Impact of Fraxinus excelsior dieback on biota of ash-associated lichen epiphytes at the landscape and community level

The dieback of European ash Fraxinus excelsior has been reported since the early 1990s. It started in north-eastern Poland and during the next years expanded to neighboring countries (Przybył 2002; Bakys et al. 2009a; Pautasso et al. 2013; Gross et al. 2014a). Recently, a severe dieback of ash, which includes trees in all age stages (Gross et al. 2014a; Needham et al. 2016; Enderle et al. 2018), can be observed in most European countries (Bakys et al. 2009a, b; Kowalski and Holdenrieder 2009a, b; Allen et al. 2010; Beck et al. 2016; Coker et al. 2019). It is caused mainly by the fungus Hymenoscyphus fraxineus, which induces wilting and loss of leaves, thinning of crowns, darkening of wood, necrosis and neoplastic changes on trunks and branches, and finally leads to the deaths of trees (Baral et al. 2014; Gross et al. 2014b), but other pathogens that weaken this tree are also reported, e.g. fungus Rigidoporus ulmarius (Lücking et al. 2019). Regardless the region, the massive extinction of ash has contributed to the loss of even more than 90% of the trees in the entire range of F. excelsior (Skovsgaard et al. 2010; Jönsson and Thor 2012; Cholewińska et al. 2018; Matisone et al. 2018). The susceptibility of ash to this disease is rising with the increasing intensity of soil water relations disturbances, linked with the water shortage caused by climate change, i.e. the increase in temperature and decrease in precipitation (Allen et al. 2010; Baral et al. 2014; Gross et al. 2014b). Therefore, due to the progressive climate change, ash recovery will be difficult and even impossible in some cases (Enderle et al. 2018). Despite the fact that some populations of ash may be resistant to infections by H. fraxineus (Enderle et al. 2015; Pušpure et al. 2017), which creates an opportunity for ash recovery, the observed regeneration of ash has not brought the expected results (Enderle et al. 2018).

The natural geographical range of European ash covers temperate lowland Europe from the Atlantic coast to the Volga River, except for most of the Mediterranean and northern Fennoscandia (Beck et al. 2016). It occurs in numerous types of forest communities as an admixture, or it can be the dominant tree species in floodplain and ravine forests, or forests growing on dry calcareous soils (Beck et al. 2016). The extinction of ash, especially in the phytocoenoses where it dominates, may result in transforming the functioning of entire forest ecosystems. It may lead, e.g. to increasing the availability of light due to gaps developing in the forest canopy and an increase in the amount of some substrates, especially deadwood. High deadwood availability drives changes in the composition of species in the forest understory, i.e. by creating a suitable microclimate and habitat conditions for the development of numerous groups of organisms, as well as increased nutrient availability in the soil (Eschtruth et al. 2006; Kurz et al. 2008; Ford et al. 2012; Mitchell et al. 2014a, b, 2016). However, the most important question connected with the ash dieback is how it will affect organisms closely associated with this tree, such as invertebrates, bryophytes, fungi, and lichen epiphytes (Jönsson and Thor 2012; Mitchell et al. 2014a, b, 2016), as F. excelsior is a foundation species for them. Growing in floodplain and ravine forests in high abundance, ash shapes the functioning and dynamics of these ecosystems and is considered to be a hotspot of biodiversity (Ellison et al. 2005). It is also a source of microhabitats (Barkman 1958) for many specialized groups of organisms, especially epiphytes (e.g. Lõhmus and Runnel 2014; Mitchell et al. 2014a; Preikša et al. 2015). Ash in primeval forests reaches its maximum age and size, and provides decaying wood after its death, which shapes unique habitats important for species conservation and preservation of the full biodiversity of organisms associated with them (Mitchell et al. 2014b; Preikša et al. 2015).

Epiphytic lichens are extremely vulnerable to any change in an environment. Apart from climate change and eutrophication, which are pointed out as main drivers of the changes in species composition of epiphytes (van Herk 2009), other important factors potentially causing these changes are those linked with shifts in the composition of tree phorophytes in forest communities (Pipp et al. 2001). However, the consequences of the ash dieback on the diversity of epiphytic lichen biota still have not been well documented. Jönsson and Thor (2012) reported that ash dieback could cause a 38% reduction in the number of ash-associated lichen epiphytes in forests in Sweden. Similar projected losses of species diversity of epiphytic lichens occurring on ash have been reported from Great Britain (Mitchell et al. 2014a, 2016). However, epiphytic lichens are not restricted to only one species of tree, but to the group of tree species with a similar complex of physical and chemical properties of their bark (Barkman 1958). Therefore lichens associated with ash may find a suitable substrate on other species of trees, such as oak, hazel, and sycamore maple, which was reported for all kinds of forests in Great Britain by Mitchell et al. (2014a, 2016). However, similar studies have not been conducted so far in old natural forests excluded from forest management and with very modest direct human influence. The only example of these ecosystems in temperate lowland Europe is Białowieża Primeval Forest, which plays an important role in the conservation and maintenance of diversity of lichen epiphytes, especially those occurring with high abundances on ash (Cieśliński et al. 1995). Therefore, taking into account the close to 98% loss of ash trees in Białowieża Forest stands (Cholewinska et al. 2018), there is an urgent need to determine the impact of the ash dieback on the preservation of the epiphytic lichen biota associated with this tree, with particular emphasis on forest communities where ash predominates.

The main aim of our work was to estimate whether the ash dieback is driving the loss of lichen epiphytes diversity amongst forest communities not affected by humans, on the same scale as was reported in Northern and Western Europe, and what other phorophyte tree species can be considered as important alternative hosts for lichens associated with ash. We hypothesized that the ash dieback effect on ash-associated lichen biota diversity would depend on the spatial scale considered: (i) moderate at the landscape scale (of the entire forest ecosystem), due to the high diversity of alternative host species, and (ii) severe at the community scale (of the ash-dominated floodplain forest) due to the absence of alternative tree hosts in the closest neighborhood.

On the landscape scale, NMDS ordination, performed for all plots containing and not containing ash, revealed a compositional continuum (high compositional similarities) of lichen biota regardless of forest community. Only plots not containing ash and representing coniferous forest showed relatively large compositional dissimilarities with other forest communities, i.e. both containing and not-containing ash (Fig. 2a). Considering results of NMDS performed only for plots containing ash, compositional dissimilarities between plots representing different forest communities were relatively low (Fig. 2b). Species composition of most plots representing a mixed deciduous fertile oak-lime-hornbeam forest was characterized by a high proportion of lichen species with higher values of the WEIV-T, WEIV-R, and Pielou evenness index (right side of NMDS plot; Fig. 2b; Table 1). Species composition of most of floodplain and swamp forest plots (left side of NMDS plot on Fig. 2b) was described by high values of WEIV-L, whereas the highest values of species richness and Shannon–Wiener index were related with some plots representing floodplain forest. Distribution of plots not containing ash revealed partial compositional discontinuum of the lichen biota across forest vegetation types, pointing out large compositional dissimilarities between most of plots representing mixed deciduous and coniferous forest (Fig. 2c). Only a few plots, representing swamp, mixed deciduous, and coniferous forest were overlapping in the middle part of the NMDS plot. Species composition of most mixed deciduous forest plots was described by high values of WEIV-T, WEIV-F, WEIV-R, and by high Pielou evenness index. Considering vectors representing species richness and Shannon–Wiener index, relationships with the same plots were less pronounced (Fig. 2c; Table 1). In contrast, species composition of coniferous forests was characterized by low values of all the above-mentioned vectors and high values of WEIV-L.

The mean Bray–Curtis compositional dissimilarity index of lichen biota was significantly higher for plots with the presence of ash than for those without this tree species (Fig. 3). Plots containing ash were characterized also by significantly higher species richness of lichens and a higher proportion of lichen species with higher light requirements (Fig. 3). On the plots without ash trees, Pielou evenness and Shannon diversity of the lichen biota were higher and the proportion of lichen species with higher temperature, higher moisture demands, and lower acidity requirements were higher than on the plots containing ash (Fig. 3).

At the tree phorophyte level we revealed that communities of lichen epiphytes growing on the bark of oak, lime, ash, and hornbeam were characterized by high values of WEIV-F, species richness, and Shannon diversity (Fig. 4a; Table 1). Considering results of hierarchical clustering, species composition of epiphytic lichens growing on ash was the most similar to those occurring on: alder, hazel, hornbeam, lime and oak (Fig. 4b), with \(\bar{x}\) ranging from 0.49 for ash and oak, and 0.61 for ash and hornbeam (Fig. 4c). We revealed the lowest compositional dissimilarities between lichen communities growing on alder and lime (\(\bar{x}\) = 0.31), as well as alder and oak (\(\bar{x}\) = 0.42). At the same time lime showed low dissimilarities of lichen epiphytes assemblages with: hazel (\(\bar{x}\) = 0.41), birch and ash (for both tress \(\bar{x}\) = 0.50), as well as hornbeam (\(\bar{x}\)= 0.51; Fig. 4c). Of the 172 epiphytes growing on the bark of ash, 152 (about 90% of the total biota) were hosted by alternative tree phorophytes (tree species with the highest lichen biotas similarity to the biota of ash), with the highest number (101 species; about 60% of the total biota) of ash-associated lichens found for oak and hornbeam (Fig. 5). A slightly lower number of lichen species typically occurring on ash was found on lime, alder, and hazel (99, 97 and 71 epiphytes, respectively). Few epiphytic species were found on aspen, birch, and spruce.

At the community scale, the lichen biota of ash and alder differed in the presence and share of individual species. Out of 71 epiphytes analysed, 44 species showed significant differences in frequency between the two tree phorophytes considered (Fig. 6). We found 23 species (about 50% of the 44 discussed species) significantly associated to ash (e.g. Alyxoria varia, Inoderma byssaceum and Pertusaria coccodes), including 13 species which occurred on this tree exclusively (e.g. Acrocordia gemmata, Lepra albescens and Lobaria pulmonaria). Twenty-one species revealed significantly higher frequencies on alder than on ash (e.g.: Arthothelium ruanum, Fuscidea arboricola and Violella fucata). This group included four species, which grew on this phorophyte exclusively: Cladonia digitata, Arthonia spadicea, Felipes leucopellaeus and Chaenotheca ferruginea.

The European ash is characterized by a very high species richness of lichen epiphytes (Cieśliński et al. 1995). From all lichen species occurring in the area studied (about 320 species; see Łubek et al. 2018), about 50% were recorded on the bark of this tree; thus, it is an important substrate providing appropriate conditions for lichens in maintaining the entire diversity of epiphytes. Our analyses revealed that even the complete extinction of ash will not pose a serious threat to the lichen biota associated with it. At the landscape scale, approximately 90% of epiphytic lichen biota connected with ash has the chance to survive on alternative tree hosts. Fifteen species occurring exclusively on ash (Supplementary Table 1) and belonging to 10% of species expected not to survive in the landscape scale, were mostly lichens of small thallus size, e.g. Bacidia biatorina, Candelariella efflorescens, Catillaria fungoides, Gyalecta flotowii, G. ulmi, Halecania viridescens, Psoroglaena abscondita. It could happen, that they were overlooked on other phorophyte species in investigated area. Among these lichens were both: species known as common in Poland (Fałtynowicz 2003) and in north-eastern Poland (Cieśliński 2003), e.g. Peltigera canina, P. polydactylon, Phaeophyscia orbicularis, as well as very rare and associated with well-preserved forests (Cieśliński 2003; Kukwa et al. 2017), e.g. Arthothelium spectabile, Catillaria fungoides, Gyalecta flotowii, G. ulmi, Schismatomma pericleum. This group of species, with the exception of Bacidia biatorina, Candelariella efflorescens, Catillaria fungoides and Phaeophyscia orbicularis, is exclusively associated with the floodplain forest community in our investigated area, thus their chances for survival are very low, no matter what spatial scale is applied.

Our analyses distinguished two distinct functional groups of ash-associated epiphytes. The first included species of higher light requirements (Fig. 2; compare Fig. 3), typical to a floodplain and to coniferous forests overexposed to light, which could be caused by natural windthrows and tree diebacks driven by insect outbreaks or fungal pathogens. In plots where ash did not occur, such species found suitable habitat conditions in coniferous communities, where they occupied mainly the branches of trees and deadwood in overexposed places (e.g. species from genera Cladonia, Hypogymnia, and Usnea). The second group included species of higher temperature requirements. These epiphytes (e.g. Thelotrema lepadinum, Ropalospora viridis, or some Pertusaria and Pyrenula species) were mostly linked to a mixed deciduous fertile oak-hornbeam forest. They were not associated strictly with ash (on which they occurred sporadically), but more often with the barks of hornbeam, lime, and hazel, which dominated in the mixed deciduous forest. These forests provide more stable temperature conditions and lower temperature fluctuations in comparison to more open floodplain forests occupying land depressions (Olszewski 1986). Due to that also epiphytes containing Trentepohlia algae as a photobiont (e.g. Arthonia, Pyrenula, and Opegrapha species), occur in the mixed deciduous forest as they have particular adaptation to environments with higher temperatures and higher humidity levels (Hametner et al. 2014). The species of the second group were also characterized by higher moisture requirements, related to stable humidity in the fertile oak-lime-hornbeam forest, as well as the lower acidity of occupied habitats (Figs. 2, 3). The presence of ash increased the species richness of the plots and their variability, which was reflected by the higher mean dissimilarity of their lichen biota, in comparison to plots without ash (Fig. 3). Epiphytes biota on plots with ash were also characterized by lower Pielou evenness and Shannon diversity, indicating the high importance of ash especially for very rare species.

Several tree species may play the role of alternative hosts for ash-associated lichens at the landscape level. The most important were oak and hornbeam. Each of these two tree species hosted close to 60% of the epiphytic lichens associated with ash (Fig. 5); however, these were not the same species groups and composition, and the share of lichen species on oak was more similar to that occurring on ash than to the species inhabiting hornbeam (Fig. 4b). Oaks and hornbeams were reported from the same site by Cieśliński et al. (1995) as the phorophytes hosting the highest number of lichen species. The importance of these trees for epiphytes is related to their high abundance in various forest communities: from shady humid oak-lime-hornbeam deciduous forest and moist oak-spruce mixed forest to the more overexposed pine-oak mixed forest. This wide habitat tolerance of oak and hornbeam, in combination with a wide spectrum of different microhabitat conditions, is suitable for hosting a high diversity of lichen epiphytes of different ecological requirements. The presence of several alternative tree species may solve the potential threat of a serious decline in lichen epiphytic biodiversity, but no single alternative tree can replace the ash as a phorophyte (Mitchell et al. 2014a). High similarity of ash and oak lichen biotas was reported from Great Britain by Ellis et al. (2013), which is congruent with our results (Fig. 4b). The next most similar phorophyte species in Great Britain were: beech, elm and sycamore, trees. At the landscape level of our study, aside from oak and hornbeam, lime, alder, and hazel also played an important role as substitute hosts for ash-associated lichens, often not occurring on the two formerly mentioned phorophyte species. Due to differences in species composition of lichens occupying different host tree species, nearly 90% of epiphyte species associated with ash are able to survive on them in the landscape scale context. Additionally, diseased or dead ash trees can still keep typical epiphytic lichens for some years, increasing the availability of their propagules and allowing them to spread. In such cases the persistence of lichen propagules in the environment increases the probability of their survival (Eaton et al. 2017; Ronnås et al. 2017; Wills et al., 2018). The extent and number of possible epiphytic lichen propagule sources within the landscape scale can prevent the species loss in the biota and facilitate their colonization on other microhabitats (Bouchard and Boudreault 2015). In this way, existing propagule sources of epiphytic lichen species maintain their dynamic and balance between mortality i.e. the extinction of local populations, and reproduction i.e. the recolonization of new microhabitats (Hanski 1998; Ruete et al. 2018), thus stabilizing entire metapopulations. This may be important for the restoration of ash-associated lichen biota after the regeneration of ash population, which recently almost entirely retreated from the forests. The reconstruction of the earlier full epiphytic biota on new generations of ash may be possible due to their refuges: i.e. alternative hosts and alternative microhabitats like dead wood. Thus, the dieback of ash at the landscape scale, where ash occurs as an admixture tree in various communities, is not problematic for the survival of most epiphytic lichens, if the forest is large enough to maintain its ecological continuity and natural regeneration processes (Angelstam and Kuuluvainen 2004; Shorohova et al. 2009). The biota of epiphytes needs not only a specific tree but also other species of phorophytes, specific microhabitats, and suitable space to maintain its continuity.

At the community scale, the dieback of ash could pose a higher threat to the ash-associated biota of epiphytes. Only about 50% of the species hosted by ash were recorded on the bark of alder (Fig. 6), the second main tree building floodplain alder-ash forests. Remaining lichen species may not find, in the limits of this community, trees with suitable microhabitats. Moreover, ash dieback at the community level could be a threat especially to very rare lichens, e.g. Opegrapha vermicellifera, Inoderma byssaceum and Alyxoria varia, realizing narrow niches and occurring within small populations (Jönsson and Thor 2012; Lõhmus and Runnel 2014). Similarly, the lack of alternative host trees for other species, e.g. Lepra albescens, Lobaria pulmonaria, and Peltigera praetextata, may reduce the chances for their survival at the community scale of floodplain forest. In our study, the bark of lying dead ash was an important substrate for some epiphytes that had previously grown on living ashes. In such a case, if the ash population will not recover, with progressive wood decomposition and the change in microhabitat conditions, they will not have any further prospects for survival (the “extinction debt” sensu Tilman et al. 1994). At the community scale, a specific ecological trap is created (Schlaepfer et al. 2002; Hale et al. 2015; Hale and Swearer 2016) due to the lower diversity of tree phorophytes, especially for lichens that exclusively or more often choose ash as a substrate for their occurrence. Thus, when there is a lack of ash phorophytes offering high-quality habitat conditions, ash-dependent lichens may colonize alders, which may be treated here as lower-quality phorophytes due to their different bark structure in comparison with ash (Barkman 1958). This may lead to lowered growth of lichen thalli, as has been recorded in Estonian forests (Lõhmus and Runnel 2014), their rapid die-off, or a complete lack of growth. Floodplain communities composed of alder can be a sink habitat for some epiphyte species, due to the limited possibilities of lichen migration to other alternative host trees.

Some authors (e.g. Ellison et al. 2005; Lõhmus and Runnel 2014; Brzeziecki et al. 2016) have suggested that changes in forest structure, caused by the massive dieback of one foundation tree species, like ash, result in irreversible changes in the species diversity of some groups of organisms, e.g. insects, bryophytes, lichens, and fungi, whose occurrence depends on the presence of this tree. Such a scenario is possible on a fine scale of the fragmented or isolated community patches composed of several species of trees, such as swamp or floodplain forests in a transformed landscape. However, such a scenario will not take place in extensive forest areas characterized by high naturalness and a large diversity of microhabitats, e.g.: trees of diverse species in age (from youngest to oldest) and deadwood in different stages of decomposition. These forests have specific dynamics, where natural disturbances like fungus or insect caused dieback of trees, are part of it. Even more, the dying of trees that dominate the canopy creates gaps, delivers dead wood, and allows forest regeneration, resulting in the development of communities of organisms dependent on this disturbance-caused habitat conditions (Kimmins 2003; Gilliam 2007; Stokland et al. 2012; Götmark 2013). Such significant changes of forest structure caused by host-specific pathogens that target a dominant forest tree species, have also been reported for several other species: oak, elm, pine, spruce, and alder (Fajvan and Wood 1996; Peterken and Mountford 1998; Kurz et al. 2008; Hicke et al. 2012; Bjelke et al. 2016). The decline in their populations causes an increase in the abundance of other tree species, which at least partly replace them as phorophytes. For instance, the loss of ash in British woodlands has increased the sycamore population (Needham et al. 2016). The same authors concluded that the long-term population dynamics of European ash are shaped by short-term pathogen-induced mortality, and even after reduction of the ash population by 75% it recovers to near pre-disturbance level—if the proportion of resistant individuals is high enough or a significant share of the population escape infection (Needham et al. 2016; Coker et al. 2019). In Białowieża National Park, a fungal pathogen caused the reduction of the ash population by over 90% during less than two decades (Cholewińska et al. 2018). The same authors have documented also the natural regeneration of ash; however, they observed that even young individuals were affected by the pathogen as well, which may pose a serious threat to ash-associated organisms. In other sites where ash regeneration is observed, e.g.: in Latvia (Pušpure et al. 2017) and South-Western Germany (Enderle et al. 2018), young trees seemed to be resistant to Hymenoscyphus fraxineus infection. If the ash in Białowieża Forest regenerate, there is a chance that the new generation will be resistant as well, and the epiphytes growing on this tree will have a chance to restore due to source propagules occurring on alternative tree phorophytes: mainly the oak and hornbeam—hot spot trees for epiphytic lichens, as well as other alternative host trees like lime and, in the longer perspective, by alder and hazel.

The extinction of ash at the scale of a floodplain streamside alder-ash forest community may result in the reduction of species diversity of epiphytic lichen associated with this phorophyte. Only about 50% of species were able to survive on the bark of alder—the second dominant tree species of the community. Forest communities composed of a few species of trees, such as floodplain forests, despite century-long strict protection and specific occurrence within an extensive forest complex, are less resistant to disturbances and thus are exposed to the risk of loss of species diversity.

The preservation of extensive forest areas, with high species diversity of trees and a high diversity of communities increases the chances for preserving species diversity of epiphytic lichens. The greater the phorophyte diversity, the greater the microhabitat diversity for lichens, and the greater the chance of survival of all species, especially those sensitive and associated with specific microhabitats.

Our results highlight the importance of a spatial scale in biological conservation. The protection of large, well preserved forest areas with a rich diversity of phorophyte trees as source habitats for epiphytic lichens will allow the continuity of lichen epiphytes metapopulations to remain dynamic and preserve their species diversity. However, the results obtained in our study should be referred with a caution in relation to other forest areas of different geographical location due to their different and specific regional climatic conditions significantly affecting species diversity of lichens.