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BIOTROPICA 42(1): 10–20 2010
10.1111/j.1744-7429.2009.00567.x
The Importance of Defining ‘Forest’: Tropical Forest Degradation, Deforestation,
Long-term Phase Shifts, and Further Transitions
Francis E. Putz1
Department of Biology, University of Florida, Gainesville, FL 32611-8526, U.S.A.
and
Kent H. Redford
Wildlife Conservation Society, 185th and Southern Boulevard, Bronx, NY 10460, U.S.A.
ABSTRACT
While research continues on the causes, consequences, and rates of deforestation and forest degradation in the tropics, there is little agreement about what exactly is
being lost, what we want back, and to whom the ‘we’ refers. Particularly unsettling is that many analyses and well-intended actions are implemented in fogs of
ambiguity surrounding definitions of the term ‘forest’—a problem that is not solely semantic; with development of markets for biomass carbon, vegetation classification exercises take on new relevance. For example, according to the basic implementation guidelines of the Kyoto Protocol, closed canopy natural forest could be
replaced by monoclonal plantations of genetically engineered exotic tree species and no deforestation would have occurred. Following these same guidelines, carbon
credits for afforestation could be available for planting trees in species-rich savannas; these new plantations would count towards a country moving towards the ‘forest
transition,’ the point at which there is no net ‘forest’ loss. Such obvious conflicts between biodiversity conservation and carbon sequestration might be avoided if ‘forest’
was clearly defined and if other vegetation types and other ecosystem values were explicitly recognized. While acknowledging that no one approach to vegetation
classification is likely to satisfy all users at all scales, we present an approach that recognizes the importance of species composition, reflects the utility of land-cover
characteristics that are identifiable via remote sensing, and acknowledges that many sorts of forest degradation do not reduce carbon stocks (e.g., defaunation) or canopy
cover (e.g., over-harvesting of understory nontimber forest products).
Key words: biodiversity; carbon sequestration; climate change; land-use change; plantation; REDD; savanna; scrub.
DESPITE LACK OF CONSENSUS OVER WHAT IS AT STAKE, multitudes of
efforts are underway to protect forests, to manage them sustainably
where outright protection is not a viable or desirable option, to
restore them where they are degraded, to re-create them where they
have been lost, and to create them anew in areas that were not historically forested. These efforts are motivated by a variety of concerns including the desire to maintain or restore ecosystem
functions (e.g., biogeochemical cycles), to protect biological diversity, to make money, to improve the welfare of rural people, to
wrest control of land from rural people, and to preserve opportunities
for research and recreation. Unfortunately, discussions about forests
and their fates are plagued by fundamental differences in what
different people mean when they invoke the term ‘forest’ (Lund
2002, Perz 2007). We argue that instead of being simply semantic,
these differences have potentially grave conservation, development,
climate, and livelihood consequences. In particular, the commodification of biomass carbon makes clarification of vegetation and
land cover classification terms critical lest biodiversity, human welfare, and other ecosystem services be jeopardized. We stress that if
people are careful to specify the ecosystem states they have in mind,
then communication and conservation will both be fostered.
While endeavoring to clarify what is meant by ‘forest’ and the
names used for other vegetation types, we realize that no one classification system is ever going to be accepted by all relevant stakeReceived 26 August 2008; revision accepted 19 April 2009.
1
Corresponding author; e-mail: [email protected]
10
holders, nor is it likely to be appropriate for all applications (e.g.,
global vegetation mapping, estimating carbon stocks, monitoring
biodiversity losses, and fostering sustainable management). Foisting
a classification system on stakeholders who do not share our predominantly Western scientific vision and our biome biases is not
our intention. Furthermore, we recognized that the boundaries between categories in any classification system are unavoidably fuzzy
and sometimes overlapping. These limitations notwithstanding, we
present an approach to ecosystem classification that can be adapted
to local conditions, that will improve communication, and that
should promote the success of conservation and development
efforts. Our focus is on forests and their derivatives, which is itself
worrisome insofar as whatever classification system is used, it is imperative that ecosystems other than forests be recognized. Our concern is that if tropical landscapes are falsely portrayed as either being
forested or nonforested, other valuable ecosystems will be put in
jeopardy. Furthermore, use of this false dichotomy could obscure
substantial losses in ecological values (e.g., carbon losses from natural forest degradation; Foley et al. 2007, Tucker et al. 2008) from
areas that still qualify as being forested. Finally, when a country or
province is reported to have passed the ‘forest transition’ (sensu
Mather 1990) from net forest loss to forest recovery, what was lost
could be very different from what is being gained if unclear or
overly simple definitions of ecosystem states are used (e.g., plantations are treated as forests).
In this paper we discuss the historical bases for classifying an
area as being forested and then examine some of the ways the term
r 2009 The Author(s)
Journal compilation r 2009 by The Association for Tropical Biology and Conservation
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Importance of Defining ‘Forest’
‘forest’ has been used in social, political, and scientific discourse.
We endeavor to show why people engaged in debates over the fates
of forests, particularly in the tropics, should be explicit about the
perceived starting conditions and desired ending conditions of the
ecosystems under consideration (Farley 2007).
WHAT IS ‘FOREST’? ENGLISH ETYMOLOGY
AND POWER POLITICS
Excavating the historical roots of the word ‘forest’ serves mostly to
illustrate the contradiction between the modern use, which implies
an area with a closed canopy of trees, and the word’s original meaning (e.g., Young 1979, Harrison 1992, Agnoletti & Anderson 2000,
Vera 2000). Apparently the term was first used in English to indicate uncultivated land of a variety of vegetation types, including
grasslands, which were claimed by royalty. In a process that continues to this day, 7th Century Frankish kings claimed rights to forestis, uncultivated hunting and swine foraging areas with no clear
owner (at least from the royal perspective). This interpretation is
based on forestis being derived from the Latin foris, which can mean
‘outside the settlement’ (Harrison 1992). Rackham (1998) made
the case that in England, ‘forest’ traditionally meant a place with
deer. When William the Conqueror of Normandy invaded in
1066, he built on this concept with his Forest Law, which proclaimed his sovereign rights to all wild animals, particularly deer.
This effort at securing the rights to game and other natural resources codified the desires of Anglo Saxon rulers for dominion
over hunting grounds (Perlin 1991). From such royal attempts at
criminalizing traditional uses and securing dominion over uncultivated lands spring the tales of Robin Hood of Sherwood Forest
(Watkins 1998), the War of the Demoiselles over forest access in
19th Century France (Sahlins 1994), the sultanization and then
federalization of Javan teak ‘forests’ (Peluso 1992), and ongoing
struggles over land-use rights in many developing countries in the
tropics (e.g., Westoby 1989, Dawkins & Philip 1998). What is too
often overlooked in discussions about forest-use rights through history is that while the basic social issue of contested access remains
relevant, uses of the term ‘forest’ have continued to evolve.
In much of the tropics where the rights of indigenous peoples
are often in contention, ‘forest’ is often defined on cultural grounds
(e.g., Brosius 1997). Given the global focus on forest conservation,
there is a degree of political expediency in being associated with
forest, even for farmers who fell forests to cultivate sun-loving
crops. Walker (2005) goes so far as to refer to this phenomenon as
the ‘arborealisation’ of conservation and development agendas.
LOOKING FOR CLARITY ABOUT ‘FOREST’ IN
THE PROFESSION OF FORESTRY
Forestry can be defined as the science of cultivating forests (Webster
1983) or more broadly, such as in the definitions adopted by the
Society of American Foresters and the International Union of Forest Research Organizations that include consideration of forest creation, use, and conservation (Helms 2002). Unfortunately, from its
inception as a self-aware discipline in Europe, forestry has been
11
motivated by the oft-conflicting wishes of stakeholders who conceive of forests in quite different ways. From the perspective of
many foresters and their official historians (e.g., Robinson 1975,
Steen 1976), forestry developed as a utilitarian profession in response to society’s need to institute rational management of a rapidly degrading resource; that is, forest management for the general
good of society. In bold contrast, a number of social historians and
radical environmentalists portray the history of the forestry profession in decidedly less favorable light by focusing on its imperial,
military, and industrial links (e.g., Shepherd 1975, Liphschitz &
Biger 1998).
The emphasis on cultivating forests, in general, and tree planting, in particular, in many definitions of forestry can be troublesome for both conservationists and foresters. For example, in
evaluations of the environmental consequences of a country passing the ‘forest transition,’ the assumed equivalence of plantations
and natural forests is fraught with environmental problems (Rudel
et al. 2005). For the purposes of land-use mapping, at least intensive plantation management should be considered an agricultural
activity. Ironically, the perceived need to plant trees can frustrate
silviculturalists entrusted with management of areas already amply
stocked with natural regeneration of the commercial species; protection and perhaps silvicultural release might be needed, but not
replacement (e.g., Wadsworth & Zweede 2006, Peña-Claros et al.
2008a, b). Ultimately, trying to ferret out what is a ‘forest’ through
consideration of the forestry profession leads only to the conclusion
that forests are what foresters manage. In defense of the profession,
the broad definition of forests endorsed by many foresters is in accord with the historical breadth of the term and is in keeping with
society’s broad expectations of their roles as land stewards.
‘FOREST’ AS DEFINED BY INTERNATIONAL
TREATIES AND NATIONAL LAWS
The multitude of ways in which ‘forest’ is defined (Lund 2002) has
myriad legal and practical implications in the contemporary world.
The fundamental problem is that any definition that is appropriate
for global (e.g., Bartholomé et al. 2002) and continental-scale land
cover mapping (e.g., Eva et al. 2004), is not likely to be appropriate
for monitoring biodiversity losses and intensities of forest use that
do not qualify as deforestation but that result in substanital carbon
fluxes. At such large scales, what is needed is a uniform set of basic
parameters that can be mapped and monitored using readily available remote sensing technology.
The vegetation classification system that the Food and Agriculture Organization of the United Nations developed for its longestablished global forest cover monitoring program (e.g., FAO
2001) provides the basics used by the Global Land Cover (Bartholomé et al. 2002) mapping project and by the United Nations
Framework Convention on Climate Change (UNFCCC) for implementation of the Kyoto Protocol. The FAO’s basic definition of
forest is an area of 4 0.5 ha with 4 10 percent tree canopy cover,
with ‘trees’ defined as plants capable of growing 4 5 m tall (FAO
2001). The implementation guidelines of the Clean Development
Mechanisms (CDM) of the Kyoto Protocol (i.e., the Marrakech
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Putz and Redford
Accords) are similar but allow each country to define ‘forest,’ at
least to the extent of being allowed to choose the minimum treecrown cover required (10–30 percent), the minimum area of this
cover (0.05–1.0 ha), and the minimum heights to which woody
plants must be able to grow to be considered trees (2–5 m). At the
next level of detail in their hierarchy of land cover types, the FAO
distinguishes closed (4 40 percent canopy cover) and open (10–40
percent canopy cover) natural forests from plantations, but this distinction is often overlooked and is not reflected in CDM guidelines.
National laws vary even more in regards to what is and what is not
considered forest. In the Philippines, ‘forests’ do not legally occur
on slopes 4 18 percent. In Spain, Finland, Norway, and Sweden,
an area is not forested if it produces o 1 m3/ha/yr of timber
whereas in Ireland, the minimum production for forest is 4 m3/ha/yr
(Lund 2002).
There is a clear need for widely accepted definitions of forest,
deforestation, forest degradation, and forest restoration that are politically expedient but culturally sensitive, ecologically reasonable,
and technologically feasible. The danger of overly simple definitions is that they can obscure substantial losses in what most people
value as forest (Kleinn 2001). The needs for simple and globally
applicable land cover classes notwithstanding, any definition based
solely on canopy cover fails to recognize that forests are more than
trees. An example of this problem is that following the rules of
the Kyoto Protocol, up to 70–90 percent of the canopy of an
initially closed-canopy forest could be destroyed and no deforestation would have occurred. Worse yet, even if the forest is entirely
razed, as long as plants (including tree seedlings, palms, and
bamboos) capable of growing to 4 2 or 5 m are present on
4 10–30 percent of the site will be allowed to regenerate, no deforestation would have occurred. Similarly, according to the FAO
or under the guidelines of the Kyoto Protocol, old growth natural
forests could be defaunated, clear-cut, or replaced by monoclonal
stands of genetically modified exotic tree species grown for oil
or fiber on 5–10 year rotations with no change in ‘forest’ cover
(Niesten et al. 2002, Bekessy & Wintle 2008). Many policy-makers
are cognizant of these sorts of threats to ecological integrity, but
tropical biologists need to make sure that policy debates are well
informed.
FOREST AND ITS DERIVATIVES
Although we realize that simple rules are needed for mapping and
monitoring land cover at large spatial scales and recognize that the
rules adopted were settled on only after extensive and continuing
international debate, we are obviously uncomfortable with the way
‘forest’ is defined by global policy-makers and land cover mappers.
To rectify this situation, we turned to our own discipline for clarity.
Unfortunately, ecologists have never agreed on an explicit definition of ‘forest’ or of any other type of vegetation for that matter,
and have often failed to differentiate between ‘potential’ and actual
land cover types (e.g., Merriam 1898; Schimper 1903; Weaver &
Clements 1929; Beard 1944, 1955; Richards 1952; Holdridge
1967; Daubenmire 1968; Walter 1971; Mueller-Dombois &
Ellenberg 1974; Tracey 1982; but see Oliveira-Filho & Ratter
2002). Nevertheless, we find a fair degree of consensus in the way
the world’s major ecosystem/vegetation/land cover types have been
categorized. In particular, along the continuum from treeless to
fully treed, the following vegetation types are typically designated
(following Jennings et al. 2009): grasslands are dominated by graminoids ( = grasses and grass-like plants) with o 5 percent treecrown
cover (note that we define a tree as a woody plant 4 5 m tall);
savannas are graminoid dominated with scattered trees (5–20 percent treecrown cover); woodlands have graminoids in the understory
and 20–40 percent treecrown cover; open forests have 40–70 percent
treecrown cover; and, closed forests have 4 70 percent cover of treecrowns. Unfortunately, tropical ecosystems are not so easily classified; perhaps the single biggest failing of a system based solely on
tree cover is lack of recognition of different sorts of scrub (e.g., cerrado, chaparral, fynbos, thicket, garrigue, mallee, maquis, and sand
heath), ecosystems dominated by woody plants that exceed the
lower height limit of 5 m but may never get much taller. There are
also some shrub-dominated areas in the tropics, particularly near
timberline and on thin, rocky, peaty, or heavy-metal-rich soils
(Huber 1989), but they are not extensive. Surprisingly, ‘shrublands’
figure prominently in many modern land cover maps (e.g., Eva et al.
2004) and, for policy-informing and ecosystem state-change-predicting purposes, are often combined with grasslands and savannas
into one category (e.g., Malhi et al. 2009). This change in map legends occurred at least partially because whereas most ecologists
use the admittedly difficult to apply definition of ‘shrub’ as a multiple stemmed woody plant, for the FAO and its followers, shrubs
are ‘woody perennial plants, generally 4 0.5 m and o 5 m in
height at maturity without a definite crown’ (Schoene et al. 2007).
Nonforest vegetation types can be ancient, rich in endemic
species, and otherwise of great ecological value, but at least in a
physiognomic sense, can also result from forest degradation. For
example, whereas from a floristic perspective it is wrong to equate
sparsely wooded cattle pastures with savannas, from physiognomic
and remote-sensing perspectives, the two are difficult to distinguish.
For this reason, we follow the established tradition of designating
such clearly anthropogenic nonforest vegetation types as derived
woodland, derived savanna, and derived scrub so as to not confuse
them with structurally similar but generally much more species-rich
communities (Richards 1952).
There are widespread fears of savannas replacing forests in the
Amazon Basin in particular and the tropics in general (Fairhead
& Leach 1996, Cox et al. 2004, Nepstad et al. 2008, Malhi
et al. 2009), but attention should also be given to the scrubs that
figure prominently in the analyses of Beard (1955), Holdridge
(1967), and Walter (1971). In these ecosystem classification systems, various sorts of scrub are expected to replace forest in lowland
tropical areas where potential evapotranspiration exceeds precipitation; savanna is taken to be a ‘dysclimax’ that only develops where
frequent fires or extreme pressure from browsing herbivores allow
grasses to invade. We wonder, for example, whether, in addition
to being concerned about ‘savannization,’ we should also address
the threat of ‘scrubification.’ Perhaps the wide dissemination of
highly productive and extremely flammable African C-4 pasture
grasses throughout the Neotropics (Daubenmire 1972, Parsons
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Importance of Defining ‘Forest’
1972) coupled with forest fragmentation, climate drying, and
substantial increases in wildfire ignition by cattle-rearing and other
people have tipped the balance toward savanna, but forest-scrub
transitions also seem worthy of study. Complicating matters is the
fact that, whereas some scrubs, woodlands, and savannas are recently
anthropogenic, others are ancient and exceedingly rich in biodiversity and other ecological values. By denigrating these imperiled
ecosystems as simply the by-products of forest degradation, we risk
losing a great deal of conservation value.
If we focus solely on forests but acknowledge that factors other
than physiognomy or carbon stocks are to be considered, then it
would help to clarify exactly what is at risk. While debates rage on
over whether a particular forest is primary, pristine, intact, natural,
virgin, or otherwise not anthropogenic (e.g., Clark 1996, Clement
1999), we need to start with an operational definition of ‘forest’
that reflects a diversity of ecological components, clarifies what is at
risk, and promotes specification of the objectives of restoration
efforts. Similarly, reference states need to be defined so that relative
degrees of degradation can be compared among dissimilar forests
(Tucker et al. 2008). An exemplary approach to defining a particular sort of reference state is provided by the U.S. Forest Service’s
National Old-Growth Task Group. They define old growth as being
‘. . . distinguished by old trees and related structural attributes. Old
growth encompasses the later stages of stand development that typically differ from earlier stages in a variety of characteristics that often include tree size, accumulations of large dead woody material,
number of canopy layers, species composition, and ecosystem function.’ Recognizing that what constitutes ‘old’ varies from forest to
forest, as do the attributes of old growth, the Task Force promoted
publication of explicit old growth definitions for a variety of forest
types in the United States. For example, old growth Douglas fir
Pseudotsuga Menziesii in the Pacific Northwest of the United States
is 4 175 years old, with the largest trees 4 100 cm dbh (diam. at
1.3 m), and with substantial stocks of coarse woody debris (Franklin et al. 1981). In stark contrast, old growth sand pine Pinus clausa
in Florida can be as young as 45 years with the largest trees
4 26 cm dbh (Outcalt 1997). A similar approach is being used for
the diverse suite of old growth vegetation types in Australia (Burgman 1996). Old growth woodlands, scrub, savanna, and grassland
also deserve consideration, but here we restrict the discussion to
forests and their derivatives.
Although we use old growth forest as the basic reference state in
our classification system (Fig. 1), what is of fundamental importance is clarity about what is being lost or gained. We acknowledge
that if conservation efforts are to be successful, reference state selection should reflect the wishes and concepts of local people (e.g.,
Moyer et al. 2008) and not be imposed by scientists, however well
intentioned. Nonetheless, we do not address this fundamental consideration other than to point out that where local and extra-local
stakeholder visions differ, appropriate compromises need to be
fairly negotiated. At least in the biophysical sense, restoration ecologists are well aware of the importance of identifying reference
states (e.g., Egan & Howell 2001, but see Pickett & Parker 1995)
even when prime ecosystem examples no longer exist (Dettman &
Mabry 2008) and when the climate and other influential factors are
13
changing (Fulé 2008). In contrast, reference states are seldom made
explicit in discussions of tropical forest degradation, deforestation,
and recovery; designating these states becomes more challenging
with the near disappearance of some formerly extensive forest types
(Ashton 2008).
Old growth forests that have lost their defining compositional
and structural attributes due to wildfires (or the converse in firemaintained ecosystems), invasion by exotic species, over hunting, or
other substantial human interventions are classified as degraded
forest; other researchers have used the term ‘nonintact’ and a
variety of other adjectives for similar state changes (Clement 1999,
Mollicone et al. 2007). Uncontrolled, repeated-entry logging with
attendant hunting followed by fire is a common sequence of forest
degrading processes in many parts of the tropics (Nepstad et al.
1999, Cochrane 2003). Defaunation by over-hunting is another
way in which old growth forest is lost; we classify the resulting ‘empty
forests’ (Redford 1992, Terborgh et al. 2008) as degraded forest. In
this admittedly coarse classification system, the ‘novel ecosystems’
that develop with incorporation of exotic species or after massive
biogeochemical changes would also be considered degraded forest if
they retain the defining characteristics of forest.
Forests that are purposefully domesticated for production of
timber and nontimber forest products should be referred to as managed forests; the ‘semi-natural forests’ of the Forest Stewardship
Council (FSC) (2009) would presumably be included in this category. We recognize that this designation suffers insofar as one person’s management is another’s degradation, that many forests
formerly considered pristine show vestiges of historical human interventions, and that considerable power derives from the right to
define these states (Tsing 2005). Furthermore, the extent of management-induced modification depends on whether maximization
of production is the sole objective of management and on the silvicultural treatments employed. And finally, modern commercial
production forestry differs substantially from the range of indigenous forest domestication processes that led Clement (1999)
to differentiate among what he referred to as pristine, promoted,
managed, cultivated, and swidden/fallow landscapes. In our more
simplistic approach, the defining differences in structure and composition between managed and old growth forest often include
reductions in the densities of adults of commercial species and,
where logging has occurred, increases in the densities of lightdemanding plant species.
Forests that develop after complete deforestation should be
referred to as secondary forests. Although some researchers do not
distinguish between secondary and degraded forest, we believe that
these types of forest are different enough in structure, composition,
dynamics, and management options to warrant distinction. Secondary forests that develop on land used for intensive agriculture and
pastures, for example, regenerate mostly from seeds that are either
dispersed in from nearby remnant forest or, to a lesser extent, germinate from buried dormant seeds. In contrast, regeneration after
logging in managed or degraded forests is typically dominated by
vegetative recovery of plants that remained on site through the episode of damage. Furthermore, although secondary forests rapidly
develop some of the structural, compositional, and functional
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Putz and Redford
FIGURE 1. Ecological state changes starting from the reference condition of tropical old growth forest. Only the principal drivers are included (in hexagons) and not all
possible transitions away from the reference condition are depicted. Back transitions that result from natural succession as well as afforestation, reforestation, forest
management, and other restoration activities are not indicated to reduce figure clutter.
characteristics of old growth forest (Chazdon 2003), other features
recover only after centuries if ever (Clark 1996).
As mentioned above, to help avoid the major losses in biodiversity that result from converting forests into plantations (e.g.,
Barlow et al. 2007, Fitzherbert et al. 2008), these two categories
should not be combined. The Dictionary of Forestry (Helms 1998)
defines plantation as ‘a stand composed primarily of trees established by planting or artificial seeding.’ Using this broad definition,
plantations include a tremendous range of management intensities,
structures, and compositions. At one end of this continuum are industrial tree farms comprised of genetically modified monoclonal
exotic tree species planted after mechanical removal or broadcast
herbicide treatment of all competing vegetation and clear-cut at
5–25 yr intervals. At the same end of this continuum, we place oil
palm plantations, intensively managed rubber estates, and other
types of tree farms. Although these stands of perennials may somewhat resemble forests in ecosystem functions such as sequestration
of carbon, nutrient cycling, and hydrology, they are very different
in structure and often completely different in composition and,
hence, we call them plantations not ‘planted forest’ and include
them among other agricultural land-uses. At the other end of the
continuum of management intensity from intensively managed
monocultures of exotic tree species are rustic coffee and cacao agroforests (Bhagwat et al. 2008), ‘forest gardens’ in Indonesia (Padoch
& Peters 1993), and the substantially enriched, promoted, and
otherwise modified ‘seminatural’ forests of the Amazon (Roosevelt
1980, Posey 1984, Balee 1994, Clement 1999, Lentz 2000). Some
of these managed communities are so ‘natural’ in structure and
composition that they are often mistaken for old growth forest and
in many cases should be classified as managed forest. Also complicating differentiation between plantations and managed forests are
planted stands of native trees with more-or-less intact understories
of native groundcovers. Examples of this intermediate condition
include the degraded forests in Malaysia that are gradually converted
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Importance of Defining ‘Forest’
into plantations of native dipterocarps by enrichment planting
(Moura Costa et al. 1996) and the intensively managed groves of
Euterpe oleraceae palms near the mouth of the Amazon (Brondizio
et al. 1994).
The issue of whether plantations should be considered a type
of forest is politically quite contentious. The World Rainforest
Movement (http://www.wrm.org.uy) and the Dogwood Alliance
(http://www.dogwoodalliance.org), for example, have mounted
substantial campaigns against the FSC’s certification of plantation
management. Although the FSC will not certify plantations that
replaced forests after 1994, it is blamed for contributing to the confusion over what is and what is not ‘forest,’ thereby inadvertently
contributing to the loss of natural ecosystems.
Narrowly defining ‘forests’ and differentiating them from
woodlands, savanna, shrublands, and scrub may make ecological
sense but it may also cause these nonforest ecosystems to be further
threatened. The arborealization of global conservation agendas can
cause conflicts when nontree values, like wildlife, and naturally treescarce ecosystems are in jeopardy. This focus on trees derives in part
from recognition of the need to consider the fate of biomass carbon
in efforts to mitigate the climate-changing effects of fossil fuel combustion. Nevertheless, ecosystems other than forests sequester carbon, and a ‘forest’ definition that may suffice from carbon and
remote sensing perspectives is not likely to satisfy concerns about
biodiversity, cultural amenities, and other ecosystem services.
DEFINING DEFORESTATION AND FOREST
DEGRADATION
Although the attention paid to deforestation in the tropics is well
justified, it is worrisome that the more extensive and often more
insidious processes of forest degradation are often overlooked. One
reason for this neglect is that ‘degradation’ denotes loss of values
that are subjective and thus often ignored by scientists. Another
reason for degradation’s disregard is that it is more difficult to
detect using remote sensing techniques. Furthermore, while deforestation can be portrayed as a binary variable (deforested or not),
degradation occurs on continua that pass through numerous nonorthogonal dimensions and follow trajectories affected by thresholds and other nonlinearities. Familiar forms of tropical forest
degradation include nonsustainable (but still ‘selective’) harvesting
of timber and nontimber forest products, over-hunting, and slowmoving low intensity understory fires that kill small trees and
shorten the life expectancies of large ones.
Improved access promotes both deforestation and forest
degradation, but the latter often proceeds much farther beyond the
reach of roads or the law. For example, given the extremely high
price per gram of gaharu Aquilaria malaccensis wood incense, we
doubt there is any place on the island of Borneo where trees of this
species are beyond the reach of harvesters. Similarly, bush meat
hunters will walk great distances for game, which renders it challenging to maintain populations of large, charismatic, edible, or
otherwise marketable species anywhere within 10–50 km of the
nearest settlement. And when the hunted species are migratory or
are drawn to saltlicks or the fresh foliage on hunted-out roadsides,
15
even stationary hunters can defaunate huge areas (Robinson &
Bennett 2004, Nasi et al. 2008).
While much attention has been paid to the deforestation that
results from perverse policies, such as those that require forest clearing before land titles are granted (Williams 2003), under a range of
circumstances, forest degradation is likely in places where outright
deforestation might otherwise be expected. For example, if remote
sensing is used for official forest monitoring and if regulations prohibiting deforestation are enforced, forest exploiters may choose to
restrict themselves to degradation rather than outright deforestation. Similarly, forest exploiters who are also guerillas, smugglers,
and drug plant cultivators might stop short of radically opening the
canopy and thus exposing their activities to aerial detection. Finally,
where land-use intensification is limited by access to capital, forest
degradation is more likely than deforestation until better capitalized
actors enter the area or the costs of conversion go down, such as
when access improves as a result of road building or paving
(Chomitz et al. 2007).
The Kyoto Protocol provides no provisions for reducing emissions of carbon dioxide and other atmospheric heat-trapping gases
by avoiding deforestation or forest degradation but does provide
carbon credits for ‘reforestation’ and ‘afforestation’ of areas that
were not forest covered on 1 January 1990 or 1 January 1940,
respectively. One of the real-world consequences of the political
process that resulted in this international agreement is that countries that selected a higher treecrown cover threshold (i.e., 30% vs.
10%) for defining forest generally have substantially more land eligible for carbon credits from tree planting (Verchot et al. 2007,
Zomer et al. 2008). What we find worrisome is that under these
guidelines, carbon funding could be made available for replacing
species-rich grasslands, savannas, and woodlands with trees or deep
rooted plants that sequester more carbon than the native ecosystem
(Fisher et al. 1994). Hopefully, the systems used in the voluntary
carbon markets to assure that forest carbon is sequestered in socially
and ecologically sound manners (e.g., Hamilton et al. 2008) will be
adopted by the regulated carbon markets expected to emerge from
international climate change convention under negotiation.
With the Kyoto Protocol set to expire in 2012, the UNFCCC
is now in the process of developing a new climate change agreement. Recognizing that the fate of tropical forest carbon cannot be
disregarded if cataclysmic climate change is to be averted, negotiators of the new convention are discussing the ‘reduced emissions
from deforestation and forest degradation’ (REDD) option (Gibbs
et al. 2007). Cost-effectiveness has drawn attention to REDD
(Stern 2006, Kinderman et al. 2008), but most emphasis in the debates has been on deforestation and not on degradation. We are
concerned that the second ‘D’ in this now familiar acronym is often
dropped altogether (e.g., Hall 2008), mentioned but then disregarded (Gibbs et al. 2007), or is interpreted as standing for ‘developing countries’ instead of ‘degradation’ (UNFCCC 2006, Putz
et al. 2008).
Whereas degradation is often obvious from the ground, it is
more difficult to detect from space. Fortunately, recent improvements in remote sensing technologies are increasing our ability to
detect the effects of selective logging on canopy closure (Asner et al.
SPECIAL SECTION
16
Putz and Redford
2005, Oliveira et al. 2007, Broadbent et al. 2008, Duveiller et al.
2008, Palace et al. 2008), the effects of understory fires (Alencar
et al. 2006), and even the understory invasion of exotic species (Asner & Vitousek 2005, Asner et al. 2008). Nevertheless, other sorts
of forest degradation are and will remain invisible from satellites
(e.g., over-hunting; Peres et al. 2006) and, thus, ground-based studies will continue to be needed if the extensive, pervasive, and pernicious impacts of forest degradation are to be avoided.
If some of the attention of environmental policy-makers can
be shifted from deforestation to forest degradation, the differences
between the two processes will need to be clarified (Southworth
et al. in press). If degradation is defined solely in terms of reductions
in canopy cover, then degradation proceeds up to the threshold of
deforestation (e.g., 10–30% treecrown cover, as set by Marrakech
Accords), but could just as well be based on the limitations of the
remote sensing technology utilized. Unfortunately, even this simplified portrayal of the conversion process fails to distinguish between forest degradation and forest management. For example,
what looks like degradation from space, as defined as canopy losses
that do not constitute deforestation, might intentionally result from
silvicultural treatments applied to promote regeneration of lightdemanding species. Similarly, removal of invasive exotic and nuisance native trees is often a step towards ecosystem restoration and
does not constitute degradation (Condon & Putz 2007).
FOREST TRANSITIONS
In the ecosystem classification system we propose, old growth forests
that lose their defining attributes (e.g., ancient trees, fauna, and
coarse woody debris) through logging, market hunting, wildfires, or
invasion by exotic species, become degraded forest (Fig. 1). In contrast, where people purposefully change forest structure and composition through silvicultural treatments to favor economically
valuable species, the result is managed forest. Where forest managers attempt to mimic natural disturbance regimes or where they explicitly manage for old growth characteristics (Landres et al. 1999,
O’Hara 2001), the resemblance between managed forest and old
growth forest can be substantial. Boundaries of managed forest are
similarly difficult to set because the degree of domestication varies
so greatly (Clement 1999). Contrast, for example, a forest that was
managed by stone-tool wielding hunter–gatherers for the production of medicinal plants and other nontimber forest products from
native species (Balee 1994) with a forest silviculturally managed for
commercial timber through heavy logging, liana cutting, and girdling of all noncommercial trees (e.g., the Malayan Uniform System; see Lee et al. 1998 for an assessments of such an area 34 years
after felling and treatment). But even where human interventions
are intensive, such as in species-rich and multi-storied agroforests
dominated by native trees species (e.g., Michon et al. 1986), many
of the structural, floristic, and functional features of old growth forest
are maintained (but see Bhagwat et al. 2008).
In response to uncontrolled multiple entry logging followed by
fire and other perturbations, ecosystem ‘phase’ shifts are likely from
old growth forest to a variety of physiognomically and floristically
distinctive states including derived scrub, derived woodland, and
derived savanna, depending on canopy cover and tree stature
(Fig. 1). Which of these derived vegetation types develops depends
mostly on the fire regime, which is in turn mostly determined by
the presence and abundance of fire-favoring grasses and social factors. In particular, in response to severe disturbances or climate
drying, where grasses are scarce, forest is most likely to be replaced
by derived scrub rather than the fire-dependent derived woodland or
derived savanna.
The fates of deforested areas left to regenerate also vary greatly
with the type, duration, and spatial extent of nonforest land-uses.
Where land uses were of low intensity and short duration, such as in
the small clearings of swidden agriculturalists, natural regeneration
can be rapid from resprouting, germination of seeds from the buried seed bank, or establishment from seeds dispersed in from nearby
forest remnants. In contrast, large and hard-used pastures and
plowed lands develop only slowly into very different sorts of secondary forests or not at all if fires recur. These secondary forests may
become old growth after decades or centuries of protection, unless
they are colonized by exotic species or if forest succession is prevented by climate change. They can also be silviculturally treated in
ways that speed the transition into old growth, managed for any of a
variety of forest products and services, or degraded by logging or
repeated fires to the extent that they become derived woodland,
derived scrub, or even derived savanna. Alternatively, if planted with
either native or exotic tree species, deforested areas can become
plantations or even urban forests, an important category we have not
discussed. Finally, plantations that are abandoned or otherwise not
intensively managed for fiber, fuel, oil, latex, or other products are
generally colonized by native and exotic species, which eventually
convert them into secondary forests. Unfortunately, despite the flurry
of publications on using plantations to facilitate forest succession
on degraded sites in the tropics (e.g., Parrotta 1992, Parrotta &
Turnbull 1997), it is not clear that this phenomenon is widespread.
Furthermore, many secondary forests are essentially forest fallows
subject to reclearing when prices for agricultural commodities rise
(Rudel et al. 2002). Finally, recovery towards old growth forest
or any other static reference state can be permanently deflected by
climate change or other alterations in abiotic conditions, species
extinctions, and the existence of persistent exotic species.
CONCLUSIONS
Vegetation classification systems, like the one we present, can serve
to foster communication but admittedly fail to capture the fluidity
of ‘forest’ and other ecosystem types both in nature and as social
constructs. This fluidity is to be expected given the extent to which
‘forest’ is a term used for political and social gains. It is also expected
given the range of ecosystem types being considered, their propensity for change, and the many ways they are perceived, remembered,
and valued. Clearly, confusion over what is and what is not forest is
rooted in substantial cultural ambivalence. As revealed by its etymology, ‘forest’ has historically been outside, beyond the walls, and
at the frontiers of civilization. Indeed, civilization is often portrayed
as pitting itself against forests (but see Frazer 1890). But while forests have always harbored outlaws (some of whom stood for justice),
SPECIAL SECTION
Importance of Defining ‘Forest’
wild men, wild beasts, and the danger of disorientation, they have
also been sources of spiritual enlightenment and tranquility as well
as refuges for the downtrodden (Harrison 1992). Furthermore, at
least since the advent of scientific forestry in Germany in the 18th
Century, forests have also been perceived as sources of raw materials
for society.
Lack of clarity in descriptions of historical ecosystems becomes
more than an academic dilemma when reference states for ecosystem restoration are selected. Were the sacred cedar groves that
King Gilgamesh of Sumeria reputedly destroyed 4000 years ago
(Harrison 1992) forests or woodlands? Could a Pre-Columbian
North American squirrel have traveled from the Atlantic seaboard
to the Mississippi River without touching the ground, or did
Amerindian burning and farming preclude this possibility (Cronon
1983)? Similarly, was the largest remaining tract of ‘pristine’ forest
in Central America a patchwork of maize fields when visited by
Balboa in the early 1500 s (Bush & Colinvaux 1994)? Mounting
archeological and ecological evidence of prior human modification has stimulated some researchers to cast doubts about the virginity of even remote forests in the Amazonian Hylea (Lentz 2000,
Heckenberger et al. 2008), but more recent assessments have suggested that while human influences were substantial, they were also
somewhat localized (Bush & Silman 2007). But while selection
of historically appropriate reference states is an acknowledged challenge, with rapid rates of global change in climate and species
distributions, it is not clear how much history should guide conservation and restoration.
Lack of consensus about what is and what is not ‘forest’ contributes to the haze in which ecological destruction continues unabated. With implementation of market-based approaches to forest
protection, clarification of terms used to describe different ecosystem states is a prerequisite for effective communication, policymaking, and conservation. It is certainly more than semantic when
forests are converted into plantations in the name of climate change
mitigation, and massive deforestation and forest degradation go
unrecorded because of the way ‘forest’ is defined. Furthermore, the
frequency and severity of conflicts between interventions motivated
by the global conventions on biodiversity and climate change
should decline when it is clear what is at stake (e.g., old growth
forest or derived scrub). Finally, in all manners of human discourse,
forests will remain something of a social construct, but one that
should be firmly rooted in ecosystem structure and composition.
ACKNOWLEDGMENTS
We cannot thank all of the people who helped shape our thinking
about this topic but we can at least acknowledge those who suggested improvements on earlier drafts of the manuscript including
V. Medjibe, S. Murphy, A. Shenkin, J. Ash, K. Didier, P. Zuidema,
C. Romero, J. Veldman, P. Brando, R. Chazdon, and an anonymous reviewer. Participation in a workshop on ‘Land Use Transitions in the Tropics’ at Rutgers University provided the final
impetus for preparing this paper.
17
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