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Vegetation structure, species diversity, and ecosystem processes as measures of restoration success

Authors:
  • Food and Agricultural Organization of the United Nations (FAO)

Abstract and Figures

Most restoration projects have focused on recovery of vegetation to assess restoration success. Nevertheless if the goal of a restoration project is to create an ecosystem that is self-supporting and resilient to perturbation, we also need information on the recovery of other trophic levels and ecosystem processes. To provide an example on how to assess restoration success, we compared four measures of vegetation structure, four measures of species diversity, and six measures of ecosystem processes among pre-reforested, reforested, and reference sites. In addition, we described how Bray Curtis Ordination could be used to evaluate restoration success. Vegetation structure recovered rapidly due to the increase in vegetation height and the decrease in herbaceous cover. Other measures such as litter cover, number of litter layers, and DBH size class values are recovering at slower rates, but they also have increased vegetation heterogeneity in the reforested site. Species diversity recovered rapidly. The increase in vegetation structure changed the local conditions in the reforested site facilitating the colonization of woody seedlings, ants, reptiles, and amphibians. Ecosystem processes, particularly litter production and turnover, have enhanced the incorporation of nutrients and organic matter in the soil. By including vegetation structure, species diversity, and ecosystem processes measures we have better information to determine the success of a restoration project. Moreover, the Subjective Bray Curtis Ordination is a useful approach for evaluating different restoration techniques or identifying measures that are recovering slowly and would benefit from additional management.
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Vegetation structure, species diversity, and ecosystem
processes as measures of restoration success
Marı
´a C. Ruiz-Jae
´n*, T. Mitchell Aide
Department of Biology, University of Puerto Rico-Rio Piedras, P.O. Box 23360,
San Juan 00931-3360, Puerto Rico
Received 25 April 2005; received in revised form 13 July 2005; accepted 14 July 2005
Abstract
Most restoration projects have focused on recovery of vegetation to assess restoration success. Nevertheless if the goal of a
restoration project is to create an ecosystem that is self-supporting and resilient to perturbation, we also need information on the
recovery of other trophic levels and ecosystem processes. To provide an example on how to assess restoration success, we
compared four measures of vegetation structure, four measures of species diversity, and six measures of ecosystem processes
among pre-reforested, reforested, and reference sites. In addition, we described how Bray Curtis Ordination could be used to
evaluate restoration success. Vegetation structure recovered rapidly due to the increase in vegetation height and the decrease in
herbaceous cover. Other measures such aslitter cover, number of litter layers, and DBH size class values are recovering at slower
rates, but they also have increased vegetation heterogeneity in the reforested site. Species diversity recovered rapidly. The
increase in vegetation structure changed the local conditions in the reforested site facilitating the colonization of woody
seedlings, ants, reptiles, and amphibians. Ecosystem processes, particularly litter production and turnover, have enhanced the
incorporation of nutrients and organic matter in the soil. By including vegetation structure, species diversity, and ecosystem
processes measures we have better information to determine the success of a restoration project. Moreover, the Subjective Bray
Curtis Ordination is a useful approach for evaluating different restoration techniques or identifying measures that are recovering
slowly and would benefit from additional management.
#2005 Elsevier B.V. All rights reserved.
Keywords: Bray Curtis Ordination; Ecosystem restoration; Karst; Limestone
1. Introduction
Most restoration projects have focused on the
recovery of vegetation (Young, 2000) to assess
restoration success. Nevertheless if the goal of a
restoration project is to create an ecosystem that is
self-supporting and resilient to perturbation (SER,
2004), we need to measure more than just vegetation.
What measures need to be assessed to determine if a
restored site is self-supporting? Vegetation structure,
species diversity, and ecosystem processes have been
identified as essential components for a long-term
www.elsevier.com/locate/foreco
Forest Ecology and Management 218 (2005) 159–173
* Corresponding author at: Department of Biology, 1205 Dr.
Penfield, McGill University, Montreal, Canada H3A 1B1.
Tel.: +514 398 3730; fax: +514 398 5069.
E-mail address: mcruizjaen@gmail.com (M.C. Ruiz-Jae
´n).
0378-1127/$ – see front matter #2005 Elsevier B.V. All rights reserved.
doi:10.1016/j.foreco.2005.07.008
persistence of an ecosystem (Elmqvist et al., 2003;
Dorren et al., 2004). Measures of vegetation structure
provide information on habitat suitability, ecosystem
productivity, and help predict successional pathways
(Jones et al., 2004; Silver et al., 2004; Wang et al.,
2004). Measures of species diversity provide informa-
tion on susceptibility to invasions (e.g., proportion of
native and exotic species), and trophic structure
necessary for ecosystem resilience (Parmenter and
MacMahon, 1992; Peterson et al., 1998; Nichols and
Nichols, 2003). Measures of ecosystem processes
provide information on biogeochemical cycles and
nutrient cycling necessary for the long-term stability
of the ecosystems (Herrick, 2000). Most restoration
projects measure some aspects of vegetation structure
or diversity, arthropod diversity or nutrient pools
(Ruiz-Jae
´n and Aide, 2005), but studies rarely assess
more than one measure of each component.
Along with assessing many measures in a restored
site, it is necessary to compare this information with
similar data from pre-restored and reference sites
(Hobbs and Norton, 1996). The pre-restored and
reference sites should occur in the same life zone,
close to the restoration project, and should be exposed
to similar natural disturbances (Hobbs and Harris,
2001; SER, 2004). If chosen correctly, these sites can
provide endpoints to evaluate the success of a project
(Passell, 2000; Purcell et al., 2002). The use of
reference points can help to identify whether the
response of the restored site is caused by the
restoration activity or by unassisted recovery (White
and Walker, 1997).
The goal of this study is to provide an example of
how to evaluate restoration success in an integrative
way using measures of vegetation structure, species
diversity, and ecosystems process. We evaluated
restoration success by comparing four measures of
vegetation structure, four measures of species diver-
sity, and six measures of ecosystem processes among
pre-reforested, reforested, and reference sites in
Puerto Rico. We addressed the following questions:
(1) how does the vegetation structure of the reforested
area changed in comparison to both the pre-reforested
and reference sites? (2) how does the change in
vegetation structure enhance species diversity in the
reforested site? and (3) how does the ecosystem
processes of the reforested site changed in comparison
to both pre-reforested and reference sites?
2. Methods
2.1. Study area
The study was conducted in Sabana Seca, Puerto
Rico (188270N, 668120W). This site is located in the
northern limestone region, and is classified as
subtropical moist forest (Holdridge, 1967). Mean
annual rainfall is 1693 mm with a rainy season
between April and December and a dry season
between January and May (Eusse and Aide, 1999).
The pre-reforested site was a park in a Karst valley
where the grass was cut on a regular basis. This site
was abandoned (i.e. no longer mowed) in May 2000.
The reforested site also was a valley previously
maintained as a park, and the grass was frequently
mowed. This site was reforested with 22 native species
from 18 families including trees and shrubs in January
2000, and was no longer mowed. The species chosen
included both pioneer and shade tolerant species
common in Karst ecosystems of Puerto Rico (Alvarez-
Ruiz et al., 1997; Rivera and Aide, 1998). A total of
516 seedlings were planted (1612.5 seedlings ha
1
).
Seedling survivorship was high (96% after 15 months
and 93% after 27 months). The major goal of the
restoration project was to recover native vegetation of
a Karst valley to provide habitat for the endangered
Puerto Rican boa, Epicrantes inornatus. The reference
forest is a secondary forest in a Karst valley that was
dominated by pastures until abandoned approximately
40 years ago. The most common tree species in this
site are Faramea occidentalis,Guarea guidonea, and
Quararibea turbinata. Although the reference site is a
secondary forest, we selected it because it was the
oldest secondary forest in the region with the same
environmental conditions as the other two sites (e.g. a
valley previously dominated by grasses).
In the past, the most common land use practices in
the Karst valleys in Puerto Rico were pastures,
shifting agriculture or coffee plantations (Rivera and
Aide, 1998). Most of these areas have been
abandoned, but previous land use can influence
present day species composition (Rivera and Aide,
1998). To control for land use history, we selected
sites that were formerly dominated by grasses based
on aerial photographs.
The three valleys are surrounded by forested hills,
and the area of the sites ranged from 0.75 to 1.0 ha
M.C. Ruiz-Jae
´n, T.M. Aide/ Forest Ecology and Management 218 (2005) 159–173160
with a minimum distance of 300 m between sites.
Each site has a different shape, therefore we used
different number of transects in each site, but a total of
200 m of transects were established in all sites.
Specifically, the pre-reforested and reforested sites
were sampled using two transects of 100 m, and the
reference site was sampled with four transects of 50 m.
In all sites transects were established systematically at
least 5 m away from the edge and 15 m between each
transect.
2.2. Vegetation structure
2.2.1. Ground cover and litter structure
Ground cover and litter structure were estimated in
twenty 1 m
2
plots in May 2003. Plots were located
every 10 m along each transect. Percent herbaceous,
litter, and bare soil cover was determined in each plot.
Herbaceous cover included grasses, vines, and herbs.
Litter structure was determined by counting the
number of leaves perforated with a needle that was
pushed down though the leaf litter layer in four points
within each plot (Vasconcelos et al., 2000).
2.2.2. Forest structure
Diameter at breast height (DBH) and height of all
woody plants 1 cm DBH were sampled within
800 m
2
along the established transects.
2.3. Species diversity
2.3.1. Woody plant seedlings
Woody seedlings from 5 to 50 cm in height were
counted and identified in 20 circular plots (1 m
diameter) in February 2003. Plots were located every
10 m along each transect.
2.3.2. Ants
Ants were collected using leaf litter samples and
pitfall traps in May 2003. Each sampling technique
was applied every 10 m along each transect. Twenty
leaf litter samples (1 m
2
) were collected. In the field
the litter was sifted (100 mm
2
-mesh) to eliminate large
debris. The sifted samples were placed in Berlese
funnels in the laboratory (modified from Agosti and
Alonso, 2000). Ants were removed from Berlese traps
after 48 h. Twenty pitfall traps (0.005 m
2
) were left in
the field for 48 h.
2.3.3. Amphibians and reptiles
Composition and abundance of the herpetofauna
were determined by diurnal and nocturnal visual and
acoustic census. The censuses were conducted
monthly in transects of 3 m 200 m in each site
from January to December of 2002. Diurnal censuses
were conducted between 08:30 and 13:30 h, and
nocturnal census was conducted between 18:30 and
00:30 h. On average it took 2.5 h to complete a diurnal
or nocturnal census.
2.3.4. Birds
Composition and abundance of birds in each site
were determined in a 10 m 100 m transect in each
site in August and September 2004. Six predawn
visual and acoustic censuses were conducted (Bibby
et al., 2000).
2.4. Ecosystem processes
2.4.1. Litter production and litter turnover
Leaf litter production was estimated by collecting
litter from 20 plastic buckets (area 0.071 m
2
/bucket) in
each site. Buckets were located on the forest floor
every 10 m along transects. Leaf litter was collected
monthly from April 2003 to March 2004 and ground
litter was collected in June 2003, October 2003, and
March 2004 from 20 plots of 0.25 m
2
. Litter samples
were separated into leaves and miscellaneous (wood,
fruits, flowers) before drying at 70 8C for 72 h. The
Olson (1963) formula k=L
f
/L
s
was used to determine
the decomposition constant (k), where L
f
is the annual
leaf litter fall (g m
2
yr
1
) and L
s
is the standing leaf
litter biomass (g m
2
).
2.4.2. Nutrient content
In each site, soil was sampled in three randomly
located plots separated by at least 30 m. In each plot,
three soil cores of 31.4 cm
3
each were collected at a
depth of 0–10 cm and three at 10–20 cm. Samples
from the same depth and plot were combined for
analyses. Soil pH, bulk density, total P, N, organic C,
Ca, K, Mg, and exchangeable cations of Ca, K, and Mg
were determined for each soil sample. Soil pH was
measured in water in a ratio of 1:5. Bulk density was
measured with a soil core of 31.4 cm
3
, samples were
oven-dry at 110 8C and weighted. Exchangeable
cations were extracted with ammonium acetate 1 M
M.C. Ruiz-Jae
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(Thomas, 1982), and by atomic absorption spectro-
meter. Total organic carbon was determined by the
colorimetric method with digestions of sulphuric acid
and potassium dichromate at 5% (Anderson and
Ingram, 1993). Nitrogen was analysed using Kjeldahl
procedure involving digestion with sulphuric acid
(Jackson, 1968). Phosphorus, calcium, potassium, and
magnesium were determined by atomic absorption
spectrometer with sulphuric and perchloric acid
digestions (Murphy and Riley, 1962; Olsen and
Sommers, 1982). Total nutrient content of P, N, Ca,
K, and Mg were measured in the monthly litter fall
samples from the three sites. Litter nutrient analysis
was based on the same procedures described for the
soil analysis. Nutrient content of monthly litter fall
was used to determine potential nutrient inputs to each
site. Potential nutrient inputs were determined by
multiplying monthly values of litter fall with its
corresponding nutrient concentration.
2.4.3. Earthworms
Earthworms were collected in 12 plots of
25 cm 25 cm 20 cm every 15 m along transects
in July 2003 (Zou and Gonzalez, 1997). Each soil
sample was separated in two profiles: 0–10 and 10–
20 cm. Each soil section was placed on a clothsheet and
earthworms were hand sorted and stored in plastic bags
in a cooler with ice. The same day, fresh weight was
determined in the laboratory after the worms have been
rinsed with water and dried with paper towels. Soil
moisture was measured in each soil sample because the
earthworm distribution is strongly dependent on water
content. Soil moisture was calculated for each site by
oven-drying 15 g of fresh soil sample at 105 8C for 48 h.
2.4.4. Carbon isotope ratios
Soil samples (0–10 and 10–20 cm) used for analysis
of nutrient content were also used to measure the carbon
stable isotope ratio (d
13
C). Samples from these depths
best reflect changes between C
3
and C
4
vegetation
(Jobba
´gy and Jackson, 2000). Leaf samples consisted
on 10 green leaves of the most abundant species within
each soil-sampling plot. Soil organic matter (SOM)
composition was assessed using the stable isotope
proportion of carbon 13 and 12 (d
13
C). Carbonates were
extracted from eachsoil sample with 0.5N HCl because
our sites were located in the limestone region of Puerto
Rico. Both soil and plant samples were oven dried at
70 8C for 48 h and grounded to powder before isotopic
proportion determination. Carbon isotope analyses for
both soil and leaf samples followed procedures
explained elsewhere (see Martin et al., 1990; Eshetu,
2002). The d
13
C values of soil organic matter are mainly
dependent on plant composition input material. Most
tropical grasses (C
4
) exhibit values of 9to19%and
shrubs and trees (C
3
) show values 23 to 40%(Smith
and Epstein, 1971). Contribution of C
3
and C
4
plants to
soil organic matter composition was determined by
using the formula in Trouve et al. (1994).C
t
=C
4
+C
3
,
and C
t
d
t
=(C
4
d
4
)+(C
3
d
3
), where d
t
is the
measure of d
13
C value of soils, and d
3
and d
4
are d
13
C
value of C
3
and C
4
plants, respectively. The relative
abundance was calculated as C
3
/C
t
(%) = [(d
t
d
4
)/
(d
3
d
4
)] 100 and C
4
/C
t
(%) = [(d
t
d
3
)/(d
4
d
3
)]
100.
2.5. Data analyses
Given that treatments (i.e. pre-reforested, reforested,
and reference) were not replicated, statistical analyses
are not included. Rank-density graphs were used to
assess difference in community dominance (e.g.
seedlings, ants, herpetofauna, and birds) among sites
(Feinsinger, 2001). These graphs ranked species in each
site from the highest to the lowest density. Moreover,
rank-density graphs provide information on species
richness in each site.
Restoration successwas estimated with a Subjective
Bray Curtis Ordination (McCune and Mefford, 1999;
Pcord4 Software). The Subjective Bray Curtis Ordina-
tion places points in relationship to selected reference
sites (i.e. endpoints). Specifically, the data from the
reforested site was arrayed relative to the endpoints(i.e.,
pre-reforested and reference sites) along a horizontal
axis by using the Sorensen coefficient of similarity as
the distance measure (Bray and Curtis, 1957; McCune
and Grace, 2002). The position of the reforested site
along this axis indicates the percent of restoration
success relative to the endpoints. In contrast to other
commonly used ordination methods (e.g. NMS, PCA,
DCA, and CCA), the Subjective Bray Curtis Ordination
is specific for evaluating data with conceptual
references points (McCune and Grace, 2002).
For the Subjective Bray Curtis analysis we used
five measures of vegetation structure, three measures
of species diversity, and five measures of ecosystem
M.C. Ruiz-Jae
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processes. For the analyses of vegetation structure, the
values (e.g., herbaceous cover, litter cover, litter
layers, DBH size classes, and plant height) were
categorized into ranges to reflect the variability in
these measures. For the analyses of woody seedlings
and herpetofauna, we used species abundance. For ant
species, presence and absence data were used to avoid
the spatial clumping of their distribution due to nesting
behaviour (Longino, 2000). For litter production we
used monthly average of each site. For litter turnover
rates we used the decomposition constant of each site.
Nutrients inputs were compared using monthly values
(litterfall nutrient concentration). For soil nutrient
content, we used the mean value of P, N, Ca, K, and
Mg for each site. Bulk density was compared using
mean values from two soil profiles (0–10 and 10–
20 cm).
Birds and earthworms were not included in the
Subjective Bray Curtis Ordination. Birds were not
included because only one species was found in the
pre-reforested site, and any increase in species
diversity would result in a higher recovery rate.
Earthworm fresh weight was also excluded because
values in the reforested site were outside the range of
the endpoints (e.g., pre-reforested and reference sites).
3. Results
3.1. Vegetation structure
Three years after planting, the growth of woody
stems in the reforested site created a diverse vegetation
structure (Fig. 1). The number of stems in the 1 to
<5 cm DBH size classes was higher in the reforested
site (171 stems) than in the pre-reforested site, which
had only one stem in this category. The reforested site
had fewer stems in the 5–10 cm DBH classes (n=8)in
comparison with the reference site (n= 156; Fig. 1a).
Nevertheless the reforested site has stems >10 cm
DBH (n= 3). Vegetation height in the reforested site
ranged from 1.4 to 12 m with a mean height of
approximately 3.0 m (Fig. 1b). Although there is some
vertical stratification in the reforested site, the
reference site had a greater range of tree heights
(e.g., 1.4–17 m). The tallest trees in the reforested site
were the pioneer species, Cecropia schreberiana,
Senna siammea, and Thespesia grandiflora.
Herbaceous cover and litter cover also varied
among sites. Herbaceous cover was highest in the pre-
reforested site (94%), lower in the reforested site
(40%), and lowest in the reference site (Fig. 1c). In
contrast, litter cover was lowest in the pre-reforested
site (6%), higher in the reforested site (47%), and
highest in the reference site (88%). Similarly, the
number of litter layers was lowest in the pre-reforested
site (1.5), higher in the reforested site (2.1), and the
highest in the reference site (3.0; Fig. 1d). The litter
layers in the reforested site were dominated by the
pioneer species, Cecropia schreberiana,Hura crepi-
tans,Senna siammea, and Thespesia grandiflora. The
litter in the pre-restored site was dominated by
herbaceous vegetation, while the reference site was
dominated by woody species (e.g. Chrysophyllum
argenteum,Faramea occidentalis,Guarea guidonea,
and Quararibea campalunata).
3.2. Species diversity
The development of a complex vegetation structure
in the reforested site has changed the microhabitat and
facilitated the colonization of other organisms
(Appendix A and Fig. 2). For example, the low
herbaceous cover in the reforested site was associated
with the colonization of 22 woody plant seedlings,
while there was low colonization of woody plants in the
pre-reforested with high herbaceous cover (Fig. 1c and
2a). The dominant species in the reforested site was
the wind-dispersed vine, Hippocratea volubilis
(Appendix A and Fig. 2a). Other species in the
reforested site included some common Karst species
(e.g., Casearia sylvestris,Guarea guidonea, and
Tabebuia heterophylla), but there were still species
common to disturbed sites (e.g., Urena sinuata and U.
lobata). Moreover, animal dispersed seeds species (e.g.,
Andira inermis,Casearia sylvestris,Cupania amer-
icana,andThespesia grandiflora) were only present in
the reforested and reference sites (Appendix A).
Ant richness and density also varied among sites.
Ant species richness was lowest in the pre-reforested
site (15 species), higher in the reforested site (21
species), and highest in the reference site (30 species;
Appendix A and Fig. 2b). Ten species were present in all
sites and seven species only occurred in the reforested
and reference sites (Appendix A). Ant densities also
differed among sites. Solenopsis geminata, an exotic
M.C. Ruiz-Jae
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invasive, had the highest density in the pre-reforested
site (6000 individuals m
2
), a lower density in the
reforested site (363 individuals m
2
), and the lowest
density in the reference site (50 individuals m
2
;
Fig. 2b). In contrast, Odontomachus ruginodis, a native
predator, has the lowest densities in the pre-reforested
site (20 individuals m
2
), the highest in the reforested
site (140 individuals m
2
), and 90 individuals m
2
in
the reference site.
The composition and density of the herpetofauna
varied among sites. Two exotic species were observed
in the pre-reforested site, nine species in the reforested
site, and eight species in the reference site
(Appendix A). Four of the nine species in the reforested
site are arboreal (Anolis cristatellus,Eleutherodactylus
cochranae,E. coqui, and Anolis cuvieri), but the three
most abundant species (Anolis pulchellus,A. krugi, and
Eleutherodactylus antillensis) are associated with herbs
and grasses (Fig. 2c). Epicrates inornatus, the Puerto
Rican boa, the target species of this restoration project,
colonized the reforested site, once prey species
increased (Rios-Lopez and Aide, unpublished data).
The overall herpetofauna density increased from
17 individuals ha
1
in the pre-reforested site to 1339
individuals ha
1
in the reforested and 1361 individuals
ha
1
in the reference site (Fig. 2c).
In a rapid assessment of the bird community, only
one species was observed in the pre-reforested site,
three species in the reforested site, and nine species in
the reference site (Appendix A and Fig. 2d). Six of the
M.C. Ruiz-Jae
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Fig. 1. Recovery of vegetation structure in the reforested site in comparison to pre-reforested, and reference sites: (a) number of woody stems in
different DBH size classes, (b) vegetation height of woody stems in meters, (c) percent of ground cover (herbaceous, litter, and bare soil), and (d)
number of litter layers. Boxes (Fig. 1b and d) represent 25–75 percentiles, lines within boxes represent the median value, and bars indicate the
90th and 10th percentiles, and points are outliers.
nine species in the reference site are endemic
(Dendroica adelaidae,Melanerpes portoricensis,
Saurothera vieilloti,Spindalis portoricensis,Todus
mexicanus), and one species is a top predator (Otus
nudipes). This rapid assessment approach provides a
good estimate of the bird diversity in the pre-
reforested and reforested sites, but failed to encounter
all species that are present in the reference site (circa
20, Acevedo, unpublished data).
3.3. Ecosystem processes
Ecosystem processes varied among sites. Litterfall
production, an indirect measure of productivity, was
higher in sites with developed vertical stratification
(i.e. reforested and reference sites; Table 1 and Fig. 1a
and b). For example, total litter fall was lowest in the
pre-reforested site (101 g m
2
yr
1
), higher in the
reforested site (467 g m
2
yr
1
), and highest in the
reference site (838 g m
2
yr
1
;Table 1). Litter turn-
over of the in the reforested site had values similar to
the reference site (Table 1). Leaf litter residence time
in both the reforested (230 d) and reference sites (139
d) are less than 8 months, while in the pre-reforested
site is it was more than 2 years (763 d).
Litterfall nutrient inputs varied among the sites
(Table 1). Phosphorus, nitrogen and calcium inputs
were lower in the pre-reforested site, higher in the in the
reforested site, and highest in the reference site
(Table 1). Potassium and magnesium inputs were much
lower in the reforested than the reference site (Table 1).
Soil nutrient content in the soil also varied among
sites, as did the other ecosystem processes. Total
phosphorus content in the soil was lower in the pre-
reforested site (3.2 g m
2
) and higher in reforested
(6.16 g m
2
) and reference (6.82 g m
2
)sites(Table 2).
Similarly, nitrogen content in the soil was lower in the
pre-reforested site (20.5 g m
2
), and higher in refor-
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Fig. 2. Species rank-density curves: (a) woody plant seedlings (20 plots each 0.78 m
2
), (b) ants (20 plots each 1 m
2
), (c) herpetofauna (n=12
census during Feb. 2001 through Feb. 2002 in 600 m
2
), and (d) birds (n= 6 census in August and September 2004 in 1000 m
2
). Species are
plotted in rank order based on density. Data for each species are means of plots or census. Note in Fig. 2bSolenopsis geminata reached densities
of 6000 individuals m
2
in the pre-reforested site, and in Fig. 1d, the pre-reforested site had only one bird species.
ested (24.5 g m
2
) and reference (26.5 g m
2
)sites.
Magnesium content and bulk density were similar in the
pre-reforested and reforested sites, but different from
the reference site (Table 2 ). Soil pH, total calcium and
exchangeable cations (Ca, K, and Mg) did not vary
among the three sites. N, C, K, and Mg content and bulk
density varied with soil depth (Table 2). N and C were
higher and bulk density, K, and Mg were lower in the 0–
10 cm soil profile in comparison with the 10–20 cm
profile.
There was no difference in earthworm fresh
weights among the three sites (Table 2). Earthworm
fresh weight in the 0–10 cm soil profile was highly
variable within sites, the pre-reforested site had
43.84 41.2 g m
2
, the reforested site had
48.22 33.3 g m
2
, and the reference site had
34.78 30.2 g m
2
. In the 10–20 cm profile, there
were no earthworms in the reforested site, but the pre-
reforested site had 0.8 1.3 g m
2
and the reference
site had 5.2 7.8 g m
2
.
After planting C
3
species (31.7%) in an area
dominated by grasses (C
4
;13.4%), there has been a
change in
13
C of SOM (Table 2). d
13
C of SOM was
higher in the 0–10 cm profile than in the 10–20 cm
profile. d
13
C values of SOM at 0–10 cm profile were
highest in the pre-reforested site, lower in the
M.C. Ruiz-Jae
´n, T.M. Aide/ Forest Ecology and Management 218 (2005) 159–173166
Table 1
Litter production, mean forest floor mass, litter turnover coefficient, residence time, an nutrient inputs for leaf litter in the pre-reforested,
reforested, and reference sites
Measures Sites
Pre-reforested Reforested Reference
Litterfall production (g m
2
yr
1
) 101.3 (78) 466.8 (279) 838.4 (322)
Mean forest floor mass (g m
2
) 121.6 (38) 212.4 (44) 296.7 (63)
Litter turnover coefficient (k) (yr
1
)
a
0.8 (0.5) 2.0 (0.9) 2.8 (0.5)
Residence time (1/k) 2.1 (1.7) 0.6 (0.4) 0.4 (0.1)
Phosphorus inputs (kg ha yr
1
) 0.6 (0.1) 3.2 (0.2) 10.9 (0.4)
Nitrogen inputs (kg ha yr
1
) 14.9 (1.1) 74.2 (4.4) 227.4 (8.4)
Calcium inputs (kg ha yr
1
) 16.9 (1.3) 75.2 (4.5) 242.8 (9.0)
Potassium inputs (kg ha yr
1
) 2.2 (0.2) 9.4 (0.6) 38.0 (1.4)
Magnesium inputs (kg ha yr
1
) 5.5 (0.4) 17.6 (1.0) 83.2 (3.1)
a
Total annual litterfall divided by the mean forest floor litter mass.
Table 2
Soil pH, water content, bulk density, nutrient content, d
13
C values, and proportion of C
3
and C
4
in organic carbon at 0–10 and 10–20 cm soil
profile in the pre-reforested, reforested, and reference sites
Measures/soil depth Pre-reforested Reforested Reference
0–10 cm 10–20 cm 0–10 cm 10–20 cm 0–10 cm 10–20 cm
pH 5.8 (0.1) 5.8 (0.1) 5.5 (0.1) 5.7 (0.1) 5.8 (0.3) 5.5 (0.3)
Water content (%) 20.2 (1.5) 18.5 (1.9) 26.4 (1.4) 23.8 (1.8) 26.6 (4.3) 26.6 (8.8)
Bulk density (g cm
3
) 1.0 (0.0) 1.2 (0.0) 0.9 (0.1) 1.1 (0.0) 0.9 (0.0) 1.0 (0.1)
P(gm
2
) 3.2 (0.3) 3.2 (0.4) 6.6 (0.8) 7.0 (0.7) 6.3 (0.6) 6.1 (0.6)
N(gm
2
) 22.6 (2.6) 18.5 (2.2) 26.4 (2.2) 26.6 (2.4) 26.8 (2.8) 22.2 (2.1)
Organic C (%) 4.1 (0.6) 3.5 (0.1) 4.4 (0.5) 3.9 (0.6) 4.3 (0.3) 2.1 (0.4)
Ca (g m
2
) 9.0 (1.1) 8.1 (0.8) 10.8 (0.8) 11.2 (1.8) 12.4 (5.6) 6.5 (2.0)
K(gm
2
) 11.3 (1.5) 12.2 (1.8) 12.8 (1.1) 17.6 (2.5) 9.5 (1.5) 11.3 (3.0)
Mg (g m
2
) 0.3 (0.1) 0.5 (0.4) 0.4 (0.3) 1.1 (0.1) 1.0 (0.2) 1.6 (0.7)
Ca cations (meq Ca/100 g) 2.3 (0.9) 1.9 (0.9) 2.7 (0.6) 2.0 (0.6) 3.3 (1.7) 1.7 (0.8)
K cations (meq K/100 g) 0.1 (0.0) 0.1 (0.0) 0.1 (0.0) 0.1 (0.0) 0.1 (0.1) 0.1 (0.0)
Mg cations (meq Mg/100 g) 0.9 (0.4) 0.7 (0.2) 0.8 (0.1) 0.6 (0.2) 0.8 (0.3) 0.5 (0.2)
d
13
C values (%
0
)18.4 (1.1) 21.4 (1.2) 22.7 (2.1) 21.7 (1.4) 26.9 (1.0) 25.9 (0.9)
Proportion of C
3
in organic carbon (%) 27.1 (5.8) 43.5 (6.7) 51.0 (11.4) 45.5 (7.9) 73.8 (5.7) 68.5 (5.0)
Proportion of C
4
in organic carbon (%) 72.9 (5.9) 56.5 (6.8) 49.0 (11.4) 54.5 (7.9) 26.2 (5.7) 31.5 (5.0)
Values are mean and standard deviation in parenthesis (n= 3).
reforested site, and highest in the reference site
(Table 2). In contrast, d
13
C values of SOM at 10–
20 cm profile were similar in the pre-reforested and
reforested sites, and the reference site had the lowest
d
13
C values (Table 2). Moreover, the proportion of
total organic carbon of C
4
origin in the 0–10 cm soil
profile was higher in the pre-reforested site, lower in
the reforested site, and the lowest in the reference site
(Table 2).
3.4. Restoration success—Bray Curtis Ordination
Vegetation structure, species diversity, and ecosys-
tem processes have responded rapidly to planting. The
Bray Curtis analyses showed that most measures of
vegetation structure and species diversity have recov-
ered >50% compared with the reference site (Table 3).
For vegetation structure, litter cover had the slowest rate
of recovery, while height of woody stems had the
fastest. For species diversity, ants had the fastest
recovery, followed by the herpetofauna, and woody
seedlings had the slowest recovery. In general,
ecosystem processes were slower to recover in com-
parison with vegetation structure and species diversity.
Litter production had the fastest recovery, while litter
turnover rates, litter nutrients, soil nutrients in 0–10 cm
soil profile, and bulk density will take longer to recover.
4. Discussion
4.1. Vegetation structure
Woody vegetation height and herbaceous cover
were the measures of vegetation structure that changed
most rapidly due to the planting and early establish-
ment of pioneer species in the reforested site. These
pioneer species (e.g., Cecropia shreberiana,Roysto-
nea borinquena, and Thespesia grandiflora, planted;
Delonix regia and Senna siammea, colonizers) quickly
accumulate biomass (Guariguata and Ostertag, 2002),
and provide a diverse vertical structure and canopy
cover necessary for arboreal faunal species to
colonized the restored site (McClanahan and Wolfe,
1993; Passell, 2000; George and Zack, 2001; DeWalt
et al., 2003). Moreover the increase in canopy cover
decreased herbaceous cover, and the presence of
pioneer species with short-lived leaves contributed to
the increase in litter cover and the increase in the
number of litter layers in the reforested site. The
recovery of these vegetation structure measures have
changed conditions in reforested site and have
facilitated the colonization of plants and animals
(species diversity) and improved nutrient cycling
(ecosystem processes).
4.2. Species diversity
The increase in woody seedlings and ant diversity
and abundance was associated with changes in ground
cover and number of litter layers. Woody seedling
recruitment was probably influenced by the decrease
in herbaceous cover in the reforested site, due to a
decrease in belowground root competition and above
ground physical barriers (Horvitz and Schemske,
1994; Otsamo, 2000; Hooper et al., 2002). Moreover,
both the pre-reforested and reforested sites were very
close to intact forest (i.e. less than 15 m), thus seed
dispersal limitation should not be a limiting factor
(Cubin
˜a and Aide, 2001; Holl, 1999). Similarly, ants
responded to the increase in litter cover and number of
M.C. Ruiz-Jae
´n, T.M. Aide/ Forest Ecology and Management 218 (2005) 159–173 167
Table 3
Percent of restoration success for measures of vegetation structure
species diversity, and ecosystem processes using Bray Curtis Ordi-
nations
Measures Recovery (%)
Vegetation structure
Woody vegetation height 74
DBH size classes 66
Herbaceous cover 65
Number of litter layers 56
Litter cover 46
Average 61
Species diversity
Ants 76
Herpetofauna 68
Woody seedlings 54
Average 66
Ecosystem processes
Litter production 70
Litter turnover rates 60
Nutrient inputs 57
Soil nutrient content (0–10 cm profile) 49
Bulk density 44
Average 56
The values indicate the position of the reforested site in comparison
with the pre-forested and reference sites, as endpoints.
litter layers in the reforested site. Ant composition has
been positively correlated with litter cover (Andersen,
1993), litter depth (Carvalho and Vasconcelos, 1999),
and litter biomass (Barberena-Arias and Aide, 2003),
which can explain the similar species composition in
the reforested and reference sites.
The herpetofauna responded positively to the
establishment of woody vegetation in the reforested
site with the colonization of arboreal species (e.g.,
Anolis cristatellus,Eleutherodactylus coqui, and E.
cochranae). This increase in species richness as well
as species abundance (Fig. 2c) have been reported
elsewhere (Pearman, 1997; Fogarty and Vilella, 2003;
Jellinek et al., 2004). Moreover, an increase in
abundance of both amphibians and reptiles in the
reforested site to levels similar to the reference site,
explains the colonization of two predators, Anolis
cuvieri and Epicrates inornatus, which depended on a
high abundance of prey.
In contrast to the herpetofauna, birds have not
responded as rapidly to the changes in vegetation
structure in the reforested site. This suggests that the
reforested site does not provide the appropriate
structural characteristics found in the surrounding
matrix. The three species found in the reforested site are
common species, which are present in many habitats in
Puerto Rico and included two insectivores (Coereba
flaveola and Dendroica adelaide) and a predator
(Coccyzus minor)(Raffaele et al., 1998; Oberle,
2003). The presence of these species suggests that
there was enough prey for insectivores and predator
species that are not habitat restricted, but therewere not
enough resources for frugivorous birds (Fig. 2b and c).
In addition to the planting effect, the short distance
to the forested limestone hills has also contributed to
the rapid recovery of species diversity. An increase in
vegetation structure enhanced the recovery of ants,
reptiles, amphibians, and woody seedlings by provid-
ing the appropriate habitat (e.g., microclimate).
Nevertheless, the rapid colonization of these groups
would not have occurred without the presence of
propagules and fauna in the surrounding limestone
hills that served as a source for the reforested site
(Hanski, 2002). Although the proximity to the forested
hills has assisted the recovery process, the poor
colonization in the pre-reforested site demonstrates
the importance of restoration (i.e. planted woody
species).
4.3. Ecosystem processes
Planting, growth, and colonization of woody plants
have enhanced aboveground ecosystem processes in
the reforested site. This is mainly due to the high litter
production of pioneer species that allocate their
energy to the leaf production and have short-lived
leaves (Brown and Lugo, 1990; Guariguata and
Ostertag, 2001). Moreover, litter turnover (k) in the
reforested site was two times faster in comparison with
the pre-reforested site. The litter turnover rates in the
reforested and reference sites are comparable with
other tropical ecosystems (Olson, 1963; Wieder and
Wright, 1995). In addition, the increase in litterfall
explained most of the difference in nutrient inputs
among sites (Burghouts et al., 1998; Herbohn and
Congdon, 1998).
Changes in aboveground ecosystem processes
contributed to the changes in belowground processes
in the reforested site. There was an increase in
phosphorus and nitrogen soil content, but there was no
change in calcium content. The increase in phosphorus
and nitrogen content in the soil is not surprising,
because these nutrients are mainly influenced by plant
inputs (Vitousek, 1984). Soil calcium content was not
influenced by plant inputs as suggested by Vitousek
(1984). In our sites, the geology (i.e., Karst) has a
much stronger influence, resulting in no difference
among sites. The change in plant composition from C
4
to C
3
contributed to changes in the composition of
SOM in the 0–10 cm profile, while no difference was
detected in 10–20 cm profile in the reforested site. The
rapid incorporation of SOM in the topsoil can be
explained by the rapid increase in litter production and
decomposition in the reforested site. This rapid change
in d
13
C values (18.4%in the pre-reforested to
22.7%in the reforested site) suggests that incor-
poration of soil organic matter can be more rapid than
previously reported. d
13
C values in studies of natural
regeneration from pastures (C
4
) to forest (C
3
) have
taken more than 15 years to reach levels similar to
those in the 3-year-old reforested site (Rhoades et al.,
1998; Guillet et al., 2001; Eshetu, 2002; Biedenbender
et al., 2004). In contrast to the topsoil, C composition
in the 10–20 cm profile is still dominated by C
4
in the
reforested site (Table 2). This can be partially
explained by the high bulk density (i.e., high soil
compaction) and the absence of earthworms in the 10–
M.C. Ruiz-Jae
´n, T.M. Aide/ Forest Ecology and Management 218 (2005) 159–173168
20 cm soil profile (Table 2 and Knoepp et al., 2000).
High bulk densities reduced earthworm activity that
can be critical for the movement of organic matter in
the soil (Herrick, 2000; Tian et al., 1997).
4.4. Bray Curtis Ordination
The Subjective Bray Curtis Ordination is a useful
approach for assessing restoration success (Table 3).
Practitioners can predefined minimum percent of
success (e.g., 70%) for the project as a whole or for
specific measures. This technique can be useful for
comparing large number of measurements and
identifying measures that are recovering slowly and
would benefit from additional management. Although
this approach is promising as a simple way to present
restoration success, it has limitations: (1) when values
in the pre-reforested site are zero, only a slight
recovery will be classified as a high recovery, and (2)
when values of the reforested site are lower than the
pre-reforested site or higher than the reference site, the
analysis will give negative or >100% recovery.
5. Conclusion
This restoration project has initiated rapid changes
in vegetation structure, species diversity, and ecosys-
tem processes, but has restoration been successful?
The rapid increase of vertical stratification, low
herbaceous cover, rapid colonization of species from
different trophic levels, high litter production, and
rapid C
3
incorporation in SOM suggest that the
reforested site could be left without further manage-
ment assistance. These results indicate that the
restoration project has been successful. Other
measures such as litter cover, bird diversity, litter
turnover, nutrient inputs, and bulk density will take
longer to recover. To accelerate the recovery rate of
these measures management efforts could focus on
planting higher densities of pioneer species in the first
years of restoration to assure an increase in litter cover
and for providing resources for birds. The recovery of
the bird community could be enhanced by planting
pioneer species (e.g., Cecropia schreberiana,Cordia
sulcata,Miconia serrulata,andSchefflera moroto-
toni) that will offer both vertical vegetation structure
and food resource in the short-term (Carlo et al.,
2003).
Acknowledgments
This project was funded by the Biology Depart-
ment Graduate Program of the University of Puerto
Rico, NASA-IRA, DEGI, and the Laboratory of Plant
Ecophysiology of the Venezuelan Institute of Scien-
tific Research. Neftali Rios-Lopez and Miguel A.
Acevedo provided the herpetofauna and bird census
data, respectively. The d
13
C values were analysed in
The Stable Isotope Ratio Facility for Environmental
Research (SIRFER) at the University of Utah. The
comments of Jim Ebersole and Carolina Monmany
improved the manuscript.
M.C. Ruiz-Jae
´n, T.M. Aide/ Forest Ecology and Management 218 (2005) 159–173 169
Appendix A
Presence/abscence of woody seedlings, ants, amphibians, reptiles, and bird species in the pre-reforested,
reforested, and reference sites
Group/scientific name Pre-reforested Reforested Reference
Woody seedlings
Asteraceae spl X
Urena lobata L. X X
Urena sinuata L. X X
Spathodea campalunata Beauv. X X X
Andira inermis (Wright) DC. X
Ardisia obovata Desv. ex Hamilt. X
Asteraceae sp2 X
Bignoniaceae spl X
M.C. Ruiz-Jae
´n, T.M. Aide/ Forest Ecology and Management 218 (2005) 159–173170
Casearia arborea (L.C. Rich.) Urban X
Casearia sylvestris Sw. X
Eugenia monticola (Sw.) DC. X
Senna siamea (Lam.) Irwin and Barnaby X
Tabebuia heterophylla (DC) Britton
a
X
Terminalia catappa L. X
Thespesia grandiflora (DC.) Urban
a
X
Casearia guianensis (Aublet) Urban X X
Chrysophyllum argenteum Jacq. X X
Dendropanax arboreus (L.) Decne. and Planch. X X
Guarea guidonea L. XX
Hippocratea volubilis L. X X
Miconia laevigata (L.) DC. X X
Quararibea turbinata (Sw.) Poir X X
Woody seedlings
Thouinia striata Radlk. X X
Bambusa vulgar is Schrad ex Wendl. X
Eugenia biflora (L.) DC X
Faramea Occident alis (L.) A. Rich. X
Ocotea leucoxylum (Sw.) Mez. X
Phyllanthus juglandifolius Wild. X
Syzygium jambos (L.) Alst. X
Trichilia pallida Sw. X
Ants
Cardiocondyla emergi Forel X
Pheidole fallax Mayr X
Pheidole sp6 X
Brachymyrmex spl X X X
Hypoponera opaciceps Mayr X X X
Monomorium ebeninum Forel X X X
Odontomachus ruginodis Smith X X X
Pheidole morens Wheeler X X X
Pheidole sp2 X X X
Solenopsis corticalis Forel X X X
Solenopsis geminata Fabricius X X X
Strumygenys rogeri Emery X X X
Wasmannia auropuctata Roger X X X
Solenopsis wagneri Santschi. X X
Wasmannia spl X X
Ants
Camponotus sexguttatus Frabricius X
Odontomachus spl X
Pheidole sp5 X
Strumygenys emmae Emery X
Anochetus mayri Emery X X
Mycocepurus smithi Forel X X
Odontomachus bauri Emery X X
Paratrechina longicornis Latreille X X
Pheidole spl XX
Pheidole sp4 XX
Pyr arnica mar gar it ae Forel X X
Cyphonomyrmex minutus Mayr X
Appendix A (Continued )
Group/scientific name Pre-reforested Reforested Reference
References
Agosti, D., Alonso, L.E., 2000. The ALL protocol: a standard
protocol for the collection of ground-dwelling ants. In: Agosti,
D., Majer, J.D., Alonso, L.E., Schultz, T.R. (Eds.), Ants: Stan-
dard Methods for Measuring and Monitoring Diversity. Smith-
sonian Institution Press, Washington, DC, pp. 204–206.
Alvarez-Ruiz, M., Acevedo-Rodriguez, P., Vazquez, M., 1997.
Quantitative description of the structure and diversity of the
vegetation in the limestone forest of Rio Abajo, Arecibo-
Utuado, Puerto Rico. Acta Cient. 11, 21–66.
Andersen, A.N., 1993. Ants as indicators of restoration success at an
uranium mine in tropical Australia. Restor. Ecol. 1, 156–
167.
Anderson, J.M., Ingram, J.S., 1993. Tropical Soil Biology and
Fertility: A Handbook of Methods, 2nd ed. CAB International,
Oxon.
Barberena-Arias, M.F., Aide, T.M., 2003. Species diversity and
trophic composition of litter insects during plant secondary
succession. Caribbean J. Sci. 39, 161–169.
Bibby, C.J., Burgess, N.D., Hill, D.A., Mustoe, S.H., 2000. Bird
Census Techniques, 2nd ed. Academic Press, London.
Biedenbender, S.H., McClaran, M.P., Quade, J., Weltz, M.A., 2004.
Landscape patterns of vegetation change indicated by soil
carbon isotope composition. Geoderma 119, 69–83.
Bray, J.R., Curtis, J.T., 1957. An ordination of the upland forest
communities of southern Wisconsin. Ecol. Monogr. 27, 325–
349.
M.C. Ruiz-Jae
´n, T.M. Aide/ Forest Ecology and Management 218 (2005) 159–173 171
Dolichoderinae X
Hypoponera puntatissima Roger X
Linepithema melleum Wheeler X
Monomorium floricola Jerdon X
Paratrechina steinheili Forel X
Pheidole exigua Mayr X
Plachycondyla spl X
Rogeria foreli Emery X
Solenopsis spl X
Strumygenys spl X
Amphibians
Bufo marinus L. X
Leptodactylus albilabris Gunther X X X
Eleutherodactylus coqui Thomas X X
Eleutherodactylus cochranae Grant X X
Eleutherodactylus antillensis Reinhardt and Lutken X
Reptiles
Anolis pulchellus Dumeril and Bribon X
Anolis krugii Peters XX
Anolis cristatellus Dumeril and Bribon X X
Anolis cuvieri Merrem X X
Epicrates inornatus Reinhardt X X
Anolis stratulus Cope X
Birds
Estrilda melpoda Vieillot X
Coccyzus minor G.K. Gmelin X
Coereba flaveola L. XX
Dendroica adelaidae Baird X X
Melanerpes portoricensis Daudin X
Otus nudipes Daudin X
Saurothera vieilloti Bonaparte X
Spindalis puertoricensis Bryant X
Todus mexicanus Lesson X
Turdus plumbeus L. X
Zenaida aurita Temminck X
a
Species planted and producing fruits.
Appendix A (Continued )
Group/scientific name Pre-reforested Reforested Reference
Brown, S., Lugo, A.E., 1990. Tropical secondary forests. J. Trop.
Ecol. 6, 1–32.
Burghouts, T.B.A., van Straalen, N.M., Bruijnzeel, L.A., 1998.
Spatial heterogeneity of element and litter turnover in a Bornean
rain forest. J. Trop. Ecol. 14, 477–506.
Carlo, T.A., Collazo, J.A., Groom, M.J., 2003. Avian fruit preferences
across a Puerto Rican forested landscape: pattern consistency and
implications for seed removal. Oecologia 134, 119–131.
Carvalho, K.S., Vasconcelos, H.L., 1999. Forest fragmentation in
central Amazonia and its effects on litter-dwelling ants. Biol.
Conserv. 91, 151–157.
Cubin
˜a, A., Aide, T.M., 2001. The effect of distance from forest
edge on seed rain and soil seed bank in a tropical pasture.
Biotropica 33, 260–267.
DeWalt, S.J., Maliakal, S.K., Denslow, J.S., 2003. Changes in
vegetation structure and composition along a tropical forest
chronosequence: Implications for wildlife. For. Ecol. Manage.
182, 139–151.
Dorren, L.K.A., Berger, F., Imeson, A.C., Maier, B., Rey, F., 2004.
Integrity, stability and management of protection forests in the
European Alps. For. Ecol. Manage. 195, 165–176.
Elmqvist, T., Folke, C., Nystrom, M., Peterson, G., Bengtsson, J.,
Walker, B., Norberg, J., 2003. Response diversity, ecosystem
change, and resilience. Front. Ecol. Environ. 1, 488–494.
Eshetu, Z., 2002. Historical C
3
-C
4
vegetation pattern on forested
mountain slopes: its implications for ecological rehabilitation of
degraded highlands of Etiopia by afforestation. J. Trop. Ecol. 18,
743–758.
Eusse, A.M., Aide, T.M., 1999. Patterns of litter production across a
salinity gradient in a Pterocarpus officinalis tropical wetland.
Plant Ecol. 145, 307–315.
Feinsinger, P., 2001. Designing Field Studies for Biodiversity Con-
servation. The Nature Conservancy and Island Press, Washing-
ton, DC.
Fogarty, J.H., Vilella, F.J., 2003. Use of native forest and eucalyptus
plantations by Eleutherodactylus frogs. J. Wildlife Manage. 67,
186–195.
George, T.L., Zack, S., 2001. Spatial and temporal considerations in
restoring habitat for wildlife. Restor. Ecol. 9, 272–279.
Guariguata, M.R., Ostertag, R., 2001. Neotropical secondary forest
succession: changes in structural and functional characteristics.
For. Ecol. Manage. 148, 185–206.
Guariguata, M.R., Ostertag, R., 2002. Sucesio
´n primaria. In: Guar-
iguata, M.R., Kattan, G.H. (Eds.), Ecologia y Conservacio
´nde
Bosques Neotropicales. Libro Universitario Regional, CR, pp.
591–623.
Guillet, B., Achoundong, G., Happi, J.Y., Kabeyene-Beyala, V.K.,
Bonvallot, J., Riera, B., Mariotti, A., Schwartz, D., 2001.
Agreement between floristic and soil organic carbon isotope
(
13
C/
12
C,
14
C) indicators of forest invasion of savannas during
the last century in Cameroon. J. Trop. Ecol. 17, 809–832.
Hanski, I., 2002. Metapopulation Ecology. Oxford University Press
Inc., New York.
Herbohn, J.L., Congdon, R.A., 1998. Ecosystem dynamics at dis-
turbed and undisturbed sites in North Queensland wet tropical
rain forest III. Nutrient returns to the forest floor through
litterfall. J. Trop. Ecol. 14, 217–229.
Herrick, J.E., 2000. Soil quality: indicator of sustainable land
management? Appl. Soil Ecol. 15, 75–83.
Hobbs, R.J., Harris, J.A., 2001. Restoration ecology: repairing the
earth’s ecosystems in the new millennium. Restor. Ecol. 9, 239–
246.
Hobbs, R.J., Norton, D.A., 1996. Towards a conceptual framework
for restoration ecology. Restor. Ecol. 4, 93–110.
Holdridge, L.R., 1967. Life Zone Ecology. Tropical Science Centre,
San Jose, CR.
Holl, K.D., 1999. Factors limiting tropical moist forest regeneration
in agricultural land: soil, microclimate, vegetation, and seed
rain. Biotropica 31, 229–242.
Hooper, E., Condit, R., Legendre, P., 2002. Responses of 20 native
tree species to reforestation strategies for abandoned farmland in
Panama. Ecol. Appl. 12, 1626–1641.
Horvitz, C.C., Schemske, D.W., 1994. Effects of dispersers, gaps
and predators on dormancy and seedling emergence in a tropical
herb. Ecology 75, 1949–1958.
Jackson, M.L., 1968. Ana
´lisis Quı
´mico de Suelos, 1st ed. Editorial
Omega, Barcelona.
Jellinek, S., Driscoll, D.A., Kirkpatrick, J.B., 2004. Environ-
mental and vegetation variables have a greater influence
than habitat fragmentation in structuring lizard commu-
nities in remnant urban bushland. Aust. Ecol. 29, 294–
304.
Jobba
´gy, E.G., Jackson, R.B., 2000. The vertical distribution of soil
organic carbon and its relation to climate and vegetation. Ecol.
Appl. 10, 423–436.
Jones, E.R., Wishnie, M.H., Deago, J., Sautu, A., Cerezo, A., 2004.
Facilitating natural regeneration in Saccharum spontaneum (L.)
grasslands within the Panama Canal Watershed: effects of tree
species and tree structure on vegetation recruitment patterns.
For. Ecol. Manage. 191, 171–183.
Knoepp, J.D., Coleman, D.C., Crossley Jr., D.A., Clark, J.S., 2000.
Biological indices of soil quality: an ecosystem case study of
their use. For. Ecol. Manage. 138, 357–368.
Longino, J.T., 2000. What to do with the data? In: Agosti, D., Majer,
J.D., Alonso, L.E., Schultz, T.R. (Eds.), Ants: Standard Methods
for Measuring and Monitoring Diversity. Smithsonian Institu-
tion Press, Washington, DC, pp. 186–203.
Martin, A., Mariotti, A., Balesdent, J., Lavelle, P., Vuattoux, R.,
1990. Estimate of organic matter turnover rate in a savanna soil
by
13
C natural abundance measurements. Soil Biol. Biochem.
22, 517–524.
McClanahan, T.R., Wolfe, R.W., 1993. Acceleration forest succes-
sion in a fragmented landscape: the role of birds and perches.
Conserv. Biol. 7, 279–288.
McCune, B., Mefford, M.J., 1999. Multivariate Analysis of Ecolo-
gical Data Version 4.25. MjM Software, Oregon.
McCune, B., Grace, J.B., 2002. Analysis of Ecological Commu-
nities. MjM Software Design, USA.
Murphy, J., Riley, J.P., 1962. A modified single solution method for
the determination of phosphate in natural waters. Anal. Chim.
Acta 27, 31–36.
Nichols, O.G., Nichols, F.M., 2003. Long-term trends in faunal
recolonization after bauxite mining in the jarrah forest of south-
western Australia. Restor. Ecol. 11, 261–272.
M.C. Ruiz-Jae
´n, T.M. Aide/ Forest Ecology and Management 218 (2005) 159–173172
Oberle, M.W., 2003. Las aves de Puerto Rico en fotografias.
Editorial Humanitas, Washington, DC.
Olson, J.E., 1963. Energy storage and the balance of producers and
decomposers in ecological systems. Ecology 44, 322–331.
Olsen, S.R., Sommers, L.E., 1982. In: Page, A.L., Miller, R.H.,
Keeney, D.R. (Eds.), Methods of Soil Analysis, Part 2 Chemical
and Microbiological Properties. 2nd ed. American Society of
Agronomy, Inc., SSSA, USA, pp. 416–418.
Otsamo, R., 2000. Secondary forest regeneration under fast-growing
forest plantations on degraded Imperata cylindrica grasslands.
New Forest 19, 69–93.
Parmenter, R.R., MacMahon, J.A., 1992. Faunal community devel-
opment on disturbed lands: an indicator of reclamation success.
In: Chambers, J.C., Wade, G.L., (Eds.), Evaluating reclamation
success: the ecological considerations. General Technical
Report NE 164, USDA Forest Service, Northeastern Forest
Experimental Station, Pennsylvania, pp. 73–89.
Passell, H.D., 2000. Recovery of bird species in minimally restored
Indonesian thin strips mines. Restor. Ecol. 8, 112–118.
Pearman, P.B., 1997. Correlates of amphibians diversity in an altered
landscape of Amazonian Ecuador. Conserv. Biol. 11, 1211–
1225.
Peterson, G., Allen, C.R., Holling, C.S., 1998. Ecological resilience,
biodiversity, and scale. Ecosystems 1, 6–18.
Purcell, A.H., Friedrich, C., Resh, V.H., 2002. An assessment of a
small urban stream restoration project in northern California.
Restor. Ecol. 10, 685–694.
Raffaele, H., Wiley, J., Garrido, O., Keith, A., Raffaele, J., 1998. A
Guide to the Birds of the West Indies. Princeton University
Press, New Jersey.
Rhoades, C.C., Eckert, G.E., Coleman, D.C., 1998. Effect of
pastures trees on soil nitrogen and organic matter: implications
for tropical montane forest restoration. Restor. Ecol. 6, 262–270.
Rivera, L.W., Aide, T.M., 1998. Forest recovery in the Karst region
of Puerto Rico. For. Ecol. Manage. 108, 63–75.
Ruiz-Jae
´n, M.C., Aide, T.M., 2005. Restoration Success: how is it
being measured? Restor. Ecol. 13 (3).
Silver, W.L., Kueppers, L.M., Lugo, A.E., Ostertag, R., Matzek, V.,
2004. Carbon sequestration and plant community dynamics
following reforestation of tropical pasture. Ecol. Appl. 14,
1115–1127.
Smith, B.N., Epstein, S., 1971. Two categories of
13
C/
12
C ratios for
higher plants. Plant Physiol. 47, 380–384.
Society for Ecological Restoration International Science & Policy
Working Group, 2004. The SER International Primer on Eco-
logical Restoration. Society for Ecological Restoration Interna-
tional, Tucson www.ser.org.
Thomas, G.W., 1982. In: Page, A.L., Miller, R.H., Keeney, D.R.
(Eds.), Exchangeable Cations. Methods of Soil Analysis,
Part 2 Chemical and Microbiological Properties. 2nd ed.
American Society of Agronomy, Inc., SSSA, USA, pp. 159–
165.
Tian, G., Kang, B.T., Brussaard, L., 1997. Effect of mulch quality on
earthworm activity and nutrient supply in the humid tropics. Soil
Biol. Biochem. 29, 369–373.
Trouve, C., Mariotti, A., Schwartz, D., Guillet, B., 1994. Soil
organic carbon dynamics under Eucalyptus and Pinus planted
on savannas in the Congo. Soil Biol. Biochem. 26, 287–
295.
Vasconcelos, H.L., Vilhena, J.M.S., Caliri, G.J.A., 2000. Responses
of ants to selective logging of a central Amazonian forest. J.
Appl. Ecol. 37, 308–514.
Vitousek, P.R., 1984. Litterfall, nutrient cycling, and nutrient limita-
tion in tropical forests. Ecology 65, 285–298.
Wang, J., Borsboom, A.C., Smith, G.C., 2004. Flora diversity of
farm forestry plantations in southeast Queensland. Ecol. Man-
age. Restor. 5, 43–51.
White, P.S., Walker, J.L., 1997. Approximating nature’s variation:
selecting and using reference information in restoration ecology.
Restor. Ecol. 5, 338–349.
Wieder, R.K., Wright, S.J., 1995. Tropical forest litter dynamics and
dry season irrigation on Barro Colorado Island. Panama. Ecol.
76, 1971–1979.
Young, T.P., 2000. Restoration ecology and conservation biology.
Biol. Conserv. 92, 73–83.
Zou, X., Gonzalez, G., 1997. Changes in earthworm density and
community structure during secondary succession in abandoned
tropical pastures. Soil Biol. Biochem. 29, 627–629.
M.C. Ruiz-Jae
´n, T.M. Aide/ Forest Ecology and Management 218 (2005) 159–173 173
... The N nutrient returned to the forest floor from nitrogen-fixing tree plantations was significantly higher than that from a nonnitrogen-fixing tree plantations and for other nutrients (Bernhard-Reversat 1996). The litter decomposition of nitrogen-fixing trees positively affect the acceleration of nutrient release and nutrient deposition to promote vegetation development and improve soil properties (Lugo 1992;Ruiz-Jaén & Aide 2005;Lanuza et al. 2018), suggesting that A. mangium plantation should be used for improving nutrient cycling in mining reclamation. Nutrient cycling is a key process for mining rehabilitation, relating to vegetation succession and soil development in the mining area. ...
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Litterfall and litter decomposition play important roles in the maintenance of nutrient cycling and rehabilitation of degraded lands. Litterfall, litter decomposition and nutrient return were investigated in a 27-year-old Acacia mangium plantation on sandy and clay sites, and in a mixed plantation at the Phangnga Forestry Research Station, Phangnga Province, Thailand. Additionally, secondary and primary forests were investigated and compared with the values obtained from the Acacia mangium and the mixed plantations. The results indicated that litter production in A. mangium plantation on sandy and clay sites, and in mixed plantations (15.47, 11.68 and 7.89 t/ha/yr, respectively) was higher than that in the secondary and primary forests (6.34 and 6.92 t/ha/yr, respectively). The rate of litter decomposition was the greatest in the secondary forest (3.01/yr) and the lowest occurred in the primary forest (1.15/yr). The decomposition rate of the mixed leaf litter between native trees and A. mangium in plantations was higher than that of only A mangium leaf, except in the mixed plantations. A high initial nitrogen concentration in A. mangium could accelerate litter decomposition and improve litter quality in the mixed litter. In addition, the nutrient return in plantations was higher than that in the secondary and primary forests, especially for N. Increased litter production, high decomposition rate and nutrient return from A. mangium plantation had important roles in nutrient cycling, suggesting that a mixed plantation consisting of A. mangium and native trees should be considered for the reclamation of mining land.
... Monitoring also ensures accountability for goals set at the beginning of the restoration process (Hobbs 2007, Suding 2011. Restoration practitioners typically use measures of vegetation to assess the quality of restored sites (Ruiz-Jaen & Aide 2005). Plant community composition, species richness, and percent non-native species cover are components of plant communities that are often incorporated into restoration monitoring (Houlahan et al., 2006). ...
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Monitoring is essential to restoration, but the standard metrics used to monitor wetland restoration do not explicitly account for function in plant communities. Functional traits may be a useful addition to the wetland monitoring toolkit, because they can represent aspects of ecosystem functioning that standard metrics may not. Our objective was to determine how abiotic factors that influence wetland community composition and structure relate to both functional leaf traits and standard vegetation monitoring metrics to determine if functional leaf traits could add a functional component to wetland monitoring. We surveyed 66 100-m² plots in 22 floodplain wetlands in Illinois that were restored between 1997 and 2010. We used plant species data to calculate the mean coefficient of conservatism (mean C) and richness, and collected leaves to determine community weighted means of specific leaf area (SLA) and leaf dry matter content (LDMC) at each plot. Hydrologic data were used to calculate variables related to frequency, depth, and duration of inundation, and soil samples were collected to determine soil pH, organic matter content, and nitrogen and phosphorous content at each plot. We used structural equation models to understand how predictor variables (hydrological variables, soil variables, canopy cover, time since restoration, and latitude) influenced each other, and ultimately how they influenced response variables (mean C, richness, percent non-native species cover, SLA, and LDMC). LDMC and SLA were poorly explained by predictor variables and had relatively few significant relationships within models. Predictor variables best explained variance in mean C, followed by percent non-native cover, richness, then LDMC and SLA. Mean C was positively influenced by canopy cover and negatively influenced by soil fertility, whereas richness was negatively influenced by latitude. There was a strong latitudinal gradient of species richness from north to south, such that southern plots were significantly more diverse and had less cover by non-native species than northern plots. Our research suggests that SLA and LDMC may not be well suited for use as wetland restoration monitoring tools in Midwestern floodplain wetlands, and that monitoring tools already in place may sufficiently reflect abiotic conditions.
... Relevant research on early vegetation diversity focused on the variation in climate and environmental conditions on vegetation types, which focused on the effects of climatic variables on different vegetation types [10][11][12]. Examination of the relationship between vegetation richness/diversity and ecosystems has continued [13][14][15]. The coexistence of different types of vegetation on Earth indicates that the formal structure of vegetation is more complex and has different degrees of impact on the stability and resilience of the regional ecology [16]. ...
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Plants play an important role in the storage of organic carbon as a carbon reservoir, mainly reflected in the absorption of carbon dioxide and the release of oxygen by plants through photosynthesis, which plays an important role in regulating regional carbon balance. This study aimed to explore the pattern of dominant vegetation species diversity in small basin ecosystems, analyze the photosynthetic carbon sequestration characteristics of dominant vegetation and its carbon sink potential, and use the Hancang River Basin as the research area to obtain species abundance data using the sampling method. The community characteristics were analyzed from the aspects of species richness, α diversity index, dominant population spatial pattern, and interspecific correlation. Fifteen typical plants were selected; the photosynthetic carbon sequestration capacity of the plants was measured and quantified, and the correlation of the nitrogen ecological measurement characteristics of the leaves were analyzed. The results showed that all 15 dominant vegetation types were clustered, and the aggregation intensity and interspecific correlation were closely related. Spearman’s analysis showed that the dominant vegetation community structure and diversity in the Hancang River Basin were unstable, and there may be strong interspecies competition and frequent species replacement in the future succession process. The daily trend of the net photosynthetic rate of dominant vegetation tended to be bimodal or nearly bimodal, and the peak generally appeared at 10:00 or 14:00. The leaf area index of the dominant species was greater than 3.0, of which the largest was Platanus acerifolia at 5.31. The smallest was Hemerocallis fulva (L.) L., which was 1.16. The average carbon sequestration per unit leaf area and unit land area of the dominant species was 10.02 g·m−2·d−1 and 42.35 g·m−2·d−1, respectively, and the average oxygen release was 7.29 g·m−2·d−1 and 30.8 g·m−2·d−1, respectively. The average maintenance breathing volume of the leaves was 2.94 μmol CO2 m−2s−1. Comprehensive analysis of the ecological restoration process of the Hancang River Basin should focus on Platanus acerifolia, Prunus serrulata var. lannesiana, Prunus cerasifera, Ligustrum lucidum, and other highly efficient carbon sequestration plants, to carry out rational planting and to build a multilayer composite configuration of forest shrub and grass models to improve the ecological problems of the basin.
... These indicators are important as the modification of micro-climate and woody-grass competition through bush clearing is expected to improve these indicators (Stephens et al. 2016). However, according to Ruiz-Jaen and Aide (2005), assessing herbaceous restoration based only on vegetation indicators is not adequate. Thus, the whole vegetation recovery process and its potential drivers should be assessed holistically in order to inform appropriate future restoration applications (Ruwanza et al. 2013). ...
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Bush thickening reduces herbaceous species diversity and forage production, leading to a decline in grazing capacity. Hence, altering the micro-climate and woody-grass competition in bush-encroached rangelands using bush clearing is instrumental for the restoration of herbaceous vegetation. The short-term recovery potential of herbaceous species composition, basal cover and biomass production was assessed following bush clearing at Makapaanstad, North West province, South Africa. The mean density was 6908 woody plants ha −1 , with Vachellia tenuispina encroaching at 6301 plants ha −1. Paired observations were conducted in bush-cleared sites (where woody plants were cut using saws, with stumps treated with picloram at 1% concentration) and uncleared sites (where the woody cover was left intact). Forty-nine different herbaceous species were recorded, with grasses being more abundant in cleared (62%) than in uncleared sites (50%). Herbaceous species showed interspecific responses, with Brachiaria eruciformis and Panicum coloratum colonising in cleared sites, whereas the frequency of other species remained comparable between cleared and uncleared sites. Graminoids, largely bunch grasses, had a significantly greater basal cover in cleared (496 cm 2 m −2) compared to uncleared (301 cm 2 m −2) sites. Likewise, total biomass production was significantly (p = 0.05) higher in cleared (760 kg DM ha −1) than in uncleared (636 kg DM ha −1) sites. The principal component analysis (PCA) showed that most herbaceous species were associated with soils having high total nitrogen and nitrite. Seedlings of forbs (r 2 = 0.81), grasses (r 2 = 0.97) and succulents (r 2 = 0.98) were highly associated with nitrate-rich soils. Our results revealed that even at a short term, bush clearing is important for the restoration of herbaceous vegetation.
... These indicators are important as the modification of micro-climate and woody-grass competition through bush clearing is expected to improve these indicators (Stephens et al. 2016). However, according to Ruiz-Jaen and Aide (2005), assessing herbaceous restoration based only on vegetation indicators is not adequate. Thus, the whole vegetation recovery process and its potential drivers should be assessed holistically in order to inform appropriate future restoration applications (Ruwanza et al. 2013). ...
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Bush thickening reduces herbaceous species diversity and forage production, leading to a decline in grazing capacity. Hence, altering the micro-climate and woody-grass competition in bush-encroached rangelands using bush clearing is instrumental for the restoration of herbaceous vegetation. The short-term recovery potential of herbaceous species composition, basal cover and biomass production was assessed following bush clearing at Makapaanstad, North West province, South Africa. The mean density was 6908 woody plants ha ⁻¹ , with Vachellia tenuispina encroaching at 6301 plants ha ⁻¹ . Paired observations were conducted in bush-cleared sites (where woody plants were cut using saws, with stumps treated with picloram at 1% concentration) and uncleared sites (where the woody cover was left intact). Forty-nine different herbaceous species were recorded, with grasses being more abundant in cleared (62%) than in uncleared sites (50%). Herbaceous species showed interspecific responses, with Brachiaria eruciformis and Panicum coloratum colonising in cleared sites, whereas the frequency of other species remained comparable between cleared and uncleared sites. Graminoids, largely bunch grasses, had a significantly greater basal cover in cleared (496 cm ² m ⁻² ) compared to uncleared (301 cm ² m ⁻² ) sites. Likewise, total biomass production was significantly ( p = 0.05) higher in cleared (760 kg DM ha ⁻¹ ) than in uncleared (636 kg DM ha ⁻¹ ) sites. The principal component analysis (PCA) showed that most herbaceous species were associated with soils having high total nitrogen and nitrite. Seedlings of forbs ( r ² = 0.81), grasses ( r ² = 0.97) and succulents ( r ² = 0.98) were highly associated with nitrate-rich soils. Our results revealed that even at a short term, bush clearing is important for the restoration of herbaceous vegetation.
... This regulation covers conservation partnerships for ecosystem restoration and community development. Ecosystem restoration is expected to improve forest biodiversity based on vegetation structure, species (flora and fauna diversity), and ecosystem processes such as litterfall production and soil nutrients [7]. In addition, the focus of ecosystem restoration has also shifted in recent years, not only to increase the number and enrichment of species and habitat improvement, but also to be associated with environmental services provided by ecosystems [8]. ...
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Synthèse bibliographique des connaissances en matière de réhabilitations minières de site primaire à ciel ouvert (préférentiellement aurifère) dans le monde et en particulier celles réalisées dans des conditions comparables à celles de la Guyane française.
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(13)C/^(12)C ratios have been determined for plant tissue from 104 species representing 60 families. Higher plants fall into two categories, those with low δ_(PDB1) ^(13)C values (-24 to -34‰) and those with high δ ^(13)C values (-6 to -19‰). Algae have δ^(13)C values of -12 to -23‰. Photosynthetic fractionation leading to such values is discussed.