Location Near Grass Patches Influences Establishment of
Native Woody Species in a Puerto Rican Subtropical Dry
Forest
Juan Gilberto García-Cancel and Jarrod M. Thaxton
Caribbean Naturalist, No. 51
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Caribbean Naturalist
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J.G. García-Cancel and J.M. Thaxton
22001188 CARIBBEAN NATURALIST No. 51N:1o–. 1531
Location Near Grass Patches Influences Establishment of
Native Woody Species in a Puerto Rican Subtr opical Dry
Forest
Juan Gilberto García-Cancel1,2,* and Jarrod M. Thaxton3
Abstract - Invasive grasses can influence the diversity and survival of native species by
changing disturbance regimes, e.g., promoting fire and altering nutrient and resource fluxes.
The goal of this study was to assess how non-native and native grasses influence the establishment
of native woody seedlings in the Guánica subtropical dry forest, Puerto Rico.
This forest has been experiencing the influence of introduced non-native plants, altered
disturbance regimes, and resultant changes in the forest community over the past decades.
We selected 2 field sites with the non-native grass Megathyrsus maximus (Guinea Grass)
or the native grass Uniola virgata (Limestone Grass). At each site, we randomly selected
20 grass clumps and 20 adjacent bare soil patches in which to transplant seedlings of 3
native woody species. We recorded seedling survival over a 23-month study period, with
soil-moisture content recorded for the first 6-month period. Seedling survival varied from
0–15%, with the best survival in all 3 species occurring in the U. virgata grass edges (15%)
as compared to other treatments (0%). Patches of M. maximus displayed drastic fluctuations
in moisture levels, which may have inhibited native-species establishment. We observed
the highest seedling survivorship in Coccoloba microstachya (Puckhout) and Erythroxylum
areolatum (Swamp-redwood) individuals, suggesting that these species are good candidates
for restoration.
Introduction
Tropical dry forests worldwide have a long history of human use due mainly to
high soil fertility and a climate that has periods of dryness, which decrease human
and crop disease organisms (Janzen 1988, Murphy and Lugo 1986). Fire is commonly
used in neotropical dry forests to clear existing vegetation and establish pasturelands
(D’Antonio and Vitousek 1992); however, it is a novel disturbance agent in these areas
and may also promote ecosystem-level changes such as conversion into grassland or
species-poor scrublands in susceptible areas, or the introduction of non-native species
(Cabin et al. 2002a, Hooper et al. 2004, Vieira and Scariot 2006).
Non-native plants such as crops, timber, and animal fodder have been introduced
in the neotropics (Parsons 1972, Williams and Baruch 2000). Once introduced,
some species may monopolize resources, even without the active presence of fires
(Olsson et al. 2012). Furthermore, non-native grasses can limit native-seedling
1Biology Department, University of Puerto Rico at Mayagüez, PO Box 5000, Mayagüez, PR
00681-5000, USA. 2Current address - Department of Natural Resources Management, Texas
Tech University, Lubbock, TX 79409, USA. 3Department of Biological Sciences, Eastern
Kentucky University, 521 Lancaster Avenue Richmond, KY 40475, USA. *Corresponding
author - yamet.colegial@gmail.com.
Manuscript Editor: Michael Oatham
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J.G. García-Cancel and J.M. Thaxton
2018 No. 51
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germination or establishment due to the thick thatch they produce, directly compete
for resources (e.g., water, light, or nutrient), or serve as shelter for animals that are
natural herbivores of plant species used in restoration efforts (Cabin et al. 2002b,
Francis and Parrotta 2006, Jackson 2005, Ortega-Piek et al. 2011, Stevens and
Fehmi 2009, Thaxton et al. 2012b). Altering disturbance regimes facilitates the arrival
and establishment of non-native grasses (D’Antonio and Vitousek 1992).
The profuse fibrous root system of some grasses increases their ability to extract
water from the soil (Rojas-Sandoval and Meléndez-Ackerman 2012) and can thereby
limit the germination and establishment of tree seedlings (Ammondt et al. 2013). In
Hawaiian dry forests, the dense, shallow root systems of the Old World, non-native
grass Pennisetum setaceum (Forssk.) Chiov. (Fountain Grass) were detrimental to
the development of native woody seedlings because these non-native root systems
competed for water following small pulses of rain (Cordell and Sandquist 2008).
Studies in South America and Australia with non-native African grasses have shown
similar evidence of these detrimental effects on local biodiversity (Baruch and Jackson
2005, Jackson 2005). Studies from Guánica Forest in Puerto Rico indicate that
the long-term survival of native woody species is diminished in non-native grassinvaded
habitats (Pérez-Martínez 2007, Ramjohn et al. 2012, Wolfe and Van Bloem
2012), yet it is not clear if native grasses with fibrous root systems also impose this
type of barrier on native woody species establishment.
Previous research indicated that native-grass patches within the Guánica Forest
had more native woody species than patches of non-native grass (García-Cancel
2013). In the Guánica sub-tropical dry forest in Puerto Rico, the most widespread
non-native grass is Megathyrsus maximus (Jacq.) B.K. Simon and S.W.L. Jacobs
(Guinea Grass) (Monsegur-Rivera 2009), a perennial, facultative, apomictic, C4
bunchgrass that was introduced from Africa into the Neotropics in cattle feed (Williams
and Baruch 2000). Megathyrsus maximus produces a great number of seeds and
tends to create large fuel loads that can readily ignite and promote fire spread (Más
and García-Molinari 2006). This grass has altered the disturbance regime in the forest,
changing nutrient and community dynamics in favor of fire-adapted species, e.g.,
creating thatch carpets capable of easily catching fire (Thaxton et al. 2012b). Previous
research also has suggested that among the native grasses present in the forest,
the patches of native bunchgrass Uniola virgata (Poir.) Griseb. (Limestone Grass)
had a higher number of native woody species present, including many native woody
seedlings, compared to the other abundant native grass Bouteloua repens (Kunth)
Scribn. & Merr. (Slender Grama) (García-Cancel 2013). Therefore, the presence of
M. maximus in the forest may indicate a barrier to effective native-plant restoration,
and the presence of U. virgata could be a catalyst for passive restoration efforts. Thus,
the goal of this study was to determine the potential role that both species may play in
altering survival of woody-plant seedlings via soil water conditions.
Field site Description
The Guánica dry forest is in the southwestern part of the Caribbean island of Puerto
Rico (17°57'56''N, 66°52' 45''W; Fig. 1). Precipitation averages 860 mm with an
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J.G. García-Cancel and J.M. Thaxton
2018 No. 51
average temperature of 25.1 °C (Murphy and Lugo 1986). Soils are derived from
a porous limestone substrate and have a notable phosphorous deficiency typical of
karstic soils (Ceccon et al. 2006, Lugo and Murphy 1986), which amplifies the effects
of water stress on the plant community (Van Bloem et al. 2004). Some areas
within the forest lack deep organic layers and experience high solar irradiation
(Castilleja 1991 and sources, Lugo et al. 1978).
Materials and Methods
Field experiment
We grew seedlings of 3 native woody species, Jacquinia berteroi Spreng (Bois
Bande; Primulaceae, subfamily Theophrastoideae), Coccoloba microstachya
Willd. (Puckhout; Polygonaceae), and Erythroxylum areolatum L. (Swampredwood;
Erythroxylaceae) and assessed their establishment and survivorship
in clumps of 2 large bunchgrasses and in nearby open-soil areas. Erythroxylum
areolatum has been suggested as a useful species for restoration of degraded areas
because it is able to survive in grass-invaded sites (Wolfe and Van Bloem 2012).
Coccoloba microstachya is a common shrub or tree in the forest, and J. berteroi
is a small tree with a widespread distribution throughout the well-drained soils
of the forest (Monsegur-Rivera 2009). We germinated seeds collected in Guánica
Forest during the summer of 2011. We grew seedlings in greenhouse conditions
for a year prior to hardening off for 2 weeks, and transplanted them into the
field from 26 to 28 October 2012. We included a total of 400 seedlings (average
height = 4.2 cm) in the experiment: 160 seedlings each for C. microstachya and
J. berteroi, and 80 seedlings for E. areolatum.
Figure 1. (A) The island of Puerto Rico, (B) the location of Guánica Subtropical Dry Forest,
and (C) the extent of the forest.
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We transplanted seedlings to 2 sites: 1 patch dominated by non-native M. maximus
and 1 dominated by native U. virgata. Coordinates for each are 17°57'40.2422''N,
66°50'56.5656''W and 17°57'21.912''N, 66°54'15.301''W, respectively. At each
site, we randomly selected 20 grass clumps and 20 adjacent bare-soil spots to receive
seedling transplants. We selected sites based on their similar fire histories
in the recent past—large burns 30 years prior to our present study (Thaxton et al.
2012b)—and for the dominant presence of 1 of the 2 grasses. We transplanted seedlings
into bare-soil spots and grass edges to discern the best woody species for initial
restoration efforts in degraded forest areas and determine if shading by grass edges
influenced the survivorship of the seedlings. At each location, the size of bare-soil
spots varied from ~1 m2 to 4 m2; though that variable was not included as a study parameter.
To limit damage to the young root systems at the transplanted sites, we did
not measure other factors, such as soil nutrients and water interactions.
We transplanted 2 J. berteroi, 2 C. microstachya, and 1 E. areolatum seedlings
into the field around each of the 20 chosen grass clumps at each site, for a total of 5
randomly selected seedlings per clump. We did the same for the 20 bare-soil spots
at each site. Seven volunteers planted the seedlings in shallow holes, taking care to
limit damage to the young root systems. Transplants around each grass clump and in
each bare-soil spot were spaced and randomized to minimize microsite effects due
to aspect and orientation relative to the grass clump. We watered all seedlings (~95
mL of water per plant per time) every 2 days for 2 weeks following transplanting
to ameliorate the effect of transplantation shock. We initially censused for seedling
survival and measured seedlings bi-weekly for 3 months (8 censuses) post-planting.
Subsequent monthly censuses were carried out until 3 May 2013 (6 months postplanting),
and 4 additional censuses were carried out until 6 September 2014 (~2 y),
for a total of 15 censuses.
We measured microclimate conditions (i.e., volumetric soil moisture) in
grass clumps near the seedlings during the initial 6-month study period to detect
associations between grass edge and available soil moisture. We set up 5 soilmoisture
sensors (Decagon EC-5; decagon.com) with attached HOBO dataloggers
(onsetcomp.com) in the field. We used the dielectric constant of the soil and the
capacitance of the sensors to measure volumetric soil moisture. We monitored
soil moisture during the initial phase of the experiment to detect any differences
between the bare soil and grass edges. Data analyzed for this study were 6:00
AM values because morning values are less affected by the plants’ transpiration
process, which in other systrems has been shown to deplete available soil waterpotential
by root action due to transpiration (Richards and Cardwell 1987). Probes
from the M. maximus site malfunctioned and stopped collecting for some dates,
so we could gather and analyze only partial data. In contrast, data from the U. virgata
site experienced no such complications and we obtained a continuous set of
data for 6 months.
Statistical analyses
The data for seedling survivorship were subject to the survival probability for
each census, which varied between 1 and 0 at each census. Data were not normal
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(Shapiro-Wilk’s test) but were homoscedastic; thus, we conducted a non-parametric
Friedman’s test (Dytham 2011). The Friedman’s test, which is similar to a repeated
measure ANOVA, is performed on binary data (i.e., 1 = survival and 0 = death)
and converted into ranks (Di Rienzo et al. 2008). Friedman’s test can account for
random sampling; observations are repeated across individuals, and the test assumes
homogeneity of variances, normality of errors, sphericity and no individual
by-treatment interaction, which satisfies the necessary assumptions to analyze our
dataset (Dytham 2011). We employed Friedman’s test to assess if seedling survivorship
was higher in grass edges or bare-soil spots, and if survivorship was higher in
the native-grass patch or the non-native grass patch. Described values are accompanied
by the standard error produced by the test.
We also carried out Friedman’s tests for the available soil-moisture data to determine
if there were statistical differences between treatments. For all tests, data were
run with α = 0.05, and adjusted a posteriori with a Bonferroni test (α = 0.01). We
conducted all analyses in the statistics program Infostat ® (Di Rienzo et al. 2008).
Results
Seedling survivorship
Initial assements after the 6-month period showed that grass-edge treatments
had higher seedling survivorship than bare-soil treatments (P = 0.0001). After 2
years, grass edges still had higher woody-seedling survival than bare-soil spots,
but the survival rate differed depending on which grass species was used as a nurse
plant (Table 1, Fig. 2). There was higher survivorship for E. areolatum seedlings
than for J. berteroi or C. microstachya seedlings (P = 0.0006; Table 1, Fig. 3).
By the end of the study, surviving seedlings were found only in the grass edge of
Table 1. Friedman test on the seedling survivorship in the native (U. v. = Uniola virgata [Sea-oats])
and non-native (M. m. = Megathyrsus maximus[Guinea Grass]) grass edges and adjacent bare-soil
treatments. Comparison among the 3 native woody species survivorship is shown. Different letters
indicate significant differences among treatments using Bonferroni a posteriori adjustments (α = 0.01).
Variable T2 P-value
Treatment (grass edge vs. bare soil) 25.79 0.0001
Native woody species 9.75 0.0006
Adjusted posterior comparisons
Site + treatment
U. v. bare soil spot A
M. m. bare soil spot A
M. m. grass edge B
U. v. grass edge B
Native woody species
Erythroxylum areolatum (Swamp-redwood) A
Coccoloba microstachya (Puckhout) B
Jacquinia berteroi (Bois Bande) B
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U. virgata grass clumps (15% ± SE 6.75); the other sites had no surviving seedlings
by September 2014. Final percent survival was 4 seedlings from the initial 80
individuals (5% ± 7.99 SE) for E. areolatum, 5 seedlings from the initial 160 individuals
(3.125% ± 9.18 SE) for J. berteroi, and 6 seedlings from the intitial 160
individuals (3.75% ± SE 9.14) for C. microstachya.
Soil moisture
Soil moisture signifcantly differed between the 4 treatments (P < 0.0001), with
higher soil moisture in bare-soil spots (on average 0.102 m³/m³ volumetric water
content or VWC) over grass edges (on average 0.082 m³/m³ VWC). Measurements
were subject to wide fluctuations over time; fluctuations were greater in the nonnative
M. maximus treatments than with U. virgata treatments (Fig. 4).
Discussion
Seedlings of native woody plants survived at higher rates near native grass clumps
compared to non-native grass clumps or bare-soil spots in this study within the
Guánica Forest, suggesting that U. virgata might be used as a nurse plant for reforestation
efforts. By the end of the experiment, 15% of transplanted seedlings survived
when adjacent to U. virgata, while no seedlings survived in the other treatments.
This result could be due to shading effects, resource and or nutrient-flux changes, or
Figure 2. Survivorship (%) of woody species seedlings during a 23-month period. Black
diamonds with dashed lines = Uniola virgata (U.v.) on bare soil, black diamonds with continuous
lines = U. v. on grass edge, grey squares with dashed lines = Megathyrsus maximus
(M.m.) on bare-soil spots, and grey squares with continuous lines = M. m. on grass edge.
Values are presented with standard error bars. Vertical solid lines indicate the start of the
longest dry season.
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other site-ameliorating effects produced by the native grass. During our experiment,
U. virgata edges showed moisture values that were more constant, which could be
the result of shading. Shading might mitigate the effects of direct sunlight exposure,
allowing more time for seedlings to absorb available soil moisture in soil-surface layers,
thus potentially becoming suitable microhabitat for germinated seedlings. Such
an environment with stable shade could also offer a cooler surface temperature, and a
slower evaporation rate for available surface water, allowing the seedlings to survive
until the next water event. Other studies have shown that different shade regimens
can affect the survivorship of seedlings in drought events (Marañón et al. 2004) and
with grass presence (Maestre et al. 2003). We carried out the experiment during 2
extremely dry seasons for 23 months; thus, even the limited shade provided by the
grasses has the potential to benefit the young plants.
Available soil-water content was highest in M. maximus bare-soil–spot
treatments during the 6-month period (Fig. 4). We recommend caution in interpreting
the data from these probes because some of them malfunctioned during the
experiment. Even so, both M. maximus bare-soil spots and grass edges showed a
highly variable soil-water content compared to those of U. virgata. Such drastic
Figure 3. Survivorship (%) of seedlings of the 3 target species under Uniola virgata grass
edges during a 23-month period. Squares represent Jacquinia berteroi seedlings, circles represent
Coccoloba microstachya seedlings, and triangles represent Erythroxylum areolatum
seedlings. Values are presented with standard error bars. Vertical lines indicate start of the
longest dry season.
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change in available soil water on M. maximus edges could be too stressful for
the survival of the seedlings. Other studies have suggested that abiotic filters are the
primary determinants in seedling germination and establishment in resource-poor
environments (Mangla et al. 2011) and that addition of shade improves seedling
establishment (Thaxton et al. 2012a).
The seedling stage is critical for tree establishment because mortality is highest
then (Baudena et al. 2010). We suggest that this period is also the critical stage
during which M. maximus can have the greatest influence on local flora; other
non-native species may exert similar effects. This dynamic is important to ecosys -
tems with pronounced dry seasons, such as Mediterranean climates, arid regions,
tropical dry forests, or habitats with a significant ecological perturbation such as
abandoned pasturelands (Holl 1999, Lugo 2004, Maza-Villalobos et al. 2011). In
highly degraded ecosystems with strong legacy effects, restoration can be an uphill
battle with established abiotic thresholds (Prober et al. 2009, Wolfe and Van Bloem
2012). Nurse plants can ameliorate some of these abiotic stresses by creating sites
with favorable microclimate, humidity, and/or shade for native woody species
(Chinea 2002, Padilla and Pugnaire 2006, Santiago-García et al. 2008).
Padilla and Pugnaire (2006) mentioned that the type of nurse plant and/or target
species, as well as the time of year for the transplantation, must be considered for
restoration efforts to be successful. Grasses have been shown to dominate early successional
ecological stages, paving the way for shrubs and trees (Kennard 2002).
Aide et al. (2000) presented evidence of facilitation of tree-seedling emergence
among grasses in abandoned pasture lands in Puerto Rico, and studies carried out
Figure 4. Average soil-water–content measurements among the 4 treatments for 6 months.
Light grey diamond icons with grey dashes = M. maximus (M.m.) bare-soil spots, Xs with
dark continuous lines are M.m. edges, dark grey circles with grey dashed lines = U. virgata
(U.v.) bare-soil spot treatment, and black squares with continuous dark lines = U.v. edge
treatment (Friedman tests, P-value < 0.0001).
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J.G. García-Cancel and J.M. Thaxton
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in the Dominican Republic have shown distinct native-plant associations with
U. virgata plants that could act as transition forms for tropical dry forests and
savannahs (Cano-Carmona et al. 2010, García-Fuentes et al. 2015). Our results
support these findings. Recent studies have also shown that both U. virgata and
M. maximus are capable of re-invading cleared areas in the Guánica Forest (Jaime
et al. 2017), making their presence a good indicator of the potential successional
trend the nearby forest might take.
The appropriateness of target plant-species for restoration must also be considered.
Jacquinia berteroi seedlings proved susceptible to transplantation shock;
their survivorship dropped in the non-native sites as well as the bare-soil native
site. Seedlings of this species still had their cotyledons when transplanted, so
low survivorship could have been partially due to seedling developmental stage.
Coccoloba microstachya seedlings had the highest mortality in bare-soil sites
regardless of native or non-native status. We planted the fewest E. areolatum seedlings
(n = 80), yet they had a higher percentage of survivorship. This result could be
because E. areolatum has a dense, fibrous root system early in development, allowing
it to absorb more water quickly, and its roots would be less exposed to damage
during transplantation (J.G. García-Cancel, pers. observ.). Erythroxylum areolatum
also occurs in a wide range of habitats (Monsegur-Rivera 2009) and it is deciduous
(Santiago-García 2010), which could be a survival strategy to conserve water
during the driest periods. In contrast to Santiago-García (2010), who observed
that E. areolatum had the highest mortality under nurse-tree treatment, our study
found this species had the lowest mortality. The greater E. areolatum survival that
we observed could be attributed to the type of nurse plant used; Santiago-García
(2010) employed the non-native tree Leucaena leucocephala (Lam.) de Wit (White
Leadtree), while we used the native grass U. virgata. Native grass and shrub species
may facilitate establishment of a range of healthier native ecosystems because they
could ameliorate many of the effects of altered disturbance regimes.
Conclusions
We found that the native U. virgata promoted higher survivorship of 3 tree
seedlings than a non-native grass or bare soil. Differences in survivorship could
be related to the type of grass used and to species-specific traits. We suggest that
U. virgata clumps could be used in restoration efforts in other tropical dry-forest
areas where the grass is native. Even with the limited survivorship of transplanted
seedlings from this study, this native grass could be useful in ecosystem restoration
by facilitating the establishment of pioneering native woody species in degraded
ecosystems. Such efforts would work towards lowering the risk of potential local
extirpations and even extinctions of native biota, and improve overall ecosystem
services at the landscape level.
Acknowledgments
We thank the Department of Biology at the University of Puerto Rico - Mayagüez for
aid with logistics and facilities, and the Howard Hughes Foundation for funding materials
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for transplant of the seedlings. We are grateful to the Departamento de Recursos Naturales
y Ambientales forest biologist M. Canals for allowing us to work in the forest, as well all
personnel from the Cabo Rojo Fish and Wildlife Refuge, specifically J.G. Martínez for the
donation of the seedlings. We thank Dr. S.J. Van Bloem, Dr. J.D. Chinea, Dr. D. Kolterman,
and Dr. W. Robles for insights into the experimental design and contributions to the
original manuscript. The authors also appreciate Z. Ascherl-Ramos, C.J. Pasiche-Lisboa,
and R.D. Cox for helpful comments during the preparation of this manuscript. We extend
special thanks to T. Velázquez-Rojas, C. Ramos-Gerena, C. Torres, V. Velez-Thaxton, R.
Carrera-Martínez, J. Lugo-Garces, L. Aponte-Díaz, P. Repollet, E. García-Roldán, I.C. Ortiz,
J. Santiago-Román, R. Almodovar, S. Rosado-Ortiz, and N. Medina-Echevarría for field
assistance during the course of the experiment, and Y. Amaya for help with coordinating
data management. This study was based on a Master’s degree experiment, which imposed
logistic restrictions due to the time frame and scale.
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