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8 What is the state of tropical montane cloud forest restoration?

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The conversion of tropical montane cloud forest (TMCF) to pastures and agricultural lands has been an important activity in this life zone for many years. Although forest clearing and grazing continues, in some areas, changing political, economic, and social drivers have led to the abandonment of marginal areas. These dynamics provide an excellent opportunity to study the rates of secondary succession and test different restoration strategies. The two major questions addressed in this review are: “What factors control rates of TMCF recovery once pastures or agricultural lands are abandoned?”, and “What restoration strategies can be used to overcome barriers to regeneration and accelerate forest recovery?” To answer these questions a literature review was carried out. Because few restoration projects have been conducted in TMCF as such, the conclusions are mainly based on studies in tropical montane forests at large. Competition with invasive grasses and ferns and poor seed dispersal appear to be the most important factors limiting natural forest recovery. To overcome these barriers, one of the most cost-effective ways to accelerate recovery is to promote the establishment of shrubs, which help to shade out invasive grasses and ferns and create more appropriate conditions for seedling growth. Although this strategy can reduce competition, planting will also be required to recover a species composition similar to intact forest because most forest species are rarely dispersed far from forest stands.
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8What is the state of tropical montane
cloud forest restoration?
T. M. Aide and M. C. Ruiz-Jaen
University of Puerto Rico, San Juan, Puerto Rico, USA
H. R. Grau
Universidad Nacional de Tucuman, Tucuman, Argentina
ABSTRACT
The conversion of tropical montane cloud forest
(TMCF) to pastures and agricultural lands has been
an important activity in this life zone for many years.
Although forest clearing and grazing continues, in
some areas, changing political, economic, and social
drivers have led to the abandonment of marginal
areas. These dynamics provide an excellent
opportunity to study the rates of secondary
succession and test different restoration strategies.
The two major questions addressed in this review
are: “What factors control rates of TMCF recovery
once pastures or agricultural lands are abandoned?”,
and “What restoration strategies can be used to
overcome barriers to regeneration and accelerate
forest recovery?” To answer these questions a
literature review was carried out. Because few
restoration projects have been conducted in TMCF as
such, the conclusions are mainly based on studies in
tropical montane forests at large. Competition with
invasive grasses and ferns and poor seed dispersal
appear to be the most important factors limiting
natural forest recovery. To overcome these barriers,
one of the most cost-effective ways to accelerate
recovery is to promote the establishment of shrubs,
which help to shade out invasive grasses and ferns
and create more appropriate conditions for seedling
growth. Although this strategy can reduce
competition, planting will also be required to recover
a species composition similar to intact forest because
most forest species are rarely dispersed far from
forest stands.
INTRODUCTION
Tropical montane ecosystems have been transformed by
human use for hundreds of years. The most common causes
of forest transformation have been agriculture and grazing,
but timber harvesting has also contributed to forest degrad-
ation. Although clearing for illegal crops – e.g. poppy,
Papaver somniferum (Cavelier and Etter, 1995) or coca,
Erythroxylum coca (Bedoya, 1997) – continues to threaten
many areas of intact montane forest in Latin America, in
many other regions, agricultural and grazing lands are being
abandoned, often because of a loss of agricultural product-
ivity associated with accelerated erosion. Soil erosion can be
even worse where fire is part of the management practice
(Cavelier et al., 1998, 1999). For example, in the Sierra
Nevada de Santa Marta in Colombia, a long history of land
use, including fire, has resulted in the loss of >50 cm of soil,
which greatly limits agriculture and forest recovery (Aide
and Cavelier, 1994).
In recent years, rural to urban migration has been another
major factor influencing the abandonment of montane lands
(Aide and Grau, 2004). In some regions, armed conflicts have
stimulated rural–urban migration. For example, in Colombia,
approximately 2 million people have been displaced by civil
unrest, and the majority of these people have migrated out of
rural mountain regions (Chernick, 2000). Economic and social
changes have also stimulated rural–urban migration in areas
where small-scale agriculture on steep marginal lands cannot
compete with large-scale industrial agriculture and imported
products (Preston, 1996). In addition, many rural inhabitants,
especially young people, are attracted to cities because of greater
opportunities for education, social services, jobs, and an urban
lifestyle (Harden, 1996; cf. Grau et al., this volume).
Tropical Montane Cloud Forests: Science for Conservation and Management, eds. L. A. Bruijnzeel, F. N. Scatena, and L. S. Hamilton. Published by
Cambridge University Press. #Cambridge University Press 2010.
101
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Where rural–urban migration continues, the associated reduc-
tion in human pressure on the ecosystems could permit forest
recovery (Aide et al., 2000; Guariguata and Ostertag, 2001; Aide
and Grau, 2004). However, even when land is abandoned or
grazing intensity decreases, some forests may not recover
because the degraded system is “resilient.” Resilience of non-
forest vegetation is due to positive feedbacks that reinforce the
maintenance of the degraded state of the system (Hobbs and
Norton, 1996; Hartig and Beck, 2003; Suding et al., 2004). For
example, in many tropical areas, the grass Melinis minutiflora
invades after fires and its resinous leaves are highly flammable,
which promotes more fires and greatly inhibits forest regener-
ation (D’Antonio and Vitousek, 1992). In the South Ecuadorian
Andes, the bracken fern Pteridium arachnoideum dominates
degraded pastures and presents a major barrier to tree establish-
ment (Hartig and Beck, 2003; Gu
¨nter et al., 2009).
This review addresses the following questions:
Once pastures or agricultural lands are abandoned, what
factors control rates of tropical montane (cloud) forest
recovery?
When sites do not recover (i.e. they are in the resilient
degraded state), what restoration strategies can be used to
overcome barriers and accelerate forest recovery?
To answer these questions, a literature review was conducted
covering articles in Biological Abstracts (1997–2003), Web of
Science (1990–2004), Science Citation Index (1997–2002),
as well as The Scholarly Journal Archive (JSTOR), Springer,
and Organization for Tropical Studies databases. Initially, the
following key words were used: tropical, montane cloud forest,
restoration, reforestation, rehabilitation, and recovery. Given that
only a few articles were found with the explicit objective of active
restoration in TMCF (as defined in Hamilton et al., 1995), the
search was expanded using the following keywords: montane
forest, secondary succession, plantations, and afforestation. In
addition, references within articles identified in these databases
were used. Articles were evaluated to determine the major factors
controlling the rate of forest recovery and what management
activities were effective in accelerating forest recovery.
FACTORS CONTROLLING RATES OF
FOREST RECOVERY AND MANAGEMENT
SOLUTIONS
Studies of tropical montane (TMF) and tropical montane cloud
forest (TMCF) have identified soils, fire, seed dispersal, colon-
ization of pioneer species, and microhabitat conditions as the
major factors controlling forest recovery. Plantations were also
considered as a factor affecting forest recovery, but in other
studies plantations have been considered mainly as a management
tool (e.g. the first step in a reforestation project). The various
observations are summarized below by first discussing how each
factor affects rates of forest recovery and then presenting restor-
ation and management responses that could ameliorate the nega-
tive effects of each factor and accelerate forest recovery.
Soils
PROBLEMS
Although saturated anoxic soils (Silver et al., 1999), low nutrient
availability (Tanner et al., 1998; Stewart, 2000; Gu
¨nter et al.,
2009; cf. Benner et al., this volume; Roman et al., this volume),
compaction (Pedraza and Williams-Linera, 2003; Zimmermann
and Elsenbeer, 2008), or extensive erosion (Aide and Cavelier,
1994; Slocum et al., 2000) all can limit forest recovery, in the
present survey, few studies have identified soils as a major
limiting factor. In the context of TMCF, the prime soil charac-
teristic reported as limiting forest regeneration was soil compac-
tion associated with cattle grazing (Pedraza and Williams-Linera,
2003; cf. Zimmermann and Elsenbeer, 2008) or road construc-
tion (Olander et al., 1998). Compacted soils have been associated
with low sapling survivorship and growth (Pedraza and
Williams-Linera, 2003). In contrast, disturbed areas along tem-
porary roads or pipelines with low soil compaction had rapid
woody vegetation recovery (Malizia et al., 2004). An additional
problem of compacted soils is that they are often colonized by
grasses, which can increase susceptibility to fire, inhibit the
colonization of woody species, and make the system resistant
to forest recovery (Rosales et al., 1997; Olander et al., 1998; cf.
Hartig and Beck, 2003; Zimmermann and Elsenbeer, 2008).
MANAGEMENT SOLUTION
If soils are a limiting factor at a regional scale, this can be a
major challenge for restoration projects (Bradshaw, 1997).
Although areas can be fertilized, and soil material added, or
plowed in the case of compaction, these are expensive activities
if large areas are to be recovered (Bradshaw, 1997). At any scale,
an immediate management goal should be to reduce soil loss
(Calle, 2003). For example, a common practice is to plant or
leave strips of vegetation along the contours of a slope to reduce
soil erosion (Calle, 2003; Pedraza and Williams-Linera, 2003).
In relatively small areas, such as shallow landslips, fertilization
with nitrogen and potassium has facilitated the establishment of
tree species (Dalling and Tanner, 1995). At present, the most
widely used technique for improving degraded soils has been to
establish plantations because they increase soil nutrients and soil
organic matter, and protect the soil from erosion (Lugo, 1992;
Vanacker et al., 2007). For gullied areas, a number of mechan-
ical, but costly, techniques have been developed, but these
are beyond the scope of this review (see e.g. Blaisdell, 1981;
Hudson, 1995).
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Fire
PROBLEMS
In general, high rainfall and moisture in TMCF limit the number
and extent of fires. Fires can occur in TMCF during extended dry
periods (Asbjornsen et al., 2005; Hemp 2005; Asbjornsen and
Garnica Sanchez, this volume; Hemp, this volume #60), but they
appear to be more important in other high-elevation habitats (e.g.
paramo and dry shrub/pine forest) where there is less precipita-
tion (Hemp and Beck, 2001; Horn et al., 2001). Although this
may change in the future because of climatic drying (Hemp,
2005; Asbjornsen and Wickel, 2009; Foster, this volume), in
most TMCF settings, fire is associated with agricultural and
grazing activities (May, 1997; Scowcroft and Jeffery, 1999;
Kessler, 2000). As in other montane systems, if fires occur
frequently they can greatly reduce forest regeneration (Aide
and Cavelier, 1994; Cavelier et al., 1998; Ramirez-Marcial
et al., 2001; Duncan and Chapman, 2003a; Grau et al., this
volume) (Figure 8.1). In addition to the direct effect of killing
most woody species, fires can indirectly affect forest recovery by
increasing soil erosion and selecting for grasses and ferns. These
effects will negatively affect forest regeneration and could lead
to a resilient degraded state (Hartig and Beck, 2003; Suding
et al., 2004; Griscom et al., 2008).
MANAGEMENT SOLUTION
Fortunately, fires are still rare in undisturbed TMCF (cf. Wa
˚rd
et al., this volume), and given that agricultural activities are the
major source of fires, if a region is being abandoned, then the
frequency of fires should diminish. Where fires are a threat, they
can be controlled with firebreaks (Aide and Cavelier, 1995), fire
lookout and response teams, and education (Asbjornsen and
Garnica Sanchez, this volume). In some montane systems fire
may be used as a restoration tool. For example, in Thailand,
surface fires enhance seedling recruitment of deciduous diptero-
carp forest and pines by reducing grasses and woody lianas
(Werner and Santisuk, 1993). Similarly, in NW Argentina, fires
of intermediate frequencies can favor the expansion of Alnus
acuminata into montane grasslands by reducing competition
between tree seedlings and grasses (Grau and Veblen, 2000).
Seed dispersal
PROBLEMS
Many studies have investigated the importance of seed dispersal
as a limiting factor for natural regeneration (Aide and Cavelier,
1994; Holl et al., 2000; Oosterhoorn and Kappelle, 2000;
Zimmerman et al., 2000; Cubin
˜a
´and Aide, 2001; Gu
¨nter et al.,
2006). In virtually all studies, there is a rapid decrease in the
diversity of species and total number of seeds with distance
from the forest edge; seeds of most species are not detected
more than 20 m from the forest edge. Small wind-dispersed species
(often shrubs belonging to the Bignoniaceae and Asteraceae fam-
ilies) are an exception as their seeds can disperse further into the
abandoned lands. Contrary to this strong pattern of dispersal limi-
tation, recruitment patterns rarely show a clear procession from the
forest edge toward the center of the abandoned lands, suggesting
that seeding from adjacent forest is not always the limiting factor.
Aide et al. (2000) suggested that some species colonize pastures
when these are still actively used. If plants are capable of sprouting
after being cut or grazed, they may survive for extended periods,
and once the land is abandoned they can immediately dominate the
area. Regardless of the dispersal patterns or time of colonization,
the species that initially establish in these abandoned agricultural
lands make up a very limited sub-set of montane forest species,
suggesting that seed dispersal is a limiting factor for species
composition (Oosterhoorn and Kappelle, 2000).
MANAGEMENT SOLUTION
Different techniques have been used to increase seed dispersal
into abandoned lands. In many systems, remnant trees or shrubs
have played an important role in attracting dispersers and
accelerating regeneration (Slocum and Horwitz, 2000). Artifi-
cial perches also can attract birds and increase the seed rain
below the perch (Aide and Cavelier, 1994), but they rarely
increase seedling density (Holl et al., 2000; Shiels and
Figure 8.1. Remnant patches of montane forest in a matrix of degraded
grasslands maintained by frequent fires in the Sierra Nevada de Santa
Marta, Colombia.
WHAT IS THE STATE OF TMCF RESTORATION? 103
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Walker, 2003). A problem with artificial perches is that they do
not produce enough shade to reduce the grass cover, which can
limit the ability of seedlings to establish. Windbreaks (Harvey,
2000), fruit trees (Sarmiento, 1997a), and pioneer shrubs
(Posada et al., 2000) also attract bird dispersers and produce
more shade than artificial perches, thus making them good
candidates for increasing seed dispersal and seedling establish-
ment. For example, in NW Argentinean cloud forests, exotic
spiny shrubs (e.g. Crataegus oxyacantha) have facilitated the
establishment of native species by attracting bird dispersers
and protecting the new recruits from browsing by domestic
animals (Malizia and Greslebin, 2000). However, if the goal
of the project is to attain a species composition similar to intact
forest, seed collection, nursery activities, and planting will all
have to be major activities of a restoration project (Lamb and
Gilmour, 2004).
Colonization of pioneer species
PROBLEMS
Although seed dispersal can be a limiting factor in many sites,
grasses in abandoned pastures or early colonizers of abandoned
agricultural lands (e.g. ferns) can inhibit forest recovery. In some
cases, these species can even arrest succession. For example,
introduced African grasses (e.g. Melinis minutiflora, Pennisetum
clandestinum, Setaria sphacelata) often dominate montane pas-
tures, and once the pastures are abandoned these grasses grow
vigorously, impeding the establishment of woody species (e.g.
Colombia – Posada et al., 2000; Ecuador – Sarmiento, 1997a,b;
Zahawi and Augspurger, 1999; cf. Gu
¨nter et al., 2009; Costa
Rica – Holl, 1999; Hawai’i – D’Antonio and Vitousek, 1992)
(Figure 8.2). When the previous land use has been logging or
agricultural, ferns are often the first species to colonize the site.
In the Dominican Republic, native ferns that colonize landslides
have greatly expanded their local distribution by colonizing
abandoned agricultural lands (Garcia et al., 1994). These ferns
stabilize soils, reduce erosion, and increase soil organic matter,
but they can also inhibit the establishment of tree species
(Slocum et al., 2000). For example, Dicranopteris pectinata
has colonized and dominated abandoned agricultural fields for
>20 years because the fern produces a thick root mat on the soil
surface, which acts as a physical barrier and inhibits the estab-
lishment of other species (Figure 8.3; Slocum et al., 2000).
A similar case has been documented for the bracken fern Pteri-
dium arachnoideum in the South Ecuadorian Andes by Hartig
and Beck (2003). At other sites, exotic trees and shrubs (e.g.
Ligustrum spp. and Citrus spp.) have colonized abandoned agri-
cultural lands and areas with a history of intense grazing (e.g.
Grau and Aragon, 2000), and these species can also dominate a
site for many generations (Tecco and Rouges, 2000; Schulenberg
and Awbrey, 1997; Lichstein et al., 2004).
Figure 8.2. A three-year-old abandoned pasture dominated by
Pennisetum clandestinum in the Cordillera Central, Colombia.
Figure 8.3. Dicranopteris pectinata fernlands in the Cordillera Central,
Dominican Republic.
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MANAGEMENT SOLUTION
A diversity of approaches will be needed to eliminate or reduce
the densities of these grasses, ferns, and exotic trees and shrubs.
This constitutes a major challenge given that they often cover
extensive areas. In Colombia, a fast-growing tree, Montanoa
quadrangularis (Asteraceae), has been used to expand forest
cover into pastures by shading grasses and enhancing seedling
recruitment of forest species (Calle, 2003). Posada et al. (2000)
demonstrated that low-density grazing reduced the cover of
Melinis minutiflora and Pennisetum clandestinum in Colombia,
and permitted the colonization of wind-dispersed shrub species
that helped to shade out the grasses and create a microhabitat
appropriate for montane forest species (Figure 8.4). In the case of
the fern Dicranopteris pectinata in the Dominican Republic, fire
(May, 2000) and manual removal (including the root mat;
Slocum et al., 2004) were techniques that allowed many native
species to colonize. Surprisingly, the fern was slow in recover-
ing, and the shade produced by shrubs was already sufficient
to limit any future dominance, demonstrating that the fern-
dominated resilient state was easily shifted to a recovering state.
Eliminating invasive trees and shrubs will be much more difficult
and costly, however, and must be accompanied by extensive
planting of native species and long-term maintenance to avoid
the recolonization of the exotic species.
Microhabitat conditions
PROBLEMS
At a more local spatial scale, microhabitat conditions can influ-
ence colonization, growth, and survivorship and, in turn, the
composition of a recovering forest. Even if seeds disperse into
a site, to colonize, seeds must avoid predators. Few studies have
identified seed predation as a major limiting factor in montane
forest, but ant (Pedraza and Williams-Linera, 2003), cattle
(Alvarez-Aquino et al., 2004; Griscom et al., 2008), and rabbit
(Holl and Lulow, 1997) herbivory can affect plant growth and
survivorship in early stages of forest recovery. Other studies have
shown that poor mycorrhizae colonization of roots has limited
plant growth (Berish and Ewel, 1988; Rosales et al., 1997; Holl
et al., 2000). Mycorrhizal limitation is most common in soils that
are highly degraded (e.g. by mining; Miller and Jastrow, 1992).
Other factors that can reduce growth and survivorship include:
freezing at high elevation (Scowcroft et al., 2000; cf. Hemp,
this volume #12), low nutrient availability (Bruijnzeel and
Veneklaas, 1998; Gu
¨nter et al., 2009) and photo-inhibition (Loik
and Holl, 1999).
MANAGEMENT SOLUTION
The easiest way to avoid seed predators is to plant seedlings.
Seedlings will still be vulnerable to herbivores, but few studies
have identified herbivores as a limiting factor in abandoned
agricultural lands (see Holl and Lulow, 1997). Furthermore, if
seedlings are being produced for a restoration project they can be
inoculated with mycorrhizae to promote better growth and sur-
vivorship. Other studies have shown that nurse plants can modify
the microhabitat conditions and promote the establishment of
more sensitive species (Rhoades et al., 1998; Posada et al.,
2000; Scowcroft et al., 2000; Calle, 2003; Weber et al., 2008).
For example, in Colombia seedling density and richness were
higher below remnant trees in pastures in contrast to open areas
in pastures (Calle, 2003), whereas Posada et al. (2000) found
that low-density grazing promoted the establishment of wind-
dispersed shrubs, which shade out the grass. In Ecuador, small
nitrogen-fixing trees in pastures ameliorated microclimate con-
ditions and increased soil nitrogen levels, which, in turn, favored
the establishment and growth of montane species (Rhoades et al.
1998; cf. Weber et al., 2008; Gu
¨nter et al., 2009). On Hawai’i,
efforts to restore Metrosideros polymorpha and other native
species in exotic grasslands have been unsuccessful because
of high seedling mortality due to freezing. At this site, Acacia
koa survived better than other species, and once established,
it helped to reduce frost damage to seedlings planted below it
(Scowcroft et al., 2000).
Figure 8.4. A three-year-old abandoned pasture adjacent to the
site depicted in Figure 8.2 which had low-intensity grazing and was
dominated by Asteraceae shrubs.
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PLANTATIONS
In the literature, plantations have been treated as both a
factor limiting forest recovery and as a technique for acceler-
ating forest recovery. Plantations can help ameliorate all the
barriers discussed above. Plantations can improve soil condi-
tions by protecting soils from erosion (Wiersum, 1984;
Vanacker et al., 2007), and improving their nutrient content
by fixing nitrogen and adding organic matter (Lugo, 1992). In
montane ecosystems, a landscape that has increased forest
cover will be less vulnerable to fires than areas dominated
by grasslands and agricultural areas. In addition, plantations
can attract dispersers, which will increase plant diversity in the
understory community (Parrotta, 1992). Furthermore, planta-
tions will eliminate most grasses and ferns by producing
shade, and thus improve the microclimate conditions for many
forest species.
Although plantations can improve conditions for forest
regeneration, the long-term impact on forest recovery depends
on the species used, the management techniques, and the goals
of the project. In Uganda, forest regeneration in pine planta-
tions was faster than in grasslands, because recurrent fires set
back forest recovery in the grasslands (Zanne and Chapman,
2001). However, when these grasslands were protected from
fire, forest regeneration was similar to that within the planta-
tion (Duncan and Chapman, 2003a,b). In Colombia, although
plantations (e.g. Pinus,Cupressus,Eucalyptus,andAlnus)
helped improve soil characteristics and native tree species
established in the understorey, species diversity in secondary
forest of similar age was much higher compared with planta-
tions (Cavelier and Tobler, 1998; Cavelier and Santos, 1999;
Murcia, 1997; cf. Ho
¨lscher et al., this volume). These studies
suggest that, in the absence of fire, natural regeneration is a
better strategy for recovering biodiversity than establishing
plantations.
If the motive of a project is to restore natural vegetation,
are single-species plantations a useful restoration tool? The
answer is positive if soils are highly degraded and plantation
species can protect and improve soil conditions. Plantations
canalsobeincorporatedinalong-termrestorationplan,where
there is a need for short-term economic benefits (timber,
carbon credits), but after harvesting, native species in the
understory should be allowed to grow. Plantations may also
be useful for shading out grasses or ferns and attracting dis-
persers of native species, but if the final goal is a diverse
montane forest, plantation trees should not be established at
high densities. In many other cases, plantations do not seem
necessary or appropriate. Secondary forest of similar age will
support a (much) greater diversity of native species in
comparison with plantations, and at a much lower cost (cf.
Ho
¨lscher et al., this volume).
SYNTHESIS
Will tropical montane clould forest recover without intervention?
In many cases it appears that TMCF will recover, and studies have
estimated the recovery time to be between 65 and 200 years
(Weaver, 2000; Kappelle, 2001; Luna et al., 2001; Silver et al.,
2001). The recovery time will clearly depend on the kind, duration,
and intensity of previous land use, but it will also depend on what is
the measure of recovery. Although plant richness may recover in 65
years, it may take centuries to recover the species composition and
fauna similar to intact forest. As shown by Ko
¨hler et al.(this
volume), this is particularly true for the epiphyte layer in TMCF.
Ruiz-Jae
´n and Aide (2005) have shown that the majority of restor-
ation projects only measure a few structural variables and plant
richness, and few studies incorporate measures of the fauna or
ecosystems processes. As the field of restoration ecology matures,
it is necessary to incorporate more variables (e.g. fauna – Murcia
et al., 2001; Medina et al., 2002; Dunn, 2004; ecosystem process –
Ruiz-Jae
´n and Aide, 2005) as measures of recovery.
In other cases, aggressive pioneer species, fire, degraded soils,
or lack of seeds will produce feedbacks that can maintain the
system in a degraded state with no or very slow recovery. How
can the recovery process be accelerated and the shift from a
resilient degraded state to a state of recovery be facilitated?
The results of this review suggest that, in the majority of cases,
competition with grass and ferns is the major factor limiting the
initial stages of forest recovery in tropical montane systems, and
seed dispersal of mature forest species limits later stages of
recovery. To overcome the barrier of aggressive herbaceous
species, low-density grazing or manual removal is often suffi-
cient to initiate secondary succession. If native shrubs do not
establish rapidly, fast-growing trees, possibly plantation species,
can be used to shade out herbaceous species and attract dis-
persers. These strategies should produce a species-diverse sec-
ondary forest. If the goal of a project is to restore the species
composition of a mature montane forest, it is recommended to
establish a nursery of mature montane forest species and to plant
these into the site once the herbaceous cover has been reduced.
HOW CAN RESTORATION ECOLOGY
CONTRIBUTE TO THE FUTURE
OF TMCF?
In many regions, old threats to TMCF, such as clearing for
agriculture or grazing, or intensive logging continue (e.g.
Mosandl et al., 2008; cf. Mulligan, this volume). Although
economic development in the lowlands may reduce land use
intensity in many montane areas, development is also accompan-
ied by expanding infrastructure, which can impact TMCF forest
(e.g. in the form of wind turbines, communication towers,
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pipelines and roads; Young, 1994; Malizia et al., 2004).
A potentially much greater concern is the impact of global
climate change on TMCF (Pounds et al., 1999; Foster, 2001;
Lawton et al., 2001; Pounds and Puschendorf, 2004; Foster, this
volume; Nair et al., this volume). These impacts should be
mitigated, but our collective restoration experience in TMCF is
extremely limited. A recent review of tropical restoration showed
that the majority of projects are located in the lowland forest
zone, with only 13% conducted in TMCF (Meli, 2003).
On the other hand, if the trend of rural–urban migration con-
tinues in Latin America, large areas of marginal pastures and
agricultural lands will be abandoned, and much of this area will
be in TMCF. Puerto Rico is a dramatic example of rural–urban
migration, agriculture abandonment, and forest recovery. Forest
cover increased from <10% in the late 1940s to >40% at
present, following socio-economic changes associated with eco-
nomic globalization (Rudel et al., 2000; Grau et al., 2003).
Although Puerto Rico is an extreme example, a similar process
is occurring in other areas of Latin America (Zweifler et al.,
1994; Preston et al., 1997; Southworth and Tucker, 2001; Rudel,
2002). How should these abandoned lands in the TMCF zone be
managed? Most of these areas will be left to natural regeneration,
if only due to financial limitations. In many cases, this will be an
appropriate “management” strategy, particularly given the low
cost. In other areas where aggressive herbaceous species have
arrested or greatly slowed succession (i.e. resilient degraded
sites), or areas of special conservation importance, active restor-
ation could help to accelerate the recovery process.
As restoration projects expand forest cover, positive feedbacks
should reinforce regeneration processes. For example, as forest
cover and connectivity increase, fire frequency will decrease, and
seed dispersal and animal populations should increase. Understand-
ing the dynamics of these feedbacks, and developing dynamic
models should be part of future restoration projects and be incorpor-
ated into the international research agenda, such as the Forest
Landscape Restoration project (Maginnis and Jackson, 2005;
www.unep-wcmc.org/forest/restoration/globalpartnership/index.
htm). To meet these present and future restoration/management
challenges, it is important to get beyond talking about restoration
or applying results from other lifezones, and begin a coordinated
effort to understand how TMCF can be restored. Payment for
environmental services is a possible approach to stimulate future
restoration efforts (e.g. Costa Rica – Rodriguez Zun
˜iga, 2003;
cf. Tognetti et al., this volume; Calvo-Alvarado et al., this volume).
ACKNOWLEDGEMENTS
During the preparation of the manuscript the authors received
support from the NASA-IRA program and the Dean of Graduate
Studies and Investigation (DEGI) of the University of Puerto
Rico. The International Institute of Tropical Forestry of the
U.S. Forest Service assisted with the literature search and travel
expenses. The comments of Becky Ostertag, Zoraida Calle,
Sampurno Bruijnzeel, and Larry Hamilton improved the
manuscript.
REFERENCES
Aide, T. M., and J. Cavelier (1994). Barriers to lowland tropical forest
restoration in the Sierra Nevada de Santa Marta, Colombia. Restoration
Ecology 2: 219–229.
Aide, T. M., and H. R. Grau (2004). Globalization, migration, and Latin
American ecosystems. Science 305: 1915–1916.
Aide, T.M., J. K. Zimmerman, J. Pascarella, J. Marcano-Vega, and L. Rivera
(2000). Forest regeneration in a chronosequence of tropical abandoned
pastures: implications for restoration ecology. Restoration Ecology 8:
328–338.
Alvarez-Aquino, C., G. Williams-Linera, and A. C. Newton (2004). Experi-
mental native tree seedling establishment for the restoration of a Mexican
cloud forest. Restoration Ecology 12: 412–418.
Asbjornsen, H., C. Gallardo-Herna
´ndez, N. Vela
´zquez-Rosas, and R. Garcı
´a-
Soriano (2005). Deep ground fires cause massive above- and below-ground
biomass losses in tropical montane cloud forests in Oaxaca, Mexico.
Journal of Tropical Ecology 21: 427–434.
Asbjornsen, H., and A. J. Wickel (2009). Changing fire regimes in tropical
montane cloud forests: a global synthesis. In Fire and Tropical Ecosys-
tems, ed. M. A. Cochrane, pp. 607–622. New York: Springer-Verlag.
Bedoya, E. (1997). Migracio
´n, falta de empleo y el impacto de la coca en los
sistemas agrı
´colas tropicales y de montan
˜a: un estudio de caso en el Peru
´.
In Desarrollo sostenible de ecosistemas de montan
˜a: Manejo de areas
fra
´giles en los Andes, eds. M. Liberman and C. Baied, pp. 83–91. La Paz,
Bolivia: The United Nations University.
Berish, C. W., and J. J. Ewel (1988). Root development in simple and com-
plex tropical successional ecosystems. Plant and Soil 106: 73–84.
Blaisdell, F. W. (1981). Engineering structures for erosion control. In Trop-
ical Agricultural Hydrology, eds. R. Lal and E. W. Russell, pp. 325–355.
New York: John Wiley.
Bradshaw, A. D. (1997). The importance of soil ecology in the restoration
science. In Restoration Ecology and Sustainable Development, eds.
K. M. Urbanska, N. R. Webb, and P. J. Edwards, pp. 33–64. Cambridge,
UK: Cambridge University Press.
Bruijnzeel, L. A., and E. J. Veneklaas (1998). Climatic conditions and tropical
montane forest productivity: the fog has not lifted yet. Ecology 79: 3–9.
Calle, Z. (2003). Restauraci
on de suelos y vegetaci
on nativa: ideas para una
ganader´a Andina sostenible. Cali, Colombia: Centro para la Investigacio
´n
en Sistemas Sostenibles de Produccio
´n Agropecuaria (CIPAV).
Cavelier, J. (1995). Reforestation with the native tree Alnus acuminata: effects
on phytodiversity and species richness in the upper montane rain forest area
of Colombia. In Tropical Montane Cloud Forests, eds. L. S. Hamilton, J. O.
Juvik, and F. N Scatena, pp. 125–137. New York: Springer-Verlag.
Cavelier, J. and A. Etter (1995). Deforestation of montane forest in Colombia
as result of illegal plantations of opium (Papaver somniferum).
In Biodiversity and Conservation of Neotropical Montane Forests, eds.
P. Churchill, H. Baslev, E. Forero, and J. L Luteyn, pp. 541–549. New
York: New York Botanical Garden.
Cavelier, J. and C. Santos (1999). Effect of abandoned exotic and native
species plantations on the natural regeneration of a montane forest in
Colombia. Revista de Biologia Tropical 47: 775–784.
Cavelier, J. and A. Tobler (1998). The effect of abandoned plantations of
Pinus patula and Cupressus lusitanica on soils and regeneration of a tropical
montane rain forest in Colombia. Biodiversity and Conservation 7: 335–347.
Cavelier, J., T. M. Aide, C. Santos, A. M. Eusse, and J. M. Dupuy (1998).
The savannization of moist forest in the Sierra Nevada de Santa Marta,
Colombia. Journal of Biogeography 25: 901–912.
Cavelier, J., T. M. Aide, J. M. Dupuy, A. M. Eusse, and C. Santos (1999).
Long-term effects of deforestation on soil properties and vegetation in a
tropical lowland forest in Colombia. Ecotropicos 12: 57–68.
WHAT IS THE STATE OF TMCF RESTORATION? 107
Comp. by: PG2693 Stage : Proof ChapterID: 8 Date:13/8/10 Time:17:07:39 Filepath:H:/3B2/Bruihnzeel-
9780521760355/Applications/3B2/Proof/8.3d
Chernick, M. (2000). Elusive peace: struggling against the logic of violence.
North American Congress on Latin America Report on the Americas 34:
34–37.
Cubin
˜a
´, A., and T. M. Aide (2001). The effect of distance from forest edge on
seed rain and soil seed bank in a tropical pasture. Biotropica 33: 260–267.
Dalling, J. W. and E. V. J. Tanner (1995). An experimental study of regener-
ation on landslides in montane forests in Jamaica. Journal of Ecology 83:
55–64.
D’Antonio, C. M. and P. M. Vitousek (1992). Biological invasions by exotic
grasses, the grass/fire cycle, and global change. Annual Review of Ecology
and Systematics 23: 63–87.
Duncan, R. S. and C. A. Chapman (2003a). Tree–shrub interactions during
early forest succession in Uganda. Restoration Ecology 11: 198–207.
Duncan, R. S. and C. A. Chapman (2003b). Consequences of plantation
harvest during tropical forest restoration in Uganda. Forest Ecology and
Management 173: 235–250.
Dunn, R. R. (2004). Managing the tropical landscape: a comparison of the
effects of logging and forest conversion to agriculture on ants, birds, and
lepidoptera. Forest Ecology and Management 191: 215–224.
Foster, P. (2001). The potential negative impacts of global climate change on
tropical montane cloud forests. Earth-Science Reviews 55: 73–106.
Garcia, R., M. Mejı
´a, and T. Zanoni (1994). Composicio
´n floristica y princi-
pales asociaciones vegetales en la Reserva Cientifica Ebano Verde,
Cordillera Central, Repu
´blica Dominicana. Moscosoa 8: 86–130.
Grau, H. R. and M. R. Aragon (2000). Arboles invasores de la sierra de San
Javier, Tucuman, Argentina. In Ecologia de arboles ex
oticos en las Yungas
Argentinas, eds. H. R. Grau and M. R. Aragon, pp. 5–20. Tucuman, Argen-
tina: Laboratorio de Investigaciones Ecologicas de las Yungas.
Grau, H. R. and T. T. Veblen (2000). Rainfall, fire, and vegetation dynamics
in subtropical montane ecosystems in northwestern Argentina. Journal of
Biogeography 27: 1107–1121.
Grau, H. R., T. M. Aide, J. K. Zimmerman, et al. (2003). The ecological
consequences of socioeconomic and land-use changes in post-agriculture
Puerto Rico. BioScience 53: 1159–1168.
Griscom, H. P., B.W. Griscom, and M. S. Ashton (2008). Forest regeneration
from pasture in the dry tropics of Panama: effects of cattle, exotic grasses,
and forest riparia. Restoration Ecology 17: 117–126.
Guariguata, M. R. and R. Ostertag (2001). Neotropical secondary forest
succession: changes in structural and functional characteristics. Forest
Ecology and Management 148: 185–206.
Gu
¨nter, S., M. Weber, R. Erreis, and N. Aguirre (2006). Influence of distance
to forest edges on natural regeneration of abandoned pastures: a case study
in the tropical montane rain forest of Southern Ecuador. European Journal
of Forest Research, doi: 10.1007/s10342–006–0156–0.
Gu
¨nter, S., P. Gonzalez, G. Alvarez, et al. (2009). Determinants of successful
regeneration of abandoned pastures in the Andes: soil conditions and
vegetation cover. Forest Ecology and Management 258: 81–91.
Hamilton, L. S., J. O. Juvik, and F. N. Scatena. (1995). The Puerto Rico
tropical cloud forest symposium: Introduction and workshop synthesis. In
Tropical Montane Cloud Forests, eds. L. S. Hamilton, J. O. Juvik, and
F. N. Scatena, pp. 1–23. New York: Springer-Verlag.
Harden, C. (1996). Interrelationships between land abandonment and land
degradation: a case from the Ecuadorian Andes. Mountain Research and
Development 16: 274–280.
Hartig, K. and E. Beck (2003). The Bracken fern (Pteridium arachnoideum
Kaulf.) dilemma in the Andes of Southern Ecuador. Ecotropica 9: 3–13.
Harvey, C. (2000). Windbreaks enhance seed dispersal into agricultural land-
scapes in Monteverde, Costa Rica. Ecological Applications 10: 155–173.
Hemp, A. (2005). Climate change-driven forest fires marginalize the impact
of ice cap wasting on Kilimanjaro. Global Change Biology 11: 1013–1023.
Hemp, A. and E. Beck (2001). Erica excelsa as a fire-tolerating component of
Mt. Kilimanjaro’s forests. Phytocoenologia 31: 449–475.
Hobbs, R. J. and D. A. Norton (1996). Towards a conceptual framework for
restoration ecology. Restoration Ecology 4: 93–110.
Holl, K. D. (1999). Factors limiting tropical moist forest regeneration in
agricultural land: soil, microclimate, vegetation, and seed rain. Biotropica
31: 229–242.
Holl, K. D., and M. E. Lulow (1997). Effects of species, habitat, and distance
from edge on post-dispersal seed predation in a tropical rainforest. Bio-
tropica 29: 459–468.
Holl, K. D., M. E. Loik, E. H. V. Lin, and I. A. Samuels (2000). Tropical
montane forest restoration in Costa Rica: obstacles and opportunities.
Restoration Ecology 8: 339–349.
Horn, S. P., L. M. Kennedy, and K. H. Orvis. (2001). Vegetation recovery
following a high elevation fire in the Dominican Republic. Biotropica 33:
701–708.
Hudson, N. W. (1995). Soil Conservation. London: Batsford.
Kappelle, M. (2001). Costa Rica. In Bosques nublados del neotropico, eds.
M. Kappelle and A. D. Brown, pp. 301–370. Santo Domingo de Heredia,
Costa Rica: Instituto Nacional de Biodiversidad.
Kessler, M. (2000). Observations on a human-induced fire event at humid
timberline in the Bolivian Andes. Ecotropica 6: 89–93.
Lamb, D. and D. Gilmour (2004). Rehabilitation and Restoration of
Degraded Forests. Gland, Switzerland: World Conservation Union, IUCN.
Lawton, R. O., U. S. Nair, R. A. Pielke Sr., and R. M. Welch (2001). Climatic
impact of tropical lowland deforestation on nearby montane cloud forests.
Science 294: 584–587.
Lichstein, J., H. R. Grau, and M. R. Aragon (2004). Recruitment limitation in
secondary forests dominated by an exotic tree. Journal of Vegetation
Science 15: 721–728.
Loik, M. E., and K. D. Holl (1999). Photosynthetic responses to light for
rainforest seedlings planted in abandoned pasture, Costa Rica. Restoration
Ecology 7: 382–391.
Lugo, A. E. (1992). Comparison of tropical tree plantations with secondary
forest of similar age. Ecological Monographs 62: 1–41.
Luna, I., A. Velazquez, and E. Velazquez (2001). Mexico. In Bosques nubla-
dos del neotropico, eds. M. Kappelle and A. D. Brown, pp. 183–229. Santo
Domingo de Heredia, Costa Rica: Instituto Nacional de Biodiversidad.
Maginnis, S., and W. Jackson (2005). Balancing restoration and develop-
ment. ITTO Tropical Forest Update 15: 4–6.
Malizia, A., N. Chacoff, H. R. Grau, and A. D. Brown (2004). Vegetation
recovery on a gas-pipeline track along an altitudinal gradient in the Argen-
tinean Yungas forest. Ecologia Austral 14: 165–178.
Malizia, L. R. and A. Greslebin (2000). Reclutamiento de especies arboreas
bajo arbustos exoticos en la sierra de San Javier, Tucuman, Argentina. In
Ecologia de arboles exot´cos en las Yungas Argentinas, eds. H. R. Grau
and M. R. Aragon, pp. 47–58. Tucuman, Argentina: Laboratorio de Inves-
tigaciones Ecologicas de las Yungas.
May, T. (1997). Fases tempranas de la sucesio
´n en un bosque nublado de
Magnolia pallescens despue
´s de un incendio (Loma de Casabalito, Reserva
Cientı
´fica Ebano Verde, Cordillera Central, Repu
´blica Dominicana).
Moscosoa 9: 117–144.
May, T. (2000). Respuesta de la vegetacio
´n en un “calimetal”de Dicranop-
teris pectinata despue
´s de un fuego, en la parte oriental de la Coordillera
Central, Repu
´blica Dominicana. Moscosoa 11: 113–132.
Medina, C. A., F. Escobar, and G.H. Kattan (2002). Diversity and habitat use
of dung beetles in a restored Andean landscape. Biotropica 34: 181–187.
Meli, P. (2003). Restauracio
´n ecolo
´gica de bosques tropicales: viente an
˜os de
investigacio
´n acade
´mica. Interciencia 29: 581–589.
Miller, R. M. and J. D. Jastrow (1992). The application of VA mycorrhizae to
ecosystem restoration and reclamation. In Mycorrhizal Functioning: An
Integrative Plant Fungal Process, ed. M. Allen, pp. 438–467. New York:
Chapman and Hall.
Mosandl, R., S. Gu
¨nter, B. Stimm, and M. Weber (2008). Ecuador suffers the
highest deforestation rate in South America. In Gradients in a Tropical
Mountain Ecosystem of Ecuador, eds. E. Beck, J. Bendix, I. Kottke,
F. Makeschin, and R. Mosandl, pp. 37–40. Berlin: Springer-Verlag.
Murcia, C. (1997). Evaluation of Andean alder as a catalyst for the recovery
of tropical cloud forests in Colombia. Forest Ecology and Management 99:
163–170.
Murcia, C., G. Kattan, and A. Galindo (2001). Recovery of Bess beetles key
to long-term restoration of Andean forest (Colombia). Ecological Restor-
ation 19: 254–255.
Olander, L. P., F. N. Scatena, and W. L. Silver (1998). Impacts of disturbance
initiated by road construction in a subtropical cloud forest in the Luquillo
Experimental Forest, Puerto Rico. Forest Ecology and Management 109:
33–49.
Oosterhoorn, M., and M. Kappelle (2000). Vegetation structure and compos-
ition along an interior–edge–exterior gradient in a Costa Rican montane
cloud forest. Forest Ecology and Management 126: 291–307.
Parrotta, J. A. (1992). The role of plantation forests in rehabilitating degraded
tropical ecosystems. Agriculture, Ecosystems, and Environment 41:
115–133.
Pedraza, R. A. and G. Williams-Linera (2003). Evaluation of native tree
species for the rehabilitation of deforested areas in a Mexican cloud forest.
New Forests 26: 83–99.
108 T. M. AIDE ET AL.
Comp. by: PG2693 Stage : Proof ChapterID: 8 Date:13/8/10 Time:17:07:49 Filepath:H:/3B2/Bruihnzeel-
9780521760355/Applications/3B2/Proof/8.3d
Posada, J. M., T.M. Aide, and J. Cavelier (2000). Cattle and weedy shrubs as
restoration tools of tropical montane rainforest. Restoration Ecology 8:
370–379.
Pounds, J. A., and R. Puschendorf. (2004). Clouded futures. Nature 427: 107–109.
Pounds, J. A., M. P. L Fodgen, and J. H. Campbell (1999). Biological
response to climate change on a tropical mountain. Nature 398: 611–615.
Preston, D. (1996). People on the move: migrations past and present. In Latin
America Development: Geographical Perspectives, ed. D. Preston,
pp. 165–187. Harlow, UK: Addison Wesley.
Preston D., M. Macklin, and J. Warburton (1997). Fewer people, less erosion:
the 20th century in southern Bolivia. Geographical Journal 163: 198–205.
Ramirez-Marcial, N., M. Gonzalez-Espinosa, and G. Williams-Linera (2001).
Anthropogenic disturbance and tree diversity in montane rain forest in
Chiapas, Mexico. Forest Ecology and Management 154: 311–326.
Rhoades, C. C., G. E. Eckert, and D. C. Coleman (1998). Effect of pastures
trees on soil nitrogen and organic matter: implications for tropical montane
forest restoration. Restoration Ecology 6: 262–270.
Rodriguez Zun
˜iga, J. M. (2003). Paying for forest environmental services: the
Costa Rican experience. Unasylva 212: 31–33.
Rosales, J., G. Cuenca, N. Ramı
´rez, and Z. de Andrade (1997). Native
colonizing species and degraded land restoration in La Gran Sabana,
Venezuela. Restoration Ecology 5: 147–155.
Rudel, T. K. (2002). Paths of destruction and regeneration: globalization and
forests in the tropics. Rural Sociology 67: 622–636.
Rudel, T. K., M. Perez-Lugo, and H. Zichal (2000). When fields revert to
forest: development and spontaneous reforestation in post-war Puerto Rico.
Professional Geographer 52: 386–397.
Ruiz-Jae
´n, M. C., and T. M. Aide (2005). Restoration success: how is it being
measured? Restoration Ecology 13: 569–577.
Sarmiento, F. O. (1997a). Arrested succession in pastures hinders regener-
ation of Tropandean forests and shreds mountain landscapes. Environmen-
tal Conservation 24: 14–23.
Sarmiento, F. O. (1997b). Landscape regeneration by seeds and succcesional
pathways to restore fragile tropandean slopelands. Mountain Research and
Development 17: 239–252.
Scowcroft, P. G. and J. Jeffrey (1999). Potential significance of frost, topo-
graphic relief, and Acacia koa stands to restoration of mesic Hawaiian
forests on abandoned rangeland. Forest Ecology and Management 114:
447–458.
Scowcroft, P. G., F. C. Meinzer, G. Goldstein, P. J. Melcher, and J. Jeffrey
(2000). Moderating night radiative cooling reduces frost damage to Metro-
sideros polymorpha seedlings used for forest restoration in Hawaii. Res-
toration Ecology 8: 161–169.
Shields, A. B., and R. F. Walker (2003). Bird perches increase forest seeds on
Puerto Rican landslides. Restoration Ecology 11: 457–465.
Schulenberg, T. S., and K. Awbrey (1997). A Rapid Assessment of the Humid
Forest of South-Central Chuquisaca, Bolivia, RAP Working Papers No. 8.
Washington, DC: Conservation International.
Silver, W. L., A. E. Lugo, and M. Keller (1999). Soil oxygen availability and
biogeochemical cycling along elevation and topographic gradients in
Puerto Rico. Biogeochemistry 44: 301–328.
Silver, W. L., E. Marin-Spiotta, and A. E. Lugo. (2001). El Caribe. In Bosques
nublados del neotropico, eds. M. Kappelle and A. D. Brown, pp. 155–181.
Santo Domingo de Heredia, Costa Rica: Instituto Nacional de
Biodiversidad.
Slocum, M. G., and C. C. Horvitz (2000). Seed arrival under different genera
of trees in a neotropical pasture. Plant Ecology 149: 51–62.
Slocum, M., T. M. Aide, J. K. Zimmerman, and L. Navarro (2000). La
vegetacio
´n len
˜osa en helechales y bosques de ribera en la Reserva Cienti-
fica Ebano Verde, Republica Dominicana. Moscosoa 11: 38–56.
Slocum, M., T. M. Aide, J. K. Zimmerman, and L. Navarro (2004). Natural
regeneration of subtropical montane forest after clearing fern thickets in
the Dominican Republic. Journal of Tropical Ecology 20: 483–486.
Southworth, J., and C. Tucker (2001). The influence of accessibility, local
institutions, and socioeconomic factors on forest cover change in the
mountains of western Honduras. Mountain Research and Development
21: 276–283.
Stewart, C. G. (2000). A test of nutrient limitation in two tropical montane
forest using root ingrowth cores. Biotropica 32: 369–373.
Suding, K. N., L. G. Gross, and G. R. Houseman (2004). Alternative states
and positive feedbacks in restoration ecology. Trends in Ecology and
Evolution 19: 46–53.
Tanner, E. V. J., P. M. Vitousek, and E. Cuevas (1998). Experimental investi-
gation of nutrient limitation of forest growth on wet tropical mountains.
Ecology 79: 10–22.
Tecco, P. A., and M. Rouges (2000). El naranjo agrio (Citrus aurantium),
exotica invasora de bosques maduros. In Ecologia de arboles exotı
´
cos en
las Yungas Argentinas, eds. H. R. Grau and M. R. Aragon, pp. 37–45.
Tucuman, Argentina: Laboratorio de Investigaciones Ecologicas de las
Yungas.
Vanacker, V., F. von Blanckenburg, G. Govers, et al. (2007). Restoring dense
vegetation can slow mountain erosion to near benchmark levels. Geology
35: 303–306.
Weaver, P. L. (2000). Elfin woodland recovery 30 years after a plane wreck in
Puerto Rico’s Luquillo Mountains. Caribbean Journal of Science 36: 1–9.
Weber, M., S. Gu
¨nter, N. Aguirre, B. Stimm, and R. Mosandl (2008). Refor-
estation of abandoned pastures: silvicultural means to accelerate forest
recovery and biodiversity. In Gradients in a Tropical Mountain Ecosystem
of Ecuador, eds. E. Beck, J. Bendix, I. Kottke, F. Makeschin, and
R. Mosandl, pp. 447–458. Berlin: Springer-Verlag.
Werner, W. L., and T. Santisuk (1993). Conservation and restoration of
montane forest communities in Thailand. In Restoration of Tropical Forest
Ecosystems, eds. H. Lieth and M. Lohmann, pp. 193–202. Dordrecht, the
Netherlands: Kluwer.
Wiersum, K. F. (1984). Surface erosion under various tropical agroforestry
systems. In Proceedings of the Symposium on Effects of Forest Land Use
on Erosion and Slope Stability, eds. C. L. O’Loughlin and A. J. Pearce,
pp. 231–239. Vienna: IUFRO, and Honolulu, HI: East–West Center.
Young, K. (1994). Roads and the environmental degradation of tropical
montane forests. Conservation Biology 8: 972–976.
Zahawi, R. A., and C. K. Augspurger (1999). Early plant succession in aban-
doned pastures in Ecuador. Biotropica 31: 540–552.
Zanne, A. and C. A. Chapman (2001). Expediting reforestation in tropical
grasslands: distance and isolation from seed sources in plantations.
Ecological Applications 11: 1610–1621.
Zimmerman, J. K., J. Pascarella, and T. M. Aide (2000). Barriers to forest
regeneration in an abandoned pasture in Puerto Rico. Restoration Ecology
8: 350–360.
Zimmermann, B., and H. Elsenberger (2008). Spatial and temporal variability
of soil saturated hydraulic conductivity in gradients of disturbance. Journal
of Hydrology 361: 78–95.
Zweifler, M. O., M. A. Gold, and R. N. Thomas (1994). Land-use evolution
in hill regions of the Dominican Republic. Professional Geographer
46: 39–53.
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... Thus far, tropical ecosystem restoration has pre-dominantly been focussed around the lowlands, where restoration ecology has been thoroughly studied and synthesized in recent years [31][32][33][34] and various restoration methods have been tested and compared 34 . On the other hand, only a handful of reviews have addressed restoration questions related to TME, most of which are specific to a single tropical mountain ecosystem type or to a specific restoration context [31][32][33][35][36][37][38][39] . ...
... Over 33% of the studies were on a patch scale followed by local-scale studies and regional scale studies. There were only 5 pantropical/global assessments, which drew comparisons of restoration processes across distant mountain ranges or across continents 38,[85][86][87][88] . These findings are in line with trends in tropical forest restoration, where "neither the scale of scientific studies nor the restoration projects being implemented have matched the ambitious forest landscape restoration plans that are being proposed" 89 . ...
... Direct seeding of species and enrichment planting was most frequently studied in montane forest. A host of additional experimental methods were tested only a few times in the TME restoration studies, such as topsoil, seed bank and hay transfers in the mountain grasslands of the Campos Rupestres 48,108 , applied nucleation, assisted migration 136,137 and inoculation of cloud forest seedlings with arbuscular mycorrhizal 38,138 (Supplementary Fig. 6a). www.nature.com/scientificreports/ ...
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Many tropical mountain ecosystems (TME) are severely disturbed, requiring ecological restoration to recover biodiversity and ecosystem functions. However, the extent of restoration efforts across TMEs is not known due to the lack of syntheses on ecological restoration research. Here, based on a systematic review, we identify geographical and thematic research gaps, compare restoration interventions, and consolidate enabling factors and barriers of restoration success. We find that restoration research outside Latin-America, in non-forested ecosystems, and on socio-ecological questions is scarce. For most restoration interventions success is mixed and generally limited by dispersal and microhabitat conditions. Finally, we propose five directions for future research on tropical mountain restoration in the UN decade of restoration, ranging from scaling up restoration across mountain ranges, investigating restoration in mountain grasslands, to incorporating socioeconomic and technological dimensions. Tropical mountain ecosystems (TME) are hotspots of biodiversity 1,2 and endemism 3 and are located in tropical latitudes between 1000 and 4000 m asl, and the elevation gradients give rise to a variety of ecosystems including montane forests, montane cloud forests, forest-grassland treelines, mountain grasslands and azonal formations (Table 1). TME span across all continents in the tropical belt, and despite their small spatial extent of just over 4 million km 2 (Table 1) they provide numerous ecosystem services to people and society, including carbon sequestration, water regulation and supply, timber and food provision, erosion control, and cultural services 4. Notwithstanding their tremendous biological importance and complexity, TME are still relatively under-studied compared to temperate mountain systems 5. In recent decades, TME have been experiencing increasing pressure from multiple external drivers and stressors, such as anthropogenic pressures due to agricultural encroachment, pasture conversion and population growth 6 , exotic plantations 7,8 , invasion by exotic animals 9 and exotic plants 10-12 , as well as accelerating climate change impacts 13. These drivers lead to severe degradation in TME, impacting all levels of ecological organization, such as disruption of ecosystem services, losses in community diversity, changes in species interactions, reductions of population sizes and lowered genetic diversity 14. Degradation in TME is far-reaching and ubiquitous: Tovar et al. 15 projected that climate change will alter 3-7% of tropical Andean biomes, resulting in a 31.4% loss in extent of high-altitudinal Páramo grasslands due to replacement by montane forests by 2039. Further, Helmer et al. 16 indicate that in the next 25-45 years, reductions in cloud immersion are estimated to diminish 57-80% of Neotropical montane cloud forests. Hall et al. 17 estimate that the Tanzanian Eastern Arc mountains have lost 25% of forested areas since 1955, with deforestation rates of 57% in sub-montane forests (800-1200 m). At the same time, socioeconomic drivers have led to migrations of people from tropical mountains to urban areas, abandoning many previously cultivated and inhabited areas 24-26 and creating a large opportunity for ecosystem recovery and restoration across many TME. Restoration of biodiverse ecosystems, such as TME, has the potential to simultaneously recover lost biodiversity and ecosystem functioning and improve local livelihoods 27 , and has recently come to the fore of global conservation efforts 28. Restoration is defined as "the process of assisting the recovery of an ecosystem that has been degraded, damaged or destroyed" 29 and, as such, encompasses a broad suite of approaches ranging from passive restoration, to assisted recovery and active restoration. The urgency for global restorative actions culminated in global restoration pledges like the 2011 Bonn Challenge and the proclamation of the UN Decade of Ecosystem Restoration. Motivations to restore damaged ecosystems include conserving biodiversity (specific habitats or species), enhancing ecosystem processes (such as nutrient cycling), combatting climate change (through carbon OPEN
... Thus far, tropical ecosystem restoration has pre-dominantly been focussed around the lowlands, where restoration ecology has been thoroughly studied and synthesized in recent years [31][32][33][34] and various restoration methods have been tested and compared 34 . On the other hand, only a handful of reviews have addressed restoration questions related to TME, most of which are specific to a single tropical mountain ecosystem type or to a specific restoration context [31][32][33][35][36][37][38][39] . ...
... Over 33% of the studies were on a patch scale followed by local-scale studies and regional scale studies. There were only 5 pantropical/global assessments, which drew comparisons of restoration processes across distant mountain ranges or across continents 38,[85][86][87][88] . These findings are in line with trends in tropical forest restoration, where "neither the scale of scientific studies nor the restoration projects being implemented have matched the ambitious forest landscape restoration plans that are being proposed" 89 . ...
... Direct seeding of species and enrichment planting was most frequently studied in montane forest. A host of additional experimental methods were tested only a few times in the TME restoration studies, such as topsoil, seed bank and hay transfers in the mountain grasslands of the Campos Rupestres 48,108 , applied nucleation, assisted migration 136,137 and inoculation of cloud forest seedlings with arbuscular mycorrhizal 38,138 (Supplementary Fig. 6a). www.nature.com/scientificreports/ ...
Article
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Many tropical mountain ecosystems (TME) are severely disturbed, requiring ecological restoration to recover biodiversity and ecosystem functions. However, the extent of restoration efforts across TMEs is not known due to the lack of syntheses on ecological restoration research. Here, based on a systematic review, we identify geographical and thematic research gaps, compare restoration interventions, and consolidate enabling factors and barriers of restoration success. We find that restoration research outside Latin-America, in non-forested ecosystems, and on socio-ecological questions is scarce. For most restoration interventions success is mixed and generally limited by dispersal and microhabitat conditions. Finally, we propose five directions for future research on tropical mountain restoration in the UN decade of restoration, ranging from scaling up restoration across mountain ranges, investigating restoration in mountain grasslands, to incorporating socio-economic and technological dimensions.
... Uno de los ecosistemas tropicales prioritarios para la conservación y restauración a nivel mundial es el Bosque Mesófilo de Montaña (BMM), dada su excepcional diversidad, capacidad de proveer servicios ecosistémicos, reducida extensión y amenazas que enfrenta (Hamilton, 1995;Toledo-Aceves et al. 2011). Se estima que anualmente se pierde ~1.1% de cobertura mundial de BMM (Scatena et al. 2010), principalmente por cambio de uso de suelo para ganadería y agricultura (Aide et al. 2010). Como resultado de estos procesos en México se calcula que sólo el 28% de la cobertura original de BMM se mantenía en 2002 (Challenger y Dirzo 2009). ...
... Las tasas de recuperación por medio de la restauración pasiva son altamente variables y en algunos casos no tiene éxito (Holl, 2007;Aide et al. 2010). Se ha reportado que en bosques tropicales húmedos el tiempo de recuperación a partir de la regeneración natural puede llevar décadas para la estructura (Kappelle et al. 1996;Aide et al. 2000;Rüger et al. 2010;Muñiz-Castro et al. 2012), riqueza y diversidad (Letcher y Chazdon, 2009;Muñiz-Castro et al. 2012), y puede requerir siglos o no lograrse la composición de especies similar a la de los bosques maduros (Guariguata y Ostertag, 2001;Chazdon, 2003). ...
... Se ha reportado que en bosques tropicales húmedos el tiempo de recuperación a partir de la regeneración natural puede llevar décadas para la estructura (Kappelle et al. 1996;Aide et al. 2000;Rüger et al. 2010;Muñiz-Castro et al. 2012), riqueza y diversidad (Letcher y Chazdon, 2009;Muñiz-Castro et al. 2012), y puede requerir siglos o no lograrse la composición de especies similar a la de los bosques maduros (Guariguata y Ostertag, 2001;Chazdon, 2003). La recuperación de los ecosistemas tropicales depende de múltiples factores que incluyen el tipo de uso e intensidad de manejo previo de la tierra, el tipo de paisaje circundante y la resiliencia del ecosistema (Guariguata y Ostertag, 2001;Chazdon, 2003;Montagnini, 2008;Aide et al. 2010;Holl y Aide, 2011;Norden et al. 2015). Particularmente, se ha reportado que los factores que limitan la regeneración de los bosques tropicales de montaña después de su conversión a potreros son, la ocurrencia de fuego , la competencia con especies pioneras (p. ...
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El Bosque Mesófilo de Montaña (BMM) es un ecosistema prioritario para la conservación y restauración a nivel mundial, dada su excepcional diversidad, capacidad de proveer servicios ecosistémicos y reducida extensión. Una de las principales amenazas al BMM en el mundo es su conversión a potreros; sin embargo, cuando el uso ganadero es abandonado algunos propietarios permiten la regeneración natural (restauración pasiva) o establecen plantaciones (restauración activa). La evaluación de la recuperación de los atributos del ecosistema bajo diferentes estrategias de restauración es fundamental para entender los factores locales y de paisaje que determinan la recuperación de la estructura y función del BMM. En este estudio, se evaluaron diferentes estrategias de restauración después de 21 años de su implementación en el centro de Veracruz, México. Las condiciones estudiadas fueron una plantación mixta con especies nativas de BMM y dos áreas de restauración pasiva adyacente (0 a 430 m) y no adyacente (314 a 1,525 m) a un fragmento de BMM conservado, el cual fue el sistema de referencia. Los indicadores de éxito evaluados fueron: la estructura, diversidad y composición de árboles adultos, juveniles y plántulas; para ello en cada condición se establecieron ~15 parcelas de 200 m2. Los resultados muestran que en comparación con la restauración pasiva, el establecimiento de la plantación mixta aceleró la recuperación de la estructura de la vegetación (mayor área basal, diámetro, altura y cobertura del dosel). Si bien, la restauración activa y pasiva no difirió en los valores de riqueza y diversidad, la plantación mixta y la condición con restauración pasiva adyacente al bosque tuvieron mayor similitud en la composición de especies con el bosque de referencia que la condición pasiva no adyacente a éste. Se encontró baja densidad de plántulas y juveniles en todas las condiciones de restauración (0.39 plántulas/m2, 0.23 juveniles/m2), lo cual podría deberse a factores que operan a nivel del sitio como la elevada cobertura de plantas pioneras en el sotobosque, lo que sugiere que se requieren nuevas intervenciones para fomentar la regeneración arbórea. Si bien, la plantación no catalizó una mayor densidad de plántulas y juveniles, si contribuyó al establecimiento temprano de un mayor número de individuos del genero Quercus. Los resultados muestran que la restauración pasiva y activa pueden complementarse como estrategias de recuperación de BMM a escala del paisaje.
... Therefore, fencing and artificial reforestation through the outplanting of nursery-grown seedlings is often required. The successful restoration of grazed sites by tree planting is inhibited by myriad factors, including costs, seasonal drought, competition from invasive species, such as non-native grasses [5][6][7][8], and increased fire susceptibility caused by inflammable introduced grasses [9,10]. ...
Article
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Anthropogenic activity has caused persistent and prominent losses of forest cover in dry tropical forests. Natural regeneration of forest trees in grazed areas often fails due to lack of seed sources and consumption by ungulates. To address this, the effective restoration of such sites often requires fencing and outplanting nursery-grown seedlings. In the degraded, dry forests of tropical Hawaii, USA, an additional challenge to restoration of native forest trees is the introduced kikuyu grass (Cenchrus clandestinus). This invasive, rapidly growing rhizomatous plant forms deep, dense mats. We studied the use of nursery cultural techniques to facilitate the establishment of koa (Acacia koa) seedlings outplanted amidst well-established kikuyu grass on a volcanic cinder cone on the dry, western side of Hawaii Island. Seedlings were grown four months in three container sizes (49, 164, 656 cm3) and with four rates (0, 4.8, 7.2, and 9.6 kg m−3) of 15–9–12 (NPK) controlled-release fertilizer incorporated into media prior to sowing. After 16 months in the field, seedling survival was > 80% for all treatments with two exceptions: the non-fertilized 49 cm3 (78%) and 164 cm3 (24%) containers. After 10 years, only these two treatments had significantly lower survival (35% and 10%, respectively) than the other treatments. One year following planting, none of the non-fertilized seedlings had transitioned to phyllodes from juvenile true leaves, regardless of container size. For the fertilized 656 cm3 container treatment, 78%–85% of seedlings had phyllodes, with mean values increasing by fertilizer rate. Phyllodes are known to confer greater drought resistance than true leaves in koa, which may help to explain the improved survival of fertilized trees on this relatively dry site. Overall, nursery fertilization was more influential on seedling height and diameter response than container size after outplanting. However, the largest container (656 cm3) with the addition of fertilizer, produced significantly larger trees than all other treatments during the early regeneration phase; early growth differences tended to fade at 10 years due to inter-tree canopy competition. Although koa is able to fix atmospheric nitrogen through rhizobium associations, our data confirm the importance of nursery fertilization in promoting regeneration establishment. Nursery cultural techniques may play an important role in forest restoration of dry tropical sites invaded by exotic vegetation.
... Restoration of forest cover to these landscapes present opportunities for increasing or restoring biodiversity (Yelenik 2017), sequestering carbon (Silver et al., 2000), and providing habitat for threatened species (Flaspohler et al., 2010). Successful restoration of these sites is challenged by many factors, including costs of the operations, invasive species such as non-native grasses (Aide et al., 2010;Pinto et al., 2015;Yelenik 2017), and changes in fire regimes with introduced grasses (Vitousek et al., 1997;Brooks et al., 2004). ...
Article
Restoration of abandoned, high-elevation pastures is needed across many ecosystems. Diverse abiotic and biotic stressors often limit establishment of native trees species, however, justifying the need for novel approaches to alleviate such stressors. Freezing damage often negatively impacts survival of planted trees across temperate landscapes and on some high-elevation tropical restoration sites, such as for Acacia koa (koa) in Hawaii, USA. Koa performs poorly under forest canopies, a potential limitation to the use of nurse trees for establishment on frost-prone sites. Using a heterogeneous canopy of a non-native conifer, Cryptomeria japonica, we underplanted koa seedlings along a simulated range of canopy shelter levels in combination with field fertilization. We tested the effect of a canopy cover gradient and nutrient availability on frost avoidance and tolerance responses, as well as the potential to harness koa's developmental plasticity to optimize growth and survival. C. japonica canopy cover provided protection from frost damage, with increased sheltering under greater canopy closure. When combined with fertilization, increasing canopy closure reduced frost damage and increased koa growth. Although we observed limited frost damage in our study, leaf-level soluble sugars increased during the winter and in more open microsites, reflecting a potential mechanism for frost tolerance in this tropical species. We conclude that frost-tolerant conifers used as nurse trees represent a potential tool to help establish native tree species on high-elevation, frost-prone sites.
... An advantage is that passive restoration is generally perceived as a low-cost alternative, although in general it has costs that are often not taken into account such as the purchase of material (fences or barriers) to isolate the ground from disturbance agents and payments for site surveillance (Zahawi et al., 2014). It has the potential to achieve similar levels of biodiversity and environmental services as an active restoration; however, it is only feasible in certain places where the disturbance was not so intense, natural communities are resilient and are far from human communities (Holl, 1999;Zahawi and Augspurger, 1999;Muñiz-Castro et al., 2006;Suding and Hobbs, 2009;Aide et al., 2010;Holl and Aide, 2011). ...
Article
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To recover the structure and functionality of a deforested ecosystem, two strategies of ecological restoration are considered: active restoration, which eliminates the disturbance agents and implements strategies to accelerate site recovery, and passive restoration, which eliminates disturbance agents, allowing natural regeneration to occur. Prior to choosing passive restoration, a field evaluation of the potential for natural regeneration is important. In this context, seedling and sapling density as well as patterns of recruitment and survival are appropriate indicators of restoration potential. In the present study, we deduced the potential of sacred fir (Abies religiosa) forest of the Monarch Butterfly Biosphere Reserve to recover by natural regeneration through seedling and sapling density and mortality, since A. religiosa is the dominant tree species in wintering sites of monarch butterfly. In 2015, we evaluated seedling density in 53 sites along an elevational gradient (3050–3550 m above sea level; m a.s.l.). There was a higher density of seedlings and saplings established in canopy gaps, compared to sites under dense forest canopy. Seedling recruitment was higher in sites at intermediate elevations (3050 to 3300 m a.s.l.) than in those at higher elevations. In a second survey, we studied A. religiosa seedling mortality over the dry season of 2016 to identify the environmental variables that cause the high seedling mortality and very low recruitment. Recently emerged seedling mortality was 49.2% at the end of the dry season (June 2016). The highest monthly mortality (14.3%) was recorded in April, a dry and warm month with the lowest values of moss thickness and soil moisture. We found no negative effects of moss layer on seedling mortality; indeed, moss appears to slow soil moisture reduction at the critical end of the warm and dry season. Soil and moss moisture values in April seem to be a critical factor for A. religiosa seedling recruitment, and we expect this condition will deteriorate under projected climatic change scenarios. Thus, the potential of MBBR A. religiosa forest to recover by passive restoration is highly constrained and will require management actions to achieve successful restoration outcomes.
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Restoration of tropical montane cloud forest landscapes is urgently required. Assisting the regeneration of endangered and shade tolerant tree species is essential for both the recovery of this vulnerable group and of ecological processes. However, there is limited species‐specific information regarding tree performance under different disturbance conditions with which to implement effective interventions. We assessed the performance of shade tolerant tree seedlings in restoration plantings under different disturbance settings and determined whether leaf mass area (LMA) and leaf dry mass content (LDMC) – functional traits typically associated with resource capture or stress tolerance – could serve as predictors of survival and growth among species. Since conservative leaf morphological traits can maximize survival, we expected species with higher LMA and LDMC to present higher survival. For a set of eight native cloud forest species, a total of 2202 seedlings were planted in four pastures, five secondary forests and three forests subjected to traditional selective logging, in tropical montane cloud forest landscapes in Eastern Mexico. Seedling survival was high after three years: 62% in pastures, 80% in secondary forests and 88% in logged forests. Growth rates were lowest in pastures, followed by secondary forests and highest in logged forests. LMA was a strong predictor of seedling survival in all of the environments; tree species with higher LMA presented greater survival. LDMC was related to seedling survival in the three environments, although to a lesser extent than LMA. In the pastures, higher LMA and LDMC were linked to lower growth. Synthesis and applications. This study supports the potential of shade tolerant tree species in restoration efforts to assist the recovery of this important functional group and to accelerate succession across altered environments. Our results support the notion that conservative leaf functional traits are linked to a higher probability of survival, not only in the shaded understorey, but also under high solar radiation in transformed habitats. Leaf Mass Area (LMA) in particular is a reliable predictor of seedling survival for shade tolerant species. Species selection based on LMA could thus improve restoration initiative outcomes: tree species with high LMA present higher survival probability and can be introduced into pastures, secondary forests and selectively logged forests.
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After one growing season of recovery, vegetation cover and height, species richness, and life forms composition were surveyed in 15 sites located along a gas-pipeline track running through an altitudinal range from 400-2000 m in the subtropical mountains of north-western Argentina (23°S). Vegetation cover was negatively correlated with altitude but was generally high at all sites (> 60%) after one year. Total species richness and maximum vegetation height did not vary significantly with altitude. Cover of grasses, and cover and species richness of trees, small shrubs and climbers were negatively correlated with altitude. Herb species richness correlated positively with altitude. Large shrub richness and cover showed no statistical relationship with altitude. The relative cover of herbs and grass species richness did not vary along the altitudinal gradient. Overall, these results indicate that in the altitudinal range studied, vegetation recovery is relatively high after this type of disturbance, probably due to low dispersal limitations and to the availability of species well adapted to intense disturbances. Vegetation recovery after gas-pipeline construction or similar perturbations may lead to relatively fast ecological restoration in subtropical montane forest ecosystems.