# Trait-dependent declines of threatened endemic trees following plant invasion on a tropical oceanic island

Robin Pouteau, Élise Adrien, Claudine Ah-Peng, Sébastien Albert, Olivier Flores, Bastian Laforgue, Christophe Lavergne, Jean-Yves Meyer, Anaëlle Regen, Margaux Rojat, et al.

<figure><img src="image_1.png" /></figure>

## To cite this version:

Robin Pouteau, Élise Adrien, Claudine Ah-Peng, Sébastien Albert, Olivier Flores, et al. Trait-dependent declines of threatened endemic trees following plant invasion on a tropical oceanic island. Biological Conservation, 2026, 315, pp.111665. 10.1016/j.biocon.2025.111665. hal-05448009

HAL Id: hal-05448009

https://hal.science/hal-05448009v1

Submitted on 8 Jan 2026

HAL is a multi-disciplinary open access
archive for the deposit and dissemination of sci-
entific research documents, whether they are pub-
lished or not. The documents may come from
teaching and research institutions in France or
abroad, or from public or private research centers.

L'archive ouverte pluridisciplinaire HAL, est
destinée au dépôt et à la diffusion de documents
scientifiques de niveau recherche, publiés ou non,
émanant des établissements d'enseignement et de
recherche français ou étrangers, des laboratoires
publics ou privés.

<figure><img src="image_2.png" /></figure>

Distributed under a Creative Commons CC BY 4.0 - Attribution - International License
# Trait-dependent declines of threatened endemic trees following plant invasion on a tropical oceanic island

Robin Pouteau<sup>a,b,*</sup>, Élise Adrien<sup>a,b</sup>, Claudine Ah-Peng<sup>c</sup>, Sébastien Albert<sup>d</sup>, Olivier Flores<sup>c</sup>, Bastian Laforgue<sup>c</sup>, Christophe Lavergne<sup>e</sup>, Jean-Yves Meyer<sup>f,g</sup>, Anaëlle Regen<sup>a,b</sup>, Margaux Rojat<sup>a,b</sup>, Sarah Roussel<sup>e</sup>

## ARTICLE INFO

### Keywords:

- Competitive interaction

- Extinction risk

- Invasive alien plant

- Population trend

- Survival

- Tree regeneration

## ABSTRACT

Biological invasions are a leading cause of the ongoing biodiversity crisis, and particularly so on islands. However, the role of invasive alien plants (IAPs) as a driver of native plant declines and extinctions remains unclear. The inherently slow and gradual nature of plant extinctions, especially that of long-lived woody species, could be a reason. Here, we examined temporal trends in subpopulations of 28 threatened endemic tree (TET) taxa. We questioned the frequency with which they decline in association with IAPs on Reunion Island (South-West Indian Ocean), and asked whether the most susceptible TET taxa exhibit characteristics that could reveal the underlying ecological mechanisms. We resurveyed 182 historically described subpopulations and tested whether observed trends of juvenile and adult TETs correlate with the abundance in IAPs using path analyses, while distinguishing TET taxa with respect to their leaf-height-seed characteristics and extinction risk. The trend of adult TETs was not affected by IAPs but that of juvenile TETs was negatively correlated with the abundance of IAPs in the understory layer. This was particularly the case for TET taxa with conservative resource-use strategies (low specific leaf area or low maximum height), whose juveniles might be particularly susceptible to competition with IAPs, but not for TET taxa with large seeds, whose metabolic reserves make seedlings more likely to attain a critical size. These findings suggest that IAPs can significantly contribute to the extinction dynamics of trees, primarily by limiting regeneration through competitive exclusion, and that their impacts vary depending on the traits of the affected taxa.

## 1. Introduction

Invasive alien species, whose introduction and spread threaten local species and ecosystems, are identified as one of the main causes of the ongoing biodiversity crisis (Roy et al., 2023; Watson et al., 2019). However, the extent to which invasive alien plants (IAPs) have brought about the decline and extinction of native plants has fuelled much debate (Denslow, 2003; Gurevitch and Padilla, 2004; Sax et al., 2002; Sax and Gaines, 2008). Available data suggest that the role of IAPs as a driver of plant extinction remains limited (Bellard et al., 2016; Sax and

Gaines, 2008). According to the 2021 version of the IUCN Red List of threatened species, only three seed-plant species, all woody and endemic (geographically restricted) to the oceanic islands of Kauai (Hawaii, USA), North Island (New Zealand) and Pitcairn (UK), have gone extinct in the wild or completely extinct as a consequence of IAPs that have been identified (Pouteau et al., 2023).

The lack of evidence that IAPs can cause widespread extinction of native plant species may stem from the inherently slow and gradual nature of extinctions (Downey and Richardson, 2016). This is likely to be exacerbated for woody plant species with long generation times whose

<footnote><sup>a</sup> AMAP, Univ Montpellier, CIRAD, CNRS, INRAE, IRD, Montpellier, France</footnote>

<footnote><sup>b</sup> AMAP, IRD, Pôle de Protection des Plantes, Saint-Pierre, Reunion</footnote>

<footnote><sup>c</sup> Université de La Réunion, UMR PVBMT, Saint-Pierre, Reunion</footnote>

<footnote><sup>d</sup> Université de Lorraine, AgroParisTech, INRAE, SILVA, Nancy, France</footnote>

<footnote><sup>e</sup> CBN-CPIE Mascarin, Saint-Leu, Reunion</footnote>

<footnote><sup>f</sup> Research Department, Government of French Polynesia, Papeete, French Polynesia</footnote>

<footnote><sup>g</sup> UPF, ILM, Ifremer, IRD, UMR SECOPOL, Punaauia, French Polynesia</footnote>

<footnote>* Corresponding author at: AMAP, Univ Montpellier, CIRAD, CNRS, INRAE, IRD, Montpellier, France.
E-mail address: robin.pouteau@ird.fr (R. Pouteau).</footnote>
extinction debt (time-delayed extinction of species following an envi-
ronmental impact) is expected to be greater than that of short-lived plant
species (Cousins and Vanhoenacker, 2011). In order to test this hy-
pothesis, long-term monitoring of tree subpopulations is required
(Downey and Richardson, 2016). Unfortunately, such long time-series of
observations are scarce, especially on islands (Borges et al., 2018). This
is unfortunate as islands, which are recognized as epicentres of both
biological invasions and species extinctions (Bellard et al., 2016;
Fernández-Palacios et al., 2021), can be useful model systems to better
understand the replacement of endemic plants by IAPs.

Among the few available island studies, a sixteen-year survey in ten
forest plots in French Polynesia revealed a 50 % increase in abundance
of IAPs, mainly *Miconia calvescens* (Melastomataceae) and *Spathodea
campanulata* (Bignoniaceae), associated with the extirpation (local
extinction) of four endemic and five native woody species, although

subpopulations of some other endemic and native species were less
affected (Meyer et al., 2025). In Mauritius, removal of IAPs, mainly
Psidium cattleyanum (Myrtaceae), on 2.5 ha of tropical forest led to the
reappearance of a presumed extinct understory island endemic tree
Ixora vaughanii (Rubiaceae) and promoted the regeneration of numerous
threatened plants within a 10-year period (Baider and Florens, 2011).
These rare examples suggest that the impact of IAPs on the extinction
dynamics of island endemic trees could be species-specific, and affect
either adult or juvenile trees. However, the underlying ecological
mechanisms remain to be clarified.

Key characteristics of island plants such as endemism, pollination and seed dispersal by vertebrates, nitrogen-fixing capacity, and woodiness were found to be closely related to their extinction dynamics and to the nature of their threats (Orihuela-Rivero et al., 2025). Similarly, identifying the characteristics of the most responsive endemic plants to

<figure><img src="image_3.png" /><figcaption>**Fig. 1.** Study sites and sampling design. A: Location of Reunion Island in the South-West Indian Ocean. B: Location of sampled sites. The map of forest types was created by combining annual rainfall data (Réchou et al., 2019) with the current forest distribution (Strasberg et al., 2005). Grey areas indicate unforested land surface. Most resurveyed sites in low elevation dry to mesic forests were in small fragments that are not perceptible at this scale. C: Sampling design. Threatened tree taxa endemic to the Mascarene Islands (in red) were inventoried in a circular area of 2500 m², invasive alien plants (in black) with a diameter at breast height (DBH) ≥ 1 cm in a concentric quadrat of 100 m² and those with a DBH < 1 cm (including juveniles of invasive alien trees and shrubs) in five quadrats of 1 m² positioned at each corner of the 100 m² quadrat and at its centre. Dark blue trees represent non-threatened native species. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)</figcaption></figure>
IAPs should provide insight into the ecological mechanisms at work (Leffler et al., 2014). In this study, we resurveyed 182 historically described subpopulations of 28 threatened endemic tree (TET) taxa on Reunion Island (South-West Indian Ocean), then tested whether observed trends of juvenile and adult TETs correlate with the abundance in IAPs, while distinguishing TET taxa with respect to their specific leaf area (SLA), maximum plant height and seed mass as well as to their IUCN Red List category of extinction risk. As IAPs tend to be fast-growers and efficient resource preempters, we expected TET taxa at the most conservative end of the resource-use spectrum of ecological strategies to decline more significantly with increasing competition from IAPs. We also hypothesised that juvenile TETs would be more responsive to IAP dominance than adult TETs due to lower life expectancy and exposure to competition from more IAP individuals and species.

## 2. Materials and methods

### 2.1. Study site

Reunion is the largest (2512 km²) and youngest island (2–3 million years old) of the Mascarene Islands, which also comprises Mauritius and Rodrigues (Fig. 1A). The climate of Reunion Island is tropical and characterised by two seasons: a warm and rainy austral summer from November to April followed by a relatively cool and dry winter from May to October (Réchou et al., 2019). Annual rainfall ranges from 500 mm on the leeward side of the island to more than 8000 mm on its windward side. Reunion Island harbours a remarkable diversity of ecosystems zonated along an elevational gradient reaching 3070 m a.s.l. (Piton des Neiges). Due to the high degree of endemism of its native vascular flora (24 % of single-island endemism, 40 % of archipelago endemism) and dramatic loss of native habitats (over two thirds of the island), Reunion Island was included in the biodiversity hotspot ‘Madagascar and the Indian Ocean Islands’ (Myers et al., 2000). In the last few years, the number of native vascular plant species threatened with extinction grew from 30 % in 2010 to 41 % in 2023 (IUCN French Committee et al., 2023).

<figure><img src="image_4.png" /><figcaption>**Fig. 2.** Hypothesised relationships linking abundance in invasive alien plants (IAPs, black circles) with a diameter at breast height (DBH) < 1 cm (i.e., in the understory layer) and with a DBH ≥ 1 cm (i.e., in the upper vegetation layer) to the temporal trend of juvenile and adult threatened endemic tree (TETs, red circles) in path analyses. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)</figcaption></figure>

### 2.2. Selected taxa

We selected 28 threatened woody taxa endemic to the Mascarene Islands, including 10 taxa strictly endemic to Reunion Island, on the basis of historical record availability, reliability and accessibility (Table 1). Extremely rare taxa, e.g. *Ceodes lanceolata* (Nyctaginaceae) and *Ruizia cordata* (Malvaceae) with less than five individuals known in the wild on Reunion Island, those restricted to very specific habitats such as high-elevation summits or with taxonomic uncertainties were excluded. Our sample accounted for over half (53 %) of all threatened shrub and tree taxa endemic to the archipelago and still occurring on Reunion Island (Table S1). Thirteen TET taxa are locally categorised as critically endangered (CR), nine as endangered (EN) and six as vulnerable (VU) (IUCN French Committee et al., 2023). A total of 182 subpopulations were resurveyed in 109 sites at elevations of 29 m to 1367 m (median = 546 m) with annual precipitation between 657 mm and 7768 mm (median = 1899 mm) (Fig. S1). Dry to mesic forests (< 1500 mm) represented 36 % of the sites and 38 % of the subpopulations, mesic to moist forests (1500–3000 mm) represented 43 % of the sites and 45 % of the subpopulations and moist to wet forests (> 3000 mm) represented 21 % of the sites and 17 % of the subpopulations (Fig. 1B). A large majority of those sites were not actively managed. Each TET taxon was present in four to eleven sites (median = six). Each site hosted from one to six TET taxa (median = one).

### 2.3. Resurvey

Historical records were botanical relevés established by various stakeholders (including the Mascarin National Botanical Conservatory, the National Forest Office, the National Park of Reunion and amateur botanists) dating from 1998 to 2018 (median = 2011, see Fig. S2 for further details on the age of the relevés). These data collected without a standard protocol described the numbers of juvenile (in the broad sense, i.e., including juveniles in the strict sense, seedlings and saplings) and adult (i.e., mature) TETs identified by botanists during their census. These inventories were highly localised as TETs were often restricted to small relic subpopulations made of an average of 3.6 adult and 0.9 juvenile trees. Based on the subpopulations that are well known and closely monitored over the long term (mainly of CR species), we evaluate the degree of confidence that can be placed in these data as good to excellent. They serve as the basis for IUCN Red List assessments, conservation master plans and national action plans.

Each site was resurveyed in 2023–2024 (which required 158 person-days) and the current numbers of adult and juvenile TETs were recorded in a circular area of 2500 m² (radius of 28 m) (Fig. 1C). At each site, we estimated the temporal trend of adult and juvenile TETs as the difference between the current and the historical numbers of individuals divided by the number of years since the historical census to account for their different dates. A positive value indicates an increasing subpopulation size, zero indicates a stable subpopulation and a negative value indicates a decreasing subpopulation. To evaluate the compliance of historical and current surveys of TET subpopulations, two expectations were investigated: i) the change in number of individuals increases with the age of the historical relevés; and ii) CR taxa decrease more than EN taxa, which decrease more than VU taxa.

A concentric quadrat of 100 m² was set up and the number of IAPs in the upper vegetation layer, i.e., with a diameter at breast height (DBH) ≥ 1 cm, was counted (Fig. 1C). Five quadrats of 1 m² positioned at each corner of the 100 m² quadrat and in its centre were used to count the number of IAPs with a DBH < 1 cm in the understory layer (including juveniles of woody IAPs). For clonal plants, we enumerated the number of ramets. The abundance of IAPs with a DBH < 1 cm in the five quadrats of 1 m² was extrapolated to 100 m² since the understory layer was in most cases evenly covered by IAPs. The taxonomical referential of the vascular flora of Reunion was used to determine the biogeographical and invasive status of each taxon (Boullet, 2020).
<caption>**Table 1**</caption>

Summary of the 28 threatened tree taxa endemic to the Mascarene Islands considered in this study. Endemism: R = Reunion, M = Mauritius, O = Rodrigues. Threat status: CR = critically endangered, EN = endangered, VU = vulnerable. Subpopulations: number of resurveyed subpopulations. Endemism and threat categories are from IUCN French Committee et al. (2023).

<table><caption>**Table 1 (continued)**</caption><thead><tr><th>Taxon</th><th>Family</th><th>Endemism</th><th>Threat status</th><th>Subpopulations</th></tr></thead><tbody><tr><td>Abutilon exstipulare (Cav.) G. Don</td><td>Malvaceae</td><td>R</td><td>CR</td><td>6</td></tr><tr><td>Coptosperma borbonicum (Hend. et Andr. Hend.) De Block</td><td>Rubiaceae</td><td>R, M</td><td>EN</td><td>8</td></tr><tr><td>Diospyros borbonica I. Richardson</td><td>Ebenaceae</td><td>R</td><td>EN</td><td>8</td></tr><tr><td>Dombeya acutangula Cav.</td><td>Malvaceae</td><td>R, M, O</td><td>VU</td><td>4</td></tr><tr><td>Dombeya blattiolens Frapp. ex Cordem.</td><td>Malvaceae</td><td>R</td><td>VU</td><td>5</td></tr><tr><td>Drypetes caustica (Frapp. ex Cordem.) Airy Shaw</td><td>Putranjivaceae</td><td>R, M</td><td>EN</td><td>5</td></tr><tr><td>Erythroxylum hypericifolium Lam.</td><td>Erythroxylaceae</td><td>R, M</td><td>VU</td><td>7</td></tr><tr><td>Eugenia mespiloides Lam.</td><td>Myrtaceae</td><td>R</td><td>VU</td><td>5</td></tr><tr><td>Fernelia buxifolia Lam.</td><td>Rubiaceae</td><td>R, M, O</td><td>EN</td><td>9</td></tr><tr><td>Foetidia mauritiana Lam.</td><td>Lecythidaceae</td><td>R, M</td><td>CR</td><td>7</td></tr><tr><td>Hernandia mascarensis (Meisn.) Kubitzki</td><td>Hernandiaceae</td><td>R, M</td><td>CR</td><td>5</td></tr><tr><td>Hibiscus columnaris Cav.</td><td>Malvaceae</td><td>R, M</td><td>CR</td><td>6</td></tr><tr><td>Indigofera ammoxylum (DC.) Polhill</td><td>Fabaceae</td><td>R</td><td>CR</td><td>5</td></tr><tr><td>Obetia ficifolia (Poir.) Gaudich.</td><td>Urticaceae</td><td>R, M, O</td><td>EN</td><td>5</td></tr><tr><td>Ochrosia borbonica J.F. Gmel.</td><td>Apocynaceae</td><td>R, M</td><td>VU</td><td>11</td></tr><tr><td>Olax psittacorum (Lam.) Vahl</td><td>Olacaceae</td><td>R, M</td><td>VU</td><td>10</td></tr><tr><td>Polyscias aemiliuineae Bernardi</td><td>Araliaceae</td><td>R</td><td>CR</td><td>4</td></tr><tr><td>Polyscias cutispongia (Lam.) Baker</td><td>Araliaceae</td><td>R</td><td>EN</td><td>5</td></tr><tr><td>Polyscias rivalsii Bernardi</td><td>Araliaceae</td><td>R</td><td>CR</td><td>5</td></tr><tr><td>Poupartia borbonica J.F. Gmel.</td><td>Anacardiaceae</td><td>R, M</td><td>CR</td><td>6</td></tr><tr><td>Pyrostria commersonii J.F. Gmel.</td><td>Rubiaceae</td><td>R</td><td>CR</td><td>7</td></tr><tr><td>Ruizia populnea (Cav.) Dorr</td><td>Malvaceae</td><td>R, M</td><td>CR</td><td>7</td></tr><tr><td>Scolopia heterophylla (Lam.) Sleumer</td><td>Salicaceae</td><td>R, M, O</td><td>EN</td><td>6</td></tr><tr><td>Sideroxylon majus (C.F. Gaertn.) Baehni</td><td>Sapotaceae</td><td>R</td><td>EN</td><td>8</td></tr><tr><td>Tabernaemontana persicariifolia Jacq.</td><td>Apocynaceae</td><td>R, M</td><td>CR</td><td>6</td></tr><tr><td>Terminalia bentzoe (L.) L. f.</td><td>Combretaceae</td><td>R, M, O</td><td>CR</td><td>4</td></tr></tbody></table>

<table><thead><tr><th>Taxon</th><th>Family</th><th>Endemism</th><th>Threat status</th><th>Subpopulations</th></tr></thead><tbody><tr><td>Volkameria heterophylla Vent.</td><td>Lamiaceae</td><td>R, M</td><td>CR</td><td>7</td></tr><tr><td>Zanthoxylum heterophyllum (Lam.) Sm.</td><td>Rutaceae</td><td>R, M, O</td><td>EN</td><td>10</td></tr></tbody></table>

## 2.4. Characteristics of threatened endemic trees

Three functional traits and the IUCN Red List category of extinction risk were tested for their ability to discriminate between TET taxa that responded to IAP abundance and those that did not. The three traits were those of the leaf-height-seed plant ecology strategy scheme (Westoby, 1998): i) the mean SLA, i.e., the amount of light-capturing surface area that is deployed with a given investment of dry mass, which expresses the fundamental acquisition-conservation trade-off; ii) the maximum plant height, which reflects a trade-off for biophysical constraints in determining water fluxes within the plant and which is related to its competitive ability; and iii) the mean seed mass, which reflects variation in dispersal capability and seedling survivorship. These traits were retrieved from the recently established TREFL database (Adrien et al., 2025). Seed mass was lacking for two TET taxa in the current version of the database. Seed mass of *Dombeya blattiolens* (Malvaceae) was considered the same as that of the closely-related *D. ficulnea* (3.2 mg). Seed mass of *Pyrostria commersonii* (Rubiaceae) was considered the same as that of *P. bibracteata* (39.2 mg; Denelle et al., 2023), a congeneric taxon distributed across islands of the South-West Indian Ocean. SLA was also missing for *Hernandia mascarensis* (Hernandiacae) but, as SLA is more dependent on the environmental conditions than the phylogeny (An et al., 2021), we did not consider this species in the SLA-based analyses. Each TET taxon was assigned to three trait-based groups depending on whether its functional trait values were below or above the median values of the 28 TET taxa: low or high SLA, low or high maximum height and low or high seed mass, and to one threat-based group: critically endangered (CR), endangered (EN) or vulnerable (VU). Each TET subpopulation was assigned to two climate-based groups depending on whether local topoclimatic conditions were below or above the median values of the 109 sites: low or high elevation and low or high annual rainfall.

## 2.5. Statistical analyses

The relationship between the trend (annual change in number of individuals) of TET subpopulations and the number of stems of IAPs was quantified using path analyses. A path analysis is a multivariate statistical analysis technique that can disentangle direct and indirect connections among multiple variables in complex networks (Grace, 2006). A major advantage of this approach is that it estimates multiple and interrelated dependencies and allows for explicit tests of relatively complex hypotheses.

Pathways represent hypothesised dependencies between predictor (abundance in IAPs) and response variables (trend of TET subpopulations). The structure of our path analysis included connections from the abundance in IAPs in the upper vegetation layer to the trend of adult and juvenile TETs and from the abundance in IAPs in the understory layer to the trend of juvenile TETs (Fig. S3). We added a connection from the abundance in IAPs in the upper vegetation layer to that in the understory layer because, when from the same species, the former were supposedly the parent plant of the latter and, when not, IAPs in the upper vegetation layer can control the establishment of IAPs in the understory layer (D'Antonio et al., 2017). We also added a connection from the trend of adult TETs to that of juvenile TETs since the subpopulation dynamics of adults most likely determines the probability of new
juvenile establishment.

We built 14 path analyses: one for the full dataset of 182 subpopulations of 28 TET taxa and one for each of the six trait-based categories (two categories for three groups), three threat-based categories (three categories for one group) and four climate-based categories (two categories for two groups). All variables were transformed using ordered quantile normalization (Beasley et al., 2009), which was found to minimize the deviations of our data from normality based on the R package 'bestNormalize' (Peterson, 2021). Ordered quantile normalization is a rank-based procedure in which the values of a vector are mapped to their percentile, which is then mapped to the same percentile of a normal distribution. Variables were then standardised (with a mean of 0 and a standard deviation of 1) to improve comparability of predictor-response associations.

Path analyses were evaluated based on a chi-square statistic used to

test the hypothesis of model-data consistency (P-value). A P-value below 0.05 indicates a significant deviation between observed and modelled values, while a P-value above this threshold indicates concordance. The significance of connections to response variables was also used to select models: only those with at least one significant connection to the trends of TETs were kept so as to address our research question. Slope estimate is expected not to be affected by methodological differences in the estimation of the number of TET individuals (between historical and current surveys) if one method consistently overestimates the other. These analyses were performed with the R package 'lavaan' (Rosseel, 2012).

<figure><img src="image_5.png" /><figcaption>**Fig. 3.** Subpopulation trend (i.e., annual change in the number of individuals) over 6 to 26 years of 28 endemic tree taxa threatened with extinction, including 13 critically endangered (A, D), 10 endangered (B, E) and 6 vulnerable (C, F), on Reunion Island (grey background). The first row (A-C) represents adult trees and the second row (D–F) represents juvenile trees. Circles on the maps show the sign and magnitude of subpopulation trends in the 109 sampled sites. The frequency distribution shows the number of decreasing (negative values), stable (zeros) or increasing subpopulations (positive values). Dashed vertical lines indicate mean values.</figcaption></figure>
# 3. Results

## 3.1. *Trends of threatened endemic trees*

The same number of individuals were enumerated in historical and current surveys in 43 % of adult subpopulations and 65 % of juvenile subpopulations. Overall, the raw (not annualised) change in number of individuals of TETs increased with the time since the historical relevés (estimate = 0.12; *P*-value = 0.03). A third (29 %) of CR subpopulations decreased in number of adults since the last census, while 55 % remained stable and 16 % increased (Fig. 3A). The most significant decline was seen in a subpopulation of 11 adult trees of the island endemic CR tree *Indigofera ammoxylum* (Fabaceae) recorded in 2004. Within two decades, the number dropped to just two adults, averaging a loss of 0.450 individual per year. CR subpopulations whose number of adults decreased the most were mainly distributed on the dry leeward coast of Reunion Island. Juveniles in CR subpopulations also decreased (9 %) more often than increased (7 %) (Fig. 3D). The same tendency was observed for the number of adults in EN subpopulations (30 % and 27 %, respectively) but the opposite was the case for juveniles (13 % and 22 %, respectively) (Fig. 3B, E). In addition, there were less decreasing than increasing VU subpopulations both for adults (16 % and 63 %, respectively) and for juveniles (19 % and 49 %, respectively) (Fig. 3C, F). As a result, CR subpopulations overall decreased (−0.013 adult individual per year and − 0.053 juvenile individual per year) while EN (0.027 and 0.053, respectively) and VU subpopulations increased (0.236 and 0.216, respectively), the latter more rapidly than the former.

## 3.2. *Abundance in invasive alien plants*

A total of 75 IAP species were recorded in the understory layer of the 109 sites and 39 species were recorded in the upper vegetation layer, with 27 species in common by both layers. The number of stems of IAPs per 100 m² ranged from 20 to 6880 with a median of 860 in the understory layer and from 0 to 1489 with a median of 21 in the upper vegetation layer. The occupancy of IAP species ranged from 1 site for 35 species to 75 sites for the small tree *Litsea glutinosa* (Lauraceae) followed by the small tree *P. cattleyanum* (53) and the shrub *Ardisia crenata* (Primulaceae) (43) in the understory layer, and from 1 site for 10 species to 58 sites, again, for *L. glutinosa* followed by *P. cattleyanum* (52) and the small tree *Leucaena leucocephala* (Fabaceae) (38) in the upper vegetation layer. The most abundant IAP species in the understory layer were *A. crenata* (272 stems per 100 m² on average) and juveniles of *L. glutinosa* (264 stems) and *P. cattleyanum* (218 stems) (Table S2). The most abundant IAP species in the upper vegetation layer were *P. cattleyanum* (28 stems per 100 m² on average), *L. leucocephala* (16 stems), and the shrub *Ligustrum robustum* (Oleaceae) (5 stems, present in 8 sites).

## 3.3. *Links between trends of threatened endemic trees and abundance in invasive alien plants*

In eight out of the 14 path analyses, for two trait-based categories (high SLA and high maximum height), all three threat-based categories (CR, EN, VU) and three climate-based categories (high rainfall, low elevation and high elevation), IAP abundance did not significantly explain trends of TET subpopulations (Table 2). However, the path analyses built with the full dataset, for subpopulations of TETs with low SLA, low maximum height, both low and high seed mass, and found in low rainfall sites revealed significant associations between abundance in IAPs and trends of TETs (Fig. 4). In such models, the abundance in IAPs in the understory layer always influenced negatively the trend of juvenile TETs, with the exception of those with a high seed mass. The trend of juvenile TETs with high seed mass was solely determined (positively) by that of adult TETs (Figs. 4E and S6D), and this connection was also strong for low-SLA TET taxa (Figs. 4B and S6B). The abundance in IAPs

<caption>**Table 2**</caption>

Results of the 14 path analyses built to explore relationships between the temporal trends of subpopulations of threatened tree taxa endemic to the Mascarene Islands and the abundance in invasive alien plants. Each of the 28 threatened endemic tree taxa was assigned to three trait-based groups: low or high specific leaf area (SLA), low or high maximum height and low or high seed mass, and to one threat-based group: critically endangered (CR), endangered (EN) or vulnerable (VU). Each of the 182 subpopulations of threatened endemic trees was assigned to two climate-based groups: low or high elevation and low or high annual rainfall. *P*-values are based on a chi-square statistic used to test for model-data consistency. They were not adjusted to control for multiple testing to be less conservative. Models in bold have at least one significant connection to the trend of subpopulations of threatened endemic trees.

<table><thead><tr><th>Groups</th><th>Path analysis</th><th>Sample</th><th>n</th><th>P-value</th><th>Signif. path to threatened taxa</th></tr></thead><tbody><tr><td rowspan="2">Trait-based</td><td>Full model</td><td>Full dataset</td><td>182</td><td>0.391</td><td>Yes</td></tr><tr><td>Low SLA</td><td>&lt; 110.9 cm2/g</td><td>98</td><td>0.946</td><td>Yes</td></tr><tr><td rowspan="2">Trait-based</td><td>High SLA</td><td>≥ 110.9 cm2/g</td><td>79</td><td>0.404</td><td>No</td></tr><tr><td>Low maximum height</td><td>≤ 10 m</td><td>104</td><td>0.468</td><td>Yes</td></tr><tr><td rowspan="4">Threat-based</td><td>High maximum height</td><td>&gt; 10 m</td><td>78</td><td>0.960</td><td>No</td></tr><tr><td>Low seed mass</td><td>≤ 0.042 g</td><td>82</td><td>0.121</td><td>Yes</td></tr><tr><td>High seed mass</td><td>&gt; 0.042 g</td><td>100</td><td>0.479</td><td>Yes</td></tr><tr><td>CR</td><td>–</td><td>75</td><td>0.997</td><td>No</td></tr><tr><td rowspan="3">Threat-based</td><td>EN</td><td>–</td><td>64</td><td>0.569</td><td>No</td></tr><tr><td>VU</td><td>–</td><td>43</td><td>0.784</td><td>No</td></tr><tr><td>Low rainfall</td><td>≤ 1673 mm</td><td>92</td><td>0.552</td><td>Yes</td></tr><tr><td rowspan="3">Threat-based</td><td>High rainfall</td><td>&gt; 1673 mm</td><td>90</td><td>0.883</td><td>No</td></tr><tr><td>Low elevation</td><td>&lt; 546 m</td><td>93</td><td>0.435</td><td>No</td></tr><tr><td>High elevation</td><td>≥ 546 m</td><td>89</td><td>0.947</td><td>No</td></tr></tbody></table>

in the understory layer was positively associated with the abundance in IAPs in the upper vegetation layer for low-SLA TET taxa, while this relationship was negative in low rainfall sites (Fig. 4F). There, the trend of juvenile TETs was also negatively correlated with the abundance in IAPs in the upper vegetation layer (Figs. 4F and S6F). In contrast, the trend of adult TETs never covaried with the abundance in IAPs.

# 4. Discussion

We examined whether the temporal trend of juvenile and adult individuals in 182 subpopulations of 28 TETs was linked to the abundance in IAPs in the understory and in the upper vegetation layers. The most consistent relationship was a negative association between the trend of juvenile TETs and the abundance in IAPs in the understory layer (Figs. 4 and S6). This was particularly the case for TETs with resource-conservative traits but not for large-seeded trees. The trend of juvenile TETs was also negatively influenced by the abundance in IAPs in the upper vegetation layer but only in the driest sites. Adults of all TETs and juveniles of the most resource-acquisitive species and those occurring in mesic to wet forests were less affected by the abundance in IAPs.

## 4.1. *Impact of understory invasions on the regeneration of threatened endemic trees*

Compared to native species, IAPs generally have higher leaf-area allocation and size (van Kleunen et al., 2010). Assuming that trait differences between TETs and IAPs can reveal impact mechanisms (Leffler et al., 2014), our finding that TETs with slow resource acquisition and low investment in height declined in proportion to the abundance of IAPs might reflect the ability of the latter to outcompete or overtop the former. In the same vein, a global analysis of the Red List extinction data
<figure><img src="image_6.png" /><figcaption>**Fig. 4.** Path analyses for direct and indirect effects of the abundance of invasive alien plants (IAPs, black circles) with a diameter at breast height (DBH) < 1 cm (i.e., in the understory layer) and with a DBH ≥ 1 cm (i.e., in the upper vegetation layer) on the temporal trend of juvenile and adult threatened endemic trees (TETs, red circles) on Reunion Island (see path diagram in Fig. 2). A: model for the full dataset of 182 subpopulations. B-E: models for a subset of TET taxa below or above the median value of three functional traits (specific leaf area, maximum height, seed mass). F: model for a subset of TET subpopulations below the median value of annual rainfall. Path-coefficient estimates provide a measure of the importance of the relationship (the larger the coefficient, the stronger the relationship). The thickness of each line is proportional to the magnitude of the coefficient. A blue line indicates a positive effect and a black line a negative effect. Only significant effects are shown. Path analyses without significant connection to the trend of juvenile or adult TETs are displayed in Figs. S3-S5. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)</figcaption></figure>

indicated that plant extinctions mostly affected island endemic tree species, and that those with a lower competitive ability were more prone to extinction, presumably because of a greater susceptibility to IAPs (Gray, 2019).

The fact that only juvenile (and not adult) TETs taxa were affected indicates that competition, if any, mostly occurs at early stages of tree

development (but see Bissessur et al., 2023). In Hawaiian seasonally dry tropical forests, the invasion by *Pennisetum setaceum* (Poaceae) was found to result in an absence of saplings while seedlings were able to germinate under the grass canopy, suggesting high mortality between the seedling and the sapling stages due to competition with the perennial grass (Litton et al., 2006). Another study in a Costa Rican abandoned
pasture demonstrated that invasive grasses strongly impacted the survival and development of tree seedlings through below-ground competition for water and nutrients, shading, and allelopathic chemicals (Holl, 2002).

In a systematic literature review on the impact of IAPs on the regeneration of native trees in European temperate forests, competition for resources was found to be by far the most commonly reported impact mechanism, particularly competition for light (Langmaier and Lapin, 2020). Our results indicate that competition for nutrients could be another major impact mechanism affecting juvenile TETs on Reunion Island, as already shown elsewhere (Moles and Westoby, 2004). First, we found that large-seeded trees with a greater quantity of metabolic reserves in their seed that make seedlings more likely to attain a critical size had their juvenile trend solely determined by their adult trend, while small-seeded trees were affected by IAPs. This is in line with recent results showing that seedlings and saplings of large-seeded trees can survive competition in a highly invaded wet rainforest of Reunion Island (Albert et al., 2023). Second, the decline of juvenile TETs was related to the number of IAPs in the understory layer, whose root system forages the topsoil as that of young trees, but not to the number of IAP in the upper vegetation layer, which prospect deeper soil horizons.

In the same literature review, the release of allelopathic compounds impacting the regeneration of tree seedlings was rarely reported among examined studies (Langmaier and Lapin, 2020). In our study, we found that the understory and the upper vegetation layers share many of the most abundant IAP species (Table S2). Furthermore, among the few most abundant IAPs that are not juveniles of the most abundant IAPs in the upper vegetation layer, only one species, *Ageratina riparia* (Asteraceae), has been reported to have an allelopathic effect on other plants (Tripathi et al., 2006). As a result, the different impact of IAPs in the understory and in the upper layers supports the view that allelopathy is not a key mechanism in our case.

The only exception in which juvenile TETs were affected as much by IAPs in the understory layer as by IAPs in the upper vegetation layer was in dry to mesic forests. In these forests, IAPs in the upper vegetation layer could compete with juvenile TETs but also with IAPs in the understory layer for the supposedly low amounts of nutrients in the shallow soil. Since IAPs in the understory layer and in the upper vegetation layer have similar impact in the driest sites, it cannot be ruled out that allelopathy, e.g., of *Lantana camara* (Verbenaceae) (Gentle and Duggin, 1997; Gooden et al., 2009) and *Leucaena leucocephala* (Xuan et al., 2006), might also be a common mechanism contributing to the decline of juvenile TETs in dry to mesic forests.

The lack of correlation between the abundance in IAPs and the trend of some functional groups of TETs may suggest that they would be able to coexist with IAPs, at least over the time scale of this study (6–26 years). Traits associated with resource-acquisitive strategies (e.g., high SLA and maximum height) or a distribution in resource-rich environments (e.g., mesic to wet forests) could confer some resistance against biological invasions to TETs, by making resource competition with IAPs less asymmetric or less intense (Kraft et al., 2015). Another explanation could be that, in the most humid forests, where trees typically bear higher-SLA leaves and attain taller mature stature, the abundance in IAPs tend to be lower than under drier conditions (Strasberg et al., 2005), and might stay below the threshold beyond which they start to affect regeneration in a number of sites. Our result that there was no association between the abundance in IAPs and the trend of TETs for all threat and elevation groups may simply reflect the poor explanatory power of these variables, plus a lower statistical power due to relatively small sample sets compared to the full SEM.

## 4.2. Management implications

Uncovering the underlying threats and processes behind species extinctions is a prerequisite to identify species, habitats and actions towards which limited financial resources should be allocated to halt or

mitigate extinction dynamics (Lee and Jetz, 2010). We found that a prime mechanism at play on Reunion Island could be IAPs inhibiting TET establishment and juvenile survival through competition in the understory layer, particularly in dry to mesic forests. This result suggests that the extinction debt caused by IAPs could be considerable given the long lifespan of trees (Gilbert and Levine, 2013).

Our study thus confirms that the very common practice of controlling IAPs of the understory layer should at least partly help restore the ecological succession (Weidlich et al., 2020). An experiment combining direct seeding and removal of IAPs on Reunion Island revealed an increase in the emergence, survival and growth rates of endemic tree seedlings in weeded plots compared to non-weeded controls (Albert et al., 2024). Removal of invasive grasses in a seasonally dry submontane forest on Hawaii also resulted in seedlings of woody taxa being significantly more abundant in weeded plots and in a higher recruitment into the sapling stage (D'Antonio et al., 1998). Many other IAP removal experiments highlighted that the survival of native tree seedlings can more than double in intensively managed plots compared to untreated plots, thereby implying that invasion was inhibiting tree recruitment (Campoe et al., 2010; Flory and Clay, 2009; Truong et al., 2021; Urgenson et al., 2012). Our approach allowed us to draw an identikit picture of the TET taxa that should be seen as a management priority (low SLA, low height at maturity, low seed mass, distributed in dry to mesic forests) since their regeneration dynamics appeared to be closely related to understory invasions. We also identified the IAPs that invade the forest understory the most on Reunion Island (Table S2).

## 4.3. Study limitations

Unsurprisingly, the trend of CR taxa was overall lower than that of EN taxa, which in turn was lower than the trend of VU taxa (Fig. 3). However, the average trends of EN and VU taxa were both found to be positive, which appears contradictory for threatened taxa. Resurvey studies using historical information usually overestimate extinction probabilities due to the imperfect detectability of plants (Martínez-Villegas et al., 2024), and because newly colonized sites are rarely considered (Holderegger and Spillmann, 2022). As we focused on large woody plants and given their limited regeneration, our results might however be less substantially affected than those of other similar studies on herbaceous plants or on plants that regenerate more efficiently (Hoarau et al., 2025). On the contrary, we assume that our estimates of subpopulation trends were overestimated, and the most likely explanation is the differences in sampling protocols between historical and current surveys. Indeed, historical data mainly came from opportunistic observations while current data were based on comprehensive inventories on a standardised surface area. Despite the great ecological value of historical data, it is therefore possible that some individuals of the least discernible taxa or in the largest subpopulations were missed, particularly seedlings and saplings. Nevertheless, this overestimation should not have significantly affected our results insofar as we searched for correlations between trends of threatened trees and abundance in IAPs. Our finding that a range of correlations were significant for some functional or climatic groups and not for others confirms that significant correlations were not an artefact of different survey methodologies, which otherwise would be evenly distributed across all the groups. On a longer term, our standardised sampling protocol could serve to monitor more accurately the focal taxa and measure the effect of restoration actions.

In this study, we used a snapshot of IAP abundance as an explanatory variable for the long-term trend of TETs, so we did not consider how this abundance changed over time in the absence of information on our study sites, although the limited available data suggest a considerable increase in alien species abundance (Barbé et al., 2015). Yet, in the likely event that there is a time lag between plant invasions and their effects on the extinction dynamics of trees (Gilbert and Levine, 2013), we may have missed a closer negative association. Moreover, we did not
distinguish between the different IAP species because there were too many. We recognize that this provides a simplistic view of the great diversity of interactions that IAPs can have with native plants or among themselves (Enders et al., 2020). The proposed approach could be used in further analyses based on the abundance of each individual IAP species or of functional groups in order to investigate the role of IAP identity and refine our understanding of the underlying ecological mechanisms. Furthermore, we measured the abundance in IAPs as the number of individuals found per plot to ensure reproducibility for future monitoring but plant cover estimates could have also been useful to account for space occupation, which sometimes differed a lot from one species to another, e.g., *Furcraea foetida* (Asparagaceae) can form a leaf rosette up to 4 m in diameter. Finally, relating the abundance in IAPs to the abundance in native and endemic species, i.e., considering how IAPs contribute to forest communities, could also provide additional insight into the impact of IAPs and reduce the potential bias arising from abundance reflecting local forest productivity.

## 5. Conclusion

Our study presented preliminary evidence that IAPs can contribute to the extinction dynamics of woody species, primarily by limiting regeneration, and that this impact varies depending on the traits of the affected taxa. The functional groups concerned offer a clearer perspective as to the potential mechanisms that underpin the decline of island endemic trees in response to biological invasions. This decline might be primarily attributable to juvenile trees being outcompeted by IAPs found in the forest understory. Exposure to competition from IAPs over multiple generations can therefore lead to widespread extinctions if actions commensurate with the scale of the challenge are not implemented. Although this study reflects features associated with Reunion Island, our results could have ramifications for other oceanic islands, where IAPs are able to invade areas with limited disturbance due to intrinsic peculiarities of native plant communities, which are relatively impoverished, unsaturated and taxonomically disharmonic with a disproportionately high number of endemic species that evolved under stable conditions. The impact of IAPs may however differ on continents where IAPs appear to be more closely associated with anthropogenic transformation of the environment.

### CRediT authorship contribution statement

**Robin Pouteau:** Writing – original draft, Project administration, Methodology, Investigation, Funding acquisition, Formal analysis, Data curation, Conceptualization. **Élise Adrien:** Data curation. **Claudine Ah-Peng:** Writing – review & editing. **Sébastien Albert:** Writing – review & editing, Data curation. **Olivier Flores:** Writing – review & editing, Project administration, Funding acquisition, Data curation. **Bastian Laforgue:** Data curation. **Christophe Lavergne:** Data curation. **Jean-Yves Meyer:** Writing – review & editing, Data curation. **Anaëlle Regen:** Data curation. **Margaux Rojat:** Data curation. **Sarah Roussel:** Project administration, Funding acquisition, Data curation.

### Declaration of competing interest

The authors have no relevant financial or non-financial interests to disclose.

### Acknowledgments

We thank the colleagues, students and volunteers who gave their time and effort to this study. We acknowledge financial support from the French National Research Agency (ANR) through the EDENE (Extinction dynamics of endemic trees in exotic-dominated ecosystems) project (ANR-22-CE32-0006), the French Office for Biodiversity (OFB) through the FRAG'ILE (Fragmentation en milieu insulaire) project, and the Fonds

Vert through the CEODES (Conservation, étude et observatoire des espèces à sauvegarder) project. We also acknowledge the Reunion National Park for providing research authorisations and collecting permits (arrêtés DIR-I-2023-030 and DIR-I-2023-398, and avis conforme 2021-030). We appreciate the helpful comments and suggestions from three anonymous referees.

## Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.biocon.2025.111665.

### Data availability

The data supporting the findings of this study are available at doi:10.23708/CDFYZ7.

### References

Adrien, E., Heymans, L., Ah-Peng, C., Albert, S., Cuenin, N., Danger, C., Flores, O., Grondin, R.-P., Hoarau, H., Parlevliet, H., Regen, A., Rivière, J.-N., Rojat, M., Pouteau, R., 2025. TREFL Database: Traits of the Reunion Flora. https://doi.org/10.23708/IIEMOE.

Albert, S., Flores, O., Rhumeur, A., Gebus, M., Ah-Peng, C., Gorissen, A., Stamenoff, P., Payet, M., Strasberg, D., 2023. Associer semis directs et lutte contre les invasions pour diversifier la régénération de la forêt tropicale humide à La Réunion. Université de La Réunion accessed 8 December 2025. https://hal.science/hal-04294200.

Albert, S., Franc, C., Solesse, R., Strasberg, D., Flores, O., 2024. Defaunated and invaded insular tropical rainforests will not recover alone: Recruitment limitation factors disentangled by hierarchical models of spontaneous and assisted regeneration. J. Ecol. 112, 265–277. https://doi.org/10.1111/1365-2745.14230.

An, N., Lu, N., Fu, B., Wang, M., He, N., 2021. Distinct Responses of Leaf Traits to Environment and Phylogeny Between Herbaceous and Woody Angiosperm Species in China. Front. Plant Sci. 12. https://doi.org/10.3389/fpls.2021.799401.

Baider, C., Florens, F.B.V., 2011. Control of invasive alien weeds averts imminent plant extinction. Biol. Invasions 13, 2641–2646. https://doi.org/10.1007/s10530-011-9980-3.

Barbé, M., Fenton, N.J., Lavergne, C., Le Péchon, T., Baider, C., Gigord, L.D.B., 2015. Changes in lowland dry-forest native and alien plant communities on Réunion Island (Indian Ocean) over 16 years. Botany 93, 843–857. https://doi.org/10.1139/cjb-2015-0112.

Beasley, T.M., Erickson, S., Allison, D.B., 2009. Rank-based inverse normal transformations are increasingly used, but are they merited? Behav. Genet. 39, 580. https://doi.org/10.1007/s10519-009-9281-0.

Bellard, C., Cassey, P., Blackburn, T.M., 2016. Alien species as a driver of recent extinctions. Biol. Lett. 12, 20150623. https://doi.org/10.1098/rsbl.2015.0623.

Bissessur, P., Dennis Reinegger, R., Baider, C., Mamoodee, R., Benjamin Vincent Florens, F., 2023. Invasive alien plant control: the priority to save one of the most rapidly declining island-endemic plant species worldwide. J. Nat. Conserv. 73, 126417. https://doi.org/10.1016/j.jnc.2023.126417.

Borges, P.A., Cardoso, P., Kreft, H., Whittaker, R.J., Fattorini, S., Emerson, B.C., Gil, A., Gillespie, R.G., Matthews, T.J., Santos, A.M., 2018. Global Island Monitoring Scheme (GIMS): a proposal for the long-term coordinated survey and monitoring of native island forest biota. Biodivers. Conserv. 27, 2567–2586.

Boulet, V., 2020. Index de la flore vasculaire de la Réunion (Trachéophytes): statuts, menaces et protections, Conservatoire botanique national de Mascarin. Ed. Saint-Leu (Réunion).

Campoe, O.C., Stape, J.L., Mendes, J.C.T., 2010. Can intensive management accelerate the restoration of Brazil's Atlantic forests? In: Forest Ecology and Management, Productivity in Tropical Plantations, 259, pp. 1808–1814. https://doi.org/10.1016/j.foreco.2009.06.026.

Cousins, S.A.O., Vanhoenacker, D., 2011. Detection of extinction debt depends on scale and specialisation. Biol. Conserv. 144, 782–787. https://doi.org/10.1016/j.biocon.2010.11.009.

D'Antonio, C.M., Hughes, R.F., Mack, M., Hitchcock, D., Vitousek, P.M., 1998. The response of native species to removal of invasive exotic grasses in a seasonally dry Hawaiian woodland. J. Veg. Sci. 9, 699–712. https://doi.org/10.2307/3237288.

D'Antonio, C.M., Ostertag, R., Cordell, S., Yelenik, S., 2017. Interactions among invasive plants: lessons from Hawai'i. Annu. Rev. Ecol. Evol. Syst. 48, 521–541. https://doi.org/10.1146/annurev-ecolsys-110316-022620.

Denelle, P., Weigelt, P., Kreft, H., 2023. GIFT—An R package to access the Global Inventory of Floras and Traits. Methods Ecol. Evol. 14, 2738–2748. https://doi.org/10.1111/2041-210X.14213.

Denslow, J.S., 2003. Weeds in paradise: Thoughts on the invasibility of tropical islands. Ann. Mo. Bot. Gard. 90, 119–127. https://doi.org/10.2307/3298531.

Downey, P.O., Richardson, D.M., 2016. Alien plant invasions and native plant extinctions: a six-threshold framework. AoB PLANTS 8. https://doi.org/10.1093/aobpla/plw047.

Enders, M., Havemann, F., Ruland, F., Bernard-Verdier, M., Catford, J.A., Gómez-Aparicio, L., Haider, S., Heger, T., Kueffer, C., Kühn, I., Meyerson, L.A., Musseau, C.,
Novoa, A., Ricciardi, A., Sagouis, A., Schittko, C., Strayer, D.L., Vilà, M., Essl, F., Hulme, P.E., van Kleunen, M., Kumschick, S., Lockwood, J.L., Mabey, A.L., McGeoch, M.A., Palma, E., Pysek, P., Saul, W.-C., Yannelli, F.A., Jeschke, J.M., 2020. A conceptual map of invasion biology: Integrating hypotheses into a consensus network. Glob. Ecol. Biogeogr. 29, 978-991. https://doi.org/10.1111/geb.13082.

Fernández-Palacios, J.M., Kreft, H., Irl, S.D.H., Norder, S., Ah-Peng, C., Borges, P.A.V., Burns, K.C., de Nascimento, L., Meyer, J.-Y., Montes, E., Drake, D.R., 2021. Scientists' warning - the outstanding biodiversity of islands is in peril. Global Ecol. Conserv. 31, e01847. https://doi.org/10.1016/j.gecco.2021.e01847.

Flory, S.L., Clay, K., 2009. Invasive plant removal method determines native plant community responses. J. Appl. Ecol. 46, 434-442. https://doi.org/10.1111/j.1365-2664.2009.01610.x.

Gentle, C.B., Duggin, J.A., 1997. Allelopathy as a competitive strategy in persistent thickets of *Lantana camara* L. in three Australian forest communities. Plant Ecol. 132, 85-95. https://doi.org/10.1023/A:1009707404802.

Gilbert, B., Levine, J.M., 2013. Plant invasions and extinction debts. PNAS 110, 1744-1749. https://doi.org/10.1073/pnas.1212375110.

Gooden, B., French, K., Turner, P.J., Downey, P.O., 2009. Impact threshold for an alien plant invader, *Lantana camara* L., on native plant communities. Biol. Conserv. 142, 2631-2641. https://doi.org/10.1016/j.biocon.2009.06.012.

Grace, J.B., 2006. Structural Equation Modeling and Natural Systems. Cambridge University Press.

Gray, A., 2019. The ecology of plant extinction: rates, traits and island comparisons. Oryx 53, 424-428. https://doi.org/10.1017/S0030605318000315.

Gurevitch, J., Padilla, D.K., 2004. Are invasive species a major cause of extinctions? Trends Ecol. Evol. 19, 470-474. https://doi.org/10.1016/j.tree.2004.07.005.

Hoarau, C., Rhumeur, A., Marie, L., Mallet, B., Triolo, J., Flores, O., Strasberg, D., Pouteau, R., 2025. A resurvey of threatened endemic tree populations in seasonally dry tropical forests of Réunion Island (Indian Ocean) suggests imminent extinctions. J. Nat. Conserv. 86, 126899. https://doi.org/10.1016/j.jnc.2025.126899.

Holderegger, R., Spillmann, J., 2022. Do revisitation studies overestimate local extinction? Conserv. Sci. Pract. 4, e12797. https://doi.org/10.1111/csp2.12797.

Holl, K.D., 2002. Effect of shrubs on tree seedling establishment in an abandoned tropical pasture. J. Ecol. 90, 179-187. https://doi.org/10.1046/j.0022-0477.2001.00637.x.

IUCN French Committee, OFB, MNHN, CBN-CPIE Mascarin, 2023. La Liste rouge des espèces menacées en France - Chapitre Flore vasculaire de La Réunion. Paris (France).

van Kleunen, M., Weber, E., Fischer, M., 2010. A meta-analysis of trait differences between invasive and non-invasive plant species. Ecol. Lett. 13, 235-245. https://doi.org/10.1111/j.1461-0248.2009.01418.x.

Kraft, N.J.B., Godoy, O., Levine, J.M., 2015. Plant functional traits and the multidimensional nature of species coexistence. Proc. Natl. Acad. Sci. 112, 797-802. https://doi.org/10.1073/pnas.1413650112.

Langmaier, M., Lapin, K., 2020. A systematic review of the impact of invasive alien plants on forest regeneration in European temperate forests. Front. Plant Sci. 11. https://doi.org/10.3389/fpls.2020.524969.

Lee, T.M., Jetz, W., 2010. Unravelling the structure of species extinction risk for predictive conservation science. Proc. R. Soc. B Biol. Sci. 278, 1329-1338. https://doi.org/10.1098/rspb.2010.1877.

Leffler, A.J., James, J.J., Monaco, T.A., Sheley, R.L., 2014. A new perspective on trait differences between native and invasive exotic plants. Ecology 95, 298-305. https://doi.org/10.1890/13-0102.1.

Litton, C.M., Sandquist, D.R., Cordell, S., 2006. Effects of non-native grass invasion on aboveground carbon pools and tree population structure in a tropical dry forest of Hawaii. For. Ecol. Manag. 231, 105-113. https://doi.org/10.1016/j.foreco.2006.05.008.

Martínez-Villegas, J.A., Pisanty, I., Martorell, C., Hernández-Apolinar, M., Valverde, T., Granados-Hernández, L.A., Rodríguez-Sánchez, M., Zúñiga-Vega, J.J., 2024. Importance of accounting for imperfect detection of plants in the estimation of population growth rates. Oikos 2024, e10708. https://doi.org/10.1111/oik.10708.

Meyer, J.-Y., Fabre Barroso, S., Mazoyer, L., Jordan, M., Fraisse, J., Fourdrigniez, M., Taputuarai, R., Pouteau, R., 2025. Vulnerability and resilience of rainforests to plant

invasions in a tropical island of the South Pacific: a sixteen-year survey. Biodivers. Conserv. 34, 3985-3999. https://doi.org/10.1007/s10531-025-03143-5.

Moles, A.T., Westoby, M., 2004. What do seedlings die from and what are the implications for evolution of seed size? Oikos 106, 193-199. https://doi.org/10.1111/j.0030-1299.2004.13101.x.

Myers, N., Mittermeier, R.A., Mittermeier, C.G., da Fonseca, G.A.B., Kent, J., 2000. Biodiversity hotspots for conservation priorities. Nature 403, 853-858. https://doi.org/10.1038/35002501.

Orihuela-Rivero, R., Morente-López, J., Reyes-Betancort, J.A., Schaefer, H., Valido, A., Menezes de Sequeira, M., Romeiras, M.M., Góis-Marques, C.A., Salas-Pascual, M., Vanderpoorten, A., Fernández-Palacios, J.M., Patiño, J., 2025. Geographic and biological drivers shape geographic extinctions in the Macaronesian vascular flora. Glob. Chang. Biol. 31, e70072. https://doi.org/10.1111/gcb.70072.

Peterson, R.A., 2021. bestNormalize: Normalizing Transformation Functions version 1.8.2. https://cran.r-project.org/web/packages/bestNormalize/index.html accessed 11 May 2022.

Pouteau, R., van Kleunen, M., Strasberg, D., 2023. Closely related aliens lead to greater extinction risk. Biol. Conserv. 284, 110148. https://doi.org/10.1016/j.biocon.2023.110148.

Réchou, A., Flores, O., Jumaux, G., Duflot, V., Bousquet, O., Pouppeville, C., Bonnardot, F., 2019. Spatio-temporal variability of rainfall in a high tropical island: patterns and large-scale drivers in Réunion Island. Q. J. R. Meteorol. Soc. 145, 893-909. https://doi.org/10.1002/qj.3485.

Rosseel, Y., 2012. lavaan: An R package for structural equation modeling. J. Stat. Softw. 48, 1-36. https://doi.org/10.18637/jss.v048.i02.

Roy, H.E., Pauchard, A., Stoett, P., Renard Truong, T., Bacher, S., Galil, B.S., Hulme, P., Ikeda, T., Sankaran, K.V., McGeoch, M., 2023. In: IPBES (Ed.), Summary for Policymakers of the Thematic Assessment Report on Invasive Alien Species and Their Control of the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services.

Sax, D.F., Gaines, S.D., 2008. Species invasions and extinction: the future of native biodiversity on islands. PNAS 105, 11490-11497. https://doi.org/10.1073/pnas.0802290105.

Sax, D.F., Gaines, S.D., Brown, J.H., 2002. Species invasions exceed extinctions on islands worldwide: a comparative study of plants and birds. Am. Nat. 160, 766-783. https://doi.org/10.1086/343877.

Trasberg, D., Rouget, M., Richardson, D.M., Baret, S., Dupont, J., Cowling, R.M., 2005. An assessment of habitat diversity and transformation on La Réunion Island (Mascarene Islands, Indian Ocean) as a basis for identifying broad-scale conservation priorities. Biodivers. Conserv. 14, 3015-3032. https://doi.org/10.1007/s10531-004-0258-2.

Tripathi, R.S., Kushwaha, S.P.S., Yadav, A.S., 2006. Ecology of three invasive species of Eupatorium: a review. Int. J. Ecol. Environ. Sci. 32, 301-306.

Truong, T.T.A., Andrew, M.E., Hardy, G.E.St.J., Pham, T.Q., Nguyen, Q.H., Dell, B., 2021. Impact of a native invasive weed (*Microstegium ciliatum*) on regeneration of a tropical forest. Plant Ecol. 222, 173-191. https://doi.org/10.1007/s11258-020-01097-y.

Urgenson, L.S., Reichard, S.H., Halpern, C.B., 2012. Multiple competitive mechanisms underlie the effects of a strong invader on early- to late-seral tree seedlings. J. Ecol. 100, 1204-1215. https://doi.org/10.1111/j.1365-2745.2012.01995.x.

Watson, R., Baste, I., Larigauderie, A., Leadley, P., Pascual, U., Baptiste, B., Demissew, S., Dzibá, L., Erpul, G., Fazel, A., 2019. Summary for Policymakers of the Global Assessment Report on Biodiversity and Ecosystem Services of the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services. Bonn, Germany, IPBES Secretariat, pp. 22-47.

Weidlich, E.W.A., Flórido, F.G., Sorrini, T.B., Brancalion, P.H.S., 2020. Controlling invasive plant species in ecological restoration: a global review. J. Appl. Ecol. 57, 1806-1817. https://doi.org/10.1111/1365-2664.13656.

Westoby, M., 1998. A leaf-height-seed (LHS) plant ecology strategy scheme. Plant Soil 199, 213-227. https://doi.org/10.1023/A:1004327224729.

Xuan, T.D., Elzaawely, A.A., Deba, F., Fukuta, M., Tawata, S., 2006. Mimosine in *Leucaena* as a potent bio-herbicide. Agron. Sustain. Dev. 26, 89-97.