Introduction
It was recognized as early as the 1990s that land protection is by no means a guarantee of the long-term species survival (Possiel et al. 1995; Maxted et al. 1997; Dopson et al. 1999). There is no doubt that legal protection is important in securing threatened species by protecting them from the immediate and most detrimental threats (e.g. logging, grazing, and poaching), but mere designation of protected areas, which has been and continues to be the primary approach to conserving biodiversity, is not enough (e.g. Brashares, 2001; Clark, 2013; Havens et al. 2014; Bridgewater 2016; Heywood 2016, 2017, 2018). There are several reasons why reliance on passive conservation through strict area protection that prohibits any modifications of the protected habitat to address threats to inhabiting its species is rather a dead end than the working strategy.
One reason why protection alone is the flawed strategy is the ongoing climate change. For many species with limited dispersal abilities, narrow environmental niches and populations scattered in the fragmented landscapes, even if all their populations are protected, the anticipated climate changes can drive them to extinction, unless some additional measures to adjust their ranges are taken. Another, even more important reason is the anthropogenic disturbance that left virtually no scrap of land (including those in protected areas) untouched, and disrupted previously existing species interactions and ecological processes (e.g., Chapman, 2010). Habitat fragmentation and deterioration that resulted from this disturbance reduced population sizes of many species below the viability threshold or made regeneration impossible. Last remaining individuals indicate that population recovery at that location is possible in principle, but only if the factors contributing to the population’s decline or failure to regenerate are identified and addressed. Elimination of these factors in the majority of situations requires interventions that will either maintain crucial ecosystems’ dynamic processes such as succession, or remove the dispersal and establishment limitations responsible for the reduced or absent recruitment.
Protection (i.e. legislating an area as protected by law) by definition does not assume a need in any modifications/alterations of the protected habitat. Therefore, passive protection alone cannot solve the problems outlined above. Unfortunately, the other conservation practices available, applied alone or together with protection, are equally unable to tackle this issue. A brief list of conservation practices available today includes:
- assessments of biodiversity summarized in IUCN species categorization and creating global, national and regional lists of threatened species;
- global and regional prioritization of species, habitats and areas for conservation;
- preservation of threatened species in ex situ seed banks and botanic garden living collections with minimal coordination between ex situ and in situ actions;
- reinforcement or reintroduction of endangered species usually conducted at single or very few locations;
- minor interventions in protected areas usually limited to control of invasive species and prescribed burning.
All these practices implicitly assume that the target habitat is intact or almost intact. If this was the case, protection with minor habitat interventions would do the job. Unfortunately, historically authentic, co-evolved biotic assemblages have largely disappeared being replaced by new combinations of species living under environmental conditions that have no historical analogs (Hobbs et al. 2006, Hobbs, 2013). In these new world realities, when virtually no habitat is intact, the listed above practices are of little use. They can become useful only if integrated into a new, creative and flexible approach able to deal with the altered habitats, and guarantee long-term species survival via restoring recruitment in existing populations and creating new viable populations. Given the lack of regeneration in many populations of threatened species, even in strictly protected areas, land protection must foster instead of forbidding the interventions that help restore species recruitment. This means that the new approach should be based on integration of conservation biology and restoration ecology.
A need for such integration has been recognized (Dobson et al. 1997; Young 2000; Burney and Burney 2007) but no attempt has been made to develop this general idea into a coherent concept.
Conservation biology focuses on processes that occur in populations of threatened species, while ecological restoration concentrates on community and ecosystem processes. The goal of the former is to ensure species persistence in their natural habitats, while the latter seeks to revitalize degraded ecosystems. However, to divert a threatened species from a path to extinction we need to identify and remove its threat, which almost certainly in any particular case is anthropogenic disturbance, be it logging, cash crop planting, invasion by alien species, changed hydrological regime, etc. Removal of these threats requires restoration of once existing conditions, and this is the field of ecological restoration. Thus, conservation biology and ecological restoration are inherently linked. Moreover, restoration of once existing conditions can be done not only FOR threatened species, but also USING threatened species when the latter satisfy certain requirements, as will be shown below.
Conservation-oriented restoration
I propose that efficient conservation of biodiversity can be achieved only by applying interventions to partly degraded habitats and that habitat restoration (instead of focal species and their populations) must become a focus of plant conservation. The concept based on these premises called conservation-oriented restoration adopts the idea of creating partly novel (i.e. having species compositions that differ from historical analogs) ecosystems, but with the goal of conservation of threatened species and their habitats, regardless of whether this will improve the ecological services for local human population or not. The concept is described in detail in (Volis, 2016, 2016a; 2018, 2019) and below I provide a summary of the distinct features of the proposed concept.
Threatened species can be used in restoration
An idea of usage of threatened species in restoration has the following logic. On one hand, the majority of the threatened species will have a future only in restored habitats. On the other hand, nowadays so many species are imperiled that it should be possible to choose species to be used in restoration plantings, which not only belong to the functionally important plant category, a category needed to restore the ecosystem integrity, but which are threatened themselves. The opponents of this idea may say that most of the threatened species have narrow ecological niche, poorly known biology and other features making them a problematic material for large-scale planting. This is (at least partly) true. However, many of the currently rare species are likely to be “anthropogenic rarities” (Fiedler and Laven 1996), i.e. species whose dramatic reduction in range and abundance, in comparison with other species, is due to higher vulnerability to alteration of once existing habitats and biotic interactions. Such species whose decline and range contraction are due to extrinsic factors (e.g. invasion of non-native species, livestock grazing, fire suppression or land conversion), may turn out to be useful for restoration of altered or partly degraded habitats both inside and outside their current range. Moreover, uncertainty about a cause of rareness regarding many threatened species paired with a high probability that they are anthropogenic rarities, should stimulate substantial broadening of lists of candidate species for habitat restoration with threatened species. If introduced into a variety of locations within their potential distribution range, they can become common or even dominant species in some of the restored ecosystems.
Comparisons of rare species with congeneric common species show similar fecundity and germination rates (Carlsen et al. 2002; Fu et al. 2009) as well as size classes distributions (Byers and Meagher 1997; Kelly et al. 2001). Establishment of introduced threatened plant species in restoration projects is similar to the establishment of non-threatened species (Morgan 1999; Shono et al. 2007; Cordell et al. 2008; Millet et al. 2013; Schneider et al. 2014; Subiakto et al. 2016). Potential limits to the utilization of threatened species in restoration projects are the requirement of large seed quantities as well as a lack of knowledge of the species’ reproductive biology and efficient methods of propagation and planting. However, the former problem of propagule supply can be efficiently solved by the quasi in situ living collections as explained below, and the crucial knowledge about threatened species propagation, although still very limited, is steadily accumulating (e.g. Iturriaga et al. 1994; Sakai et al. 2002; Danthu et al. 2008; De Motta 2010; Herranz et al. 2010; Kay et al. 2011; Ratnamhin et al. 2011; Koch and Kollmann 2012; Castellanos-Castro and Bonfil 2013; Gratzfeld et al. 2015; Lu et al. 2016). Once the necessary knowledge comparable to those for common species is acquired and protocols are available, the cost per seedling will make restoration practitioners more likely to incorporate rare and threatened species into their plans (Rodrigues et al. 2011).
Area prioritization should be based on presence of threatened species and degree of degradation
Hobbs et al. (2013) recognized three types of ecosystems: those remaining within their natural range of variability, those where anthropogenically caused changes are reversible, and those where such changes are irreversible (historical, hybrid, and novel ecosystems, respectively). Among these three ecosystem types, only hybrid ecosystems (beside historical ones) can be a home for threatened species, but not all hybrid ecosystems will have the same conservation value. The priority should be given to the least altered habitats, which still need interventions to restore altered structure and some missing ecological functions but which have a reasonable chance of approaching a once-existing habitat, and to those habitats in which threatened species still grow. I propose the following rankings of the areas targeted for conservation-oriented restoration by their conservation value:
- habitats in which highly endangered plant species still have populations and these populations exhibit natural regeneration;
- habitas in which highly endangered plant species still have populations but natural regeneration in these populations has not been observed or is depressed;
- habitats which are least degraded among other similar systems and which can potentially support establishment of endangered species currently not growing there;
- habitats of varying degree of disturbance that are located within protected areas or are important for their connectivity; and
- habitats of varying degree of disturbance that have a low probability of supporting establishment of endangered species but have a good chance of approaching (after restoration) historical habitats regarding species structure and composition.
Multiple stable states for an ecosystem can be targeted as reference conditions
Any restoration project requires reference conditions, which are used for the comparison with a contemporary ecosystem to evaluate the changes, design of the management actions, and measuring success of ecological restoration. A reference in a conservation-oriented project should not be a single ecosystem state but the historical range of variability in ecosystem composition, structure, and function. Such view is based on the assembly rules theory of theoretical community ecology according to which community assembly is deterministic in the composition of trait-based functional groups, but stochastic in terms of species composition. Besides, extant populations of threatened species usually are located in small size remnants of a natural habitat representing only a subset of the habitat’s original abiotic and biotic variation.
In working out a reference, it is important to take into account climatic fluctuations that occurred in the past, and especially important to consider climate changes that happen nowadays. This means that the restored ecosystems should be re-aligned with current and expected future conditions rather than with a single pre-disturbance state.
Choice of species should be based on “dark diversity” concept
According to Helm et al. (2015), the observed community diversity represents two species pools of different historical backgrounds. The first one, called characteristic diversity, consists of species that historically evolved in a region and represent a habitat-specific regional species pool, and the second one, called derived diversity, represents aliens whose presence is due to intended or unintended human impact. The regional species pool should be the major, and in many cases the only source of species to be introduced. However, in some cases, species from the second group can be selected, for example, if a functionally important species went extinct and needs a replacement by a functional analog not presented in the regional species pool; or when an endangered species has no suitable habitat in the whole region to which it belongs.
Compiling species lists for conservation-oriented restoration should adopt an idea of “dark diversity” (Partel et al. 2011). The species from the regional species pool absent in the characteristic diversity pool can be considered representatives of the “dark diversity” group, that is, the set of species in a region that currently do not inhabit a site due to dispersal or establishment limitations. Many of these species could have disappeared from a site due to human-caused alteration of abiotic and biotic conditions, or direct exploitation. Introduction of the species from the regional species pool can convert dark diversity into characteristic diversity.
Conservation interventions should be done in an experimental manner
To succeed, active interventions should be applied in an experimental manner rather than as a single “optimal” treatment for a number of reasons. For protected areas, to enable working out the efficient management scheme it would be desirable to apply a variety of experimental treatments among and within protected areas, so that some areas remain untouched while others are managed, and managed differently. A comparison of the outcomes will make it possible to identify the best treatment(s) to facilitate the transition of ecosystems along desired trajectories (Radeloff et al. 2015). For example, a site may include the creation of favorable micro-sites for the target species such as canopy gaps, deadwood, mounds or pits. Artificial creation of such micro-sites has a lot of uncertainty because the optimal levels of required intervention differ among species. Thus, the variety of disturbance types and their levels need to be tested for working out the optimal intervention required to create suitable conditions for the target species.
Experimentation is also needed for mitigating climate change effects. Potential species niche can be determined through species distribution modeling and used to predict the anticipated range shifts, but no modeling can predict the presence of suitable microsites, mutualistic biota or detrimental herbivores. Because responses to climate are usually species-specific, climatic changes will result in complex and difficult-to-predict novel species combinations. As a result of range shifts and competition, species with previously non-overlapping ranges will under new conditions reassemble into presently not existing communities and ecosystems (Williams and Jackson 2007; Hobbs et al. 2009; Gilman et al. 2010). While many of these new ecosystems will be unsuitable for imperiled species, limited scale translocation trials will help to identify those in which they may find a new home.
Conservation-oriented restoration projects should always be preceded by experiments investigating species- and treatment-specific responses. This can be done by applying mosaics of replicated treatments within mosaics of habitats (Howe and Martinez-Garza 2014). Modified in this way, the introduced micro-sites will differ in species composition, mostly in presence and abundance of rare species, and will serve as sources of colonization for each other. Thus, broadening the list of species introduced in different combinations and treatments (Howe and Martinez-Garza 2014) and replicating introduced populations over time and space (Guerrant 1996; Dani Sanchez et al. 2018) is a way to maximize the likelihood of reintroduction success in projects using threatened species because introductions of such species are often unsuccessful (Maunder 1992; Seddon et al. 2007; Godefroid et al. 2011; Dalrymple et al. 2012; Drayton and Primack 2012).
Ex situ and in situ approaches must be integrated
Restoration of a habitat may not be limited to, but as a rule includes introducing plant material, predominantly in a form of seedlings or saplings. Plant germplasm maintained and propagated ex situ can be used for this purpose but limitations of ex situ collections in botanical gardens and arboreta for producing outplants are well known (Simmons, 1976; Hamilton 1994; Schoen and Brown 2001; Maunder et al. 2004; Volis and Blecher 2010). A way to bridge ex situ and in situ to make the former a source of material for the latter is to create living collections of needed capacity outside botanic gardens and arboreta in natural or seminatural settings. Such quasi in situ collections (Volis and Blecher 2010; Volis 2016d), besides preserving species genetic variation, can be a reliable source of seeds for in situ conservation and restoration projects. Seed banks cannot fulfill this task due to space limitations and problems with storing non-orthodox seeds, while collecting large quantities of seeds in natural populations is either impossible or undesirable due to the negative impacts of seed harvesting on local population dynamics. For species whose seed output is low or varies greatly from year to year, or for species represented by small populations suffering from inbreeding depression, quasi in situ collections can be a solution because in these collections cross-pollination of plants originated from several populations will result in the production of healthy, well-performing offspring (Volis 2016c). These offspring can be produced with certainty, at no or very low cost, and collected in the large quantities required by nurseries producing seedlings of rare and threatened species.
Legislation must allow active interventions
Habitat protection is a vital component of the proposed strategy because it prohibits activities that can damage, destroy or modify the target habitats. However, the strictest protection does not guarantee a halt to further degradation of the habitat and species loss. This halt often is impossible to achieve without well-organized interventions and clear recovery criteria to follow. However, the interventions must be allowed by the protection status of the target site. Unfortunately, virtually all interventions which may require restoration of a habitat (e.g. introduction of a suite of functionally important for the ecosystem species, creation of deadwood, thinning of pioneer in favor of late-successional tree species, liberation of juveniles of threatened species from competing vegetation, and various forms of translocation) are not allowed in strictly protected areas (Categories I-II) (Dudley 2008).
To make implementation of conservation-oriented restoration possible, the current categories must be re-defined to permit i) management through active interventions while forbidding any unauthorized activities, and ii) introduction of critically endangered species based on predictions of species distribution modeling even if there are no records of their past occurrence.
Bibliography
Brashares, J.S., Arcese, P., Sam, M.K., 2001. Human demography and reserve size predict wildlife extinction in West Africa. Proceedings of the Royal Society of London, Series B 268, 1-6.
Bridgewater, P., 2016. The Anthropocene biosphere: do threatened species, Red Lists, and protected areas have a future role in nature conservation? Biodiversity and Conservation 25, 603-607.
Burney, D.A., Burney, L.P., 2007. Paleoecology and "inter-situ" restoration on Kauai, Hawaii. Frontiers in Ecology and Environment 5, 483-490.
Byers, D.L., Meagher, T.R., 1997. A comparison of demographic characteristics in a rare and a common species of Eupatorium. Ecological Applications 7, 519-530.
Carlsen, T.M., Espeland, E.K., Pavlik, B.M., 2002. Reproductive ecology and the persistence of an endangered plant. Biodiversity and Conservation 11, 1247-1268.
Castellanos-Castro, C., Bonfil, C., 2013. Propagation of three Bursera species from cuttings. Botanical Sciences 91, 217-224.
Chapman, C.A., Chapman, L.J., Jacob, A.L., Rothman, J.M., Omeja, P., Reyna-Hurtado, R., Hartter, J., Lawes, M.J., 2010. Tropical tree community shifts: Implications for wildlife conservation. Biological Conservation 143, 366-374.
Clark, N.E., Boakes, E.H., McGowan, P.J., Mace, G.M., Fuller, R.A., 2013. Protected areas in South Asia have not prevented habitat loss: a study using historical models of land-use change. PLoS ONE 8, e65298.
Cordell, S., McClellan, M., Yarber Carter, Y., Hadway, L.J., 2008. Towards restoration of Hawaiian tropical dry forests: the Kaupulehu outplanting programme. Pacific Conservation Biology 14, 279-284.
Dalrymple, S.E., Banks, E., Stewart, G.B., Pullin, A.S., 2012. A meta-analysis of threatened plant reintroductions from across the globe, in Plant reintroduction in a changing climate: promises and perils. eds J. Maschinski, E.H. Haskins, pp. 31-50. Island Press, Washington, DC.
Dani Sanchez, M., Manco, B.N., Blaise, J., Corcoran, M., Hamilton, M.A., 2019. Conserving and restoring the Caicos pine forests – the first decade. Plant Diversity 41, 75-83.
Danthu, P., Ramaroson, N., Rambeloarisoa, G., 2008. Seasonal dependence of rooting success in cuttings from natural forest trees in Madagascar. Agroforestry Systems 73, 47-53.
De Motta, M.J., 2010. A history of Hawaiian plant propagation. Sibbaldia 8, 31-43.
Dobson, A.P., Bradshaw, A.D., Baker, A.J.M., 1997. Hopes for the future: Restoration ecology and conservation biology. Science 277, 515-522.
Dopson, S.R., de Lange, P.J., Ogle, C.C., Rance, B.D., Courtney, S.P., Molloy, J., 1999. The conservation requirements of New Zealand’s nationally threatened vascular plants. Threatened Species Occasional Publication No. 13. New Zealand Biodiversity Recovery Unit Department of Conservation, Wellington.
Drayton, B., Primack, R.B., 2012. Success rates for reintroductions of eight perennial plant species after 15 years. Restoration Ecology 20, 299-303.
Dudley, N. ed., 2008. Guidelines for applying protected area management categories. IUCN, Gland, Switzerland.
Fiedler, P.L., Laven, R.D., 1996. Selecting reintroduction sites, in Restoring Diversity: strategies for reintroduction of endangered plants. eds D.A. Falk, C.I. Millar, M. Olwell, pp. 157-170. Island Press, Washington, DC.
Fu, Y., Chen, S., Wu, T., Hao, J., Sima, Y., 2009. Comparison of seed modality and germination characteristics of compare of Manglietia hookeri and endangered M. grandis. Journal of Northwest Forestry College 24, 33-37.
Gilman, S.E., Urban, M.C., Tewksbury, J., Gilchrist, G.W., Holt, R.D., 2010. A framework for community interactions under climate change. Trends in Ecology & Evolution 25, 325-331.
Godefroid, S., Piazza, C., Rossi, G., Buord, S., Stevens, A.D., Aguraiuja, R., Cowell, C., Weekley, C.W., Vogg, G., Iriondo, J.M., Johnson, I., Dixon, B., Gordon, D.R., Magnanon, S., Valentin, B., Bjureke, K., Koopman, R., Vicens, M., Virevaire, M., Vanderborght, T., 2011. How successful are plant species reintroductions? Biological Conservation 144, 672-682.
Gratzfeld, J., Kozlowski, G., Fazan, L., Buord, S., Garfi, G., Pasta, S., Gotsiou, P., Fournaraki, C., Dimitriou, D., 2015. Whither rare relict trees in climate of rapid change? BGjournal 12, 21-25.
Guerrant, E.O.J., 1996. Designing populations: demographic, genetic and horticultural dimensions, in Restoring diversity. Strategies for reintroduction of endangered plants. eds D.A. Falk, C.I. Millar, M. Olwell, pp. 171-207. Island Press, Washington, DC.
Hamilton, M.B., 1994. Ex situ conservation of wild plant species: time to reassess the genetic assumptions and implications of seed banks. Conservation Biology 8, 39-49.
Havens, K., Kramer, A.T., Guerrant, E.O., 2014. Getting plant conservation right (or not): the case of the United States. International Journal of Plant Sciences 175, 3-10.
Helm, A., Zobel, M., Moles, A.T., Szava-Kovats, R., Pärtel, M., 2015. Characteristic and derived diversity: implementing the species pool concept to quantify conservation condition of habitats. Diversity and Distributions 21, 711-721.
Herranz, J.M., Ferrandis, P., Martinez-Duro, E., 2010. Seed germination ecology of the threatened endemic Iberian Delphinium fissum subsp. sordidum (Ranunculaceae). Plant Ecology 211, 89-106.
Heywood, V.H., 2016. In situ conservation of plant species – an unattainable goal? Israel Journal of Plant Sciences 63, 211-231.
Heywood, V.H., 2017. Plant conservation in the Anthropocene – challenges and future prospects. Plant Diversity 39, 314-330.
Heywood, V.H., 2019. Conserving plants within and beyond protected areas – still problematic and future uncertain. Plant Diversity 41, 36-49.
Hobbs, R.J., Arico, S., Aronson, J., Baron, J.S., Bridgewater, P., Cramer, V.A., Epstein, P.R., Ewel, J.J., Klink, C.A., Lugo, A.E., Norton, D., Ojima, D., Richardson, D.M., Sanderson, E.W., Valladares, F., Vila, M., Zamora, R., Zobel, M., 2006. Novel ecosystems: theoretical and management aspects of the new ecological world order. Global Ecology and Biogeography 15, 1-7.
Hobbs, R.J., Higgs, E., Harris, J.A., 2009. Novel ecosystems: implications for conservation and restoration. Trends in Ecology & Evolution 24, 599-605.
Hobbs, R.J., Higgs, E.S., Hall, C.M. eds., 2013. Novel ecosystems: intervening in the new ecological world order. Wiley-Blackwell, Oxford, UK.
Howe, H.F., Martinez-Garza, C., 2014. Restoration as experiment. Botanical Sciences 92, 459-468.
Iturriaga, L., Jordan, M., Roveraro, C., Goreux, A., 1994. In vitro culture of Sophora toromiro (Papilionaceae), an endangered species. Plant Cell Tissue and Organ Culture 37, 201-204.
Kay, J., Strader, A.A., Murphy, V., Nghiem-Phu, L., Calonje, M., Griffith, M.P., 2011. Palma corcho: A case study in botanic garden conservation horticulture and economics. Horttechnology 21, 474-481.
Kelly, C.K., Smith, H.B., Buckley, Y.M., Carter, R., Franco, M., Johnson, W., Jones, T., May, B., Ishiwara, R.P., Perez-Jimenez, A., Magallanes, A.S., Steers, H., Waterman, C., 2001. Investigations in commonness and rarity: a comparative analysis of co-occurring, congeneric Mexican trees. Ecology Letters 4, 618-627.
Koch, C., Kollmann, J., 2012. Clonal re-introduction of endangered plant species: The case of German false tamarisk in pre-Alpine rivers. Environmental Management 50, 217-225.
Lu, Y., Ranjitkar, S., Xu, J.-C., Ou, X.-K., Zhou, Y.-Z., Ye, J.-F., Wu, X.-F., Weyerhaeuser, H., He, J., 2016. Propagation of native tree species to restore subtropical evergreen broad-leaved forests in SW China. Forests 7, 12.
Maunder, M., 1992. Plant reintroduction – an overview. Biodiversity and Conservation 1, 51-61.
Maunder, M., Hughes, C., Hawkins, J.A., Culham, A., 2004. Hybridization in ex situ plant collections: conservation concerns, liabilities, and opportunities, in Ex situ plant conservation: Supporting species survival in the wild. eds E.O.J. Guerrant, K. Havens, M. Maunder, pp. 325-364. Island Press Washington, DC.
Maxted, N., Ford-Lloyd, B.V., Hawkes, J.G., 1997. Complementary conservation strategies, in Plant genetic conservation, the in situ approach. eds N. Maxted, B.V. Ford-Lloyd, J.G. Hawkes. Chapman and Hall, London.
Millet, J., Tran, N., Ngoc, N.V., Thi, T.T., Prat, D., 2013. Enrichment planting of native species for biodiversity conservation in a logged tree plantation in Vietnam. New Forests 44, 369-383.
Morgan, J.W., 1999. Have tubestock plantings successfully established populations of rare grassland species into reintroduction sites in western Victoria? Biological Conservation 89, 235-243.
Partel, M., Szava-Kovats, R., Zobel, M., 2011. Dark diversity: shedding light on absent species. Trends in Ecology & Evolution 26, 124-128.
Possiel, W., Saunier, R.E., Meganck, R.A., 1995. Chapter 2 – in-situ conservation of biodiversity, in Conservation of biodiversity and the new regional planning. eds R.E. Saunier, R.A. Meganck. Organization of American States and the IUCN – The World Conservation Union. Department of Regional Development and Environment Executive Secretariat for Economic and Social Affairs General Secretariat, Organization of American States.
Radeloff, V.C., Williams, J.W., Bateman, B.L., Burke, K.D., Carter, S.K., Childress, E.S., Cromwell, K.J., Gratton, C., Hasley, A.O., Kraemer, B.M., Latzka, A.W., Marin-Spiotta, E., Meine, C.D., Munoz, S.E., Neeson, T.M., Pidgeon, A.M., Rissman, A.R., Rivera, R.J., Szymanski, L.M., Usinowicz, J., 2015. The rise of novelty in ecosystems. Ecological Applications 25, 2051-2068.
Ratnamhin, A., Elliott, S., Wangpakapattanawong, P., 2011. Vegetative propagation of rare tree species for forest restoration. Chiang Mai Journal of Science 38, 306-310.
Rodrigues, R.R., Gandolfi, S., Nave, A.G., Aronson, J., Barreto, T.E., Vidal, C.Y., Brancalion, P.H.S., 2011. Large-scale ecological restoration of high-diversity tropical forests in SE Brazil. Forest Ecology and Management 261, 1605-1613.
Sakai, C., Subiakto, A., Nuroniah, H.S., Kamata, N., Nakamura, K., 2002. Mass propagation method from the cutting of three dipterocarp species. Journal of Forest Research 7, 73-80.
Schneider, T., Ashton, M.S., Montagnini, F., Milan, P.P., 2014. Growth performance of sixty tree species in smallholder reforestation trials on Leyte, Philippines. New Forests 45, 83-96.
Schoen, D.J., Brown, A.D.H., 2001. The conservation of wild plant species in seed banks. BioScience 51, 960-966.
Seddon, P.J., Armstrong, D.P., Maloney, R.F., 2007. Developing the science of reintroduction biology. Conservation Biology 21, 303-312.
Shono, K., Davies, S.J., Chua, Y.K., 2007. Performance of 45 native tree species on degraded lands in Singapore. Journal of Tropical Forest Science 19, 25-34.
Simmons, J.B., Beyer, R.I., Brandham, P.E., Lucas, G.L., Parry, V.T.H. eds., 1976. Conservation of Threatened Plants. Plenum Pub Corp, New York and London.
Subiakto, A., Rachmat, H.H., Sakai, C., 2016. Choosing native tree species for establishing man-made forest: A new perspective for sustainable forest management in changing world. Biodiversitas Journal of Biological Diversity 17.
Volis, S., 2016a. Conservation-oriented restoration – how to make it a success? Israel Journal of Plant Sciences 63, 276-296.
Volis, S., 2016b. Conservation meets restoration – rescuing threatened plant species by restoring their environments and restoring environments using threatened plant species. Israel Journal of Plant Sciences 63, 262-275.
Volis, S., 2016c. How to conserve threatened Chinese species with extremely small populations? Plant Diversity 38, 53-62.
Volis, S., 2016d. Species-targeted plant conservation: time for conceptual integration. Israel Journal of Plant Sciences 63, 232-249.
Volis, S., 2018. Plant conservation in the Anthropocene: Definitely not win–win but maybe not lose–lose?, in Reference Module in Earth Systems and Environmental Sciences. eds D.A. DellaSala, M.I. Goldstein, pp. 389-397. Elsevier, Oxford.
Volis, S., 2019. Conservation-oriented restoration – a two for one method to restore both threatened species and their habitats. Plant Diversity 41, 50-58.
Volis, S., Blecher, M., 2010. Quasi in situ – a bridge between ex situ and in situ conservation of plants. Biodiversity and Conservation 19, 2441-2454.
Williams, J.W., Jackson, S.T., 2007. Novel climates, no-analog communities, and ecological surprises. Frontiers in Ecology and Environment 5, 475-482.
Young, T.P., 2000. Restoration ecology and conservation biology. Biological Conservation 92, 73-83.