# Fifty years of regional-scale vegetation change on Atlantic heathlands

M. Dano, V. Jung, G. Thiebaut, S. Chollet

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## To cite this version:

M. Dano, V. Jung, G. Thiebaut, S. Chollet. Fifty years of regional-scale vegetation change on Atlantic heathlands. Biological Conservation, 2025, 308, pp.111210. 10.1016/j.biocon.2025.111210. hal-05105199

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# Fifty years of regional-scale vegetation change on Atlantic heathlands

M. Dano*, V. Jung, G. Thiébaut, S. Chollet

UMR 6553 ECOBIO, University of Rennes, France

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## ARTICLE INFO

**Keywords:**

Afforestation

Biodiversity conservation

Ellenberg indicator values

Habitat conservation status

Plant community

Temporal changes

Vegetation resurvey

## ABSTRACT

Heathlands are semi-natural habitats of high conservation interest in Europe. The Armorican Massif (north-west of France) is a prime area for heathlands conservation, although they have declined sharply during the last century. In order to understand the extent of the changes in heathland plant communities over the last 50 years in this region, we carried out a resurvey of 444 vegetation plots historically sampled between 1947 and 1992. After identifying the qualitative changes in habitat types between the historical and current periods, we analysed the response of various plant community parameters reflecting species diversity, the presence or abundance of particular species (typical heathland species, insect-pollinated species, rare species), and Ellenberg indicators of ecological conditions. We also determined whether the responses of these parameters depended on the location in coastal or inland areas, or in protected or unprotected areas. Almost half (46 %) of the most typical historic heathlands (i.e. open ericaceous vegetations) is now dominated by *Molinia* and bracken or has evolved into wooded vegetations. Responses for all community parameters indicate a heathland habitat degradation. For species richness and insect-pollinated plant species richness, this decline was more pronounced for inland than coastal heathlands. Also, our results show that the designation of sites as protected areas has not prevented heathland degradation. By identifying the changes that have taken place over the last 50 years at regional level, our results highlight the importance of conserving the last remaining heathlands by applying appropriate management methods.

## 1. Introduction

Heathlands are among the oldest agropastoral landscapes in Europe, shaped by millennia of human activities (Ellis, 2021; Fagúndez, 2013; Loidi et al., 2010; De Graaf et al., 2009; Webb, 1998). These ecosystems developed during the forest clearance of the Neolithic period and expanded until mid-19th century (Fagúndez, 2013; Loidi et al., 2010; Webb, 1998). Historically, these semi-natural habitats have been maintained by traditional agropastoral practices, such as extensive grazing, burning, mowing and sod-cutting (Webb, 1998; Bignal and McCracken, 1996). These land-uses practices were crucial in preserving the ecological conditions, physiognomy and floristic composition of heathlands (Webb, 1998). Atlantic heathlands are semi-open shrubby formations that develop on nutrient-poor and acidic soils under a temperate oceanic climate in the European Atlantic zone (Schellenberg and Bergmeier, 2020; De Graaf et al., 2009). The vegetation is dominated by ericaceous species and gorse (genus *Ulex*) (Gimingham, 1972). Heathlands are a prime habitat for pollinating insects because of the

characteristics of their vegetation: a high nutritional value and a long flowering period, which guarantees access to resources throughout the year (Moquet et al., 2017). In addition, heathlands support a unique and rare diversity of plants and animals and provide important ecosystem services, such as carbon sequestration, water quality and cultural heritage (Walmsley et al., 2021; De Graaf et al., 2009). Consequently, heathlands have been identified as a matter of Community interest in accordance with the European Union's Habitats Directive (Directive 92/43/EEC) and their conservation status has emerged as a significant concern (Olmeda et al., 2020; Schellenberg and Bergmeier, 2020).

Despite their conservation interest, heathlands have been in general decline across Europe over the last century as a result of global anthropogenic changes (Fagúndez, 2013), including land-use changes, climate change and atmospheric nitrogen deposition (Bobbink et al., 2010). Land-use changes toward intensified agricultural practices have disrupted the traditional management and the preservation of heathlands (Ellis, 2021; Fagúndez, 2013). Considered unproductive, heathlands have either been abandoned or converted to more intensive

<footnote>* Corresponding author at: UMR 6553 ECOBIO, Campus de Beaulieu, University of Rennes, 35700 Rennes, France.</footnote>

<footnote>E-mail addresses: mathilde.dano@univ-rennes.fr (M. Dano), vincent.jung@univ-rennes.fr (V. Jung), gabrielle.thiebaut@univ-rennes.fr (G. Thiébaut), simon.chollet@univ-rennes.fr (S. Chollet).</footnote>
agrosystems (Ellis, 2021; Fagúndez, 2013). Traditional farming practices abandonment has led to heathland shrub encroachment and consequently to the loss of the species specialist of open habitats (Prévosto et al., 2011; Rose et al., 2000). In this context, large areas of heathland have been converted to pine plantations in order to enhance their economic value (Fagúndez, 2013; Donald et al., 1998). Also, agricultural intensification (mechanization, fertilization; Ellis, 2021) has led to conversion of some heathlands to arable lands (crops, grasslands, Fagúndez, 2013; Hooftman and Bullock, 2012). In addition to land-use changes, heathlands are sensitive to the effects of climate change and atmospheric nitrogen deposition (Bråthen et al., 2024; Britton et al., 2017; Bobbink et al., 2010). Rising global temperatures leads to thermophilization of plant communities (Britton et al., 2017) and atmospheric nitrogen deposition leads to the eutrophication of heathland ecosystems (Gaudnik et al., 2011; De Graaf et al., 2009). In these ecosystems with nutrient-poor soils, nutrient enrichment favours the dominance of fast-growing species over heathland plants, specifically adapted to oligotrophic conditions (Bobbink et al., 2010; Diemont, 1996). The degradation of heathlands over time can be reflected qualitatively, i.e., by changes in habitat type, such as historic heathlands now being converted to woodlands or arable lands. However, in a less visible way, degradations may also have affected areas that are today still heathland habitats. For these remnant heathlands, an important issue is therefore to know to what extent plant communities have changed compared with the previous situation. In this case, plant communities can be analysed through different parameters such as species diversity, the presence or abundance of particular species (typical heathland species, rare species), or plant-based indicator values for ecological conditions. In addition, these shifts in species composition can alter the role of heathland in supporting biodiversity, especially for insect pollinator populations that are dependent of floral resource in heathlands (Moquet et al., 2017).

Previous studies have shown that the response of heathland vegetation to the various facets of global change can be strongly modulated by local abiotic conditions. In particular, heathlands located near the coastline, and therefore subject to arch environmental conditions (e.g., wind, salinity), are known to be less subject to shrub and tree encroachment after abandonment of agricultural practices (Glemarec and Bioret, 2023). Another crucial factor that may influence the response of heathland vegetation is related to the management practices applied in protected areas. In fact, due to their ecological and cultural importance, some heathland areas are managed (e.g., by cutting ligneous plants, controlling bracken [Pteridium aquilinum] and grazing), in order to maintain them in a favourable conservation status (Olmeda et al., 2020; Diaz et al., 2013).

Long-term resurvey studies are very useful tools for documenting changes in vegetation and landscapes over time (Kapfer et al., 2017). In order to understand the extent of the changes that occurs in heathland plant communities over the last 50 years at a regional scale, we carried out a vegetation resurvey at a large number of sites throughout the Armorican massif (north-west of France). In this region, heathlands were estimated to cover approximately one million hectares in the mid-19th century, representing almost 40 % of the land area. However, although considered as a major component of the region's cultural identity, they today account for only 1 % (Sellin, 2022) and the current state of vegetation degradation is not precisely known. The main objectives of our study were (1) to reveal the qualitative changes in habitat types that have occurred over the last five decades at regional scale; (2) to analyse the responses of various community parameters, both for all the sites and only for those that had remained heathland over time; (3) to determine whether the responses of plant communities are influenced by the location of sites (coastal vs. inland) and by their designation as protected areas.

## 2. Material and methods

### 2.1. Study area

The Armorican Massif is a natural geographical area located in north-western France (Fig. 1a), extending over ca. 62000km² with an elevation ranging from 0 to 384 m, characterized by acidic soils (granite, schist, and sandstone; Clement and Touffet, 1990). The elevation ranges from 0 to 384 m. The climate is typically Atlantic, dominated by oceanic influences which results in a low temperature range (8.5 °C–15 °C) and rainfall distributed throughout the year (800-1200 mm/year) (Météo-France, 2024; Corillion, 1971). Historically, the massif is characterized by the widespread presence of heathland habitats, which are favoured by the acidic soil and past farming practices (Corillion, 1971). These heathlands include various types of habitats that depend mainly on soil moisture conditions (dry, mesic and wet heathlands) and are considered to be habitats of Community interest (Annex I of the Habitats Directive 92/43/EEC; Bensettiti et al., 2005). In the Armorican Massif heathlands are present both inland and on the coast. In this study we considered as ‘coastal heathlands’ only the heathlands located within the 250 m margin from the coastline that are directly exposed to sea spray (Glemarec and Bioret, 2023; Fig. 1). The post-war decades were a pivotal period for the natural habitats of the region, characterized by significant and rapid changes in land-use, marked by the expansion of urban areas, the intensification of agricultural practices and the abandonment of traditional uses (Dalton and Canévet, 1999). To counter the decline of heathlands habitats, some sites have been designated as protected areas and have thus benefited from management practices dedicated to heathland conservation. In this study only strictly protected areas were considered, i.e., zones with only limited and well-controlled human activities and where management measures can be taken to restore or conserve habitat (Cazzolla Gatti et al., 2023). On a regional scale these protected areas account for only 1 % (755 km²) of the total surface area.

### 2.2. Vegetation resurvey

To assess changes in heathlands plant communities between two time periods, historical data were extracted from previous vegetation studies. These studies were carried out from 1947 to 1998 (most of them between 1971 and 1975, Text A1, Fig. A1), with the main aim of describing the plant composition of the different heathland habitat types in the Armorican massif. Vegetation data were collected in plots ranged in size from 2 to 400 m², within which all the vascular plant species were inventoried and their cover-abundance estimated according to the Braun-Blanquet scale (Braun-Blanquet et al., 1932). These plots were not permanently marked in the field, but information on their location was available in the form of annotated paper maps, GPS coordinates or precise written indications. For resampling purposes, the historical plots can therefore be considered as ‘quasi-permanent’ plots, as they can be relocated approximately using plot-specific geographic positions, making them suitable for one-to-one comparisons between two time periods (Kapfer et al., 2017). We were thus able to relocate 444 plots, 87 of which were on the coastal margin (vs. 357 inland) and 183 in protected areas (vs. 261 unprotected). All the historical plots were resampled from late April to late June over three years from 2021 to 2023. The vegetation in each plot was sampled using the same protocol as described above for the historical data, keeping the same initial plot size. The taxonomic standardization of the various data sources was carried out using the TaxRef v16 taxonomic reference system (TaxRef, 2022).

### 2.3. Qualitative habitat changes

In order to characterize the qualitative habitat changes between the historical and current periods, we distinguished five main habitat types in our historical and resurvey datasets: Erica heathlands, Molinia/ bracken-dominated heathlands, encroached heathlands, wooded
<figure><img src="image_5.png" /><figcaption>**Fig. 1.** Location of the Armorican massif (grey) in the northwest of France. The points correspond to the location of the vegetation surveys (N = 444). The white dots correspond to coastal heathlands (i.e. located <250 m from the coastline) and black dots correspond to inland heathlands. The dots circled in red represent heathlands located in a protected area. The resolution of the map may not allow to distinguish points that are very close to each other. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)</figcaption></figure>

heathlands and woodlands. Erica heathlands are composed of ericaceous
species without non-typical forest woody species. While Erica heath-
lands correspond to typical heathland habitats, Molinia and/or bracken-
dominated heathlands, encroached heathlands and wooded heathlands
can be considered as degraded heathland habitats. Molinia and bracken-
dominated communities are considered as variants of heathland habitat
due to their similar environmental conditions and open physiognomy,
but over-dominated by one of the two species due to previous land-use
(Bensettiti et al., 2005). Encroached heathlands refer to communities in
the initial stage of natural succession from Erica heathlands, character-
ized by the colonization of non-typical forest woody species but with
heathland shrubs remaining dominant. Wooded heathlands are defined
as an intermediate stage in the succession of heathlands to forests, where
ericaceous species, generally at low abundance, persist under the tree
cover. Woodlands are habitats that have completely lost their heathland
identity (i.e., no longer hosting any typical heathland species). Each
vegetation plot (both historical and resampled) was assigned to one of
these types according to the total cover of ericaceous species (Erica cil-
iaris, E. cinerea, E. tetralix, Calluna vulgaris), typical shrubs (Ulex minor,
U. europaeus, U. gallii, Cystisus scoparius, Myrica gale, Frangula alnus, Salix
repens) (Bensettiti et al., 2005; Table A2) and non-typical forest woody
species (e.g., Quercus spp., Castanea sativa, Pinus spp.). The decisions
rules for habitat assignment are available in Appendix A (Fig. A2). The
evolution of proportion of each habitat type between historical and
current sampling period was visualized through a Sankey diagram using
the ‘ggsankey’ (Sjoberg, 2024) R-package. For wooded heathlands and
woodlands, we distinguished between pine plantations and deciduous
vegetation, but we did not include them as sub-categories on the Sankey

diagram in order to optimize its readability.

## 2.4. *Changes in plant community parameters*

For the subsequent quantitative analyses, the dataset was split into
two parts. The first dataset (n = 429), named ‘All Historic Heathlands’
(AHH, Fig. A3, Table A1), integrated all plots except those converted to
urban or crop areas (n = 15). The second dataset was a sub-sampling of
the first (n = 132) and integrated only those plots of Erica heathland that
had since remained Erica heathland. This subset was hereafter referred
to as ‘Remnant Erica Heathlands’ (REH, Fig. A3, Table A1).

The following metrics were calculated for each plot in each period:
species richness, richness and abundance of obligately insect-pollinated
plants species, vegetation conservation status (VCS; Jung et al., 2021),
total abundance of ericaceous species, range-rarity richness (RRR; Wil-
liams et al., 1996) and community-weighted mean Ellenberg indicator
values (CWM EIV; Ellenberg, 1974).

Species richness corresponds to the number of plant species recor-
ded. Obligately insect-pollinated plants correspond to species that are
exclusively pollinated by insects (Pan et al., 2024) and concerned 132
species of our dataset (Table A2). The richness and total abundance of
these species were calculated to assess the role of heathland in sup-
porting pollinator diversity. Pollination data were extracted from
Baseflor (Julve, 1998).

The VCS index is a tool for assessing the habitat conservation status
through plant species composition, and is based on the concept of the
‘habitat-specific pool’ (Zobel, 2016). The VCS takes into account the
richness and the abundance of both ‘typical’ and ‘non-typical’ species
(Jung et al., 2021). The typical species refers here to plant species that belong to the habitat species pool of heathlands, i.e., all species that can be found in heathlands in the absence of habitat degradation. In contrast, non-typical species do not belong to the habitat species pool of heathlands and thus reflect heathland degradation or a drift toward another habitat. The VCS index is formulated as:

$$VCS = \left[ 1 - \sum (n_j/N_T)^2 \right] \times (N_T/N)^2$$

where $n_j$ is the abundance of each typical species j, $N_T$ is the sum of the abundance of all typical species, and N is the sum of the abundance of all species including both typical and non-typical species.

As suggested by Helm et al. (2015), we employed a combined expert-based approach and information provided in regional flora (des Abbayes et al., 1971) and in Natura 2000 habitats manuals (Bensettiti et al., 2005) to check the typicity of the species. Some species that could not be classified as either typical or non-typical – and with low occurrences or incomplete identification – were considered ‘neutral’, i.e., not considered in the calculation of the VCS index. Among the 351 species recorded, 82 species were classified as typical species, 226 as non-typical species and 37 as neutral species (Table A2).

As ericaceous species are a marker of heathland identity and contribute to both soil properties and overall ecosystem function (Liu et al., 2021), particular emphasis was placed on these species by comparing their total abundance between the two sampling periods.

The range-rarity richness (RRR) was used to measure the concentration of rare species within plots. Here, we defined ‘rare species’ as species with a restricted geographical distribution relative to other species of the same taxon (Gaston, 1994). The rarity score of a species is characterized as the inverse of the number of geographical units in which it occurs in a given area (Williams et al., 1996). In our study, species rarity scores were calculated from occurrence data within 10 km × 10 km UTM grid cells in France over the period 1990–2019 (Just et al., 2015; available at https://siflore.fcbn.fr/). The RRR of each community was calculated as the mean of the rarity scores of species (Williams et al., 1996). The RRR index was formulated as:

$$RRR = \left( \sum_{1}^{s} 1 / Q_{j} \right) / S$$

where $Q_j$ is the number of grid cells occupied by each species j and S is the number of species in the community.

Community-weighted Mean Ellenberg indicator values (CWM EIV; Ellenberg, 1974) were used to compare environmental conditions between historical and resurveyed plots. EIVs serve as biotic proxies that provide insight into abiotic environmental characteristics and are widely used in European resurvey studies (Zolotova et al., 2023). The effect of encroachment was assessed with CWM of light availability (canopy closure). The effect of atmospheric N deposition was assessed with CWM of soil nutrient availability (eutrophication). To detect climate change effects, CWM of soil moisture and temperature were calculated (desiccation, thermophilization). As soil acidity is a determinant environmental factor influencing heathland development, the CWM of soil reaction was also calculated. The Ellenberg indicator values adapted to the Atlantic domain were extracted from Baseflor (Julve, 1998).

## 2.5. Statistical analysis

First, we analysed the response of each community parameter between the two sampling periods (historical vs. current) using linear model (LM). To take into account that historical surveys are spread over several decades, we performed a second analysis to assess whether the responses of community parameters between the historical and current periods depend on the time interval between the two sampling dates. We therefore used a LM to test the effect of the time interval between paired

sampling dates and the difference in each community parameter between the current period and the historical period.

Finally, we analysed whether the responses of community parameters between the historical and current periods are influenced by the location on coastal vs. inland area, or by protection status (protected vs. unprotected). For each community parameter, we used a LM including the following factors as independent categorical variables: the period (historical vs. current), the location (coastal vs. inland), the protection status (protected vs. unprotected), as well as the period*location and the period*protection interactions. Particular attention was paid to these two interaction factors, as they indicate the effects of location or protection status on the response of community parameters from historical to current periods. Variables retained in each final model were selected using a multi-model inference based on the bias-corrected Akaike information criterion (AICc), using the ‘MuMln’ (Barton, 2024) R-package. From all possible candidate models, we retained the ones with a ΔAICc <2, according to Burnham and Anderson (2002). When more than one model was selected, we performed a model-averaging approach to account for uncertainty in the selection process (Burnham and Anderson, 2002). The full-model averaged estimates and 95 % confidence intervals (CI) were generated, and the effects of variables were considered significant if their 95 % CI did not overlap zero (‘MuMln’ R-package).

For each LM constructed, spatial autocorrelation in residuals was examined using the Moran index (R-package ‘spdep’, Bivand et al., 2024a). When significant spatial autocorrelation was found, we chose autoregressive spatial models (R-package ‘spatialreg’, Bivand et al., 2024b) selected following a Lagrange-multiplier diagnostic (Bivand et al., 2024a, see Fig. B1, Text B1 and Table B1 for the used analysis workflow).

# 3. Results

## 3.1. Qualitative habitat changes

The historical dataset comprised 62 % of typical Erica heathlands and 38 % of other heathland habitats (Molinia/bracken-dominated heathlands, encroached heathlands and wooded heathlands). Today these proportions are 33 % and 42 % for Erica heathlands and other heathland habitats respectively, while woodlands, urban areas and arable lands (initially missing from the historical dataset) account for 22 %, 2 % and 1 % respectively (Fig. 2). Within the current wooded heathlands and woodlands, 56 % correspond to deciduous vegetation and 44 % to pine plantations. Almost half (46 %) of the historic Erica heathlands is now dominated by Molinia or bracken or has evolved into habitats representative of later successional stages.

## 3.2. Response of community parameters

For the All Historic Heathlands dataset (AHH), a significant decline was observed between the two sampling periods for all six metrics (Fig. 3). By restricting the analysis to the Remnant Erica Heathlands dataset (REH) only, no further significant declines appear for abundance of obligately insect-pollinated plants and rarity index (Fig. 3).

For the AHH dataset, CWM of Ellenberg indicator values decreased significantly over time for light availability and temperature, and increased for soil nutrient availability and soil reaction (Fig. 4). For the REH dataset, only the decrease for light and the increased for soil reaction are still observed.

For species richness, species richness of insect-pollinated plants and abundance of ericaceous species, the decrease observed between the historical and the current periods for the AHH dataset (Fig. 3) was more pronounced when the time interval between sampling dates increased (Table B4). For all other community parameters, the time interval between sampling periods showed no significant effect (Tables B4 & B5).
<figure><img src="image_6.png" /><figcaption>**Fig. 2.** Sankey diagram of changes between the historical (left) and the current (right) habitat type based on a categorization of the surveys using species composition. Blocks correspond to habitat types (*Erica heathlands*, *Molinia caerulea* and/or bracken-dominated heathlands, encroached heathlands, wooded heathlands, woodlands, arable lands and urban areas). The flows reflect the trajectory of each historical survey. The height of the blocks and the width of the flows are proportional to the number of surveys.</figcaption></figure>

### 3.3. Effects of location in coastal zone and protected areas

For both AHH and REH datasets, species richness of insect-pollinated plants was higher in coastal than in inland locations (Table 1, Fig. B2), while the opposite was observed for abundance of insect-pollinated plants. For the REH dataset, CWM EIV for light and temperature were higher in coastal than in inland locations and CWM EIV for soil reaction was lower in coastal locations. A significant effect of the period\*location interaction was found for species richness and species richness of insect-pollinated plants measured in the AHH dataset (Table 1). This interaction reflects the fact that the decrease in these two parameters between the two sampling periods was more pronounced for inland than for coastal heathlands (Fig. B2). Also, the increase of CWM EIV for soil nutrient concentration and the decrease of CWM EIV for temperature between sampling periods were more pronounced in coastal than in inland areas, as reflected by the significant effects of the period\*location interaction (Table 1, Fig. B2). For the REH dataset, a significant interaction persists only for species richness of insect-pollinated plants.

Significant effects of protection were found for most community parameters from the AHH dataset (Table 1). In particular, species richness, species richness of insect-pollinated plants, VCS index, abundance of ericaceous species and RRR index were higher in protected than in unprotected areas (Fig. B3). Community weighted mean EIV for soil moisture and soil reaction were lower in protected than unprotected areas (Table 1, Fig. B3). Significant effects for the period\*protection interaction were found only for species richness and species richness of insect-pollinated plants from the AHH dataset (Table 1), and indicates that the decrease in these two parameters were more pronounced in protected areas (Fig. B3).

## 4. Discussion

### 4.1. Habitat shifts are driven by afforestation and abandonment of traditional uses

On a regional scale, our qualitative results showed an important decline of *Erica* heathlands. We also found that a fifth of former

heathlands (all types combined) became woodland that no longer host any typical heathland species. Afforestation is therefore the most significant land-use change that has taken place over the study period in the Armorican Massif, occurring at rates comparable to those reported in other European studies (Rose et al., 2000; Diemont, 1996). This afforestation has its origins in two very different mechanisms that have occurred at comparable frequencies on a regional scale: conifer plantation and ecological succession. Indeed, the loss of agronomic interest in heathland in the decades following the Second World War has led to either their conversion to tree plantations (mainly *Pinus pinaster* and *Pinus sylvestris*) or to their natural colonization by deciduous forest following the abandonment of cultivation practices. Historically, traditional land-uses (e.g., mowing, grazing, burning; Ellis, 2021; Webb, 1998) exported organic matter, keeping soils nutrient-poor while preventing the colonization of woody plants (Prévosto et al., 2011; Webb, 1998). The abandonment of these traditional uses has naturally led to colonization by forest species.

Another notable change was the rise of heathland dominated primarily by *Pteridium aquilinum*, *Molinia caerulea* or a combination of both species. This dominance can be explained by suitable site conditions (e. g., wind-sheltered areas, slope, soil acidity) and by the abandonment of some land-uses (e.g., burning, grazing, ploughing), that are known to favour the development of *M. caerulea* and/or *P. aquilinum* (Marrs and Watt, 2006).

In contrast to afforestation, very few of the historic heathlands in our study have been converted to crops or urban areas. This is in line with long-term resurvey studies conducted during the second half of the 20th century, which usually only report residual rates of conversion of these type of land-uses (<5%; Diaz et al., 2013; Rose et al., 2000). In contrast, studies covering the early 20th century indicate significant conversion of heathland into urbanized and cultivated areas (Hooftman and Bullock, 2012; Piessens and Hermy, 2006). The low proportion of heathland converted to arable land observed in our study is therefore due to the fact the great heathland conversion period took place well before the date of our historical data, mainly in the second half of the 19th century (Dalton and Canévet, 1999). In addition, it can be argued that during the period covered by our study, laws have emerged to
<figure><img src="image_7.png" /><figcaption>**Fig. 3.** Boxplots showing (a) species richness, (b) richness of obligately insect-pollinated plants species and (c) abundance of obligately insect-pollinated plants species, (d) vegetation conservation status index (VCS index), (e) total abundance of ericaceous species and (f) range-rarity richness (RRR index) between the two survey periods (historical in white and current in grey) for the All Historical Heathlands dataset (AHH, *N* = 429) and for the Remnant *Erica* Heathlands dataset (REH, *N* = 132). Significance was obtained using autoregressive spatial models (results available in Tables B2 and B3 - Appendix B). (*** *p* < 0.001, ** *p* < 0.01, * *p* < 0.05, ns not significant).</figcaption></figure>

regulate urbanization, particularly in coastal areas (e.g., French Coastal Act in 1986).

## 4.2. Dramatic changes in community parameters

The major changes in land use that we have observed are concomitant with a dramatic decline in heathland conservation on a regional scale. Over the last five decades, species richness has declined by 19%, vegetation conservation status (VCS) by 39%, abundance of ericaceous species by 54% and average rarity by 25%. These declines collectively highlight the deterioration of biodiversity support and conservation status of heathlands. The strong decline of the vegetation conservation status (VCS) we found reflects a decrease in the richness and abundance of typical heathland species, such as ericaceous species, and an increase in the abundance of non-typical species, mainly woody species. These declines could be explained by some changes in ecological conditions and soil properties caused by afforestation. As shown by EIV value changes, the closure of the canopy led to shaded and semi-shaded conditions that favoured species tolerant to these new habitat conditions (e. g., *Hedera helix*, *Rubus* spp.). This shift resulted in a decline in

heliophilous species (e.g., *Jasione montana*, *Aira praecox*, *Cistus umbellatus*). In addition, regarding soil conditions, woody vegetation encroachment led to less oligotrophic and less acidic soils, altering the conditions required for heathland habitats. Similar changes linked to afforestation have been observed in several European studies (Diaz et al., 2013; Prévosto et al., 2011) and our results therefore confirm the importance of this phenomenon. The diversity of exclusively insect-pollinated plants has declined by 30% in richness and 20% in abundance, indicating reduced resource availability for pollinators. These findings are in line with those reported by Pan et al. (2024) in Dutch landscapes during the same period. These changes in vegetation composition and physiognomy can affect the fauna diversity dependent on open habitats (e.g., Silva-Monteiro et al., 2021; Moquet et al., 2017; Schirmel et al., 2011).

An unexpected result of this study was that even the historical heathlands that remain heathlands today are in a worse conservation status than they were. This degradation was revealed by a decline in the vegetation conservation status by 9%, species richness by 17%, the abundance of insect-pollinated and ericaceous species by 13% and 22% respectively. The decline of ericaceous species is particularly
<figure><img src="image_8.png" /><figcaption>**Fig. 4.** Boxplots showing community-weighted means of the Ellenberg indicators values (CWM EIV) of (a) soil moisture, (b) light availability, (c) soil nutrient availability, (d) soil reaction and (e) temperature between the two survey periods (historical in white, current in grey) for the All Historical Heathlands dataset (AHH, N = 429) and the Remnant *Erica* Heathlands dataset (REH, N = 132). Significance was obtained using autoregressive spatial models (results available in Tables B2 and B3 - Appendix B). (*** p < 0.001, **p < 0.01, *p < 0.05, ns not significant).</figcaption></figure>

concerning, given their vital role in heathland functioning, particularly nutrient cycling (Liu et al., 2021). In addition, *Ericaceae* are known for their resilience, attributed to their prolonged survival mechanisms (e.g., long-term persistent seed bank, long life-span), even under deteriorating environmental conditions (Schellenberg and Bergmeier, 2020; Piessens and Hermy, 2006). Consequently, the decline of these species may be a first sign of an even greater risk of heathland decline in the future (Bråthen et al., 2024; Ransijn et al., 2015). Nevertheless, it should be noticed that, compared to all historical heathlands, remnant heathlands are less severely degraded with the average rarity and abundance of insect-pollinated plants remaining stable. As highlighted in previous studies, preserving these few remnant heathlands is crucial, as they play an important role for biodiversity support (Ransijn et al., 2015; Diaz et al., 2013; Piessens and Hermy, 2006) and also because of the difficulty of restoring heathlands that have experienced important abiotic changes.

## 4.3. Coastal and protected heathlands conservation

The decreases in species richness and species richness of insect-pollinated plants observed in the coastal zone were mitigated compared to the inland area. In contrast, the increase in Ellenberg indicator value for soil nutrient was more pronounced in the coastal zone. These effects of coastal area location can be attributed to the combined influence of stressful environmental conditions and distinct successional dynamics compared to inland heathlands (Glemarec and Bioret, 2023; Olmeda et al., 2020). Encroachment in coastal heathlands – especially in areas sheltered from wind and sea spray – is expressed by the development of highly nitrophilous species like *Rubus* spp. or *Pteridium aquilinum*, resulting in a more pronounced increase in soil nutrient availability (EIV, Marrs and Watt, 2006). These successional dynamics tend to maintain the open physiognomy of coastal heathlands more effectively than in inland heathlands, where succession tends to favour trees and shrubs. These changes have also resulted in a less severe decline in species richness, including exclusively insect-pollinated plants, due to the persistence of species (especially herbaceous) adapted to these environmental conditions.

We found that community parameters corresponding to species richness, vegetation conservation status and plant species rarity were

higher in protected areas than in unprotected areas. This can be explained by the fact that historically, the designation of protected areas has targeted the best-preserved heathlands and this more favourable condition remains today. However, a key finding of our study is that protection did not mitigate the decline in these parameters over time, as indicated by the few interactive effects between period and protection. The only interactive effects observed for total and insect-pollinated species richness even indicate an accentuated decline in these parameters within protected areas. Although we do not have precise plot-specific information, some heathlands located in protected areas in the Armorican Massif have benefited from conservation management practices, mainly scrub clearance (Chollet et al., 2025). Our study therefore suggests that such (occasional) actions cannot substitute for traditional agricultural practices that have lasted for centuries, and are therefore not sufficient to ensure the long-term conservation of heathlands (Webb, 1998; Clément, 1978).

## 4.4. Heathlands show resistance to global changes but face rising long-term risks

Contrary to what we might have expected, our results do not show a trend that could demonstrate an effect of global changes, in particular atmospheric deposition or temperatures rising. Despite the documented increase in mean temperatures (about 1 °C increase) and atmospheric nitrogen deposition, reaching 18 kg/N/ha y⁻¹ in the study region (Gaudnik et al., 2011; Bobbink et al., 2010), we did not find significant trend related to these changes. While the thermophilization of the communities was expected, the observed decrease in EIV temperature contradicts this expectation. This reduction can be attributed to canopy closure, which was associated with cooler conditions leading to the decline of thermophilic species. The lack of response may be due to the sclerophyllous nature of heathland vegetation, which is well-adapted to extreme conditions. For example, dry heathland vegetation is particularly tolerant to elevated temperatures and pronounced water deficits, imposed by the shallow and heat-absorbing soils (Fagúndez, 2013; Bennetiti et al., 2005).

Similarly, although an increase in EIV for soil nutrient availability was expected in response to atmospheric nitrogen deposition, the observed increase would rather be attributable to the effect of
<caption>**Table 1**</caption>

Summary of the model-averaged estimates ±95 % confidence interval (CI) of the effects of sampling period (current relative to historical), location (coastal relative to inland) and protection (protected relative to unprotected), Period*Location and Period*Protection interactions on community parameters for the AHH (All Historic Heathlands) and the REH (Remnant Erica Heathlands) datasets. Stars indicates 95 % confidence intervals that did not encompass zero. Dashes indicates variables excluded by the selection model.

<table><thead><tr><th rowspan="2">Parameter</th><th rowspan="2">Term</th><th colspan="2">AHH</th><th colspan="2">REH</th></tr><tr><th colspan="2">Estimate ± 95% CI</th><th colspan="2">Estimate ± 95% CI</th></tr></thead><tbody><tr><td rowspan="6">Species richness</td><td>(Intercept)</td><td>2.368 ± 0.064</td><td>*</td><td>2.361 ± 0.104</td><td>*</td></tr><tr><td>Period</td><td>-0.194 ± 0.07</td><td>*</td><td>-0.259 ± 0.117</td><td>*</td></tr><tr><td>Location</td><td>-0.108 ± 0.127</td><td>ns</td><td>-0.119 ± 0.213</td><td>ns</td></tr><tr><td>Protection</td><td>0.111 ± 0.097</td><td>*</td><td>0.008 ± 0.064</td><td>ns</td></tr><tr><td>Period*Location</td><td>0.29 ± 0.141</td><td>*</td><td>0.173 ± 0.282</td><td>ns</td></tr><tr><td>Period*Protection</td><td>-0.255 ± 0.112</td><td>*</td><td>-</td><td>-</td></tr><tr><td rowspan="6">Species richness of insect-pollinated plants</td><td>(Intercept)</td><td>1.334 ± 0.19</td><td>*</td><td>1.976 ± 0.16</td><td>*</td></tr><tr><td>Period</td><td>-0.28 ± 0.099</td><td>*</td><td>-0.302 ± 0.16</td><td>*</td></tr><tr><td>Location</td><td>-0.201 ± 0.143</td><td>*</td><td>-0.361 ± 0.243</td><td>*</td></tr><tr><td>Protection</td><td>0.14 ± 0.113</td><td>*</td><td>0.165 ± 0.182</td><td>ns</td></tr><tr><td>Period*Location</td><td>0.569 ± 0.201</td><td>*</td><td>0.508 ± 0.278</td><td>*</td></tr><tr><td>Period*Protection</td><td>-0.18 ± 0.16</td><td>*</td><td>0.028 ± 0.162</td><td>ns</td></tr><tr><td rowspan="6">Abundance of insect-pollinated plants</td><td>(Intercept)</td><td>4.265 ± 0.599</td><td>*</td><td>2.948 ± 1.053</td><td>*</td></tr><tr><td>Period</td><td>-1.416 ± 0.399</td><td>*</td><td>-0.823 ± 1.154</td><td>ns</td></tr><tr><td>Location</td><td>1.142 ± 0.613</td><td>*</td><td>1.031 ± 0.847</td><td>*</td></tr><tr><td>Protection</td><td>0.03 ± 0.22</td><td>ns</td><td>0.452 ± 0.959</td><td>ns</td></tr><tr><td>Period*Location</td><td>0.15 ± 0.69</td><td>ns</td><td>-0.082 ± 0.698</td><td>ns</td></tr><tr><td>Period*Protection</td><td>-</td><td>-</td><td>0.544 ± 1.37</td><td>ns</td></tr><tr><td rowspan="6">VCS index</td><td>(Intercept)</td><td>0.258 ± 0.036</td><td>*</td><td>0.082 ± 0.025</td><td>*</td></tr><tr><td>Period</td><td>-0.189 ± 0.032</td><td>*</td><td>-0.007 ± 0.018</td><td>ns</td></tr><tr><td>Location</td><td>-0.004 ± 0.024</td><td>ns</td><td>0.001 ± 0.006</td><td>ns</td></tr><tr><td>Protection</td><td>0.044 ± 0.035</td><td>*</td><td>0.019 ± 0.016</td><td>*</td></tr><tr><td>Period*Location</td><td>-</td><td>-</td><td>-</td><td>-</td></tr><tr><td>Period*Protection</td><td>0.002 ± 0.029</td><td>ns</td><td>-0.011 ± 0.024</td><td>ns</td></tr><tr><td rowspan="6">Abundance of ericaceous species</td><td>(Intercept)</td><td>3.398 ± 0.469</td><td>*</td><td>1.749 ± 0.967</td><td>*</td></tr><tr><td>Period</td><td>-2.687 ± 0.449</td><td>*</td><td>-0.251 ± 1.131</td><td>ns</td></tr><tr><td>Location</td><td>0.015 ± 0.263</td><td>ns</td><td>0.686 ± 0.981</td><td>ns</td></tr><tr><td>Protection</td><td>0.852 ± 0.482</td><td>*</td><td>1.837 ± 1.072</td><td>*</td></tr><tr><td>Period*Location</td><td>-</td><td>-</td><td>-</td><td>-</td></tr><tr><td>Period*Protection</td><td>-0.034 ± 0.412</td><td>ns</td><td>-0.848 ± 1.549</td><td>Ns</td></tr><tr><td rowspan="6">RRR index</td><td>(Intercept)</td><td>0.001 ± 8.97e-5</td><td>*</td><td>-3.687 ± 4.989</td><td>ns</td></tr><tr><td>Period</td><td>-2e-4 ± 6.7e-5</td><td>*</td><td>-0.024 ± 0.086</td><td>ns</td></tr><tr><td>Location</td><td>2.2e-5 ± 7.8e-5</td><td>ns</td><td>0.018 ± 0.082</td><td>ns</td></tr><tr><td>Protection</td><td>7.7e-5 ± 7.1e-5</td><td>*</td><td>0.004 ± 0.041</td><td>ns</td></tr><tr><td>Period*Location</td><td>-</td><td>-</td><td>-</td><td>-</td></tr><tr><td>Period*Protection</td><td>-</td><td>-</td><td>-</td><td>-</td></tr><tr><td rowspan="6">Moisture</td><td>(Intercept)</td><td>1.31 ± 0.042</td><td>*</td><td>1.622 ± 1.959</td><td>ns</td></tr><tr><td>Period</td><td>0.001 ± 0.011</td><td>ns</td><td>-0.003 ± 0.021</td><td>ns</td></tr><tr><td>Location</td><td>-0.047 ± 0.082</td><td>ns</td><td>-0.025 ± 0.079</td><td>ns</td></tr><tr><td>Protection</td><td>-0.069 ± 0.048</td><td>*</td><td>-0.008 ± 0.041</td><td>ns</td></tr><tr><td>Period*Location</td><td>-</td><td>-</td><td>-</td><td>-</td></tr><tr><td>Period*Protection</td><td>-</td><td>-</td><td>-</td><td>-</td></tr><tr><td rowspan="6">Light</td><td>(Intercept)</td><td>1.19 ± 0.177</td><td>*</td><td>1.389 ± 0.342</td><td>*</td></tr><tr><td>Period</td><td>-0.041 ± 0.017</td><td>*</td><td>-0.068 ± 0.048</td><td>*</td></tr><tr><td>Location</td><td>0.016 ± 0.023</td><td>ns</td><td>0.099 ± 0.065</td><td>*</td></tr><tr><td>Protection</td><td>-0.001 ± 0.013</td><td>ns</td><td>-</td><td>-</td></tr><tr><td>Period*Location</td><td>-0.001 ± 0.014</td><td>ns</td><td>0.012 ± 0.068</td><td>ns</td></tr><tr><td>Period*Protection</td><td>-0.004 ± 0.02</td><td>ns</td><td>-</td><td>-</td></tr><tr><td rowspan="6">Nutrient</td><td>(Intercept)</td><td>0.938 ± 0.038</td><td>*</td><td>0.671 ± 0.722</td><td>ns</td></tr><tr><td>Period</td><td>0.179 ± 0.035</td><td>*</td><td>0.032 ± 0.05</td><td>ns</td></tr><tr><td>Location</td><td>0.074 ± 0.077</td><td>ns</td><td>-</td><td>-</td></tr><tr><td>Protection</td><td>-0.014 ± 0.047</td><td>ns</td><td>-0.004 ± 0.03</td><td>ns</td></tr><tr><td>Period*Location</td><td>0.096 ± 0.077</td><td>*</td><td>-</td><td>-</td></tr><tr><td>Period*Protection</td><td>-0.005 ± 0.033</td><td>ns</td><td>-</td><td>-</td></tr><tr><td rowspan="6">Reaction</td><td>(Intercept)</td><td>3.5 ± 0.122</td><td>*</td><td>0.856 ± 0.224</td><td>*</td></tr><tr><td>Period</td><td>0.448 ± 0.131</td><td>*</td><td>0.057 ± 0.086</td><td>Ns</td></tr><tr><td>Location</td><td>-0.058 ± 0.195</td><td>ns</td><td>-0.143 ± 0.084</td><td>*</td></tr><tr><td>Protection</td><td>-0.419 ± 0.179</td><td>*</td><td>-0.108 ± 0.082</td><td>*</td></tr><tr><td>Period*Location</td><td>0.016 ± 0.125</td><td>ns</td><td>0.011 ± 0.075</td><td>ns</td></tr><tr><td>Period*Protection</td><td>0.05 ± 0.182</td><td>ns</td><td>0.029 ± 0.105</td><td>ns</td></tr><tr><td rowspan="6">Temperature</td><td>(Intercept)</td><td>1.388 ± 0.013</td><td>*</td><td>1.373 ± 0.024</td><td>*</td></tr><tr><td>Period</td><td>-0.017 ± 0.014</td><td>*</td><td>-</td><td>-</td></tr><tr><td>Location</td><td>0.029 ± 0.028</td><td>*</td><td>0.044 ± 0.036</td><td>*</td></tr><tr><td>Protection</td><td>0.001 ± 0.01</td><td>ns</td><td>0.007 ± 0.025</td><td>ns</td></tr><tr><td>Period*Location</td><td>-0.041 ± 0.031</td><td>*</td><td>-</td><td>-</td></tr><tr><td>Period*Protection</td><td>-</td><td>-</td><td>-</td><td>-</td></tr></tbody></table>
afforestation. Indeed, the encroachment of woody vegetation led to higher soil nutrient availability due to improved litter quality and subsequent mineralization (Prévosto et al., 2011; Mitchell et al., 1997). This effect of afforestation seems to be confirmed by the fact that the remnant heathlands show no signs of eutrophication (no change in EIV for soil nutrient). However, while the heathlands in our study appear to be resistant to global change, this resistance may not be indefinite. In fact, the increasing intensity and duration of heatwaves, combined with chronic nitrogen inputs, could threaten heathlands and their resistance in the future (Britton et al., 2017).

## 4.5. Conclusion

Our long-term regional study has shown that 25 % of heathland has disappeared as a result of urbanization, agricultural intensification and, above all, afforestation, whether artificial (pine plantation) or spontaneous. We have also shown that many sites that still have heathland today are more degraded than they were in the past. Moreover, site-specific conditions (coastal or protected location) only marginally compensate for the overall degradation. Nevertheless, it seems necessary to continue monitoring the evolution of heathland habitats in the future, as the acceleration of global change is likely to exacerbate an already deteriorating situation in a region recognized for this emblematic cultural landscape.

## CRediT authorship contribution statement

**M. Dano:** Writing – review & editing, Writing – original draft, Methodology, Investigation, Funding acquisition, Formal analysis, Data curation, Conceptualization.
**V. Jung:** Writing – review & editing, Validation, Supervision, Methodology, Investigation, Funding acquisition, Data curation, Conceptualization.
**G. Thiébaut:** Writing – review & editing, Validation, Supervision, Resources.
**S. Chollet:** Writing – review & editing, Validation, Supervision, Methodology, Investigation, Data curation, Conceptualization.

## Funding

This work was supported by two grants: the “Défis scientifiques” grant of the University of Rennes, obtained by Simon Chollet, and the “bourse de campagne de terrain” grant of the Société Française d'Écologie et d'Évolution (SFE²), obtained by Mathilde Dano.

## Declaration of competing interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

## Acknowledgments

The authors would like to thank Tanguy Doaré and Léo Drapier for their help in the field. We would also like to thank the editor (Orsolya Valkó) and two anonymous reviewers for their very constructive comments on the manuscript.

## Appendices A-B. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.biocon.2025.111210.

## Data availability

Data will be made available on request.

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