# Can ecological engineering restore Mediterranean rangeland after intensive cultivation? A large-scale experiment in southern France

Renaud Jaunatre, Elise Buisson, Thierry Dutoit

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Renaud Jaunatre, Elise Buisson, Thierry Dutoit. Can ecological engineering restore Mediterranean rangeland after intensive cultivation? A large-scale experiment in southern France. Ecological Engineering, 2014, 64, pp.202-212. 10.1016/j.ecoleng.2013.12.022. hal-01326838

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# Can ecological engineering restore Mediterranean rangeland after intensive cultivation? A large-scale experiment in southern France

Renaud Jaunatre*, Elise Buisson, Thierry Dutoit

Université d'Avignon et des Pays de Vaucluse, UMR CNRS IRD 7263/237 Institut Méditerranéen de Biodiversité et d'Ecologie, France

## ABSTRACT

Ecological restoration has been identified as one of the possible ways to replace biodiversity loss. While ecological engineering methods can successfully restore some ecosystem attributes, restoration is generally incomplete, mainly due to lack of target species propagule dispersal, unsuitable abiotic conditions and negative biotic interactions, especially after intensive cultivation. How best to restore reference ecosystems is therefore a vital research objective, not only on a small experimental scale but also on a large scale. This study aimed to determine which ecological engineering techniques are the most suitable for large-scale restoration of a low-productive species-rich ecosystem after intensive cultivation. Experiments were carried out at La Crau in southern France, within a 357 ha land rehabilitation project whose aim was to recreate a herbaceous sheep-grazed habitat. We investigated: (i) covering species seeding, (ii) topsoil removal, (iii) hay transfer, and (iv) soil transfer as methods to restore a steppe plant community in this rehabilitated area, using the last remaining French Mediterranean steppe as a reference ecosystem. Species-richness, diversity and composition of vascular plant communities were monitored over three years. The rehabilitation made it possible to recover a large area dominated by grasses but with vegetation different from that of the steppe. Hay transfer was successful in transferring some target species, but the number of target species did not significantly increase by the end of the third year. Covering species seeding seemed to provide a suitable area for target species colonization, but probable competition with grasses needs to be monitored. The greatest similarity in richness, diversity, and composition with the reference ecosystem was obtained with topsoil removal and soil transfer, which made it possible to recover the species-richness and, to some extent, the composition of the steppe. Our findings show that current ecological restoration and ecological engineering techniques can lead to at least partial restoration of some ecosystem attributes. However, they also underline the importance of in situ conservation of natural habitats rather than restoring them after their destruction.

**Keywords:**

Ecological restoration

Plant community composition

Hay transfer

Covering species seeding

Soil transfer

Topsoil removal

## 1. Introduction

In 2011, the Conference of the Parties to the Convention on Biological Diversity (Convention on Biological Diversity, 2011) stated that 15% of European degraded ecosystems should be restored before 2020. Ecosystem restoration is becoming an increasingly common mitigation measure or management tool in environmental conservation. Restoration techniques are therefore an important research goal, not only on a small experimental scale but also on a large scale involving the application of ecological engineering techniques (Jones, 2012; Mitsch, 2012). Large-scale experiments have already been conducted in some ecosystems (e.g., Donath et al., 2007 in flood meadows; Shaish et al., 2008 in coral reefs);

however, there is clearly a need to implement such experiments in Mediterranean rangelands, which today still cover more than 3.7 M ha (Dehesas and Montado ecosystems) in the Iberian peninsula and more than 63 M ha in North Africa (Le Houérou, 1995).

The abandonment of intensively cultivated areas provides ideal locations for large-scale experiments seeking to restore semi-natural ecosystems with extensive land-use, integrated into current agricultural landscapes (Cramer et al., 2007; Buisson et al., 2009). Ecosystems which have evolved with long-term severe environmental constraints, either biotic (e.g., grazing) or abiotic (e.g., dryness), and which have subsequently been cultivated, often exceed biotic and abiotic thresholds (Whisenant, 1999), because this type of disturbance is long-lasting and/or large-scale. Even though spontaneous succession may offer many advantages for ecological restoration (Prach and Hobbs, 2008; Jirova et al., 2012), it can lead to ecosystems totally different from the chosen reference ecosystem (Manchester et al., 1999; Prach and Pyšek, 2001; Török et al., 2011), especially when this reference ecosystem

<footnote>* Corresponding author at: IUT, Site Agroparc, BP 61207 84911 Avignon Cedex 9
– France. Tel.: +33 4 90 84 38 58; fax: +33 4 90 84 38 07.
E-mail address: renaud.jaunatre@yahoo.fr (R. Jaunatre).</footnote>
was created and maintained by former human land uses (i.e., cultural ecosystems (Clewell and Aronson, 2007)). This is especially true when soil has been nutrient-enriched, leading to increased competition (Marrs, 2002), and when target species propagules are no longer available, either because the seed bank has been depleted (Hutchings and Booth, 1996; Bossuyt and Honnay, 2008) or because of dispersal limitation (Bakker et al., 1996; Münzbergová and Herben, 2005). These disturbed ecosystems may need active restoration (Cramer et al., 2008; Prach and Hobbs, 2008; Hölzel et al., 2012), which must focus on lowering non-target species abundance and on improving target species dispersal (Walker et al., 2004; Baer et al., 2008; Dickson and Busby, 2009; Kiehl et al., 2010). Numerous ecological engineering techniques have been developed for these purposes (see Fagan et al., 2008; Kiehl et al., 2010; Török et al., 2011 for reviews).

Apart from fire and overgrazing disturbances (D'Antonio et al., 2003), Mediterranean ecosystem restoration issues have been poorly addressed. Yet these systems are particularly threatened by anthropogenic disturbances (Underwood et al., 2009). In the present study on Mediterranean rangeland areas, we assessed the efficiency of four ecological engineering techniques, as well as rehabilitation alone, applied on a large scale (3–60 ha) with the aim of restoring a Mediterranean species-rich steppe community. Experiments were carried out within a 357 ha rehabilitation project whose objective was to recreate a herbaceous sheep-grazed habitat for steppe birds. We tested the following ecological engineering techniques on the rehabilitated area: (i) covering species seeding, (ii) topsoil removal, (iii) hay transfer, and (iv) soil transfer, aiming to restore a Mediterranean rangeland plant community, with the last remaining French Mediterranean steppe as a reference ecosystem. These four ecological engineering techniques, applied for the first time on a large-scale Mediterranean herbaceous ecosystem, were monitored over three years.

One way of lowering non-target species abundance is by changing the facilitation-competition balance (i.e., increasing or decreasing the positive or negative effect of plants on other plants) by introducing new species (Gómez-Aparicio, 2009), and this has led to the use of covering species. For example, the species-richness of former arable land was increased by seeding hemiparasitic species of the *Rhinanthus* genus to reduce dominant species density (Davies et al., 1997; Pywell et al., 2007). Seeding covering species can use up nutrients without the species being competitive for several years after restoration, depending on the environmental conditions of the site to be restored. Once established, the sown mixture can drive succession from an inhibition to a tolerance model of succession (Connell and Slatyer, 1977). Rapid cover and nutrient consumption can inhibit arable weed species density in the first years via priority effects (i.e., a competitive advantage for the first-arriving species) (Ross and Harper, 1972; Kardol et al., 2013). We therefore tested the effects of seeding covering species to rapidly inhibit the dense cover of non-target species.

Another approach consists in decreasing nutrients and removing the permanent seed bank of weeds (Davy, 2002) through topsoil removal. In former agricultural areas, this has been shown to favor low-productive plant communities (Verhagen et al., 2001). Here, we tested topsoil removal on a large scale in a Mediterranean environment for the first time.

Dispersion has been identified as a major limiting factor for spontaneous succession of target communities (Hutchings and Booth, 1996; Bischoff, 2002). Strengthening dispersion has thus become a major topic in restoration ecology (Kiehl et al., 2010; Hölzel et al., 2012). As commercial regional seed mixtures (Jongepierová et al., 2007) are rarely available for species-rich communities, the reintroduction of propagules gathered from the reference ecosystem can be a very efficient solution (Kiehl et al.,

2010). Here, we tested the transfer of both hay material and soil material. Hay transfer is a well-known technique, widely used in northern European restoration experiments (Hölzel and Otte, 2003; Rasran et al., 2006; Kiehl et al., 2010). However, it has never before been tested on a large scale in drier ecosystems. Hay is used instead of commercial seed mixtures, which can be very expensive and can lead to genetic pollution if the origin of seeds has not been verified (Bischoff et al., 2010; Lencová and Prach, 2011). We vacuum-harvested seeds that had already fallen on the ground, this seed-harvesting method having proved successful in north-western Europe (Stevenson et al., 1997; Riley et al., 2004) as well as in small-scale experiments in Mediterranean plant communities (Coiffait-Gombault et al., 2011).

Soil transfer can be used to transfer propagules either via intact turves, fragmented turves or bulk soil, and has already proved successful in recreating species-rich plant communities (Pywell et al., 1995; Bullock, 1998; Vécrin and Muller, 2003). Bulk soil is cheaper and easier to gather, and restoration of species richness and composition is as successful as with whole-turf transfer (Good et al., 1999; Sparks et al., 2013). Although this technique involves destroying an area, it should be a very efficient way of transferring not only seeds but also propagules and associated microorganisms. It should also lower nutrient content, since soil from the reference ecosystem is mixed with soil from the degraded site. This can be considered an ecological engineering approach only inasmuch as it saves a soil resource already scheduled to be destroyed, using it for restoration purposes.

These four ecological engineering techniques, which can modify one or several filters (dispersion, abiotic, or biotic), were expected to be suitable for restoring a Mediterranean ecosystem degraded after intensive cultivation.

The objectives of the present study were to assess the possibility of restoring large areas of low-productive species-rich ecosystem following intensive cultivation and to determine which ecological engineering techniques provide the best restoration results in the short term. In addition to the assessment of suitability, large-scale testing of these ecological engineering methods allowed us to address the question of whether plant community ecology theories can be applied to ecological restoration (e.g., importance of filters in plant community dynamics (Whisenant, 1999)). Two main barriers to the spontaneous recolonization of target plants have been identified: dispersal limitation of target species and high dispersal and establishment potential of non-target species, in particular due to increased fertility in the former cultivation area (Buisson, 2005; Buisson et al., 2006). We therefore hypothesized that rehabilitation involving neither target species dispersal strengthening nor abiotic condition restoration would lead to a rather species-poor community, different from the reference community. Topsoil removal and covering species seeding were expected to decrease non-target species abundance by changing abiotic conditions and biotic interactions, respectively. However, these methods were not expected to lead to a rapid increase in target species due to the lack of dispersal strengthening, contrary to hay transfer, and soil transfer. Soil transfer, which is supposed to both partially restore abiotic conditions and strengthen dispersal of target species propagules, was expected to lead to a rapid increase in target species, with results closest to the reference plant community composition.

## 2. Materials and methods

### 2.1. Study site

La Crau area, the last remaining French Mediterranean steppe, has been shaped by thousands of years of interactions among soil,
<figure><img src="image_2.png" /><figcaption>**Fig. 1.** Location of study site and experimental design. (A) Location of the La Crau area in France and location of the remnant patches of steppe (light grey), rehabilitated orchard (RO, black), hay donor site (HDS; dark grey) and soil donor site (SDS; dark grey); (B) experimental design of restoration treatment on former Cossure orchard (C: control; CSS: covering species seeding; T: topsoil removal; REH: rehabilitated area; H: hay transfer; and S: soil transfer) (Jaunatre et al., 2012).</figcaption></figure>

climatic conditions, and sheep grazing (Fig. 1A). The 40 cm deep soil is composed of 50% silicaceous stones and lies on a calcareous conglomerate which cannot be penetrated by plant roots (Devaux et al., 1983). The climate is Mediterranean, featuring high interan- nual variability, with an average of 540 mm yearly precipitation, mainly in spring and autumn and 110 days per year with winds above 50 km h⁻¹ (Devaux et al., 1983). Traditional extensive sheep grazing has taken place in the La Crau area for more than 2000 years (Badan et al., 1995; Henry et al., 2010), as in many typical Mediter- ranean rangelands (Le Houérou, 1995). This xeric steppe, located in southern France, is a unique species-rich plant community com- posed mainly of annuals and dominated by Brachypodium retusum Pers. and Thymus vulgaris L. Although the steppe of La Crau is a unique ecosystem, experiments of this kind can provide insights into restoration technique efficiency applicable to other relatively dry Mediterranean rangeland ecosystems, like Dehesas in Spain, Montado in Portugal and the steppes of north Africa (Le Houérou, 1995).

## 2.2. Restoration goals within the rehabilitation project

All the ecological engineering techniques tested were applied within the framework of a larger rehabilitation project: a 357 ha orchard located approximately in the center of the steppe area (Fig. 1A) and adjacent to the largest remnant patch of steppe (6500 ha), was abandoned in 2006. Around the dead, or almost dead, fruit trees, the vegetation was mainly dominated by grasses such as Avena barbata Link and Bromus madritensis L., with some forbs such as Galium aparine L., Crepis foetida L., or Lactuca seriola L. The overall rehabilitation project began in 2009 with the aim of re-creating a herbaceous steppe-like habitat for steppe birds. Veg- etation and soil characteristics were studied before rehabilitation, the whole area being found to be homogenous (multivariate ordi- nation results not shown). Rehabilitation consisted in: (i) cutting down and exporting trees from the abandoned orchard (200,000 peach trees and 100,000 poplars used for hedgerows, (ii) level- ing soils, and (iii) reintroducing traditional sheep grazing in spring 2010 (Jaunatre et al., 2012). This rehabilitation was applied over the

whole area, which thus can be considered spatially homogenous regarding soil characteristics and potential vegetation. The current study focuses on four additional ecological engineering techniques applied to the rehabilitated area: covering species seeding, top- soil removal, hay transfer, and soil transfer (Fig. 2). The short-term objective of this restoration experiment is to limit the coloniza- tion of unwanted plant species and to improve the establishment of characteristic species just after the end of the rehabilitation phase. The mid-term objective is to re-direct the plant community towards the desired successional pathway so as to recreate a plant community with steppe-like characteristics of species-richness, composition, and structure, over a much longer term (>30 years).

## 2.3. Ecological engineering techniques

Four techniques were tested: (i) covering species seeding, (ii) topsoil removal, (iii) hay transfer, and (iv) soil transfer.(i) The three covering species (*Lolium perenne* L., *Festuca arundinacea* Schreb., and *Onobrychis sativa* Lam) were chosen for their palatability, their commercial availability, their ability to rapidly cover bare ground, as well as their low competitive ability in Mediterranean envi- ronmental conditions and their current presence in the artificial meadows of La Crau area (Devaux et al., 1983). They were sown on 60 ha (Fig. 1B). (ii) Topsoil removal consisted in removing the nutrient-rich upper soil layer, down to a depth of 20 cm over a 3250 m² area (Fig. 1B). (iii) The hay material used for hay trans- fer was gathered by a leaf-vacuum truck in summer at a donor site located less than 5 km away from the restoration site (Fig. 1A). The material was then spread at a 1:3 ratio (spreading seeds gath- ered from 1 ha over 3 ha) over two 10 ha areas in the rehabilitated orchard (Fig. 1B). (iv) The material for soil transfer was the 20 cm upper soil layer of a 1 ha steppe patch previously scheduled for destruction to allow for building, located less than 2 km away from the restoration site (Fig. 1A). Gathering of bulk soil, transport, and spreading at 1:3 ratio over three 1 ha areas (Fig. 1B) were carried out within one day in early September a few hours before the first significant autumn rains. Hay and soil donor sites were connected by sheep grazing from Neolithic times until the establishment of
<figure><img src="image_3.png" /><figcaption>**Fig. 2.** Schematic representation of terms used in the manuscript. Dark grey represents the most degraded states: abandoned orchard which is not studied in the present paper, control without dead trees and rehabilitated after soil leveling; light grey represents the four tested restoration techniques: covering species seeding, topsoil removal, hay transfer, and soil transfer; white represents the steppe, which is the reference ecosystem.</figcaption></figure>

cultivation in the La Crau area in the 1970s (Fabre, 1997). As control, we used a 2 ha area (Fig. 1B) where trees had been removed for safety purposes but where soils had not been leveled (Fig. 2) (Jaunatre et al., 2011).

Because of the constraints imposed by the multiple stakeholders of the project (i.e., costs reduction, homogeneous areas, etc.), we were obliged to apply each treatment to a few large areas rather than using the classic scientific experimental design of many small areas. However, the only difference between multiple sampling from several large treatment areas and single sampling from numerous small treatment areas is that when small areas are treated, the areas between samples are not included in the treatment. Designing a restoration project which leaves areas untreated would appear counter-productive, making it difficult to obtain approval within a large-scale restoration project (Jaunatre et al., 2011).

## 2.4. Soil seed bank and seed sources

Germination potential was assessed from five types of sample: four types of soil seed bank and one hay seed bank. The four types of soil seed bank were collected from the 0–10 cm layer from control, rehabilitated area, topsoil removal, and soil donor sites. Each seed bank was estimated according to the concentrated seedling emergence method (Ter Heerdt et al., 1996) using ten 2l soil samples. Ten hay seed bank samples were spread on the same substrate as the soil seed bank samples: a 1:4 compost:vermiculite mix. All the samples were randomly placed in a greenhouse and germinations were counted, identified, and removed every week for 3 months. These data were used to estimate germinated seed species-richness and the numbers of germinated seeds from target and non-target species. Target species were species found in the reference steppe communities (Molinier and Tallon, 1950). Species names for identification from the seed bank or above-ground vegetation are according to the data base of metropolitan French tracheophytes (Tela Botanica, 2012).

## 2.5. Vegetation survey

On the steppe, control, rehabilitated area, covering species seeding, hay transfer, and soil transfer sites, 18 2 × 2 m quadrats were surveyed. For topsoil removal, which covered too small an area for such an extensive survey, 9 2 × 2 m quadrats were surveyed. Quadrats were all placed at least 20 m from the edge of the area under treatment. Every May from the first (2010) to the third (2012) year after treatment, on each quadrat, plant species were identified, a Braun-Blanquet abundance-dominance coefficient was determined for each recorded plant species (Braun-Blanquet et al., 1952) and vegetation height and vegetation cover were measured.

## 2.6. Soil analyses

Analyses were carried out on 30 samples of soil from the abandoned orchard before rehabilitation (in 2008), five samples from the soil donor site (in 2009) and six samples from each of the following sites: control, rehabilitated area, covering species seeding, topsoil removal, and soil transfer (in 2012). For each sample, three 70 g subsamples of soil were randomly gathered from a 35 m² area before being pooled and sieved with 2 mm mesh sieve for analyses carried out by INRA (Institut National de la Recherche Agronomique). Nutrient analysis (organic C (loss-on-ignition), total N, C:N, Olsen total P (Olsen et al., 1954) and exchangeable cations Ca and K) and water pH were measured on these samples (Baize, 2000).

## 2.7. Data analysis

As data on species-richness of germinated seeds and number of germinated seeds from target or non-target species did not conform to parametric conditions (model residuals were not normal), treatments were compared via non-parametric tests: a Kruskal–Wallis test, followed by a pairwise Wilcoxon comparison with a *p*-value adjustment according to Benjamini–Hochberg's method if a significant difference was found (Benjamini and Hochberg, 1995).

Soil characteristics were ordinated by a principal component analysis (PCA), a suitable method for continuous variables (Borcard et al., 2011), and vegetation composition characteristics were ordinated by a non-metric multidimensional scaling (NMDS) based on Bray–Curtis dissimilarity, a suitable method for contingency tables with many double zeros (Borcard et al., 2011). In order to assess plant community restoration success, we used the normalized Community Structure Integrity Index (CSII<sub>norm</sub>) and Higher Abundance Index (HAI). The CSII<sub>norm</sub> measures the proportion of the species abundance of the steppe community present in the restored community, and HAI measures the proportion by which the species abundance in the restored community exceeds that of the steppe community (Jaunatre et al., 2013).

Parametric conditions were not reached for the testing of interactions between years and treatments, so we did not test this interaction (model residuals were not normal). However, as data for above-ground vegetation (species-richness, Shannon index, Shannon evenness (Pielou, 1969), height, vegetation cover, CSII<sub>norm</sub>, and HAI) met parametric conditions with respect to the effect of year or treatment, ANOVA, and Tukey honest significant differences post-hoc tests were performed to compare means between treatments in the third year or within a treatment between years Table 1.

All the analyses were conducted using R 2.13.0 (R Development Core Team, 2011), univariate analyses with the “stats” package and
<caption>**Table 1**</caption>

Plant community characteristics for each treatment after 3 years (2012): control, rehabilitated area (rehab.), covering species seeding (CSS), topsoil removal (topsoil), hay transfer (hay T.), soil transfer (soil T.) and steppe. The given values are means±standard errors, df, *F*, and *p* correspond to the degree of freedom, the *F* value and *p* value resulting from ANOVAs testing for the effect of treatment on each variable. Within a row, two cases with a common letter have significantly different values according to Tukey honest significant differences post hoc tests.

<table><thead><tr><th></th><th>ANOVA</th><th>Control</th><th>Rehab.</th><th>CSS</th><th>Topsoil</th><th>Hay T.</th><th>Soil T.</th><th>Steppe</th></tr></thead><tbody><tr><td rowspan="3">Species-richness (4 m²)</td><td>df=6</td><td>11.78±1.09</td><td>15.61±1.41</td><td>9.61±0.81</td><td>27.11±1.33</td><td>13.91±0.87</td><td>28.72±1.8</td><td>32.94±1.33</td></tr><tr><td><i>F</i>=51.43</td><td>ab</td><td>b</td><td>a</td><td>c</td><td>ab</td><td>c</td><td>c</td></tr><tr><td><i>p</i> &lt; 0.001</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="3">Shannon index</td><td>df=6</td><td>1.03±0.04</td><td>1.19±0.06</td><td>0.98±0.03</td><td>1.56±0.02</td><td>1.17±0.04</td><td>1.54±0.04</td><td>1.68±0.02</td></tr><tr><td><i>F</i>=44.68</td><td>ab</td><td>ab</td><td>a</td><td>c</td><td>b</td><td>cd</td><td>d</td></tr><tr><td><i>p</i> &lt; 0.001</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="3">Shannon evenness</td><td>df=6</td><td>0.425±0.005</td><td>0.444±0.007</td><td>0.441±0.004</td><td>0.474±0.003</td><td>0.451±0.004</td><td>0.464±0.004</td><td>0.483±0.002</td></tr><tr><td><i>F</i>=16.02</td><td>ab</td><td>ab</td><td>a</td><td>c</td><td>b</td><td>cd</td><td>d</td></tr><tr><td><i>p</i> &lt; 0.001</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="3">Average height (cm)</td><td>df=6</td><td>15±1.28</td><td>9.33±0.53</td><td>9.56±0.65</td><td>10.67±0.82</td><td>9.91±0.57</td><td>10.94±0.81</td><td>6.22±0.68</td></tr><tr><td><i>F</i>=10.17</td><td>a</td><td>b</td><td>b</td><td>b</td><td>b</td><td>b</td><td>c</td></tr><tr><td><i>p</i> &lt; 0.001</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="3">Vegetation cover</td><td>df=6</td><td>93.11±1.78</td><td>68.33±4.39</td><td>66.94±2.83</td><td>47.56±8.63</td><td>73.64±2.61</td><td>74.17±2.22</td><td>58.61±2.77</td></tr><tr><td><i>F</i>=10.31</td><td>a</td><td>bc</td><td>bc</td><td>d</td><td>b</td><td>b</td><td>cd</td></tr><tr><td><i>p</i> &lt; 0.001</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="3">Normalized Community Structure Integrity Index</td><td>df=6</td><td>0.034±0.005</td><td>0.111±0.014</td><td>0.039±0.006</td><td>0.257±0.019</td><td>0.099±0.012</td><td>0.442±0.042</td><td>1.00±0.036</td></tr><tr><td><i>F</i>=197.24</td><td>a</td><td>b</td><td>a</td><td>b</td><td>c</td><td>d</td><td>e</td></tr><tr><td><i>p</i> &lt; 0.001</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="3">Higher Abundance Index</td><td>df=6</td><td>0.946±0.007</td><td>0.872±0.012</td><td>0.935±0.008</td><td>0.783±0.008</td><td>0.872±0.012</td><td>0.674±0.019</td><td>0.334±0.009</td></tr><tr><td><i>F</i>=314.78</td><td>a</td><td>b</td><td>a</td><td>c</td><td>b</td><td>d</td><td>e</td></tr><tr><td><i>p</i> &lt; 0.001</td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr></tbody></table>

multivariate analyses with the “ade4” (Chessel et al., 2004; Dray and Dufour, 2007; Dray et al., 2007) and “vegan” (Oksanen et al., 2008) packages.

# 3. Results

## 3.1. Seed germination potential

The species richness of the seed bank significantly decreased from control to rehabilitated area to topsoil removal (Fig. 3A). Average species-richness was below 8 for all types, and maximum species-richness was 15 in one control sample. Seed bank species-richness from the soil donor site was similar to that of the rehabilitated area and seed bank species-richness of the hay material was similar to that of control. The number of germinated seeds varied considerably, both for target and non-target species (Fig. 3B-C). The control site showed the highest number of non-target germinated seeds, with a mean of 177.7 ± 77.3 germinated seeds per sample, mainly *Chenopodium album* L. and *Cardamine hirsuta* L. However, control showed a very low mean number of target germinated seeds: 1.2 ± 0.5 seeds (*Linaria arvensis* (L.) Desf. or *Trifolium campestre* Schreber). The rehabilitated area showed a slightly higher, but not significantly different, number of non-target germinated seeds than the topsoil removal area, in both cases these germinated seeds being the same species as in the control. None of the target species germinated in either the rehabilitated area or the topsoil removal treatment. The soil donor and hay material sites contained a relatively low number of non-target germinated seeds, mainly *Anagallis* sp. for the soil donor site and *Senecio vulgaris* L. for the hay material site. However, a significant number of target species seeds germinated in the soil donor (19.5 ± 5.9 germinated seeds) and hay material (34.4 ± 7.1 germinated seeds) sites, mainly *P. bulbosa* and *Brachypodium distachyon* (L.) P. Beauv. for the soil donor site and *B. distachyon* and *Plantago lagopus* L. for the hay material site.

## 3.2. Effects of rehabilitation and restoration on soil properties

The ordination based on soil properties showed a nutrient content gradient on the first axis (62.4%) (Fig. 4). Samples from the

<figure><img src="image_4.png" /><figcaption>**Fig. 3.** Species-richness of germinated seeds (A) and number of germinated seeds of non-target (B) and target (C) species from the seed bank. Assessed from control (dark grey), rehabilitated area (dark grey) and topsoil removal (light grey) as well as from the seed bank, soil donor site (white) and from hay material (white). Kruskal-Wallis chi-squared test and its corresponding *p*-value are shown above each graphic. Bars showing common letters do not have any significant differences according to pairwise multiple comparisons with Benjamini-Hochberg *p* adjustment (Benjamini and Hochberg, 1995).</figcaption></figure>
<figure><img src="image_5.png" /><figcaption>**Fig. 4.** PCA ordination based on analysis of 71 soil samples. Samples from each treatment (2012) are identified by solid lines: Steppe in white, Soil transfer, Topsoil removal, and Covering species seeding in light grey and rehabilitated area and control in dark grey. Both samples from the abandoned orchard before rehabilitation (2008) and samples from the soil donor site (2009) are identified by dashed lines. Arrows represent soil variables (C:N: Carbon:Nitrogen ratio; C: total carbon; N: total nitrogen; K₂O: potassium, P₂O₅: Olsen phosphorus (Olsen et al., 1954); CaO: Calcium oxide and pH).</figcaption></figure>

control in 2012 and from the abandoned orchard before rehabilita-
tion in 2008 were in the same range of relatively high nutrient
content (carbon, nitrogen, potassium, and phosphorus). At the
opposite end were the steppe, soil donor site, and topsoil removal
site. The rehabilitated area, covering species seeding site, and soil
transfer site were in an intermediate position, ordered along this
gradient from highest nutrient content (rehabilitated area) to low-
est nutrient content (soil transfer). The ordination showed a pH
and C:N gradient on the second axis (18.1%), with lower pH and
higher C:N in the steppe samples and higher pH and lower C:N in
the topsoil removal samples (Appendix 1).

3.3. *Effects of rehabilitation and restoration on plant community characteristics*

In the control, rehabilitated area, and hay transfer site, species-
richness declined from the first to the second year and recovered
in the third year (Fig. 5A). In the covering species seeding and
soil transfer sites, species-richness did not show any significant
differences over the years. The topsoil removal site showed a sig-
nificant increase in species-richness from the first to the third
year, whereas the steppe showed a significant decrease during
these years. In the third year, the highest species-richness was
found in the steppe (32.94 ± 1.33) but was not significantly differ-
ent from the species-richness in the topsoil removal (27.11 ± 1.33)
or soil transfer (28.72 ± 1.8) treatments. The rehabilitated area and
the hay transfer treatment showed intermediate species-richness
(around 15), whereas the covering species seeding treatment and
the control showed the lowest species-richness (around 10).

Height followed the same dynamics in all treatments over the three years, except for the steppe, which was stable (Fig. 5B): it increased from the first to the second year, reaching mean values above 20 cm (with a maximum of 48 cm in the control), then significantly decreased in the third year. In the third year, the greatest height was in the control (15 ± 1.28 cm) and the lowest in the steppe (6.22 ± 0.68 cm), the other treatments showing heights of around 10 cm. Vegetation cover tended to decrease from the first to the third year, except in the topsoil removal area (Fig. 5C). In the third year, vegetation cover was highest in control (93.11 ± 1.78%) and lowest in topsoil removal (47.56 ± 8.63%) and

the steppe (58.61 ± 2.77%), showing intermediate values of around
70% in the other treatments.

3.4. Effects of rehabilitation and restoration on plant community composition

The NMDS ordination showed a gradient in plant community
composition from the steppe to the rehabilitated area, control
and covering species seeding communities (Fig. 6A). The control
community was very similar to the rehabilitated area commu-
nity: dominated by grasses and thistles such as Bromus diandrus
Roth, B. madritensis, A. barbata, Carduus pycnocephalus L., or Galac-
tites elegans (All.) Soldano. The trajectories of their barycenter (i.e.,
the mean coordinates of all the communities) showed very slow
dynamics from the first year to the third year (Fig. 6B). Cover-
ing species seeding was well discriminated on the ordination plot,
characterized by the sown species: L. perenne, F. Arundinacea, and
O. sativa. The trajectory showed a marked shift from the first year
to the second year, characterized by a decrease in O. sativa, Alopecu-
rus muysiroides Hudson, and C. album. The hay transfer community
was confounded with the rehabilitated area community, but on
the steppe side of the gradient, characterized by more target forbs,
such as P. lagopus, Sixalix atropurpurea (L.) Greuter et Burdet, Sher-
ardia arvensis L. The trajectory of its barycenter was like that of
the rehabilitated area: very small changes occurred in composition
and abundance of species from the first to the third year. The top-
soil removal community showed the strongest dynamics (Fig. 6B),
going from a community dominated by ruderals, such as Senecio
vulgaris L. or Polycarpon tetraphyllum (L.) L. or Polygonum aviculare
L. to a community very close to the soil transfer community. In the
third year, these two communities were the closest to the steppe
community and were characterized by numerous target species
such as Plantago bellardii All., Poa bulbosa L., B. distachyon, Eryn-
gium campestre L., or Logfia gallica (L.) Coss. & Germ. The steppe
was still well discriminated on the ordination by its characteristic
species, such as B. retusum, Stipa capillata L., T. Vulgaris, or Aspho-
delus ayardii Jahand. & Maire., and showed almost no changes in
plant composition and abundances from the first to the third year
(Fig. 6B).
<figure><img src="image_6.png" /><figcaption>**Fig. 5.** Means and standard errors of species-richness (4 m²) (A), height (cm) (B) and vegetation cover (%) (C) for the first three years (1st: 2010; 2nd: 2011; 3rd: 2012) for each treatment: steppe (white), for restoration techniques (Soil transfer (Soil T.), Hay transfer (Hay T.), topsoil removal (topsoil) and covering species seeding (covering); light grey), for rehabilitated area (rehab.; dark grey) and control (dark grey). The *F* and *p* value of ANOVA performed within each treatment to compare years are shown above the bars. Within a treatment, bars showing common letters do not have any significant differences according to Tukey honest significant differences post-hoc tests.</figcaption></figure>

CSII<sub>norm</sub> was not significantly different from one year to another in the control and the rehabilitated area, remaining relatively low (below 0.15) (Fig. 7A). The covering species seeding site showed a slight but significant increase from the first to the second year and the hay transfer site from the second to the third year, but both remained below 0.15 as well. At the topsoil removal site, CSII<sub>norm</sub> increased throughout the three years, reaching 0.25 ± 0.02 in the third year. Soil transfer was the restoration treatment with the highest CSII<sub>norm</sub> from the first to the third year (0.44 ± 0.04), nevertheless remaining significantly different from the steppe (1.0 ± 0.03). HAI was relatively high in the control and the rehabilitated area (above 0.8) and did not change over the years (Fig. 7B). In covering species seeding, topsoil removal and hay transfer, HAI decreased, reaching its lowest value in the third year. In soil

transfer, HAI increased after the first year and decreased from the second to the third year. The highest HAI in the third year was found in the control and covering species seeding (above 0.93); the rehabilitated area and hay transfer had intermediate values (around 0.87), while topsoil removal (0.783 ± 0.008) and soil transfer (0.674 ± 0.019) had the lowest values, although still higher than the steppe (0.334 ± 0.009).

## 4. Discussion

Spontaneous succession potential can be estimated by seed bank studies (Willems and Bik, 1998; Cui et al., 2013). Our results confirm that the seed banks of previously cultivated areas rarely contain seeds from the target community, but rather from ruderal species (Hutchings and Booth, 1996; Buisson et al., 2006). Via priority effects (Ross and Harper, 1972; Kardol et al., 2013), this non-target seed bank can form dense cover that may affect the establishment of recolonizing target species (Baer et al., 2008; Standish et al., 2008). This pattern was observed on the rehabilitated area, where grasses such as *A. barbata* or *Bromus* spp. dominated right from the first year (Jaunatre et al., 2012). Moreover, the soil nutrient content here is still affected by the cultivation history. Nitrogen is very labile and decreased to values similar to those of the steppe in three years, while potassium and phosphorus remained significantly higher. After fertilization cessation, the long-term effect persists on plant communities (Smits et al., 2008; Královec et al., 2009), even when soil parameters have recovered to pre-fertilization values (Semelová et al., 2008). These environmental conditions resulted here in high vegetation height and cover, which may have had a negative effect on germination and growth of less competitive light-demanding target species (Hautier et al., 2009). As spontaneous succession did not provide successful results with respect to the reference plant community restoration objective, more active restoration needs to be considered in order to accelerate or permit its recovery (Cramer et al., 2008; Prach and Hobbs, 2008; Hölzel et al., 2012).

By only changing biotic interactions, we expected to decrease non-target species abundances without a rapid increase in target species. The results obtained with covering species seeding confirm this hypothesis: non-target species that dominated in the rehabilitated area, such as *A. barbata* and *Bromus* spp., showed very low abundance in this treatment. However, covering species seeding had the lowest species-richness and its composition differed most from that of the steppe. Two of the sown species (non-target) dominated: *F. arundinacea* and *L. perenne*. Despite this dominance after three years, the relatively low vegetation height and cover created by this technique could provide safe sites for further colonization by target species, especially if the sown grasses decrease under certain environmental conditions (Mitchley et al., 1996). This pattern justifies considering these covering species as nurse species. Long-term success would depend: (i) on a future decrease in abundance of the sown species (thus creating safe sites) and (ii) on non-saturation of the safe sites by non-target species. Without competitively totally excluding the non-target grasses, sown covering species can provide sufficient competition to prevent them from dominating and to allow safe sites to be established by target species (Davies et al., 1997; Pywell et al., 2007). These hypothetical interactions need to be monitored and confirmed both by models and in the field, but could provide an interesting and innovative way to modify spontaneous succession in ecosystems where the dynamics is interrupted by early succession competitive species (Connell and Slatyer, 1977).

It was hypothesized that topsoil removal would decrease non-target species by recreating nutrient-poor abiotic conditions and by removing the seed bank. This hypothesis is confirmed by
<figure><img src="image_7.png" /><figcaption>**Fig. 6.** NMDS ordination based on 324 2 × 2 m vegetation samples and the 149 species present in at least 3 samples. The A plot shows only the 2012 samples (open circles) grouped by treatments: steppe in white, soil transfer, hay transfer, topsoil removal and covering species seeding in light grey and rehabilitated area and control in dark grey. For clarity, only the 53 most correlated species are shown. The B plot shows the succession of vegetation for each treatment, according to the position of their barycenter: dashed lines represent the succession from the first year (1; 2010) to the second year (2; 2011), full arrows represent the succession from the second year to the third year (3; 2012).</figcaption></figure>

our results: the number of non-target species in the seed bank decreased to that of the steppe seed bank. Moreover, nutrient content was significantly lowered, as expected from results in other ecosystems (Verhagen et al., 2001; Klimkowska et al., 2010). As in Allison and Ausden's experiment (2004), pH and calcium content were increased by topsoil removal. This can be explained by the crumbling of the calcareous conglomerate top and/or the breaking of some calcareous stones during topsoil removal. Like Patzelt et al. (2001), in the early stage of restoration we found low species-richness and ruderal species on the areas where topsoil was removed (Jaunatre et al., 2012). In the following years, species-richness and target vegetation composition recovered as nutrient availability decreased (Temperton et al., 2012). This resilience was probably enhanced by the almost total absence of competition with a dense cover of grasses (Allison and Ausden, 2004; Kiehl et al., 2006). Nevertheless, while a suitable habitat for the characteristic species of the reference steppe seems to have been restored by this method; the slow increase in species-richness between 2010 and 2013 shows that dispersal limitation is also an active filter limiting the natural colonization of topsoil removal sites.

It was hypothesized that the number of target species would increase with methods enhancing their dispersion: hay and soil

transfer. In fact, the two methods yielded contrasting results. In the hay transfer area, the first year only some target forbs, such as *P. lagopus*, *S. arvensis*, or *S. atropurpurea* occurred (Jaunatre et al., 2012). After three years, the composition was unchanged: species-richness was still no higher than in the rehabilitated area and the composition was still very close to that of the rehabilitated area. These results differ from those obtained in other hay transfer experiments, where richness of target species increased significantly during the first years (Hölzel and Otte, 2003; Edwards et al., 2007; Coiffait-Gombault et al., 2011). Possible explanations for the poor success of this hay transfer include: (i) only one hay harvesting date, a problem which can be solved by multiple harvesting to maximize the number of target species, (ii) low quantities of hay, although the same amount was used as in other experiments (Kiehl et al., 2010), and (iii) unsuitable abiotic conditions. This last point seems to be the main explanation, as the same hay spread in another experiment after topsoil removal showed a higher number of target species (Jaunatre et al., 2013). A seed mixture containing many species with low competitive ability should be applied on appropriate substrate; for example, topsoil removal prior to hay transfer has shown successful results (Patzelt et al., 2001; Hölzel and Otte, 2003; Klimkowska et al., 2010). A lower-cost solution may be to wait for
<figure><img src="image_8.png" /><figcaption>**Fig. 7.** Means and standard errors of normalized Community Structure Integrity Index (A) and Higher Abundance Index (B) for the first three years (1st: 2010; 2nd: 2011; 3rd: 2012) for each treatment: steppe (white), restoration techniques (Soil transfer (Soil T.), hay transfer (Hay T.), topsoil removal (topsoil) and covering species seeding (covering) (light grey), rehabilitated area (rehab.) (dark grey) and control (dark grey). The *F* and *p* value of ANOVA performed within each treatment to compare years are shown above the bars. Within a treatment, bars showing common letters do not have any significant differences according to Tukey honest significant differences post-hoc tests.</figcaption></figure>

a decrease in the high nutrient content due to cultivation legacies, as well as a decrease in vegetation cover and height (Bartha et al., 2003) due to grazing, before transferring hay. However, staggering restoration implementation over several years in this way may be difficult to accept by managers of large-scale projects.

In contrast, soil transfer partly confirmed our hypothesis. Soil nutrient content was not affected by soil transfer, contrary to what sometimes occurs in habitat or turf translocation (Trueman et al., 2007). Only pH and calcium were higher than at the soil donor site. Such increases have already been observed when soils are moved in the La Crau area (Coiffait-Gombault et al., 2012) and can be attributed to the calcareous matrix of the conglomerate breaking during soil gathering. Right from the first year, species-richness recovered to values similar to those of the steppe (Jaunatre et al., 2012), and did not significantly change subsequently. Moreover, the composition was close to that of the steppe, including many target species, as observed in previous soil transfers in other ecosystems (Bullock, 1998; Good et al., 1999; Box, 2003; Trueman et al., 2007). However, vegetation structure differed slightly from the steppe, especially in the second year, which was characterized by marked rainfalls. This structure change is reflected in the restoration indices: normalized Community Structure Integrity Index did not change over the three years, revealing stability in target species abundances, whereas Higher Abundance Index increased, revealing an increase in non-target abundances compared to the steppe. These higher abundances are again explained by the density of grasses, such as *A. barbata* and *Bromus* spp. The decrease in HAI in 2012 is a more positive result, especially if this tendency can be

maintained with grazing in the future (Gibson and Brown, 1992; Stroh et al., 2002). Several previous restoration projects using soil transfer failed to maintain high conservation value due to unsuitable management practices (Bullock, 1998; Box, 2003). The substantial difference we found between the low potential which might have been expected based on the seed bank study and the largely positive results obtained in the field vegetation with soil transfer has previously been observed, and was attributed either to differences in volume (Willems and Bik, 1998; Jaunatre et al., 2012) or to differences in environmental conditions (Stevenson et al., 1997). As soil transfer involves destroying a portion of the reference area, it cannot be considered as a substitute for in situ conservation (McLean, 2003), but only as a kind of ecological engineering technique allowing a soil resource already scheduled for destruction to be conserved for restoration purposes.

When grazing is used to rehabilitate land, controlling grazing pressure on a large scale is not an easy task. Usually, there is a trade-off between expected biomass production and the sheep breeders' ability to increase or decrease numbers of sheep. This trade-off does not always satisfy restoration objectives, especially in view of climate variability. Given the relatively high spring precipitation in 2010 (191% of average precipitation January–April) and 2011 (116%), the grazing pressure was relatively light for the first year (399 day sheep⁻¹ ha⁻¹) and increased in 2011 and 2012 (618 day sheep⁻¹ ha⁻¹). There was a significant difference in vegetation height and cover between the grazed and ungrazed control only in 2012 (results not shown), added to both higher grazing pressure and lower precipitation in spring (48% of average precipitation January–April). Grazing has been proved to be an efficient method of removing preferentially nutrient-rich biomass (Stroh et al., 2002; Jacquemyn et al., 2003). However, deposits of faeces may redistribute phosphorus (Henkin et al., 2010), so sheep should only be allowed to graze during the day and kept in an enclosure at night (Marrs, 1985), as at our study site. As this has been the traditional sheep-herding practice in the La Crau area for more than 2000 years (Badan et al., 1995; Fabre, 1997; Henry et al., 2010), this beneficial management approach will be applied long-term to the whole restored area (Meffre et al., 2011). The results obtained on vegetation argue for continuing high grazing pressure and thereby favoring target species establishment (Gibson and Brown, 1992).

These experiments in La Crau show that the current trend in ecological restoration towards using ecological engineering techniques leads to at least partial restoration of some ecosystem components. It also leads to enhanced biodiversity and habitat quality, compared to the former intensive cultivation or to natural resilience. Here, rehabilitation yielded a large area dominated by grasses and constituting a favorable habitat for numerous steppe birds (Wolff, 2011). Yet our findings confirm our hypothesis, namely that without changing abiotic conditions or strengthening target species dispersal, the vegetation remains different to that of the steppe. Contrary to what was hypothesized, however, even though some target species were successfully transferred by hay transfer, the number of target species did not increase by the end of the third year. This result highlights the fact that dispersal strengthening alone is not enough; it needs to be combined with suitable abiotic conditions. Our hypotheses concerning covering species seeding are confirmed: the density of non-target grasses is not very high after three years; therefore, this treatment seems to create a suitable area for target species colonization. In accordance with our hypotheses, the best results are obtained by topsoil removal, which allowed the restoration of abiotic conditions and thus the spontaneous settlement of some target species, and by soil transfer, which both partially restored abiotic conditions and strengthened target species dispersal, thus enabling species-richness and, to some extent, composition to be recovered.
Our findings are based on a short-term assessment of this eco-
logical engineering project. Given the relatively slow dynamics of
natural ecosystems, the three-year period considered here could
be viewed as far too short to draw conclusions as to the effi-
ciency of each method. However, in a world where more and more
restoration projects are being implemented to compensate for the
destruction of ecosystems, it is important to discuss the efficiency
of such methods at time scales similar to those involved in changes
in land-use. Moreover, it is important to assess the short-term
dynamics of such ecological engineering projects, so as to high-
light what is required in terms of management to reach restoration
objectives. Our relatively positive results were obtained at a high
economic and ecological cost and led to lower biodiversity com-
pared to the reference, which supports the findings of a review
of 89 restoration projects (Benayas et al., 2009). This mixed suc-
cess underlines the importance of in situ conservation of natural
habitats, rather than restoring them after their destruction.

Acknowledgements

This study was supported by CDC Biodiversité, CEEP Ecomusée de Crau, the Réserve Naturelle des Coussols de Crau, the Conseil Régional de Provence Alpes Côtes d'Azur, and CNRS RTP Ingécotech funding. We would like to thank Michael Paul and Marjorie Sweetko for checking the English. We are also grateful to the Société des Carrières de La Ménudelle for its collaboration and to students and field assistants for help with field work and seed bank study: Jean-François Alignan, Noëllie Fonvieille, Alexandre Fuentes, Laurie Gariglio and Daniel Pavon.

**Appendix A. Supplementary data**

Supplementary data associated with this article can be
found, in the online version, at http://dx.doi.org/10.1016/j.ecoleng.
2013.12.022.

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