Research Paper

INTRODUCTION

The Philippines is categorized as one of the world’s megadiverse countries (Mittermeier et al. 1997), which is reflected in its classification as a global biodiversity hotspot (Myers et al. 2000), and as an important centre of bird endemism (Convention on Biological Diversity 2020). It has one of the highest proportions of globally threatened bird species in the world: 13.6% of 735 species (IUCN 2023), while 29.1% of 247 country endemics are globally threatened with extinction (Brinkman et al. 2023). Despite this, conservation concerns arise from some of the world’s highest deforestation rates, having lost 93% of its original forest cover (Hughes 2017), coupled with considerable pressure from unsustainable and illegal wildlife hunting (Gray et al. 2018).

The Philippines’ Western Visayas faunal region (Heaney 1985) has experienced the highest rates of deforestation in the country, resulting in a reduction in old growth forest to < 6%, 4%, and 0.05% of the region’s main islands of Panay, Negros, and Cebu, respectively (Lorica and Oliver 2006). It is likely that most of the remaining forest in the region is found in the west and northwest of Panay (Lorica and Oliver 2006), and particularly the Northwest Panay Peninsula Natural Park (NPPNP), which is thought to include the largest remaining area of low elevation forest in the Western Visayas (BirdLife International 2024a). Understanding how bird communities respond to habitat degradation and land-use change, and which species may be more sensitive (Gray et al. 2006), remains a conservation priority in the species-diverse and highly threatened forest landscapes of the Western Visayas.

Avian species richness, or taxonomic diversity, tends to decrease along gradients of increasing habitat modification (Schultze et al. 2004), yet species responses to the impacts of anthropogenic land-use change can differ widely (Şekercioḡlu et al. 2002, Sodhi et al. 2004) across species and species groups or guilds defined by shared ecological traits (Lambert 1992, Canaday 1996, Schultze et al. 2004, Waltert et al. 2004, Tvardíková, 2010). For example, habitat degradation may differentially impact species based on dietary type and feeding strata (Flynn et al. 2009, Luck et al. 2012, Corbelli et al. 2015), while larger species may be more impacted than smaller species, which has wider ramifications relating to home range size, lifespan, population viability, and the provision of ecosystem services (Cofre et al. 2007, Flynn et al. 2009, Vandewalle et al. 2010, Luck et al. 2012, Corbelli et al. 2015).

The efficacy of taxonomic diversity as a measure of community responses to, or recovery from habitat degradation (Cianciaruso et al. 2009) or the state and resilience of a community’s ecological functionality (Petchey et al. 2004, Mouillot et al. 2013) is limited by the fact that all species are considered equally (Cianciaruso et al. 2009). Incorporating into diversity measures information on the functional traits represented within a community can demonstrate how that community may respond to habitat change (Cosset and Edwards 2017). Understanding how these changes influence functional diversity, the degree of (trait-based) ecological differences among species within a specified community (Tilman 2001), has wider implications for the functionality of a system and its processes (Tilman 2000, Loreau et al. 2001). Additionally, although highest species diversity may be evidenced in non-equilibrium tropical forest communities subject to perturbations and an absence of disturbance leading to equilibrium communities of lower diversity, as posited by the intermediate disturbance hypothesis (Connell 1978), this implies all species are equivalent, so species diversity may not represent absolute conservation value (Sheil et al. 1999). Although conservation action needs to consider how disturbance events drive species diversity in tropical forests (e.g., Sheil and Burslem 2003), which may run counter to conservation goals, functional diversity can aid in conservation planning, particularly when managing or attempting to restore functioning ecosystems (Cadotte et al. 2011).

Although previous research in the Western Visayas has focused on the responses of single bird species (Jakosalem et al. 2013, Mynott et al. 2020, 2021, Paguntalan et al. 2023) and non-avian communities (Alcala et al. 2004, Peque and Hölscher 2014) to habitat alteration, we look to address the significant knowledge gap in understanding habitat-related responses of tropical forest bird communities in this threatened faunal region. We use three measures of community diversity, taxonomic (species richness), species, and functional, to quantify patterns in bird community assemblages across a habitat gradient of primary forest, secondary forest, plantation, and open areas in NPPNP. Understanding such responses to habitat disturbance is essential for identifying priorities and applying limited conservation resources that improve habitat quality or guide habitat connectivity most effectively (Whytock et al. 2018), and this novel work provides a significant baseline for informing species and community-focused recommendations for conservation management decision making in the park. We anticipate this will provide empirical support to wider discussions on conserving forest biodiversity in the Western Visayas.

METHODS

Study site

With consent from the Protected Area Management Board, this study took place from January to April 2020 in the Northwest Panay Peninsula Natural Park (NPPNP hereafter; central coordinates 11° 49′ 31″ N, 121° 58′ 02″ E; Fig. 1), which is situated in the provinces of Aklan and Antique on the island of Panay, the Philippines. The NPPNP was designated a protected area in 2002 under the National Integrated Protected Areas System (NIPAS) Act, Philippines (UNEP-WCMC 2022) and comprises 120 km² of tall dipterocarp forest, limestone karst forest, lower montane forest, and bamboo forest across an elevational range of 27 to 915 m (BirdLife International 2024a). With up to 5000 ha of old growth forest remaining, it is one of the largest contiguous protected lowland forests in the Western Visayas (Key Biodiversity Areas Partnership 2022) and probably the most extensive lowland forest remaining in the Negros and Panay endemic bird area (EBA; Stattersfield et al. 1998). Within NPPNP, there has been some legal land conversion for farming on plots specified before the protected area was established, while illegal logging and slash-and-burn agriculture (kaingin) persist alongside the collection of forest products and illegal hunting (Kerhoas et al. 2019, Mynott et al. 2020, BirdLife International 2024a, Key Biodiversity Areas Partnership 2022).

The NPPNP is an Important Bird and Biodiversity Area (IBA; BirdLife International 2024a) due to the presence of five globally threatened species: Negros Bleeding-heart (Gallicolumba keayi; CR), Writhe-billed Hornbill (Rhabdotorrhinus waldeni; CR), Visayan Hornbill (Penelopides panini; EN), White-winged Cuckooshrike (Analisoma ostenta; VU), and Green-faced Parrotfinch (Erythrura viridifacies; VU; BirdLife International 2024b). These species also trigger the park’s qualification as a key biodiversity area (KBA; Key Biodiversity Area Partnership 2022). The IBA designation pre-dates the taxonomic overhaul of del Hoyo and Collar (2014, 2016), which now includes the presence of the additional threatened species of Yellow-faced Flameback (Chrysocolaptes xanthocephalus; EN) and Negros Scops-owl (Otus nigrorum; VU), while the under-recorded Negros Jungle Flycatcher (Vauriella albigularis; EN) is also present (eBird 2023). The combination of the site’s regionally important forest cover, existing pressures, and presence of several key species make NPPNP a conservation priority (BirdLife International 2024a).

Sampling design and bird surveys

We divided NPPNP into a grid of 2.5 km * 2.5 km cells (n = 33) and randomly selected 24 of these cells to survey (72.7% of the park; Fig. 1). However, once on site, the terrain of three of the selected cells was found to be inaccessible. In these instances, the nearest accessible grid cell was surveyed instead (Mynott et al. 2021). Each cell contained between one and four line transects of varying distance (0.3 to 2.5 km, x̅ = 2.3 km ± 0.53 SD) that were positioned randomly, or along narrow trails where the terrain was particularly difficult. Any trails used were < 1 m wide and therefore were not considered to have affected the representativeness of the data (Cornils et al. 2015) or violated any sampling design assumptions (Bibby et al. 2000). Nine cells included transects positioned in part or entirely in adjacent habitat outside the park’s boundary.

We positioned survey points ≥ 200–250 m apart along each transect to minimize the chance of double-counting individual birds from adjacent points (Bibby et al. 1998, Marsden 1999). We completed 10-minute distance sampling point transect surveys with no settling down period and recorded all birds seen or heard (Lloyd et al. 2000, Lee and Marsden 2008). We used Kennedy et al. (2000) to aid with bird identification in the field and followed the taxonomy of Clements et al. (2023).

Surveys were completed between 08:00–18:00 and, although bird detectability may be (aurally) higher in early morning/late afternoon (Lynch 1995, Bibby et al. 2000), survey points were visited throughout the day in each location and habitat, minimizing any temporal detectability bias while potentially capturing fine-scale variation in species availability for detection and their detectability (Verner and Ritter 1986). We did not complete any surveys in periods of high wind or rainfall (Marsden and Pilgrim 2003). When possible, we used a laser rangefinder to record radial distances to individual birds or to the central location of a species group. However, if a clear line of sight was unobtainable, we estimated distance to a bird detection. Pre-survey training in distance estimation ensured radial distances were not consistently over- or under-estimated (Bibby et al. 1998). We recorded the group size of each detection. If we could not confirm this visually, we used a habitat-specific mean group size calculated from visual detections of the species instead (Lee and Marsden 2008). All cells were visited once, with surveys repeated in seven cells from mid-February to the start of April (29.2%; access restrictions in response to the coronavirus pandemic prevented repeats in the remaining 17 cells). In total, we surveyed 278 points, of which 80 were repeated, across an elevational range of 17 to 907 m.

We classified the habitat type at each point into one of four categories based on floristic composition and structural attributes, local knowledge of plot histories, and supported by satellite imagery (PhilGIS 2020): primary (old growth forest; n = 111), secondary (previously disturbed) forest (n = 105), plantation (n = 41), and open habitat (n = 21). Habitat categorization was supported by plot-based surveys as described in Mynott et al. (2021). Plantations had a more open canopy than forest habitat, mainly consisting of coconut palm trees with the occasional presence of regenerating forest and some newly planted areas of young mahogany. Open habitats had low canopy cover (< 11%), were dominated by scrub, and generally lacked trees with ≥ 25 cm diameter at breast height. In this study, we considered change between primary forest, secondary forest, plantation, and open habitat as a gradient of habitat disturbance or degradation.

Statistical analyses

For overall site diversity estimates, we used all species records, including aerial detections (e.g., Collocalia species), with an unlimited detection radius. For all other (habitat-focused) analyses, we excluded aerial detections and right truncated bird detections at a radial distance of 50 m to ensure species records were attributed to a habitat type spatially commensurate with the respective survey point. For each habitat type, we used species richness (S_r) as a measure of taxonomic diversity, the Shannon-Wiener (H') index for species diversity, and two indices for functional diversity (FD); functional dispersion index (FDis), which reflects deviations in species-level trait values from a functional centroid, and Rao’s quadratic entropy (Q), which considers the variance of species-species trait dissimilarities (Laliberté and Legendre 2010). We selected these two measures of FD because they allow for the inclusion of qualitative traits and are unaffected by species richness (Laliberté and Legendre 2010). Greater differences in species traits within communities are reflected in larger FD (Petchey and Gaston 2002). We also generated a simple endemicity index based on the proportion of endemic species, from the total pool of endemic species recorded in the study, detected at each point. We used Brinkman et al. (2023) for defining species’ endemicity status.

We estimated species richness using iNEXT software (Chao et al. 2016) as a sample size-based inter- and extrapolation for comparisons across habitat-specific communities from incomplete sample data. We produced rarefaction and extrapolation sampling curves for each habitat type using species incidence data, which are less sensitive than abundance data to aggregations of individuals (Chao et al. 2020), while also minimizing habitat-based detectability differences associated with relative abundances. We set the confidence interval (CIs) at 84% (for a Type I error rate of p < 0.05; Payton et al. 2003) for each curve with 999 bootstrap replications and compared diversity estimates for common sample sizes and sample coverages with Ĉ as an estimator of sample coverage in each habitat (Chao and Jost 2012). We calculated habitat-specific species diversity indices in PRIMER 7 (Clarke and Gorley 2006).

We extracted species-level foraging attributes from Wilman et al. (2014), and references therein, to classify species into guilds based on the functional traits of body size, foraging location, and diet. For diet, we assigned species to the majority type reported (frugivorous, granivorous, insectivorous, omnivorous, or vertivorous). We allocated species to their most frequently recorded foraging strata, i.e., ground, understorey, understorey-midstorey (lower storeys), midstorey-canopy (upper storeys), general-level (all vegetation levels), or aerial. We classified body size into three categories: small (< 25 g), medium (25–125 g), and large (> 125 g). In combination, this resulted in 31 functional guilds (Appendix 1). Using these functional traits alongside a species incidence matrix, we calculated FDis and Q using the FD package (Laliberté et al. 2014) in R (R Development Core Team 2017).

We used the glm2 package (Marschner 2011) in R to construct generalized linear models (GLMs) for comparing point values of species richness and diversity, FD, and endemicity (response variables) between habitat types. We used a Poisson distribution and log link function for species richness, gaussian distribution and identity link function for species and functional diversities, and endemicity. We tested overall model significance using likelihood ratio tests (LRT; p < 0.05) and used the emmeans function, which calculates estimated marginal means among groups given a fitted model (in this case, GLMs), in the emmeans R package (Lenth et al. 2022) to test for significant pairwise differences between habitats.

We used PRIMER 7 (Clarke and Gorley 2006) to analyze patterns in bird communities at the functional guild level within and between the four habitat types; Clarke (1993), and Clarke and Warwick (2001) fully describe these techniques. We weighted bird encounters by point-based survey effort, pre-treated species abundances with a square root transformation to down-weight the importance of the most abundant guilds, and factorized each point transect by habitat type (Clarke and Warwick 2001). Based on a Bray-Curtis coefficient resemblance matrix, we performed an analysis of similarity (ANOSIM) to test for spatial differences between bird communities of different habitat types. We then conducted a similarity of percentages analysis (SIMPER) to identify which functional guilds (and member species) contributed most to any dissimilarities in community composition between habitat types. We compared the relative abundances of key contributing guilds and selected species (those that contributed ≥ 5% to at least one pairwise habitat comparison) in the different habitats using Kruskal-Wallis testing with p values adjusted using a Bonferroni correction for multiple tests.

RESULTS

Overview

In total, we recorded 61 bird species, including 33 endemic to the Philippines and 7 listed on the IUCN Red List as threatened or near threatened: Negros Bleeding-heart, Writhe-billed Hornbill, Visayan Hornbill, Yellow-faced Flameback, Negros Jungle Flycatcher, White-winged Cuckooshrike, and Pink-bellied Imperial-pigeon (Ducula poliocephala; NT). All 7 were recorded in primary forest, although the single Writhe-billed hornbill record was outside the 50 m truncation distance. Four of these species were also recorded in secondary forest and none in plantation or open habitat. Nine species (14.8% of those recorded) were unique to primary forest, including Negros Bleeding-heart and Yellow-faced Flameback (and Writhe-billed Hornbill), and two additional Philippine endemics, Brown-breasted Kingfisher (Halcyon gularis) and Yellow-breasted Fruit-dove (Ptilinopus occipitalis; Appendix 1). Two species, Plaintive Cuckoo (Cacomantis merulinus) and Gray-streaked Flycatcher (Muscicapa griseisticta), were only recorded in plantations, and another four were only recorded in open habitat (Chestnut Munia Lonchura atricapilla, Common Kingfisher Alcedo atthis, Gray Wagtail Motacilla cinerea, and Lesser Coucal Centropus bengalensis).

Taxonomic, species, and functional diversity, and endemicity

Estimated site species richness was 62 species (62.3 +/- 84% CI [57.4, 67.2]) with 100% sample coverage obtained with 400 points (99.8% coverage with observed survey effort, n = 278). Observed species richness was highest in primary forest (38 species, 84% CI [30.3, 45.7]) and decreased across a gradient of increasing habitat disturbance (Table 1, Fig. 2). Shared species richness was higher between habitats least separated along a disturbance gradient, i.e. primary and secondary forest (60.9%), plantation and open habitat (46.9%), than habitats most separated, i.e., open habitat and primary (27.7%) and secondary forests (38.1%).

Including habitat as a factor (rather than a single predictor) improved model fit for community measures except FDis and Q (LRT_FDis = 1.498, p = 0.473; LRT_Q = 2.412, p = 0.299). All habitats had a significant effect on species richness and diversity, and endemicity (Table 2). Primary forest was the only habitat that had a positive effect on these measures whereas open habitat had the most negative effect. Species richness was significantly higher in primary forest than secondary forest (Tukey post-hoc pairwise contrasts; z_{PF, SF} = 3.584, p = 0.002) and open habitat (z_{PF, O} = 3.177, p = 0.008). Primary forest also had significantly higher species diversity than open habitat (z_{PF, O} = 3.165, p = 0.009). Endemicity (E) was significantly higher in primary forest than in secondary forest (z_{PF, SF} = 2.812, p = 0.032) and open habitat (z_{PF, O} = 3.834, p < 0.001); mean E_PF = 0.127 ± 0.079 SD, E_SF = 0.103 ± 0.051, E_P = 0.100 ± 0.050, E_O = 0.069 ± 0.051.

Only primary forest had a significant (positive) effect on both FDis and Q (Table 2); open habitat had a negative effect on both functional diversity measures. Highest functional diversity was described in plantations (FDis = 0.209 ± 0.015; Q = 1.59 ± 0.110), then primary (FDis = 0.200 ± 0.011; Q = 1.48 ± 0.077) and secondary forests (FDis = 0.206 ± 0.011; Q = 1.44 ± 0.078), and finally open habitat (FDis = 0.174 ± 0.027; Q = 1.25 ± 0.187), with no significant pairwise differences. Species richness and diversity, and endemicity (E) were highly correlated with FDis and Q across habitats (r_{p Sr-FDis} = 0.617, r_{p H-FDis} = 0.757, r_{p E-FDis} = 0.610; r_{p Sr-Q} = 0.732, r_{p H-Q} = 0.855, r_{p E-Q} = 0.717; p < 0.001 in all cases).

Community compositions

Although overall functional diversity did not differ across habitats, community compositions defined by functional guild were significantly different across the four habitat types (R = 0.142, p = 0.001), with differences detected between all pairwise habitat assemblages. Medium general-level frugivores and small lower-storey insectivores were the greatest contributors to community assemblage similarities within each of the four habitats (Table 3). Small omnivores of lower storeys and small and medium general-level insectivores were important contributors to primary forest community similarities, as were small general-level omnivores to plantation and open habitat similarities.

All guild compositions were at least 67.4% dissimilar between habitats (primary and secondary forest communities), with primary forest and open habitat communities most dissimilar (79.2%; Table 4). Five guilds contributed consistently to these dissimilarities. Of these, the relative abundances of medium general-level insectivores (H₃ = 15.70, p = 0.001), small general-level insectivores (H₃ = 20.03, p < 0.001), small general-level omnivores (H₃ = 114.27, p < 0.001), and small lower-storey omnivores (H₃ = 53.46, p < 0.001) were significantly different across habitats.

On average, eight key species contributed 68.8% to community composition dissimilarities between different habitats (Table 5). All eight species are endemic to the Western Visayas or the Philippines. Visayan Bulbul (Hypsipetes guimarasensis), Visayan Tailorbird (Orthotomus castaneiceps), and Magnificent Sunbird (Aethopyga magnifica) were the three most important species for explaining assemblage dissimilarities. The relative abundances of four species were significantly different across habitat types. Magnificent Sunbird was more frequent in plantations than all other habitats, and in open areas than forest habitats (H₃ = 141.21, p < 0.001); Maroon-naped Sunbird (Aethopyga guimarasensis) was more frequent in forest habitats than plantations and open habitats (H₃ = 50.83, p < 0.001); Balicassiao (Dicrurus balicassius) was more frequent in primary forest than secondary forest (H₃ = 24.97, p < 0.001); and Elegant Tit (Periparus elegans) was more frequent in primary forest than all other habitats (H₃ = 19.97, p < 0.001).

DISCUSSION

Our study is the first to quantify extensively the bird communities of the Northwest Panay Peninsula Natural Park, a KBA (Key Biodiversity Areas Partnership 2022) and IBA (BirdLife International 2024a) of the Western Visayas faunal region. Alongside the autecological work of Mynott et al. (2020, 2021), it confirms the presence of four of the IBA’s five trigger species (BirdLife International 2024a); Negros Bleeding-heart, Visayan Hornbill, Writhe-billed Hornbill, and White-winged Cuckooshrike. Although the final trigger species, Green-faced Parrotfinch, was not detected and has only been recorded at one site on Panay since 2006 (de Soye 2006, eBird 2023), it is generally found > 1000 m and only irruptively in the lowlands (BirdLife International 2024b). Our records of the sparsely distributed White-winged Cuckooshrike (Birdlife International 2016a) in primary and secondary forests are important, while that of Yellow-faced Flameback in primary forest adds to the increasing records of this species on Panay, and in the park particularly (eBird 2023). We also recorded 13 of 14 diurnal species endemic to the Western Visayas (Brinkman et al. 2023) and previously recorded in NPPNP (eBird 2023).

Species richness and diversity decreased along the gradient of habitat disturbance, which is comparable to similar studies in southeast Asia (Thiollay 1995, Waltert et al. 2004), including reduced avian richness in plantation forest (Zurita et al. 2006, Mahiga et al. 2019). The taxonomic diversity we report in natural forests and plantations is similar to that recorded in equivalent elevational forests (Amoroso et al. 2018) and broadly analogous agroforestry communities (Tanalgo et al. 2015) in other faunal regions in the Philippines, although these studies do not report on sample completeness relative to survey effort, which we encourage for the evaluation and comparison of future community studies.

Although secondary forest and plantations may help support a subset of bird diversity (Schultze et al. 2004), there are significant species losses when compared to primary forests in the park and specifically for globally threatened and endemic species. This is despite the fact that these habitats are found in close proximity within the site’s heterogeneous matrix (Lawton et al. 1998). Primary forest supported the highest taxonomic diversity and endemicity and is essential for at least a subset of the park’s bird community; eight species, including three threatened species and one additional Philippine endemic (Yellow-breasted Fruit-dove), were only recorded in undisturbed forest. The most disturbed open areas appear unsuitable for all seven threatened and near threatened species detected as well as nearly 40% of endemic species recorded, which is of significant conservation concern. Any landscape management should not overlook the necessity of focusing efforts on conserving primary forest, even if some key species are at least also detected in secondary forests (Brooks et al. 1992, Kinnaird and O’Brien 2007, Mynott et al. 2021), which itself does not confirm habitat reliance or population viability.

Functional diversity may be a more sensitive metric than taxonomic diversity, which can mask changes in composition and functionality (Lawton et al. 1998), when considering the influence of habitat disturbance (Zurita and Bellocq 2010). We found important differences in both functional diversity and community composition across a gradient of habitat modification that could have consequences for ecosystem functioning (Sodhi et al. 2004) alongside conservation implications for individual species and representative endemicity. For instance, most bird species described as important frugivores in the Western Visayas (Hamann and Curio 1999) were only recorded in forest habitat, including large- (e.g., both hornbill species, Pink-bellied Imperial-pigeon, Yellow-breasted Fruit-dove), medium- (e.g., Black-naped Oriole Oriolus chinensis, Coleto Sarcops calvus), and small-bodied (e.g., Yellowish White-eye Zosterops nigrorum) species. All these species, other than Coleto and Black-naped Oriole, are endemic to the Philippines.

As with previous work (Canaday 1996, Gray et al. 2006, Tvardíková 2010, Mahiga et al. 2019), we found that insectivorous and frugivorous guilds were particularly affected by habitat disturbance. Large general-level insectivores (e.g., Yellow-faced Flameback, White-bellied Woodpecker Dryocopus javensis) were only recorded in primary forest, and large frugivores (e.g., both hornbill species, Pink-bellied Imperial-pigeon, Yellow-breasted Fruit-dove) were detected considerably less frequently in non-primary forest habitat. Additionally, medium upper-storey frugivores (e.g., Philippine Hanging-parrot Loriculus philippensis, White-eared Brown-dove Phapitreron leucotis), small general-level frugivores (Dicaeum spp.), insectivores (Elegant Tit), and lower-storey insectivores (e.g., Visayan Fantail Rhipidura albiventris) appeared especially susceptible to habitat disturbance (Stouffer and Bierregaard 1995, Şekercioḡlu et al. 2002, Lees and Peres 2010, Tvardíková 2010) and the conversion of (any) forest to open habitat. Similar negative responses are exhibited by congeners of the endemic Visayan Fantail and Visayan Shama (Copsychus superciliaris; Mallari et al. 2011, Pabico et al. 2021), classified here as small lower-level insectivores.

Although we recognize bird-habitat associations are often complex (Meents et al. 1983) and may or may not be nonlinear in nature (Radford et al. 2005, Española et al. 2016), the negative responses exhibited may result from food scarcity (Zanette et al. 2000, Mahiga et al. 2019), changes in microclimate (Karr and Freemark 1983, Canaday 1996), habitat structure and specificity (Canaday 1996, Sodhi et al. 2004, Mahiga et al. 2019), or limited dispersal (Stouffer and Bierregaard 1995, Şekercioḡlu et al. 2002, Sodhi et al. 2004). In the case of insectivorous guilds, declines may be linked to changes in microhabitat characteristics or the abundance of food resources (Gray et al. 2006, Mammides et al. 2015) and the specialized nature of these requirements (Sodhi et al. 2008). The greater prevalence of large trees, indicative of primary forest in the park (Mynott et al. 2021), and their provision of suitable understorey microclimate, is an important determinant of preferred habitat for some (closely related) understorey insectivorous species in the region (e.g., White-vented Shama Copsychus niger; Mallari et al. 2011), while the presence of regenerating forest adjacent to primary forest in the park’s heterogeneous matrix may facilitate recovery of forest insectivores in the landscape (Stouffer and Bierregaard 1995). This reduction in large trees and associated changes in floristic resources may also explain declines in large frugivores across the habitat disturbance gradient (Española et al. 2016). These larger-bodied species may be more vulnerable to habitat change due to their occurrence at naturally low densities (Sodhi et al. 2004). Consequently, there is a need for larger areas of suitable primary forest and stepwise improvements in habitat quality for their sustained conservation (Española et al. 2016).

At the species level, our findings support current understanding of some broad habitat associations. For example, Negros Jungle Flycatcher is reported in disturbed forest (Brooks et al. 1992, BirdLife International 2016a); Pink-bellied Imperial-pigeon is a species of lowland primary forest and old secondary growth (BirdLife International 2016b); Negros Bleeding-heart is only recorded in primary forest (Brooks et al. 1992, Mynott et al. 2020); and Visayan Hornbill is recorded at significantly higher density in primary forest than in secondary forest (Mynott et al. 2021). Our records of Yellow-faced Flameback only in primary forest emphasize the need to evaluate the specific habitat requirements of this threatened, priority woodpecker (Lammertink 2014).

Both hornbill species (large frugivores) are considered at high risk of extinction based on the fragmented forest habitat configuration within their limited geographic ranges (Kinnaird and O’Brien 2007). Despite the recent suggestion that Writhe-billed Hornbill is locally extirpated in NPPNP (Reintar et al. 2022), our single record of the species (in primary forest) further confirms the importance of this protected area. Although we do not evidence breeding in the park (and there are no breeding records since 1997; BirdLife International 2018) or the park as a source population, it serves to emphasize the potential importance of a site containing considerable primary forest cover characterized by ~3x more trees of nest cavity size (Klop et al. 2000) than in secondary forests (Mynott et al. 2021). It also highlights the importance of ensuring site connectivity across the landscape for a species that may be nomadic in response to spatio-temporal variation in fruit availability (BirdLife International 2018), increasing their vulnerability to habitat change (Sodhi et al. 2004), and for which the current protected area system remains inadequate (Kinnaird and O’Brien 2007). Enhancing ecological connectivity with the Central Panay Mountain Range, where there appears to be a healthy breeding population of the Writhe-billed Hornbill (A. Alabado, R. Lestino, J. Venus, M. Ibabao, T. Kuenzel, and E. Curio 2009, unpublished report), and ensuring the park is not isolated by > 10 km for the Visayan Hornbill and > 20 km for the Writhe-billed Hornbill, enabling dispersal between forest patches in the wider landscape, will contribute to future proofing both hornbill species and the ecosystem services they provide as seed dispersers (Kinnaird and O’Brien 2007), providing hunting pressures are also addressed synchronously (E. Klop, 1998, unpublished manuscript). This concern may contribute to national discussions on the Philippines’ commitment to the targets of the High Ambition Coalition for Nature and People’s 30 x 30 initiative, especially when only 15% of Philippine forests are currently protected (Wolf et al. 2021).

Our results indicate an expected richness of 62 species in the park and includes detecting 85–92% of forest species, which suggests there may be 3–6 diurnal forest species that remained undetected in the specific survey season. Despite this, eBird records report 84 diurnal species at the same time of year in the park (eBird 2023). Although such bird lists may result from a spatially targeted approach to recording specific species, including the use of playback, and maximizing species detection (Watson et al. 2018), this discrepancy in species richness highlights some considerations regarding survey implementation.

Some of our surveys were restricted by the severity of the local terrain where point transects could not be accessed safely and/or surveyed around dawn (e.g., Lynch 1995, Bibby et al. 2000, Lee and Marsden 2008). Temporally then, some species may have been under recorded, e.g., Negros Jungle Flycatcher, or gone undetected. For example, we did not detect Gray-breasted Brown-dove (Phapitreron amethystinus maculipectus), a near threatened Philippine endemic generally found above 1000 m, although recorded below 500 m (del Hoyo et al. 2020a). Although reported as “numerous in March/April” in the park in 2014 (BirdLife International 2023), there is only one eBird record since 2016 (eBird 2023). Similarly, we failed to detect Visayan Rhabdornis (Rhabdornis rabori), which is a scarce and local resident of submontane and montane forests above 800 m (del Hoyo et al. 2020b), although our sampling effort included only one point in this elevational range.

Implementing complementary survey approaches that can capture peaks in all species’ detectability and without spatial restrictions will help circumvent local challenges with accessing specific locations for conducting point transect surveys immediately following sunrise (e.g., Lynch 1995, Bibby et al. 2000). Consequently, we recommend the systematic inclusion of passive bioacoustic recording to increase the likelihood of detecting all species (e.g., Metcalf et al. 2022) and support robust, constant-effort surveys necessary for monitoring population trends and conservation planning (Thompson 2002). Caveating this is the understanding that aural detection and spatial coverage for some species may be comparatively small, e.g., Negros Jungle Flycatcher, which is only audible “at close range” (Clement 2020), which may contribute to it being under recorded (eBird 2023).

Our survey effort was sufficient in capturing most species of the avian assemblage in primary forests, but less so in more disturbed habitats. Although this may relate to inter-habitat differences in daily patterns in bird activity and, hence, detectability (Fontúrbel et al. 2021), our sampling design minimized any temporal detectability differences across habitats (Verner and Ritter 1986). Instead, the higher proportion of unique species (those only detected from a single point count) recorded in open habitat (63.6%) compared to primary forest (18.4%) may contribute to lower estimated sample coverage and more variable measures of observed and expected species richness for this habitat (Chao and Jost 2012, Chao et al. 2020). To obtain similar sample completeness in plantations and open habitats requires 2.0 and 3.1 times more survey effort, respectively. This is unlikely in the park in which our lower survey effort in these habitats proportionally reflected the smaller areas of these habitats in the landscape. Including surveys of representative plantations and open habitats adjacent to the park may then contribute to fully capturing species richness with improved precision and provide a more robust temporal dataset for monitoring bird communities in these anthropic habitats. Although reducing survey effort in natural forests in the park can support more cost-effective, long-term monitoring (Thompson 2002), we are cautious in making such a recommendation if the objective is to monitor all species and entire communities because not all forest species were detected. However, 75 and 83 points in primary and secondary forests, respectively, still captured 90% of predicted diurnal species present, while reducing survey effort by 22–36 points, or 2–4 days surveying, in each forest type, and this may be a sufficient subset of species richness for conservation monitoring objectives (Vellend et al. 2007).

Since 2001, the park has lost 2.2% of its overall tree cover (249 ha; Global Forest Watch 2023a, b) including 1.3% (49 ha) of remaining primary forest (Hansen et al. 2013). This rate of primary forest loss is similar to, but still higher than the surrounding areas (mean = 1.1 ± 0.41%), whereas overall tree cover loss is noticeably higher outside the park (mean = 4.6 ± 1.93%; Global Forest Watch 2023a, b), which is comparable to the national trend of higher forest loss outside protected areas (Wolf et al. 2021). This suggests that despite its protected status, primary forest in the park is as vulnerable to clearance as unprotected primary forest remnants outside the park’s boundaries. Specifically, the southwest edge of the park is a sporadic hotspot of primary forest loss (defined as on-off annual activity, with < 12 of the last 14 years showing significant forest loss; Harris et al. 2017). Although overall tree cover loss is lower in the park, our findings demonstrate the importance of primary forest over secondary habitats in maintaining avian endemicity and supporting globally threatened species in the heavily impacted Western Visayas faunal region (Lorica and Oliver 2006). To this end, it is essential to prevent further primary forest losses and conversion through the main drivers of commercially driven and, depending on permanence and intensity, shifting agriculture (kaingin; Curtis et al. 2018, BirdLife International 2024a) while also considering empirically the potentially synergistic impacts of illegal hunting (e.g., Sathya Chandra Sagar et al. 2023). Although there is no current forest restoration in place in the park or environs, accompanying passive (e.g., natural regeneration of secondary forest) and/or active (e.g., tree planting strategies in degraded areas) restoration approaches (e.g., Chazdon 2019) will improve the condition and area of crucial forest habitat for key bird species and support wider ecosystem co-benefits in this conservation priority landscape (BirdLife International 2024a).

CONCLUSIONS

The Northwest Panay Peninsula Natural Park is one of the few protected areas in the highly deforested Western Visayas with intact lowland primary forest cover remaining (Lorica and Oliver 2006), and in a region where most primary forest is now at submontane and montane elevations (Hamann et al. 1999). Our findings highlight that declines in the extent of primary forest (or forest quality) will lead to reductions in community diversity and regional endemicity, including those species of most conservation concern, within the park’s heterogeneous habitat matrix. This is potentially indicative of the wider region and emphasizes the high conservation value of primary forest for avian diversity, however defined, and endemicity, paralleling that of other faunal regions in the Philippines (Mallari et al. 2011). The different responses of functioning guilds to habitat disturbance suggests this affects the trophic organization of bird communities and, consequently, their ecosystem functioning (Gray et al. 2006). We also demonstrated the need for robust ecological understanding and consideration of different species groups when designing conservation strategies that ensure intact, functioning communities are maintained (Canaday 1996) and support protected area management strategies (Sodhi et al. 2004, Mallari et al. 2011). Although the park remains threatened by small-scale selective logging, conversion to agriculture (e.g., coconut groves) and hunting (Kerhoas et al. 2019, Mynott et al. 2020), adjacent local communities have a high level of awareness of the importance of local endangered species, but little monetary capacity to help actively conserve them (Subade and Jugado 2010). Based on our study and previous recent work (Mynott et al. 2020, 2021), we believe that an effective future-proofed conservation strategy for bird populations of this protected area, and a key representation of the wider Western Visayas, requires strong collaboration between city and provincial Environment and Natural Resources offices and local communities to protect primary forest areas from anthropogenic threats, to enable secondary forest areas to recover, and to support measures that retain and enhance connectivity with forest in the wider landscape.

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