Recovery of tree community composition in restored tropical forests relies on successful recruitment of later-successional species. However, the long-term effects of different restoration interventions on establishment success of arriving seeds are poorly understood. We evaluated the effects of three restoration treatments on the seed-to-seedling transition for later-successional tree species in a fragmented agricultural landscape in southern Costa Rica. Restoration plots (0.25 ha) were established in a block design nearly two decades prior and spanned a gradient of intervention intensity: natural regeneration (not planted), applied nucleation (planted tree clusters), and plantation (fully planted). We conducted seed addition experiments from 2021-2023 using eight species at seven replicate restoration sites and in four nearby remnant forests.

Study region

We conducted this study from June 2021 to September 2023 in southern Costa Rica near Las Cruces Biological Station. The forests in this region are at the boundary between Tropical Premontane Wet and Rain Forest zones (Holdridge et al., 1971), and the study sites range from ~1100 to 1200 m in elevation. Soils are volcanic in origin, mildly acidic (pH ~5.5), low in P, and high in organic matter. Mean annual temperature is ~21°C, and annual precipitation is between 2700 and 5000 mm, with a dry season from December to March. The landscape was largely deforested between 1960 and 1980 and is now a fragmented mosaic of remnant forest, secondary forest, cattle pastures and agricultural fields.

Design of long-term restoration experiment

We used seven restoration sites set up in 2004-2006 (Figure 1), a subset of the 15 sites described in Holl et al. (2020). All sites were previously used for ≥ 18 years for grazing or coffee production, and most are steeply sloped (15-35º). Minimum distance between sites is 700 m. Each restoration site had three 50 × 50 m (0.25 ha) plots: natural regeneration, applied nucleation, and plantation. Natural regeneration plots were not planted, whereas applied nucleation and plantation plots were planted with seedlings of four tree species: Erythrina poeppigiana (Walp.) Skeels, Inga edulis Mart., Terminalia amazonia (J.F. Gmel.) Exell, and Vochysia guatemalensis Donn. Sm. E. poeppigiana and I. edulis are naturalized, fast-growing N-fixers commonly used in agricultural intercropping, whereas T. amazonia and V. guatemalensis are native timber species. In applied nucleation plots trees were planted in six nuclei or patches, two each of 4 × 4, 8 × 8, and 12 × 12 m, whereas plantation plots were planted uniformly throughout; the number of trees planted in applied nucleation plots was 27% of the total in plantation plots. Four sites also have paired remnant forest (2 to >300 ha) located 5-50 m from restoration plots, which serve as a reference system for comparing the conditions and ecological processes of restoration plots. On average, percent canopy cover was lower and cover of exotic pasture grasses was much greater in natural regeneration than planted treatments (Kulikowski et al., 2023).

Seed addition experiment

We selected eight tree species (Table 1; Figure S1) for seed addition experiments, based on: (a) their later-successional status, categorized by expert opinion as occurring exclusively in late-successional forest or both early- and late-successional forest (Schubert et al., 2025); (b) a continuum of seed sizes; and (c) regional availability of seeds (>700) for the experiment. Study species were all animal-dispersed and ranged from 2.4 to 24.8 mm seed width.

We collected seeds from at least three mother trees per species, encountered in conditions that differed by species: (1) mature fruits/nuts from the ground (Pseudolmedia mollis, Quercus benthamii, Trophis mexicana, and some Erythroxylum macrophyllum and Otoba novogranatensis), (2) seeds with pulp already removed by birds from the ground (Ocotea puberula and some O. novogranatensis), and (3) ripe fruit directly from trees (Lacistema aggregatum, Palicourea padifolia, some E. macrophyllum). We manually removed any fruit pulp from seeds and briefly submerged all seeds in water to check for insect damage. Seeds that floated were assumed to be non-viable and not used. We measured fresh mass and width for ≥50 seeds per species (Table 1; Figure S1). We thoroughly mixed conspecific seeds to avoid bias in seed source or quality among sites or habitat types (hereafter ‘treatments’). We conducted concurrent germination trials to confirm that no species completely failed to germinate in a nursery setting.

Within each restoration plot and remnant forest, we established four sets (hereafter ‘stations’) of three 1-m² quadrats separated by ≥15 m. In natural regeneration, plantation, and reference forest plots, one station was located in each quarter of the plot (Figure 2a). In applied nucleation plots, stations were systematically distributed with one station at each of four positions relative to the initial planting design (Figure 2b): (1) within a large nucleus; (2) at the edge of a medium nucleus; (3) between two nuclei, and (4) far from any nucleus. Nuclei have spread considerably beyond the planted area due to crown expansion and natural recruitment; however, these positions spanned a representative range of microsite conditions. Two quadrats per station were used for seed additions to accommodate all eight species, and one non-sown control quadrat was surveyed to assess background abundance of naturally recruited seedlings of the study species (Figure 2c). Each control quadrat was located 2 m from a seed addition quadrat and on the uphill side if on a slope. We placed quadrats to avoid large tree trunks but did not remove existing vegetation or leaf litter.

At each station, we sequentially added seeds of three species (P. mollis, O. novogranatensis, and O. puberula) to one quadrat (26 total seeds m⁻²) and five species (Q. benthamii, E. macrophyllum, L. aggregatum, T. mexicana, and P. padifolia) to the second quadrat (47 total seeds m⁻²) as seeds of each species became available between July 2021 and August 2022. The number of seeds added per quadrat ranged from 7 to 11 seeds per species (Table 1) because of varying seed availability. Realized densities of total added individuals were lower than the cumulative densities because of (a) temporal staggering of additions by species and (b) seed mortality. We placed seeds on top of the litter or soil to simulate how seeds would naturally be deposited from primary dispersal, slightly pressing down to minimize rolling on steeper slopes. We secured a roll of fine mesh ~5 cm high on the downhill side of each seed addition quadrat to catch seeds washed downslope by runoff. Seeds were placed in rows with 10-20 cm minimum spacing and assigned an identifier based on grid position. For the five smallest-seeded species, locations were marked with popsicle sticks to facilitate monitoring.

We monitored seedling emergence, survival, and height (to nearest 0.5 cm) for each species at least four times. Census intervals were timed to: (a) to capture seedling emergence approximately 1-2 months post seed addition and every 2-4 months thereafter, and (b) measure the first-year survival percentage of emerged seedlings 12-15 months after seed addition. Because some seedlings may have emerged and died between censuses or prior to the first census, we may have underestimated emergence and overestimated seedling survival. This uncertainty was mitigated by our observations of obviously dead seeds (indicating failure to emerge) or dead seedlings (indicating emergence but subsequent mortality). For the four larger-seeded species for which ungerminated seeds could be reliably re-found, missing seeds were interpreted as pre-emergence mortality. Assuming that missing seeds were predated was reasonable because a companion study found that almost all seeds removed by vertebrates were eventually predated (Joyce et al., 2024). Variable durations of the “first year” period are typical in tropical tree seed addition experiments involving multiple species with differing phenologies (e.g. Svenning & Wright, 2005). One of the seven sites where we added seeds in 2021 was then cleared by the landowner, so in 2022 we added seeds of the remaining six species to an alternate site. The other six sites were used for the duration of the study. Accordingly, P. mollis and Q. benthamii, were tested at six sites rather than seven.

We did not tag seedlings but were able to distinguish between natural recruits and seedlings that recruited from added seeds based on the location of seedlings relative to the grid and the size of seedlings compared to the experimental cohort. We sampled the non-sown control quadrat at each station at a single timepoint (July 2023) for seedlings similarly sized to our added individuals.

References

Holdridge, L. R., Grenke, W. C., Hatheway, W. H., Liany, T., & Tosi Jr, J. A. (1971). Forest environments in tropical life zones: A pilot study. Pergamon Press.

Holl, K. D., Reid, J. L., Cole, R. J., Oviedo‐Brenes, F., Rosales, J. A., & Zahawi, R. A. (2020). Applied nucleation facilitates tropical forest recovery: Lessons learned from a 15-year study. Journal of Applied Ecology, 57(12). https://doi.org/10.1111/1365-2664.13684

Joyce, F. H., Ramos, B. M., Zahawi, R. A., & Holl, K. D. (2024). Vertebrate seed predation can limit recruitment of later-successional species in tropical forest restoration. Biotropica, 56(6), e13381. https://doi.org/10.1111/btp.13381

Kulikowski, A. J., Zahawi, R. A., Werden, L. K., Zhu, K., & Holl, K. D. (2023). Restoration interventions mediate tropical tree recruitment dynamics over time. Philosophical Transactions of the Royal Society B: Biological Sciences, 378(1867), 20210077. https://doi.org/10.1098/rstb.2021.0077

Schubert, S. C., Zahawi, R. A., Oviedo-Brenes, F., Rosales, J. A., & Holl, K. D. (2025). Active restoration increases tree species richness and recruitment of large-seeded taxa after 16–18 years. Ecological Applications, 35(1), e3053. https://doi.org/10.1002/eap.3053

Svenning, J.-C., & Wright, S. J. (2005). Seed limitation in a Panamanian forest. Journal of Ecology, 93(5), 853–862. https://doi.org/10.1111/j.1365-2745.2005.01016.x