keeley JE, Lubin D, Fotheringham CJ (2003) Fire and grazing impacts on plant diversity and alien
plant invasions in the southern Sierra Nevada. Ecol Appl 13(5):1355–1374
Keeley JE, Baer-Keeley M, Fotheringham CJ (2005) Alien plant dynamics following fi re in
Mediterranean-climate California shrublands. Ecol Appl 15:2109–2125
Keenleyside C, Tucker G, McConville A (2010) Farmland abandonment in the EU: an assessment
of trends and prospects. Institute for European Environmental Policy, London
Lambin EF, Geist HJ, Lepers E (2003) Dynamics of land-use and land-cover change in tropical
regions. Annu Rev Environ Resour 28:205–241
Leadley P, Proença V, Fernandez-Manjarréz, Pereira HM, Alkemade R, Biggs R, Bruley E,
Cheung W, Copper D, Figueiredo J, Gilman E, Guénette S, Hurtt G, Mbow C, Oberdorff T,
Revenga C, Scharlemann JPW, Scholes R, Stafford-Smith M, Sumaila R, Walpole M (in press)
Interacting regional-scale regime shifts for biodiversity and ecosystem services. BioScience
Leadley P, Pereira HM, Alkemade R, Fernandez-Manjarres JF, Proença V, Scharlemann JPW (eds)
(2010) Biodiversity scenarios: projections of 21st century change in biodiversity and associ-
ated ecosystem services. Convention on Biological Diversity, Montreal, QC
Lloyd AH, Bunn AG, Berner L (2011) A latitudinal gradient in tree growth response to climate
warming in the Siberian taiga. Glob Change Biol 17:1935–1945
Mace GM, Bateman I (coord.) (2011) Chapter 2: conceptual framework and methodology.
Technical report. UK National Ecosystem Assessment, Cambridge, pp 11–25.
http://uknea.
unep-wcmc.org/
Maes J, Teller A, Erhard M, Liquete C, Braat L, Berry P, Egoh B, Puydarrieux P, Fiorina C, Santos
F, Paracchini ML, Keune H, Wittmer H, Hauck J, Fiala I, Verburg PH, Condé S, Schägner JP,
San Miguel J, Estreguil C, Ostermann O, Barredo JI, Pereira HM, Stott A, Laporte V, Meiner
A, Olah B, Royo Gelabert E, Spyropoulou R, Petersen JE, Maguire C, Zal N, Achilleos E,
Rubin A, Ledoux L, Brown C, Raes C, Jacobs S, Vandewalle M, Connor D, Bidoglio G (2013)
Mapping and assessment of ecosystems and their services. An analytical framework for eco-
system assessments under action 5 of the EU biodiversity strategy to 2020. Publications offi ce
of the European Union, Luxembourg.
Mankiw NG (2008) Principles of economics. South-Western College Publishing, Mason, OH
Martín-López B, Goméz-Baggethun E, González JA, Lomas PL, Montes C (2009) The assessment
of ecosystem services provided by biodiversity: re-thinking concepts and research needs. In:
Aronoff JB (ed) Handbook of nature conservation: global, environmental and economic issues.
Nova Science, Hauppauge, NY, pp 261–282
Martín-López B, Iniesta-Arandia I, García-Llorente M, Palomo I, Casado-Arzuaga I, García Del
Amo D, Gómez-Baggethun E, Oteros-Rozas E, Palacios-Agundez I, Willaarts B, González JA,
Santos-Martín F, Onaindia M, López-Santiago C, Montes C (2012) Uncovering ecosystem
service bundles through social preferences. PLoS One 7:e38970
MEA (2005) Ecosystems and human well-being: biodiversity synthesis. World Resources Institute,
Washington, DC
Merlo M, Croitoru L (2005) Valuing Mediterranean forests: towards total economic value. CABI,
Wallingford, Oxfordshire
Morse-Jones S, Luisetti T, Turner RK, Fisher B (2011) Ecosystem valuation: some principles and
a partial application. Environmetrics 22:675–685
Nahuelhual L, Donoso P, Lara A, Núñez D, Oyarzún C, Neira E (2006) Valuing ecosystem ser-
vices of Chilean temperate rainforests. Environ Dev Sustain 9:481–499
C. Marta-Pedroso et al.

135
Nobre C, Leadley P, Fernandez-Manjarres J (2010) Amazonian forest (Appendix 3). In: Leadley P,
Pereira HM, Alkemade R, Fernandez-Manjarres JF, Proença V, Scharlemann JPW (eds)
Biodiversity scenarios: projections of 21st century change in biodiversity and associated eco-
system services. Technical Series no. 50. Convention on Biological Diversity, Montreal,
Canada, pp 68–75
Nowak DJ, Crane DE, Stevens JC (2006) Air pollution removal by urban trees and shrubs in the
United States. Urban For Urban Greening 4:115–123
Ojea E, Ruiz-Benito P, Markandya A, Zavala MA (2012) Wood provisioning in Mediterranean
forests: a bottom-up spatial valuation approach. For Policy Econ 20:78–88
Oksanen M (1997) The moral value of biodiversity. Ambio 26:541–545
Olschewski R, Klein A (2011) Ecosystem services between sustainability and effi ciency. Sustain
Sci Pract Policy 7(1):69–73
Olschewski R, Bebi P, Teich M, Wissen Hayek U, Grêt-Regamey A (2012) Avalanche protection
by forests—a choice experiment in the Swiss Alps. For Policy Econ 17:19–24
Ostrom E (2009) A general framework for analyzing sustainability of social-ecological systems.
Science 325:419–422
Pagiola S, Bishop J, Landell-Mills N (2002) Selling forest environmental services: market-based
mechanisms for conservation and development. Earthscan Publications, London
Pearce DW (1993) Economic values and the natural world. Earthscan Publications, London
Pearce DW (2001) The economic value of forest ecosystems. Ecosyst Health 7:284–296
Pearce DW, Marandya A, Barbier EB (1989) Blueprint 1: for a green economy. Earthscan
Publications, London
Pereira HM, Leadley PW, Proença V, Alkemade R, Scharlemann JPW, Fernandez-Manjarrés JF,
Araújo MB, Balvanera P, Biggs R, Cheung WWL, Chini L, Cooper HD, Gilman EL, Guénette
S, Hurtt GC, Huntington HP, Mace GM, Oberdorff T, Revenga C, Rodrigues P, Scholes RJ,
Sumaila UR, Walpole M (2010) Scenarios for global biodiversity in the 21st century. Science
330:1496–1501
Pereira HM, Navarro LM, Martins IS (2012) Global biodiversity change: the bad, the good, and the
unknown. Annu Rev Environ Resour 37:25–50
Phillips OL, Aragão LEOC, Lewis SL, Fisher JB, Lloyd J, López-González G, Malhi Y,
Monteagudo A, Peacock J, Quesada CA, van der Heijden G, Almeida S, Amaral I, Arroyo L,
Aymard G, Baker TR, Bánki O, Blanc L, Bonal D, Brando P, Chave J, Alves de Oliveira AC,
Dávila Cardozo N, Czimczik CI, Feldpausch TR, Aparecida Freitas M, Gloor E, Higuchi N,
Jiménez E, Lloyd G, Meir P, Mendoza C, Morel A, Neill DA, Nepstad D, Patiño S, Peñuela
MC, Prieto A, Ramírez F, Schwarz M, Silva J, Silveira M, Sota Thomas A, ter Steege H, Stropp
J, Vásquez R, Zelazowski P, Alvarez Dávila E, Andelman S, Andrade A, Chao KJ, Erwin T, Di
Fiore A, Honorio EC, Keeling H, Killeen TJ, Laurance WF, Peña Cruz A, Pitman NCA, Núñez
Vargas P, Ramírez-Angulo H, Rudas A, Salamão R, Silva N, Terborgh J, Torres-Lezama A
(2009) Drought sensitivity of the Amazon rainforest. Science 323:1344–1347
Poyatos RJ, Latron J, Llorens P (2003) Land use and land cover change after agricultural abandon-
ment. Mt Res Dev 23:362–368
Proença V, Pereira HM (2010) Mediterranean forest (Appendix 2). In: Leadley P, Pereira HM,
Alkemade R, Fernandez-Manjarres JF, Proença V, Scharlemann JPW (eds) Biodiversity sce-
narios: projections of 21st century change in biodiversity and associated ecosystem services.
Technical Series no. 50. Convention on Biological Diversity, Montreal, Canada, pp 60–67
Proença VM, Pereira HM, Guilherme J, Vicente L (2010a) Plant and bird diversity in natural for-
ests and in native and exotic plantations in NW Portugal. Acta Oecol 36:219–226
Proença V, Pereira HM, Vicente L (2010b) Resistance to wildfi re and early regeneration in natural
broadleaved forest and pine plantation. Acta Oecol 36:626–633
Reino L, Beja P, Osborne PE, Morgado R, Fabião A, Rotenberry JT (2009) Distance to edges, edge
contrast and landscape fragmentation: interactions affecting farmland birds around forest plan-
tations. Biol Conserv 142:824–838
Rey Benayas J, Newton AC, Diaz A, Bullock JM (2009) Enhancement of biodiversity and ecosys-
tem services by ecological restoration: a meta-analysis. Science 325:1121–1124
5 Changes in the ecosystem services provided by forests and their economic…

136
Robertson M (2011) Ecosystems services. In: Nriagu JO (ed) Encyclopedia of environmental
health. Elsevier, Burlington, pp 225–233
Rodrigues ASL, Ewers RM, Parry L, Souza C, Veríssimo A, Balmford A (2009) Boom-and-bust
development patterns across the Amazon deforestation frontier. Science 324:1435–1437
Rodríguez JP, Balch JK, Rodríguez-Clark KM (2007) Assessing extinction risk in the absence of
species-level data: quantitative criteria for terrestrial ecosystems. Biodivers Conserv
16:183–209
Rolfe J, Bennett J, Louviere J (2000) Choice modelling and its potential application to tropical
rainforest preservation. Ecol Econ 35:289–302
Rudel TK, Coomes OT, Moran E, Achard F, Angelsen A, Xu J, Lambin E (2005) Forest transi-
tions: towards a global understanding of land use change. Glob Environ Change 15:23–31
Sala OE, van Vuuren D, Pereira HM, Lodge D, and Alder J (2005) Biodiversity across scenarios.
In: Ecosystems and human well-being: scenarios: fi ndings of the Scenarios Working Group.
Island Press, Washington, DC, pp 375–408
Sander HA, Haight RG (2012) Estimating the economic value of cultural ecosystem services in an
urbanizing area using hedonic pricing. J Environ Manage 113:194–205
Sattout EJ, Talhouk SN, Caligari PDS (2007) Economic value of cedar relics in Lebanon: an appli-
cation of contingent valuation method for conservation. Ecol Econ 61:315–322
Schaafsma M, Morse-Jones S, Posen P, Swetnam RD, Balmford A, Bateman IJ, Burgess ND,
Chamshama SAO, Fisher B, Green RE, Hepelwa AS, Hernández-Sirvent A, Kajembe GC,
Kulindwa K, Lund JF, Mbwambo L, Meilby H, Ngaga YM, Theilade I, Treue T, Vyamana VG,
Turner RK (2012) Towards transferable functions for extraction of non-timber forest products:
a case study on charcoal production in Tanzania. Ecol Econ 80:48–62
Schaefer K, Lantuit H, Romanovsky VE, Schuur EAG (2012) Policy implications of warming
permafrost. United Nations Environment Programme, Geneva
Shvidenko A, Barber CV, Persson R, Gonzalez P, Hassan R (2005) Forest and woodland systems.
In: Millennium Ecosystem Assessment (ed) Ecosystems and human well-being: current state
and trends. Island Press, Washington, DC, pp 585–622
Sileshi G, Akinnifesi FK, Ajayi OC, Chakeredza S, Kaonga M, Matakala PW (2007) Contributions
of agroforestry to ecosystem services in the Miombo eco-region of eastern and southern Africa.
Afr J Environ Sci Technol 1:68–80
Soja AJ, Tchebakova NM, French NHF, Flannigan MD, Shugart HH, Stocks BJ, Sukhinin AI,
Parfenova EI, Chapin FS, Stackhouse PW (2007) Climate-induced boreal forest change: pre-
dictions versus current observations. Global Planet Change 56:274–296
Stocks BJ, Fosberg MA, Lynham TJ, Mearns L, Wotton BM, Yang Q, Jin JZ, Lawrence KM,
Hartley GR, Mason JA, McKenney DW (1998) Climate change and forest fi re potential in
Russian and Canadian boreal forests. Clim Change 38:1–13
TEEB (2010), The Economics of Ecosystems and Biodiversity Ecological and Economic
Foundations. Edited by Pushpam Kumar. Earthscan, London
Ten Brink P, Alkemade ME, Bakkenes M, Eickhout B, de Heer M (2006) Cross-roads of planet
Earth’s life. Exploring means to meet the 2010-biodiversity target. Netherlands Environmental
Agency, The Netherlands
Thompson ID, Okabe K, Tylianakis JM, Kumar P, Brockerhoff EG, Schellhorn NA, Parrotta JA,
Nasi R (2011) Forest biodiversity and the delivery of ecosystem goods and services: translating
science into policy. BioScience 61:972–981
Tietenberg TH (1996) Environmental and natural resource economics, 4th edn. HarperCollins
College Publishers, New York
UNEP (2007) Global environment outlook 4. United Nations Environment Programme, Valleta,
Malta
USDA (1999) Roads analysis: informing decisions about managing the national forest transporta-
tion system. Report FS-643. United States Department of Agriculture, Washington, DC
Van Kooten GC (2007) Economics of forest ecosystem carbon sinks: a review. Int Rev Environ
Resour Econ 1:237–269
C. Marta-Pedroso et al.

137
Vergara W, Scholz SM (2011) Assessment of the risk of Amazon dieback. World Bank Study. The
World Bank, Washington, DC
Visseren-Hamakers IJ, McDermott C, Vijge MJ, Cashore B (2012) Trade-offs, co-benefi ts and
safeguards: current debates on the breadth of REDD+. Curr Opin Environ Sustain 4:646–653
Vorra V, Barg S (2008) Pimachiowin Aki World Heritage Project Area—ecosystem services valu-
ation assessment. Prepared for the Pimachiowin Aki Corporation. International Institute for
Sustainable Development, Winnipeg, MB
Wang S, Fu B (2013) Trade-offs between forest ecosystem services. For Policy Econ 26:145–146
Wilson MA, Hoehn JP (2006) Valuing environmental goods and services using benefi t transfer: the
state-of-the art and science. Ecol Econ 60:335–342
Wilson MA, Howarth RB (2002) Discourse-based valuation of ecosystem services: establishing
fair outcomes through group deliberation. Ecol Econ 41:431–443
Wise M, Calvin K, Thomson A, Clarke L, Bond-Lamberty B, Sands R, Smith SJ, Janetos A,
Edmonds J (2009) Implications of limiting CO2 concentrations for land use and energy.
Science 324:1183–1186
Zerbe RO, Bellas AS (2006) A primer for benefi t-cost analysis. Edward Elgar Publishing,
Cheltenham, UK
5 Changes in the ecosystem services provided by forests and their economic…