Caribbean Naturalist 1 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 Caribbean Naturalist J.G. García-Cancel and J.M. Thaxton 2018 No. 51 2 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 Caribbean Naturalist 3 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. Caribbean Naturalist J.G. García-Cancel and J.M. Thaxton 2018 No. 51 4 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 Caribbean Naturalist 5 J.G. García-Cancel and J.M. Thaxton 2018 No. 51 (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 Caribbean Naturalist J.G. García-Cancel and J.M. Thaxton 2018 No. 51 6 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. Caribbean Naturalist 7 J.G. García-Cancel and J.M. Thaxton 2018 No. 51 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. Caribbean Naturalist J.G. García-Cancel and J.M. Thaxton 2018 No. 51 8 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). Caribbean Naturalist 9 J.G. García-Cancel and J.M. Thaxton 2018 No. 51 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 Caribbean Naturalist J.G. García-Cancel and J.M. Thaxton 2018 No. 51 10 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. Literature Cited Aide, T.M., J.K. Zimmerman, J.B. Pascarella, L. Rivera, and H. Marcano-Vega. 2000. Forest regeneration in a chronosequence of tropical abandoned pastures: Implications for restoration ecology. Restoration Ecology 8(4):328–338. Ammondt, S.A., C.M. Litton, L.M. Ellsworth, and J.K. Leary. 2013. Restoration of native plant communities in a Hawaiian dry lowland ecosystem dominated by the invasive grass Megathyrsus maximus. Applied Vegetation Science 16:29–39. Baruch, Z., and R.B. Jackson. 2005. Responses of tropical native and invader C4 grasses to water stress, clipping, and increased atmospheric CO2 concentration. Oecologia 145:522–532. Baudena, M., F. D’Andrea, and A. Provensale. 2010. An idealized model for tree–grass coexistence in savannas: The role of life-stage structure and fire disturbances. Journal of Ecology 98:74–80. Búrquez-Montijo, A., M.E. Miller, and A. Martí-Yrizar. 2002. Mexican grasslands, thornscrub, and the transformation of the Sonoran Desert by invasive exotic Buffelgrass (Pennisetum ciliare). Pp 126–146, In B. Tellman (Ed.). Invasive Exotic Species in the Sonoran Region. University of Arizona Press and The Arizona-Sonora Desert Museum, Tucson, AZ, USA. 424 pp. Cabin, R.J., S.G. Weller, D.H. Lorence, S. Cordell, and L.J. Hadway. 2002a. Effects of microsite, water, weeding, and direct seeding on the regeneration of native and alien species within a Hawaiian dry-forest preserve. Biological Conse rvation 104:181–190. Cabin, R.J., S.G. Weller, D.H. Lorence, S. Cordell, L.J. Hadway, R. Montgomery, D. Goo, and A. Urakami. 2002b. Effects of light, alien grass, and native-species addition on a Hawaiian dry forest restoration. Ecological Applications 12(6):1595–1610. Cano Carmona, E., A. Velóz, and A. Cano-Ortiz. 2010. The habitats of Leptochloopsis virgata in the Dominican Republic. Acta Botanica Gallica 157 (4):645–658. Castilleja, G. 1991. Seed germination and early establishment in a sub-tropical dry forest. Ph.D. Disseration. Yale University, New Haven, CT, USA. Ceccon, E., P. Huante, and E. Rincón. 2006. Abiotic factors influencing tropical dry forest regeneration. Brazilian Archives of Biology and Technology 49(2):305–312. Chinea, J.D. 2002. Tropical forest succession on abandoned farms in the Humacao municipality of eastern Puerto Rico. Forest Ecology and Management 167(1–13):195–207. Cordell, S., and D.R. Sandquist. 2008. The impact of an invasive African bunchgrass (Pennisetum setaceum) on water availability and productivity of canopy trees within a tropical dry forest in Hawaii. Functional Ecology 22:1008–1017. Caribbean Naturalist 11 J.G. García-Cancel and J.M. Thaxton 2018 No. 51 D’Antonio, C.M., and P.M., Vitousek. 1992. Biological invasions, the grass–fire cycle, and global change. Annual Review of Ecology and Systematics 23:63–87. Decagon Devices, Inc. 2016. EC-5 Soil Moisture Sensor Operator’s Manual. Pullman WA. Di Rienzo, J.A., F. Casanoves, M.G. Balzarini, L. Gonzalez, M. Tablada, and C.W. Robledo. (2008). InfoStat, versión 2008. Grupo InfoStat, FCA, Universidad Nacional de Córdoba, Argentina. Dytham, C. 2011. Choosing and Using Statistics: A Biologist’s Guide. 3rd Edition. Wiley– Blackwell, Oxford, UK. 320 pp. Francis, J.K., and J.A. Parrotta. 2006. Vegetation response to grazing and planting of Leucaena leucocephala in a Urochloa maximum-dominated grassland in Puerto Rico. Caribbean Journal of Science 42(1):67–74. Franco, A.C., and P.S. Nobel. 1988. Interactions between seedlings of Agave deserti and the nurse plant Hilaria rigida. Ecology 69(6):1731–1740. Holl, K.D. 1999. Factors limiting tropical rain forest regeneration in abandoned pasture: Seed rain, seed germination, microclimate, and soil. Biotropica 31(2):229–242. Hooper, E.R., P. Legendre, and R. Condit. 2004. Factors affecting community composition of forest regeneration in deforested, abandoned land in Panama. Ecology 85(12):3313–3326. García-Cancel, J.G. 2013. Effects of native and non-native grasses on woody species regeneration in a Puerto Rican subtropical dry forest. M.Sc. Thesis. University of Puerto Rico-Mayagüez. Mayagüez, PR, USA. Jackson, J. 2005. Is there a relationship between herbaceous species richness and Buffelgrass (Cenchrus ciliaris)? Austral Ecology 30:505–517. Jaime, X.A., S.J. Van Bloem, F.H. Koch, and S.A. Nelson. 2017. Spread of common native and invasive grasses and ruderal trees following anthropogenic disturbances in a tropical dry forest. Ecological Processes 6 (38):1–14. Janzen, D.H. 1988. Tropical dry forests: The most endangered major tropical ecosystem. Pp. 130–137, In E.O. Wilson and F.M. Peter (Eds.). Biodiversity. National Academies Press, Washington, DC. 521 pp. Kennard, D.K. 2002. Secondary forest succession in a tropical dry forest: Patterns of development across a 50-year chronosequence in lowland Bolivia. Journal of Tropical Ecology 18:53–66. Lugo, A.E. 2004. The outcome of alien tree invasions in Puerto Rico. Frontiers in Ecology and the Environment 2(5):265–273. Lugo, A.E., and P.G. Murphy. 1986. Nutrient dynamics in a subtropical dry forest. Journal of Tropical Ecology 2:55–72. Lugo, A.E., J.A. González-Liboy, B. Cintrón, and K. Dugger. 1978. Structure, productivity, and transpiration of a subtropical dry forest in Puerto Rico. B iotropica 10(4):278–291. Maestre, F.T., S. Bautista, and J. Cortina. 2003. Positive, negative, and net effects in grass– shrub interactions in Mediterranean semiarid grasslands. Ecolog y 84(12):3186–3197. Magla, S., R.N. Sheley, J.J. James., and S.R. Radosevich. 2011. Role of competition in restoring resource-poor arid systems dominated by invasive grasses. Journal of Arid Environments 75:487–493. Marañón, T., R. Zamora, R. Villar, M.A. Zavala, J.L. Quero, I. Pérez-Ramos, I. Mendoza, and J. Castro. 2004. Regeneration of tree species and restoration under contrasted Mediterranean habitats: Field and glasshouse experiments. International Journal of Ecology and Environmental Sciences 30(3):187–196. Mas, E.G., and O. García-Molinari. 2006. Manual de yerbas de Puerto Rico. USDA NRCS Technical Report, 2nd Edition. University of Puerto Rico-Mayagüez College of Agricultural Sciences, Mayagüez, PR, USA. 300 pp. Caribbean Naturalist J.G. García-Cancel and J.M. Thaxton 2018 No. 51 12 Maza-Villalobos, S., P. Balvanera, and M. Martínez-Ramos. 2011. Early regeneration of tropical dry forest from abandoned pastures: Contrasting chronosequence and dynamic approaches. Biotropica 43(6):666–675. Monsegur-Rivera, O. 2009. Vascular flora of the Guánica fry forest. M.Sc. Thesis. University of Puerto Rico-Mayagüez. Mayagüez, PR, USA. Murphy, P.G., and A.E. Lugo. 1986. Ecology of tropical dry forest. Annual Reviews of Ecological Systems 17:67–88. Olsson, A.D., J. Betancourt, M.P. McClaran, and S.E. Marsh. 2012. Sonoran Desert ecosystem transformation by a C4 grass without the grass/fire cycle. Diversity and Distributions 18:10–21. Ortega-Piek, A., F. López-Barrera, N. Raírez-Marcial, and J.G. García-Franco. 2011. Early seedling establishment of two tropical, montane cloud-forest tree species: The role of native and exotic grasses. Forest Ecology and Management 261:13 36–1343. Padilla, F.M., and F.I. Pugnaire. 2006. The role of nurse plants in the restoration of degraded environments. Frontiers in Ecology and the Environment 4 (4):196–202. Parsons, J.J. 1972. Spread of African grasses to the American tropics. Journal of Range Management 25(1):12–17. Pérez-Martínez, F.O. 2007. Effect of exotic canopy on understory species composition in degraded subtropical dry forest of Puerto Rico. M.Sc. Thesis. University of Puerto Rico- Mayagüez. Mayagüez, PR, USA. Prober, S.M., I.D. Lunt, and J.W. Morgan. 2009. Rapid internal plant–soil feedbacks lead to alternative stable states in temperate Australian grassy woodlands. Pp. 156–168, In R.J. Hobbs and K.N. Suding (Eds.). New Models for Ecosystem Dynamics and Restoration. Island Press, Washington, DC, USA. 366 pp. Ramjohn, I.A., P.G. Murphy, T.M. Burton, and A.E. Lugo. 2012. Survival and rebound of Antillean dry forests: Role of forest fragments. Forest Ecology and Management 284:124–132. Richards, J.H., and M.M. Caldwell. 1987. Hydraulic lift: Substantial nocturnal transport between soil layers by Artemisia tridentata roots. Oecologia 73:486–489. Rojas-Sandoval, J., and E. Meléndez-Ackerman. 2012. Effects of an invasive grass on the demography of the Caribbean cactus Harrisia portoricensis: Implications for cacti conservation. Acta Oecologica 41:30–38. Santiago-García, R.J. 2010. Performance of native tree species planted under nurse trees for forest restoration in Puerto Rico. M.Sc. Thesis. University of Puerto Rico-Mayagüez. Mayagüez, PR, USA. Santiago-García, R.J., S. Molina-Colón, P. Sollins, and S.J. Van Bloem. 2008. The role of nurse trees in mitigating fire effects on tropical dry-forest restoration: A case study. Ambio 37(7):604–608. Stevens, J.M., and J.S. Fehmi. 2009. Competitive effect of two nonnative grasses on a native grass in southern Arizona. Invasive Plant Science and Management 2(4):379–385. Thaxton, J.M., S. Cordell, R.J. Cabin, and D.R. Sandquist. 2012a. Non-native grass removal and shade increase soil moisture and seedling performance during Hawaiian dryforest restoration. Restoration Ecology 20(4):4756–482. Thaxton, J.M., S.J. Van Bloem, and S. Whitmire. 2012b. Fuel conditions associated with native and exotic grasses in a subtropical dry forest in Puerto Rico. Fire Ecology 8(3):9–17. Van Bloem, S.J., P.G. Murphy, and A.E. Lugo. 2004. Tropical dry forests. Pp. 1767–1775, In J. Burley, J. Evans, and J. Youngquist (Eds.). Encyclopedia of Forest Sciences. Elsevier, Oxford, UK. 2400 pp. Caribbean Naturalist 13 J.G. García-Cancel and J.M. Thaxton 2018 No. 51 Vieira, D.L.M., and A. Scariot. 2006. Principles of natural regeneration of tropical dry forests for restoration. Restoration Ecology 14(1):11–20. Williams, D.G., and Z. Baruch. 2000. African grass invasion in the Americas: Ecosystem consequences and the role of ecophysiology. Biological Invasions 2:123–140. Wolfe, B.T., and S.J. Van Bloem. 2012. Subtropical dry-forest regeneration in grass-invaded areas of Puerto Rico: Understanding why Leucaena leucocephala dominates and native species fail. Forest Ecology and Management 267:253–261.