Continuous Degradation of Forest Structure and Composition Causes a Shift in Phorophytes and Population Genetics in Aphyllous Vanilla Species

Abstract

Background and Research Aims

The natural habitats of aphyllous vanillas in Madagascar have been extensively degraded by human activities. The implications of this degradation are complex given their intricate interactions with other organisms. This study highlights possible approaches to the conservation of aphyllous vanillas after examining the impact of forest degradation on phorophyte diversity, reproductive modes, and genetic diversity.

Methods

We conducted comparative studies of conserved and degraded forests in two contrasting bioclimatic zones. We characterized the changes in horizontal and vertical forest structures and assessed the ecological importance of the tree species. We identified the preferred phorophytes of two aphyllous Vanilla species, Vanilla madagascariensis and Vanilla decaryana, and investigated their population genetics.

Results and Conclusions

Our findings confirmed a declining trend in tree diversity, number of trees, and cover rate in degraded forests. Canopy openness is associated with an increase in the number of phorophytes and the asexual reproductive mode of Vanilla species. Dominant tree species were more likely to become phorophytes, and there was a positive correlation between phorophyte diameter and number of hosted vanilla individuals. Significant deviations from the Hardy-Weinberg equilibrium were observed in each population, particularly in the dry forest and degraded sites, compared with the conserved sites. Additionally, slight differentiation was observed between vanilla populations in conserved and degraded forests, with an alarming effective population size (Ne) that did not exceed 13 individuals.

Conservation Implications

We recommend the introduction of new genotypes to promote genetic enrichment, the ban on logging potential phorophytes, and the inclusion of some degraded forests comprising Vanilla species in conservation.

Keywords

climbers deforestation genetics orchids reproduction vanilla

Introduction

Plant taxonomic groups respond differently to anthropogenic disturbances across ecoregions (Irwin et al., 2010; Sfair et al., 2018), causing notable changes in the forest structure and composition (Haq et al., 2019). These alterations significantly affect plant pollination by modifying the microenvironment and pollinator availability and behavior (Torres-Vanegas et al., 2021; Xiao et al., 2016). Consequently, deforestation, which is the conversion of a forest into different land-use types or permanent reduction of tree canopy cover below the 10% threshold (Carter et al., 2017; FAO, 2015), can decrease plant reproductive success (Torres-Vanegas et al., 2021). Numerous studies have indicated that forest fragmentation into smaller, isolated patches surrounded by non-forest land, significantly affects the genetic diversity of plant and animal populations (González et al., 2020; Schlaepfer et al., 2018). Fragmentation leads to reduced population sizes and changes in species richness and composition, and creates residual habitat patches of varying shapes and sizes that disrupt population dynamics (Schlaepfer et al., 2018; Torres-Vanegas et al., 2021). The spatial isolation of populations theoretically erodes genetic variability and increases genetic divergence due to enhanced random genetic drift, high inbreeding, and reduced gene flow between populations (Cheptou et al., 2017; Schlaepfer et al., 2018).

Over the past 60 years, Madagascar's forest cover has declined drastically, leaving only 15% of its land forested (Bond et al., 2008; Harper et al., 2007; Vieilledent et al., 2018). Ecological degradation is primarily driven by economic activities, poverty, and population growth. Poverty forces many people to rely heavily on natural resources for survival, which, in turn, weakens livelihoods and exacerbates poverty (Barbier & Hochard, 2018; Miyamoto, 2020). Key activities contributing to forest degradation, which is a progressive decline of forest structure and composition resulting in a loss of functions (Vásquez-Grandón et al., 2018), include selective logging, targeting of commercially valuable tree species, and slash-and-burn agriculture, which involves clearing unmarketable woody vegetation (Kaneko et al., 2014; Ralimanana et al., 2022; Zaehringer et al., 2015).

Madagascar harbors a significant diversity of orchids, with 1002 species, including several in the genus Vanilla (Cribb & Hermans, 2009; Goodman, 2022). Aphyllous Vanilla species, wild relatives of cultivated vanillas, exhibit drought adaptation (Botomanga et al., 2024). Conserving wild vanilla resources and recognizing their genetic value are crucial for addressing the current agronomic challenges (Flanagan et al., 2018; Watteyn et al., 2023). Some species are used in traditional medicine as aphrodisiacs (Rakotoarivelo et al., 2019; Randriamiharisoa et al., 2015), raising concerns regarding resource depletion due to overharvesting. Aphyllous Vanilla species are semi-epiphytes that initially germinate in the soil before growing on phorophytes, which are host trees that serve as supports for epiphytic plants (Derzhavina, 2019). The flowers attract bees for potential pollination (Gigant et al., 2014, 2016; Petersson, 2015). Similar to many lianas, they can act as bridges for primates and arboreal animals (Gerolamo et al., 2022; Rowe, 2018). Although epiphytic orchids use trees as phorophytes, not all trees serve effectively as supports (Clemente-Arenas et al., 2023; González-Orellana et al., 2024; Shen et al., 2022). Such selective behaviors can significantly affect the distribution and abundance of epiphytic populations (Cortés-Anzúres et al., 2020; Gowland et al., 2011; McCormick & Jacquemyn, 2014; Rubio-Salcedo et al., 2015; Wagner et al., 2015). The rough bark of phorophytes offers better support and moisture retention, thereby enhancing the germination rates (González-Orellana et al., 2024; Timsina et al., 2016; Zarate-García et al., 2020). Preferences may also be linked to specific fungi involved in orchid mycorrhizal associations (Pecoraro et al., 2021; Rasmussen et al., 2015; Wang et al., 2017).

Vanilla species typically reproduce through a mixed mating system involving vegetative propagation and pollinator-mediated sexual reproduction (de Oliveira et al., 2022; Grisoni & Nany, 2021). In Puerto Rico, the clonality rates for the three Vanilla species ranged from 6 to 38% (Nielsen & Siegismund, 1999). For Vanilla humblotii in Mayotte, the rate was 12.5%, while seven aphyllous Vanilla species in Madagascar averaged 9% (Andriamihaja et al., 2022; Gigant, De Bruyn, et al., 2016a). Forest degradation may enhance the vegetative propagation of vanilla vines due to increased light access. Canopy openness significantly affects liana density (Gianoli et al., 2012; Miranda et al., 2022; Schnitzer et al., 2021), and co-occurrence of liana species (Silva et al., 2016). Consequently, forest degradation could influence liana reproductive modes, justifying the investigation of the impact of canopy openings on the development and reproduction of aphyllous Vanilla species.

Highly polymorphic markers have revolutionized population genetics research (Garrido-Cardenas et al., 2018; Nadeem et al., 2018). Vegetative reproduction, often linked to repeated multilocus genotypes (MLG) (Stoeckel et al., 2021a), enhances clonality estimation accuracy. F-statistics (Wright, 1949), widely used in population genetics (Cui et al., 2022; Stoeckel & Masson, 2014; Zimmerman et al., 2020), describe gene flow and breeding structures, especially deviations from random mating. The fixation index of individuals in subpopulations (FIS) assesses genetic divergence within a subpopulation relative to the total population, ranging from -1 to +1. Negative values signify a surplus of heterozygotes, positive values indicate a heterozygote deficit, and zero denotes the Hardy-Weinberg equilibrium, where the genetic distribution is considered stable (Arnaud-Haond et al., 2007; Mishra et al., 2021). Essentially, FIS reflects intra-population differentiation. FST, which ranges from 0 to 1, measures the proportion of the total genetic diversity attributable to differences between subpopulations. Higher FST values indicate greater genetic divergence among the subpopulations (Arnaud-Haond et al., 2007; Mishra et al., 2021).

The effects of habitat degradation on population genetic diversity have been studied in both conserved and degraded environments (Aguilar et al., 2019; Chung et al., 2014; Depecker et al., 2023). However, these studies are often difficult because of the limited availability of the pristine sites. Traditional societies protect sacred natural sites from various threats (Frascaroli et al., 2019; Ormsby, 2021; Zannini et al., 2021). These areas are typically remnant forests linked to spirits, temples, or burial sites, providing significant protection against logging, agriculture, and other forms of habitat destruction (Frascaroli et al., 2019; Ormsby, 2021; Zannini et al., 2021). To assess the impact of forest degradation on the population of aphyllous Vanilla species, we used undisturbed sacred forests as reference sites.

This study aimed to assess the impact of forest degradation on aphyllous vanilla populations. Specifically, we evaluated the effects of degradation on (i) the diversity of phorophytes and tree flora, (ii) the mode of reproduction, and (iii) the genetic diversity of aphyllous Vanilla species. We compared vanilla populations in sacred forests with those in the surrounding degraded forests to test our hypothesis that aphyllous vanilla populations in sacred sites would have greater genetic diversity owing to a less disturbed environment. Conversely, we expected reduced floristic diversity and higher clonality in the degraded forests. We also hypothesized that forest degradation would significantly affect vanilla genetics, leading to differentiation between populations within a locality over time. Ultimately, this study provides insights into the optimal strategies for replanting programs with aphyllous Vanilla species, depending on whether the forest is degraded or well-conserved.

Methods

Study Sites

The study was conducted in April and May 2018 at two sites in northern and southern Madagascar, including (i) a sacred forest with relatively intact vegetation and (ii) a degraded forest of secondary formations or ‘savoka’. The Ambatofaly sacred forest and Ampitolova degraded forest are in the north (Figure 1, Table 1), whereas the Ampotaka sacred forest and Vohitsara degraded forest are in the south (Figure 1, Table 1). The northern sites are in a sub-humid bioclimatic zone, whereas the southern sites have sub-arid conditions (Table 1).

Figure 1.

Study areas and locations of Vanilla species populations in Madagascar. Map displaying the Ambatofaly sacred forest and Ampitolova degraded forest in the north (up). Map displaying the Ampotaka sacred forest and the Vohitsara degraded forest in the south of Madagascar (Down). The pentagon indicates the geographical positions of the four localities where Vanilla population was sampled.

Table 1.

Area, location, geographic information, ecological status, vegetation, and climatic conditions of four study sites in northern (ABF and APL) and southern (APTK and VHT) Madagascar.

Site	Code	Area (ha)	Longitude	Latitude	Altitude (m)	Status	Vegetation	Bioclimate	T (°C)	R (mm)
Ambatofaly	ABF	5	48°34'40.10”E	13°27'9.18”S	20	Conserved	Humid forest	Sub-humid	26	2095
Ampitolova	APL	125	48°32'34.86”E	13°27'53.37”S	26	Degraded	Savoka	Sub-humid	26	2095
Ampotaka	APTK	157	43°56'41.38”E	23°54'13.65”S	108	Conserved	Thicket	Sub-arid	24	547
Vohitsara	VHT	112	44°14'29.63”E	23°53'22.95”S	258	Degraded	Thicket	Sub-arid	24	551

Climatic data were obtained from https://www.worldclim.org, a widely used source of global climate data (data for 1970 – 2000). T: Annual Mean Temperature, R: Annual precipitation.

The Sacred Forests of Ambatofaly and Ampotaka

The Ambatofaly sacred forest (ABF), a relictual forest on the northwestern coast, spans five hectares and is classified as lowland humid tropical forest (Moat & Smith, 2007) (Table 1). In 1825, the forest was consecrated through a ritual that sought blessings from ancestors via the largest tree, an Andansonia madagascariensis Baill. This ceremony initiated the sacredness of the Ambatofaly forest, governed by taboos and rules.

The Ampotaka (APTK) sacred forest (Table 1) is situated in the Androy region of southern Madagascar (Figure 1) on the Mahafaly Plateau. The vegetation is classified as a dry, spiny thicket (Moat & Smith, 2007), covering an area of ca. 157 ha. The forest is home to numerous ancestral tombs called ‘kibory’, and is managed by local communities. Although our study could not determine the exact date, most southern sacred forests predated the colonial era (Tengö et al., 2007). Reports on turtle reintroduction into the APTK forest suggest that it remains largely undisturbed, except for minor traces from goat herds (Randrianjafizanaka, 2014).

The Degraded Forests of Ampitolova and Vohitsara

The Ampitolova (APL) ‘savoka’ (Figure 1, Figure 2b) is situated approximately 4.5 km away from the ABF forest and is surrounded by small villages. The APL undergoes extensive logging of timber and firewood. The 6–7-year-old successional vegetation primarily consists of young shrubs 1–3 m tall.

Figure 2.

Overview of conserved (left) and degraded (right) forests. (a) Ambatofaly sacred forest; (b) Ampitolova degraded forest; (c) Ampotaka sacred forest; (d) Vohitsara degraded forest; (e) vanilla stem (arrow) at Ambatofaly (semi-closed canopy); (f) vanilla vines (arrow) at Vohitsara (open canopy); (g) vanilla inflorescence (arrow) at Ambatofaly; and (h) vanilla inflorescence and capsules (arrows) at Ampitolova (open canopy).

The degraded Vohitsara forest (VHT) (Table 1) is situated 30 km east of the sacred APTK forest (Figure 1, Figure 2d). Located in one of the most deprived regions of Madagascar, the VHT is frequently affected by the devastating effects of drought, which limits land-use options (Andriamparany et al., 2014). The local population relies heavily on forest products for subsistence, exacerbating resource overexploitation and further worsening their situation (Andriamparany et al., 2014). The lack of government protection has left VHT forests accessible to the local communities for years.

Study Species

Aphyllous Vanilla species (Orchidaceae) are characterized by their succulent stems, which are typically green in color (Figure 2e-f) but may turn brown (Figure 2h) or yellow when exposed to excessive direct sunlight. There are seven known aphyllous Vanilla species in Madagascar: Vanilla bosseri L. Allorge, V. decaryana H. Perrier, V. humblotii Rchb.f., V. madagascariensis Rolfe, V. perrieri Schltr, V. allorgeae Andriamihaja & Pailler, and V. atsinananensis Andriamihaja & Pailler (Allorge-Boiteau 2005, 2013; Andriamihaja et al. 2022; Cribb & Hermans 2009; Portères 1954). White and yellow flowers bloom only once a year and are ephemeral. Based on the research of Andriamihaja et al. (2021, 2022), the two northern sites, ABF and APL, likely had V. madagascariensis, while APTK and VHT were likely home to V. decaryana.

Floristic Samplings

In an area with uniform vegetation and aphyllous vanilla vines, a 50-by-20-meter plot was established using the Braun-Blanquet method (Westhoff & Van Der Maarel, 1978). Three comparable plots were established at each site. Trees and shrubs with a diameter at breast height (DBH at 1.3 meters) of over 5 cm were sampled. The DBH, maximum height, and local name of each specimen were recorded. The total number of trees and phorophytes per plot and the number of vanilla plants per phorophyte were documented. A vanilla plant was defined as a vine with all its branches, and phorophytes were specifically trees hosting aphyllous vanilla vines. A total of 501 herbarium specimens were collected from every tree or shrub species to match the reference specimens and provide precise scientific names. This method offers a systematic, comprehensive, and quantitative approach to studying vegetation structure, composition, and ecological interactions.

Forest Vertical and Horizontal Structures

Gautier's method (Gautier et al., 1994) was used to analyze the vertical structure of the forest (Figure S1A), profile the foliage distribution, and estimate the coverage of key species. This involved recording vegetation contact points with a vertical graduated pole placed at one-meter intervals along a transect through the plot center, noting the height of these contacts. These points were projected onto an Excel spreadsheet to create a structural vegetation profile (Figure S1B). Analyzing contact points by height class, quantified coverage, expressed as a percentage of the continuity of forest cover. The coverage scale developed by Godron et al. (1983) was used to interpret the results (Figure S2). The horizontal forest structure, assessed by density, dominance, and species frequency, was determined by evaluating the ecological weight and diameter distribution of all individuals of each species, following Lamprecht (1989). The importance value index (IVI) was calculated based on the collected data for each species as described by Melo and Vargas (2003). IVI was derived as follows: IVI = RA + RD + RF, where RA is the relative density, RD is the relative dominance, and RF is the relative frequency. The relative density was calculated as RA = (n/N) × 100, where n is the number of individuals of each species and N is the total number of individuals. Relative dominance was calculated as RD = (Ga/GT) × 100, where Ga is the basal area of each species, and GT is the total basal area. The basal area was calculated using Ga = (π/4) × (DBH)². The relative frequency was obtained by considering RF = (aF/TF) × 100, where aF is the absolute frequency, and TF is the total frequency. This index aids in identifying the species with the highest ecological value and disturbance indicators in forests (Catalá, 2011).

Flora Diversity

This study used online databases, such as the Madagascar Catalogue of Vascular Plants by Tropicos (2023) and Plants of the World (Royal Botanic Gardens Kew 2023), to determine plant species endemism. Tree species were classified as exotic, indigenous, or endemic. EstimateS software (Colwell & Elsensohn, 2014) was used to calculate the expected number of species at each site, considering rarefaction and extrapolation. We conducted 100 randomization iterations to estimate the number of nodes in each sample by combining the data collected from each site. Classical calculations for Chao1 and Chao2 were used to determine estimators and indices. The extrapolated 'S' values and associated standard deviations were used to construct comparable plot species curves for the different sites. The means of the Chao1 species richness estimator indices were derived using EstimateS, which considers the bias stemming from undetected species and serves as a minimum estimate of the total species richness at each site (Chao, 1984, 1987). To quantify the specific diversity at each site, the Shannon diversity index (H’) was computed using EstimateS and the Pielou evenness index using the formula J = H'/ln(S), where S is the number of species found in one locality.

Data Analysis

For statistical analysis, we conducted site-by-site tests using R version 4.3.1 (R Core Team, 2023). To evaluate the number of families, genera, and species present at each site, we combined the data from three separate plots and analyzed their distribution frequency. The percentage of families was calculated considering all individuals. The biodiversityR package (Kindt & Kindt, 2019) was used to calculate the ecological importance value index (IVI, RA, RD, and RF). Two main criteria were assessed to evaluate vanilla preference for phorophytes. First, the most frequently observed phorophyte species among all tree species were identified using the chi2 test to compare the tree species frequencies. Second, the tree species hosting most of the vanilla individuals were determined by excluding phorophytes that hosted vanilla only once. The potential effects of phorophyte species on vanilla abundance were tested using one-factor ANOVA. Mean number of vanilla individuals per phorophyte was then compared using Tukey's post hoc test with the “glht” function from the R package “multcomp” (Hothorn et al., 2015) at a significance level of p < 0.05. Pearson’s correlation tests were performed to examine the relationship between species density and their identification frequency as phorophytes, as well as the correlation between the number of vanilla individuals and phorophyte diameter. These analyses offer insights into the phorophyte preference of vanillas. Additionally, a split-plot factorial ANOVA was used to assess the average tree height across localities in both the conserved and degraded forests.

Sampling and DNA Extraction

We employed a site-specific sampling approach, in which we collected 25 samples from each study site. We excised 15 cm of stems or young buds and preserved them in silica gel, resulting in 100 individuals across the four populations. DNA collection from young buds allowed us to obtain high quality, uncontaminated, and undegraded DNA. To prevent collection from the same clonal shoot, we ensured a minimum distance of 2 m between collected specimens. DNA was extracted from 100 specimens using the DNEASY® plant kit (https://www.qiagen.com).

Microsatellite Genotyping

Seven microsatellite markers developed by Gigant et al. (2011) (Table S1) were used to amplify and genotype each sample. Polymerase chain reaction (PCR) was carried out using fluorescently labeled primers in a total volume of 25 μl using a Promega kit (Promega, 2007). The mixture contained 12.5 μl of GoTaq Colorless Master Mix 2X, 2 μl of DNA solution (5 < DNA < 250 ng), 0.4 μM of each primer, and 8.5 μl of nuclease-free water. Amplifications were carried out under the following conditions: an initial denaturation step at 95°C for 2 min, followed by 45 cycles of denaturation at 95°C for 30 s, annealing at 57°C for 45 s, and extension at 72°C for 1 min, and a final extension step at 72°C for 7 min. Capillary electrophoresis with an automated 3130XL ABI Genetic Analyzer (Applied Biosystems) was used to genotype PCR fragments. GENEIOUS PRIME 2019.1 (https://www.geneious.com) was used for allele size coding.

Genetic Diversity

Of the 25 samples collected per site, 23 were considered for ABF, APL, and VHT, and 21 for APTK due to missing data. Vanilla species at each site were identified through Bayesian genotypic clustering analysis using STRUCTURE software (Pritchard et al., 2000) on all sampled individuals, and six aphyllous Vanilla species (Table S2) from Andriamihaja et al. (2021, 2022). Ten independent replicates were performed for each K value (1 to 10), without admixture and independent allelic frequency models. The analysis was conducted with a burn-in length of 100 000 iterations and 1 000 000 Markov chain Monte Carlo (MCMC) replications. We determined the optimal number of clusters based on the posterior probabilities of K “Ln P (X/K),” and the summary likelihood statistic ΔK (Evanno et al., 2005). Two datasets were generated: one with all sampled individuals (“ramets”) and another excluding identical multilocus genotypes (“genets”), identified using multilocus clonality analysis in GenAlex 6.51 (Peakall & Smouse, 2006). The probability of encountering a specific multilocus genotype through sexual reproduction under random mating was calculated using GenAlex 6.51 to identify clones from vegetative reproduction. Genotypic diversity was assessed using the ratio of “genets” (G) to “ramets” (N) (Ellstrand & Roose, 1987). Genetic variability parameters and Hardy-Weinberg equilibrium (HWE) deviations were analyzed using the genet dataset. The average number of alleles per locus, total alleles per population, and private alleles was estimated using GenAlex for each polymorphic locus across populations. Observed heterozygosity (H_O), expected heterozygosity (H_E), and fixation indices (F_IS) were calculated using the Genepop 4.7.5 (Rousset, 2008). Exact p-value estimates (Guo and Thompson, 1992) were determined at the population level to test for HWE deviations using default parameters. Linkage disequilibrium between loci was tested using Fisher’s exact test in Genepop, comparing the results with and without repeated multilocus genotypes. To estimate the effective population size, we used the “genet” data, excluding any missing values. We applied the linkage disequilibrium (LD) method, assuming random mating, using NeEstimator v2.1 software (Do et al., 2014). NeEstimator calculated the 95% confidence interval (CI) by using the JackKnife procedure for each Ne estimate. The LD method requires setting a critical value (PCrit) at a sufficiently high level to exclude alleles present in a single copy (heterozygotes) within the sample. We used PCrit > 1/(2S), where S is the number of individuals with available data for both loci. With our population of 23 to 11 genets, this required PCrit values greater than 1/46=0.02 to 1/22=0.045. We report the results obtained using a PCrit value of 0.05.

Differentiation of Populations

Population differentiation was analyzed by calculating the pairwise F_ST index using Genepop, which considered only the 'genet' dataset. Bayesian clustering analysis using STRUCTURE software was performed on the dataset to assign multilocus genotypes to different groups. Bayesian analysis uses probabilistic models to estimate the number of genetic groups (K) present in the data, and assigns each individual to these groups based on their multilocus genotypes. This approach is particularly effective in detecting genetic differentiation patterns, even when the differences between populations are subtle. Ten independent replicates were conducted for each K value, ranging from 1 to 10, to detect any sub-structuring within the samples. The no admixture and correlated allelic frequency models were used, with a burn-in length of 100,000 iterations and 1,000,000 MCMC replications. The optimal number of clusters was determined using the online software Structure Analyses (https://pophelper.com/), which calculates the summary likelihood statistics ΔK.

Results

Changes in Forest Structure Following Degradation and Impacts on Vanilla Species Colonization

A total area of 1.2 hectares was surveyed, including 4 412 vanilla individuals, 1 404 trees, and shrubs, of which 627 were phorophytes. Table 2 compares the forest structure parameters of the four sites located in northern and southern Madagascar. The sacred forest of ABF stood out because of its average tree height of 9.2 ± 3.1m (Table 2). There is a clear trend of a decreasing number of trees in degraded forests compared to conserved forests (Table 2). In addition, a reduction in forest cover was observed in degraded forests, although cover loss was less pronounced in the southern forest of the VHT (7.7%). In contrast, forest degradation caused a significant increase in the number of phorophytes hosting vanilla vines (Table 2). Specifically, there was an 80% increase in phorophytes in the degraded APL forest compared to the sacred ABF forest, and a 48% increase in the degraded VHT forest compared to the sacred APTK forest (Table 2). The number of vanilla individuals was 10 to 16 times higher in open sites, such as the APTK and VHT forests in the south and the degraded APL forest in the north, than in the ABF site where the canopy was semi-closed (Table 2).

Table 2.

Comparison of forest structure parameters, with abundance of Vanilla species between conserved and degraded forests located in the northern (ABF and APL) and southern regions (APTK and VHT) of Madagascar.

Site	H_max	C_co (%)	C_co scale	N_t	N_p (%)	V
ABF	9.2 ± 3.1^b	64.6	Semi-open stratum	363	24 (6.6)	118
APL	4.4 ± 1.7^a	28.6	Open stratum	290	253 (87.2)	1958
APTK	4.7 ± 1.7^a	37.5	Open stratum	532	174 (32.7)	1160
VHT	4.7 ± 2.4^a	29.8	Open stratum	219	176 (80.3)	1176

H_max: Tree maximal height, C_co: Canopy cover rate, C_co scale: Canopy cover scale, N_{t :} Total number of trees, N_p : Total number of phorophyte trees, with the percentage in parentheses indicating the proportion of the total number of trees present on the site, V: number of vanilla individuals. ABF: Ambatofaly, APL: Ampitolova, APTK: Ampotaka, VHT: Vohitsara. Means followed by the same letter within the same column are not significantly different (p > 0.05).

Changes in Tree Species Composition Following Degradation and Impacts on Vanilla Species Colonization

In total, 104 tree species belonging to 47 families and 83 genera were identified in the study area. Table 3 presents a summary of the flora diversity parameters at the four sites located in northern and southern Madagascar. From these data, it can be seen that the number of recorded species (S) is always lower in degraded forests. Additionally, the extrapolated accumulation curves (Figure S4) indicated that there were more expected species (Chao 1) in conserved forests (Table 3). The Shannon diversity index (H') observed between conserved and degraded forests revealed significant differences (Table 3), with higher diversity index values for conserved forests. However, Piélou's evenness was statistically similar across all study sites, with values ranging from 0.76 to 0.85, indicating relatively high uniformity of species across the study sites (Table 3). A comprehensive floristic list including all tree species, phorophytes, and non-phorophytes is provided in Table S3. Of these 104 species, 72 were confirmed to be endemic to Madagascar, 14 were indigenous and seven were introduced. The remaining 11 species were not identified in the present study. Among the 627 phorophyte trees (Table 2), 73 species were identified, representing 70% of all the recorded species. Interestingly, 4% of these phorophyte trees hosted only one individual of vanilla.

Table 3.

Comparison of parameters describing flora diversity and ecology between conserved and degraded forests in the northern (ABF and APL) and southern (APTK and VHT) regions of Madagascar.

Site	F	G	S	H’	J	Chao1	Php
ABF	22	31	35	2.48 ± 0.1^a	0.76 ± 0.3^a	56.4^a	11
APL	17	24	24	2.19 ± 0.2^b	0.84 ± 0.6^a	25.12^b	23
APTK	28	39	46	2.68 ± 0.4^a	0.77 ± 0.1^a	53.3^a	35
VHT	10	15	17	2.12 ± 0.7^b	0.85 ± 0.1^a	17.2^b	17

F: number of families, G: number of genera, S: species richness, H’: Shannon’s diversity index, J: Pielou’s evenness index, Chao1: Species richness estimates, Php: Number of species identified as phorophytes. ABF: Ambatofaly, APL: Ampitolova, APTK: Ampotaka, VHT: Vohitsara. Means followed by the same letter within the same column are not significantly different (p > 0.05).

The ABF inventory identified 35 tree species across 31 genera and 22 families, with Loganiaceae being the most prevalent, accounting for 22% of trees. Among these species, 22 were endemic, five indigenous, four introduced, and four unidentified (Figure 3). The most dominant species were Strychnos decussata (Pappe) Gilg (IVI = 48.11), Diospyros gracilipes Hiern (IVI = 40.88), and Trilepisium madagascariensis DC. (IVI = 27.77) (Table 4). Additionally, the inventory allowed the recording of 118 individuals of V. madagascariensis hosted by 11 different phorophyte species (Table 2 and Table 3), six of which were endemic, two autochthonous, two introduced, and one unidentified (Table S3). On average, 4.9 ± 2.6 vanillas were observed on each phorophyte. Among the 24 surveyed phorophyte trees, Diospyros gracilipes was the most commonly encountered, with seven trees observed, followed by Trilepisium madagascariensis, which was observed four times as phorophytes. However, none of the phorophyte species showed significant affinity for vanillas (χ² = 16.33; df = 10; p = 0.09). It is noteworthy that Trilepisium madagascariensis had the highest number of hosted vanilla plants, with an average of 7.2 ± 3.5 individuals per tree, but did not differ significantly from the other phorophytes. No significant correlation was observed between the number of aphyllous vanilla individuals and phorophyte diameter (r = 0.2; p > 0.05). However, a significant positive correlation was observed between the tree species abundance and their role as phorophytes (r = 0.7, p < 0.01).

Figure 3.

Number of tree species per locality and according to endemism. Endemism: End (endemic), Exo (exotic) et Ind (indigenous), Uni (Unidentified), Locality: ABF (Ambatofaly), APL (Ampitolova), APTK (Ampotaka), VHT (Vohitsara).

Table 4.

List of the five species with the highest ecological importance value index (IVI) at the studied sites.

Locality	Species	Dis	Den	RF (%)	RA (%)	RD (%)	IVI (%)	Php	PhpS	AvV
ABF	Strychnos decussata (Pappe) Gilg	Nat	80	4.05	22.04	22.02	48.11	Yes	3	3.66
ABF	Diospyros gracilipes Hiern	End	67	4.05	18.46	18.37	40.88	Yes	7	4.42
ABF	Trilepisium madagascariense DC.	Nat	53	4.05	14.60	9.11	27.77	No
ABF	Albizia lebbeck Benth.	Exo	30	4.05	8.26	14.10	26.41	Yes	1	8
ABF	Schizolaena parviflora (F. Gérard) H. Perrier	End	24	4.05	6.61	4.22	14.88	Yes	2	5
APL	Ravenala agathea Haev. & Razanats.	End	51	8.11	17.59	58.65	84.35	Yes	32	17,56
APL	Schizolaena parviflora (F. Gérard) H. Perrier	End	45	8.11	15.52	7.12	30.75	Yes	40	5,27
APL	Gaertnera arenaria Boulanger	End	40	8.11	13.79	4.43	26.33	Yes	37	6,27
APL	Mascarenhasia arborescens A. DC.	Nat	33	8.11	11.38	6.20	25.69	Yes	33	6,3
APL	Dypsis ambanjae Beentje	End	12	5.41	4.14	6.01	15.55	Yes	13	7,2
APTK	Gyrocarpus americanus Jacq.	Nat	91	3.23	17.11	13.74	34.07	Yes	23	6,5
APTK	Commiphora lamii H. Perrier	End	57	3.23	10.71	14.27	28.21	Yes	14	11,9
APTK	Euphorbia fiherenensis Poiss.	End	62	3.23	11.65	10.87	25.75	Yes	22	6,63
APTK	Croton geayi Leandri	End	51	3.23	9.59	12.02	24.83	Yes	3
APTK	Cedrelopsis gracilis J.-F. Leroy	End	41	3.23	7.71	10.92	21.85	Yes	12	8
VHT	Euphorbia antso Denis	End	72	9.09	32.88	38.04	80.01	Yes	55	7,2
VHT	Croton geayi Leandri	End	21	6.06	9.59	9.01	24.66	Yes	19	6,68
VHT	Croton salviformis Baill.	End	18	9.09	8.22	6.31	23.62	Yes	14	5.2
VHT	Terminalia ulexoides H. Perrier	End	16	9.09	7.31	6.48	22.88	Yes	15	5,9
VHT	Ehretia cymosa Thonn.	Exo	13	3.03	5.94	5.44	14.40	Yes	12	5

Ambatofaly (ABF), Ampitolova (APL), Ampotaka (APTK), and Vohitsara (VHT). Within each site, the list was arranged in decreasing order of IVI. Den: Density of each species; Dis: Distribution, End: Endemic, Exo: Exotic, Ind: Indigenous, Uni: Unidentified, IVI: Importance Value Index, RA: Relative density; RD: Relative dominance; RF: Relative frequency; Php: Phorophyte status; PhpS: Number of individual phorophytes; AvV: Average number of aphyllous vanilla individuals per phorophyte.

The degraded forest in APL exhibited 17 plant families, 24 genera, and 24 species, including one unidentified species. Among the species, two were introduced, seven were indigenous, and 14 were endemic. Four endemic species were common to the neighboring ABF-sacred forest. The Sarcolaenaceae family was the most represented (21.4%), with three species, followed by the Rubiaceae (20.4%) (Table S3). The most dominant species at the APL site was Ravenala agatheae Haevermans & Razanatsoa (IVI = 84.35), followed by Schizolaena parviflora (F. Gérard) H. Perrier (IVI = 30.75) and Gaertnera arenaria Boulanger (IVI = 26.33) (Table 4). A total of 1958 V. madagascariensis and 23 phorophyte species were recorded, representing 95.8% of the potential phorophytes in the APL. Endemic species represented 14 of the 23 phorophyte species, with Schizolaena parviflora being the most frequent (15.8%), followed by Gaertnera arenaria (14.6%). Analysis of the frequency distribution of these phorophytes showed a significant difference (χ² = 308.32; df = 22; p < 0.001), highlighting a pronounced preference for these two species for aphyllous vanillas over other phorophytes. APL differed from the other three sites in having a high number of vanillas per host tree, with an average of 7.7 ± 7.2 individuals. This was illustrated by the species Ravenala agatheae, which harbored an average of 17.5 ± 11.5 vanilla individuals per tree. A significant positive correlation was found between the number of aphyllous vanilla individuals and the diameter of their phorophytes (r = 0.7; p < 0.001), and between the density of a tree species and its role as a phorophyte (r = 0.9; p < 0.001).

The APTK forest comprised 46 tree and shrub species from 39 genera and 28 families, including 41 endemic, four indigenous, and one introduced species. The Euphorbiaceae family was the most prevalent, representing 24.8% of the recorded species. Gyrocarpus americanus Jacq., Commiphora lamii H. Perrier, and Euphorbia fiherenensis Poiss. were the most dominant species, with IVI values of 34.07, 28.21, and 25.75, respectively (Table 4). A total of 1160 individuals of V. decaryana individuals hosted by 35 different phorophyte species were recorded (Table 2). According to these findings, aphyllous vanilla was hosted by 76% of the tree species found in APTK, with endemic trees dominating the landscape. Of the 35 phorophyte species, 28 were endemic, two were indigenous, and one was introduced. The remaining four species were not identified in this study. The most common phorophyte for vanilla was Securinega seyrigii Leandri (14.9%), followed by Gyrocarpus americanus (13.2%). These species exhibited significantly higher affinity for vanillas (χ² = 310.13; df = 33; p < 0.001). Analysis of the number of vanillas hosted by phorophytes indicated an average of 6.6 ± 4.6 vanillas per host tree. Commiphora lamii hosted an elevated number of vanillas, with an average of 12.1 ± 9.2 per tree. The analysis revealed significant positive correlations between the number of aphyllous vanilla individuals and phorophyte diameter (r = 0.6; p < 0.001) as well as between the abundance of a tree species and its role as a phorophyte (r = 0.8; p < 0.001).

The VHT vegetation included ten families, 15 genera, and 17 tree species: 15 endemic, one indigenous, and one introduced. Euphorbiaceae constituted 50.7% of the recorded families. Euphorbia antso Denis was the dominant species (IVI = 80) followed by Croton geayi Leandri (IVI = 24.66). Of the endemic species, eight were common to the APTK sacred forest. Vanilla recorded at the VHT revealed 1176 individuals of V. decaryana, with the majority being associated with Euphorbia antso, accounting for 31.3% of the phorophytes, followed by Croton geayi at 10.8%. Analysis of the frequency distribution of these phorophytes showed a significant difference (χ² = 244.94; df = 16; p < 0.001), underscoring the distinct preference of Euphorbia antso for aphyllous vanilla over alternative phorophytes. On average, each phorophyte hosted 6.6 ± 4.3 of the vanilla individuals, and no species exhibited a significantly higher number of vanilla individuals compared to others. Significant positive correlations were found between the number of aphyllous vanilla individuals and the diameter of the phorophytes (r = 0.6; p < 0.001), and between the abundance of a tree species and its role as a phorophyte (r = 0.9; p < 0.001).

Genetic Diversity of Aphyllous Vanillas

The results of STRUCTURE analysis for species determination in studied sites revealed that individuals sampled from ABF and APL sites belong to the species V. madagascariensis, while those from APTK and VHT sites belong to the species V. decaryana (Figure S3). Clonal individuals were found exclusively in southern populations of V. decaryana. In total, 16 clones were identified, with six belonging to the VHT population (G/N = 0.52) and ten to the APTK population (G/N = 0.78) (Table 5), which were distributed across five distinct multilocus genotypes (MLGs). Our Psex analysis (Psex < 0.001) confirmed that all MLGs originated from vegetative reproduction and we observed no common MLGs between the different populations. Following the elimination of clones, we identified 80 unique ‘genets’ out of the 96 examined samples. We identified 97 alleles across seven loci, with the number of alleles per locus ranging from seven to 18 (Table S1). The average number of alleles per locus varied between 4.5 to 7.5, depending on the population (Table 5). Allelic diversity was slightly lower in the conserved forests than in the degraded forests, with an average of 6.8 and 4.5 alleles per locus compared to 7.5 and 5 alleles per locus in the degraded forests, with V. madagascariensis and V. decaryana, respectively (Table 5). Additionally, the degraded forests in the north and south exhibited a higher number of private alleles (1.7 for APL and 1.8 for VHT) than the conserved forests, with 1.2 for ABF and 1.4 for APTK, respectively (Table 5). The multilocus exact test indicated significant deviations from the Hardy-Weinberg equilibrium (p < 0.05) for each population (Table 5). Nei heterozygosity (H_E) values showed relatively high genetic diversity (H_E > 0.6) in all four populations (Table 5). When comparing the two northern sites, the expected heterozygosity (H_E) of the degraded forest (H_E APL = 0.72) was higher than that of the conserved forest (H_E ABF = 0.63). Conversely, for the two southern sites populated by V. decaryana, the expected heterozygosity was higher at the conserved forest (APTK), with a value of H_E = 0.65, whereas the degraded forest (VHT) had a slightly lower value of H_E = 0.62. All four populations had H_O values above H_E, resulting in negative F_IS values, indicating an excess of heterozygotes. Heterozygote excess was more significant in V. decaryana populations but also in degraded versus conserved forests (Table 5). The ABF population exhibited a significant excess of heterozygotes at only one locus, whereas the APL and APTK populations exhibited a significant excess at five of the seven loci (Table S4). In the VHT population, all studied loci showed a significant excess of heterozygotes (Table S4). In this study, we observed effective population size (Ne) values below N, indicating the accuracy of the estimates obtained, as suggested by England et al.(2006). The aphyllous vanilla populations yielded very low Ne values, suggesting an effective population size of < 13 individuals. Moreover, populations in degraded forests exhibited higher Ne values than those in sacred forests (Table 5). When evaluating possible linkage disequilibrium among the 21 loci combinations, significant disequilibrium (p < 0.05) was observed for nine loci pairs in the 'genets' file and 11 loci pairs in the 'ramets' file. However, linkage disequilibrium was not due to physical linkage between loci, as previous studies have demonstrated that all loci are independent from each other (Andriamihaja et al., 2021; Gigant, De Bruyn, et al., 2016a; Gigant, Rakotomanga, et al., 2016b).

Table 5.

Summary of genetic diversity at seven microsatellite loci in four populations (Ambatofaly, Ampitolova, Ampotaka, and Vohitsara) of aphyllous Vanilla species from Madagascar.

POP	Species	N	G	G/N	Na	Pa	H_O	H_E	F_IS	HWE	Ne [95%CI]
ABF	V. madagascariensis	23	23	1	6.8 ± 0.8	1.2 ± 0.4	0.73	0.63	- 0.1724	**	6.9 [3 - 13.3]
APL	V. madagascariensis	23	23	1	7.5 ± 0.9	1.7 ± 0.4	0.85	0.72	- 0.1721	***	9.3 [5.7 - 15.2]
APTK	V. decaryana	21	11	0.52	4.5 ± 0.7	1.4 ± 0.5	0.85	0.65	- 0.3187	***	6.5 [2.1 - 47]
VHT	V. decaryana	23	18	0.78	5 ± 0.9	1.8 ± 0.9	0.90	0.63	- 0.4356	***	12.1 [3.2 - 36.6]

N: number of individuals sampled (ramet), G: number of unique genotypes (Genet), Na: mean number of alleles per locus, Pa: mean number of private alleles per locus, H_O: observed heterozygosity for all loci, H_E: expected heterozygosity for all loci, F_IS: fixation index, HWE: results of deviation tests from Hardy-Weinberg, Ne: effective population (LD method) with 95% confidence interval (CI) for allelic frequency threshold of 0.05, obtained from JackKnife method.

Genetic Differentiation Between Aphyllous Vanilla Populations

Bayesian assignment analysis using STRUCTURE showed a distribution of ΔK values, with a maximum for K = 2, indicating the presence of two distinct genetic clusters (Figure 4). Populations of V. madagascariensis in the north (ABF and APL) and V. decaryana in the south (APTK and VHT) exhibited genetic homogeneity in terms of their structure. However, analysis of genetic differentiation based on F_ST revealed significant differences (p < 0.001) between the four populations (Table 6). Consistent with the STRUCTURE results, a strong differentiation was observed between the populations of V. madagascariensis and V. decaryana, with F_ST values ranging from 0.24 0.32. However, a slight yet significant differentiation was noted between the populations of V. madagascariensis growing in conserved and degraded humid forests (F_ST = 0.04 between ABF and APL) as well as between the populations V. decaryana growing in conserved and degraded dry forests (F_ST = 0.07 between APTK and VHT).

Figure 4.

Genetic structuring of the four aphyllous vanilla populations. (A) DeltaK value (max K= 2) obtained by Structure Analyses (https://pophelper.com/). (B) Genetic assignment plot of individuals evaluated by the STRUCTURE program (K = 2). Each individual is represented by a vertical bar partitioned into segments of K colors, which represent the probability of the individual belonging to the cluster of this color.

Table 6.

Results of genetic differentiation analyses (F_ST values) between four populations (Ambatofaly, Ampitolova, Ampotaka, and Vohitsara) of aphyllous Vanilla species from Madagascar.

Pop	ABF	APL	APTK
APL	0.0413 ***
APTK	0.2525 ***	0.3236 ***
VHT	0.2436 ***	0.3110 ***	0.0697 ***

p-value for each population pair across all loci (Fisher's method) p <0.001, ***.

Discussion

Forest Degradation Alters Tree Communities and Dominant Species

Our findings confirmed a decrease in woody species richness due to prolonged forest degradation. This decline in tree numbers significantly affects forest structure, leading to biodiversity loss and altering ecological dynamics by changing habitats, resource competition, and species interactions (Arroyo-Rodríguez et al., 2017; Lohbeck et al., 2015). Endemic species are highly prevalent in floristic studies, which is unsurprising given the high flora endemism rate of Madagascar (Binggeli, 2003; Goodman, 2022; Ralimanana et al., 2022). Despite degradation at the APL and VHT sites, exotic and shrub species remained low. However, habitat degradation significantly alters floristic composition. For instance, Ravenala agatheae, common in degraded forests, dominates the APL (Irwin et al., 2010; Randrianasolo et al., 2019). Continuous pressure, such as firewood and timber collection by riparian communities and gemstone mining, likely explains the low number of common species between the conserved and degraded forests.

Larger and Most Abundant Tree Species are the Preferred Phorophytes

We documented the total number of phorophyte species corresponding to 70% of the tree and shrub flora in the study area. Host specificity for epiphytic orchids has been widely discussed (Huang et al., 2018; Otero et al., 2007; Zarate-García et al., 2020), and our study revealed that some phorophyte species show significant affinities with aphyllous vanilla. Specifically, Securinega seyrigii and Commiphora lamii at the APTK site, Schizolaena parviflora and Ravenala agatheae at the APL site, and Euphorbia antso at the VHT site exhibited strong connections with vanilla. Surprisingly, these phorophyte species were the most ecologically significant at the study sites. Because of their prevalence, these species play key ecological roles, including carbon storage (Måren & Sharma, 2021; Speak et al., 2020), and influence hydrological regimes, nutrient cycles, and critical ecosystem processes (Lindenmayer & Laurance, 2017). Among the three phorophyte species significantly associated with vanilla in the south, only S. seyrigii was commonly used for charcoal production, whereas C. lamii and E. antso were not (Ranaivoson et al., 2017). Similarly, R. agathea is a tree species that remains unutilized for charcoal production, thus safeguarding it from deforestation. Moreover, our observations revealed that, in degraded forests, phorophytes represented 95% of both shrubs and large tree species. These findings suggest that vanillas can colonize a diverse range of available tree species. Except for the ABF site, where the number of vanilla individuals was low and most phorophytes had large diameters, the other three sites displayed significant correlations between the phorophyte diameter and the number of aphyllous vanilla individuals they hosted. Larger phorophytes were able to support a greater number of vanilla individuals, which explains the high density of vanilla observed on Ravenala agatheae at APL. These results reflect those of Shen et al. (2022), who suggested that large trees offer more surface than small trees and have been available for a longer time than younger hosts for epiphyte colonization. This also accords with several studies that reported root climbers’ ability to colonize phorophytes of different sizes (Botomanga et al., 2024; Mori et al., 2016; Orihuela & Waechter, 2010). Interestingly, most phorophyte species hosted more than one individual vanilla. However, this outcome is contrary to those of several authors (Orihuela & Waechter, 2010; Visser et al., 2018), who found that isolated occurrence was by far the most common. It is possible that the newly established aphyllous vanilla on phorophytes tended to propagate rapidly via ramification before splitting into different individuals. Another explanation is that degraded remnant forests are less densely populated, which increases the chance of a tree becoming a phorophyte of several vanilla individuals. The phorophyte preference of aphyllous vanilla plants may have evolved with the ongoing degradation of forests in Madagascar. The scarcity of large trees in secondary vegetation limits phorophyte options for smaller trees. However, owing to the high plasticity of their attachment roots, aphyllous vanilla plants can effectively colonize these smaller trees (Botomanga et al., 2024).

Canopy Openness is Associated With Increasing Aphyllous Vanilla Species Colonization

Except for the ABF sacred forest, which had a semi-open canopy, all other forests exhibited a moderately open stratum. These sites had a significantly higher number of vanillas, indicating that canopy openness positively influences the propagation of aphyllous vanillas. These findings align with existing knowledge that forest gaps and low tree density promote the establishment of heliophilous species and the proliferation of lianas (He et al., 2023; Schnitzer et al., 2021). The closed canopy of the ABF forest resulted in a lower density of V. madagascariensis, thereby reducing the number of phorophytes at this site. However, considering the “sparse and deciduous” nature of southern Madagascar's dry forests (Seddon & Tobias, 2007), the increase in light due to gaps in the degraded forests of VHT likely led to only a marginal increase in light intensity compared to the conserved APTK forest. This may explain the similar numbers of V. decaryana individuals recorded in both the conserved and degraded forests in the south.

Aphyllous Vanilla Populations Displayed High Genetic Diversity and an Excess of Heterozygotes

The results of our structural analysis indicated that vanilla populations in the north correspond to V. madagascariensis, whereas those in the south belong to V. decaryana. These findings corroborate those of the earlier studies conducted by Andriamihaja et al. (2021, 2022). Information on species reproductive patterns and the level of habitat degradation is crucial for understanding the genetic structure of populations and their evolutionary trends. Our results indicate a high genetic diversity of aphyllous Vanilla species at all studied sites (H_E > 0.6). This high diversity corroborates the results reported by Andriamihaja et al. (2021) for 22 locations where aphyllous Vanilla species are present, with an H_E values that can exceed 0.8. However, in contrast to our results for the northern sites, conserved forests, mainly corresponding to national parks, tended to have higher H_E values, which also appears to be the case for the southern sites investigated in our study. Nevertheless, it is important to note that Andriamihaja et al. (2021) did not examine the degree of openness of these sites, and tree density in protected forests can also be highly variable, with some being sparser than others. In addition, the ABF sacred forest has a total area of only 5 ha, much smaller than the national park average. Forest fragment size can have a dramatic bottleneck effect on orchid species diversity (Minasiewicz et al., 2018; Yamashita et al., 2023).

Our study found that all vanilla populations had negative F_IS results, suggesting an excess of heterozygotes. One possible explanation for this is the small size of the breeding population (Balloux, 2004; Templeton, 2018). Another explanation is the theory of overdominance, which suggests that selection against recessive alleles increases the frequency of heterozygotes (Hedrick et al., 2016; Mérot et al., 2020; Schöfl et al., 2018). Negative F_IS values can also occur due to negative assortative mating between individuals carrying different alleles (Fishman & Sweigart, 2018; Glémin et al., 2001; Schoen & Baldwin, 2023). Another plausible explanation for negative F_IS is asexual reproduction, which does not involve mixing of the DNA of the two parents, as in sexual reproduction (Arnaud-Haond et al., 2020; Probowati et al., 2023; Stoeckel & Masson, 2014). However, our results revealed a frequently observed paradox in which high genotypic diversity coexists with a significant excess of heterozygotes (Dia et al., 2014; Orantes et al., 2012). These two observations appear contradictory; high genotypic diversity suggests active sexual reproduction, whereas an excess of heterozygotes is generally associated with near-strict clonality. Stoeckel et al. (2021a) suggested that this apparent contradiction could be due to the strong prevalence of clonal reproduction in partially clonal organisms, which may explain the observed duality. Lack of genetic recombination can increase mutational heterozygosity over generations, with each new mutation immediately inherited by all descendants (Arnaud-Haond et al., 2020; Stoeckel, Arnaud-Haond, et al., 2021a; Stoeckel, Porro, et al., 2021b). Our results are consistent with those of Stoeckel et al. (2006, 2021b), who observed an excess of heterozygosity, but noted strong variations in F_IS values between loci and populations. Given the reproductive modes of aphyllous Vanilla species and the presence of genome-wide linkage disequilibria, we strongly suspect that the negative F_IS observed in the examined populations indicates the asexually mediated propagation of aphyllous Vanilla species.

The vanilla populations in the southern region exhibited a significantly greater proportion of heterozygotes than those in the north, which is supported by the more significant P values obtained in the HWE H1 test for excess heterozygosity (Table S5). For the southern populations, the (G/N) index was less than 1, the observation of linkage disequilibrium between the pairs of loci, and the high negative value of F_IS suggested a strong clonal structure in the populations (Halkett et al., 2005; Stoeckel, Porro, et al., 2021b; Stoeckel & Masson, 2014). Indeed, ten of the 21 APTK individuals and six of the 23 VHT individuals were identified as clones. Consequently, the significant increase in heterozygosity in the southern populations is due to this unusually high rate of asexuality. In support of this hypothesis, our results regarding the regeneration of aphyllous vanillas revealed the complete absence of new recruitments within populations in southern Madagascar (Botomanga A., unpublished data). Moreover, the HWE H1 test for excess heterozygosity was always more significant in degraded forests than in the corresponding conserved forests (non-significant for ABF) (Table S5), suggesting increased asexual reproduction in degraded forests. We observed significant and non-significant variations in F_IS values across loci in the APL and APTK populations (Table S4), which also suggested the occurrence of infrequent sexual reproduction events in these populations (Balloux et al., 2003; Halkett et al., 2005; Stoeckel, Porro, et al., 2021b; Stoeckel & Masson, 2014). Altogether, these results confirm that canopy openness promotes aphyllous Vanilla species colonization via asexual reproduction

Forest Degradation Leads to Slight Differentiation of Aphyllous Vanilla Species Populations

The genetic differentiation between populations from the conserved and degraded forests, as measured by a F_ST value of 0.04 for V. madagascariensis, was comparable to those reported by Andriamihaja et al. (2021) across four localities (Ambanja, Ankarana, Analamerana, and Maromandia) with F_ST values ranging from 0.01 to 0.04. Similarly, the genetic differentiation between populations of V. decaryana from conserved and degraded forests (F_ST = 0.7) aligned with F_ST values from three localities (Anja, Antandroka, and Kirindy Mitea) reported by Andriamihaja et al. (2021). The significant pairwise genetic differentiation detected between the populations of the conserved and degraded forests in our study might be attributed to the loss of natural habitats in Madagascar over the decades, which may have resulted in the erosion of genetic variability linked to changes in the vanilla reproductive system (Fajardo et al., 2017; Jump & Peñuelas, 2006; Li et al., 2020; Pandey et al., 2015). This translates into the increased vegetative propagation of aphyllous vanilla plants and a reduction in gene flow rates between populations, ultimately leading to population differentiation. Previous studies have shown that the effects of habitat degradation on plant genetic diversity are discernible only after 50–100 years of environmental fragmentation (Aguilar et al., 2008). The sacredness of the two conserved forests fell within this period, which could explain the subtle distinctions between the vanilla populations in the sacred and degraded forests. However, it is possible that the number of generations owing to habitat fragmentation and degradation is insufficient to generate high values for genetic differentiation (Ehlers & Pedersen, 2000).

Limitations

The shortcoming of this study lies in the insufficient number of true replicates for each species within a single bioclimatic zone. This might have hindered the statistical power of our findings. Nevertheless, we noticed a similar trend in the responses of the two unrelated Vanilla species to forest degradation, although they evolved in distinct bioclimatic zones.

Conclusion

Our results showed that aphyllous Vanilla species thrived best in areas with open canopies. Thus, conserved forests with closed canopies are not necessarily appropriate refugia. Forest degradation forces the growing vanillas to climb the remaining uncut trees and is accompanied by a reduction in genetic diversity that may originate from excessive clonality but also from a low recruitment rate. The lessons drawn from this study could be useful for the conservation of orchids that rely on phorophytes for their development.

Implications for Conservation

To prevent the erosion of genetic diversity and inbreeding, a population size (Ne) of over 100 individuals is crucial (Frankham et al., 2014; Hoban et al., 2020). Our study found that the effective population size (Ne) of the four aphyllous vanilla populations did not exceed 13 individuals. Thus, we recommend introducing new genotypes from the surrounding subpopulations to enhance their genetic diversity. Alternatively, replanting seedlings raised in vitro is an option to compensate for low in situ recruitment rates. Passive restoration, seeding, and mixed sources can significantly increase the genetic diversity of the restored plant populations (Bucharova et al., 2019; Wei et al., 2023). Nevertheless, restocking should be conducted cautiously to reduce the risk of outbreeding depression (Bucharova et al., 2019; Rogers et al., 2024) and prevent the introduction of genotypes that might promote invasiveness (Le Roux et al. 2007; Matesanz & Sultan, 2013). The inclusion of degraded forests in conservation areas could be a suitable option for forest management. In some degraded forest spots, halting wood harvesting may be beneficial for increasing the chance of a tree becoming a phorophyte. Additionally, stricter regulations on the wildcrafting of vanilla stems for trade should be enacted via the establishment of a quota system to ensure the sustainability of isolated populations.

Supplemental Material

Supplemental Material - Continuous Degradation of Forest Structure and Composition Causes a Shift in Phorophytes and Population Genetics in Aphyllous Vanilla Species

Supplemental Material for Continuous Degradation of Forest Structure and Composition Causes a Shift in Phorophytes and Population Genetics in Aphyllous Vanilla Species by Alemao Botomanga, Cathucia F. Andriamihaja, Pascale Besse, Vololoniaina H. Jeannoda, Michel Grisoni, Sedera E.F. Ny Aina Ranaivoson, Goum Antsonantenainarivony, Nicola Fuzzati and Aro Vonjy Ramarosandratana in Tropical Conservation Science

Footnotes

Acknowledgments

Our work received research permit No. 004/17/MEEF/54/DGF/DSAP/SCB.Re. from the Ministry of the Environment and Sustainable Development of Madagascar. We are grateful to the communities of Bevoay and Betioky for their warm welcome and support during fieldwork. The vanilla conservation program was supported by Chanel Parfums Beauté. We thank Tabita Randrianarivony and Nivo Rakotoarivelo for their assistance in species identification, and Johnson Randrianantenaina for technical support.

Declaration of Conflicting Interests

The author(s) declared no potential conflicts of interest with respect to the research, authorship, and/or publication of this article.

Funding

The author(s) disclosed receipt of the following financial support for the research, authorship, and/or publication of this article: This work was supported by this research received financial support from Chanel Parfums Beauté.

Data Availability Statement

Data utilized to derive the findings in this study have been deposited in the Science Data Bank (ScienceDB: ) for open access and availability: - Genetic data from four leafless vanilla populations. DOI: 10.57760/sciencedb.09416. - Ecological data from four leafless vanilla populations. DOI: 10.57760/sciencedb.09419

ORCID iDs

Alemao Botomanga

Aro Vonjy Ramarosandratana

Supplemental Material

Supplemental material for this article is available online.

References

Aguilar

Cristóbal-Pérez

E. J.

Balvino-Olvera

F. J.

de Jesús Aguilar-Aguilar

Aguirre-Acosta

Ashworth

Lobo

J. A.

Martén-Rodríguez

Fuchs

E. J.

Sanchez-Montoya

. (2019). Habitat fragmentation reduces plant progeny quality: a global synthesis. Ecology Letters, 22(7), 1163–1173. https://doi.org/10.1111/ele.13272

Aguilar

Quesada

Ashworth

Herrerias‐Diego

Lobo

(2008). Genetic consequences of habitat fragmentation in plant populations: susceptible signals in plant traits and methodological approaches. Molecular Ecology, 17(24), 5177–5188. https://doi.org/10.1111/j.1365-294X.2008.03971.x

Allorge-Boiteau

(2005). Les vanilles succulentes de Madagascar. Succulentes, 2, 3–11.

Allorge-Boiteau

(2013). Une nouvelle espèce de vanille à Madagascar. Hommes Et Plantes, 85, 4–5.

Andriamihaja

C. F.

Botomanga

Misandeau

Ramarosandratana

A. V

Grisoni

Da Silva

Pailler

Jeannoda

V. H.

Besse

(2022). Integrative taxonomy and phylogeny of leafless Vanilla orchids from the South‐West Indian Ocean region reveal two new Malagasy species. Journal of Systematics and Evolution, 61(1), 80-98. https://doi.org/10.1111/jse.12858

Andriamihaja

C. F.

Ramarosandratana

A. V

Grisoni

Jeannoda

V. H.

Besse

(2021). Drivers of population divergence and species differentiation in a recent group of indigenous orchids (Vanilla spp.) in Madagascar. Ecology and Evolution, 11(6), 2681–2700. https://doi.org/10.1002/ECE3.7224

Andriamparany

J. N.

Brinkmann

Jeannoda

Buerkert

(2014). Effects of socio-economic household characteristics on traditional knowledge and usage of wild yams and medicinal plants in the Mahafaly region of south-western Madagascar. Journal of Ethnobiology and Ethnomedicine, 10, 82. https://doi.org/10.1186/1746-4269-10-82

Arnaud-Haond

Duarte

C. M.

Alberto

Serrão

E. A.

(2007). Standardizing methods to address clonality in population studies. Molecular Ecology, 16(24), 5115–5139. https://doi.org/10.1111/j.1365-294X.2007.03535.x

Arnaud‐Haond

Stoeckel

Bailleul

(2020). New insights into the population genetics of partially clonal organisms: When seagrass data meet theoretical expectations. Molecular Ecology, 29(17), 3248–3260. https://doi.org/10.1111/mec.15532

10.

Arroyo‐Rodríguez

Melo

F. P. L.

Martínez‐Ramos

Bongers

Chazdon

R. L.

Meave

J. A.

Norden

Santos

B. A.

Leal

I. R.

Tabarelli

(2017). Multiple successional pathways in human‐modified tropical landscapes: new insights from forest succession, forest fragmentation and landscape ecology research. Biological Reviews, 92(1), 326–340. https://doi.org/10.1111/brv.12231

11.

Balloux

(2004). Heterozygote excess in small populations and the heterozygote‐excess effective population size. Evolution, 58(9), 1891–1900. https://doi.org/10.1111/j.0014-3820.2004.tb00477.x

12.

Balloux

Lehmann

de Meeûs

(2003). The Population Genetics of Clonal and Partially Clonal Diploids. Genetics, 164(4), 1635–1644. https://doi.org/10.1093/genetics/164.4.1635

13.

Barbier

E. B.

Hochard

J. P.

(2018). Land degradation and poverty. Nature Sustainability, 1(11), 623–631. https://doi.org/10.1038/s41893-018-0155-4

14.

Binggeli

(2003). Introduced and invasive plants. In Goodman

S.M.

Benstead

J.P.

(Eds.), The Natural History of Madagascar (pp. 257–268). University of Chicago Press.

15.

Bond

W. J.

Silander Jr

J. A.

Ranaivonasy

Ratsirarson

(2008). The antiquity of Madagascar’s grasslands and the rise of C4 grassy biomes. Journal of Biogeography, 35(10), 1743–1758. https://doi.org/10.1111/j.1365-2699.2008.01923.x

16.

Botomanga

Jeannoda

V. H.

Fuzzati

Ramarosandratana

A. V.

(2024). Morpho-anatomical responses of leafless Vanilla spp. roots to drought and habitat degradation. Flora, 317, 152562. https://doi.org/10.1016/j.flora.2024.152562

17.

Bucharova

Bossdorf

Hölzel

Kollmann

Prasse

Durka

(2019). Mix and match: regional admixture provenancing strikes a balance among different seed-sourcing strategies for ecological restoration. Conservation Genetics, 20(1), 7–17. https://doi.org/10.1007/s10592-018-1067-6

18.

Carter

Herold

Avitabile

de Bruin

De Sy

Kooistra

Rufino

M. C.

(2017). Agriculture-driven deforestation in the tropics from 1990–2015: emissions, trends and uncertainties. Environmental Research Letters, 13(1), 1–13. https://doi.org/10.1088/1748-9326/aa9ea4

19.

Catalá

E. I.

(2011). Los conceptos de especies indicadoras, paraguas, banderas y claves: su uso y abuso en ecología de la conservación. Interciencia, 36(1), 31–38. https://www.redalyc.org/articulo.oa?id=33917727005

20.

Chao

(1984). Nonparametric estimation of the number of classes in a population. Scandinavian Journal of Statistics, 11(4), 265–270. https://www.jstor.org/stable/4615964

21.

Chao

(1987). Estimating the population size for capture-recapture data with unequal catchability. Biometrics, 43(4), 783–791. https://doi.org/10.2307/2531532

22.

Cheptou

P.-O.

Hargreaves

A. L.

Bonte

Jacquemyn

(2017). Adaptation to fragmentation: evolutionary dynamics driven by human influences. Philosophical Transactions of the Royal Society B: Biological Sciences, 372(1712), 1. https://doi.org/10.1098/rstb.2016.0037

23.

Chung

M. Y.

Nason

J. D.

López-Pujol

Yamashiro

Yang

B.-Y.

Luo

Y.-B.

Chung

M. G.

(2014). Genetic consequences of fragmentation on populations of the terrestrial orchid Cymbidium goeringii. Biological Conservation, 170, 222–231. https://doi.org/10.1016/j.biocon.2013.12.005

24.

Clemente‐Arenas

E. R.

Trujillo‐Rodriguez

L. R.

Hilário

R. R.

Irume

M. V.

Zartman

C. E.

de Toledo

J. J.

(2023). Phorophyte size and soil profiles differentially correlate with community structure among hemiepiphytes and nomadic vines. Biotropica, 55(2), 368–381. https://doi.org/10.1111/btp.13191

25.

Colwell

R. K.

Elsensohn

J. E.

(2014). EstimateS turns 20: statistical estimation of species richness and shared species from samples, with non‐parametric extrapolation. Ecography, 37(6), 609–613. https://doi.org/10.1111/ecog.00814

26.

Cortés-Anzúres

B. O.

Corona-López

A. M.

Damon

Mata-Rosas

Flores-Palacios

(2020). Phorophyte type determines epiphyte-phorophyte network structure in a Mexican oak forest. Flora, 272, 151704. https://doi.org/10.1016/j.flora.2020.151704

27.

Cribb

Hermans

(2009). Field guide to the orchids of Madagascar. Royal Botanic Gardens.

28.

Cui

Qin

Huang

Gan

Jiang

Huang

Yang

Xiang

(2022). High‐throughput sequencing‐based microsatellite genotyping for polyploids to resolve allele dosage uncertainty and improve analyses of genetic diversity, structure and differentiation: A case study of the hexaploid Camellia oleifera. Molecular Ecology Resources, 22(1), 199–211. https://doi.org/10.1111/1755-0998.13469

29.

de Oliveira

R. T.

da Silva Oliveira

J. P.

Macedo

A. F.

(2022). Vanilla beyond Vanilla planifolia and Vanilla tahitensis: Taxonomy and Historical Notes, Reproductive Biology, and Metabolites. Plants, 11(23), 3311. https://doi.org/10.3390/plants11233311

30.

Depecker

Verleysen

Asimonyio

J. A.

Hatangi

Kambale

J.-L.

Mwanga Mwanga

Ebele

Dhed’a

Bawin

Staelens

(2023). Genetic diversity and structure in wild Robusta coffee (Coffea canephora A. Froehner) populations in Yangambi (DR Congo) and their relation to forest disturbance. Heredity, 130(3), 145–153. https://doi.org/10.1038/s41437-022-00588-0

31.

Derzhavina

N. M.

(2019). Experience of a synthetic approach to an ecological classification of vascular epiphytes. Contemporary Problems of Ecology, 12, 434–443. https://doi.org/10.1134/S1995425519050044

32.

Dia

Guillou

Mauger

Bigeard

Marie

Valero

Destombe

(2014). Spatiotemporal changes in the genetic diversity of harmful algal blooms caused by the toxic dinoflagellate Alexandrium minutum. Molecular Ecology, 23(3), 549–560. https://doi.org/10.1111/mec.12617

33.

Waples

R. S.

Peel

Macbeth

G. M.

Tillett

B. J.

Ovenden

J. R.

(2014). NeEstimator v2: re‐implementation of software for the estimation of contemporary effective population size (Ne) from genetic data. Molecular Ecology Resources, 14(1), 209–214. https://doi.org/10.1111/1755-0998.12157

34.

Ehlers

B. K.

Pedersen

H. Æ.

(2000). Genetic variation in three species of Epipactis (Orchidaceae): geographic scale and evolutionary inferences. Biological Journal of the Linnean Society, 69(3), 411–430. https://doi.org/10.1111/j.1095-8312.2000.tb01214.x

35.

Ellstrand

N. C.

Roose

M. L.

(1987). Patterns of genotypic diversity in clonal plant species. American Journal of Botany, 74(1), 123–131. https://doi.org/10.1002/j.1537-2197.1987.tb08586.x

36.

England

P. R.

Cornuet

J.-M.

Berthier

Tallmon

D. A.

Luikart

(2006). Estimating effective population size from linkage disequilibrium: severe bias in small samples. Conservation Genetics, 7, 303–308. https://doi.org/10.1007/s10592-005-9103-8

37.

Evanno

Regnaut

Goudet

(2005). Detecting the number of clusters of individuals using the software STRUCTURE: a simulation study. Molecular Ecology, 14(8), 2611–2620. https://doi.org/10.1111/j.1365-294X.2005.02553.x

38.

Fajardo

C. G.

de Almeida Vieira

Felix

L. P.

Molina

W. F.

(2017). Negligence in the Atlantic forest, northern Brazil: a case study of an endangered orchid. Biodiversity and Conservation, 26, 1047–1063. https://doi.org/10.1007/s10531-016-1285-5

39.

Fishman

Sweigart

A. L.

(2018). When two rights make a wrong: the evolutionary genetics of plant hybrid incompatibilities. Annual Review of Plant Biology, 69(1), 707–731. https://doi.org/10.1146/annurev-arplant-042817-040113

40.

Flanagan

N. S.

Chavarriaga

Mosquera‐Espinosa

A. T.

(2018). Conservation and sustainable use of Vanilla crop wild relatives in Colombia. In Havkin-Frenkel

Daphna

Faith

C. B.

(Eds.), Handbook of Vanilla Science and Technology, (pp. 85–109). John Wiley & Sons Ltd. https://doi.org/10.1002/9781119377320.ch6

41.

Food and Agriculture Organization of the United Nations (2015). Global Forest Resources Assessment 2015: How are the World’s Forests Changing?.

42.

Frankham

Bradshaw

C. J. A.

Brook

B. W.

(2014). Genetics in conservation management: revised recommendations for the 50/500 rules, Red List criteria and population viability analyses. Biological Conservation, 170, 56–63. https://doi.org/10.1016/j.biocon.2013.12.036

43.

Frascaroli

Zannini

Acosta

A. T. R.

Chiarucci

d’Agostino

Nascimbene

(2019). Sacred natural sites in Italy have landscape characteristics complementary to protected areas: Implications for policy and planning. Applied Geography, 113, 102100. https://doi.org/10.1016/j.apgeog.2019.102100

44.

Garrido-Cardenas

J. A.

Mesa-Valle

Manzano-Agugliaro

(2018). Trends in plant research using molecular markers. Planta, 247, 543–557. https://doi.org/10.1007/s00425-017-2829-y

45.

Gautier

Chatelain

Spichiger

R.-E.

(1994). Presentation of a releve method for vegetation studies based on high resolution satellite imagery. In Proceedings of XIII th plenary meeting of AETFAT, Seyani

J. H.

Chikuni

A. C.

(Eds.), (pp. 1339-1350). National Herbarium and Botanic Gardens of Malawi.

46.

Gerolamo

C. S.

Fogaça

M. D.

Bastos

C. L.

(2022). How dehydration affects stem bending stiffness and leaf toughness after sampling of the liana Amphilophium crucigerum (L.) LG Lohmann (Bignoniaceae). Hoehnea, 49, e592021. https://doi.org/10.1590/2236-8906-59-2021

47.

Gianoli

Saldaña

Jiménez-Castillo

(2012). Ecophysiological traits may explain the abundance of climbing plant species across the light gradient in a temperate rainforest. PLoS One, 7(6), e38831. https://doi.org/10.1371/journal.pone.0038831

48.

Gigant

R. L.

Bory

Grisoni

Besse

(2011). The Dynamical Processes of Biodiversity: Case Studies of Evolution and Spatial Distribution. In Grillo

Oscar

Venora

Gianfranco

(Eds), Biodiversity and evolution in the Vanilla genus, (pp. 1–26). InTech.

49.

Gigant

R. L.

De Bruyn

Church

Humeau

Gauvin-Bialecki

Pailler

Grisoni

Besse

(2014). Active sexual reproduction but no sign of genetic diversity in range-edge populations of Vanilla roscheri Rchb. f.(Orchidaceae) in South Africa. Conservation Genetics, 15(6), 1403–1415. https://doi.org/10.1007/s10592-014-0626-8

50.

Gigant

R. L.

De Bruyn

M’sa

Viscardi

Gigord

Gauvin-Bialecki

Pailler

Humeau

Grisoni

Besse

(2016a). Combining pollination ecology and fine-scale spatial genetic structure analysis to unravel the reproductive strategy of an insular threatened orchid. South African Journal of Botany, 105, 25–35. https://doi.org/10.1016/j.sajb.2016.02.205

51.

Gigant

R. L.

Rakotomanga

Citadelle

Silvestre

Grisoni

Besse

(2016b). Microsatellite markers confirm self-pollination and autogamy in wild populations of Vanilla mexicana Mill. (syn. V. inodora) (Orchidaceae) in the island of Guadeloupe. In Abdurakhmonov

Ibrokhim Y.

(Ed.), Microsatellite Markers (pp. 73–94). BoD – Books on Demand.

52.

Glémin

Bataillon

Ronfort

Mignot

Olivieri

(2001). Inbreeding depression in small populations of self-incompatible plants. Genetics, 159(3), 1217–1229. https://doi.org/10.1093/genetics/159.3.1217

53.

Godron

Daget

Long

Sauvage

Emberger

Le Gloch

Poissonnet

Wacquant

J. P.

(1983). Relevé méthodologique de la végétation et du milieu, code et transcription sur carte perforée. Centre National de la Recherche Scientifique, Paris, 292p.

54.

González

A. V

Gómez‐Silva

Ramírez

M. J.

Fontúrbel

F. E.

(2020). Meta‐analysis of the differential effects of habitat fragmentation and degradation on plant genetic diversity. Conservation Biology, 34(3), 711–720. https://doi.org/10.1111/cobi.13422

55.

González-Orellana

Mendoza

A. S.

Tremblay

R. L.

Ackerman

J. D.

(2024). Host suitability for germination differs from that of later stages of development in a rare epiphytic orchid. Lankesteriana: International Journal on Orchidology, 93–114. https://doi.org/10.15517/lank.v24i1.59617

56.

Goodman

S. M.

(2022). The new natural history of Madagascar. Princeton University Press.

57.

Gowland

K. M.

Wood

Clements

M. A.

Nicotra

A. B.

(2011). Significant phorophyte (substrate) bias is not explained by fitness benefits in three epiphytic orchid species. American Journal of Botany, 98(2), 197–206. https://doi.org/10.3732/ajb.1000241

58.

Grisoni

Nany

(2021). The beautiful hills: half a century of vanilla (Vanilla planifolia Jacks. ex Andrews) breeding in Madagascar. Genetic Resources and Crop Evolution, 68(5), 1691–1708. https://doi.org/10.1007/s10722-021-01119-2

59.

Guo

S.-W.

Thompson

E. A.

(1992). Performing the exact test of Hardy-Weinberg proportion for multiple alleles. Biometrics, 48(2), 361–372. https://doi.org/10.2307/2532296

60.

Halkett

Simon

J.-C.

Balloux

(2005). Tackling the population genetics of clonal and partially clonal organisms. Trends in Ecology & Evolution, 20(4), 194–201. https://doi.org/10.1016/j.tree.2005.01.001

61.

Haq

S. M.

Rashid

Khuroo

A. A.

Malik

Z. A.

Malik

A. H.

(2019). Anthropogenic disturbances alter community structure in the forests of Kashmir Himalaya. Tropical Ecology, 60, 6–15. https://doi.org/10.1007/s42965-019-00001-8

62.

Harper

G. J.

Steininger

M. K.

Tucker

C. J.

Juhn

Hawkins

(2007). Fifty years of deforestation and forest fragmentation in Madagascar. Environmental Conservation, 34(4), 325–333. https://doi.org/10.1017/S0376892907004262

63.

Peng

Yang

Zhang

(2023). The response of seedlings and saplings to canopy structure and light in different gaps in a spruce-fir mixed stand in Changbai Mountains, China. Forest Ecology and Management, 546, 121365. https://doi.org/10.1016/j.foreco.2023.121365

64.

Hedrick

P. W.

Hellsten

Grattapaglia

(2016). Examining the cause of high inbreeding depression: analysis of whole‐genome sequence data in 28 selfed progeny of Eucalyptus grandis. New Phytologist, 209(2), 600–611. https://doi.org/10.1111/nph.13639

65.

Hoban

Bruford

Jackson

J. D.

Lopes-Fernandes

Heuertz

Hohenlohe

P. A.

Paz-Vinas

Sjögren-Gulve

Segelbacher

Vernesi

(2020). Genetic diversity targets and indicators in the CBD post-2020 Global Biodiversity Framework must be improved. Biological Conservation, 248, 108654. https://doi.org/10.1016/j.biocon.2020.108654

66.

Hothorn

Bretz

Westfall

(2015). R package multcomp: Simultaneous Inference in General Parametric Models. Published Online in the CRAN Repository .

67.

Huang

X.-M.

Lin

Gao

J.-Y.

(2018). Host-specificity of symbiotic mycorrhizal fungi for enhancing seed germination, protocorm formation and seedling development of over-collected medicinal orchid, Dendrobium devonianum. Journal of Microbiology, 56, 42–48. https://doi.org/10.1007/s12275-018-7225-1

68.

Irwin

M. T.

Wright

P. C.

Birkinshaw

Fisher

B. L.

Gardner

C. J.

Glos

Goodman

S. M.

Loiselle

Rabeson

Raharison

J.-L.

(2010). Patterns of species change in anthropogenically disturbed forests of Madagascar. Biological Conservation, 143(10), 2351–2362. https://doi.org/10.1016/j.biocon.2010.01.023

69.

Jump

A. S.

Peñuelas

(2006). Genetic effects of chronic habitat fragmentation in a wind-pollinated tree. Proceedings of the National Academy of Sciences, 103(21), 8096–8100. https://doi.org/10.1073/pnas.0510127103

70.

Kaneko

Yoshiura

Kobayashi

(2014). Sustainable living with environmental risks. Springer Nature.

71.

Kindt

M. R.

(2019). Package ‘BiodiversityR.’ Package for Community Ecology and Suitability Analysis, Version 1.10-1. The comprehensive R archive network. https://cran.r-project.org/web/packages/BiodiversityR/index.html

72.

Lamprecht

(1989). Silviculture in the tropics. Tropical forest ecosystems and their tree species-possibilities and methods for their long-term utilization. TZ-Verlag.

73.

Le Roux

J. J.

Wieczorek

A. M.

Wright

M. G.

Tran

C. T.

(2007). Super-genotype: global monoclonality defies the odds of nature. PloS One, 2(7), e590. https://doi.org/10.1371/journal.pone.0000590

74.

(2020). Effect of human disturbance on genetic structure of rare and endangered Paphiopedilum micranthum implied the habitat status. Tropical Conservation Science, 13, 1940082920942012. https://doi.org/10.1177/1940082920942012

75.

Lindenmayer

D. B.

Laurance

W. F.

(2017). The ecology, distribution, conservation and management of large old trees. Biological Reviews, 92(3), 1434–1458. https://doi.org/10.1111/brv.12290

76.

Lohbeck

Poorter

Martínez-Ramos

Bongers

(2015). Biomass is the main driver of changes in ecosystem process rates during tropical forest succession. Ecology, 96(5), 1242–1252. https://doi.org/10.1890/14-0472.1

77.

Madagascar Catalogue . (2023). Catalogue of the vascular plants of Madagascar. Missouri Botanical Garden. [Internet]. [accessed 2023 Jan 11]https://www.tropicos.org/Project/Madagascar

78.

Måren

I. E.

Sharma

L. N.

(2021). Seeing the wood for the trees: Carbon storage and conservation in temperate forests of the Himalayas. Forest Ecology and Management, 487, 119010. https://doi.org/10.1016/j.foreco.2021.119010

79.

Matesanz

Sultan

S. E.

(2013). High‐performance genotypes in an introduced plant: insights to future invasiveness. Ecology, 94(11), 2464–2474. https://doi.org/10.1890/12-1359.1

80.

McCormick

M. K.

Jacquemyn

(2014). What constrains the distribution of orchid populations? New Phytologist, 202(2), 392–400. https://doi.org/10.1111/nph.12639

81.

Melo

Vargas

(2003). Evaluación ecológica y silvicultural de ecosistemas boscosos. Ibagué: Universidad Del Tolima, CRQ, Carder, Corpocaldas, Cortolima, 4–7.

82.

Mérot

Llaurens

Normandeau

Bernatchez

Wellenreuther

(2020). Balancing selection via life-history trade-offs maintains an inversion polymorphism in a seaweed fly. Nature Communications, 11(1), 670. https://doi.org/10.1038/s41467-020-14479-7

83.

Minasiewicz

Znaniecka

J. M.

Górniak

Kawiński

(2018). Spatial genetic structure of an endangered orchid Cypripedium calceolus (Orchidaceae) at a regional scale: limited gene flow in a fragmented landscape. Conservation Genetics, 19, 1449–1460. https://doi.org/10.1007/s10592-018-1113-4

84.

Miranda

B. C.

Sansevero

J. B. B.

Amorim

T. de A.

Rodrigues

P. J. F. P.

Nascimento

M. T.

Braga

J. M. A.

(2022). Are liana communities in linear canopy openings subject to edge effects? Plant Ecology, 223(8), 1023–1034. https://doi.org/10.1007/s11258-022-01253-6

85.

Mishra

Kumari

Choudhary

Gondhali

(2021). Evaluation and comparison of population genetics software in Rabari Tribe of Gujarat population. Egyptian Journal of Forensic Sciences, 11, 1–12. https://doi.org/10.1186/s41935-021-00239-1

86.

Miyamoto

(2020). Poverty reduction saves forests sustainably: Lessons for deforestation policies. World Development, 127, 104746. https://doi.org/10.1016/j.worlddev.2019.104746

87.

Moat

Smith

P. P.

(2007). Atlas of the vegetation of Madagascar. Royal Botanic Gardens, Kew.

88.

Mori

Kamijo

Masaki

(2016). Liana distribution and community structure in an old-growth temperate forest: the relative importance of past disturbances, host trees, and microsite characteristics. Plant Ecology, 217, 1171–1182. https://doi.org/10.1007/s11258-016-0641-6

89.

Nadeem

M. A.

Nawaz

M. A.

Shahid

M. Q.

Doğan

Comertpay

Yıldız

Hatipoğlu

Ahmad

Alsaleh

Labhane

(2018). DNA molecular markers in plant breeding: current status and recent advancements in genomic selection and genome editing. Biotechnology & Biotechnological Equipment, 32(2), 261–285. https://doi.org/10.1080/13102818.2017.1400401

90.

Nielsen

L. R.

Siegismund

H. R.

(1999). Interspecific differentiation and hybridization in Vanilla species (Orchidaceae). Heredity, 83(5), 560–567. https://doi.org/10.1038/sj.hdy.6885880

91.

Orantes

L. C.

Zhang

Mian

M. A. R.

Michel

A. P.

(2012). Maintaining genetic diversity and population panmixia through dispersal and not gene flow in a holocyclic heteroecious aphid species. Heredity, 109(2), 127–134. https://doi.org/10.1038/hdy.2012.21

92.

Orihuela

R. L. L.

Waechter

J. L.

(2010). Host size and abundance of hemiepiphytes in a subtropical stand of Brazilian Atlantic Forest. Journal of Tropical Ecology, 26(1), 119–122. https://doi.org/10.1017/S0266467409990496

93.

Ormsby

A. A.

(2021). Diverse values and benefits of urban sacred natural sites. Trees, Forests and People, 6, 100136. https://doi.org/10.1016/j.tfp.2021.100136

94.

Otero

J. T.

Aragón

Ackerman

J. D.

(2007). Site variation in spatial aggregation and phorophyte preference in Psychilis monensis (Orchidaceae). Biotropica, 39(2), 227–231. https://doi.org/10.1111/j.1744-7429.2006.00258.x

95.

Pandey

Richards

Sharma

(2015). Microsatellite-based genetic diversity patterns in disjunct populations of a rare orchid. Genetica, 143, 693–704. https://doi.org/10.1007/s10709-015-9867-9

96.

Peakall

R. O. D.

Smouse

P. E.

(2006). GENALEX 6: genetic analysis in Excel. Population genetic software for teaching and research. Molecular Ecology Notes, 6(1), 288–295. https://doi.org/10.1111/j.1471-8286.2005.01155.x

97.

Petersson

, (2015). Pollination biology of the endemic orchid Vanilla bosseri in Madagascar. Master Thesis. Uppsala University, Disciplinary Domain of Science and Technology, Biology Education Center.

98.

Pecoraro

Rasmussen

H. N.

Gomes

S. I. F.

Wang

Merckx

V. S. F. T.

Cai

Rasmussen

F. N.

(2021). Fungal diversity driven by bark features affects phorophyte preference in epiphytic orchids from southern China. Scientific Reports, 11(1), 11287. https://doi.org/10.1038/s41598-021-90877-1

99.

Probowati

Koga

Harada

Nagano

A. J.

Ishimaru

Ohshima

Fukuda

(2023). RAD-Seq analysis of wild Japanese garlic (Allium macrostemon Bunge) growing in Japan revealed that this neglected crop was previously actively utilized. Scientific Reports, 13(1), 16354. https://doi.org/10.1038/s41598-023-43537-5

100.

Promega (2007). Usage information of Gotaq Colorless Master Mix. [accessed 2023 Jan 14]https://www.promega.com/resources/protocols/product-information-sheets/g/gotaq-colorless-master-mix-m713-protocol/

101.

Portères

(1954). Le genre Vanilla et ses espèces. In Lechevalier

(Ed.), Le vanillier et la vanille dans le monde (pp. 94–290). Paris: Editions Paul Lechevalier.

102.

Pritchard

J. K.

Stephens

Donnelly

(2000). Inference of population structure using multilocus genotype data. Genetics, 155(2), 945–959. https://doi.org/10.1093/genetics/155.2.945

103.

Rakotoarivelo

N. H.

Randrianarivony

T. N.

Rakotoarivony

Randrianasolo

(2019). “Mangidy”: Malagasy Folk Beverages Sold in Madagascar’s Market Places. Ethnobotany Research and Applications, 18, 1–14. https://ethnobotanyjournal.org/index.php/era/article/view/1565

104.

Ralimanana

Perrigo

A. L.

Smith

R. J.

Borrell

J. S.

Faurby

Rajaonah

M. T.

Randriamboavonjy

Vorontsova

M. S.

Cooke

R. S. C.

Phelps

L. N.

(2022). Madagascar’s extraordinary biodiversity: Threats and opportunities. Science, 378(6623), eadf1466. https://doi.org/10.1126/science.adf1466

105.

Ranaivoson

Rakouth

Buerkert

Brinkmann

(2017). Wood biomass availability for smallholder charcoal production in dry forest and savannah ecosystems of south-western Madagascar. Journal of Arid Environments, 146, 86–94. https://doi.org/10.1016/j.jaridenv.2017.07.002

106.

Randriamiharisoa

M. N.

Kuhlman

A. R.

Jeannoda

Rabarison

Rakotoarivelo

Randrianarivony

Raktoarivony

Randrianasolo

Bussmann

R. W.

(2015). Medicinal plants sold in the markets of Antananarivo, Madagascar. Journal of Ethnobiology and Ethnomedicine, 11, 1–13. https://doi.org/10.1186/s13002-015-0046-y

107.

Randrianasolo

Z. H.

Razafimahatratra

A. R.

Razafinarivo

R. N. G.

Randrianary

Rakotovololonalimanana

Rajemison

A. H.

Mamitiana

Andriamanalina

R. L.

Rakotosoa

Ramananantoandro

(2019). Which allometric models are the most appropriate for estimating aboveground biomass in secondary forests of Madagascar with Ravenala madagascariensis? Scientific African, 6, e00147. https://doi.org/10.1016/j.sciaf.2019.e00147

108.

Randrianjafizanaka

S. T.

(2014). Development of Technical Reintroduction and Population Monitoring of Radiated Tortoises Astrochelys radiata in Southern Madagascar: Ampotaka. Open Access Library Journal, 1(6), 1–8. https://doi.org/10.4236/oalib.1100279

109.

Rasmussen

H. N.

Dixon

K. W.

Jersáková

Těšitelová

(2015). Germination and seedling establishment in orchids: a complex of requirements. Annals of Botany, 116(3), 391–402. https://doi.org/10.1093/aob/mcv087

110.

R Core Team (2023). R: A Language and Environment for Statistical Computing_. R Foundation for Statistical Computing, Vienna, Austria. <https://www.R-project.org/>

111.

Rousset

(2008). A complete re-implementation of the GENEPOP software for software for teaching and research. Molecular Ecolology Resources, 8, 103–106.

112.

Rogers

D. L.

Washburn

L. K.

Birker

Labbé

M. A.

Campbell

M. A.

Schreier

A. D.

(2024). Genomic and common garden data reveal significant genetic differentiation in the endangered San Fernando Valley spineflower Chorizanthe parryi var. fernandina. Conservation Genetics, 1–18. https://doi.org/10.1007/s10592-024-01610-0

113.

Rowe

(2018). Lianas. Current Biology, 28(6), R249–R252. https://doi.org/10.1016/j.cub.2018.01.028

114.

Rubio-Salcedo

Merinero

Martínez

(2015). Tree species and microhabitat influence the population structure of the epiphytic lichen Lobaria pulmonaria. Fungal Ecology, 18, 1–9. https://doi.org/10.1016/j.funeco.2015.08.002

115.

Schlaepfer

D. R.

Braschler

Rusterholz

Baur

(2018). Genetic effects of anthropogenic habitat fragmentation on remnant animal and plant populations: A meta‐analysis. Ecosphere, 9(10), e02488. https://doi.org/10.1002/ecs2.2488

116.

Schnitzer

S. A.

DeFilippis

D. M.

Visser

Estrada‐Villegas

Rivera‐Camaña

Bernal

Peréz

Valdéz

Aguilar

(2021). Local canopy disturbance as an explanation for long‐term increases in liana abundance. Ecology Letters, 24(12), 2635–2647. https://doi.org/10.1111/ele.13881

117.

Schoen

D. J.

Baldwin

S. J.

(2023). Self‐incompatibility and the genetic architecture of inbreeding depression. New Phytologist, 237(3), 1040–1049. https://doi.org/10.1111/nph.18550

118.

Schöfl

Sauter

Klasberg

Schmidt

A. H.

Lange

(2018). P061Allele-level genotyping of KIR2DL4 in large european population samples reveals highly significant heterozygote excess for 9A/10A allelic variants. Human Immunology, 79, 106. https://doi.org/10.1016/j.humimm.2018.07.120

119.

Seddon

Tobias

J. A.

(2007). Population size and habitat associations of the Long-tailed Ground-roller Uratelornis chimaera. Bird Conservation International, 17(1), 1–12. https://doi.org/10.1017/S095927090600058X

120.

Sfair

J. C.

de Bello

de França

T. Q.

Baldauf

Tabarelli

(2018). Chronic human disturbance affects plant trait distribution in a seasonally dry tropical forest. Environmental Research Letters, 13(2), 025005. https://doi.org/10.1088/1748-9326/aa9f5e

121.

Shen

Song

Collart

Guisan

H.-X.

Dong

J.-L.

Vanderpoorten

(2022). What makes a good phorophyte? Predicting occupancy, species richness and abundance of vascular epiphytes in a lowland seasonal tropical forest. Frontiers in Forests and Global Change, 5, 1007473. https://doi.org/10.3389/ffgc.2022.1007473

122.

Silva

I. A.

van Melis

Sfair

J. C.

Martins

F. R.

Ullah

(2016). Phylogenetic interactions among lianas in a southeastern Brazilian semideciduous tropical forest. South African Journal of Botany, 103, 108–125. https://doi.org/10.1016/j.sajb.2015.09.011

123.

Speak

Escobedo

F. J.

Russo

Zerbe

(2020). Total urban tree carbon storage and waste management emissions estimated using a combination of LiDAR, field measurements and an end-of-life wood approach. Journal of Cleaner Production, 256, 120420. https://doi.org/10.1016/j.jclepro.2020.120420

124.

Stoeckel

Arnaud-Haond

Krueger-Hadfield

S. A.

(2021a). The combined effect of haplodiplonty and partial clonality on genotypic and genetic diversity in a finite mutating population. Journal of Heredity, 112(1), 78–91. https://doi.org/10.1093/jhered/esaa062

125.

Stoeckel

Grange

Fernández-Manjarres

J. F.

Bilger

Frascaria-Lacoste

Mariette

(2006). Heterozygote excess in a self-incompatible and partially clonal forest tree species - Prunus avium L. Molecular Ecology, 15(8), 2109–2118. https://doi.org/10.1111/J.1365-294X.2006.02926.X

126.

Stoeckel

Masson

J.-P.

(2014). The exact distributions of F_IS under partial asexuality in small finite populations with mutation. PLoS One, 9(1), e85228. https://doi.org/10.1371/journal.pone.0085228

127.

Stoeckel

Porro

Arnaud‐Haond

(2021b). The discernible and hidden effects of clonality on the genotypic and genetic states of populations: improving our estimation of clonal rates. Molecular Ecology Resources, 21(4), 1068–1084. https://doi.org/10.1111/1755-0998.13316

128.

Templeton

A. R.

(2018). Human population genetics and genomics. Academic Press.

129.

Tengö

Johansson

Rakotondrasoa

Lundberg

Andriamaherilala

J.-A.

Rakotoarisoa

J.-A.

Elmqvist

(2007). Taboos and forest governance: informal protection of hot spot dry forest in southern Madagascar. AMBIO: A Journal of the Human Environment, 36(8), 683–691. https://doi.org/10.1579/0044-7447(2007)36[683:TAFGIP]2.0.CO;2

130.

The Royal Botanic Gardens Kew (2023). Plants of the world online. [Internet]. [accessed 2023 Jan 11].https://www.plantsoftheworldonline.org/

131.

Timsina

Rokaya

Münzbergová

Kindlmann

Shrestha

Bhattarai

Raskoti

. (2016). Diversity, distribution and host-species associations of epiphytic orchids in Nepal. Biodiversity and Conservation, 25, 2803–2819. https://doi.org/10.1007/s10531-016-1205-8

132.

Torres‐Vanegas

Hadley

A. S.

Kormann

U. G.

Jones

F. A.

Betts

M. G.

Wagner

H. H.

(2021). Tropical deforestation reduces plant mating quality by shifting the functional composition of pollinator communities. Journal of Ecology, 109(4), 1730–1746. https://doi.org/10.1111/1365-2745.13594

133.

Vásquez-Grandón

Donoso

P. J.

Gerding

(2018). Forest degradation: when is a forest degraded? Forests, 9(11) 726. https://doi.org/10.3390/f9110726

134.

Vieilledent

Grinand

Rakotomalala

F. A.

Ranaivosoa

Rakotoarijaona

J.-R.

Allnutt

T. F.

Achard

(2018). Combining global tree cover loss data with historical national forest cover maps to look at six decades of deforestation and forest fragmentation in Madagascar. Biological Conservation, 222, 189–197. https://doi.org/10.1016/j.biocon.2018.04.008

135.

Visser

M. D.

Muller‐Landau

H. C.

Schnitzer

S. A.

de Kroon

Jongejans

Wright

S. J.

(2018). A host–parasite model explains variation in liana infestation among co‐occurring tree species. Journal of Ecology, 106(6), 2435–2445. https://doi.org/10.1111/1365-2745.12997

136.

Wagner

Mendieta-Leiva

Zotz

(2015). Host specificity in vascular epiphytes: a review of methodology, empirical evidence and potential mechanisms. AoB Plants, 7. https://doi.org/10.1093/aobpla/plu092

137.

Wang

Song

Meng

Zhu

Zhao

(2017). Influence of host tree species on isolation and communities of mycorrhizal and endophytic fungi from roots of a tropical epiphytic orchid, Dendrobium sinense (Orchidaceae). Mycorrhiza, 27, 709–718. https://doi.org/10.1007/s00572-017-0787-7

138.

Watteyn

Reubens

Bolaños

J. B. A.

Campos

F. S.

Silva

A. P.

Karremans

A. P.

Muys

(2023). Cultivation potential of Vanilla crop wild relatives in two contrasting land use systems. European Journal of Agronomy, 149, 126890. https://doi.org/10.1016/j.eja.2023.126890

139.

Wei

Lyu

Xiao

Wang

Yang

Jiang

(2023). Impacts of ecological restoration on the genetic diversity of plant species: A global meta‐analysis. Journal of Applied Ecology, 60(6), 1149–1160. https://doi.org/10.1111/1365-2664.14390

140.

Westhoff

Van Der Maarel

(1978). The Braun-Blanquet Approach. In Whittaker

R.H.

(Eds), Classification of Plant Communities (pp. 287–399). Springer, Dordrecht. https://doi.org/10.1007/978-94-009-9183-5_9

141.

Wright

(1949). The genetical structure of populations. Annals of Eugenics, 15(1), 323–354. https://doi.org/10.1111/j.1469-1809.1949.tb02451.x

142.

Xiao

Cao

Dong

(2016). The diverse effects of habitat fragmentation on plant–pollinator interactions. Plant Ecology, 217, 857–868. https://doi.org/10.1007/s11258-016-0608-7

143.

Yamashita

Satoh

Kurosawa

Kaneko

(2023). Genetic diversity and structure of the endangered lady’s slipper orchid Cypripedium japonicum Thunb. (Orchidaceae) in Japan. Population Ecology, 65(1), 54–63. https://doi.org/10.1002/1438-390X.12134

144.

Zaehringer

J. G.

Eckert

Messerli

(2015). Revealing regional deforestation dynamics in North-Eastern Madagascar—Insights from multi-temporal land cover change analysis. Land, 4(2), 454–474. https://doi.org/10.3390/land4020454

145.

Zannini

Frascaroli

Nascimbene

Persico

Halley

J. M.

Stara

Midolo

Chiarucci

(2021). Sacred natural sites and biodiversity conservation: a systematic review. Biodiversity and Conservation, 30(13), 3747–3762. https://doi.org/10.1007/s10531-021-02296-3

146.

Zarate‐García

A. M.

Noguera‐Savelli

Andrade‐Canto

S. B.

Zavaleta‐Mancera

H. A.

Gauthier

Alatorre‐Cobos

(2020). Bark water storage capacity influences epiphytic orchid preference for host trees. American Journal of Botany, 107(5), 726–734. https://doi.org/10.1002/ajb2.1470

147.

Zimmerman

S. J.

Aldridge

C. L.

Oyler-McCance

S. J.

(2020). An empirical comparison of population genetic analyses using microsatellite and SNP data for a species of conservation concern. BMC Genomics, 21, 1–16. https://doi.org/10.1186/s12864-020-06783-9

Supplementary Material

Please find the following supplemental material available below.

For Open Access articles published under a Creative Commons License, all supplemental material carries the same license as the article it is associated with.

For non-Open Access articles published, all supplemental material carries a non-exclusive license, and permission requests for re-use of supplemental material or any part of supplemental material shall be sent directly to the copyright owner as specified in the copyright notice associated with the article.

0.00 MB

0.96 MB