1 INTRODUCTION

Recent decades have brought about a significant negative impact on the environment due to anthropogenic pressures such as land use change, overexploitation/over-harvesting, pollution, and climate change, all of which have contributed to a massive global loss in biodiversity (Ceballos et al., 2020; Dirzo et al., 2014). The expansion of land conversion for agriculture and livestock production has led to exponential losses of primary vegetation and natural habitats, which in turn has drastically reduced species' population sizes and overall biodiversity (Curtis et al., 2018; Lanz et al., 2018; Steffen et al., 2015). The scope and magnitude of the increasing human footprint (HFP) (i.e., human population growth which will increase demand and consumption of natural resources at both regional and global scales) threatens ecosystems and accelerates the global mass extinction crisis (Ceballos et al., 2017; Jaureguiberry et al., 2022). Global policy frameworks aimed at mitigating negative effects on biodiversity (e.g., the United Nations Sustainable Development Goals, post-2020 Global Biodiversity Framework) should be based on sound scientific knowledge to understand the value of human-altered ecosystems. This understanding will help address the new challenges in sustaining the health of biodiversity and ecosystem functions (Chan et al., 2023; Leadley et al., 2022; Leclère et al., 2020).

Among terrestrial mammals, non-human primates (hereafter primates) are a diverse and widespread taxonomic group. Critically, they play an important role in regulating community-wide ecological functions, processes, and services (e.g., seed dispersal, pollination, and predator–prey relationships) that promote forest regeneration and carbon sequestration, and provide benefits to local human populations (Chapman et al., 2013; Estrada et al., 2017; Galán-Acedo, Arroyo-Rodríguez, Andresen, et al., 2019; Galán-Acedo, Arroyo-Rodríguez, Cudney-Valenzuela, et al., 2019). Relative to most mammals, primates typically have slow life histories, are gregarious, and have larger body masses, all of which, unfortunately, are traits that make them particularly vulnerable to extinction (Estrada et al., 2017; Purvis et al., 2000). Threats to primates are diverse and include anthropogenically driven factors related to increasing human population growth rates, such as habitat fragmentation and degradation, agricultural expansion, and expanding road and rail networks (Ascensão et al., 2021, 2022; Branch et al., 2022; Estrada et al., 2020). These pressures often lead to direct (e.g., hunting and wildlife trade) and indirect (e.g., loss of genetic diversity) mortalities, ultimately contributing to population declines and local extirpations, increasing species exposure to threats and extinction risks (de Almeida-Rocha et al., 2017; Estrada et al., 2017; Fa et al., 2003). Approximately 68% of all primate species have been listed as threatened (i.e., Critically Endangered [CR], Endangered [EN], or Vulnerable [VU]) by the International Union for the Conservation of Nature (IUCN, 2022) and are at risk of local or range-wide extinction.

Due to their diversity and richness, primate species maintain various ecological niches, thus their ability to cope with anthropogenic pressures is equally diverse (Galán-Acedo, Arroyo-Rodríguez, Andresen, et al., 2019; Galán-Acedo, Arroyo-Rodríguez, Cudney-Valenzuela, et al., 2019; Machado et al., 2022; Mittermeier et al., 2013). Whereas some primates might have little tolerance for anthropogenic pressures and cannot persist in modified habitats, others may be more ecologically flexible, allowing them to persist in anthropogenic landscapes (Eppley et al., 2020; Galán-Acedo et al., 2021; Hoffman & O'Riain, 2012; Siljander et al., 2020). Regardless of the ability to cope within anthropogenic landscapes, both protected areas (PAs) and Indigenous Peoples' lands (IPLs) tend to provide stable habitats for the conservation of primate species and other terrestrial vertebrates (Estrada et al., 2022; Torres-Romero et al., 2020; Torres-Romero & Giordano, 2022), but PAs are not always effective for primate conservation (see e.g., Appleton et al., 2022; Chapman & Peres, 2021; Hoskins et al., 2020; Junker et al., 2020; Nijman, 2005). Despite these species' strongholds, however, our overall lack of knowledge of how anthropogenic variables differentially impact primate diversity at different spatial scales limits efforts to manage their long-term viability (Garnett et al., 2018; Junker et al., 2020).

While the impact of the ongoing sixth mass extinction on earth's biodiversity, including primates, has been explored (see e.g., Ceballos et al., 2017, 2020; Estrada et al., 2017, 2022), only a few studies have evaluated the exposure risk of primate ranges to land use and climate changes (e.g., Carvalho et al., 2019; de Almeida-Rocha et al., 2017; Fernández et al., 2022). However, there is currently a lack of detailed analyses that explore the influence of global- and regional-scale landscape modifications on primate diversity, population trends, and threatened species status. Identifying the effects of anthropogenic pressures on species richness spatial patterns may assist in establishing targeted conservation strategies. Such information may better facilitate the development of regional and transnational conservation actions aimed at mitigating extinction risk across increasingly human-dominated landscapes. Here, we employ a path analysis framework to contrast the dependent and independent associations of human-altered landscape variables, as well as PAs, and IPLs, on primate diversity across and between large-scale spatial gradients, that is, major biogeographic regions (Table 1). Using the most recent IUCN Red List assessments (IUCN, 2022), we specifically examined the following: (i) diversity of primates (i.e., overall species richness), (ii) number of threatened primate species (i.e., CR, EN, and VU) versus non-threatened species (i.e., LC = Least Concern, NT = Near Threatened), and (iii) primate species with a “decreasing” population trend versus non-decreasing (i.e., stable and/or increasing population).

2 METHODS

The order Primates (including strepsirrhines, tarsiers, monkeys, and apes) is the third most specious mammal radiation, constituting approximately 520 extant taxa (IUCN, 2022). They are represented by 16 families within four major geographic regions: (1) The Americas (Aotidae, Atelidae, Callitrichidae, Cebidae, and Pitheciidae), (2) Africa (Cercopithecidae, Galagidae, Lorisidae, and Hominidae), (3) Madagascar (Cheirogaleidae, Daubentoniidae, Indriidae, Lemuridae, and Lepilemuridae), and (4) Asia (Cercopithecidae, Hominidae, Hylobatidae, Lorisidae, and Tarsiidae).

3 RESULTS

3.1 Global primate species patterns

According to our compiled data, primate species richness was highest in the Americas (n = 178), followed by Asia (n = 129), Africa (n = 107), and Madagascar (n = 106) regions (Figure 1). When considering threatened species (i.e., CR, EN, and VU) and species with declining populations, the highest richness occurred in Asia, followed by Africa, Madagascar, and the Americas (Figure 1). This same pattern was also observed for trade and trafficking (Figure 1). Furthermore, our results indicated that 84% of primate species are affected by wildlife trade. In fact, primate species that are traded were listed in higher IUCN Red List threat categories and constitute more than half of all known primate taxa (Figure 2b).

Of all primate taxa we examined (n = 520), 67% (n = 348) were listed by the IUCN as threatened (i.e., CR, EN, or VU), ~30% (n = 158) as non-threatened (i.e., LC and NT), and only ~3% (n = 14) of data are listed as Data Deficient. Approximately 86% (n = 446) of primate taxa have declining populations, while ~6% (n = 30) had stable and/or increasing populations, and ~8.5% (n = 44) are listed as unknown (Figure 1). We also found similar patterns in the Americas, Africa, Madagascar, and Asia, where the proportion of species threatened are greatest compared to non-threatened, as well as the primates with declining populations showed a proportion more pronounced versus those primates that exhibit stable or increasing populations (Figure 1).

Our dataset shows that primate population trends are heavily declining, but the percentage greatly varies across families, for example, Cercopithecidae (143 spp.), Cheirogaleidae (40 spp.), Callitrichidae (39 spp.), Pitheciidae (38 spp.), Cebidae (27 spp.), Lepilemuridae (25 spp.), and Atelidae (22 spp.) (Figure 2a). Furthermore, among 86% of primate taxa with declining populations, 77% were listed as threatened (i.e., CR, EN, or VU), 21% were non-threatened (i.e., LC and NT), and only 2% were classified as Data Deficient. For primate taxa with stable population trends (~5.5%), 50% were categorized as threatened (i.e., CR, EN, or VU) and 50% were non-threatened (i.e., LC and NT). Unfortunately, only about ~0.5% of the primates studied were estimated to have an increasing population trend, with 7% of these listed as threatened and 83% as non-threatened, while 8.5% were listed as unknown. Among those with unknown trends, 2% were listed as threatened, 85% as non-threatened, and 13% were classified as Data Deficient.

3.2 Path analysis models

Based on our SEMs analyses, we found that PAST, HFP, and ROAD were the most important factors explaining primate species richness patterns on a global scale; all three of these variables had a strong direct negative effect on primate richness (Figure 3). Similarly, HFP, PAST, and CROP had strong direct negative effects on both threatened and non-threatened species richness (Figure 3a,b). For primate species with decreasing population trends, the same variables (i.e., HFP, PAST, and CROP) exhibited direct negative effects, while PAST and CROP also showed a direct negative effect for species with non-decreasing populations (Figure 3c,d). In contrast, PAs emerged as important positive predictors for species that are threatened, and species with declining populations, but IPLs had a negative association for non-decreasing species (i.e., stable and increasing populations) (Figure 3a,c,d). Globally, this should be interpreted as both large and small patches having a disproportionately high value in protecting the world's primate fauna.

Across the Americas, we found similar patterns where the pressures of humans (i.e., roads, HFP, and CROP) had the most significant direct negative impact on species richness patterns (Figure 4). Again, for threatened and non-threatened species richness, as well as for primate species populations (both declining and stable/increasing), we also found the same pattern, where the expansion of human activities (i.e., roads, HFP, and PAST) showed direct negative effects (Figure 4a–d). However, we found that landscape areas considered to be IPLs had negative and positive associations on threatened and non-threatened species richness, respectively (Figure 4a,b). Specifically, this means that across the region, small IPLs tended to protect more threatened species than larger ones. Concerning primate species populations (both declining and stable/increasing), PAs coverage, had a direct positive impact (Figure 4c,d).

For Africa, PAST and CROP had the most direct negative effect on species richness patterns, whereas IPLs and PAs showed a negative and positive association, respectively (Figure 4). For threatened and non-threatened species, as well as for primate species populations (both declining and stable/increasing), human activities had direct negative effects on spatial patterns of these primate species (Figure 4e–h), whereas PAs emerged as an important predictor with s direct positive effect for non-threatened species (Figure 4f). In addition, IPLs had a direct negative effect on primate populations (both those decreasing and stable/increasing), however, PAs also emerged as an important predictor for both primate population categories (i.e., decreasing and stable/increasing) with positive and negative effects, respectively (Figure 4g,h).

For Madagascar, PAs emerged as the single strongest predictor with a direct positive effect on primate species richness (Figure 5). HFP, PAST, and ROAD infrastructure, followed in importance as negatively impacting species richness. It should be noted that there are no areas classified as IPLs in Madagascar (Figure 5). Furthermore, the path analysis included all Madagascar primate because 95% and 100% are threatened with extinction and with a “decreasing” population trend, respectively.

Within the Asia region, IPLs and PAs were the most important factor predicting higher primate species richness, as both ecological landscapes had direct positive effects; whereas human activities (i.e., HFP, CROP, and ROAD) exhibited negative direct effects on richness (Figure 5), with strong direct negative effects for threatened and non-threatened species, as well as for declining, stable, and increasing primate populations (Figure 5a–d). Indeed, IPLs and PAs emerged as the strongest predictor effects for all primate species categories and for declining, stable, and increasing primate populations, but IPLs had a negative association for stable and increasing primate populations (Figure 5a–d). This means that those small IPLs and large PAs tended to protect more stable and increasing (i.e., non-decreasing) species compared to larger IPLs and smaller PAs (Figure 5d).

4 DISCUSSION

Our results provide quantitative evidence for the various ways that human impacts and landscape conditions influence the global distribution patterns of primates' species richness. Regions experiencing high rates of land transformation through direct human interactions (e.g., development and expansion of various infrastructures, agriculture, etc.) showed pronounced negative effects on primate species richness patterns. In fact, our results showed that these negative effects were for both species with decreasing population trends and species with stable populations, revealing the critically dire situation for primates worldwide. Although recent modeling projections by Carvalho et al. (2019) suggested that approximately a quarter of Asian and African primates will experience extensive agricultural expansions within their ranges that lead to a 98% increase in disturbed habitats, our study indicates that the presence of small or large core areas can serve as potential refuges (i.e., IPLs or PAs) as these were directly associated with primate richness. Previous studies have demonstrated the importance of protected habitats which, for example, may buffer animal populations against the impact of anthropogenic pressures by providing essential resources, including food, suitable resting/sleeping locations, and breeding opportunities (Eppley et al., 2022; Estrada et al., 2018; Laurance, 2005). Although many protected forest lands are experiencing significant degradation due to human activity (Chapman & Peres, 2021; Laurance et al., 2012), global species richness patterns are shaped by the presence of IPLs, showing that they play an important role in determining the long-term persistence of primate species (Corrigan et al., 2018; Estrada et al., 2022), as well as threatened species versus non-threatened, and decreasing versus non-decreasing, both globally and regionally (Bruner et al., 2001; Griffiths et al., 2023).

In Africa and Madagascar, IPLs and PAs cover approximately 6.6 and 4.5 million km², respectively (Estrada et al., 2022). However, our results show that overall; primate species are at risk of extinction from the rapid and relatively recent expansion of the HFP. This appears to be a response to not only local but also global economic demands, which are accelerating the expansion of agricultural and livestock activities. In turn, this results in deforestation and creates a potent economic impetus for constructing roads through forested areas, driving massive biodiversity loss (Ascensão et al., 2022; Estrada et al., 2018). Other studies have found that the expansion of land use changes frequently contributes to significant declines in many African mammal populations, including primates (Carvalho et al., 2019; Chapman et al., 2006; Di Marco et al., 2014; McKinney et al., 2023). Similarly, Morelli et al. (2020) reported on the severe effect of human activities on rainforest habitat losses, specifically, that the eastern rainforest habitat of Madagascar could be almost entirely lost before the year 2080. In fact, many lemur taxa are unable to cope with varying degrees of habitat fragmentation, so as viable habitat contracts; the resulting extirpation may be widespread (Eppley et al., 2020; Mittermeier et al., 2023; Ralimanana et al., 2022; Schwitzer et al., 2014).

Within the Americas, we found that increased land use is driven by the global need to feed and support a rapidly expanding human population, which negatively impacts primate species (Estrada et al., 2012, 2020). Thus, the expansion of pasture and cropland degrades and fragments forest habitat, causing significant primate range contractions and population declines. Furthermore, primate populations living in these relatively well-preserved habitats, for instance, the Amazon rainforest, will have additional challenges as anthropogenic pressures increase, such as accessing limited resources, exposure to novel zoonotic diseases, and predation (Estrada et al., 2012, 2018; Fernández et al., 2022; Han et al., 2016). Similar to other regions, large forested areas, including the Amazon, are important for primate diversity and have been well demonstrated (Gouveia et al., 2014).

Throughout the Americas, IPLs and PAs cover approximately 3.0 and 5.2 million km², respectively (Estrada et al., 2022). We found that IPLs are critical to protecting higher numbers of threatened primates, despite the reality that these forests are increasingly the most vulnerable to climate change, agriculture conversion and development, and wildfire (Feng et al., 2021; Garnett et al., 2018; Zhang et al., 2019). In contrast, we did not find an effect of PAs on primate species richness, suggesting that in most cases, in the Americas current PA management and law enforcement are insufficient for effective wildlife protection, and/or that PAs have only limited overall with primate population distributions (Agostini et al., 2022).

Within Asia, IPLs and PAs formally manage approximately 4.4 and 1.0 million km² respectively, and have generally been considered critical to the sustainability and protection of ecosystems, the conservation of global biodiversity, and climate change mitigation strategies (Estrada et al., 2022; Garnett et al., 2018; Torres-Romero et al., 2020; Zhang et al., 2021). Indeed, our results for Asia indicate that IPLs were positively associated with all primate species, including threatened and non-threatened primate richness, and both declining and non-declining species. Importantly, HFP had a strong direct negative effect on these same three primate groupings, except for non-declining species. However, the cumulative pressure on IPLs from land use change and encroachment (e.g., from roads, expansion of pasture and cropland, etc.) may further compromise what remains of these important landscapes and increase the extinction risk of primates (Carvalho et al., 2019; Kennedy et al., 2019; Lechner et al., 2018).

Overall, while we found positive effects of IPLs and PAs on primate conservation globally, the regional variation of efficacy highlights the need for evidence-driven, bottom-up approaches when gazetting land and developing management plans for PAs (Oldekop et al., 2016). According to Estrada et al. (2022), approximately 71% of primate taxa worldwide have distributions that overlap with IPLs. Thus, the early involvement of local communities in identifying areas to be protected, while integrating their use of natural resources into effective and sustainable management practices, not only provides positive effects on wildlife conservation but also ensures the long-term support of these communities (Estrada et al., 2022; Wali et al., 2017). It should be noted, however, that while our results revealed a positive effect of PAs on primate diversity, a recent global analysis showed that compared to unprotected areas, PAs were ineffective at reducing human land conversion pressures (Geldmann et al., 2019). In fact, the number of personnel employed by PAs globally, as well as their enforcement capacity, are insufficient for effectively safeguarding these areas and their biodiversity (Appleton et al., 2022). These trends coupled with additional climate change and land conversion projections (Carvalho et al., 2019) underscore the critical urgency with which mitigation measures for conserving primates need to be implemented.

Lastly, we found that trafficking and trade (e.g., in parts and live individuals) are important and unsustainable threats that have increased in recent years (Nijman et al., 2011; Norconk et al., 2020). Primates continue to be traded in large volumes but the effects on wild populations often remain unknown, and what species are traded and in what volumes change over time. Much of this trade is for domestic markets. Thus, in Indonesia, over a 25-year period, Nijman et al. (2017) reported on the trade in ~2500 live primates of 17 species for sale as pets in the animal markets with clear changes over time (e.g., fewer orangutans, more slow lorises). For Liberia, Covey and McGraw (2014) estimated that close to 10,000 primates were traded a year for wildmeat in one district in eastern Liberia, and Fa et al. (2006) estimated that ~150,000 primates were traded a year at 100 sites in Nigeria and Cameroon. Several primates are traded for medicinal purposes, for some species are clearly at unsustainable levels. Nijman et al. (2014) showed that in one market in Myanmar annually ~1000 Bengal slow lorises (Nycticebus bengalensis) were killed and traded for their parts to be used as medicine. While these reports are all from physical markets, the primate trade increasingly has moved online (Nijman et al., 2021; Siriwat et al., 2019), and here also, the effect this trade may have on wild populations (if any) remains unclear.

Primates are also traded internationally, some of them illegally. For example, ~4000 wild-caught night monkeys (Aotus nancymaae, A. vociferans, and A. nigriceps) from Brazil–Colombia–Peru tri-border area were illegally sold to a biomedical research company in 2007–2008 (Maldonado et al., 2009). Much of the international live primate trade involves captive-bred individuals that are mostly bred in countries to which they are not native (Hansen et al., 2022; Nijman et al., 2011).

5 CONCLUSION

Overall, our results show that human-altered landscapes, due to the expansion of road networks, urban areas, CROP, PAST, and human population densities, and the projected uncontrolled future growth of these aforementioned factors, are accelerating the global primate extinction crisis (Chapman & Peres, 2021; Estrada et al., 2017, 2019; Fernández et al., 2022). It is likely that if current deforestation and land conversion trends continue (FAO, 2016), primate species strongly associated with the integrity of forest wildlands may become locally extirpated or even extinct. Alarmingly, IPLs are also at risk due to urban expansion (Kennedy et al., 2022), but those that still persist are uniquely suited to preventing the extirpation of primate-rich areas (Estrada et al., 2022; Fa et al., 2020).

While seemingly dire, our findings provide a unique understanding of how human-altered landscape variables influence primate species richness across global anthromes. To curtail trends, we advocate for community-led PAs and IPLs that integrate the use that local human populations make and need of those habitats, creating sustainable livelihoods (e.g., Bersacola et al., 2023; Eppley et al., 2023), as well as stronger governance to better implement environmental policies on both PAs and unprotected areas, ensuring they provide long-lasting benefits to both, wildlife and people.

CONFLICT OF INTEREST STATEMENT

The authors declare no conflict of interest.