Introduction
Species face numerous threats, principally related to human activities, and biodiversity continues to be lost (Pimm et al., Reference Pimm, Russel, Gittleman and Brooks1995; Butchart et al., Reference Butchart, Walpole, Collen, van Strien, Scharlemann and Almond2010; Ceballos et al., Reference Ceballos, Ehrlich, Barnosky, García, Pringle and Palmer2015). Halting, or at least significantly reducing, the loss of biodiversity requires adequate investment and a comprehensive and reliable measure of conservation status (e.g. Balmford et al., Reference Balmford, Bennun, ten Brink, Cooper, Côté and Crane2005; Fenu et al., Reference Fenu, Bacchetta, Giacanelli, Gargano, Montagnani and Orsenigo2017; Orsenigo et al., Reference Orsenigo, Montagnani, Fenu, Gargano, Peruzzi and Abeli2018). Target II of Objective I of the Global Strategy for Plant Conservation 2011–2020 of the Convention on Biological Diversity (CBD; GSPC, 2008) is the preliminary assessment of the conservation status of the Earth's flora. Evaluation of species conservation status is required not only to evaluate progress towards the CBD's Aichi Targets of the Strategic Plan for Biodiversity 2011–2020 (Pimm et al., Reference Pimm, Jenkins, Abell, Brooks, Gittleman and Joppa2014) but also to identify and develop effective conservation strategies (Rodrigues et al., Reference Rodrigues, Pilgrim, Lamoreux, Hoffmann and Brooks2006; Mace et al., Reference Mace, Collar, Gaston, Hilton-Taylor, Akçakaya, Leader-Williams and Stuart2008; Fenu et al., Reference Fenu, Cogoni, Pinna and Bacchetta2015a; Rossi et al., Reference Rossi, Orsenigo, Montagnani, Fenu, Gargano and Peruzzi2016; Collen et al., Reference Collen, Dulvy, Gaston, Gärdenfors, Keith and Punt2016; Orsenigo et al., Reference Orsenigo, Montagnani, Fenu, Gargano, Peruzzi and Abeli2018). The IUCN Red List criteria (IUCN, 2001, 2012a) are the accepted standard for assessing the extinction risk of species (Rodrigues et al., Reference Rodrigues, Pilgrim, Lamoreux, Hoffmann and Brooks2006; Mace et al., Reference Mace, Collar, Gaston, Hilton-Taylor, Akçakaya, Leader-Williams and Stuart2008; Collen et al., Reference Collen, Dulvy, Gaston, Gärdenfors, Keith and Punt2016; Orsenigo et al., Reference Orsenigo, Montagnani, Fenu, Gargano, Peruzzi and Abeli2018).
Habitat Directive 92/43/EEC is the core strategy for nature conservation in Europe (Balmford et al., Reference Balmford, Bennun, ten Brink, Cooper, Côté and Crane2005; Pullin et al., Reference Pullin, Baldi, Can, Dieterich, Kati and Livoreil2009; Beresford et al., Reference Beresford, Buchanan, Sanderson, Jefferson and Donald2016; Fenu et al., Reference Fenu, Bacchetta, Giacanelli, Gargano, Montagnani and Orsenigo2017). Through the implementation of cogent conservation policies, the Directive promotes the maintenance of a favourable conservation status for a group of key species and habitats (European Commission, 1992; Rossi et al., Reference Rossi, Orsenigo, Montagnani, Fenu, Gargano and Peruzzi2016; Fenu et al., Reference Fenu, Bacchetta, Giacanelli, Gargano, Montagnani and Orsenigo2017). It is mandatory for EU member states that have full responsibility for their conservation efforts to monitor and report the conservation status of all species listed in the Directive (European Commission, 1992; Rossi et al., Reference Rossi, Orsenigo, Montagnani, Fenu, Gargano and Peruzzi2016; Fenu et al., Reference Fenu, Bacchetta, Giacanelli, Gargano, Montagnani and Orsenigo2017).
One of the key species listed in the Habitat Directive is the rock rose Helianthemum caput-felis Boiss. (Cistaceae), which is protected by several international, national and regional regulations. Helianthemum caput-felis is a thermophilous long-lived half shrub that grows in coastal environments under the direct influence of the sea, mostly on calcareous rocky cliffs with garrigues or scrublands; some populations grow on sand dunes in Majorca, fossil dunes in Morocco and rocky slopes bordering inland ravines (Fenu et al., Reference Fenu, Cogoni, Sulis and Bacchetta2015b). The species occurs in several disjunct and fragmented populations throughout the coasts of the western Mediterranean Basin (south-east Iberian Peninsula, Balearic Islands, Sardinia and north-west Africa; Fenu et al., Reference Fenu, Cogoni, Sulis and Bacchetta2015b and references therein). The global conservation status of H. caput-felis was previously unknown because only regional or local assessments were available, and there was little information on the species distribution and conservation status in Africa. Helianthemum caput-felis is categorized as Endangered in Europe (Bilz et al., Reference Bilz, Kell, Maxted and Lansdown2011) and Spain (Agulló et al., Reference Agulló, Juan, Alonso, Crespo, Bañares, Blanca, Güemes, Moreno and Ortiz2010), and as Critically Endangered in Italy (Fenu et al., Reference Fenu, Cogoni, Sulis and Bacchetta2015b) and Algeria (Agulló et al., Reference Agulló, Juan, Crespo, Alonso and Terrones2017). In addition, according to the European Habitat Directive H. caput-felis has an inadequate conservation status in Italy (Fenu et al., Reference Fenu, Bacchetta, Giacanelli, Gargano, Montagnani and Orsenigo2017).
Here, to help establish a conservation plan for H. caput-felis, we evaluated the species' global conservation status using the Red List criteria. We aimed to: (1) describe the species' range based on extensive field surveys, (2) quantify the size and structure of populations, (3) identify the main threats to the species’ persistence, (4) assess the species' global conservation status, and (5) recommend appropriate conservation measures.
Methods
Data collection
The geographical distribution of H. caput-felis was determined through field surveys during 2012–2017 in all localities for which there were herbarium specimens or database records (Supplementary Table 1) and/or published data (Agulló et al., Reference Agulló, Juan, Alonso, Crespo, Bañares, Blanca, Güemes, Moreno and Ortiz2010; Fenu et al., Reference Fenu, Cogoni, Sulis and Bacchetta2015b; Sulis, Reference Sulis2016). Additionally, all sites along the coasts of Sardinia, the Balearic Islands and the Mediterranean coasts of Spain and Morocco with suitable ecological conditions for the species were surveyed. No surveys were made on the Algerian coast, for which information was obtained from Agulló et al. (Reference Agulló, Juan, Crespo, Alonso and Terrones2017).
In each locality where occurrence of the species was confirmed or discovered, the following analyses were undertaken. The geographical limits of confirmed localities were mapped and their area estimated using ArcGis 9.2 (Esri, Redlands, USA). Localities separated by > 1 km were considered to be geographically separate. At each locality the altitudinal range, slope, aspect and habitat type according to the Habitat Directive were determined. Major threats to H. caput-felis were identified through field observations (except for Algerian localities), and categorized following the IUCN Threats Classification Scheme (IUCN, 2012b).
Where possible, population size was determined by a direct count of the total number of mature plants. In extensive localities, population size was estimated from a count of all individuals in 3–15 plots (each plot was 2 m2; the number of plots depended on the location); in these cases population size (sensu IUCN, 2017) was estimated as d × A × p, where d is an estimate of density within the sampled plots, A is an estimate of the area occupied by the population, and p is an estimate of the proportion of mature individuals within the sampled plots. Populations were categorized in three size classes, defined a priori as small (< 100 individuals), medium (101–2,000) and large (> 2,000). The presence or absence of seedlings was also recorded in each locality by surveying the population several times during the recruitment season. Demographic data were collected on 98 permanent plots in six populations across the species range that were representative of the range of ecological conditions in which the plant grows (Table 1; Sulis et al., Reference Sulis, Bacchetta, Cogoni and Fenu2018). In each population, after excluding areas with marginal conditions for the species, permanent plots (2 × 1 m) were placed randomly in the area where the species was found; within the plots all plants present (a total of 821 in the 98 plots in the first census) were marked, mapped and monitored over a 3-year period.
Localities of Helianthemum caput-felis surveyed in this study, with altitudinal range, area, mean density ± SD, population size, whether seedlings were present, protection status, current status, source and the main threats (IUCN, 2012b). Localities in bold are those selected for study of detailed population dynamics (Sulis et al., Reference Sulis, Bacchetta, Cogoni and Fenu2018).
1Determined by counting (superscript C) or estimating (superscript E) the total number of mature plants and categorizing as small (< 100 individuals), medium (101–2,000) or large (> 2,000) population.
2SCI, Site of community importance; PMR, Plant micro-reserve sensu Laguna et al. (Reference Laguna, Deltoro, Pérez-Botella, Pérez-Rovira, Serra, Olivares and Fabregat2004); *protection only partially covers the population area.
Data analysis
To assess the extinction risk of H. caput-felis we used the Red List criteria (IUCN, 2001) and the guidelines for their application (IUCN, 2017), considering the criteria A, B and E. To determine whether H. caput-felis fulfilled criterion A we assessed any potential population reduction (observed, estimated or inferred) in the last 10 years or three generations, based on the area of occupancy (AOO; the area within the extent of occurrence occupied by the taxon, see below; IUCN, 2001).
To apply criterion B, the extent of occurrence (EOO; the area contained within the shortest continuous imaginary boundary that could be drawn to encompass all known sites of occurrence of a taxon, excluding cases of vagrancy) and the AOO were calculated from the distribution records. Extinctions that predated 1950 were excluded from these calculations. To estimate EOO, the minimum convex polygon that included all the occurrences was drawn (IUCN, 2017), with unsuitable areas excluded by deriving the correspondent α-hull using the Delauney triangulation (Burgman & Fox, Reference Burgman and Fox2003; Gargano et al., Reference Gargano, Fenu, Medagli, Sciandrello and Bernardo2007; IUCN, 2017). The AOO was calculated by generating a 2 × 2 km grid with ArcGis.
To apply criterion E the estimated quasi-extinction risks of populations were calculated based on a demographic study that used both an integral projection model and a matrix population model (Sulis, Reference Sulis2016; Sulis et al., Reference Sulis, Bacchetta, Cogoni and Fenu2018; Supplementary Material 1). The matrix projection model was used to calculate the estimated quasi-extinction risk using the popbio package (Stubben & Milligan, Reference Stubben and Milligan2007) in R 3.1.2 (R Core Team, 2014). The mean generation time (T) extracted from the integral projection model was 15.97 years (3 generations = 47.91 years) and the mean value extracted from the matrix projection model was 28.44 years (3 generations = 85.32 years; Sulis, Reference Sulis2016). Following the precautionary principle (IUCN, 2017), the smaller value of generation length (i.e. 15.97 years) was retained. Extinction risk was calculated considering three alternative assessments (10 years or three generations, 20 years or five generations, and 100 years). Quasi-extinction probabilities were calculated by 500 model iterations (van der Meer et al., Reference van der Meer, Dahlgren, Mildén and Ehrlén2014). Matrices were selected at random with replacement (each matrix had an equal probability of selection; Morris & Doak, Reference Morris and Doak2002). A quasi-extinction threshold of 20 individuals was designated a priori, to help minimize the demographic stochasticity associated with small population size (Morris & Doak, Reference Morris and Doak2002).
We also modelled three separate scenarios, one for each population size class, to examine extinction risk. We developed each model using the global stochastic growth rate (λs = 1; Sulis, Reference Sulis2016; Sulis et al., Reference Sulis, Bacchetta, Cogoni and Fenu2018) and the effective population size.
Results
We located records of H. caput-felis in 49 localities along the western Mediterranean coasts, mainly in Spain (32 localities, including two in Melilla), followed by Morocco (nine localities), Italy and Algeria (four localities in each country; Table 1). We confirmed the presence of H. caput-felis in 41 localities, 40 of which we verified in the field, and one of which was documented in Agulló et al. (Reference Agulló, Juan, Crespo, Alonso and Terrones2017; no. 49 in Table 1). In two localities on the Balearic Islands (Cala Conta in Ibiza, Maioris in Majorca) and one in Morocco (near Hidoun, Nador), field surveys carried out by ourselves and other researchers failed to locate the species. Following a precautionary approach, we reported the species as ‘not retrieved’ in these localities. We were, however, able to confirm recent extinctions at two localities in Europe (Santa Pola in Spain, Seu in Italy) and three localities in Algeria.
All sites where occurrence was confirmed are in coastal environments, with the exception of one population in Morocco that is 3.44 km from the coast. The altitudinal range at which we located the species was 0–220 m. In Europe, we found the plant below 110 m, whereas in North Africa the species occurred in some cases above 150 m. The slope of localities is 0–45°. The area occupied by the species in each locality was variable, from 10 m2 at Is Arutas in Sardinia to 116 and 101 ha at Taxdirt and Cap Blanc, respectively. Mean plant density was 3.19 ± SD 2.13 plants/m2, varying from 1.10 ± SD 0.65 plants/m2 (Cabo Cervera, Alicante) to 9.6 ± SD 1.85 plants/m2 (Sa Ràpita, Majorca). Population size ranged from a few mature plants (Is Arutas, Cabras) to more than tens of thousands (Cap Blanc, Llucmajor). More than half of localities (56.1%) had a population size > 2,000 mature plants. The population size was ≤ 100 individuals in 26.8% of localities. The population structure was mainly characterized by reproductive and juvenile plants (> 95%). Despite repeated surveys in the recruitment season, we observed seedlings in only 47.5% of localities, with recruitment absent in all small populations and in the 36.7% of medium and large populations (Table 1).
The main threats affecting the persistence of H. caput-felis populations are related to human activities (Fig. 1), including recreational activities (i.e. disturbance effects posed by recreation; 63.45% of localities), housing and urban areas (48.78%) and tourism and recreation (i.e. habitat effects of tourism and recreation sites with a substantial footprint; 39.00%). Other threats were invasive alien species (i.e. Carpobrotus sp., Agave sp., Acacia sp. and Ricinus communis L.; 41.46%), roads and railroads (26.83%), wood and pulp plantations (24.39%), and the presence of rubbish (IUCN threat classification ‘garbage and solid waste’; 21.95%). In 26.66% of localities, all in Europe, avalanches or landslides appear to threaten the persistence of H. caput-felis (Fig. 1, Table 1).
Number of localities subject to each of the main threats (Table 1) to the rock rose Helianthemum caput-felis populations; data for Algeria was obtained from Agulló et al. (Reference Agulló, Juan, Crespo, Alonso and Terrones2017).
According to the generation time extracted from the integral projection model, there was no extinction risk for three generations (Fig. 2), but there was an extinction probability of c. 30% for a five generation period (79.85 years; Fig. 2). Considering the generation time extracted from the matrix population model, the length of three generations was 85.31 years, which corresponded to a quasi-extinction risk of c. 45% (Fig. 2).
Simulated cumulative distribution functions of the number of years for populations of H. caput-felis to reach a quasi-extinction threshold of 20 individuals: a, three generations (integral projection model); b, five generations (integral projection model); c, three generations based on generation time from the matrix population model. The lines are separate estimates of the cumulative distribution of extinction probabilities based on 500 iterations of population growth over 100 years (see Stubben & Milligan, Reference Stubben and Milligan2007 for further details).
The quasi-extinction risk models for the three population size classes, considering each locality to be isolated, showed that seven small populations (< 50 plants) are likely to become extinct within the next 10 years, small populations with < 100 individuals had a quasi-risk extinction probability of 100% in 67 years, and medium populations had quasi-extinction probabilities of 20 and 100% in 100 and 160 years, respectively. Only large populations (> 2,000 individuals) are not at risk of extinction.
Conservation status assessment
Considering the mean distance between localities (392 km) and the results of the Delauney triangulation, the current EOO of H. caput-felis, including all confirmed localities, is 100,682 km2 (Fig. 3), and the AOO is 168 km2 (42 2 × 2 km cells). Since 1999 the extinction of H. caput-felis has been documented in five localities (Table 1). However, the reduction in EOO was negligible (0.007%) because these localities were within the convex polygon or close to the edge. However, AOO decreased by three cells (12 km2), which gave a decline 6.67% since 1999. Such values do not reach the minimum threshold for threatened taxa under Red List criterion A.
Extent of Occurrence (EOO) calculated for H. caput-felis.
Under Red List criterion B the species distribution has ‘severe fragmentation’ (sensu IUCN, 2017). Our quantitative models indicate c. 46% of confirmed localities (eight medium and 11 small populations) are below the viability threshold and therefore could be prone to extinction, close to the 50% threshold for ‘severe fragmentation’. The Red List guidelines recommend integrating this evaluation with a species' biological traits, in particular dispersal ability. A population genetic study in the Alicante (the core of the species range, on the Iberian Peninsula), Melilla (North Africa) and Balearic Islands localities, indicated significant genetic divergence, which would indicate genetic isolation and limited gene flow among these populations (Agulló et al., Reference Agulló, Juan, Guilló, Alonso and Crespo2011), suggesting that localities separated by longer distances (i.e. those in Algeria and Sardinia) may be functionally isolated. Taken together, these findings suggest the species is severely fragmented. Helianthemum caput-felis could therefore be categorized as Endangered based on B2ab(ii, iii, iv, v); i.e. its small AOO (2), highly fragmented distribution (a) and calculated/observed decline (b) in AOO (ii), habitat quality (iii), number of localities (iv) and number of mature plants (v).
Helianthemum caput-felis could also be categorized as Endangered under criterion E as a result of a quasi-extinction risk probability of c. 30% in five generations (79.85 years based on the integral projection model) and c. 45% in three generations (85.32 years based on the matrix population model).
Discussion
Although many countries are contracting parties to international conventions and other international regulations, such as the European Habitat Directive, that encourage monitoring and protection of wild flora, global protection remains insufficient, with c. 13% of known vascular plant species threatened with extinction (Royal Botanic Gardens, 2016). Moreover, the number of global species assessments, as required by the targets of the Global Strategy for Plant Conservation and the CBD, is low, and such assessments particularly challenging when a species occurs in several countries. We believe that our research presented here is the first global assessment of the range and conservation status of H. caput-felis.
The distribution of H. caput-felis is concentrated in the westernmost Mediterranean Basin along the eastern Iberian coasts, reaching its easternmost limit in Su Tingiosu (Italy), northernmost in Capo Mannu (Italy), westernmost in Beni Chiker and southernmost in Beni Sidel (both in Morocco). The status of the species in North Africa remains unclear, however. In three Algerian localities where occurrence was documented by recent herbarium specimens Agulló et al. (Reference Agulló, Juan, Crespo, Alonso and Terrones2017) were unable to confirm presence but did locate the species in a previously undocumented locality. The species' status also requires clarification in Morocco, where we could not confirm presence in one previously documented locality. Considering that this species can grow in locations that may be difficult to explore (e.g. vertical coastal cliffs, sandy and fossil dunes, inland ravines), additional surveys are needed.
Our global assessment of H. caput-felis as Endangered is consistent with previous regional assessments (Agulló et al., Reference Agulló, Juan, Alonso, Crespo, Bañares, Blanca, Güemes, Moreno and Ortiz2010, Reference Agulló, Juan, Crespo, Alonso and Terrones2017; Bilz et al., Reference Bilz, Kell, Maxted and Lansdown2011; Fenu et al., Reference Fenu, Cogoni, Sulis and Bacchetta2015b; Rossi et al., Reference Rossi, Orsenigo, Montagnani, Fenu, Gargano and Peruzzi2016) based only on distribution data (i.e. criterion B). Assessments based on geographical data are the most common for plants, because distribution data are the easiest to obtain. However, our integration of criteria B and E provides a more complete conservation assessment. In addition, the absence of seedling recruitment and the high juvenile mortality rate recorded in several localities were further negative indicators (Sulis, Reference Sulis2016; Sulis et al., Reference Sulis, Bacchetta, Cogoni and Fenu2018). The null or limited recruitment rate in some populations could be partially explained by the effect of drought during 2013–2017 in eastern Spain (García de la Serrana et al., Reference García de la Serrana, Vilagrosa and Alloza2015; Laguna & Ferrer, Reference Laguna and Ferrer2016).
A combination of some recent extinction events and the fact that we were unable to relocate the species at some previously documented locations indicates that the AOO of H. caput-felis is contracting, probably as a result of continuing loss of habitat quality and reproductive individuals. Habitat reduction and degradation appear to be mainly a result of the expansion of infrastructure, a key element in biodiversity loss (e.g. Newbold et al., Reference Newbold, Hudson, Hill, Contu, Lysenko and Senior2015). In particular, as described for several Mediterranean plant species (e.g. Fenu et al., Reference Fenu, Mattana and Bacchetta2011; Ballantyne & Pickering, Reference Ballantyne and Pickering2013; Fois et al., Reference Fois, Bacchetta, Cuena-Lombraña, Cogoni, Pinna, Sulis and Fenu2018; Orsenigo et al., Reference Orsenigo, Montagnani, Fenu, Gargano, Peruzzi and Abeli2018), the main threats to H. caput-felis are related to the degradation of the species' habitat from tourism and recreational activities and the expansion of housing and urban areas. Both the Spanish (Giménez et al., Reference Giménez, Sánchez, Padilla, Marco, Bosque and Rodríguez-Espinosa2008; Agulló et al., Reference Agulló, Juan, Alonso, Crespo, Bañares, Blanca, Güemes, Moreno and Ortiz2010; Marco et al., Reference Marco, Giménez, Padilla and Sánchez2011, Reference Marco, Giménez, Padilla, Sánchez, Gómez Zotano, Arias, Olmedo and Serrano2016; Zaragozí et al., Reference Zaragozí, Giménez, Navarro, Dong and Ramón2012) and Algerian (Agulló et al., Reference Agulló, Juan, Crespo, Alonso and Terrones2017) coasts have experienced extensive urban development linked to tourism. We did not find that climate change, although a major driver of species extinctions (Gómez et al., Reference Gómez, González-Megías, Lorite, Abdelaziz and Perfectti2015; Fenu et al., Reference Fenu, Bacchetta, Giacanelli, Gargano, Montagnani and Orsenigo2017; Orsenigo et al., Reference Orsenigo, Montagnani, Fenu, Gargano, Peruzzi and Abeli2018), is a threat to H. caput-felis but this could be because of the difficulty of detecting its effects (Fenu et al., Reference Fenu, Bacchetta, Giacanelli, Gargano, Montagnani and Orsenigo2017). However, the long-term consequences of climate change, especially drought or irregular rainfall, need to be considered for coastal plants such as H. caput-felis (García de la Serrana et al., Reference García de la Serrana, Vilagrosa and Alloza2015; Laguna & Ferrer, Reference Laguna and Ferrer2016).
Although several populations of H. caput-felis (c. 41% of all localities and c. 62.5% of those in Europe) are currently protected (i.e. within a Site of community importance and/or Plant micro-reserve), our findings nevertheless indicate that H. caput-felis faces a substantial risk of extinction over the short to medium term in the absence of additional management actions. As demonstrated for other species (e.g. Aguilella et al., Reference Aguilella, Fos and Laguna2010; Rossi et al., Reference Rossi, Orsenigo, Montagnani, Fenu, Gargano and Peruzzi2016; Fenu et al., Reference Fenu, Bacchetta, Giacanelli, Gargano, Montagnani and Orsenigo2017), our data indicate that current legal protection and passive conservation measures are insufficient to guarantee the persistence of H. caput-felis. This is contrary to EU legislation, which states that conservation of this species is mandatory and that member states are responsible for its conservation.
Helianthemum caput-felis requires a transnational conservation strategy focusing on protection of each locality in which it occurs, to avoid further decline or extinctions, habitat restoration in degraded localities, mainly in Spain and Algeria, and reduction of the impacts of recreational activities and urban sprawl. Furthermore, as already underway in Melilla (M. Tapia, pers. comm.), translocations could be carried out in suitable areas and reintroductions at sites from which the species has recently disappeared. Previous experiences with threatened Mediterranean coastal plants have demonstrated that these activities would be low-cost projects with good chance of success (e.g. Cogoni et al., Reference Cogoni, Fenu, Concas and Bacchetta2013; Fenu et al., Reference Fenu, Cogoni and Bacchetta2016; Laguna et al., Reference Laguna, Navarro, Pérez-Rovira, Ferrando and Ferrer-Gallego2016). Finally, monitoring is required of all known localities in which the species persists. Although it could be challenging to sustain long-term monitoring, it is fundamental for assessment of conservation status and for effective local conservation.
Acknowledgements
We thank Sergio Sulis, Iván Ramos Torrens and Eva Moragues Botey (Servicio de Protección de Especies de la Consejería de Medio Ambiente, Agricultura y Pesca de las Islas Baleares), Jaime Güemes (Jardín Botánico Valencia), Manuel Tapia Claro (Guelaya-Ecologistas en Acción Melilla) and Jaime X. Soler for their help with fieldwork, and the anonymous reviewers and the Editor for their constructive comments.
Author contributions
Conception, design, drafting text: ES, GF; Data collection: GB, DC, GF; data analysis: ES, DC, GD, GF; finalizing text: all authors.
Conflicts of interest
None.
Ethical standards
This research abided by the Oryx guidelines on ethical standards.
