Numerous industrial processes disturb land of which the principal ones are mining, extraction of sand, gravel and clay, rock and limestone quarries, deposition of waste products including landfill sites, road and railway construction (Savill et al. 1997). The substrate to be reclaimed is almost always derived from mining or earth moving, and it is largely undeveloped subsoil or rock or it is polluted. The nature of reclaimed sites necessitates the use of species which are tolerant of exposure and undemanding nutritionally, characteristics often associated with pioneer species including alien trees (Savill et al. 1997). Non-native plants are widely used for revegetation in many parts of the world (D’Antonio and Meyerson 2002, Li 2006) if they fulfil a temporary successional role to colonize and ameliorate severely degraded sites and facilitate colonization and eventual dominance by native flora (Seo et al. 2008). Species with exceptional physiological tolerances are needed to improve site conditions and initiate soil-forming processes; species of Acacia, Alnus, Betula, Eucalyptus, Pinus, Salix and other pioneers are frequently employed for this purpose (Evans 2009a).

Short-rotation forestry is the practice of cultivating fast-growing trees that reach their economically optimum size between eight and 20 years old; each plant produces a single stem that is harvested at around 15 cm diameter. The crops tend to be grown on lower-grade agricultural land, previously forested land, or reclaimed land; they typically do not compete directly with food crops for the most productive agricultural land (McKay 2011). Fast-growing poplars and willows can be cultivated in short-rotation forestry (SRF) cycles of 15–18 years, but in short-rotation coppice (SRC) this is reduced further by cut-back/coppicing at 3–5-year intervals (Karp and Shield 2008).

Of the approximately 400 species of willows, the shrub willows (especially Salix viminalis in Europe) are deemed most suitable as bioenergy crops (Kuzovkina et al. 2008). Other species that are used include S. dasyclados, S. schwerinii, S. triandra, S. caprea, S. daphnoides and S. purpurea, and many clonal varieties are interspecific hybrids (e.g. S. schwerinii × S. viminalis; Karp et al. 2011, Raslavičius et al. 2013). Among poplar species, Populus nigra, P. alba and their hybrids (e.g., P. maximowiczii × P. nigra, P. maximowiczii × P. trichocarpa, P. trichocarpa × P. deltoides) are most suitable for bioenergy (Karp and Shield 2008, Faasch and Patenaude 2012). Many other alien species, including clones, hybrids and genetically modified trees, are used or are being tested for SRF/SRC, e.g., Robinia pseudoacacia in Albania, Italy, Germany, Hungary and Spain (Grünewald et al. 2009, González-García et al. 2011, Rédei et al. 2011a, Kellezi et al. 2012, Ciccarese et al. 2014), Acacia saligna in Israel (Eggleton et al. 2007), and Eucalyptus spp. in Portugal (Knapic et al. 2014) and in the UK (Evans 1980, Leslie et al. 2012, Keith et al. 2015).

The European Union has agreed to ambitious targets in terms of renewable energy that will probably promote a dramatic increase in the use of biofuels including short-rotation forestry and short-rotation coppice. This expansion and the continuous search for new species or genotypes may cause several direct and indirect undesired effects on biodiversity, including an increase in the introduction of additional invasive alien tree species into the region (Genovesi 2011).

Agroforestry systems include both traditional and modern land-use systems where trees are managed together with crops and/or animal production systems in agricultural settings. Agroforestry is practiced in both tropical and temperate regions, for both wood and non-wood products, including food and fibre for improved food and nutritional security (Jama and Zeila 2005). The potential of agroforestry to contribute to sustainable development has been recognized in international policies, including the United Nations Framework Convention on Climate Change (UNFCCC) and the Convention on Biological Diversity (CBD), justifying increased investment in its development (FAO 2013). Agroforestry (or “silvoarable agroforestry”) has traditionally formed important elements of European and Mediterranean landscapes, has the potential to contribute towards sustainable agriculture in Europe in the future, and it is supported by the Common Agricultural Policy (Eichhorn et al. 2006).

Nevertheless, many agroforestry systems, particularly those that depend on tree planting in or near treeless landscapes, rely heavily on alien plant taxa. As is the case in all endeavours based largely on non-native species, problems arise when these alien trees spread from sites of introduction and cultivation to invade areas where their presence is, for various reasons, deemed inappropriate. In some areas, problems caused by the spread of agroforestry trees from sites set aside for this land use pose a serious threat to biodiversity that may reduce or negate any biodiversity benefit of the agroforestry enterprise (Richardson et al. 2004).

Plantations in the Mediterranean have a long history. In mountainous areas, coniferous plantations were once limited to land at risk from erosion, but these now cover large areas of pastoral land and even agricultural land, either as a result of the establishment of plantations (e.g., Pinus nigra) or through colonization of abandoned land. Pinus radiata was planted in more than 300,000 has of old fields in Spain during the second half of the 20th century, mainly in Atlantic areas. More recently, the species has also been planted in acidic soils of the wet Mediterranean area in former agricultural lands with lime-free soils and annual rainfall exceeding 700 mm (Romanyà and Vallejo 2004). Plantations dominated by pines (Pinus halepensis, P. pinaster, P. pinea) are very common in coastal areas and are increasing in extent, despite an increase in major forest fires. Traditional forest activities (e.g., cork extraction, P. pinaster sawmills) have been replaced by multiple uses linked to tourism, hunting, and recreational activities (Etienne 2000).

In Turkey, afforestation with P. pinaster was undertaken by the French for the protection of sand dunes around Terkos Lake in 1880 (Deniz and Yildirim 2014). Italian foresters developed successful techniques for stabilizing sand dunes, and as a result of their efforts several thousand hectares of dunes were fixed and afforested in Italy in the 1940s with Pinus spp., Acacia spp. and Eucalyptus spp. (Messines 1952).

Diverse biotechnological methods are being intensively pursued to support plantation forestry with alien trees. These include clonal propagation (e.g., Rédei et al. 2002, 2011a, 2011b), interspecific hybridization, the use of a variety of molecular tools to intensify the selection of superior genotypes (DNA fingerprinting, genome mapping, gene identification and genome sequencing) and transformation (Grattapaglia and Kirst 2008, Strauss et al. 2009). However, of this diverse array of technologies, only transformation, defined by the use of direct modification and asexual insertion of DNA into organisms in the laboratory (that is, genetic engineering or modification), engenders attention from the Convention on Biological Diversity, strong government regulation and controversy over its use, even for research (Strauss et al. 2009).

Traits introduced to genetically modified (GM) trees include modification (quality and quantity) of lignin and cellulose composition, optimised biomass for biofuel production, resistance to pests and diseases, herbicide tolerance, altered growth and reproductive development, among others (Strauss et al. 2009). Hence, GM technology is clearly part of the toolbox for breeding of trees for agriculture and forestry use (Aguilera et al. 2013, Ledford 2014). Ecological risks associated with commercial release range from transgene escape and introgression into wild gene pools to the impact of transgene products on other organisms and ecosystem processes. Evaluation of those risks is confounded by the long life span of trees, and by limitations of extrapolating results from small-scale studies to larger-scale plantations (Frankenhuyzen and Beardmore 2004).

Many tree species are the focus of GM research. Frankenhuyzen and Beardmore (2004) identified 33 species of forest trees that had been successfully transformed and regenerated and additional species are reported by Häggman et al. (2013). Although most field trials have involved Populus spp. because of the status of poplar as a model organism for tree genomics and biotech (e.g., Jansson and Douglas 2007), and most have occurred in the United States, field tests have also been conducted in a number of other tree species and geographies around the world. In Europe 44 confined field trials for Populus spp. (30), Betula pendula (6), Eucalyptus spp. (4), Picea abies (2), Pinus sylvestris (2) have been approved (Council Directive 90/220/EEC of 23 April 1990, Strauss et al. 2009, Häggman et al. 2013).

Those engaged in the planted forest sector need to be aware of their obligations under regulations and legislation. The Regulation (EU) no. 1143/2014, the Plant Health Directive 2000/29/EC, the Wildlife Trade Regulations (338/97/EC and 1808/2001/EC) and the Habitats Directive (92/43/EEC) only apply to the 28 member countries of the European Union. Many other international conventions addressing issues of invasive alien species have been ratified by European and Mediterranean Countries (Shine 2007, Srivastava 2011, Table 2). At the national (or subnational) level, some countries have legislation and/or regulations aimed at preventing possession, transport, trade or release in the wild of specific invasive alien trees (Suppl. material 1). For example, in Norway, the 2005 white paper on the Government’s environmental policy and the state of the environment in Norway (Report No. 21 – 2004-2005 - to the Storting), the new Forestry Act (Act of 27 May 2005, no. 31, relating to forestry), the Nature Diversity Act (Act of 16 June 2009, no. 100), the Regulation on non-native trees (Regulation of 15 March 2013, no. 284), the national Strategy on Invasive Alien Species (published in May 2007) and the Norwegian Black List (Gederaas et al. 2012), are the main national specific documents referring to non-native trees. The Guidelines on trees, shrubs and plants for planting and landscaping in the Maltese Islands limit the use of alien trees in afforestation projects on agricultural land (MEPA 2002). The Iceland Forest Service has put forth a set of guidelines to afforestation planners: planting of aliens trees within natural woodlands is discouraged (Gunnarsson et al. 2005). Planting in treeless land must be carefully assessed considering the phenomenal and unique importance of the Icelandic breeding waterfowl populations which are at risk from the forestry. The Swedish Forestry Act placed restrictions on the planting programme of Pinus contorta in 1987, 1989 and 1991 due to extensive infection by Gremmeniella abietina in high elevation areas in northern Sweden after periods of extreme weather conditions from 1984 to 1987 (Karlman 2001).

Over 430 alien tree species worldwide are known to be invasive, and the list is growing as more tree species are moved around the world and become established in novel environments (Rejmánek and Richardson 2013, van Wilgen and Richardson 2014). Increasing awareness of problems associated with invasive forestry trees means that information on invasive species and ways of dealing with them is becoming more easily accessible - on the Internet, in scientific and popular publications, and via special interest groups. Ignorance is no longer an excuse for disseminating invasive alien trees (Richardson 2011). Global lists of invasive alien trees are available (Richardson and Rejmánek 2011, Rejmánek and Richardson 2013). “Invasive elsewhere” is one of the most robust predictors of invasiveness in trees, and there is strong evidence that species replicate invasive behaviour in environmentally-similar conditions in different parts of the world (Wilson et al. 2011).

The fact that some alien forestry trees have not yet spread from given planting sites should not be taken as evidence that invasions will not occur in the future. Experience with the same species in other parts of the world, including areas where the species have long residence times, should be evaluated to assess the extent of “invasion debt” (Richardson et al. 2015; Rouget et al. 2016).

Some countries have national or sub-national black lists (Suppl. material 1), identifying those alien species whose introduction is prohibited or discouraged due to their potential adverse effects on the environment or human, animal or plant health. Alien tree species that appear on black-lists should not be used for new plantations. An alternative approach used in other countries relies on a “white list” approach (or red, green and amber, see Perrings et al. 2005, Simberloff 2006) for identifying alien species that pose low invasion risk. Both listing systems have pros and cons (Simberloff 2006). For example, black-lists should only be considered as guides and one should not assume that non-listed alien tree species are safe. Additionally, in a huge country the translocation of a species from one part to another is just as likely to lead to invasions as are trans-continental introductions. For this reason, Notov et al. (2011) propose the adoption of three-level system of sub-national lists called “black books” for Russia.

Nevertheless, lists offer a useful approach for both companies and government agencies and could be used to fast-track approval of species or to reduce liability for forest owners when using low-risk non-native trees for plantations. Only in a few European countries are such lists supported by dedicated legislation (Essl et al. 2011); in other cases they are not legally binding even if scientifically sound, with priorities based on a rigorous risk assessment process. There are over 100 risk assessment models for invasive plant species (Leung et al. 2012), with some decision schemes developed specifically for trees or woody plants (Reichard and Hamilton 1997, Pheloung et al. 1999, Haysom and Murphy 2003, Widrlechner et al. 2004, Křivánek and Pyšek 2006, Gordon et al. 2011, 2012, Kumschick and Richardson 2013, Wilson et al. 2014). At the same time, only a few risk assessment methods are in line with the requirements of the Regulation (EU) No 1143/2014 (Roy et al. 2014).

The efficacy of any strategy to address invasive alien trees, including the capacity to produce reliable risk assessment reports (see principle 1.2), depends on the available information, and on the sharing of data, knowledge and experience. Information sharing systems would greatly improve the ability of authorities to prevent the introduction and spread of invasive tree species (e.g., Katsanevakis et al. 2014). Also, invasive species management requires specialist knowledge and skills which can only be developed over time. The capacity and awareness of land owners, forestry officials and other stakeholders are crucial for the effective implementation of the principles of the Code. There is a need to strengthen training institutions and to revisit the training curricula of forestry personnel and other stakeholders in silviculture, species and provenance identification, reduced impact logging, resource assessment, and in the management of both natural forests and non-native tree plantations.

The use of native species or non-invasive alien or less-invasive alien tree species as alternatives for highly invasive alien species in planted forest should be always considered (Richardson 1998, FAO 2010c, Gordon et al. 2012, Lorentz and Minogue 2015, Peltzer at al. 2015), as should the precise provenance of seeds and germplasm (Aarrestad et al. 2014). For example, Lorentz and Minogue (2015) remark that trait selection during breeding is potentially a very effective containment approach for managing Eucalyptus invasion risk. The likelihood of spread can be reduced by decreasing fecundity or by increasing the age to maturity, although the later method may negatively influence productivity (Gordon et al. 2012). This strategy has been successfully implemented in other taxonomic groups, including a triploid Leucaena hybrid in Hawaii (Richardson 1998). Likewise, elimination of seed production is thought to be a feasible goal for Eucalyptus (Gordon et al. 2012), and elimination of fertile pollen production has already been accomplished in the transgenic hybrid E. grandis × E. urophylla (AGEH427) (Hinchee et al. 2011). Ensuring containment of genetically modified trees through sterility could be significant because it eliminates the need for costly, uncertain and complex ecological research to understand and predict the impacts (FAO 2010d). However, the major limitation to this approach is that the permanence of containment technology is still uncertain (FAO 2010d, Lorentz and Minogue 2015). An additional obstacle to this solution is that FSC regulations currently expressly forbid any use of GM trees (Strauss et al. 2004, Brunner et al. 2007, Meirmans et al. 2010, Richardson 2011). In addition, some invasive alien tree species (Ailanthus altissima, Populus spp., Robinia pseudoacacia) also spread by vegetative propagation. Plantations of non-native species of Acacia, Eucalyptus and Pinus and have typically been relatively free of pest problems during the early years of establishment due to a separation from their natural enemies. This situation has however changed dramatically recently, as pests are accidentally introduced, but also as native organisms have started to infect and infest alien trees (Payn et al. 2015, Wingfield et al. 2015).

Best-practice methods relating to species and provenances of seed (Karlman 2001), seedling production, weed, pest and disease control should be adopted (FAO 2011). Weeds should be identified, recorded, and eradicated where possible, before planting. The EPPO standard PP 1/141 (3) describes the conduct of trials for the efficacy evaluation of herbicides in tree and shrub nurseries including nurseries within forest stands (EPPO 2009). Nurseries can act as important sources of alien species into plantation sites. Many forest pests, both insects and pathogens, have also entered new lands via nursery stock. Nurseries have a fundamental role in promoting the use of native trees, stocking suitable provenances, and proposing alternative native tree species in place of alien species (principle 2.1).

Containment of alien trees to areas set aside for their cultivation must become an integral part of silviculture and must be incorporated in best-management practice guidelines and certification schemes (e.g., Engelmark et al. 2001, Richardson and Rejmánek 2004, Richardson 2011, Dodet and Collet 2012, Felton et al. 2013). Silvicultural practices can either enhance or hamper biological invasions (e.g. Sitzia et al. 2016). Wingfield et al. (2015) have called for a global strategy to promote the health and sustainability of planted forests. Practices to reduce problems with invasive forestry trees need to be incorporated in such a strategy.

Decision-support schemes and research findings should be applied to identify the most appropriate sites for cultivation within landscapes; biodiversity issues and ecosystem services must be always considered in plantation design and site selection (e.g., Veldman et al. 2015). While some of these rules can be considered of general utility, some other good practices refer to specific alien tree species and aim to mitigate specific impacts, as in the case of the practices suggested by Finch and Szumelda (2007) for Douglas fir in temperate forests of Central and Western Europe, by Ledgard (2002) for the same species in New Zealand, by Engelmark et al. (2001) for lodgepole pine in Sweden, by Rejmánek and Richardson (2011), Calviño-Cancela and Rubido-Bará (2013), Lorentz and Minogue (2015) for Eucalyptus.

To avoid natural spread, eucalypts should not be planted near rivers and streams. Temporarily flooded or eroded banks are suitable habitats for spontaneous establishment of their seedlings. Moreover, their seeds can be dispersed over long distances by running water (Lorentz and Minogue 2015). Calviño-Cancela and Rubido-Bará (2013) suggest the establishment of a safety belt around eucalypt plantations in Spain to reduce eucalypt spread from plantations in the absence of fire. This measure would require the elimination of all newly recruited individuals in this safety belt (e.g. a 15-m wide belt could reduce the probability of eucalypt spread in more than 95%) before they mature and start producing their own seeds, thus hindering the advance of the front line of invasion. For this purpose, Calviño-Cancela and Rubido-Bará (2013) recommend interventions at 1-2-year intervals to uproot saplings and resprouts. Their results refer to a situation without fire. Fire stimulates regeneration (Gill 1997) and could increase dispersal distances, so that additional measures would probably be needed to control E. globulus spread after fires. In addition, Catry et al. (2015) suggest planting sterile Eucalyptus trees and prioritizing control in regions with the highest probabilities of recruitment.

In many cases, options exist for managing plantations of non-native trees and adjoining areas (invaded or potentially invasible) by manipulating disturbance regimes (e.g., fire cycles, grazing levels) to impede invasion (e.g. van Wilgen et al. 1994). The management of planted forests should also promote biodiversity (e.g., Zapponi et al. 2014), both within the planted forest itself and in areas of natural forest that are retained within the planted forest landscape (e.g. establish planted forests on degraded sites and retain areas of high biodiversity value protected) as recommended by the Secretariat of the Convention on Biological Diversity (2009). Managers can modify the silviculture of plantations in other ways to enhance diversity. For example, small variations in the timing and type of site preparation can affect the development and composition of the understory (Carnus et al. 2006).

Specific attention and management practices should be followed in the case of genetically modified tree plantations, such as hybrid or transgenic poplars and conifers (Engelmark et al. 2001, FAO 2006, 2010c, 2011, Brunner et al. 2007, Strauss et al. 2009, Di Fazio et al. 2012, Häggman et al. 2013). In Canada and many other countries, regulatory guidelines have been created regarding the introduction of such plants with novel traits (which in Canadian regulation includes alien species and transgenics; Bonfils 2006, Meirmans et al. 2010).

Forest plantation owners should be aware of those forestry activities that favour the spread of invasive alien tree species (Sitzia et al. 2016). For example, coppicing was found to be a driver of invasions by Ailanthus altissima and Robinia pseudoacacia in South Tyrol, Northern Italy. Radtke et al. (2013) concluded that currently applied coppice management, involving repeated clear cuttings every 20–30 years, favours the spread of both invasive tree species. They suggested an adaptation of the management system to avoid further invasion.

The risk of promoting the spread of fire-tolerant or pyrophytic alien trees must be taken into account when planning the use of prescribed burning in plantation forests. For example, the resprouting ability and pyrophytic seeds of Acacia dealbata allows this species to establish after fires in the northwestern Iberian Peninsula (Sanz Elorza et al. 2004, González-Muñoz et al. 2011). Maringer et al. (2012) describe the colonization of burned patches by Ailanthus altissima and Robinia pseudoacacia on the southern slopes of the Alps. Todorović et al. (2010) suggest that the post-fire invasive potential of Pauwlonia tomentosa can, at least partly, be explained at the germination level.

Finally, tailored management practices should be followed in plantations for bioenergy production (SRF/SRC) to ensure the careful choice of new planting sites for favouring biodiversity (Weih 2008, Framstad 2009), protecting hydrology (Christen and Dalgaard 2012), conserving landscape values and for the restoration of the site after the cultivation cycle (Hardcastle 2006, McKay 2011, Neary 2013, Caplat et al. 2014). In Austria 10 principles for short-rotation forestry systems, from the viewpoint of nature protection and environment, have been declared since 1998 (Trinkaus 1998). Principle 2 states that “ … Indigenous plants should play an important part, because non-indigenous plants (e.g., Robinia pseudoacacia and Ailanthus altissima) often show an undesirable tendency to spread”.

Harvesting activities such as road construction and movement of harvesting equipment are well known to disperse seeds or propagules of invasive species and to cause disturbances that help them to flourish (Kaplan et al. 2014).

Harvesting and transport of non-native trees should be planned, supervised and undertaken by appropriately trained personnel. Good practices should minimise the risk of further spread of invasive alien species, and the disturbance that could promote the establishment of other invaders. Careful planning will substantially reduce the road density required within a forest, the number of temporary timber extraction tracks, and minimise adverse environmental impacts such as soil disturbance, compaction and erosion. Whenever feasible, alien trees should be harvested individually or in small groups, to limit the risk of creating suitable habitats for other invaders.

Forest personnel should be trained to recognize and report unusual pests and symptoms of diseased or infested trees, and to carry out practices that reduce the risk of pest and weeds populations moving to other locations. Personnel should wear outer layers of clothing and footwear that are not “seed friendly” to minimise the risk of spreading alien species accidentally.

Specific guidelines for the restoration of sites previously occupied by plantations with alien trees need to be adopted. Restoration objectives can be broadly classified into overarching strategies, such as rehabilitation, reconstruction, reclamation, and replacement (see Stanturf et al. 2014). Only native plant species should be used for habitat restoration in areas affected by plantations. Native tree species can grow in the understory of alien tree plantations established for timber production or a variety of other forestry purposes. Not all alien tree plantations develop species-rich understories; some remain as tree monocultures. Low light intensity below the canopy, distance to seed sources, inhospitability to seed dispersers, poor soil or litter conditions for seed germination or seedling growth, intensive root competition with the planted alien species, chemical inhibition and other forms of allelopathy and plant interactions, plantation design, or periodic disturbances by organisms or any external factor are likely causes that require careful consideration (Lugo 1997).

Guidelines for restoration of sites previously occupied by plantations of Robinia pseudoacacia have been produced in the Piedmont region of Italy. Sturgess and Atkinson (1993) suggested management strategies for the restoration of near-natural sand-dune habitats following the clearfelling of Pinus plantations in Britain, and Brown et al. (2015) proposed approaches for plantations of alien conifers on ancient woodland sites. Szitár et al. (2014) assessed the recovery of open and closed grasslands over five years following the removal of alien pine plantations through burning at an inland sand dune system in Hungary. Arévalo and Fernández-Palacios (2005) proposed continuous elimination of Pinus radiata and enrichment with new individuals of P. canariensis on Tenerife, Canary Islands (Spain). Hughes (2003) and Moss and Monstadt (2008) propose management guidelines for the restoration of floodplain forests in Europe.

Early detection and initiation of management can make the difference between being able to employ feasible offensive strategies (eradication) and facing the necessity of retreating to a more expensive defensive strategy (mitigation, containment, etc.). Proactive measures to reduce the chances of invasions and to deal with problems at an early stage must be incorporated in standard silvicultural practices. Developing watch lists of possible new tree invaders can also enable more rapid reaction (Richardson 2011, Faulkner et al. 2014).

The relatively long initial lag phase between introduction and naturalization/invasion and slow dynamics observed in many forest plantation tree species compared with other plant species, offers opportunities to control the alien species while escaped populations are still small (Finnoff et al. 2007, Dodet and Collet 2012). Any signs of invasiveness reported inside the forest plantation or in its proximity should be carefully monitored so as to avoid serious problems developing.

Conifer wildings are relatively easy to control in the very early stage of invasion, as they are relatively easy to detect (most invasions are into grasslands and shrublands), and their direction of spread (downwind), and age when significant seed production begins (usually 10-15 years) is very predictable. There are therefore good opportunities to intercept the spread sequence very early in the cycle, and prevent wildings becoming dominant and uncontrollable outside the forest plantation (Froude 2011).

However, experience with introduced conifers in new environments indicates that spread events could begin at any time, even if little significant spread had been observed up to that time. Possible reasons could be synchronisation of all factors needed for successful spread (e.g. plentiful seed, low herbivores/ pathogens, good germination and seedling establishment conditions), arrival of suitable symbionts (notably mycorrhizae) to aid early establishment, and climatic change to conditions more suited to the planted alien trees (Despain 2001; Engelmark et al. 2001). Widespread natural establishment of Eucalyptus globulus plants in Portugal was recently documented by Águas et al. (2014) and Catry et al. (2015).

The idea of having a network of sentinel sites for monitoring or detecting biological changes or phenomena is not new and has been most widely applied to monitoring the spread of infectious diseases (e.g., Sserwanga et al. 2011, Vettraino et al. 2015). This approach has also been advocated for detecting the arrival or initiation of spread of alien species (Richardson and Rejmánek 2004, Meyerson and Mooney 2007) and a national system for detecting emerging plant invasions was proposed in the United States (Westbrooks 2003), but has yet to be implemented.

Plantations of alien trees should form part of any sentinel site network for monitoring alien tree invasions. Other areas that are likely to act as sources of propagules and sites of entry for new invasions are areas of human habitation where gardens have been established, especially where these adjoin natural vegetation (Alston and Richardson 2006), and experimental plantings, arboreta or botanical gardens containing alien tree species. Visser et al. (2014) have shown that Google Earth can be an useful tool for establishing a global sentinel site network for tree invasions, because imagery is continuously being updated, is free and low-tech. The wide availability of Google Earth could enable monitoring of this network of sentinel sites as part of ‘‘citizen science’’ efforts which could help to: (1) identify emerging trends in tree invasions; (2) provide valuable locality information for particular alien tree species; (3) monitor changes in alien tree species abundance and distribution over time; (4) help ensure legislative compliance of land managers and plantation owners; and (5) track management efforts over time (Visser et al. 2014). Besides such sentinel sites, new technologies such as smartphone application software (apps) are increasingly used to reach a wider audience on the subject of invasive alien species and to involve the public in recording them (Adriaens et al. 2015).

The general public is one of the most important stakeholder groups in national issues of forests and forestry (e.g., Hemström et al. 2014). The active and informed participation of communities and stakeholders affected by plantation forest management decisions is critical for the credibility and sustainability of management processes. Social learning (Leys and Vanclay 2011), public awareness-raising and communication activities are crucial for informing and educating the public, thereby allowing them to more effectively participate in decision making. Public participation GIS and related methods can be effectively used for decision-making processes related to planted forests (Brown et al. 2015). Public support for control efforts directed at invasive alien trees must be sought through carefully planned, long-term outreach initiatives involving, among other things, meetings with stakeholders, local village leadership, employment of villagers from areas adjacent to invaded sites, and the effective use of media outlets (Andreu et al. 2009, McNeely 2001, Marchante et al. 2010, Schreck Reis et al. 2011). Forestry has become more complex over the years. This form of land use now impacts on a wider stratum of people and environments than ever before, and is subject to many social and environmental demands.

Furthermore, an increasing number of tourists are interested not only in experiencing unique natural and cultural environments and forest landscapes but also in learning more about them. Forest-based tours are an ideal opportunity to share information about different types of forest environments, native and non-native tree species, restoration actions, wildlife and landscapes, and how they function.

Invasion biology is a complex multidisciplinary field and public and private plantations of alien trees are good places to conduct research on topics such as the spread, control, management and risks posed by invasive alien trees in collaboration with national or local environment agencies, research centres and appropriate regional or European bodies. Great Britain, for instance, with its long history of tree introductions and large plantings of many alien species (e.g. Picea sitchensis, the commonest British tree, Peterken 2001), is a good natural laboratory for studies of the determinants of naturalization and invasion in conifers and its consequences (Richardson and Rejmánek 2004). It would be very informative to revisit as many sites as possible in Europe where many alien tree species were planted long ago, e.g. the experimental plantings of many conifers in Italy (Nocentini 2010), Portugal and Spain, and abandoned plantations (Richardson and Rejmánek 2004). The exchange of information on the management experiences is another key aspect.

Forest management and conservation are expected to be strongly influenced by global change. Besides forest species, strategies and references for environmental management and conservation will be affected by global change trends (Jackson et al. 2005, Aitken et al. 2008, Canadell and Raupach 2008, Diaz et al. 2009, Heller and Zavaleta 2009, Thompson et al. 2009, Strassburg et al. 2010, Milad et al. 2013). For example, rapidly changing climate patterns, altered disturbance and nutrient regimes, and increased fragmentation are likely to favour the expansion of pine invasions worldwide (e.g., Higgins and Richardson 1999, Richardson and Rejmánek 2004).

Bernier and Schoene (2009) propose three possible approaches for adapting forests to climate change: no intervention, reactive adaptation and planned adaptation. Unfortunately, most current management belongs to the first or at best to the second category. No intervention means business as usual, with tree species and site selection, management targets and practices based on the premise that the planted forest will adapt more or less as it has in the past. Reactive adaptation is action taken after the fact. Planned adaptation, on the other hand, involves redefining planted forest goals and practices in advance in view of climate change-related risks and uncertainties.

In planted forest, climate change could affect the dynamics of alien tree invasions in many interacting ways, for example by: (a) causing modification in the native ecosystems, promoting range changes, naturalisation and spread of both native and alien trees (e.g., Iverson et al. 2008, McKenney et al. 2011); (b) favouring individual traits of particular alien trees (e.g. Capdevila-Argüelles and Zilletti 2008, Kawaletz et al. 2013, Castro-Díez et al. 2014); and (c) modifying introduction pathways and promoting increased use of certain alien tree taxa (Courbet et al. 2012, Lindenmayer et al. 2012), including a process of re-thinking the importance of the “always choosing native species” principle. Managed relocation has been proposed as a means of maintaining forest productivity, health, and ecosystem services under rapid climate change (e.g., Schwartz et al. 2012). Discussion is intensifying in many countries on whether and, if so, then to what extent, alien tree species should be used for afforestation, especially when native species are no longer able to fulfil essential forest functions. For example, in this regard, for the first time the growth potential of Cedrus libani was evaluated under climatic conditions in Central Europe (Bayreuth, Germany) by Messinger et al. (2015).

Finally, it is important to incorporate climate change into risk models for an anticipatory evaluation of scenarios for invasiveness of alien trees. Risk maps that incorporate the effects of climate change should help land managers and forest stakeholders with longer-term planning activities. Management plans of nature reserves should incorporate changes to invasion risk driven by global warming more explicitly. For example, Kleinbauer et al. (2010) suggest that the area suitable for invasions by Robinia pseudoacacia will increase considerably in Europe under a warmer climate. They argue that management plans for European nature reserves should incorporate such changes to invasion risk by species such as this one more explicitly. Reducing propagule pressure by avoiding plantings of R. pseudoacacia close to protected areas and sensitive habitats would be a simple way of reducing the risk of further invasions of this species under future climates. On the contrary, González-Muñoz et al. (2014) found no evidence that climate change will cause substantial changes to the invasion dynamics of A. dealbata in Spain.