Robinia pseudoacacia L. (black locust) is a tree, but as a heliophilous and short-lived species, it is a weak competitor. This limitation is balanced by its easy and fast propagation (mainly through root suckers), tolerance of disturbance, rapid growth and tolerance of a wide range of habitats including extreme conditions. On the other hand, Robinia is robust and persistent, therefore it is able to persist in a site once colonized for several decades largely independent of the environment, which the tree itself modifies by changing the availability of nutrients in the soil and light conditions (Pyšek et al. 2012, Chytrý 2013, Vítková et al. 2015, Schiffleithner and Essl 2016).

Current landscape is characterized by habitat fragmentation which causes large areas of ecotones and boundary line stands, i.e. optimal conditions for Robinia. Serious large-scale disturbances (e.g. mining) provide a lot of open, well aerated and nutrient-rich substrata. Rotation of such disturbance events resulting in decades of successional development at abandoned sites enables Robinia to spread, establish and play a key role in succession. Moreover, transport of large volumes of soil containing Robinia propagules effectively compensates for the low ability of its large seeds to disperse over great distances.

Although most data comes from Central and Southern Europe, we considered for our assessment the whole of Europe (Table 1). Czechia (the Czech Republic) was used as the model area for the description of the management approaches as there is a lot of field data for this country (Vítková and Kolbek 2010, Vítková et al. 2015, 2016, 2017, our unpublished data) and Robinia is included in the Black List of IAS (Pergl et al. 2016c). We used also some data on the consequences of its planting from other parts of the world (e.g. China – Zhang 2014, Kou et al. 2016; Korea – Lee et al. 2004, Kolbek and Jarolímek 2008) to extend the applicability of suggested management strategies.

Information for our paper, illustrating the approach for a major IAS in our study area, was obtained from (i) more than 100 research articles and local papers referring or applicable mostly to European countries (Table 1), (ii) hundreds of phytosociological relevés of Robinia stands growing in Europe (Table 1), (iii) inquiries addressed to European specialists (see Acknowledgments) in nature conservation, invasion ecology and management of Robinia, (iv) tens of results of local projects (often unpublished) testing different methods of Robinia removal and aftercare (e.g. Halassy and Török 1996, Novák 2005, Böcker and Dirk 2007, Trylč 2007, Bogdan 2008, Bělař 2011, Bělohlávková 2014); (v) practical experience of Czech private companies and administrations of protected areas involved in Robinia management, including unpublished data (e.g. Čechová 1998, Veverková 2009), and (vi) our unpublished long-term research on the ecology and impact of Robinia stands in various European countries. Although it might seem that there is a great body of quantitative data on, e.g. yield, growing stock, forest regeneration or eradication, in fact the available information is surprisingly poor and rather gappy. Moreover, it does not allow for comparing among individual categories of Robinia stands in our model area of Czechia, and even less so in other European countries. The total growing stock and yield of both planted and spontaneous Robinia stands could be determined only on forest land belonging to the state (not private owners) in some countries. Robinia stands growing on non-forest land, such as on arable land, in parks, urban and mining areas are mostly planted for other purpose than economic profit, therefore both their extent and biological parameters are not known. Robinia stands growing in protected areas are usually only monitored in a preparatory phase for eradication. For these reasons, it is not possible to make a rigorous statistical analysis of our general model.

According to Dickie et al. (2014) we consider a dichotomy between positive and negative effects on ecosystem services resulting from planting of Robinia which currently causes conflicts of interest between different groups of stakeholders (e.g. nature conservation, forestry, urban landscaping, beekeepers and the public). These conflicts are often viewed only in a local context therefore we propose a complex management strategy on European level taking into account both economic benefits and environmental risks associated with Robinia cultivation (van Wilgen and Richardson 2014). Based on Holmes et al. (2008), Shafroth et al. (2008) and Gaertner et al. (2016), we suggest practical decision framework for sustainable Robinia management (Figure 1). Such framework has to be based on rigorous cost-benefit analysis (Naidoo et al. 2006, Hanley et al. 2009), leading to identification of potential conflicts. At first the potential threats associated with the presence of Robinia have to be identified, including threats resulting from inappropriate management of stands. If no conflict is identified, a standard management should continue (management of plantations, ornamental trees). In presence of any conflict the recommended management depends on the intensity of the threat ranging from slow conversion by succession to fast eradication. In addition, the decision scheme needs to be accompanied by categorization of stands with Robinia into eight groups (Table 2) reflecting the variation in habitat conditions and character of stands, in order to make context-dependent decisions relevant to local conditions. For each group, the distribution and source of Robinia, its history, ecological characteristics (habitat, structure, plant composition) and currently used management are summarized.

Figure 1.

Decision framework for selecting suitable Robinia management. Width of arrows indicate importance of the management. Shading indicates the number of potential sites covered (white – relatively few occurrences, black – most of the sites). Data come from the reviewed literature and project reports.

Based on links between ecological traits such as habitat, vegetation structure, origin, utilization, benefits and environmental risks we distinguish eight types of Robinia stands (Table 2). Each type includes four management practices, which are effective in various combinations depending on local conditions: (i) establishment of stands, (ii) maintenance of the existing state or utilization, (iii) tolerance of natural succession without major human interventions, and (iv) conversion, i.e. management or measures targeted at changing a stand into another unit or type of vegetation. The advantages and risks of particular management practices are discussed in the context of different initial conditions. Relations among the types of Robinia stands distinguished, successional trends and suitable management practices are shown in Figure 2.

Deep, well-aerated, nutrient-rich mesic soils in warm areas are optimal for the growth of Robinia since trees reach up to 35 m, form straight trunks and provide high-quality timber (Figure 3A). However, most Robinia forests are in dry habitats such as nutrient-poor sandy or rocky pastures on originally infertile soils threatened by soil erosion (Vadas 1914, Hegi 1924, Göhre 1952, Kolbek et al. 2004), where trees hardly reach 10 m and are often used for firewood or making poles (Vítková and Kolbek 2010, Vítková et al. 2015, 2017). In wooded areas, light-demanding Robinia does not spread into dense forests, but is able to colonize forest margins or disturbed sites, such as fresh clear-cuts or post-fire sites. Spontaneously, it spreads also into other open habitats in the vicinity, for example abandoned vineyards, orchards and fields.

Biodiversity value of such Robinia forests is mostly low, however, certain groups of organisms prefer them (e.g. macrofungi or habitat generalists among birds; Ślusarczyk 2012, Hanzelka and Reif 2015a, 2015b). The undergrowth is often dense and rich in species (~20 to 45 plant species/200 m², similar to that in climax forests), but it is dominated by widely distributed nitrophilous species sharing a wide range of nemoral and ruderal habitats, e.g. Bromus sterilis, Galium aparine, Urtica dioica, Hedera helix and Sambucus nigra. Species-poor Robinia forests growing in dry habitats are dominated by grasses, the dense cover of which may slow down the establishment of native trees.

Figure 2.

Main successional / intentional dynamic changes among the types of Robinia stands. Numbers of vegetation units correspond to stand categorization in the text.

Establishment and maintenance: Most European production of Robinia wood comes from these plantations. In the Pannonian basin in particular they are the main type of forest and their yield varies between 80 and 280 m³/ha and have an average rotation age of 30 years (Rédei et al. 2008). New stands are still being established, for example in Hungary, Italy and Romania (Rédei et al. 2008, 2012, Enescu and Dănescu 2013, Ciuvăt et al. 2015, Meloni et al. 2016) but not in Czechia, Poland and Switzerland (e.g. MZE 2014, Wojda et al. 2015). Producing saplings from seed is a relatively simple and low cost method, although germination must be facilitated by mechanical scarification (Rédei et al. 2012), soaking in concentrated sulphuric acid or boiling water (Huntley 1990). Propagation from root cuttings is suitable for producing articular clones or special cultivars (Keresztesi 1988, Rédei et al. 2012). Regeneration from root suckers produces a higher yield than from seedlings at a harvest age of 35–37 years. Robinia forests need more management than climax tree species (e.g. oaks), as without regular silvicultural treatments the quality of wood deteriorates due to an unshaped crown and deformed trunk (Bělař 2011, Rédei et al. 2012).

Tolerance of natural succession: Not remarkable due to economic value of these forests.

Conversion: Restoration of native vegetation is mostly not profitable, being costly and time-consuming. Because of the high sprouting ability of Robinia, it is very risky to stop eradication before totally removing all sprouts (Novák 2005, Pergl et al. 2016b, Vítková et al. 2017). There is nothing to be gained by restricting conventional silviculture, especially in early deforested lowlands or suburban zones where Robinia has been domesticated for a long time, forms extensive stable metapopulations and where native trees suitable for afforestation are lacking and there are no issues with nature conservation. However, establishment of new stands must be specially assessed, notably those to be established in close proximity of dry or mesic open natural habitats, due to the high sprouting ability of Robinia.

Admixture of Robinia with cover of up to 10% is the most common type of its occurrence (Vítková et al. 2017). It is a frequent spontaneous admixture in drier parts of hardwood floodplain forests on well-drained and fertile soils, mainly consisting of Quercus robur, Carpinus betulus, Ulmus minor and alien Ailanthus altissima (Essl et al. 2003, Terwei et al. 2016). Dry deforested slopes in Czechia were stabilized using Robinia and locally alien Pinus nigra at the turn of the 19^th century.

The environmental impact of Robinia growing in mixed stands is considerably less than in monocultures. In closed mature forests, it survives only as individual trees or groups of trees in areas that were previously disturbed. The composition of shaded undergrowth is dependent on the proportion of the canopy that consists of Robinia (Essl et al. 2011). Birds benefit from its presence in mixed forests up to approximately equal proportions between Robinia and native trees, but its higher share causes shifts in bird community compositions toward a dominance of generalist species at the expense of specialists. This invasive species affects birds by altering structural components of the habitat and related supply of food and cavities for hole-nesting birds (Kroftová and Reif 2017). Mixed Robinia forests occur mostly close to native forests and thus Robinia does not pose danger for local or surrounding vegetation.

Establishment and maintenance: Reasons for the establishment of these forests were either to supplement natural sparse stands, e.g. forest-steppes with Quercus spp. or to improve soil quality, yield and species diversity after logging of native forests and in inter-cropping plantations (Figure 3B; Groninger et al. 1997, Mosquera-Losada et al. 2012). Mixed forests with Robinia can be managed as a standard part of current silviculture if some conditions are fulfilled. It is important to reduce light availability inside the forest. Traditional management with regular clear-cuts recurring every 20–30 years creates sunny sites which are suitable for reproduction and vegetative regeneration of Robinia and thus drives its invasion into native deciduous forests (Radtke et al. 2013). Such invasion can be accompanied by spread of other weedy or invasive species. Natural disturbances forming light gaps in closed forest canopies, such as trees dying, fire or windthrow are other factors facilitating Robinia invasion as the species is highly adapted to disturbance. Under unfavourable light conditions, it develops a persistent bud bank on roots, stems and branches, allowing a rapid reaction to canopy opening following disturbance resulting in the establishment of compact clonal colonies covering areas up to several hundred square meters (Kowarik 1996, Chang et al. 1998, Krízsik and Körmöczi 2000, Schiffleithner and Essl 2016).

Tolerance of natural succession: Natural decline in Robinia abundance during succession was observed only in forests without large-scale disturbances, where Robinia finally occurs only as an admixture restricted to more open sites (Motta et al. 2009, Somodi et al. 2012, Terwei et al. 2013).

Conversion: Selective cutting that reduces light availability (Radtke et al. 2013) and favours native tree species is needed in such mixed forests. However, efforts to eradicate all Robinia trees would be fruitless because of economic demands and risk of failure.

Figure 3.

Closed forests in natural habitats (A–C) and small-scale stands in man-made habitats (D). A Robinia forest regenerating and managed by coppicing in stripes B Planted mixed forest with native Fraxinus excelsior and alien Robinia pseudoacacia C Old Robinia forest overgrown by Fraxinus excelsior and Acer platanoides as a result of spontaneous succession D A spontaneously established mixed stand with Robinia growing in a quarry.

Protective monodominant forests 12–16 m tall and over 50 years old on steep slopes pose a big problem in terms of their stability. Trees gradually die, are prone to windthrow and damage and the forest becomes more open. The shrub layer is rich in species. The herb layer consists of dominating grasses, relicts or pioneers of natural forest communities and nitrophytes with cover depending on water regime of topsoil. Such protective forests provide excellent honey (Vítková et al. 2017).

Establishment: In some countries (e.g. Czechia and Switzerland), this species was used to stabilize deforested steep eroded hillsides along rivers that were threatened by soil erosion (former pastures) and transport corridors (Vítková et al. 2017). Because of inaccessible terrain, old Robinia plantations have remained without management for several decades.

Maintenance: Maintenance or restoration with native species is mostly not profitable. Old trees are often unstable, therefore logging is difficult and risky, and profit is low. Moreover, logging may trigger soil erosion and regeneration of Robinia.

Tolerance of natural succession: During spontaneous succession, Robinia is replaced by shade-tolerant competitive trees such as Fraxinus excelsior, Acer pseudoplatanus, A. platanoides, A. campestre (Figure 3C) or tall shrubs such as Crataegus monogyna on dry sites (Vítková 2014). The rate of this succession is greater if populations of native competitors already occur in the understory or in the neighbourhood. Under closed canopies of native species, Robinia does not sprout spontaneously or only slightly (Vítková et al. 2016).

Conversion: Slow conversion to natural forest by means of natural succession is recommended, if there is no risk to biodiversity (adjacent natural habitats) or human infrastructure (traffic corridors or built-up sites; Pergl et al. 2016b). To prevent recovery of Robinia, it is important to avoid all interventions that induce sprouting, even leaving dead wood at a site after disturbance (e.g. wind break). If necessary to proceed faster, the natural step-wise canopy opening can be supported by killing of vital trees using combination of cutting and incomplete girdling deep into the phloem followed by application of herbicides (Böcker and Dirk 2007). Very slow decay of felled Robinia trunks (Schwarze 2007) may be utilized to stabilize slopes. However this is costly and time-consuming and should be used only when other methods or natural succession fail.

A common feature of this rather heterogeneous type is a ruderal environment in urban, agrarian, industrial or mining areas (Figure 3D), and a high proportion of aliens including cultivated ornamental woody species in the canopy or understory. The stands are widespread across Europe and differ in their origin (spontaneous vs. planted), structure (forest vs. shrubs or semi-open stands) and composition (pure or mixed stands with different types of undergrowth). Most stands are young with either prevailing isolated tree clumps or strips growing in the peripheries of towns and agrarian landscapes or larger disconnected groves in reclaimed mining areas.

Establishment and tolerance of natural succession: As early as in the 1970’s, Robinia was used for the biological recultivation of the post-mining landscapes and landfills (e.g. Bellon et al. 1977) as it is still used in many countries in Europe, South Korea and China (Kim and Lee 2005, Grünewald et al. 2009, Wang et al. 2012, Wojda et al. 2015). In mining areas e.g. in Poland, Germany and Czechia, Robinia forms planted or spontaneous stands with native pioneer species such as Betula pendula, Pinus sylvestris, or alien Populus hybrids. In urban areas, Robinia is at first cultivated, often escapes and overgrows wasteland and public greenery. These Robinia stands are accompanied by native nitrophilous trees such as Acer platanoides and Fraxinus excelsior, and many aliens such as Prunus cerasifera, Lycium barbarum and Parthenocissus quinquefolia. In agricultural Pannonian lowland, spontaneous and planted Robinia stands along roads are commonly admixed with thermophilous alien trees such as Ailanthus altissima, Gleditsia triacanthos, Celtis occidentalis and Morus alba.

Maintenance and conversion: Active management is needed since rapid spontaneous changes tend to occur in this habitat. Consideration of the local context (e.g. role of surroundings, ornamental or utility value, claims of owner or public) is necessary, especially in urban areas. Therefore, different parts of the same stand may be managed differently, including e.g., removal of Robinia or whole stands. However, there is no reason for eradicating or banning the planting of Robinia in urban areas (Sjöman et al. 2016). Some stands with alien species can even be developed within a novel system of urban nature (e.g. in Berlin; Kowarik and Langer 2005). Planting Robinia in mining areas does not pose a problem providing its dispersal does not threaten surrounding valuable habitats. Its gradual decrease during natural succession or mechanical control followed by conversion of stands to vegetation with native species is recommended.

Most of these stands originated from unsuccessful planting combined with spontaneous spread in dry habitats. Robinia survives in very dry habitats where it occurs as small and twisted trees (~5–10 m in height) or even shrubs forming sparse semi-open stands with an admixture of native xerophilous shrubs, e.g. Crataegus spp., Prunus spinosa and Rosa spp. This type is common in the Pannonian lowland (Hungary and adjacent parts of Austria, Czechia, Slovakia and Slovenia) and in Southern and South-eastern Europe.

Establishment and tolerance of natural succession: In some European countries (e.g. Slovakia, Slovenia, Italy), there is a long historical tradition in Robinia planting for vineyard poles and wine barrels (at least since the late 19^th century; Vítková et al. 2017). Such plantations have been established at sunny and dry sites of low quality – often low stony knolls surrounded by farmland, where ploughing of fields or mowing of meadows have prevented the vegetative spread and survival of Robinia seedlings and sprouts (Figure 4A). Slow growth and propagation of Robinia together with weak nitrification and low shading effect ensure the survival of these stands and of some plants, fungi, invertebrates and birds of sunny habitats (Vítková and Kolbek 2010, Ślusarczyk 2012, Hanzelka and Reif 2015b). Such stands form stable patches increasing the local biodiversity of deforested land; with some of them having over 60 species/200m². Some rare plant species are specifically linked to these stands, such as perennial grasses (Melica ciliata, M. transsilvanica), geophytes (Anthericum liliago, Ranunculus illyricus, many species of Allium, Gagea, Muscari, and Ornithogallum genera) and xerophilous herbaceous plants (Hesperis tristis, Verbascum phoeniceum). Despite high levels of potential nitrification, nitrophytes typical of Robinia stands occur only rare, probably due to drought (Vítková et al. 2015).

Maintenance and conversion: It should be left to the nature conservationists to decide whether to tolerate or remove these stands. However, most of these stands are very old and unlike those in mesic habitats, their shrubby growth does not indicate they are young plants with a potential for future growth, but are usually full-grown with their propagation greatly constrained by stress (Vítková et al. 2017). As in previous units, eradication of Robinia and restoration of native vegetation would be expensive and very risky. Monitoring succession and restricting spread into surrounding habitats, possibly combined with grazing or mowing seems to be the optimal management strategy.

Figure 4.

Non-forest habitats (A–C) and Robinia in urban environment (D). A Agrarian landscape with small-scale and semi-open Robinia stands. The spread of this species is suppressed by regular use of farming practices B Root suckers of Robinia invading a thermophilous grassland, which is the habitat of protected plant species C Intensive short rotation plantation regenerated by coppicing D Avenue of flowering Robinia in Prague (Czech Republic).

This type, which complements the previous one, refers to current invasion of natural habitats by Robinia (Figure 4B). Spontaneous occurrence of the young stages of Robinia poses serious threat to the conservation of dry to mesic grasslands and open dry forests as they are the habitats most endangered by this species invasion (Vítková et al. 2017).

Establishment and tolerance of natural succession: Compared to native trees, Robinia has a high sprouting ability and is extremely resistant to disturbance. It produces numerous root suckers that enable it to disperse at up to 1 m per year (Central Europe; Kowarik 1996) or 2 m per year (South Europe; Crosti et al. 2016) in non-forest ecosystems. Especially after disturbance of a tree its roots produce sprouts that grow up to 4m in height per year. On shading by Robinia the light regime, microclimate and soil conditions change and endangered light-demanding plants and invertebrates disappear (e.g. Kowarik 1994, Greimler and Tremetsberger 2001, Matus et al. 2003, Vítková and Kolbek 2010). Based on above mentioned reasons, it is not possible to tolerate establishment of Robinia plantations and their natural succession on vulnerable habitats, especially dry to mesic grasslands (including sandy steppes and rocky outcrops) and open dry forests as well as areas within a radius of 500 m from them (consistently with http://neobiota.bfn.de).

Maintenance and conversion: The spread should be restricted if Robinia stands occur in or adjacent to fallow land, grassland or other habitats with rare native plants, such as those on rocky slopes. The eradication should be rapid and persistent although expensive and risky due to use of herbicides and the disturbance causing vigorous regeneration of Robinia and erosion resulting in the release of nutrients and growth of weeds. For detailed list of suitable and unsuitable methods see (Silva et al. 2014, Schmiedel et al. 2015, Pergl et al. 2016b). However, no universally efficient and widely acceptable method seems to exist, because the stem- and root-sprouting ability of Robinia is affected by the eradication method as well as by local site conditions. Application of herbicides is necessary, otherwise resprouting of Robinia overcomes the effect of grazing or mowing and suckers appear even 30 years after the felling of Robinia (Trylč 2007).

Whole Robinia clones must be removed as the roots of the individual plants are connected. For quick eradication the best choice is felling followed immediately by spraying the area felled with herbicide. Removal by incomplete girdling (Böcker and Dirk 2007, Schiffleithner and Essl 2016), though demanding and time-consuming, is suitable for inaccessible sites. It is more efficient if combined with herbicide application at the end of summer, when assimilates are translocated to the roots. Elimination of new suckers and seedlings is necessary for at least 3–5 years. Well-proven is long-term grazing by goats once or twice a year, which also prevents the spread of tall weedy grasses. It is also best to remove all the Robinia biomass in order to prevent its sprouting and nutrient release. Due to the high dispersal rate of Robinia, control should also concentrate on populations adjacent to valuable habitats, at least to the distance of 500 m (consistently with http://neobiota.bfn.de).

Planting short-lived Robinia plantations for renewable bioenergy production (Figure 4C) is currently fashionable. Short-lived Robinia plantations occur in many countries worldwide, such as Albania, Austria, China, Italy, Germany, Greece, Hungary, Poland, Slovakia, Spain, South Korea and the United States (e.g. Grünewald et al. 2009, Rédei and Veperdi 2009, Stolarski et al. 2013, Zhang 2014, Straker et al. 2015). Other forms of utilization are rare, for example forage (Papanastasis et al. 1998). Energy production is profitable due to its high, early and easily produced dense, fast drying and combustible wood (Rédei et al. 2008). In the reclamation of heaps of industrial waste in post-mining landscapes one can add other benefits of using Robinia, such as high drought tolerance and ability to fix nitrogen (Grünewald et al. 2009).

Establishment and maintenance: These plantations should be established only in areas where an abundant metapopulation of Robinia already exists. The most common methods are either planting seedlings or rooted cuttings, however, a more environmental friendly and cheaper method is to transform Robinia forests at low quality sites (Rédei and Veperdi 2009). Because of its short coppicing period (average 4–5 years), Robinia grows to 5–6m in height (Rédei et al. 2010), nutrients in topsoil are depleted (Vasilopoulos et al. 2007) and undergrowth is species-poor and dominated by undemanding weeds. It is important to prevent further spread of Robinia (Crosti et al. 2016). Although closed forests are invasion-resistant, the establishment of new plantations in open land, especially at abandoned sites, close to roads or navigable rivers, is not recommended. As a barrier against Robinia invasion buffer zones of non-invasive plants (e.g. vineyards, orchards or fields) can be used, because periodic ploughing or harrowing suppress both the vegetative and generative reproduction of Robinia (Crosti et al. 2016).

Tolerance of natural succession and conversion: Extreme caution should be taken when such plantations are abandoned. There is a great risk of an intensive growth of suckers of Robinia, especially as the spontaneous succession of native vegetation is very slow. In northeastern Greece, succession to near natural riparian forest was not recorded even 14 years after abandonment. Site preparation for establishment of plantations as well as relatively low production of litter and periodic removal of organic matter through wood cutting caused a long-term changes in availability of soil nutrients and light, thereby affected species composition in behalf of ruderal species (Vasilopoulos et al. 2007). Another limitation often is a low pool of native trees in the vicinity and lack of serious natural enemies (Vítková et al. 2017). For successful conversion it is important to eliminate competition from Robinia and assist with reforestation using native tree species.

This type includes individual Robinia trees occurring solitarily or in groups in parks, gardens and at sites such as chapels or crossroads (Pergl et al. 2016d), furthermore in lines along roads, streets and rivers, in windbreaks, vineyard boundaries, hedgerows, gullies etc. (Figure 4D). Their function is mainly ornamental, together with protection against dust, noise or wind. Such structures are currently used to protect crops and livestock against weather extremes, for example in Hungary (Takács and Frank 2009). In Germany, “open orchards” consist of belts of vegetables or cereal fields separated by lines of fast-growing trees including Robinia, which are coppiced for biomass production and also used for improvement of soil quality and biodiversity (Mosquera-Losada et al. 2012, Medinski et al. 2014). As Robinia is a favourite horticultural tree, there are many interesting cultivars that are generally less invasive than the typical form (Hillier and Lancaster 2014).

Establishment, maintenance, tolerance of natural succession and conversion: Planting is usually easy. Trees need to be pruned and suckers removed regularly to prevent invasion into surrounding habitats. Consideration of the local context is necessary, Robinia should not be planted close to vulnerable natural habitats. Old trees are desirable because they provide shade and habitat for, e.g., rare saprophytic fungi or saprophagous beetles (Ślusarczyk 2012, Stejskal and Vávra 2013).