Invasive species have been considered one of the main causes of habitat degradation ( Gurevitch & Padilla, 2004 ). This invasion is closely linked to global change and globalization ( Kaluza et al., 2010 ) and has had detrimental economic impacts on human enterprises caused by the spread of plant diseases and competition with agricultural species ( Shacklenton et al., 2007 ; Arévalo & Fernández-Palacios, 2005 ; Arévalo et al., 2017 ). Specific elements of globalization that explain the spread of invasive species around the planet are: global changes (e.g. global warming, nitrogen deposition, habitat fragmentation, Dukes & Mooney, 1999 ); socioeconomic changes, such as gross domestic product (GDP) in relation to global trade and intercontinental trade ( Lin et al. , 2007 ; Sharma et al., 2010 ; Arévalo et al., 2017 ); transport ( Westphal et al., 2008 ) and tourism ( Sutherst, 2000 ) among others. Indeed, globalization has been one of the most substantial drivers behind the homogenization of insular biotas ( Arévalo et al., 2010a ) and is linked to a worldwide alien species expansion ( Levine & D´Antonio, 2003 ; Westphal et al., 2008 ; Pysek et al., 2010 ). Ultimately, the consequences of invasive species affect different spheres such as the economy ( Pimentel et al., 2005 ), health ( Levine & D´Antonio, 2003 ) and ecological processes ( Reaser et al., 2007 ). The cost of their eradication is also increasing as many more species are introduced into new habitats. The risks of new diseases and problematic weeds for agriculture and how these introduced species disrupt ecological processes in the native plant community have been widely reported ( Pimentel et al., 2005 ).

On islands, infrastructure development is also directly related to the number of invasive species, whose impact can be either immediately evident or can lag for a period over 100 years ( Weber & Li, 2008 ). Moreover, the extinction of native island biota and dismantling of island ecosystems worldwide have been some of the consequences of the high numbers of non-native species introduced ( Donlan & Wilcox, 2008 ; Kueffer et al., 2010 ). Although we can consider non-native plants as important components of plant communities in many parts of the world ( Mack et al., 2000 ), they have also become a global problem that requires analysis and evaluation of the ecosystems under threat. In some cases, the impact can even reach levels of ecosystem modification that completely alter the plant community, as is the case of Myrica faya in Hawaii ( Lenz & Taylor, 2001 ) or Bromus tectorum in California ( DiTomaso, 2000 ), while in others they only act by basically occupying disturbed areas and being more competitive than native species ( Arévalo et al., 2010b ).

The objective of this study is to test the recovery of native plant communities under the closed canopies of exotic species. Our hypothesis is that species composition of the understory below the canopy of the exotic Quercus suber may be favored by the environmental conditions and will be dominated by the native plant community, the laurel forest, following a "catalytic" effect of this exotic stand on succession. This effect can be revealed at different levels, such as a increases in soil humidity, organic matter or shade that favor shade intolerant native species ( Arévalo et al., 2005 ), or other processes related to the provision of benign microhabitats that are more favorable for seed germination and/or seedling recruitment than their surrounding environment ( Harris & Harris, 1997 ; Lugo, 1997 ; Parotta, 1995 ; Parotta et al., 1997 ; Whitmore, 1999 ).

We also aim to put forward proposals to manage the forest following a landscape, ecological, forestry approach involving the manipulation of stands in order to help them recover from natural or anthropogenic disturbances such as afforestation with an exotic species ( Oliver & Larson, 1996 ).

The plots did not present substantial environmental differences, with similar altitudes, slopes, canopy cover and litter cover. Some variability can be considered part of the microsite heterogeneity in the area (Table 1). In the case of the soil nutrient composition, there was greater variability in the case of some nutrients. For example %OM presented a range of variations from 4.29 to 13.8 and Polsen varied from 6 to 40 ppm (Table 2).

The tree canopy composition was completely dominated by Quercus suber . In some plots, this was almost the only species (>99% total basal area of the plots QS2, 6, 7, 12 and 13). Even the plot with the lowest dominance in basal area of Q. suber reached almost 75% of the total (Average basal area of Q. suber of all the plots 93.1 ± 8.3 %; Table 3).

In the case of canopy regeneration, Q. suber also dominated in all plots with over 80% of density, with the exception of one plot (QS10). However, three more species were also present: Pinus canariensis , Laurus novocanariensis and Olea cerasiformis . In four plots, the woody regeneration of the plots was only Q. suber . The average regeneration of all plots for Q. suber was more than 92% including the non-dominant plot of the species (Fig. 2 represents the decimal logarithm of regeneration density in order to visualize species with low densities).

The plant community analysis of the trees and regeneration shows the species scores in Fig. 3. These species are enclosed in the same polygon depending on whether their values are from basal area species or sapling species. Discrimination among polygons (as long as this discrimination is related to differences in relative abundance in the plots) revealed differences in the regeneration community and canopy tree com-munity. As the analysis shows, the polygon enclosing the regeneration species is included in the polygon of the tree species. There was much more variability in the canopy tree community than in the regeneration one (as we have seen previously, with only four species). This indicates that species in the canopy will be partially replaced and no major changes are expected based on regeneration. We should note the important impact on the regeneration plant community of Olea cerasiformis , a species favored by the present environmental conditions created by the canopy of Q. suber with respect to the others present in the canopy.

A total of 93 species were found in the study (Table S1 [suppl.]), and there was some variation in the number of species among plots, from 16 (plot QS13) to 43 (plot QS8). However, Smith and Wilson evenness revealed high values on the plots and low differences among these (from 0.72 to 0.84).

The CCA for all the plant community revealed that the most significant environmental characteristics for the plant community (Fig. 4) were C/N, %Limestone, %Silt, NO ₃ and Na. There was also substantial vari-ability in the plant community, such as plots QS13, QS12 and QS11 had low levels C/N and limestone and were characterized by species such as Opuntia maxima , Erica arborea , Rubus ulmifolius of Hy-pericum canariense , whereas plots QS7, QS4 and QS5 had high values of NO ₃ and Na and were mainly characterized by species like Viburnum rigidum , Phagnalon purpurascens or Andryala integrifolia . Plots QS8 and QS6 had high C/N values and were characterized by species such as Carlina salicifolia , Teline microphylla or Castanea sativa (Fig. 4).

Common species in all the plots were located as the origin of coordinates such as Quercus suber , Laurus novocanariensis , Sonchus acaulis , Oxalis pes-caprae or Olea cerasiformis . With respect to native or invasive species, there was no distribution specific pattern on the defined bi-dimensional space of CCA. Some of the species were common for all the plots like Q. suber , Oxalis pes-caprae , and others were only found near to some plots such as Scorpiurus muricatus , Brunsvigia rosea (QS8) or Opuntia maxima (QS11 and QS12).

Conflicting evidence exists as to whether exotic monocultures exert a predominately negative or positive effect on the regeneration of the native forest beneath their canopies. In our case, the exotic canopy did not favor the recovery of the natural plant community. Exotic tree monocultures ad-versely affect native ecosystems species composition through competition with native species ( Attiwill & Leeper, 1987 ; Jurgensen et al., 1986 ; Fimbel & Fimbel, 1996 ). However, some studies in other areas have indicated that exotic tree plantations in the Canary Islands have facilitated a more rapid restoration of the native forest community, with a significant advance in regeneration of the canopy and establishment of native plant community species ( Arévalo & Fernández-Palacios, 2005 ; Arévalo et al., 2005 ; Arévalo et al., 2011 ). Such results also agree with several studies in which exotic species have facilitated forest succession to a native plant community in their understory on sites where dis-turbances have prevented recolonization by native forest species ( Parrotta, 1995 ; Fimbel & Fimbel, 1996 ; Loumetto & Huttel, 1997 ; Arévalo et al., 2005 ). In these cases, afforestation with exotic species has provided an improvement in environmental con-ditions favoring the establishment of the native shade-tolerant species. In addition, it has been shown that native forests are resilient to invasion by exotic species ( Guldenhuys, 1996 ).

In this study, we have identified a high number of species in the understory of a stand dominated by the exotic species Quercus suber , with one of the plots having 43 species (Table 2), showing high evenness in all the plots (almost over 0.80 in all the plots). However, the DCA analysis was unable to discriminate canopy tree plant community vs . sapling plant community. This means there were no differences at plant community level and the dominant canopy species persisted.

Regeneration has lower species richness than the canopy with 14 species for trees and only 4 in the case of sapling species. This is common in sub-tropical and tropical forests, where seedling dis-tribution is determined by the distribution of parent trees producing seeds, the seed rain, the presence of dispersers and predators and the availability of adequate sites for germination ( Schupp, 1995 ). Fu-rthermore, it is consistent with previous studies of laurel forest regeneration, in which human activities (wood extraction, agriculture, charcoal production, etc.) or natural disturbances have promoted the occurrence of pioneer species in the forest. Seedlings from these pioneer species only germinate under light conditions, so they cannot be found under a closed canopy ( Fernández-Palacios & Arévalo, 1998 ; Arévalo & Fernández-Palacios, 2007 ). So, in spite of such differences, canopy and sampling composition are characterized by the same dominant species: Quercus suber and Laurus novocanariensis favouring also Olea cersiformis .

Adult trees of pioneer species are likely remnants of early successional stages of the forest, mainly related to human management. These trees may persist within the forest for decades, Erica arborea , or other introduced species such as Pinus halepen-sis , Myoporum laetum of Eucalyptus camaldulensis among others, unable to establish under a close ca-nopy (Fig. 4). After several decades, the canopy composition cannot be totally explained by rege-neration, but the most dominant species will persist in the canopy. Pioneer species are now excluded from the regeneration, and they will need natural or anthropogenic disturbances to become established in order to persist in the present canopy species composition.

The understory community, with a high number of exotic species (around 20%; (Table S1 [suppl.])), is another indicator of the disturbance level of the area, with high variability depending on the microsite variability in nutrient composition and environmental conditions, with some species present in all the plots as Quercus suber or Oxalis pes-caprae (Fig. 3).

The laurel forest of the Canary Islands is the most emblematic community of the Canary Island Archipelago. Thus, restoration programs are being developed to increase the laurel forest area on the islands, such as on Tenerife. Today less than 1% of the original laurel forest on Gran Canary Island remains unaltered ( Santos, 1990 ). Based on these results, we suggest the management of the Quercus suber stand, since there is no evidence of a recovery of the native plant community or of a "catalytic" effect in this exotic stand.

Although the management of these areas is not an urgent task, we propose some management practices to favor the regeneration of native species. Treatments of thinning in the Quercus suber stand should be carried out, because plantations' closed canopies are preventing the establishment of shade-intolerant species ( Ashton et al., 1997 ). Other Quercus species have been identified as invasive ( Lemke et al., 2013 ), and will require some monitoring. The ability of the natural forest to dominate the stand should also be promoted with stand management activities, followed by enrichment with plantations of containerized individuals of laurel forest, especially the more heliophytic species ( Erica arborea , Morella faya and Ilex canariensis ). Continuous mon-itoring of the area will also be required to ensure the establishment of native species.

We thank the personnel and administration of the Servicio de Medio Ambiente of the Cabildo de Gran Canaria (Carlos Velázquez, Benjamín Artíles and Santiago Cabrera) for the support to conduct this study. In addition, we acknowledge the support given by the Agricultural Experimental Farm (Cabildo de Gran Canaria) with the soil analyses, we are grateful to the students of the Department of Geography (University of Las Palmas de Gran Canaria) for their help with the fieldwork and the associated editor and two anonymous reviewers for providing valuable comments that significantly improved the manuscript.